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[email protected] doi:10.1016/S0304-3894(11)01319-7
Journal of Hazardous Materials 196 (2011) 1–15
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Review
Effects of the presence of sulfonamides in the environment and their influence on human health b ´ Wojciech Baran a , Ewa Adamek a,∗ , Justyna Ziemianska , Andrzej Sobczak a,b a b
Silesian Medical University, Department of General and Analytical Chemistry, Jagiello´ nska 4, 41-200 Sosnowiec, Poland Institute of Occupational Medicine and Environmental Health, Ko´scielna 13, 41-200 Sosnowiec, Poland
a r t i c l e
i n f o
Article history: Received 10 March 2011 Received in revised form 22 July 2011 Accepted 31 August 2011 Available online 6 September 2011 Keywords: Sulfonamides Biotransformation Ecotoxicity Environmental risk Drug resistance
a b s t r a c t World production and consumption of pharmaceuticals has been steadily increasing. Anti-infectives have been particularly important in modern therapy of microbial infection. Sulfonamides have been widely used for a long time as anti-infectives and are still widely prescribed today. This review presents the most common types of sulfonamides used in healthcare and veterinary medicine and discusses the problems connected with their presence in the biosphere. Based on the analysis of over 160 papers, it was found that small amounts of sulfonamides present in the environment were mainly derived from agricultural activities. These drugs have caused changes in the population of microbes that could be potentially hazardous to human health. This human health hazard could have a global range, and administrative activities have been ineffective in risk reduction. © 2011 Elsevier B.V. All rights reserved.
Contents 1. 2. 3. 4. 5. 6. 7. 8.
9. 10. 11. 12.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physicochemical properties of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mechanism of antibacterial activity of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimated usage of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Occurrence of SNs in the environment and food . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecotoxicity of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Degradation of SNs in organisms and in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.1. Metabolism of SNs in mammals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2. Biodegradability of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3. Physicochemical methods of degradation of SNs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal of SNs from wastewater . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The environmental risk assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Generation of drug resistance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Introduction World production and consumption of pharmaceuticals has been steadily increasing at a rate higher than the rate of global
∗ Corresponding author. Tel.: +48 032 364 15 62. E-mail address:
[email protected] (E. Adamek). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.082
1 2 2 3 4 5 6 7 7 7 8 8 8 9 12 12 12
population growth. After use, large amounts of drugs have been discharged into the environment in the form of human and animal excretions and unused waste [1]. The persistence of pharmaceuticals in the environment, the rate of their spreading and their ability to accumulate in the biosphere has differed. However, their high biological activity indicates that these drugs, even in trace amounts, could cause significant changes in the biosphere. An example of such changes in the last decade of the 20th century is
2
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
O 4
H2N
S
N1 R
O
H
Fig. 2. Chemical structure of SNs with bacteriostatic properties.
• excessive amounts of SNs are introduced to the biosphere (a common practice is illegal and/or without control administration of SNs to healthy farm animals), • a local concentration of SNs in the environment and risk associated with this issue are very high, • SNs can remain in the environment for a long time, • SNs present in the environment are active in the generation of drug resistance in bacterial cells (including cross resistance to drugs). 2. Physicochemical properties of SNs Fig. 1. The possible fates of SNs residues and resistance genes (SNsR) in the environment.
the phenomenon of feminization of fish by sex hormones caused by anthropogenic pollution of European rivers [2]. For these reasons, pharmaceuticals have been classified as particularly dangerous pollutants for the environment. As a result, research and multinational projects (e.g., REMPHARMAWATER [3], POSEJDON [4], KNAPPE [5], ERAPHARM [6], and ECO-SENS [7]) have been carried out to find answer to the following questions: • Which pharmaceuticals have the greatest environmental risk? • How can we effectively control the amounts and effects of drugs on the environment? • How can we successfully reduce their release into the environment? Antibiotics are a group of pharmaceuticals with effects on the environment that could be particularly harmful to human health. Unfortunately, their frequency in environmental samples is very high [1,5,8–18]. Historically, sulfonamides (SNs) have been used as synthetic antibiotics the longest. Recently the large quantities of SNs are used in animal husbandry in particular as veterinary medicines. Based on these drugs, we can obtain a reliable estimation of the effects and consequences of prolonged use of anti-infectives on people’s health and on the environment. A report for State Office for Nature, Environment and Consumer Protection of North Rhine-Westphalia (Germany) published in 2007 has been presented the literature review on effects of the introduction of SNs to the environment [1]. In the majority of published articles, the authors assessed the risk caused by SNs almost exclusively on the basis of their use, toxicity, and removal efficiency from the environment. The data presented in this context led to the conclusion that the presence of drugs in the environment was a negligible problem regarding quality of life. However, in the majority of articles, the effect of antiinfectives in the generation of drug resistance in microbes was not considered. The effect of antibiotics occurring in the environment to the generation and prevalence of drug-resistant microorganisms is essential from the human health point of view (Fig. 1.). This influence has been much more widespread in the last decades due to the globalization process.The aim of the work is to show that:
Since the early 1940s, over 150 SNs, sulfanilamide derivatives, have been applied in human and veterinary medicine as antibacterial drugs [19]. The formula of structure presented in Fig. 2 corresponds to the synthetic antimicrobial agents that contain the sulfonamide group. Such a molecule should have a free amino group (–N4 H2 ) at one end. SNs are a group of synthetic bacteriostatic drugs classified by the Anatomical Therapeutic Chemical (ATC) classification index as a group of antibacterial drugs for systemic use (the subgroup J01E) [20]. Many SN derivatives have also been used as antiprotozoal agents [21] and herbicides [19], and complexes of SNs with Ag+ and Zn2+ have been used as antifungals [22]. Moreover, SNs have been the most commonly used components of more composite drugs with trimethoprim (TMP). The characteristics of commonly used SNs are presented in Table 1. SNs are polar molecules with amphoteric properties. Their amino nitrogen (N4 ) is protonated at pH 2–3, while the amide nitrogen (N1 ) is deprotonated at pH 4.5–11 [10,23]. The SNs presented in this text are small molecules (molar mass 177–300 g mol−1 ), are water soluble (with the exception of SGM and sulfasalazine) and have low Henry’s constant (1.3 × 10−12 –1.8 × 10−8 ) values [9,10,24]. They are slightly sorbed by soil (the soil partition coefficient values are 0.6–7.4 l kg−1 ) [9]. Thus, these SNs are easily and quickly spread in the environment, but their properties should limit their accumulation in defined biotopes. SNs do not easily adsorb onto activated carbon [1,4]. They are classified as photo- and thermally stable substances at the degradation half-life (DT50 ) >1 year [24]. They can undergo alkaline hydrolysis and coupling reactions with phenols and amines and easily react with the hydroxyl radical HO• [10,11,25]. 3. Mechanism of antibacterial activity of SNs As shown in Fig. 3, antibacterial SNs act as competitive inhibitors of the enzyme dihydropteroate synthease (DHPS) which catalyses the conversion of para-aminobenzoate (PABA) to dihydropteroate (AHHMD), a precursor of folate synthesis. Tetrahydrofolic acid (THF) participates in the synthesis of nucleic acids that are essentials as building blocks of DNA and RNA.A mechanism of action of herbicidal SNs is similar.As a result, it is possible to inhibit the synthesis of nucleic acids and thus proteins [18,27]. SNs also inhibit the permeability of the bacterial cell wall for glutamic acid, which is also an essential component in folic acid synthesis. However, SNs do not inhibit the growth of microorganisms that:
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
3
Table 1 Common names, CAS number and structure of selected SNs. Common name of SNs
CAS number
ATC classification index
Abbreviation
–R
Sulfanilamide Sulfacetamide Sulfacarbamide Asulam (herbicide) Carbutamide Sulfathiourea Sulfaguanidine
63-74-1 144-80-9 547-44-4 3337-71-1 339-43-5 515-49-1 57-67-0
J01EB06, D06BA05, QJ01EQ06 S01AB04 J01EC20 (with SDZ and SDM) – A10BB06 J01EB08 A07AB03
SAD SCT SC
–H –COCH3 –CONH2 –COOCH3 –CONH(CH2 )3 CH3 –CSNH2 C(NH2 )2
Sulfathiazole
72-14-0
D06BA02, J01EB07, QJ01EQ07
STZ
STU SGM
S N
H3C Sulfafurazole, Sulfisoxazole
127-69-5
J01EB05, S01AB02, QJ01EQ05
CH3
SSZ
N
O N Sulfamethoxazole
723-46-6
J01EC01, QJ01EQ11
O
SMX
CH3
Sulfamoxole
Sulfapyridine
729-99-7
144-83-2
J01EC03
SMM
J01EB04, QJ01EQ04
SPY
O
CH3
N
CH3
N N
Sulfadiazine
68-35-9
J01EC02, QJ01EQ10
SDZ
N
N Sulfamethoxine, Sulfamethoxydiazine
651-06-9
J01ED04
OCH 3
SMO
N
N Sulfamerazine
127-79-7
J01ED07
SMR
N CH3 CH 3
N Sulfamethazine, Sulfadimidine
57-68-1
J01EB03, QJ01EQ03, QP51AG01
SDM
N CH 3 OCH3
Sulfadimethoxine
122-11-2
J01ED02, QJ01EQ09, QP51AG02
SDT
N N OCH3 N N
Sulfamethoxypyridazine
80-35-3
Sulfachloropyridazine
80-32-0
J01ED05, QJ01EQ15
OCH 3
SMP
N N SCP
Cl
N N Sulfadoxine
2447-57-6
QJ01EQ13
SDX
H3CO
• need the presence of folic acid in the environment, • possess a high concentration of PABA, or • have modified metabolic pathways (drug resistance). 4. Use of SNs SNs are active against a broad spectrum of Gram-positive and many Gram-negative bacteria including species of the genus
OCH3
Streptococcus, Staphylococcus, Escherichia, Neisseria, Shigella, Salmonella, Nocardia, Chlamydia and Clostridium. Moreover, SNs have used against protozoa (e.g., Toxoplasma gondii), parasites (e.g., Plasmodium malariae), and fungi (e.g., Pneumocystis carinii). SMX, SCT or sulfasalazine belong to SNs commonly used in human medicine while SDM, SDT, SMR, SDZ, STZ are used most frequently in veterinary medicine (different SNs have been used in different countries). Moreover, SNs have been added to animal feed
4
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
N
H2N N 2-Amino-4-hydroxy-6-hydroxymethyl7,8 dihydropteridine diphosphate
H N
H H
N
CH2
H2N
O
H N
N N
P ~P
N
ATP AMP
OH
H H CH2
OH
OH (2-Amino-4-hydroxy6-hydroxymethyl-7,8 dihydropteridine
(AHHMD)
(AHHMP) dihydropteroate synthetase (DHPS) EC 2.5.1.15 - P-P and H2O
SO2NHR
H 2N
Sulfonamide
COOH
H2N
4-Aminobenzoic acid (PABA)
X
H N
N
H2N N
H H
Dihydropteroic acid COOH
CH2 HN
N OH
CH2
CH2
COOH
H2N CH
Glutamic acid
ATP
COOH H N
N
H2N N
N
H H
O
CH2
CH2 HN
C HN
CH
CH2
COOH [O] [H]
OH
COOH
Dihydrofolic acid (DHF) dihydrofolate reductase (DHFR) EC 1.5.1.3
Folic acid N-[p-[[(2-Amino-4-hydroxy-6-pteridinyl) methyl]amino]benzoyl]-L-glutamic acid
[O]
[H]
N
H2N N
OH
H N N H
H H H CH2 HN
C
CH2
CH2
O
HN CH
COOH
Tetrahydrofolic acid (THF)
COOH
Fig. 3. The schema of SNs pathways, based on Wilson & Gisvold’s Textbook of Organic Medicinal and Pharmaceutical Chemistry [26].
premix used in young animals feeding. For example in Denmark in 2009 the consumption of SNs with TMP per kg of meat produced was as follows [28]:
However, the use of Asulam could lead to the contamination of honey with SNs residues [29]. In 2008, it has been withdrawn from use in the EU countries.
• • • •
5. Estimated usage of SNs
pigs 4.82 mg, cattle 17.2 mg, broilers 0.033 mg, farmed fish (aquaculture) 58.5 mg.
Moreover, SNs can be used in commercial beekeeping (they protect honey bees against bacterial diseases e.g., American foulbrood). In agriculture, sulfonamide Asulam has been widely used as a herbicide. It is effective against dicotyledonous weeds e.g., barnyard grass (Echinochloa crus-galli), velvet grass (Holcus lanatus), wild oat (Avena fatua) and broadleaf dock (Rumex obtusifolius).
An accurate assessment of the global consumption of all drugs would be difficult, if not impossible. The authors of the KNAPPE project have estimated that the global consumption of pharmaceuticals used in human and veterinary medicine has reached 100,000 tonnes per year [5]. Based on information from the Union of Concerned Scientists, Sarmach et al. indicated that, at the beginning of the 21st century, Americans consumed 16,000 tonnes of antibiotics per year [9]. SNs used in veterinary medicine accounted for approximately 2.3% of the total amount of antibiotics. In European
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
18 16
active compounds (t)
14 12 10 8 6 4 2 0 1990 1992 1994 1996 1998 2000 2002 2004 2006 2008 2010
year Fig. 4. The dynamics of consumption of SNs and TMP in Denmark in the years 1990–2009 [28,31].
