Advances in MARINE BIOLOGY VOLUME 37
This Page Intentionally Left Blank
Advances in MARINE BIOLOGY Edited by
A. J. SOUTHWARD Marine Biological Association, The Laboratoty, Citadel Hill, Plymouth, England, UK
I? A. TYLER School of Ocean and Earth Science, University of Southampton, England, UK
and
C . M. YOUNG Harbor Branch Oceanographic Institution, Fort Pierce, Florida, USA
ACADEMIC PRESS A Horcourt Science and Technology Company
San Diego San Francisco New York Boston London Sydney Tokyo
This book is printed on acid-free paper Copyright 0 2000 by ACADEMIC PRESS Chapter entitled 'Population Structure and Dynamics of Walleye Pollock, Theragru chalcogrumma' is a US Government work in the public domain and not subject to copyright. All Rights Reserved No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopy, recording, or any information storage and retrieval system, without permission in writing from the publisher. Academic Press 24-28 Oval Road, London NW1 7DX, UK http:llwww.hbuk.co.uWap/ Academic Press A Harcourt Science and Technology Company 525 B Street, Suite 1900, San Diego, California 92101-4495, USA http:llwww.apnet.com ISBN 0-12-026137-5 A catalogue record for this book is available from the British Library
'Ehpeset by Keyset Composition, Colchester, Essex Printed in Great Britain by MPG Books Ltd., Bodmin, Cornwall
99 0001 02 03 04 MP 9 8 7 6 5 4 3 2 1
K . M . BAILEY,Resource Assessment and Conservation Engineering Division, Alaska Fisheries Science Center, 7600 Sand Point Way NE, Seatde W A 98115, USA R. BEIRAS, Area de Ecoloxia, Universidade de Vigo, 36200 Galicia, Spain P. BEmzm, Marine Molecular Biology Laboratory, School of Fisheries, University of Washington, Seattle W A 98195, USA W. S. GRANT,Conservation Biology Division, Northwest Fisheries Science Center, 2725 Montlake Blvd., Seattle W A 98112, USA E. H I S ,IFREMER, Quai Silhouette, 33120 Arcachon, France T. J. QUINN 11, Juneau Center, School of Fisheries and Ocean Sciences, University of Alaska Fairbanks, 1120 Glacier Highway, Juneau A K !?98018677, U S A M . N . L. SEAMAN, Institute of Marine Research, 24105 Kiel, Germally
This Page Intentionally Left Blank
CONTENTS
CONTRIBUTORS TO VOLUME 37...........................................
v
The Assessment of Marine Pollution .Bioassays with Bivalve Embryos and Larvae E. His. R. Beiras and M. N. L. Seaman 1. Introduction .................................................................. 3 2. Relevant Aspects of Bivalve Biology....................................... 10 3. Bioassay Methodology ....................................................... 40 4. Testing the Toxicity of Marine Pollutants to Bivalve Embryos and Larvae 87 5 . Assessing Marine Environmental Quality with Bivalve Embryo and Larval Bioassays..................................................................... 125 130 6. Summary and Discussion.................................................... 138 Acknowledgements ........................................................... References .................................................................... 139
Population Structure and Dynamics of Walleye Polllock. Theragra chalcogramma K. M. Bailey. T. J. Quinn II. I? Bentzen and W. S. Grant 1. 2. 3. 4. 5.
Introduction .................................................................. Background The Fishery. Life History and Ecosystem Interactions ....... Population Ecology.......................................................... Population Structure ......................................................... Management Implications ................................................... Acknowledgements ........................................................... References ....................................................................
Taxonomic Index .............................................................. Subject Index .................................................................. Cumulative Index of Titles .................................................... Cumulative Index of Authors .................................................
180 184 189 206 238 242 242 257 261 269 278
This Page Intentionally Left Blank
The Assessment of Marine Pollution . Bioassays with Bivalve Embryos and Larvae E. His'. R . Beiras' and M . N. L . Seaman3
'IFREMER. Quai Silhouette. 33120 Arcachon. France 2 h e a de Ecoloxia. Universidade de Ego. 36200 Galicia. Spain 31nstitute of Marine Research. 24105 Kiel. Germany Correspondence to: Edouard His. IFREMER. Quai Silhouette. 33120 Arcachon. France. Fax: 05 56 83 89 80. Tel. 05 56 83 78 17. e-mail:
[email protected]
1. Introduction ......................................................................... 3 1.1. Generalities on pollution ...................................................... 3 1.2. Bioassays: advantages and limitations ...................................... 4 6 1.3. Bivalve larvae and pollution .................................................. 2. Relevant Aspects of Bivalve Biology ............................................. 10 2.1. Species used in bioassays ................................................... 10 2.2. Reproduction .................................................................. 16 2.3. Larval rearing in the laboratory ............................................. 25 3. Bioassay Methodology ............................................................ 40 3.1. General methods .............................................................. 40 50 3.2. Bioassay procedures ......................................................... 3.3. Bioassay applications: toxicity tests and environmental bioassays ....... 74 82 3.4. Statistical methods ........................................................... 4. Testing the Toxicity of Marine Pollutants to Bivalve Embryos and Larvae .... 87 87 4.1. Pollutants ...................................................................... 117 4.2. Intrinsic (biological) factors affecting toxicity ............................... 4.3. Extrinsic (environmental) factors affecting toxicity ......................... 121 122 4.4. Interactions between different toxicants .................................... 5. Assessing Marine Environmental Quality with Bivalve Embryo and Larval 125 Bioassays ........................................................................... 5.1. Algal and bacterial toxins .................................................... 125 127 5.2. Urban and industrial effluents ............................................... 5.3. Receiving waters .............................................................. 128 129 5.4. Sediments ..................................................................... 6. Summary and Discussion ......................................................... 130
ADVANCES IN MARINE BIOLOGY VOL. 37 ISBN 0-12-026137-5
Copyright 0 1999 Academic Press All rights of reproduction in any form reserved
2
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
6.1. Sensitivity of bioassay organisms ........................................... 131 6.2. Assessment of the toxicity of various contaminants ....................... 132 6.3. Bivalve embryo and larval bioassay methodology ........................ 134 6.4. Perspectives in future research on bivalve larval bioassays .............. 137 6.5. Concluding remarks .......................................................... 138 Acknowledgements ................................................................... 138 References ............................................................................. 139
Tens of thousands of synthetic substances are in existence today and hundreds of new compounds are being introduced every year. Because of the complexity of the physico-chemical interactions between pollutants and the marine environment, the potential toxicity of contaminants can be assessed adequately only by means of bioassays with living organisms. From a practical point of view, a bioassay needs to be sensitive and scientiJically valid, yield rapid results at moderate cost, and the organism in question must be readily available. Ecotoxicological bioassays with bivalve embryos and larvae fulfil these criteria better than most other tests. They have increasingly come into use during the past three decades and are now commonly employed to ascertain the biological effects of pure chemicals, as well as to determine the quality of efluents, coastal waters and sediments sampled in the field. There do not appear to be very great differences between bivalve species with regard to larval sensitivity to toxicants. The principal species f o r bioassays are oysters (Crassostrea gigas and C. virginica), and mussels (Mytilus edulis and M . galloprovincialis). Bioassays are conducted with gametes and larvae of all ages: sperm and unfertilized eggs, embryos, young D-larvae, intermediate umboned larvae, and pediveligers towards the end of the pelagic period. Embryos are usually the most sensitive stage. Recent advances now also permit bioassays on metamorphosing pediveligers, a method particularly suited to investigate the effects of adsorbate-contaminated surfaces. There are various criteria for the assessment of toxic effects, including embryogenesis success (abnormalities), larval growth, mortality, physiology (e.g. feeding or swimming activity), and metamorphosis success. Chronic toxicity studies may be carried out over periods of several weeks, but larval rearing in the laboratory requires considerable effort (e.g. cultivation of algal food). The method of choice f o r investigations of acute toxicity and for routine monitoring studies is the embryo bioassay because it is very sensitive, relatively simple, and produces results within 24 or 48 hours. The data obtained by different investigators are often dificult to compare, however, because of differences in methodology. There is no firmly established procedure, and further simplification and standardization of techniques is required. In bioassays with a single pollutant, the effective toxic concentration may
THE ASSESSMENT OF MARINE POLLUTION
3
span several orders of magnitude, depending on bioassay procedures, larval stage and choice of response. Tributyl-tin (TBT) is the most toxic compound ever bioassayed with bivalve larvae, with effective concentrations (EC,,) as low as a few nanograms per litre (i.e. lo--’ ppb). Heavy metals (particularly mercury, silver and copper) are next in order of toxicity, with ECSovalues between a few micrograms per litre (ppb) and several hundred ppb. Chlorine and some organochlorine pesticides may also have ECso values of less than 100 ppb, while detergents and petroleum products are generally less toxic.
1. INTRODUCTION
1.1. Generalities on pollution
Although human activities have always impacted on coastal areas, it is only within the last two centuries that the effects of industrialization, intensive agriculture and coastal engineering (including tourism) have seriously begun to threaten marine life. Some of these impacts lead to environmental pollution, i.e. the introduction by man into the environment of substances or energy which may put human health or living resources at risk (Holdgate, 1979). Pollutants are defined as substances present in the natural environment which are (at least in part) of anthropogenic origin, and which may have deleterious effects on living organisms (Moriarty, 1990). According to this definition it is necessary to distinguish between contaminants, which encompass all substances of anthropogenic origin introduced into the environment, and pollutants, which consist of those contaminants with presumably negative biological effects. Moriarty recognizes, however, that the term “pollutant” is generally applied in the sense of “contaminant”. The assessment of marine pollution is not restricted to the study of the water. Most of the pollutants - heavy metals and hydrophobic organic compounds in particular - have a tendency to adsorb to sediment particles, and their concentrations in marine sediments may be several orders of magnitude higher than in the water column (Livett, 1988). The association between pollutants and sediments can be of very long duration and it may have deleterious effects on organisms living on, and within, the bottom. The assessment of the toxicity of marine bottom sediments has, therefore, also come into focus since the late 1970s. According to Moriarty (1990) some 63000 chemical products are presently in use, and 3000 of these constitute 90% of the total mass of
4
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
compounds being produced industrially. In addition, between 200 and 1000 new synthetic compounds are being introduced every year. It would obviously be desirable to screen their possible ecological effects. As already pointed out by Cunningham (1979), ecotoxicological studies require several levels of investigation. Depending on the mode of action of the pollutants in question, these may be at the subcellular, cellular, organismic, population or ecosystem level (see also Bayne, 1985; Haux and Forlin, 1988). 1.2. Bioassays: advantages and limitations
Anthropogenic impacts on the aquatic environment may be viewed from a physical, chemical or biological perspective. The biological effects of pollutants in the environment are more important than the mere presence of pollutants; with regard to environmental quality criteria, the data from chemical analyses of pollutants can only be interpreted within a biological context. It is therefore logical to use biological systems for the assessment of environmental quality (Anonymous, 1989). The following points apply to ecotoxicological bioassays in general:
1. Detection of new pollutants: reliance on chemical analyses alone would presuppose that the potentially important pollutants are known and already being monitored. Examples such as tributyl-tin (TBT) demonstrate that this is not always the case and that biological systems provide the means to detect and identify the presence of new or unexpected pollutants. 2. Bioavailability: chemical data often do not reflect the bioavailability of pollutants, e.g. owing to the speciation of organic compounds or the bonding state of metals. By definition, biological systems can only respond to what is effectively available to them, and test organisms therefore provide the best indication of bioavailable pollutants exceeding toxicity thresholds. 3. Integration of toxic effects: toxic contaminants typically do not occur singly and environmental quality is generally determined by their combined effects. Biological systems respond to the totality of environmental pollution, thus providing an integrated response to the totality of pollutants present, as well as to their interactions. 4. Cost: the continuous increase in the number of contaminants being introduced entails a constant rise in the cost of chemical monitoring programmes. This underscores the importance of using biological techniques as reconnaissance systems, because they help to focus the effort of chemical analysis on situations of demonstrated biological
THE ASSESSMENT OF MARINE POLLUTION
5
relevance. Therefore, “bioassays ... may be used before any other testing commences as a cost-effective screening tool” (Chapman and Long, 1983, p. 83). Any bioassay used for routine monitoring purposes should meet a variety of criteria (Moore, 1966; Butler et al., 1971; Bryan et al., 1980, 1985; Stebbing et al., 1980): it should be easy to learn, affordable, and not require very sophisticated equipment it should be of short duration (a few hours or days, rather than weeks) the organisms employed should, if possible, be available for bioassay purposes year-round the organisms should be of ecological or economic importance the organisms, whether they originate from laboratory cultures or from field sampling, should be identical as far as possible, in order to reduce variability resulting from age, size etc. genetically homogeneous, or cloned, organisms would be preferable, to reduce the effects of genetic variability the data should be of a type that can be analysed by standard statistical methods. Most bivalves of commercial interest generally fulfil these criteria, particularly those living in the marine and brackish water environment, such as oysters of the genus Crassostrea and mussels of the genus Mytilus. They are eurytypic organisms with higher tolerance to environmental fluctuations and therefore better suited for studies on the evaluation of environmental quality (Bayne, 1985), than stenotypic organisms with narrow tolerance. The different ontogenetic stages in marine species may differ in their sensitivity to pollutants (Coglianese and Martin, 1981). Among the many methods employed in bioassays, those using meroplanktonic stages (such as sea urchin plutei or bivalve veligers) appear to be the most promising to obtain reliable biological responses with regard to coastal water quality (Klockner et al., 1985). Many authors have proposed the use of early life stages of bivalves for marine toxicological studies, because it is clearly established that they are more sensitive to toxic substances than are the adults (Wisely and Blick, 1967; Granmo, 1972; Brereton et al., 1973; Hrs-Brenko et al., 1977; Coglianese and Martin, 1981; Bourne et al., 1981; His and Robert, 1981, 1982; Ringwood, 1991). The success of a species depends on its performance during passage through successive life stages, and it is therefore realistic to use the most sensitive stage for the purpose of testing environmental quality (Stebbing et al., 1980; Calabrese, 1984).
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
1.3. Bivalve larvae and pollution
Bivalve molluscs are an exclusively aquatic taxon of extremely wide distribution in fresh, brackish and marine waters around the globe. They are particularly abundant in the highly productive marine coastal areas. They have been exploited for food since prehistoric times and they have been cultivated systematically in various parts of the world for many centuries. The first problems in bivalve culture with regard to pollution were encountered at the beginning of the twentieth century. In studies on the American oyster, Crassostrea virginica,Prytherch (1924) asserted that “the rapid decline of this valuable industry has been brought about by a constant depletion of the oyster beds from various factors, such as pollution . . .” (p. 1). Prytherch was also the first to attempt, and to succeed in, the “artificial reproduction” of oysters. Besides intending to obtain larvae for culture purposes and thus offset the irregularities of natural recruitment, he also intended to conduct basic studies on larval biology and “to produce larvae for tests in regard to the effect upon them of various chemicals in solution” (op. cit., p. 2). It was already evident from Prytherch’s studies that bivalve larvae (veligers) are particularly suited for investigating environmental impacts, because in oyster cultivation areas the first environmental effects to be observed concerned the species’ natural reproduction. The same observation was made again 60 years later by Stanley and DeWitt (1983) who declared that the absence of a population of Mercenaria mevcenaria in the species’ normal area of distribution is an ecological indicator of environmental disturbance. The efforts of Loosanoff and his associates at the Milford Laboratory in Connecticut led to the development of practical and reliable methods for the culture of bivalve larvae (reviewed by Loosanoff and Davis, 1963) and the first toxicological bioassays with such larvae. A serious constraint in working with oyster larvae, however, was the limitation to the species’ reproductive period, from June to September. The invention of “conditioning”, i.e. techniques to obtain ripe adults at all seasons (Loosanoff, 1945) made it possible to study larval nutrition throughout the year. The work conducted by Cole (1937,1939) in Great Britain, made it possible to study the nutritional value of marine unicellular algae, first isolated by Parke at Plymouth; certain species, especially Isochrysis galbana, permitted the rearing of larvae to metamorphosis (Davis, 1953; Loosanoff, 1955; Walne, 1956; Davis and Guillard, 1958). It thus became possible to grow veligers in the laboratory as a matter of routine, and to study their principal environmental requirements. This was followed by studies on the effects of micropollutants on growth and survival in Crassostrea virginica larvae (Davis, 1961; Calabrese and Davis, 1967).
THE ASSESSMENT OF MARINE POLLUTION
7
At the same time, the concept of the “biological quality” of seawater was introduced by Wilson (1951) and Wilson and Armstrong (1961), who found that pluteus larvae of the sea urchin Echinus esculentus were capable of developing in seawater obtained from the Celtic Sea, but not in water from the vicinity of Plymouth. Subsequently Woelke (1961, 1966, 1967, 1972), Okubo and Okubo (1962) and Dimick and Breese (1965) suggested the use of oyster embryos and larvae to study the general effects of pollutants in water samples from the natural environment, and not just for toxicological laboratory investigations of specific pollutants. This ecotoxicological approach, i.e. the study of samples obtained from the natural environment, has since been expanded to include assays with marine sediments (Chapman and Morgan, 1983; Swartz, 1989; Phelps and Warner, 1990). Most bivalve embryo-larval bioassays can be conducted within a relatively short time period (24 to 48 hours after fertilization). Assessments of sublethal toxic effects of pollutants are based on the percentage of normal D-shaped larvae obtained at the end of embryogenesis (determination of the concentration that inhibits larval development in 50% of the fertilized eggs). These types of studies have been conducted by Woelke and coworkers (Woelke, 1960a,b, 1961, 1966, 1967, 1968, 1972; Cardwell et al., 1977a,b), Davis and coworkers (Davis, 1961; Calabrese and Davis, 1967), and many others. Experiments of longer duration, based on the assessment of larval growth, have also been conducted in order to study the effects of sublethal micropollutant concentrations (e.g. Davis and Hidu, 1969a; His and Robert, 1981 and 1982; Watling, 1982). These types of studies represent a better simulation of conditions prevailing in the natural environment, but they also necessitate the cultivation of algae to feed the larvae during the experiment, which may last several days, or several weeks if it is continued until the end of the veliger’s pelagic life stage. Finally, recent work has focused on the use of eyed larvae several days old, in order to investigate the effects of pollutants on the pediveliger stage and on metamorphosis, because these stages are considered to be especially sensitive (Stafford, 1913; also Watling, 1978, 1983; Sheffrin, 1982; Sheffrin et al., 1984; Beiras and His, 1994, 1995a; His et al., 1997b). Most bioassays with bivalve larvae and embryos have been conducted with various species of oysters (57% of the publications listed in Table l), followed by mussels (22%). Only two other species have been used regularly in larval bioassays - the quahog Mercenaria (Venus) rnercenaria and the coot clam Mulinia lateralis. The first step in bioassays with bivalve embryos and larvae is to obtain spawning adults and rear the fertilized gametes. We will therefore review the present state of knowledge on bivalve reproduction, methods to obtain
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
8
Table 1 Bivalve species used in embryo and larval bioassays. Mytilus
Crassosfrea
c gigas Beiras and His 1994, 1995b Beiras ef al., 1998 Bourne ef al., 1981 Boyden el al., 1975 Brereton el ol., 1973 Butler el ni., 1992 Cardwell, 1976,1978 Cardwell ef al., 1976, 1977a.b. 1979a.b Chang ef al., 1996 Chapman and Morgan, 1983 Chapman ef al., 1991, 1992 Chien and Chou, 1989 Cleary ef al., 1993 Coglianese, 1982 Coglianese and Martin, 1981 Connell ef al.. 1997 Coon era[., 1990 Crecelius, 1979 Eertman ef 01.. 1W3 Rtt el al., 1990 Garland el al., 1986 Geffard, 1997 Glickstein. 1978 Helm ef al., 1974 Hi%1996 His and Robert, 1980, 1981,1982,1987b His and Seaman, 1993 His ef al., 1983, 1996 Klockner ef al., 1985 Konar and Stephanson, 1995 Le Gore, 1974 Lourens el a/., 1995 McFadzen, 1992 McFadzen and Cleary. 1994 Maimstone el al., 1989 Martin ef al., 1981 Nelson ef al., 1983 Okubo and Okubo, 1962 Pbelps and Warner, 1990 Renard, 1991 Renzoni, 1973a,b Robert and His, 1981, 1985 Robert ef al., 1986 Smith, 1968 Stewart and Blogoslawski, 1985
C virginica Baker and Mann, 1992. 1994a.b Brown. 1981 Brown and Roland, 1984 Butler and Lowe, 1978 Calabrese, 1972 Calabrese and Davis, 1964,1%7,1970 Calabrese ef a[.. 1973, 1977a.b Capuzzo, 1979 Chapman et nl., 1987 Davis 1958,1960,1961 Davis and Calabrese,
1964 Davis and Hidu, 1%9a,b Dim, 1973 Hidu, 1965 Hidu el ni., 1974 Ho and Zubkoff, 1979, 1980 Mann and Rainer, 1990 Mchnes, 1981 McInnes and Calabrese, 1978,1979 Nelson ef al., 1983 Noyes ef al., 1978 Phelps and Mihursky, 1986 Phelps and Warner, 1990 Prytherch, 1931, 1934 Reuzoai, 1915 Richardson ef ol., 1982 Ringwood and Brouwer, 1995 Roberts 1980,1987 Roberts and Casey, 1985 Roberts and Gleeson, 1978 Roberts ei al., 1975,1977 Roesijadi ef al., 19% Roosenburg ef ~ l .1980b , Sigler and Leibovitch, 1982 Stewart er al., 1979 Stiles-Jewel],1994 Stiles and Blogoslawski, 1993 Tagatz and hey, 1981 Ukeles and Sweeney, 1969 Widdows ef al., 1989 Wikfors and Ukeles 1982 Wirtb eta/., 1996 Wolfe ef al., 1993
M. edulis Akberali ef a/., 1985 Armstrong and Millemann, 1974 Beaumont and Budd, 1982,1984 Beaumont and Tserpes, 1984 Beaumont el al., 1981 Breese eral., 1963 Brunetti ef a/., 1989 Butler ef al., 1990 Chapman and Long, 1983 Chapman ef al., 1993, 1996 Courtright el a/., 1971 Dimick and Breese, 1965 Dixon and Prosser, 1986 Granmd, 1972 Granmo and Jorgensen, 1975 Granmo ef 01.. 1988, 1989 Hansen ef 111.. 1997 Hoare er al., 1995a,b Johnson, 1988 Knezovicb ei al., 1996 Lapota er al., 1993 Le Pennec and Prieur, 1972 Le Pennec ef al., 1973 Long ef al., 1990 Magnusson el al., 1996 Martin ef d., 1981 Mitchell ef al., 1985 Morgan ef al., 1986 Okubo and Okubo, 1962 Pavicic, 1980 Sheffrin and Williams, 1984 Sheffrin ef al., 1984 Stewart ef al., 1%7 Strdmgren and Nielsen, 1991 main, 1983 Wisely and Blick, 1967 Wolfe er al.. 1995 Zhadan er al., 1992
M.galloprovincialis Beiras and Hq 1995a Brunetti ef ai., 1989 Bucaille and Kim, 1979 His and Beiras, 1995 €in-Brenko er a[., 1977 Ix Pennec and Le Roux, 1979 Le ROUK,1977 Lucu ef a/., 1980 Pagano ef ol., 1996 Pavicic, 1976 Pavicic and Pihlar,1982 Pavidc ef al., 1984a, 1994a,b Renzoni, 1973a Robert and His, 1981 Seaman el al., 1991
9
THE ASSESSMENT OF MARINE POLLUTION
Table 1 -(Continued). Crnsroslrea
C. gigas Slewart el al., 1967,1983, 1991 Thain and Watts, 1987 Thain eral., 1990 Utting and Helm, 1985 Van den Hurk, 1994 Van den Hurk el al., 1997 Wang ef 01.. 1985 Warder al., 1992a,b Watling, 1978,1981,1982, 1983 Wikfors el al., 1993 Williams el 01.. 1986 Woelke, 1960,1961,1967, 1968, 1972 Wolfe er el., 19% Zhadan ei a/., 1992
Mytilus
C. virginica
M. edulis
M. galloprovincialis
Wright el al., 1983 Zaroogian and Morrison, 1981
Other species Argopeclen irradians: Nelson and Siddall, 1988; Wright el al., 1983 Callisfa brevisiphonata: Zhadan el a/., 1992 Cerastodema edule: Tunmemans el al., 19% Chlamys asperrima: Krassoi, 1995; Krassoi el 01.. 1996, 1997: Pablo el al., 1997; Stauber er a/., 1996 Clinocardium nultalli: Stewart el al., 1967 Crassoszreu ungulata: Renzoni, 1973a Crassosrrea nrccullaru: Watling, 1981,1982 Crassosrrea iredalei: Ramachandran ef a/., 1997 Crassostrea margaritacea: Watling, 1981, 1982 Crassosrrea rhizophorae: Chung, 1980; Nascimento, 1989; Pereira er al., 1998 Isognomon californicurn: Ringwood, 1990,1991,1992a.h, 1993 Macoma balthica: Tunmemans el al., 1996 Macrru chinensix Zhadan el a/., 1992 Mercenaria mercenaria: Brown, 1974;Byme and Calder, 197R Calabrese, 1972; Calabrese and Davis 1966,1970; Calabrese and Nelson, 1974; Calabrese el al., 1977a.b Davis, 1958, 1%0; Davis and Calabrese, 1964; Davis and Hidu, 1969a.b:
Hidu, 1965;Huntington and Miller, 1989; Laughlin el a/., 1988.1989; Miller, 1989; Pavicic. 1980, Roberts, 1987; Roberts 1983; Wright el a/.. 1983 Meretrix lusoria: Tzong-Shean and Chen, 1993 Mizuhopecfen yessoensix Karaseva and Medvedeva, 1993; Zhadan era/., 1992 Mulinia larerdix Burgess and Momson, 1994, Calabrese, 1970b; Calabrese and Rhodes 1974; Diaz 1975; Gormly e1 al., 19%; Hall el a/., 1995; Ho and Zubkoff, 1980, 1982, 1983; Mann el al., 1991; Morrison and Petrocelli, 1990 Pelletier el al., 1997; Renzoni, 1975; Roberts 1980:Wright eral., 1983 Mya arenaria: Roosenburg el al., 1980a Mytilus califomianus: Cherr er al., 1990; Spangenberg and Cherr. 1996 Mytilus lrossulus: Karaseva and Medvedeva, 1993 Osrreu edulix Connor, 1972 Davis 1961;Davis and Hidu. 1969b Helm. 1971; Millar and Scott, 1968, Nottage and Birkbeck, 1987a.b Renzoni. 1973b: Smith, 1968; Thain. 1983 Osfreaplicalula: Xu el al., 1994 Peaen marimus: Beaumont and Budd, 1982; Beaumont el al., 1987 Prolothaca staminea: Cardwell el a/., 1979b Rangiu cuneala: Mann el al., 1983 Saccosfrea commercialis (= Crassostrea commercialis): Krassoi. 1995; Nell and Holliday, 1986; Wilson and Hynes 1997; Wiseley and Blick, 1967 Scrobicularia plma: Ruiz er al., 1994,1995a.b.c Spisula sukhalinensis: Zhadan el a l , 1992 SpisuIu solidissima: Thurberg er al., 1975; Calabrese et a/., 1982: Eyster and Morse, 1984; Mann et al., 1983; Wright er a/., 1983 Tapesphilippinarum ( = bjaponica): Cardwell e1 al., 1979b: Cleary el al., 1993; McFadzen, 1992; McFadzen and Cleary, 1994 Tresus capar and 7: nuttalli: Cardwell. 1976; Cardwell el a/., 1978, 1979a el al., 1975; Robinson,
10
E. HIS, R. BEIRAS AND M . N. L. SEAMAN
and fertilize the gametes, and cultivation of the larvae to metamorphosis, with emphasis on the principal species used to date in toxicological assays and environmental monitoring.
2. RELEVANT ASPECTS
OF BIVALVE BIOLOGY
2.1. Species used in bioassays
The species used most often in bioassays is the Pacific or Japanese oyster Crassostrea gigas (30%), followed by the American or Eastern oyster C. virginica (23%), the blue mussel Mytilus edulis (15%), the quahog or hard clam Mercenaria mercenaria (13%), the Mediterranean mussel Mytilus galloprovincialis (7%), the European oyster Ostrea edulis (3%), and the coot clam Mulinia lateralis (3%) (Table 1). The biology of these species is reviewed below. 2.1 .l. The Japanese oyster Crassostrea gigas The Pacific or Japanese oyster originates from Eastern Asia, where it is distributed from the Sea of Okhotsk southward along the coasts of Korea and Japan to China (Ahmed, 1975; Andrews, 1980; Ventilla, 1984). Following the studies by Imai and Sakai (1961) four varieties or races are generally recognized; these are, distributed from north to south, Hokkaido, Miyagi, Hiroshima and Kumamoto (see Quayle, 1969). They are characterized, among other points, by different spawning temperatures. The Miyagi variety has repeatedly been introduced to the Pacific coast of North America since the beginning of the twentieth century, and it has formed an important natural population which spawns irregularly, but intensively, in British Columbia. Kumamoto oysters have been introduced in smaller numbers to Washington State (Quayle, 1969; Andrews, 1980). Owing to the scarcity of naturally setting oysters, the spat for aquaculture in British Columbia and Washington State are mostly produced by hatcheries. Miyagi, but also Hiroshima and Kumamoto oysters, have been introduced to Australia, and now form important natural populations on the southwest coast and around Tasmania (Thompson, 1951; Summer, 1980a,b). They have spread accidentally to the coast of New South Wales, where oyster farmers consider them a nuisance (Medcof and Wolf, 1975). They were also introduced accidentally to New Zealand in the 1960s, where they quickly formed abundant natural populations (Dinamani, 1971, 1987).
THE ASSESSMENT OF MARINE POLLUTION
11
In Europe (discounting the probable introduction centuries ago of Crassostrea angulata, often considered the same species as C. gigas, but never used in bioassays, except by Renzoni (1973a,b), Pacific oysters were first introduced to Great Britain in the form of hatchery-produced spat from British Columbia in 1972 and 1975 (Gosling, 1981). Natural spatfall has been observed in Britain after warm summers (Askew, 1972). After the decline of the stocks of C. angulata in France, massive imports of Miyagi spat from Japan, as well as spat and spawners from British Columbia, quickly led to the formation of important natural stocks in the early 1970s (His, 1976). C. gigas reproduces very successfully in most oyster culture areas of southwest France, and occasionally in Brittany. Natural spatfall has also led to the establishment of significant wild populations in the Netherlands (Andrews, 1980) and Germany. The latter probably originated from spawning by the Dutch stocks (Seaman and Ruth, 1997), because successful local reproduction of the German stocks appears improbable (Neudecker, 198l), even though very ripe adults are found in some years. Finally, Japanese oysters also reproduce in the Adriatic part of the Mediterranean from Triest (Valli et al., 1979) to the Canal of Lim, where spawning occurs from July to October (Filic and Krajnovic-Ozretic, 1978; Hrs-Brenko, 1982). Andrews (1980) considers C. gigas to be the only species of oyster to have been introduced as a successful member of coastal communities around the world; “this oyster has adapted well to a wide range of environmental conditions and is probably the most globally widespread and ubiquitous oyster species in the world” (Chew, 1990). This explains why it is one of the species used most frequently in marine ecotoxicology. Gametogenesis in C. gigas is activated when the water temperature attains 14°C and continues as temperatures rise; Mann (1979) showed that not only are temperatures of 15°C to 18°C required, but that gametogenesis also depends on the duration at different temperatures (day-degrees). In addition, a temperature of at least 18°C to 20°C is necessary for spawning to occur (Fujiya, 1970; Mann, 1979). Buroker (1985), on the other hand, mentions a spawning temperature of 19°C to 24”C, and a salinity of 11 to 32 psu. With respect to the reproductive cycle in the natural environment, Imai and Sakai (1961) showed the differences between the four varieties in transplantation experiments at Miyagi province. The Hokkaido and Miyagi oysters spawned in late August to early September, those from Hiroshima in early or mid-September; those from Kumamoto spawned in late July when water temperatures attained 22°C to 23°C and began to mature again, with mature gametes still present in November and December. More generally, Koganezawa (1978) reports that Pacific oysters in Japan spawn from June to August. On the Pacific coast of Siberia (Vostok Bay,
12
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Sea of Japan), they generally spawn in July (Yakovlev, 1978). In the United States there are no natural populations except at Washington State, where they spawn in July and August (Katkansky and Sparks, 1966; Perdue and Erickson, 1984). In Australia the oysters spawn from midJanuary or mid-February on, after water temperatures exceed 22°C (Thompson, 1951). In New Zealand the reproductive season may be very extended, with light spawning in mid-spring and successive spawnings in summer and early fall (Dinamani, 1987). In Europe, spawning is occasional in British, Dutch and German waters, occurring only in exceptionally warm summers (Askew, 1972; Andrews, 1980; Neudecker, 1981). Pacific oysters spawn intensively on the Atlantic coast of France from the Bidassoa estuary on the French-Spanish border, to the Loire estuary, which is the region with the most important centres of oyster culture. Further north, in Brittany and Normandy, the oysters attain sexual maturity, but spawn only in unusually warm summers; spatfall is light, except in the nearly closed bay of the Rade de Brest. Gametogenesis begins at the end of winter and spawning takes place after the water temperature reaches 20°C or more. At Arcachon the oysters spawn every 2 or 3 weeks from mid-June to mid-August with intermittent periods of gonad restoration (His, 1976); from mid-August to the end of September, bioassays can be conducted here without laboratory conditioning by using unspawned ripe oysters from Brittany and Normandy. 2.1.2. The Eastern oyster Crassostrea virginica The Eastern oyster has its natural distribution on the Atlantic coast of North America from the Gulf of St Lawrence to the Gulf of Mexico and the West Indies (Ahmed, 1975; Hayes and Menzel, 1981; Buroker, 1983 and 1985; Kennedy, 1983). It has repeatedly been introduced for aquaculture purposes to the Pacific coast of North America (Andrews, 1980) and to the Baltic coast of Denmark and Germany since the late nineteenth century (Seaman and Ruth, 1997), but has never become established in Europe. It has, however, been introduced successfully to Hawaii. Loosanoff and Engle (1942) already suspected the existence of different physiological races, because of variations in spawning behaviour. Stauber (1950) and Loosanoff and Nomejko (1951) subsequently established the existence of three different races in the northern, central and southern areas of its distribution, with spawning thresholds at 17"C, 20°C and 25"C, respectively. Similarly, C. virginica from Long Island Sound may mature at 12"C, whereas those from Florida do not ripen at 18°C (Loosanoff, 1968). The spawning season is in late June and early July in the Gulf of St Lawrence, and from mid-June to early August at New Haven (Stiles and
THE ASSESSMENT OF MARINE POLLUTION
13
Longwell, 1973). In Chesapeake Bay it may last until the end of August (Kennedy and Krantz, 1982), and in Florida it lasts from March to October (Ingle, 1951). In Hawaii, C. virginica reproduces from February to November (Sakuda, 1966). 2.1.3. The European oyster Ostrea edulis The European oyster was originally distributed from Norway to Morocco and the Mediterranean (Walne, 1965; Ahmed, 1975; see also Korringa, 1941, 1958; Cole, 1941, 1942). It was introduced experimentally in 1949 to various estuaries in Maine (United States), where it reproduced, notably establishing itself at Boothbay Harbor (Loosanoff, 1955). It was also transplanted to the western coast of the United States with moderate success (Matthiessen, 1970). European oysters generally reproduce at temperatures of 15°C to 20°C (Buroker, 1985). According to the differences in spawning temperature, various physiological races exist (Korringa, 1958), but their distribution is not necessarily linked to geographic latitude. In Galicia (northwest Spain) some stocks spawn at temperatures as low as 12°C to 13”C, whereas in certain Norwegian fjords (where a superficial freshwater layer creates a greenhouse effect) the European oyster spawns at 25°C (Yonge, 1960; Wilson and Simons, 1985). Along the Atlantic coast the spawning season may last from early spring to November or December (Yonge, 1960). In the Adriatic, its larvae are encountered from April to October (HrsBrenko, 1982). 2.1.4. The mussels Mytilus edulis and M. galloprovincialis The genus Mytilus is widely distributed throughout the world (see Figure 1.6 in Gosling, 1992); its taxonomic status has often been investigated and discussed (Seed, 1971, 1976; Gosling, 1984, 1992; Lubet et al., 1984; Brock, 1985; MacDonald et al., 1991). Mytilus edulis and M . galloprovincialis are often sympatric, characterized by considerable morphological variability. They frequently interbreed, giving rise to intermediate phenotypes; where they coexist, the former is more frequently found in exposed, and the latter in more protected areas (Lubet, 1959). Their spawnings rarely coincide (Lubet, 1973; Seed, 1976). According to Gosling (1984) morphological, cytological, immunological, electrophoretic and hybridization studies show that M. galloprovincialis is “a form or ecotype” of M . edulis. There are no major differences between the chromosomes of the two forms (Dixon and Flavell, 1986). For reasons of convenience, however, they are still often referred to as separate species in the literature.
14
E. HIS,
R. BEIRAS AND M. N. L. SEAMAN
Both forms are very widely distributed. M. edufis is boreal, but eurythermic (Lubet, 1973), and is distributed from northern Norway (where summer temperatures hardly surpass 4°C to 5°C) to the Bay of Biscay (where summer temperatures may reach 23°C at Arcachon). It is found on both coasts of the North Atlantic, including the North and Baltic Seas (Gosling, 1992), but it is rare in Spain and Portugal, and very infrequent in the Mediterranean (Lubet et af.,1984; McDonald et af.,1991; Gosling, 1992). M. galloprovincialis is distributed from southern Britain to the northern coast of Africa, and the Mediterranean and Black Seas. According to Gosling (1992) M. trossufus,another ecotype of M. edufis, has a more northerly distribution than either M. edufis or M. gaffoprovincialis, and is found only in the northern hemisphere. The range of M. cafifornianus, considered to be a true species distinct from the M. edufis group, is restricted to the northeast Pacific, but is continuous over a latitudinal range of about 30 degrees. M. trossufus and M. cafifornianusare almost never used in bioassays. In many areas mussels may maintain sexual maturity and spawn repeatedly throughout the year. It is generally accepted that the reproductive period is shorter in cooler regions, and that it is more extended in M . galloprovincialis than in M . edulis (Seed, 1971, 1976). Spawning may generally be obtained in M. galloprovincialis throughout the year, except in unusually cold winters, when reproductive activity may cease altogether. The spawnings of greatest relevance for recruitment are in early spring and in fall, although the fall spawning may be late, or lacking altogether, in the northern range of its distribution (Lubet, 1973). In M. edufis on the coasts of Europe, the principal spawnings take place from March to early June. This is followed by a non-reproductive period during the summer, and gametogenesis resumes in October or November and continues through the winter (Bayne et af., 1975). Mussels from warmer waters generally spawn earlier, although this was not found to be the case in seven M. edulis populations on the east coast of the United States; Bayne’s rule on day-degrees did not apply either (Newel1 et af., 1982). 2.1.5. The hard clam Mercenaria mercenaria The northern quahog, or hard clam, is distributed on the Atlantic coast of North America from the Gulf of St Lawrence to the northern Gulf of Mexico, and it is particularly abundant from Maine to Virginia. In the southern part of its range, it is sympatric with M. campechiensis, with which it may interbreed. M. mercenaria has been introduced to the coast of California (Ansell, 1968; Stanley and DeWitt, 1983). It has also been introduced to Europe, forming wild stocks in the areas of Portsmouth and Southampton in Britain (Mitchell, 1974), but despite various introductions
THE ASSESSMENT OF MARINE POLLUTION
15
to France, significant populations have only become established in the Marennes-OlCron area and in Brittany. Small stocks also exist on the Dutch and Belgian coasts (Tebble, 1966). The reproductive cycle of this species has been reviewed by Stanley and DeWitt (1983), with additional descriptions by Keck et al. (1975), Dalton and Menzel (1983), Knaub and Eversole (1988), Manzi et al. (1985), and Walker and Heffernan (1995). The sexual cycle has been described for populations of Long Island Sound (Loosanoff, 1937a,b), Delaware Bay (Keck et al., 1975), North Carolina (Porter, 1964), South Carolina (Eversole et aE., 1980), and northwest Florida (Dalton and Menzel, 1983; Eversole, 1989). M . rnercenaria generally has two breeding peaks in North and South Carolina, where spawning occurs at 27°C to 30°C (Porter, 1964; Eversole et al., 1980). This bimodal pattern is not found at Delaware Bay, where spawning takes place at 25°C to 27°C (Keck et al., 1975), nor at Long Island Sound, where the temperature for spawning is 23°C to 25°C (Loosanoff, 1937a,b). In Britain the species has adapted to the colder climate and spawns at 18°C to 20°C (Mitchell, 1974). In contrast, Dalton and Menzel (1983) describe a trimodal spawning pattern in northern Florida, with peaks in autumn, winter and spring. Laughlin et al. (1988) assert that reproduction occurs year-round and that mature spawners are available without laboratory conditioning throughout the year. The existence of different spawning temperatures for different stocks has led Porter (1964) and Keck et al. (1975) to hypothesize the existence of different races. In summary, M. mercenaria does not seem to have a sexually inactive season, and individuals at various stages of gametogenesis can be found at all times, although the proportions vary greatly between locations. 2.1.6. The coot clam Mulinia lateralis The coot clam is also distributed on the eastern coast of North America from Canada to the eastern Gulf of Mexico (Kennedy and Mihursky, 1971; Calabrese and Rhodes, 1974; Morrison and Petrocelli, 1990; Burgess and Morrison, 1994). The species is not very abundant, except at certain favourable locations (Calabrese, 1969a,b, 1970a). According to the presence of the larvae in the plankton, the reproductive period is from mid-July to early December at Prince Edward Island, Canada (Sullivan, 1948); the larvae are found from May to November at Tred Avon River, Maryland (Shaw, 1965). At Long Island Sound, gametogenic activity continues throughout the year, with principal spawning peaks in late July and mid- to late August, the first peak being at temperatures near 20°C (Calabrese, 1970a; Calabrese and Rhodes, 1974).
16
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
This confirms the observations of Sullivan (1948) and Loosanoff et al. (1966).
2.1.7. Other species At least 25 other bivalve species have also been used in bioassays (see Table l), but most of them do not fully satisfy the criteria required (Stebbing et al., 1980). Although some of these species are of economic importance, their geographic distribution is generally limited. This applies in particular to the mussels Mytilus californianus and M . trossulus, the oysters Crassostrea angulata, C. cuccullata, C. margaritacea, C. rhizophorae, Saccostrea commercialis and Ostrea plicatula, as well as to the Hawalan species, Zsognomon californicum, which has been studied in detail by Ringwood (1990, 1991, 1992a,b, 1993) with regard to the micropollutant sensitivity of adults and larval stages.
2.2. Reproduction
Experiments with larvae depend first of all on the availability of spawning individuals of both sexes. The bivalve species commonly in use for bioassays have a very wide geographical distribution, but the availability of mature adults may nevertheless be highly variable. Following the studies by Orton (1927) on Ostrea edulis, Coe (1932a,b) on Crassostrea virginica and Loosanoff (1937a,b) on Mercenaria mercenaria, there have been numerous publications on bivalve reproduction (reviews by Galtsoff, 1964; Raven, 1964; Purchon, 1968; Sastry, 1975,1979; Seed, 1976; Andrews, 1979; Mackie, 1984; Gosling, 1992). 2.2.1. Gametogenesis The reproductive anatomy of bivalves is rather simple. The gonad consists of a mass of follicles which develop fully at the period of sexual maturity, at which time the sexual products make up a significant part of the body. The sexual products grow within genital ducts, the diameters of which increase progressively during gametogenesis; the various small ducts converge on larger gonoducts, through which the gametes are expelled into the exhalent part of the pallial cavity at the time of spawning. The simplicity of the reproductive organs, notably the absence of specialized structures (such as a penis in the male or accessory glands in the female) facilitates the change of sex observed in many bivalve species. In oysters, during the winter, when there is no reproductive activity, the
THE ASSESSMENT OF MARINE POLLUTION
17
gonad mass is replaced by a mass of connective tissue with vesicle cells containing lipids and glycogen. Embedded in this and close to the internal organs is a duplicate system of branching tubules, one on each side, beginning near the anterior end of the body, uniting into a single tube on either side and ending in the genital pore. These pores open into the suprabranchial chamber at the posterior base of the adductor muscle, in close proximity to the opening of the urinal ducts (Quayle, 1969). During the breeding season the reproductive organs form at least 50% of the body volume. In a fully ripe oyster, the gonadal tubules, small in diameter at the anterior end of the oyster and thickening as they approach the genital opening, may be clearly seen on the surface of the soft body of the oyster. At this time, the two gonadal systems are almost completely inseparable, except at the genital pores. In mussels the gonad extends in a diffuse manner throughout the mantle. It is made up of a multitude of acini which are grouped around tubules, which are in turn clustered around the gonoducts. These fuse to form the terminal gonoducts on either side of the body, which then unite and open into the genital papilla near the posterior adductor muscle. In venerid clams such as Mercenaria mercenuria, the gonad is very diffuse and located around the digestive gland, sometimes covering it entirely and extending into the foot by the time maturity is attained. The genital pores open into the exhalent part of the pallial cavity and the gametes are released by the exhalent siphon. In Muliniu lateralis, the gonad of ripe animals forms a uniform and continuous mass around the digestive tube and gland (Calabrese, 1970a, b). Most lamellibranchs (96%) are gonochoric, i.e. they have separate sexes (Coe, 1943), but there are various hermaphroditic species, including oysters. With regard to the hermaphrodites, the following classifications have been proposed (Table 2). Crussostrea gigus and C. virginica are both alternative hermaphrodites (Amemiya, 1929; Coe, 1932a,b), changes in sex occurring irregularly during the course of life. Most individuals only emit sexual products corresponding to one sex in the course of the reproductive season, but there is a small proportion of simultaneous hermaphrodites (Loosanoff, 1965a). Crassosfreu is usually protandric, about 70% of the individuals male in their first year, and about half in the second year, with females dominant in the older age groups. In Ostreu edulis, which is larviparous, protandric sex changes are the rule (Sparks, 1925; Orton, 1927). The time lag between the maturation of spermatocytes and oocytes is small, however, and European oysters may alternately function as males and females in the course of one reproductive season. The frequency of sex changes increases with temperature, so
18
E. HIS, R. BEIRAS AND M . N. L. SEAMAN
Table 2 nYo systems of classifying hermaphroditisms in Bivalvia (from Fretter and Graham, 1964).
Coe Functional or simultaneous Consecutive Rhythmical consecutive Alternative
Bacci
Simultaneous (with synchronous ripening) Successive (with asynchronous ripening) or consecutive Successive (with asynchronous ripening) or alternate Successive (with separate ripening) or alternate
it is related to latitude (Cole, 1942). The change from the female sex to the male is more rapid than vice versa. Mytilus edulis and M. galloprovincialis are gonochoric, even though rare cases of simultaneous hermaphroditism have been reported by Coe (1943) and Lubet (1959). According to the latter, hermaphroditism occurs in about 0.02 to 0.1% of the population. Changes of sex by the same individual have never been observed. Mercenaria rnercenaria is a protandric hermaphrodite (Eversole, 1989), but the simultaneous development of oocytes and spermatocytes is common during the first year of life (Loosanoff, 1937a,b), and a small percentage of simultaneous hermaphrodites is always present in the population. Nevertheless, 98% of the individuals are male during the first year, but the sex ratio becomes even with advancing age. Mulinia lateralis is gonochoric. The reproductive cycle is very brief, the egg-to-egg cycle being only 39 to 135 days with an average generation time of 60 days (Calabrese, 1970a). Individuals as small as 2.7 mm in length are already capable of spawning (Calabrese and Rhodes, 1974). Another practical aspect of this species is that the sexes are readily distinguishable, because the orange-red female gonad and the whitish male gonad are discernible through the shell near the umbo. Galtsoff (1964), Bayne (1976) and Mackie (1984) have described the structure of bivalve gametes in detail. In the males the chromosomal reduction (first order spermatocytes with 2n to second order spermatocytes with n chromosomes) is achieved by an equal division during spermatogenesis. In the female, however, the gametes are spawned at the germinal vesicle stage and only develop to the metaphase of the first meiotic division: meiosis is arrested until the egg is activated by sperm. Penetration of the ovum by the sperm is made possible by the acrosomal reaction (i.e. dissolution of the egg membrane by the sperm’s acrosome), after which the sperm cytoplasm fuses with the egg cytoplasm; this occurs within about 5 minutes in Mytilus edulis. Meiosis in the egg (Figure 1)then
THE ASSESSMENT OF MARINE POLLUTION
19
Figure 1 Normal development in the bivalve egg. For simplification only one pair of chromosomes is shown. (a) egg at release at metaphase of meiosis I, activation by sperm; (b) meiosis I is complete, first polar body extruded, sperm nucleus has entered egg; (c) meiosis I1 completed, second polar body extruded, male and female pronucleus unite; (d) first cleavage perpendicular to point of polar body extrusion. (From Beaumont and Fairbrother, 1991.)
continues with the formation of the first and second polar bodies, after which the egg’s chromatin forms the chromosome vesicle. This is followed by the fusion of the male and female chromosome vesicles and the formation of the 2n pronucleus (Longo and Anderson, 1969a,b; Bayne, 1976; Mackie, 1984; Cherr et al., 1990). 2.2.2. Sexual maturation in the field Spawning in marine invertebrates may occur year-round in regions with little seasonal variability (Sanders and Hessler, 1969, cited after Dalton and Menzel, 1983). According to Rand (1973), cool climates are characterized by species with a single annual spawning season, temperate zones by species with two separate spawning periods, and tropical areas by species with year-round spawning. The determination of sexual cycles has been based on direct observation of spawning activity in the natural environment, on determinations of the state of maturation of the gonad, on the appearance in the plankton of larvae of the species in question, and on the settlement of the juveniles (Seed, 1976). The reproductive cycle of bivalves can generally be divided into three phases: gametogenesis and vitellogenesis, spawning and fertilization, and larval growth and development (Newel1 et al., 1982). With regard to the maturation state of the gonad, a variety of scales have been proposed by Chipperfield (1953), Lubet (1959), Seed (1969) in mussels, Loosanoff (1942) and Kennedy and Battle (1964) in C. virginica, and Keck et al. (1975) and Eversole et al. (1980) in Mercenaria mercenaria. 2.2.2.1. Environmental factors influencing gametogenesis. Mackie (1984) has reviewed the principal factors governing sexual maturation in bivalves. They are either exogenous, such as temperature, lunar cycle, and,
20
E. HIS, R. BEIRAS AND
M. N.
L. SEAMAN
particularly, nutritional factors, or endogenous, such as genetic or hormonal factors. Among the external factors, Orton (1920) considered temperature the most important influence. Loosanoff and Nomejko (1951) were the first to advance a hypothesis on the existence of physiological races in Crassostrea virginica with different temperature requirements for spawning. The issue was taken up again by Ahmed (1975) with respect to oysters. As will become apparent in the next section, laboratory “conditioning” of tachydictic species, i.e. the possibility of inducing sexual maturation out of season, largely depends on an elevation of the temperature. Even though this applies to most bivalve species used in bioassays, it must be emphasized that mussels, to the contrary, reproduce at low temperatures. Bayne (1975) established a day-degree function for the relationship between temperature and duration of gametogenesis in Mytilus edulis. The same observation was made by Mann (1979) with regard to C. gigas, and by Knaub and Eversole (1988) in Mercenaria rnercenaria. In eurythermic species with high temperature tolerance, the duration of the reproductive cycle often varies with geographical latitude. Nutritional factors are also of paramount importance. Various authors have shown that bivalves accumulate nutritional reserves in their connective tissue at the time of sexual inactivity, in order to mobilize them towards the formation of sexual products during gametogenesis (see Mackie, 1984). The nutritional needs of mussels increase during gonad maturation (Bayne, 1975); stress and bad nutritional conditions, on the other hand, are accompanied by a reduction in fertility, with a decrease in the number of gametes produced and a modification of their biochemical composition (Bayne et al., 1975,1978; Bayne, 1985). A rise in temperature and stress from a simultaneous decrease in available food, and the influence of pollutants such as hydrocarbons also lead to lower gamete quality and larval viability. In C.virginica starvation leads to a decrease in the proportion of females (Bahr and Hillman, 1967). An increase in the proportion of males in a population of oysters may therefore be an indication of environmental disturbance (Kennedy, 1983). Among the endogenous factors, neurosecretion was first shown to be important in mussels by Lubet (1959), and in oysters by Nagabhushanam (1963). Genetic factors may also be important (Lannan et al., 1980). 2.2.2.2. Environmental factors influencing spawning behaviour. Male oysters are more sensitive to stimulation and usually spawn first. The first males induce spawning in their male neighbours, then the first females follow suit, and finally the entire population spawns simultaneously. This results in the formation of “spawn streaks” described by Quayle (1969), milky masses of water several hundred metres in length spreading in the channels of bays and estuaries. Among the environmental factors which
THE ASSESSMENT OF MARINE POLLUTION
21
induce spawning behaviour are the thermal stimulation resulting from abrupt differences in temperature between the water masses of the falling and the rising tide, changes in salinity during the tidal cycle, mechanical agitation by waves and currents, and differences in pressure between high and low water and during the falling tide. Spawning in the field is never provoked by one single stimulus; when an appropriate combination of stimuli occurs, spawning is most common during the 3 hours following high tide (58% of all spawnings), and it is more frequent (69%) and more prolonged (76% of the duration of all spawnings) at spring tides (His, 1976). In flat oysters, “swarming” (i.e. the release of larvae incubated within the mother’s pallial cavity) is also related to the lunar cycle: from July to September larvae are most abundant in the plankton following a full moon (Orton, 1926), or 10 days after a full or new moon (Korringa, 1941; Knight-Jones, 1952).
2.2.3. Cleavage, embryogenesis and larval development Bioassays may be conducted with gametes and larvae of all stages, from sperm and unfertilized eggs to metamorphosis of the pediveliger at the end of pelagic life. Many researchers indicate the time passed between fertilization and the fertilized gametes’ first exposure to the substance being investigated, as this may have a bearing on the interpretation of the results. The rates of development and cleavage, and the duration of the successive embryonic and larval stages differ widely between the various bivalve species, and they also depend largely on environmental factors, notably on the temperature. The development of bivalves consists of an embryonic phase followed by a larval phase, and has been described by various authors (Erdmann, 1935 in 0. edulis; Cahn, 1950 in C. gigas; Galtsoff, 1964 in C. virginica; Lubet, 1973 and Bayne, 1976 in M. edulis). More recent work by Elston (1980) on C. virginica and by Waller (1981) on 0. edulis and C. gigas have provided additional details of larval morphology, particularly demonstrating the complexity of various larval organs. The common terminology employed for the various development stages is shown in Figure 2. 2.2.3.1. Embryonal stages. The fertilized egg (see Figure 1) completes meiosis by expelling the first and second polar bodies, and then begins to divide. The first division is unequal, giving rise to the blastomeres AB and CD, as well as a polar lobe which fuses with CD. The second division (4-cell stage) leads to the formation of three blastomeres (A, B and C) at the animal pole and one very large blastomere (D) at the vegetal pole. During the third division, the cleavage becomes spiral and subsequent
22
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
TERM EGG BLASTULA GASTRULA TROCHOPHORE VELIGER 0-SHAPED STAGE UMBO STAGE PEDIVELIGER SPAT
PRESHELLED PRODISSOCONCH I STAGE (less (about 7 days) than 1 day)
DISSOCONCH (remainder of life)
PRODISSOCONCH II (10-15 days)
--
--...
...
--...
....
....-
-
Relationship of stages of development of the prodissoconch to other common terms used to describe the larval shell and body of Ostrea edulis (dashed lines indicate uncertainty or transition; duration of stages may be highly variable). (From Waller, 1981.)
Figure2
divisions lead to the formation of the morula. Cilia appear, and the embryo becomes motile, gyrating slowly. Gastrulation begins as the micromeres of the animal pole cover the blastula, followed by a slight invagination that gives rise to the archenteron and to a small blastopore. 2.2.3.2. Larval stages. The first larval stage is the trochophore, which is covered by short cilia; in mussels and some other species it also has a flagellum. A dorsal thickening of the ectodermis is secreted by the shell gland, forming the initial organic cuticle which eventually spreads to cover the entire body. A circular band of cilia, the prototroch, also forms, enveloping the apex. The second larval stage is the veliger, which is formed 24 hours after fertilization in Crussostrea, and after 48 hours in Mytilus. It has a straight dorsal hinge giving the larva the characteristic shape of a capital letter D (hence the synonymous terms “straight-hinge larva” and “D-larva”), and it is some 60 to 70 pm in size in most bioassay species (see below). At this stage the larva begins to feed. The shell, consisting of two valves, begins to calcify, forming the prodissoconch I. The prototroch has continued its development to form the velum. As the larva grows beyond 100 pm, secretion of the secondary shell (the prodissoconch 11) begins and the umbo begins to form, extending over the dorsal hinge (hence the terms “veliconcha” and ‘‘urnboned larva”). In oysters and mussels, as the larva grows to a size of 250 or 300 pm, an “eye spot” appears within the shell and the foot is formed (hence the terms “eyed larva” and “pediveliger” for this stage). As the ciliated foot grows and becomes functional, the larva becomes capable of both pelagic and benthic
THE ASSESSMENT OF MARINE POLLUTION
23
modes of life, either swimming by use of the velum (e.g. to feed), or creeping along a hard substrate with its foot (e.g. to explore for a suitable site for settlement). Once it has attained this stage it is ready to metamorphose. All of the various stages mentioned above are used in bioassays. 2.2.3.3. Settlement and metamorphosis. The first photographs of settlement and metamorphosis in Crassostrea virginica were made by Prytherch during the 1930s, and some of his images were reproduced by Medcof (1961). The fixation of pediveligers of Ostrea edulis has been described by Cranfield (1973). The first attachment stage after metamorphosis is the plantigrade or juvenile (Figure 3). Oysters settle by glueing their shell to the substratum with a proteinaceous cement produced by the pallial gland (Figure 4). Mussels and clams attach themselves to the substratum with the help of byssal threads produced by the byssal gland at the base of the foot. During metamorphosis the velum disappears, and in oysters the foot is resorbed as well; in mussels and clams, the foot continues to develop and specialize to the adult form. Labial palps and gills are developed to replace the larval feeding apparatus, the velum, and the final shell, the dissoconch, is secreted. The anatomical changes of metamorphosis may make mussel larvae unable to feed for 1 to 3 days (Bayne, 1976). In Crassostrea virginica, Baker and Mann (1994b, p. 239) reported that “velar feeding occurred during the searching and crawling stages, but not during cementation”, and subsequently, Baker and Mann (1994a, see Figure 1, p. 94) identified four separate phases of metamorphosis: settler, prodissoconch postlarva, dissoconch postlarva and juvenile. Various studies have been devoted to the problems of settlement and metamorphosis in bivalves (see Ritchie and Menzel, 1969;Lutz et al., 1970; Hidu and Haskin, 1971; Keck et al., 1971; Veitch and Hidu, 1971; Andrews, 1979; Mackie, 1984; Baker and Mann, 1994a,b). Prytherch (1924,1931 and 1934) found that copper induces metamorphosis in larvae of C. virginica. Since then, a great number of publications have documented natural and artificial chemical substances capable of inducing metamorphosis (see Hadfield, 1984; Pawlick, 1992). The special role of neuroactive compounds (L-DOPA, dopamine, serotonin, epinephrine, norepinephrine) has been shown by Coon and Bonar (19861, Coon et al. (1985, 1986, 1990), Shpigel et al. (1989) and Beiras and Widdows (1995), and the importance of biofilms has been demonstrated by Weiner et al. (1989). In oysters, neurotransmitters may induce metamorphosis without fixation, producing “cultchless spat” which are anatomically identical to naturally set spat (see photographs in Coon et al., 1985, p. 217; Coon and Bonar, 1986). These advances are of considerable practical significance in toxicological bioassays with pediveligers (see Section 3.2.4.5).
24
E. HIS, R. BEIRAS AND M .
N. L. SEAMAN
Figure 3 Crassostrea gigas pediveliger metamorphosing (a), and metamorphosed post-larva (b). Gr: gill rudiment; F foot; P d prodissoconch shell; D: dissoconch shell. (Experiments by His et al., 1997b.) Metamorphosis was induced by epinephrine.
25
THE ASSESSMENT OF MARINE POLLUTION
2 Larva begins swimlcrawl
Larva searches for surface cue
Larva detects
Cementation to substratum
Metamorphosis to juvenile oyster
Figure 4 Two-cue model of microbial induction of oyster settlement (1-3) and metamorphosis (45). A soluble cue induces searching behaviour and then a surface cue induces attachment. (From Weiner et al., 1989.)
2.3. Larval rearing in the laboratory
2.3.1. Laboratory conditioning of spawners To conduct routine bioassays with bivalve embryos and larvae it is desirable to be able to obtain ripe adults year-round. In some species, mature adults can be found in the field during many months, and in others it is possible to condition them out of season. This was first discovered by Loosanoff (1945), when he attempted to clear Eastern oysters from a growth of sponges by keeping them in running seawater at elevated temperatures and found that by doing so he had induced gametogenesis in winter. He subsequently demonstrated that the duration of conditioning can be prolonged or shortened by manipulating the temperature, and that the technique can be applied to Mercenaria mercenaria as well (Loosanoff and Davis, 1950, 1951, 1952, reviewed 1963). In the case of Crassostrea virginica,gonad maturation can be obtained in the winter by gradually raising the water temperature to 20°C or 25°C; it is even possible to transfer oysters directly from a temperature of 5°C in the field to 20°C in the laboratory. The duration of conditioning depends on the temperature, as long as it is maintained within the range of 15°C to 30°C. Prolonged incubation at 27°C to 28°C can, however, inhibit or delay spawning (Ruddy et al., 1975). At the end of the reproductive season, after the adult oysters have used up their reserves, it is difficult or impossible to induce them to mature
26
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
again immediately. To avoid this inconvenience, Loosanoff and Davis (1963) took maturing oysters from Long Island Sound and transferred them to the cooler waters of Maine, where their maturation was arrested. They were subsequently able to induce final maturation and spawning by manipulating the temperature in the laboratory. Spawning can be delayed this way for 6 to 8 weeks; beyond this time limit, progressive resorption of the gonad makes spawning impossible. To circumvent the problem, Loosanoff and Davis induced oysters to spawn early and then transferred them also to Maine, where they resorbed the residual sexual products and reconstituted their glycogen reserves, thus beginning a new reproductive cycle and becoming susceptible to conditioning again. This two-sided approach to conditioning (acceleration and inhibition of gametogenesis) can be applied to C. virginica and M. mercenaria (Davis and Chanley, 1955a; Loosanoff and Davis, 1963), and also to C. gigus, and it is of great practical relevance to toxicological bioassays. In Mufinia faterafis it is equally possible to obtain mature adults throughout the year by delaying or accelerating sexual maturation with the same manipulations of temperature (Calabrese and Rhodes, 1974); apparently it works even better with the coot clam than with the other species (Burgess and Morrison, 1994). Conditioning has also been achieved with Ostrea edufis (Dannevig, 1951; Abou-Ela, 1960), where adults can be transferred from environmental temperatures of 1°C to 5°C to conditioning tanks at 21°C (Walne, 1966). Supplementary feeding with cultured algae (Tetrasefmis suecica) during conditioning improves the subsequent survival of the larvae (Helm et af., 1973). It is equally possible to delay spawning in mussels after initiation of gametogenesis by transferring them to 4°C or 5°C one month before the time of spawning and by supplementing their nutrition with cultured phytoplankton (Bayne, 1965, 1975; Riisgard et af., 1980; Dixon and Prosser, 1986). Mussels have been thermally conditioned by Mitchell et al. (1985) by maintaining them for 3 weeks in unfiltered seawater at 14°C. Spawning was obtained by Bayne (1975) one month early in adults which had initiated gametogenesis by keeping them at 15°C (5°C higher than the ambient temperature) and by feeding them with Tetrasefmis suecica (at least 2.2% of the meat weight per day). The time required for gonad maturation depends on the stage of maturation at the onset of conditioning as well as on the conditioning temperature. On the other hand, stress and sub-optimum rearing conditions (particularly with respect to nutrition) reduce the fertility of mussels, as well as the quality of the gametes and larvae (Bayne, 1975, 1985). The methods for conditioning scallops, oysters and clams in hatcheries have been reviewed by Utting and Millican (1997): “of particular
27
THE ASSESSMENT OF MARINE POLLUTION
Table 3 The fecundity of bivalves used in bioassays. Species
Crassostrea gigas Crassostrea virginica Mytilus edulis
Mytilus galloprovincialis Mercenaria mercenaria
Mulinia lateralis Ostrea edulis
Fecundity (eggs per female; x106) 15 to 114 15 to 114.8 10 to 66.4 (M = 28.8) 10 >0.5 1.2 to 7.6 2 no data 0.38 to 18.83 0.6 to 13.2 (A4 = 6.3) 8 to 39.5 ( M = 24.6) 1.4 >7 (M = 3 to 4) 0.09 to >1 0.616 to 1.155 0.1 (1year old) to 1.5 (7 years old)
Reference Galtsoff, 1964 Galtsoff, 1964 Davis and Chanley, 1955a Lubet, 1959 Bayne, 1975 Bayne et al., 1978 Thompson, 1979 Lubet, 1973 Ansell, 1967 Bricelj and Malouf, 1980 Davis and Chanley, 1955a Knaub et al., 1987 Calabrese, 1969a,b Cole, 1941 Millar, 1961 Walne. 1974b*
M: mean value; *estimated from the number of incubated larvae.
importance is the optimization of the quantity and the quality of microalgae diets that are provided during the broodstock conditioning”. Flow-through systems are generally better than recirculation systems, because “natural phytoplankton is a valuable component of the diet during broodstock conditioning” (p. 46). Gametogenesis is improved by preconditioning in high food regimes before the actual conditioning at elevated temperature. Egg quality and larval viability depend largely on an adequate accumulation of lipid reserves in the ova. 2.3.2. Spawning
To have a suitable number of replicates, or to be able to study a range of concentrations in toxicity tests, a great number of larvae are often required in bioassays. Most of the bivalves used in bioassays are characterized by high fertility. Flat oysters are an exception; being larviparous, not only are they less fecund, but they also have much larger oocytes (100 to 150pm in diameter in Ostrea edulis) than the other species. The data on fecundity (Table 3) are incomplete, especially with regard to Mytilus galloprovincialis. The data on oysters have been reviewed by Galtsoff (1964), those on Mercenaria mercenaria by Davis and Chanley
28
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
(1955a) and by Eversole (1989), Mytilus edulis by Thompson (1979), and Mulinia lateralis by Calabrese (1969a,b). The number of eggs released by any individual always depends on its size and physiological condition, as evidenced by significant correlation between size and number of eggs spawned in Crassostrea virginica and M. mercenaria. Stress and pollution, however, are known to reduce the fertility in mussels (Bayne, 1972; Zaroogian et al., 1979; Zaroogian and Morrison, 1981). The reproductive physiology of Crassostrea virginica has been studied in detail by Galtsoff (1938a,b, 1940, 1964). He determined that spawning in female oysters involves complex mechanisms with the participation of the nervous system, the gills, the mantle and the adductor muscle. The genital pore is situated near the anus. Oocytes are expelled into the exhalent part of the pallial cavity, then pass through the gills and accumulate between the gills and the mantle before being expelled by muscular action through the inhalant part of the pallial cavity into the surrounding water. The shell movements associated with spawning are “so characteristic that they cannot be mistaken for any other type of muscular activity” (Galtsoff, 1964), and they can serve to monitor the frequency and duration of reproduction in the field (His, 1975, 1976). Ova are violently expelled at regular intervals, forming whitish clouds near the spawning female. In the male, however, the sperm leave as a steady stream with the exhalent current, almost without contraction of the adductor muscle (Galtsoff, 1938a,b). In the water column, the gametes emit pheromones (termed “fertilizins” by Galtsoff, 1938a,b, 1940) which induce spawning in members of the opposite sex. Sperm also release a nucleoprotein termed diantline which facilitates spawning in both sexes by relaxing the adductor muscle, enlarging the gill ostiae, augmenting ciliary activity, and thus increasing the pumping rate (Nelson and Allison, 1940). These observations probably apply to all oysters of the genus Crassostrea (Galtsoff, 1964). The presence of similar pheromones has been demonstrated in mussels by Lubet (1959). In Ostrea edulis the ova also pass through the gills into the inhalant part of the pallial cavity, but this does not give rise to any visible change in behaviour (Yonge, 1960). The ova are fertilized by sperm inhaled by the female and are subsequently retained within the pallial cavity for about one week, after which they are released (a phenomenon termed “swarming” by Dutch and German authors). In Mercenaria mercenaria and in Mytilus, sperm are liberated with the exhalent current in a whitish thread-like stream which rapidly dissipates and gives the water a milky appearance. The ova are similarly liberated by the females without particular valve movements. In the laboratory, spawning can be induced in mature bivalves by a
THE ASSESSMENT OF MARINE POLLUTION
29
variety of physical and chemical stimuli (Table 4, after Le Pennec, 1981; see also Mackie, 1984). In the case of Eastern oysters from South Carolina and Florida it is not possible to induce spawning in the laboratory by the usual methods (thermal stimulation and addition of sperm suspension), because they are often subjected to these stimuli for extended periods in their habitat (Maurer and Price, 1968). Gametes may be obtained from individuals that refuse to spawn naturally by “stripping” (i.e. teasing the gonad with a forceps), however this is not generally recommended (Woelke, 1961, 1966; ASTM, 1989; Widdows, 1993). These methods are discussed in Sections 3.1.3.2 and 6.3.2. 2.3.3. Fertilization In the oviparous species, with the notable exception of Mercenaria mercenaria, the eggs are somewhat irregular in shape at the time of release, becoming spherical after a few minutes. Egg sizes (Table 5) do not vary much among species, with diameters of 50 to 55 pm in Crassostrea, 60 to 70 pm in Mytilus, 70 to 73 pm in M. mercenaria, and about 50 pm in Mulinia lateralis (Lubet, 1959; Yonge, 1960; Loosanoff and Davis, 1963; Galtsoff, 1964; Fretter and Graham, 1964; Calabrese and Rhodes, 1974; Purchon, 1968; Mackie, 1984; Eversole, 1989). In M. mercenaria the egg has a gelatinous envelope which swells to a diameter of 163 to 170 pm,and which may persist to the blastula stage (Loosanoff and Davis, 1950). Fertilization is practically instantaneous when the egg meets the spermatozoa (Allen et al., 1988). According to Lu (1986, in Allen et al., 1988) in Crassostrea gigas the second polar body is extruded 50 minutes after fertilization at a temperature of 18”C, after 43 minutes at 2VC, and after 32 minutes at 25°C. 2.3.3.1. Polyspermy. Galtsoff (1964, p. 343) states, “A few seconds after the sperm head touches the egg surface a thin transparent fertilization membrane is elevated”. Alliegro and Wright (1983) have, however, questioned the existence of this phenomenon: “It is possible that the fertilization envelope reported by Galtsoff can be attributed to the refractile nature of the fertilized egg surface seen with light microscopy.” Galtsoff also notes that two or more sperm (“supernumeraries”) may enter the egg before the barrier is developed when sperm suspensions are thick; this leads to polyspermy and results in an abnormal embryo (Turner, 1958; Loosanoff and Davis, 1963; Galtsoff, 1964). Determination and use of appropriate sperm to egg ratios significantly enhance the yields of viable larvae (Stephano and Gould, 1988). According to Stiles and Longwell (1973), in C. virginica the number of sperm has to approach 20 per egg before the incidence of chromosome and division abnormalities increases significantly, but this has been questioned by
30
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Table 4 Experimental methods used to induce spawning in marine bivalves (completed after LE Pennec, 1981). Method
Species
Reference
Stimulation due to transport
Crassostrea gigas
Imai, 1967
Unspecified temperature fluctuations
Crassostrea virginica C. virginica Mercenaria mercenarin Mulinia lateralis
Prytherch, 1924 Wells, 1927 Loosanoff, 1937b Calabrese, 1969b
Specified temperature fluctuations
Mytilus edulis C. gigas Mytilus galloprovincialis M. edulis
Bayne, 1965 Imai, 1967 Masson, 1975 Lutz and Hidu, 1979
Temperature fluctuations and addition of gametes
C. virginica C. virginica, M. mercenaria M. mercenaria several bivalves M. lateralis M. lateralis
Davis, 1953
Constant temperature and addition of gametes
C. virginica, C. gigas
Galtsoff. 1930
Salinity fluctuations
tsognomon califomicum
Ringwood, 1990,1991,1992
Electrical stimulation
M. edulis M. edulis M. edulis
Aboul-Ela, 1960 Iwata, 1950 Sugiura, 1962
M. edulis M. edulis
Iwata, 1951a,b Morse et al., 1977; Garland et al., 1986 Loosanoff and Davis, 1963 Loosanoff and Davis, 1963 Castagna et al., 1985
Chemical stimulation KCI 0.5 M
HzOz NbOH Injection of serotonin Pricking the adductor muscle
M. mercenaria C. virginica M. mercenaria, M. edulis M. edulis
Loosanoff, 1954 Chesnut et al., 1957 Loosanoff and Davis, 1963 Kennedy et al., 1974 Calabrese and Rhodes, 1974
M. galloprovincialis
Hrs-Brenko and Calabrese, 1969 Masson, 1975
Use of "Kraft mill effluent"
M. edulis
Breese et al., 1963
Addition of algae
M. edulis, Mytilus californianus, C. gigas, M. rnercenaria M. californianus
Breese and Robinson, 1981 Smith and Strehlow, 1983
Table 5 Sue of egg, size of early veliger, sue at time of metamorphosis and duration of larval development under laboratory conditions.
Species Crassostrea gigas Crassostrea virginica Mytilus edulis Mytilus galloprovincialis Mercenaria mercenaria Mulinia lateralis Ostrea edulis
Temperature ("C) 25 24 22 20 30-32.5 16 (11)
Egg diameter
Size of early veliger
Size at time of metamorphosis
(Pm)
(Pm)
(CLm)
-
76.8 68 68-75
303-320 310 275-315
96 100-120 76
260 215-300 298 ? 23 258
50-55
?
?
18 24*
77,8 ? 4
18 ?
20-25 20-22 18-20
-
73-75 70-73 ? 50 114-126
86 60 ?
170-190*** 208
175-240 21Cb230 20fA220 240-350 280-300
Duration of larval development (4 1619 17-21 12** 36-40 10-12 1 6 2 0 (34-38) ? ?
17-19 16 (6-8**) 6 6-8 10-16 15-16
Reference Gerdes, 1983b Beiras and His, 1994 Loosanoff and Davis, 1963 Davis and Calabrese, 1964 Davis and Calabrese, 1964 Bayne, 1965 Loosanoff and Davis, 1963 Sprung, 1984 Beiras and His, 1995a Aguirre, 1979 Loosanoff and Davis, 1963 Loosanoff and Davis, 1963 Calabrese and Rhodes, 1974 Walne, 1974b Loosanoff and Davis, 1963
*egg incubation at 20°C; **33"C. Data from references in italics. ***larval sue at liberation. Mytilus edulis (Bayne, 1965): temperature and corresponding duration of pelagic life in parenthesis.
32
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Staeger and Horton (1976) in C. gigas and Alliegro and Wright (1983) in C. virginica. In the Japanese oyster, the mean percentages of larvae developing to the D-shape stage increased until 7.3 X 10’ s p e d 1 0 0 eggs were used (Staeger and Horton, 1976), and in C. virginica the number of sperm entering eggs is restricted to one per fertilized egg at a sperm:egg ratio as high as 1OOO:l (Alliegro and Wright, 1983). In both cases gametes were obtained by stripping the gonads. In mussels, viable larvae can be obtained with variable sperm to egg ratios. The recommended sperm to egg ratio is in the order of 103:1in M. edulis (Sprung and Bayne, 1984) and M . galloprovincialis (Sedan0 et al., 1995). In M . californianus it varies between 25 and 200 sperm per egg, depending on the physiological state of the adults of both sexes (Cherr et al., 1990). In Mercenaria mercenaria the optimum ratio is between 1.3 and 2.5 X 103:1 (Bricelj and Malouf, 1980). Finally, Calabrese (1984) states that polyspermy is to be avoided in Mulinia lateralis, but without speclfying sperm densities. In their study on polyspermy in C. gigas Stephano and Gould (1988) found that, compared to eggs spawned naturally, those obtained by stripping are very susceptible to polyspermy upon immediate fertilization. This susceptibility decreases greatly when the eggs are held in seawater for 90 minutes before fertilization. The authors conclude that stripped eggs lack a “sperm block” present in naturally spawned eggs. Togo et al. (1995) obtained normal (monospermic) fertilization in Mytilus edulis when the eggs were fertilized at a sperm ratio of 5 X 103:1 within 30minutes after spawning. They recognized three mechanisms for blocking polyspermy in mussels. The first depends on a rapid depolarization of the egg’s plasma membrane, the second consists of the suppression of the acrosomal reaction, and the third is the blocking of contact or fusion between the plasma membranes of sperm and egg. 2.3.3.2. Ageing of gametes affer spawning. One of the prerequisites for bivalve embryo and larva bioassays is to begin with the best possible biological material, i.e. with a proportion of normal larvae in the controls as close to 100% as possible. “Oyster eggs undergo aging and lose their ability to be fertilized” (Galtsoff, 1964). Except in Mulinia lateralis (see Section 2.1.6), it is impossible to know in advance to which sex the individual adult belongs. Successful fertilization requires simultaneous spawning of at least one member of each sex; in general, however, one or more males spawn first, and only very rarely does a female spawn before any of the males has done so. According to Galtsoff (1964), concentrated sperm (presumably obtained by stripping) of C. virginica conserves its fertilizing capacity for at least 24 hours at lO”C,whereas diluted sperm at room temperature [probably 18°C to 20°C] loses its capacity within 4 or 5 hours. The number of dividing eggs drops to 60% after 5 hours, and to 20% after 10 hours. In C. gigas the sperm
33
THE ASSESSMENT OF MARINE POLLUTION 6-
I
0
I
20
I
I
40
L
I
I
M)
I
80
I
I
100
Hours from fertilization
Figure 5 The rates of cleavage and early development of embryos of Mytilus edulis at different temperatures. A,8°C (Bayne, 1965); 0, 18°C (Bayne, 1965); X , 20°C (Field, 1922, cited by Bayne, 1976); U, 19-22°C (Rattenbury and Berg, 1954, cited by Bayne, 1976). Stages of development as follows: 1, first polar lobe; 2, first cilia; 3, trochophore; 4,appearance of velar cilia; 5, appearance of shell gland; 6, prodissoconch I. (From Bayne, 1976.)
remain very motile for 1.5 hours after release at temperatures of 21°C to 25°C (Stephano and Gould, 1988);however, "if fertilization was delayed for more than 60 to 90 minutes after gamete liberation, the proportion of larvae which developed was greatly reduced" (Walne and Helm, 1974, p. 1; Helm and Millican, 1977, p. 2). In Mytilus edulis, diluted sperm lose their motility within 1 or 2 hours at room temperature, but maintain their fertilizing capacity for several hours at 15°C;fertilization of eggs is possible after 4 to 6 hours at 18"C, but at 15°C the fertilization rate drops to 40% within 6 to 11 hours after spawning (Sprung and Bayne, 1984). In M. galloprovincialis, fertilization was still possible within 8,7 and 4 hours after spawning at 10°C, 14°C and 18"C, respectively (Sedan0 et al., 1995).
2.3.4. Physical requirements of bivalve larvae The range of environmental conditions over which embryogenesis is possible will depend on the location of the adult population (Loosanoff, 1954; Bayne, 1976). The duration of the embryonal and larval phases depends mainly on temperature (Figure 5, Tables 5 and 6), and also on salinity and other environmental factors, including food availability
Table 6 Embryonic and larval development event times of oviparous marine bivalves. Crassostrea gigas, 25°C (Cahnn, 1950); in parenthesis, 20°C (Tazawa et al., 1985). Crassostrea virginica, 23-25°C (Galtsoff, 1964). Mytilus edulis, 17°C and 2 5 0 ~(Armstrong and Millemann, 1974). Mytilus galloprovincialis, 10-17°C (Aguirre, 1979); in parenthesis, 20°C (Masson, 1975). Mercenaria mercenaria, 22°C (Loosanoff and Davis, 1950). Mulinia lateralis, 20-25°C (Calabrese and modes, 1974).
Developmental stage
Crassostrea gigas
Crassostrea virginica
Mytilus edulis
Fertilized egg 1st polar body 2nd polar body + polar lobe 2-cell 4-cell 8-cell 16-cell 32-cell 64-cell ciliated blastula trochophore veliger
0 min 50-60 min 70 rnin
Omin 25-52 rnin 40-65 min
20 min
100 min
45-72 min 52-120 rnin 55-195 min
(420) 480 min 24h (15) 48 h
390 min 8-9 h 24 h
Omin
Mytilus galloprovincialis
Mercenaria mercenaria
Mulinia lateralis
0 min (80 min)
0 min
Omin
90 min 180-210 min
45 min 90 min
(24 h) 48 h
360 rnin 12h 24-36 h
40 min
65 min 90 min 120 min 150 rnin 180 min 220 min 450 min 16-19 h 40 h
9h 15h
Table 7 Effects of temperature, salinity and combined effects of temperature and salinity on survival and growth of bivalve 1arvae. Salinity (psu)
Temperature ("C) Species
Min.
Max.
Crassostrea gigas
15"
34"
Crassostrea virginica
17.7"
Mytilus edulis
10" 14" 5"
Opt.
Min.
Max.
15
39
30" 30-32.5"
15
15-20'
15 24 15
Opt. 25 19-27 30 2&27
T/S opt.