countries, this value ranged from 11 to 23% [9]. According to other authors, the worldwide consumption of antibiotics (anti-infective drugs) ranged from 100,000 to 200,000 tonnes per year, including 50–75% that were used in veterinary medicine and animal husbandry [1,24]. It has been possible that each year more than 20,000 tonnes of SNs, with bacteriostatic properties, have been introduced into the biosphere (not counting drugs introduced as herbicides). Since the end of 20th century, Scandinavia and other countries in Europe and North America have imposed restrictions on the use of antibiotics (including SNs) in animal husbandry. The use of antibiotics as growth promoters in animal husbandry in the European Union has been banned since January 1, 2006 [30]. However, reports on the consumption of pharmaceuticals in different countries have not shown a reduction in the use of these drugs. Fig. 4 presents the dynamics of SN with TMP consumption in Denmark in the years 1990–2009 [28,31]. A decrease in the use of SNs in animal husbandry occurred in the mid-1990s and is associated with the introduction of administrative restrictions related to the application of these drugs in animal feed. Although the ban has been still in place, the use of SNs in agriculture is similar as in 1994. In our opinion, the plots in Fig. 4 illustrate global trends in consumption of SNs in livestock farming and medicine. 6. Occurrence of SNs in the environment and food The first publication containing quantitative data about the presence of SNs in river water was published in 1982 [9]. However, systematic studies on the quantitative determinations of SNs in environmental matrices became possible after the development of highly sensitive analytical methods. According to data from the U.S. Environmental Protection Agency the limit of detection during routine analytical procedures using SPE/HPLC-MS/MS techniques for the selected SNs was below 10−9 g l−1 (e.g., for SDT, the limit of detection was 1 × 10−10 g l−1 ). A detailed statement of the analytical techniques and limits of detection of drugs (including SNs) in environmental samples has been discussed by García-Galán et al. [18] and Seifrtová et al. [32]. At the described level of detection, SNs were detected in 27% of rivers and streams in the USA [11], in almost all surface waters in France and Tajwan [33,34], and in 100% of wastewater samples [13,35,36]. According to Vulliet and CrenOlivé, in the Rhônee Alpes region of Frances the frequencies of SMX in surface and groundwater were 37 and 66%, respectively [37]. In commercially available, Italian natural mineral water the frequency
5
of SNs was 50% (in 4 of the 8 investigated samples) [38]. GarcíaGalán et al. [36] described in detail the frequency of occurrence of 19 selected SNs in wastewater. Moreover, metabolites of SNs, mainly N4 -acetyl sulfonamides (N4 -AcSNs), were also identified in environmental samples [11,39]. SNs concentrations in the environment underwent significant fluctuations, which were mainly dependent on the type of matrix and the type of SN [36]. Additionally, the results obtained may have depended on the sampling site, the day of the week [40] and even the time of day [41]. However, it was important to note that the data concerning the determination of SNs in environmental samples could contain significant errors. The cause of this may be imperfection of the analytical procedure used and the incorrect (incomplete) extraction of samples. For example, the recovery of SNs from soil samples ranged from 5 to nearly 294% but the authors have found that the “presented method is characterized by good selectivity” [42]. The recovery efficiency depends on various parameters including extraction/purification strategies [43] and the type of matrix [44]. A summary of the occurrence of SNs, depending on the matrix, is shown in Fig. 5, and the maximal values are given in Table 2. The presented data are based on the maximum values described in the literature. SNs concentrations in samples increased as follows: seawater < ground water < surface water < treated sewage < untreated (raw) municipal sewage < hospital sewage < activated sludge < soil < runoff from farmland < leachates from landfill < manure. Due to the low concentrations and low abundance of SNs, the presence of trace amounts of these drugs in drinking water was not considered a significant problem. The maximum concentrations were found in freshly removed bedding [58] and manure from pigs fed diets that contained SNs, mainly SDM [59]. This SN occurred in almost 50% of samples (the average concentration of the drug was 7 mg kg−1 ). Additionally, other SNs were identified in tested samples (e.g., for SDZ, the maximum concentration was 35.2 mg kg−1 ). Fortunately, even short-term storage of manure could result in a significant reduction in the concentration of SNs [58]. The highest allowed concentrations of SNs in food were established in administrative regulations. The European Union adopted a maximum SN concentration of 100 g kg−1 in animal foodstuffs [61]. In Poland, the maximum permitted concentration of Asulam in fruits and vegetables is 0.5 mg kg−1 [62]. The occurrence of SNs in tissues of farmed fish has been incidental, e.g., in Slovenia, SNs residues were found in 14 of the 2363 samples [63]). SNs residues in edible marine food were detected rarely, however the concentration of SNs in tissue of common eel (Anguilla anguilla) was above 5 mg kg−1 [64]. In EU countries, the occurrence of SNs residues in edible tissues of farm animals has been insignificant. According to “Report for 2006 on the results of residue monitoring in food of animal origin in the Member States” SNs, at concentrations above their maximum allowed limits, were detected in 0.006, 0.05, 0.08, 0.97 and 3.86% of samples of poultry, bovines, pigs, eggs and rabbits, respectively [65]. Although the use of SNs in beekeeping is banned in the EU, the frequency of these drugs in honey samples is high. In Poland, it has been estimated that almost 10% of honey samples contain excessive amounts of SNs i.e., above the allowed maximum concentration. The results reported by the Chinese researchers are much less optimistic. High concentrations of SNs were determined in pig offal (almost 74 mg kg−1 of SDT and 73 mg kg−1 of STZ) and poultry offal (46 mg kg−1 of SDZ) [66]. Even more worrying is the fact that 75% of meat samples contained SNs at the total concentration >100 g kg−1 [67]. SNs could be absorbed and accumulated by plants fertilized with manure (the highest concentrations of SNs are determined in roots and leaves [9,68–71]. For example, the maximum concentration
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W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
1000 000 100 000
-1
SNs concentration (ug l )
10 000 1 000 100 10 1 0,1 0,01
Drinking Bottled Groundwater waters mineral water
SeawaterLeachate
Wastewater /influent
Hospital wastewater
Effluent
SDZ
Ac-SDM
SNs
Biosolid Soil /sludge
SDM
SPY
STZ
SMX
SDZ
SPY
SDZ
SMX
SMX
STZ
SDZ
SDT
SDM
SMX
SCP
SDM
SDM
SDZ
Surface
SMX
STZ
SDM
SMX
SDT
SDM
SMX
SDT
STZ
SMX
SMX
0 ,0 001
Ac-SDM
0,001
Manure
Fig. 5. Occurrence of the selected SNs depending on the matrix.
of SDM determined in corn, tomatoes and lettuce was 0.1 mg kg−1 [69]. Migliore et al. [71] reported that cosmopolitan weeds (Amaranthus retroflexus and Plantago major) showed a high tendency to bioaccumulate. In the tissues of these plants cultured in the medium containing SDT the accumulation rates were 2314 and 6065 mg kg−1 , respectively [71]. In our opinion, although the data presented in this section are based on the maximum values, it is possible that they can be underestimated. A commonly routine practice is the excessive and prophylactic use of antibiotics in animal husbandry and the use of manure as a fertilizer. Therefore, the local real concentrations of
SNs in the biosphere are much higher. Additionally, this effect is difficult to control due to the high mobility of SNs in the environment. For these reasons, the excessive use of manure containing SNs as fertilizer should be banned. 7. Ecotoxicity of SNs The toxicity of SNs to higher organisms (vertebrates) is not high. According to the EU directive 93/67/EEC, SNs under investigation can be classified as non-toxic or harmful [72]. The results described in the literature indicate that SNs do not exhibit mutagenic or
Table 2 Concentrations of SNs in the environment. nb
Maximal values
2.1 (0–8.5 [8]) g l (SMX); 0.011 g l (SMX) [37] 0.164 ng l−1 (SDT) [38] 0.047 (0.013–0.080) g l−1 (SMX) [38] 0.80 (0.0099–1.11 [14]) g l−1 (SMX) 0.053 (0.0002–0.09148 [45]) g l−1 (SDT) 0.87 (0.015–18 [8]) g l−1 (SMX) 2.26 (0.0108–19.2 [46]) g l−1 (SDM) 0.0475 g l−1 (SMX) [47] 379.78 (0.66–703.2 [48]) g l−1 (SCP) 46.58 (0.05–1340 [33]) g l−1 (SMX)
4 2 11 3 39 12 1 7 31
61.11 (0.0269–500 [49]) g l−1 (SDM)
17
17.78 (0.3–79.9 [51]) g l−1 (SMX) 1.28 (0.353–2.2 [51]) g l−1 (SDZ) 0.517 (0.00366–6.0 [53]) g l−1 (SMX) 1.26 (0.005–4.27 [50]) g l−1 (STZ)
6 2 30 4
Soil
22.56 (0.01–113 [54]) g kg−1 (SMX) 99.1 (1.2–197 [54]) g kg−1 (SPY) 211.6 (0.16–860 [56]) g kg−1 (SNs)
6 2 10
Manure
27.30 (0.23–167 [1]) mg kg−1 (SDM)
7
59.07 (35.2–91 [1]) mg kg−1 (SDZ)
3
8.5 g l−1 (PECc for SMX) [8] 0.080 g l−1 (SMX) [38] 3.461 g l−1 (SCT) [45] <1.11 g l−1 (SMX) [14] 19.2 g l−1 (SDM) [46] >25 g l−1 (all SNs) [43] 0.0475 g l−1 (SMX) [47] 703.2 g l−1 (SCP) [48] 1340 g l−1 (SMX; from pharmaceutical production) [33] 1158.68 g l−1 (STZ; agricultural wastewater) [50] 12.8 g l−1 (SMX) [52] PEC 92.8 g l−1 (all SNs) [51] 6.0 g l−1 (SMX) [53] 4.27 g l−1 (STZ; effluent of agricultural WWTP) [50] 197 g kg−1 dwd (SPY) [54] 31 g kg−1 (SDM) [55] 400 g kg−1 (STZ; agricultural soil) [57] PEC 860 g kg−1 (SCP; soil pore water estimation) [56] 395.730 mg kg−1 (SDT; in bedding – day 0) [58] 167 mg kg−1 (SDM) [59] 1600 g l−1 (all SNs) [60]
Matrix Drinking waters Bottled mineral water Ground water Surface water Sea water Drainflow/leachate Influent/wastewater
Hospitals wastewater Effluent (after WWTP)
Sludge (after WWTP)
Meana /the most described SNs −1
Waste landfill a b c d
Calculated based on maximal values given in tables. Number of papers. Predicted environmental concentration. Dry weight.
−1
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
carcinogenic (teratogenic) activity [73]. On the other hand, in the report “Environmentally Classified Pharmaceuticals 2009”, SNs were considered as highly toxic drugs [74]. The discrepancies between these reports probably result from different criteria used to define a risk. Directive 93/67/EEC is based on the environmental risk posses by pharmaceutical substances while “Environmentally Classified Pharmaceuticals” report assesses both environmental risk (based on the acute toxic risk to the aquatic environment) and additionally persistence and bioaccumulation of SNs in the environment (based on the information published by the Swedish Association of the Pharmaceutical Industry [74]). Fig. 6 illustrates the toxicity of SMX to selected test organisms. Important data on the SNs ecotoxicity were summarized in articles by García-Galán et al. [18] and Isidori et al. [73]. SNs are practically non-toxic to most microorganisms tested [4,18,73,75], including selected strains of bacteria, such as Vibrio fischeri, Enterococcus faecalis, Escherichia coli, Pseudomonas aeruginosa, and Staphylococcus aureus. For example, the L(E)C50 values determined using the Microtox® test (V. fischeri) ranged from 16.9 to 118.7 mg l−1 (for SMX) to >1000 mg l−1 (for STZ) [73,76,77]. Strong bacteriostatic properties caused by the SNs could significantly change the functioning of microorganisms living in the environment, for example a significant reduction of their microbial activity [78]. Additionally the number of less sensitive (resistant) strains has increased and the number of strains sensitive to SNs has decreased. Thiele-Bruhn and Beck showed that the disposing of urine that contained even a low concentration of SPY (0.02 mg kg−1 ) into the soil resulted in a significant reduction of microbial activity [78]. It was found that, in the case of SPY, the EC10 values for soil organisms ranged from 0.00014 to 0.16 mg kg−1 (the microbial Fe(III) reduction test) and from 0.0071 to 0.056 mg kg−1 (the substrate-induced respiration test) [79]. However, the most sensitive assays for the presence of SNs are bioindicators containing chlorophyll [9,18,73]. A highly toxic effect of SMX on Synechococcus leopoliensis (EC50 = 0.0268 mg l−1 ) was described by Ferrari et al. [77]. In the case of SMX, the no observed effect concentrations (NOECs) for algae (Pseudokirchneriella subcapitata and S. leopoliensis) and gibbous duckweed (Lemna gibba) were 0.090 [77], 0.0059 [77] and 0.01 mg l−1 [4], respectively. This indicates that even low concentrations of SNs may significantly affect the growth and development of plants. SNs can accumulate in various organisms in the food chain, and this accumulation could lead to a local increase in toxic effects induced by these drugs [9,10,70,71]. In addition, the toxic effects of SNs and other pollutants could exhibit a synergism [11,80,81]. At environmental exposure levels (samples contained 13 micropolutans, including SMX at the concentration of 46 ng l−1 ) the drug mix inhibited the growth of human embryonic cells HEK293, with the highest effect observed as a 30% decrease in cell proliferation compared to controls [81]. Since there has not been sufficient extensive experiments in patients with a single overdose of SNs the maximum tolerated dose in humans are unknown [82]. In laboratory experiments, acute oral overdoses of SNs in animals (LD50 ) were as follows: • in rats 10,000 mg kg−1 (SSZ), • in rabbits 2000 mg kg−1 (SSZ), • in mice 5700 mg kg−1 (SSZ), 16,500 mg kg−1 3700–4200 mg kg−1 (SAD), 4500 mg kg−1 (STZ) [83].
(SCT),
Exemplary adverse effects associated with overdosage of SNs in humans include nausea and cutaneous hypersensitivity reactions. Other adverse effects e.g., stomatitis, hemolysis, methemoglobinemia, hepatotoxicity and renal toxicity occur rarely. SNs can cause interaction with other drugs, for example with methotrexate,
7
sulfonylureas, wafarin, mercaptopurine, cyclosporine or didanosine [84]. In our opinion, direct toxic effects caused by SNs occurring in the environment do not appear to be a significant threat to public health. Potential possible cases of direct toxic effects of SNs on human may be sporadic. On the other hand, the occurrence of SNs residues in food, particularly in the case of illegal or improper use of these drugs, can be a more serious problem. According to Dolliver et al., SNs residues in food products do not pose a threat and/or adverse effect to human health but “development and spread of antibiotic resistance, which is a major problem globally” [69]. 8. Degradation of SNs in organisms and in the environment Possible products of the biotransformation and degradation of SNs are shown in Fig. 6. A detailed discussion of these processes is presented in the next sections. 8.1. Metabolism of SNs in mammals A large part of the SNs dose is excreted from organisms as unchanged compounds. For example, 75% of SMR could be excreted from the body in its parent form [1]. However, in general, over 80% of an SN dose undergoes biotransformation in mammals. The degree of transformation of each SN depends both on its type and the features of the organism. Biotransformation of SNs is mainly based on oxidation, acetylation or hydroxylation at the N4 nitrogen atom or glucuronidation of the N1 - or N4 -nitrogen atoms [1,10,11]. It is assumed that, after oral administration, 50–70% of the dose is excreted in urine as N4 -AcSNs, and 15–20% as N1 glucuronides [1,10]. The metabolites of SNs do not possess high biological activity as unchanged SNs. However, this activity could be easily restored during in vitro conditions [11,85]. The concentrations of metabolites other than those listed above are small and are likely not significant in the environment. Reviews of possible paths of SNs biotransformation were described in the papers of Sukul and Spiteller [10] and García-Galán et al. [11]. 8.2. Biodegradability of SNs The opinions of researchers on the biodegradability of SNs have been divided [1,4,10,17,24,86]. The cause of this may be the differences in microbial activity of the matrix, the inoculum used, and the applied methods used to assess SN degradation (Table 3). The stability of various SNs is also different; for example, SDM is more (10x) resistant to biodegradation than STZ. The results of standardized tests, such as the ISO 11734:1995 and OECD 301D, and the assessment of soil microbial activity suggest that most of the SNs do not undergo natural biodegradation. One of the most often described SNs in the literature is SMX, which has been regarded as a non-biodegradable compound (in pure water, seawater, natural water and wastewater or active sludge) in 9 of 24 articles [1,4,24,86,89,91,97–99]. According to Weifen et al. [114], in the presence of shrimp (Penaeus chinensis), the DT50 value for SMX is 5.68 h. Ingerslev and Halling-Sørensen [92] found that, in the presence of microorganisms in activated sludge, the DT50 of SNs is only ∼7 h. De Liguoro et al. [58] stated that, in the case of SDT, the DT50 for microbial degradation in fresh bedding is ∼1 day. Similarly, equally rapid degradation of SDT has been described by Wang et al. [87]. These authors have observed an increase in the DT50 value with increasing initial concentrations of SDT in fresh and sterile manure. In these cases, most of the SNs were incorporated into microorganisms and/or underwent only reversible transformations, such as acetylation [11,85]. The rapid disappearance of SNs in soil and manure could be an effect of binding between SNs and organic or mineral particles [85,88,93] or could be caused by
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100000
L(E)C50 (mg l-1) or (mg kg -1)
10000 1000 100 10 1 0,1 0,01
Bacterium
Diatom
Algae
Aqua Gnitic daria plant
Crustacean
Rotifer
In verte brate
Fish
Children (PNEC)
D. rerio (96h)
O. latipes (96h)
O. mykiss (EA)
T. platyurus (24h) B. calciflorus (24h) B. calciflorus (48h) M. macrocopa (48h)
C. dubia (7d)
C. dubia (48h)
H. attenuata (96h) D. magna (24h) D. magna (48h) D. magna (48h)
L. gibba (7d)
Actived sludge (20d) AMES test (72h) C. meneghiniana S. capricornutum P. subcapitata (96h) C. vulgaris (48h) S. subspicatus (72 h) S. leopolensis (96h)
C. freudii (24h)
P. agglomerans P. aeruginosa (48h) V.fischeri (5min) V. fischeri (15min) V. fischeri (30min)
0,001
Mam mals
Fig. 6. Comparison of SMX toxicity to selected test organisms.
photochemical processes (on the soil surface, in the presence of Fe compounds and nitrates) [107,115]. Most researchers recognized SNs as poor or non-biodegradable compounds in the environment (in pure water, surface water and in soil with a DT50 > 30 days) [24,74]. The fact that SNs occurred so often in test samples could also be considered as evidence of their persistence in the environment. In our unpublished study on the biodegradation of SAD, STZ, SMX and SDZ applied to natural matrices, we found that, in individual cases (STZ under aerobic conditions, in wastewater and water from the swamp), the DT50 was less than 2.5 days. In the remaining cases, the DT50 values for SDZ, SAD and SMX were >5, >8 and 31 days, respectively. In our opinion the read stability data related to SNs residues in the environment (especially for SMX) are generally much higher than the data reported by other researchers in the cited articles. The above-described high frequency of SNs in environmental samples can confirm this assumption. 8.3. Physicochemical methods of degradation of SNs The efficiency of SNs degradation using the most commonly used chemical and physicochemical methods is presented in Table 3. The high degradation efficiency of SNs in wastewater was obtained using various advanced oxidation processes (AOP) [1,4,101,105,108,116,117], such as the use of O3 , Cl2 , and ClO2 [1,101,116–118], the Fenton reaction [105,116] or photocatalytic processes [85,105,116,117]. Unfortunately, the application of these methods is costly and could be harmful to the environment due to the formation of highly toxic intermediates [118]. Moreover, a decrease in the efficiency of AOP with an increase in overall wastewater pollution was observed [108]. This fact made it difficult to apply these methods directly to remove SNs from manure. In Table 3, examples of other methods used to remove contaminants from the aquatic environment without their degradation or transformation (non-destructive methods) are presented. SNs can be removed from wastewater with nearly 100% efficiency by reverse osmosis [1,4,24,109,110]. However, with this method, there could be a problem with wastewater containing concentrated solutions
of toxins (including SNs) [110]. In the case of substances resistant to biodegradation, there could be a local, risky increase in the concentration of these toxins in a small area [119]. The physico-chemicals methods (particularly AOP) can be effective and very useful in the degradation of SNs. In our opinion, it is not excluded that their environmental degradation is the result of photochemical reactions initiated by sunlight in the presence of natural photosensitizers, and not only of biodegradation processes. 9. Removal of SNs from wastewater Opinions on the efficiency of SN removal in conventional biological-mechanical treatment plants are divergent. Similar differences occur during the assessment of the biodegradability of drugs (Table 3). Onesios et al. [120] analyzed 49 cases of the removal of SDZ, SDM, SMX and SPY from wastewater in wastewater treatment plants (WWTPs). Based on the analysis of recent publications, −280 to 100% of SNs were removed using activated sludge (AS) (Table 4). The mean degree of SN removal in these cases was ∼24%. According to the data published in 2010, SMX was removed from the selected WWTPs in Spain in the range of 30–92% [35]. However, there are also cases in which the concentration of SNs in effluent was higher than that in the influent [4,86,134,138]. This effect was described in a pilot WWTP in Austria [90] and in Switzerland [138]. This effect is likely caused by hydrolysis of the N4 -AcSNs present in wastewater to the parent SNs [11]. A conclusion of this problem may be found in the data from the study by Turkdogan and Yetilmezsoy [109]. These authors have estimated that 80% of used antibiotics enter the environment despite the use of various processes in WWTPs (based on the data from Turkey, without regard to SNs). Importantly, a large part of SNs may be adsorbed in WWTPs by biomass [132] and could return again to the environment. 10. The environmental risk assessment The majority of researchers have used the method recommended by the European Medicines Evaluation Agency (EMEA) for environmental risk assessment. This method uses the results of
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
9
Table 3 Biotransformation, degradation and other methods of SNs removal. Matrix
Methods
Biotransformation Human Animal Manure
Efficiency
DT50
<90% (SNs) [1,11] 50–90% (SNs) [11] 0% after 28 days (SDZ) [1]; 10% after 11 weeks (SDM) [1]; 25% after 15 days (SDT)
Bedding Wastewater Membrane bioreactor Activated sludge Soil Sediment Surface water
99.5% after 28 days (SDT) [58] 41 (0–90)% (SNs) [1,89] 70–90% (SMX) [1,90] 42 (−138 to 99)% (SMX) [86,91] kbiol = 0.1–10 l g SS−1 d−1 (SMX)a [4] 0% after 28 days (SDZ) [1]; 0.2–0.3% after 64 days (SDM) [1] 0–90% after 28 days (SDZ) [1,94]; 20% after 180 days (in marine sediment) [1] 24 (0–82)% (SNs) [95,96]; practically non-biodegradable (SMX) [1,91,97]
Pond water
Physicochemical degradation Pure water
Pure water Wastewater Pure water Wastewater Non-destructive methods Permeate Concentrate
a
0.3–4.1 days (SNs) [92] 2.8–21.3 days (SCP) [88,93]; <10 days (SDM) [1]; <15 days (SDZ) [1] 0.7 day (SDM) [94]; 4.9–10.1 days (SMX) [94]
15.8 (1.7–47.6) days (SNs) [94] OECD 301D test
Non-biodegradable after 40 days [1,98]; non-biodegradable in manure (SDT) [97]; 4% after 28 days (SMX) [24]; BOD5 /TOD <1.5% (SNs) [99] BOD5 /COD ≈ 0 (SMX) [89]
Photolysis HClO Fe(VI) ClO2 Cl2 HO• O3
>90% (at 254 nm; 2768 mJ cm−2 ) [100]
O3 /H2 O2 Photo Fenton Fenton TiO2
TiO2 /FeCl3
Reverse osmosis Microfiltration Bank filtration Adsorption
Hospitals wastewater Pure water Effluent
>30 days (SNs) [1] 1.36–2.56 days (SDT) [87]; 7 days (SDM) [1]; 61 days (SDT) [58], 127 days (SCP) [88] 1 day (SDT)[58]
Coagulation Al2 (SO4 )3 Ionic treatment MIEX® resin
6–181 s [101] 91–241 s [101] k = 2.2–103 l mol−1 s−1 [101] <88% after 2 h; k = 0.00025–0.0347 s−1 [102] k = 3.7–7.1 109 l mol−1 s−1 (SNs) [25] >90% (SMX) [1]; <99% after 60 min [103]; k = 2.5–106 l mol−1 s−1 [101] <99% after 20 min [104] <100% (at 5 Einstein m−3 ) [89] >90% after 10 min (SNs) [105] ∼100% after 180–300 min (SNs) [99]; 88% after 360 min [106]; 15–30% after 60 min (SNs) [108] <90% after 90 min (SAD) [107] 62–84% after 60 min (SNs) [108]
∼20 min (SAD)[107]
∼4 min (SAD) [107]; 11–42 min (SNs) [108]
∼100% [109,110] −58% [110] −18% [109]; 0–90% (SMX) [1] 25% (SMX) [111] 89–98% Micelles; 45–58% Activated carbon [112]; ∼0% [1,4] ∼0 (−50 to 21.3)% [112] >90% (SNs) [113] 40–90% (SNs) [113]
kbiol kinetic constant for pseudo first order biodegradation (l gSS−1 d−1 ), suspended solids concentration (gSS l−1 ) [4].
toxicological studies and is based on calculating the hazard quotient (HQ) as the ratio of the PEC value to the predicted no-effect concentration (PNEC) [1,4,13,17,18,24,72–76,109,141]. The method for the determination of these values was described in detail by Koschorreck et al. [24], Park and Choi [75], Kim et al. [76] and Lopes de Souza et al. [141]. A similar method is based on a calculation of the MEC/ PNEC ratio in which MEC is the maximum environmental concentration [1,8,35,73,75]. Typically, values of HQ < 1 indicate that the substance analyzed could be considered environmentally safe. A comprehensive review of the data on the HQ values for 5 selected SNs was published by García-Galán et al. [18]. Although the presented HQ values were mainly obtained for SMX, they are significantly different. The selected data on the HQ values calculated based on the available literature and this review [18] are shown in Table 5. However, the maximum HQ values have probably negligible importance. According to Schwab et al. [8], the concentrations of SNs in the environment do not pose a risk to human health. Moreover, according to Environmentally Classified Pharmaceuticals
(2009), the environmental risk of SNs is specified as insignificant [74]. On the other hand, data on the quantity of these drugs in matrices such as manure, wastewater from agricultural fields and pharmaceutical industries indicate that, in these cases, SNs could cause serious problems for the environment. The potential negative effects on the soil microorganisms of SNs present in manure and bedding are especially alarming (Table 5). Moreover, changes in the genotypes of microorganisms are often not taken into account. In contrast to the toxic effects, these changes could easily be transferred, even to species in other biocenoses. 11. Generation of drug resistance In populations of bacteria that are sensitive to specific antibiotics, there are intrinsically occurring strains that are resistant to at least one drug (natural resistance). As a result, these resistant bacteria can survive, multiply and spread to others in the family [7,9,12,30]. In Nicole Kemper’s article [12] on the influence of veterinary antibiotics on the environment, the author formulated
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W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
Table 4 Efficiency of WWTPs (full scale, pilot and lab scale). WWTP
SNs
Influent (g l−1 )
Efficiency (%)
Ref.
Guangzhou, China (AS, F, S or Cl2 )
SDZ SMX SDM SMX SDM
5.10–5.15 5.45–7.91 0.11; 0.21 0.13–1.25 0.52
∼100
[121]
∼100 17.8–100 >93
[122]
SMX SNs/SDZ SMX SDT SDM SMX SDT SDM SMX SMX SMX
0.68 0.06–0.51 0.179–1.760 10
SDT SMX
0.25–1.3 0.18 0.44 0.07 0.72–0.88
38 20–82 26–88 66 46 76 93 72 90 74 ± 13 26 26 62 36–77
SMX
0.01–0.118
0–64
SDZ SDM SMX SMX SAD
0.072 0.696 0.3 0.7 0.82 max 1.2
50 50 2 92 24 54–91
SMO SCP Sulfaquinoxaline SNs
max 0.215 max 0.057 max 0.103 0.362
60–89 60–82 73–85 25
SDT SPY SMX SPY SMX SPY SMX SMX SMX
0.0047 0.495 0.104 0.495 0.104 0.495 0.104
>67 4.6 62 18.6 72 95 96 −280 to 61 0–91
[132] [133]
−138 to 99 (33) 57–90 (73) kbiol = 5.9–7.6 l gSS−1 d−1 kbiol = 3.2–5.0 l gSS−1 d−1 kbiol < 0.1 l gSS−1 d−1 43–98 (69) 30–92 (74) 62 27 25 −107 to 72 −138 to 60 85–96 41; 52 −61; 29 81; 86 −21 to 0 −29 to 21 9–21 68.2 95.7 72.8 75
[86]
Wisconsin, USA (AS) Fort Collins, USA (Drake water reclamation facility) Taiwan (AS or trickling filter, Cl2 or UV) Laboratory scale up-flow bio-reactors
Terrassa, Spain (CAS, MBR) Japan (AS, Cl2 )
Erie County, USA (extended aeration; rotating contactors; pure oxygen aeration) Pearl River Delta, South China, (AS, oxidation ditch, Cl2 , UV)
ˇ Croatia Cakovec, Braunschweig, Germany (AS) Slaughterhouse, Beijing, China, (anaerobic–anoxic–oxic or anaerobic–oxic processes)
Brisbane, Australia (CAS, advanced wastewater treatment) Northwest Ohio, USA, Tokyo, Japan, (AS)
AS+F As+F+O3 Austria (MBR pilot plant or CAS) Kloten/Opfikon or Altenrhein Swiss, (MBR pilot plant or CAS) CAS (n = 20) MBR (n = 3) CAS MBR CAS and MBR Spain (AS or biologic filters)
SMX N4 -AcSMX
Stanley and Shatin, Hong Kong
SMX/SDM SDZ SMX SMX SMX SNs SPY SMX N4 -AcSMX SPY SMX N4 -AcSMX SPY/SMX SPY/SMX N4 -AcSMX SMX
Luxemburg
SDZ SMX
Jamaica Bay Albuquerque, USA Turkey (CAS) Swiss (CAS)
(Fixed-bed reactor)
(Sand filtration) (Primary wastewater treatment)
2
0.06–0.15 0.23–0.57 0.85–1.6
0.1465 0.3555 0.0730 max = 0.155
[123]
[124] [125]
[126] [127]
[128] [55]
[129] [130] [131]
[110]
[90] [134]
[135]
[35] [136] [137] [109] [138]
[139]
[140]
AS, activated sludge; CAS, conventional activated sludge; F, filtration; S, sedimentation; Cl2 , chlorination; MBR, membrane bioreactor; O3 , ozonation; UV–UV illumination.