30"/30 30"/18-35
Mytilus galloprovincialis
22" 20" 20"
Mercenaria mercenaria
20" 12.5" 25-27.5" 2CL25" 10.5-15 25-30" 12.5"
Mulinia lateralis
20"
30"
27.5"
30"
30" 25-27"
20"
Ostrea edulis 17.5"
40 33 40
30-32 25-30 25-30 35 30-35 27
20
30
7
33
32.5 25-30
20
22.5-27
18"/27 20"/25-30 20'135 20°/30 21.5-33'122-31 22.5-27.5"/20-35 20-26'123-32 25"130
References Helm and Millican, 1977 Nell and Holliday, 1988 His et al., 1989 Davis and Calabrese, 1964 Lough, 1975 Bayne, 1965 Lough, 1974 Hrs-Brenko and Calabrese, 1969 His et al., 1989 Hrs-Brenko, 1977 Davis and Calabrese, 1964 Lough, 1975 Calabrese, 1969c Lough, 1975 Morrison and Petrocelli, 1990 Robert et al., 1989 Davis and Calabrese, 1969 Davis and Ansell, 1962
36
E. HIS, R . BEIRAS AND
M. N. L. SEAMAN
(Calabrese, 1969c; Mackie, 1984; Widdows, 1991). The time required to attain the D stage in different species is shown in Table 6. It must be pointed out that the data of Cahnn (1950) are not very accurate, because C. gigas attains the veliger stage within 24 hours, the same as C. virginica, at least at 25°C (see also Loosanoff et al., 1966; Helm and Millican, 1977; Tazawa et al., 1985). Moreover, it should be kept in mind that in bivalves, particularly in oysters, development is not highly synchronous (Stephano and Gould, 1988). The influence of various environmental effects on Crussostrea virginica larvae has been the subject of a modelling study by Dekshenieks et al. (1993). 2.3.4.1. Temperature. Embryos are generally more sensitive to environmental factors than larvae (see Section 4.2.2). Thus with regard to temperature the cleavage stages in M . edulis require a slightly narrower temperature range than the shelled larvae (Bayne, 1976; see Sastry, 1979); similar observations have been made concerning Mercenaria mercenaria (Loosanoff, 1954; Kennedy et al., 1974). In Mytilus galloprovincialis the embryos do not develop normally above 20"C, although this temperature represents the optimum for larval growth (Hrs-Brenko, 1977; His et al., 1989). In general, the rate of development increases with temperature as long as it remains within the range of tolerance (Sastry, 1979). 2.3.4.2. Salinity. As in the case of temperature, development of the eggs requires a narrower salinity range than survival and growth of the larvae (Bayne, 1976,1983). In the case of C. virginica the minimum salinity at which the eggs will develop is determined by the salinity at which the broodstock was kept prior to spawning (Davis, 1958; Davis and Calabrese, 1964). Salinity tolerance also depends on genetic factors (Newkirk et al., 1977; Newkirk, 1978; Widdows, 1991) and on interactions with temperature (Table 7). Studies on the combined effects of temperature and salinity show that, compared to nutrition and temperature, salinity has relatively little effect on larval growth and development. Significant interactions occur only at the extreme limits of the tolerance ranges of temperature and salinity. Within these limits, growth depends on temperature and food (Widdows, 1991). 2.3.4.3. Oxygen. The respiration of bivalve larvae has been studied by Gerdes (1983b), Tazawa et al. (1985) and Hoegh-Guldberg and Manahan (1995) in Crassostrea gigas, by MacInnes and Thurberg (1973), Widdows et al. (1989), and Baker and Mann (l992,1994a,b) in Crassostrea virginica, by Riisgard et al. (1980, 1981), and Wang and Widdows (1991) in Mytilus edulis, and by Morrison (1971) in Mercenaria mercenaria. According to Morrison (1971), hypoxia does not affect larval growth except at oxygen concentrations below 4 mg 1-I. Hyperoxia, on the other hand, negatively affects growth at 13.7 mg 1-I (180% saturation) in larvae
THE ASSESSMENT OF MARINE POLLUTION
37
of Mercenaria mercenaria (Huntington and Miller, 1989). Settlement was reduced significantly in hypoxic treatments, as compared to normoxic treatments, and no settlement took place in anoxic conditions in Crassostrea virginica larvae (Baker and Mann, 1992, 1994a,b). In mussels, both embryos and early prodissoconch larvae developed and grew normally at Po, values greater than or equal to 3.16 kPa (Wang and Widdows, 1991; see also Widdows, 1991, p. 158). Aeration in the presence of an antibiotic (1 mg 1-' erythromycin) increased larval survival in Crassostrea rhizophorae larvae at high temperature (30"C), but reduced individual growth and total biomass (Lemos et al., 1994). Hoegh-Guldberg and Manahan (1995) found that the conditions to which bivalve larvae are subjected in the small respirometers used in respiration studies negatively affect larval metabolism. From a practical point of view, artificial oxygenation of the small chambers used in these tests is not recommended, because resulting turbulence inhibits larval growth and survival (Helm and Spencer, 1972; see also Loosanoff and Davis, 1963, p. 38). 2.3.4.4. Turbidity. The effects of turbidity on bivalve larvae have been studied by Davis (1960), Robinson (1983), and Huntington and Miller (1989) in Mercenaria mercenaria; by Davis and Hidu (1969b) and Dekshenieks el al. (1993) in C. virginica; by Davis and Hidu (1969b) in Ostrea edulis; and by Seaman et al. (1991, and unpublished data) in Mytilus edulis, M. galloprovincialis, Crassostrea gigas and 0. edulis. Natural turbidity values, up to a few hundred mg 1-' tend to be beneficial to larval growth (excepting polluted sediments), because adsorption and desorption processes on the surface of suspended particles may buffer and stabilize the water (Koke, 1993). In bioassays, however, the water used is usually filtered (mesh of 1pm, or less), and turbidity is not a factor of influence in most experiments (excepting sediment bioassays). 2.3.4.5. p H . The data on pH requirements of larvae are scarce. The pH should not be less than 6.75 for C. virginica, and 7 for Mercenaria mercenaria, and both species are unable to reproduce when the pH remains above 9. M. lateralis requires a pH between 7.25 and 8.25 for reproduction (Calabrese and Davis, 1966,1970; Calabrese, 1970b). Krassoi et al. (1996) maintained the pH between 7.8 and 8.4 in tests on Chlamys asperrima.
2.3.5. Nutritional requirements of bivalve larvae Nutrition plays a decisive role in the development of bivalve larvae, even though they are capable of surviving a few days without food (Loosanoff, 1954; Bayne, 1965; Millar and Scott, 1967; Ukeles, 1975; His and Seaman, 1992). With regard to the combined effects of salinity, temperature and
38
E. HIS, R. BEIRAS AND M.
N. L.
SEAMAN
nutrition, the last has the greatest effect, explaining 85 to 88% of the variance in growth in Ostrea edulis larvae (Robert et al., 1989), 64 to 75% in Mytilus galloprovincialis, and 54 to 70% in Crassostrea gigas (His et al., 1989). A great number of studies have been devoted to the nutrition of bivalve larvae (reviews by Ukeles, 1975; Bayne 1976, 1983; Sastry, 1979; Widdows, 1991; Boidron-Metarion, 1995). As far as bivalve larval bioassays are concerned, nutrition is a topic of paramount importance, because although it is not difficult to maintain adequate temperature and salinity ranges for an experiment, providing adequate food may be more difficult to put into practice. Imai et al. (1950) and Imai and Sakai (1961), for instance, succeeded in rearing larvae of C. gigas to metamorphosis in tanks by feeding them Monas sp., a colourless flagellate, at Milford. However, Loosanoff (1969) was unable to obtain any growth in larvae fed with the strain he received from Imai, and this species is no longer used in larval culture. 2.3.5.1. Phytoplankton. Marine bivalve larvae feed mostly on nanoplankton and are thus termed planktotrophic (Thorson, 1950). Veligers pass through three different trophic stages: the endotrophic period, during which nutrition depends exclusively on vitelline reserves; the mixotrophic period, in which both vitelline reserves and exogenous resources are used; and the exotrophic period with exclusive use of external food sources (Lucas et al., 1986; Boidron-Metairon, 1995). Despite studies with other feeds, it has become clear since the pioneering work by Loosanoff and Davis (1963) and Walne (1963) that phytoplankton is the principal food source for bivalve larvae; no food other than unicellular algae has been found to be entirely satisfactory for bivalve cultures (Webb and Chu, 1981). Food algae for culture of veligers must meet three criteria: adequate size (limited by the diameter of the mouth and oesophagus of the larvae), good nutritional quality and ease of cultivation. Some 50 species of algae have been tested for the purpose but only a dozen are generally used in bivalve larval culture (ChrCtiennot-Dinet et al., 1986). Not only does the nutritional quality vary among different algal species, but food quality can also vary during cultivation, and some may produce substances which are toxic to the larvae (Davis and Chanley, 1955b). For instance, Nannochloris cells may excrete a growth-inhibiting substance in great concentrations during the stationary phase of algal growth (Bayne, 1965). Methods to feed bivalve larvae were developed at Milford by Davis (1953) and Davis and Guillard (1958), and reviewed by Loosanoff and Davis (1963). The food consisted of a mixture of live algal flagellates (Calabrese and Davis, 1966) given at a rate of 0.01 ml packed cell volume per litre culture per day (Hidu, 1965). According to Calabrese
THE ASSESSMENT OF MARINE POLLUTION
39
(1970a) the algae used were Zsochrysis galbana, Monochrysis Zutheri and ChZoreZZa sp. Veligers are capable of feeding selectively (Davis, 1953). The work of Davis and Guillard (1958) and Walne (1963) has shown that bivalve larvae grow best when they are fed a mixed diet of two or more algal species (see also Bayne, 1983). Their quantitative and qualitative dietary requirements may change in the course of development (Davis and Guillard, 1958 Loosanoff and Davis, 1963). Veligers of Crassostrea virginica, for instance, are incapable of using Chlorella sp. during the first days of life (Babinchak and Ukeles, 1979), although they do utilize them at the age of about 1 week, at a size of 110pm (Loosanoff and Davis, 1963). Oyster larvae of the genus Crassostrea have particularly narrow requirements, and the number of algal species suitable as food for them is limited, Mytilus and Mercenaria are rather tolerant and Ostrea is intermediate. The methods of rearing larvae of Crassostrea gigas most frequently used nowadays follow Walne and Helm (1974), Helm and Millican (1977), and Utting and Spencer (1991). These methods may be used for rearing most species of bivalve larvae used in ecotoxicological bioassays. The best results were obtained by rearing D-larvae without aeration at densities of about 10 per ml in 1 litre of filtered and ultraviolet-sterilized seawater, and feeding them with an algal mixture of Zsochrysis galbana Parke and Chaetoceros calcitrans (Paulsen) Takano (50 cells of each per pl of larval incubation volume). The water is changed every 2 days and larvae up to a size of 120 pm are retained on a 45 pm nylon mesh-based PVC sieve, rinsed with freshly filtered seawater and put back into incubation. After the umbo has developed (usually at the age of 6 to 8 days) the cultivation densities may be reduced to 5000 or 6000 per litre and the feeding regime is changed to 3.3 cells per pl of culture of Tetraselmis suecica (Kylin) Butch plus 33 cells per p1 each of 1. galbana and C. calcitrans (representing approximately equal volumes of all three species of algae). At Arcachon we follow the same methods, but we have found that ultraviolet sterilization of the water is not indispensable, and we use stainless steel sieves which are sterilized at 180°C. 2.3.5.2. Bacteria and microencapsulated diets. Davis (1953), Hidu and Tubiash (1963), Millar and Scott (1967), Mengus (1978) and Prieur (1981), among others, have studied the importance of bacteria for larval nutrition. Widdows (1991, p. 151) concluded “there is little evidence that they play a significant role in meeting the nutritional requirements during larval growth”. Douillet (1993) found that oyster larvae fed on a bacterial strain received only 41% of their carbon requirements. There have also been experiments with inert foods (see Robert and Trintignac, 1997). The first investigations by Ukeles (1975) were not very encouraging. Chu er al. (1982, 1987) have been able to rear larvae of Crassostrea
40
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
virginica to metamorphosis, but in their experiments only 2 to 25% of the larvae attained the eyed stage, and 2 to 20% achieved metamorphosis. The development of a reliable artificial diet would represent an enormous progress in bivalve larval culture, because it would vastly reduce the time and effort required (Ukeles, 1975), while standardizing feeding of the larvae between laboratories. 2.3.5.3. Dissolved organic matter. Finally, bivalve embryos and larvae are also able to take up dissolved organic substances (review by Widdows, 1991). This may well be beneficial, but Widdows (1991) concludes that there is no evidence that larvae are able to grow and develop solely on dissolved organics. The issue of the uptake of dissolved organic matter is important with regard to bioassays because it may also be a factor in the action of various pollutants.
3. BIOASSAY METHODOLOGY As Calabrese (1984) has pointed out, the various techniques and methods of ecotoxicological investigation with embryos, larvae and adult marine bivalves have resulted from advances in aquaculture. There are various manuals and reviews describing in detail the technical precautions necessary for conducting ecotoxicological tests with seawater (types of material, cleaning and sterilization, preparatory steps, etc.). Recommended literature on bioassays with bivalve larvae includes the work of Woelke (1972), the standards of ASTM (1980, 1989), Calabrese (1984), and more recently Widdows (1993) and Krassoi et al. (1996). The general principle is to eliminate the influence of toxic materials and any type of hazardous conditions, which could invalidate the bioassay, other than the condition under study. A broad theoretical and practical knowledge of larval rearing and larval biology is an important prerequisite for performing embryo and larval bioassays. Valid ecotoxicological studies and proper interpretation of their results presuppose an understanding of the normal development of the test species and a mastery of the technical procedures for optimal laboratory rearing. 3.1. General methods
3.1.1. Seawater quality 3.1.1.1. Natural seawater. One common element for all types of toxicity tests is the need for seawater of excellent quality. The first studies with marine bivalve embryos and larvae were carried out at biological
THE ASSESSMENT OF MARINE POLLUTION
41
laboratories specialized in shellfish aquaculture and located on bays and estuaries where running seawater was obtained from an area in which oyster populations reproduced naturally. This was notably the case at Milford (Connecticut), where bioassays were pioneered by Davis (1961), Loosanoff and Davis (1963), and Hidu (1965). As pointed out by Loosanoff and Davis (1963) and Woelke (1972), the source of seawater must be located at a site which is not contaminated in any way, and where the salinity is stable year-round (or can be kept stable by pumping water at appropriate periods of the tidal cycle). In addition, materials that might release toxic substances must be scrupulously avoided. Most coastal areas are nowadays increasingly subject to anthropogenic impacts resulting in a deterioration of the “biological quality” of the seawater. One of the first examples was noted by Millar and Scott (1968) at the laboratory of Millport (U.K.), where an unknown toxic substance slowed the growth of Ostrea edulis larvae. Shortly thereafter, Helm (1971, p. 8) remarked that “at Conway, sea water varies considerably in quality both in the short term and seasonally”, and the biological quality of the seawater was improved by adding 1mg 1-l of EDTA (Utting and Helm, 1985). Brereton et al. (1973) and Boyden et al. (1975) showed that the poor growth of Crassostrea gigas larvae and the irregularities in oyster recruitment resulted from the presence of zinc in the water, the source of which was mines in the catchment area that drain into the bay at which the biological station was located. This corroborated the observations of Wilson (1951) and Wilson and Armstong (1961) concerning the variability of the “biological quality” of coastal waters (cf. Section 1.3). The volume of the vessels employed in bioassays has been greatly reduced and can now be as small as 3ml (see 3.2.3), so that large quantities of seawater are no longer required and bioassays are not restricted to laboratories located on the seashore. In addition, seawater may be obtained offshore beyond the region of anthropogenic impact in order to obtain it “from an area known to support a healthy naturally reproducing population of bivalves” (ASTM, 1989, p. 339). Seawater filtration. Following the work of Loosanoff and Davis (1963) and Walne (1966, 1974b), most authors agree that bioassays should be conducted only with filtered seawater sterilized with ultraviolet light (Beaumont and Budd, 1982; Beaumont and Terpes, 1984; ASTM, 1989; Hoare et al., 1995a,b; Krassoi et al., 1996). The porosity of the filter is usually equal to, or less than, 1 pm. The ASTM proposes 0.45 pm, but many authors prefer 0.2 pm (Utting and Spencer, 1991; Ringwood, 1991; Hoare et al., 1995a,b; studies by His and coworkers at Arcachon). Filters of 0.1 pm have been used as well (Lemos et al., 1994). Simple filtration systems may prove equally useful. For example, Armstrong and
42
E. HIS, R. BEIRAS AND
M. N. L. SEAMAN
Millemann (1974) were able to cultivate various larval stages of Mytilus edulis in seawater which had merely been passed through a sand filter. In contrast to almost all other workers, Garland et al. (1986) found that 0.2 pm filtered seawater led to poor growth in bivalve larvae. lbenty years of experience with bivalve embryos and larvae of Crassostrea gigas, Ostrea edulis and Mytilus galloprovincialis at Arcachon have shown clearly, however, that 0.2 pm Htration of seawater is perfectly adequate for toxicity bioassays of 24 or 48 hours duration, as well as for tests of larval growth over periods of several weeks, and for studies of metamorphosis, as long as the seawater is obtained from adequate sites. Filtration of the water should be carried out immediately prior to the incubation, no more than 1 or 2 hours before the test is initiated. In the case of monitoring of in situ water quality, on the other hand, “ideally, filtration is not recommended, because it may remove some of the dissolved contaminants” (Thain, 1991, p. 4), and because it will definitely remove fine suspended particles to which pollutants are likely to adsorb. In such cases good results can be obtained by simply sieving the water through a 32 pm stainless steel sieve in order to eliminate the largest suspended particles and possible predators of the larvae (His and Beiras, 1995). Bacteria and antibiotics. Because the comparatively small volume of the incubation vessels (usually 1 litre at most) favours bacterial growth, and the bacterial populations eliminated by filtration are rapidly reconstituted (Walne, 1958; Lemos et al., 1994), many authors have added antibiotics to the seawater even after filtration and ultraviolet sterilization (cf. Davis and Chanley, 1955b; Walne, 1958). Hidu and Tubiash (1963) recommended the systematic addition of antibiotics (dihydrostreptomycin-streptomycin sulphate = Combistrep at 2 ppm) in larval cultures of Crassostrea virginica and Merceaaria mercenaria. They thought (p. 25, op. cit.), “an antibioticinduced bacterial flora. . . may be utilized by larvae as a food source” (see also Section 2.3.5.2). Similarly, Calabrese and Davis (1967, p. 12) used sulfamethazine soluble powder at 0.33 ppm, Brereton et al. (1973) added 0.3g1-1 of penicillin to the seawater, and Lemos et al. (1994) added 1mg 1-’ of erythromycin. The latter, however, found that although growth is thus improved by antibiotics in Crassostrea rhizophorae cultured at high temperatures and low salinity (30°C and 12 to 20 psu), this is later followed by increases in mortality. Chloramphenicol (5 mg 1-1 of 1pm filtered seawater) was used in larval cultures of Spisula solidissima by Thurberg et al. (1975). Le Pennec and coworkers (Le Pennec and Prieur, 1972; Le Pennec et al., 1973) have shown that antibiotics (aureomycin, erythromycin, chloramphenicol and sulfamethazine) may influence the metabolism and growth of bivalve larvae. This may affect the accuracy of
THE ASSESSMENT OF MARINE POLLUTION
43
bioassays designed to test a pollutant’s possible reduction of larval growth. Although Woelke (1966) stated that the seawater should be used within 3 hours of sampling, we have obtained excellent results at Arcachon with seawater sampled offshore and stocked for 24 hours at ambient temperature (18°C to 19OC), but filtered immediately before use (see above). On the other hand, Thain (1991) apparently did not obtain good results with seawater sampled offshore during winter and stocked at -18°C in acid-washed bottles, as he states that “even 60 percent abnormalities - in controls - is acceptable” (Thain, 1991, p. 7). We shall see below that most workers do not accept such high levels of abnormalities in the controls. Klockner et al. (1985) also encountered high levels of abnormality in the controls when they reared larvae in “aged (4years, filtered - Seitz K 15) and pasteurized (8OOC for 1 h) sea water” (p. 2). 3.1.1.2. Artificial seawater. Seawater quality problems can be overcome with artificial seawater made by dissolving the mineral constituents of oceanic water in distilled water. This also prevents the complex interactions between pollutants and organic matter that may affect toxicity. Contrary to the opinion of Thain (1991), artscial seawater is perfectly suited for bivalve embryo bioassays (e.g. Krassoi, 1995; His et al., 1997a). The seawater formula of Zaroogian et al. (1979), used successfully by Calabrese et al. (1973), Calabrese and Nelson (1974), Utting and Helm (1985) and recommended by ASTM (1989), gives excellent results in embryo-larva bioassays of 24 hours duration in Crassostrea gigas, or 48 hours in Mytilus galloprovincialis. It may be prepared using reagent-grade chemicals and deionized water aerated before conducting the test. EDTA, which chelates metals and some pollutants, may be omitted (His et al., 1997b, p. 352). Besides Zaroogian’s formulation, various other seawater media have been used to rear bivalve larvae (Table 8), and Krassoi (1995) has recently used several different formulations to grow larvae of Chlamys asperrima and Saccostrea commercialis. Kester et al. (1967), who proposed a modification of the formula by Lyman and Fleming, stressed that artificial seawater should be made up with reagents of known composition; this is not always the case, as the composition of the salt used by Courtright ef al. (1971) is unknown. The formula used by Chang et al. (1996) is a simplification of the medium of Zaroogian et al. (1979), containing only six mineral salts (NaC1, KC1, CaCl,, MgCl,, MgS04 and NaHCO,), instead of eleven. 3.1.2. Broodstock Spawners in good physiological condition are indispensable; they must be able to mature in an unpolluted environment and in excellent nutritional
Table 8 Artificial seawater (ASW) formulations used for the rearing of bivalve larvae.
Artificial seawater Lyman and Fleming, in Sverdrup et al. (1949) Wood (1961) Zaroogian et al. (1969)
Leslie coarse hide salt Instant Ocean sea salt mixture Instant Ocean Aquarium Systems, Mentor OH Different ASW formulations
Bivalve species Ostrea edulis 0. edulis Crassostrea virginica Mercenaria mercenaria Crassostrea gigas C. gigas Mytilus galloprovincialis C. virginica Mytilus edulis C. virginica C. gigas C. virginica Chlamys asperrima Saccostrea commercialis
Reference Helm (1971) Millar and Scott (1968) Calabrese et al. (1973) Calabrese and Nelson (1974) Utting and Helm (1985) His et al. (1997a) His et al. (1997a) La Roche et al. (1970) Courtright et al. (1971) Sigler and Leibovitz (1982) Chang et al. (1996) Chang et al. (1996) Krassoi (1995) Krassoi et al. (1996)
Remarks Rearing Rearing Bioassay Bioassay Rearing Bioassay Bioassay Bioassay Bioassay Bioassay Bioassay Bioassay Bioassay Bioassay
THE ASSESSMENT OF MARINE POLLUTION
45
conditions, whether in the field or in the laboratory (ASTM, 1980, 1989; Krassoi et al., 1996). The best results are usually obtained during the regular spawning season by using adults from areas where bivalves are cultivated (see Section 2.2). On the other hand, when bioassays need to be conducted out of season, ripe adults may be obtained by conditioning techniques (see Section 2.3.1). Most commercial hatcheries are capable of supplying mature oysters, Crassostrea spp., which are the ones used most extensively in bioassays (see Table l ) , to laboratories lacking the facilities for conditioning bivalves. A word of caution is in order, however, because the adults from commercial hatcheries are sometimes overripe, i.e. they have been conditioned for too long without spawning. “. . . Oysters resorbing their unspawned products . . . result in poor bioassay material” (Woelke, 1972, p. 31), with abnormality levels near, or above 20% in the controls, thus generating results of questionable validity (see Section 3.1.3.2). As noted by Eertman et al. (1993, p. 38) in the case of C. gigas embryos and larvae, “despite the oysters being conditioned for spawning, the autumn and winter periods seem less suitable for performing toxicity tests” as larval abnormalities reached 40% in October and March. Moreover, many hatcheries clean out their facilities in autumn and winter and are therefore unable to provide spawners during this time. If the spawners are subjected to unfavourable environmental conditions, this will affect the quality of the gametes and influence the results of the test (Widdows, 1993; see Section 2.2.2). In Mytilus rrossulus and Mizuhopecten yessoensis, exposure of the adults to sublethal doses of heavy metals (Cu and Zn) prevents the formation of gametes capable of normal embryonic development (Karaseva and Medvedeva, 1993). On the other hand, adaptive mechanisms such as the formation of metalloproteins have also been found, particularly in the case of cadmium (see Webb, 1979), and these metalloproteins are later found in the larvae upon their exposure to the contaminant (e.g. Pavicic et at., 1984, 1994a,b; Roesijadi et al., 1996, 1997; Ringwood and Brouwer, 1995). Similarly, Hoare et al. (1995a,b) found that Mytilus edulis embryos obtained from spawners originating from a site polluted by heavy metals had higher tolerance to copper than those spawned by adults from a non-polluted area; tolerance appeared to be “maternally determined”. There have been attempts to cryopreserve bivalve larvae in order to dispense with the need to obtain mature adults, but little progress has been made since the first experiments by Lannan (1971; see also Renard, 1991). It has usually not been possible to obtain adequate percentages of normal D-larvae (80% or higher) from cryopreserved gametes. Nevertheless, McFadzen (1992), Cleary et al. (1993) and McFadzen and Cleary (1994) have performed bioassays with veligers of C. gigas cryopreserved 24 hours
46
E. HIS, R. BEIRAS AND
M. N . L. SEAMAN
after fertilization, and of Tapes philippinarum 48 hours old. Contrary to the authors’ contention, it is not at all certain that 48-hour cryopreserved larvae are more sensitive as test organisms than the embryonic stages favoured by most other workers. More recently, Chao et al. (1997) have proposed the use of C. gigas and Meretrix lusoria cryopreserved late embryos and early larvae in aquaculture management programmes on the basis of survival rates from 62 to 84%. Future progress may yet facilitate the development of routine bioassays with cryopreserved material, making it possible, for example, to exchange genetically homogeneous material between laboratories. At present, however, cryopreservation techniques are not sufficientlyreliable and widespread for routine employment. We agree with Widdows (1993, p. 153), that the use of cryopreserved D-larvae “still remains to be fully evaluated”. 3.1.3. Spawning and fertilization Spawners must be used within 24 hours after being obtained from the field or the hatchery; they need to be carefully cleaned of fouling organisms and scrubbed in seawater. Two-year-old animals are preferable in the case of oysters, in order to have a fair chance of obtaining both males and females (see Section 2.2.1). 3.1.3.1. Induction of spawning. Gametes may be obtained either by stimulating the animals to spawn, or by “stripping” the gonad. The former method, recommended by Woelke (1972), has been adopted by ASTM (1989) in Crassostrea gigas, C. virginica, Mytifus spp. and Mercenaria mercenaria. Mature adults are placed in 1 litre beakers with filtered seawater and stimulated to spawn by varying the temperature. According to ASTM (1989, Table 2, p. 341) the temperature may be raised by 5°C to 10°C above the temperature of conditioning, but it is not supposed to exceed 20°C for mussels. On the other hand, bivalves may be subjected in the field to temperatures which fluctuate from more than 30°C (at low tide and strong solar irradiation) to 20°C at high tide. His and Beiras (1995) and His et al. (1997a) have found that 30 minute periods at temperatures alternating between 18°C and 28°C serve very well to induce spawning in both C. gigas and Mytilus galloprovincialis. Dixon and Prosser (1987) induced M . edufis to spawn at 30”C, and C. Bittkau (personal communication) obtained healthy gametes of the same species from the Baltic in winter (ambient temperature 5°C) after exposing them to fluctuations between 18°C and 28°C immediately after field sampling. An additional spawning stimulus may be provided by adding a suspension of sperm in filtered seawater to the beakers containing the mature adults (obtained by stripping the gonads of a male). To avoid
THE ASSESSMENT OF MARINE POLLUTION
47
undesired fertilizations in such a case, it is also possible to use a suspension of previously frozen sperm which has been brought to ambient temperature. Ringwood (1992b) used heat-killed sperm to stimulate spawning in Zsognomon californicum. 3.1.3.2. Stripping the gonad. With regard to the production of gametes by stripping the gonad, Allen et al. (1988, p. 7) states, “during our research, we investigated the conventional wisdom that stripped eggs of Pacific oysters are inferior to naturally spawned ones. We found nothing to support that belief”. Alliegro and Wright (1983) stress that stripping avoids the often tedious induction of spawning by other methods, which may take hours. According to Thain (1991), there is no important difference between the use of larvae obtained by stripping or by natural spawning in the conduction of ecotoxicological bioassays. Most other authors, however, disagree (Woelke, 1961, 1966; Loosanoff and Davis, 1963; Loosanoff, 1965b, 1969; Wilson, 1981; ASTM, 1989; Widdows, 1993). Although stripping of mature females can provide some uninjured ripe eggs that can be fertilized and develop normally, among the eggs obtained by induced spawning there are always considerably fewer abnormal ones than among eggs obtained by stripping (Loosanoff and Davis, 1963). As early as 1961, Woelke stated that “this practice [stripping] was dropped” because of “the frequency with which unsatisfactory results were achieved, apparently due to immature eggs” (Woelke, 1961, p. 115). Moreover, in Mercenaria mercenaria it is not possible to fertilize stripped eggs until they have been treated by a weak ammonia solution “to cause the germinal vesicle to break and the eggs become physically prepared for fertilization” (Loosanoff, 1969, p. 21). In Mytilus edulis “attempts to obtain mature (fertilizable) eggs by removing the ovary of apparently ripe females have met with unsatisfactory results, because it is only during their stay in the ovary, and during and following the act of spawning, that the ova become physiologically mature” (Longo and Anderson, 1969a, p. 73). In C. gigas, spawned oocytes are considerably less susceptible to polyspermy (fertilization of one egg by more than one sperm) than oocytes mechanically removed from the ovary (Stephano and Gould, 1988; Konar and Stephenson, 1995). In short, we agree with ASTM, that “use of eggs stripped from female bivalves is not recommended because it often results in an excess of poorly developed and malformed embryos” (ASTM, 1989, p. 343). 3.1.3.3. Number of parents. Opinions differ with regard to the number of parents whose spawn should be used in bioassays. Thain (1991) states that, for Crassostrea gigas, after stripping the gonads, “three batches of eggs are pooled and two batches of sperm are also pooled”. Johnson (1988), in Mytilus edulis, and Krassoi et al. (1996) in Chlamys
48
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
asperrima mixed the gametes of three males and three females; Beaumont et al. (1987) also preferred to use gametes of several spawners, both in M. edulis and Pecten maximus, “in order to avoid working with the very restricted genetic variation inherent in full or half sibs” (p. 300), as did Spangenberg and Cherr (1996) in M. californianus and Mercenaria mercenaria, and Laughlin et al. (1988, 1989) in M. mercenaria. Cherr et al. (1990) conducted separate fertilizations of several females of Mytilus californianus with sperm of two or three males, and Roberts (1987), in Crassostrea virginica and Mercenaria mercenaria, fertilized the eggs of several females with the sperm of one or several males. Finally, Ringwood (1991, 1992a,b) mixed the gametes of several males and females in Zsognomon californicum. ASTM (1980) recommended using gametes from two individuals of each sex in order to obtain a genetic mix of embryos. In contrast, ASTM (1989) recommends in order to verify differences in sensitivity between parental couples “ideally, the test should be conducted by subjecting progeny from each of at least three individual male-female pairings to each of the one or more control treatments” (p. 242). Woelke (1972) and Chapman et al. (1992) in Crassostrea gigas, and Granmo et al. (1988, 1989) in Mytilus edulis, recommended the use of a single male and a single female in a given bioassay; the same approach was chosen by Butler et al. (1990) in M. edulis: “eggs and sperm from two individuals were selected to give a single pairing, to minimize genetic variability in the test embryos” (p. 214). 3.1.3.4. Elimination of impurities in the spawn. Loosanoff and Davis (1963) recommended sieving the gametes to eliminate tissue fragments emitted by the bivalves along with the gametes. Butler et al. (1990) passed the gametes through a 1OOpm mesh, and Thain (1991) through 90pm. ASTM (1989) recommends a mesh of 75pm for the eggs and 37 pm for the sperm, the fertilized eggs being retained on a sieve of 37 pm for rinsing and counting. Krassoi et al. (1996) retained the eggs on a sieve of 45 pm, after passing them through a mesh of 100 pm. Granmo (1972), on the other hand, rinsed the eggs three times, allowing them to decant between rinsings, prior to fertilization. Spangenberg and Cherr (1996) simply rinsed the eggs in 0.45 pm filtered seawater after fertilization, and Ringwood (1992a,b) eliminated tissue debris by rinsing and gentle centrifugation. His and Beiras (1995) and His et al. (1997a) obtained excellent results by passing the eggs and sperm through sieves of 100 p m and 32 pm respectively, before proceeding to fertilize, without further rinsing or sieving after fertilization. 3.1.3.5. Counting the eggs. After spawning, and sieving or rinsing, the first step in a bioassay is to count the eggs. The method proposed for
THE ASSESSMENT OF MARINE POLLUTION
49
this purpose by Loosanoff and Davis (1963) has been adopted by most of the workers in the field (e.g. Woelke, 1972). It consists in placing the fertilized eggs in a graduated cylinder of 1 litre volume filled with filtered seawater and gently mixing the water column with a “plunger constructed by drilling holes in a disc of acrylic plastic or fibreglass of suitable diameter and attaching it to a PVC or acrylic rod of suitable length” (ASTM, 1989, p. 343; see Figure 7 in Woelke, 1972). Several samples are taken by pipette and counted under the microscope in a haematocytometer or well slide. A sample (usually 100 p1, or less) should consist of about 50 eggs, and, as the number of eggs in the cylinder is initially unknown, it may be necessary to adjust the volume of the samples accordingly, or to take a large sample from the graduated cylinder and dilute it according to need. The total number of eggs is computed by extrapolating the average number of eggs in the samples to the volume of the cylinder. 3.1.3.6. Fertilization. The sperm suspension should be obtained at approximately the same time as the eggs are collected: “Fresh actively-moving sperm are used to assure a normal fertilization of eggs” (Loosanoff and Davis, 1963, p. 35). To obtain a concentrated suspension, males should be made to spawn in the smallest possible volume of filtered seawater; some authors propose to count the sperm with a haematocytometer (Cherr et al., 1990; Ringwood, 1992b; Konar and Stephenson, 1995; Krassoi et al., 1996), particle counter (Thain, 1991), or by “turbidimeter readings previously calibrated with a hematocytometer” (Eyster and Morse, 1984, p. 642). ASTM (1989) admits that “precise sperm counts are unnecessary after one gains a little experience” (p. 343), even though optimum fertilization depends on the ratio of eggs to sperm (Section 2.3.3.1). This differs between species (and between authors), and it is also strongly influenced by gamete quality (potency of the sperm and receptiveness of the eggs, which are difficult to appraise). Most authors add a few millilitres of dense, freshly spawned sperm suspension and verify fertilization success under the microscope; 5 to 10 sperm (24 at most) should be visibly attached to each egg membrane (Konar and Stephenson, 1995; His and Beiras, 1995; His et al., 1997a). Because the biological quality of the gametes deteriorates rapidly after spawning (Section 2.3.3.2), fertilizations should preferably be undertaken within 30 minutes. After eggs and sperm have been united in one container, the actual fertilization takes place within minutes and the polar bodies are expelled within half an hour (Allen et al., 1988; see also Section 2.2.3). As a rule, fertilization of the gametes for bioassays should be undertaken as soon as the first females in the batch have spawned.
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
3.2. Bioassay procedures In assays with bivalve embryos and larvae, we may distinguish between acute toxicity tests, when response is recorded after short-term exposure (a few days or hours), and chronic toxicity tests, when the exposure period covers the majority of the larval life span (several weeks). The former tests frequently record embryogenesis success or larval mortality as the response, whereas the latter record sub-lethal responses, usually larval growth. Both kinds of tests are complementary. Acute toxicity tests are easier to standardize and provide a rapid assessment, whereas chronic toxicity tests are more sensitive and better approximate the environmental conditions. The execution of longer-term sub-lethal tests requires perfect capabilities in larval rearing - particularly with regard to the culture of food algae - as well as the availability of relatively sophisticated equipment, including a unit for the cultivation of algae. Sub-lethal tests cannot, therefore, be performed as a routine method in most marine biological laboratories, yet they remain an integral part of more detailed studies on particular chemicals or environmental hot spots. The first step in all bioassays - whether they are acute tests with embryos, chronic exposure studies lasting several days with larvae, or investigations on pediveligers and toxic effects on metamorphosis - is the production of gametes, embryos and D-larvae of excellent quality and the elimination of all artefacts resulting from any type of deficiency in the biological material. We have already reviewed the methods to obtain mature adults, induce spawning, and perform in vitro fertilization in Section 2.3. The principal causes of failure in bioassays with Chlamys asperrima have been summarized by Krassoi et al. (1996, Table 7). These include loss of broodstock, spawning of broodstock in holding tanks, poor spawning, poor fertilization success, low percentage of normal larval development (<70%), termination of development at the trochophore stage, and finally poor preservation of larvae preventing proper assessment of the biological response. These problems are common to all bivalve species used in marine ecotoxicological bioassays. 3.2.1. Spermiotoxicity test The use of sperm, previously exposed to a contaminant, to fertilize eggs in the absence of that contaminant has mostly been tested using sea urchins (e.g. Dinnel et al., 1982, 1987; Dinnel, 1995). Spermiotoxicity tests have rarely been used with bivalves. Renzoni (1973a, p. 11, experiment 4) “inseminated Crassostrea gigas and C. angulata eggs in clean filtered sea water with sperm previously maintained (3 h) in sea water with various concentrations of oil/emulsilier mixtures”. This author also studied the
THE ASSESSMENT OF MARINE POLLUTION
51
effects of water soluble extracts of three crude oils on sperm of Mulinia lateralis by exposing them for 3 hours and then using them to fertilize untreated eggs (Renzoni, 1975). His and Robert (1980) subjected sperm of C. gigas for 30 minutes to various concentrations of tributyl-tin (TBT), and subsequently fertilized eggs with that sperm in a toxicant-free medium. Eyster and Morse (1984) separately exposed sperm and eggs of the surf clam, Spisula solidissima, to silver salts, and then conducted fertilization experiments with untreated eggs and treated sperm, and with untreated sperm and treated eggs. Finally, Ringwood (1992b) studied the effects of 1 hour exposure to TBT, and to cadmium and copper salts on sperm of Isognomon californianum, using the method of Dinnel et al. (1987). Unlike sea urchin tests, spermiotoxicity tests with bivalves are not more sensitive than embryo-larval tests (see also Section 4.2.2).
3.2.2. Embryo-larval bioassay The progress of embryogenesis from fertilized egg to straight-hinge D-larva is the biological response most frequently used in bivalve bioassays. Therefore, the methodological details of the embryo development bioassays will be discussed in depth. 3.2.2.1. Age of the fertilized eggs at the beginning of the test. There is some disagreement as to the time that should pass between fertilization and the beginning of the bioassay. The ASTM standard states, “embryos should be tested within an hour of spawning and kept suspended during the period by frequent agitation with a perforated plunger” (ASTM, 1989, p. 252). Woelke (1972), Cardwell et al. (1977b), Morgan et al. (1986), Johnson (1988) and Thain (1991) used an interval of between 1 and 2 hours, and Laughlin et al. (1988) used an interval of 4 hours between fertilization and exposure to the toxicant. Widdows (1993) also proposes waiting 4 hours. Some authors wait until the first polar body is expelled (Chien and Chou, 1989; Spangenberg and Cherr, 1996). Ringwood (1992a,b) initiates the test within 30 minutes after fertilization, as is the case at the Arcachon laboratory (e.g. His and Beiras, 1995; His et al., 1997a). In our view, two ideas should be highlighted. First, the rapidity of embryonic development to the D-larva stage allows restriction of the duration of the test to less than 24 hours in Crassostrea spp., and less than 48 hours in Mytilus spp. (see next Section). Second, the ideal test (that closest to natural environmental conditions) would perform fertilization already in the presence of the toxicant. As explained below (see Section 3.2.3), in practice it is difficult to perform simultaneous fertilizations in separate treatments, and this difficulty may be overcome by beginning incubation as soon as possible after the sperm suspension
52
E. HIS, R. BEIRAS A N D M. N. L. SEAMAN
has been added to the eggs, i.e. after about 15 minutes, in order to maximize the exposure period. 3.2.2.2. Incubation conditions: duration, temperature and egg density. The duration of bioassays also varies between authors, and it cannot be discussed independently of the temperature, which determines the rapidity of the development. Duration depends on the species used as well. We shall also review here another basic incubation parameter: egg density. In the case of Crassosfrea gigas, ASTM (1989) recommends incubation at 20°C for 48 hours. The same values were suggested by Woelke (1967, 1972), Cardwell et al. (1977a, 1979a) and Konar and Stephenson (1995), whereas Thain (1991) and His et al. (1997a) propose incubating the eggs at 24°C for 24 hours. Van den Hurk (1994) followed the incubation temperature and period recommended by ASTM, but subjected the embryos to a light-dark cycle of 16 hours of light to 8 hours of dark. For Crassostrea virginica, ASTM (1989) also recommends that embryos be incubated at 20°C to 25°C for 48 hours. Calabrese and Davis (1967), Calabrese et al. (1973) and Calabrese and Nelson (1974) used temperatures of 20°C to 26"C, salinity of 25 psu and incubation periods of 48 hours in C. virginica and Mercenaria mercenaria. Roberts (1987) conducted the tests at 28"C, and obtained D-larvae in 48 hours. For Mytilus edulis, ASTM (1989) recommended 16"C, Granmo (1972) and Granmo et al. (1988) used 20"C, while Johnson (1988) incubated at 15°C for 72 hours. Butler et al. (1990) incubated for 48 hours in the dark at 20°C. For M. galloprovincialis, His and Beiras (1995) incubated embryos at 25 psu salinity and 18°C 2 1°C for 48 hours, while Brunetti et al. (1989), used 20°C for 40 hours. Cherr et al. (1990) incubated M. californianus in the dark for 96 hours at 12°C and Spangenberg and Cherr (1996) for 48 hours to 64 hours at 15°C. With Mulinia lateralis, Calabrese (1970a,b), Calabrese and Rhodes (1974) and Morrison and Petrocelli (1990) performed their experiments at 21°C for 48 hours, while Burgess and Morrison (1994) used 25°C to 26°C for 48 hours. Finally, in Zsognomon californicurn, Ringwood (1992b) conducted 48 hour incubations, and Krassoi et at. (1996) recommended incubation of Chlamys asperrima at 18°C 5 05°C for 48 hours. Regarding egg density, Calabrese and coworkers (Calabrese and Davis, 1967; Calabrese, 1970b; Calabrese et al., 1973; Calabrese and Nelson, 1974; Calabrese and Rhodes, 1974) used 10 to 20 eggs per ml, while Woelke (1967, 1972), Cardwell et al. (1977a, 1979a) and ASTM (1989) recommended a maximum density of 30 embryos per ml, even though normal development is possible at densities of 100 embryos per ml. Granmo (1972), Granmo et al. (1988), Johnson (1988), Burgess and Morrison (1994), Butler et al. (1990), Cherr et al. (1990), Ringwood (1992b), His and coworkers (His and Beiras, 1995; His et al., 1997a), Konar and Stephenson (1995) and Krassoi et al. (1996) followed the ASTM recommendation,
THE ASSESSMENT OF MARINE POLLUTION
53
whereas Roberts (1987) and Brunetti et al. (1989) conducted the tests with 40 embryos per ml. Thain (1991) used 50 embryos per ml, and Morrison and Petrocelli (1990) used up to 75 embryos per ml. As a rule, we think embryo density should not exceed 30 embryos per ml. The incubation period depends on the species; bioassays should not last longer than the shortest possible period required to complete embryogenesis. At optimal temperature, an exposure period of 24 hours is sufficient for most species, with some exceptions, such as Mytilus species, where 48 hours are usually necessary. 3.2.2.3. Volume of the test vessels. Acute toxicity tests are frequently used in biological monitoring of waters and sediments, often requiring numerous vials for numerous replicates to permit adequate statistical evaluation. There has been a continuing trend towards simplification of the bioassay procedures, with a concomitant reduction of incubation volumes. The first toxicity bioassays by Davis (1961), Hidu (1965), Calabrese and Davis (1966) and Woelke (1960 to 1972) were conducted in glass and polypropylene beakers of 1 to 2 litre volumes. Since 1980 ASTM has found that “very small vessels, containing as little as 10 to 30 ml may be used if data are provided demonstrating these systems have negligible effects on water quality” (ASTM, 1980, p. 249). Similarly, Byrne and Calder (1977) studied the toxicity of the water-soluble fraction of various petroleum products on larvae of Mercenaria rnercenaria in 10ml tubes, Morrison and Petrocelli (1990) used scintillation vials of 10 ml, and Ringwood (1992b) incubated lsognomon californicum embryos in tubes of 12 ml. Volumes of 100 ml have been used by Watling (1981) and Johnson (1988); Laughlin et al. (1988) used 50ml glass finger bowls; Butler et al. (1990) and Spangenberg and Cherr (1996) conducted their studies with glass jars of 30ml (cf. Widdows, 1993), whereas His and Beiras (1995), Beiras and His (1995a,b) and His et al. (1997a) carried out their incubations in transparent polypropylene Coulter Counter accuvettes of 25ml permitting the assessment of the larvae under an inverted microscope without further manipulation of the test organisms. Finally, Cherr et al. (1990), Chang et al. (1996) and Gormly et al. (1996) conducted tests in 3 ml wells commonly used for tissue cultures, which also permit analysis under an inverted microscope. Transparent flat-bottomed multi-well tissue culture plates of 3ml volume make it possible to study simultaneously four samples, with five to six replicates each, under an inverted microscope; the results are absolutely comparable to those obtained with 30 ml expendable vials or with beakers of 1 litre (His and Beiras, unpublished data). It is important to carefully define the criteria and end points upon which an assessment of toxicity is based, regardless of the physiological activity being studied (feeding activity, respiration, swimming behaviour,
54
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
(4 Figure 6 Effects of a contaminant (CuS04)on Crassostreagigas larval growth. (a) 1%-day-oldlarvae in controls. (b) 12-day-old larvae at 50 pg Cu 1-I; only a few larvae grew normally. Growth inhibition increased with increasing copper sulphate concentrations;some larvae which had not reached the umboned stage were still alive on day 12. (From His and Robert, 1982.)
calcium uptake, metamorphosis). The criteria used most frequently in embryo bioassays are mortality and larval abnormalities. 3.2.2.4. Larval mortality as the end point. It is not always easy for an inexperienced biologist to recognize moribund and recently dead larvae, particularly when the samples have been treated with formalin for later evaluation. On the other hand, it is very easy to obtain mortality in the larval and algal cultures unless various precautions are taken. Mass mortality is usually caused by bacteria, which can affect several incubations of an experimental series without relation to pollutant concentration. The addition of antibiotics is ineffective once mass mortality has begun. In contrast, mortality as a result of a contaminant appears gradually over several days and increases progressively with pollutant concentration; there are always some larvae which remain apparently unaffected. At certain concentrations growth may be inhibited and D-larvae, for instance, will not attain the umboned stage except for some individuals whose number decreases with increasing concentration. The small larvae are continuously decimated in the course of the experiment (e.g. His and Robert, 1982; see Figure 6). Loosanoff and Davis (1963, pp. 39-40, Figure 18), noted that in C. virginica and M. mercenaria, “it has
THE ASSESSMENT
OF MARINE POLLUTION
55
been clearly demonstrated that even though larvae originate from the same spawning and sometimes from the same parents and are kept in the same vessel under identical conditions, individuals grow at widely different rates . . .”. This is particularly true when larvae are exposed to contaminants, and in experiments with algae which do not satisfy the nutritional needs of the larvae very well (e.g. His et al., 1985). In both cases the size distributions of the larvae are heterogeneous, even though the mortality after 10 or 12 days is rarely more than 10% higher than in the controls. Various authors have described dead larvae. In fresh material, “because decomposition of larval tissue is slow in clean culture conditions, it is not easy to distinguish between moribund and recently dead bivalve larvae. In general, in larvae approaching death, velum cilia cease beating first and oesophagae cilia last” (Beaumont and Budd, 1984, p. 403). Similarly, “criteria for death in setting M y a arenaria larvae (and in straight-hinge larvae) included lack of internal movement and disrupted internal organization” (Roosenburg et al., 1980a, p. 107, and 1980b, p. 94). Larvae which have been dead for more than 24 hours, however, are easily distinguishable, because bacterial degradation and feeding of protozoa on the dead tissue leave nothing but a transparent empty shell. Various authors have devised methods to identify dead individuals in larval cultures. Byrne and Calder (1977) stained larvae of Mercenaria rnercenaria with neutral red, following the work of Loosanoff and Davis (1947; see also ASTM, 1980, p. 254). Test tubes of 10ml containing 300 larvae were then refrigerated at 4°C (or even frozen) and conserved for several days prior to evaluation. Richardson et al. (1982) noted in C. virginica that “the neutral red dye . . - was picked up by the tissue of living larvae, causing them to appear pink. This aided determination of live and dead larvae” (p. 102). Phelps and Warner (1990) distinguished between live and dead pediveligers “by cilia movement, clarity of internal organs and transparency. Dead larvae are often gray and opaque with no evidence of internal organization or movement.” They also facilitated identification by staining with neutral red at a final concentration of 1/10000 for 3 hours. In “lethal” embryotoxicity bioassays lasting 24 to 48 hours it is unnecessary to take mortality into account, because the contaminants affect embryogenesis and either prevent its completion or induce abnormalities. The appropriate criterion is thus the percentage of normal straight-hinge D-larvae obtained. If mortality occurs in the D-larvae, it may well be owing to causes other than the contaminant being investigated. Mortality, growth and larval behaviour are, in our opinion, valid criteria to be used in sub-lethal bioassays lasting several days. 3.2.2.5. Abnormal development as the end point. In embryo-larva
56
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
bioassays, the end point of the experiment is attainment of the straighthinge D-larva stage. A precise definition of what constitutes a normal D-larva, i.e. the correct identification of abnormalities, is therefore of paramount importance. Loosanoff and Davis (1963) have pointed out that it is difficult to distinguish among D-larvae of different species simply by observing their external morphology. The photographs of about 20 species reared by Loosanoff et al. (1966), including the species commonly used in bioassays (C. virginica, C. gigas, Mytilus edulis, Mercenaria mercenaria and Mulinia lateralis), show that straight-hinge veligers of all species have the same general shape of a capital D. Identification of the species at this stage is generally possible only by observation of the hinge (position of the ligament, number and form of the teeth; cf. Le Pennec, 1978) under an electron microscope. The larval abnormalities in oyster and mussel larvae described below generally apply to all bivalve species employed in bioassays, Several authors have described abnormalities, but the criteria they employed differ. Loosanoff and Davis (1963), in descriptions applying mostly to C. virginica and M. mercenaria, attributed the presence of “winged” larvae (velum protruding from the shell; p. 31), and of “saddleback” and “humpback” larvae (concave and convex hinges; p. 33) in their cultures to a variety of factors, including excessive densities, poor gamete quality and presence of contaminants. Courtright et al. (1971), while attempting to devise an artificial seawater adequate for larval rearing, also defined various abnormalities in D-larvae of Mytilus edulis (hollow type, ear muffed, black spotted, exploded but swimming embryos; see their Figure 3, p. 88, op. cit.). In the same species, Le Pennec and Le Roux (1979) described six types of abnormalities of the prodissoconch identified by scanning electron microscopy (Photograph 3, p. 5). Renzoni (1975, Figure 1, p. 127) described the abnormalities induced by petroleum products in embryos and D-larvae of C. virginica and M. lateralis as “protruding velum and small larvae without a shell”. Larval abnormalities caused by lead salts have been described by Hrs-Brenko et al. (1977, see Figure 2, p. 110) in M. galloprovincialis and by Karaseva and Medvedeva (1993) in M. trossulus. Finally, Eyster and Morse (1984) photographed various abnormalities resulting from the exposure of gametes and embryos of Spisula solidissima to silver, under an optical microscope (their Figures 1 , 2 and 3) and scanning electron microscope (Figures 4 and 5, p. 643). The criteria employed by Eyster and Morse (1984) to distinguish between normal and abnormal larvae are nearly identical to those used at Arcachon, illustrated in Figures 7 and 8, for Crassostrea gigas (His and Beiras, 1995; His et al., 1997a). In contrast, the D-larvae of Mulinia lateralis photographed by Gormly et al. (1996, see Figure l A , p. 231) are abnormal,
THE ASSESSMENT OF MARINE POLLUTION
57
Figure 7 The different abnormalities observed in D-larvae of oysters, Crassostrea gigus; (a) normal D-larva; (b) convex hinge; (c) indented shell margin; (d) incomplete shell; (e) protruding mantle. (From His et al., 1997a.)
because the visceral mass is not completely retracted into the shell and the hinge is not completely straight. In order to achieve scientifically accurate assessments of embryo-larval bioassays, certain prerequisites must be met:
1. Larval abnormalities must be clearly defined and the definition must not be based upon observations of in vitro larvae, but rather on studies of larvae isolated from the natural environment (compare Figures 7 and 8). 2. The experimental protocol (duration of the bioassay and evaluation methods) must not introduce any biases with regard to the results. 3. The maximum percentage of abnormalities permissible in the control incubations (i.e. resulting from experimental artefacts) must be specified. (a) Definition of abnormalities. Woelke (1972) states, “normal larvae as referred to here, are those which are fully shelled, even though many may be misshapen or undersized”. D-larvae defined in this manner are a far cry from those sampled in situ, e.g. in Figure 8. Woelke himself admitted that he adopted these criteria in order to simplify the evaluation, but his definition greatly diminishes the quality of the bioassay. Similarly, Cardwell et al. (1977a,b, 1979b)
58
E. HIS,
R. BEIRAS AND M . N.
L. SEAMAN
Figure 8 One-day-old D-larvae isolated from the plankton of Arcachon Bay in July 1985, after an intense spawning of Crassostrea gigas oysters. The growth and chemical composition of the larvae were followed through to metamorphosis. (His and Maurer, 1988.)
determined larvae developing abnormally, as those failing to complete shell development typical of the prodissoconch I stage (p. 5). Abnormality, as defined by ASTM (1980, p. 254), is: “incomplete or abnormal development is based upon the proportion of larvae failing to develop a complete shell after 48 hours (abnormal), relative to the total number of normal plus abnormal larvae. Larvae possessing misshapen or otherwise malformed shells are considered as normal, provided shell development has been completed” (emphasis added), and the same is the case in ASTM (1989, p. 346): “all larvae with complete developed shells containing meat must be considered as normal”. This obviously classifies as normal the “winged”, “saddleback” and “humpback” larvae considered abnormal by Loosanoff and Davis (1963, p. 33), and the various embryonal stages (normal or not) failing to complete development to the D-larva stage are excluded from the analysis. In contrast, Thain (1991, p. 51) defined as abnormal “those eggs which were not
THE ASSESSMENT OF MARINE POLLUTION
59
fertilized and those which died at an early stage of development or became malformed”, without, however, defining precisely what constitutes an abnormality. We concur with Klockner et al. (1985, p. 7), who defined normal veligers of C. gigas as follows: “24-hour-old normal D-shaped larvae exhibited (1) a fully developed symmetrical shell and (2) complete soft structures including the velum. Any retardation resulting in incomplete soft parts or deformed shells (asymmetrical shape, concave D-line, carved shells) were considered abnormal.” Similarly, Krassoi et al. (1996) in Chlamys asperrima, considered that 48 hours after fertilization “normal larvae are defined as those possessing a perfect D-veliger shell. Those larvae without the D-veliger shell are defined as abnormal (including trochophores, blastopores, larvae with misshapen shell plates)” (7.16, p. 29). At Arcachon (e.g. His and Beiras, 1995; His et al., 1997a), we have defined abnormality on the basis of a study of larvae of Crassostrea gigas isolated from the natural environment (His and Maurer, 1988). Figure 8 shows one-day-old D-larvae from the Bay of Arcachon after a mass spawning of the in situ oyster population, isolated from the plankton in the waters of the bay for laboratory investigations on growth and biochemical composition of wild larvae (His and Maurer, 1988). All of these larvae may be classified as perfectly normal straight-hinge D-larvae (compare with Figure 7b, c, d and e). Because of its narrow connection with the open sea the Bay of Arcachon represents a veritable natural hatchery (cf. Yonge, 1960, p. 165) with mass spawning of Pacific oysters occurring almost every year in mid-July. In more than 30 years of observations of larvae from the natural environment, we have never discovered any of the abnormalities shown in Figure 7. These observations form the basis for the definition of abnormality in Mytilus galloprovincialis (His and Beiras, 1995) and C. gigas (His et al., 1997a). (b) Recording the response. With regard to the mathematical assessment of abnormality levels, Davis (1960) and Davis and Hidu (1969a,b), referring to Loosanoff and Davis (1963), determined the response (R) simply as the percentage of eggs which develop into D-larvae:
R = Average number of larvae in experimental cultures x100 (1) Average number of larvae in control cultures
The larvae were counted after 48 hours by retaining them on a 44pm sieve, re-suspending them in a graduated cylinder with
60
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
filtered seawater, and taking sub-samples by pipette after mixing the water column in the cylinder. This procedure has been adopted by numerous authors, with or without use of equation 1,particularly by Calabrese in his studies on Mulinia lateralis. Woelke (1960a,b), working with C. gigas, also adopted the methods developed at Milford, retaining the larvae on a 37 pm sieve after 48 hours and counting sub-samples of about 150 to 250 larvae for the determination of the percentage of abnormalities (Woelke, 1967, p. 115). Cardwell et al. (1977a,b, 1979a,b) basically used the same methods, but omitted the sieving procedure after 1977, asserting that “concentration of veliger-stage bivalve larvae with nylon screening having a porosity of 37 pm or greater may bias toxicity test results significantly” (Cardwell ef al., 1978, p. 132; see also Bourne et al., 1981, and Chapman and Morgan, 1983). The recommendations of ASTM (1980, 1989) are based on the work of Cardwell et al. (1977a,b, 1979a,b). This requires counting the number of fertilized gametes placed into the containers, to compare with the number of “normal” D-larvae obtained after 48 hours. These counts are based on sub-samples of 10ml taken directly from the containers upon mixing, and sieving is thus avoided. ASTM (1989) proposes an alternative method with preparation of three beakers in addition to the experimental incubations for the sole purpose of making the initial counts. Klockner et al. (1985) sub-sampled 400 to 500 D-larvae of C. gigas and directly determined the percentage of abnormalities, whereas Thain (1991) sub-sampled aliquots of 2 ml from his 30 ml recipients, counting the larvae after addition of two drops of 8% formalin. Finally, Krassoi et al. (1996) added formalin to their 30 ml cultures 48 hours after fertilization; after decantation of the larvae they eliminated the top 28.5 ml by pipette and transferred the remaining 1.5 ml to a Sedgewick-Rafter cell for microscopic examination. None of these authors practised sieving, nor did they compare the number of D-larvae obtained to the original number of fertilized gametes introduced. In contrast His and Beiras (1995) and His et al. (1997a), have adopted a technique using transparent polypropylene vials and 3 ml tissue culture microwells (after Cherr et al., 1990), eliminating any need for sub-sampling or preliminary counts, because all of the biological material introduced in the test is assessed at the end of the experiment. Moreover, sieving and sub-sampling even with an automatic pipette may introduce errors, because some of the material originally introduced (e.g. decomposing segmented eggs and malformed embryos) may be destroyed by these procedures (cf. the observations by Cardwell et al., 1978).
61
THE ASSESSMENT OF MARINE POLLUTION
(c) Permissible percentages of larval abnormalities in control cultures. Woelke (1966, 1972) considered that abnormalities in control incubations should be less than 2.5%. They state, “between 5 and 10% we are extremely cautious in making firm conclusions; we totally reject any bioassay where our controls have in excess of 10% abnormal” (Woelke, 1961, p. 119 and 1966, p. 34). It must be kept in mind, however, that he sieved his larvae after 48 hours and used a very restricted definition of abnormality. Woelke expressed his results as a percentage of abnormalities when his controls exhibited less than 2.5% abnormal larvae. When his controls had a higher percentage of abnormalities, he calculated “percent net risk” (sic) - originally percentage net response (PNR) - by subtracting the percentage of abnormal controls from the percentage of abnormalities in the test incubations (originally Abbot’s equation in Emmens, 1948; Finney, 1971): PNR =
(% treatment abnormal - % control abnormal) X 100
100 - YOcontrol abnormal
(2)
Cardwell et a/. (1977a,b, 1979a’b) and ASTM (1980,1989) adopted Woelke’s definition of abnormality and Abbot’s equation (equation 2), but using an acceptable level of 10% abnormal larvae in the controls. Finally, Klockner et al. (1985) were much more restrictive in their definition of normality and based their evaluation on the Net Risk (NR): NR (%)
= (NT-NR) X
100NR
(3)
where “NR [is the] net risk - value of the test water sample (YO), NT [is the] mean relative number of normally developed larvae in replicates of test water samples, [and] NR [is the] mean relative number of normally developed larvae in replicates of reference water samples. NT and NR refer to the rates of normal larvae (raw data) on the basis of the total number o f . . . normal and abnormal larvae hatched together” (p. 9). The validity of their assessments is limited, however, by the fact that the percentage of normal larvae in their controls was merely 65.3% 5 3.3%. In our direct counts in culture vials (His and Beiras, 1995; His et al., 1997a), conducted 24 hours after fertilization in oysters and after 48 hours in mussels, we considered a test invalid when larval abnormalities attained 20% in the controls, whereas Thain (1991, p. 7) judged that “control abnormalities up to 50 or even 60 per cent are acceptable”.
62
E. HIS, R. BEIRAS AND
M. N. L. SEAMAN
In publications of bivalve embryo-larval bioassays, we think it is indispensable to mention the number of abnormally developed larvae in the controls (including the 95% confidence interval), to enable readers to appreciate the validity of the test. These data are usually absent in publications where the data are evaluated according to equation 2. 3.2.3. Tests of larval growth
These tests consist of determining sub-lethal responses by measuring the growth of the larval shell (cf. Widdows, 1993). It is interesting to note that the first anomalies detected in the field or in hatcheries did not concern embryogenesis, which proceeded normally, but larval growth. As mentioned above, toxicants usually induce a highly heterogeneous size distribution compared to healthy growing larvae (see Figure 6). The tests of larval growth are necessarily of longer duration (a few days, to several weeks if the pediveliger stage is to be studied) and require feeding the larvae. They are usually performed in beakers of 1litre (static tests), even though flow-through tests in running seawater are preferable (Calabrese, 1984, p. 460; cf Roberts and Gleeson, 1978; Roberts, 1980). These tests are even less standardized than embryo-larval bioassays. The size or growth of larvae during toxicological experiments is assessed by measuring the prodissoconch I and 11. The anatomically correct definition of these measurements (Figure 9) has been described by Galtsoff (1964) for adult oysters, and it applies equally to bivalve larvae (see also Bayne, 1976, p. 83). The following definitions are given by Galtsoff (1964, p. 18): “The height is the distance between the umbo and the ventral valve margin. The length is the maximum distance between the anterior and posterior margin measured parallel with the hinge axis. The width is the greatest distance between the outsides of the closed valves measured at right angles to the place of the shell commissure.” These measurements are nevertheless often specified incorrectly in the literature, even to this day (e.g. shell height is termed “width” by Loosanoff et al. (1966), and it is commonly termed “length” by many authors). It is therefore very important that authors, when reporting results using shell dimensions, do not limit themselves to using these terms (whether correctly or incorrectly); they must unequivocally specify how the determination was made in order to avoid confusion. In tests of larval survival and larval growth, it is necessary to distinguish between studies in which fertilization and embryogenesis have taken place in unpolluted conditions (the investigation is initiated exclusively with healthy D-larvae), and studies in which fertilization has taken place in the
63
THE ASSESSMENT OF MARINE POLLUTION
height height Prodissoconch I Urnboned (Prodissoconch 11)
height Dissoconch
Figure 9 The dimensions of the bivalve larval shell (prodissoconch I - D-larva and prodissoconch I1 - umboned larva), and of the adult shell (dissoconch). C-C’ = hinge axis. (From Le Pennec, 1978.)
presence of the contaminant. The importance of such differences to sensitivity has been recognized by all authors who have conducted ecotoxicological studies of growth in bivalve larvae. We have found toxic effects on larval growth and survival at copper concentrations five times lower when larvae were reared in the presence of metal from fertilization on, compared to larvae developed from unexposed gametes (His and Robert, 1981, 1982). An intermediate possibility is to begin exposure to the toxicant minutes after the start of fertilization. 3.2.3.1. Studies on growth inhibition with previously unexposed larvae. The first studies of bivalve larval growth in the presence of pollutants were conducted by Davis and Chanley (1955b), who investigated the effects of various antibiotics, and by Davis (1961), who demonstrated the toxicity of 32 pesticides on larvae of Crassostrea virginica and Mercenaria mercenaria. These experiments were concerned first with embryogenesis and the formation of D-larvae, and then with the subsequent development of D-larvae obtained in the absence of contaminants during a further 12 days of contaminant exposure. Growth and mortality were determined only on the twelfth day. Further laboratory studies were conducted at Milford by Hidu (1965) on the effects of synthetic surfactants on M. mercenaria; Calabrese and Davis (1966,1967) studied the effects of pH on C. virginica and M. mercenaria and of linear alkylate sulphonate (LAS), a surfactant, on C..virginica; Calabrese (1972), Calabrese and Davis (1967), Calabrese and Nelson (1974) and Calabrese et al. (1977a) then investigated the effects of heavy metals and various other pollutants such as detergents and pesticides. Finally, Calabrese and Rhodes (1974) applied the same techniques in studies with embryos and larvae of the coot clam, Mulinia lateralis, demonstrating the advantages of using this species in bioassays. The results obtained at Milford laboratory were summarized by Calabrese et al. (1977b, 1982). In studies of the effects of contaminants on larval growth using previously unexposed (healthy) larvae, the embryos were raised in an
64
E. HIS,
R. BEIRAS AND M. N. L. SEAMAN
uncontaminated mass culture, collected on a sieve 48 hours after fertilization, counted and redistributed into the experimental beakers at concentrations of 5000 to 15 000 per litre. Incubation generally lasted about 10 days, except in the case of M. lateralis, which attains settlement competence after 6 to 8 days (cf. Calabrese, 1970b). The incubation water was changed every 2 days, always maintaining the appropriate contaminant concentration. To this end, the larvae were first collected on a sieve and sampled to determine mortality, possible abnormalities and shell size. Larval size was measured by ocular micrometer to a precision of 2 5 pm (cf. Calabrese, 1970b). Measurements can also be made by photographing samples of larvae and later measuring the diameter of the shell under a binocular microscope, either on the photographs or on the negatives (e.g. His and Robert, 1981, 1982). This is usually done with a video imaging system (e.g. Widdows et al., 1989; McFadzen, 1992). Larval weight is rarely measured (Widdows et al., 1989; Beiras and His, 1994, 1995a) despite the fact that it is required to express lethal concentrations on a mass-specific basis, enabling comparison of data among larvae of different sizes (see Section 4.2.3). In the ecotoxicological studies with Mytilus edulis and Pecten maximus by Beaumont et al. (1987), and those with Mercenaria mercenaria by Byrne and Calder (1977), larval feeding was based on the methods reviewed by Loosanoff and Davis (1963) (see Section 2.3.5). Renzoni (1975) studied the effects of the water soluble fraction of three petroleum products on growth in Mulinia lateralis during a period of 2 weeks, changing the water (and renewing the pollutant concentration) every 2 days and feeding the same algae as used by Calabrese (1970b). Laughlin et al. (1989) in studies of the effects of TBT on Mercenaria mercenaria fed the larvae exclusively with Isochrysis galbana (Tahiti strain). Finally, Gormly et al. (1996) studied the resistance of Mulinia lateralis to microbial pest control agents during 48 hours in microwells of 3 ml (15 to 20 larvae per well); the larvae were fed during these brief experiments on mortality and growth of the larvae. Some authors obtain their larvae from commercial hatcheries and this may pose various problems. Such larvae are generally the progeny of a haphazard mix of parents, and they may also represent a mix of incubations of different ages which have all been passed through the same sieves to obtain a homogeneous size; i.e., they may be the result of the culling procedure employed by commercial hatcheries to rid themselves of slow growers. Watling (1978, p. 126, and 1982) mentions that in the hatchery-reared larvae used “the range of individual sizes was restricted at the beginning of the experiments”, which may pose a problem in studies of larval growth.
THE ASSESSMENT OF MARINE POLLUTION
65
3.2.3.2. Studies on growth inhibition with D-larvae already exposed to contaminants. Studies in which fertilization and embryogenesis were conducted in the presence of the contaminant being bioassayed are rare, even though they represent the best approximation of in situ conditions. Granmo (1972) first distributed eggs of Mytilus edulis at a density of 10 per ml into vials with various concentrations of LAS (a surfactant) before proceeding to fertilize them. The water was subsequently changed daily during the first 8 days, the larvae being retained on a mesh of 60 pm at each change of water. Fertilization success was assessed after 20 to 24 hours, and larval mortality and size (“measured microscopically, the greatest length attained being recorded”, p. 356) were subsequently determined after 72 hours, 96 hours and 240 hours. Brereton et al. (1973) studied the effects of several concentrations of zinc sulphate on freshly fertilized eggs of Crassostrea gigus (15 000 per beaker of 1 litre) at a temperature of 20°C to 22°C and a salinity of 29 psu. Larval abnormalities were counted after 48 hours, and the D-larvae were subsequently passed through a 100 pm sieve to remove impurities, retained on a 45 pm sieve and incubated at a density of 5000 per litre at 26°C. They were grown for 5 days in the presence of the contaminant and another 5 days in unpolluted seawater, and fed daily with Zsochrysis galbana, Chaetoceros calcitrans, Monochrysis lutheri and Cyclotella nana at 100 cells per pl of larval culture volume. The water was changed every 48 hours, and at this time swimming behaviour was recorded, and mortality and growth were assessed. Shell size (“width across the valve”) was determined in 25 individuals of each of four replicates of each concentration. Laughlin er al. (1988) studied the toxicity of TBT to Mercenaria mercenaria; the eggs were exposed to the contaminant 4 hours after fertilization at densities of 150 per bowl of 50 ml, and the incubation lasted a total of 14 days. The solutions were renewed daily except during the first 48 hours, and the larvae were fed with Zsochrysis galbana (Tahiti strain) at 40 000 cells per ml. Size was determined daily in sub-samples of 25 larvae per incubation. The first studies of larval growth undertaken at Arcachon followed the methodology of Brereton et al. (1973) and Helm and Millican (1977). They were initially applied to investigate the effects of TBT on Crassostrea gigas and Mytilus galloprovincialis, showing the threat posed by this compound to coastal shellfish culture areas (His and Robert, 1980; Robert and His, 1981). These experiments were carried out with larvae that had been reared through embryogenesis in the presence of the toxicant, as well as with larvae reared to the D-larva stage in unpolluted water. The eggs were distributed into 2 litre beakers at a density of 60 per ml; oyster embryos were incubated for 24 hours at 20°C and mussel embryos for 48 hours at 19°C. After this period the contents of each beaker were passed through a 100 pm sieve to remove impurities, and the larvae were collected on a
66
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
sieve with 40 pm mesh. After assessment of larval abnormalities the larvae were re-suspended in 250 ml graduated cylinders for recounting and incubation then continued at densities of eight individuals per ml for 7 days (only 5 days in the case of TBT, owing to the high toxicity of the concentrations tested). The water (0.2 pm filtered seawater) was changed daily and the larvae were fed daily as well. Mortality was assessed in each incubation, and larval size was determined with an ocular micrometer in sub-samples of 50 individuals. Studies of growth were subsequently extended to 10 or 12 days, and the investigations with larvae formed in unpolluted water were abandoned because of their lesser sensitivity. The feeding regimes of Helm and Millican (1977) were used at all times. The water was changed only every second day, and mortality and growth were therefore equally assessed in intervals of 2 days; shell size was determined by photographing sub-samples of larvae, rather than directly under the microscope. After studying the effects of TBT on larvae of Crassostrea gigas at concentrations around 1pg per litre, further studies were conducted on the effects of copper sulphate, cadmium chloride and salinity, and atrazine-simazine, a herbicide (His and Robert, 1982; His et al., 1983; Robert and His, 1985; Robert et al., 1986). The method has lately been modified further because of the impossibility of conducting numerous fertilizations in many different contaminant concentrations almost simultaneously. We therefore now prefer to add the sperm to the graduated cylinder immediately after determining the density of the egg suspension, distributing the gametes into the experimental vials 15 to 30 minutes after fertilization (Section 3.1.3.6). This method has been used to study the effects on C. gigas of 12 pesticides (His and Seaman, 1993), of mercury (Beiras and His, 1994) and of aluminium-plant effluents (His et al., 1996), as well as the effects of mercury on Mytilus galloprovincialis (Beiras and His, 1995a). 3.2.4. Tests on larval physiology These bioassays concern the effects of pollutants and physico-chemical environmental impacts on larval behaviour or activity, such as feeding activity, respiration, resistance to hypoxia, swimming behaviour and metamorphosis. Physiological analyses of bivalve larvae are not easy to conduct and may produce doubtful results, particularly in toxicity assays accompanied by larval mortalities and ensuing bacterial proliferation; for instance, in a study on the feeding behaviour of oyster larvae (Seaman and Ueberschar, unpublished), the samples in which all the larvae were dead had by far the highest activities of trypsin, a digestive enzyme produced by animals and bacteria alike.
THE ASSESSMENT OF MARINE POLLUTION
67
3.2.4.1. Feeding activity. After studying the effects of copper chloride, cadmium chloride and zinc chloride on the growth of algae (Monochrysis lutheri, Isochrysis galbana, Dunaliella euchlora and Phaeodactylum tricornutum), Wikfors and Ukeles (1982) proceeded to study the transfer of these contaminants to the next trophic level by feeding larvae of Crassostrea virginica with axenic cultures of 1. galbana grown in sub-lethal concentrations of copper chloride (10 mg% = 47.3 ppm Cu) and zinc chloride (2.5 mg% = 15.3ppm). The digestive tract of the larvae was observed by epifluorescence microscopy (study of the fluorescence of the algae ingested, by the method of Babinchak and Ukeles, 1979), showing that the larvae fed normally and digested the algal cells. Larval growth, however, was strongly affected, showing the potential for bioaccumulation of pollutants beginning with the level of primary production: “The phytoplankton can exhibit a limited increase in tolerance and adaptation to metals. These populations are then potentially toxic for grazing species at higher trophic levels” (Babinchak and Ukeles, 1979, p. 191). Wikfors et al. (1993) and Wikfors and Smolowitz (1995) also studied the effects of the toxic dinoflagellate Prorocentrum minimum on larval nutrition in C. virginica, again using epifluorescence microscopy. They were unable to conclude from their results that the reduction of larval growth and survival were a direct result of the toxic effect of the dinoflagellate. Widdows et al. (1989) used fluorescent microspheres (3.4 pm diameter) derived from autoclaved algal extracts to study the effects of hypoxia and anoxia on the ingestion rate of prodissoconch larvae (73 to 140 pm) and pediveligers (300 to 376 pm); the microspheres ingested were counted under an inverted epifluorescence microscope. In contrast to the pediveligers, the prodissoconch larvae continued to feed under anoxic conditions. 3.2.4.2. Respiration. There are few studies on respiration in bivalve larvae beyond those of Morrison (1971) and the investigations of hypoxia resistance by Widdows et al. (1989) and Baker and Mann (1992, 1994a), particularly with regard to the effects of pollutants. Thurberg et al. (1975), using a microrespirometer developed by MacInnes and Thurberg (1973), studied respiration rates in larvae of Spisula solidissima exposed to 50 ppb of silver (added as AgN03) at 2-day intervals during a 15-day period (initial and final size of the controls 73 and 243 pm, respectively). Larvae exposed to the pollutant had respiration rates one-third higher than the controls. Pavicic (1980) studied the combined effects of cadmium and zinc on respiration in embryos and larvae of Mytilus galloprovincialis. Oxygen consumption was measured during embryogenesis and into the D-larva stage at temperatures ranging from 10°C to 20°C. Fertilized eggs were placed in hermetically sealed incubation vessels of 500 ml volume at
68
E. HIS, R. BEIRAS AND M. N .
L. SEAMAN
densities of 30 to 100 eggs per ml, and dissolved oxygen levels were determined by the Winkler method after 48 hours to 67 hours. The results were inconclusive “due to the difficulties of measuring low respiratory rates of bivalve larvae with sufficient accuracy to detect physiological changes resulting from the different treatments” (p. 63). 3.2.4.3. Swimming activity. Chang et al. (1996) studied the effects of copper and zinc fixed on biofilms in cell culture dishes (1.5 ml, each with 20 to 30 larvae) by observing under the microscope “speed of swimming, ciliary activity and the number of quiescent or inactive larvae” (p. 590) at different metal concentrations. The most sophisticated work on swimming behaviour of bivalve larvae has been undertaken by Mann and Rainer (1990) in larvae of C. virginica subjected to low oxygen concentrations, and by Mann et al. (1991) who studied the response of larvae of Spisula solidissima, Mulinia lateralis and Rangia cuneata to salinity discontinuities. These authors observed the behaviour of veligers in glass cylinders 30 cm high and 0.6 cm in diameter (with an approximate volume of 10.8ml), either by recording the motion of the larvae with a high-resolution video camera (Mann and Rainer, 1990), or by simple observation under a binocular microscope and recording the movements on graph paper (Mann et at., 1991). Van den Hurk (1994) studied the distribution of larvae of Crassostrea gigas in the water column of 1litre jars, in order to compare the behaviour of larvae exposed to a toxic sediment with that of larvae in unpolluted filtered seawater. In the absence of sediment the larvae remained at the bottom of the beakers, but moved towards the surface in the jars with sediment. The author concluded that this represented a strategy for avoidance of the sediment. We conducted preliminary trials to study the response of C. gigas oyster larvae to mercury. Setting up eight 25 ml vessels with a sub-lethal mercury concentration of 16 pg 1-1 and eight controls with seawater, we added 5000 larvae to each vessel and frequently sampled the upper water layer with a Id pipette. Figure 10 shows the mean numbers of larvae in these samples. During the first few hours there was a marked difference in the vertical distribution of mercury-exposed and unexposed veligers. The former tended to concentrate in the upper part of the experimental vessels, perhaps as a result of increased swimming activity. However, this kind of result must be interpreted with caution. Healthy oyster larvae (C. gigas as well as Ostrea edulis) often have a tendency to remain in the upper part of the water column when kept in beakers of 1 or 2 litres, as shown by our own observations in 20 years of larval culture and those of Wilson (1981). The umboned stages tend to float at the water-air interface until the appearance of the eye spot, forming characteristic dense mats. This phenomenon, termed “rafting” by Wilson, is “an
69
THE ASSESSMENT OF MARINE POLLUTION
140
120
F
100
Q)
2
Y-
80
O
b
Q
E 3
2
60 40
20
0 I
I
I
I
I
0
10
20
30
40
I
50
I
60
Time (h) Figure 10 Density (I&’) of Crassostrea gigus larvae actively swimming near the surface of the 25 ml vessels when exposed to 16 pg Hg 1-’ (0) and in seawater controls (0).Values are means ( 2 SD) of eight samples from each different vessel. Note the sigmficantly increased upward swimming activity in the mercury-exposed larvae during the first hours of the experiment.
indication of optimum conditions, and in our experiments such cultures grow and settle well” (Wilson, 1981, p. 21). On the other hand, moribund D-larvae and umboned larvae of oysters tend to accumulate at the bottom of the incubation vessels (cf. Galtsoff, 1964, p. 376). These observations do not apply to mussel larvae, which tend to remain at the bottom of the beakers when they are doing well. Therefore, little is known on the sub-lethal effects of toxicants on the swimming activity of veliger larvae, but this is a promising field for future studies. 3.2.4.4. Calcium uptake. Ho and Zubkoff (1979, 1980, 1982) have proposed “a sensitive and rapid bioassay”, using 45Caisotope uptake by
70
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
larvae (Ho and Zubkoff, 1982, p. 409). Their first study (1979) concerned calcium uptake in larvae of C. virginica over a period of 68 hours in the presence of the toxic dinoflagellate Cochlodinium heterolobatum, following a phytoplankton bloom. The algae caused mortality in the larvae, as well as a reduction in calcium uptake. The method, described in detail in Ho and Zubkoff (1980, p. 144) was also used to study the effects of mercury, copper and zinc on Mercenaria mercenaria, showing that copper has the highest, and zinc the lowest toxicity of the three metals, both with respect to mortality and to calcium uptake; they suggested that the method be used for toxicity monitoring. In another study (Ho and Zubkoff, 1982), they found that calcium uptake by larvae was depressed by nickel, but stimulated by cadmium. 3.2.4.5. Metamorphosis. The last larval stage of bivalves, the pediveliger, considered particularly sensitive to pollutants (Watling, 1983), has been used in some 20 ecotoxicological studies. In the 1930s Prytherch was the first to record metamorphosis in Crassostrea virginicu on film; some of his films were recovered by Medcof (1961). Prytherch (1931,1934) showed that concentrations of 0.15 to 0.50 mg Cu 1-1 induce settlement in competent larvae of C. virginica. He interpreted this as an avoidance reaction to the contaminant, thus making the first ecotoxicological observation using pediveligers. Most studies with pediveligers have investigated the effects of heavy metals (Boyden et al., 1975; Watling, 1978, 1983; Nell and Holliday, 1986; Beiras and His, 1994). Sheffrin et al. (1984) worked with plantigrades, the first post-metamorphic stage. Chang et al. (1996) and Phelps and Mihursky (1986) investigated the effects of metals in the microbial layer of settlement surfaces. Ruiz et al. (1995b) studied the effects of TBT on metamorphosis in Scrobicularia plana. Smith and Hackney (1989) conducted a field study on the effects of hydrocarbons on in situ settlement in C. virginica. Various bioassays have been conducted on the environmental impacts of nuclear-energy plants, particularly with regard to the various treatments employed for the prevention of biofouling in tubes and canals. Studies focused mainly on the toxic effects of chlorination and ozonation on oyster and mussel pediveligers (Bucaille and Kim, 1979; Roosenburg et al., 1980a,b; Richardson et al., 1982; Roberts and Casey, 1985). The ability of the pediveligers to undergo metamorphosis has also been employed to assess the quality of marine sediments (e.g. Phelps and Warner, 1990), since the pediveliger is the earliest and most sensitive phase of the oyster that comes in contact with the substratum. The same idea motivated His et al. (1997b) to investigate the negative effects of a sediment polluted by polycyclic aromatic hydrocarbons (PAH) on recruitment in C. gigus.
THE ASSESSMENT OF MARINE POLLUTION
71
Two types of studies may be distinguished with regard to their methodology: those conducted before the importance of chemical inducers of metamorphosis had been unravelled, and those which were able to profit from the work of Coon et al. (1985,1986,1990), Coon and Bonar (1986) on metamorphosis in C. gigas and C. virginica, and of Shpigel et al. (1989) on Ostrea edulis. Metamorphosis is induced by L-Dopa (dihydroxyphenylalanine), and epinephrine, making it possible to obtain cultchless spat (metamorphosis without attachment to a settlement surface) in oysters. Cultchless spat are important in commercial and laboratory culture of oysters, because it is easier to collect post-larvae from the water column than from the walls of the incubation vessel. It should be pointed out that various authors (Phelps and Mihursky, 1986; His et al., 1997b) have conducted their studies with late veligers obtained from a commercial hatchery and raised within days to the pediveliger stage in the laboratory. Some authors (e.g. Watling, 1983) have worked with hatchery larvae having settlement competence, as these (eyed) larvae can easily be “shipped by overnight express under moist and cold conditions” (Chang et al., 1996, p. 590). Tests on pediveligers without chemical induction of metamorphosis. Boyden et al. (1975) and Watling (1978,1983) exposed hatchery larvae of Crassostrea gigas (16 to 23 days old, 240 pm to 300 pm in size) to heavy metals for 5 days at densities of 1000 to 15 000 larvae per 12 litres. The bottom of each beaker was covered by a ground glass plate and “a black PVC disc slightly arched to make available both upper and lower surfaces for settlement” (Boyden et al., 1975, p. 228; Watling, 1983, p. 345). The collectors were placed in unpolluted flowing seawater after 5 days, and the number of metamorphosed larvae and the mortality 30 days after contaminant exposure were assessed. Phelps and Mihursky (1986) placed 30000 eyed larvae of C. virginica in 1.5 litre beakers with 0.2pm filtered seawater. The bottom was covered with shell debris and coppercontaminated microbial films. After 4 hours the larvae were transferred to flowing estuarine water, filtered at 0.5 pm to allow the passage of food algae. The number of spat was assessed after 4 to 7 days. Richardson et al. (1982) reared their own larvae to the pediveliger stage in 9 to 21 days (sic). As soon as most of them had attained the eyed stage and the first settlements were observed, the larvae were placed in “shallow setting trays lined with black polyethylene sheeting and the entire bottom covered with numbered 2.5 by 7.6 cm glass microscope slides which are used as experimental substrate” (p. 101). They were exposed to ozone produced oxidants (OPO) and chlorine produced oxidants (CPO) as soon as 100 to 200 larvae “in the late crawling stage - not definitively fixed” were observed on the slides. Mortality was determined after 12, 24, 48 and 96 hours by staining the
72
E. HIS,
R. BEIRAS AND M. N. L. SEAMAN
larvae with neutral red (only live larvae acquire the stain). They do not mention data on metamorphosis success. The effects of CPO on Mya arenariu and C. virginicu larvae were also studied by Roosenburg et al. (1980a,b) using a similar experimental protocol. D-larvae and pediveligers were exposed to CPO in a flowthrough setup. The latter were exposed to the contaminant after having settled; as in the case of Richardson et al. (1982), the larvae were collected when they attained the “swim-crawl phase prior to settlement” and placed at densities of 500 per ml in 8 litre fibreglass tanks, the bottom of which was covered with microscope slides. The tanks were filled with filtered seawater, and the larvae were allowed to settle on the slides. The number of fixed spat was then counted and the slides were cleaned except for sections of 4 cm*containing an average of 100 to 162 spat. The slides were then exposed to CPO in a flow-through system. Dependent variables were larval resistance to the pollutant and detachment of the spat from the settlement surfaces, as a function of concentration and duration of the exposure (8,24,48,72 and 96 hours in oysters; 4 hours, 6 hours, and the same intervals to 72 hours in clams). The authors mention percentage mortality and detachment, but they do not give clear data on the percentage of metamorphosed larvae. Studies on the effects of chlorinated seawater on fixation and detachment of mussels (Mytilus sp.) have been conducted by Bucaille and Kim (1979). Their experimental protocol is not described in detail, but they too worked with a flow-through system with continuous and discontinuous chlorination, first observing settlement in comparison to the controls (which had 90% settlement), and then the detachment of larvae which had initially settled in the absence of chlorination. Nell and Holliday (1986), working with Saccostrea commercialis in an experimental hatchery, investigated settlement behaviour in the presence of K+ and Cu2+in larvae retained on a 300 pm mesh (diagonal measurement) and placed at a density of 2000 larvae per litre in 8 litre aquaria painted black (four tanks per treatment). Grey polyvinyl sheets (300 X 150 X 0.8 mm) held 5 mm from the bottom of the tanks were used as a settlement surface. The aquaria were dimly lit at 100 lux, and weakly aerated. The water was changed daily, and the larvae were fed daily with 75 000 cells per ml of Zsochrysis gatbana and Monochrysis Eutheri. “The number of live settled spat that could not be washed off the sheets by gentle rinsing with sea water was counted and the viability of the larvae checked daily” (p. 264). Finally, Ruiz et ul. (1995a), attempting to demonstrate the effects of TBT on pediveligers of Scrobicularia plana, confronted technical difficulties which make their method hardly applicable in routine bioassays. The rates of successful embryogenesis were only about 10%. One
THE ASSESSMENT OF MARINE POLLUTION
73
advantage of S. plana, on the other hand, is that pediveligers of the “swimming and crawling” stage can be obtained in the laboratory in little less than a week, making it possible to conduct metamorphosis assays more rapidly than with other species. Competent larvae require sediment in order to metamorphose, and the authors used “acid washed sand ground with a mortar and pestle until it passed through a 1OOpm pore mesh, and autoclaved” (Ruig et al., 1995a, p. 120), deposited as a layer 3mm in thickness at the bottom of the experimental vessels of 90ml volume. They placed 70 larvae (10-day pediveligers about 240pm in length) in each vessel. The spat were exposed to four concentrations of TBT (ranging from 10 to 100 ng Sn 1-‘) for 30 days, during which the water was changed every 2 days. Some of the difficulties encountered by the authors concerned the use of sediment: “the definitive acceptance of sand as settling substrate seemed to be delayed in controls” (p. 121), “and the sand may not have been optimum substratum” (p. 123). It was also difficult to separate the larvae and the sand, and the alcohol used as a solvent for TBT provoked an excessive secretion of mucus leading to aggregation of the larvae and ensuing mortality. The use of solvents is sometimes unavoidable, but at least in the case of oyster larvae, alcohol is much more toxic than dimethylsulphide (DMSO) and acetone (His, unpublished data; cf. His and Seaman, 1993). The authors based their assessment on the growth and survival of the larvae and spat of S. plana. Tests on pediveligers with chemical induction of metamorphosis. The chemically mediated induction of settlement in oyster larvae pioneered by Coon and coworkers (op. cit.) has opened new possibilities for bioassays involving metamorphosis of pediveligers. Thus, Phelps and Warner (1990) studied the potential toxicity of contaminated sediments on metamorphosis of hatchery-reared C. gigus and C. Virginia. The stock of veligers was kept in aerated 2.7 litre beakers with artificial seawater (ASW) treated with antibiotics; the water was changed every 2 days (larvae retained on 149 pm sieves) and the larvae fed daily with Monochrysis and Zsochrysis. For the bioassays the larvae were placed in ASW with 0.001mol epinephrine for 20 hours, and the sediments were tested at 21°C in 2.75 ml wells with 1ml of sieved sediment and 1ml of water in tissue culture plates (24 wells per plate). After 96 hours the contents of each well were removed by automatic pipette and sieved through a 149 pm mesh. Larvae were classified as live, metamorphosed or dead after staining with neutral red: “Dead larvae were often grey and opaque with no evidence of internal organization or movement. Some dead larvae showed retraction or partial decomposition of tissues and some had invasion by protozoa. Metamorphosis was determined by the transition from rounded veliger to flat shape with new shell growth” (p. 199).