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
11
Table 5 Ecotoxicological data on the HQa value (based on the available literature). Matrix
Maximal values of HQ calculated based on data from Table 2
Maximal values of HQ presented in the literature (only for SMX) Comments
Drinking water
8.5b /0.05c = 170 (SMX)
0.0097 [8]
Surface water
59.30 [77]
Wastewater Aquatic environment
18/0.05c = 360 (SMX) 19.2/201d = 0.9552 (SDM) 1340/0.05c = 26800 (SMX) –
Hospital wastewater Soil
12.8/0.05c = 256 (SMX) 395.73e /0.00014f ∼ = 2.8 × 106
15.1 [51]
a b c d e f
22.96 [35] 6.3 [76]
For child drinking water and fish consumption, US Acute toxicity test, Germany For algae, Spain The PEC of test pharmaceuticals was estimated based on several conservative assumptions, Korea For hospital effluent, Germany
HQ = P(M)EC/PNEC. PEC. PNEC = NOEC/10 for S. leopoliensis. For D. magna. SDT; in fresh bedding. The microbial Fe(III) reduction test for SPY.
the following thesis: “Resistance is provoked by repeated exposition of bacteria to sub-lethal dosages of antibiotics, as realized by continuing manuring with contaminated faeces on land used agriculturally”. Although the natural resistance to pathogenic bacteria has not been transferred between strains, the formation of drug resistance by the transfer of “resistance” genes between bacterial cells belonging to different strains, or even genera, during one recombination process (horizontal gene transfer) may have contributed to the dissemination of drug-resistant bacterial species on a large scale. As a result, these strains may occur in ecosystems theoretically not exposed to chemotherapeutics [142–144]. For example, Pallecchi et al. described the occurrence of drug resistance in 67% of members of the Guaraní Indian community of Alto Los Athletic (Bolivia) [143]. Drug resistance against one group of drugs may favour the generation of drug resistance to other drugs or disinfectants [145]. Due to the importance of pathogenic resistance to human health, programs for monitoring microorganism resistance in Europe and the Americas have been implemented [7,142,146]. For example, the ECO-SENS project has collected and
analyzed drug resistance data in 17 European and American countries since the 1960s [7,142]. The resistance of pathogenic bacteria to SNs may be due to structural changes in dihydropteroate synthase (DHPS, Fig. 3) that are the effect of point mutations in the DHPS gene (folP) [147]. Thus, these mutations affect the expression of the DHPS enzyme with has a lower affinity for SNs. The spontaneous mutants of E. coli, showing resistance to SNs as a result of the substitution of one or more base pairs in the DHPS gene have been isolated in laboratory [147]. SNs resistance may be also distributed on mobile genetic elements, such as plasmids, transposons and integrons [148,149]. There have been three known genes encoding resistance to SNs [144]. The sul1 gene is usually located on the 3 conserved region of class 1 integrons, which are a part of large conjugative plasmids and Tn21-like transposons. The dissemination of this gene increased with the prevalence of class 1 integrons in bacterial pathogens. The sul2 gene was first identified on the RSF 1010 plasmid in E. coli. It is frequently located on large conjugative plasmids, e.g., pGS05,
Table 6 Dissemination of SNs resistance genes (sul1, sul2 and sul3). Matrix
SNs-resistant isolates positive for sul1-3 genes (%)
SNs-resistant isolates (%)
Ref.
Pigs Swine Cattle Dogs and cats Laying hens Pigs Wild small mammals Danish broiler faeces, and meat Broiler meat Foodstuffs Wastewater directly from swine farms Shrimp ponds City canal/fish ponds Water-sediment and Manure Faecal samples Urine UK 1991 UK 1999 UK 2004 Europe before 1990 Europe 1999–2000 Healthy humans Humans Animal, food and human
11-84(sul1), 19-54(sul2), 3-46(sul3)
50–97 81 22 20 26 50 6 15–18 45 92.5
[152] [153]
18(sul1), 20(sul2), 18(sul3) 5(sul1), 1(sul2) 11(sul1), 82–100(sul2) 26(sul1), 61(sul2), 8(sul3) 69.8(sul1), 36.9(sul2), 1.4(sul3) 92 43 72 14(sul1), 96(sul2) 100(sul1–3) 43 53.9 57.5
33(sul1), 91(sul2), 5(sul3) 16(sul1), 97.5(sul2)
(total of SNs resistant isolates positive for sul genes)
[154] [155] [156] [157]
[158] [159] 39.7 46 45.5 0–5 9–26 74 100
[160]
[142] [155] [143] [31]
12
W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1–15
and on small non-conjugative plasmids, such as pBP1, pHD148 and RSF 1010. The last two plasmids also carried genes associated with resistance to streptomycin, and therefore, the resistance to SNs and streptomycin are strongly linked together [150,151]. The sul3 gene has been found in pathogenic E. coli isolated from swine. The dissemination of each sul1-3 gene depends on the location of sampling sites and bacterial species. Among Gram-negative isolates resistant to SNs, mainly E. coli and Salmonella, the sul1 and sul2 genes are often found at almost an equal frequency [144]. Resistant bacterial species commonly carried single genes, but in recent years, an increased number of pathogens that possess three SNs-resistant genes have been observed (Table 6). In environmental matrices, the presence of organisms resistant to SNs could be determined by detection of the genes described above. Most often, bacterial resistance to SNs has been described in E. coli, Salmonella enterica and Shigella spp. from the manure of farm animals, from meat and meat products, from healthy humans with urinary infections and from wastewater (Table 6). However, all SNs-resistant bacterial species positive for the sul genes and plasmids mentioned above were identified and classified as belonging to thirteen genera, namely Acinetobacter, Aeromonas, Arthrobacter, Bacillus, Brachybacterium, Cellulosimicrobium, Enterobacter, Escherichia, Pseudoalteromonas, Pseudomonas, Shigella, Vitreoscilla and Wautersiella [157].The most important facts related to drug-resistance include the following: • the use of antibiotics in veterinary medicine increases the drugresistance of microorganisms, including cross-resistance [9,161], • the presence of SNs in the environment increases the antimicrobial resistance of microorganisms [9,12], • the number of bacterial strains resistant to SNs increases systematically in recent years [7,160], • SNs have shown the highest drug resistance, almost twice as high as tetracyclines and many times higher than other antibiotics [153].
12. Conclusions • Antibacterial SNs are a group of drugs still commonly used in human and veterinary medicine. Used SNs could be spread into the environment in an almost entirely biologically active form or could recover activity. • Opinions on the possibility of SNs removal in conventional WWTPs are divergent. There are known technologies that could completely degrade SNs in WWTP. However, nearly 80% of used SNs have reached the biosphere. This indicates that some existing, modern technologies are not able to manage the degradation of SNs. • SNs introduced into the environment likely remain there for a long time and could spread easily and infiltrate groundwater. • The frequency of SNs in tested environmental samples is very high. • SNs have a very low toxicity to higher organisms (vertebrates) and are highly toxic to microorganisms, algae and certain plants. • SNs occurring in the environment favour the generation of drug resistance. SNs resistance genes may be transferred into the environment. High concentrations of SNs in the environment occur incidentally (in manure from livestock), but due to a gene transfer process, their relevance to the global change of drug resistance may be much larger than expected. The risk caused by the generation of drug resistance by anti-infectives drugs is much higher than the risk caused by their toxicity.
These facts indicate that the problem presented here has a serious global importance in ecology, and limitations of antibiotic consumption in individual countries will not solve this problem. The data also pointed to the need to search for effective and inexpensive methods of removing pollutants from the environment.
Acknowledgements This work was supported by Medical University of Silesia in Katowice (Poland), Contract No. KNW-1-015/10.
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Journal of Hazardous Materials 196 (2011) 16–21
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Elimination of water pathogens with solar radiation using an automated sequential batch CPC reactor ˜ a,∗ , E. Ubomba-Jaswa b , C. Navntoft c,d , I. García-Fernández a , M.I. Polo-López a , P. Fernández-Ibánez f f P.S.M. Dunlop , M. Schmid , J.A. Byrne f , K.G. McGuigan e a
Plataforma Solar de Almería – CIEMAT, PO Box 22, 04200 Tabernas, Almería, Spain Natural Resources and the Environment, CSIR, PO Box 395, Pretoria, South Africa c Instituto de Investigación e Ingeniería Ambiental, Universidad Nacional de San Martín (3iA-UNSAM), Peatonal Belgrano 3563, B1650ANQ San Martín, Argentina d Universidad Tecnológica Nacional – Facultad Regional Buenos Aires – Departamento de Ingeniería Civil - Laboratorio de Estudios sobre Energía Solar, (UTN-FRBA-LESES), Mozart 2300, (1407) Ciudad Autónoma de Buenos Aires, República Argentina e Nanotechnology and Integrated BioEngineering Centre, University of Ulster, Shore Road, Newtownabbey, Northern Ireland BT37 0QB, United Kingdom f Department of Physiology and Medical Physics, Royal College of Surgeons in Ireland, Dublin 2, Ireland b
a r t i c l e
i n f o
Article history: Received 24 May 2011 Received in revised form 12 August 2011 Accepted 16 August 2011 Available online 10 September 2011 Keywords: Solar disinfection Escherichia coli Compound parabolic collector
a b s t r a c t Solar disinfection (SODIS) of water is a well-known, effective treatment process which is practiced at household level in many developing countries. However, this process is limited by the small volume treated and there is no indication of treatment efficacy for the user. Low cost glass tube reactors, together with compound parabolic collector (CPC) technology, have been shown to significantly increase the efficiency of solar disinfection. However, these reactors still require user input to control each batch SODIS process and there is no feedback that the process is complete. Automatic operation of the batch SODIS process, controlled by UVA-radiation sensors, can provide information on the status of the process, can ensure the required UVA dose to achieve complete disinfection is received and reduces user work-load through automatic sequential batch processing. In this work, an enhanced CPC photo-reactor with a concentration factor of 1.89 was developed. The apparatus was automated to achieve exposure to a predetermined UVA dose. Treated water was automatically dispensed into a reservoir tank. The reactor was tested using Escherichia coli as a model pathogen in natural well water. A 6-log inactivation of E. coli was achieved following exposure to the minimum uninterrupted lethal UVA dose. The enhanced reactor decreased the exposure time required to achieve the lethal UVA dose, in comparison to a CPC system with a concentration factor of 1.0. Doubling the lethal UVA dose prevented the need for a period of post-exposure dark inactivation and reduced the overall treatment time. Using this reactor, SODIS can be automatically carried out at an affordable cost, with reduced exposure time and minimal user input. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Lack of access to a reliable and safe source of potable water is a significant problem in developing countries. Each year, there are approximately 4 billion cases of diarrhoea resulting in an estimated 1.8 million fatalities. Every day approximately 4500 children die of dehydration due to diarrhoea [1]. Water treatment processes which are robust, easy to use and low cost could be readily deployed at point-of-use and may also find application in emergency situations, where access to safe potable water is a primary concern.
∗ Corresponding author. Tel.: +34 950 387957; fax: +34 950 365015. E-mail addresses:
[email protected] (M.I. Polo-López),
[email protected] ˜
[email protected] (E. Ubomba-Jaswa), (P. Fernández-Ibánez),
[email protected] (C. Navntoft),
[email protected] (I. García-Fernández),
[email protected] (P.S.M. Dunlop),
[email protected] (J.A. Byrne),
[email protected] (K.G. McGuigan). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.052
Solar disinfection (SODIS) is a water treatment method suitable for use at household level. Normally SODIS is carried out by placing water in transparent containers (usually 2 L plastic PET bottles) and exposing to sunlight (≥6 h) [2,3]. The synergistic effect of mild thermal heating and solar UV radiation is responsible for the inactivation of pathogens in the water. The inactivation rate depends on the temperature reached during the process and also on the type of microorganism present in the water [4,5]. This basic SODIS practice has significant limitations which include, (a) the recommended time for SODIS treatment is 6 h in full sunshine or, two consecutive days in cloudy conditions; (b) the volume of water treated is small, typically 1.5 to 2 L in bottles; and (c) the user has no feedback indicating treatment efficacy or completion. SODIS in glass tube photo-reactors (with and without photocatalyst), incorporating compound parabolic collectors (CPC’s), has been shown to be effective for the inactivation of a range of microorganisms, including bacteria (E. coli) and fungi (Fusarium spp.) [6–8].
M.I. Polo-López et al. / Journal of Hazardous Materials 196 (2011) 16–21
17
Our recent contribution shows a new low cost SODIS reactor for purifying 25 L-batches of water, utilising CPC enhancement and low cost materials. This system was tested for six months under natural sunlight and was demonstrated to be efficient against E. coli [9]. Even with improvements in reactor efficiency, the SODIS process is both dependant upon, and controlled by users, i.e., a person must check that treatment is carried out under the recommended protocol for a minimum treatment time of 6 h. For example, the user must pay attention to the local weather, note the exposure time and trust that process will improve the microbiological safety of the treated water. These limitations may contribute to low levels of compliance in the use of SODIS. As the treatment time is dependent on the ambient solar irradiance, the measurement of the UVA dose may be used to indicate treatment completion, or preferably, provide feedback control of the process. The UVA dose can be calculated as follows:
Dose(J m−2 ) =
UVA(W m−2 )dt(s)C
(1)
where UVA is the solar irradiance (320–400 nm) incident upon the reactor; dt is the exposure time; and C is the concentration factor of the mirror [6]. C is a dimensionless number that defines the multiplication factor by which sunlight is concentrated at the absorber/receiver. In this case, the absorber is the glass tube of the photo-reactor. We have recently demonstrated that SODIS relies upon the receipt of a minimum and uninterrupted UVA dose, defined as the “lethal UVA dose”. For 106 CFU mL−1 of E. coli K-12 in 2.5 L of well-water in a CPC reactor with C = 1, this dose was found to be ≥108 kJ/m2 . The lethal dose depends on the total amount of water treated per batch. This means that the amount of solar UVA energy per unit of volume that has to be delivered in an uninterrupted manner into the system is 8.6 kJ/L (where the irradiated collector surface is 0.2 m2 ; and the total volume is 2.5 L). This lethal dose also depends on the level and nature of the microbiological contamination, and on the physical and chemical properties of the water. For example – and this applies for all water treatments – the more resistant the microorganism the more energy will be required to disinfect the water. For this reason, the lethal dose must be experimentally determined for very different natural water sources such as open water (rivers, lakes, streams, ponds, etc.), underground sources (wells, aquifers) or rainwater. The lethal UVA dose was also demonstrated to be independent of UVA irradiance, for solar UVA irradiance between 14 and 40 W m−2 [10]. CPC enhanced SODIS reduced the time needed for complete inactivation (below the detection limit) of bacteria on both cloudy and sunny days. However, following receipt of the lethal UVA dose, a period of approximately 2 h post-exposure was necessary before complete disinfection (i.e. 6-log unit reduction) was accomplished [10]. For example, a 3-log kill was observed if the water was tested immediately following the lethal dose (1 h in sunny conditions), but a 6-log kill was later observed after the water was left to stand for 2 h following exposure. Therefore, the total treatment time for a 6-log kill was 3 h. In an attempt to address the practical problems associated with SODIS, a novel sequential batch photo-reactor was designed with the aim of decreasing the treatment time required and reducing user-dependency. The new photoreactor incorporated two major improvements over traditional CPC photo-reactors. Firstly, to reduce the solar exposure time required to receive the lethal UVA dose, the concentration factor C of the CPC was increased from 1.00 to 1.89, i.e. the glass tube receives almost twice the quantity of UV solar radiation in comparison to a C = 1 CPC system. Secondly, the treatment time was automatically controlled by an electronic UVA sensor. The feedback sensor system controlled the gravity-filling of the reactor from an untreated water reservoir, and
Fig. 1. Schematic of the sequential batch system.
controlled the discharge of the treated water into a clean reservoir tank following receipt of the pre-defined UVA dose. The full sequence was then automatically repeated for as many times as permitted by the solar UVA intensity during daylight hours. The reactor was tested using E. coli as the model pathogen in well water under real sun conditions.
2. Materials and methods 2.1. Sequential batch photo-reactor The sequential batch photo-reactor consisted of a glass tube positioned at the focus of a CPC mirror; two 25 L reservoir tanks (the untreated water tank (UWT) and the treated water tank (TWT)); a control system consisting of a UVA photodiode, electronic valves to control fluid flow and the necessary hardware/software to automate the device (Fig. 1). The electronic control system measured the solar intensity and calculated the solar UVA dose. When the preprogrammed dose had been acquired, a series of electronic valves opened to dispense the treated water into the TWT. The tube was subsequently refilled from the UWT and the treatment cycle automatically re-started. The system also included water level sensors in the UWT and TWT. These sensors were incorporated to stop the cycle if the UWT level was too low or the TWT level was too high. The photoreactor tube (1.50 m length, 0.05 m outer diameter, 1.8 mm wall thickness, and 2.5 L illuminated volume) was made of borosilicate glass (Schott-Duran, Germany). The glass had a transmittance of 89–90% in the UVA range. The tube was sealed with PTFE (Polytetrafluoroethylene) end caps connected to two electronic valves (Betavalve, UK), which were regulated by the control system. The CPC mirrors were made from highly reflective aluminium sheets (type 320G ALANOD anodized aluminium of 0.5 mm thickness, Alanod Aluminium GmbH, Ennepetal, Germany). The manufacturer reports a reflectivity of 82% for the UV and 85% for the rest of the solar spectrum. CPC mirrors with C = 1.00 and C = 1.89 were used in these experiments. A major advantage of CPC systems is that the concentration factor remains constant for all values of sun zenith angle within the acceptance angle limit, whereas conventional parabolas or flat mirrors require sun tracking to maintain the same concentration factor. On the other hand, CPC mirrors require almost 2–4 times the reflective area of a conventional parabola. Due to the inherent characteristics of non-imaging optics used by CPC reflectors, the area of reflectors can be truncated to almost 50% of their actual length with a loss of less than 10% in the concentration ratio [11].
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29.70cm
Height = 6.43cm
Aperture width = 15.70cm
19.37cm
(b)
(a)
C=1
C=1.89
Fig. 2. Diagram of CPC mirrors with concentration factor 1 (a) and 1.89 (b).
In this way, the total reflector area is reduced to half its original length and there is very little loss in radiation concentration. For the case of C = 1, the acceptance angle is c = 90◦ and the result is an involute reflector with an aperture width of 15.70 cm which is shown in Fig. 2(a). For the case of C = 1.89, the acceptance angle is c = 30◦ . For C = 2 total reflector height must be 36.13 cm and 31.4 cm of aperture width. To simplify manufacturing and lower reflector area, the CPC was truncated to almost half it’s height, 19.37 cm, an aperture width of 29.70 cm and C = 1.89, as can be seen in Fig. 2(b). Hence, the mirror area was reduced by nearly 50%, but this only reduced the concentration factor by 5%. As mentioned earlier, only solar rays with an incidence angle lower than the acceptance angle are useful for concentration purposes. In the case where C = 1, c = 90, the concentrator accepts all sun rays from sunrise until sunset. For C = 1.89, only rays with < 30◦ will be accepted. For the fixed and inclined system used in this work, such incidence angles are obtained approximately ±2 h either side of solar noon, yielding approximately, between 4 and 7 h of useful concentrated sunlight, depending on the season of the year. In the case of fixed systems (non-tracking) equipped with CPC mirrors, the available hours of sun within the acceptance angle diminishes as the concentration factor rises. The mathematical relationships between the available hours of sunshine and the acceptance angle of CPC have been explained in detail previously, e.g., by Rabl [12] (Fig. 3). The UVA dosage is determined only by exposure time (t, s) and irradiance (UVA, W m−2 ), as explained in the introduction (Eq. (1)). The size of the reactor is affected by two design parameters: (1) concentrating factor of the solar mirror, and (2) total volume of treated water. In this study, we used two solar reactor systems, one with a concentration factor of 1.0 and the other with 1.89. The total volume and irradiated volume in both reactors was the same. That means that UVA irradiance collected by the mirror and delivered to the water only depends on exposure time and concentration factor.
The experiments were started at different local times so the system received different UVA dosages during irradiation. 2.2. Measurement of solar radiation Solar UVA radiation was measured with a global UVA radiometer described elsewhere [9]. The radiometer had the same inclination as that of the platform where experiments were conducted. UVA irradiance was measured outside the tube. It is recognised that photonic losses may occur due to absorption and scattering effects within the reactor. Quantification of efficient radiation levels inside the reactor cannot be easily determined, and is a matter of independent theoretical and experimental study elsewhere. Nevertheless, our studies supporting the lethal dose concept are based on UVA dose measurements done also outside the tube with same type of well water (equal turbidity and bacterial load), therefore the correlation between UVA dose received and disinfection results, as determined in our previous work [10], can be considered as valid for the present study. 2.2.1. Calibration of UVA control sensor within the sequential batch system The UVA photodiode (TW30SX, Sg-lux, Germany) and control electronics were calibrated against a spectral radiometer (Gemini 180, Jobin Yvon, UK) using a 1 kW Xenon source fitted with AM1 filter. A linear response was observed within a UVA range of 5–60 W m−2 described by the following relationship: Output voltage (V) = 0.0069 × UVA irradiance (W m−2 ) + 0.0045; R2 = 0.999. The sensor’s response was also validated against global solar UVA radiation at Plataforma Solar de Almería (PSA) using the global UVA radiometer (295–385 nm, Model CUV3, Kipp & Zonen, Netherlands). Fig. 4 shows the response observed during full sun
Solar field UVA radiometer (case 1) Sensor (case 1)
-2
UVA irradiance (W m )
40
30
20
10 Solar field UVA radiometer (case 2) Sensor (case 2) 0 11:00
12:00
13:00
14:00
15:00
16:00
Local time (hh:mm) Fig. 3. CPC collector diagram showing the acceptance angle ( c ) and aperture.
Fig. 4. Response of the Sg-lux sensor (dashed lines) and global UVA radiometer at PSA (solid line) during sunny (case 1) and cloudy weather (case 2).
M.I. Polo-López et al. / Journal of Hazardous Materials 196 (2011) 16–21 Table 1 Summary of physical and chemical properties of the well-water batch used for the experiments. Natural well-water at PSA Cl− NO3 − SO4 2− F− Br− PO4 3− pH Turbidity TOC
285 ± 2 mg/L 8.2 ± 0.5 mg/L 205.0 ± 0.5 mg/L 0.9 ± 0.3 mg/L ND ND 7.8 1.5 NTU 5 mg/L
Na+ NH4 + K+ Mg2+ Ca2+ HCO3 − Conductivity Bacteria COD
501.1 ± 0.8 mg/L ND 9.4 ± 0.3 mg/L 64.5 ± 0.6 mg/L 79.1 ± 0.5 mg/L 495 ± 15 mg/L 2805 S/cm 0 CFU mL−1 45 mg/L
(27th February 2008) and during cloudy weather (8th April 2008). In both weather conditions the response was accurate, however, when the sun was at a low angle (early morning, late afternoon) shading of the sensor’s active area by the diode casing occurred and the accuracy decreased slightly. Therefore, the sensor was calibrated between 11.00 and 16.00 h local time.
19
agar plates, incubated at 37 ◦ C overnight and counted the following day. To determine the initial bacterial concentration in the reactor, a sample of water was taken for bacterial enumeration before the system was exposed to sunlight. This sample was maintained in the dark at laboratory temperature (25 ◦ C) for the duration of the solar exposure experiment (“no treatment control”) and the bacterial concentration determined as described above. Volumes of 250 L of undiluted samples were plated when bacterial concentration was expected to be below 1 CFU per plate; therefore, the detection limit for this quantification method was 4 CFU mL−1 . Analysis for bacterial re-growth was undertaken for all experiments by leaving the last two samples taken from the reactor at room temperature for 24 h and 48 h. Bacterial concentration was determined using the plate count method described above with samples plated onto both LB agar and Endo agar (Sigma–Aldrich, USA) plates with samples taken after 24 and 48 h. All experiments were conducted in triplicate, and each bacterial sample was plated in triplicate. Statistical data analysis was carried out as described in Ref. [9]. Data points in figures represent the average of data analysis and the error bars show the standard deviation.
2.3. Solar disinfection experiments In a typical experiment, the UWT was filled with 25 L of well water inoculated with E. coli to give an initial bacterial loading of 106 colony forming units per mL (CFU mL−1 ). The control cycle was initialised which filled the photoreactor with 2.5 L. Following exposure to the pre-defined UVA dose, the system automatically discharged the water from the photoreactor into the TWT. Samples were taken from the UWT and the TWT for bacterial analysis. Water temperature and UVA irradiance were monitored during the experiments. 2.4. Well water In order to simulate naturally contaminated water and to avoid osmotic stress on the bacteria, natural well-water was used for the experiments. Water was collected from a well situated on the PSA site at a depth of approximately 200 m. A single batch of well water (approximately 100 L) was withdrawn to ensure the same stock of water was used for all the experiments. Table 1 shows the values of water quality parameters of the well water. To preserve the chemical integrity of the well water it was not autoclaved before each experiment. The concentration of naturally occurring organisms was determined by plate count enumeration technique using both LB agar and Endo agar and was found to be less than the detectable limit (DL) of 4 CFU mL−1 . Turbidity measurements were performed using a turbidimeter (model 2100N, Hach, USA). For all experiments, turbidity values between 1 and 2 NTU were obtained. Iron was not present in the water (UV–vis measurements, DL 0.05 mg/L), however, a high concentration of HCO3 − , ∼500 mg/L, was determined (5050A TOC analyser, Shimadzu, Japan). The ions present in the water were analyzed with ion chromatography (Dionex DX-600, USA). This well water has been used in previous solar disinfection research [7,9,10]. 2.5. Bacterial strain and quantification E. coli K12 (ATCC 23631) was generated and grown as described elsewhere [9]. All disinfection experiments were conducted by adding bacterial stock to water in the UWT to obtain an initial concentration of 106 CFU mL−1 . Samples were taken at different time intervals over the 4 or 5 h total experiment time from both the UWT and the TWT. Samples were diluted in PBS (Phosphate Buffer Solution) and enumeration of bacteria was carried out using the standard plate count method. Volumes of 20 L were plated onto LB
3. Results and discussion 3.1. Comparison of SODIS in CPC 1.00 and CPC 1.89 SODIS experiments using the CPC photo-reactor equipped with either C = 1.00 or C = 1.89 were carried out under real sunlight conditions using 2.5 L of well water containing 1 × 106 CFU mL−1 E. coli. The reactor with CPC 1 was exposed to sunlight at 10:30–12:30 local time receiving 229 kJ m−2 of solar UVA; and CPC 1.89 was exposed at 12:00–13:00 to achieve 245 kJ m−2 of UVA dose. Both were covered after exposure to examine post-treatment inactivation in the dark. Samples (10 mL) were taken at regular intervals for bacterial analysis during SODIS treatment and also during the post-exposure period. In the C = 1.00 CPC (Fig. 5(a)) a 3-log kill was observed after 60 min exposure, and complete bacterial inactivation (until detection limit) was achieved after 2 h exposure. Therefore, the total treatment time to achieve a 6-log inactivation was 2 h. For C = 1.89 CPC a 6-log kill was observed after 60 min exposure and during the dark period bacterial regrowth was not detected (Fig. 5(b)). As expected, the total treatment time observed for CPC 1 was halved when the CPC 1.89 was used, i.e. the time required to receive a similar UVA dose in the CPC 1 is almost twice the time needed for CPC 1.89. Therefore, the system with a CPC = 1.89 will permit treatment of double the volume of water in the same time period as compared to that with a CPC = 1. It is thought that photolytic bacterial inactivation proceeds via photon damage followed by subsequent reactions leading to cell death [10]. The sequence of disruption to normal bacterial cell function during solar disinfection has been described by Berney et al. [13]. One of the important effects observed during irradiation of cells is the damage of DNA, where interaction with UV-radiation produces cyclobutane dipyrimidine dimmers preventing mRNA translation and cell reproduction. Bacteria have evolved a number of defence mechanisms and can initiate a complex enzyme system to repair genetic damage [14]. Bohrerova and Linden [15] examined the DNA photo-repair rate of E. coli during exposure to four different fluorescent lamps and natural sunlight. During studies using fluorescent lamps photo-repair was observed, however, they concluded that the initiation of the photo-repair process in E. coli did not take place above a critical level of exposure to solar radiation [15]. The recent contribution of Bosshard et al. [16] showed that the first targets on the way to cell death were found to be the
20
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(a)
CPC = 1.89 7
50
10
Exposure
Dark period
4
10
30
Solar disinfection Control Temperature UV irradiance UVA
3
10
2
10
20
10
1
10
-2
5
10
UVA irradiance (W m ) Temperature (ºC)
40
-1
E. coli K12 (CFU mL )
6
10
DL
0
0
10
12:00
13:00
14:00
15:00
Local time (hh:mm)
CPC = 1.89 7
Exposure
50
Dark period
6
40
-1
E. coli (CFU mL )
10
5
10
30
4
10
Solar disinfection control solar UVA irradiance water temperature
3
10
20
2
10
10 1
10
-2
10
UVA irradiance (W m ) Temperature (ºC)
(b)
DL
0
0
10
12:00
13:00
14:00
15:00
16:00
Local Time (hh:mm) Fig. 5. Inactivation of E. coli in well water during natural sunlight exposure using the sequential batch reactor (a) C = 1.00 (UVA dose = 229 kJ m−2 ); (b) C = 1.89 (UVA dose = 245 kJ m−2 ).
respiratory chain and even the cells’ potential to generate ATP were inhibited. 3.2. Increasing the lethal UVA dose Our previous results [10] demonstrated that “an uninterrupted minimum lethal UVA dose” of 108 kJ m−2 , was necessary to disinfect 2.5 L of well-water polluted with E. coli K-12 (initial concentration ∼106 CFU mL−1 ) in the C = 1.00 solar CPC reactor. Nevertheless, we observed a 3–4-log kill during solar exposure and complete inactivation 2 h after treatment when the reactor was kept in the dark. A similar result was observed in Fig. 6(a), where the CPC 1.89 system received 108 kJ m−2 (corresponding to 35 min of solar exposure). This graph shows a 2-log decrease under illumination with complete disinfection attained following 2 h of dark treatment. Treatment following receipt of the minimum lethal UVA dose therefore resulted in a total batch treatment time of 2 h and 35 min for 2.5 L of water. In order to remove the need for a dark inactivation period, and allow faster batch processing, the solar exposure can be lengthened thereby increasing the UVA dose. The effect of increasing the UVA dose upon the total treatment time required for complete disinfection was investigated in the CPC 1.89 photo-reactor. Complete disinfection (3 × 106 CFU mL−1 to DL) was observed within 1 h of solar exposure without the need of post-exposure dark treatment (Fig. 5(b)) (UVA dose equal to 245 kJ m−2 ). The water temperature in the reactor remained below 35 ◦ C at all times, therefore inactivation of bacteria cannot be attributed thermal effects but
Fig. 6. Inactivation curve of E. coli in well water during natural sunlight exposure using the sequential batch reactor with C = 1.89 and dark post-irradiation effect after deliver 108 kJ m−2 (a) and 245 kJ m−2 (b).
to the synergistic effects of mild heat and UVA light observed during SODIS [17] where the main bacterial photo-inactivation mechanism depends on the generation of reactive oxygen species (ROS) [10]. These results demonstrate that exposure to a UVA dose of approximately double the minimal lethal UVA dose halves the total treatment time required to process 2.5 L in the CPC 1.89 photoreactor. In addition, potential health risks associated with bacterial recovery in the dark are significantly reduced. These findings support our initial results, where following receipt of the minimum uninterrupted lethal UVA dose, the concentration of viable E. coli K-12 cells was decreased to below the detection limit. In addition, bacterial re-growth was not evident at 24 or 48 h following SODIS treatment, indicating that photo-repair mechanisms had not been activated and/or were not effective. 3.3. Sequential batch processing In order to treat water using SODIS in sequential batches, complete disinfection must be observed before the treated water can be dispensed into the treated water tank. If a post-exposure dark inactivation period is required in the photo-reactor, this would significantly increase the total treatment time. The results in Fig. 5(b) confirm that solar exposure corresponding to UVA dose equal to 245 kJ m−2 (received in approximately 1 h in the CPC 1.89 system) is sufficient to ensure complete bacterial inactivation and therefore permit sequential batch processing based upon receipt of that UVA dose.