74
E. HIS, R. EEIRAS AND M . N. L. SEAMAN
The same method was used by Beiras and His (1994) to determine the effects of mercury on metamorphosis in C. gigas, with epinephrine being used at a final concentration of lop4M. Following the methods of Beiras and Widdows (1995), competent eyed larvae (size greater than 290 pm) were placed in 3 ml wells of tissue culture plates at a density of 11 per ml. The number of post-larvae (see Figure 3) was assessed under the microscope after 2 days of exposure to epinephrine and toxicant. In the same way, the toxicity of sediment contaminated with PAH was evaluated (His et al., 1997b). These two tests are prime examples of sub-lethal bioassays, as the larval mortality was extremely low (less than 3% in all cases). The criteria used to determine metamorphosis success were: loss of the velum and foot; plainly visible outline of the gills; secretion of the dissoconch, which was often slight (presumably owing to pollutant effects) but nevertheless visible as a translucent veil, as opposed to the strongly coloured prodissoconch I1 (see Figure 3).
3.3. Bioassay applications: Toxicity tests and environmental bioassays
3.3.1. Toxicity tests with pure chemicals Toxicity is primarily a function of the amount of toxicant available to the organism, i.e. the dose, and the duration of exposure. In aquatic organisms the toxicant is often present in solution and its concentration is taken as the dose. Toxicity can then be evaluated either by fixing an exposure time and varying the concentration or by fixing a concentration and varying exposure time. In the first case the percentage of test individuals exhibiting a biological response at each concentration is determined after the exposure period, and the theoretical concentration that would produce a response in 50% of the individuals is calculated (see below). This is the median effective concentration or EC50.In the second case the biological response after different periods of exposure is measured and the theoretical time at which a response occurs in 50% of the individuals at a particular concentration is calculated, this is the median effective time or ETSO.When lethal toxicity is studied the equivalent parameters obtained are the LC50 and LT,,, respectively. The difficulties associated with holding marine animals in the laboratory for long periods and maintaining a constant concentration of toxicant dissolved in the water have made the and LCS0)more popular. first approach (calculation of The toxicity of pure chemicals to marine organisms has been investigated in the laboratory by measuring a biological response, e.g. successful embryogenesis, at different concentrations of a toxic compound dissolved
THE ASSESSMENT OF MARINE POLLUTION
75
in seawater (see Section 3.4.1). The results can be represented by a concentration-response curve. For non-polar chemicals a non-toxic carrier (e.g. acetone or DMSO) is recommended and carrier controls must be included in the experimental design. Acute toxicity studies are those studies which investigate the effects of toxicants after short exposure periods (from a few hours to a few days), and which usually assess fertilization, embryo development or mortality; whereas chronic toxicity studies are those which use exposure periods lasting for most of the larval life span (at least one week), and which assess larval growth as the measure of toxicity. When data on the toxicity of the chemical to be studied are not available, it is advisable to conduct preliminary trials, testing concentrations in log" steps to find the range of no effect to 100% mortality. Once this range is known Concentrations should be tested in smaller steps, depending on the reliability of the nominal concentrations (i.e. measured actual concentrations must be significantly different among treatments). Since extrapolation in regression models is strongly ill-advised, a good test should always include concentrations causing both more than 50%, and less than 50%, response, even though concentrations giving 0% and 100% response are not to be included in the calculations of the LC50and related parameters. On the other hand, partial responses are necessary for most methods of calculation of the LCSo(see Section 3.4.1). Many toxicants tend to disappear from the water column during the bioassay because they are volatile, adsorb to the walls of the test vessels, or react with other compounds in the water. Therefore, the actual concentrations of the pollutants must always be measured at the beginning and at the end of the test, to make sure they do not deviate from the nominal concentrations. Dilution water should be 35psu artificial seawater made up under axenic conditions, or filtered (0.45 pm or lower) seawater of oceanic characteristics (see Section 3.1.1). When natural oceanic seawater is used, the concentrations of the main chemical toxicants should be analysed first. The containers used to make up the toxicant solution and to conduct the bioassays depend on the chemical properties of the toxicant and should be selected to minimize possible losses. Plastic containers (polypropylene or PTFE) should be used when assaying metals, while glass is preferred for organic compounds. Containers should be capped with an airtight seal. If the density of the test organisms is too high in static bioassays, spurious toxicity may result from low oxygen concentrations or ammonia excretion by the larvae (see Section 3.2.2.2). Experimental concentrations should be made up from a stock solution of the toxicant dissolved in distilled water at a concentration at least ten times higher than the maximum concentration to be tested. This limits the
76
E. HIS,
R. BEIRAS AND
M. N.
L. SEAMAN
reduction in salinity to 10% or less when the required amount of stock solution is added to the incubation seawater. A 10% reduction from oceanic salinity (i.e. to 31.5 psu) does not affect the gametes, embryos or larvae of the marine bivalves reviewed here. The rest of the experimental concentrations can be prepared by diluting in seawater the higher concentration, using volumetric flasks. Replicates without animals should be run to permit the removal of samples for chemical analyses. The toxicity tests described here aim to ascertain the EC5* of the chemical being investigated. The EC5o has several applications:
1. It directly describes the degree of toxicity of a single chemical; the lower the EGO,the more toxic the chemical. 2. It can be used to study interactions between two chemicals, providing a common unit to compare the effects of toxicants acting singly or in combination. One Toxicity Unit (TU) is the concentration of the chemical divided by its EC5o. If a mixture of a certain number of TUs of two chemicals causes the same toxicity as the total number of TUs of each single chemical alone, then their interaction is additive. If the toxicity of the mixture is greater, the interaction is synergistic, and if the toxicity of the mixture is significantly lower, the interaction is termed antagonistic (see also Section 4.4). 3. Toxicity tests are an integral part of environmental risk assessment of chemicals likely to be released into the environment. Safe concentrations can be determined by multiplying the EC50 by a safety factor (generally 0.01). The safety factor is designed to account for the disparity between the results of an acute toxicity test in the laboratory and the situation in the natural environment, including possible effects of long-term (chronic) toxicity, and possible synergistic interactions with other pollutants in the natural environment.
3.3.2. Bioassays with samples from the natural environment. Rationale for the use of bioassuys in monitoring marine pollution Most contaminants originating from human activities enter the marine environment via atmospheric deposition, fluvial inputs, or direct spillage into the sea. Field samples (e.g. of water or sediments) generally contain a wide range of unidentified potential pollutants and their complete individual quantification is unrealistic. Even when chemical identification of the main toxicants included in the ever-growing “black lists” is achieved, this does not allow an inference on actual toxicity, because chemical speciation, complexation, interaction and other physico-chemical processes affect bioavailability. Therefore, biological assays are strongly
THE ASSESSMENT OF MARINE POLLUTION
77
recommended as tools to provide biological criteria for the assessment of pollution effects. The toxicity of bioassayed field samples may be classified according to the arbitrary scale recommended by Woelke (1966) who considers a sample which gives less than 5% abnormal larvae in excess of the controls as non-toxic, 5 to 15% abnormalities as slightly toxic, and over 15% as toxic. Irrespective of the source, contaminants in the oceans may accumulate in four compartments: water surface microlayer, water column, sediments and biota. Accumulation of pollutants within organisms is beyond the scope of this review. 3.3.2.1. Water quality bioassays. For assessing water quality, bulk seawater may be used immediately after collection for incubation of the fertilized eggs, as described by Woelke (1967). He recommended this biological procedure for the monitoring of estuaries, and pointed out that a chemical criterion (the Pearl-Benson index for pulp-mill wastes) was not correlated with toxicity to the test organisms. In order to allow comparisons between experiments performed with different biological material, Woelke (1972) calibrated the experimental data with a control incubated in clean seawater, thereby calculating the net response (see Section 3.2.2.5). This general approach has been followed by Lloyd and Thain (1981), Nelson et al. (1983), Klockner et al. (1985), Nascimento (1989), Thain (1992), Thain and Kirby (1994), and Konar and Stephenson (1995). Konar and Stephenson (1995) did not find significant differences in toxicity between sub-surface and bottom waters. The sea surface is an important concentration area of pollutants from the atmosphere (such as lead), as well as for hydrophobic substances, including crude oil and its derivatives, PAH, and chlorinated hydrocarbons. The surface microlayer therefore contains higher concentrations of metal and organometal contaminants than the sub-surface bulkwater (Hardy and Cleary, 1992), and is usually more toxic to bivalve larvae (McFadzen, 1992). This has motivated the development of methods for sampling the top few micrometres of the sea surface. While these methods are extremely useful for the detection of pollutants which can be below detection limits, the ecological relevance of the derived toxicity data is questionable (Thain, 1992). The use of euryhaline species allows the extension of the bivalve embryogenesis bioassay to fresh and brackish water. Morrison and Petrocelli (1990) conducted a bioassay using Mulinia lateralis at salinities as low as 1Opsu. The species commonly used in Europe, Mytilus galloprovincialis and C. gigas are also suitable for assaying coastal waters of different salinities after salinity is adjusted to 25 psu by adding either 0.2 pm filtered clean oceanic seawater or deionized water to the samples (His and Beiras, 1995; His et al., 1997a).
78
E. HIS, R. BEIRAS AND M .
N. L. SEAMAN
3.3.2.2. Sediment quality bioassays. Pollutants are deposited and concentrated in sea-floor sediments by precipitation of dissolved chemicals, and by adsorption to the surface of particles which then sink to the sea floor (either directly, or after passing through the digestive system of an animal). Organic pollutants may persist in sea-floor sediments for long time periods owing to anoxic conditions which reduce the rate of microbial degradation. While bulk sediment analytical chemistry may provide an estimate of the degree of pollution, the fraction of the total that is bioavailable and affects the organisms can only be determined by using biological assays. Sediment characteristics such as particle size distribution, organic content, redox potential and acid volatile sulphides (AVS), affect the quantity of bioavailable pollutants (Anderson et al., 1987; Lamberson et al., 1992; Ankley et al., 1996b). The concentration of metals in the interstitial water, rather than total dry weight metal content of the sediment, is often correlated with toxicity (see Ankley et al., 1996a, and citations therein). Interstitial water can be obtained from the bulk sediment, thereby providing the opportunity to evaluate sediment quality by means of water-column organisms such as bivalve embryos and larvae. Sediment toxicity bioassays are widespread tools in environmental quality assessment (reviewed by Lamberson et al., 1992; Luoma and Ho, 1993), and most international agencies concerned with environmental management are developing marine bioassay methods and will promote their use to a greater extent in the future. These tests, according to Long and Chapman (1985), are one of three steps in a complete evaluation of the aquatic environment: (1) analytical chemistry, (2) toxicity bioassays, and (3) community structure (e.g. diversity indices) and other field studies (see Malueg et al., 1986; Chapman, 1990). The bivalve embryo bioassay has been adapted to deal with sediments by Chapman and Morgan (1983) and it has subsequently been validated for oceanic, coastal and estuarine environmental quality assessment (Williams et al., 1986; Long et al., 1990; US-EPA, 1991; Butler et al., 1992; Chapman et al., 1992; His et al., 1996; Van den Hurk et al., 1997). Carr (1998) found that an embryo-larval liquid phase bioassay is, on average, one order of magnitude more sensitive than the widely used whole-sediment adult amphipod bioassay. However, sediment bioassays have important limitations because of problems associated with the collection, storage and manipulation of the samples, which may affect the final toxicity (Anderson et al., 1987; Lee and Jones, 1987; Samoiloff, 1989; Swartz, 1989; Burton, 1992; Luoma and Ho, 1993; Beiras et al., 1998; Carr, 1998). Standardization of collection and storage procedures is overdue (Butler et al., 1992; Chapman et al., 1992; Lamberson et al., 1992; Beiras and His, 1995b). Different methods (including different test species) may lead to large differences between the
THE ASSESSMENT OF MARINE POLLUTION
79
estimates of toxicity for the same samples (e.g. Long et al., 1996; but see also Van den Hurk et al., 1997), and even contradicting results have been found (Phelps and Warner, 1990; Butler et al., 1992). Variability of the results is markedly reduced when the techniques to obtain the experimental matrix are standardized. Carr et al. (1996) found excellent agreement (10 of 11 cases) among the results for the same sediment pore waters assayed with three different tests (sea urchin embryogenesis, polychaete reproduction and copepod nauplii survival). The optimum method for the collection of samples depends on the purpose of the study. For biological assessment of the present state of pollution, only the surface layer (e.g. the top 2cm) should be collected. For heavy metal and hydrocarbon analyses, wooden sampling equipment is preferred. For more detailed studies on the speciation, sorption or mobilization of a toxicant, the integrity of the sediment's vertical profile should be preserved and Teflon-lined box-cores may be necessary. Further research is needed to determine the optimum preservation methods and the maximum period during which a sediment may be stored without affecting its toxicity. Fresh storage (4°C) should be limited to few days (2 to 7 days according to Anderson et al., 1987; Swartz, 1989; Phelps and Warner, 1990; Burton, 1992; Luoma and Ho, 1993). A significant change, often a decrease in sediment toxicity, may occur during prolonged storage (Phelps and Warner, 1990; Ho and Quinn, 1993; Beiras and His, 1995b; Beiras et al., 1998), presumably resulting from one or more of the following processes: (1) biodegradation of organics (e.g. Ho, cited by Luoma and Ho, 1993), (2) complexation with organic matter of inorganics and coagulation of the complexes (Landrum et al., 1987), (3) oxidation of reduced irodmanganese species to insoluble oxides and co-precipitation (scavenging) of trace metals (Lee and Jones, 1987; Luoma and Ho, 1993), and (4) reaction of metal ions with acid-volatile sulphides and precipitation of the resulting insoluble sulphides (e.g. CdS) (Di Tor0 et al., 1990). An optimum sample preservation method should prevent these processes from taking place. Drying or freezing the sediments is generally not advised because the structure of the sediment matrix is presumed to affect the toxicity. Unpredictable changes in sediment toxicity resulting from freezing have been reported by Schuytema et al. (1989). We have also found spurious toxicity (false positives) in control sediment samples deep-frozen at -200°C and frozen at -2O"C, but in the long term, freezing was found to be preferable to fresh storage (Beiras et al., 1998). More attention has been focused on the manipulation of the sediment prior to exposure to the biological material. Two basic techniques have been developed: exposure to bulk sediment, and incubation in a liquid phase. The liquid phase may consist of the interstitial (pore) water or it can be obtained by mixing the sediment with control water, i.e. elutriation,
80
E. HIS, R. BEIRAS AND M . N.
L. SEAMAN
with subsequent decantation, filtration or centrifugation. Extraction of the sediment with an organic solvent has also been performed (Wolfe et al., 1995). Bulk sediments are generally more toxic than their elutriates (Van den Hurk, 1994; Geffard, 1997). Even in bulk sediment bioassays, water must be added and some sort of agitation provided because bivalve embryos and larvae are free-swimming organisms. Therefore these methods will be discussed along with elutriate bioassays. The difference is the presence or absence of the sediment during the incubations. Pore water bioassays. In these tests the liquid matrix to be assayed is obtained directly from the sediment. Interstitial water is separated by filtration under pressure (Carr et al., 1989; Long et al., 1996), or by centrifugation (Ankley and Schubauer-Berigan, 1994; Magnusson et al., 1996). The toxicity may depend on the method chosen (Carr, 1998). These are the “softest” methods to obtain a liquid phase, and therefore those which are supposed to represent the actual toxicity of the sediment in the natural environment with the highest accuracy. Nevertheless, elutriation has been used more frequently. Elutriate and sediment bioassays. In these tests, the assayed liquid matrix consists of unpolluted water which is mixed with the sediment. Several different mixing techniques have been used; rotary shaking (Cardwell et al., 1976; Chapman and Morgan, 1983; Williams et al., 1986; Becker et al., 1990), orbital shaking (Thain, 1992), magnetic stirring (Beiras and His, 1995b; Fichet et al., 1998), mechanical plunging (His et al., 1996), mechanical rolling (Butler et al., 1992; Van den Hurk, 1994; Van den Hurk et al., 1997), even hand shaking (Long et al., 1990). Daniels et al. (1989) claim rotary mixing produces the most consistent results. The proportion of sediment to water may vary by more than two orders of magnitude, between 2 and about 500 g 1-’ (sediment weight to total volume). The duration of mixing should be sufficient to saturate the liquid phase with the mobilizable fraction of the toxicants, and this is apparently achieved in as little as 3 minutes of rotation in stoppered bottles at 4rpm (Cardwell et al., 1976), or, according to Van den Hurk (1994), even just hand shaking for a few seconds. Van den Hurk (1994) did not find a significant effect of duration of mixing on the toxicity of the elutriate. Fertilized eggs can be incubated with the bulk sediment providing continuous agitation (e.g. Cardwell et al., 1976), but in this case deleterious effects resulting merely from physical abrasion cannot be ignored. Turbulence and turbidity generally reduce survival of bivalve larvae by 1 to 5 % , depending on the method of agitation (Davis, 1960; Davis and Hidu, 1969b; Seaman et al., 1991). The effects of turbulence and turbidity on larval growth rates are more complex, and depend on biological species, sediment type, and incubation method. At concentrations up to 1g 1-’, the growth rates may be increased or reduced by as
THE ASSESSMENT OF MARINE POLLUTION
81
much as 20% on average, but the variability is very high and may attain between 5% and 260% of the growth rate in controls (Davis, 1960, Davis and Hidu, 1969b; Seaman et al., 1991). At sediment concentrations of 2 to 5 g 1-' average growth rates may be comparable to those of the controls (Seaman et al., 1991), or they may be 40 to 60% lower (Davis, 1960; Davis and Hidu, 1969b). To assess and compare the degree of pollution in different sediments directly from toxicity data, the general geo-physico-chemical characteristics (such as particle size and organic matter content) in the sediments to be compared should be similar. These general characteristics affect the quantity and bioavailability of pollutants bound in the sediment matrix (Anderson et al., 1987; Lamberson et al., 1992). Without such similarity, inter-site toxicity comparisons have very limited value. Fine-silt sediments have a high surface-to-volume ratio, thus facilitating the sorption of a wide range of micropollutants, and their usually high organic content further facilitates the sorption of hydrophobic pollutants (DeGroot, 1973; Adams, 1987; Lee and Jones, 1987; Lamberson et ai., 1992). Long et al. (1990) found that clay and total organic carbon content of the sediments are more strongly correlated with sediment toxicity than heavy metal or hydrocarbon concentrations. Furthermore, it is desirable to record general water quality parameters, such as temperature, salinity, pH and dissolved oxygen in the liquid phase assayed, to guarantee suitable values for larval development. The optimum conditions of temperature, salinity, oxygen and pH for larval rearing in the laboratory have been discussed above (Section 2.3.4). Bivalve larvae are quite resistant to hypoxic conditions. There is no increase in mortality with decreases in oxygen concentration between 3 and 16% saturation (Huntington and Miller, 1989; Wang and Widdows, 1991; Bittkau, 1993). However, embryos are more sensitive to hypoxia than larvae (Bittkau, 1993). Highly hypoxic or nearly anoxic conditions do cause larval mortality, particularly in younger larvae. At a dissolved oxygen concentration of 0.2% saturation, Wang and Widdows (1991) found that survival decreases to 16%. We have found that such extremely low levels of dissolved oxygen may occur in elutriates from reduced sediments with a high organic content, despite thorough stirring during elutriation (R. Beiras, unpublished data). Other potential sources of false-negatives are ammonia and, in particular, sulphide, especially when highly reduced sediments and/or anoxic elutriation conditions are used. Reported no-observed-effect concentrations for ammonia and hydrogen sulphide are 4.7 and 0.1mg1-' respectively (SAIC, cited by Van den Hurk, 1994), while Cardwell and coworkers (1979b) reported ECS0of 15 and <0.01 mg 1-l respectively, and Bittkau (1993) working with 15- and 30-hour-old mussel
82
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
embryos found 50% abnormal development at c. 0.3 mg H2S1-l. Knezovich et al. (1996) found the 48-hour LCso for mussel embryos to be between 29 and 205pg total sulphide per litre, and they expected hydrogen sulphide to be approximately 10% of total sulphide at the pH of the test solutions. Sulphides showed the strongest correlation with toxicity among a number of physico-chemical variables (including heavy metals and oils) in the experimental samples, with sulphide concentration accounting for 81% of the variation in larval mortality (Cardwell et al., 1976). Matthiessen et al. (1998) found toxic effects on C. gigas embryos in elutriates from sediment collected upstream of industrial effluent inputs, and they estimated H2S or NH3 toxicity as a source of false positives in evaluation programmes with elutriate bioassays (see also Sections 4.2 and 4.3). 3.4. Statistical methods
3.4.1. Dose-response curves The relationship between the dose of a potential toxicant and the biological response of the test organism is generally described by doseresponse curves. Preferred mathematical methods to describe this relationship are regression models, in which dose is the independent variable and the biological response is the dependent variable. When we are concerned with lethal toxicity, the response is the proportion of dead individuals (p), a binary discontinuous variable with a binomial distribution. In the case of sub-lethal toxicity, the response is often a normally distributed continuous variable such as growth, although in the context of the present review discontinuous binomially distributed responses are common. Examples include proportion of embryos developing into normal larvae or proportion of metamorphosed larvae. In both cases, a linear model is generally desirable to simplify curve fitting, model fit checking and prediction of the concentration causing a given biological response (Forbes, 1993). For proportion data, linearization of the dose-response curve is achieved by plotting concentrations on a log scale and transforming response values (p) into arcsine pin, normal equivalent deviates ( Z ) , probits (2 5 ) or logits (ln[pl(l - p ) ] ) . Transformed data can be fitted to a straight line, for example,
+
arcsine pl"
=a
+ b log C
(4)
where C is the concentration of toxicant. As mentioned above, all 0% and 100% responses are generally discarded. It may, however, prove impossible to find a mathematical formula that will yield a linear relationship.
THE ASSESSMENT OF MARINE POLLUTION
a3
Once the data have been fitted to a linear regression model, effective concentrations (EC, - values corresponding to an X% reduction in the biological response), or lethal concentrations (LC,), can be calculated by interpolation. LCsois the most commonly used summary statistic, because according to the laws of regression its confidence interval (CI) is narrower than that for any other value of X.The LCso95% confidence intervals can be calculated according to the Litchfield-Wilcoxon method (Newman, 1995) as: Upper 95% CI limit = LCsof (5) Lower 95% CI limit = LCS0.f-' where
f=
LC84 I LC5012,77'" LC50 LC16
and N' is the total number of results between the 16% and the 84% predicted responses. The Litchfield-Wilcoxon method requires partial responses (between 16% and 84%). When such responses are not available the SpearmanKarber method can be applied (Finney, 1964). This procedure, whose advantages have been pointed out by Hamilton et al. (1977), is nonparametric (model-free) and only requires symmetry of the theoretical dose-response curve. This last method, though, tends to underestimate the LCso when all experimental concentrations cause mortalities higher than the control. Guidance for choice between these and other statistical methods is provided by Gelber et al. (1985). Despite efforts to standardize methods, biological material still carries a considerable amount of variability. Differences among stocks of adults and differences in gamete quality are well-known sources of undesired variability in bivalve larval bioassays, and unexposed control replicates are indispensable in the experimental design. In order to overcome this difficulty, the use of a positive control exposed to a reference toxicant such as sodium chloride, sodium dodecyl sulphate (SDS) or dissolved heavy metals, has also been suggested (e.g. Krassoi et al., 1996, 1997). Cardwell and co-workers discussed this subject in the context of bivalve larvae (Cardwell et al., 1977b) and advocate the use of a natural substance such as sodium chloride (Cardwell et al., 1979a). A reference toxicant should be easily available and measurable, highly soluble in water, chemically stable, and should elicit a universal response. Mercury chloride has all of the former properties except for ease of analysis, and, in contrast with sodium chloride, it can be used with euryhaline estuarine animals (see also Beiras and His, 1994, 1995a).
84
E. HIS,
R. BEIRAS AND M. N. L. SEAMAN
Comparisons between experiments are only permissible after correcting the percentages of biological responses with the blanks (without toxicant) using the following equation (Emmens, 1948; Newman, 1995, p. 136, Abbot’s formula): P = (Po - P,)*100/(100 - Pc)
where P is the corrected response (%), Po is the non-corrected original response (%) and P, is the control response (%). This expression has been used to compare responses from experimental and control samples in seawater quality assessment bioassays (see also equation 2, Section 3.2.2.5). 3.4.2. Qualitative assessment of environmental samples
When environmental samples of unknown composition are tested (either water or sediment liquid phases) we can either attempt a quantitative assessment as explained above or we can obtain a qualitative “yes or no” criterion of toxicity, this criterion being a significant reduction in the biological response (e.g. 24-hour embryogenesis success) compared to a non-toxic control. First approaches were based on arbitrary scales such as that by Woelke (1966) (described in Section 3.3.2) who considered as toxic all samples which increased control mortality by more than 15%. We can proceed with a more consistent approach by means of a statistical comparison between a control of water (or sediment liquid phase) from well-known unpolluted sources and the samples whose biological quality we wish to assess. Statistical analysis should then include the following steps: 1. Angular transformation when the biological response is measured in percentages (this is usually the case). 2. Checking the distribution for normality (e.g. chi-square test) and homogeneity of variance (e.g. Bartlett’s test). 3. If the distribution is normal and the variances are homogeneous, ANOVA test for significance of the differences in biological response and a posteriori multiple range tests for identification of problem samples significantly different to the controls (often Dunnett’s test). 4. In case assumptions for ANOVA (2) are clearly unwarranted, nonparametric rank tests (e.g. Wilcoxon) should be applied. When this procedure is used in toxicity tests with known concentrations of single toxicants, it is usually referred to as the NOEC (no observed effect concentration)/ANOVA approach. Strong criticism of the
THE ASSESSMENT OF MARINE POLLUTION
85
NOEC/ANOVA approach arises from the fact that NOEC must be one of the experimental treatments, thus depending to a large extent on the experimenter’s choice of concentrations. Therefore, when toxicity of isolated identified compounds is studied, the dose-response approach and ECx calculation described above are preferred (OECD, 1998). When comprehensive analytical chemistry data for the main pollutants of concern at the different sampling sites are available, identification of the chemicals causing the observed toxicity may be attempted. One possible procedure is to identify toxicants responsible for the observed toxicity by non-parametric Spearman correlation between chemical (concentration) and biological (toxicity) data. Non-parametric rank correlation is preferred because the amount of pollutant is not expected to be normally distributed along the sampling sites. Rather a skewed distribution with a few extremely polluted sites is often the case. Another more ambitious approach would use step-wise multiple linear regression aiming to describe toxicity as a function of the relevant toxicant concentrations weighed by the corresponding coefficients. However, nonlinear interactions between different toxicants (among other causes) prevent these deterministic models from having any predictive value. 3.4.3. Minimum sample size When designing a biological assay with bivalve larvae, two questions arise regarding sample size: how many larvae should be observed in each experimental vessel, and how many replicates (vessels) should be set up for each experimental treatment. To answer these questions adequately we need a preliminary estimate of the variability of the biological response we are to record, both within replicates and among replicates. Let us take first the question of number of observations per vessel. In the case of a variable with a theoretical normal distribution, such as larval size, we can use the following iterative procedure (Zar, 1984). Given an estimation S2 of the population variance obtained from n data, the number of observations, m,needed to estimate the parametric mean with an error s d is:
where 1- a is the confidence level and 1- p is the power of the test. When consulting the statistical tables the t value should be obtained from the two-tailed column while the F corresponds to single-tailed values. In the case of discontinuous variables with only two possible values (normal D-larvae versus other forms, or alive versus dead individuals), we take from a population of N individuals a sample of size n, and record the
86
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
number of individuals, c, with a certain characteristic (e.g. normal D-larvae). The proportion p ’ = c/n is an estimate of the parametric proportion, p , of individuals in the whole population with that characteristic. Burstein (1971) provides us with tables and explanations appropriate for knowing the sample size (n) necessary to estimate p with a certain error margin at a given confidence level. The tables are based on a Poisson approximation when p ‘ 0.25 and a normal approximation when 0.25 < p ‘ s 0.5. (When p ‘ > 0.5, simply find sample size for q‘ = 1- p ‘ . ) This procedure assumes sampling without replacement from an infinite population. In the particular case when sample size is a non-negligible fraction of the total population (>5% N), then n is overestimated, and a correction for finite populations can be applied. The corrected sample size (nF)is in that case (Burstein, 1971):
==
nF= nN/(n+ N - 1)
(10)
For example, given a 25 ml vessel with 20 embryos per ml ( N = 500), the minimum sample size ( n ) necessary to estimate the proportion of normal D-larvae (p) with an error of 0.05 at 95% confidence level, expecting p values of 1, 0.90, 0.75, 0.5, 0.25, 0.10 and 0, would be respectively n = 58, n = 200, n = 334, n = 400, n = 334, n = 200, and n = 58. After correcting for finite populations (since n > 5%N in all cases), then nF = 52, nF = 143, nF = 200, nF= 222, nF= 200, nF= 143, nF = 52. Therefore a sample size of 200 usually allows 95% confidence in our estimate with a low (0.05) error. Moreover, in embryo-larval bioassays, responses often follow a “yes or no” pattern; i.e., extreme values demanding lower sample sizes are more common than proportions close to 0.5. Concerning the question of number of replicates per treatment, we are interested now in the minimum sample size required to detect differences between two population means. Each population consists of the mean for each vessel; therefore they can be assumed to be normally distributed, irrespective of the recorded response (either percentage of mortality or larval size). Again, given an estimation, S2, of the population variance needed obtained from n data, the number of replicates per treatment (m), to estimate the parametric mean with an error s d is (Sokal and Rohlf, 1969):
where 1- a is the confidence level and P is the power of the test. When consulting the statistical tables, the t values should be obtained from the two-tailed column. Applying the methods just described, and aiming for P = 90% and
THE ASSESSMENT OF MARINE POLLUTION
87
a = 0.05, we recommend an experimental design for the bivalve embryolarval bioassays with five replicates per treatment. When replicates are 25 ml accuvettes, 200 observations per vessel to record percentages, and 25 observations per vessel to record normally distributed variables such as size are recommended. This enables estimation of the responses with errors as low as 5% in percentages and 7.5 pm in sizes.
4. TESTING THE TOXICITY OF MARINE POLLUTANTS TO BIVALVE EMBRYOS AND LARVAE 4.1. Pollutants 4.1.lMetals .
Tables 9 and 10 show that the most sensitive biological response to the most toxic heavy metals, such as mercury, silver and copper, is the inhibition of normal embryogenesis, followed by growth reduction. Sub-lethal responses are generally more sensitive than larval mortality. For instance, in the case of copper, the metal most frequently studied, the average values for EC5* are: 39.8 pg 1-' (sd 51.0, n = 26) for embryogenesis inhibition, 63.6 pg 1-l (sd 60.0, n = 17) for growth reduction, and 86.5 pg 1-l (sd 76.1, n = 16) for larval mortality. The heterogeneity of the data, however, prevents these differences from reaching stathical significance. Pediveliger settlement as a criterion of toxicity appears similar in sensitivity to embryogenesis, but data are scarce. According to Watling (1983) about 40 pg Cu 1-l reduces settlement of Crassostrea by half, and Chang et al. (1996) estimate that the value is 5 to 15 pg 1-1 (but see also Phelps and Mihursky, 1986). Other responses, such as respiration and genotoxicity, require a considerable experimental effort and are not suited for routine studies. 4.1.1.1. Embryogenesis. Table 9 summarizes the data on metal concentrations which cause 50% inhibition of normal embryogenesis (ECS0). Despite important variation in test procedures among the different studies, two conclusions can be drawn from a detailed inspection of the table: (1) there is a remarkable consistency in the ranking of the toxicity of the different metals, and (2) the species most frequently used in the bioassays display a similar sensitivity to the various pollutants. For instance, the EC50of mercury to C. gigas is 10.4 pg 1F' (26.7, n = 5) and to Mytilus spp. it is 10.1 (k7.8, n = 4). The average ECS0in pg of metal per litre are (in order from most toxic to least toxic) the following. Mercury, 10 (55.7 sd, n = 14); silver, 19 (29.3, n = 11); copper, 39 (k50, n = 27); barium, 189 (n = 1); zinc, 320 (5497,
Table 9 Toxicity of heavy metals to bivalve embryos. ECS0:toxicant concentration causing 50% abnormal embryogenesis. Chlorides were used for Hg, Cu, Zn, Ni, Cd,nitrates for Ag, Pb, oxide for As and chromate for Cr, except when otherwise stated. Test species
Initial stage (time after fertilization)
Exposure conditions (time, temperature, salinity, density, food, seawater)
EC.
(pg metal ion I-')
HE
Ae
Cu
Zn
Ni
Pb
Cd
Reference and notes As
Cr
Mn ~~
C. gigar
fertilized egg
27°C
C. gigas
fertilized egg
C. gigas
fertilized egg
C. gigas (and 3 other SPP.)
fertilized egg
48 h, 2&2TC, 5 pm FSW uv sterilized 48 h, 2O"C, Mpsu. 24-28ml-l, 5.7 (5.5*) 1 pm FSW uv sterilized 48 h, 20% %25psu, 35 d-', FSW uv sterilized
C. gigas
ferrilized egg
48 b, 20°C 33 psu, 2 M 6 I&', 1 pm FSW 48 h, 20°C. 3Opsu. H)ml-*, low FSW uv sterilized 48 b. 20°C. 24-28 I&'. 1um FSW uv sterilized ' 48h, 25'C, 3635psu. 5 p m FSW
c.g1glLF
fertilized egg
C. gigas
fertilized egg
C. gigas
fertilized egg (15 min)
C margariracea C. cucullafa C. gigas
egg
C gigas C. gigas
egg
C. gigas
fertilized egg
C. gigas (and 5 other
fertilized egg
32-100 1 O C G 3200 -250
1&32
>2500,920* 13900. 1290' 370,60* 1700.590* 1618
C. virginica
fertilized egg (15 mi.) fertilized egg
C. virginica
fertilized egg
(1h) C virginica
fertilized egg (2 h)
Coglianese and Martin, 1981; Cu nitrate; abnormal larvae excluded Knezovich ef a/., 1981: *humic matter 2mgI-', **EDTA 1 0 - 6 ~ .
10 12, 40,* 100"
6.7
48 h. 26°C 28 psu, 30 d-', 0.8 pm FSW 24, 48 h, 24% 32psu, 20 n P , 0.2 um FSW 48 h, b C %36 d-', 33psu 1 pm FSW uv sterilized 23psu 24 h, 24T, 20, 25,3Opsu, 0.2 pm FSW 24 to 72 h, methods not given
22
180*
>loo >loo >loo >loo
>loo
Martin ef al.. 1981; Cu,Zn and Ni sulphates; abnormal larvae excluded Wathng, 1981; *calculated by
160* 130* 13
1100 >loo >1w >loo >loo >loo >loo >loo
>1W >100
His and Robert, 1981; abnormal larvae
5.3
119
349
758
611
326
4538
excluded
larvae excluded Coghanese, 1982; Ag nitrate; abnormal larvae excluded
218 5-6.5 W13.5 5 4 . 5
>so 44
1&100
24 b, 24°C 24 d-', 0.2 pm FSW
13
42-48 h, 26°C 15-17 ml-', ASW 44.48 h, 26psu. 20-22d-', 1 pm FSW uv sterilized 48h, 20. 25, 30°C 22.5, 27.5psu, 1 pm FSW
5.6
5.8
103
310
20°C 11.4
24.2 35.3 32.2
15.1 18.7 18.3
206 325 230
25'C 12.6 30°C 10.2
1&20
extrapolation.
His and Robert, 1982; sulphate; abnormal
10
SPP.)
C. gigas
Okubo and Okubo. 1%2; Cu and Zn sulphates Brereton d al., 1973: most larvae abnormal from Zn 150. Glickstem, 1978; abnormal larvae excluded Se decreases Hg toxicity; *nitrate. Cardwell ef ul., 1979a; all sulphate; 'fully shelled regardless of shape.
5w1000
1180 2450 3800 7500* 10300
Robert and His, 1985; abnormal larvae excluded Zhadan ef al., 1992; approximate ranges; sulphates; abnormal larvae excluded Beuas and His,1994; abnormal larvae excluded 16000 Calabrese ef al., 1973; *arsenite MacInnes and Calabrese. 1978; abnormal larvae excluded MacInnes and Catabrese. 1979; 220 at 2% 27.5 psu; abnormal larvae excluded
48h, 25psu. 60ml-, FSW sterilized
C. iredalei
fertilized egg
M. edulis
fertilized egg
1?-17"C
M. edulis
fertilized egg
M. edulrr
M. galloprovincialis
fertilized egg (2 h) fertilized egg (0.5-2 h) fertilized egg
48 b, 17°C. 2 4 2 8 ml-l,1 pm FSW uv sterilized 48 h, 19°C. 26psu, 151&', 5 pm FSW uv sterilized 2 d. W C , 35 psu, 30 I&'
M. galloprovincialis
fertilized egg
M.galloprovincialk
fertilized egg
M. mercenaria
fertilized egg
M. galloprovmcialis
S.soltdissimo
I. califomicum
fertllued egg (1h) fertilized egg (2 h) fertilized egg (4 h) fertilized egg (15min) sperm + egg (1 h later)
I. califomicum
fertilized egg
M. lateralis
S. plano I. californicum
(M Chlamys asperrima Chlamys aspemma
fertilized egg (>I h) fertilized egg (30 min)
10-32 5.8
14
32-100 3.2-10* 5.8
24
3.5
891
476
13'7
us*
4426 4743**
145
4055
Hrs-Brenko el al., 1977; abnormal larvae excluded Pavicic. 1980: Z n sulohate. *60h at 16°C; "60h at 1 6 k abnormal larvae excluded Pavicic er al... 1994a:. abnormal larvae excluded Beiras and Hi$ 1995a; abnormal larvae excluded Calabrese and Nelson, 1974
10 4.8
21
166
310
780
Eyster and Morse, 19W, abnormal larvae excluded
6.4-9.5 c. 16
48 h, 21°C. 10, 3Opsu, 75 I&'
18,17
48 h, ZOS'C, 24psu. 25 I&', 0.45 pm FSW 48 h, 2 4 T, 34psu (24psu), 2 ml-', 0.22 wm FSW 2 h. temnerature not eiven. 34psd 200 I&', 0.8 FSW 48 h, temperature not given, 0.45 pm 34psu, 17 d-', FSW 48 h, 18°C 33psu, 30ml-', 1pm FSW uv sterilized 48h, 18'C, 33psu, U)ml-', 5 sm FSW uv sterilized
10-20
&
320loo0 175
250-500
48 h, 20"C, 37-38psu 48 h, 20°C 37-38 psu, 3 W m1-l. marselv FSW 48 h, 20°C, ml-', 0:2 pm FSW 42-48 h, 26"C, 15-17 m1-I. ASW 48 h, 20"C, 3Opsu, 30 ml-l, 0.22 pm FSW uv sterilized
Ramachandran er al., 1997; Cu sulphate; abnormal larvae excluded 3 2 W >loo00 Okubo and Okuho. 1962; Cu,Zn and too00 Cr sulphate; *acetate. 1200 >3000 4469 Martin el al., 1981; Cu,Zn and Ni sulphate; abnormal larvae excluded 30000 Morgan er al., 1986. 459
81
Momson and Petrocelli, 1990. sulphate. Ruiz el 111.. 1995~:nitrate; abnormal larvae excluded, control embryogenesis success <4% Ringwood, 1990; results for 24 psu salinity in brackets
55
Ringwood, 1992b; endpoint: cleavage. Ringwood, 1992b; abnormal larvae excluded
7
48*
39'
266
I26
4.5.
45
295
I350
Krassoi et al., 1996; Cu and Zn sulphate; abnormal larvae excluded Krassoi er al., 1997, Cu sulphate: abnormal larvae excluded
Bold 95% CI provided. ASW artificial seawater; FSW filtered seawater. Other metals: Al C.virginica 42-48 h EC50>7500 (Calabrese er 01.. 1973: chloride); M.galloprovincialis 48 h E C , 54-162 (Pagano ef al.. 1996, sulphate): Saccosrrea commercialis 48 h E C , 200 (Wilson and Hyne, 1997; sulphate). E a Mytilus colifominnus 48 h E G O= 189 (abnormal larvae as judged by polarization o p t i q Spangenberg and Cherr, 1996). Sr Mytilus californianus 48 h ECso >50o00 (abnormal larvae as judged by polarization o p t i q Spangenberg and Cherr, 1996). Se C. gigas and M. edulis 48 h ECS, >loo00 (Martin er al., 1981); C. gigar 48 h EC, >loo00 for SeOl and Na2Se03(Glickstein. 1978). Fe M. edulis E C , 1OMxI-32000 (Okubo and Okubo, 1962). M o M. edulis 48 h E C , = 147000 (Morgan er al., 1986).
Table 10 Effects of heavy metals on mortality, growth and metamorphosis of bivalve larvae. LC50:toxicant concentration causing 50% mortality; otherwise EC50: toxicant concentration causing 50% reduction in the end-point. Nitrate for Ag and chlorides for the other metals were used except when otherwise stated. No attempt has been made to calculate accurate LCsOvalues from original data, because of variability between trials within an experiment and, particularly, differences between nominal and actual concentrations. Rather, concentrations above andlor below 50% effect are provided. LC, or EC, fw metal ion 1.')
Exposure conditions hitial agekize
Test species
C. gigas C gigas C gigas
21 d, 250-320 pm 5d,88pm 16 d, 260 pm 6d,135~ 16 d, 309 pm
(time, temperature, salinity,
density, food, seawater) 5 d, 21°C 1ID-', mix algal food. 2-5 pm FSW uv
sterilized 4 d. U25"C 2.5 ml-l (young) 1.5 ID-' (old), mix algal food, 5 pm FSW 4 d, 22-23'C 2.5 ml" (young) 1.5 ID-' (old), mix algal food, 5 pm FSW
End-point
Hg
Ag
Cu
Zn
50
80
>loo
>loo
>lW
85 1100 75 >lo0
mortality
>lW >loo
C cucullata
3d,75pm 13d, 254 pm 24 h
60 >1M) 2447
>loo
c.gigas
1d,Mpm 211 pm 3 1 0 ~
C. virginica
2d
C. virginica
2d
C virginicn
48 h, 70 mn
0. edulis
1-3 d
12 d. 25°C. 10-12 I&', mix algal food, 1pm FSW 48h, 17.5.22.5,27.Jpsu, 20 d-'mix , algal food, 1pmFSW 8 d. 26°C. 27-28 psu, 20 I&', monoalgal food, 1 pm carbon FSW uv sterilized 4.2 h, 15°C0.8 ID-',no food
Cr
Referenceandnotes
Watling, 1978
200
60 >loo
monoalgal food. 0.8pm FSW 48 h, 24°C 24 I&', no food, 0.2 pm FSW
Cd
Boyden er d., 1975; sulpbate
mortality
3d,80pm 13d, 239 pm
5 d, 26°C. 28psu. 8 DIP,
Pb
>500
mortality
C margaritam
C gigas
Ni
mortality mortality (tissue degeneration)
m
mortality
12.0
Watling, 1982
80
>loo
>lo0
His and Robert, 1981 1994 & Beiras and €I
33
115 25.0
32.8
1200
Calabrese el a/., 1977a
>90 60-90
mortality
Macnes and Calabrese, 1979: *3060 at 1 7 . 5 salinity ~ ~ ~
&!XI*
mortality
mortality (no response touching with needle)
>86
3.3
227
Wikfors and Ukeles 1982
Connor, 1972
M. edulis
3d
15 d, 15% 32psu. 5 ID-'.mix algal food, 0.2 pm FSW
M. edulis
150 pm
10d, 1SoCc, 32psu, l o r n - ' , monoalgal food
M.galloprovincialis 2 d. 76 um 141 pm 225 pm 258 p m
48 h, 20"C, 24 m1-l. no food, 0.2 rn FSW
mortality (tissue degeneration) mortality
M. rnercenaria
2d
b 1 0 d. 25°C. 10-12 ID-',mix algal food, 1 pm FSW
I. californicurn
3d 10 d 24 d 36 d
48 h. 24'C. 34psu ( 2 4 ~ s ~ ) . mortality 1ID-],no fbod,'O.ZZ pm FSW
P mnxirnus
3d
C gigas
2 d, 50-70 pm
C. gigas
5d,88pm 16 d, 260 pm
15 d, 15°C. 32 psu. 5 ml-'. mix mortality (inactive algal food, 0.2 pm FSW oesophageal cilia) growth (length) 5 d + 5 d recovery, 26"C, 5 ID-',mix algal food, 5 pro FSW uv sterilized 7 d , 2 5 T , 2.5 d-' growth (length) (young), 5 d, 23°C 1.5 ID-' (old), mix algal food, 5 pm FSW growth (length) 4 d, 22-23", 2.5 ID-' (young) 1.5ml-l (old), mix algal food, 5 pm FSW
C. gigus
6 d, 135 pm
C cucullnra
51 164 322 383 14.7 32.4
Stromgren and Nielsen. 1991; sulphate; silicon-coated culture jars Beiras and His,1995a
16.4
195
Calahrese e t a / . , 1977a
5700
Ringwood, 1990.results at 24psu salinity in brackets
100
Beaumont er 01.. 1987 Brereton er a/., 1973; sulphate
-150
>1M)
75 120
3d,@Jw 13d, 239 pm
35 85
45
40
85
1W
3 d, 75 pm 13 d, 254 pm
40 85
50
45 120
egg 24 h
C. gigas
egg 24 h fertilized egg
7 d, 26°C 28psu. 6 ml-', growth (height) monoalgal food, 0.8 pm FSW mix growth (height) 12 d, 24% 32psu. 8 d-', algal food, 0.2 pm FSW 10d, 24°C 20,25,30psu, growth (height) 8 mix algal food, 0.2 pm FSW
I&',
Watling, 1978
1W7.00 80
C. gigas
C gigas
10
50 75
16d,309~ C margariracee
Beaumont er al., 1981
2W500
mortality (inactive oesophageal cilia) mortality
95
85
Watling, 1982
-12 12-24
His and Robert, 1981
613 -64
His and Robert, 1982: sulphate Robert and His, 1985
250
~~
~
Exposure wnditions (time, temperature, salinity, density, food, seawater)
Initial age/size
Test species
LC, or ECx (pg metal ion I-') End-point
Hg
c. gigar
ld,60&m
10 d, 24°C 8 I&', mix algal food 0.2 pm FSW
growth (height)
7.1
C virginica
2d
growth (length)
11.8
C. virginica
48 h, 70 ~ L I J
, 12 d, 25°C 10-12 d-'mix algal food, 1f l FSW 8 d, 26'C. 27-28psu. 20 ml-', monoalgal food, 1 pm carbon FSW uv sterilized
growth (length)
0.edulis
179,181pm
%h
3d
M. edulis
150 m
15 d, 15% 32psu. 5 d-l,mix growth (height) algal food 0.2 ~ L IFSW J 10d, 15°C32psu. lOml-', growth (length) monoalgal food
M.mercenaria
2d
b 1 0 d, 25°C. 10-12 d-'mix , algal food, 1pm FSW
growth (length)
21.4
fl maximur
3d
15 d. 15°C 3 2 ~ s5~d.-'. mix growth (height) algal food, 0.2 pm FSW
1. californicum
3d.85~1~
28 d, 24°C 34psu. l O d - ' , mix algal food, 0.45 pm FSW 4 d. temperature not given, 34 psu. 10 d-', mix algal food, 0.45 pm F'SW
growth (dry weight) (tissue dry weight)
72 h. 2C-22pC. 34psu. Mml-', 1~ L I FSW J
Ca uptake
5 d, 21°C. 1ml-', mix algal
settlement
21 d, 250-320 eyed
~LIJ,
food. 2-5 sterilized
~ L I FSW J uv
Cd
Cr
Referenceandnotes
30.2
47.1
Calabrese ei al., 1977a
1078 >27
Wikfors and Ukeles 1982
Stromgreu and Nielsen, 1991: sulphate; siliwnsoated culture jan 7.4
C. gigas
Pb
Walne. 1970
growth (height)
9d, 118pm
Ni
Beaumont et al., 1981
10d, 20°C. 8 d-'mix . algal food, 0.2 pm FSW
M. laiernlir
Zn
>86
M. galloprovincialis 2d,75fim
3d
Cu
Beiras and His 1994
growth (length)
M. edulis
I. californicum
Ag
Beiras and His 1995a 42.2
16.9
232
<110 0
Calabrese et a[.. 1977a
20-100
Beaumont el al., 1987 >20
Ringwood, 1992a
-20 -100
>20
growth
Ringwood, 1992b
(dry weight) 26.5
18.5
176 <125* >500'*
Ho and Zuhkoff, 1982 Boyden et al., 1975; sulphate; ' 5 d continuous exposure; **5 d exposure C5 d rewvery in clean FSW; at Z n 125 and 250 no marked reduction in settler number but 2 d delayed settlement
C. gigas
16 d
C. gigas
19 d
c. gigas
2290 pm, eyed
C. gigas
2280. eyed
C. virginica
2230, eyed
C virginica
eyed
6 d exposure f 3 d recovery, 22-u"C 1.5 ml-'. mix algal food, 5 pm FSW 20 d
3445
settlement*
1&20*
idem
4 d, 24T, 11 n f ' , no food, 0.2 pm FSW, eplnephrine lO-'mol. added d 2 4 d, 37 psu, ASW, epinephrine mol. added at different tunes
metamorphosis* metamorphosis
c. 1wo
c. loo0
4 d, 25 psu, ASW, mol. epinephrine added at different times 20 I&', flowing 0.5 p m FSW
metamorphosis
c. 500
c. 5M)
settlement (4 h) +spat mortality (4-7d)*
534 ppm**
ASW, artificial seawater; FSW, filtered seawater.
2&25
3c-35
3244
1&20 1&20*
1c-20
Watling, 1983; *fixation to bard substrate 1&20
Watling, 1983 (only approximate figures given); '1 d delayed settlement Beiras and Hi$1994 *adulr shell formation G a n g ef ul., 1996: nominal concentrations; estimated actual concentrations 2 orders of magnitude lower for cu
Phelps and Mihursky, 1986: *fan-shaped non-empty shells; **Cuconcentration in the setting surface micro-organism laver; nitrate
94
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
n = 16); nickel, 683 (2425, n = 4); lead, 968 (2847, n = 5); cadmium, 2219 (23312, n = 18); arsenic 3913 (25073, n = 2); chromium, 4564 (23663, n = 6); and manganese, 23 000 (*9899, n = 2). In the case of copper toxicity the estimation of the EC5,, is much lower and more precise if the EC5o values extrapolated from Watling's (1981) data are considered as outliers and disregarded: 24 2 28 (n = 24). According to their embryotoxicity, metals can be classified into three groups. Dissolved mercury, silver and copper are toxic at the 10ppb (i.e. lo-*) level, zinc, nickel and lead are usually toxic at levels from 100 to 1000ppb to and cadmium, arsenic and chromium often show embryotoxicity to level only. Other bivalves at concentrations above the 1 ppm elements occasionally tested and found virtually non-toxic are selenium (EC50 > 10ppm), iron (ECSO> 10 ppm), strontium (ECS0> 50 ppm), and molybdenum (EC50 > 50 ppm) (Okubo and Okubo, 1962; Glickstein, 1978; Martin et al., 1981; Morgan et al., 1986; Spangenberg and Cherr, 1996). Data on aluminium toxicity (not included in Table 9) to bivalve embryos are scarce and contradictory. Calabrese et al. (1973) found no effects on embryogenesis at concentrations up to 7500 pg A1 1-', while recently Pagano et al. (1996) found toxic effects between 54 and 162 pg 1-', and Wilson and Hyne (1997) report ECSovalues of about 200 pg 1-l. In all three cases pH was within the limits of tolerance for bivalve embryos and larvae (27, Calabrese and Davis, 1966,Wilson and Hyne, 1997). However, Calabrese et al. (1973) used aluminium chloride, whereas Pagano et al. (1996) and Wilson and Hyne (1997) used aluminium sulphate; the differences in toxicity may therefore be a result of differences in speciation (see Section 4.3) of the salts in seawater. 4.1.1.2. Larval mortality and growth. The lethal and effective concentrations shown in Table 10 are even more difficult to compare than those for embryos (see Table 9) because of two additional variables, age of the larvae and duration of the experiment. As discussed below, larval tolerance increases with size in terms of survival (Watling, 1978;Beiras and His, 1994, 1995a) and growth (Watling, 1982). Also, detectable effective concentrations impairing larval growth depend on the duration of the experiment, and in the studies reviewed here this ranged from 4 to 28 days. In a number of cases, larval growth is a more sensitive response to toxicants than inhibition of embryogenesis (e.g. Beiras and His, 1994, 1995a). Cadmium, for instance, has been shown to inhibit larval growth at significantly lower concentrations than those which inhibit embryogenesis (compare Tables 9 and 10). Irrespective of its sensitivity, however, a bioassay using growth as the biological endpoint is neither simple nor quick thus, it is not suited for routine screening tests. Nevertheless, effective pollutant concentrations at which larval growth is significantly reduced are still worth studying because of their ecological relevance.
THE ASSESSMENT OF MARINE POLLUTION
95
Ho and Zubkoff (1982) studied the effective concentrations of some heavy metals which reduce calcium uptake in bivalve larvae, and found these concentrations to be similar to those affecting growth (see Table 10). Therefore, calcification of the organic matrix of the shell may be one of the processes impaired by heavy metals and can therefore be used as a measure of sub-lethal toxicity to bivalve larvae. 4.1.1.3. Settlement. Data on the effects of heavy metals on settlement and metamorphosis in competent pediveligers are scarce (Table lo), but they do indicate that the sensitivity of this ecologically relevant stage is similar to that of embryos and larvae. Bioassays with settlement as the endpoint might be particularly relevant for testing the toxicity of surfaceassociated toxicants, such as those that adsorb readily to sediment particles. A shortcoming for the development of settlement bioassays is the limited knowledge of the physico-chemical factors that naturally promote attachment to the substratum and morphogenic change in bivalves. Some advances have been made with the use of epinephrine, a potent inductor of metamorphosis in bivalve larvae, allowing massive and synchronous settlement in the controls (see Section 3.2.4.5). On the other hand, heavy metals do not necessarily inhibit larval settlement. Sub-lethal concentrations of copper are known to stimulate settlement in oyster larvae (Prytherch, 1934, in C. virginica; Nell and Holliday, 1986, in Saccostrea commercialis; P. Salkeld, Plymouth Marine Laboratory, personal communication, in C. gigas). 4.1.1.4. Respiration. Heavy metals are known to increase respiration in all developmental stages of bivalves, including eggs and larvae. Oxygen consumption increased by 1.2 times to 2.7 in Spisula solidissima larvae reared at 50 pg Ag 1-' (Thurberg et al., 1975) and by about 2 times in unfertilized eggs of Mytilus edulis exposed to 30 000 pg Cu 1-' (Akberali et al., 1985). Mercury also increased oxygen consumption of 5-day-old C. virginica larvae after 24-hour exposure to 1-100 pg 1-' (Cunningham, 1972, cited by Cunningham, 1979). However, zinc (100 pg 1-') and cadmium (2000 pg 1-l) have been reported to cause slight (1620%) decreases in respiration in mussel embryos (Pavicic, 1980). The sensitivity of this response is not substantially different from the test of embryogenesis inhibition, whereas the latter is much more simple, rapid and accurate to measure. Moreover, these tests have yielded contradictory results (generally an increase, but also some cases of decrease in respiration) which are difficult to interpret. Therefore respiration does not seem to be a suitable response for embryo and larval bioassays. 4.1.1.5. Genotoxicity. Mercury has been shown to be genotoxic, as exposure to 30 pg HgC121-1 doubles the mean number of sister-cromatid exchanges (SCE, a sensitive indicator of chromosomal damage), in fertilized mussel eggs (Brunetti et al., 1986). Nitrilotriacetate (a com-
96
E. HIS, R.
BEIRAS AND M. N. L. SEAMAN
ponent of some phosphate-free detergents), which is suspected of having mutagenic effects, did not increase SCE at the concentration tested (5mg1-'), nor did it interact with the genotoxicity of mercury chloride (Brunetti et al., 1986). TBT-oxide (up to 1pgl-') was not found to be genotoxic to 12-hour-old mussel larvae either, based on the results of SCE and chromosomal aberrations (Dixon and Prosser, 1986). Fractions extracted with an organic solvent from oil-polluted sediments have also been assayed for genotoxicity in bivalve larvae (Wolfe et at., 1995), but the results are difficult to interpret, because there was significant SCE induction by the reference sediment as well. 4.1.2. Biocides
The available data on biocide concentrations causing 50% reduction in normal embryogenesis, larval survival, growth and settlement are summarized in Table 11. Chemicals are listed in tentative order of decreasing toxicity, based on the highest toxicity values reported. In order to focus on the most relevant environmental pollutants, we are limiting this review to the most toxic biocides for bivalve larvae, i.e. those with ECS0values of less than 10 ppm (10 000 pg 1-') in at least one experiment. Substances of lesser toxicity, for which all published studies give an ECS0higher than lOppm, are not considered here regardless of the fact that they may be highly toxic to organisms other than bivalve larvae. Tributyl-tin (TBT) is the most toxic chemical assayed to date with bivalve embryos and larvae, with an EC50 below the unit ppb level in several instances. This is not surprising, considering that oysters and other molluscs are target species for this active component of antifouling paints. Halogens (chlorine, bromine) and halogen-produced oxidants (such as monochloramine, chloroform, bromamines, bromide, bromate, halomethanes, etc.) are also extremely toxic. Accurate measurement of EC50 values is difficult owing to the high volatility of these substances, and experimental designs should employ flow-through systems. Identification of individual halogen compounds is usually not attempted, but Stewart et at. (1979) found high toxicity of bromate and comparatively low toxicity of chloroform and bromoform. Data on ozone-produced oxidants are shown in Table 11 for comparison. Again, bivalve larvae are target species of these compounds, used in energy plants to prevent their cooling-water pipes clogging. Cyanide (CN-) is highly toxic to all aerobic organisms, and early developmental stages of bivalves are no exception, with ECsovalues in the 10-100 pg 1-' range (Okubo and Okubo, 1962; Pavicic and Pihlar, 1982; Pablo et al., 1997).
THE ASSESSMENT OF MARINE POLLUTION
97
Intensive agriculture increasingly relies on the use of phyto-sanitary chemical products for pest control. In 1992 the world consumption of these products was approximately 1.4 million tons (Andral, 1996). In coastal areas these products are leached from the soil by precipitation and drained into the sea. The toxicity of a great number of insecticides, herbicides and other biocides has been assayed with early developmental stages of bivalves, but the most important contribution still remains the pioneering study by Davis and Hidu (1969a). More recently, His and Seaman (1993) extended these studies of lethal and sub-lethal toxicity in embryos and larvae to 12 commercially available pesticides. Some organochlorinated pesticides such as trichlorocarbanilide (TCC), 2,3dichloro-l,4-naphthoquinone(Phygon), dichlorodiphenyltrichloroethane (DDT), pentachlorophenol (PCP), dehydroabyetilamine compounds (Delrad, Rosin Amine D), and cyanide can be toxic at concentrations below 100 pg 1-l. Toxicity of chlorophenols increases with the number of chlorine substituents, from an ECsO in the order of 20000pg1-' for 4-chlorophenol (4-CP, one chlorine atom) (Krassoi et al., 1997), to values around 1000 for tetrachlorophenol (TCP, four chlorine atoms) (Davis and Hidu, 1969a), and 27-55 pg 1-' for pentachlorophenol (PCP) (Woelke, 1972). Another group of persistent chlorinated insecticides (dieldrin, endrin, lindane, methoxychlor, toxaphene, aldrin) and copepodicides (dichlorvos) can be toxic at 100-500ppb. Dichlorvos, a highly toxic pesticide used in salmon farming for the control of ectoparasitic copepods, is particularly relevant in marine toxicology, because it is deliberately introduced into the marine environment in appreciable quantities. Crustacean larvae are ten times more sensitive to Dichlorvos than molluscs (LCs0: 5-50 pg 1-' for lobster larvae, according to McHenery et al., 1991). Organo-phosphorated pesticides are generally less toxic, with EC50values in the range of 500-1000 pg 1-' (Parathion, Fenitrothion, Guthion, etc.), or higher (Malathion). Herbicides (e.g. 2,4-D, atrazine-simazine, Neburon, Diuron, etc.) are usually toxic to bivalve embryos and larvae only at concentrations above the ppm level (e.g. Robert ef al., 1986), with the exception of Dinoterbe, a highly toxic phenolic herbicide used in maize culture, which has an ECsOof 50-100 pg 1-'. As in the case of metals, pesticides affect embryonal development more than larval survival. Growth reduction may take place at concentrations below those inhibiting embryogenesis, although this is not always the case (see Table 11). The sensitivity of larval growth is particularly dramatic in the case of TBT, which reduces the rate of growth at concentrations in the order of 1ng 1-l (His et al., 1983; Lapota et al., 1993). In those cases in which there are sufficient data for comparison, the various bivalve species (oysters, mussels and clams)
Table 11 Toxicity of pesticides, biocides and miscellaneous chemicals to early developmental stages of bivalves. LCso: toxicant concentration causing 50% mortality; otherwise ECSo:toxicant concentration causing 50% reduction in the end-point. Chemicals are tentatively ranked in decreasing toxicity order. Except for TBT,data on other toxicants are mostly from the classic work by Davis and Hidu (1969a); these authors refer to Loosanoff and Davis (1963) for details on experimental procedures. Davis and Hidu (1969a) and His and Seaman (1993) used acetone as a carrier when appropriate. Pesticide TBT
Test sp.
Initial age/&
Exposure conditions
End-voint
Notes
C. gigas
30 min fertilized egg
24h, F W 24% ZEpsu, 0.8 pm
embryogenesis
<5
His and Robert, 1980
C gigas
30 min fertilized egg
24 h, 24°C 28psu. 0.8 pm
embryogenesis
(1
Robert and fig1981
c gigas
30 min fertilized egg
FSW 48 h, 24'C 28psu, 0.2 pm FSW
embryogenesis
0.5-1.
C. virginica
4-16 cell embryos
M.galloprovincialis 3 0 m h fertilized egg M. mercenaria
4-16 cell embryos
I. cnlifomicum
sperm + egg (1 b later) fertilized egg
1. difomicum
- - ..
48 h, 20-24'C, 1&22psu, 40,'&I 1pm carbon FSW 24 b, 1YC 28psu, 0.8 pm FSW 48 b, 20-24T, 1&22psu, 40d-', 1pm carbon FSW 2 b, 34psu, ux)ml-', 0.45 pm FSW 48 h, 34psu, 17d-', 0.45 @n FSW 48 h, 24'C 28psu, 0.8 pm FSW 24% 28psu. mix algal food, 0.8 @n FSW
His ef al., 1983
embryogenesis
1.30.
embryogenesis
<1
Robert and Hig1981
embryogenesis
1.13
Roberts. 1987
cleavage
4.0
Ringwwd, 1992b
Roberts. 1987
(acetate) (abnormal larvae excluded) (acetate) (abnormal larvae excluded) (acetate) (abnormal larvae excluded except visceral mass abnonnalities) '(0.71 acetone as carrier) (acetate) (abnormal larvae excluded)
embryogenesis
0.5
Ringwood,1992b
(abnormal larvae excluded)
mortality
<5
His and Robert, 1980
mortality
M (4 d)
(acetate, glacial acetic acid as carrier) (acetate, glacial acetic acid as carrier)
C gigas
egg
24°C 28psu, 8 d-l,mix algal food, 0.2 w FSW
mortality
c gigas
not given
48 h, up to 1I&'
mortality
C virginicn
24 h
mortality
M. edulis
not given
48 h. 2&24T, lS-22psu. 10ml-', mix algal food, 1 pm carbon FSW 48 h. up to 1
M. edulis
7 d, 133J.UU
mortality
15 d, WC. 33psu. 4 d-', mortality (oesophageal mix algal food, 0.2 pm FSW uv sterilized cilia ceased heating)
Robert and His 1981 1-3 (5 d) <1 (9 d) 0.5-1 (6 d) His ef al., 1983 0.24.5 (8 d) 0.0210.05 (12 d) Thain, 1983 1.6 3.%
Roberts 1987
2.3
Tbain, 1983
0.100
Beaumont and Budd. 1984
(acetate, glacial acetic acid as carrier) (bisTBT oxide, acetic acid 01 ethanol as carrier) (cbloride, glacial acetic acid as camer) (bisTBT oxide, acetic acid or ethanol as carrier) (bisTBT oxide, carrier glacial acetic acid)
M. edulis
8d, 140pm
M. galloprovincialis 1d(?), 73 pm M. mercenaria
24 h
M. mercenaria
24-48 h
S. plana
10 d, 240 pm
C. gigas
24 h, 61pn
c. gigm M. edulis
7 d, 133pm
M.edulis
8 d , 140pm
M.galloprovincialis 24 h
CI
(?), 73 pm
M. mercenaria
4 h fertilized egg
S. plana
2d, loOpm
s. p l M 4
10d, 239pn
I. californicum
3d
C. virginica
gastrulaltrochophore
M.mercennria
25 d, 19T, 2 d-', mix algal food, 50 pm FSW
mortality (empty valves)
4 d, lPC, 28psu, mix algal mortality food, 0.8 pm FSW mortality 48 h, 2&24°C, 18-22psu, 10 ml-', mix algal food, 1 pm carbon FSW mortality 8 d, temperature not given. 3Oosu. 0.1 dP.monodeal food 30 d, 20.5"C, 24psu. 0.8 d-',mortality' monoalgal food, 0.45 wm FSW 48 h, W C , Bpsu, mix algal growth (height) food, 0.8 pm FSW 8 d, 24°C. 28psu. 8 d-'. mix growth (height) algal food, 0.2 pm FSW 15d, WC, 33psu, 4 d - ' , growth (length) mix algal food, 0.2 pm FSW uv sterilized 25 d, 19T. 2 mi-', mix algal growth (length) food, 50 pm FSW growth (height) 24 h, 19'C, 28psu, mix algal food, 0.8 prn FSW growth (length) 14 d, 25"C, 32psu, 3 ml-', monoalgal food 6 d, 22.5"C, 24psu, growth (length) monoalgal 0.4-3 d-', food, 0.45 pn FSW 30 d, 20.5"C24psu. 0.8 d-',growth (length) monoalgal food, 0.45 pm FSW 4 d, temperature not given, growth 34psu, lOd-', mix algal (dry weight) food, 0.45 pn FSW 17-28'C*, lMOpsu
mortality (48 h)
>0.13* 5
1.65 -0.6. 0.13-0.31
0.1-1
Lapota er 01.. 1993
Robert and Hi$1981 Roberts 1987 Laughlin ef al., 1989
*1-2.5 at 48 h @isTBT oxide in acetone)
Ruiz ef al.. 199%
(bisTBT oxide, ethanol as camer) 'nowexistent or inactive soft puts inside gaping valves (acetate, glacial acetic acid as carrier) [acetate, glacial acetic acid as carrier) (hisTBT oxide. glacial acetic acid as camer)
Robert and Hi% 1981
His et al., 1983 Beaumont and Budd. 1984
0.W64.050 Lapota ef al., 1993 <1 0.0~.10
Robert and
His 1981
Laughlin el al., 1988
(TBT salt and camer not given) [acetate, glacial acetic acid as carrier) (hisTBT odde, acetone as carrier) (bisTBT oxide, ethanol as carrier)
<0.13
Ruiz ef al., 1995a
<0.13
Ruiz er al., 199%
(bisTBT oxide, ethanol as carrier)
>0.5
Ringwood, 1 W b
(TBT chloride)
Roberts er a/., 1975
'those ambient depending on season, no more than e2'C within a single experiment **depending on methods
<s-110** <5
(TBT salt and camer not given) 'making allowance of control mortality (acetate, glacial acetic acid as carrier) (chloride, glacial acetic acid as carrier)
Test sp.
Pesticide
Phygon
Delrad
End-point
ECsa or L% I-')
Reference
7 d larva
open-flow system
C gig-
0.5-2 h 1st polar 2onl-'
19, 22, 25,ZB"C 18,26,34psu
3-7 h blastula 20 tn-' 6 1 6 b trocophore 20 tn-' 1 6 3 0 b larva 5 d-', monoalgal diet 24-50 h lama 5 mlP, monoalrral diet pediveliger
19, 22, 25, 28'C 18,26,34psu 19.22, 25. 28°C 18,26, 34psu 19.22, 25, 28'C 18,26, 34psu 19, 22, 25, 28°C 18, 26, 34mu
1B"c
120 (20°C) Capulzo, 1979 80 (U'C') 10 (ZOT)** <10(25T*)** no cleavage (1 h) 1031, 1303, Chien and Chou, 1989 1322.1070 852, 1273, 1425 456, 524,606, 654 mortality 458, 571,651 (1 hY 384,362,502,926 mortality 430, 524, 677 (1 h)' 556,729. 867, 950 mortality 688,766,872 (1 b)' 607,589,569,550 mortality 525, 573,638 (1 h)* 1500' Bucaille and Kim, 1979 settlement
fertilized egg (usually 2-cellstage) 2d
48h
embryogenesis
10 d, 24T, mix algal food
mortality growth
C. virginica
fertilized egg (usually 2ceU stage) 2d
48h
embryogenesis
14
Davis and Hidu, 1%9a
12 d, 24°C mix algal food
M. mercenaria
fertilized egg 2d
48 h 10 d. ZaT, mix algal food
mortality growth embryogenesis mortality growth
41 5-10 14 1750 25-50
Davis and Hidu, 1%9a
C.virginica
fertilized egg 2d fertilized egg 2d
12 d, 24°C mix algal food
M. mercenaria
M. mercenaria
CN-(NaCN)
Exposure conditions
C. virginica
Mytilus sp.
TCC
Initial agdsize
10 d, 24T, mix algal food
mortality (48 h, 30 min exposure)
mortality
grad,
mortality growth embryogenesis
Cgigas M. edulis M. galloprovincialis fertilized egg D-larva D-larva
48 h, 37-38psu
Chlamys arperrimn 30-min-old embryo
lVC, 32psu, 30d-', 5 pm FSW uv sterilized
embryogenesis* mortality (48 h) growth (48 h)
embryogenesis (48 h)
32
Notes 'back to U1"C after exposure *+a added as chloramine
*no movement, immobile cilia *no movement, immobile d i a
'no movement, immobile cilia *no movement, immobile cilia 'continuous chlorination
Davis and Hidu, 1%9a
37 2.5-5
31 50 72
Davis and Hidu, 1%9a Davis and Hidu, 1%9a
10-50
32-100 32-100
lo.@* w
Okuho and Okubo, 1962 Pavicic and Pihlar, 1982
'abnormal larvae excluded **l5,11.7 with Cd and zinc cyanide ***c.16,10 with Cd and zinc cyanide
Pablo er al., 1997
*U8,686with iron-cyanide complexes
c. ll***
286*
Rosin A m i e D C. virginica
fertilized egg 2d
48 h 12 d, ?AT, mix algal food
embryogenesis mortality
<250 27-55
Woelke, 1972
1250 71
Davis and Hidu, 1%9a
PCP'
C. gigus
f e r t i i i d egg
20T, P25psu, 2&30ml-', FSW uv sterilized
embryogenesis (48 h) (fully shelled*)
Pcp
C. virginica
fertilized egg 2d
48h 12 d, 24°C. mix algal food
embryogenesis mortality growth
C virginica
fertilized egg 2d
48h 12 d, 24°C. mix algal food
embryogenesis mortality growth
4250 <25
PCP'
M. edulis
fertilized egg
48 h, 24 psu
embryogenesis
300-400
CI-produced oxidants
C. Virginia
W
C Virginia
eyed, >14 d fertilized egg
24°C. 12-14psu, no food, 1pn mortality' FSW, flow-through system (48, .% h) Bow-through system mortahty (48 h)
M. lateralis C. virginica
M. arenaria
97-117 pm 308 pn 100% eyed
27-28°C. 15-16psu. diluted sewage, no food*, flow-through system
3648 h
17-20-C 13psu, 1wm FSW, flow-through system
18 d
Davis and Hidu, 1969a
ZpentachJornphenyl acetate
R. Brown, cited by Dimick and Breese, 1965
'Na pentachloropheuate
Roosenhurg et a!., 1980b
*no movement and disturbed
>25
300, 64 -300, -300 27,46** two separate experiments
internal organization *author mentions monochloramine, bromamine. bromide, halomethanes and bromate
Roberts 1980
38 mortality (96 h) <20 Roberts and Casey, 1985 metamorphosis 30-60** (every 24 h until end of spatfall) mortality (no c. 500 (8 h) Roosenburg et al., 1980a movement and c. 300 (16 h) >500 (16 h) disturbed internal W 5 0 0 (24 h) <200 (48 h) organization)
BrCl-produced oxidants
C. virginica
4 h fertilized egg
ZOpsu, flow-through system
mortality (48 h)
100, 210** two separate experiments
CIBr-produced oxidants*
C virginica
fertilized egg
flow-through system
mortality (48 h)
220
03-produced oxidants
C virginica
48 h D-larva 9-21d pediveliger
25°C. 12psu. 1pm FSW. flow-through system
mortality* (48, 96 h)
DDT
C. gigas
fertilized egg
2O"C, XSpsu, 2&30 F'SW uv sterilized
I&',
*even misshapen or undersized 'Na pentachlorophenate
>so
Pcp2
h
Davis and Hidu, 1%9a
<25
embryogenesis (48 h) (h11~ shelled*)
-300, 125
Roberts and Gleeson, 1978
Roberts 1980
Richardson
ef
>300, >300 100-1WO
'mixed diet for mortality exp. **1&30 in other experiments with younger pediveligers
'author mentions bromamine, bromide, halomethanes and bromate a[., 1982
*no vital stain, empty shells lack of movement in D-larvae
Wnelke, 1972
'even misshapen or undersized
EC, or LG, Pesticide
Test sp.
Initial agelsize
Exposure conditions
End-point
(fig 1-7
Reference
C virginica
2d
12 d. 24°C mix algal food
mortality growth
34 >25
Bromate
C. Virginia
fertilized egg
48 b, ZMPC ZSpsu, 15 ml-',
mortality
50-100
Stewart el d..1979
Dinoterbe'
C gigm
fertilized egg
24!%?!,!!8~l-1, FSW
embryogenesis (24 h)
5C-100
His and Beirns
70
Lucu cf d.,1980
78 340
Davis and Hidu. 1%9a
0.2 pm
Slimicide'
M.g d o p r o v i n c i d u fertilized egg
96 h, 13°C37psu
embryogenesis
Omazene
C Virginia
48 h 12 d, 24°C mix algal food
embryogenesis mortality growth embryogenesis mortality
M.merenaria
Co-Ral
Dichlorvos
fertilid egg 2d fertilized egg
Zd
48h 10 d, W C , mix algal food
C virginiua
fertilized egg 2d
48h 12 d, 24°C mix algal food
M. mercenaria
fertilized egg 2d
48h 10 d. 24T, mxi algal food
C gigar
fertilized egg (1-2 h)
48 h, U'C 50 n P , 0.45 pm
Notes
Davis and Hidu, 1%9a
*phenolic herbicide
(unpublished data)
growth
1lx-250 81 378 250-500
embryogenesis mortality growth embryogenesis mortality growth
110 >loo0 1lx-250 9120 5210 500-5000
embryogenesis
165
*bis-trichloromethyl-sulphone (20%), methylene bis-thiocyanate(5%)
Davis and Hidu, 1969a
Davis and Hidu, 1%9a Davis and Hidu, 1%9a
ll~ainel a/., 1990
FSW Dieldrin
C virginiua
fertilized egg 2d
48h 12 d, 24°C mix algal food
embryogenesis mortality growth
640 >10 000 1lx-250
Davis and Hidu. 1%9a
Endrin
C. Virginia
fertilized egg 2d
48h 12 d, 24'C mix algal food
embryogenesis mortality growth
>lo 000
790
Davis and Hidu, 1%9a
Lindane
C. gigm
fertilized egg
20% 325psu, 20-30 ml-', FSW uv sterilized
embryogenesis
5000
C gigm
fertilized egg (30 min)
9d, 24°C Bpsu, 3 O d - l (egg), 8 d-' (larva), mix algal food, 0.2 pm FSW
(48 h)
1Cn-250 Woelke, 1972
*even misshapen or undersized
His and Seaman, 1993
'6 d
(fUW
shelled') mortahty growth
170 100-250*
M. mercstaria
fertilized egg fertilized egg 2d
c5 P
fertilized egg
C virginica
Methoxychlor
48h 48h 10 d. 24°C mix algal food
embryogenesis embryogenesis mortality growth
9100 >loo00 >loo00 >loo00
Davis and Hidu. 1%9a
embryogenesis
Mo (430) 170 (260) 250 (440) 210 (160)
Cardwell er al., 1979a
190
Davis and Hidu, 1969a
(and 3 other spp.)
Roccal
M.mercenaria
fertilized egg 2d
48h 10 d, 24"C, mix algal food
embryogenesis mortality growth
140 >lo0
Dowcide G
M. rnercenarin
fertilized egg 2d
48h 10 d. 24°C. mix algal food
embryogenesis mortality
<250 <250
Davis and Hidu, 1%9a
Toxaphene
M.memenaria
fertilized egg 2d
48h 10 d, 24°C mix algal food
embryogenesis mortality growth
1120
Davis and Hidu. 1%9a
<250
fertilized egg 2d
48h 10 d, 2 4 T mix algal food
embryogenesis mortality
>loo00 410
Aldrin
M. mercenada
<250
growth
<250
Davis and Hidu, 1%9a
Griseofulvin
M.mercenaria
fertilized egg 2d
48h 10 d. 24T, mix algal food
embryogenesis mortality growth
<250 <1wO >lwO
Davis and Hidu, 1%9a
Propen-1-01-3
M.mercenarin
fertilized egg 2d
48h 10 d. 24'C, mix algal food
embryogenesis mortality
t250
1030
Davis and Hidu, 1%9a
C. virginica
fertilized egg 2d
48h 10 d 24°C mix algal food
embryogenesis mortality growth
5900
Davis and Hidu. 1%9a
25C-5500
fertilized egg 2d
48h 12 d, 24°C mix algal food
embryogenesis mortality growth
25C-5500
Silvex
Nemagon
Isoproturon
N-3452
M. mercenada
C gigas
C virginica
fertilized egg (30 min)
9 d, 24°C 28psu. 30 ml-l (egg). 8 tn-' (larva). mix algal food, 0.2 c ~ FSW m
mortality growth
fertilized egg 2d
48h 12 d. 24°C mix algal food
embryogenesis mortality
710 10 Ooo 780 370
Davis and Hidu, 1969a
His and Seaman. 1993
2500
Davis and Hidu. 1969a
'fully shelled regardless of shape (total)
Pesticide
Test sp.
Initial agehize
Exposure conditions
End-point
EC, or LC& ( M1.')