M.I. Polo-López et al. / Journal of Hazardous Materials 196 (2011) 16–21
This study was carried out in Southern Spain where the average daily UVA irradiation dose is (1180 ± 20) kJ m−2 (yearly average of 2007 to 2010), which would permit treatment of 6 batches of water per day. Standard sunny days in this area have an average UVA irradiance of 30 W m−2 . Therefore, the use of an automated C = 1.89 CPC photo-reactor would permit processing of 6 sequential batches of 2.5 L each day, with the single tube photo-reactor producing 15 L of solar purified water each per day. The sequential batch system is modular and could be scaled up to allow several CPC photoreactors to be used under the control of a single UVA sensor. For example, six C = 1.89 CPC modules could theoretically produce around 90 L of potable water per day, which would be a suitable volume of drinking water for several households. Allowing for maintenance and non-optimal solar conditions, each 6-tube system could produce approximately 31,500 L during a typical year. A preliminary cost-based analysis, using parameters previously described by Clasen et al. [18], indicated that a 6-tube automated sequential batch system, with a predicted life span of ten years, could provide solar disinfected water at a total treatment cost equivalent to $0.23 per 100 L. This compares favourably with commonly used point-of-usewater treatment processes, such as chlorine solutions and P&G PUR® sachets, which have been estimated to cost $0.045 and $1.00 per 100 L respectively [18]. Research is ongoing to further reduce the initial cost of the automated SODIS system through the use of alternative materials for CPC’s and low power electronics in the control apparatus. 4. Conclusions The use of a CPC photo-reactor with a C of 1.89 approximately halves the time taken to acquire the lethal UVA dose, in comparison to a CPC with a C of 1.00. However, a dark inactivation period, following the solar exposure, is required to achieve a 6-log kill. This dark inactivation period introduces uncertainty in relation to the SODIS treatment and increases the total treatment time. Doubling the UVA dose was demonstrated to give a 6-log kill without need for a dark inactivation period, permitting batch treatment in approximately 1 h (under typical solar conditions). The addition of simple, low cost electronic control apparatus to SODIS photoreactors allows sequential processing of batch SODIS. The system described has a number of advantages including: (1) ensuring that double the lethal dose is received; (2) providing feedback to the user during the treatment process (i.e. process not complete); and (3) removing user input with respect to control of the SODIS process. Cost-based analysis of the sequential batch CPC solar disinfection reactor shows that it compares favourably with other point-of-use water purification systems. Acknowledgements This work was funded by the European Union under contract no. FP6-2006-INCO-DEV-031650-SODISWATER and by the Spanish
21
Ministry of Science and Innovation under the Consolider-Ingenio 2010 programme (Project CSD2006-00044 TRAGUA). CN was supported by ANPCyT and BECAS MAE-AECI. References [1] G. Hutton, L. Haller, J. Bartram, Economic and health effects of increasing coverage of low cost household drinking-water supply and sanitation interventions to countries off-track to meet MDG target 10, in: WHO Press, Geneva, Switzerland, 2007. [2] M. Wegelin, S. Canonica, K. Mechsner, T. Fleischmann, F. Pesaro, A. Metzler, Solar water disinfection: scope of the process and analysis of radiation experiments, Aqua: J. Water Supply Res. Technol. 43 (1994) 154–169. ˜ [3] J.A. Byrne, P. Fernandez-Ibanez, P.S.M. Dunlop, D.M.A. Alrousan, J.W.J. Hamilton, Photocatalytic enhancement for solar disinfection of water: a review, Int. J. Photoenergy (2011) 1–12 (art. ID 798051). [4] M. Berney, H.U. Weilenmann, A. Simonetti, T. Egli, Efficacy of solar disinfection of Escherichia coli, Shigella flexneri, Salmonella Typhimurium and Vibrio cholerae, J. Appl. Microbiol. 101 (2006) 828–836. [5] K.G. McGuigan, T.M. Joyce, R.M. Conroy, J.B. Gillespie, M. Elmore-Meegan, Solar disinfection of drinking water contained in transparent plastic bottles: characterising the bacterial inactivation process, J. Appl. Microbiol. 84 (1998) 1138–1148. [6] P. Fernández, J. Blanco, C. Sichel, S. Malato, Water disinfection by solar photocatalysis using compound parabolic collectors, Catal. Today 101 (2005) 345–352. ˜ [7] C. Navntoft, E. Ubomba-Jaswa, K.G. McGuigan, P. Fernández-Ibánez, Effectiveness of solar disinfection using batch reactors with non-imaging aluminium reflectors under real conditions: natural well-water and solar light, J. Photochem. Photobiol. B: Biol. 93 (2008) 155–161. ˜ [8] P. Fernández-Ibánez, C. Sichel, M.I. Polo-López, M. de Cara-García, J.C. Tello, Photocatalytic disinfection of natural well-water contaminated with Fusarium solani using TiO2 slurry in solar CPC photo-reactors, Catal. Today 144 (2009) 62–68. ˜ [9] E. Ubomba-Jaswa, P. Fernández-Ibánez, C. Navntoft, M.I. Polo-López, K.G. McGuigan, Investigating the microbial inactivation efficiency of a 25 L batch solar disinfection (SODIS) reactor enhanced with a compound parabolic collector (CPC) for household use, J. Chem. Technol. Biotechnol. 85 (2010) 1028–1037. ˜ [10] E. Ubomba-Jaswa, C. Navntoft, M.I. Polo-López, P. Fernandez-Ibánez, K.G. McGuigan, Solar disinfection on drinking water (SODIS): an investigation of the effect of UVA dose on inactivation efficiency, Photochem. Photobiol. Sci. 8 (2008) 587–595. [11] M.J. Carvalho, M. Collares Pereira, Truncation of CPC solar collector and its effect on energy collection, Solar Energy 35 (1985) 393–399. [12] A. Rabl, Comparison of solar concentrators, Solar Energy 18 (1976) 93–111. [13] M. Berney, H.U. Weilenmann, T. Egli, Flow-cytometric study of vital cellular functions in Escherichia coli during solar disinfection (SODIS), Microbiology 152 (2006) 1719–1729. [14] P.S. Rajeshwar, D.P. Häder, UV-induced DNA damage and repair: a review, Photochem. Photobiol. Sci. 1 (2002) 225–236. [15] Z. Bohrerova, K.G. Linden, Standardizing photoreactivation: comparison of DNA photorepair rate in Escherichia coli using four different fluorescent lamps, Water Research 41 (2007) 2832–2838. [16] F. Bosshard, M. Bucheli, Y. Meur, T. Egli, The respiratory chain is the cell’s Achilles’ heel during UVA inactivation in Escherichia coli, Microbiology 156 (2010) 2006–2015. [17] K.G. McGuigan, T.M. Joyce, R.M. Conroy, J.B. Gillespie, M. Elmore-Meegan, Solar disinfection of drinking water contained in transparent plastic bottles: characterizing the bacterial inactivation process, J. App. Microbiol. 84 (1998) 1138–1148. [18] T. Clasen, L. Haller, D. Walker, J. Bartram, S. Cairncross, Cost-effectiveness of water quality interventions for preventing diarrhoeal disease in developing countries, J. Water Health 5 (2007) 599–608.
Journal of Hazardous Materials 196 (2011) 22–28
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Treatment of low level radioactive liquid waste containing appreciable concentration of TBP degraded products T.P. Valsala ∗ , M.S. Sonavane, S.G. Kore, N.L. Sonar, Vaishali De, Y. Raghavendra, S. Chattopadyaya, U. Dani, Y. Kulkarni, R.D. Changrani Nuclear Recycle Board, Bhabha Atomic Research Centre, Tarapur 401 502, India
a r t i c l e
i n f o
Article history: Received 12 November 2010 Received in revised form 26 August 2011 Accepted 26 August 2011 Available online 2 September 2011 Keywords: Low level radioactive liquid waste Co-precipitation Vitrification Ion exchange Nickel sulphide Barium sulphate
a b s t r a c t The acidic and alkaline low level radioactive liquid waste (LLW) generated during the concentration of high level radioactive liquid waste (HLW) prior to vitrification and ion exchange treatment of intermediate level radioactive liquid waste (ILW), respectively are decontaminated by chemical co-precipitation before discharge to the environment. LLW stream generated from the ion exchange treatment of ILW contained high concentrations of carbonates, tributyl phosphate (TBP) degraded products and problematic radio nuclides like 106 Ru and 99 Tc. Presence of TBP degraded products was interfering with the co-precipitation process. In view of this a modified chemical treatment scheme was formulated for the treatment of this waste stream. By mixing the acidic LLW and alkaline LLW, the carbonates in the alkaline LLW were destroyed and the TBP degraded products got separated as a layer at the top of the vessel. By making use of the modified co-precipitation process the effluent stream (1–2 Ci/L) became dischargeable to the environment after appropriate dilution. Based on the lab scale studies about 250 m3 of LLW was treated in the plant. The higher activity of the TBP degraded products separated was due to short lived 90 Y isotope. The cement waste product prepared using the TBP degraded product was having good chemical durability and compressive strength. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Classification of waste is very important from safety as well as process consideration point of view. In India apart from IAEA categorization, the radioactive liquid waste is broadly classified into three categories based on the radioactivity concentrations, low (<1 mCi/L), intermediate (1 mCi/L–1 Ci/L) and high level (>1 Ci/L) waste [1]. As a waste management practice, vitrification, ion exchange treatment and chemical co-precipitation are the conditioning processes adopted for high, intermediate and low level waste, respectively with the objective of safeguarding the health of present and future generations [2]. Reprocessing of spent fuel from reactor generates broadly two categories of radioactive liquid waste streams viz high level liquid waste (HLW) and intermediate level radioactive liquid waste (ILW). HLW in acidic conditions is stored in stainless steel tanks. Vitrification process based on sodium borosilicate glass matrix [3,4] has been accepted and adopted for immobilisation of HLW using induction heated metallic melter and joule heated ceramic melter at Trombay and Tarapur, respectively [5]. Prior to vitrification, the HLW is concentrated with respect to salt content and activity.
∗ Corresponding author. E-mail address:
[email protected] (T.P. Valsala). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.065
During the concentration operation of HLW, nitric acid is recovered through fractionator and subsequently the off gases are condensed to collect as condensate stream. The condensates obtained during concentration operation have activity in the range of LLW. Major radionuclides present in this stream are 137 Cs, 90 Sr, 125 Sb and 106 Ru. ILW is generated during U/Pu purification cycles and waste evaporation in PUREX process flow sheet. At Tarapur site this ILW is made alkaline using sodium carbonate and sodium hydroxide prior to storage in carbon steel tanks. This IL radioactive liquid waste contains carbonates and large amount of salts. It also contains radiation degradation products of TBP like Na-dibutyl phosphate, Na-monobutyl phosphate etc. Major radionuclide present in this stream is 137 Cs. Other radio nuclides present are 90 Sr, 106 Ru, 125 Sb, 99 Tc and trace concentration of actinides. A treatment process using resorcinol formaldehyde polycondensate (RF) resin with iminodiacetic chelating group is followed for the effective removal of Cs and Sr from this ILW at Tarapur [6,7]. The alkaline effluent from the RF column is LLW in nature and has high salt content mainly in the form of sodium nitrate. Carbonates and TBP degraded products are getting removed during the pretreatment step prior to ion exchange. The effluent comprises of small amount of 137 Cs, 90 Sr, 125 Sb, 106 Ru and 99 Tc which are not retained by the RF column. LLW generated from the concentration operation of HLW and ion exchange treatment of ILW are managed by chemical coprecipitation [8]. The radio nuclides in the LLW are removed from
T.P. Valsala et al. / Journal of Hazardous Materials 196 (2011) 22–28
23
Table 1 Characteristic of the low level radioactive liquid waste. Sr. No.
Property
LLW from IX process
1 2 3 4
Colour pH TBP degraded products as DBP, g/l Gross  (mCi/L) Gross ␣ (mCi/L)
Yellow 12.5 1.8 0.20 5 × 10−3
5 6 7 8 9 10 11
137
Cs (mCi/L) Cs (mCi/L) Sr-90 Y (mCi/L) 106 Ru (mCi/L) 125 Sb (mCi/L) 99 Tc (mCi/L) 154 Eu (mCi/L)
0.12 5 × 10−4 2 × 10−3 7.5 × 10−3 1.5 × 10−3 5.1 × 10−2 BDL
12 13
Sodium (g/L) Uranium (g/L)
56 1.8
134 90
Condensate LLW from concentration of HLW
Colourless 0.43M acidic BDL 0.43 BDL Isotopic constituents 0.18 8 × 10−3 0.25 4 × 10−2 2.86 × 10−3 BDL 3.04 × 10−4 Elemental analysis 0.10 0.05
MW (1:1 mixing) Pale yellow <1 NA 0.30 2.0 × 10−4 0.14 4.1 × 10−3 0.11 2.1 × 10−2 2.0 × 10−3 2.4 × 10−2 BDL 28 0.85
BDL, below detection limit; NA, not analysed. Minimum detectable conc. for ˇ = 3 × 10−5 mCi/L and for ˛ = 5 × 10−6 mCi/L.
the bulk solution as sludge by co-precipitation or adsorption, using insoluble compounds like hydroxides, sulphate, phosphates and ferrocyanides [9]. Subsequent to precipitation, liquid phase is discharged after providing necessary dilution, in order to meet the regulatory limits for radioactive discharge. The sludge containing bulk of the radioactivity is conditioned by immobilizing in cement matrix before disposal. Fig. 1 shows the schematics for this process. Presently a modified flow sheet has been adopted for treatment of ILW where the pretreatment step is avoided to improve the processing capacity of the plant [10]. In addition to high salt content the effluent LLW from this process have carbonates, degraded TBP products and relatively high concentrations of 106 Ru and 99 Tc. The presence of carbonates and TBP degraded products caused interference in the chemical co-precipitation process during the effective removal of radionuclides. The present study details the formulation of a modified chemical treatment scheme for the LLW, containing carbonates, TBP degraded products and relatively higher concentrations of 106 Ru and 99 Tc. The performance of this modified scheme in the plant scale treatment of about 250 m3 of LLW was also assessed. Conditioning of the separated TBP degraded products in cement matrix was also studied.
technical grade. All experiments were carried out in triplicate and the average values are presented. 2.2. Characterization of the radioactive waste The LLW streams generated from concentration cycle of the HLW and from modified ion exchange treatment process of ILW were characterized for radiochemical nature and isotopic constituents. The values are given in Table 1. The activity spectrum of the waste streams was determined using GM counter (Nucleonix, GM tube 72314) for gross , ZnS detector based scintillation counter (Para Electronics Pvt. Ltd. India) for gross ␣ and HPGe detector coupled with 4K MCA (GMBh Germany) for ␥- spectrum. Na was analysed using flamephotometer (Elico CL361). U and Fe were analysed by ammonium thiocyanate method at wavelengths 360 nm and 565 nm, respectively [11] using spectrophotometer (Chemito, UV2600). The acidic and alkaline LLW streams were mixed to get pH <2 and kept for 5–6 h to separate the TBP degraded products. The characteristics of this mixed waste stream (MW) are also given in Table 1. The MW was further taken up for its treatment using chemical co-precipitation for removal of various radionuclides.