Reference
Notes
Nabam
C virginica M . rnercenana
fertilized egg fertilized egg 2d
48 h 48h 10 d, 24°C mix algal food
embryogenesis embryogenesis mortality growth
Davis and Hidu, 1969a Davis and Hidu. 1969a
TCP
C. virginica
fertilized egg 2d
48h 12 d. 24% mix algal food
embryogenesis mortality growth
Davis and Hidu, 1969a
Parathion'
C gigas
fertilized egg (30 d)
9d, 24% 28psu, 3Oml-' (egg), 8 d - ' (larva), mix algal food, 0.2 I . L ~FSW
mortality
His and Seaman, 1993
'methyl-parathion
C gigas
fertilized egg
20°C. *25psu, 20,30ml-', FSW uv sterilized
Woelke, 1972
*even misshapen or undersized
C virginica M. edulis
2d fertilized egg
12 d, 24% mix algal food 1>17"C
embryogenesis (48 b) (fully shelled) growth embryogenesis
Davis and Hidu, 1%9a Okubo and Okubo, 1962
Fenitrothion
c gigar
fertilized egg (30 min)
9d, 24°C 28psu, 30~1.' (egg). 8ml" (larva). mix algal food, 0.2 pm FSW
growth
His and Seaman. 1993
Guthion
C. virginica M.mercenaria
fertilized egg fertilized egg 2d
48h 48h 10 d, 24T, mix algal food
embryogenesis embryogenesis mortality
Davis and Hidu, 1%9a Davis and Hidu. 1%9a
2;J-D ester
C. virginica
fertilized egg 2d
48 h 12 d, 24°C mix algal food
embryogenesis mortality growth
Davis and Hidu. 1%9a
Dowcide A
M. rnercenana
fertilized egg 2d
48 h 10 d, 24°C mix algal food
embryogenesis mortality growth
Davis and Hidu, 1969a
C. virginica
fertilized egg 2d fertilized egg 2d
48h 12 d, 24°C mix algal food 48h 10 d, 2 f C . mix algal food
embryogenesis mortality embryogenesis mortality
Davis and Hidu, 1%9a
fertilized egg 2d
48 h 12 d, 24% mix algal food
mortality
Davis and Hidu, 1969a
fertilized egg
48 h, 2&29"C, ZSpsu, 15In-', 0.45 m FSW
Parathion
N-3514
M. mercenaria Dipterex
C. virginica
Uranin (stain)
M.edulis
Bromoform
C virginica
Chloroform
growth
Davis and Hidu, 1%9a
growth embryogenesis
Okubo and Okubo, 1962
mortality
Stewart et af., 1979
Di-syston
C. virginica
fertilized egg 2d
48 h 12 d, 24°C. mix algal food
M. mercenaria
fertilized egg 2d
48 h 10 d, 24°C mix algal food
AtrazineSimazine'
c. grga
Mecoprop
C gigas
Sevin
24 h, 2 4 T 28psu. 81n-' (larva), mix algal food (larva), 0.2 pm FSW
embryogenesis* mortality (10 d) growth** (5 d)
fertilized egg (30 min)
9 d, 24°C 28psu, 30 d-l (egg), 8ml-' (larva), mix algal food, 0.2 pm FSW
mortality growth
48 h, 20°C. 25psu. 5 0 n t ' emhryogenesis (Dimick and Breese, 1965) 2 0 PZPSU. ~ 2 6 3 0 ml-1, embryogenesis FSW uv sterilized (48 h) (fully shelled') 48 h embryogenesis 12 d, 24% mix algal food mortality prowth 48 h, 20°C 25psu. 5 0 d P embryogenesis 1 h exposure + 48 h recovery, embryogenesis 17°C. 25psu, 1 0 d - ' , (abnormal sand-FSW, no food larvae excluded)
c gigm
egg
C. gigas
fertilized egg
C. virginica
fertilized egg 2d
M. edulu
M. edulis
M. mercenario
embryogenesis mortality growth embryogenesis mortality growth
5860 3670
Davis and Hidu, 1969a
1W2500
5280 1390 1W2500 5W10000
Davis and Hidu, 1969a
Robert ef al., 1986
<2500 1CQO-2500 4200 1W2500 2200 (803:) lEO&3700 wx)
'25% Atrazine-25% Simazine *abnormal larvae excluded **height
His and Seaman,1993
Stewart ef 01.. 1967 Woelke, 1972
*values for I-naphthol, Sevin hydrolytic product *even misshapen or undersized
Davis and Hidu, 1%9a
3000
>m
egg egg 1st polar body (25 min) 2ceU (1h) 64cell or more (4 h) blastula (8 h) trocophore (20 h) veliger (32 h) fertilized egg 2d
48h 10 d, 24°C. mix algal food
embryogenesis mortality growth
3820 2500 1W2500
2300 (1300:) Stewart ef al., 1967 20 700 (24 5 0 0 3 Armstrong and 5300 (5200') Millemann, 1974
7000 8300 16 000 19 000
2Aooo Davis and Hidu, 1969a
Fenvalerate
C. virginica
not given
48 h, not given
emhryogenesis
>loo0
P. W. Borthwick, cited by
Neburon
M. mercenan'a
fertilized egg 2d
48 h 10 d, 24°C. mix algal food
embryogenesis mortality
<2400 <2m
Davis and Hidu, 1%9a
Diuron
M. mercenaria
fertilized egg 2d
48h 10 d, 24°C. mix algal food
emhryogenesis mortality growth
2530
Davis and Hidu, 1969a
>5000 >So00
Tagatz and hey. 1981
*values for I-naphthol, Sevin hydrolytic product
Exposure conditions
End-point
C. virginica
fertilized egg 2d
48h 12 d. 24"C, mix algal food
embryogenesis mortality growth
Davis and Hidu. 1%9a
C. virginica M. mercenaria
fertilized egg fertilized egg 2d
48 h 48h 10 d, 24"C, mix algal food
embryogenesis embryogenesis mortality growth
Davis and Hidu, 1%9a Davis and Hidu, 1%9a
embryogenesis
Okubo and Okubo, 1%2
Test sp.
Pesticide Malathion
Rhodamine B (stain)
M. edulis
T m i c acid
M. edulis
Initial age/size
20°C ==25psu,20-30 ml-', FSW uv sterilized
C gigas KKIO,
c
embryogenesis embryogenesis (48 h*)
3200-10000 210 000
Okuho and Okubo, 1962 Cardwell er 01.. 1979a
embryogenesis
3200-10 000 3200-10 000
Okubo and Okubo, 1962
M. edulis
Dicapthon
Phosphamidon Monuron
M. mercenarin
C. gigas M. mercenaria
Reference
fertilized egg 2d
48 h
10 d, 24°C mix algal food
embryogenesis mortality growth
3340 5740 >2000
fertilized egg
20% ~ 2 5 p s u 20-30 . ml-', FSW uv sterilized
embryogenesis
4lmo-mm
48h 10 d, 24'C. mix algal food
embryogenesis mortality
fertilized egg 2d
Notes
*fully shelled, even misshapen or undersized
Davis and Hidu, 1%
Woelke, 1972
(48h*)
*fully shelled, even misshapen or undersized
Davis and Hidu. 1%9a
growth
Fenuron
M. mercenaria
fertilized egg 2d
48h 10 d, 24°C mix algal food
embryogenesis mortality
Davis and Hidu, 1%9a
Carbofuran
C. gigas
fertilized egg (30 min)
9d. 24% Bpsu, 3 O d P (egg), 6 d - l (larva), mix algal food, 0.2 p m FSW
mortality growth
His and Seaman, 1993
Bromoxynil
c gigas
fertilized egg (30 min)
9 d, 24°C 28psu, 30 I&' (egg), E d - ' (larva), m h algal food, 0.2 pm FSW
mortality growth
7000 5000-10000
His and Seaman, 1993
Metaldehide
C. gigas
fertilized egg (30 min)
9 d, 24% 28psu, 30 m1-l (egg), 8 d P (larva), mix algal food, 0.2 pm FSW
mortality growth
7400 5000-10 000*
His and Seaman. 1993
Carbetamide
c. gigas
fertilized egg (30 min)
9 d, 24°C 28psu. 30 ml" (larva), mix (egg), 6 d-' algal food, 0.2 FSW
mortality growth
93w
His and Seaman, 1993
210000
Bold: 95% CI provided. FSW, filtered seawater.
*6 d
*6 d
107
THE ASSESSMENT OF MARINE POLLUTION
Table 12 Interspecific comparison of sensitivity of bivalve embryos to Sevin insecticide. Median effective concentrations (EC,,) inhibiting normal embryogenesisare given.
Species
EGO(mg 1-7 ~~
Crassostrea gigas C. virginica Mytilus edulis
Mercenaria mercenaria
2.2 1.8-3.7 2.04.0
1.6-3.2 2.3 2.5-4.0
Reference ~
~~
Stewart et al., 1967 Woelke, 1972 Davis and Hidu, 1969b Dimick and Breese, 1965 Stewart et al., 1967 Davis and Hidu, 1969b
show similar sensitivity to biocides (e.g. Sevin, Table 12); this corresponds to the similarity in their sensitivity to metals. The effects of biocides on larval settlement are poorly documented. DDT did not greatly affect settlement in C. gigas (Loosanoff, 1954), but DDT and other insecticides did affect settlement in Ostrea edulis (Waugh et al., 1952). Cardwell et al. (1979b), working with C. gigas embryos at 20°C, reported a 48-hour EC50ammonia of 15 mg NH3-N1-’. Gormly et al. (1996) assayed the toxicity of microbial pest-control agents (Bacillus thuringiensis, B. alvei, Metharyzium anisopliae) and of a virus of the gipsy moth to D-larvae of Mulinia lateralis, finding strong effects of B. thuringiensis upon this non-target species. 4.1.3. Detergents and oil Data on the toxicity of detergents and oil to bivalve embryos and larvae are given in Table 13. Detergents are usually a mixture containing at least one surfactant and one builder. The surfactant is the active component that reduces the surface tension of the water and dissolves the organic molecules of a stain. The first synthetic detergents included non-linear surfactants, which were poorly degradable. Owing to environmental concerns, these so-called “hard” detergents have largely been replaced by “soft” detergents, which use readily degradable compounds with a linear alkyl group, such as the linear alkylate sulphonates (LAS) as surfactants. The builder of a detergent contributes to the cleaning action by sequestering cations that would otherwise interfere in the action of the surfactant. Polyphosphates used to be the typical cation-sequestering agents in detergents, but concerns about their role in water eutrophication have prompted their replacement. Nitrilotriacetate (NTA) was one of the alternative builders, but its innocuity to mammals has been under discussion (Stoker and Seager, 1976), and zeolites (sodium aluminium
Table 13 Toxicity of detergents, oil and detergent-oil mixtures to early developmental stages of bivalves. LC5,,: toxicant concentration causing 50% mortality; otherwise ECS0:toxicant concentration causing 50% reduction in the end-point. Several detergents were assayed as mixtures: in these cases, only the active components are mentioned and their percentages in the assayed mixture provided. Concentrations are expressed as p,g I-’ of active cornponenth except when otherwise stated. Detergent!Oil
Test species
Cationic surfactants para diisobutyl C.virginicn phenoxyethoxy ethyl dimethyl benzyl ammonium chloride monohydrate 98.8%
lauryl pyridinium chloride
Initial agelsize
Exposure wnditions
EC& or L C ,
Reference
490
Hidu, 1965
abnormal larvae excluded
Hidu, 1965
abnormal larvae excluded
fertilized egg
24°C 2C-30 I&’, 15 pm FSW uv sterilized
embryogenesis
2d
24°C 1C-15 d-l,mix algal food, 15 pm FSW uv sterilized
mortality (12 d) growth (12 d) (length) embryogenesis mortality (10 d) growtb (10 d) (length) embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortalilty (10 d) growth (10 d) (length)
8.5 100-250 -250
M.mercenarin
idem
idem
C virginiur
idem
idem
M. rnercenaria
End-point
>So0 100-250 1270
>So0
100-250
90
100-250
50-100
idem
idem
Id
not given
mortality (6 h) growth (1week) settlement (6 h expo.)
1ooO* 50” 1ooO***
26°C 13 UP, FSW uv sterilized
embryogenesis (48 h)
25-250
Linear anionic
Notes
SUrfaclanlS
h e a r akylate sulphonate (LAS) dodecylbenzene (12 C) (anionic biodegradable surfactant)
0.edulis
Linear alkylate sulphonate’ 60.8% (LAS) (anionic biodegradable surfactant)
C virginica
8-10 d
fertilized e g
‘referred to as “lethal concentration”; “slightly superior” for I-d-old C. gigar larva “conc. mentioned as “seriously affecting” growth for 0. edulis and C gigas ***cOnc mentioned as “significantly reducing” settling and metamorphosis Calabrese and Davis, *described as “compo&e of a 1%7 number of wmmercially available product$ typical of the LAS presently (1%5) being marketed” Renzoni, 1973b
48h linear alkylate sulphonate ( + A S ) (12 C) (anionic hiodeeradable surfactant)
M. edufir
LAS (12 C)
M. edulis
egg
324 pm larvae
26°C 1&12 ml-l, mix algal food, FSW uv sterilized ZOT, 10 I&', food not given, paper FSW. uv sterilized
22% 20%, 1 1 0 ml-', monoalgal diet, 20 Fm FSW
mortality (10 d) 500-1000 growth (10 d) (length) 250-500 fertilization (2&24 h) embryogenesis (% h) mortalitv (10 d) growth (10 d) iength mortality* (24 h) (48 h) (72 h)
(96h)
C. gigas LAS 10 C 11 c 12 c 13 C 14 C short chain blend medium chain blend long chain blend LAS degradation products: DTW lienosulvhonic acid orher pioducts sodium dodecyl sulphate 85% C. gigas
Granm6, 1972
>m >4M) 7500 4500
34cQ
Hansen er of., 1997 *no ciliary movement **length increase and C content abnormal swimming at 35 ppb
38M) 1400 8206500
growth** (9 d) settling (7 d) <80 embryogenesis (48 h) 1700 (2025)
fertilized egg
580 (800) 260 (280)
Cardwell er af.. 1979a
fully shelled (total)
;: g? 460 (470) 250 (7.90) 110 (120)
*diethy1 tetralin indane sulphonate
1270 (2030) 4333 (>10000) fertilized egg
(SDSI* SDS
-300 -50
20°C 29% 2&30
I&',
unfiltered SW C gigas Protorhacu staminea Tresus capax
>iom
embryogenesis (48 h) 840*sd0.09 (normalshell development) 910*sd0.14 mortality (48 h) embryogenesis (48 h) 930 (1ooO) 450 (870)
Cardwell er af.,
197%
Cardwell er ul., 1979a
*14% 10 C, 72% 12 C,14% 14 C *means of 20 batches
fully shelled (total)
Trous nurralfi SDS
M. larerutis
2 h fertilized egg
21°C 75 ml-'
M. lateralis
embryogenesis (48 h) 8200 (10%) 5800 (30%)
SDS
48 h D-larva
22-25°C %28%, 5-8ml-', 0.22 pm FSW uv sterilized
mortality* (48 h)
SDS
Chlamys ospem'ma fertilized egg ('1 h) Chfamys arpewima fertllized egg
SDS
(30 min)
18°C 33%, 30 ml-', 1pm embryogenesis (48 h) FSW uv sterilized 18°C 33%, 3Oml-', 5 pm embryogenesis (48 h) FSW uv sterilized
6300
845
1wO
Momson and Petrocelli, 1990 G o d y er al., 19% *loss of respiratory circulation and ciliary motion, disfiguration and eventual decomposition Krassoi er af., 1996 abn. lar. exc.; LOEC 6-56, NOEC 500 Krassoi er of., 1997 abn. lar. exc.: LOEC 950, NOEC 530
DetergenUOil
Test species
Non-linear anionic surfactanrs alkyl naphthalene sulphonate’ C. virginica
M. memotarin
Initial age/size
Exposure conditions
fertilized egg
24‘C 2&30ml-’, 15 pm FSW uv sterilized
2d
24°C 10-15 de1. mix algal food, 15 pm FSW uv sterilized
idem
idem
End-point
embryogenesis (48h) (abn. lar. exc.)
mortality (12 d) growth (12 d) (length) embryogenesis (48 h) (abn. l a . exc.)
ECm or LC, (I% 1-7 1630
Reference
Notes
Hidu, 1965
abnormal larvae excluded; *(ANS+polyphosphate+fatty alcohol sulphates; percentages unknown; w n c expressed as gross product)
Hidu, 1%5
abnormal larvae excluded
Hidu, 1965
abnormal larvae excluded
Hidu, 1965
abnormal larvae excluded
>2500
looo-2500 5830
mortality (10 d) 210 000 growth (10 d) (length) 5wO-10000 alkyl benzene sulphonate 54.8%
C virginica
M.mercenaria triethanolamine dodeeyl benzene sulphonate 609.
alkyl sulphate 26.&28.9%, free fatty alwhol 1.&2.5%
tetrapropylene benzene sulphonate
idem idem
idem idem
C virginica
idem
idem
M.mercenarin
idem
idem
C virginica
idem
idem
M. mercenaria
idem
idem
Id
not given
0. edulir
b10 d
embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortality (1Od) growth (10 d) (length) embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortality (10 d) growth (10 d) (length) embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortality (10 d) growth (10 d) (length) mortality (6 h) growth (1 week) settlement (6 h expo.)
270 500-1000
25&500
940 >2500 1ooO-2500 390 1wO-2500 -1000 1030
25W5000 -2500 370 2500-5000 2500-5000 470 5000-10000 5000-10000 2000‘ 50.’ l000***
Renzoni. 1973b
‘referred to as “lethal wncentration”; “slightly superior” for I-d-old C.gigas larva **cone mentioned as “seriously affecting” growth for 0.edulis and C. gigas ***wnc mentioned as “sigm5cantly reducing” settling and metamorphosis
Non-ionic su@ctrrnts iso-octyl phenoxy polyethoxy C virginica ethanol
fertilized egg 2d
24°C. 2C-30 d - l , 15 pm FSW uv sterilized 24"C, 1C-15 tn-', mix algal food, 15 pm FSW uv sterilized
M. mercenaria alkyl polyether alcohol tallow alcohol decaethyleneglycolether (non-ionic biodegradable surfactant) Builden sodium nitrilotriacetate W.4)
Oil dispersants Chevron
E-314 Hollchem Aquaclean Houghtosol Polyclens Janslov Spillex Seasweep Polywmplex W1439 Emsol Slik Gamlen Dasic
idem
idem
idem idem egg
idem idem 72 h, 18°C. 33%.
C virginicrr
M. mercenaria M. edulis
M. galloprovincialis fertilized egg
20% 35%. 40 d-'
860
embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortality (10 d) growth (10 d) (length) embryogenesis mortality (12 d) growth (12 d) (length) embryogenesis mortality (10 d) growth (10 d) (length) fertilization (22 h) embryogenesis (46 h) embryogenesis (72 h)
770 1000-2500 2500-5000 1600 1000-2500 1000-2500 1750 2500-5000 2500-5000 100-500 100-500 >loo0
embryogenesis (40 h)
>20 ooO*
(30 min)
C gigas
C. gig?
0.cdulrr C gigas 0. edulis
0. edulrr
c, gigm
0.edulrr
C gigas
fertilized egg
developing egg 1 week developing egg 1 week
1 week developing egg 1 week developing egg
20T, a5%, 2&3l d-' embryogenesis' FSW uv sterilized (4 h)
23°C 23°C
23°C
embryogenesis (24 h) growth (2 d) embryogenesis (24 h) growth (2 d)
growth (2 d) embryogenesis (24 h) erowth (2 d) kmbryogenesis (24 h)
Hidu, 1965
abnormal larvae excluded
1000-2500 1000-2500
Hidu, 1965 abnormal larvae excluded
Granmo and J@rgensen, 1975
Brunetti et 01.. 1989 'increasing embryogenesis success with increasing NTA mnc.
0.1-1
Woelke. 1972
1-10 1&100 la1000 400 c. 1 m <500
Woelke, 1972 Woelke, 1972 Woelke, 1972 Walne'
'cited by Smith, 1968
1000-5000 -8M m15w 800-1500 6c&1500 800-1500
800-1500 c. 1000 c. 1000 c. 1000 c. 1000
*embryogenesis is considered successful if larvae are fully shelled, even misshapen or undersized
Walne'
'cited by Smith, 1968
Woelke, 1972 Woelke, 1972 Woelke, 1972 Woelke, 1972 Woelke, 1972 Woelke, 1972 Walne' Walne'
'cited by Smith, 1968 'cited by Smith, 1968
Walne'
'cited by Smith, 1968
DetergentlOil Slipclean Clock 0639 C-1575-A BP 1002 F O -300-B Kudos Teepol Farrells Slickaway
Test species
Initial agehue
C gigas
developing egg
c. gigas 0.edulis
developing egg fertilized egg lama
0. edulis 0. edulis
1 week 1 week
Exposure conditions
End-point embryogenesis (24 h)
c. 1000 1m1ooOo 10mlOoOo
23°C
embryogenesis (24 h) embryogenesis (48 h) growth (2 d)
10W-3000 <3ooO
23°C 23'C
growth ( 2 d) growth (2 d)
Reference Walne' Woeke, 1972 Woelke, 1972 Walne' Smith, 19702
Notes
'cited by Smith, 1968 'cited by Smith, 1968 'cited by Woelke, 1972
2.500-7500 4ow8cm
Woelke, 1972 Walne' 5 ~ 1 0 0 0 0 Walne' c. 5000
'cited by Smith, 1968 'cited by Smith, 1968
EQLnJ-8ooO
100m20000 2oow4oooO 40fMIN0000 -1CQooO
Actusol
Corexit Sorbent-C Oil
Mercenaria sp.
refined oil (No.1 fuel oil)
C gigas fertilized egg (1 h), 22-25°C. 30%. C anguhta 6 h exposure FSW M. galloprovinciulir egg (-1 h). 1 h exp. sperm (-3 h), 4 b exp.
Mercenoria sp.
2nd cleavage (45 min) 2d
25T, 27%, 30 d-', 0.45 pm FSW
used motor oil
refined oil (No. 2 fuel oil, Venezuelan Bunker C)
E M G o (Bg I-')
2nd cleavage (45 min) 2d
25°C 27%. 30 d-', 0.45 pm FSW
40
embryogenesis (48 h) (vital staining) mortality (2 d) (vital staining) embryogenesis (6 h)
>loOo*
embryogenesis (6 b) embryogenesis (6 h)
0.1-1 ppt**
100'
>lW**
embryogenesis (48 h) 1000 (Ven) 430 (No. 2 ) (vital staining) mortality (10 d) (vital staining)
1600 (Ven) 530 (No.2)
loo0 (Ven) 570 (No. 2) 440 (450) (No. 6) 450 (410) (No. 2) mortality (4 d) (larval 1700 (d 1) density in the water 1700 (d 2) column) 200 (d 3) 1wO (d 4)
Woelke, Woelke, Woelke, Woelke,
1972 1972 1972 1972
Byme and Calder, WSF 12 h gyratory shaking (200 rpm)t 24 h equilibrium+glass 1911 wool filtration '100% mortality at d 6 Renzoni, 1973a *also with a 4 1 mixture of oil and dispersant (Corexit 8666). **oiI+dispersant, oysters only Conc. mentioned are those used to make up experimental solutions (30min agitation); actual oil wnc. accommodated in SW not measured Byme and Calder. WSF: 12 h gyratory shaking (2CQrpm)+24h 1977 equilibrium+glass wool filtration
length growth (10 d)
relined oil (No. 2 and No. 6 fuel oil)
C. gigas
fertilized egg
relined oil (No. 2 fuel oil)
C virginicu
-120 pm
embryogenesis (4s h)
22.5% 21%. mix algal food. ASW
CardweU era[., 1979a
water-accommodated fraction
Sigler and Leibovitz, 1982
WSF k100 oil:water 20 h magnetic stirring, resulting 8.65 ppm fuel oil
refined oil (No. 2 fuel oil)
refined oil (No. 2 fuel oil, No. 6 fuel oil) refined oil (diesel oil) nude oil (Venezuelan)
M. lateralis
M. edulis
48 h D-larva
22-25"C, 2&28%, 5 4 I&'. 0.22 pm FSW uv sterilized
crude oil (Algeria, Libya, Iraq, Kuwait, Indonesia)
Gormly el a/..
1996
fertilized egg
21.5'C. 3 0 8 , 30 d-', embryogenesis (48 h)
Pelletier e t a / . , 1997
150 pm
15°C3 2 8 , loml-',
Strfimgren and Nielsen, 1991 Woelke. 1972 Renzoni. 1973a
monoalgal food fertilized egg (1 h), 6 h exposure egg (-1 h), 1 h exposure M. galloprovincdis sperm (-3 h), 4 h exposure
22-25"c, 30%. FSW
C. virginica
fertilized egg
FSW uv sterilized
M. laterulis
M.edulis
C. gigas
length growth (10 d) mortality (10 d) embryogenesis (6 h)
25-30 30-35 25-50 >1ooO* ppt
embryogenesis (6 h)
>lWO**
embryogenesis (6 h)
0.1-1 ppt**
fertilization embryogenesis (48 h) mortality (48 h)
>loo0 PlooO -loo0
fertilized egg
fertilization embryogenesis (48 h) mortality (48h) growth (length) (14 d)
>looO* >loo0 >loo0 -loo0
blastuldtrowphore 1 h, 18T, 3520 d-'. 3% 0.2 pm FSW (20 h) 100-110 pm (5 d)
embryogenesis (48 h) (abn. l a . exc.) mortality (5 d)
2-4 ppt**
growth (length) (5 d)
2-4 ppt***
C anguluta
crude oil (Alaska, Nigeria, Kuwait)
mortality* (48 h)
2oo0*
Renzoni. 1975
Lucas and Le Roux,1975
WSF *lossof respiratory circulation and ciliary motion. disfiguration and eventual decomposition water-accommodated fraction *under uv light; **under fluorescent light supplied in 1-10 pm acacidgelatine microcapsules 'WSF *also with a 41 mixture of oil and dispersant (Corexit 8666). **oil+dispersant, oysters only. Conc. given are those used to make up experimental solutions (30 min agitation); actual oil conc. accommodated in SW not measured Conc. given are those used to make up experimental solutious (30 min agttation); actual oil conc. accommodated in SW not measured 'similar results when 1 h exposed eggs were inseminated with untreated sperm in clean SW or 1 h exposed sperm inseminated untreated eggs in clean SW *cox. yielding 10-70% of control embryogenesis success depending on crude type. **Lib>4: Irq<2. ***data for Lib only. Conc. mentioned are those used to make up experimental solutions (10 min agitation); actual oil wnc. accomodated in SW not measured
DetergenVOil
Test species
Initial age/size
Exposure wnditions
End-mint
ECJLC,
Reference
Notes
-
crude oil Mercenario sp. (Kuwait, Louisiana. Florida)
2nd cleavage (45 min) 2d
I&',
embryogenesis (48 h) 12000 (Kwt) Bryae and Calder, WSF 12 h gyratory shaking (200 5700 (Lou) 1977 rpm)+24 b equilibrium+glass (vital staining) 230 (no) wool Ptration mortality (10 d) 2000 (KG) (vital staining) 2100 (Lou)
WC, 27%, 30 0.45 .um FSW
50 (no)
growth (10 d) (length)
crude oil (Kuwait) (Alaska) crude oil (Arabian light, Prudboe Bay medium) PAH Anthracene Fluoranthene Pyrene
C. gigas
fertilized egg
M. lafernlis
fertilized egg
M. lnterulb
fertilized egg
42W (Kwt) 11M) (Lou) 220 (no) embryogenesis (48 h) 260 (380) 5200 (5800) 21.5"C. 3%, 30 d-', embryogenesis (48 h) 40.1 g * '
FSW
21.5"C. 3%. 30 ml-l, FSW
Cardwell er al., 1979a Pelletier ef al., 1997
embryogenesis (48 h) 6' (4260**) Pelletier ef al., 1* (69") 1997 0.2' (>11 m a * )
Miscellaneous Sudbury's bilge cleaner (composition not given)
1900,2400* 1600,2800* 200,1700* 1800,1700* 1800.1800' 1800,1800' 11400,M)o* 8700, 1600' n.m., 47700.
(dl) (d2) (d3) (d4) (dl) (d 2) (d 3) (d 4) (d 1) nm., 10600* (d 2) n.m., 4100. (d 3) nm.. 4600' (d 4)
Amway LOC bilge cleaner (n-akyl diethanolamide, 10% +primary alcohol akoxylate, 10%) Boatlife bilge cleaner (composition not given)
~~~
~
~~
~
~
~
fully shelled (total) water-acmnunodated fraction *under uv light acetone used as carrier *under uv light; **under fluorescent light Sigler and Leibovitz, 1982 'mixed with N' 2 fuel-oil 1:0.2 vol 'mixed with No 2 fuel-oil 1:0.2 VOl
'mixed with No 2 fuel-oil 10.2 vol
~~~~~~
Values in bold when ae. for the median effective concentration is provided. FSW fdtered seawater; ASW artificial seawater; WSF "water-soluble fraction" (sic); mnc.: concentration; nm.: no mortality.
THE ASSESSMENT OF MARINE POLLUTION
115
silicates) are now the predominant builders used by detergent producers (Henries-Morgan and de Oude, 1994). Products assayed with early developmental stages of bivalves include cationic, anionic and non-ionic surfactants, NTA, oil dispersants, crude oil, refined oils (mainly fuel oils), individual polyaromatic hydrocarbons and oil-detergent mixtures. Hidu (1965) was the first to study the toxicity of surfactants on early developmental stages of bivalves. His work covers “hard” detergents based on cationic, anionic and non-ionic surfactants. Subsequent work by other authors focused on the biodegradable linear surfactants (LAS) (mainly dodecylbenzene sulphonate) and sodium dodecyl sulphate (SDS). In agreement with the results of toxicity studies with heavy metals and biocides, embryos are frequently affected more than larvae, and among larvae growth is a more sensitive indicator of toxicity than mortality. The cationic surfactant lauryl pyridinium chloride is the most toxic compound in this group, with an embryogenesis EC50 of 8.5 to 90 pg 1-’. The new linear anionic surfactants are generally more toxic than the old non-linear ones. The biodegradable surfactant LAS has an embryo EC50 ranging from 50 to 1700 (mean ? sd: 383 2 493, n = lo), but according to Cardwell et al. (1979a) its degradation products are mostly non-toxic (EC,, from 1270 to >10000). SDS, sometimes used as reference toxicant, has consistent effects on embryogenesis, with the exception of the study by Morrison and Petrocelli (1990), who report EC50 values an order of magnitude higher than the other studies. When these high values are excluded, the average embryo EC50 of SDS is 7412233 (meanksd, n = 7). The toxicity of anionic surfactants is directly related to the length of the alkyl chain. The 14-C compounds are some 30 times more toxic than 10-C compounds (Cardwell et al., 1979a; see also Abel, 1996). Embryo-larval toxicity tests have frequently been performed with commercial products which are a mixture of molecules with different and sometimes unspecified chain lengths (e.g. Calabrese and Davis, 1967). This, along with the use of other additives, explains part of the variability in the results reviewed in Table 13. Non-linear anionic, and non-ionic surfactants are less toxic than the other surfactants, with embryogenesis EC5, varying between 300 and 8200 pg 1-’ (mean 2 sd: 1229 2 1416, n = 14). The only builder assayed to date, NTA, was found to be non-toxic (Bruvetti er al., 1989). In the comprehensive work of Woelke (1972), oil dispersants showed a very wide range of toxicities, in many cases higher than that of the oil itself. Toxicity studies with petroleum and its derivatives suffer from important handicaps, such as their mixed and variable chemical nature and the
116
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
hydrophobicity of their main components. Most authors chose to mix up oil and water, testing the so-called “water-soluble’’ (sic), or wateraccommodated fraction, but sometimes the data refer to the total concentrations of the oil added (Renzoni, 1973a, 1975; Lucas and LeRoux, 1975). Some authors have assayed mixtures of oil and detergents, an approach that may be more relevant from an environmental standpoint. Finally, Stramgren and Nielsen (1991) used acacidgelatine microcapsules containing the test product for their assays. Despite the methodological heterogeneity, the data in Table 13 permit some general conclusions. Refined oil is generally more toxic than crude oil. However, both are rarely toxic in environmentally realistic situations, i.e. at concentrations below 0.2 ppt. In contrast, the toxicity of used refined oil is higher (embryo ECS0of 40 pg l-’, Byrne and Calder, 1977), probably owing to the presence of highly toxic pyrogenic compounds such as polyaromatic hydrocarbons (PAH). This emphasizes the importance of adequate disposal of the residues from motor vehicle oil. Pelletier et al. (1997) have recently demonstrated the high toxicity of PAH to early developmental stages of marine organisms, placing emphasis on the increase in toxicity caused by ultraviolet radiation, owing to the formation of highly oxidizing intermediates. Pyrene under ultraviolet light had an EC5,, as low as 0.23 pg 1-l in Mulinia lateralis embryos. The effect of detergents and oil in inhibiting larval settlement is greater in Crassostrea virginica ‘%an in barnacles (Smith and Hackney, 1989). Renzoni (1973b) found in 0. edulis that larval settlement and larval mortality were affected by oil at similar concentrations (1-2 ppt). 4.1.4. General conclusions Despite the variability among the experiments reviewed so far, they generally agree with respect to various points. The tentative ranking of the toxicity of the main marine micropollutants to bivalve larvae is: tributyl tin > mercury > silver > copper > zinc > nickel > lead > cadmium > organochlorines > organophosphates > other pesticides = detergents > used oil > refined oil > crude oil. A more complete (and complex) picture can be obtained by inspection of Tables 9-11 and 13. TBT (see Table 11) is the most toxic compound reported for bivalve larvae, with ECS0values below the ppb level. With respect to heavy metals (see Tables 9 and lo), bivalves appear to be particularly resistant to cadmium, compared to other species such as decapod crustaceans (e.g. Marijio-Balsa, 1998). The Gironde estuary (southwest France) with its high cadmium levels and reproductively active Crassostrea populations provides natural evidence. Embryos and larvae are only affected by cadmium concentrations of the order of 1 and 0.1 mg l-l, respectively. In
THE ASSESSMENT OF MARINE POLLUTION
117
addition to the data in Tables 9 and 10, Zaroogian and Morrison (1981) did not find effects on larval growth (3 weeks) at 15 pg 1-1 of cadmium. As might be expected, bivalves and their larvae are not particularly sensitive to insecticides, as compared to crustaceans (arthropods) and larvae of other species with more complex nervous systems. Detergents, oils and their mixtures are usually toxic to early developmental stages of bivalves at concentrations in the order of l p p m or higher, with some significant exceptions (see Table 13).
4.2. Intrinsic (biological) factors affecting toxicity
4.2.1. Species
In the literature, studies comparing different toxicants are much more frequent than those comparing different bivalve species, but when studies use the same standardized methods, then the general conclusion is that various bivalve species display similar sensitivity to pollutants. For example, the median effective concentrations of the insecticide Sevin inhibiting normal embryogenesis in oysters, mussels and clams fall within a relatively narrow range of 2 to 4mgl-', approximately (see Table 12). The sensitivity of C. gigas oyster embryos to sulphite pulp-mill effluent is quite similar to that of four clam species (Cardwell et al., 1977a; see also Woelke, 1972, pp. 9-13, Table 8). We have also previously shown that, given uniform criteria to record the response, mercury embryotoxicity took place at similar concentrations in different species (see Beiras and His, 1994, 1995a, and citations therein). Therefore, interspecific variability in sensitivity to pollutants seems limited, allowing broad comparison between the results of embryo-larval bioassays performed with different bivalves. The moderate coefficients of variation for the average EC5,, of the most toxic metals (see Section 4.1.1) further support this conclusion. One notable exception to this rule is hydrogen sulphide, one of the oldest molecules in the biosphere, against which many species have evolved highly effective detoxification mechanisms (e.g. Oeschger and Vetter, 1992). Bittkau (1993, 1996) found significant species-specific differences in H2S tolerance in bivalve larvae: those of Mucoma had the highest, those of Mytilus edulis and M . galloprovincialis had intermediate, and those of Crassostrea gigas had the lowest sulphide resistance; moreover, she found that in Baltic clams (Macoma balthica), adults living at sulphidic locations produced larvae with significantly higher tolerance than clams from well-oxygenated sites.
118
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
4.2.2. Age and developmental stage As a general trend, a higher sensitivity to toxicants is found at earlier life stages. The scale of relative sensitivity in bivalves is therefore embryo > veliger > pediveliger > juvenile > adult (Woelke, 1972; Glickstein, 1978; Roosenburg et al., 1980a; Robert and His, 1981; Hansen et al., 1997; Richardson et al., 1982; Ringwood, 1990, 1993; Bittkau, 1993). Embryos are also known to be less tolerant than larvae to extreme environmental conditions in general, such as temperature and salinity (e.g. Davis and Calabrese, 1964). Kennedy et al. (1974) found that the sensitivity to temperature decreases in the order cleavage > trochophore > D-larva. This inverse relation of age to sensitivity apparently does not apply to the eggs, which are more resistant to pollutants than embryos (e.g. His and Robert, 1980). Sperm is even less sensitive than unfertilized eggs (His and Robert, 1980). Results from Pavicic et al. (1994b), who exposed mussel embryos to zinc and cadmium at 8-hour intervals from 0 to 48 hours of age, confirm the general trend. However, maximum sensitivity was found at the ages of 16 hours and 24 hours, corresponding to the tissue differentiation that precedes the formation of the trochophore. Trochophores were also consistently more sensitive than veligers to chlorine in the study by Chien and Chou (1989), but fertilized eggs (first polar body stage) were the most resistant stage. The valuable and detailed studies by Zhadan et al. (1992) concluded that the gastrula is the most sensitive stage to heavy metals in bivalve ontogenesis, since developmental abnormalities which will be conspicuous in the veliger appear at that stage, and not earlier. They identify the abnormalities as affecting the organs developing from the vegetative pole (shell gland, front abductor muscle), while those originating from the animal pole, such as the velum, developed almost normally. Similarly, Ringwood (1992b) found normal D-larva formation to be an end point approximately one order of magnitude more sensitive to copper than cleavage in the bivalve Isognomon californicum, again suggesting that it is organogenesis, rather than cell division, that is the most delicate process in embryonic life. 4.2.3. Body size The general increase in resistance to pollution with age is very closely correlated to the concomitant increase in body size. In our work on the effects of mercury on bivalve larvae (Beiras and His, 1994, 1995a) we recorded the larval growth (organic weight increase), in order to calculate the amount of toxicant theoretically available to the test organisms per unit body weight (i.e. dosage of the toxicant per organic mass of the larva).
119
THE ASSESSMENT OF MARINE POLLUTION
loo0
0)
3.
r----7
100
1
0
0
9 10
I 10
100
1000
10 000
Larval AFDW (ng) Figure If Double logarithmic plot relating mercury concentrations that cause 50% larval mortality after 48 hours (LCs0) to larval ash-free dry weight (AFDW). Crassostrea gigas H; Mytilus galloprovincialis 0.
The results showed that the increase in weight could account for the increase in larval resistance described above. The organic-weight-specific sensitivity increased during larval development, as shown by the slope (1 in the regression log LCs0 versus log organic weight (Figure 11). This finding is consistent with the allometric pattern of increase exhibited by most physiological traits. 4.2.4 Hormesis Low levels of potentially toxic compounds often stimulate growth (or physiological functions) in microorganisms, plants and animals, a phenomenon termed “hormesis” (reviewed by Stebbing, 1982). In marine bivalve larvae, hormesis effects have been described, for instance, during exposure to pesticides (Davis, 1961; Davis and Hidu, 1969a; His and Seaman, 1993) and sediment suspensions (Davis and Hidu, 1969b; Seaman et al., 1991). Stebbing (1982, pp. 230-231) suggested that “growth hormesis may be the consequence of regulatory overcorrections by biosynthetic control mechanisms to low levels of inhibitory challenge, resulting in growth that is greater than normal”. One aspect usually ignored in ecotoxicological testing is that, even
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
120
5 E 90 .-0) (D
c
0
5
10
15
20
25
Age of larvae (days)
Figure 12 Growth of Mytilus galloprovincialis larvae in three concentrations of suspended sediment (days 3 to 12) and during subsequent recovery in non-turbid water (days 12 to 25), showing hormesis at 0.2 g 1-’ and reduced growth at the other concentrations from days 6 to 12, as well as ulterior hormesis at all three concentrations.The differences in size between experimentals and controls were sighcant (0.95 level) at days 12, 16 and 25. (Seaman et al., unpublished.)
though the growth rate of larvae cultivated under sub-optimum conditions may fall behind the growth rate of the control, the experimentals may well recover and overtake the controls after the inhibiting influence is removed. We have found this to occur subsequent to a temporary absence of food (His and Seaman, 1992) and after temporary incubation of larvae in turbid water (Figure 12), and we propose to use the term “ulterior hormesis” for this phenomenon. Toxicological bioassays generally consist of exposing the test organisms to a toxicant for a limited time period; the recovery of the test organism after sub-lethal exposure is almost never considered. But since pollutants are often progressively diluted in the marine environment, and the organisms often suffer only temporary exposure, a full ecotoxicological evaluation would need to take possible ulterior hormesis effects into account. In this context, it would be wrong to regard growth enhancement due to low pollution levels as something “good” (or “bad”), particularly as its physiological basis is not well understood, but it is a phenomenon of ecological relevance.
-
121
THE ASSESSMENT OF MARINE POLLUTION
loo
20
7r-----l
16
20
24
28
loo
0
32
Temperature ("C)
a
10 20 Salinity (ppt)
30
b
Figure 13 mica1 temperature (a) and salinity (b) tolerance curves of bivalve embryos. Embryogenesis success decreases as temperature departs from the optimum. Dependence on salinity, in contrast, is generally asymmetrical, with increasing tolerance at higher salinities throughout the whole range of realistic values. The effects of these environmental variables on resistance to toxicants are expected to follow similar trends. (Data from Davis and Calabrese, 19684, in Crassostrea virginica.)
4.3. Extrinsic (environmental) factors affecting toxicity
Environmental factors may modify the toxic effects of pollutants, either because of their effects on the test organisms, or on the pollutant itself. In the first case, physiological stress caused by sub-optimal conditions is likely to reduce the resistance of the organisms to toxicants, thus decreasing the lethal and sub-lethal response thresholds. Reviewing the effects of temperature and salinity on metal toxicity, McLusky et al. (1 986) concluded that marine animals near their osmotic or thermal limit:; are more susceptible to heavy metals. Therefore, prior knowledge of the optimum environmental conditions for embryogenesis and larval development is required, and bioassays are only valid if these conditions are met. Qpical tolerance curves of bivalve embryos to temperature and salinity are depicted in Figure 13. Tests aimed at determining the toxicity of a pure chemical must be performed within the optimum range of environmental variables for the test species, otherwise the animals may be hypeasensitized and effective concentrations may be underestimated. In the second case, variables such as temperature, salinity and pH may affect the toxicity of a pollutant by altering its physical and/or chemical state. In the case of metals the dissociated ionic form is generally more toxic. Different mercury salts, for instance, increase in toxicity as ionization level decreases (Von Burg and Greenwood, 1991). In hydrogen cyanide, on the other hand, the molecular form is more toxic (Mason,
122
E. HIS, R. BEIRAS AND M.
N. L. SEAMAN
1991). The toxicity of a polar compound may therefore strongly depend on its dissociation constant, and thus be indirectly influenced by environmental factors such as temperature, salinity and pH. With independence of environmental conditions, different salt forms having intrinsically different dissociation constants, and yielding different patterns of speciation when dissolved in seawater, might thus exhibit different toxicity. However, with the exception of aluminium (see Section 4.1.1), the experimental evidence does not show important differences in toxicity between different salt forms. Okubo and Okubo (1962) reported a higher toxicity of copper acetate than copper sulphate in mussel embryos, but not in other species. Glickstein (1978) reports oyster embryo EC5,, values of 5.7 and 5.5 pgl-' for mercury chloride and nitrate, respectively. MacInnes (1981) did not find marked differences in toxicity between chloride and nitrate salts of heavy metals either. When expressed on a copper ion basis, copper chloride and pentahydrated sulphate also show similar toxicity to oyster embryos, with approximate 48-hour values of 13 pg 1-' for the former and 10 pg 1-1 for the latter (His and Robert, 1981, 1982). Oxidation states do affect toxicity; for instance Cr6+ is more toxic to marine organisms than C9' (Deslous-Paoli, 1982). Studying the effects of chemical speciation on the toxicity of pollutants is a complex subject which is beyond the scope of the present review. Finally, dissolved organic matter may act as a chelator, decreasing the proportion of metal made bioavailable (free ions), thus reducing its toxicity (e.g. Knezovich et al., 1981). 4.3.1. Temperature
Toxicity generally increases with temperature, and this holds true for embryo-larval bioassays, as well as for other types of toxicological studies (e.g. Philips, 1980; Taylor, 1981). Higher temperatures enhance the solubility of metal salts, as well as increasing the rate of water and solute movement across cell membranes (MacInnes and Calabrese, 1978). Temperatures above optimum are energetically costly for an ectothermic organism, and the resulting physiological stress may increase the sensitivity of the test organism. Higher toxicity at higher temperatures has been reported by Hrs-Brenko et al. (1977; see also McLusky et al., 1986). Chlorine shows an opposite pattern, probably owing, as Chien and Chou (1989) point out, to lower persistence of chlorine in the water at higher temperatures. MacInnes and Calabrese (1978, 1979) found increased toxicity to embryos not only at supra-optimal but also at sub-optimal temperatures. This pattern of response fits well with the symmetrical curve of tolerance to temperature
THE ASSESSMENT OF MARINE POLLUTION
123
in many biological responses, including embryogenesis (see Figure 13). Resistance to metals is influenced less by temperature and salinity in larvae than in embryos (MacInnes and Calabrese, 1979). 4.3.2. Salinity The sensitivity of marine organisms to toxicants increases as salinity decreases (e.g. Philips, 1980; Taylor, 1981). For early developmental st ages of bivalves, higher toxicity at lower salinities has been reported by Hrs-Brenko et al. (1977), MacInnes and Calabrese (1979), Coglianese (1982) and Chien and Chou (1989). In the case of heavy metals., the proportion of free ion exponentially increases as salinity decreases (e.g. Rainbow el al., 1993). Besides, aqueous solubility of some organometals such as TBT is inversely related to salinity (Inaba et al., 1995).
4.4. Interactions between different toxicants
Given a toxicant, A, present at a concentration [A], we define toricity units (TU) as: TU
=
[A]/LC,o(A)
where LC50(A) is the toxicant’s median lethal concentration. If two or more toxicants act together with the same intensity as they do singly, their interaction is termed additive, and the toxicity of the mixture can be predicted by adding the toxicity units (TU) of each chemical present in the mixture. In other words, for two toxicants A and B with additive interaction, the median lethal concentration of the mixture (LCSO(M))would be: LCsO(M)= iLCso(A)+ iLCSo(B) where LCsO(A) and LC,,(B) are the median lethal concentrations measured for A and B individually. If the presence of one chemical enhances the toxicity of another, then the mixture will be more toxic than the additive toxicity of the two chemicals. This potentiation of the toxicity is termed synergy. For two synergistic chemicals the toxicity of the mixture will be significantly higher than expected by simple addition of the TIJs of each one:
Finally, the presence of one chemical may reduce the toxicity of another
124
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
and then the mixture would be less toxic than expected by simple addition of toxic units for each chemical individually. Thus: LGO(M)> $LCSO(A) + iLCso(B). In this case the effects are antagonistic and the interaction is termed antagonism. The same rationale applies when sub-lethal effects are studied. Various combinations of toxicants, like copper and zinc (MacInnes and Calabrese, 1978), mercury and silver (MacInnes and Calabrese, 1978), copper and silver (Coglianese and Martin, 198l), and manganese and molybdenum (Morgan et al., 1986), are thought to interact additively. The type of interaction may change depending on the absolute or even relative quantities of each metal. MacInnes (1981) found that the type of interaction might depend on the concentrations of the toxicants, with a trend towards synergism as the toxic units increased, at least for copper/zinc mixtures. This might be a result of saturation of the complexing capacity of the seawater (MacInnes, 1981). In contrast, Pereira et al. (1998) found less interaction (more additivity) in copper/zinc and copper/mercury combinations at higher concentrations, but the interactions were very weak and never departed more than 25% from the effects expected by simple addition. Johnson (1988) found 50% net response in embryos incubated in mesocosms at 5 pg of copper and 31.5 pg of diesel oil per litre. These concentrations are about one order of magnitude lower than the average individual embryogenesis EC50of these chemicals (see Tables 9 and 13, data for fuel oil), and thus appear to represent a case of synergy between a heavy metal and a hydrocarbon. Yet the author refers to the copper EC50of 5.8 pg 1-l found by Martin et al. (1981), and interprets the results as being caused by copper toxicity alone. This appears doubtful, however, because Martin's copper EC50 is one of the lowest values found for this metal (see Table 9). Instances of antagonistic interaction have also been described. Pavicic (1976) and Pavicic et al. (1994a) reported that zinc toxicity to Mytilus galloprovincialis embryos was reduced when they were simultaneously exposed to cadmium. Glickstein (1978) found that the toxicity of mercury to C. gigas embryos decreased in the presence of 10 to lOOOpgSel-', although the statistical significance was not verified. MacInnes and Calabrese (1978) did find a statistically significant less-than-additive interaction for a combination of mercury and silver on oyster embryos at two of the three temperatures assayed, but the authors explain this as owing to silver loss by precipitation of silver chloride. Finally, Pereira et al. (1998) reported significant but weak antagonism in some (though not all) mercury/copper and mercuryizinc combinations.
THE ASSESSMENT OF MARINE POLLUTION
125
In short, only interactions between certain heavy metals have been studied with bivalve embryo-larval tests, and it is mostly additive or quasi-additive effects that have been revealed. 5. ASSESSING MARINE ENVIRONMENTAL QUALITY WITH BIVALVE EMBRYO AND LARVAL BIOASSAYS
Besides being used for basic research on specific (clearly identified and dosed) chemical compounds, bivalve embryo and larval bioassays are increasingly being used to monitor the quality of environmental samples of unknown composition. The results of these investigations are rarely published in reviewed journals, as they often concern specific situations that are difficult to generalize with respect to other regions or seasons of the year. Frequently the results of these studies show important smallscale variations (Lourens et al., 1995), which further complicate monitoring programmes, and stress the need for simple, rapid and affordable screening techniques. The summary presented here may nevertheless provide an overview of the use of embryo and larval bioassay in environmental monitoring. 5.1. Algal and bacterial toxins
The deleterious effect of planktonic flagellates, the blooms of which may cause “red tides”, on early developmental stages of bivalves has been reported occasionally. Davis and Chanley (1955b) observed that fertdized eggs failed to develop and larvae did not grow at a normal rate in seawater sampled during a bloom of dinoflagellates. Ukeles and Sweeney (1969) suggested that reduced larval growth in bloom conditions might be a result of inhibition of feeding by dinoflagellate trichocysts. Cardwell et al. (1979b) linked the toxicity of South Puget Sound (Washington State, USA) water on bivalve larvae to the occurrence of red tides produced mainly by Ceratium fusus and Gymnodinium splendens. Toxicity was attributed to the dinoflagellates, because it disappeared when the cells were removed by a 1Opm screen, and was restored when the filtered cells were resuspended in laboratory seawater. However, these authors found that dinoflagellates were markedly more toxic when obtained from the wild (LCs0200 dinoflagellate cells per ml) than when obtained from laboratory cultures (L& 3000 cells per ml). Similarly, cultured Prorocentrum minimum dinoflagellates did not have toxic effects on Crassostrea virginica embryos (Wikfors et al., 1993; Wikfors and Smolowitz, 1995). In contrast, another dinoflagellate obtained from
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
126
Table14 Effect of extracts from red-tide-toxic and non-toxic mussels on embryogenesis success and larval growth of Mytilus galloprovincialis. Means of five replicates. Mouse units = quantity lethal to a mouse. Acetone controls had 11.2% abnormal D-Iarvae and 11.1pmd-' larval growth rate. (Beiras and His, unpublished.)
Extract concentration Abnormal D-larvae (%) (mouse units) Non-toxic Toxic 0.001 0.002 0.005
0.01 0.02 0.05 0.1
10.8 7.6 10.8 10.4 15.4 15.2 100
13.2 14.8 18.4 29.4 33.8 100 100
Larval growth rate (pm d-') Non-toxic
Toxic
11.9 11.0 11.8 12.2 11.4 11.6
11.4 10.3 8.6 7.1 7.4
-
-
-
natural waters, Cochlodinium heterolobatum, markedly reduced calcium uptake and increased mortality in Crassostrea virginica larvae at densities around 500 per ml (Ho and Zubkoff, 1979), although the authors claim that physical factors, rather than the toxin, were responsible for the adverse effects on the larvae. Variability in the toxicity of laboratoryreared algae may be explained by fluctuations in the production of toxin depending on culture conditions (Pan et al., 1996a,b). Thain and Watts (1987) found reduced embryogenesis success in natural seawater during a Gyrodiniurn aureolum bloom, although embryogenesis and G. aureolum cell density did not correlate. Granmo et al. (1988) found inhibited fertilization and development at environmentally realistic concentrations of Chrysochromulina polylepis. Gentien et al. (1991) found that cultures of G. aureolum were toxic to mussel embryos at densities above 2 cells per pl; above this threshold, toxicity was independent of cell concentration. During a red-tide episode at Arcachon Bay, we (Beiras and His, unpublished) reared mussel larvae in different concentrations of extracts from stocks of toxic and non-toxic mussels (as determined by the mouse test performed by IFREMER). Extracts were made up from 50gof digestive gland in 200 ml of acetone. Adult mussels were induced to spawn by thermal shock and the fertilized eggs were incubated at the different experimental treatments for 48 hours at 19°C 2 1°C. The larvae were then reared according to His and Robert (1982) at 19°C for 12 days. Extracts from toxic mussels inhibited embryogenesis and larval growth at lower concentrations than those obtained from non-toxic mussels (Table 14). A red marine bacterium has also been proven to be toxic to bivalve
THE ASSESSMENT OF MARINE POLLUTION
127
embryos (Brown, 1974), owing to the prodiginine pigment it produced (Brown, 1981). Brown and Roland (1984) purified an exotoxin produced by a shellfish-pathogenic Vibrio species with an ECso on oyster emlbryos of 46.6 pg toxin 1-'. Blooms of non-toxic algae may also result in bivalve larval mortalities because of their poor value as food, presumably as a consequence of low digestibility (Nelson and Siddall, 1988). Wikfors and Smolowitz ( 1995) confirmed the low food value of a dinoflagellate. 5.2. Urban and industrial effluents
Bioassays with early developmental stages of bivalves have been used to evaluate the toxicity of industrial and urban wastes intended for disposal at sea, such as pulp-mill wastes (Okubo and Okubo, 1962; Woelke, 1960a,b, 1967, 1972; Dimick and Breese, 1965; Courtright et al., 1971; Stauber et al., 1996), aluminium-plant effluents (His et al., 1996), mine tailings (Mitchell et al., 1985), effluents from recently dredged sediments (Wirth et al., 1996) and sewage sludge (Butler et al., 1990). These types of studies normally form part of pollution monitoring and abatement programmes, and therefore focus on practical aspects. The investigations often do not conform to the standards commonly applied in basic research and a comparison between studies is difficult. A more extensive review would therefore be of limited value. Wastes are a composite of numerous potential pollutants, sometinies in variable proportions (e.g. urban sewage) and their quantitative chemical composition is often unknown (e.g. Woelke, 1967; Courtright et al., 1.971). Therefore, the potential biological response cannot be attributed to any single toxicant. Even when the effluent is analysed chemically, toxicity can be assigned only tentatively to a certain component of the mixture, unless every component is assayed individually and synergistic and antagonistic interactions are accounted for (His el al., 1996). Moreover, effluents may prove to be toxic even when the concentrations of the individual toxicants screened by analytical chemistry are below the effective toxicity threshold of each single toxicant (e.g. Mitchell et al., 1985). This result emph,asizes the requirement for biological assays to evaluate pollution. Nevertheless, Okubo and Okubo (1962) and Cardwell et al. (1979a; as well as previous work by Woelke and co-workers) found that toxicity in pulp-mill effluents depended mostly on their biological oxygen demand (BOD) and chemical oxygen demand (COD) (e.g. embryotoxicity at 13ppm COD in Okubo and Okubo, 1962). Dimick and Breese (1965) showed that the toxicity of Kraft-mill effluent decreased with time of stabilization, particularly during the first two days, when the BOD declined by 90%.
128
E. HIS, R. BEIRAS AND M . N. L. SEAMAN
5.3. Receiving waters
Even when the source point of pollution can be identified, it may be preferable to bioassay the receiving water rather than the effluent, an approach pioneered by Woelke (1967, 1968, 1972). As mentioned above, the fact that industrial effluents are frequently complex mixtures of several potential pollutants in variable concentrations precludes quantification of effluent toxicity in absolute terms. Once released into the seawater, these mixtures undergo physico-chemical changes that are extremely difficult to predict or reproduce in the laboratory. These limitations provide a motivation to conduct receiving-water toxicity tests rather than tests of the effluent itself (Cardwell et al., 1977a). Bioassaying the quality of the receiving water does not overcome the difficulties of waste toxicity measurement, but it provides a straightforward and useful method for monitoring pollution at different points in space and time. This approach has also been followed in studies of waste treatment efficacy and pollution abatement (Cardwell et al., 1977a, 1979b). In one example, a 67% reduction in BOD loading in pulp-mill effluent within 3 years resulted in a 98% reduction of the area of receiving waters causing abnormal development in Crassostrea gigas embryos; in other cases, however, pollution abatement measures appeared unsuccessful or yielded contradictory results, e.g., “there appeared to be major problems in predicting receiving water toxicity on the basis of effluent bioassays with oyster larvae” (Cardwell et al., 1979a, p. 58). The bivalve embryo-larval bioassay is a valuable tool for monitoring fresh, brackish and saline water quality, because it provides a rapid and inexpensive method to identify “hot spots” on which further, more costly, research may then be focused. These assays have therefore been used recently in monitoring studies of rivers, estuaries, bays and coastal areas (Thain and Kirby, 1994; Hall et al., 1995; His and Beiras, 1995; His et al., 1997a). Other uses of bivalve embryo-larval bioassays in monitoring seawater include the work by Lloyd and Thain (1981) and Klockner et al. (1985), who found toxicity in seawater taken from the proximity of marine grounds used for dumping sewage sludge. Nelson et al. (1983) also demonstrated a decrease in water quality after the dumping of dredge spoils. Bourne et at. (1981, cited by Klockner et at. 1985) and Konar and Stephenson (1995) were able in some instances to identify spatial gradients of pollution in US harbours, the inner areas of the harbours generally being more polluted. McFadzen (1992) also identified a pollution gradient along a transect in the North Sea by using cryopreserved larvae. Nascimento (1989) bioassayed the seawater quality of a bay in Brazil which received heavy industrial discharges and identified sampling
THE ASSESSMENT OF MARINE POLLUTION
129
stations at which the water caused abnormal oyster embryo development. 5.4. Sediments
Depending on their purpose, sediment bioassays may be classified into two groups, (1) a priori studies of risk assessment with particulate material intended to be disposed at the sea (e.g. dredge spoils) and (2) apos~~eriori studies for monitoring the quality of natural seabed sediments. Permission to dump solid and liquid wastes, dredged material in particular, should depend on a prior evaluation of their quality, including toxicity bioassays. Bivalve embryo-larval bioassays are simple and sensitive candidates to be included in these evaluations, as first shown by Cardwell et al. (1976) and recently advocated by Lourens et al. (1995) and Van den Hurk et al. (1997). Coastal monitoring has frequently resorted to bivalve embryo-larval bioassays as biological criteria to detect pollution (McFadzen, 1992; 'hain, 1992; Butler et al., 1992). These bioassays are demonstrably useful in detecting sediment toxicity in the vicinity of oil rigs (Chapman et al., 1991), as well as in harbours (Fichet et al., 1998); in many cases, however, the overall assessment of contamination is rendered questionable by the spatial variability of the pollution, including patchiness on a scale of a few metres (Matthiessen et al., 1998). The toxicity of oil-polluted sediments decreases with time, because of weathering of the spilled oil. Oil-pslluted sediments maintain moderate and low residual toxicity after 8 months and 1.5 years of weathering, respectively (Beiras and His, 1995b, 50 g 1-'; Wolfe et al., 1995, 0.1% organic solvent extracts). The most toxic aromatic components of the crude oil are probably effective only during the first few days after a spill, because they are readily lost by evaporation and other weathering processes. More often, pollution sources are diffuse or unknown, and in these cases the bioassayed sediments are best investigated by general analytical chemistry. In those few cases that include chemical analysis, there are weak correlations between the toxicity and the chemical components identified in the bulk sediment (Long et al., 1990; Magnusson et al., 1996; Van den Hurk et al., 1997; Matthiessen et al., 1998) or in the pore water (Magnusson et al., 1996). But most importantly, the results of bivalve embryo-larval bioassays agree with data on the alteration of impacted benthic communities (Becker et al., 1990), i.e. in contrast with raw analytical chemistry the results of the bivalve bioassay are ecologically relevant. Sediment toxicity studies are difficult to compare because of differences
130
E. HIS, R. BEIRAS AND M. N.
L. SEAMAN
in methodology (see Section 3.3.2.2). Chapman and Morgan (1983) advocate the use of a single sediment concentration (20 g l-'), and this approach has been followed by a number of investigators. Alternatively, several elutriates of the same sediment at different concentrations may be tested, to identify the range of effective concentrations and to calculate the EC,, of the sediment in question. More frequently different dilutions of the same elutriate have been assayed (see Butler et al., 1992). Preliminary trials indicate that both approaches (either diluting the media before or after elutriation) yield the same results (R. Beiras, unpublished data with sea-urchin embryos), but more research on this and other aspects of elutriation is necessary. Methodological standardization is urgently needed; this is a common conclusion in various reviews of sediment toxicity bioassays (see Burton, 1992; Butler et al., 1992; Chapman et al., 1992; Lamberson et al., 1992; Luoma and Ho, 1993). Regardless of the procedure chosen for standardization, testing different sediment concentrations in order to establish a median effective concentration is strongly advisable, because it would help to describe the toxicity of different sediments on a quantitative basis. Taking into account the serious limitations mentioned above, sediments inhibiting bivalve embryogenesis at concentrations in the range of 1 to 2OOg1-1 may be considered toxic, and sediments with effective concentrations below 1g 1-' may be considered extremely toxic. We have obtained contradictory results and some instances of toxicity in unpolluted and otherwise non-toxic sediments at concentrations higher than 200 g 1-I; therefore we do not recommend elutriation above those concentrations. 6. SUMMARY AND DISCUSSION
The present-day status of bivalve embryo and larval ecotoxicological bioassays is largely the result of almost a century of research on bivalve aquaculture, and specifically of the work by Loosanoff and co-workers at Milford (Connecticut, USA) and by Walne and co-workers at Conwy (Wales, UK) since the mid-twentieth century. On the other hand, aquaculture activities have equally profited from the development of bioassays monitoring water quality in shellfish growing areas. One important example is the resolution of the crisis of the oyster industry of the Bay of Arcachon in the late 1970s, which caused the demise of one-half of the oyster farms at the time. Bioassays conducted in 1980 demonstrated that the TBT in antifouling paints of the numerous recreational boats in the bay inhibited embryogenesis at levels low enough to escape detection by the chemical analysis methods available at the time. In addition, follow-up studies demonstrated that TBT induced shell
THE ASSESSMENT OF MARINE POLLUTION
131
malformation in adult oysters (Alzieu, 1991), imposex in gastropods (Gibbs and Bryan, 1996), and a variety of other toxic effects in the marine environment. Protective regulations (notably the prohibition of TBT on recreational boats) were introduced as a result of these studies, all’owing oyster culture in the Bay of Arcachon to recover during the 1980s.
6.1. Sensitivity of bioassay organisms
The marine species used most often in assessments of pollutant toxicity and aquatic environmental quality include sea urchins (e.g. Paracenifrotus, Strongylocentrotus, Arbacia), used in laboratory studies since th,e late nineteenth century; rotifers (Brachionus plicatilis); crustaceans such as copepods (e.g. Acartia tonsa), brine shrimps (Artemia salina), mysids (e.g. Mysidopsis bahia), and various barnacles; and bivalves such as oysters (Crassostrea gigas, C. virginica and Ostrea edulis), mussels (Mytilus edulis and M. galloprovincialis), and clams (Mercenaria mercenaria and k’ulinia lateralis). Another method frequently used in environmental monitoring is the Microtox test kit, based upon bioluminescent bacteria. Each of these groups of organisms presents a variety of advantages and disadvantages with regard to year-round availability, ease of handling, sensitivity to contaminants, reliability and accuracy of evaluation of the response to the pollutant. According to Widdows (1993, p. 152), “the advantages of using bivalve embryos as an acute lethal test are: (j) the short exposure times (24-48 hours); (ii) the biological end point is easily determined (i.e. shelled larvae); (iii) the range of salinity tolerance (<15 to 35 psu); (iv) the moderate sensitivity; (v) the relatively low cost; and (vi) bivalve species are both commercially and ecologically important. There are also disadvantages. To achieve a high survival of embryos through to the larval stage requires: (i) a supply of high quality conditioned brood stock; and (ii) considerable care at all stages prior to the actual test (i.e. handling, conditioning, spawning and fertilization).” With regard to point (iv) above, the sensitivity of the bivalve embryogenesis bioassay depends on the choice of criterion. If the toxicity of the compound being assayed is judged on the basis of larval abnormality as described in Section 3.2.2.5 (i.e. by comparison to larvae from the natural environment), then the bivalve embryo test is one of the most sensitive in existence today. Ultimately, sensitivity depends on the toxicant tested, which can selectively affect certain taxa, but overall, bivalve larvae have frequently been proved to be more sensitive to rnarine pollutants than other test species. Heavy metals, some organics and pulp waste were toxic to bivalve embryos at lower concentrations than those that were toxic to Artemia and Balanus adults and nauplii (Okubo and
132
E. HIS. R. BEIAAS AND M. N . L. SEAMAN
Okubo, 1962). Chlorine affected bivalve embryos more than Acartia, Brachionus, Homarus larvae and Onchorhynchus (Capuzzo, 1979); bromate affected them more than adult mysids and juvenile salmon (Crecelius, 1979). Higher sensitivity to cyanides than adult mysids and fish was found by Pavicic and Pihlar (1982), and higher sensitivity to pulp-mill effluents than Microtox, microalgal growth, macroalgal fertilization, fish larval survival and sea urchin fertilization was found by Nelson et al. (1983) and Stauber et al. (1996). Finally, higher sensitivity to polluted sediments was reported by Long et al. (1990) compared to the amphipod survival test and the polychaete (Dinophilis) reproduction test (but see also Williams et al., 1986; and Becker et al., 1990). There are exceptions; as expected, bivalve larvae are less affected than crustaceans by insecticides and other biocides that target arthropods [e.g. Tagatz and Ivey (1981) for Fenvalerate; Thain et al. (1990) for Dichlorvos]. The choice of bivalve species depends largely on practical aspects (see below), as the various bivalves commonly employed in toxicity tests do not differ greatly with regard to sensitivity and handling. The choice of a biological response to be measured in test individuals depends on a compromise between sensitivity and feasibility. The scale of relative lethal sensitivity in bivalves is embryo > veliger > pediveliger > adult (Figure 14). Among the various possible toxicity tests with bivalve larvae the embryogenesis bioassay will usually be the method of choice for both pure chemical toxicity tests and routine environmental monitoring. Metamorphosis success (i.e. settlement of pediveligers) and measurement of chronic effects on larval growth may be similar in sensitivity to the criterion of embryogenesis success, but they require rearing the larvae for weeks, entailing considerable effort and cost (see Section 3.2.3) and rendering these bioassays impractical for routine investigations. They may well be preferable in studies of particular hot spots and for risk assessment of new chemical products likely to enter the marine environment. Tests with gametes (spermiotoxicity, unfertilized eggs) may be simple from a methodological point of view, but are not commonly employed, and their sensitivity is lower than that of the embryogenesis test. 6.2. Assessment of the toxicity of various contaminants
Any discussion on ecotoxicological bioassays must make a careful distinction between toxicological investigations with identified compounds on one hand, and routine monitoring techniques with environmental samples on the other. The former mostly concern basic research on specific toxicants (e.g. new industrial compounds) and on methodological aspects (e.g. comparison of bioassay techniques); the latter focus on general
133
THE ASSESSMENT OF MARINE POLLUTION
I
1
SUB-LETHAL EFFECTS
LETHAL EFFECTS Lc50
EC50
larva growth embryogenesis D-larva metamorphosis
adult (Connor, 1972)
1
ybyrd:eliger
estuaries :P
I
I 0.01
+
I
I
0.1
1
&:-:a Iw w p J 21 8 PfJl 0
,c
A
A
v
10
v
100
1
1000
v
10 000
Hg (P9. I-’)
Figure 14 Sub-lethal (0)and lethal (+) effects of mercury upon the oyster. Larval growth and embryogenesis are the most sensitive responses, but the latter is more easily and rapidly assessed.Data from Beiras and His (1994) except for the adult. Usual mercury concentrations in estuarine waters are also shown.
pollution monitoring and abatement. With regard to both, we need to keep in mind that biological systems are extremely diverse and complex, and that no single method or species could ever be adequate to measure pollution “as such”. A battery of bioassays may often be required for a fairly adequate assessment of a particular pollution phenomenon. B a k e r et al. (1990), for instance, advocate the use of the oyster embryo bioassay in samples determined as toxic by the Microtox bioassay, which is supposedly more sensitive, but less reliable. The most toxic compound in the marine environment bioassayed to date with bivalve embryos and larvae is TBT, the toxicity of which is generally one or two orders of magnitude greater than that of any other compound bioassayed to date. The further ranking of pollutants is heavy metals (especially mercury, silver and copper), chlorine and derived oxidants, organics and pesticides (especially organo-chlorine), detergents, petroleum products and, depending on their composition, industrial1 and urban effluents and sediments (Figure 15). Apart from a considerable amount of variation between the results of different authors, the toxicity of many compounds is affected by chemical and physical conditions and interactions of a complex nature, often defying precise analysis. ‘These include bioassay methodology, chemical speciation and complexation, synergistic and antagonistic effects, and seasonality in the bioassay organisms (see also Sections 4.2 and 4.3). As Figure 15 shows, the range of reported effective concentrations spans at least two orders of magnitude, even in the case of precisely defined
134
E. HIS, R. EElRAS AND M. N. L. SEAMAN
1 PPt 0
y
In
1PPm
1 PPb
Materials Figure 15 Toxicitv of various materials to bivalve embryos and larvae in the publiiations reviewed here (range of median effective concentrations;EC50).TBT tributyl-tin; HM: heavy metals; C1: halogens and halogen-produced oxidants; C1-Pest: organochlorinated pesticides; Pest: other pesticides; Det: surfactants and other components of detergents; Petr: petroleum products; Effl: industrial and urban effluents; Sed: sediments.
compounds, depending on biological and chemical species, larval stages, experimental regimes and criteria of toxicity. On the other hand, scientists are often asked to provide clear and simple data as a basis for political decision making. This underscores the continuing need for standardization. In the case of many substances (contaminated sediments in particular), there are considerable problems with regard to the methodology of sampling and the handling of samples; standardization of these methods is urgently needed, but it will not be easy to achieve. Correspondingly,the use of bioassays for the purpose of enacting and monitoring pollution abatement measures may yield contradictory results for some time yet. 6.3. Bivalve embryo and larval bioassay methodology
In the case of using bivalve embryos and larvae for the purpose of pollution assessment, decades of experience as well as recent improvements and simplifications in methodology do render standardization achievable (e.g. ASTM, 1980, 1989; Thain, 1991; Krassoi et al., 1996; His et al., 1997a). With respect to routine investigations of environmental quality, the embryogenesis bioassay is generally regarded as the method of
THE ASSESSMENT OF MARINE POLLUTION
135
choice, even though differences of opinion may exist concerning certain details of methodology. The following aspects should be considered in future standardization efforts.
6.3.1. Bivalve species The various species commonly used in embryotoxicity bioassays (Crassostrea gigas, C. virginica, Mytilus edulis, M. galloprovincialis, Mercenaria mercenaria, Mulinia lateralis) appear to have similar sensitivity to environmental contaminants (see Section 4.2.1 and Table 12). The choice of species to be used in a bioassay therefore depends mostly on practical considerations. In studies concerning a shellfish-culture area, for instance, the species of choice will be the one being cultivated on site. Profound knowledge of its biology and maintenance in the laboratory is essential to prevent artefacts in the bioassay. Other aspects include availability (e.g. use of a species during its spawning season rather than using conditioned broodstock), sensitivity, commercial interest and ecological relevance. With regard to more sophisticated basic research (e.g. chronic and persistent toxicity, physiological and genetic adaptation to pollution) the American coot clam Mulinia lateralis offers interesting perspectives. Its small size and the brevity of its life cycle make it relatively easy to cultivate several generations per year in adequately equipped laboratories, although these advantages have not been exploited fully in toxicological research to date. Doubtless this is largely because of its limited natural distribution (the Atlantic coast of North America) and the present-day difficulties and necessary precautions required for importing and holding exotics in the laboratory. Possibly, species with similar Characteristics may yet be found in Europe and other parts of the world.
6.3.2. Broodstock availability and gamete production Ripe broodstock of Crassostrea and/or Mytilus is generally available almost year-round because conditioning techniques extend the seasonal availability. Even though it may be considered controversial by solme of the researchers in the field, we propose that “stripping” be abandoned as a method of obtaining gametes for embryotoxicity bioassays. Most authors who have worked with bivalve larvae will undoubtedly concur, because this method results in unacceptably high levels of mortality and abnormality in the controls, introducing gamete quality as an undesirable factor of variation in the test. When inducing spawning by thermal stimulation of the broodstock, rigid temperature limits are useless; adult bivalves should be stimulated to
136
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
spawn by temperature variations of about lWC, for example 5°C above and below the ambient water temperature at the time of sampling. 6.3.3. Incubation water At many marine laboratories the quality of the water is inadequate for rearing bivalve larvae, even though it may be sufficient for holding the adults of various species. In such cases water should be obtained 1 or 2 days before the experiment from areas known to be unpolluted. Filtration of the water at 0.2 pm is generally recommended. Alternatively, artificial seawater is adequate, and it also eliminates any possibility of contaminants in natural seawater (for example dissolved organic matter); the cost is negligible when only small volumes are used (see below).
6.3.4. Toxicant exposure and observation of the response Fertilization should be achieved as soon as possible after the first females spawn, in order to avoid deterioration of gamete quality. Toxicant exposure should begin as soon as fertilization of the gametes has been achieved because this enhances the sensitivity of the test; there is usually no particular reason for delay. Similarly, the bioassay should be terminated as soon as embryogenesis is achieved (24 hours in most bivalve species, 48 hours in Mytifus).There is no particular reason to continue a test of embryogenesis beyond the time required for embryogenesis, as prolonged incubation introduces the risk of mortalities from causes other than the toxicant under study (e.g. bacterial proliferation resulting from decomposition of undeveloped eggs and dead embryos). For the sake of accuracy and precision, further manipulation of the biological material (sieving, sub-sampling, etc.) should be kept to a minimum once a bioassay has been initiated. Therefore, small volume incubations in suitable vessels permitting a direct assessment of the material under the microscope (e.g. transparent 30 ml vials, and 3 ml microwells) are particularly suitable. 6.3.5. Assessment of embryogenesis success Many workers have in the past employed definitions of larval normality which do not reflect the appearance of larvae from the natural environment; by definition, however, the criteria of normality should conform to the natural situation. A strict and accurate definition of larval abnormality (see Section 3.2.2.5, Figure 7) has yet to become generally accepted. At the same time, a valid acceptable level of abnormalities in the controls needs to be agreed upon; we suggest that embryogenesis bioassays are invalid if
THE ASSESSMENT OF MARINE POLLUTION
137
the level of larval abnormality in the controls is higher than 20%. The percentage of abnormal controls must always be stated when publishing an investigation. 6.3.6. Statistical evaluation of bioassays In embryo-larval bioassays end points are often measured as proportions. Angular transformation is generally sufficient to overcome the difficulties of binomially distributed variables, but this must not be taken for gr.anted and alternative methods are available (see Section 3.4.1). Besides, very little attention is often paid to the question of sample size required for precise estimation of the biological responses and powerful detection of differences among them (see Section 3.4.3). The assessment of environmental samples by means of bivalve Iiioassays has relied too often on arbitrary scales (e.g. Woelke, 1966), or at best, calculation of “percent net risk” (PNR; equation 2, Section 3.2.2.5) The PNR is highly influenced by the control response, and comparison of PNR values calculated with controls that exhibit highly different responses is unacceptable. As mentioned above, we consider that the PNR values have limited value when the level of larval abnormality in controls is :.20%. Alternatively, a more systematic approach based on ANOVA-multiple range tests is recommended (see Section 3.4.2). 6.4. Perspectives in future research on bivalve larval bioassays
Despite the fact that bivalve embryogenesis and larval bioassays are presently approaching the status of routine methods, there is still ample room for improvement of the techniques. Research in the immediate future should focus on the following aspects:
1. The definition of a “standard” organism that is easy to cultivate in the laboratory (e.g. Mulinia lateralis) as a reference species, particularly if a genetically homogeneous laboratory strain is created. This would help in the standardization of toxicological investigations and routine monitoring studies, as well as in the inter-calibration between different laboratories, rendering all types of studies directly comparable. 2. Cryopreservation of fertilized or unfertilized gametes and the evaluation of gamete quality immediately after spawning. In conjunction with (1) above, this would eliminate the effects of genetic, seasonal and inter-annual variability, and facilitate the performance of bioassays at all times of the year and at laboratories lacking sophisticated conclitioning and larval-rearing equipment.
138
E. HIS, R. BEIRAS AND
M. N. L. SEAMAN
3. Development of simple methods for performing long-term bioassays on larval performance (e.g. growth, swimming behaviour, feeding behaviour) to assess sub-lethal effects of toxicity. One central element is the development of an adequate, commercially available and standardized artificial diet (e.g. freeze-dried algae), to liberate long-term tests from the slavery of algal production. In this respect, toxicological research again shares a common interest with aquaculture (commercial hatcheries and nurseries). 4. Automatization of the assessment responses to toxicant exposure. Small incubation volumes in standardized incubation recipients, as well as technological progress, may transform this aspect of marine ecotoxicological research in the near future. Automatic image analysis systems may soon permit the evaluation of larval abnormalities in acute toxicity tests and larval growth in chronic toxicity studies with a minimum of human intervention and with a maximum of comparability.
6.5. Concluding remarks
Bivalve embryo and larval bioassays have shown that in various cases where environmental concerns have led to the replacement of toxic substances by new compounds (e.g. linear instead of non-linear surfactants), the replacement may be more hazardous to the environment than the original compound. This may also be the case when a remedy is used to combat an environmental threat (e.g. the greater toxicity of oil dispersants in comparison to the oil itself). This not only raises questions with regard to environmental regulation and pollution abatement actions but also emphasizes the necessity of assessing the biological effects of new compounds before they are mass produced and before they are introduced into the natural environment in appreciable quantities. Finally, after all of this is said and done, we should keep in mind that there is one decisive aspect preferable to each and every refinement of toxicity assessment, environmental monitoring methodology, pollution abatement and bioassay techniques, and that is to introduce fewer contaminants, and in lesser quantities, into the natural environment.
ACKNOWLEDGEMENTS We thank M. L'Excellent and A. Radenac (central library of IFREMER, Nantes) and S. Robinson (library of Plymouth Marine Laboratory) for
THE ASSESSMENT OF MARINE POLLUTION
139
their invaluable help in obtaining rare specimens of literature, and C. Cantin (IFREMER, Arcachon) for his important technical assistance with respect to many of the illustrations.
REFERENCES Abel, I? D. (1996). “Water Pollution Biology”. 2nd edn, Longman, Essex, 286 pp. Aboul-Ela, I. A. (1960). Conditioning Ostrea edulis L. from LimfjoId for reproduction out of season. Meddelelser fra Kommissionen for Havundersogelser, Kobenhavn New Series 2(25), 1-14. Adams, J. A. (1983). Effects of PCB (Aroclor 1254) on development and mortality in Arbacia eggs. Water, Air and Soil Pollution 20, 1-5. Adams, W. J. (1987). Bioavailability of neutral lipophilic organic chemicals contained in sediments: a review. In “Fate and Effects of Sediment-13ound Chemicals in Aquatic Systems” (K. L. Dickinson, A. W. Maki and W. A. Brungs, eds), pp. 219-244. Pergamon Press, New York. Aguirre, M. P. (1979). Biologia del mejill6n (M. edulis) de cultivo de la Ria de Vigo. Boletin del Instituto espaiiol de Oceanografia 5(276), 109-159. Ahmed, M. (1975). Speciation in living oysters. Advances in Marine Biology 13, 357-397. Akberali, H.B. and Trueman, E. R. (1985). Effects of environmental stress on marine bivalve molluscs. Advances in Marine Biology 22, 102-198. Akberali, H. B., Earnshaw, M. J. and Marriott, K. R. M. (1985). The action of heavy metals on the gametes of the marine mussel, Mytilus edulis (L.). I1 Uptake of copper and zinc and their effects on respiration in the sperm and unfertilized eggs. Marine Environmental Research 16,37-49. Allen, S. K. Jr, Downing, S. L. and Chew, K. K. (1988). “Hatchery manual for producing triploid oysters”. Washington Sea Grant Program, University of Washington Press, Seattle, WA, pp. 1-27. Alliegro, M. C. and Wright, D. A. (1983). Polyspermy inhibition in the oyster, Crassostrea virginica. Journal of Experimental Zoology 227, 127-137. Alzieu, C. (1991). Environmental problems caused by TBT in France: assessment, regulation prospects. Marine Environmental Research 32,7-17. Amemiya, I. (1929). On the sex-change of the Japanese common oyster, iOstrea gigas Thunberg. Proceedings of the Imperial Academy of Japan 7 , 284286. Amiard, J. C. (1983). Les tests de toxicit6 sublCtale: approches biochimiques et Bthologiques. OcCanis 9(6), 465480. Amiard-Triquet, C. (1983). Les tests de toxicite aigue en milieu aquatique: mCthodologie, standardisation, interprktation, limites d’application. Octanis 9(6), 451-453. Anderson, J., Birge, W., Gentile, J., Lake, J., Rodgers, J. Jr. and Swartz, R. (1987). Biological effects, bioaccumulation, and ecotoxicology of sediment-associated chemicals. In “Fate and effects of sediment-bound chemicals in aquatic sys tems” (K. L. Dickson, A. W. Maki and W. A. Brungs, eds), pp. 267-296. Pergamon Press, New York. Andral, B. (1996). DonnCes sur le comportement et les effets des produits phytosanitaires dans I’environnement. Institut FranGais de Recherche pour
140
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
L‘Exploitation de la Mer, Rapport Interne de la Direction de YEnvironnement et de l’Am6nagement littoral DEW94 03, Issy-les Moulineaux, France, 106 pp. Andrews, J. D. (1979). Pelecypoda: Ostreidae. Zn “Reproduction of Marine Invertebrates” (A. C. Giese and J. S. Pearse, eds), vol. 5, pp. 293-341. Academic Press, New York. Andrews, J. D. (1980). A review of introduction of exotic oysters and biological planning for new importations. Marine Fishery Research 42, 1-11. Ankley, G. T. and Schubauer-Berigan, M. K. (1994). Comparison of techniques for the isolation of sediment porewater for toxicity testing. Archives of Environmental Contamination and Toxicology 27, 507-512. Ankley, G. T., Di Toro, D. M., Hansen, D. J. and Berry, W. J. (1996a) Assessing the ecological risk of metals in sediments. Environmental Toxicology and Chemistry 15,2053-2055. Ankley, G. T., Di Toro, D. M., Hansen, D. J and Berry, W. J.. (1996b) Technical basis and proposal for deriving sediment quality criteria for metals. Environmental Toxicology and Chemistry 15,2056-2066. Anonymous (1989). Report of the Working Group on the Biological Effects of Contaminants. Conseil International pour 1’Exploration de la Mer, C.M. 1989/E: 19, 16 pp. Ansell, A. D. (1967). Egg production of Mercenaria mercenaria. Lirnnology and Oceanography 12, 172-176. Ansell, A. D. (1968). The rate of growth of the hard clam, Mercenaria mercenaria (L.) throughout the geographical range. Journal du Conseil International pour 1’Exploration de la Mer, Copenhagen, 31, 364-409. Armstrong, D. A. and Millemann, R. E. (1974). Effects of the insecticide Sevin and its first hydrolytic products, 1-naphthol, on some early developmental stages of the Bay mussel Mytilus edulis. Marine Biology 28, 11-15. Askew, C. G. (1972). The growth of oysters Ostrea edulis and Crassostrea gigas in Emsworth Harbour. Aquaculture 1, 237-259. ASTM (1980). Standard practice for conducting static acute toxicity tests with larvae of four species of bivalve molluscs. American Society for Testing and Materials (ASTM) E 724 -80, Philadelphia, PA, pp. 246262. ASTM (1989). Standard guide for conducting static acute toxicity tests starting with embryos of four species of saltwater bivalve molluscs. American Society for Testing and Materials, Philadelphia, (ASTM) E 724-89, PA, pp. 334-351. Babinchak, J. and Ukeles, R. (1979). Epifluorescence microscopy, a technique for the study of feeding in Crassostrea virginica veliger larvae. Marine Biology 51, 69-76. Bacci, G. (1951). Ermafroditismo ed intersessualita nei Gasteropodi e Lamellibranchi. Archivio Zoologico Ztaliano, Supplemento 7 , 57-151. Bagarino, T. (1992). Sulfide as an environmental factor and toxicant: tolerance and adaptations in aquatic organisms. Aquatic Toxicology 24(1,2), 21-62. Bahr, L. M. and Hillman, R. E. (1967). Effects of repeated shell damage on gametogenesis in the American oyster, Crassostrea virginica (Gmelin). Proceedings of the National Shellfisheries Association 57, 59-62. Baker, S. M. and Mann, R. (1992). Effects of hypoxia and anoxia on larval settlement, juvenile growth and juvenile survival of the oyster, Crassostrea virginica. Biological Bulletin, Marine Biological Laboratory, Woods Hole 182, 265-269. Baker, S. M. and Mann, R. (1994a). Description of metamorphic phases in the oyster, Crassostrea virginica and effects of hypoxia on metamorphosis. Marine
THE ASSESSMENT OF MARINE POLLUTION
141
Ecology Progress Series 104, 91-99. Baker, S. M. and Mann, R. (1994b). Feeding ability during settlement and metamorphosis in the oyster Crassostrea virginica (Gmelin, 1971) and the effects of hypoxia on post-settlement ingestion rates. Journal of Experimental Marine Biology and Ecology 181, 239-253. Baldwin, B. S. (1995). Selective particle ingestion by oyster larvae (Crassoatrea virginica) feeding on natural seston and cultured algae. Marine Biology 123, 95-108. Bayne, B. L. (1965). Growth and the delay of metamorphosis of the larvae of Mytilus edulis L. Ophelia 2, 1 4 7 . Bayne, B. L. (1971). Some morphological changes that occur at the metamorplhosis of the larvae of Mytilus edulis. In “4th European Marine Biology Symposium” (J. Crisp, ed.), pp. 259-279. Cambridge University Press, Cambridge. Bayne, B. L. (1972). Some effects of stress in the adult on the larval development of Mytilus edulis. Nature 237, 459. Bayne, B. L. (1975). Reproduction in bivalve molluscs under environmental st.ress. In “Physiological ecology of estuarine organisms” (F. J. Vernberg, ed.)., pp. 259-277. University of South Carolina Press, Columbia, SC. Bayne, B. L. (1976). The biology of mussel larvae. In “Marine mussels, their ecology and physiology” (B. L. Bayne, ed.), pp. 81-119. Cambridge University Press, Cambridge. Bayne, B. L. (1983). Physiological ecology of marine molluscan larvae. In “The Mollusca” (W. Wilbur, ed.), vol. 3, pp. 299-343. Academic Press, London. Bayne, B. L. (1985). Ecological consequences of stress. In “The Effects of Stress and Pollution on Marine Animals” (B. L. Bayne, D. A. Brown, K. Bums, D. R. Dixon, A. Ivanovici, D. R. Livingstone, D. M. Lowe, M. N. Moore, A. R. D. Stebbing and J. Widdows, eds), pp. 141-157. Praeger Publishers, New Yorl;. Bayne, B. L., Gabbott, P. A. and Widdows, J. (1975). Some effects of stress in the adult on the eggs and larvae of Mytilus edulis. Journal of the Marine Biological Association of the United Kingdom 55, 675689. Bayne, B. L., Holland, D. H., Moore, M. N., Lowe, D. M. and Widdows, J. (1978). Further studies on the effects of stress in the adult on the eggs of Mytilus ed.ulis. Journal of the Marine Biological Association of the United Kingdom 58, 825-841. Bayne, B. L., Clarke, K. R. and Moore, M. N. (1981). Some practical considerations in the measurement of pollution effects on bivalve molluscs, and some possible ecological consequences. Aquatic Toxicology, 1, 1-16. Beaumont, A. R. and Budd, M. D. (1982). Delayed growth of Mussel, (Mytilus edulis) and Scallop (Pecten maximus) veligers at low temperatures. Ma.rine Biology 71, 97-100. Beaumont, A. R. and Budd, M. D. (1984). High mortality of the larvae of the common mussel at low concentrations of tributyltin. Marine Pollution Bulletin 15,402-405. Beaumont, A. R. and Fairbrother, J. E. (1991). Polyploidy manipulation in molluscan shellfish: a review. Journal of Shelljish Research 10, 1-18. Beaumont, A. R. and Tserpes, G. (1984). The effects on Mytilus edulis larvae of toxins leaching from paints. Aquaculture 38, 365-369. Beaumont, A. R., Tserpes, G. and Budd, M. D. (1987). Some effects of copper on the veliger larvae of the mussel Mytilus edulis and the scallop Pecten maxiinus (Mollusca Bivalvia). Marine Environmental Research 21, 299-309. Beaumont, A. R., Newman, I? B., Mills, D. K., Waldock, M. J., Miller, D. and
142
E. HIS, R. BEIRAS AND M . N. L. SEAMAN
Waite, M. E. (1989). Sandy-substrate microcosm studies on tributyl tin (TBT) toxicity to marine organisms. Topics in Marine Biology (J. D. Ros, ed.), Scientia Marina 53(2-3), 737-743. Becker, D. S., Bilyard, G. R. and Ginn, T. C. (1990). Comparison between sediment bioassays and alterations of benthic macroinvertebrate assemblage at a marine Superfund site: Commencement Bay, Washington. Environmental Toxicology and Chemistry 9, 669685. Beiras, R. and His, E. (1994). Effects of dissolved mercury on embryogenesis, survival, growth and metamorphosis of Crassostrea gigas oyster larvae. Marine Ecology Progress Series 113, 95-103. Beiras, R. and His, E. (1995a). Effects of dissolved mercury on embryogenesis, survival, growth and metamorphosis of Mytilus galloprovincialis mussel larvae. Marine Ecology Progress Series 126, 185-189. Beiras, R. and His, E. (1995b). Toxicity of fresh and freeze-dried hydrocarbonpolluted sediments to Crassostrea gigas embryos. Marine Pollution Bulletin 30, 47-49. Beiras, R. and Widdows, J. (1995). Induction of metamorphosis in Crassostrea gigas larvae using neuroactive compounds. Marine Biology 123, 327-334. Beiras R.,His, E. and Seaman, M. N. L. (1998). Effects of storage temperature and duration on the toxicity of sediments assessed by the Crassostrea gigas oyster embryo bioassay. Environmental Toxicology and Chemistry 17, 2100-2105. Bittkau, C. (1993). Auswirkungen von Schwefelwasserstoff-Exposition auf Larven ausgewwter mariner Wirbelloser. M.Sc. Thesis, University of Bremen, 61 pp. Bittkau, C. (1996). Effects of sulfide on reproduction of marine molluscs. In “Abstracts and program of the workshop on processes and structures in marine methane and sulfide biotopes” (M. K. Grieshaber and U. Fischer, eds), pp. 71-72. Shaker Verlag, Aachen. Boidron-MCtairon, I. F. (1995). Larval nutrition. In “Ecology of Marine Invertebrate Larvae” (L. McEdward, ed.), pp. 223-248. CRC Press, Boca Raton. Boudou, A. and Ribeyre, F. (1989). Aquatic ecotoxicology: fundamental concepts and methodologies. CRC Press, 2 volumes, 332 pp and 314 pp. Bourne, N., Rogers, H., Mahood, H. and Neil, D. (1981). Water quality study in Ladysmith Harbour, British Columbia. Canadian Technical Report of Fisheries and Aquatic Sciences 1026, 1-57. Boyden, C. R., Watling, H. and Thornton, I. (1975). Effect of zinc on the settlement of the oyster Crassostrea gigas. Marine Biology 31, 227-234. Breese, I? and Robinson, A. (1981). Razor clams, Siliqua patula (Nelson): gonadal development, induced spawning and larval rearing. Aquaculture 22, 27-33. Breese, W. P., Milleman, R. E. and Dimick, R. E. (1963). Stimulation of spawning in the mussels, Mytilus edulis Linnaeus and Mytilus californianus Conrad, by Kraft mill effluent. The Biological Bulletin 125, 197-205. Brereton, A., Lord, H., Thornton, I. and Webb, J. S. (1973). Effect of zinc on growth and development of larvae of the Pacific oyster Crassostreagigas. Marine Biology 19, 96-101. Bricelj, V. M. and Malouf, R. E. (1980). Aspects of reproduction of hard clam (Mercenaria mercenaria) in Great South Bay, New York. Proceedings of the National Shellfisheries Association 70, 216-229. Brock, V. (1985). Immuno-electrophoretic studies of genetic relationship between populations of Mytilus edulis and M. galloprovincialis from the Mediterranean, the Baltic, the East and West Atlantic, and the East Pacific. In “Proceedings of the 19th European Marine Biology Symposium, Plymouth, 1984” (P. E. Gibbs,
THE ASSESSMENT OF MARINE POLLUTION
143
ed.), pp. 515-520. Cambridge University Press, Cambridge. Brown, C. (1974). A pigment-producing pseudomonad which discolors culture containers of embryos and bivalve mollusks. Chesapeake Science 15, 17-21. Brown, C. (1981). A prodigine pigment toxic to embryos and larvae of Crassostrea virginica. Journal of Invertebrate Pathology 38, 281-293. Brown, C. and Roland, G. (1984). Characterization of exotoxins produced by a shellfish-pathognenic Vibrio sp. Journal of Fish Diseases 7 , 117-126. Brunetti, R., Gola, I. and Majone, F. (1986). Sister-chromatid exchange in developing eggs of Mytilus galloprovincialis Lmk (Bivalvia). Mutation Research 174, 207-211. Brunetti, R., Bressan, M., Marin, M. G. and Fuolega, M. (1989). Sodium nitrilotriacetate (NTA) influences the larval development and metamorphosis of marine invertebrates. Bulletin of Environmental Contamination and Toxicology 42, 234-240. Bryan, G. W., Langston, W. J. and Hummerstone, L. C. (1980). The use of biological indicators of heavy metals contamination in estuaries. With special reference to an assessment of the biological availability of metals in estuarine sediments from South-West Britain. Occasional Publication, Marine Biolcigical Association, no. 1. Bryan, G. W., Langston, W. J., Hummerstone, L. C. and Burt, G. R. (1985). A guide to the assessment of heavy-metals contamination in estuaries using biological indicators. Occasional Publication, Marine Biological Association, no. 4. Bucaille, D. and Kim, N. (1979). Influence de la chloration sur la fixation, le dktachement, la croissance des larves pCdivCligkres et le jeune naissain de nioule et sur la gamCtogCnbse des adultes. In “2””” JournCes de la Thermo-Ccologie, Institut Scientifique et Technique des P&chesMaritime, 14-15 Novembre 1979” (ElectricitC de France, eds), pp. 386-411. ElectricitC de France, Paris. Burgess, R. M. and Morrison, G. E. (1994). A short-term exposure, sublethal, sediment toxicity test using the marine bivalve Mulinia lateralis:statistical design and comparative sensitivity. Environmental Toxicology and Chemistry 13, 571-580. Buroker, N. E. (1983). Population genetics of the American oyster Crmsostrea virginica along the Atlantic coast-and the Gulf of Mexico. Marine Biology 75, 99-112. Buroker, N. E. (1985). Evolutionary patterns in the family Ostreidae: larviparity and oviparity. Journal of Experimental Marine Biology and Ecology 90, 233-247. Burstein, H. (1971). “Attribute sampling - tables and explanations”. McGraw-Hill, New York, 464 pp. Burton, G. A. Jr (1992). “Sediment toxicity assessment”. Lewis Publishers, London, 211 pp. Burton, G. A. J r and Scott, K. J. (1992). Sediment toxicity evaluations. Their niche in ecological assessments. Environmental Science and Technology 26, 20682075. Butler, P. A. and Lowe, J. I. (1978). Flowing sea water toxicity tests using oyster (Crassostrea virginica). In “Bioassay Procedures for the Ocean Disposal Permit Program”, EPA-600/9-78-010, pp. 25-27. Butler, P. A., AndrCn, L., Bonde, G. J., Jernelov, A. and Reish, D. J. (1971). Monitoring organisms. In F A 0 Fisheries Reports, no. 99 (suppl. l ) , pp. 101-112. Butler, R., Roddie, B. D. and Mainstone, C. I? (1990). The effects of sewage sludge
144
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
on two life-history stages of Mytilus edulis. Chemistry in Ecology 4, 211-219. Butler, R., Chapman, P. M., van den Hurk, P., Roddie, B. and Thain, J. E. (1992). A comparison of North American and West European oyster embryo-larval toxicity tests on North Sea sediments. Marine Ecology Progress Series 91,245-251. Byme, C. J. and Calder, J. A. (1977). Effect of water soluble fractions of crude, refined and waste oils on the embryonic and larval stages of the quahog clam Mercenaria sp. Marine Biology 40,225-231. Cahn, A. R. (1950). Oyster culture in Japan. Fishery Leaflet, Fisheries and Wildlife Service 383,140. Calabrese, A. (1969a). The early life history and larval ecology of the coot clam, Mulinia lateralis (Say) (Mactridae; Pelecypoda). Dissertation, University of Connecticut, University of Rhode Island, Auburn University, 101 pp. Calabrese, A. (1969b). Mulinia lateralis: Molluscan fruit Ay? Proceedings of the National Shelljkheries Association 59,6546. Calabrese, A. (1969~).Individual and combined effects of salinity and temperature on embryos and larvae of the coot clam, Mulinia lateralis (Say). The Biological Bulletin 137,417-428. Calabrese, A. (1970a). Reproductive cycle of the coot clam, Mulinia lateralis (Say), in Long Island Sound. The Veliger 12,265-269. Calabrese, A. (1970b). The pH tolerance of embryos and larvae of the coot clam, Mulinia lateralis (Say). The Veliger 13, 122-126. Calabrese, A. (1972). How some pollutants affect embryos and larvae of American oyster and hard-shell clam. Marine Fisheries Review 34,66-77. Calabrese, A. (1984). Ecotoxicological testing with marine mollusks. In “Ecotoxicological Testing for the Marine Environment” (E. Jaspers and C. Clauss, eds), vol. 1,pp. 455477. State University of Ghent and Institute of Marine Scientific Research, Bredene, Belgium. Calabrese, A. and Davis, H. C. (1966). The pH tolerance of embryos and larvae of Mercenaria mercenaria and Crassostrea virginica. Biological Bulletin, Marine Biological Laboratory (Woods Hole) 131,427453. Calabrese, A. and Davis, H. C. (1967). Effects of soft “detergents” on embryos and larvae of the American oyster (Crassostrea virginica). Proceedings of the National ShellfisheriesAssociation 57, 1-16. Calabrese, A and Davis, H. C. (1970). Tolerance and requirements of embryos and larvae of bivalve molluscs. Helgolander Meeresuntersuchungen20,553-564. Calabrese, A. and Nelson, D. A. (1974). Inhibition of embryonic development of the hard clam, Mercenaria mercenaria, by heavy metals. Bulletin of Environmental Contamination and Toxicology 11, 92-97. Calabrese, A. and Rhodes, E. W. (1974). Culture of Mulinia lateralis and Crepidula fomicata embryos and larvae for studies of pollution effects. ThalassiaJugoslavica 10,89-102. Calabrese, A., Collier, R. S., Nelson, D. A. and MacInnnes, J. R. (1973). Toxicity of heavy metals to embryos of the American oyster, Crassostrea virginica. Marine Biology 18, 162-164. Calabrese, A., MacInnes, J. R., Nelson, D. A. and Miller, J. E. (1977a). Survival and growth of bivalve larvae under heavy-metal stress. Marine Biology 41, 179-184. Calabrese, A., Thurberg, P. F. and Gould, E. (1977b). Effects of cadmium, mercury, and silver on marine animals. Marine Fisheries Review 39(4), 5-11. Calabrese, A., Gould, E. and Thurberg, F. I? (1982). Effects of toxic metals in marine animals of the New York Bight: some laboratory observations. In “Ecological stress and the New York Bight: Science management” (G. E Mayer, ed.).
THE ASSESSMENT OF MARINE POLLUTION
145
Estuarine Research Federation, Columbia, S.C., 715 pp. Capuzzo, J. M. (1979). The effect of temperature on the toxicity of chlorinated cooling waters to marine animals. A preliminary review. Marine Pollution Bulletin 10,4547. Cardwell, R. D. (1976). Relative acute toxicity of a pesticide, heavy metal and anionic surfactants to marine organisms. Proceedings of the National Shelljisheries Association 66, 107. Cardwell, R. D. (1978). Oyster larvae mortality in south Puget Sound. Proceedings of the National Shellfisheries Association 68, 88-89 Cardwell, R. D., Woelke, C. E., Carr M. 1. and Sanborn, E. W. (1976). Sediment and elutriate toxicity to oyster larvae. In “Proceedings of the Special Conference on Dredging and its Environmental Effects”, (P. A. Krenkel, J. Harrissoni and J. C. Burdick, 111, eds). American Society of Civil Engineers, New York. Cardwell, R. D., Woelke, C. E., Carr, M. I. and Sanborn, E. W. (1977a). Evaluation of the efficacy of sulfite pulp mill pollution abatement using oyster larvae. Aquatic Toxicology and Hazard Evaluation, American Society for Testing and Materials, Philadelphia, PA, STP 634, 281-295. Cardwell, R. D., Woelke, C. E., Carr, M. I. and Sanborn, E. W. (1977b). Appraisal of a reference toxicant for estimating the quality of oyster larvae. Bulletin of Environmental Contamination and Toxicology 18(6), 719-725. Cardwell, R. D., Woelke, C. E., Carr, M. I. and Sanborn, E. W. (1978). Variation in toxicity tests of bivalve molluscs larvae as a function of termination technique. Bulletin of Environmental Contamination and Toxicology 20, 128-134. Cardwell, R. D., Olsen, S., Carr, M. I. and Sanborn, E. W. (1979a). Causes of oyster larvae mortality in South Puget Sound. NOAA Technical Memorandum ERL MESA-39, NOAA Environmental Research Laboratories, 73 pp. Cardwell, R. D., Woelke, C. E., Carr, M. I. and Sanborn, E. W. (1979b). Toxic substance and water quality effects on larval marine organisms. Washington Department of Fisheries, Technical Report 45, 1-71. Carlson, B. K. (1982). Settlement and subsequent survival of commercially-reared eyed-pediveliger larvae of the Pacific oyster Crassostrea gigas (Thunherg). Journal of Shellfish Research 2, 116 (abstract). Carr, M. L. (1975). Bivalve embryo water quality monitoring bioassays of Olympia Harbor dredge spoil disposal. In “Evaluations of Effects of Channel Maintenance Dredging and Disposal on the Marine Environment in Southern J’uget Sound, Washington”, Washington Department Fisheries, Technical Report 15, 23-40. Carr, R. S. (1998). Marine and estuarine porewater toxicity testing. In “Microscale Testing and Aquatic Toxicology Advances, Techniques and Practice” (I? G. Wells, K. Lee and C. Blaise, eds), pp. 523-537. CRC Press, Boca Raton, FL. Carr, R. S., Williams, J. W. and Fragata, C. T. B. (1989). Development and evaluation of a novel marine sediment pore water toxicity test with polyclhaete Dinophilus gyrociliatus. Environmental Toxicology and Chemistry 8, 533-543. Carr, R. S., Chapman, D. C., Presley, B. J., Biedenbach, J., Robertson, L., Boothe, P. N., Kilada, R., Wade, T. L. and Montagna, I? (1996). Sediment porewater toxicity assessment studies in the vicinity of offshore oil and gas production platforms in the Gulf of Mexico. Canadian Journal of Fisheries and Aquatic Sciences 53, 2618-2628. Carriker, M. (1961). Interrelation of functional morphology, behavior, and autoecology in early stages of the bivalve Mercenaria mercenaria. Journal Elisha Mitchell Scientific Society 77, 168-241.
146
E. HIS, R. BEIRAS AND
M. N. L. SEAMAN
Castagna, M., Gibbons, M. C. and Goodsell, J. (1985). A rapid method to induce spawning by injection of serotonin solution. Conseil International pour I'Exploration de la Mer, Mariculture Committee, C.M. 1985/F:61, 8 pp. Chang, Y., Coon, S. L., Walch, M. and Weiner, R. (1996). Effects of Hypomonas PM-1 biofilms on the toxicity of copper and zinc to Crassostrea gigas and Crassostrea virginica larval set. Journal of Shellfish Research 15, 589-595. Chanley, F? E. (1954). Possible causes of growth variations in clam larvae. Proceedings of the National Shellfisheries Association 45, 84-94. Chao, N.-H., Lin, T.-T., Chen, Y.-J., Hsu, H.-W. and Liao, I.-C. (1997). Cryopreservation of late embryos and early larvae in the oyster and hard clam. Aquaculture 155, 31-44. Chapman, P. M. (1990). The sediment quality triad approach to determining pollution-induced degradation. Science of the Total Environment 97/98, 815825. Chapman, P. M. and Long, E. R. (1983). The use of bioassays as part of a comprehensive approach to marine pollution assessment. Marine Pollution Bulletin 14, 81-84. Chapman, P. M. and Morgan, J. D. (1983). Sediment bioassays with oyster larvae. Bulletin of Environmental Contamination and Toxicology 31, 438-444. Chapman, P. M., Dexter, R. N. and Long, E. R. (1987). Synoptic measures of sediment contamination, toxicity and infaunal community composition (the Sediment Quality Triad) in San Francisco Bay. Marine Ecology Progress Series 37, 75-96. Chapman, P. M., Power, E. A., Dexter, R. N. and Andersen, H. B. (1991). Evaluation of effects associated with an oil platform, using the sediment quality triad. Environmental Toxicology and Chemistry 10,407424. Chapman, P. M., Swartz, R. C., Roddie, B., Phelps, H. L., Van den Hurk, P. and Butler, R. (1992). An international comparison of sediment toxicity tests in the North Sea. Marine Ecology Progress Series 91, 253-264. Chapman, P. M., Arthur, A. D., Paine, M. D. and Taylor, L. A. (1993). Sediment studies provide key information on the need to treat sewage discharged to sea by a major Canadian city. Water Science and Technology 28(8-9), 255-261. Chapman, P. M., Paine, M. D., Arthur, A. D. and Taylor, L. A.(1996). A triad study of sediment quality associated with a major, relatively untreated marine sewage discharge. Marine Pollution Bulletin 32, 47-64. Chen, T. T., Roesijadi, G., Olmi, E. J., Hens, B., Hill, P. and Sanders, J. G. (eds) (1995). Effects of trace metals and organic pollutants on stress-induced proteins in oyster larvae and spat: A molecular approach. 1994 Workshop Report Chesapeake Bay Environmental Effects Studies, Toxic Research Programme, VSG-95-10, 184 pp. Cherr, G. N., Shoffner-McGee, J. and Shenker, J. N. (1990). Methods for assessing fertilization and embryonic larval development in toxicity tests using the California mussel (Mytilus californianus). Environmental Toxicology and Chemistry 9, 1137-1145. Chesnut, A., Fahy, W. and Porter, H. (1957). Growth of Venus mercenaria, Venus campechiensis and their hybrids. Proceedings of the National Shellfisheries Association 47, 50-56. Chew, K. K. (1990). Global bivalve shellfish introductions: Implications for sustaining a f i s h e j or strong potential for economic gains? World Aquaculture 21. 9-22. Chien, Y.-H. and Chou, Y.-H. (1989). The effects of residual chlorine, tempera-
THE ASSESSMENT OF MARINE
POLLUTION
147
tures and salinities on the development of Pacific oyster (Crassostrea gi,gas). Asian Fisheries Science 3, 69-84. Chipperfield, P. (1953). Observations on the breeding and settlement of Mytilus edulis L. in British waters. Journal of the Marine Biological Association o,f the United Kingdom 32, 449-476. Chretiennot-Dinet M. J., Robert, R. and His, E. (1986). Utilisation des “alguesfourrage” en aquaculture. Annke biologique 25, 97-119. Chu, L. E., Hepworth, K. L. and Roberts, D. (1982). The acceptability and digestibility of microcapsules by larvae of Crassostrea virginica. Journd of Shellfish Research 2, 29-34. Chu, F. L., Webb, K. L., Hepworth, D. A. and Beverly, B. (1987). Metamorplhosis of larvae of Crassostrea virginica fed microencapsulated diets. Aquaculture 64, 185-1 97. Chung, K. S. (1980). Acute toxicity of selected heavy metals to mangrove oyster, Crassostrea rizophorae. Bulletin of the Japanese Society of Scientific Fisheries 46, 777-780. Cleary, J. J., McFadzen, I. R. B. and Peters, L. D. (1993). Surface micro’layer contamination and toxicity in the North Sea and Plymouth nearshore waters. International Council for the Exploration of the Sea, Marine Environmental Quality Committee, ICES-CM-1993E: 28. Coe, W. R. (1932a). Sexual phases in the American oyster (Ostrea virgirrica). Biological Bulletin, Marine Biological Laboratory, Woods Hole 63, 4 1 9 4 1. Coe, W. R. (1932b). Histological basis on sex changes in the American oyster (Ostrea virginica). Science 76, 125-127. Coe, W. R. (1943). Sexual differentiation in mollusks. I. Pelecypods. Quarterly Review of Biology 18, 154164. Coe, W. R. (1945). Development of the reproductive system and variations in sexuality in Pecten and other pelecypod molluscs. Transactionsof the Connecticut Academy of Arts and Sciences 36, 673-700. Coglianese, M. P. (1982). The effects of salinity on copper and silver toxicity to embryos of the Pacific oyster. Archives of Environmental Contamination and Toxicology 11, 297-303. Coglianese, M. P. and Martin, M. (1981). Individual and interactive effects of environmental stress on the embryonic development of the Pacific Oyster, Crassostrea gigas. I. The toxicity of copper and silver. Marine Environmental Research 5, 13-27. Cole, H. A. (1937). Experiments in the breeding of oysters (Ostrea edulis) in t.anks, with special reference to the food of the larva and spat. Fishery Znvestigai’ions, series 2 , London 15(4), 1-28. Cole, H. A. (1939) Further experiments on the breeding of oysters (Ostrea edulis) in tanks. Fishery Investigations London, series 2, 16(4), 1-47. Cole, H. A. (1941). The fecundity of Ostrea edulis. Journal of the Marine Biological Association of the United Kingdom 25, 243-260. Cole, H. A. (1942). Primary sex phases in Ostrea edulis. Quarterly Journal of Microscopical Science 83, 317-356. Connell, L., Welling, K. A. and Cattolico, R. A. (1997). Algal organic metabolites affect survival of Pacific oyster, Crassostrea gigas larvae. Journal of Shellfish Research 16(2), 493-502. Connor, P. M. (1972). Acute toxicity of heavy metals to some marine larvae. Marine Pollution Bulletin 3, 190-192. Coon, S. L. and Bonar, D. B. (1986). Norepinephrine and dopamine content of
148
E. HIS. R. BEIRAS AND M . N. L. SEAMAN
larvae and spat of the Pacific oyster, Crassostrea gigas. Biological Bulletin, Marine Biological Laboratory, Woods Hole 171, 632-639. Coon, S. L., Bonar, D. B. and Weiner, R. M. (1985). Induction of settlement and metamorphosis of the Pacific oyster, Crassostrea gigas (Thunberg) by L-DOPA and catecholamines. Journal of Experimental Marine Biology and Ecology 94, 211-221. Coon, S. L., Bonnar, D. B. and Weiner, R. M. (1986). Chemical production of cultchless oyster spat using epinephrine and norepinephrine. Aquaculture 58, 255-262. Coon, S. L., Walch, M., Fitt, W. K., Weiner, R. M. and Bonnar, D. B. (1990). Ammonia induces settlement behavior in oyster larvae. Biological Bulletin, Marine Biological Laboratory, Woods Hole 179(3), 297-303. Courtright, R., Breese, W. l? and Krueger, H. (1971). Formulation of a synthetic seawater for bioassays with Mytihs edulis embryos. Water Research 5, 877888. Cranfield, H. J. (1973). Observations on the behaviour of the pediveliger of Ostrea edulis during attachment and cementing. Marine Biology 22, 203-209. Crecelius, E. A. (1979). Measurements of oxidants in ozonized seawater and some biological reactions. Journal of the Fisheries Research Board of Canada 36, 1006-1008. Crisp, D. J., Yule, A. B. and White, K. N. (1985). Feeding by oyster larvae - the functional response, energy budget and a comparison with mussel larvae. Journal of the Marine Biological Association of the United Kingdom 65, 759-784. Cunningham, I? A. (1979). The use of bivalve molluscs in heavy metal pollution research. In “Marine Pollution: functional responses” (W. B. Vernberg, F. l? Thurberg, A. Calabrese and F. J. Vernberg, eds), pp. 183-221. Academic Press, New York. Dalton, R. and Menzel, W. (1983). Seasonal gonadal development of young laboratory-spawned southern (Mercenaria campechiensis) and northern (Mercenaria mercenaria) quahogs and their reciprocal hybrids in northwest Florida. Journal of Shelljish Research 3, 11-17. Daniels, S. A., Munawar, M. and Mayfield, C. I. (1989). An improved elutriation technique for the bioassessment of sediment contaminants. Hydrobiologia 1881189, 619-631. Dannevig, A. (1951). Lobster and oyster in Norway. Rapports et Proc2s-Verbaux du Conseil Permanent International pour I’Exploration de la Mer 128,92-96. Davies, A. G. (1978). Pollution studies with marine plankton. Advances in Marine Biology 15,381-508. Davis, H. C. (1953). On food and feeding of larvae of the American oyster, Crassostrea virginica. Biological Bulletin, Marine Biological Laboratory, Woods Hole 104,334-350. Davis, H. C. (1958). Survival and growth of clam and oyster larvae at different salinities. Biological Bulletin, Marine Biological Laboratory, Woods Hole U4, 296-307. Davis, H. C. (1960). Effects of turbidity-producing material in sea water on eggs and larvae of the clam (Venus (Mercenaria) mercenaria). Biological Bulletin, Marine Biological Laboratory, Woods Hole 118, 48-54. Davis, H. C. (1961). Effects of some pesticides on eggs and larvae of oysters (Crassostrea virginica) and clams (Venus mercenaria). Commercial Fisheries Review 23. 8-23.
THE ASSESSMENT OF MARINE POLLUTION
149
Davis, K. C. and Ansell, A. D. (1962). Survival and growth of larvae of the European oyster, 0. edulis at lowered salinities. Biological Bulletin, Marine Biological Laboratory, Woods Hole 122, 32-39. Davis, H. C. and Calabrese, A. (1964). Combined effects of temperature and salinity on development of eggs and growth of larvae of Mercenaria rnercrnaria and Crassostrea virginica. Fishery Bulletin US. 63, 643-655. Davis, H. C. and Calabrese, A. (1969). Survival and growth of larvae of the European oyster (Ostrea edulis L.) at different temperatures. Biological Bulletin, Marine Biological Laboratory, Woods Hole 136, 193-199. Davis, H. C. and Chanley, P. E. (1955a). Spawning and egg production of oysters and clams. Proceedings of the Nationai Shellfisheries Association 46, 40-58. Davis, H. C. and Chanley, P. E. (1955b). Effects of some dissolved substances on bivalve larvae. Proceedings of the National Shelljisheries Association 46, 59-74. Davis, H. C. and Guillard, R. R. (1958). Relative value of ten genera of micro-organisms as food for oyster and clam larvae. Fishery Bulletin US. 58, 293-304. Davis, H. C. and Hidu, H. (1969a). Effects of pesticides on embryonic development of clams and oysters and on survival and growth of the larvae. Fishery Bulletin US. 67, 393404. Davis, H. C. and Hidu, H. (1969b). Effects of turbidity-producing substances in sea water on eggs and larvae of three genera of bivalve mollusks. The Veliger 11, 316-323. DeGroot, A. J. (1973). Occurrence and behavior of heavy metals in river $deltas with special reference to the Rhine and Ems rivers. In NATO North Sea Science Conference (E. D. Goldberg, ed.), pp. 219-244. MIT Press, Cambridge, h4A. Dekshenieks, M. M., Hofmann, E. E. and Powell, E. N. (1993). Environmental effects on the growth and development of eastern oyster, Crassostrea virginica (Gmelin, 1791), larvae: a modeling study. Journal of Shellfish Research 12, 241-254. Deslous-Paoli, J.-M. (1982). ToxicitC des ClCments mCtalliques dissous pour les larves d’organismes marins: donnCes bibliographiques. Revue des Travau de l’lnstitut des Pgches Maritirnes 45, 73-84. Diaz, R. J. (1973). Effects of brief temperature increases on larvae of the American oyster (Crassostrea virginica). Journal of the Fisheries Research Board of C,unada 30(7), 991-993. Dimick, R. E. and Breese, W. P. (1965). Bay mussel embryo bioassay. Proceedings of the 12th Pacific Northwest Industrial Waste Conference, pp 165-175. Department of Civil Engineering, College of Engineering, University of Washington. Dinamani, P. (1971). Occurrence of the Japanese oyster, Crassostrea gigas (Thunberg), in Northland, New Zealand. New Zealand Journal of Marine and Freshwater Research 5, 352-357. Dinamani, P. (1987). Gametogenetic patterns in populations of Pacific oyster, Crassostrea gigas, in Northland, New Zealand. Aquaculture 64, 65-76. Dinnel, P. A. (1995). Evolution and present status of sea urchin sperm test. Soviet Journal of Marine Biology 21, 344-350. Dinnel, P. A., Stober, Q. J., Crumley, C. and Nakatani, R. E. (1982). Development of a sperm cell toxicity test for marine waters. In “Aquatic Toxicology and Hazard Assessment: 5th Conference” (J. G. Pearson, R. B. Foster, W. E. Bishop, eds), ASTM STP 766, pp. 82-98. American Society for Testing and Materials, Philadelphia.
150
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Dinnel, €! A., Link, J. M. and Strober, A. J. (1987). Improved methodology for a sea urchin sperm cell bioassay for marine water. Archives of Environmental Contamination and Toxicology 16,23-32. Di Toro, D. M., Mahony, J. D., Hansen, D. J., Scott, K. J., Hicks, M. B., Mayr, S. M. and Redmond, M. S. (1990). Toxicity of cadmium in sediments: the role of acid volatile sulfide. Environmental Toxicology and Chemistry 9, 1487-1502. Dixon, D. R. and Flavell, N. (1986). A comparative study of the chromosomes of Mytilus edulis and Mytilus galloprovincialis. Journal of the Marine Biological Association of the United Kingdom 66, 219-228. Dixon, D. R. and Prosser, H. (1986). An investigation of the genotoxic effect of organotin antifouling compound (bis tributyl-tin oxide) on the chromosomes of Mytilus edulis. Aquatic Toxicology 8, 185-195. Dos Santos, A. E. and Nascimento, I. A. (1985). Influence of gamete density, salinity and temperature on the normal embryonic development of the mangrove oyster Crassostrea rizophorae Guilding, 1928. Aquaculture 47, 335352. Douillet, P. A. (1993). Carbon contribution through bacterivory in larvae of the Pacific oyster Crassostrea gigas. Marine Ecology Progress Series 102, 303-314. Eertman, R. H. M., Groenink-van Emstede, M. and Sandee, B. (1993). The effects of the polycyclic aromatic hydrocarbons fluoranthene and benzo(a)pyrene on the mussel Mytilus edulis, the amphipod Bathyporeia sarsi and larva of the oyster, Crassostrea gigas. Netherlands Institute of Ecology, Center for Estuarine and Coastal Ecology, Yerseke, Rijkswaterstaat, Tidal Waters Division. BEONEffecten, DGW-NIOO, 50 pp. Elston, R. (1980). Functional anatomy, histology and ultrastructure of the soft tissues of the larval American oyster, Crassostrea virginica. Proceedings of the National Shellfisheries Association 70, 65-93. Emmens, C. W. (1948). “Principles of biological assays”. Chapman and Hall, London, 205 pp. Erdmann, W. (1935). Untersuchungen iiber die Lebensgeschichte der Auster. No 5. Uber die Entwicklung und die Anatomie der aansatzreifen, Larve von Ostrea edulis mit Bemerkungen iiber die Lebensgeschichte der Auster. Wissenschaftliche Meeresuntersuchungen,new series 19(6), 25 pp. Eversole, A. G. (1989). Gametogenesis and spawning in North American clam populations: implications for culture. In “Clam Mariculture in North America” (J. J. Manzi and M. Castagna, eds), Elsevier, N.Y., 19, 75-110. Eversole, A. G., Michener, W. K. and Eldridge, P. J. (1980). Reproductive cycle of Mercenaria mercenaria in a South Carolina estuary. Proceedings of the National Shellfisheries Association 70, 22-30. Eyster, L. S. and Morse, M. I? (1984). Development of the surf clam, (Spisula solidissima) following exposure of gametes, embryos, and larvae to silver. Archives of Environmental Contamination and Toxicology 13, 641-646. Falk-Petersen, I. B. and Lonning, S. (1984). Effects of hydrocarbons on eggs and larvae of marine organisms. In “Ecotoxicological Testing for the Marine Environment” (E. Jaspers and C. Claws, eds), vol. 2, pp. 197-217. State University of Ghent and Institute of Marine Scientific Research, Bredene, Belgium. Fichet, D., Radenac, G. and Miramand, P. (1998). Experimental studies of impacts of harbour sediments resuspension to marine invertebrates larvae: bioavailability of cadmium, copper, lead and zinc and toxicity. Marine Pollution Bulletin 36, 509-518.
THE ASSESSMENT OF MARINE POLLUTION
151
Filic, 2. and Krajnovic-Ozretic, M. (1978). Sur la presence de I’huitre Crassostrea gigas (Thunberg) dans le canal du Lim (Adriatique Septentrionale). Ichthyologia 10,41-54. Finney, D. J. (1964). “Statistical Methods in Biological Assay, 2nd edn”. Griffin, London. Finney, D. J. (1971). “Probit analysis, 3rd edn”. p. 1-333. Cambridge University Press, Cambridge, Great Britain. Fitt, W. K., Coon, S. L., Walch, M., Weiner, R. M., Colwell, R. R. and Bonar, D. B. (1990). Settlement behavior and metamorphosis of oyster larvae (Crassostrea gigas) in response to bacterial supernatants. Marine Biology 106(3), 389-2194. Forbes, T. L. (1993). The design and analysis of concentration-response experiments. In “Handbook of Ecotoxicology”, 1 (F! Calow, ed.), pp. 438-459, Blackwell, Oxford. Franc, A. (1960). Classe des Bivalves. In “Trait6 de Zoologie” (P. GrassC, ed.), vol. V(II), Masson Publisher, Paris, pp. 845-2133. Fretter, V. and Graham, A. (1964). Reproduction. In “Physiology of Mollusca” (K. M. Wilbur and C. M. Yonge, eds), vol. 1, pp. 127-156. Academic ]Press, London. Fujiya, M. (1970). Oyster farming in Japan. Helgolander Wissenschajiliche Meeresuntersuchungen 20, 464-479. Galtsoff, P. S. (1930). The fecundity of the oyster. Science 72, 97-98. Galtsoff, P. S. (1938a). Physiology of reproduction of Ostrea virginica. I. Biological Bulletin, Marine Biological Laboratory, Woods Hole 74(3), 461436. Galtsoff, l? S. (1938b). Physiology of reproduction of Ostrea virginica. 11. Biological Bulletin, Marine Biological Laboratory, Woods Hole 75(2), 286-307. GaltsoK P. S. (1940). Physiology of reproduction of Ostrea virginica. 111. Biological Bulletin, Marine Biological Laboratory, Woods Hole 78(l), 117-135. Galtsoff, P. S. (1964). The American oyster, Crassostrea virginica Gmelin. Fishery Bulletin US.64, 480 pp. Gardner, J. P. A. (1997). Hydridization in the sea. Advances in Marine Biology 31, 1-78. Garland, C. D., Cooke, S. L., McMeekin, T. A. and Valentine, J. E. (1986). E,ffects of 0.2 pm membrane-filtered seawater as a culture medium on fertilizedl eggs and larvae of the Pacific oyster, Crassostrea gigas. Australian Journal of Marine and Freshwater Research 37, 713-720. Geffard, 0. (1997). Mise en evidence de la toxicite potentielle des sCdiiments marins a l’aide des embryons et des larves de Crassostrea gigas et de Paracentrotus lividus. D.E.A. Toxicologie de l’Environnement, UniversitC de Metz-Rouen-Strasbourg, France, 48 pp. Gelber, R. D., Lavin, P. T., Mehta, C. R. and Schoenfeld, D. A. (1985). Statistical analysis. In “Fundamentals of aquatic toxicology” (G. M. Rand and S. R. Petrocelli, eds), pp. 11C123. McGraw-Hill, Auckland. Gentien, l?, Arzul, G. and Toularastel, F. (1991). Modes of action of the toxic principle of Gyrodinium cf. Aureolum. In “Actes du Colloque sur les Biotoxines Marines”. (J. M. Fremy, ed.), pp. 83-86. Centre National d’Etudes VktCrinaires et Alimentaires Publisher. Gerdes, D. (1983a). The Pacific oyster, Crassostrea gigas. Part I. Feeding behaviour of larvae and adults. Aquaculture 31, 195-219. Gerdes, D. (1983b). The Pacific oyster, Crassostrea gigas. Part 11. Oxygen consumption of larvae and adults. Aquaculture 31,221-231. Gibbons, M. C. and Castagna, M. (1985). Responses of the hard clam, Mercenaria
152
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
mercenaria (LinnC) to induction of spawning by serotonin. Journal of Shellfish Research 5, 65-67. Gibbs, P. E. and Bryan, W. G. (1996). Reproductive failure in the gastropod Nucella lapillus associated with imposex caused by tributyltin pollution: a review. Zn “Organotin, environmental fate and effects” (M. Champ and F! E Seligman, eds), pp. 259-280. Chapman and Hall, London. Glickstein, N. (1978). Acute toxicity of mercury and selenium to Crassostrea gigas embryos and Cancer magister larvae. Marine Biology 49,113-117. Gormly, N. M., Singer, S. and Genthner, F. J. (1996). Nontarget testing of microbial pest control agents using larvae of the coot clam, Mulinia lateralis. Diseases of Aquatic Organisms 26, 229-235. Gosling, E. M. (1981). Genetic variability in hatchery-produced Pacific oyster (Crassostrea gigus Thunberg). Aquaculture 26, 273-287. Gosling, E. M. (1984). The systematic status of Mytilus galloprovincialis in western Europe: a review. Malacologia 25, 551-568. Gosling, E. M. (1992). “The mussel Mytilus: ecology, physiology, genetics and culture”. Elsevier, Amsterdam, 589 pp. Granmo, A. (1972). Development and growth of eggs and larvae of Mytilus edulis exposed to a linear dodecylbenzenesulphonate, LAS. Marine Biology 15, 356-358. Granmo, A. and Jorgensen, G. (1975). Effects on fertilization and development of the common mussel, Mytilus edulis, after long-term exposure to a non-ionic surfactant. Marine Biology 33, 17-20. Granmo, A., Havenhand, J., Magnusson, K. and Svane, I. (1988). Effects of the planktonic flagellate Chrysomulinapolylepis Manton et Park on fertilization and early development of the ascidian Ciona intestinalis (L.) and the blue mussel, Mytilus edulis L. Journal of Experimental Marine Biology and Ecology 124, 65-71. Granmo, A,, Ekelund, R., Magnusson, K. and Berggren, M. (1989). Lethal and sublethal toxicity of 4-nonylphenol to the common mussel, (Mytilus edulis L.). Environmental Pollution 59, 115-127. Hadfield, M. G. (1984). Settlement requirements of molluscan larvae: new data on chemical and genetic roles. Aquaculture 39,283-298. Hall, L.W. Jr, Ziegenfuss, M. C., Anderson, R. D. and Killen, W. D. Jr. (1995). Use of estuarine water column tests for detecting toxic conditions in ambient areas of the Chesapeake Bay watershed. Environmental Toxicology and Chemistry 14(2), 267-278. Hamilton, M. A., Russo, R. C. and Thurston, R. V. (1977). Trimmed SpearmanKarber method for estimating median lethal concentrations in toxicity bioassays. Environmental Science and Technology 11(7), 714-719. Hansen, B., Fotel, F. L., Jensen, N. J. and Wittrup, L. (1997). Physiological effects of the detergent linear alkylbenzene sulphonate on blue mussel larvae (Mytilus edulis) in laboratory and mesocosm experiments. Marine Biology 128, 627637. Hardy, J. T. and Cleary, J. (1992). Surface microlayer contamination and toxicity in the German Bight. Marine Ecology Progress Series 91,203-210. Haux, C. and Forlin, L. (1988). Biochemical methods for detecting effects of contaminants on fish. Ambio 17,376-380. Hayes, P. F. and Menzel, R. W, (1981). The reproductive cycle of early setting Crassostrea virginica (Gmelin) in the northern gulf of Mexico, and its implications for population recruitment. Biological Bulletin, Marine Biological
THE ASSESSMENT OF MARINE POLLUTION
153
Laboratory, Woods Hole 160, 8C88. Helm, M. M. (1971). The effects of seawater quality on the laboratory culture of Ostrea edulis L. larvae. Conseil International pour exploration de la mer, (C.M. 1971K) 28, 1-11. Helm, M. M. and Millican, F? F. (1977). Experiments in the hatchery rearing of Pacific oyster larvae (Crassostrea gigas Thunberg). Aquaculture 11, 1-12. Helm, M. M. and Spencer, B. E. (1972). The importance of the rate of aeration in hatchery cultures of the larvae of Ostrea edulis. Journal du Conseil 34, 244-255. Helm, M. M., Holland, D. L. and Stephenson, R. R. (1973). The effects of supplementary algal feeding of a hatchery breeding stock of Ostrea edulis L on larval vigour. Journal of the Marine Biological Association of the United Kingdom 53, 673-684. Helm, M. M., Hepper, B. T., Spencer, B. E. and Walne, F? R. (1974). Lugworm mortalities and bloom of Gyrodinium aureolum Hulburt in the Eastern Irish Sea, Autumn 1971. Journal of the Marine Biological Association of the United Kingdom 54, 857-869. Hennes-Morgan, E. C. and Oude, N. T. (1994). Detergents. In “Handbook of Ecotoxicology”, 2, pp. 130-154. Blackwell, Oxford. Hidu, H. (1965). Effects of synthetic surfactants on the larvae of clams (Mercenaria. mercenaria) and oysters (Crassostrea virginica). Journal Water Pollution Control Federation 31, 262-270. Hidu, H. and Haskin, H. H. (1971). Settling of the American oyster related to environmental factors and larval behavior. Proceedings of the Natr onal Shellfisheries Association 61, 35-50. Hidu, H. and Tubiash, H. S. (1963). A bacterial basis for the growtlh of antibiotic-treated bivalve larvae. Proceedings of the National ShellJisheries Association 54, 25-39. Hidu, H., Roosenburg, W. H., Drobeck, K. G., McErlean, A. J. and Mihurski, J. A. (1974). Thermal tolerance of oyster larvae, Crassostrea virginicu Gmelin, as related to power plant operation. Proceedings of the National ShellJisheries Association 64, 102-110. His, E. (1975). La dCtection des pontes dans le milieu naturel: application de I’ostrCographie B l’ttude de la reproduction des huitres. Haliotis 5 , 206-21 5 . His, E. (1976). Contribution 2 1’Ctude biologique de I’huitre dans le Bassin d’Arcachon. Activitk valvaire de Crassostrea gigas, application B l’ttude tie la reproduction de l’huitre japonaise. Thbse 3e cycle, Bordeaux I, 63 pp. His, E. (1992). Embryogenesis and larval development in Crassostrea gigas: experimental data and field observations on the effects of tributyltin compounds. In “Organotin, Environmental fate and effects” (M. Champ and F? F. Seligman eds.), pp. 239-258. Chapman and Hall, London. His, E. and Beiras, R. (1995). Monitoring fresh and brackish water quality around shellfish farming areas with a bivalve embryo and larva simplified bioassay method. Oceanologica Acta 18, 591-595. His, E. and Maurer, D. (1988). Shell growth and gross chemical compositicn of oyster larvae (Crassostrea gigas) in the field. Aquaculture 69, 185-194. His, E. and Robert, R. (1980). Action d’un sel organo-mttallique, l’acttate de tributyle-Ctain, sur les oeufs et les larves D de Crassostrea gigas (Thunberg). Conseil International pour 1’Exploration de la Mer, Mariculture Committee, C.M. 1980E 27-10. His, E. and Robert, R. (1981). Effects of copper chloride on the eggs and D larvae
154
E. HIS, R. BEIRAS AND M . N. L. SEAMAN
of Crassostrea gigas (Thunberg). Preliminary results. Conseil International pour 1’Exploration de la Mer, Mariculture Committee, C.M. 1981E 43, 14 pp. His, E. and Robert, R. (1982). Le danger des traitements par le sulfate de cuivre en zone conchylicole: toxicite vis-8-vis des oeufs et des jeunes larves de C. gigas. Revue des Travaux de l’lnstitut des PEches Maritimes 42(2), 117-125. His, E. and Robert, R. (1987a). L‘isolement des vkligkres de Crassostrea gigas du milieu naturel: un nouveau mode d’investigation sur la nutrition des larves d’huitres. Haliotis 16, 573-575. His, E. and Robert, R. (1987b). Comparative effects of two antifouling paints on the oyster Crassostrea gigas. Marine Biology 95(1), 83-86. His, E. and Seaman, M. (1992). Effects of temporary starvation on the survival, and subsequent feeding and growth of oyster (Crassostrea gigas) larvae. Marine Biology 114,277-279. His, E. and Seaman, M. (1993). Effects of twelve pesticides on larvae of oysters (Crassostrea gigus) and two species of unicellular marine algae (Zsochrysis galbana and Chaetoceros calcitrans). Conseil International pour 1’Exploration de la Mer, Marine Environmental Quality Committee, CM 1993E22, 8 pp. His, E., Maurer, D. and Robert, R. (1983). Estimation de la teneur en ac6tate de tributyleCtain dans l’eau de mer par une mCthode biologique. Journal of Molluscan Studies l2A, 60-68. His, E., Robert, R. and Chretiennot-Dinet, M. J. (1985). Nouvelle mCthode pour Ctudier la nutrition de jeunes larves de Crassostrea gigas (nunberg) en milieu naturel. Premibres donnCes exp6rimentales. Comptes Rendus de 1’Acadimie des Sciences Paris 300,sCrie 111 no. 8, 319-321. His, E., Robert, R. and Dinet, A. (1989). Combined effects of temperature and salinity on fed and starved larvae of the Mediterranean mussel Mytilus galloprovincialis and the Japanese oyster Crassostrea gigas. Marine Biology 100, 455463. His, E., Beiras, R., Seaman, M. N. L., Trieff, N. M. and Pagano, G. (1996). Sublethal and lethal toxicity of aluminium industry effluent to early development stages of the Crassostrea gigas oyster. Archives of Environmental Contamination and Toxicology 30, 335-339. His, E., Seaman, M. N. L. and Beiras, R. (1997a). A simplification of the bivalve embryogenesis larval development bioassay method for water quality assessment. Water Research 31, 351-355. His, E., Budzinski, H., Geffard, 0. and Beiras, R. (1997b). Effects of a hydrocarbon-polluted sediment on Crassostrea gigas (Thunberg) metamorphosis. Comptes Rendus de I’Acadtmie des Sciences Paris, Strie Sciences de la Vie 320, 797-803. His, E., Heyvang, I., Geffard, 0. and de Montaudouin, X. (1999). A comparison between oyster (Crassostrea gigas) and sea urchin (Paracentrotus lividus) larval bioassays for toxicological studies. Water Research 33, 1706-1718. Ho, K. T. Y and Quinn, J. G. (1993). Physical and chemical extraction and fractionation that influence toxicity, as evaluated by microtox. Environmental Toxicology and Chemistry 12, 615-625. Ho, M.-S. and Zubkoff, l? L. (1979). The effects of Cochlodinium heterolobaturn bloom on the survival and calcium uptake by larvae of the American oyster, Crassostrea virginica. In “Toxic Dinoflagellate Blooms. Proceedings of the International Conference (2nd)” (D. L. Taylor and H. H. Seliger, eds), Elsevier, North Holland, 409412. Ho, M.-S. and Zubkoff, l? L. (1980). The calcium uptake by larvae of the oyster,
THE ASSESSMENT OF MARINE POLLUTION
155
Crassostrea virginica and the clam, Mulinia lateralis. Comparative Biochernistry and Physiology 65A,143-146. Ho, M.-S. and Zubkoff, P. L. (1982). The effects of mercury, copper, and zinc on calcium uptake by larvae of the clam, Mulinia lateralis. Water, Air and’ Soil Pollution 17,409-414. Ho, M.-S. and Zubkoff, P. L. (1983). The opposing effects of cadmium and nickel on calcium uptake by larvae of the clam Mulinia lateralis. Comparative Biochemistry and Physiology 74C, 337-339. Hoare, K., Davenport, J. and Beaumont, A. R. (1995a). Effects of exposure and previous exposure to copper on growth of veliger larvae and survivorship of Mytilus edulis juveniles. Marine Ecology Progress Series 120, 163-168. Hoare, K., Beaumont, A. R. and Davenport, J. (1995b). Variation arnong populations in the resistance of Mytilus edulis embryos to copper: adaptation to pollution? Marine Ecology Progress Series 120, 155-162. Hoegh-Guldberg, 0. and Manahan, D. T. (1995). Coulometric measurement of oxygen consumption during development of marine invertebrate embryos and larvae. Journal of Experimental Biology 198, 19-30. Holdgate, M. W. (1979). “A perspective of environmental pollution”. Cambridge University Press, Cambridge. Hrs-Brenko, M. (1977). The relationship of temperature and salinity to larval development in mussel (Mytilus galloprovincialis Lamark). In “Physiology and Behaviour of Marine Organisms, Proceedings of the 12th European Symposium on Marine Biology”, Stirling, Scotland, September 1977 (D. S. McLusky and A. J. Berry, eds). Pergamon Press, Oxford, pp. 359-365. Hrs-Brenko, M. (1981). The growth and survival of larvae of several bivalve species at high temperatures and the practicability of their culture in heated effluent waters. Ichthyologia 13, 29-37. Hrs-Brenko, M. (1982). Ostrea edulis Linnaeus and Crassostrea gigas (Thunberg) larvae in the plankton of Limski Kana1 in the Northern Adriatic Sea. .4cta Adriatica 23, 399407. Hrs-Brenko, M. and Calabrese, A. (1969). The combined effects of salinity and temperature on larvae of the mussel Mytilus edulis. Marine Biology 4, 224-226. Hrs-Brenko, M., Claus, C. and Bubic, S. (1977). Synergistic effects of lead, salinity and temperature on embryonic development of the mussel, Mytilus galloprovincialis. Marine Biology 44, 109-115. Huntington, K. M. and Miller, D. C. (1989). Effects of suspended sediment, hypoxia, and hyperoxia on larval Mercenaria mercenaria (Linnaeus, 1958). Journal of ShellJishResearch 8, 37-42. Imai, T. (1967). Mass production of molluscs by means of rearing the larvae. Tohoku Journal of Agricultural Research 1, 69-86. Imai, T. and Sakai, S. (1961). Study of breeding of Japanese oyster, Crassostrea gigas. Tohoku Journal of Agricultural Research 12, 125-171. Imai, T., Hatanaka, M., Sato, R., Sakai, S. and Yuki, R. (1950). Artificial breeding of oysters in tanks. Tohoku Journal of Agricultural Research 1,69-86. Imai, T., Sakai, S., Okada, H. and Yoshida, T. (1954). Breeding of the Olympia oyster in tanks. Tohoku Journal of Agricultural Research 12, 125-163. Inaba, K., Shiraishi, H. and Soma, Y. (1995). Effects of salinity, pH imd temperature on aqueous solubility of four organotin compounds. Water Research 29(5), 1415-1417. Ingle, R. M. (1951). Spawning and setting of oysters in relation to seasonal
156
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
environmental changes. Bulletin of Marine Science of the Gulf and Caribbean 1, 11-115. Iwata, K. S. (1950). Spawning of Mytilus edulis. 11. Discharge by electrical stimulation. Bulletin of the Japanese Society of Scientific Fisheries 15, 443-446. Iwata, K. S. (1951a). Spawning of Mytilus edulis. IV. Discharge by KC1 injection. Bulletin of the Japanese Society of Scientific Fisheries 16, 393-394. Iwata, K. S. (1951b). Spawning of Mytilus edulis. VII. Comparison of abilities of salts of alkali metals and of alkali earth metals to induce spawning. Bulletin of the Japanese Society of ScientiJc Fisheries 17, 94-95. Jensen, R. A., Morse, D. E., Petty, R. L. and Hooker, N. (1990). Artificial induction of larval metamorphosis by free fatty acids. Marine Ecology Progress Series 67(1), 55-72 Jespersen, H. and Olsen, K. (1982). Bioenergetics in veliger larvae of Mytilus edulis. Ophelia 21, 101-113. Johnson, D. (1988). Development of Mytilus edulis embryos. A bioassay for polluted waters. Marine Ecology Progress Series 46, 135-138. Karaseva, E. M. and Medvedeva, L. A. (1993). Morphological and functional changes in the offspring of Mytilus trossulus and Mizuhopecten yessoensis after parental exposure to copper and zinc. Russian Journal of Marine Biology 19(4), 276-280. Katkansky, S. C. and Sparks, A. K. (1966). Sexual patterns in the Pacific oyster Crassostrea gigas in Washington State. Fisheries Research Papers, Washington Department of Fisheries 2, 80-89. Keck, R. T., Maurer, D., Kauer, J. C. and Sheppard, W. A. (1971). Chemical stimulants affecting larval settlement in the American oyster. Proceedings of the National Shellfisheries Association 61, 24-28. Keck, R. T., Maurer, D. and Lind, H. (1975). A comparative study of the hard clam gonad developmental cycle. Biological Bulletin 148, 243-258. Kennedy, A. V. and Battle, H. I. (1964). Cyclic changes in the gonad of the American oyster, Crassostrea virginica (Gmelin). Canadian Journal of Zoology 42, 305-321. Kennedy, V. S. (1983). Sex ratios in oysters, emphasizing Crassostrea virginica from Chesapeake Bay, Maryland. The Veliger 25(4), 329-338. Kennedy, V. S. and Krantz, L. B. (1982). Comparative gametogenic and spawning patterns of the oyster Crassostrea virginica (Gmelin) in central Chesapeake Bay. Journal of Shellfih Research 2, 133-140. Kennedy, V. S. and Mihursky, J. A. (1971). Upper temperature tolerances of some estuarine bivalves. Chesapeake Science 12, 193-204. Kennedy, V. S., Rosenburg, W. H., Zion, H. H. and Castagna, M. (1974). Temperature-time relationship for survival of embryos and larvae of Mulinia lateralis (Mollusca Bivalvia). Marine Biology 24, 137-145. Kester, D. N., Duedall, I. W., Connors, D. N. and F'ytkowicz, R. M. (1967). Preparation of artificial seawater. Limnology and Oceanography 12, 176-179. Kingston, P. (1974). Some observations on the effects of temperature and salinity upon the growth of Cardium edule and Cardium glaucum in the laboratory. Journal of the Marine Biological Association of the UK 54, 309-317. Kitajima, F., Satuito, C. G., Hirota, H., Katsuyama, I. and Fusetani, N. (1995). A new screening method for antifouling substances against the young mussels Mytilus edulis galloprovincialis. Fisheries Science 61, 578-583. Klockner, K., Rosenthal, H. and Willfiihr, J. (1985). Invertebrate bioassays with North Sea water samples. I. Structural effects on embryos and larvae of
THE ASSESSMENT OF MARINE POLLUTION
157
serpulids, oysters and sea urchins. Helgolander WissenschaftlichenMeere.wntersuchungen 39, 1-19. Knaub, R. S. and Eversole, A. G. (1988). Reproduction of different stocks of Mercenaria mercenaria. Journal of Shel&sh Research 7 , 371-376. Knaub, R. S , Eversole, G. and Goodsell, J. G. (1987). Fecundity of different stocks of hard clams. Journal of the World Aquaculture Society 18, 11A. Knezovich, J. F!, Harrison, F. L. and Tbcker, J. S. (1981). The influence of organic chelators on the toxicity of copper to embryos of the Pacific oyster, Crassostrea gigas. Archives of Environmental Contamination and Toxicology 10, 241- 249. Knezovich, J. l?, Steichen, D. J., Jelinski, J. A. and Anderson, S. L. (1996). !sulfide tolerance of four marine species used to evaluate sediment and pore-water toxicity. Bulletin of Environmental Contamination and Toxicology 57, 450457. Knight-Jones, E. W. (1952). Reproduction of oysters in the rivers Crouch and Roach, Essex, during 1947, 1948 and 1949. Fishery Investigations, London 18, 48 PP. Kobayashi, N. (1984). Marine ecotoxicological testing with echinoderms. In “Ecotoxicological testing for the marine environment” (G. Personne, E. Jaspers and C. Clauss, eds), 1, pp. 341405. Laboratory of Biological Research in Aquatic Pollution, State University, Ghent. Koganazewa, A. (1976). The status of Pacific oyster culture in Japan. Food and Agricultural Organization Technical Conference on Aquaculture, Kyoto, Japan, FIR AQ/conf./76/E.69, pp. 1-9. Koganezawa, A. (1978). Ecological study of the production of seeds of the Pacific oyster, Crassostrea gigas. Bulletin of the Japanese Sea Regional Fisheries Research Laboratory 29, 1-88. Koke, E. 0. (1993). Contributions by the sediment to the Wadden Sea DOC budget. Ph.D. Thesis, University of Oldenburg, Germany, 140 pp. Konar, B. and Stephenson, M. D. (1995). Gradients of subsurface water toxi,city to oyster larvae in bays and harbors in California and their relation to mussel watch bioaccumulation data. Chemosphere 30, 165-172. Kon-Ya, K., Shimidzu, N., Adachi, K. and Miki, W. (1994). 2,5,6-Tribromo-lmethylgramine, an antifouling substance from the marine bryozoan Zoobotryon pellucidum. Fisheries Science 60(6), 773-775. Korringa, P. (1941). Experiments and observations on swarming, pelagic life and settling in the European flat oyster. Archives Nierlandaises de Zoologie 5, 1-249. Korringa, P. (1958). Water temperature and breeding throughout the geographical range of Ostrea edulis. Annie Biologique 33, 109-116. Krassoi, R. (1995). Salinity adjustment of effluents for use with marine bioaissays: effects on the larvae of the doughboy scallop Chlamys asperrimus and the Sydney Rock oyster Saccostrea commercialis. Australian Journal of Ecotoxicology 1, 143-148. Krassoi, R., Everett, D. and Anderson, I. (1996). Protocol for using doughboy scallop Chlamys asperrima (Mollusca: Pectinidae) L. to test the sublethal toxicity of single compounds and effluents. Technical Report Series 17, National Pulp Mills Research Program, Pulp and Paper Research Advisory Eloard, Canberra, Australia, 46 pp. Krassoi, R., Anderson, I. and Everett, D. (1997). Larval abnormalities in doughboy scallops Chlamys (Mimachlamys) asperrima 1. In response to test conditions and six reference toxicants. Australian Journal of Ecotoxicology 3, 65-74. Lamberson, J. O., DeWitt, T. T. and Swartz, R. C. (1992). Assessment of sediment
158
E. HIS,
R. BEIRAS AND M. N . L. SEAMAN
toxicity to marine benthos. In “Sediment Toxicity Assessment” (A. Burton Jr, ed.), pp. 183-211. Lewis Publishers, Boca Raton. Landner, L. (1984). Experimental procedures for hazard assessment in the marine environment. In “Ecotoxicological testing for the marine environment” (G. Personne, E. Jaspers and C. Clauss, eds) vol. 1, pp. 657-687. Laboratory of Biological Research and Aquatic Pollution, State University, Ghent. Landrum, l? F., Mihart, S. R., Eadie, B. J. and Herche, L. R. (1987). Reduction in bioavailability of organic contaminants to the amphipod Pontoporeia hoyi by dissolved organic matter of sediment interstitial waters. Environmental Toxicology and Chemistry 6, 11-20. Lannan, J. E. (1971). Experimental self-fertilization of the Pacific oyster, Crassostrea gigas, utilizing cryopreserved sperm. Genetics 68, 599-601. Lannan, J. E., Robinson, A. and Breese, W. F! (1980). Broodstock management of Crassostrea gigas. 11. Broodstock conditioning to maximize larval system. Aquaculture 21,337-345. Lapota, D., Rosenberg, D. E., Platter-Rieger, M. F. and Seligman, F! F. (1993). Growth and survival of Mytilus edulis larvae exposed to low levels of dibutyltin and tributyltin. Marine Biology 115,413-420. La Roche, G., Eider, R. and Tarzwell, C. M. (1970). Bioassay procedures for oil and oil dispersant toxicity evaluation. Journal of the Water Pollution Control Federation 42,1982-1989. Laughlin, R. B. Jr, Gustafson, R. and Pendoley, P. (1988). Chronic embryo-larval toxicity of tributyltin (TBT) to the hard shell clam Mercenaria mercenaria. Marine Ecology Progress Series 48, 29-36 Laughlin R. B. Jr, Gustafson, R. and Pendoley, F? (1989). Acute toxicity of tributyltin (TBT) to early life history stages of the hard clam, Mercenaria mercenaria. Bulletin of Environmental Contamination and Toxicology 42, 352358. Lee, G. F. and Jones, R. A. (1987). Water quality sigmficance of contaminants associated with sediments: an overview. In “Fate and Effects of Sediment Bound Chemicals in Aquatic Systems” (K. L. Dickson, A. W. Maki and W. A. Brungs, eds), pp. 3-34. Pergamon Press, New York. Le Gore, R. S. (1974). The effects of Alaska crude oil and selected hydrocarbon compounds on embryonic development of the Pacific oyster, Crassostrea gigas. Ph.D. thesis, University of Washington, Seattle, WA, 186 pp. Lemos, M. B. N., Nascimiento, I. A., de Araujo, M. M. S., Pereira, S. A., Bahia, I and Smith, D. H. (1994). The combined effects of salinity, temperature, antibiotic and aeration on larval growth and survival of the mangrove oyster, Crassostrea rizophorae. Journal of Shellfish Research 13,187-192. Le Pennec, M. (1978). Genkse de la coquille larvaire et postlarvaire chez divers bivalves marins. Thkse de Doctorat ks Sciences, Sene C no. 9, no. d’ordre 52,21 DCcembre 1978, FacultC des Sciences, Brest, France, 229 pp. + annexes. Le Pennec, M. (1981). Les mkthodes expkrimentales induisant la ponte chez les mollusques bivalves manins. Haliotis 11,139-155. Le Pennec, M. and Le Roux, S. (1979). Effets d’un petrole brut sur la formation de la coquille de Mytilus edulis (L.) (Mytilidae, Bivalvia). Revue Internationale d’ocdanographie Mddicale LV, 49-55. Le Pennec, M. and Masson, M. (1976). MorphogknBse de la coquille de Mytilus galloprovincialis (Lmk) ClevC au laboratoire. Cahiers de Biologie Marine 17, 113-118. Le Pennec, M. and Prieur, D. (1972). DCveloppement larvaire de Mytilus edulis
THE ASSESSMENT OF MARINE POLLUTION
159
(L.) en presence d’antibiotiques. lkre partie: dttermination des concentrations actives non toxiques de quatre antibiotiques, aurkomycine, trythromycine, chloramphenicol et sulfamkrazine. Revue Internationale d’Ockanographie Mkdicale XXVII, 167-179. Le Pennec, M., Prieur, D. and Chardy, P. (1973). Developpement larvaire de Mytilus edulis (L.) en prksence d’antibiotiques. 22me partie: action sur la croissance de quatre antibiotiques, aurkomycine, trythromycine, chloramphenicol et sulfamkrazine. Revue Znternationale d’Octanographie Mkdicale XXX, 115-137. Le Roux, S. (1977). The toxicity of pure hydrocarbon to mussel larvae. Rapports et Process-Verbauxdu Conseil Permanent International pour 1’Exploration de la Mer 171, 189-190. Le Roux, S. and Lucas, A. (1978). Technique d’etude des effets des hydrocarbures sur la physiologie des larves de moules. Revue Znternationale d’Ockanog,raphie Mkdicale XLM, 75-79. Livett, E. A. (1988). Geochemical monitoring of atmospheric heavy metal pollution: theory and applications.Advances in Ecologicaf Research 18,65-177. Lloyd, R. and Thain, J. E. (1981). The use of the oyster embryo bio-assay to monitor coastal water quality. Paper presented at the 1981 Meeting of the ICES Working Group on Marine Pollution Baseline and Monitoring Studies in the North Atlantic, 4 pp. Long, E. R. and Chapman, P. M. (1985). A sediment toxicity triad: measures of sediment contamination, toxicity, and infaunal community composition in Puget Sound. Marine Pollution Bulletin 16, 405-415. Long, H.R., Buchman, M. F., Bay, S. M., Bretler, D. J., Chapman, l? M., Hose:, J. E., Lissner, A. L., Scott, J. and Wolte D. A. (1990). Comparative evaluation of five toxicity tests with sediment from San Francisco Bay and Tomales Bay, California. Environmental Toxicology and Chemistry 9, 1193-1214. Long, E. R., Roberston, A., Wolfe, D. A., Hameedi, J. and Sloane, G. M. (1996). Estimates of the spatial extent of sediment toxicity in major U.S. estuaries. Environmental Science and Technology 30,3585-3592. Longo, F. J. and Anderson, E. (1969a). Cytological aspects of fertilization in the lamellibranch, Mytilus edulis. I. Polar body formation and development of the female pronucleus. Journal of Experimental Zoology 172, 69-96. Longo F. J. and Anderson, E. (1969b). Cytological aspects of fertilization in the lamellibranch, Mytilus edulis. 11. Development of the male pronucleus and the association of maternally and paternally derived chromosomes. Journal of Experimental Zoology 172, 97-119. Loosanoff, V. L. (1937a). Development of the primary and sexual phases in, Venus mercenaria (L.). Biological Bulletin, Marine Biological Laboratory, Woods Hole 72, 349405. Loosanof€, V. L. (1937b). Seasonal gonadal changes of adult clams, Venus mercenaria (L.). Biological Bulletin, Marine Biological Laboratory, Wood,yHole 72, 406-416. Loosanoff, V. L. (1942). Seasonal gonadal changes in the adult oyster, Ostrea virginica, of Long Island Sound. Biological Bulletin, Marine Biological Laboratory, Woods Hole 82, 195-206. Loosanoff, V. L. (1945). Precocious gonad development in oysters induced in midwinter by high temperature. Science, New York 102 No. 2640, 124-1;!5. Loosanoff, V. L. (1954). New advances in the study of bivalve larvae. American Scientist 42, 607-624.
160
E. HIS. R. BEIRAS AND M. N. L. SEAMAN
Loosanoff, V. L. (1955). The European oyster in American waters. Science, New York 121 no. 3135, 119-121. Loosanoff, V. L. (1965a). Gonad development and discharge of sperm in oysters of Long Island Sound. Biological Bulletin, Marine Biological Laboratory, Woods Hole 122, 86-94. Loosanoff, V. L. (1965b). The American or eastern oyster. Fish and Wildlife Service, Bureau of Commercial Fisheries, United States Department of the Interior circular 205, 36 pp. Loosanoff, V. L. (1968). Maturation of gonads of oysters, Crassostrea virginica of different geographical areas subjected to relatively low temperatures. The Veliger ll,153-163. Loosanoff, V. L. (1969). Development of shellfish culture techniques. Proceedings of the Conference on Artificial Propagation of Commercially Valuable Shellfish, October 22-23, College of Marine Studies, University of Delaware, Newark, DE. Loosanoff, V. L. and Davis, H. C. (1947). Staining of oyster larvae as a method for studies of their movements and distribution. Science, New York 106,597-598. Loosanoff, V. L. and Davis, H. C. (1950). Conditioning V mercenaria for spawning in winter and breeding its larvae in the laboratory. Biological Bulletin, Marine Biological Laboratory (Woods Hole) 98, 60-65. Loosanoff, V. L. and Davis, H. C. (1951). Delaying spawning of lamellibranchs by low temperature. Journal of Marine Research 10, 197-202. Loosanoff, V. L. and Davis, H. C. (1952). Temperature requirements for maturation of gonads of northern oysters. Science, New York 115,675-676. Loosanoff, V. L. and Davis, H. C. (1963). Rearing of bivalve mollusks. Advances in Marine Biology 1,1-136. Loosanoff, V. L. and Engle, J. B. (1942). Accumulation and discharge of spawn by oysters living at different depth. Biological Bulletin, Marine Biological Laboratory, Woods Hole 82,413422. Loosanoff, V. L. and Nomejko, C. A. (1951). Existence of physiologically different races of oysters, Crassostrea virginica. Biological Bulletin, Marine Biological Laboratory, Woods Hole 191,151-156. Loosanoff, V.L., Davis, H. C. and Chanley, P.E.(1955). Food requirements of some bivalve larvae. Proceedings of the National Shelljisheries Association 45, 66-83. Loosanoff, V. L., Davis, H. C. and Chanley, P. E. (1966). Dimensions and shapes of larvae of some marine bivalve molluscs. Malacologia 4, 351-435. Lough, R. G. (1974). A reevaluation of the combined effects of temperature and salinity on survival and growth of Mytilus edulis larvae using response surface techniques. Proceedings of the National Shelljisheries Association 64, 73-76. Lough, R. G. (1975). A reevaluation of the combined effects of temperature and salinity on survival and growth of bivalve larvae using response surface techniques. Fishery Bulletin 7, 86-94. Lourens, J. M., Vonck, A. I?, Guchte, C. V., Hartnack, J. and Stronkhorst, J. (1995). Sediment toxicity testing of lightly contaminated dredged material in The Netherlands. Journal of Aquatic Ecosystem Health 4, 271-275. Lu, J. K. (1986). The effects of temperature and salinity on meiosis, fertilization and first mitosis of the Pacific oyster (Crassostrea gigas) eggs. M.Sc. thesis, University of Washington, Seattle, WA, 110 pp. Lubet, F? (1959). Recherches sur le cycle sexuel et 1’Cmission des gamktes chez les Mytilidks et les Pectinid& Revue des Travaux de l’lnstitut des PCches Maritimes 23, 389-548.
THE ASSESSMENT OF MARINE POLLUTION
161
Lubet, P. (1973). Expos6 synoptique des donnCes biologiques sur la mode Mytilus galloprovincialis (Lamark, 1819) (Atlantique et MCditerranCe).Organisation des Nations Uines pour 1’Alimentation et l’Agriculture, Rome, janvier 1973. Synopsis F A 0 sur les P6ches No. 88 (SAST-Mode, 3, n16(10) 028,08, pag. var.). FAO, Rome. Lubet, P. (1980). Influence des facteurs externes sur la reproduction des lamellibranches. Oceanis 6 , 469-489. Lubet, P., Prunus, G., Masson, M. and Bucaille, D. (1984). Recherches exPCrimentales sur l’hybridation de Mytilus edulis L. et M . galloprovincialis Lmk (Mollusques Lamellibranches). Bulletin de la Societk Zoologique de France 109, 87-98. Lucas, A. (1975). Remarques methodologiques sur l’emploi des larves de rnoules comme tests biologiques. Haliotis 5, 126-132. Lucas, A. and Le Roux, S. (1975). Mise en Cvidence de la toxicit6 de divers pCtroles bruts vis-a-vis des larves de moules. Cornptes Rendus des Seances de I’Acadirnie des Sciences, Paris 280, 2381-2384. Lucas, A. and Le Roux, S. (1979). Les embryons et larves de bivalve organismes - tests vis-B-vis de la toxicit6 des hydrocarbures. Conseil International pour YExploration de la Mer, Marine Environmental Quality Committee, CM 1979lE, 4 pp. Lucas, A., Chebab-Chalabi, L. and Aldana Aranda, D. (1986). Passaige de l’endotrophie B l’exotrophie chez les larves de Mytilus edulis. Oceanologica Acta 9, 97-103. Lucu, C., Pavicic, J., Skreblin, M. and Mastrovic, M. (1980). Toxicological effects of biocides Slimicide C-30 on some marine invertebrates. Marine Pollution Bulletin 11, 294-296. Lund, D. S. (1972). Feeding studies with Pacific oyster larvae. Proceedings of the National Shellfisheries Association 63, 8 (abstract). Luoma, S. N. and Ho, K. T. (1993). Appropriate uses of marine and estuarine sediment bioassays. In “Handbook of Ecotoxicology” (€?Calow, ed.), vol. 1, pp. 193-226. Blackwell, Oxford. Lutz, R. and Hidu, H. (1979). Hinge morphogenesis in the shells of larval and early post-larval mussel (Mytilus edulis L. and Modiolus modiolus L.). Journal of the Marine Biological Association of the United Kingdom 59, 111-121. Lutz, R. A., Hidu, H. and Drobeck, K. G. (1970). Acute temperature increase as a stimulus to setting in the American oyster, Crassostrea virginica (Grnelin). Proceedings of the National Shelljisheries Association 60, 68-71. McDonald, J. H., Seed, R. and Koehn, R. K. (1991). Alloenzyme and morphometric characters of three species of Mytilus in the Northern and Southern hemispheres. Marine Biology 111, 323-335. MacDonald, R. W. and Thomas, D. J. (1983). Assessing environmentally acceptable levels of metals in drilling products. 12th annual Arctic Environmental Workshop, Fairmont, British Columbia, May 8th-11th 1983, 139-154. McFadzen, I. R. B. (1992). Growth and survival of cryopreserved oyster arid clam larvae along a pollution gradient in the German Bight. Marine Ecology Progress Series 91, 215-220. McFadzen, I. R. B. and Cleary, J. J. (1994).Toxicity and chemistry of the sea-surface microlayer in the North Sea using a cryopreserved larval bioassay. Marine Ecology Progress Series 103, 103-109. McHenery, J. G., Saward, D. and Seaton, D. D. (1991). Lethal and sub-lethalleffects of the salmon delousing agent dichlorvos on the larvae of the Lobster (Hornarus
162
E.
HIS, R. BEIRAS AND M. N. L. SEAMAN
gammarus L.) and herring (Clupea harengus L.). Aquaculture 98, 1-47. McHugh, J. L. and Sumner, M. W. (1988). Annotated bibliography I1 of the hard clam. United States Department of Commerce, National Oceanic and Atmospheric Administration, National Marine Fisheries Service, Scientific Publication Office, BLN C15700, Seattle, WA 98115 NOAA Technical Repoi? NMFS 68 (Contribution no 584 of the Marine Sciences Research Center of the State University of New York, Stony Brook), September 1988, 1-59. MacInnes, J. R. (1981). Response of embryos of the American oyster, Crassostrea virginica, to heavy metal mixtures. Marine Environmental Research 4, 217-227. MacInnes, J. R. and Calabrese, A. (1978). Response of embryos of the American oyster, Crassostrea virginica, to heavy metals at different temperatures. In “Proceedings of the 12th European Marine Biology Symposium” (D. S. McLusky and A. J. Berry, eds), pp. 195-202. Pergamon Press, Oxford. MacInnes, J. R. and Calabrese, A. (1979). Combined effects of salinity, temperature, and copper on embryos and early larvae of the American oyster, Crassostrea virginica. Archives of Environmental Contamination and Toxicology 8,553-562. MacInnes, J. R. and Thurberg, F. P. (1973). A new technique for measuring the oxygen consumption of larvae of the American oyster, Crassostrea virginica. Proceedings of the National Shellftsheries Association 63, 6 M 2 . Mackie, G. L. (1984). Bivalves. Zn “The Mollusca” (A. S. Tompa, N. H. Verdonk and J. A. M. den Biggelaar, eds), vol. 7, pp. 351-402. Academic Press, London. McLusky, D. S., Bryant, V. and Campbell, R. (1986). The effects of temperature and salinity on the toxicity of heavy metals to marine and estuarine invertebrates. Oceanography and Marine Biology Annual Review 24, 481-520. Magnusson, K., Ekelund, R., Dave, G., Granmo, A., Foerlin, L., Wennberg, L., Samuelsson, M.-O., Berggren, M. and Brostroem-Lunden, E. (1996). Contamination and correlation with toxicity of sediment samples from the Skagerrak and Kattegat. Journal of Sea Research 35, 223-234. Mainstone, C. €!, Butler, R. and Roddie, B. D. (1989). The effects of sewage sludge on two life-history stages of marine bivalve mollusks. 8th International Oceanography Disposal Symposium, Dubrovnik, Yugoslavia, 9-13 October 1989, Abstract, p. 8. Malueg, K. W., Schuytema, G. S., Gakstatter, J. H. and Frawczyk, D. F. (1983). Effect of hexagenia on Daphnia response in sediment toxicity tests. Environmental Toxicology and Chemistry 2, 73-82. Malueg, K. W., Schuytema, G. S., Frawczyk, D. F. and Ganstatter, J. H. (1986). Laboratory sediment toxicity test. Sediment chemistry and distribution of benthic macroinvertebrates in sediments from the Keweenaw Waterway. Environmental Toxicology and Chemistry 3, 233-242. Mann, R. (1979). Some biochemical and physiological aspects of growth and gametogenesis in Crassostrea gigas and Ostrea edulis at sustained elevated temperatures. Journal of the Marine Biological Association of the U.K. 59, 95-110. Mann, R and Rainer, J. (1990). Effect of decreasing oygen tension on swimming rate of Crassostrea virginica (Gmelin, 1791) larvae. Journal of Shellfish Research 9, 323-327. Mann, R., Campos, B. M. and Luckenbach, M. W. (1991). Swimming rate and responses of larvae of 3 mactrid bivalves to salinity discontinuities. Marine Ecology Progress Series 68(3), 257-270.
THE ASSESSMENT OF MARINE POLLUTION
163
Manzi, J. J., Bobo, M. Y. and Burrell, V. G. (1985). Gametogenesis in a population of the hard clam, Mercenaria mercenaria (Linnaeus) in North Santee Bay, South Carolina. Veliger 28, 18&194. Mariiio-Balsa, J. C. (1998). Toxicidade comparativa de metais pesados sobre as primeiras fases larvarias de crustaceos decapodos de interese comercial: Palaemon serratus, Maja squinado e Homarus gammarus. M.Sc. thesis, Lhiversity of Vigo, Galicia, 69 pp. Martin, M., Osborn, K. E., Billig, P. and Glickstein, N. (1981). Toxicities of ten metals to Crassostrea gigas and Mytilus edulis embryos and Cancer magister larvae. Marine Pollution Bulletin l2(9), 305-308. Mason, C. F. (1991). “Biology of freshwater pollution” 2nd edn., Longman, Essex, 351 pp. Masson, M. (1975). Etude exPCrimentale de la croissance et de la nutrition de la larve de Mytilus galloprovincialis Lmk (Mollusque PClCcypode). T h h e de Troisikme Cycle, Caen, 105 pp. Matthiessen, G. C. (1970). A review of oyster culture and oyster industry in North America. Woods Hole Oceanographic Institution, Contribution 2528, 1-52, Woods Hole, MA. Matthiessen, P., Bifield, S., Jarrett, F., Kirby, M. F., Law, R. J., McMinn, W. R., Sheahan, D. A., Thain, J. E. and Whale, G. F. (1998). An assessment of sediment toxicity in the Riber Tyne estuary, UK, by means of bioassays. Marine Environmental Research 45, 1-15. Maurer, D. and Price, K. S. Jr. (1968). Holding and spawning Delaware Bay oysters (Crassostrea virginica) out of season. 1. Laboratory facilities for retarding spawning. Proceedings of the National Shellfisheries 58, 71-77. Medcof, J. C. (1961). Oyster farming in the Maritimes. Bulletin of the Fisheries Research Board of Canada 131, 158 pp. Medcof, J. C. and Wolf, P. H. (1975). Spread of Pacific oyster worries NSW culturists. Australian Fisheries 45, 1-7. Mengus, B. (1978). R61e des bactkries dans l’alimentation des larves de mollusques bivalves marins en Clevages expCrimentaux. Bulletin de I’Observatoire de !a Mer, Le Brusc, Var, France, 3 (suppl. 7), 141 pp. Menzel, W. (1989). The biology, fishery and culture of quahog clam, Mercenaria. In “Developments in Aquaculture and Fisheries Sciences” (J. J. Manzi and M Castagna, eds), vol. 19, pp. 201-242. Elsevier, New York. Millar, R. H. (1961). Scottish oyster investigations 1946-1958. Department of Agriculture and Fisheries for Scotland. Marine Research 3, 76 pp. Millar, R. H. and Scott, J. M. (1967). The larvae of the oyster Ostrea edulis during starvation. Journal of the Marine Biological Association of the UK 4’7, 475484. Millar, R. H. and Scott, J. M. (1968). A n effect of water quality on the growth of cultured larvae of the oyster Ostrea edulis L. Journal du Conseil International Permanent pour I’Exploration de la Mer 32, 123-130. Miller, D. C. (1989). Effects of suspended sediments, hypoxia, and hyperoxia on larval Mercenaria mercenaria (Linnaeus, 1758). Journal of Shellfish Research 8, 37-42. Minchin, D., Duggan, C. B. and King, W. (1987). Possible effects of organotins on scallop recruitment. Marine Pollution Bulletin 18, 604-608. Mitchell, D. G., Morgan, J. D., Vigers, G. A. and Chapman, P. M. (1985). Acute toxicity of mine tailings to four marine species. Marine Pollution Bulletin 16, 450-455.