2. Experimental
2.3. Lab scale co-precipitation experiments
2.1. Material
Large numbers of co-precipitation experiments were carried out with the MW. Various combinations and concentrations of the precipitating chemicals were tried. The optimum chemical combination and concentration selected for further studies is given in Table 2. All the experiments were performed on 25 ml batch size.
As this chemical treatment method is to be adopted on plant scale, all the chemicals used for the treatment process were of
Chemical addition
Chemical addition
Alkaline LLW
Table 2 Concentration and sequence of chemicals for 1st and 2nd step treatment.
Acidic LLW S
Sludge for cementation
Chemicals added
S
Decant for discharge
S - MS Settlers Fig. 1. Schematics for chemical co-precipitation of alkaline and acidic LLW.
Initial pH Na2 SO3 Fe2+ as Fe(SO4 )·7H2 O (PO4 )3− as Na3 (PO4 )·12H2 O S2− as Na2 S NaOH to adjust pH to Cu2+ as CuSO4 ·5H2 O [Fe(CN)6 ]4− as K4 Fe(CN)6 ·3H2 O Ba2+ as BaCl2 Fe3+ as Fe(NO3 )3 ·9H2 O NaOH to adjust pH to
Concentration (ppm) 1st step
2nd step
2.0 1260 600 – 300 7–7.5 40 60 400 150 7–7.5
– – 300 400 – – – – 800 100 9.5–10
24
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Table 3 DF obtained in lab scale chemical co-precipitation of MW. Properties
pH Gross  (mCi/L) Gross ␣ (mCi/L) 137 Cs (mCi/L) 106 Ru (mCi/L) 125 Sb (mCi/L) 90 Sr-90 Y (mCi/L) 99 Tc (mCi/L)
Initial values
<1 0.30 2.0 × 10−4 0.14 2.1 × 10−2 2.0 × 10−3 0.11 2.4 × 10−2
With organics
Without organics
1st step
2nd step
7–7.5 1.6 × 10−2 BDL 4.5 × 10−4 8.6 × 10−3 4.2 × 10−4 6.3 × 10−3 1.4 × 10−3
9.5–10 8.5 × 10−3 BDL 1.5 × 10−4 6.8 × 10−3 BDL 1.1 × 10−3 3.2 × 10−4
DF
1st step
2nd step
DF
35 50 1000 7 220 120 80
7–7.5 5.6 × 10−3 BDL 2.6 × 10−4 1.1 × 10−3 BDL 2.8 × 10−3 4.8 × 10−4
9.5–10 1.5 × 10−3 BDL 3.1 × 10−5 6.5 × 10−4 BDL 6.1 × 10−4 1.4 × 10−4
200 50 5000 40 220 215 185
BDL, below detection limit. Minimum detectable conc. for ˇ = 3 × 10−5 mCi/L and for ˛ = 5 × 10−6 mCi/L. Table 4 DF obtained in plant scale chemical co-precipitation of MW (average of 10 batches). Initial activity of MW
Activity and DF after two step chemical treatment 1st step
pH Gross  (mCi/L) Gross ␣ (mCi/L) 137 Cs (mCi/L) 106 Ru (mCi/L) 125 Sb (mCi/L) 90 Sr-90 Y (mCi/L) 99 Tc (mCi/L)
<1 0.38 2.5 × 10−4 0.11 9.2 × 10−2 1.8 × 10−3 0.12 2.6 × 10−2
7–7.5 4.2 × 10−3 BDL 3.3 × 10−4 1.0 × 10−3 BDL 2.1 × 10−3 4.8 × 10−4
DF
2nd step
Overall DF
90 50 333 92 180 57 54
9.5–10 1.6 × 10−3 BDL 5.4 × 10−5 6.6 × 10−4 BDL 5.2 × 10−4 1.2 × 10−4
238 50 2040 140 180 230 216
BDL, below detection limit. Minimum detectable conc. for ˇ = 3 × 10−5 mCi/L and for ˛ = 5 × 10−6 mCi/L.
To see the effect of the TBP degraded products on the coprecipitation of radionuclides from the MW two different sets of experiments were carried out. For the first step chemical precipitation, 25 ml of the clear decant of the MW with out any organic content was taken in a beaker. Similarly in another beaker 25 ml of the MW was taken after mixing the solution thoroughly with separated organic mass. Required concentrations of the chemicals were then added in the sequence as given in Table 2 step-1. During addition of each chemical the content was mixed thoroughly using magnetic stirrer. The pH of the solution was adjusted to 7–7.5 using NaOH solution. After complete settling of the sludge the activity spectrum of the clear decants was determined and the decontamination factor (DF) was calculated using the equation: DF =
Ai Af
(1)
where Ai and Af are the initial and final specific activity (activity per unit volume), respectively. For the second step precipitation, the clear decant from the first step was given chemical treatment as per the scheme given in Table 2 step-2. The pH of the final solution was adjusted to 9.5–10 using NaOH solution. Here also the activity spectrum of the clear decants was determined and the DF was calculated. The DF values achieved in both steps are given in Table 3.
done by recirculation. The decant solution was analysed for activity spectrum and was transferred to another MS settler. There, chemicals for second step treatment as given in Table 2 step 2 were added and allowed to settle overnight. Decant was analysed for activity spectrum and DF was calculated. The details are given in Table 4. The schematic for this MW treatment in plant scale is given in Fig. 2. Further batches of treatment were carried out in the same tanks and settlers without removing the organic mass/sludge from them. This was based on the lab scale multiple batch precipitation studies (Table 5). After about 10 batch treatment the sludge from the settlers was cleared and fresh batches were received. However in the carbonate destruction tank the organic mass was removed only after completing the treatment campaign of the LLW from the particular waste storage tank. 2.5. Analysis of the organic mass separated Solubility of the separated organic mass in acid and alkali medium were tried. The clear decant of the alkali dissolved solution was analysed. The reddish precipitate of ferric hydroxide was dissolved in acid and analysed. The activity spectrum of the organic mass calculated from these analysis are presented in Table 6.
Chemical for 1st step chemical treatment
2.4. Plant scale chemical treatment Based on the lab scale treatment results, about 250 m3 of MW was treated in the plant. The batch size of the plant scale treatments was 12 m3 . Acidic LLW was mixed with alkaline LLW in a SS tank till pH 1–2 was obtained and kept for about 5–6 h. The degraded products of TBP got separated and were floating on the top of the tank. The clear decant below the organic layer was transferred to another SS tank and chemicals such as Na2 SO3 , FeSO4 and Na2 S as given in Table 2 step1 were added and pH was adjusted to 7. The total mass was then transferred to a MS settler and remaining chemicals for first step treatment were added and allowed to settle overnight. During the addition of chemicals thorough mixing was
Alkaline LLW
Acidic LLW
MW
S
Sludge for cementation
Chemicals for 2nd step chemical treatment
S
Decant for discharge
S - MS Settlers Fig. 2. Schematics for plant scale treatment of MW.
T.P. Valsala et al. / Journal of Hazardous Materials 196 (2011) 22–28
25
Table 5 DF obtained during lab scale multiple batch treatment. Chemical Treatment
DF with respect to initial activity Batch I
Ist step SR Ist step CT IInd step SR IInd step CT
9 160
Batch II
Batch III
Batch IV
Batch V
Batch VI
10 40 130 220
12 45 120 230
13 40 115 190
12 37 110 180
11 50 125 200
Initial gross  activity of the waste solution = 0.30 mCi/L. CT, chemical treatment; SR, sludge recycle. Table 6 Analysis of organic solid mass dissolved in Na2 CO3 . Sl.No
Properties of organic mass
1 2 3 4 5 6 7 8 9 10
Gross  (Ci/g) Gross ␣ (Ci/g) 137 Cs (Ci/g) 106 Ru (Ci/g) 125 Sb (Ci/g) 90 Sr (Ci/g) 154 Eu (Ci/g) (PO4 )3− (mg/g) U (mg/g) Fe (mg/g)
Value calculated from Supernatant (1)
Fe(OH)3 residue (2)
Total (1 + 2)
57 0.012 0.693 0.670 0.072 4.401 BDL 240 <10 NA
4.260 0.063 0.282 0.687 0.766 BDL 0.102 NA NA 58
61.260 0.075 0.975 1.357 0.838 4.401 0.102 – – 58
NA, not analysed; BDL, below detection limit. Minimum detectable conc. for ˇ = 3 × 10−5 mCi/L and for ˛ = 5 × 10−6 mCi/L.
The NaOH dissolved organic mass was further cementised to see the feasibility of fixation in cement matrix [12]. For this, ordinary Portland cement and alkaline solution of the organic mass was mixed in 1.3:1 w/w ratio and the cement block formed was cured for a period of 28 days in humid condition. Chemical durability study of the cement block was carried out by conventional leaching method at ambient temperature using distilled water as leachant [13]. The details of the cement block and leaching conditions are presented in Table 7. The leachant was replenished periodically and the leachate was analysed for gross beta activity. Compressive strength of the cement block was also determined using Compression testing machine provided by Lawrence & Mayo (India) Pvt. Ltd. 3. Results and discussion In the new scheme for co-precipitation, mixing of acidic and alkaline LLW was considered to minimise the use of concentrated acid and alkali for the neutralization of these waste streams. It was observed that during carbonate destruction, the degraded products of TBP is getting separated as solid mass and floats on the top of the liquid. Since the waste was containing relatively higher concentrations (0.02–0.03 mCi/L) of 106 Ru and 99 Tc, TBP degraded products
Table 7 Details of cement block and leaching conditions. Details of cement block 21.924 g (1.924 g organic mass) Weight of waste (organic mass dissolved alkali) Cement (∼1.3 time waste) 26 g 2.0 cm Height of cement block Diameter 3.8 cm 47.51 g Weight 46.56 cm2 Surface area (S.A.) Leaching conditions Distilled water Leachant Temperature Ambient 600 ml Volume of leachant ∼0.1 cm−1 S.A./vol. of leachant
and carbonates, normal chemical co-precipitation method [10] followed for the treatment of LLW streams did not provide required DF with respect to gross beta activity. Hence a new scheme with two step chemical treatment as given in the Table 2 was formulated after extensive experimental trials in the laboratory. The precipitation of Ru, Tc and Cs are effective in acidic/neutral pH and that of Sr and Sb in alkaline pH. Therefore a two step chemical treatment was carried out at two different pH conditions. Ru and Tc form various anionic, cationic and neutral species in aqueous medium. Due to the complex chemistry of these ions the mechanism of the reductive co-precipitation of them is not clearly known. However it may be explained as follows. Na2 SO3 is reducing the perruthanate and pertechnate ions (VII) to ruthanate and technate ions (VI) [14]. These ions are then co-precipitated along with FeS and Fe(OH)2 . 2(RuO4 )− + 2[O] → 2(RuO4 )2− + O2 −
2−
2(TcO4 ) + 2[O] → 2(TcO4 )
+ O2
(2) (3)
Cs is precipitated along with potassium copper ferrocyanide and Sr with barium phosphate. Sb is precipitated as hydroxide along with ferric hydroxide. Co-precipitation of 106 Ru, 99 Tc 134,137 Cs, 90 Sr and 125 Sb occurred during first step treatment. In the second step treatment coprecipitation of remaining 106 Ru, 90 Sr and 125 Sb was there (Table 3). Separation of the TBP degraded products and reductive coprecipitation of 106 Ru and 99 Tc at acidic pH improved the DF values. However the DF for Ru is low compared to other nuclides. This may be due the inefficient co-precipitation of Ru ions because of its complex chemistry in aqueous medium. The effluent after the two step chemical treatment is dischargeable to environment. Table 3 shows that the DF obtained is high for MW with out TBP degraded product compared to that with TBP degraded products. The TBP degraded products are soluble in alkaline medium and hence interferes with the co-precipitation process. Table 4 shows the average DF obtained in ten batches of plant scale treatment of MW. The values are comparable or better than
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T.P. Valsala et al. / Journal of Hazardous Materials 196 (2011) 22–28
Na2 CO3 solution resulted in a clear solution. Here also reddish precipitate was observed. In alkali medium the TBP degraded products namely DBP, MBP and butyl alcohol form respective soluble sodium salts. (C3 H7 )2 HPO4 + NaOH → (C3 H7 )2 NaPO4 + H2 O
(4)
(C3 H7 )H2 PO4 + 2NaOH → (C3 H7 )Na2 PO4 + 2H2 O
(5)
(C3 H7 )OH + NaOH → (C3 H7 )ONa + H2 O
(6)
Table 6 shows the activity spectrum of the supernatant and residue of Na2 CO3 dissolved organic mass. Fig. 4 shows the gamma ray spectra of the supernatant and residue. From Table 6 it can be seen that there is appreciable concentration of 154 Eu in the organic mass whereas in the MW solution its concentration was below detection limit (Table 1). This may be due to the extraction of trace levels of 154 Eu by the organic mass. The source of 154 Eu is the acidic LLW. The gross  activity of the sample is much more than the sum total of all individual isotope activities (Table 6). This shows that there are unknown isotopes. The unknown isotopes have to be pure  emitters as there is no unknown line in the ␥ spectrum. It was also observed that the gross beta activity of the organic solid mass was fast decreasing with time. The fast decrease in beta activity shows that the pure beta emitting isotope is short-lived. To identify the unknown isotope, a fresh sample of organic mass was dissolved in sodium carbonate and its activity trend with respect time was followed (Fig. 5). (Activity of the unknown isotope was taken as ‘gross beta activity – sum total of activity of other isotopes’.) The initial activity of the unknown isotope was 110 mCi/g. After 28 days the activity got decreased to 5.7 mCi/g. Considering the presence of 90 Sr in the waste solution, the short lived beta isotope may be 90 Y (T1/2 = 64 h). The activity decay of an isotope is represented by the equation:
Fig. 3. Organic layer separated at the tope of the SS tank in the plant.