164
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Mitchell, R. (1974). Aspects of the ecology of the 1amelLibranch Mercenaria mercenaria in British waters. Hydrobiological Bulletin 8, 124-138. Moore, N. W. (1966). A pesticide monitoring system with special reference to the selection of indicator species. Journal of Applied Ecology 3 (suppl.) 261-269. Morgan, J. D., Mitchell, D. G. and Chapman, I? M. (1986). Individual and combined toxicity of manganese and molybdenum to mussel, M y t h edulis, larvae. Bulletin of Environmental Contamination and Toxicology 37, 303-307. Moriarty, F. (1990). “Ecotoxicology. The study of pollutants in Ecosystems”. Academic Press, London, 289 pp. Morrison, G. (1971). Dissolved oxygen requirements for embryo and larval development of the hard shell clam, Mercenaria mercenaria. Journal of the Fisheries Research Board of Canada 28, 379-381. Morrison, G. E. and Petrocelli, E. A. (1990). Suitability of Mulinia lateralis as a euryhaline toxicity test species. Canadian Technical Report of Fisheries and Aquatic Sciences 1 4 , 337-340. Morse, D., Hooker, N. and Morse, A. (1977). Hydrogen peroxide induces spawning in mollusks with activation of prostaglandin endoperoxide synthetase. Science 196,298-300. Muranaka, M. S. and Lannan, J. E. (1984). Broodstock management of Crassostrea gigas: environmental infiuences on broodstock conditioning. Aquaculture 39, 217-228. Nagabhushanam, R. (1963). Neurosecretory cycle and reproduction in the bivalve Crassostrea virginica. Indian Journal of Experimental Biology 1, 161-162. Nascimento, I. A. (1989). Bioassays of water quality in Aratu Bay, Bahia, Brazil using embryonic development of the mangrove oyster Crassostrea rizophorae (Guilding, 1828). Proceedings of the International Symposium on Utilization of Coastal Ecosystems: Planning, Pollution and Productivity, Rio Grande, RS, Brazil, November 21-27, 1982, 2, 195-212. Nell, J. A. and Holliday, J. E. (1986). Effects of potassium and copper on the settling rate of Sydney Rock oyster (Saccostrea commercialis) larvae. Aquaculture 58, 263-267. NeU, J. A. and Hofiday, J. E. (1988). Effects of salinity on the growth and survival of Sydney Rock oyster (Saccostrea commercialis) and Pacific oyster (Crassostrea gigas) larvae and spat. Aquaculture 68, 3 9 4 . Nelson, C. L. and Siddall, S. E. (1988). Effects of an algal bloom isolate on growth and survival of bay scallop (Argopecten irradians) larvae. Journal of Shelrfish Research 7 , 683-694. Nelson, D., Miller, J., Pereira, J. and Calabrese, A. (1983). Monitoring water quality at a dredge spoil dump site using oyster larvae. International Council for the Exploration of the Sea, Marine Environmental Quality Committee C.M. 1983lE: 59. Nelson, T. C. and Allison, J. B. (1940). On the nature and action of diantlin, a new hormone-like substance carried by the spermatozoa of the oyster. Journal of Experimental Zoology 85, 299-338. Neudeker, Th. (1981). Sex determination and maturation of the Pacific oyster (Crassostrea gigas) of the German coast. European Mariculture Society Special Publication 6, 249-254. Newell, R. I. E., Hilbish, T. J., Koehn, R. K. and Newell, C. J. (1982). Temporal variation in the reproductive cycle of Mytilus edulis L. (Bivalvia, Mytilidae) from localities on the east coast of the United States. Biological Bulletin, Marine Biological Laboratory, Woods Hole 162, 299-310.
THE ASSESSMENT OF MARINE POLLUTION
165
Newkirk, G. (1978). Interaction of genotype and salinity in larvae of the oyster, Crassostrea virginica. Marine Biology 48, 227-234. Newkirk, G. F., Waugh, D. L. and Haley, L. E. (1977). Genetics of larval tolerance to reduced salinities in two populations of oysters, Crassostrea virginica. Journal of Fishery Research Board of Canada 34, 384-387. Newman, N. C. (1995). Quantitative methods in aquatic ecotoxicology. Advances in Trace Substances Research, Lewis Publishers, Boca Raton, 426 pp. Nottage, A. S. and Birkbeck, T. H. (1987a). Production of proteinase during experimental infection of Ostrea edulis L. larvae with Vibrio alginolyticus NCMB 1339 and the antigenic relationship between proteinases produced by marine vibrios pathogenic for fish and shellfish. Journal of Fish Diseases 10(4), 265-273. Nottage, A. S. and Birkbeck, T. H. (1987b). Purification of a proteinase profduced by the bivalve pathogen Vibrio alginolyticus NCMB 1339. Journal of Fish Diseases lO(3) 211-220. Noyes, G. S., Haskin, H. H. and Van Dover, C. (1978). Oil and the oyster in Delaware Bay. Proceedings of the National Shellfisheries Association 68, 84-85 OECD (1998). Report of the OECD workshop on statistical analysis of atquatic toxicity data. OECD Environmental Health and Safety Publications. Series on Testing and Assessment no. 10. 133 pp. Organization for Economic Co-operation and Development, Paris. Oeschger, R. and Vetter, R. (1992). Sulfide detoxification and tolerartce in Halicryptus spinulosus (Priapulida): a multiple strategy. Marine Ecology Progress Series 86, 167-179. Okubo, K. and Okubo, T. (1962). Study on the bio-assay method for the evaluation of water pollution. I1 Use of the fertilized eggs of sea urchin and bivalves. Bulletin of the Tokai Regional Fishery Research Laboratory 32, 131-140. Orton, J. H. (1920). Sea temperature, breeding and distribution in marine animals. Journal of the Marine Biological Association of the United Kingdom 12, 339-366. Orton, J. H. (1926). On lunar periodicity in spawning of normally grown Falmouth oysters (0. edulis) in 1925, with a comparison of the spawning capacity of normally grown and dumpy oysters. Journal of the Marine Biological Association of the U.K. 14, 199-225. Orton, J. H. (1927). Observations and experiments on sex-change in the European oyster (Ostrea edulis). Part I. The change from female to male. Journal of the Marine Biological Association of the U.K. 14, 967-1045. Pablo, F., Buckney, R. T. and Lim, R. P. (1997). Toxicity of cyanamide complexes and a blast furnace effluent to larvae of the doughboy scallop, Chlamys asperrimus. Bulletin of Environmental Contamination and Toxicology 58( l), 93-100. Pagano, G., His, E., Beiras, R., Iaccarino, M., Oral, R., Quiniou, F., Warnau, M. and Trieff, N. (1996). Cytogenetic, developmental, and biochemical effects of aluminium, iron and their mixture in sea urchins and mussels. Archives of Environmental Contamination and Toxicology 31(4), 466477. Pan, Y. D., Rao, V. S., Mann, K. H., Brown, R. G. and Pocklington, R. (l996a). Effects of silicate limitation on production of domoic acid, a neurotoxin, by the diatom Pseudonitzschia multiseries. I. Batch culture studies. Marine Ecology Progress Series 131, 225-233. Pan, Y. D., Rao, V. S., Mann, K. H., Li, W. K. and Hamson, W. G. (1996b). Effects of silicate limitation on production of domoic acid, a neurotoxin, by the ,diatom
166
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
Pseudonitzschia multiseries. 11. Continuous culture studies. Marine Ecology Progress Series 131, 235-246. Pavicic, J. (1976). Combined cadmium-zinc toxicity on embryonic development of Mytilus galloprovincialis Lam. (Mollusca, Mytilidae). 3bmes JournCes d’Etudes sur les Pollutions (Split). Commission Internationale pour 1’Exploration Scientifique de la Mer MCditerranCe, 79-80. Pavicic, J. (1980). Interaction of Cd and Zn in relation to oxygen consumption in early stages of marine bivalve molluscs. Commission Internationale pour 1’Exploration Scientifique de la Mer MCditerranCe, 5bmes JournCes d’Etudes sur les Pollutions (Cagliari). pp. 627-634. Pavicic, J. and Pihlar, B. (1982). Toxic effects of cyanides (including complex metal cyanides) on marine organisms. Commission Internationale pour I’Exploration Scientifique de la Mer MCditerranCe, 6bmes JournCes &Etudes Pollutions (Cannes). pp. 817-820. Pavicic, J., Skreblin, M., Kregar, I., ”bsek-Znidaric, M. and Stegnar, F! (1984). Formation of inducible Cd-binding proteins similar to metallothioneins in selected organs and developmental stages of Mytilus galloprovincialis. Commission Internationale pour I’Exploration Scientifique de la Mer MCditerranCe, 7bmes JournCes d’Etudes Pollutions (Lucerne). pp. 699-705. Pavicic, J., Smodis, B., Skreblin, M., Kregar, I. and Stegnar, P. (1994a). Embryolarval tolerance of Mytilus galloprovincialis, exposed to elevated seawater metal concentrations. 2. Stage-specific fluctuations in sensitivity toward Zn and Cd and their bioaccumulation into veliger larvae. Comparative Biochemistry and Physiology 109(1), 3746. Pavicic, J., Skreblin, M., Kregar, I., Tusek-Znidaric, M. and Stegnar, P. (1994b). Embryo-larval tolerance of Mytilus galloprovincialis, exposed to the elevated sea water metal concentrations. 1.Toxic effects of Cd, Zn and Hg in relation to the metallothionein level. Comparative Biochemistry and Physiology 107(2), 249-257. Pawlik, J. R. (1992). Chemical ecology of the settlement of benthic marine invertebrates. Oceanography and Marine Biology, Annual Review 30,273-335. Pelletier, M. C., Burgess, R. M., Ho, K. T., Kuhn, A., McKinney, R. A. and Ryba, A. (1997). Phototoxicity of individual polycyclic aromatic hydrocarbons and petroleum to marine invertebrate larvae and juveniles. Environmental Toxicology and Chemistry 16,2190-2199. Perdue, J. A. and Erickson, G. (1984). A comparison of the gametogenic cycle between the Pacific oyster Crassostrea gigas and the Simunoe oyster, Crassostrea rivularis in Washington State. Aquaculture 37, 231-237. Pereira, S. A., Nascimento, I. A., Smith, D. H., De Aratijo, B. N. L., Milena,. M. S. and Marcia, M. A. (1998). The combined effects of temperature and metals copper, zinc and mercury on the embryological development of the mangrove oyster, Crassostrea rizophorae. Ecotoxicology and Environmental Restoration 1, 21-32. Phelps, H. L. and Mihursky, J. A. (1986). Oyster (Crassostrea virginica, Gmelin) spat settlement and copper in Aufwuchs Estuaries 9, 127-132. Phelps, H. L. and Warner. K. A. (1990). Estuarine sediment bioassay with oyster pediveliger larvae (Crassostrea gigas). Bulletin of Environmental Contamination and Toxicology 44(2), 197-204. Philips, D. J. H. (1980). Toxicity and accumulation of cadmium in marine and estuarine biota. In “Cadmium in the environment” (J. 0.Nriagu, ed.). part I, pp. 426-569. John Wiley, New York.
THE ASSESSMENT OF MARINE POLLUTION
167
Porter, H. (1964). Seasonal gonadal changes of adult clams, Mercenaria mercenaria (L.) in North Carolina. Proceedings of the National Shellfisheries Associatiton 55, 35-42. Prieur, D. (1981). Nouvelles donnCes sur les relations entre bactkries et bivalves marins. Haliotis 11,251-260. Prytherch, H. F. (1924). Experiments in the artificial propagation of oysters. Report of the US. Commissioner of Fisheries for the fiscal year 1923, appendix XI (document 961), Washington, 14 pp. Prytherch, H. F. (1931). The role of copper in the setting and metamorphosis of the oyster. Science, New York 73, 429-431. Prytherch, H. F. (1934). The role of copper in the setting, metamorphosis and distribution of the American oyster, Ostrea virginica. Ecological Monogrqphs 4, 49-107. Purchon, R. D. (1968). The biology of the Mollusca. In “International Series of Monographs in Pure and Applied Biology, Division Zoology” (G. A. Kcrkut, ed.), 57, 1-560. Pergamon Press, Oxford. Quayle, D. B. (1969). Pacific oyster culture in British Columbia. Bulletin of the Fisheries Research Board of Canada 169,1-193. Rainbow, I? S., Malik, I. and O’Brien, P. (1993). Physicochemical and physiological effects on the uptake of dissolved zinc and cadmium by the amphipod crustacean Orchestia gammarellus, Aquatic Toxicology 25, 15-30. Ramachandran, S., Patel, T. R. and Colbo, M. H. (1997). Effect of copper and cadmium on three Malaysian tropical estuarine invertebrate larvae. Ecotoxicology and Environmental Safety 36, 183-188. Rand, W. M. (1973). A stochastic model of the temporal aspect of breeding strategies. Journal of Theoretical Biology 40,337-351. Raven, C. P. (1964). Development. In “Physiology of Mollusca” (K. M. Wilbur and C. M. Yonge, eds), vol. I, pp. 165-188. Academic Press, London. Renard, F! (1991). Cooling and freezing tolerances in embryos of the F’acific oyster, Crassostrea gigas Methanol and sucrose effects. Aquaculture ’92(1), 43-58. Renzoni, A. (1973a). Influence of crude oil, derivatives and dispersants on larvae. Marine Pollution Bulletin 4(1), 9-13. Renzoni, A. (1973b). The influence of some detergents on the larval life of marine bivalves larvae. (Actes) 5Cme Colloque International d’OcCanographie Mkdicale, Messina. Italy, pp. 101-104. Renzoni, A. (1975). Toxicity of three oils to bivalve gametes and larvae. Marine Pollution Bulletin 6, 125-128. Rhodes, E. W. and Landers, W. S. (1976). Growth of oyster larvae Crassostrea virginica of various sizes in different concentrations of the chryso phyte Isochrysis galbana. Proceedings of the National ShellJisheriesAssociation 63, 53-59. Rhodes, E. W., Calabrese, A., Gable, W. D. and Landers, W. S. (1972). The development of methods for rearing the coot clam, Mulinia lateralis, and three species of coastal bivalves in the laboratory. In “Culture of Marine Invertebrate Animals” (W. L. Smith and M. H. Chanley, eds), pp. 273-282. Plenum Publishing Corporation, New York. Richardson, L. B., Burton, D. T. and Stavola, A. M. (1982). A comparison of ozone and chlorine toxicity to three life stages of the American oyster Crassostrea virginica. Marine Environmental Research 6 , 99-113. Riisgard, H. U., Randlov, A. and Kristensen, P. S. (1980). Rates of water
168
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
processing, oxygen consumption and efficiency of particle retention in veligers and young post-metamorphic Mytilus edulis. Ophelia 19, 3747. Riisgard, H. U., Randlov, A. and Hamburger, K. (1981). Oxygen consumption and clearance as function of size in Mytilus edulis L. veliger and larvae. Ophelia 20, 179-183. Ringwood, A. H. (1990). The relative sensitivities of different life stages of Isognomon californicum to cadmium toxicity. Archives of Environmental Contamination and Toxicology 19,338-340. Ringwood, A. H. (1991). Short-term accumulation of cadmium by embryos, larvae and adults of an Hawdian bivalve, Isognomon californicum. Journal of Experimental Marine Biology and Ecology 149, 55-66. Ringwood, A. H. (1992a). Effects of chronic cadmium exposures on growth of larvae of an Hawayan bivalve, Isognomon californicum. Marine Ecology Progress Series 83, 63-70. Ringwood, A. H. (1992b). Comparative sensitivity of gametes and early development stages of a sea urchin (Echinometra mathaei) and a bivalve species (Isognomon californicum) during metal exposures. Archives of Environmental Contamination and Toxicology 22, 288-295. Ringwood, A. H. (1993). Age-specific differences in cadmium sensitivity and bioaccumulation in bivalve molluscs. Marine Environmental Research 35, 35-39. Ringwood, A. H. and Brouwer, M. (1995). Patterns of metalloprotein expression in oyster embryos. Marine Environmental Research 39, 101-105. Ritchie, T. P. and Menzel, R. W. (1969). Influence of light on larval settlement of American oyster. Proceedings of the National Shellfiheries Association 59, 116-120. Robert, R. and His, E. (1981). Action de l’acktate de tributyle-Btain sur les oeufs et les larves D de deux mollusques d’intBr&tcommercial: Crassostrea gigas (Thunberg) et Mytillus galloprovincialis Lmk. Conseil International pour 1’Exploration de la Mer, Mariculture Committee, C.M. 1981E 42. Robert, R. and His, E. (1985). Combined effects of salinity and cadmium chloride upon embryos and larvae of the Japanese oyster, Crassostrea gigas. Marine Environmental Research 15,303-312. Robert, R. and Trintignac, P. (1997). Substitute for live microalgae in mariculture: a review. Aquatic Living Resources 10, 315-327. Robert, R., His, E. and Maurer, D. (1986). ToxicitB d’un dksherbant, l’atrazinesimazine, sur les jeunes stades larvaires de Crassostrea gigas et sur deux algues fourrages, Zsochrysis aff-galbana et Chaetoceros calcitrans. Haliotis 15, 319325. Robert, R., His, E. and Dinet, A. (1989). Combined effects of temperature and salinity on fed and starved larvae of the European flat oyster Ostrea edulis. Marine Biology 97, 95-100. Roberts, M. H. Jr (1980). Flow-through toxicity testing system for molluscan larvae as applied to halogen toxicity in estuarine waters. In “Aquatic Invertebrate Bioassays, American Society for Testing and Materials, Special Technical Publication 715” (A. L. B. Buikema Jr, and John Cairns Jr, eds), pp. 131-139. American Society for Testing and Materials, Philadelphia. Roberts, M. H. Jr (1987). Acute toxicity of tributyltin chloride to embryos and larvae of two bivalve mollusks, Crassostrea virginica and Mercenaria mercenaria. Bulletin Environmental Contamination Toxicology 39, 1012-1019. Roberts, M. H. Jr and Casey, B. B. (1985). Depression of larval growth and
THE ASSESSMENT OF MARINE POLLUTION
169
metamorphosis of oysters exposed to chlorinated sewage. Proceedings of the 5th Conference on Water Chlorination: Environmental Impact and Health Effects, Williamsburg, Virginia, June 3-8, 1984 (R. L. Jolley, R. J. Bull, W. €! Davis, S. Katz, M. H. Roberts and V. A. Jacobbs eds), Lewis Publishers Inc., Chapiter 40, 509-520. Roberts, M. H. Jr and Gleeson, R. A. (1978). Acute toxicity of bromochlorinated seawater to selected estuarine species with a comparison to chlorinated seawater toxicity. Marine Environmental Research 1, 19-30. Roberts, M. H. Jr, Diaz, R. J., Bender, M. E. and Huggett, R. J. (1975). Acute toxicity of chlorine to selected estuarine species. Journal of Fishery Raearch Board of Canada 32, 2525-2527. Roberts, M. H. Jr, Buikema, A. L. Jr and Cairns, J. Jr (eds) (1977). Flow-through toxicity system for molluscan larvae as applied to halogen toxicity in estuarine water. Aquatic invertebrate bioassays. A symposium sponsored by ASTM Committee D-19 on Water, American Society for Testing and Materials, Blacksburg, VA, 27-29 September 1977. Special Technical Publication of The American Society for Testing and Materials, v. 1980, published by The American Society for Testing and Materials, Baltimore, MD, pp. 131-139. Robinson, A. and Breese, W. €? (1984). Gonadal development and hatchery rearing techniques for the Manilia clam, Tapes philippinarum (Adams and Reeve). Journal of Shellfish Research 4(2), 161-163. Robinson, W. E. (1983). Quantification of ingestion by Mercenaria mercenaria (L.) veligers feeding on mixed suspensions of inert material using microspectrofluonmetry. Journal of Molluscan Studies, Supplement l2A, 167-171. Roesijadi, G., Hansen, K. M. and Unger, M. E. (1996). Cadmium-imduced metallothionein expression during embryonic and early larval development of the mollusc Crassostrea virginica. Toxicology and Applied Pharmacology 140, 356-363. Roesijadi, G., Hansen, K. M. and Unger, M. E. (1997). Metallothionein nnRNA accumulation in early developmental stages of Crassostrea virginica following pre-exposure and challenge with cadmium. Aquatic Toxicology 39, 185-1 94. Roosenburg, W. H., Drobeck, K. G., Hidu, H., McErlean, A. J. and Mihursky, J. A. (1969). Acute temperature tolerance of oyster larvae as related to power plant operation. Proceedings of the National Shellfisheries Association 60, 11 (abstract). Roosenburg, W. H., Rhoderick, J, C., Block, R. M., Kennedy, V. S. and Vreenegoor, S. M. (1980a). Survival of Mya arenaria larvae (Mollusca: Bivalvia) exposed to chlorine-produced oxidants. Proceedings of the National Shelljirheries Association 70, 105-1 1. Roosenburg, W. H., Rhoderick, J. C., Block, R. M., Kennedy, V. S., Gullans, S. M., Vreenegoor, S. R., Rosenkranz, A. and Collette, C. (1980b). Effects of chlorineproduced oxidants on survival of larvae of the oyster Crassostrea virginica. Marine Ecology Progress Series 3, 93-96. Ruddy, F. M., Feng, S. Y. and Campbell, G. S. (1975). The effects of prolonged exposure to elevated temperatures on the biochemical constituents, gonadal development and shell deposition of the American oyster, Crassostrea virginica. Comparative Biochemistry and Physiology 51, 157-164. Ruiz, J. M., Bryan, G. W. and Gibbs, l? E. (1994). Bioassaying the toxicity of tributyltin (TBT) polluted sediment to spat of the bivalve Scrobicularia plana. Marine Ecology Progress Series 113, 119-130. Ruiz, J. M., Bryan, G. W. and Gibbs, P. E. (1995a). Effects of tributyltin (TBT)
170
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
exposure on the veliger larvae development of the bivalve Scrobicularia plana (da Costa). Journal of Experimental Marine Biology and Ecology 186, 53-63. Ruiz, J. M., Bryan, G. W. and Gibbs, F! E. (1995b). Acute and chronic toxicity of tributyltin (Tl3T) to pediveliger larvae of the bivalve Scrobicularia plana. Marine Biology W, 119-126. Ruiz, J. M., Bryan, G. W., Wigham, G. D. and Gibbs, P. E. (1995~).Effects of copper on the 48-h embryonic development of the bivalve Scrobicularia. Journal of the Marine Biological Association of the U K . 76, 829-832. Sakuda, H. M. (1966). Reproductive cycle of the American oyster, Crassostrea virginica in West Loch, Pearl Harbor, Hawali. Transactions of the American Fisheries Society 95, 216-218. Samoiloff, M. R. (1989). Toxicity testing of sediments: problems, trends, and solutions. In “Aquatic Toxicology and Water Quality Management” (J. 0. Nriagu and J. S. S. Lacksminarayana, eds), pp. 143-152. Wiseley, New York. Samoilloff, M. R. and Wells, P. G. (1984). Future trends in marine ecotoxicology. In “Ecotoxicological testing for the marine environment” (G. Personne, E. Jaspers and C. Claws, eds), 1, 733-749. Laboratory of Biological Research and Aquatic Pollution, State University, Ghent. Sanders, H. L. and Hessler, R. R. (1969). Ecology of the deepsea benthos. Science, New York 163, 1419-1424. Sastry, A. N. (1975). Physiology and ecology of reproduction in marine invertebrates. In “Physiological Ecology of Estuarine Organisms” (F. J. Vernberg, ed.), pp. 279-299. University of South Carolina Press, Columbia, SC. Sastry, A. N. (1979). Pelecypoda (excluding Ostreidae). In “Reproduction of Marine Invertebrates” (A. C. Giese and J. S. Pearse, eds), 5,113-265. Academic Press, New York. Schuytema, G. S., Nebeker, A. V., Griffis, W. L. and Miller, C. E. (1989). Effects of freezing on toxicity of sediments contaminated with DDT and endrin. Environmental Toxicology and Chemistry 8, 883-891. Seaman, M. N. L., His, E., Keskin, M. and Reins, T. (1991). Influence of turbulence and turbidity on growth and survival of laboratory-reared bivalve larvae. International Council for the Exploration of the Sea, Shellfish Committee C.M. 1991: K:56, 6 pp. Seaman, M. N. L. and Ruth, M. (1997). The molluscan fisheries of Germany. NOAA. NMFS Technical Report, 129, pp. 57-84. Sedano, F. J., Rodriguez, J. L., Ruiz, C., Garcia-Martin, L. 0. and Sanchez, J. L. (1995). Biochemical composition and fertilization in the eggs of Mytilus galloprovincialis Lmk. Journal of Experimental Marine Biology and Ecology 192, 75-85. Seed, R. (1969). The ecology of Mytilus edulis L. (Lamellibranchiata) on exposed rocky shores. 1. Breeding and settlement. Oecologia 3, 277-316. Seed, R. (1971). A physiological and biochemical approach to the taxonomy of Mytilus edulis L. and M. galloprovincialis Lmk. from the S. W. England. Cahiers de Biologie Marine, Roscoff, 12, 291-322. Seed, R. (1976). Ecology. In “Marine mussels: their ecology and physiology” (€3.L. Bayne, ed.), pp. 13-66. Cambridge University Press, Cambridge. Shaw, W. N. (1965). Seasonal setting patterns of five species of bivalves in Tred Avon River, Maryland. Chesapeake Science 6, 33-37. Sheffrin, N. M. H. (1982). The effects of impaired water quality on the behaviour of Mytilus edulis plantigrades. Ph.D. thesis, University of Sheffield. Sheffrin, N. M. H., Reller, N. R. J. and Williams, E. E. (1984). A behavioural
THE ASSESSMENT OF MARINE POLLUTION
171
bioassay for impaired sea-water quality using the plantigrades of the cornmon mussel Mytilus edulis L.: the response to copper. Aquatic Toxicology 5, 77-91. Shpigel, M., Coon, S. L. and Kleinot, P. (1989). Growth and survival of cultchless spat of Ostrea edulis Linnaeus, 1750 produced using epinephrine and shell chips. Journal of Shellfish Research 8, 355-357. Sigler, M. and Leibovitz, L. (1982). Acute toxicity of oil and bilge cleaners to larval American oysters (Crassostrea virginica). Bulletin of Environmental Contamination and Toxicology 29, 137-145. Smith, C. M. and Hackney, C. T. (1989). The effects of hydrocarbons on the setting of the oyster, Crassostrea virginica, in intertidal habitats in Southern North Carolina. Estuaries 12, 42-48. Smith, J. E. (1968). “Torrey Canyon pollution and marine life”. Cambridge University Press, Cambridge, 195 pp. Smith, J. R. and Strehlow, D. R. (1983). Algal-induced spawning in the marine mussel Mytilus californianus. International Journal of Invertebrate Reproduction 6 , 123-133. Spangenberg, J. V. and Cherr, G. N. (1996). Developmental effects of barium exposure in a marine bivalve (Mytilus californianus). Environmental Toxicology and Chemistry 15, 1769-1774. Sparks, R. (1925). Studies on the biology of oyster (Ostrea edulis) in the Limlfjord, with special reference to the influence of temperature on the sex change. Report of the Danish Biological Station to the Board of Agriculture 30, 84 pp. Sprung, M. (1984). Physiological energetics of mussel larvae (Mytilus edur‘is).I. Shell growth and biomass. Marine Ecology Progress Series 17, 283-293. Sprung, M. and Bayne, B. L. (1984). Some practical aspects of fertilising the eggs of the mussel Mytilus edulis L. Journal du Conseil International Permament pour I’Exploration de la Mer 41, 125-128. Staeger, W. H. and Horton, H. F. (1976). Fertilization method quantifying gamete concentrations and maximizing larvae production in Crassostrea gigas. Fishery Bulletin National Marine Fishery Service, US. 74, 698-701. Stafford, J. (1913). The Canadian oyster. Its development, environment and culture. Report to the Canadian Commission of Conservation. Committee on Fisheries, Game and Fur-bearing Animals, 159 pp. Mortimer Company, Ottawa. Stanley, J. G. and DeWitt, R. (1983). Species profiles: life histories and environmental requirements of coastal fishes and invertebrates (North Atlantic) Hard clam. U.S. Fish and Wildlife Service, Bureau of Commercial Fisheries, FWS/OBS-82/11.18. U.S. Army Corps of Engineers, TR EL-82-4, 19 pp. Stauber, J., Gunthorpe, L., Woodworth, J., Munday, B. M., Krassoi, R. and Simon, J. (1996). Comparative toxicity of effluents from ECF and TCF bleaching of eucalypt kraft pulps. Appita 49, 184-188. Stauber, L. A. (1950). The problem of physiological species with special reference to oysters and oyster drills. Ecology 31, 109-118. Stebbing, A. R. D. (1982). Hormesis - The stimulation of growth by low levels of inhibitors. The Science of the Total Environment 22, 213-234. Stebbing, A. R. D. (1985). Bioassay. In “The effects of stress and polluti’on on marine animals” (B. L. Bayne, D. A. Brown, K. Bums, D. R. Dixoln, A. Ivanovici, D. R. Livingstone, D. M. Lowe, M. N. Moore, A. R. D. Stebbing and J. Widdows, eds), pp. 133-140. Praeger Publishers, New York. Stebbing, A. R. D., Akesson, B., Calabrese, A., Gentile, J. H., Jensen, A. and Lloyd, R. (1980). The role of bioassays in marine pollution monitoring. Rapports
172
E. HIS, R. EEIRAS AND M. N. L. SEAMAN
et Proc2s-verbaux des Rtunions du Conseil International Permament pour 1’Exploration de la Mer 179,322-332. Stephano, J. L. and Gould, M. (1988). Avoiding polyspermy in the oyster, (Crassostrea gigas). Aquaculture 73,295-307. Stewart, M. E. and Blogoslawski, W. J. (1985). Effect of selected chlorine-produced oxidants on oyster larvae. In Water Chlorination Chemistry, Environmental Impact and Health Effects. (R. L. Jolley, R. J. Bull, W. P. Davis, S. Katz, M. H. Roberts Jr. and V. Jacobs, eds), Lewis Publishers Incorporation, Chelsea, Michigan, Volume 5, Chapter 41, 521-532. Stewart, M. E., Blogoslawski, W. J., Hsu, R. Y. and Helz, G. R. (1979). By-products of oxidative biocides: toxicity to oyster larvae. Marine Pollution Bulletin 10, 166-169. Stewart, N. E., Milleman, R. E. and Breese, W. P. (1967). Acute toxicity of insecticide Sevin and its hydrolytic product 1-naphthol to some marine organisms. Transactions of the American Fishery Society 96(1), 25-30. Stiles, S. and Blogoslawski, W. (1993). Viability and genetic effects on oyster embryos exposed to bacteria and effluent from diseased juvenile oysters from a Long Island hatchery. Journal of Shellfish. Research 12, 114. Stiles, S. S. and Longwell, C. (1973). Fertilization, meiosis and cleavage in eggs from large mass spawnings of Crassostrea virginica Gmelin, the commercial American oyster. Caryologia 26, 253-262. Stiles-Jewell, S. (1994). Cytogenetic and viability effects of petroleum aromatic and PCB hydrocarbons, temperature and salinity, on early development of the American oyster, Crassostrea virginica Gmelin. Ph.D. thesis, University of Massachusetts, Amherst, No. 9420694, 336 pp. Stocker, H. S. and Seager, S. L. (1976). Environmental Chemistry: air and water pollution, 2nd edn. Scott Foresman and Company, Glenview, 231 pp. StrGmgren, T. and Nielsen, M. V. (1991). Spawning frequency, growth and mortality of Mytilus edulis larvae, exposed to copper and diesel oil. Aquatic Toxicology 21, 171-180. Sugiura, Y . (1962). Electrical induction of spawning in two marine invertebrates (Unchis unicintus and hermaphroditic Mytilus edulis). Biological Bulletin, Marine Biological Laboratory, Woods Hole 123,203-206. Sullivan, C. M. (1948). Bivalve larvae of Malpeque Bay, P.E.I. Bulletin Fishery Research Board of Canada 77, 1-36. Sumner, C. E. (1980a). Growth of Pacific oysters, Crassostrea gigas Thunberg, cultivated in Tasmania. I. Xntertidal stick culture. Australian Journal of Marine and Freshwater Research 31, 129-135. Sumner, C. E. (1980b). Growth of Pacific oysters, Crassostrea gigas Thunberg, cultivated in Tasmania. 11. Subtidal culture. Australian Journal of Marine and Freshwater Research 32,411416. Sverdrup, H. U., Johnson, M. W. and Fleming, R. H. (1949). “The oceans”. Prentice-Hall, New York, 1087 pp. Swartz, R. C. (1989). Marine sediment toxicity tests. In “Proceedings of the National Research Council Symposium on Contaminated Marine Sediments”, (K. Kamlet, ed.), Washington, D.C. National Academic Press, pp. 115-129. Tagatz, M. E. and hey, J. M. (1981). Effects of fenvalerate on field- and laboratory-developed estuarine benthic communities. Bulletin of Environmental Contamination and Toxicology 27(2), 256-267. Taylor, D. (1979). A review of the lethal and sub-lethal effects of mercury on aquatic life. Residue Review 72,33-70.
THE ASSESSMENT OF MARINE POLLUTION
173
Taylor, D. (1981). “A summary of the data on the toxicity of various metals to the aquatic life, 2nd edn”. Vol. 2, 31lAl2073. Brixham Laboratory Report, 46 pp. Tazawa, E., Fujiwara, A. and Yasumasu, I. (1985). Change in glycogen level in the oyster eggs during development. Zoological Science 2, 349-357. Tebble, N. (1966). “British bivalves seashells”. Trustees of the British Museum (Natural History), Alden Press, Oxford, 212 pp. Thain, J. E. (1983). The acute toxicity of bis (Tributyl tin) oxide to the adult and larvae of some marine organisms. International Council for the Exploration of the Sea, Marine Environmental Committee, CM 1983/E:13, 5 pp. Thain, J. E. (1991). Biological effects of contaminants: oyster (Crassostrea gigas) embryo bioassay. Techniques in Marine Environmental Sciences, ICES 11, 1-12. Thain, J. E. (1992). Use of oyster Crassostrea gigas embryo bioassay on water and sediment elutriate samples from the German Bight. Marine Ecology Progress Series 91, 211-213. Thain, J. E. and Kirby, M. (1994). Improving the sensitivity of biological water quality measurements in marine waters. In “Scientific Symposium on the 1993 North Sea Quality Status Report”, (Danish Environmental Protection Agency, ed.), pp. 152-156. Ministry of Environment and Energy, Denmark. Thain, J. E. and Watts, J. (1987). The use of bioassay to measure changes in water quality associated with a bloom of Gyrodinium aureolum Hulburt. Rapports et Procks-verbaux des Rtunions du Conseil International Permament pour !’Exploration de la Mer 187, 103-107. Thain, J. E., Matthiessen, l? and Bifield, S. (1990). The toxicity of dichlorvos to some marine organisms. International Council for the Exploration of the Sea, Copenhagen (Denmark) Marine Environmental Quality Committee, CM1990/E: 18, 15 pp. Thompson, J. M. (1951). The acclimatization and growth of the Pacific oyster, (Gryphaea gigas) in Australia. Australian Journal of Marine and Freshwater Research 3, 64-73. Thompson, R. J. (1979). Fecundity and reproductive effort in the blue mussel (Mytilus edulis), the sea urchin (Strongylocentrus droebachiensis) and the snow crab (Chinoecetes opilio) from populations in Nova Scotia and Newfoundland. Journal of Fishery Research Board of Canada 36, 955-964. Thornton, I. (1975). Geochemical parameters in the assessment of estuarine pollution. In “The Ecology of Resource Degradation and Renewal” (hf. J. Chadwick and G. T. Goodman, eds) 15th Symposium of the British Ecological Society, pp. 157-169. Blackwell Scientific, Oxford. Thorson, G. (1950). Reproductive and larval ecology of marine bottom invertebrates. Biological Reviews of the Cambridge Philosophical Society 2, 1-45. Thurberg, F. F,! Cable, W. D., Dawson, M. A., MacInnes, J. R. and Wenzloff, D. R. (1975). Respiratory response of larval, juvenile and adult surf clam, Spisula solidissima, to silver. In “Respiration of Marine Organisms” (J. J. Cech, Jr, Di. W. Bridges and D. B. Horton, eds), pp. 41-52. TRIGOM Publications, South Portland, ME. Timmermans, B. M. H., Hummel, H. and Bogaards, R. H. (1996). The effect of polluted sediment on the gonadal development and embryogenesis of bivalves. The Science of the Total Environment 187(3), 231-237. Togo, T., Osanai, K. and Morisawa, M. (1995). Existence of three mechanisms of blocking polyspermy in ovocytes of the mussel, Mytilus edulis. Biological Bulletin, Marine Biological Laboratory, Woods Hole 189, 330-339.
174
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
lhmer, H. J. (1958). Abnormal development of quahog larvae resulting from multiple fertilization. In Report on Investigations of the Shellfisheries of Massachusetts for 1957, pp. 7-8. Woods Hole Oceanographic Institution, Department of Natural Resources, Division of Marine Fisheries. Woods Hole, Massachusetts. Tzong-Shean, C. and Chen, H.-C. (1993). Toxic effects of mercury on the hard clam, Meretrix lusoria, in various salinities. Comparative Biochemistry and Physiology 105,501-507. Ukeles, R. (1975). Views on bivalve larvae nutrition. Proceedings of the 1st International Conference on Aquaculture Nutrition, Sea Grant Program University, Delaware, October 14-15, 1975 (in cooperation with the U.S./Japan Aquaculture Panel), 127-162. Ukeles, R. and Sweeney, B. M. (1969). Influence of dinoflagellate trichocysts and other factors on the feeding of Crassostrea virginica larvae on Monochrysis lutheri. Limnology and Oceanography 14,403410. US-EPA (1991). Evaluation of dredged material proposed for ocean disposal Testing manual. U.S. Environmental Protection Agency/U.S. Army Corps of Engineers. EPA-503-8-91/001, Washington, D.C. Utting, S. D. and Helm, M. M. (1985). Improvement of sea water quality by physical and chemical pre-treatment in a bivalve hatchery. Aquaculture 44, 133-144. Utting, S. D. and Millican, P F. (1997). Techniques for the hatchery conditioning of bivalve broodstocks and the subsequent effect on egg quality and larval viability. Aquaculture 155, 45-54. Utting, S. D. and Spencer, B. E. (1991). “The hatchery culture of bivalve mollusc larvae and juveniles”. Ministry of Agriculture, Fisheries and Food Laboratory Leaflet, Lowestoft 68,31 pp. Valli, G., Zanelli, R. and Pittinelli, P. (1979). Aspects de la reproduction et de I’accroissement de Crassostrea gigas (Thunberg) en Clevage dans la lagune de Grado (Golfo di Trieste). Rapports de la Commission Internationale pour 1’Exploration Scientifique de la Mer Mkditerrante 25/26,231-232. Van der Hurk, P. (1994). Effects of natural sediment properties on test results in bioassays with oyster larvae (Crassostrea gigas) on sediment elutriates. Journal of Aquatic Ecosystem Health 3, 185-191. Van der Hurk, P, Eertman, R. H. M. and Sronkhorst, J. (1997). Toxicity of Harbour Canal sediments before dredging and after off-shore disposal. Marine Pollution Bulletin 34, 244-249. Vanhaecke, P. and Bogaert, T. (1984). Standardization of marine ecotoxicological test methods. Summary of a round-table discussion held during the MARTOXN Symposium (Ghent, Belgium, 1983). In “EcotoxicologicalTesting for the Marine Environment” (G. Personne, E. Jaspers, and C. Clauss, eds) 1, 753-756. Laboratory of Biological Research on Aquatic Pollution, State University, Ghent. Veitch, F. P. and Hidu, H. (1971). Gregarious setting in the American oyster, Crassostrea virginica (Gmelin). 1. Properties of a partially purified “setting factor”. Chesapeake Science 12, 173-178. Ventilla, R. F. (1984). Recent development in the Japanese oyster culture industry. Advances in Marine Biology 21, 1-57. Von Burg, R. and Greenwood, M. R. (1991). Mercury. In “Metals and their compounds in the environment” (E. Merian, ed.) VCH, Weinheim, 1439 pp. Walker, R. L. and Heffernan, P. B. (1995). Sex ratio of the northern quahog
THE ASSESSMENT OF MARINE POLLUTION
175
according to age, size, and habitat in coastal waters of Georgia. Transactions of the American Fisheries Society 124, 929-934. Waller, T. R. (1981). Functional morphology and development of veliger larvae of the European oyster, Ostrea edulis L. Smithsonian Contributions to Zoology 328, 70 PP. Walne, P. R. (1956). Experimental rearing of the larvae of Ostrea edulis L. in the laboratory. Fishery Investigations, London 20(9), 1-23. Walne, P. R. (1958). The importance of bacteria in laboratory experiments on rearing the larvae of Ostrea edulis. Journal of the Marine Biological Association of the U.K. 37,415425. Walne, P. R. (1963). Observations on the food value of seven species of algae to the larvae of Ostrea edulis L. Feeding experiments. Journal of the Marine Biological Association of the U.K. 43, 767-784. Walne, P. R. (1965). Observations on the influence of food supply and temperature on the feeding and growth of the larvae of Ostrea edulis L. Fishery Znvestigai’ions, London 24,45 pp. Walne, P. R. (1966). Experiments in the large-scale culture of the larvae of Ostrea edulis L. Fishery Investigations, London 25(4), 53 pp. Walne, P. R. (1970). Present problems in the culture of the larvae of Ostrea etiulis. Helgolander Wissenchafrliche Meeresuntersuchungen 20, 514-525. Walne, P. R. (1974a). The culture of marine bivalve larvae. In “Physiology of Mollusca I” (K. M. Wilbur and C. M. Yonge, eds), pp. 197-207. Academic Press, London. Walne, P. R. (1974b). “Culture of bivalve molluscs. 50 years experience at Conwy”. Fishing News (Books), Surrey, England, 173 pp. Walne, P. and Helm, M. M. (1974). “The routine culture of the Pacific oyster (Crassostrea gigas) at Conwy during (1973)”. Shellfish Information Le.aflet, Ministry of Agriculture, Fisheries and Food 32, 10 pp. Wang, W. X. and Widdows, J. (1991). Physiological responses of mussel larvae Mytilus edulis to environmental hypoxia and anoxia. Marine Ecology Prol;ress Series 70, 223-236. Ward, J. A., Word, J. Q., Pinza, M. R., Mayhew, H. L. and Barrows, E. S. (1992a). Ecological evaluation of proposed discharge of dredged material from Oakland Harbor into ocean waters. (Phase 3A of 42-foot project). Technical Report of the Battelle Pacific Northwest Laboratories 1, PNL-8302, analyses and di.,C’CUSsion 169 pp. Ward, J. A., Word, J. Q., Pinza, M. R., Mayhew, H. L. and Barrows, E. S. (1992b). Ecological evaluation of proposed discharge of dredged material from Oakland Harbor into ocean waters (Phase 3A of 42-foot project). Technical Report of the Battelle Pacific Northwest Laboratories 2, 685 pp. Waterman, A. J. (1937). Effect of salts of heavy metals on development of the. sea urchin, Arbacia punctulata. Biological Bulletin, Marine Biological Laborai!ory, Woods Hole 73, 401420. Watling, H. R. (1978). Effect of cadmium on larvae and spat of the oyster Crassostrea gigas (Thunberg). Transactions of the Royal Society of South AIFrica 43, 125-134. Watling, H. R. (1981). Effects of metals on the development of oyster embryos. South African Journal of Sciences 77, 134-135. Watling, H. R. (1982). Comparative study of the effects of zinc, cadmium and copper on the larval growth of three oyster species. Bulletin of Environmental Contamination and Toxicology 28, 195-201.
176
E. HIS, R. BEIRAS AND M.
N. L. SEAMAN
Watling, H. R. (1983). Comparative study of the effects of metals on the settlement of Crassostrea gigas. Bulletin of Environmental contamination and Toxicology 31, 344-351. Waugh, G. D., Hawes, F. B. and Williams, F. (1952). Insecticides for preventing barnacle settlement. Annuls of Applied Biology 39, 407415. Webb, K. L. and Chu, F. E. (1981). Phytoplankton as food source for bivalve larvae. In “2nd International Conference on Aquaculture, Nutrition: Biochemical and Physiological Approaches to Shellfish Nutrition” (G. D. Pruder, C. J. Langdon and D. E. Conklin, eds), pp. 272-291. Louisiana State University, Baton Rouge, LA. Webb, M. (1979). The metallothioneins. In ‘The Chemistry, Biochemistry, and Biology of Cadmium” (M. Webb, ed.), pp. 195-266. Elsevier, Amsterdam. Weiner, R., Walch, M., Labare, M. P., Bonar, D. B. and Colwell, R. R. (1989). Effect of biofilms of the marine bacterium Alteromonas colwelliana (LST) on set of the oysters Crassostrea gigas (Thunberg, 1793) and Crassostrea virginica (Gmelin, 1791). Journal of Shellfish Research 8, 117-123. Wells, W. F. (1927). Report of the experimental shellfish station. Report ofthe New York Station Conservation Department 16, 1-22. Widdows, J. (1991). Physiolog&al ecology of mussel larvae. Aquaculture 94, 147-163. Widdows, J. (1993). Marine and estuarine invertebrate toxicity tests. In “Handbook of Ecotoxicology” (P. Calow, ed.), 1, pp. 145-166. Blackwell, Oxford. Widdows, J., Newell, R. I. E. and Mann, R. (1989). Effects of hypoxia and anoxia on survival, energy metabolism and feeding of oyster larvae (Crassostrea virginica Gmelin). Biological Bulletin, Marine Biological Laboratory, Woods Hole 177,154-166. Wikfors, G. H. and Smolowitz, R. (1995). Experimental and histological studies of four life history stages of the eastern oyster, Crassostrea virginica, exposed to a culture strain of the dinoflagellate Prorocentrum minimum. Biological Bulletin, Marine Biological Laboratory (Woods Hole) 188, 313-328. Wikfors, G. H. and Ukeles, R. (1982). Growth and adaptation of estuarine unicellular algae in media with excess copper, cadmium or zinc, and effects of metal-contaminated algal food on Crassostrea virginica larvae. Marine Ecology Progress Series 7 , 191-206. Wikfors, G. H., Smolowitz, R. M. and Smith, B. C. (1993). Effects of a Prorocentrum isolate upon the oyster, Crassostrea virginica: a study of three life-history stages. Journal of Shellfish Research 12, 114-115. Wilde, €? A. W. J. de and Berghuis, E. M. (1977). Laboratory experiments on the spawning of Macoma balthicap: its implication for production research. In Proceedings of the 12th European Symposium on Marine Biology, Stirling, Scotland (D. S. McLusky and A. J. Berry, eds), pp. 355-384. Pergamon Press, Oxford. Williams, L. G., Chapman, P. M. and Ginn, T. C. (1986). A comparative evaluation of marine sediment toxicity using bacterial luminescence, oyster embryo and amphipod sediment bioassays. Marine Environmental Research 19, 225-249. Wilson, D. F! (1951). A biological difference between natural waters. Journal ofthe Marine Biological Association of the U.K. 30,1-21. Wilson, D. P. and Armstrong, F. A. T. (1961). Biological differences between sea waters: experiments in 1960. Journal of the Marine Biological Association of the U.K. 41, 663-681.
THE ASSESSMENT OF MARINE POLLUTION
177
Wilson, J. (1981). Hatchery rearing of Ostrea edulis and Crassostrea gigas. Aquaculture Technical Bulletin, Ireland 4, 1-34. Wilson, J. H. and Simons, J. (1985). Gametogenesis and breeding of Ostrea edulis on the west coast of Ireland. Aquaculture 46, 307-321. Wilson, S. l? and Hyne, R. V. (1997). Toxicity of acid-sulfate soil leachate and aluminium to embryos of the Sydney Rock oyster. Ecotoxicology and Environmental Safety 37, 30-36. Wirth, E. F., Scott, G. I., Fulton, M. H., Van Dolah, R. F., Maier, P. l?, Hadley, N., Daugomah, J. W. and Key, P. B. (1996). In situ monitoring of dredged material spoil sites using oyster Crassostrea virginica. Archives of Environmental Contamination Toxicology 30, 340-348. Wiseley, B. and Blick, R. A. P. (1967). Mortality of marine invertebrate larvae in mercury, copper and zinc solutions. Australian Journal of Marine and Freshwater Research 18,63-72. Woelke, C. E. (1960a). The effects of sulfite waste liquor on the development of eggs and larvae of two marine molluscs and three of their food organisms. Washington State Department of Fisheries Research Bulletin 6, 86-106. Woelke, C. E. (1960b). Effects of sulfite waste liquor on the normal developiment of Pacific oyster (Crassostrea gigas) larvae. Washington State Departmetzt of Fisheries, Research Bulletin 6 , 149-161. Woelke, C. E. (1961). Bioassay: the bivalve larvae tool. In “Proceedings of the Northwest Symposium on Water Pollution Research”, US Department HEWF’HS, Portland, OR, pp. 113-123. Woelke, C. E. (1966). Bioassay with bivalve larvae. In “18th annual report of the Pacific Marine Fisheries Commission for the year 1965”. Pacific Marine Fisheries Commission, Portland, OR, Washington Department of Fisheries, pp. 33-35 Woelke, C. E. (1967). Measurement of water quality with the Pacific oyster embryo bioassay. Water Quality Criteria, ASTM STP 416, American Society for Testing and Materials, pp. 112-120. Woelke, C. E. (1968). Application of shellfish bioassay results to the Puget Sound pulp mill pollution problem. Northwest Science 42(4), 125-133. Woelke, C. E. (1972). Development of a receiving water quality bioassay criterion based on the 48-hour Pacific oyster (Crassostrea gigas) embryo. Washington Department of Fisheries Technical Report 9, 1-93. Wolfe, D. A., Scott, K. J., Clayton, J. R. Jr., Lunz, J., Payne, J. R. and Thompson, T. A. (1995). Comparative toxicities of polar and non-polar organic fractions from sediments affected by the Exxon Valdez oil spill in Prince William Sound, Alaska. Chemistry and Ecology lO(1-2), 137-156. Wolfe, D. A., Krahn, M. M., Castillas, E., Sol, S., Thompson, T.A., Lunz, J., Scott, K. J., Rice, S. D. and Wright, B. A. (eds) (1996). Toxicity of intertidal sediments contaminated by Exxon Valdez oil spill. Proceedings of the Exxon Valdez Oil Spill Symposium, AFS, Bethesda, MD, American Fisheries Society Symposium 18,121-139. Wood, l? G. (1961). The principle of water sterilisation by ultraviolet light, and their application in the purification of oysters. Fishery Investigations, London series 2, 23(6), 48 pp. Wright, D. A., Kennedy, V. S., Rooseburg, W. H., Castagna, M. and Mihursky, J. A. (1983). Temperature tolerance of embryos and larvae of five species under simulated power plant entrainment conditions: a synthesis. Marine Biology 77, 271-278, Xu, Z., Hong, L. and Zheng, B. (1994). Toxic effects of heavy metals on several
178
E. HIS, R. BEIRAS AND M. N. L. SEAMAN
bivalves and crustaceans. Journal of Oceanography, Taiwan Strait, Taiwan Haixia 13, 381-387. Yakovlev, Y. M. (1978). Reproductive cycle of the Pacific oyster (Crassostrea gigas Thunberg) in the Sea of Japan. Soviet Journal of Marine Biology 3, 85-87. Yonge, C. M. (1960). “Oysters”. Collins, London. Zar, J. H. (1984). “Biostatistical analysis. 2nd edn”. Prentice-Hall, London. Zaroogian, G. E. (1980). Crassostrea virginica as an indicator of cadmium pollution. Marine Biology 58, 275-284. Zaroogian, G. E. and Morrison, G. (1981). Effect of cadmium body burdens in adult Crassostrea virginica on fecundity and viability of larvae. Bulletin of Environmental Contamination and Toxicology 27, 344-348. Zaroogian, G. E., Pesch, G. and Morrison, G. (1969). Formulation of an artificial sea water media suitable for oyster larvae. American Zoologist 9, 1144. Zaroogian, G. E., Morrison, G. and Heltshe, J. E (1979). Crassostrea virginica as an indicator of lead pollution. Marine Biology 52, 189. Zhadan, l? M., Vashchenko, M. A., Medvedeva, V. V. and Gareyeva, R. V. (1992). The effect of environmental pollution, hydrocarbons and heavy metals on reproduction of sea urchins and bivalves. In “Oceanic and anthropogenic controls of life in the Pacific Ocean” (V. I. Ilyichev and V. V. Anikiev, eds), pp. 267-286. Kluwer Academic Publishers. The Netherlands.
Population Structure and Dynamics of Walleye Pollock. Theragra chalcogramma K . M. Bailey'. T. J. Quinn 11'. F! Bentzen3 and W. S. Grant4 'Resource Assessment and Conservation Engineering Division. Alaska Fisheries Science Center, 7600 Sand Point Way NE. Seattle W A 98115 Email:
[email protected]. Phone: 206.526.4243 . Fax: 206-526-6723 2Juneau Center, School of Fisheries and Ocean Sciences, University of Alaska Fairbanks, 1120 Glacier Highway, Juneau A K 99801-8677 3MarineMolecular Biology Laboratory. School of Fisheries. University of Washington. Seattle W A 98195 4ConservationBiology Division. Northwest Fisheries Science Center, 2725 Montlake Blvd., Seattle W A 98112 1. Introduction ......................................................................... 2. Background: The Fishery. Life History and Ecosystem Interactions 2.1. The fishery for walleye pollock 2.2. Life history 2.3. Predator-prey interactions and ecosystem considerations 3 Population Ecology 3.1. Macroecology 3.2. Population dynamics 3.3. Recruitment 4. Population Structure 4.1. Methods for estimating stock structure ...................................... 4.2. Phenotypic population structure 4.3. Genetic population structure 4.4. Metapopulationstructure ..................................................... 4.5. Populationstructuring mechanisms 5. Management Implications Acknowledgements References
.
180
............. 184 ............................................... 184 .................................................................... 185 ................. 187 ................................................................ 189 ................................................................. 189 .......................................................... 195 ................................................................... 201 ............................................................... 206 208 ............................................. 210 ................................................. 217 226 ......................................... 230 ......................................................... 238 ................................................................... 242 ............................................................................. 242
The population biology of walleye pollock. Theragra chalcogramma, is described including its life history. population dynamiq genetic structure
ADVANCBS IN MARINE BIOLOGY VOL. 37 ISBN &12-(126137-5
180
K. M. BAILEY ETAL.
and metapopulation structure. Walleye pollock is an important species in the ecosystems of the subarctic PaciJic Ocean, and is one of the world’s largest fisheries. The population dynamics of pollock is driven by recruitment, which is associated with environmental variability. Management of pollock stocks is based on harvests from large geographic regions. However, lumping stocks within these regions may be adverse to conservation and management goals. Historical genetic studies of pollock have produced some conflicting results and comprehensive genetic studies are needed. A summary view of genetic structure in walleye pollock to date suggests a pattern of geographic stock structure, with varying levels of gene flow between major regions. Phenotypic differences between stocks, elemental composition of otoliths and parasite studies indicate restricted mixing of juveniles and adults. Genetic differences appear between broad regions, but resolution between adjacent stocks, especially within the eastern Bering Sea, is currently lacking. Recent studies indicate genetic differentiation among pollock in the Gulf of Alaska and Bering Sea, possibly resulting from reduced gene flow owing to larval retention mechanisms or strong natal homing. The global population of pollock does not fit into a strict metapopulation framework, but some neighbouring populations may be considered as metapopulations. Whether there is either density-driven migration of strong recruitment cohorts, or population sinks, is controversial and more information is needed. Stock mixing problems can be best addressed by means of high resolution genetic techniques in conjunction with tagging and the use of natural environmental markers. 1. INTRODUCTION
Walleye pollock, Theragra chalcogramma (see cover picture) is a dominant groundfish in many ecosystems across the North Pacific Ocean and plays an important role in the dynamics of higher trophic levels (National Research Council, 1996). For example, about 70% of the groundfish biomass in the eastern Bering Sea consists of pollock (Wespestad, 1993). Pollock is the target of one of the world’s largest fisheries, with annual harvests ranging from 4 to 7 million metric tons (mt) in the North Pacific over the past decade. This species represents about 5 % of the world’s harvest of fishes. In US waters, catches have been in the order of 1.5 million mt, about 40% of US fisheries, with an ex-vessel value of 200 to 400 million dollars and a post-primary processing value of 600 to 900 million dollars from 1992-96 (Kinoshita et al., 1998). It has been suggested that declines in abundance of pollock are associated with declines in the abundances of other groups of animals, such as marine mammals and birds in the North Pacific (Springer, 1992; Merrick et al., 1997).
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
181
Many issues in the management of pollock have not been resolved. These include the delineation of population boundaries and the dynamics of pollock populations, discoveries of new populations, and the effects of fishing on migratory individuals in areas of population mixing. Fishery resource management is based on the concept of sustainable stocks. The idea behind this concept is the identification of groups of fish whose demographics (population statistics) and abundances are largely independent of other groups. Ideally, each group produces a surplus of fish that when harvested does not threaten the ability of the stock to persist, nor does its harvest affect the abundances of fish in other stocks. The identification of these idealized units, however, is often controversial because the motivations of fishery managers may be variously influenced by biological, practical or political considerations (Carvalho and Hauser, 1994). In the case of pollock in US waters, populations in the Gulf of Alaska are presently managed independently of populations in the Bering Sea, as are populations in the eastern and western Bering Sea and in the Aleutian Basin. However, demographically independent sub-populations may exist within some of these areas, for example, there are spawning aggregations within the Gulf of Alaska in Shelikof Strait, Prince William Sound, and the regions around the Shumigan Islands (Figure 1). A further complication is that inshore fisheries in the US are currently managed by state agencies, whereas offshore fisheries are under federal or international jurisdiction. In addition to the identification of population boundaries, little is known about the amount of mixing of fish from different populations on summer feeding grounds, or of the sources of fish for the recolonization of depleted areas. In fisheries, attempts to define local populations date back t o the conceptual developments of Schmidt, Heincke and Hjort in the early 1900s (reviewed by Sinclair, 1988). Information on the spatial structure of populations, their intermixing and the ecology and life-history dynamics of individual populations is vital to formulating a sound policy for the harvest management of walleye pollock, as well as for other species. The identification of population boundaries and the measurement of migration between populations are especially important in pollock because high levels of harvest may rapidly deplete individual segments of a population. Such information can potentially be used to address several questions. If a population is depleted by overharvesting, will it recover quickly by replenishment from neighbouring populations through migraticon of juveniles or adults or through transport of eggs and larvae? How much migration is required for the recovery of a depleted population on a time scale that is of interest to the fishing industry? How can levels of migration between populations be measured? We know that pollock populations in previously glaciated areas have been transient over a scale of tens of
Figure 1 The distribution range (:::) of walleye pollock and major spawning locations ( 0 ) .
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
183
thousands of years, but how stable are populations on shorter time scales? What is the effect of intense harvests on natural cycles of population variability? The role of spatial factors in population dynamics, and especially the importance of habitat subdivision and spatially distributed dynamics is relatively unknown for most animals (Kareiva, 1990). However, a growing body of theory and empirical observations exists in the fields of landscape ecology, macroecology, population dynamics, population genetics and metapopulation dynamics that can be used to address the above questions and to understand the dynamics of walleye pollock populations and guide future research. Individuals are not scattered uniformly within the distributional range of a species, and the theories in these disciplines provide a framework for understanding the effects of population subdivision and demography on the harvest management of walleye pollock. Terminology is somewhat different in ecology and population gen etics. Population sub-units are sometimes referred to as populations, local populations, sub-populations, stocks, or sub-stocks. In this review, we define a population as a group of individuals in an area that is distinguishable from other groups in other areas. Populations may have different demographic trajectories; that is, they may have different age structures that result from different birth, recruitment, death and dispersal rates. Populations may also be reproductively isolated from one another to some degree, and may show genetic differentiation if they have been isolated for a long time. The term stock has been variously used in the literature to designate collections of fish. These collections have variously been defined as fish occurring in a particular locality, for example, a current system or within political boundaries, or merely as fish harvested by a partiicular method. The term stock is sometimes used in place of population and may or may not imply genetic discreteness. Indeed, fisheries scientists and managers usually use the term stock for a population component for which assessment information can be determined and effective management regulations can be developed. We use the word here in a more general sense indicating a group of fish. In this chapter, we review the natural history of walleye pollock with particular reference to population dynamics, including what is known about early life-history stages, demography, and population structure. We attempt to integrate and revise current knowledge linking dynamics and population structure of pollock (Bailey el al., 1997, 1998) as a framework for future research. We have included studies on North Pacific and B,ering Sea oceanography, and on larval life history patterns as they influence population structure and dynamics, as well as studies of the popullation genetics of pollock based on recently developed molecular methods. The scope of the review is to concentrate on pollock stocks most familiar to us,
184
K.
M. BAILEY ETAL.
those of the northeastern Pacific Ocean, but we refer to other stocks, such as those around the Japan Archipelago, when appropriate. Pollock has a relatively short history as a harvested species and has not been studied as thoroughly as other demersal species, for example Atlantic cod, Gudus morhua, which has a rich literature on population and catch histories, and population structure. Thus, comparisons with cod are frequently made. 2. BACKGROUND: THE FISHERY, LIFE HISTORY AND ECOSYSTEM INTERACTIONS
In this section, background information is presented on the fishery and life history of walleye pollock as it applies to population structure and dynamics. The role of pollock as a predator and as a prey in its ecosystem is also reviewed. 2.1. The Fishery for Walleye Pollock
In the waters around Japan, a fishery for pollock existed as early as the eighteenth century, although it remained at a low level until the 1970s (Saito, 1972; Tsugi, 1989). In the northeastern Pacific Ocean, commercial catches of pollock are recorded as early as 1954, but it was the development of at-sea processing of surimi (see below) in the early 1960s that led to a large-scale directed fishery. The fishery in the eastern Bering Sea rapidly expanded to a peak catch of 1.9 million mt in 1972. Bilateral agreements between nations followed soon afterwards and then the passage of the 1976 Magnuson Fisheries Conservation Management Act (MFCMA) extended the US fishery management jurisdiction to 200 miles offshore, and led to regulated harvest levels. In the early history of eastern Bering Sea and Gulf of Alaska fisheries, harvests were made by foreign nations (mainly Japan, USSR, Poland and Korea), but in the 1980s a domestic US fishery developed. Concentrations of pollock were discovered in the 1970s in the central Bering Sea in an area outside the US and USSR exclusive economic zones (EEZs), the so-called “Donut Hole”. Fish in this area were outside national jurisdiction and harvests were unregulated until 1988, when international treaties and cooperative efforts curtailed fishing (Wespestad, 1993). In the early period of development of pollock fisheries in the northeast Pacific, the major gear used was bottom trawls, but targeted pollock fisheries have used mainly large midwater trawls deployed close to the bottom (Megrey, 1989). In Funka Bay, Japan, walleye pollock are caught with bottom gillnets (Kendall and Nakatani, 1992), and in Nemuro Strait, Japan, fish are caught with longlines and gillnets (Yoshida, 1989).
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
185
Pollock is utilized in a varied array of products. In 1996, about 51% of pollock landed off Alaska was used to make surimi, a minced fish product that is further modifed into artifical crab, scallop, or shrimp meat, fish sausage (kameboko) and other products. Another 20% of the landed Alaska catch was used to make fillets, and the remainder was used for roe, other minced fish, whole fish, fish meal and other products. About 8% of the catch was discarded because fish were too small or pollock wals an undesirable bycatch. The overall biomass and catch of pollock on the US side of the North Pacific Ocean is relatively stable, but some populations have experienced declines and fisheries closures in recent years, including stocks in F’uget Sound, Shelikof Strait, the Donut Hole and near Bogoslof Island. More recently, there are reports that the western Pacific stocks are in a state of decline. Eastern Bering Sea continental shelf populations have berm at healthy levels in the past, although there is concern about the sustainability of present harvest levels. Pollock harvests were especially high from the mid-1980s through about 1992 resulting from relatively strong recruitment, high abundance levels and unrestricted high seas fisheries. Wespestad (1996) lists 12 geographically distinct (although not necessarily genetically distinct) stock groupings and their catch trends (Table 1). Biomass and catch trends for the major stock groupings indicate generally declining levels since the late-1980s in the major fishing grounds (Figure 2a and b). The greatest relative declines in biomass and catches for pollock have been away from the centire of pollock’s distribution. For example, in the Gulf of Alaska catches peaked at 307 thousand mt in 1984 and have declined to 55 thousand mt in ’1996. In this region catches are regulated by quotas that reflect biomass levels. At the extreme southern end of its range in the eastern North Pacific, pollock in southern Puget Sound may almost be extinct (Palsson et ul., 1996). Likewise for the southern end of its range in the western North Pacific around northern Japan, recent catches are reduced by three to four times from their maximal values in the 1970s (H. Yoshida, personal communication, 1994). 2.2. Life History
Walleye pollock is one of seventeen gadiform species represented by four families in the northeastern Pacific Ocean and is one of five species iin the family Gadidae (Dunn and Matarese, 1987). There is one other member of the Therugru genus which is 7: jinmurchicus, a small fish infrequently found in the Barents Sea (Pethon, 1989). Most pollock populations spawn at predictable times, in the late winter
186
K. M. BAILEY ETAL.
and early spring, in the same locations year after year, usually in sea valleys, canyons, or indentations in the outer margin of the continental shelf. They are also known to spawn in fjords or deep-water bays (such as Puget Sound) and in some deep-water locations over the Aleutian Basin. Since observations began about 15 years ago the bulk of spawning of the Shelikof Strait population has occurred in a 2-week period in April, within a few kilometres of Cape Kekurnoi. In the Bogoslof Island region in the Aleutian Basin, pollock spawn from early to mid-March, and over the southeastern Bering Sea shelf, most fish spawn from April to mid-May. In Funka Bay, Japan, spawning occurs in January and February (Kendall and Nakatani, 1992). Some populations of pollock apparently spawn at a particular time of year following a spawning migration to a specific location; presumably the eggs and larvae can then reach nursery areas that are favourable for survival (Kim, 1987; Kendall and Nakatani, 1992). This presumed adaptive strategy may vary among populations. For example, in Funka Bay fish spawn near the entrance of the bay in an area where eggs and larvae drift into the bay and in Shelikof Strait fish spawn in an area where larvae are retained on the shelf and are transported into coastal nurseries (Kendall and Nakatani, 1992). Pollock have complex pairing and mating behaviour during spawning (Sakurai, 1982; Baird and Olla, 1991). Female pollock spawn numerous batches of eggs over a relatively short time (Sakurai, 1982; Hinckley, 1987). Eggs are spawned at depth, and in most areas remain deep in the water column (100 to 400m) (Kendall et al., 1994). However, they are shallower in Funka Bay and over the Bering Sea shelf. The deep distribution of eggs presents a problem for interpreting spawning distributions from historical ichthyoplankton surveys that usually sample only to depths of 100 to 200 m. For example, in the Aleutian Basin, most eggs are found at 300 to 400 m (Dell'Archiprete, 1992). Eggs take from 7 to 30 days from fertilization to hatch, depending on ambient temperature. After hatching, the larvae are located in the upper portion of the water column (generally from 20 to 60 m depth); larvae undertake limited die1 migrations (Kendall et al., 1994). Larvae grow relatively slowly (about 0.10 to 0.20 mm.d-l; Nishimura and Yamada, 1984; Kendall et al., 1987; Bailey et al., 1996a). Larvae metamorphose to juveniles at about 18 mm and undergo associated life-history changes (Bailey, 1989; Grover, 1990;Merati and Brodeur, 1996; Brodeur, 1998). Young-of-year juveniles grow about 1mm.d-l, reaching 80 to 100 mm by 6 months and 120 to 140 mm by the end of their first year. Juveniles mature sexually at about age 4 and a length of 40 to 45cm. Pollock can live as long as 20 years and attain a maximum length of 75 cm (Wespestad, 1993).
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
187
Table I Geographical distribution of walleye pollock stocks, modified from Wespestad (1996).
Stock
Stock characteristics
North American
Southeast Alaska-Canada Westem-Central Gulf of Alaska (GOA) Eastern Bering Sea (EBS) Aleutian Basin (AB) Aleutian Islands (AI)
Small stock, minor fisheries Variable stock, 50-200 thousand t catch Large stock, 1-2 mt catch Variable stock (0.1-1.4 mt catch) Small stock, minor fisheries
Asian
Northwest Bering Sea Western Bering Sea East Kamchatka West Kamchatka North Sea of Okhotsk Sakhalin Kurd Islands Japan Sea Japan Pacific
Mix of US and Russian fish, 0.5-1 mt catch Medium stock, 0.5-1 mt catch Small-medium stock, 100-300 thousand t catch Large stock, near 1 mt catch Medium stock, 0.5-1 mt catch Small stock, 65 thousand t average catch Small-medium stock Heavily fished Moderate catch to 0.5 mt
2.3. Predator-Prey Interactions and Ecosystem Considerations
Early-stage larvae feed mainly on copepod nauplii (Nakatani, 1988; Canino et al., 1991), the success of this behaviour is related to survival (Nakatani, 1988; Bailey et al., 1995; Paul et al., 1997). Between larval and juvenile life, pollock become crepuscular feeders (Merati and Brodeur, 1996). Juveniles prey mostly on euphausiids, decapod larvae and copepods (Grover, 1990; Brodeur, 1998). As adults, pollock feed mainly on euphausiids, small fishes, copepods and amphipods but are capable of eating all manner of smaller marine organisms. Cannibalism is a particularly important aspect of adult pollock feeding in the eastern Bering Sea (Dwyer et al., 1987; Livingston, 1989, 1993). In some years, during the autumn and winter, up to 80% of the mean stomach contents of an adult pollock may be composed of age-0 juvenile pollock. Cannibalism is also prevalent in other regions, but is noticeably less so in the Aleutian Basin and Gulf of Alaska. Pollock eggs and larvae are preyed upon by a wide assortment of animals, including euphausiids and amphipods. Small invertebrate predators may consume from 4 to 17% of the total number of eggs present in the water column (Bailey et al., 1993; Brodeur and Merati, 1993). The impact of predation by small fishes on eggs and larvae is not well known
188
K. M. BAILEY ETAL.
16 14 12 E
.-.-
10
E .-
8
E
6
E In
._ m
4
2 0 1984
1887
1970
1973
1976
1879
1982
lQs5
1988
IS91
1994
Year
Figure 2 Trends in major stock groupings of walleye pollock (a) catch trends for Asian and North American stocks, (b) abundance trends for major stocks (after Wespestad, 1996).
(Brodeur et al., 1991; Brodeur and Merati, 1993). Egg cannibalism occurs but is a fairly small component of the total egg mortality, with consumpto 3% of the total egg production tion estimates ranging from 4% (Brodeur et al., 1991). Pollock are prey to many groundfishes and are a critical prey component of the ecosystem both in the Gulf of Alaska and eastern Bering
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
189
Sea (Springer, 1992; Livingston, 1993). Marine mammals and seabirds depend on strong year classes of pollock in the eastern Bering Sea. Juvenile pollock, primarily age 0 and age 1, are the main prey of northern fur seals (about 80% of stomach content is pollock), and feeding on pollock varies with the recruitment level of juveniles (Sinclair et al., :1996). Swartzman and Haar (1983) proposed an interaction between the fishery harvests, cannibalism and fur seal feeding that includes the fishery removing older cannibalistic fish, thus reducing mortality of young pollock and making them more available to fur seals.
3. POPULATION ECOLOGY Geographic range considerations and aspects of macroecology (Brown, 1995) are an important population characteristic of species. Species with broad niches may become both widespread and locally abundant (Brown, 1984), and large ranges, abundance and invasion ability are linked characteristics within a species (Lawton et al., 1994). Species with extraordinary invasion abilities are generally those best adapted for marginal habitats (MacArthur and Wilson, 1967). As described below, these concepts are especially relevant to pollock population biology., Our treatment of population dynamics covers stock assessment, harvest management, and recruitment. Stock assessment is the tool used to detect changes in abundance in the population, estimate natural and fishing mortality and make harvest recommendations. Since this study is key to understanding the pollock population, this section is covered in detail. Recruitment and mortality are the main factors causing changes in abundance in the population.
3.1. Macroecology
Walleye pollock has a broad geographical niche. Pollock is comrrionly associated with the outer shelf and slope regions of coastal waters, but they are capable of utilizing a wide variety of habitats including nearshore eelgrass beds (Sogard and Olla, 1993), large estuaries like Puget Sound, coastal embayments, and open ocean basins such as the Aleutian Basin of the Bering Sea. Adults are often described as semi-demersal, although in some areas they are strictly pelagic (Bakkala, 1993). In some regions pollock is even considered a benthic fish (Tsuji, 1989). As noted previously, pollock commonly feed on a wide assortment of prey from pelagic copepods to epibenthic organisms and pelagic and demersal fishes.
K. M. BAILEY ETAL.
190 Larvae Posehatchage (days) 0
1020
I
I 1
I
40
50
MI
70
I
I
I
I
,
/*
, , , I
I
..
i
I
. J
.,:*. . ' !W : *
-
1 ; :; ; : f i; s 1 1
2
3
4
Egg stages
5
6
4
12
20
I
I
20
I
I
36
,
I
,
44
Larval length (mm)
Figure 3 The spatial patchiness of walleye pollock eggs and larvae as indicated by Lloyd's index of patchiness (after Stabeno et al., 1996a).
Pollock are distributed from Puget Sound to the northern Bering Sea and across the North Pacific Ocean and are most abundant in the eastern Bering Sea and the Sea of Okhotsk. The local abundance of pollock is usually high and they often dominate regional groundfish communities. Given its ecological plasticity, broad range and high levels of abundance, pollock appears to be a classical generalist species capable of invading and adapting to marginal habitats. Spatial patchiness of the early life stages varies with size (and age) (Stabeno et al., 1996). Based on Lloyd's index of patchiness (the ratio of mean crowding of a population to its mean density; Lloyd, 1967), patchiness increases from the egg stage to newly hatched larvae, and tends to decrease through the late larval stage (Figure 3). By the early juvenile stage, patchiness increases again as fish begin to school. Three types of larval pollock patches were described in the above study: those created by the interaction of larvae with time-dependent currents, those associated with eddies, and those associated with geographic structure, such as islands. In the laboratory, the formation and maintenance of larval pollock aggregations has been related to prey patches (Davis and Olla, 1995), and as well, in field studies the distribution of pollock larvae often coincides with that of their prey (Nakatani and Maeda, 1983). The geographic distribution of pollock varies with ontogeny. The distribution patterns of several year classes in the eastern Bering Sea were tracked to examine ontogenetic dynamics in distribution of different year
191
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK -170'00'
-160"OO'
-17O"OO'
-15O"OO'
60"OO'
60"OO'
60"OO'
5O"OO
5O"OO'
5O"OO'
-170"OO'
-17O"OO'
-16O"OO'
-160"OO'
50"OO'
,
-170~00'
-15O"OO'
-150"CIO'
.60"00'
-17O"OO'
-15O"OO'
-16O"OO'
-16O"OO'
-16O"OO'
-15O"OO'
-15O"OO'
60"OO'
6O"OO'
60"OO'
- 6O"OO'
5O"OO'
!5O"OO'
5O"OO
5O"OO'
- 170"OO'
- 16O"OO'
- 150"OO'
-17O"OO'
-160"OO'
-150"OO'
Figure 4 Relative distribution patterns of the 1982 year class of walleye pollock in the eastern Bering Sea: (a) larvae in June from ichthyoplankton surveys, (b) age-0 juveniles in autumn from midwater trawl surveys, (c) age-1 in summer from bottom trawl surveys, (d) age-3 in summer from bottom trawl surveys. 0 shows area of highest abundance. --- shows approximate region surveyed.
classes (Bailey et al., in press). For example, the 1982 year class was .found predominantly in the outer-shelf region of the southeastern Bering Sea as larvae (Figure 4a); as age-0 juveniles they had moved northward and inshore (Figure 4b). As age-1 fish, they had distributed themselves farther northward and also a large portion of the population was found shoreward (Figure 4c). As age-3 fish in summer, a portion of the 1982 year class returned to the southern outer-shelf region, but a large number of fish remained in the northeastern outer shelf (Figure 4d). Similar patterns were observed with other year classes. Overall, these distribution patterns indicate generally northward movements of age-0 and age-I1 fish.
192
K. M. BAILEY ETAL.
However, distributions vary considerably between years and sometimes these age groups move shoreward. By a g e 3 it appears that pre-spawning fish move southward again. Hydrographic fronts, temperature, prey availability and depth influence the distribution of adults and juveniles (Bailey, 1989; Swartzman et al., 1994; Brodeur et al., 1997). The interacting effects of temperature, light and prey levels on the vertical migratory behaviour of pollock have been well-studied in the laboratory (Olla et al., 1996; Sogard and Olla, 1996a and b). In the sea, the geographical distribution of pollock is limited by low bottom temperatures, from 0°C to 2”C, as shown by distribution of commercial catches (Figure 5 ) , and catches of juveniles and adults in research surveys (Francis and Bailey, 1983; Wyllie-Echeverria and Wooster, 1998). However, when bottom temperatures are very low, concentrations of pollock may reside in warmer water above the cold pool (Swartzman et al., 1994), but juveniles are sometimes found in the cold pool (Francis and Bailey, 1983; Wyllie-Echeverria and Wooster, 1998). Kihara and Uda (1969) and Maeda (1972) believed that pollock in the Bering Sea are associated with the Alaska Stream “extension water mass” (temperature 3 4 ° C and salinity 32 to 34 ppt). Detailed studies have not been conducted at high temperatures, but the range of pollock appears to be limited by temperatures of 10°C to 12°C. Changes in spatial distribution, such as patchiness, geographic distribution and ontogenetic changes in vertical distribution play a key role in ecological interactions. Patchiness of larvae may influence the aggregation of predators, a potential density-dependent regulating factor. The mechanisms that cause larval patchiness, such as eddies, may increase retention of larvae in favourable nursery areas. In daytime, juvenile pollock aggregate in dense concentrations near the bottom and are more diffuse and shallow at night (Brodeur and Wilson, 1996). Aggregating behaviour of juvenile pollock has been shown from laboratory studies to have energetic implications that may influence growth rates (Ryer and Olla, 1997). The geographic distribution of juveniles may affect cannibalism as a result of horizontal overlapping of juveniles and adults, and thus be a factor in recruitment (Francis and Bailey, 1983). The vertical overlap of juveniles and adults, influenced by water column structure and prey availability, is also an important factor in cannibalism (Bailey, 1989). Characterization of pollock as a colonizing species lends some support to suggestions that the rapid increase in pollock occurring in the late 1960s was related to the harvesting of Pacific ocean perch (Somerton, 1978) and Pacific herring, thus reducing competition and initiating an “ecological release” (MacArthur and Wilson, 1967). Pollock could be considered a classic r-selected species with opportunistic rapid growth, early maturity and high fecundity that is able to rapidly occupy a niche opening.
193
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
60N
55N
50N
60N
55N
50N
Figure 5 Monthly catch rates by Japanese fishing vessels in 1"X 1" statistical squares during June 1976 and June 1978. Shaded area is the boreal bottom water mass (sometimes known as the cold pool, less than 2°C and 31-32 ppt). 0 represent catches >lo00 mt and 0 are catches between 16999 mt).
194
K. M. BAILEY ETAL.
In the Gulf of Alaska, an increase of pollock in the mid-1980s coincided roughly with a regime shift occurring in 1977 to 1978 (Hollowed and Wooster, 1995) and possibly was associated with good conditions for juveniles of the 1976 to 1978 year classes. Recently, however, pollock stocks in the Gulf of Alaska declined markedly until 1995 with a small increase since then. Curiously, the age of first maturity has increased (Megrey, 1988) in spite of declining density. Since the mid-1980s juvenile survival has been relatively poor; from 1980 to 1985 age 0-2 mortality was lower than the 1980 to 1991 mean value for five of six year classes. By contrast, from 1986 to 1991, five of six year classes had higher than the mean long-term mortality (from data in Bailey et al., 1996~).These data tend to indicate trends of increasing predation pressure on juveniles or eroding environmental conditions for juveniles and adults. Range expansions may be limited by physical impediments, such as temperature, salinity and substratum availability, and by biological factors including the presence of competitors and predators. There is also historical structure in the environment, such as changes in the occupation of niches owing to disease and environmental events. Some populations may expand their range as they become more abundant, although others do not show this trend but show increases in local density. For example, species with highly specialized niches may not expand readily compared with species that have more generalized requirements. Ontogenetic changes in the relationship of range expansiodcontraction and abundance may also exist; for example, Schneider et al. (1997) found that the range of juvenile Atlantic cod did not contract with decreasing abundance, whereas the opposite is true for adult cod (Swain and Wade, 1993; Atkinson et al., 1997). The relationship of distributional range and population abundance has not been formally examined for most pollock stocks, as historical records over a long time span and over the whole range of pollock are not currently available. However, based on the historical declines of abundance in populations outside the major fishery area, while the central area of pollock biomass has remained relatively stable, pollock can be characterized as a species with a main central range and numerous fringe populations; the overall range of the population may not shrink with population decreases unless local fringe populations become extinct. Within local populations there may be a positive range and abundance association, similar to that found for Atlantic cod. Tsugi (1989) asserted that during times of increasing commercial catch levels (and therefore abundance) pollock expand into adjacent waters. Likewise, Stepanenko (1997) reports range expansions of pollock in the Bering Sea related to increasing abundance and warming temperatures. However, range con-
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
195
tractions with declining abundance may not occur for juvenile pollock due to their dispersal potential. Instead, the range extent may be controlled by transport and other environmental factors. Wyllie-Echeverria (1995) related invasions of juvenile pollock into the Chukchi Sea with warm ocean conditions. As described below, density-driven migrsations into new or previously colonized habitats have been proposed, but definitive data supporting these movements are lacking. 3.2. Population Dynamics
Most knowledge about walleye pollock on a stock-scale comes from stock assessment information and modelling. This information is used to monitor the status of the stocks and to evaluate the effects of fisheries upon them. Detailed history of the pollock fisheries and stock assessment procedures are found in Bakkala et al. (1987), Quinn and Collie (1990), Marasco and Aron (1991), Hollowed et ai. (1997) and Wespestad et al. (1997). Analysis of the population dynamics of pollock involves using fisheries survey data and catch-at-age population models that make some assumptions about stock structure and natural mortality to estimate trends in fishing mortality, recruitment and population abundance. There are two types of survey used in stock assessment of pollock: a hydroacoustic (sonar) survey assesses the midwater component o f the population and a bottom trawl survey assesses the near bottom (within 3 m off bottom) component. Surveys provide a consistent sampling of fish from year to year, give information on abundance in areas not commercially fished, and contribute information on small, pre-recruit fish that are not caught by the commercial fishery. National Marine Fisheries Service (NMFS) survey data used in stock assessment date back to the 1970s. Hydroacoustic surveys occur annually in the Bogoslof area and in Shelikof Strait. The eastern Bering Sea is surveyed with hydroacoustics every 3 years. In hydroacoustic assessment, electrical energy from a transmitter is converted into acoustic energy by an underwater tramducer. The energy reflected from fish is converted back into electrical energy and the signal is processed. Generally, NMFS survey transect lines are run 5 to 10 nautical miles apart. Midwater trawls are used to collect samples for species identification and length and age composition. Typically, for NMFS bottom trawl surveys stations are located 20 nautical miles apart. The gear is a US east coast-type otter trawl with a 31.4 m long footrope, equipped with a small mesh liner (3.2 cm str,etched mesh). The catch from each tow is sorted by species, weighed and counted. Otoliths are collected for age and growth information. Total biomass is
196
K. M. BAILEY ET AL.
estimated using an area-swept method (length of the net opening multiplied by the distance the net is towed). The density of fish from all survey stations is averaged and extrapolated to the surveyed area to provide a total biomass estimate. In the Bering Sea, bottom trawl surveys are annual, while in the Gulf of Alaska, they are triennial. Stock assessment from the commercial fishery is based on catch-at-age data. Standard virtual population analysis (VPA, also known as cohort analysis) (Pope, 1972) was originally used. VPA is a back-calculation procedure where catch-at-age data from the fishery is used in Baranov catch equations (Ricker, 1975) along with assumptions of natural mortality and terminal fishing mortality to estimate to numbers-at-age, exploitation rates and biomass levels over time (Quinn and Deriso, 1999). A basic assumption of VPA methods is that the catch-at-age data are measured without error. More recently, statistical catch-at-age models, summarized in Quinn and Deriso (1999) have been utilized. In the statistical age-structured (SAM) and stock synthesis models, the catch-at-age in any given year is assumed to be measured with an amount of error. The stock assessment model contains parameters for recruitment, fishing and natural mortality, selectivity of the gear, and possible stock-recruitment relationships. These stock assessment methods are also fitted to the catch, survey and other information to estimate parameters. In the more recent forms of the analysis hundreds of parameters are estimated by the model, mostly dealing with year-to-year and age-specific deviations in coefficients such as selectivity and catchability. The main advantage of the current stock assessment models is their ability to integrate the various sources of information into a single framework. For the purpose of stock assessment and management, walleye pollock in the US EEZ is divided into four stocks: eastern Bering Sea (EBS), Aleutian Islands (AI), Bogoslof Island-Aleutian Basin (AB) (Figure 6), and Gulf of Alaska (GOA), (Wespestad et al., 1997; Hollowed et al., 1997). Catches in the eastern Bering Sea have been far higher than those in other regions, except during a period in the late 1980s when Aleutian Basin (Donut hole) catches were very high. Different assessments are made for these stocks because of differences in spawning time, weight-at-age, fecundity and other characteristics. Furthermore there are seasonal allocations in some fisheries. There are assessment problems related to stock structure for each of the stocks. The stock structure of the Bering Sea as a whole is poorly understood for assessment purposes (Wespestad et al., 1997). On the Russian side, the western Bering Sea shelf has a unique walleye pollock stock of lesser abundance than the eastern Bering Sea. In the northern Bering Sea (mainly in Russia), there is thought to be intermingling of
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
197
3000000
2000000
.-c% f- : m c
1000000
i
.
0 1984 1966 1968 1970 1972 1974 1976 1978 1980 1982 1984 1986 1988 1990 1992 1994 1996
’
Year . _
---
Figure 6 Time trends of catches of walleye pollock in waters off the coast of
Alaska. eastern and western Bering Sea pollock. With current large catches coming out of this area, there is concern about impacts on the eastern Bering Sea population. For current management considerations, the Aleutian Basin (or Donut Hole or Central Bering Sea) is thought to contain fish originating both in the east and the west (Wespestad et al., 1997). These fish can be either a separate stock or stocks with some intermingling with the shelf populations or a spillover of immigrants from strong cohorts (such as 1978,1982 and 1984) from the shelves. Some component of the Aleutian Basin “stock” travels to the vicinity of Bogoslof Island each year. Large fislheries in the Aleutian Basin and Bogoslof region in the late 1980s, led to substantial reductions in these populations. Recent surveys have shown little biomass in the Bogoslof or the Aleutian Basin proper. One exception was a relatively large hydroacoustic survey biomass of 1.1million mt in the Bogoslof region in 1995, which was confirmed by a replicate survey. It is suspected that some rebuilding of this population occurred as a result of a strong 1989 year class. The Aleutian Islands region, a broad area stretching from 170W to 170”E, contains a small pollock population (Wespestad et al., 1997). Trawl surveys taken at roughly three-year intervals (1980,1983,1986,1991,1994,
198
K. M. BAILEY ETAL.
1997) covers this area during the summer. The winter-roe fishery has been moving farther to the west and possibly catches Bogoslof fish as well as Aleutian Islands fish. Wespestad et al. (1997) suggest that much of the 1994 Aleutian Islands catch was Bogoslof fish. The proximity of eastern Bering Sea fish to the Aleutian Islands region and the variable movement of eastern Bering Sea fish create further problems for the Aleutian Islands assessment, in that data in this region contain contributions from both the Aleutian BasidBogoslof and eastern Bering Sea stocks. The Gulf of Alaska pollock population is subdivided into WesterdCentral and Eastern Gulf stocks (Hollowed et al., 1997). Little information is available about the latter so recommendations are inferred from the WesternKentral Gulf based on relative biomass distributions in the two areas. The most recent assessment has shown a large increase in biomass owing to strong appearances of the 1988/89 and 1994 year classes in the WesterdCentral Gulf. The 1989 year class did not appear there until 1994 but has been steady since then, raising the possibility of migration from the eastern Bering Sea (which had an exceptionally strong 1989year class), among other possible reasons. Alternatively, a portion of the strong adjacent 1988 year class has been mis-aged as these fish have become older than 6 years of age. Recommended catches are partitioned among the Shumagin, Chirikof and Kodiak areas in relation to relative biomass in the most recent bottom trawl survey. These distributions are variable over time. Another stock issue relates to a population of pollock discovered in Prince William Sound (PWS). It is not known whether these fish are accounted for in the regular bottom trawl survey as a part of the Westendcentral or Eastern stock or whether they are a separate stock component altogether. A stock/spatial issue emerges in these assessments relative to marine mammal and seabird populations. In the eastern Bering Sea, a Catcher Vessel Operational Area (CVOA) was installed in 1992 to provide catcher vessels with a better opportunity to harvest pollock (compared to larger factory trawlers and motherships). This action has tended to increase fishing effort in areas adjacent to endangered Steller sea lion and seabird populations. At the current time, there is little scientific evidence about the effects of concentrated fishing in time or space on pollock, marine mammal or seabird populations. Nevertheless, concern has been expressed and further management actions may be taken to reduce any risk. A multiple-area age-structured model was developed to provide abundance and exploitation information for the entire Bering Sea (Quinn and Wespestad, unpublished reports to the International Pollock Workshops 1992, 1994). The model is based on a back-calculation procedure similar to cohort analysis. It has three regions (western Bering Sea, Aleutian Basin, and eastern Bering Sea) with interchange between the western
.
199
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
-
1.5E + 07
-EBS
1.OE + 07
A
--
I
m m
0
5.OE
I O.OE
Basin Elomass
-8-Bogoslof Stmey Biomass Model Biomas
5
C
+ 06
0
EBSSurvey Biomass WBS Mod~l Biomass Japanese ( F U E
I
I
+ 00 1978
1980
1982
ISM
1-
1988
1990
1992
Figure 7 Biomass estimates from a multi-area, cohort analysis model (Quinn and
Wespestad, unpublished data). Some of the data series used in fitting the model are also shown: biomass estimates from the Bogoslof hydroacoustic survey (expected to be 77% of the modelled Basin biomass shown), biomass estimates from the eastern Bering Sea triennial trawVhydroacoustic survey, and Japanese CPUE data from trawlers operating in the Basin (adjusted by the model's estimated catchability for this gear).
Bering Sea and Aleutian Basin and the eastern Bering Sea and Aleutian Basin. Biomass in Bogoslof is assumed to be a constant proportion P E of the Aleutian Basin biomass and is also interpreted as the proportion of Aleutian Basin fish which originated in the eastern Bering Sea. Migration is defined in terms of the proportion of Aleutian Basin fish at age a which moved from the shelf to the Aleutian Basin at the end of the previous year. This proportion is estimated within the model as a linear function starting at the value 1 at age 5 and ending at the value 0 at age 1:l. The rationale for this function is that Aleutian Basin fish are older than shelf fish and do not appear in the Aleutian Basin before age 5 and that there does not seem to be much recruitment of older fish to the Aleutian Basin. The model integrates catch, (catch per unit effort) and survey information from all three regions. Estimates of biomass (ages 3-t) shown in Figure 7 show a maximum Aleutian Basin population of about 7 million mt in 1986 declining to under 1million mt by 1992. In contrast the eastern Bering Sea population was at least twice as large in 1986 and its relative decline was much smaller than that for the Aleutian Basin. The western Bering Sea population was about one-half the size of the Aleutian Basin population in 1986 but became larger than the Aleutian Basin population in 1990,
200
K. M. BAILEY ET AL.
because its biomass has been fairly stable. The great decline in the Aleutian Basin population is attributed to a much larger exploitation rate than those on parts of the shelf. Biomass estimates from the model are quite sensitive to choices for model parameters and several combinations of parameters explain the data comparably. Thus, the single-area assessment models are considered superior at the present time, pending better information on stock structure, movement and natural mortality. Natural mortality for eastern Bering Sea pollock is traditionally assumed to be 0.45 at age 2 and 0.3 at older ages. In the Bering Sea multi-area model, natural mortality of 0.9 at age 1, 0.45 at age 2 and 0.2 at other ages was assumed. A lower natural mortality is needed in the multi-area model because part of natural mortality in the traditional model is essentially the loss of fish from the shelf to the Basin. For the Gulf of Alaska, the traditional choice for natural mortality is 0.3 for ages 2 and older. A multi-species predation model has been developed for Gulf of Alaska pollock to investigate this assumption (Hollowed et at., 1997). Using consumption rates of pollock by arrowtooth flounder, Pacific halibut and Steller sea lions, the authors extended the use of the stock synthesis model to include predation mortality. Their results showed that estimated natural mortality ranged from 0.63 to 0.99 at age 1, to 0.25 to 0.49 at ages 3 and older. However, their results were based on limited data about consumption rates and were viewed as preliminary. Nevertheless, their results show that natural mortality is probably not constant by age or by year. Furthermore, changes in natural mortality are confounded with changes in recruitment, so better understanding of multi-species interactions would improve understanding of pollock population dynamics. In particular, current high abundance of flatfishes in the Gulf of Alaska and eastern Bering Sea may be influencing biomass and recruitment of pollock. Fishing mortality is well understood for the eastern Bering Sea and Gulf of Alaska pollock populations because catch and biomass are both measured directly. Catches are kept low via conservative procedures used by the North Pacific Fishery Management Council, wherein scientists make recommendations about conservative harvest levels and managers recommend allowable catches under these levels. For Gulf of Alaska pollock, fishing mortalities for fully recruited fish has generally been less than 0.25 (Hollowed et al., 1997), meaning that average exploitation rates have been of the order of 10 to 15% or less. For eastern Bering Sea pollock, average exploitation rates have ranged between 7 and 22%, below the maximum sustainable yield (MSY) level of 30% (Wespestad et al., 1997). Furthermore, these rates are lower than those found in several other fisheries on gadids (Wespestad et al., 1997).
201
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK 4.5
,
1
2.5
4
2
3.5
-.-.-0 CI
c
- 3
E
.-
1.5
z
2.5
.-
2
8
t
E
6
.-a
1 .-P c
B
1.5
v)
1
0.5
0.5 0
0
1967
1972
1977
1982
1987
Year Class
Figure 8 Year class abundance as age-2 recruits (bars) compared to spawning biomass (line) of walleye pollock in the Gulf of Alaska, showing the dramatic effects of year class strength on population biomass.
3.3. Recruitment For all of the major pollock groupings stock fluctuations are strongly influenced by intermittent recruitment of strong year classes. For example, Figure 8 shows the ‘‘logged’’ impact of a series of strong year classes on stock abundance in the Gulf of Alaska, as well as the drop in abundance related to subsequent recruitment of relatively poor year classes. In the eastern Bering Sea the 1978 year class comprised 67% of the pollock population in 1981 and 53% of the population in 1982 (Figure 9). Many regions share the same strong year classes, for example 1978 was a strong year class in the Gulf of Alaska, Aleutian Basin, eastern Bering Sea, western Bering Sea and Sea of Okhotsk. Likewise 1982, 1984 and I989 were strong across the Bering Sea although not necessarily in the Gulf of Alaska. Strong year classes in the Gulf of Alaska including 1976, 1977, 1979 and 1988 did not appear strong in the Bering Sea. Therefore, although some strong year classes are shared, there is not a consistent association of strong year classes among the Bering Sea and Gulf of Alaska populations that would clearly indicate density-dependent dispersal between these geographic regions or large-scale conditions favourable
202
K. M. BAILEY ETAL.
35 30
25
1
p 20
.-
P
.=a 2 8 a
15
10
5 0 1964
1969
1974
1979
1984
1989
Year class
Figure 9 Year class abundance as age-2 recruits (bars) compared to spawning biomass (line) of walleye pollock in the eastern Bering Sea.
to recruitment in all areas. Within the Bering Sea, there appears to be an association of strong year classes among the different regions. The occurrence of similar strong year classes across the Bering Sea has been cited as evidence of panmixia within the Bering Basin (Dawson, 1994). However, Francis and Bailey (1983) showed some evidence for shifts in dominance of year classes between north and south portions of the eastern Bering Sea shelf. At the extreme ends of the range of pollock (e.g. Puget Sound), the year classes of good recruitment (1972 to 1975) were quite different from those in the Gulf of Alaska and Bering Sea. Age-specific life tables for walleye pollock in the western Gulf of Alaska for the 1980 to 1991 year classes were compiled to perform exploratory key factor analyses (Bailey et al., 1996~).Early larval mortality was significantly correlated with generational mortality (-In(Recruits/Eggs)), but patterns in juvenile mortality were also similar to generational mortality, and in some years were clearly dominant in determining the fate of a cohort (Figure 10). Density-dependent mortality was indicated only for the late larval to early juvenile stage. Time trends in juvenile mortality were associated with the increasing abundance of arrowtooth flounder, a major predator. These authors proposed that pollock recruitment levels could be influenced at any life stage, but
203
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
PerallK
1
Juvenile k k
! =Larval
Year Class Figure 10 Results of key factor analysis, comparing annual values of generational mortality ( K ) , larval mortality (kz + k3) and juvenile mortality (k5 + k6) (from Bailey et al., 1996~).
depends partly on a sufficient supply of individuals from earlier stages of development. Forecasts of future abundance and biomass in the eastern Bering, Sea and Gulf of Alaska are currently made by a linear prediction of recruitment from relative estimates of age-1 abundance from the bottom trawl survey and hydroacoustic survey respectively. Recently, these fisheries have been supported by only a few strong year classes, which has made forecasts of total abundance more dependent on the forecasts of recruitment. This approach is advantageous in that a direct measurement of the year class at age 1 is used and the linear relationship has lbeen strong based on R2 values. Disadvantages include the presumably large measurement errors in age-1 abundance from a bottom trawl survey and the likelihood of inter-annual variability in the linear relationship. These problems stem from a finding that pollock at age 1 in the Bering Sea are only partially vulnerable to bottom trawl gear, because the survey is conducted in deep water, although many young pollock typically reside higher in the water column (Smith, 1981). In the Gulf of Alaska, the survey is conducted to assess abundance of the spawning stock, and only a portion of the age-1 stock may be assessed if their distribution does not overlap completely with spawners. The major factors that affect recruitment to a fish population
204
K. M. BAILEY ETAL.
20 4
+ Observed Expected -Standard
1978
h
16
14
OY 0
I
1
2
3
4
5
6
7
8
9
10
Spawning Biomass ( a t )
Figure 11 Ricker spawner-recruit relationship for pollock in the eastern Bering Sea without and with environmental data. The predicted recruitment using environmental data is shown as a smoothed line. The strong 1978 year class is noted (after Fair 1994).
are egg production effects, biological effects on early life survival, environmentaYoceanographic effects on early life survival (Heath, 1992) and overlapping of juvenile fish with their predators. The first factor relates to maturity, fecundity and spawning characteristics. Spawning biomass can be used as a proxy for egg production as long as fecundity is roughly proportional to fish weight. Biological effects on early life history are generally referred to as density-dependent effects and include such mechanisms as cannibalism, density-dependent predation, and within-species competition for food and/or space. Such effects can be explored by fitting spawner-recruit relationships. Since adult pollock in the eastern Bering Sea are known to be cannibalistic (Dwyer et al., 1987; Livingston, 1993), there should be a dome-shaped or Ricker-like spawner-recruit relationship if the effect is strong (Quinn and Deriso, 1999). Early analyses showed that a Ricker relationship can be fitted to eastern Bering Sea pollock spawner-recruit data (Quinn and Collie, 1990; Wespestad, 1995; Quinn and Niebauer, 1995), although the spread of data around the modelled curve is broad. Figure 11 shows a fit of the Ricker relationship for estimates of eastern Bering Sea pollock recruitment and spawning biomass obtained from catch-age analysis using data through 1994 (after Fair, 1994). The Ricker fit is statistically significant, suggesting that density dependence is present.
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
205
The most recent stock assessment (Wespestad et al., 1997), however, did not show a strong Ricker relationship with data up to 1997. Part of the problem is a result of using only part of the time series (data starting in 1978 since there is some question of the accuracy of prior data), but some of the discrepancy is owing to recent strong year classes in 1989 and 1996 being spawned from somewhat large spawning biomasses. Perhaps recent compression of the age composition of the pollock population has made cannibalism less of a factor in recent pollock dynamics. In the Gulf of Alaska, no clear spawner-recruit relationship emerges, and it is of interest that cannibalism is not a major factor there. The effects of density dependence on pollock dynamics were modelled by Wespestad and Quinn (1996). They performed deterministic, retrospective simulations of the eastern Bering Sea population under different levels of density dependence and fishing mortality. Their results showed that fishing has little effect on pollock recruitment, but when it did have an effect it tended to increase recruitment. The reason for this phenomenon in the simulations is that very high levels of pollock spawning biomass would result in decreased recruitment because of increased mortality from cannibalism. Environmental and oceanographic effects on recruitment are also well studied. Larval mortality has been loosely linked to temperature (Bailey et al., 1996c), and storms (Bailey and Macklin, 1994). Larval growth and survival has also been associated with prey levels for early feeding lairvae (Haldorson et al., 1989; Canino et al., 1991; Bailey et al., 1995; Paul et al., 1997). Better environmental conditions may thus be conducive to better development and growth, and perhaps greater ability to avoid predation. But mismatches between the emergence of larval pollock and their food (Brase, 1996) and episodes of catastrophic mortality (Bailey et al., 1995) have also been shown. Since the bulk of spawning in some stocks, such as Shelikof Strait, takes place over a short 2-week period, an episode of high larval mortality can have a significant effect on the survival of an annual cohort. Correlation studies indicate that environmental factors appear to significantly affect eastern Bering Sea pollock recruitment (Quinn and Niebauer, 1995). Seven environmental variables were used in the analyses: air temperature, bottom temperature, sea surface temperature, ice colver, wind, the Southern Oscillation Index and the Pacific North American pressure index. High pollock recruitment coincided with above average air temperatures and reduced ice cover when pollock are about age 1, suggesting that oceanographic conditions associated with warmer temperatures in the Bering Sea during the early life history are conducive to higher recruitment (Quinn and Niebauer, 1995). Ohtani and Azurriaya (1995) found similar results with temperature during the first winter and
206
K. M. BAILEY ET AL.
fecundity closely related to recruitment. These results are in accord with the current understanding of the physical processes of the eastern Bering Sea (Niebauer and Day, 1989). As these processes are atmospherically driven, air temperature is likely to be a proxy variable representing those physical processes, which follows empirically that high correlations occur among these environmental variables. Francis and Bailey (1983) proposed a complex recruitment process where the interaction between temperature over the shelf and spawning location puts eggs and larvae in different current regimes, affecting larval transport patterns, and the overlap of cannibalistic adults and juveniles. Consistent relationships between environmental variables and pollock recruitment occur when using monthly, quarterly or annual environmental breakdowns. The most consistent relationships occur with annual averages of environmental data, suggesting that an integration of the effects of environmental variables occurs in determining pollock recruitment. Figure 11 shows predicted recruitments from fitting a generalized Ricker spawner-recruit model including environmental data, showing that both biological and environmental factors appear to be important. Similar studies with environmental and early life history variables have been conducted in the Gulf of Alaska and other regions as well. Megrey et al. (1995, 1996) showed that pollock recruitment in the Gulf was correlated with precipitation (as an indicator of eddies), an index of atmospheric sea-level pressure gradient, and wind mixing energy. Balykin (1996) correlated recruitment indices of pollock in the western Bering Sea to adult stock abundance and temperature during the first year of life, and Vasil’kov and Glebova (1984) associated recruitment variations of pollock of western Kamchatka to thermal variability during egg and larval stages. 4. POPULATION STRUCTURE
On short time scales, differences between populations may arise because of environmental influences on larvae and juveniles. Elemental composition of hard parts, parasite load, morphology or meristics may differ between areas because of differences in temperature, elemental composition of sea water, food and growth. Many of these differences result from plastic responses to environmental variability and may reflect only short-term population structure that is not related to the level of genetic differentiation among populations. Genetic differentiation is determined by rates of mutation, random genetic drift, natural selection and gene flow between populations, and is influenced as well by demographic structure (such as the type of connections among populations, population size,
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE
POLLOCK
207
extinction rates, etc.). Although mutation is the ultimate source of genetic diversity, recurrent mutation by itself is not a strong force in creating differences between populations. When the number of breeding individuals in a population is small, random drift may lead to genetic changes because of the incomplete representation of parental genes in offspring. Recent studies of marine organisms indicate that population sizes may be highly variable for many species because of high larval mortality and because variability in ocean currents and food availability leads to highly variable recruitment (Hedgecock, 1994a and b). When the reproductive variance among families is large, even species with large census population sizes may have small effective population sizes. Random drift in populations of these species may therefore be important in creating differences between populations. Natural selection may lead to different results depending upon the kind of selection operating. For example, selection for local adaptation in low gene flow species may produce a mosaic of genetically distinct populations in coarse-grain environments, but the same mode of selection may produce genetic homogeneity in high gene flow species which experience the environment as finely grained. The genetic variants detected by most molecular techniques appear to be quasi-neutral relative to levels of natural selection and are therefore useful for estimating rates of gene flow between populations under some circumstances (see Waples, 1998). Gene flow tends to reduce genetic differences between populations and to counter random drift by homogenizing allelic frequencies between populal.ions and by increasing the effective size of local populations. An important consideration in the use of genetic methods for stock identification for management is that only a small amount of gene flow is needed to produce genetic similarity among populations that nevertheless may be demographically distinct (Waples, 1998). Many of the early models of genetic population structure incorporated migration between partially isolated sub-populations, but assumed that sub-population sizes were constant (e.g. Wright, 1931; Kimura and Weiss, 1964). These models have been used to a large extent to interpret the results of empirical studies and to estimate levels of gene flow between populations from molecular data. The importance of local extinctions and recolonizations, in addition to gene flow, random drift and selection, in influencing population structure has been recognized in the last few years and has been incorporated recently into population genetic models (Hanski and Gilpin, 1991). In this view, a species consists of a collection of sub-populations (a metapopulation), which are tied together by gene flow, but which respond to local environmental variability independently of other sub-populations. High rates of local extinction and recolonization may lead to an accelerated loss of genetic variability relative to a group of
208
K. M. BAILEY ETAL.
stable populations and may tend to genetically homogenize the subpopulations (Gilpin, 1991; Grant and Leslie, 1993). 4.1. Methods for Estimating Stock Structure
Several methods have been used to infer population structure. Nongenetic methods measure traits that reflect environmental influences during the lifetime of a fish and are often useful for identifymg short-term population structure. For example, tagging provides evidence of migration from one locality to another, while the analysis of parasites and trace element composition in calcified parts of a fish may provide evidence of separate areas of origin for adults on feeding grounds. Methods that measure phenotypic variables, such as meristics, morphology or behaviour, provide population information that reflects both environmental and genetic influences. Even when purely genetic information is available, understanding morphological and life-history variability is essential for comprehending the mechanisms shaping population structure (Grant et al., 1999). Several molecular genetic methods have been applied to delineating the genetic boundaries of fish populations. Since the 1970s, the electrophoretic analysis of enzymatic proteins has been widely used to infer population structure. This method detects Mendelian variants (allozymes) at a single gene locus and provides genotypic and allelic data to infer breeding structure within populations and to measure the amount of genetic differentiation between populations (see Ryman and Utter, 1987). Analysis of allozymes has been useful for studying levels of divergence on a scale of hundreds to a few million years. However, this method is limited in its ability to infer short-term population structure in species of marine fishes with high levels of gene flow. Small amounts of gene flow tend to produce genetic homogeneity among populations, even though the populations may be demographically independent. The direct analysis of DNA often reveals greater levels of genetic variability within populations than the analysis of the encoded proteins. One class of DNA that has been used to study fish populations occurs as a circular loop in mitochondria and carries a unique set of genes not found in nuclear DNA. Mitochondria1 DNA (mtDNA) lacks recombination during reproduction and is inherited from the maternal parent in fishes, so that gene lineages may be inferred from RFLP (restriction fragment length polymorphism) analysis with restriction enzymes, or from sequences of fragments that have been amplified with the polymerase chain reaction (PCR). In some cases, analysis of mtDNA may be more powerful than the analysis of allozymes because the smaller female effective
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
209
population size and haploid condition is expected to lead to greater amounts of random drift, and hence to larger differences between populations. The analysis of VNTRs (variable number of tandem repeats) shows promise for distinguishing populations. One class of VNTR genes, microsatellites, consists of repeated 1-4 nucleotide motifs with high rates of length mutation. Microsatellite sequences may be the most powerful Mendelian population markers yet developed for detecting short-term population events (Jarne and Logoda, 1996). Microsatellites typically exhibit much higher levels of variability than do other genetic marlkers, such as allozymes and mtDNA. In marine fishes, polymorphic microsatellite loci typically have 10 or more alleles and may have heterozygosities of 50 to 95% (Wright and Bentzen, 1994; Bentzen et al., 1996; Garcia de Le6n et al., 1997). These high levels of polymorphism, which reflect high mutation rates, may confer a greater ability to resolve some aspects of population structure than do the levels of polymorphism revealed by the analysis of allozymes or mtDNA (Bentzen et al., 1996; Ruzzante e,t al., 1996 a and b). Mutation rates at microsatellite loci appear to be three to five orders of magnitude greater than those for non-repetitive DlNA. Mutations are usually a result of replication slippage errors that cause size polymorphisms, rather than point mutations resulting in single base pair changes. The large numbers of low-frequency alleles at a locus, however, point to the necessity of using large sample sizes for population analysis (Ruzzante, 1998). Furthermore, microsatellite variants may be less prone to selection than are allozymes, which increases their value for estimating gene flow rates (Wright and Bentzen, 1994). The identification of demographically independent management units is often complicated by the use of different phenotypic and genetic met hods that individually may be capable of measuring only limited aspects of population structure. For example, vertebral number, which is generally elevated in fish developing at high latitudes in cold water, may indicate environmental conditions during early life history stages, but may not reflect adult migration or gene flow patterns. As a specific example, meristics in Atlantic cod vary significantly among populations on the Newfoundland and Labrador shelves in the northwestern Atlantic (Lear and Wells, 1984; Pepin and Carr, 1993; Templeman, 1981). The results of an analysis of mtDNA variability, however, failed to find genetic differences between these populations (Pepin and Carr, 1993). Yet the analysis of other kinds of DNA, such as mini- (Galvin et al., 1995) and microsatellite loci (Bentzen et al., 1996), and anonymous nuclear loci (Pogson et al. 1995), showed significant frequency differences between samples that were collected over distances as small as several hundred kilometres. The analysis of parasite distributions, elemental composition
210
K. M. BAILEY ETAL.
and morphology may also indicate the existence of stock structure that may or may not correspond to patterns of genetic differentiation between populations. The present-day population structure of a fish may reflect mechanisms and events operating on several different temporal and spatial scales. Some events, such as Pleistocene coastal glaciation, may have led to large-scale geographic differences in morphology, life history or demography that may be reflected in temporally stable genetic differences between populations. Other events, such as annual or decadal shifts in temperatures and current systems, food availability, predation and harvesting may lead to changes in geographic distributions and demography. Some of these population events may produce genetic variability between groups of fish that may be detectable with genetic methods. For example, populations of Atlantic haddock (Melanogrammus aeglefinus) are characterized by a large amount of variability in annual recruitment and by a recent decrease in population size in the northwestern Atlantic most likely resulting from overharvesting (Clark et al., 1982). An analysis of mtDNA variability in archived scales collected in 1975 and 1985 and in fresh samples collected in 1995 revealed significant temporal variability in haplotype frequencies in populations on Georges Bank in the northwest Atlantic (Purcell et al., 1996). These results were interpreted to indicate that fish from genetically distinct populations episodically contribute to the Georges Bank population. Measurable genetic differences have been found on even smaller temporal and spatial scales. Ruzzante et al. (1996b) found frequency differences at microsatellite loci between larval cohorts of Atlantic cod in a retentive gyre of warm mixed water off Nova Scotia. The low level of differentiation, however, indicates that the samples were drawn from different components of a single population and not from different populations. 4.2. Phenotypic Population Structure
Phenotypic characteristics of pollock, as reflected in meristic and morphometric variability, both within small geographic regions and across much broader areas indicate population structure (Table 2). For example, Koyachi and Hashimoto (1977) and Hashimoto and Koyachi (1977) used differences in allometrics, vertebral, gill raker and fin ray counts to distinguish 11 to 12 groups of pollock across its range (Sea of Japan: western Hokkaido, two populations in northwestern Honshu, and Pormorskaya; Pacific coast of Japan: southern Hokkaido, northern Honshu, southern Kuril; southwestern and northern Sea of Okhotsk: Kamchatka Peninsula, eastern Bering Sea, Gulf of Alaska and west coast of Canada).
Table 2 Summary studies of pollock stock structure using phenotypic characteristics.
Author
Method
Area
Results
Ogata (1959)
meristic - vertebral counts
Sea of Japan and Pacific Ocean side of Japan
Iwata and Hamai (1972)
meristic - vertebral counts
Sea of Japan, Okhotsk Sea and Pacific Ocean near Hokkaido
Hashimoto and Koyachi (1969)
northern Japan
Janusz (1994) Temnykh (1991)
morphometrics - body length and other morphological features meristics and morphometrics morphometrics
Sea of Japan has 3 different stocks. Sea of Japan differs from the Pacific Ocean side 8 “local forms” 2 groups in the Sea of Japan 3 groups in the Okhotsk Sea 3 groups in the Pacific Ocean 3 groups discriminated
Temnykh (1994)
morphometrics
western Bering Sea and eastern Kamchatka
Ishida (1954)
morphometrics - otoliths
northern Sea of Japan, Okhotsk Sea and northern Pacific Ocean coast of Japan
Shaw and McFarlane (1986) morphometrics length-at-age
Sea of Okhotsk Sea of Okhotsk
British Columbia - Dixon Entrance, Strait of Georgia
3 stocks distinguished Southern Kurils population distinguished from northern Sea of Okhotsk western Bering and eastern Kamchatka stocks distinguished otolith size is larger in Sea of Japan pollock than Okhotsk Sea. Otoliths are similar between Sea of Japan and Pacific Ocean pollock 2 stocks discriminated - Strait of Georgia pollock are smaller. Little interaction between pollock north and south of Queen Charlotte Sound
Author
Method
Area
Results
Thompson (1981)
morphornetrics length-at-age
Saunders et al. (1989)
morphornetrics, life history
British Columbia - Dixon Entrance, Strait of Georgia and Queen Charlotte Sound British Columbia
L p d e et al. (1986)
morphometrics length-at-age
eastern Bering Sea and Bering Sea basin
Hinckley (1987)
spawning time and location; morphornetrics length-at-age; fecundity
Aleutian Basin and eastern Bering Sea shelf and slope
Mulligan et al. (1989)
spawning time and location
eastern Bering Sea and Aleutian Basin
Serobaba (1977)
morphometrics and rneristics
Dawson (1994)
morphometrics
northern, western, eastern and southern Bering Sea Bering Sea
3 separate stocks - each area contains its own distinct stock. Little mixing occurred between them separate stocks in Strait of Georgia, Hecate StraWDixon Entrance, Queen Charlotte Sound and western Vancouver Island northeastern slope and Aleutian Basin represent 1 stock distinct from other regions of the eastern Bering Sea 3 spawning stocks in the eastern Bering Sea - basin, northeastern slope and eastern shelf and slope 3 spawning areas separated in space and time: eastern Bering Sea southeastern Shelf, eastern Bering Sea northwestern shelf, Aleutian Basin different stocks occupy each region 3 stocks - eastern Bering Sea sheK Aleutian Basin, and Aleutian Islands
Janusz et al. (1989)
meristics and morphometrics
Nitta and Sasaki (1990)
morphometrics
Donut Hole and eastern Bering Sea shelf Donut Hole, eastern Bering Sea, near Japan
Gong et al. (1990)
meristics
Asian and Bering Sea
Wilimovsky et al. (1967)
meristics - fin ray and vertebral counts and morphometrics
entire Pacific Ocean
Koyachi and Hashimoto (1977)
meristics - fin ray, gill raker and vertebral counts
entire Pacific Ocean
2 stocks distinguished in Donut Hole and eastern Bering Sea characteristics distinguish 3 stocks, with about 90% classification accuracy Asian stock and Bering Sea stocks distinguished but stocks within these regions not distinguished morphometric - no strong evidence for discrete stocks meristic - no differences between Bering Sea and Puget Sound pollock. No differences between Gulf of Alaska and northern British Columbia pollock 12 sub-populations, including the Bering Sea and Gulf of Alaska
K. M. BAILEY ETAL.
214
55t
/ .....
40f;p 3!
4
5
6
8
7
9
10
Age
Figure 12 Mean length-at-age for male pollock in different regions of the Bering Sea, 1978 to 1983. Age is in years (after Lynde et al., 1986).
Within much smaller geographic regions there are also distinct groups. For example, around the islands of Japan and Sea of Okhotsk, Iwata and Hamai (1972) identified eight groups, and in the Sea of Okhotsk three stocks were distinguished (Janusz, 1994). Within the eastern Bering Sea including the eastern Aleutian Basin, three to five stocks have been distinguished using morphometrics and life history characters (Hinckley, 1987; Dawson, 1994). Lynde et al. (1986) showed differences in length-atage among fish collected in summer on the southeastern Bering Sea shelf and fish collected in the Aleutian Basin or northeastern Bering Sea slope region (Figure 12). Fish collected in the southeastern slope were intermediate in growth characteristics, possibly indicating mixing, and fish from the northeastern shelf were relatively small when young and large when older, perhaps resulting from ontogenetic migrations of Aleutian Basin fish. Within a fairly small region around Hokkaido Island of Japan, lwata and Hamai (1972) found eight local forms based on vertebral counts. Naturally acquired tags such as elemental composition of otoliths and parasite characteristics indicate restricted mixing among pollock juveniles and adults of different sub-populations (Table 3). For example, the chemical “fingerprint” of otoliths near the nucleus (deposited during early larval life) can be utilized to assign fish to their capture location as juveniles with 70 to 85% accuracy over broad regions of the eastern Bering Sea (Mulligan et al., 1989), indicating limited movement and
Table 3 Summary of studies of pollock stock structure using acquired characteristics.
Author Nakano et al. (1991) Mulligan et al. (1989)
Severin et al. (1995)
Method
Area
otolith chemistry: adults, whole otolith homogenates otolith chemistry: juveniles, inner early life otolith increments
eastern Bering Sea, western Bering Sea, Donut Hole eastern Bering Sea, (southeastern shelf, northwestern shelf, Aleutian Basin) eastern Bering Sea (Bristol Bay), Gulf of Alaska British Columbia - Strait of Georgia, west side of Vancouver Island, Queen Charlotte Sound and Dixon Entrance Sea of Okhotsk, Kommander Islands., Kamchatka Peninsula
Arthur (1983)
otolith chemistry: juveniles, outer otolith increments parasites
Avdeev and Avdeev (1989)
parasites
Miscellaneous Authors, see Figures 14 and 15
tagging studies
western and eastern Bering Sea, Japan
Results differences in 3 areas, little mixing differences in 3 areas, some mixing distinguish 5 areas, some mixing 3 stocks in this area: Strait of Georgia, Vancouver Island and Queen Charlotte Sound/ Dixon Entrance seven distinct groups within the northeastern Sea of Okhotsk, distinct groups in Kommander Islands, eastern Kamchatka and Shirshov Ridge broad movements, homing migrations to spawning site
216
K.
M. BAILEY H A L .
65 N
60 N
55 N
I
I
165 E
/
I
1
I
I
I
1
170 E
175 E
180
175W
17DW
165 W
I
Figure 13 Movement of walleye pollock tagged by Japanese scientists in the Bering Sea (from Dawson 1994). The dashed line represents the US-Russia conventional boundary. The area in the centre of the figure designated by a solid line is the area outside the exclusive economic zone of either country, known as the “donut hole”. Note: most tagging and recoveries occurred during the summer/autumn feeding season.
mixing of fish from different geographic regions. However, results of otolith composition studies should be viewed with some caution because of potential artefacts caused by sample preparation (Proctor and Thresher, 1998). Using parasite frequencies, adult pollock caught on the south side of Vancouver Island can be distinguished from those on the west side with about 75% accuracy (Arthur, 1983). In the Sea of Okhotsk, several different populations of pollock were distinguished based on parasite frequencies (Avdeev and Avdeev, 1989). By contrast, mark-recapture studies where pollock were tagged in summer indicate broad movement of individuals across areas of the Bering Sea (Figure 13). However, critical studies of spawning populations are rare. Tagging results thus far indicate potential for dispersal during the summer feeding period. Historical tagging studies do not indicate whether dispersal patterns in the Bering Sea are part of ordered seasonal migrations, or whether individual tag
217
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK -136"OO'
-138"OO'
-140"OO'
-142"OO'
-144"OO'
-146"OO'
-148"OO'
48"OO'
46"OO'
44"OO'
42"OO'
-136"OO'
-138"OO'
-14O"OO'
-142"OO'
-144"OO'
-146"OO'
-148"OO'
Figure 14 Recapture distribution for an experiment in which 666 pollock were tagged and released (A) in the northern Sea of Japan on 17 April, 1968. Loca.tions and dates of recaptures are shown by ( 0 )(redrawn after Tsugi, 1989).
returns represent fish moving within large schools as migrating populations or as sole immigrants mixing with local populations. However, tagging studies around Japan support a model of dispersed feeding migrations and homing migrations to specific spawning areas (Figure 14; Tsuji, 1989).
4.3. Genetic Population Structure
Different molecular methods are variously suited to detecting population structure on different temporal and spatial scales, and the results of the various genetic studies of pollock must be interpreted in this light (Table 4). On a scale of thousands of years, allozyme studies have detected genetic subdivision across the North Pacific for several marine fishes that apparently was the result of geographic isolation by coastal glaciation during Pleistocene ice ages. During the last ice age, which peaked about 18000 years ago, much of the coastline of south central Alaska and the Kamchatka Peninsula was covered with ice and much of the present-day Bering Sea was dry land (CLIMAP, 1976). Allozyme divergence between
Table 4 Summary of stock structure studies on walleye pollock using biochemical genetics characteristics. Author
Method
Area
Mulligan et al. (1992)
mtDNA RFLP
eastern Bering Sea basin, Aleutian Islands and Gulf of Alaska
Shields and Gust (1995)
mtDNA sequencing
across Bering Sea and Gulf of Alaska
Grant and Utter (1980)
allozyme
southeastern Bering Sea and Gulf of Alaska
Johnson (1977)
allozyme
Iwata (1973)
allozyme
Iwata (1975a and b)
allozyme
Efremov et al. (1989)
allozyme
eastern Bering Sea and Gulf of Alaska northern Sea of Japan and north Pacific coast of Japan northern Sea of Japan and eastern Bering Sea northern Sea of Okhotsk
Powers, unpublished data, see text
mtDNA RFLP and DNA microsatellite
Bering Sea and Gulf of Alaska
Results Aleutian Islands and Donut HoleBogoslof, Gulf and Donut HoleBogoslof have informative differences, but sample sizes are small minor differences between eastern and western Bering Sea minor genetic differences between the two areas. No differences within the areas no significant differences found no differences found significant differences found between the two areas allozyme variability suggesting that aconitase could be genetic marker eastern and western Bering Sea distinguished using mtDNA, Gulf of Alaska and eastern Bering Sea stocks have informative differences using microsatellite DNA
POPULATION STRUCTURE AND DYNAMICS
OF WALLEYE POLLOCK
219
western and eastern North Pacific populations of Pacific herring (Crrant and Utter, 1984), Pacific cod (Grant et al., 1987), and chum salmon (Seeb et al., in review) appear to have originated from this glaciation. FsT (an index of population differentiation, the variance of allele frequencies among populations) values among populations of these species across the North Pacific are 0.15 or larger (Table 5 ) , however, the boundaries of the eastern and western groups differ among species. The genetic demarcations between eastern and western races of Pacific herring and chum salmon are around the Alaska Peninsula, but those for Pacific cod and pollock are in the western North Pacific Ocean or western Bering Sea. An east-west subdivision among pollock populations was detected at a siingle allozyme locus (Sod) in the combined studies of Iwata (1973,1975a arid b) and Grant and Utter (1980) and appears to be located on the Asian side of the Bering Sea or in the Okhotsk Sea. The precise location of the boundary is uncertain, however, because of the paucity of samples in these areas. Genetic population structure arising on shorter time scales may als80be apparent in the distributions of allozyme frequencies. This variation is difficult to measure, however, because the magnitude of sample error is often the same as the level of genetic differentiation between populations of high gene-flow species (Waples, 1998). Within the eastern North Pacific race of pollock, no significant allele-frequency differences were found between samples collected within the southeastern Bering Sea or within the Gulf of Alaska (Grant and Utter, 1980). However, a small, but significant, amount of allele-frequency heterogeneity was detected between populations in the southeastern Bering Sea and the Gulf of Alaska. FsT among samples within each region was 0.021, and is typical of values for several other marine fishes (Table 5; Waples, 1998) with apparently high equilibrium levels of gene flow between populations. Although the analysis of mtDNA is expected to reveal a greater amount of genetic population structure because of its maternal inheritance, little genetic structure has been found between populations of pollock in some studies of mtDNA variability. Mulligan et al. (1992) found 65 RFLP haplotypes in 168 fish from four localities: 1) the Gulf of Aliiska, 2) the Donut Hole in the middle of the Bering Sea, 3) Bogoslof Island in the southeastern Bering Sea, and 4) Adak Island in the Aleutian Archipelago. Two haplotypes occurred in 36% of the individuals and 51 haplotypes were represented by a single fish. Monte Carlo chi-square tests and cluster analysis of sequence divergence showed that the sample from Adak Island was distinct from the other three samples, which were. not distinct from each other after correction for multiple tests (Figure 15). Haplotypic diversity was 0.918, and FST among samples was 0.019 (Table 6), which is similar to that estimated from allozyme frequencies.
Table 5 Estimates of allozyme Hs (the mean sub-population heterozygosity over all populations), and population differentiation FsT (the variance of allele frequencies among populations (From Seeb et al., in preparation)).
No. of samples
Mean sample size
No. of loci
HS
HT
FS T
Walleye pollock
14
44.4
28
0.045
0.046
0.021
Yellowfin sole
3 16
93.3 68.9
31 31
0.089 0.051
0.090 0.053
0.004 0.043
Pacific herring Western Pacific
21 11
95.7 101.5
40 40
0.083 0.073
0.159 0.012
Eastern Pacific
10
89.3
40
0.098
0.007
Pacific ocean Perch
27
53.1
25
0.069
0.023
Pacific cod
11
71.2
41
0.025
9
75.2
41
5 9
95.6 97.8
19 10
Species
Atlantic cod
0.032
0.189 0.031
0.071 0.021 0.014
Geographical range
Reference
1
Southeastern Bering Sea, Gulf of Alaska same Japan, Bering Sea, Gulf of Alaska North Pacific Japan to eastern Bering Sea Gulf of Alaska to California Eastern Bering Sea to Washington Westem-eastern North Pacific Bering Sea to Gulf of Alaska North Atlantic
7
(recalculated)
8
2 3
4
5 6
1. Calculated from data in Grant and Utter (1980); 2. Seeb, unpublished manuscript; 3. Grant et al. (1983); 4. Grant and Utter (1984); 5. Seeb and Gunderson (1988); 6. Grant et al. (1987); 7. Mork er al. (1985); 8. Pogson et al. (1995).
221
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
r
Gulf of Alaska
4
Donut Hole
Bogoslof Island
I
I
0.6
A
d
1
0.45
a
I
0.3
k Island
1
0.15
1
0.0
Divergence
Figure 15 UPGMA clustering of genetic distances among walleye pollock stocks in the eastern Bering Sea and Shelikof Strait (after Mulligan et al., :1992).
Nucleotide diversity (n-) was 0.75. This relatively low level of divergence among haplotypes is typical of species of marine fishes that have had recent population origins (Grant and Bowen, 1998). A second study of mtDNA also points to low levels of genetic differentiation among pollock populations. Shields and Gust (1995) examined nucleotide sequence variability in a 76 base pair (bp) spacer region and in a 250 bp segment of the control region of mtDNA in 162 pollock from 32 localities. However, the seasonal timing of sample collection is unknown and samples may have been collected when fish from different spawning groups mix. Samples were grouped into six regions for analysis: 1) western Bering Sea, 2) northwestern Bering Sea, 3) the Donut Hole, 4) Aleutian Island chain, 5) southeastern Bering Sea, and 6) Gulf of Alaska. Data for the two segments were analysed separately since not all fish were successfully sequenced for both mt DNA segments. Twenty spacer-region haplotypes were found in 110 fish. A total of 83 fish (75%) had the same haplotype, and 17 of the 20 haplotypes each occurred in a single fish. A Monte Carlo chi-square test of haplotypic frequencies in the six regions was non-significant. Seventeen controlregion haplotypes were found in 140 fish, of which 114 had the same haplotype. Eleven of the 17 haplotypes each occurred in a single fish. A Monte Carlo test of the 11 most frequent haplotypes among the six
Table 6 Estimates of mtDNA haplotypic (h) and nucleotide (n,*) diversities and population differentiation, F,, among populations of demersal fishes. (From Seeb et al., in preparation.) h Species Walleye pollock Atlantic haddock Atlantic cod Greenland halibut Red drum Black drum Red snapper
No. of localities
Mean size
No. of haplotypes
4 5 5 4 7
42.0 10.8 26.6 25.8 40.0 63.0 37.5 46.8
65 21 22 10 22 99 37 68
11 8 9
TI,*
(%)
Mean
Pooled
F,,
Mean
Pooled
Reference
0.900 0.888 0.769 0.335 0.692 0.954
0.918 0.867 0.789 0.330
0.019
0.69
0.75 0.61
0.025 0.003 0.003 0.002 0.011 0.0
1 2 2 3
0.25 0.59
0.951 0.780 0.750
~
4 0.58 0.48 0.50
5
6 6 ~
Data from: 1. Mulligan et al. (1992); 2. Zwanenburg et al. (1992); 3. Pepin and Carr (1993); 4. Morgan et al. (1997); 5. Gold et al. (1993); 6. Gold et al. (1994).
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
223
regions also failed to detect significant heterogeneity. However, an a posteriori test of regions 1 and 2 versus regions 4 and 5 showed a significant difference for combined spacer- and control-region haplo types. The finding of very low levels of population differentiation is similar to the results of Mulligan et al. (1992), except for the lack of differentiation of Aleutian Island samples. In contrast, another study of mtDNA variability appears to have detected high levels of variability among populations in the Bering Sea and Gulf of Alaska. Quattro and Powers (unpublished data, reported in Macklin, S. A. 1999, Bering Sea FOCI Final Report, Environmental Research Laboratory Special Report, Pacific Marine Environmental Laboratory, Seattle, WA (in press); see also Bailey et al., 1997) sequenced two 300 bp segments each in coding regions of cytochrome-b and ATPase and found three haplotypes that were subsequently screened with PCRRFLP methods in 262 pollock from nine regions. These regions included 1) the Gulf of Alaska, 2) southeastern and 3) northeastern Bering Sea, 4) southeastern and 5 ) northeastern Aleutians, 6), Aleutian Basin, 7) western Bering Sea, 8) east Kamchatka Peninsula and 9) Sea of Japan. Regions 1, 3 and 5 were tested for temporal variability between samples collected in 1988 and 1991 or 1994, and none of these comparisons was signilicant. These data were pooled by region and the nine regions were then tested for frequency heterogeneity. The amount of differentiation among these regions was considerable; an analysis of molecular variance (AN'OVA, equivalent to FsT) showed that 14.5% of the total variability resulted from differences among samples. A cluster analysis of sequence divergences between samples indicated that the greatest amount of differentiation was between Gulf of Alaska/eastern Bering Sea populations and western Bering Sea/Sea of Japan populations. This major subdivision is consistent with allozyme data but not with some studies of mtDNA. Bering Sea populations and those in the Gulf of Alaska also appear to be differentiated from each other, but to a lesser extent. The analysis of microsatellite loci in pollock has apparently dei ected differences among genetically discrete groups of fish on finer temporal and spatial scales than has the analysis of allozyme loci or mtDNA. Pollock primers for two microsatellite loci found in Atlantic cod (Gmo-2, Gmo-145) were used by Villa and Powers (unpublished data, reported in: Macklin, S. A. 1999, Bering Sea FOCI Final Report, Environmental Research Laboratory Special Report, Pacific Marine Environmental Laboratory, Seattle, WA (in press); see also Bailey et al., 1997) to arialyse pollock from six localities that had been collected from 1990 to 1996. These included 1) western Bering Sea off Kamchatka Peninsula, 2) southeastern Bering Sea, 3) Pribilof Islands and 4) Shelikof Strait in the Gulf of Alaska. Eleven alleles were detected for Gmo-2 and 20 alleles
224
K. M. BAILEY ETAL.
were detected for Gmo-145. ANOVAs for both loci showed significant differences among the four regions, and paired a posteriori comparisons using 'lhkey's test showed significant differences between all pairs of samples except between regions 2 and 3. However, the time scale over which the samples were taken and the lack of temporal sampling make it difficult to interpret these results. Furthermore, in the case of Gmo-145 a null allele was responsible for the major east-west differences. The presence of a null allele indicates that the nucleotide sequences in the flanking regions of microsatellite markers were not identical between these populations, resulting in no PCR product for those individuals bearing the null allele. Cluster analyses of two genetic distances indicated that the sample from Shelikof Strait was more closely related to samples from the western Bering Sea than they were to samples in the southeastern Bering Sea. These results differ to some degree with the analysis of rntDNA and with results from allozymes. In a study of allozyme, mtDNA and microsatellite variability in three samples collected off Bogoslof Island, Shelikof Strait and in Prince William Sound, Seeb et al. (unpublished data) found significant differences among various combinations of samples for each method. Eleven of 31 allozyme loci were polymorphic (common allele frequency 0.095 or less) and the average heterozygosity among samples was 0.08. The sum of chi-square values over these 11 polymorphic loci showed a significant difference between the Gulf of Alaska samples and the Bogoslof sample, but no significant difference was found between the two Gulf of Alaska samples. Monte Carlo comparisons of frequencies of 68 composite PCR-RFLP haplotypes for the mtDNA genes cytochrome b, cytochrome oxidase and NADH Dehydrogenase 5/6 also revealed significant differences between the Bogoslof Island sample and the two samples from the Gulf of Alaska, but not between the two Gulf of Alaska samples. The results for the microsatellite loci were less certain because of the apparent presence of null alleles at two of the loci, Gmo-9 and Gmo-132, as revealed through inheritance studies. However, significant differences between Prince William Sound and Bogoslof Island ( P = 0.038) were detected for the Gmo-1 locus, which appeared to give better results than the other microsatellite loci. Together, these results reflect other studies in showing little differentiation among Gulf of Alaska populations and moderate amounts of differentiation between Gulf of Alaska and southeastern Bering Sea populations. These genetic studies permit a few tentative conclusions about the temporal and spatial patterns of genetic structure in pollock. As with several other fishes in the North Pacific, the imprint of isolation by Pleistocene glaciation appears to be visible in allozyme differences between Asian and North American populations and in mtDNA dif-
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
225
ferences in one of two studies where the extent of sampling permits this comparison. Microsatellite data, however, failed to reflect this major east-west subdivision. These data appear to indicate that Gulf of Alaska populations are more closely related to western Bering Sea populations than they are to geographically nearby southeastern Bering Sea populations. This is similar to microsatellite data for Pacific herring, which also failed to detect strong east-west differences that appeared in allozyme (Grant and Utter, 1984) and mtDNA data (P. Bentzen, unpublished data). The reason for the apparent genetic similarities between east-west populations across the North Pacific of both pollock and Pacific herring could be convergence in the microsatellite length alleles and not longdistance gene flow. Within the North Pacific races, small genetic differences have been detected between the Gulf of Alaska and the southeastern Bering Sea for pollock in a study of allozymes, in three studies of mtDNA variability, and in a study of microsatellite variability. The Alaska Peninsula and the Aleutian Archipelago are obvious barriers to gene flow between these two areas. Genetic differences between localities have been detected within these areas in some studies, but no clear picture of geographically stable populations has so far emerged. Several reasons may contribute to the lack of clear population definition. One reason is that for some molecular techniques, such as allozymes and mtDNA, only a small amount of gene flow between populations is required to produce the appearance of genetic homogeneity (Waples, 1998). The genetic structure detected by these methods is the result of random drift, which is small in large populations, and of gene flow, which is potentially large in marine species. Microsatellite loci, on the other hand, are often characterized by elevated mutation rates, which may produce population-specific frequency profiles that depict fine-scale population structure. Because of the nature of length mutatialns at microsatellite loci, however, allelic convergence is common SCI that same-length alleles in two fish may not be identical by descent, an assumption that is important for inferring genetic relatedness between populations from molecular data. More studies on the temporal/spatial stability of these markers in populations are also needed. Another reason for the lack of genetic definition may be that the samples for some studies were collected out of the spawning season. If the homing of pollock to specific spawning localities leads to genetically discrete populations, samples collected in mixed-stock (feeding) areas maly not resolve the genetic structure of breeding populations. Another poss4ibility is that the genetic differences detected in some of the studies reviewed here may result from fine-scale structure related to short-term isadation of population segments by physical oceanographic features.
226
K. M. BAILEY ETAL.
4.4. Metapopulation Structure Metapopulations are broadly defined as a collection of partially isolated populations that are connected by migration and gene flow (Harrison and Taylor, 1997). The metapopulation concept can provide a link between landscape ecology, population dynamics and population genetics. Metapopulation dynamics may be important for marine fishes because episodes of climatic and oceanic change may influence local population abundances. A metapopulation approach requires that: 1) local populations have spatial structure where local populations are separated by unsuitable habitat, and 2) migration has an effect on local dynamics (Hanksi and Simberloff, 1997). Pollock dispersed from the Atlantic through the Arctic Ocean and into the Pacific with the submergence of the Bering Strait during the Pliocene (Svetovidov, 1948). Some populations have apparently colonized available habitats relatively recently. For example, Prince William Sound, Puget Sound and other fjords and bays were either ice covered, or in the process of being formed during the last period of heavy glaciation about 10 to 18 thousand years ago. At this time, sea levels could have been as much as 150 m lower than present-day levels, and the major basins would have been substantially more isolated with strong geographical barriers between populations. New local populations are in the process of being discovered, such as the Aleutian Basin population found in the 1970s and the Shumgin Island population discovered in the early 1990s. Whether these are newly established populations or whether they existed previously at low levels without being recognized is unknown. Local pollock populations also appear to be susceptible to extinction. A recent example is in south Puget Sound, where a healthy charter boat recreational fishery existed in the 1970s to 1980s. Since about 1985 pollock are rarely caught in south Puget Sound by either recreational fishers or scientists, and for all practical purposes are considered to be extinct in this part of their range (Palsson et al., 1996). It is not always clear whether application of the metapopulation concept is appropriate for marine fish populations because of the issues of non-continuous distributions separated by uninhabitable areas and the extent of exchange of individuals between populations. Genetically distinct populations would indicate little exchange of individuals between sub-populations on an evolutionary scale, much less on an ecological scale. However, even if there is movement between sub-populations, life-history differences in spawning time or mating behaviour may prevent gene flow. In the case of pollock, local spawning populations are patchily distributed, but during the summer feeding period they may mix and are not clearly separated by uninhabitable areas, although there appear to be
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
227
less populated habitats between major aggregations. The concept of metapopulations was derived from land animals where the landscape is discontinuous, whereas the ocean is clearly different. Furthermore, the distribution of fish at different stages changes with ontogeny, arid the exchange of individuals among populations may occur at differeint life stages. Movement of individuals among walleye local pollock populations on an ecological time scale remains a subject of controversy. In spite of these issues, the framework of metapopulation dynamics may have utility for assessing population dynamics of neighbouring local populations where there is a large potential for exchange, for examiple in the Bering Sea. It is likely that some groups of local pollock populations have characteristics that may fit into a metapopulation framework. However, within the metapopulation concept there may be different types of connections between local populations including: mainland-aland, source-sink, non-equilibrium and stepping stone models. A strawman conceptual model of population structure for walleye pollock is presented in Figure 16. This hypothetical view shows some populations (e.g. Puget Sound) that are relatively isolated from other distant populations, and groups with varying degrees of population exchange ranging from a lot (within the eastern Bering Sea) to a little (across the Bering Sea). However, there are correspondences between stock groupings, diffeirentiation and the potential for mixing among different current systems. Most pollock stock members appear to complete their life cycle within specific current systems (Figure 17). Another way of conceptualizing pollock population structure is through MacCall’s (1990) geographic basin model. In this model, a population is analogous to liquid in a basin, where the volume of the basin represents abundance, the depth is the density, and the extent of shoreline is the range. Population growth rate is set by in sits growth (birth and death rates) and dispersal (which is determined by viscosity and habitat suitability). Characterizing pollock in particular, local stocks may be represented as shallow-basin populations (tending to increase range as the population grows), with low viscosity (high mobility). Across its range pollock may be viewed as a multi-basin array (multi-stock) connected by sills of varying heights (limiting dispersal between basins to varying degrees). Whether there are source-sink relationships among stocks of pollock is controversial. The existence of commercial concentrations of pollock in the Aleutian Basin is commonly considered to represent density-driven outflow of fish from the eastern Bering Sea continental shelf. Furthermore, as surplus fish they could be harvested severely without affecting their source population (Wespestad, 1993). Unfortunately, hypotheses about Aleutian Basin fish originate from sparse scientific observations. The
228
K. M. BAILEY ETAL.
Figure 16 Hypothetical model of the population structure of walleye pollock. Filled circles represent spawning populations scaled to approximate abundance. Hatched regions represent the proposed distributions of several major populations (Okhotsk, east Bering, west Bering, Shelikof) during the feeding season, indicating overlap. Lines show some of the potential connections among populations scaled to relative gene flow through migration or larval transport.
pollock in the Aleutian Basin are largely composed of fish older than 5 years. Juveniles are believed to be rare in the Aleutian Basin (Mulligan et at., 1989), and only recently have they been found along the northeastern side of the basin (Tang et al., 1996). Spawning occurs in the central basin and in the southeastern part of the basin as well (Hinckley, 1987; Sasaki, 1988; Mulligan et al., 1989). These offshore spawning aggregations are separated from continental shelf spawning groups by as much as 900 km. Pollock in the basin region also have different length-at-age and fecundity characteristics compared with shelf fish (Hinckley, 1987; Dawson, 1994), and these differences indicate the existence of different spawning groups. Eggs and larvae from the southeastern part may drift with prevailing currents onto the outer continental shelf and slope of the eastern Bering Sea (Bailey et at., 1997). In contrast to the idea of separate spawning populations, Shuntov (1992) and Dawson (1994) proposed that large
Figure 17 Major currents in the North Pacific Ocean.
230
K. M. BAILEY ETAL.
numbers of pollock found in the Aleutian Basin make seasonal migrations from the eastern and western Bering Sea shelves to graze there. However, grazing conditions over the basin region cannot be especially good, because size-at-age and fecundity of fish caught there are lower than over the shelf regions (Hinckley, 1987), and stomach contents indicate poor feeding (Dwyer, 1984; Yoshida, 1994). Density-driven migrations have important implications for harvesting strategies, demographic structure and genetic structure of the population. However, most of the support for the source-sink concept in the Aleutian Basin comes from fragmented or anecdotal observations. We believe that scientific data collected within a solid theoretical framework is needed. Density-dependent movement of the 1989 year class from the Bering Sea into the Gulf of Alaska also merits further examination. The 1989 year class was believed to be weak in the Gulf of Alaska from observations of juvenile abundance, so its increasing abundance in the Gulf in subsequent years has been interpreted as evidence of outflow of the strong 1989 year class from the Bering Sea. Alternatively, larvae and juveniles may have .been transported out of the normal area of juvenile surveys in the Gulf, followed by homing of adults. A similar situation has been observed in haddock (Frank, 1992). A portion of the strong 1988 year class in the Gulf may also be misclassified to another year class owing to imprecision in assigning fish ages as they get older and their otoliths become more difficult to interpret. 4.5. Population Structuring Mechanisms
In marine fishes, the importance of larval dispersal on gene flow and population structure has been shown by several studies (Waples, 1987; Waples and Rosenblatt, 1987; Doherty et al., 1995; Shulman and Bermingham, 1995; also see Avise, 1994). However, high dispersal potential may not always translate into high gene flow and genetic homogeneity (Palumbi, 1995). For example physical impediments, such as fronts and eddies (Iles and Sinclair, 1982), may limit dispersal when early life stages are not passively drifting but are actively and purposefully swimming, and when later stages have natal homing behaviour. Pollock have been the subject of studies of larval transport in relation to oceanographic conditions in the Gulf of Alaska since 1981 (Kim and Kendall, 1989). Pollock eggs and larvae are found in concentrations that can be tracked and monitored (Bailey et al., 1996c) and their distributions can be predicted reasonably well from modelled ocean currents (Hermann et al., 1996). Pollock larvae are often associated with oceanographic features such as eddies (Vastano et al., 1992; Schumacher et at., 1993;
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
231
Bailey et al., 1997). In the following section we discuss the rode of variability in larval transport and retention by features, such as eddies, in structuring the walleye pollock population, and we explore the possible linkages between them. Other mechanisms of population structuring in pollock are discussed with the main emphasis on populations in the Bering Sea and Gulf of Alaska. Considerable knowledge of ocean currents and their variability is required to understand patterns of gene flow in pollock populations. In the Gulf of Alaska, the major spawning of pollock is in Shelikof !$trait, a deep sea valley (>250 m) penetrating the continental shelf between the Alaska Peninsula and Kodiak Island. Flow in this region is dominated by the Alaska Coastal Current, which is one of the most vigorous coastal currents in the world with speeds of 25 to lOOcm.s-' (Stabeno et al., 1996b; Figure 18). The Alaska Coastal Current bifurcates east of Sutwik Island, with one branch continuing along the shallow continental shelf (-100 m) along the Alaska Peninsula. This branch has relatively weak flow, with speeds of about 10cm.s-'. The other branch flows seaward through the sea valley with a portion that joins the Alaska Stream (with speeds of 50 to 100cm.s-') and the remainder recirculates onto the shallow shelf. Frontal features, meanders and eddies are prominent in the Shelikof sea valley (Vastano er al., 1992; Napp et al., 1996). Horizontal density gradients and vertical shear in the flow result in baroclinic instabilities which generate eddies in the region. The location of eddy formation coincides with the area of pollock spawning (Schumacher and Kendall, 1995) and in the springtime spawning period three to four eddies form per month (Bograd et al., 1994), some of which may remain stationary for weeks (Schumacher et al., 1993). About 70 satellite-tracked drifters drogued at 40 m to simulate larval drift have been released in the Shelikof sea valley. The drifters show the pattern of bifurcation of the Alaskan Coastal Current and the variability in the fate of larvae during the 50 days or so when they are planlktonic (Figure 19). Residence time of the drifters on the shelf ranges from 35 to 122 days, with a mean of 55 days. Drifters can exit the sea valley as early as 15 days and enter the Alaska Stream where they are quickly transported westward. In Shelikof Strait, pollock show a mean downstream progression of their centre of abundance over time (Figure 20). However, these patterns vary greatly between years. Larval transport patterns show reasonably good correspondence with satellite-tracked drifter patterns. In years when larvae are transported onto the shallow shelf west of the sea valley, larvae drift downstream at speeds of 4 to 6 cm.s-' (Hinckley et al., 1991). Although passive physical transport seems important, larvae are able to
-175O00'
-175"Oo'
-17PoO'
- 16S000'
-1WW
-155000'
-165"00'
-1W00'
-155"00'
-150"00'
Figure 18 Currents in the Gulf of Alaska and Bering Sea. Hatched areas are known pollock spawning locations.
233
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
iww
170-w
i6aw
166~1 1 6 4 ~ 1i
ww
i6oow
i5avv
1 7 2 ~
Figure 19 Composite of representative satellite-tracked drifter trajectories in the Gulf of Alaska. Drifters were drogued at 40m,released near larval pollock aggregations, and tracked for 50 days, approximating the duration of larval drift. Dashed line is the 1OOOm isobath. Solid dark lines are drifter paths and circles mark position at 10 day intervals (after Bailey et al., 1997).
control their vertical movement and thus may have an active influence in their distribution (Olla et al., 1996), especially for older larvae and juveniles. Eddies in the Shelikof Sea Valley can retain larvae for several weeks in the region. For example, in 1990 an eddy was found with an aggregation of larvae. Satellite-tracked drifters remained in the eddy for about 22 days, until the eddy moved westward, interacted with sh.allow topography of the shelf and disintegrated (Bailey et al., 1995). In some years with strong and frequent storm activity, wind-driven transport increases and flushes most larvae from the sea valley into the Alaskan Stream. An example of this was in 1991 when winds were extremely high, with three storm events during the late April early May period, and drifter trajectories indicated that most water vigorously flowed out of the sea valley (Bailey et al., 1995). Under these conditions, primary production and microzooplankton levels were depressed, 1 arval feeding and nutritional conditions were poor and larval mortality was high (Bailey et al., 1995). In general, pollock larvae experiencing
234
K. M. BAILEY ETAL.
0 March 29-April
13
&lApril 15-April 29
@ @
April 30-May 15 May 16-June 2
Figure 20 Centroids of larval pollock distribution tracked by 2-week periods, averaged over 1987 to 1992 (after S. Picquelle, unpubl. data).
conditions associated with these high transport conditions will have low survival. In the Bering Sea, a cyclonic gyre dominates circulation, with the Kamchatka Current flowing southward along the western side, and the Bering Slope Current northward along the eastern side of the central Aleutian Basin (Figure 17). Alaska Stream water can enter the Aleutian Basin through any of the deep passes through the Aleutian Islands, but predominately through Amchitka, Amutka and Buldiur Passes and Near Strait. The transport through the passes varies on time scales of months. Over the main area of our interest, the southeast basin and the shelves, the major flows are the Bering Slope Current (BSC) and Aleutian North Slope Flow (ANSF) (Figure 18). The Aleutian North Slope Flow goes eastward along the north slope of the Aleutian Islands and connects the flow through the passes with the Bering Slope Current. Instabilities in the Bering Slope Current can result in onshelf flow from the basin onto the shelf. The shelf is divided into three regions, the inner shelf (depth <50m),
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
235
the middle shelf (50-90 m), and the outer shelf (10Cb170 m). Flow om the middle shelf is weak (mean flow <1 cm.s-'). Stronger flows occur a t the transition zones between the inner middle shelf and the middle outer shelf. Flow along the 100 m isobath is generally northwestward at a bout 5 to 10cm.s-', while the northeastward flow along the 50m isobath is weaker at about 2 to 3cm.s-' (Schumacher and Stabeno, 1997). Flow along the 100m isobath originates from onshelf flow from the basin and from flow through Unimak Pass. The flow along the 50m iscibath originates from the flow through Unimak Pass. Eddies in the eastern Bering Sea are relatively common (Schumacher and Stabeno, 1994) and are formed in regions of high current shear, or by the interaction of inflowing Alaskan Stream water with topography. Eddies that may transgress the continental shelf break up in sh-11 cl ow water over a period of days. Eddies may circulate vigorously (25 cm.s-I), but their net movement is generally slow (1 to 5cm.s-'). Diameters of eddies may vary from 5 to 80 km. In the eastern Bering Sea, larval pollock concentrations have not lbeen tracked as effectively as in the Shelikof region. Repeated spawning is known to occur in several areas and larvae are frequently concentrated in these same areas. From drifter studies, it appears that larvae from different areas can readily mix during the larval drift period. About 80 satellite-tracked drifters have been released in the Bering Sea to study potential larval drift patterns (typical patterns are shown in Figure 21). On the middle shelf area (50 to 100m depth), the flow is weak and all drifters remained resident there for more than 60 days. Drifters released in the Unimak Pass spawning region are either transported along the 100m isobath until reaching the Pribilof Islands, at which time they may circle the islands, flow eastward across the shelf along the 50m isobath, or continue northward to S t Matthew Island. Residence time in the slope region is 30 to 90 days, with a mean northward velocity of 5cm.s-'. Drifters released near Bogoslof Island usually follow the Aleutian North Slope Flow into the Bering Slope Current, although about 15% of drifters go onto the shelf where they may re-circulate in a small sub-gyre in the southeast corner of the Aleutian Basin. Eddies in the Bering Sea are similar to those in Shelikof Strait and they may act as larval retention mechanisms (Figure 22). An eddy found in 1993 over the southeastern corner of the Aleutian Basin was associated with very high numbers of larvae, however, it later entered the outer shelf and disintegrated, providing a vehicle for mixing of larvae between spawning regions. The Aleutian Basin region may act as a geographical barrier to gene flow across the Bering Sea. Microzooplankton concentrations are low and larval feeding levels are also low compared with other regions
236
K. M. BAILEY ET AL. I
I
I
I
172"w
172W
172"
I
I
I
172W
I
I
I
172"w
I
I
1
I
172W
Figure 21 Composite of representative satellite-tracked drifter trajectories in the Bering Sea. Drifters were drogued at 40m, released near larval pollock aggregations or on spawning grounds, and tracked for 50 days, approximating the duration of larval drift (after Bailey ei al., 1997).
(Hillgruber et al., 1995; Theilacker et al., 1996). Furthermore, studies of larval mortality in the basin region indicate low larval growth rates and high levels of mortality (Dell'Arciprete, 1992). Other mechanisms of maintaining population structure include philopatry (a tendency to remain in the native locality) and natal homing (returning to birth site). Local sub-populations are connected by dispersal, which includes larval dispersal and colonizing movements of adults. Migrations, or coordinated movements of individuals that can be ontogenetic, seasonal or daily, are distinguished from dispersal or nomadic behaviour, which is undirected and out of the natal or home range. Dispersal can occur due to environmental factors, such as El Niiio events, and also due to density-dependent effects. As demonstrated for some fishes, fidelity to home range depends on habitat quality (Matthews, 1990) and this can depend on density. The advantages of dispersal are colonization of new habitats and avoidance of unfavourable local conditions (Quinn, 1993). The effectiveness of movement between
-1 70000'
-168% I
I
-1 6POO' I
-165'30' I
n
55%
55"3O'
N
Driftertracks
Number per 10 m2 0-1000 1000.1-5000 5000.1-35000 0No Data
54"M
0 I
-17O"OO'
1
-168"30'
-16POO'
I
-1 65"
Figure 22 Distribution of walleye pollock larvae in April 1993 in the eastern Bering Sea. Also shown are satellite-tracked drifter trajectories showing the location of an eddy in the region. Inset shows the area of coverage. Latitudes are ON and longitudes are "W (redrawn from Bailey et al., 1997).
238
K. M. BAILEY ETAL.
populations depends on the extent of the movement and the mortality of the dispersing individuals. One of the major issues of the sub-population concept is natal philopatry. Through natal homing, animals that stray away from their natal site may return for reproduction. In many fishes where juveniles and adults make seasonal and ontogenetic migrations related to feeding or reproduction, they may utilize either one of four recognized mechanisms, or a combination thereof, to return to their natal site and maintain philopatry including: 1) imprinting on environmental cues, 2) learning from other fishes (social tradition), 3 ) genetically based, innate, homing and 4) passive transport by currents. m e nature of the homing mechanism, when it occurs, has some interesting implications for management issues, as discussed below. Pollock show repeated and predictable spawning at specific sites and times, which is one piece of evidence for natal philopatry. Among the best-studied spawning aggregations is the one in Shelikof Strait which has been monitored since 1981. The Shelikof Strait spawning aggregation is the largest spawning biomass in the Gulf of Alaska, and usually occurs within a limited area (40km by SOkm) during the first week in April (Kendall et al., 1996). The geographic and temporal consistency of spawning in the Shelikof population argues against the notion that the time and place of spawning are random. Mark-recapture studies around Japan, where pollock were tagged during the spawning season (April), showed dispersed feeding migrations but homing to specific previous spawning locations (Tsuji, 1989;Figure 14). In other studies, pollock tagged during the summer feeding periods were later recaptured across a wide geographical expanse (Figure 13). Those recaptured around the time of the spawning season were located near well-established spawning grounds. Pollock appear to disperse widely after spawning, and the location of summer feeding grounds may vary. The fact that fish tagged in summer in the western Bering Sea were recaptured in the eastern Bering Sea, along with the significant population differentiation between eastern and western Bering Sea spawning pollock discussed above, indicates isolation of spawning groups even though they may mix during the summer feeding period. While not unequivocally proving natal spawning, the persistence of spawning locations over time and repeat spawning are two elements of natal philopatry. 5. MANAGEMENT IMPLICATIONS Populations have two components of structure (Slatkin, 1993); demographic (or ecological) structure is determined by birth, death and
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
239
dispersal processes, and genetic structure is caused by mutation, selection, drift and dispersal. In fisheries management, both of these components of population structure are important. Demographic structure provides guidelines for local harvesting practices and indicates the potentiall for locally adapted stocks, while genetic structure reflects deep demographic structure and requires consideration in conservation of genetic variability (e.g. Carvalho, 1993). Since the time of Hjort (1914) fisheries populations have been treated as closed homogeneous systems. Hjort realized that the dynamics of populations are a result of local variability in birth and death rates rather than large-scale movements of populations across ocean basins (Sinclair, 1988). A contemporary issue is whether our perception of fisheries dynamics is confounded by lumping individual populations whose: individual population trajectories may be quite different from the aggregate. Several recent studies have emphasized the need to re-examine the large-scale spatial patterns underlying the dynamics of fish populations (MacCall, 1990; Frank, 1992; Purcell et al., 1996). Prior to the collapse of the northern cod population in the north west Atlantic Ocean it was believed that genetically based population structure was absent because of extensive egg and larval drift, followed by opportunistic and non-philopatric recruitment of juveniles to adult assemblages (deYoung and Rose, 1993). Studies of allozymes and mtDNA had little success in discriminating among northwest Atlantic cod stocks. However, Bentzen et al. (1996) examined nuclear DNA microsatellite distributions and found that the northern cod population does not represent a single panmictic assemblage, but is composed of genetically distinguishable sub-units, each of which is affiliated with a distinct spawning area. It is of great interest that considerable fine-structure has been found in Atlantic cod populations, because cod and pollock are phylogenetically related and have similar life histories. Among the major basin regions of the Sea of Japan, Sea of Okhotsk, Bering Sea and Gulf of Alaska, structure in pollock populations is supported by phenotypic and life-history data, and to some degree by genetic studies. Probably these major basin regions were largely isolated from each other during the last major glaciation, as well as at other times during pollock’s history in the North Pacific. Considering the colonizing and migration ability of pollock, the occurrence of genetic structure among populations is somewhat surprising, but perhaps can be attributed to larval retention mechanisms, differences in reproductive characteristics (timing and temperature adaptations), natal homing, or outbreeding depression of immigrants that reproduce with historical local populations. Furthermore we would expect that local and heritable life-history adaptations may evolve rapidly through selection and may not necessarily be
240
K. M. BAILEY ETAL.
apparent in the geographic distributions of selectively neutral genetic markers. Several issues are important to consider if genetically or geographically discrete sub-populations are lumped into a single management unit. Daan (1991) indicated that if catch-at-age analysis is applied to a heterogeneous unit stock (a metapopulation), then fishing effort is underestimated and fishing mortality deviates from the recent true trend, leading to the possibility of overfishing. But the depletion of unidentified local subpopulations by heavy fishing and their possible extinction is a key concern (Ovenden, 1990). The rate of recovery of geographically, but not genetically, discrete populations will depend on the level of gene flow between sub-populations. Genetically distinct sub-populations, by definition, have low levels of gene flow among them, and specific stock adaptations to local habitat conditions can impede recovery by colonization from extrinsic source populations. There are some other areas where better stock resolution would be useful in management of pollock harvests. A central issue in the case of pollock management is whether broad-scale migrations occur that result in seasonal mixing of sub-populations. For example, if eastern or western Bering Sea fish migrate across or around the basin (as proposed by Dawson, 1994), they may be harvested on both sides of the Bering Sea basin. Therefore, a migratory population could experience two independent sources of fishing mortality in two management regions. Recently, new local populations of pollock have been discovered. We do not know if they are self-sustaining populations or transitory stocks. For example, in recent years a spawning population of 300000 mt was found around the Shumagin Islands in the western Gulf of Alaska (C. Wilson, personal communication, 1996). This population is largely composed of the 1989 year class, which was strong in the Bering Sea but weak in the rest of the Gulf of Alaska. Did this aggregation originate from density-dependent dispersal from the Bering Sea, or is it an increase in a previously unknown local sub-population? Another explanation may be that vagrants from the Bering Sea could have joined a small local spawning population. Similar issues exist in the Aleutian Basin to Bering Sea populations. The accuracy of spawner-recruitment relationships also depends on the definition of stock structure among populations. For example, consider the following hypothetical but realistic example: if two genetically distinct and asynchronously fluctuating populations, A and B, are lumped together in a VPA analysis, dominance shifts could result in a spawner-recruit relationship in which recruits from population A are compared to egg production from population B. Shifts in dominance between putative stocks appears to be a situation particularly applicable to pollock in the
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
241
Bering Sea (Francis and Bailey, 1983). In this case, knowledge of the underlying population structure would vastly improve the understanding of spawner-recruit relationships. Another confounding situation may occur when new recruits are immigrants, this sometimes occurs with pollock in the Gulf of Alaska. All of these circumstances can result in misinterpretation of the factors underlying the dynamics of these populations. The mechanism of natal-site fidelity has broad implications to fisheries management and sustainable harvesting of these populations. For example, if fish learn their migration routes from older fish, a speculated mechanism for cod (Rose, 1993) and herring (Corten, 1993), then removing older fish from the population can lead to increased straying of younger fish owing to loss of social tradition. Furthermore, this mechanism implies that vagrant larvae can adopt new local populations. The mechanism of an innate sense of natal migration implies that if a local population goes extinct, recolonization may depend on natural selection and adaptation to local conditions which may take a very long time, and vagrants from other populations are unlikely to spawn with local populations. The mechanism of imprinting on a natal location implies that harvesting will not have an impact on migratory routes and heavily depleted populations can be restocked. It is not known which homing mechanisms are utilized by pollock during spawning migrations. In fact, we suspect marine fish species probably use a combination of all these mechanisms for natal fidelity, and the response to any one stimulus is weighed against other stimuli. Based on the complex range of behaviours pollock are capable of exhibiting (Olla et al., 1996), it should nalt be surprising that they show flexibility in their homing behaviour. New paradigms for managing fisheries in a sustainable manner are needed, and these concepts may depend upon better definitions of stock distributions and interrelationships. However, problems exist in implementing new information on stock structure. For example, in a review of stock structure methods, Pawson and Jennings (1996) suggest that despite the many studies of genetic structure for marine fish populations, only a few have provided information useful to managers. They note that the exception is Pacific salmon, where genetic studies have been very useful. Certainly, information indicating multiple genetic-based stocks adds another element of concern to fisheries managers and makes their task more complex (Lincoln, 1994). Good communication between fisheries managers and geneticists is needed for effective implement i it'l0n of stock structure information, including clear objectives for proposed work and utilization of appropriate techniques (Carvalho and Hauser, 1994). Given the high potential for gene flow in marine species including
242
K. M. BAILEY ET AL.
pollock, genetic approaches alone may not completely resolve stock structure issues at the level of harvest stocks or ecological populations, and tagging studies offer another way to study linkages between fish populations. Although not necessarily showing levels of gene flow between areas, appropriate tagging and/or transplant experiments would offer a powerful approach to studying the effects of migration on stock structure. Tagging pollock has been problematic in the past owing to issues of barotrauma, tag-recovery bias and distribution of effort. A large number of tags is necessary to mark a high proportion of these large populations to get effective returns. Technologies should be improved to minimize these problems including environmental tags, such as otolith chemical fingerprints (Campana and Gagne, 1995), and in situ tagging. Historically, most tagging of pollock has been done opportunistically on adults in summer when they are dispersed, which has not been very effective. Experiments that tag juveniles or spawning adults, as well as tag and transplant experiments, could be more informative. Studies of fish dispersal that are coordinated with modern genetic studies should contribute significantly to understanding linkages between population dynamics and structure of marine fish populations.
ACKNOWLEDGEMENTS Arthur Kendall, Susan Sogard, Carol Sibley, Robin Waples, Ric Brodeur, Phyllis Stabeno, Gary Stauffer, Gary Duker, Dennis Powers, John Blaxter and Mike Canino reviewed either some sections or the whole of this manuscript. In some cases (D.P. and P.S.), they co-authored manuscripts that were pre-cursors to this effort. We also thank Dennis Powers, Phyllis Stabeno, Steven Porter, Annette Brown, Akira Nishimura, Jim Traynor, Gary Walters, Jim Ianelli, Gary Villa, Joseph Quattro, Allen Macklin, James Seeb, Lisa Seeb and Susan Picquelle for contributing information used in this text.
REFERENCES Arthur, J. R. (1983). A preliminary analysis of the discreteness of stocks of walleye pollock, Theragra chalcogramma, from the northeastern Pacific Ocean o f f Canada based on their parasites. Canadian Technical Report of Fisheries and Aquatic Sciences 1184, 1-15. Atkinson, D. B., Rose, G. A., Murphy, E. F. and Bishop, C. A. (1997). Distribution changes and abundance of northern cod (Gadus morhua), 1981-1993. Canadian Journal of Fisheries and Aquatic Sciences 54, 132-138.
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
243
Avdeev, V. V. and Avdeev, G. V. (1989). A study of walleye pollock population structure and migration routes using parasitological indicators. In “Proceedings of the International Symposium on the Biology and Management of Walleye Pollock”, pp. 569-590. Sea Grant Report No 89-1, University of Alaska, Fairbanks, AK. Avise, J. C. (1994). “Molecular Markers, Natural History and Evolution”. Chapman and Hall, New York. Bailey, K. M. (1989). Interaction between the vertical distribution of juvenile walleye pollock Theragra chalcogramma in the eastern Bering Sea, and cannibalism. Marine Ecology Progress Series 53, 205-213. Bailey, K. M. and Macklin, S. A. (1994). Analysis of patterns in larval walleye pollock Theragra chalcogramma survival and wind mixing events in Shelikof Strait, Gulf of Alaska. Marine Ecology Progress Series 113,1-13. Bailey, K. M., Brodeur, R. D., Merati, N. and Yoklavich, M. M. (1993). Predation on walleye pollock (Theragra chalcogramma) eggs and yolk-sac larvae by pelagic crustacean invertebrates in the western Gulf of Alaska. Fisheries Oceanography 2, 30-39. Bailey, K. M., Canino, M. F., Napp, J. M., Spring, S. M. and Brown, A. L. (1995). Contrasting years of prey levels, feeding conditions and mortality of larval walleye pollock Theragra calcogramma in the western Gulf of Alaska. Marine Ecology Progress Series 119,11-23. Bailey, K. M., Brown, A. L., Yoklavich, M. M. and Mier, K. L. (1996a). Interannual variability in growth of larval and juvenile walleye pollock Theragra chalcogramma in the western Gulf of Alaska, 1983-91. Fisheries Oceanogruphy 5 (Suppl. 1), 137-147. Bailey, K. M., Brodeur, R. D. and Hollowed, A. B. (1996b). Cohort survival patterns of walleye pollock, Theragra chalcogramma, in Shelikof Strait, Alaska: a critical factor approach. Fisheries Oceanography 5 (Suppl. l ) , 179-188. Bailey, K. M., Stabeno, €! J., and Powers, D. A. (1997). The role of larval retention and transport features in mortality and gene flow of walleye pollock. Journcil of Fish Biology 51, 135-154. Bailey, K. M., Powers, D. A., Quattro, J. M., Villa, R. G., Nishimura, A., Traynor, J. J. and Walters, G. (1999). Population Ecology and Structural Dynamics of Walleye Pollock, Theragra chalcogramma. In “Dynamics of the Bering Sea” (T. Loughlin and K. Ohtani, eds.). Alaska Sea Grant Press, University of Alaska, Fairbanks, AK (in press). Baird, T. A. and Olla, B. L. (1991). Social and reproductive behaviour of a captive group of walleye pollock Theragra chalcogramma. Environmental Biology of Fishes 30, 295-301. Bakkala, R. G. (1993). “Structure and historical changes in the groundfish comlplex of the eastern Bering Sea”. NOAA Technical Report NMFS 114,1-91. Bakkala, R. G., Wespestad, V. G. and Low, L.-L. (1987). Historical trends in abundance and current condition of walleye pollock in the eastern Bering Sea. Fisheries Research 5, 199-215. Balykin, P. A. (1996). Dynamics and abundance of western Bering Sea pollock.. In “Ecology of the Bering Sea: A Review of Russian Literature” (0.A. Mathiesen and K. 0. Coyle, eds), pp. 177-182. Alaska Sea Grant Report. No 96-01, University of Alaska, Fairbanks, AK. Bentzen, P., Taggart, C. T., Ruzzante, D. E. and Cook, D. (1996). Microsatellite polymorphism and the population structure of Atlantic cod (Gadus morhua) in
244
K. M. BAILEY ETAL.
the northwest Atlantic. Canadian Journal of Fisheries and Aquatic Sciences 53, 2706-2721. Bograd, S. J., Stabeno, P. J. and Schumacher, J. D. (1994). A census of mesoscale eddies in Shelikof Strait, Alaska, during 1989. Journal of Geophysical Research 99, 18243-18254. Brase, A. L. J. (1996). Temporal variation in feeding of larval walleye pollock Theragra chalcogramma in the SE Bering Sea region. MSc thesis, University of Alaska, Fairbanks, AK. Brodeur, R. D. (1998). Prey selection by age-0 walleye pollock, Theragra chalcogramma, in nearshore waters of the Gulf of Alaska. Environmental Biology of Fishes 51,175-186. Brodeur, R. D. and Merati, N. (1993). Predation on walleye pollock (Theragra chalcogramma) eggs in the western Gulf of Alaska: the roles of vertebrate and invertebrate predators. Marine Biology 117,483493. Brodeur, R. D. and Wilson, M. T. (1996). Mesoscale acoustic patterns of juvenile walleye pollock (Theragra chalcogramma) in the western Gulf of Alaska. Canadian Journal of Fisheries and Aquatic Sciences 53, 1951-1963. Brodeur, R. D., Bailey, K. M. and Kim, S. (1991). Cannibalism on eggs by walleye pollock Theragra chalcogramma in Shelikof Strait, Gulf of Alaska. Marine Ecology Progress Series 71,207-218. Brodeur, R. D., Wilson, M. T., Napp, J. M., Stabeno, P. J. and Salo, S. (1997). Distribution of juvenile pollock relative to frontal structure near the Pribilof Islands, Bering Sea. In “Ecology of Juvenile Walleye Pollock, Theragra chalcogramma” (R. D. Brodeur, P. A. Livingston, T. R. Loughlin and A. B. Hollowed, eds), pp. 573-588. NOAA Technical Report NMFS 126. Brown, J. H. (1984). On the relationship between abundance and distribution of species. American Naturalist 123, 253-279. Brown, J. H. (1995). “Macroecology”. University of Chicago Press, Chicago. Campana, S. E. and Gagne, J. A. (1995). Cod stock discrimination using ICPMS elemental assays of otoliths. In “Recent Developments in Fish Otolith Research” (D. H. Secor, J. M. Dean and S. E. Campana, eds), pp. 670-691. University of South Carolina Press, Columbia, SC. Canino, M. F., Bailey, K. M. and Incze, L. S. (1991). Temporal and geographic differences in feeding and nutritional condition of walleye pollock larvae Theragra chalcogramma in Shelikof Strait, Gulf of Alaska. Marine Ecology Progress Series 79,27-35. Carvalho, G. R. (1993). Evolutionary aspects of fish distribution: genetic variability and adaptation. Journal of Fish Biology 43, 53-73. Carvalho, G. R. and Hauser, L. (1994). Molecular genetics and the stock concept in fisheries. Reviews in Fish Biology and Fisheries 4, 326-350. Clark, S. H., Overholtz, W. J. and Hennemuth, R. C. (1982). Review and assessment of the Georges Bank and Gulf of Maine haddock fishery. Journal of Northwest Atlantic Fishery Science 3, 1-27. CLIMAP (1976). The surface of the ice-age Earth. Science 191,1131-1137. Corten, A.(1993). Learning processes in herring migrations. International Council for the Exploration of the Sea, Council Meeting, 1993/H:18. Daan, N. (1991). Bias in virtual population analysis when the unit stock assessed consists of sub-stocks. International Council for the Exploration of the Sea Council Meeting, 1991/D:17. Davis, M. W. and Olla, B. L. (1995). Formation and maintenance of aggregations in walleye pollock, Theragra chalcogramma, larvae under laboratory conditions:
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
245
role of visual and chemical stimuli. Environmental Biology of Fishes 44, 385-392. Dawson, €! (1994). The stock structure of Bering Sea walleye pollock (Theragra chalcogramma). M.S. thesis: University of Washington, Seattle, WA. Dell’Arciprete, 0. P. (1992). Growth, mortality, and transport of walleye pollock larvae (Theragra chalcogramma) in the eastern Bering Sea. M.S. thesis: University of Washington, Seattle, WA. deYoung, B. and Rose, G. A. (1993). On the recruitment and distribution of Atlantic cod (Gadus morhua) off Newfoundland. Canadian Journal of Fisheries and Aquatic Sciences 50, 2729-2741. Doherty, l? J., Planes, S. and Mather, I? (1995). Gene flow and larval duration in seven species of fish from the Great Barrier Reef. Ecology 76, 2373-2391 Dunn, J. R. and Matarese, A. C. (1987). A review of the early life history of northeast Pacific gadoid fishes. Fisheries Research 5, 163-184. Dwyer, D. A. (1984). Feeding habits and daily ration of walleye pollock (Theragra chalcograrnma)in the eastern Bering Sea. M.S. thesis: University of Washington, Seattle, WA. Dwyer, D. A., Bailey, K. M. and Livingston, €! A. (1987). Feeding habits and daily ration of walleye pollock (Theragra chalcograrnma) in the eastern Bering Sea, with special reference to cannibalism. Canadian Journal of Fisheries and Aquatic Sciences 44,1972-1984. Efremov, V. V., Mikhajlova, E. V. and Molokanov, V. B. (1989). Aconitase, a new and promising marker to be used in genetic studies of Theragra chalcograrnma populations. Genetika 25, 2263-2265 (English abstract only). Fair, L. F. (1994). Eastern Bering Sea walleye pollock: revised estimates of population parameters, relation of recruitment to biological and environmental variables, and forecasting. M.S. thesis: University of Alaska, Fairbanks, A K. Francis, R. C. and Bailey, K. M. (1983). Factors affecting recruitment of selected gadoids in the northeast Pacific and east Bering Sea. In “From Year to Year” (W. S. Wooster, ed.), pp. 3540. Washington Sea Grant, University of Washington, Seattle, WA. Frank, K. T. (1992). Demographic consequences of age-specific dispersal in maarine fish populations. Canadian Journal of Fisheries and Aquatic Sciences 49, 2222-2231. Galvin, P., Sadusky, T., McGregor, D. and Cross, T. (1995). Population genetics of Atlantic cod using amplified single locus minisatellite VNTR analysis. Journal of Fish Biology 47 (Suppl. A), 200-208. Garcia de Leon, F. J., Chikhi, L. and Bonhomme, F. (1997). Microsatellite polymorphism and population subdivision in natural populations of European sea bass Dicentrarchus labrax (Linnaeus, 1758). Molecular Ecology 6 , 51-62. Gilpin, M. (1991). The genetic effective size of a metapopulation. Biological Journal of the Linnean Society 42, 165-175. Gold, J. R., Richardson, L. R., Furman, C. and King, T. L. (1993). Mitochondrial DNA differentiation and population structure in red drum (Sciaenops ocellatus) from the Gulf of Mexico and Atlantic Ocean. Marine Biology 116, 175-185. Gold, J. R., Richardson, L. R., Furman, C. and Sun, F. (1994). Mitochondrial DNA diversity and population structure in marine fish species from the Gulf of Mexico. Canadian Journal of Fisheries and Aquatic Sciences 51 (Suppl. l), 205-214. Gong, Y., Hur, Y. H. and Kim, S. S. (1990). Comparison of meristic characters of Alaska pollock, Theragra chalcogramma, from seven geographic areas of’ the
246
K. M. BAILEY ETAL.
North Pacific. In “Compilation of Papers Presented at the International Symposium on Bering Sea Fisheries”, pp. 77-82, April 2-5 1990 in Khabarovsk, USSR. Unpublished report, Alaska Fisheries Science Center, Seattle, WA. Grant, W. S. and Bowen, B. W. (1998). Shallow population histories in deep evolutionary lineages of marine fishes: insights from sardines and anchovies and lessons for conservation. Journal of Heredity 89,415-426. Grant, W. S. and Leslie, R. W. (1993). Effect of metapopulation structure on nuclear and organellar DNA variability in semi-arid environments of southern Africa. South African Journal of Science 89,287-293. Grant, W. S. and Utter, F. M. (1980). Biochemical genetic variation in walleye pollock, Theragra chalcogramma: population structure in the southeastern Bering Sea and Gulf of Alaska. Canadian Journal of Fisheries and Aquatic Sciences 37, 1093-1100. Grant, W. S. and Utter, F. M. (1984). Biochemical population genetics of Pacific herring (Clupea pallasi). Canadian Journal of Fisheries and Aquatic Sciences 41, 85-64. Grant, W. S., Bakkala, R., Utter, F. M., Teel, D. J. and Kobayashi, G. (1987a). Biochemical genetic population structure of yellowfin sole, Limanda aspera, of the North Pacific Ocean and Bering Sea. Fishery Bulletin, US.81, 667-677. Grant, W. S., Zhang, C. I., Kobayashi, T. and StAhl G. (1987b). Lack of genetic stock discretion in Pacific cod (Gadus macrocephalus). Canadian Journal of Fisheries and Aquatic Sciences 44,490498. Grant, W. S., Garcia-Marin, J. L. and Utter, F. M. (1999). Defining population boundaries for fishery management. In “Genetics in Sustainable Fisheries Management” (S. Mustafy, ed.). Blackwell Scientific, Oxford. Grover, J. J. (1990). Feeding ecology of late-larval and early juvenile walleye pollock Theragra chalcogramma from the Gulf of Alaska in 1987. Fishery Bulletin, US. 88, 463-470. Haldorson, L., Paul, A. J., Sterritt, D. and Watts, J. (1989). Annual and seasonal variation in growth of larval walleye pollock and flathead sole in a southeastern Alaskan bay. Rapports et Proc2s-verbaux des RCunions. Conseil international pour I’Exploration de la Mer 191,220-225. Hanski, I. and Gilpin, M. (1991). Metapopulation dynamics: brief history and conceptual domain. In “Metapopulation Dynamics” (M. Gilpin and I. Hanski, eds), pp. 3-16. Academic Press, San Diego, CA. Hanski, I. and Simberloff, D. (1997). The metapopulation approach, its history, conceptual domain, and application to conservation. In “Metapopulation Biology” (I. Hanski and M. Gilpin, eds.), pp. 5-26. Academic Press, San Diego, CA. Harrison, S. and Taylor, A. D. (1997). Empirical evidence for metapopulation dynamics. In “Metapopulation Biology” (I. Hanski and M. Gilpin, eds), pp. 27-42. Academic Press, San Diego, CA. Hashimoto, R. and Koyachi, S. (1969). Biology of the Alaska pollock, Theragra chalcogramma (Pallas), distributed on the fishery grounds of the Tohoku districts and the Pacific coast of Hokkaido, southward from the Erimo ground. I. The morphological differentiation of the three types and the comparison with the other fishery ground’s groups. Bulletin of the Tohoku Regional Fisheries Research Laboratory 29,37-92. (In Japanese with English summary.) Hashimoto, R. and Koyachi, S. (1977). Geographical variation of relative growth of walleye pollock Theragra chalcogramma (Pallas). Bulletin of the Tohoku Regional Fisheries Research Laboratory 38, 41-74. (In Japanese with English summary.)
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
247
Heath, M. R. (1992). Field investigations of the early life stages of marine: fish. Advances in Marine Biology 28, 2-173. Hedgecock, D. (1994a). Temporal and spatial genetic structure of marine animal populations in the California Current. California Cooperative Oceanic Fisheries Investigative Report 35, 73-81. Hedgecock, D. (1994b). Does variance in reproductive success limit effective population sizes of marine organisms? In “Genetics and Evolution of Aquatic Organisms” (A. R. Beaumont, ed.), pp. 122-134. Chapman and Hall, London. Hermann, A. J., Rugen, W. C., Stabeno, P. J. and Bond, N. C. (1996). Physical transport of young pollock larvae (Theragra chalcogramma) near Shelikof .Strait as inferred from a hydrodynamic model. Fisheries Oceanography 5, 58-70’. Hillgruber, N., Haldorson, L. J. and Paul, A. J. (1995). Feeding selectivity of llarval walleye pollock Theragra chalcogramrna in the oceanic domain of the Bering Sea. Marine Ecology Progress Series 120, 1-10. Hinckley, S. (1987). The reproductive biology of walleye pollock, Theragra chalcogramma, in the Bering Sea, with reference to spawning stock structure. Fishery Bulletin, US 85, 481498. Hinckley, S., Bailey, K. M., Picquelle, S., Schumacher, J. D. and Stabeno, P. J. (1991). Transport, distribution and abundance of larval and juvenile walleye pollock (Theragra chalcogramma) in the western Gulf of Alaska. Canadian Journal of Fisheries and Aquatic Sciences 48, 91-98. Hjort, J. (1914). Fluctuations in the great fisheries of northern Europe. Rapports et Procss-verbaux des Re‘unions. Conseil international pour I’Exploration de la Mer 20, 1-228. Hollowed, A. B. and W. S. Wooster. (1995). Decadal-scale variations in the ea.stern subarctic Pacific 11: responses of northeast Pacific fish stocks. Canadian Special Publication of Fisheries and Aquatic Sciences 121, 373-385. Hollowed, A. B., Brown, E., Ianelli, J., Livingston, P., Megrey, B. and Wilson, C. (1997). Walleye Pollock. In “Stock assessment and fishery evaluation report for the groundfish resources of the Gulf of Alaska regions as projected for 1998”. North Pacific Fishery Management Council, Anchorage, AK. Iles, T. D. and Sinclair, M. (1982). Atlantic herring: stock discreteness and abundance. Science 215, 627-633. Ishida, T. (1954). On the age determination and morphometrical differences o f the otolith of Alaska pollack in the Hokkaido coast. Bulletin of the Hokkaido Regional Fisheries Research Laboratory 11, 3647. (In Japanese with English summary.) Iwata, M. (1973). Genetic polymorphism of tetrazolium oxidase in walleye pollock. Japanese Journal of Genetics 48, 147-149. Iwata, M. (1975a). Population identification of walleye pollock, Theragra chalcogramma (Pallas), population in the vicinity of Japan. Memoirs of the Fuculty of Fisheries, Hokkaido University 22, 193-258. Iwata, M. (1975b). Genetic identification of walleye pollock, Theragra chalcogramma (Pallas), populations on the basis of tetrazolium oxidase polymorphism. Comparative Biochemistry and Physiology 50B,197-201. Iwata, M. and Hamai, I. (1972). Local forms of walleye pollock, Theragra chalcogramma (Pallas) classified by number of vertebrae. Bulletin of the Japanese Society of Scientific Fisheries 38, 1129-1142. Janusz, J. (1994). How many pollock (Theragra chalcogramma) stocks support the fishery in the Sea of Okhotsk. Bulletin of the Sea Fisheries Institute 132, 6748. (English abstract.)
248
K. M. BAILEY ET AL.
Janusz, J., Linkowski, T. B. and Kowalewska-Pahlke, M. (1989). Results of the population studies of walleye pollock Theragra chalcogramma in the Bering Sea. In “Proceedings of the international scientific symposium on Bering Sea fisheries”, pp. 216-230. NOAA Technical Memorandum, NMFS F/NWC-163, Alaska Fisheries Science Center, Seattle, WA. Jarne, I? and Logoda, l? J. L. (1996). Microsatellites, from molecules to populations and back. Trends in Ecology and Evolution 11, 424-429. Johnson, A. G. (1977). A survey of biochemical variants found in groundfish stocks horn the North Pacific and Bering Sea. Animal Blood Groups Biochemistry and Genetics 8, 13-19. Kareiva, P. (1990). Population dynamics in complex environments. In “Population Regulation and Dynamics”. (M. P. Hassell and R. M. May, eds). Philosophical Transactions of the Royal Society of London B 330, 175-190. Kendall, A. W. and Nakatani, T. (1992). Comparisons of early-life-history characteristics of walleye pollock Theragra chalcograrnma in Shelikof Strait, Gulf of Alaska, and Funka Bay, Hokkaido, Japan. Fishery Bulletin, US. 90, 129-138. Kendall, A. W., Clarke, M. E., Yoklavich, M. M. and Boehlert, G. W. (1987). Distribution, feeding, and growth of larval walleye pollock, Theragra chalcogramma, from Shelikof Strait, Gulf of Alaska. Fishery Bulletin, US. 85, 499-521. Kendall, A. W., Incze, L. S., Ortner, F! B., Cummings, S. and Brown, P. K. (1994). Vertical distribution of eggs and larvae of walleye pollock, Theragra chalcogramma, in Shelikof Strait, Gulf of Alaska. Fishery Bulletin, US. 92, 540-554. Kendall, A. W., Schumacher, J. D. and Kim, S. (1996). Walleye pollock recruitment in Shelikof Strait: applied fisheries oceanography. Fisheries Oceanography 5 (Suppl. l ) , 4-18. Kihara, K. and Uda, M. (1969). Studies on the formation of demersal fishing grounds: 1. Analytic studies on the mechanism concerning the formation of demersal fishing grounds in relation to the bottom water masses in the eastern Bering Sea. Journal of Tokyo University of Fisheries 55, 83-90. Kim, S. (1987). Spawning behaviour and early life history of walleye pollock, Theragra chalcogramma, in Shelikof Strait, Gulf of Alaska, in relation to oceanographic factors. Ph.D. thesis: University of Washington, Seattle, WA. Kim, S. and Kendall, A. W. (1989). Distribution and transport of larval walleye pollock (Theragra chalcograrnma) in Shelikof Strait, Gulf of Alaska, in relation to water movement. Rapports et Procks-verbaux des RCunions. Conseil international pour 1’Exploration de la Mer 191, 127-136. Kimura, M. and Weiss, G. H. (1964). The stepping stone model of population structure and the decrease of genetic correlation with distance. Genetics 49, 561-576. Kinoshita, R. K., Grieg, A. and Terry, J. M. (1998). Economic status of the groundfish fisheries off Alaska, 1996. NOAA Technical Memorandum NMFSAFSC-85. Seattle, WA. Koyachi, S. and Hashimoto, R. (1977). Preliminary survey of variations of meristic characters of walleye pollack Theragra chalcograrnma (Pallas). Bulletin of the Tohoku Regional Fisheries Research Laboratory 38, 1740. (In Japanese with English summary.) Lawton, J. H., Nee, S., Letcher, A. J. and Harvey, F! H. (1994). Animal distributions: patterns and processes. In “Large-scale Ecology and Conserva-
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
249
tion” (F! J. Edwards, R. M. May and N. R. Webb, eds.), pp. 41-58. Blackwell Scientific, Oxford. Lear, W. H. and Wells, R. (1984). Vertebral average of juvenile cod, Gadus morhua, from coastal waters of eastern Newfoundland and Labrador as indicators of stock origin. Journal of Northwest Atlantic Fisheries Science 5 , 23-31. Lincoln, R. (1994). Molecular genetics applications in fisheries: snake oil or restorative. Reviews in Fish Biology and Fisheries 4, 389-392. Livingston, P. A. (1989). Interannual trends in walleye pollock, Theragra chalcogramma, cannibalism in the eastern Bering Sea. In “Proceedings of the International Symposium on the Biology and Management of Walleye Pollock”, pp. 275-296, November 14-16, 1988. Alaska Sea Grant Report 89-1, University of Alaska, Fairbanks, AK. Livingston, F! A. (1993). The importance of predation by groundfish, marine mammals, and birds on walleye pollock, Theragra chalcogramma, and Pacific herring, Clupea pallasi, in the eastern Bering Sea. Marine Ecology Progress Series 102,205-215. Lloyd, M. (1967). Mean crowding. Journal of Animal Ecology 36, 1-30. Lynde, C. M., Lynde, M. V. H. and Francis, R. C. (1986). Regional and temiporal differences in growth of walleye pollock (Theragra chalcogramma) in the eastern Bering Sea and Aleutian Basin with implications for management. Unpublished report, Northwest and Alaska Fisheries Center, Seattle, WPL. MacArthur, R. H. and Wilson, E. 0. (1967). ‘‘The Theory of Island Biogeography”. Princeton University Press, Princeton, NJ. MacCall, A. (1990). “Dynamic Geography of Marine Fish Populations”. University of Washington Press, Seattle, WA. Maeda, T. (1972). Fishing grounds of the Alaska pollock. Bulletin of the Japanese Society of Scientific Fisheries 43, 39-45. (In Japanese with English summary.) Marasco, R. and Aron, W. (1991). Explosive evolution - the changing Alaska groundfish fishery. Reviews in Aquatic Sciences 4, 299-315. Matthews, K. R. (1990). An experimental study of the habitat preferences and movment patterns of copper, quillback, and brown rockfishes (Sebastes spp.). Environmental Biology of Fishes 29, 161-178. Megrey, B. A. (1988). On the contribution of density-dependent factors to year class variations in Alaskan walleye pollock. In “Proceedings from workshcop on year class variations as determined from pre-recruit investigations”, pp. 425-456. Institute of Marine Research, Bergen, Norway. Megrey, B. A. (1989). Exploitation of walleye pollock resources in the Gulf of Alaska, 1964-88: portrait of a fishery in transition. In “Proceedings of the International Symposium on the Biology and Management of Walleye Polllock”, pp. 33-58, November 1416,1988. Alaska Sea Grant Rept. No 89-1, University of Alaska, Fairbanks, AK. Megrey, B. A., Bograd, S. J., Rugen, W. C., Hollowed, A. B., Stabeno, F! J., Macklin, S. A., Schumacher, J. D. and Ingraham, W. J. (1995). An exploratory analysis of associations between biotic and abiotic factors and year-class strength of Gulf of Alaska walleye pollock (Theragra chalcogramma). Canadian Special Publication of Fisheries and Aquatic Sciences 121,227-243. Megrey, B. A., Hollowed, A. B., Hare, S. R., Macklin, S. A. and Stabeno, l? J. (1996). Contributions of FOCI research to forecasts of year-class strength of walleye pollock in Shelikof Strait, Alaska. Fisheries Oceanography 5 (Suppl. l), 189-203.
250
K.
M. BAILEY ET AL.
Merati, N. and Brodeur, R. D. (1996). Feeding habits and daily ration of juvenile walleye pollock, Theragra chalcogramma, in the western Gulf of Alaska. In “Ecology of Juvenile Walleye Pollock, Theragra chalcogramma” (R. D. Brodeur, I? A. Livingston, T. R. Loughlin and A. B. Hollowed, eds), pp. 65-80. NOAA Technical Report NMFS 126. Merrick, R. L., Chumbley, M. K. and Byrd, G. V. (1997). Diet diversity of Steller sea lions (Eumetopias jubatus) and their population decline in Alaska: a potential relationship. Canadian Journal of Fisheries and Aquatic Sciences 54, 1342-1348. Morgan, L. V., Carr, S. M., Bowering, W. R. and Davidson, W. S. (1997). Greenland halibut (Reinhardtius hippoglossoides) in the North Atlantic are genetically homogeneous. Canadian Journal of Fisheries and Aquatic Sciences 54, 1813-1821. Mork, J., Ryman, N., StBhl, G., Utter, F. and Sundnes, G. (1985). Genetic variation in Atlantic cod (Gadus morhua) throughout its range. Canadian Journal of Fisheries and Aquatic Sciences 42, 1580-1587. Mulligan, T. J., Bailey, K. M. and Hinckley, S. (1989). The occurrence of larval and juvenile walleye pollock, Theragra chalcogramma, in the eastern Bering Sea with implications for stock structure. In “Proceedings of the International Symposium on the Biology and Management of Walleye Pollock”, pp. 471490, November 1616,1988. Alaska Sea Grant Report no 89-1, University of Alaska, Fairbanks, AK. Mulligan, T. J., Chapman, R. W. and Brown, B. L. (1992). Mitochondria1 DNA analysis of walleye pollock, Theragra chalcogramma, from the eastern Bering Sea and Shelikof Strait, Gulf of Alaska. Canadian Journal of Fisheries and Aquatic Sciences 49, 319-326. Nakano, H., Takahashi, M. and Kikuchi, K. (1991). Geographical differences of walleye pollock caught from three Bering Sea areas from standpoint of metal and amino acid contents of otoliths. Unpublished report: International Pacific Fisheries Commission Document No. 3668, Fisheries Agency of Japan, National Research Institute of Far Sea Fisheries, 7-1 Orido 5 chome, Shimizu, Shizuoka, Japan. Nakatani, T. (1988). Studies on the early life history of walleye pollock Theragra chalcogramma in Funka Bay and vicinity, Hokkaido. Memoirs of the Faculty of Fisheries, Hokkaido University 35, 1-46. Nakatani, T. and Maeda, T. (1983). Distribution of walleye pollock larvae and their food supply in Funka Bay and the adjacent waters, Hokkaido. Bulletin of the Japanese Society of Scientific Fisheries 49, 183-187. Napp, J. M., Incze, L. S., Ortner, I? B., Siefert, D. L. W. and Britt L. (1996). The plankton of Shelikof Strait, Alaska: standing stock, production, mesoscale variability and their relevance to larval fish survival. Fisheries Oceanography 5 (Suppl. l), 19-38. National Research Council (1996). “The Bering Sea Ecosystem”. National Academy Press, Washington, DC. Niebauer, H. J. and Day, R. H. (1989). Causes of interannual variability in the sea ice cover of the eastern Bering Sea. GeoJournall8, 45-59. Nishimura and Yamada (1984). Age and growth of larval and juvenile walleye pollock Theragra chalcogramma (Pallas), as determined by otolith daily growth increments. Journal of Experimental Marine Siology and Ecology 82, 191-205. Nitta, A. and Sasaki, T. (1990). Study on stock identification of walleye pollock based on morphometric data. In “Compilation of papers presented at the
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
251
international symposium on Bering Sea fisheries”, pp. 74-76, April 2-5, 1990 in Khabarovsk, USSR. Unpublished report, Alaska Fisheries Science Center, Seattle, WA. Ogata, T. (1959). Population studies of the Alaska pollack in the Sea of Japan I. On the variation in the vertebral count. Annual Report of the Japan Sea Regional Fisheries Research Laboratory 5, 119-125. (In Japanese with English summary.) Ohtani, K. and Azumaya, T. (1995). Influence of interannual changes in ocean conditions on the abundance of walleye pollock Theragra chalcogramma in the eastern Bering Sea. In “Climate change and northern fish populations”. (R. J. Beamish, ed.). Canadian Special Publication in Fisheries and Aquatic Sciences 121, 87-95. Olla, B. L., Davis, M. W., Ryer, C. H. and Sogard, S. M. (1996). Behavioral determinants of distribution and survival in early stages of walleye pollock, Theragra chalcogramma: a synthesis of experimental studies. Fisheries Ckeanography 5 (Suppl. l), 167-178. Ovenden, J. R. (1990). Mitochondria1 DNA and marine stock assessment: a review. Australian Journal of Marine and Freshwater Research 41, 835-853. Palsson, W. A., Hoeman, G. G., Bargmann, G. G. and Day, D. (1996). “1995 Status of Puget Sound Bottomfish Stocks”. Unpublished report, Washington D,epartment of Fisheries and Wildlife, Olympia, WA. Palumbi, S. R. (1995). Using genetics as an indirect estimator of larval dispersal. In “Marine Invertebrate Larvae” (L. McEdward, ed.), pp. 369-387. CRC Press, New York. Paul, A. J., Paul, J. M. and Clayton, L. A. (1997). Copepod nauplii abundance in an Alaskan glaciated fjord during spring walleye pollock hatch in 1988 and 1991. In “Forage Fishes in Marine Ecosystems”, pp. 633-643. Alaska Sea Grant Press Report, no. 97-01, University of Alaska, Fairbanks, AK. Pawson, M. G. and Jennings, S. (1996). A critique of methods for stock identification in marine capture fisheries. Fisheries Research 25, 203-217. Pepin, l? and Carr, S. M. (1993). Morphological, meristic, and genetic analysis of stock structure in juvenile Atlantic cod (Gadus morhua) from the Newfoundland shelf. Canadian Journal of Fisheries and Aquatic Sciences 50, 1924-1933. Pethon, l? (1989). “Aschehougs Store Fiskebok”. H. Aschehoug, Stockholm, Sweden. Pogson, G. H., Mesa, K. A. and Boutilier, R. G. (1995). Genetic population structure and gene flow in the Atlantic cod Gadus morhua: A comparison of allozyme and nuclear RFLP loci. Genetics 139, 375-385. Pope, J. G. (1972). An investigation of the accuracy of virtual population arialysis using cohort analysis. Research Bulletin of the International Commission of Northwest Atlantic Fisheries 9, 65-74. Proctor, C. H. and Thresher, R. E. (1988). Effects of specimen handling and otolith preparation on concentration of elements in fish otoliths. Marine Biology 131, 681694. Purcell, M. K., Kornfield, I., Fogarty, M. and Parker, A. (1996). Interdecadal heterogeneity in mitochondria1 DNA of Atlantic haddock (Melanogrammus aeglefinus) from Georges Bank. Molecular Marine Biology and Biotechno1,ogy5, 185-1 92. Quinn, T. J., I1 and Collie, J. S. (1990). Alternative population models for eastern Bering Sea pollock. INPFC Symposium on application of stock assessment
252
K. M. BAILEY ETAL.
techniques to gadids. International North Pacific Fisheries Commission Bulletin 50,243-258. Quinn, T. J., 11, and Deriso, R. B. (1999). “Quantitative Fish Dynamics”. Oxford University Press, New York, 542 pp. Qu~M,T. J., I1 and Niebauer, H. J. (1995). Relation of eastern Bering Sea walleye pollock (Theragra chalcogramma) recruitment to oceanographic variables. Canadian Special Publication of Fisheries and Aquatic Sciences 121, 497-507. Quinn, T. €? (1993). A review of homing and straying of wild and hatcheryproduced salmon. Fisheries Research 18, 29-44. Ricker, W. E. (1975). Computation and interpretation of biological statistics of fish populations. Bulletin of the Fisheries Research Board of Canada 191,382 pp. Rose, G. A. (1993). Cod spawning on a migration highway in the north-west Atlantic. Nature 366,45&461. Ruzzante, D. E. (1998). A comparison of several measures of genetic distance and population structure with microsatellite data: bias and sampling variance. Canadian Journal of Fisheries and Aquatic Sciences 55, 1-14. Ruzzante, D. E., Taggart, C. T., Cook, D. and Goddard, S. (1996a). Genetic differentiation between inshore and offshore Atlantic cod (Gadus morhua L.) off Newfoundland microsatellite DNA variation and antifreeze level. Canadian Journal of Fisheries and Aquatic Sciences 53, 634-645. Ruzzante, D. E., Taggart, C. T. and Cook, D. (1996b). Spatial and temporal variation in the genetic composition of a larval cod (Gadus morhua) aggregation: cohort contribution and genetic stability. Canadian Journal of Fisheries and Aquatic Sciences 53, 2695-2705. Ryer, C. H. and Olla, B. L. (1997). Altered search speed and growth social versus independent foraging in two pelagic juvenile fishes. Marine Ecology Progress Series 153,273-281. Ryman, N. and Utter, F. (eds.) (1987). “Population Genetics and fishery Management”. University of Washington Press, Seattle, WA. Saito, T. (1972). Symposium on the Alaska pollock fishery and resources: Introduction. Bulletin of the Japanese Society of Scientific Fisheries 38, 413. Sakurai, T. (1982). Reproductive ecology of walleye pollock Theragra chalcogramma (Pallas). Ph.D. thesis, Hokkaido University (English translation available: Alaska Fisheries Science Center, Seattle, WA). Sasaki, T. (1989). Synopsis of biological information on pelagic pollock resources in the Aleutian Basin. In “Proceedings of the International Scientific Symposium on Bering Sea Esheries”, pp. 90-182, July 19-21, 1988, Sitko, Alaska. NOAA Technical Memorandum NMFS F/NWC 163. Saunders, M. W., McFarlane, G. A. and Shaw, W. (1989). Delineation of walleye pollock (Theragra chalcogramma) stocks off the Pacific coast of Canada. In “Proceedings of the International Symposium on the Biology and Management of Walleye Pollock”, pp. 379-401, November 14-16, 1988. Alaska Sea Grant Report, no 89-1, University of Alaska, Fairbanks, AK. Schneider, D. C., Methven, D. A. and Dalley, E. L. (1997). Geographic contraction in juvenile fish a test with northern cod (Gadus morhua) at low abundances. Canadian Journal of Fisheries and Aquatic Sciences 54, 187-199. Schumacher, J. D. and Kendall, A. W. (1995). An example of fisheries oceanography: walleye pollock in Alaskan waters. Reviews in Geophysics July 1995, 1153-1 163. Schumacher, J. D. and Stabeno, €? J. (1994). Ubiquitous eddies of the eastern
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
253
Bering Sea and their coincidence with concentrations of larval pollock. Fisheries Oceanography 3, 182-190. Schumacher, J. D. and Stabeno, F! J. (1997). The continental shelf of the Bering Sea. The Sea 7, (in press). Schumacher, J. D., Stabeno, P. J. and Bograd, S. J. (1993). Characteristics of an eddy over a continental shelf Shelikof Strait, Alaska. Journal of Geophysical Research 98, 8395-8404. Seeb, L. W. and Gunderson, D. R. (1988). Genetic variation and population structure of Pacific ocean perch (Sebastes alutus). Canadian Journal of Fisheries and Aquatic Sciences 45, 78-88. Serobaba, I. I. (1977). Data on the population structure of the walleye pollock, Theragra chalcogramma (Pallas), from the Bering Sea. Voprosy Ikhtiologii 17, 247-260. (In Russian.) (English translation 1978. Journal of Ichthyo1ol;y 17, 219-23 1.) Severin, K. P., Carroll, J. and Norcross, B. L. (1995). Electron microprobe analysis of juvenile walleye pollock, Theragra chalcogramma, otoliths from Alaska: a pilot stock separation study. Environmental Biology of Fishes 43, 269-283. Shaw, W. and McFarlane, G. A. (1986). Biology, distribution and abundance of walleye pollock (Theragra chalcogramma) off the west coast of Canada. International North Pacific Fisheries Commission Bulletin 45, 262-283. Shields, G. F. and Gust, J. R. (1995). Lack of geographic structure in mitochondria1 DNA sequences of Bering Sea walleye pollock, Theragra chalcogrmnma. Molecular Marine Biology and Biotechnology 4, 69-82. Shulman, M. J. and Bermingham, E. (1995). Early life histories, ocean currents, and the population genetics of Caribbean reef fishes. Evolution 49, 897-5110. Shuntov, V. P. (1992). Functional structure of the distribution area of walleye pollock in the Bering Sea. Russian Journal of Marine Biology (transll.) 17, 189-198. Sinclair, E. H., Antonellis, G. A., Robson, B. W., Ream, R. R. and LoughIin, T. R. (1996). Northern fur seal, Callorhinus ursinus, predation on juvenile walleye pollock, Theragra chalcogramrna. In “Ecology of Juvenile Walleye Pollock, Theragra chalcogramma” (R. D. Brodeur, F! A. Livingston, T. R. Loughljn and A. B. Hollowed, eds.), pp. 167-178. NOAA Technical Report, NMFS 126. Sinclair, M. (1988). “Marine Populations: An Essay on Population Regulation and Speciation”. Washington Sea Grant Press, Seattle, WA. Slatkin, M. (1993). Gene flow and population structure. In “Ecological Genetics” (L. A. Real, ed.), pp. 3-17. Princeton University Press, Princeton, NJ. Smith, G. B. (1981). The biology of walleye pollock. In “The Eastern Bering Sea Shelf Oceanography and Resources” (D. Hood and J. Calder, eds), pp. 235-261. US Government Printing Office, Washington, DC. Sogard, S. M. and Olla, B. L. (1993). The influence of predator presence on utilization of artificial seagrass habitats by juvenile walleye pollock, Theragra chalcogramma. Environmental Biology of Fishes 37, 5745. Sogard, S. M. and Olla, B. L. (1996a). Die1 patterns of behaviour in juvenile walleye pollock, Theragra chalcogramma. Environmental Biology of Fishes 47, 379-386. Sogard, S. M. and Olla, B. L. (1996b). Food deprivation affects vertical distribution and activity of a marine fish in a thermal gradient: potential energy-conserving mechanisms. Marine Ecology Progress Series 133, 43-55. Somerton, D. (1978). Competitive interaction of walleye pollock and Pacific Ocean perch in the northern Gulf of Alaska. In “Fish Food Habit Studies” (S. J.
254
K. M. BAILW ETAL.
Lipovsky and C. A. Simenstad, eds.), pp. 163 (Abstract). Washington Sea Grant, University of Washington, Seattle, WA. Springer, A. M. (1992). A review: walleye pollock in the North Pacific. How much difference do they really make? Fisheries Oceanography 5, 205-223. Stabeno, P. J., Schumacher, J. D., Bailey, K. M., Brodeur, R.D. and Cokelet, E. D. (1996a). Observed patches of walleye pollock eggs and larvae in Shelikof Strait, Alaska: their characteristics, formation and persistence. Fisheries Oceanography 5 (Suppl. 1), 81-91. Stabeno, P. J., Reed, R. K. and Schumacher, J. D. (1996b). The Alaska coastal current: continuity of transport and forcing. Journal of Geophysical Research 100,2477-2485. Stepanenko, M. A. (1997). Variations from year to year in the spatial differentiation of the walleye pollock, Theragra chalcogramma, and the cod, Gadus macrocephalus, in the Bering Sea. Journal of Ichthyology 37, 14-20. Svetovidov, A. N. (1948). “Fauna of USSR fishes. Gadiformes. Volume 9, Number 4”. Zoology Institute of the Academy of Science. Moscow, USSR. Swain, D. I? and Wade, E. J. (1993). Density-dependent geographic distribution of Atlantic cod (Gadus morhua) in the southern Gulf of St Lawrence. Canadian Journal of Fisheries and Aquatic Sciences 50, 725-733. Swartzman, G. and Haar, R. T. (1983). Interactions between fur seal populations and fisheries in the Bering Sea. Fishery Bulletin, US.81, 121-132. Swartzman, G., Stuetzle, W., Kulman, K. and Powojowski, M. (1994). Relating the distribution of pollock schools in the Bering Sea to environmental factors. ICES Journal of Marine Science 51, 481492. Tang, Q., Jin, X., Li, F., Chen, J., Wang, W., Chen, Y., Zhao, X. and Dai, F. (1996). Summer distribution and abundance of age-0 walleye pollock, Theragra chalcogramma, in the Aleutian Basin. In “Ecology of Juvenile Walleye Pollock, Theragra chalcogramma” (R.D. Brodeur, P. A. Livingston, T. R. Loughlin and A. B. Hollowed, eds), pp. 3546. NOAA Technical Report, NMFS 126. Temnykh, 0. S. (1991). The population structure of Alaska pollock in the Okhotsk Sea. Rybnoye Khozyaistvo 6, 28-31. (English abstract only.) Temnykh, 0.S. (1994). Morphological differentiation of Alaskan pollock, Theragra chalcogramma, in the west Bering Sea and Pacific waters of Kamchatka. Journal of Ichthyology 34, 204-211. Templeman, W. T. (1981). Vertebral numbers in Atlantic cod, Gadus morhua, of the New Foundland and adjacent seas, 1947-1971, and their use in delineating cod stocks. Journal of Northwest Atlantic Fishery Science 2, 2145. Theilacker, G. H., Bailey, K. M., Canino, M. F. and Porter, S. M. (1996). Variations in larval walleye pollock feeding and conditions: a synthesis. Fisheries Oceanography 5 (Suppl. l ) , 112-123. Thompson, J. M. (1981). Preliminary report on the population biology and fishery of walleye pollock (Theragra chalcogramma) off the Pacific coast of Canada. Canadian Technical Report of Fisheries and Aquatic Sciences 1031, 157 pp. Tsugi, S. (1989). Alaska pollack population, Theragra chalcogramma, of Japan and its adjacent waters, I: Japanese fisheries and population studies. Marine Behavior and Physiology 15, 147-205. Vasil’kov, V.P. and Glebova, S. Yu. (1984). Factors determining year-class strength of the walleye pollock, Theragra chalcogramma (Gadidae), of Western Kamchatka. Journal of Ichthyology 24, 80-90. Vastano, A. C., Incze, L. S. and Schumacher, J. D. (1992). Observation and analysis
POPULATION STRUCTURE AND DYNAMICS OF WALLEYE POLLOCK
255
of fishery processes: larval pollock at Shelikof Strait, Alaska. Fisheries Oceanography 1, 20-31. Waples, R. S. (1987). A multispecies approach to the analysis of gene flow in marine shore fishes. Evolution 41, 385-400. Waples, R. S. (1998). Separating the wheat from the chaff patterns of !genetic differentiation in high gene flow species. Journal of Heredity (In Press). Waples, R. S. and Rosenblatt, R. H. (1987). Patterns of larval drift in southern California marine shore fishes inferred from allozyme data. Fishery Bulletin, US. 85, 1-11. Wespestad, V. G. (1993). The status of Bering Sea pollock and the effect of the “Donut Hole” fishery. Fisheries 18, 18-24. Wespestad, V. G. (1995). Walleye Pollock. In “Stock assessment and lishery evaluation report for the groundfish resources of the Bering SealAleutian Islands regions as projected for 1996”. North Pacific Fishery Management Council, Anchorage, AK. Wespestad, V. G. (1996). Unpublished document prepared for Nor-fishing ’96, Trondheim, Norway, 15-16 August 1996. See: www.wrc.noaa.gov/-jianellil home.htm1. Wespestad, V. G. and Quinn, T. J. 11. (1996). Importance of cannibalism in the population dynamics of walleye pollock, Theragra chalcograrnma. NOAA Technical Report NMFS 126, 212-216. Wespestad, V. G., Ianelli, J. N., Fritz, L., Honkalehto, T., Williamson, PJ. and Walters, G. (1997). Bering Sea - Aleutian Islands Walleye Pollock Asse,;c sment for 1998. In “Stock assessment and fishery evaluation report for the groundfish resources of the Bering Sea/Aleutian Islands regions as projected for 1998”. North Pacific Fishery Management Council, Anchorage, AK. Wilimovsky, N. J., Peden, A. and Peppar, J. (1967). Systematics of six deimersal fishes of the North Pacific Ocean. Fisheries Research Board of Canada Technical Report 34, 95 pp. Wright, J. M. and Bentzen, P. (1994). Microsatellites: genetic markers f,or the future. Reviews in Fish Biology and Fisheries 4, 384-388. Wright, S. (1931). Evolution in Mendelian populations. Genetics 16,97-159. Wyllie-Echeverria, T. (1995). Sea-ice conditions and the distribution of walleye pollock (Theragra chalcogramma) on the Bering and Chukchi Sea shelf. Canadian Special Publication in Fisheries and Aquatic Sciences 121, 131-136. Wyllie-Echeverria, T. and Wooster, W. S. (1998). Year-to-year variations in ]Bering Sea ice cover and some consequences for fish distributions. Fisheries Oceanography 7 , 159-170. Yoshida, H. (1989). Walleye pollock fishery and fisheries management in the Nemuro Strait, Sea of Okhotsk, Hokkaido. In “Proceedings of the International Symposium on the Biology and Management of Walleye Pollock”, pp. 59-78, November 1416,1988. Alaska Sea Grant Report, no 89-1, University of Alaska, Fairbanks, AK. Yoshida, H. (1994). Food and feeding habits of pelagic walleye pollock in the central Bering Sea in summer, 197680. Scientific Report of the Hokkaido Fisheries Experimental Station 45, 1-35. Zwanenburg, K. C. T, Bentzen, P. and Wright, J. M. (1992). Mitochondria1 DNA differentiation in western North Atlantic populations of haddock (Melanogrammus aeglefinus). Canadian Journal of Fisheries and Aquatic Sciences 49, 2527-2537.
This Page Intentionally Left Blank
Taxonomic Index Page references in italics refer to Figures; those in bold refer to Tables
M Acartia tonsa 131, 132 algae see under subject index American oyster see Crassostrea virginica Arbacia 131 Argopecten irradians 10n Artemia salina 131, 132 Bacillus alvei and B. thuringiensis 107 Balanus 132 Brachionus plicatilis 131, 132 Callista brevisiphonata 10n Cerastoderma edule 10n Ceratium fusus 125 Chaetoceros calcitrans 39, 65 Chlamys asperrima 10, 37 bioassay methodology 43, 44,48 bioassay procedures 50, 531, 59 toxicity testing 91,100,110-1 Chlorella 39 Chrysochromulina polylepis 125 clam coot see Mulinia lateralis hard see Mercenaria mercenaria surf see Spisula Clinocardium nuttalli 10n Cochlodinium heterolobatum 69, 125 coot clam see Mulinia lateralis copepods 131 Crassostrea generally 5, 87, 135 bioassays 45, 53, 116 reproduction, biology of 22, 25 Crassostrea angulata 10, 11, 18 bioassay procedures 50-1 toxicity testing 113-4 Crassostrea cucculata 10, 16 toxicity testing 90,92,93 Crassostrea gigas 2, 8-10, 10, 131 bioassay methodology 45, 46, 48
seawater quality 41, 42, 43, 4.4 bioassay procedures 50, 65, 66 embryo-larval 51, 52, 54, 56, 57-8, 59, 60 larval physiology tests 68, 69, 71, 73 described 10-13 larval rearing in laboratory 26 nutritional requirements 38, 39, 40 physical requirements 36, 36-7, 37 spawning and fertilization 29-30, 29, 32,32 reproduction biology of 18, 20, 21, 24 spawning and fertilization 2!3-30, 29, 32, 32, 46, 48 toxicity testing 77, 87, 118, 128, 135 biocides 98-106,97, 107 detergents and oil 110,1U-5 metals 90,92-5, 120 Crassostrea iredalei 10,91 Crassostrea margaritacea 10, 16 toxicity testing 90,92,93 Crassostrea rhizophorae 10, 16, 37, 42 Crassostrea virginica 2, 6, 8-10, ILO, 131 bioassay methodology 42, 44,46, 48 bioassay procedures 63 embryo-larval 52, 54-6 larva[ physiology' tests 66-7, 69, 70. 71. 73 described 12-13 larval rearing in laboratory 25-6, 39 physical requirements 36-7, 37 spawning and fertilization 2 7 , 29-30, 32,3 2 4 reproduction biology of 16, 17, 19-20, 21, 23 spawning and fertilization 27, 29-30, 29, 32, 3 2 4 , 46, 48 toxicity testing 122, 125, 135
258 Crassostrea virginica bioassay methodology (cont.) biocides 98-106, 107 detergents and oil 110, 1124, 116 metals 90, 94, 95, 95 crustaceans see under subject index Cyclotella nana 65 Dinophilus 132 Dunaliella euchlora 66 Eastern oyster see Crassostrea virginica Echinus esculentus 7 European oyster see Ostrea edulis Gadus morhua see cod, Atlantic under subject index Gymnodinium splendens 125 Gyrodinium aureolum 125 Homarus 132 Zsochrysis galbana 6, 39 Isognomon californicum 10, 16, 30, 47, 48 bioassay procedures 51, 52, 53 toxicity testing 91, 93-4, 98-9 Japanese oyster see Crassostrea gigas Macoma balthica 1On Mactra chilensis 10n Mediterranean mussel see Mytilus galloprovincialis Melanogrammus aeglefinus 208, 220 Mercenaria generally 115 Mercenaria campechiensis 14 Mercenaria mercenaria 6, 8-9, 10, 131 bioassay methodology 42, 44,46, 47, 48 bioassay procedures 70 embryo-larval 52, 53, 54-6 larval physiology tests 63, 64, 65 described 14-15 larval rearing in laboratory 26-7 physical requirements 36, 36-7, 37 spawning and fertilization 27, 29-30,29,32, 32 reproduction 46, 47-8 biology 16-17, 18, 19-20 toxicity testing 135 biocides 98-100, 1024, 106, 107
TAXONOMIC INDEX
detergents and oil 109, 111-2 metals 91, 93-4 Mercenaria (Venus) mercenaria 7 Meretrix lusoria 1011, 46 Metharyzium anisopliae 107 Mizuhopecten yessoensis 10n, 45 Monochrysis lutheri 39 bioassay procedures 65, 66, 72, 73 Mulinia lateralis 7, 8-9, 10, 131 bioassay applications 77 bioassay procedures 51, 52, 56, 60, 63-4,68 described 15-16 future research 137 larval rearing in laboratory 26 physical requirements 36-7, 37 spawning and fertilization 27, 29-30, 29, 32, 32 reproduction 17, 18 spawning and fertilization 27, 29-30, 29, 32, 32 toxicity testing 91, 94, 101, 107, 135 detergents and oil 110, 113-5, 116 mussel blue see Mytilus edulis Mediterranean see Mytilus galloprovincialis Mya arenaria 10n, 55, 71,101-2 Mysidopsis bahia 131 Mytilus generally 5, 29, 46, 135 bioassay procedures 51, 72 larval rearing in laboratory 39 reproduction biology of 22 toxicity testing 100, 136 Mytilus califomianus 9, 15, 17, 52 larval rearing in laboratory 32, 48 Mytilus edulis 2, 8-10, 10, 125, 131 bioassay methodology 42, 44, 45 bioassay procedures 52, 56, 59, 64, 65, 66, 69 described 13-14 larval rearing in laboratory physical requirements 36, 36, 35437 spawning and fertilization 27, 29-30, 32, 32-6,46,47-8 reproduction biology of 18, 19, 20, 21 spawning and fertilization 27, 29-30, 31, 32-6, 4 6 , 4 7 4
259
TAXONOMIC INDEX
Mytilus edulis (cont.) toxicity testing 113, 127, 135 bioasay applications 77 biocides 98-100, 102, 10&5, 107 detergents and oil 110, 11214 metals 91, 93-4, 95, 120 Mytilus galloprovincialis 2, 8-9, 131 bioassay methodology 42, 43, 44 described 13-14 larval rearing in laboratory 46 nutritional requirements 38 physical requirements 36, 5 7 , 37 spawning and fertilization 27, 29-30, 32, 32-6 reproduction 18 spawning and fertilization 27, 29-30, 32, 32-6 Mytilus trossulus 9, 14, 16, 45, 56 Nannochloris 38 Onchorhynchus 132 Ostrea edulis 8, 131 bioassay methodology 41, 42, 44 bioassay procedures 68, 71 described 13 larval rearing in laboratory 26 nutritional requirements 38, 39 physical requirements 37 spawning and fertilization 27, 29, 32 reproduction biology of 16, 17-18, 21, 22, 23 spawning and fertilization 27, 28, 32 toxicity testing 97 detergents and oil 109, 111, 113, 116 metals 92, 94 Ostrea plicatula 8-9, 16
oyster Americafiastern see Crassosi’rea virginica European see Ostrea edulis Hawaian see Zsognomon californicum JapanesePacific see Crassostrea gigas see also Saccostrea Pacific oyster see Crassostrea gigos Paracentrotus 131 Pecten maximus 10n, 48, 64, 9 3 4 Phaeodactylum tricornutum 66 polychaetes 130 Prorocentrum minimum 67, 125 Protothaca staminea 10n, 110 quahog (hard clam) see Mercenaria mercenaria
Rangia cuneata lOn, 68 Saccostrea commercialis 10n, 16, 95 bioassays 43, 44,72 Saccostrea sakhalinensis 9 Scrobicularia plana 10 bioassay procedures 70, 72-3 toxicity testing 91, 99 sea urchins 7, 131 Spisula solidissima 10, 43 bioassay procedures 52, 56, 67-8 toxicity testing 91, 95 Strongylocentrotus 131 Tapes philippinarum 10n, 46 Tetraselmis suecica 26, 39 Theragra chalcogramma see walbeye pollock in subject index Theragra finmarchicus 183 Tresus capax and I: nuttali lOn, 110
This Page Intentionally Left Blank
Subject Index Notes 1. Page references in italics indicate diagrams and those in bold indicate tables. 2. Walleye pollock in sub-entries is abbreviated to pollock abnormality of bivalve larvae at end of bioassay 55-6, 57-8,59-62 acquired characteristics of pollock 213, 214,215 Adak Island, pollock near 217, 219 age of bivalve egg at beginning of test 51-2 of bivalve larva and toxicity reaction 118-19 see also body sue; eggs; juveniles; larvae Alaska see Gulf of Alaska aldrin toxicity and bivalves 103, 97 Aleutian Basin, pollock in 179, 180, 184,187 population dynamics 194-9 population structure genetic 216,217,219, 219,221,223 mechanisms 232-4 metapopulation 225-6,228 phenotypic 210,211,212-13 stocks 185 Aleutian Islands, pollock near 185 population dynamics 194, 195-6, 199 population structure 216,217,219, 221,223,232 algae/dinoflagellatesand bivalves 6 bioassay procedures 64,65,66-8, 69-70,72,73 larval rearing in laboratories 26,38-9 toxicity 125-6 allozyme studies of pollock 207, 215-16, 218,222,223,237 aluminium toxicity and bivalves 87,94 ANOVA approach 84,221-2 antibiotics and bivalves 42-3,54, 63 arsenic toxicity and bivalves 87-8, 90-1 artificial seawater quality and bivalves 43, 44,136
atrazine-simazine toxicity and bivalves 66,97, 104 bacteria and bivalves 39-40,42-3, 125-6 Bering Sea, pollock in 178-9, 180,184 fisheries and management 182-3, 237,238-9 macroecology 187-8, 189,190,192 population dynamics 193-9 population structure genetic 215-18, 219,221-3 mechanisms 22g9, 23@1,232-3, 2345,236 metapopulation 226-7 phenotypic 208-11, 212,213, 214 prey and predators 186-7 recruitment 199, 200,201, 202, 203-4 stocks 185, 186 see also Aleutian Basin bioassay of bivalves advantages and limitations 4-5 applications 74-82 elutriate water 80-2 pore water 80 rationale for use of 76-82 sediment quality 77-80 toxicity tests with pure chemicals 74-6 water quality 77 broodstock availability and gamete production 135-6 methodology 40-50,133-7 broodstock 47-8 seawater quality 4 3 4 spawning 46-9 species 135 procedures 50-74 spermiotoxicity test 50-1
262 bioassay of bivalves (cont.) see also embryo-larval bioassay; physiology tests; tests under growth statistical evaluation 136 biocide toxicity (herbicides, pesticides and insecticides) and bivalves 63,66,95, 98-106,97-107,133 see also in particular TBT biological factors affecting bivalves see intrinsic bivalve embryos and larvae 1-175 biology 10-25 see also reproduction; species future research 137-8 see also bioassay; laboratory rearing; pollutants; statistical methods; toxicity tests “blooms” 125-7, 127 body size of bivalve larva shells 63 and toxicity reaction 118-19 see also age Bogoslof Island, pollock near 180,183, 184 population dynamics 193, 194, 195, 196,197 population structure 217,219,221, 233 bottom trawl surveys of pollock 193-6 British Columbian waters, pollock in 209-10,213,214,227 bromate toxicity and bivalves 96, 102 broodstocklspawners, bivalve availability and gamete production 134-5 and bioassay methodology 45-6 conditioning of 25-7 cadmium toxicity and bivalves 45,66-7 tests 87-8, 90-5, 116 calcium uptake of bivalve larvae 69-70, 94 cannibalism of pollock 185,202,203 catch-at-age data on pollock 194,238 chlorine and chlorophenol toxicity and bivalves 3,77, 96,116, 123, 132 bioassay procedures 70-2 in biocides 101, 102-3,97 chloroform toxicity and bivalves 96. 105
SUBJECT INDEX
chromium toxicity and bivalves 87-8, 90-5 clams see under taxonomic index cleavage of bivalves 22-6,36 climate and pollock 203-4 see also temperature cod, Atlantic: comparison with pollock 182,192,237,239 population structure 207,218,220, 221 cod, Pacific 217-18 conditioning of spawners 25-7 copepods 107,185,187 copper toxicity and bivalves 3,45,54, 132 tests 88, 90-5116, 119, 122, 123 Iarval physiology 66-7,68,70,72 crustaceans 97,131-2 currents, ocean, pollock and 190,227, 229,230-1 cyanide toxicity and bivalves 96, 97, 120,132 DDT toxicity and bivalves 97, 101 dehydro-abyetilamines 97,100, Delrad toxicity and bivalves 97, 100 detergents and oil toxicity to bivalves 3, 90,107, 10%15,115-16, I33 and bioassay procedures 63,64 and sediments 77,129 developmental stage see age dichloronaphthoquinone see Phygon dichlorvos toxicity and bivalves 97, 102 dieldrin toxicity and bivalves 97, 102 Dinoterbe toxicity and bivalves 97, 102 Diuron toxicity and bivalves 97, 104 D-larvae, bivalve, growth tests and
64-6
DNA analysis of pollock 200-7, 216 see also mtDNA Donut Hole, pollock in I80 fisheries 182,183 population structure 211,217,219 pupulation dynamics 194,195 dose-response curves and bivalves 82-4 Dowcide toxicity and bivalves 103, 105 drifter tracking and Dollock 2 3 4 5 drum, red andYblack’220
SUBJECT INDEX
dynamics of population of pollock 193-9 ecosystems of pollock 185-7 effluents see urban and industrial eggs bivalve normal development 19 s u e 32 pollock 188,228 see also fertilization; spawning elemental composition and pollock 206, 207 elutriate water and bioassay of bivalves 8&2 embryogenesis of bivalves 21-5 and metals 87-9 success assessed 134-5 and toxicity 89-4,97,9-106,136 embryo-larval bioassay of bivalves 5142 age of fertilized egg at beginning of test 51-2 embryo stages 19,21-2, 35 incubation temperature, duration and egg density 52-3 larva abnormality at end 554,574, 5942 mortality at end 54-5 stages 22,23,29, 35 volume of test vessels 53-4 embryos, bivalve see bivalve; embryo-larval; embryogenesis endrin toxicity and bivalves 97, 102 environment and bivalve bioassays 125-30 algal and bacterial toxins 125-6 receiving waters 128-9 urban and industrial effluents 127-8 see also extrinsic extinctions see under mortality extrinsic (environmental) factors affecting toxicity 119-23 salinity 123 temperature 122-3 see also environment fecundity of bivalves 28 feeding see nutrition
263 Fenitrothion toxicity and bivalves, 97, 105 fertilization bivalves 136 and bioassay methodology 4!) laboratory rearing 30-3 pollock 184 see also reproduction; spawning filtration of seawater 41-2 fisheries, pollock 178,182-3,191, 193, 196 management 23-0 time trends of catches 195 flounder, arrowtooth 198 food see nutrition Funka Bay, pollock in 180,182, 184 future bivalve research 137-8 gametes, bivalve ageing after spawning 32-3 gametogenesis 12-13, 16-9 production and bioassay methodology 135-6 genetics of pollock 181,208,215, 216, 217-23,237 gene flow 2065,224,23940 molecular see mtDNA mutation 204-5, 207 random drift 204-5 genotoxicity and bivalves 95-6 geographic basin model of pollock 225 geographical distribution bivalves 10-18 pollock see in particular Aleutian Basin; Bering Sea; British Columbia; Gulf of Alaska; Japan; Okhotsk gonads, bivalve, stripping of 47, 135-6 growth of bivalve larvae 94 tests 6 2 4 inhibition with already unexposed 646 inhibition with previously unexposed 63 and toxicity 97, 98-106 Gulf of Alaska, pollock in 178, 1'79, 184,192 fisheries and management 182--3, 237,238-9 population dynamics 1934,1!2.5,196, 198,199
264 Gulf of Alaska, pollock in (cont.) population structure genetic 215-18,219,221-3 mechanisms 227,228-9,23&1,236 phenotypic 208,213 prey and predators 186-7 recruitment 200-3,204 stocks 185 Guthion toxicity and bivalves 97, 105 haddock, Atlantic 208,220 halibut 198,220 halogen toxicity and bivalves 133 see also chlorine heavy metals see metal toxicity herbicides see biocide toxicity hermaphroditism 17, 18 herring, Pacific: comparison with pollock 190,217,218,223,239 hydroacoustic surveys of pollock 193, 197 incubation of bivalve eggs 52-3, 136 industry see urban and industrial insecticides see biocide toxicity interaction between toxicants 123-5 intrinsic (biological) factors affecting toxicity 118-19 age and developmental stage 118-19 body size 119 species 118 Japan, Sea ofiJapanese waters, pollock in 180,182,184 fisheries and management 183,191, 237 population structure 227,236 genetic 216, 218,221 phenotypic 209,212-13,215 stocks 185 juvenile pollock 188,189,192-3,197, 199,200-3,226 Kamchatka, pollock near population structure 208-9, 213,215, 221,227 stocks 185,186 laboratorv rearing of bivalve larvae 2520
SUBJECT INDEX
conditioning of spawners 25-7 fertilization 29-3 nutritional requirements 37-40 spawning 27-30 see also physical requirements larvae bivalve see bivalve; embryo-larval; growth pollock 184-5,188-9,201-3 migration 228-9,231-2 LAS (linear alkylate sulphonates) toxicity and bivalves 107, 10%10,115 lead toxicity and bivalves 77, 88-9, 904,116 lindane toxicity and bivalves 97, 102-3 linear alkylate sulphonates see LAS macroecology of pollock 187-93 Malathion toxicity and bivalves 97, 106 manganese toxicity and bivalves 88, 9&1,123 mechanisms of population structure of pollock 228-36 mercury toxicity and bivalves 3 sub-lethal and lethal effects 132 tests 88, 9 0 4 9 5 , 116,120, 123, 125 larval physiology 66,68, 69,70, 73-4 meristics and pollock 206,207,208-11 metal toxicity and bivalves 3,88-96, 904132,133 embryogenesis 87-8 genotoxicity 95-6 growth, larval 94 mortality, larval 94 respiration 95 settlement 94,95 tests 88-9, 90-5, 95, 116,120, 123-4 see also in particular cadmium; copper; lead; mercury; silver; zinc metamorphosis of bivalve embryos and larvae 23,24 physiology tests for 70-4 and toxicity 70-1, 95 metapopulation of pollock 224-8 methoxychlor toxicity and bivalves 103,107 microencapsulated diets for bivalve larvae 40-1
SUBJECT INDEX
microsatellite analysis of pollock 207, 216,222-3 migration of pollock 184 and population structure 206, 21415,224,2345 larvae 228-9,231-2 management implications 236, 237-40 vertical 190-1 mitochondria1 DNA analysis see mtDNA molecular genetics see mtDNA molybdenum toxicity and bivalves 89, 123 morphometrics of pollock 206,208, 209-11,212 mortality bivalve larvae 94 at end of bioassay 54-5 and toxicity 98-106 pollock 198,200,231,237 extinctions 224 mtDNA (mitochondrial DNA analysis) of pollock 2067,208,216,217, 219, 220,221-3,237 mussels see under taxonomic index mutation of pollock 2065,207 natal homing of pollock 234,236,237 natural seawater quality and bivalves 41-3 natural selection 204-5 Neburon toxicity and bivalves 97, 104 nickel toxicity and bivalves 87-8, 90-5, 116 nitrilotriacetate see NTA NOEC (no observed effect concentration)lANOVA approach 85 North Pacific see pollock NTA (nitrilotriacetate) toxicity 107,
112
nutrition bivalve larvae 20, 37-40, 6 6 7 pollock 185,187,202,231 oil toxicity and bivalves 113-5 see also detergents and oil Okhotsk, Sea of, pollock in 180, 188, 199,237
265 population structure 208-9,212-14, 216-17,226 stocks 185,186 see also Kamchatka organo-phosphate toxicity and bivalves 97,116 oxygen availability and respiration of bivalves 81 and laboratory rearing 34,37 and metals 95 physiology tests 67,68 oysters see under taxonomic index ozonation and bivalves 70, 71-2 Pacific see pollock parasites of pollock 206, 207 Parathion toxicity and bivalves 97, 105 PCP (pentachlorophenol) toxicity and bivalves 97, 101 PCR (polymerase chain reaction) 206, 222 pentachlorophenol see PCP “percent net risk” (PNR) 137 perch, Pacific 190,218 pesticides see biocide toxicity petroleum see detergents and oil pH and bivalves 37,63,81,120 phenotypic population structure of pollock 208, 209-11, 212-15, 237 philopathy of pollock 234,236 Phygon (dichloronaphthoquinone) toxicity and bivalves 97, 100 physical requirements of bivalve larvae reared in laboratory 33-7 oxygen 34,37 pH 37 salinity 34,36 temperature 36-6,36 turbidity 37 physiology tests for bivalve larvae 6 6 7 4 calcium uptake 69-70 feeding activity 6 6 7 metamorphosis 70-4 respiration 69 swimming activity 68-9 phytoplankton 38-9,70 PNR (“percent net risk”) 137 pollock see walleye pollock pollutants and bivalves 3 4 6 - 1 0 , 87-118
SUBJECT INDEX
pollutants and bivalves (cont.) biocides 96107, 98-106 detergents and oil 107-16, 109-15 see also bioassay; biocide; detergents and oil; metal; toxicity polymerase chain reaction (PCR) 206, 222 polymorphism and pollock 206,207, 216,222 polyspermy and bivalves 30, 32 population ecology of pollock 187-204 dynamics 193-9 macroecology 187-93 recruitment 199,200,201,202,2034, 238-9 population structure of pollock 204-36 estimation methods 206-8 mechanisms 228-36 metapopulation 224-8 phenotypic 208, 209-11,212-15,237 see also genetics potassium toxicity and bivalves 72, 106 predator-prey interactions of pollock 185-7,190,202,203 Prince William Sound, pollock in 179, 180,196,222,224 Puget Sound, pollock in 180,183,184, 1874,224,225 qualitative assessment of environmental samples 8 4 5 quality of seawater and bivalves 41-3, 44, 77,137 of sediment and bioassay of bivalves 77-80 receiving waters and bivalve bioassays 128-9 recruitment of pollock 199,200,201, 202,2034,238-9 “red tides” 125-6, 127 reproduction of bivalve embryos and larvae 1 6 2 5 cleavage, embryogenesis and larval development 21-5 embryonal stages 21-2 gametogenesis 1 6 1 9 larval stages 22-3 settlement and metamorphosis 23-5
sexual maturation in field 19-21 see also fertilization; gametes; spawning respiration see oxygen RFLP (restriction fragment length polymorphism) 206,216,222 Rosine Amine D toxicity and bivalves 97, 101 rotifers 103, 132 salinity and bivalves 66, 68, 81 laboratory rearing 36, 37 spawning 30 toxicity reaction 120,122, 123 and pollock 192 salmon 217,239 SDS see under sodium sea lions, Steller 198 sea urchins 7,130,131 seawater quality and bivalves 41-3, 44, 77,136 sediments and bivalve bioassays 3, 77-80,128-7,132,133 sensitivity of bivalves 131-2 see also bivalve embryos and larvae settlement of bivalves 23,25 and metals 88, 95 and toxicity 72, 95, 97-107 Sevin toxicity and bivalves 104, 107 sexual maturation of bivalve embryos and larvae 19-21 Shelikof Strait, pollock in 179,180, 183, 184,193 population structure 203,226 genetic 219,221-2 mechanisms 229,231,236 shell size of bivalves 63 Shumagin Island, pollock near 179,180, 224,238 silver toxicity and bivalves 3, 132 tests 87, 904116,123-5 snapper, red 220 sodium toxicity and bivalves 83, 100 SDS (sodium dodecyl sulphate) 109-10,115 sole, yellowfin 218 sonar see hydroacoustic spawner-recruit relationship of pollock 202,203,238-9
267
SUBJECT INDEX
spawning bivalves 136 ageing of gametes after 32-3 behaviour 20-1 and bioassay methodology 46-49 counting eggs 49 impurities eliminated 48 induction of 29, 30, 46-7 laboratory rearing 27-30 number of parents 47-8 stripping gonad 47,135-6 see also broodstock; eggs; fertilization; reproduction pollock 1 8 0 , 1 8 3 4 , 2 0 3 4 species of bivalves 10-16 Iisted 8-10 type and toxicity 118 see also taxonomic index spermiotoxicity test and bivalves 50-1 statistical methods and bivalves 84-6 dose-response curves 8 2 4 minimum sample size 85-6 qualitative assessment of environmental samples 84-5 stocks of pollock 179,181, 185,186, 192 “stripping”, gonad 47, 135-6 sub-populations of pollock 205,238 surfactants see detergents and oil surimi 182,183 swimming activity of bivalves 68,69 tagging pollock 206,214-15,240 TBT (tributyl-tin) toxicity and bivalves 3, 4, 13, 116, 132,133 salinity and 123 tests 96, 98-100 larval physiology 66,70,72-3 TCC (trichlorocarbanilide) toxicity and bivalves 97, 100 TCP (tetrachlorophenol) toxicity and bivalves 97, 105 temperature and bivalves incubation 52-3
laboratory rearing 3 3 4 3 6 spawning 13-16,20,29,31,46 toxicity reactions 118-19, 120, 122 and pollock 190,192,2034 tetrachlorophenol see TCP toxaphene toxicity and bivalves 97, 103 toxicity tests and bivalves 88-125 assessment 1 3 2 4 exposure and observation of response 136 factors affecting see extrinsic; intrinsic interaction between toxicants 123-5 with pure chemicals 74-6 spermiotoxicity 50-1 toxicity defined 123-4 see also bioassay; pollutants tributyl-tin see TBT trichlorocarbanilide see TCC turbidity and turbulence and bivalves 37.80 urban and industrial effluents 127-9, 132,133 see also biocide; detergents and oil; metal; toxicity VPA (virtual population analysis) 194 walleye pollock 177-253 background 182-7 life history 183-5 predator-prey interactions and ecosystem 185-7,190,202,2.03 see also fisheries; population eco Logy; population structure wastes see urban and industrial year class abundance of pollock 199-200,201 zinc toxicity and bivalves 45 tests 88-9, 90-5,95, 116, 1 2 3 4 larval physiology 6 6 7 , 68, 70
This Page Intentionally Left Blank
Cumulative Index of Titles Note: TCtles of papers have been converted into subjects and a specific article may therefore appear more than once Abyssal and hadal zones, zoogeography, 32, 325 Abyssal macrobenthos, trophic structure, 32, 427 Abyssal organisms, population genetics, 35, 1 Acetabularia, marine alga, recent advances in research, 14, 123 Algal-invertebrate interactions, 3, 1 Antarctic benthos, 10, 1 Antarctic fishes, comparative physiology, 24, 321 Ascidians biology, 9, 1 physiology, 12, 2 Assessment of marine pollution - bioassays with bivalve embryos and larvae, 31, 177 Atlantic, Northeast, meiobenthos, 30,1 Baltic Sea, autrophic and heterotrophic picoplankton, 29, 73 Barnacles, growth, 22, 199 Bathyal organisms, population genetics, 35, 1 Bathyal zone, biogeography, 32, 389 Benthic marine infaunal studies, development and application of analytical methods, 26, 169 Benthos abyssal macrobenthos, trophic structure, 32, 427 Antarctic, 10, 1 Northeast Atlantic meiobenthos, 30, 1 sampling methods, 2, 171 Biogeography, hydrothermal vent fauna, 34,353 Blood groups, marine animals, 2, 85 Blue whiting, North Atlantic, population biology, 19, 257 Brachiopods, living, biology, 28, 175 Bryozoans, marine, physiology and ecology, 14, 285 Bullia digitalis, 25, 179 Calanoid copepods, biology of, 33 Cephalopods flotation mechanisms in modern and fossil, 11, 197 recent studies on spawning, embryonic development, and hatching, 25, 85 Chaetognaths, biology, 6,271
270
CUMULATIVE INDEX OF TITLES
Cladocerans, marine, reproductive biology, 31, 80 Climatic changes, biological response in the sea, 14, 1 Clupeid fish behaviour and physiology, 1, 262 biology, 20, 1 parasites, 24, 263 Copepods association with marine invertebrates, 16, 1 calanoid, biology of, 33 respiration and feeding, 11, 57 Coral reefs adaptations to physical environmental stress, 31, 222 assessing effects of stress, 22, 1 biology, 1, 209 communities, modification relative to past and present prospective Central American seaways, 19, 91 ecology and taxonomy of Halimeda: primary producer of coral reefs, 17, 1 Crustaceans, spermatophores and sperm transfer, 29, 129 Ctenophores, nutritional ecology, 15, 249 Die1 vertical migrations of marine fishes: an obligate or facultative process?, 26, 115 Donax serra, 25, 179 Echinoids, photosensitivity, 13, 1 Eels, North Atlantic freshwater, breeding, 1, 137 Effluents, effects on marine and estuarine organisms, 3, 63 Environmental simulation experiments upon marine and estuarine animals, 19, 133 Euphausiidas, biology, 7, 1, 18, 373 Fish alimentary canal and digestion in teleosts, 13, 109 Antarctic fish, comparative physiology, 24, 321 artificial propagation of marine fish, 2, 1 clupeid behaviour and physiology, 1, 262 clupeid biology, 10, 1 clupeid parasites, 24, 263 die1 vertical migrations, obligate or facultative process?, 26, 115 diseases, 4, 1 egg quality, 26,71 gustatory system, 13, 53 migration, physiological mechanisms, W, 241 North Atlantic freshwater eels, 1, 137 nutrition, 10, 383 parasites in deep-sea environment, 11, 121 photoreception and vision, 1, 171 predation on eggs and larvae of marine fishes and the recruitment problem, production and upwelling, 9, 255 year class strength, and plankton production, update of matchmismatch
CUMULATIVE INDEX OF TITLES
271
hypothesis, 26, 249 Fish farming, estuarine, 8, 119 Fish larvae appraisal of condition measures for marine fish larvae, 31, 217 field investigations of the early life stages of marine fish, 28, 1 turbulence and feeding ecology, role of microscale, 31, 170 Fish migration, physiological mechanisms in the migration of marine and amphihaline fish, 13, 248 Fisheries, lagoons, constraints on, 34, 73 Fisheries, management of resources, 6, 1 Fisheries and seabird communities, competition, 20, 225 Fishing effects, on marine ecosystems, 34, 201 Frontal systems, aspects of biology, 23, 163 Gastropods comparison between Donax serra and Bullia digitalis, bivalve molluscs, 25, 179 intertidal, ecology, 16, 111 marine, burrowing habit, 28, 389 Gonatid squids, subarctic North Pacific, ecology, biogeography, niche diversity and role in ecosystem, 32, 243 Gonionemus, erratic distribution: the occurrence in relation to oyster distribution, 14, 251 Growth performance, and mortality, aquatic macrobenthic invertebrates, 35, 153 Habitat selection by aquatic invertebrates, 10, 271 Halibut Hippoglossus hippoglossus, biology, 26, 1 Halimeda, 17, 1 Hemng Clupea harengus L. and other clupeids behaviour and physiology, 1, 262 biology, 20, 1 relationships with its parasites, 24, 263 Human affairs, marine biology, 15, 233 Hybridization in the sea, 31, 2 Hydrothermal vent communities, deep sea, ecology, 23, 301 Hydrothermal vent fauna, mid-Atlantic ridge, ecology and biogeography, 32, 93 Hydrothermal vents, biogeography of, 34,353 Indo-West Pacific region, mangrove swamps and forests, fauna and flora, 6, 74 Isopoda, oniscid, biology of the genus Tylos, 30, 89 Japan, scallop industry, 20, 309 Japanese oyster culture industry, recent developments, 21, 1 Lagoon fisheries, constraints on, 34, 73 Learning by marine invertebrates, 3, 1 Mangrove swamps and forests of Indo-West Pacific region, general account of fauna and flora, 6, 74 Marine animals
272
CUMULATIVE INDEX OF TITLES
blood groups, 2, 85 neoplasia, 12, 151 Marine ecosystems, effects of fishing on, 34, 201 Marine toxins venomous and poisonous animals, 3,256 venomous and poisonous plants and animals, 21, 59 Meiobenthos of the Deep Northeast Atlantic, 30,1 Mesoplankton distribution patterns, 32, 9 and macroplankton, some problems of vertical distribution in ocean, 32, 1 Metabolic energy balance in marine invertebrates, influence of temperature on maintenance, 17,329 Microbiology, marine, present status of some aspects, 2, 133 Mo11uscs hosts for symbioses, 5, 1 marine, mucus in, 34,1 wood-boring teredinids, biology, 9, 336 Molluscs, bivalve effects of environmental stress, 22, 101 and gastropods: comparison between Donax serra and Bullia digitalis, 25, 179 rearing, 1, 1 scatological studies, 8, 307 Mortality, and growth performance, aquatic macrobenthic invertebrates, 35, 153 Molluscs, marine, mucus in, 34,1 Mucus, in marine molluscs, 34,1 15N, natural variations in the marine environment, 24, 389 Nazca submarine ridge, composition and distribution of fauna, 32, 145 Nitrogen cycle, and phosphorus cycle, plankton, 9, 102
Oil pollution of the seas, problems, 8, 215 Oniscid isopods, biology of the genus Tylos, 30, 89 Oysters, living, speciation, 13, 357 Particulate and organic matter in sea water, 8, 1 Pelagic invertebrates, social aggregation, 30, 155 Penaeida, biology, 27, 1 Petroleum hydrocarbons and related compounds, 15, 289 Phoronida, biology, 19, 1 Phosphorus cycle, and nitrogen cycle, plankton, 9, 102 Pigments of marine invertebrates, 16, 309 Plankton distribution, vertical, in the ocean, 32, 4 laboratory culture of marine holozooplankton and its contribution to studies of planktonic food webs, 16, 211 and nitrogen and phosphorus cycles of sea, 9, 102 oceanic phytoplankton, outline of geographical distribution, 32, 527 parasitology of marine zooplankton, 25, 112 phytoplankton blooms, harmful or exceptional, 31, 302 phytoplankton, circadian periodicities in natural populations, 12, 326 picoplankton, Baltic Sea, 29, 73
CUMULATIVE INDEX OF TITLES
273
production, and year class strength in fish populations: update of the matchmismatch hypothesis, 26,249 Pollution studies with marine plankton Part 1: Petroleum hydrocarbons and related compounds, 15, 289 Part 2: Heavy metals, 15, 381 Population genetics, bathyal and abyssal organisms, 35, 1 Population structure and dynamics of Walleye Pollock, Theragra chalcogramma,
37,1
Pseudocalanus, biology, 15, 1 Pycnogonida, biology, 24, 1 Sala y G6mez submarine ridge, composition and distribution of fauna, 32, 145 Salmon, acclimatization experiments, Southern Hemisphere, 17,398 Scallop industry in Japan, 20, 309 Sea anemones, nutrition, 22, 65 Seabird communities, and fisheries, competition, 20,225 Seamount, biology, 30,305 Seaweeds of economic importance, aspects of biology, 3, 105 population and community ecology, 23, 1 Shrimps, pelagic, biology of, 12, 233 Siphonophore biology, 24, 97 Sole (Solea solea), Bristol Channel, 29, 215 Spermatophores and sperm transfer in marine crustaceans, 29, 129 Squids gonatid squids, subarctic North Pacific, ecology, biogeography, niche diversity and role in ecosystem, 32, 243 oceanic, review of systematics and ecology, 4, 93 Taurine in marine invertebrates, 9, 205 Teredinid molluscs, wood-boring, biology of, 9, 336 Tropical marine environment, aspects of stress, 10,217 Turbulence feeding ecology of larval fish, role of microscale, 31, 170 phytoplankton cell size, and structure of pelagic food webs, 29, 1 Tylos biology, 30, 89
This Page Intentionally Left Blank
Cumulative Index of Authors
Ahmed, J., 13, 357 Akberali, H. B., 22, 102 Allen, J. A., 9, 205 Ansell, A. D., 28, 175 Arakawa, K. Y., 8, 307 Arnaud, F., 24, 1 Bailey, K. M., 25, 1, 37, 177 Bailey, R. S., 19, 257 Balakrishnan Nair, M., 9, 336 Bamber, R. N., 24,1 Beiras, R., 37, 1 Bentzen, P., 37,177 Bett, B. J., 30, 1 Blaxter, J. H. S., 1, 262, 20, 1 Boletzky, S. V., 25, 85 Boney, A. D., 3, 105 Bonotto, S., 14, 123 Bourget, E., 22, 200 Branch, G. M., 17,329 Brey, T., 35, 153 Brinkhurst, R. O., 26, 169 Brown, A. C., 25, 179, 28, 389, 30, 89 Brown, B. E., 22, 1, 31, 222 Bruun, A. F., 1, 137 Burd, B. J., 26, 169 Campbell, J. I., 10, 271 Carroz, J. E., 6, 1 Chapman, A. R. O., 23, 1 Cheng, T. C., 5, 1 Clarke, M. R., 4, 93 Collins, M. J., 28, 175 Corkett, C. J., 15, 1 Comer, E. D. S., 9, 102, 15, 289 Cowey, C. B., 10, 383 Creasey, S. S., 35, 1 Crisp, D. J., 22, 200 Curry, G. B., 28, 175
Cushing, D. H., 9, 255, 14, 1, 26, 2.49 Cushing, J. E., 2, 85 Dall, W., 27, 1 Davenport, J., 19, 133 Davies, H. C., 1, 1 Davies, M. S., 34,1 Davis, A. G., 9, 102, 15, 381 Dell, R. K., 10, 1 Denton, E. J., 11, 197 Dickson, R. R., 14, 1 Dinet, A., 30,1 Dower, J. F., 31, 170 Edwards, C., 14, 251 Egloff, D. A., 31, 80 Emig, C. C., 19, 1 Evans, H. E., 13, 53 Ferrero, T., 30, 1 Ferron, A., 30, 217 Fisher, L. R., 7, 1 Fofonoff, I? W., 31, 80 Fontaine, M., 13, 241 Furness, R. W., 20, 225 Galkin, S. V., 32, 93 Gardner, J. I? A., 31, 2 Garrett, M. R., 9, 205 Gebruk, A. V., 32, 93 Ghirardelli, E., 6, 271 Gilpin-Brown, J. B., 11, 197 Glynn, l? W., 19, 91 Gooday, A. J., 30, 1 Goodbody, I., 12, 2 Gotto, R. V., 16, 1 Grant, W. S., 37, 177 Grassle, J. F., 23, 301 Gulland, J. A., 6, 1
276 Harris, R. P., 16, 211 Haug, T., 26, 1 Hawkins, S. J., 34, 1 Heath, M. R., 28, 1 Hickling, C. F., 8, 119 Hill, B. J., 27, 1 Hillis-Colinvaux, L., 17, 3 His, E., 37, 1 Holliday, F. G. T., 1, 262 Holme, N.A., 2, 171 Holmefjord, I., 26, 71 Horwood, J., 29, 215 Houde, E. D., 25, 1 Howard, L. S., 22, 1 Hunter, J. R., 20, 1 James, M. A., 28, 175 Jennings, S., 34, 201 Joyeux, J.-C., 34, 73 Kaiser, M. J., 34, 201 Kapoor, B. G., 13, 53,l3, 109 Kennedy, G. Y., 16,309 Kiarboe, T., 29, 1 KjZrsvik, E., 26, 71 Kuosa, H., 29, 73 Kuparinen, J., 29, 73
CUMULATIVE INDEX OF AUTHORS
Moore, H. B., 10, 217 Moskalev, L. I., 32, 93 Naylor, E., 3, 63 Neilson, J. D., 26, 115 Nelson-Smith, A,, 8, 215 Nemec, A., 26, 169 Nesis, K. N., 32, 144, 32, 243 Newell, R. C., 17, 329 Nicol, J. A. C., 1, 171 Noble, E. R., 11, 121 Odendaal, F. J., 30, 89 Omori, M., 12, 233 OnbC, T., 31, 80 Owens, N. J. F!, 24,389 Paffenhofer, G. A., 16, 211 Parin, N. V., 32, 144 Peck, L. S., 28, 175 Perry, R. I., 26, 115 Pevzner, R. A., 13, 53 Pfannkuche, O., 30, 1 Pugh, P. R., 24, 97 Purcell, J. E., 24, 97 Quinn 11, T. J., 37, 177
Lambshead, P. J. D., 30, 1 Le Fkvre, J., 23, 163 Leggett, W. C., 30, 217, 31, 170 hosanoff, v. L., 1, 1 Lurquin, P., 14, 123 McArthur, A. G., 34, 353 Macdonald, J. A., 24, 321 McHugh, D., 34,353 Mackenzie, K., 24, 263 Mackie, G. O., 24, 97 McLaren, I. A., 15, 1 Macnae, W., 6, 74 Mangor-Jensen, A., 26, 71 Marshall, S. M., 11, 57 Mauchline, J., 7, 1, 18, 1, 33, 1-660 Mawdesley-Thomas, L. E., 12, 151 Mazza, A,, 14, 123 Meadows, F! S., 10, 271 Millar, R. H., 9, 1 Miller, T. J., 31, 170 Millot, N., 13, 1 Mironov, A. N., 32, 144 Montgomery, J. C., 24, 321
Reeve, M. R., 15, 249 Rhodes, M. C., 28, 175 Richardson, K., 31, 302 Riley, G. A., 8, 1 Ritz, D. A., 30, 155 Rogers, A. D., 30, 305, 35, 1 Rothlisberg, P. C., 27, 1 Russell, F. E., 3, 256, 21, 60 Russell, F. S., 15, 233 Ryland, J. S., 14, 285 Saraswathy, M., 9, 336 Sargent, J. R., 10, 383 Scholes, R. B., 2, 133 Seaman, M. N. L., 37, 1 Semina, H. J., 32, 527 Shelbourne, J. E., 2, 1 Shewan, J. M., 2,133 Sindermann, C. J., 4, 1 Smit, H., W, 109 Sokolova, M. N., 32, 427 Soltwedel, T., 30, 1 Sournia, A., 12, 326
CUMULATIVE INDEX
277
OF AUTHORS
Southward, A. J., 32, 93 Staples, D. J., 27, 1 Stenton-Dozey, J. M. E., 25, 179 Stewart, L., 17, 397 Subramoniam, T., 29, 129 Taylor, D. L., 11, 1 ThCodoridks, J., 25, 117 Trueman, E. R., 22, 102, 25, 179, 28, 389 Tunnicliffe, V., 34,353
Vanreusel, A., 30, 1 Ventilla, R. F., 20, 309, 21, 1 Vereshchaka, A. L., 32, 93 Verighina, I. A., 13, 109 Vincx, M., 30, 1 Vinogradov, M. E., 32, 1 Vinogradova, N. G., 32, 325 Walters, M. A., 15, 249 Ward, A. B., 34,73 Wells, M. J., 3, 1 Wells, R. M. G., 24, 321
Underwood, A. J., 16,111
Yonge, C. M., 1, 209
Van-Praet, M., 22, 66
Zezina, 0. N., 32, 389
This Page Intentionally Left Blank