the lab scale values. The better value may be the advantage of large scale treatment. During multiple batch treatment, the sludge obtained in the previous batch operation was giving advantage to the next batch operation (Table 5). This is due to the ion exchange property of the precipitate formed for the uptake of different radionuclides. At the end of treatment of ten batches of MW, a thick layer of solid organic mass was observed at the top of the carbonate destruction SS tank (Fig. 3). The total organic mass separated was about 0.2–0.3% of the total waste treated. Though the volume of the organic mass is very less, treatment of the same is of concern due to its organic nature. In view of this detailed analysis of the organic mass was carried out. The separated solid mass was off-white in colour and was fluffy in nature. Dissolution of this mass in boiling concentrated nitric acid resulted in hazy solution with finely distributed fine particle. Perchloric acid digestion also gave a hazy solution. In both these cases the organic part is only partially getting destroyed. With 1 M NaOH solution the organic part was almost soluble and there was some reddish precipitate of Fe (OH)3 at the bottom, which was soluble in dilute nitric acid. Dissolution of the organic mass with 1 M
A = A0 et =
(7)
0.693 T1/2
(8)
where A0 is the initial activity, A is the activity after a time period of t, is the decay constant and T1/2 is the half life of the radioactive isotope. 9000 b
7000 6000
c a :125Sb 106 b : Ru 137 c : Cs
b
8000
152
a : Eu 125 b : Sb 106 c : Ru 137 d : Cs
7000 b 6000
Counts
Counts
5000 4000 3000
c b
5000 b
b 4000
d
3000
b
2000
2000 1000
a a a
a 0 200
c
a
1000
a
a
a
a
0
300
400
500
600
700
100
200
300
400
Channel
500
600
700
800
900
Channel
Gamma Spectrum of dissolved portion of organic mass in Na2CO3
Gamma Spectrum of Undissolved Residue of organic mass
Fig. 4. Gamma ray spectra of organic mass dissolved in Na2 CO3 .
a aa 1000 1100
T.P. Valsala et al. / Journal of Hazardous Materials 196 (2011) 22–28
120
27
2.5
0.01
Leach Rate Cumulative fraction leached
Leach Rate (gm/cm /cay)
Acivity, micro Ci/g
2
80
60
40
2.0 1E-3
1.5 1E-4
20
1.0
0
0 0
5
10
15
20
25
20
40
60
80
Cumulative fraction leached (%)
100
100
Time period, days
30
Time, days
Fig. 7. Leaching behaviour of the cement product.
Fig. 5. Decrease in  activity of the organic mass with time.
the solid mass is also well with in the permissible range for NSDF disposal.
The decay pattern of the unknown beta isotope was tested by plotting the logarithm of activity of unknown isotope against time of decay (Fig. 6). The T1/2 of the unknown beta isotope calculated from the slope of the straight line in Fig. 5 (70 h) matches with the T1/2 of 90 Y with in the error limits and hence confirms the presence of 90 Y. This shows that yttrium is behaving similar to 154 Eu and is getting extracted from the waste solution in to the degraded TBP products. The high beta activity of the organic mass was mainly due to the short lived isotope 90 Y present in it. After about one month time the activity got decreased to about twenty times less. Fig. 7 shows the leaching behaviour of the cementised product of organic mass. As seen from Fig. 6, the stabilized leach rate with respect to gross beta for the cementised waste product after a period of 100 days is 4.80 ×10−5 g cm−2 day−1 and the cumulative activity leached from the cement block is around 2.4%, which is below the internationally accepted leach rate for cementised waste product [5,15]. The compressive strength of the cement block (15–20 M Pa) was also with in the acceptable limit of cement waste products [16]. Hence fixation of organic mass in cement matrix can be adopted for its conditioning and shallow land disposal. The alpha burden of
4. Conclusion In the new scheme for chemical treatment of low level effluent stream, separation of TBP degraded products provided very good DF with respect gross beta activity. The reductive co-precipitation of 106 Ru and 99 Tc resulted in effective removal of these radio nuclides to get sufficient decontamination of the effluent stream for its discharge into environment. Based on the lab scale treatment results of the new chemical treatment scheme, about 250 m3 of LLW was successfully treated in the plant. As the leach rate data and compressive strength of the cement product are with in the internationally accepted value, fixation of organic mass in cement matrix can be adopted for its conditioning and shallow land disposal. Acknowledgements Authors are thankful to Shri. Kanwar Raj, Head, Waste Management Division BARC Trombay and Shri. S.D. Mishra, Director, Nuclear Recycle group, BARC Trombay for their guidance and interest in the work.
100
References Slope = -0.10349 R = 0.99896
Activity, µCi/g
10
1
0
5
10
15
20
25
Decay time, days Fig. 6. Plot of  activity (log scale) versus decay time.
30
[1] IAEA Technical Reports Series no. 101, IAEA, Vienna, Austria, 1970. [2] IAEA Safety Series 111- F, 1995. [3] R.G. Yeotikar, M.S. Sonavane, J.G. Shah, R. Kanwar, Development of vitrified matrix for high level waste and its characterization – experience at WIP, Tarapur, SMART-93, 1993. [4] R. Kanwar, M.T. Samuel, Modified pot glass process for vitrification of high level radioactive waste – process engineering aspects, in: XIV International Congress on Glass, vol. II, 1986, pp. 399–409. [5] K. Raj, K.K. Prasad, N.K. Bansal, Nucl. Eng. Des. 236 (2006). [6] Y. Kulkarni, S.K. Samanta, S.Y. Bakre, K. Raj, M.S. Kumra, Process for treatment of intermediate level radioactive waste based on radionuclides separation, WM96, 1996. [7] R.G. Yeotikar, C.P. Kaushik, J. Gabriel, R. Kanwar, Treatment of alkaline intermediate level radioactive waste, NUCAR 95 (1995). [8] IAEA, Tech. Report. Series No. 222 (1983) b. [9] L. Berak, E. Uher, M. Marhol, At. Energy Rev. 13 (1975). [10] P.D. Ozarde, S.K. Samanta, K. Raj, Proceedings of the IAEA international conference on issues and trends in radioactive waste management – IAEA-CN-90/51, Vienna, Austria, 2002. [11] G.H. Jeffery, J. Bassett, J. Mendham, R.C. Denney, Vogel’s Text Book of Quantitative Chemical Analysis, sixth ed., p. 452. [12] U.S. Singh, A. Mishra, R.G. Yeotikar, R. Kanwar, Immobilisation of intermediate level alkaline radioactive liquid waste in cement matrix, NUCAR 95 (1995).
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[13] United State Department of Energy, Nuclear Waste Management Material Hand book (test methods), Rep. DOE/TIC-11400 (1981). [14] F.A. Cotton, G. Wilkinson, Advanced inorganic chemistry. A Comprehensive Text, Third ed. [15] N.K. Bansal, S. Kumar, C.P. Kaushik, R.R. Rakesh, S.M. Galande, Safety assessment of radioactive waste packages for disposal in near surface disposal facilities,
long term behaviour of low and intermediate level waste packages under repository conditions, IAEA-TECDOC-1397. pp. 101–118 (2004). [16] C. Verghese, A.K. Govindan, P.K. Wattal, T.K. Theyyunni, Cementitious matrices for intermediate level aqueous radioactive wastes BARC report BARC/1995/E/004 (1995).
Journal of Hazardous Materials 196 (2011) 29–35
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Recyclable magnetic photocatalysts of Fe2+ /TiO2 hierarchical architecture with effective removal of Cr(VI) under UV light from water S.C. Xu, Y.X. Zhang, S.S. Pan, H.L. Ding, G.H. Li ∗ Key Laboratory of Materials Physics, Anhui Key Lab of Nanomaterials and Nanostructure, Institute of Solid State Physics, Chinese Academy of Sciences, Hefei 230031, PR China
a r t i c l e
i n f o
Article history: Received 1 July 2011 Received in revised form 19 August 2011 Accepted 26 August 2011 Available online 2 September 2011 Keywords: Magnetic Photocatalyst Hierarchical architecture Fe doped TiO2 tubes Removal of Cr(VI)
a b s t r a c t We report the synthesis and photocatalytic removal of Cr(VI) from water of hierarchical micro/nanostructured Fe2+ /TiO2 tubes. The TiO2 tubes fabricated by a facile solvothermal approach show a three-level hierarchical architecture assembled from dense nanosheets nearly vertically standing on the surface of TiO2 microtube. The nanosheets with a thickness of about 20 nm are composed of numerous TiO2 nanocrystals with size in the range of 15–20 nm. Ferrous ions are doped into the hierarchical architecture by a reduction route. The Fe2+ /TiO2 catalyst demonstrates an effective removal of Cr(VI) from water under UV light and the removal effectiveness reaches 99.3% at the initial Cr(VI) concentration of 10 mg L−1 . The ferrous ion in the catalyst serves not as the photo-electron trap but as an intermedium of a two-step reduction. The TiO2 photoreduces the Fe2+ ions to Fe atoms firstly, then the Fe atoms reduce the Cr(VI) to Cr(III), and the later is removed by adsorption. The hierarchical architecture of the catalyst serves as a reactor for the photocatalytic reaction of Cr(VI) ions and an effective absorbent for the removal of Cr(III) ions. The catalyst can be easily magnetically separated from the wastewater after photocatalytic reaction and recycled after acid treatment. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Many toxic heavy metals, such as Cr6+ , Ni2+ , Cu2+ , Cd2+ and Pb2+ , have been discharged into the environment as industrial wastes, causing serious soil and water pollution [1–4]. Among them, the Cr(VI) is more dangerous due to its highly toxic, extremely soluble and mobile, especially, it is carcinogenic and potentially mutagenic to human being [5,6]. Different techniques have been reported in literatures for the treatments of Cr(VI), such as electrocoagulation [7,8], physical and biological adsorption [9–11], ion exchange [12], membrane separation [13] and photocatalytic reduction [14,15]. The most advisable catalyst for the removal of Cr(VI) from the wastewater should be not only capable of removing the Cr(VI) thoroughly, but also easily separated from the wastewater and economically recyclable. As an important semiconductor, TiO2 has been extensively investigated for degrading organic pollutions and removing heavy metal ions from water due to its high photocatalytic activity, chemical/photocorrosion stability, low cost and safety to environment [16–19]. Since the photocatalytic reaction and adsorption easily occurs on or around material surfaces, enlarging the surface area has been proved to be an effective route to realize high performance photocatalysis. It was found that the hierarchical structures
with hollow interior and porous surface exhibit an improved photocatalytic activity due to their large surface area and special micro/nanostructures, especially for those pollutants tending to be absorbed by the catalyst [20–23]. Doping with nonmetal atoms, such as N, C, S, and F, can improve the photocatalytic activity of TiO2 by narrowing its forbidden band gap, upon which enhance the absorption of visible light and generation of photoelectrons [24–27]. Recently, the research activities to improve the photocatalytic activity of TiO2 by doping with transition metals, such as Ag, Fe and Pt, have been flourished. The doped transition metal ions can serve as the photo-electron trap and thus facilitate the separation of electrons and holes and inhibit their recombination reactions [28–30]. In this paper, we report a strategy to synthesize Fe2+ ions doped TiO2 catalyst with a three-level hierarchical architecture, the catalyst demonstrates an effective photocatalytic removal of Cr(VI) from water under UV light. Different from previous reports, our study indicates that the ferrous ion in the catalyst serves not as the photo-electron trap but as an intermedium of a two-step reduction. We also found that the catalyst can be magnetically separated and recycled conveniently and economically. 2. Experimental 2.1. Preparation of samples
∗ Corresponding author. Tel.: +86 0551 559 1437; fax: +86 0551 559 1434. E-mail address:
[email protected] (G.H. Li). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.068
All the chemicals were of analytic grade and used without further purification. In a typical procedure, an emulsion
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containing TiOSO4 ·2H2 O (1.8013 g), glycerol (12.1 mL), and methanol (27.9 mL) was first mixed and stirred for 5 min, and then ethyl ether (8.6 mL) was added to the emulsion and stirred for another 2 min. The emulsion was then moved into a 70 mL autoclave and kept at 110 ◦ C for 24 h, followed by natural cooling to room temperature. Finally, the precipitate was taken out from the autoclave, stirred, ultrasonic dispersed, centrifuged and washed thoroughly with alcohol, dried at 80 ◦ C and calcined at 550 ◦ C in flow air for 3 h. 2.2. Characterization The final products were characterized by X-ray diffraction (XRD, X’Pert Pro MPD), X-ray photoelectron spectroscopy (XPS, Thermo ESCALAB 250), field emission scanning electron microscopy (FESEM, Sirion 200) and high resolution electron microscopy (HRTEM, JEM 2010). Low-temperature nitrogen sorption–desorption isotherms were measured at −195.7 ◦ C using a gas adsorption apparatus (model: Omnisorp 100CX). Specific surface area was evaluated using the Brunauer–Emmett–Teller (BET) equation and the plot of the pore-diameter distribution was determined by using the Barrett–Joyner–Halenda (BJH) method from the desorption branch of the isotherm. The pH values were measured with Mettler Toledo pH meter (FG2/EL2). The ions concentrations were measured by ICP emission spectrometer (ICP 6000). The photocatalytic reactions were carried out under Xe lamp (XBO 500 W/H OFR VS1). 2.3. Reduction of Fe onto TiO2 tubes In a typical experiment, as-prepared TiO2 tubes (80 mg) were immersed in a 40 mL glass bottle fill with FeCl2 ·4H2 O solution (20 mL, 8 mM), ultrasonic dispersed for 1 min, and then ice bathed and stirred for 10 min. Then NaBH4 (5 mL, 100 mM) was dropped in the solution in 5 min and kept stirring for 30 min. Initially, the solution turned primrose yellow from milk white, and then to silver gray. Finally, the Fe nanoparticles doped TiO2 tubes were centrifuged, washed thoroughly with deionized water and dried at 70 ◦ C. The as-prepared sample turns to dark green due to the oxidation of Fe nanoparticles. For small Fe nanoparticles, the resulted Fe oxide is mainly FeO with some Fe2 O3 , while for large ones the pure Fe core might still exist inside Fe oxides. The corresponding sample was named as Fe2+ /TiO2 catalyst. For comparison, some of the Fe2+ /TiO2 catalyst were annealed in flow air at 500 ◦ C for 60 min to transform all the FeO and Fe to Fe2 O3 (the sample has a red brown color and is named as Fe3+ /TiO2 catalyst. Accordingly, the pure TiO2 tubes are named as TiO2 catalyst for simplicity). 2.4. Removal of Cr(VI) A series of experiments was performed to investigate the removal capacity of each catalyst under different illumination conditions. The TiO2 , Fe2+ /TiO2 and Fe3+ /TiO2 catalysts (each 5 mg) was respectively added in nine quartz glass tubes containing 20 mL, 20 mg L−1 K2 Cr2 O7 stock solution. The suspensions were ultrasonic dispersed for 1 min, then agitated tenderly and kept in different illuminated conditions (in darkness, illuminated with visible light (>400 nm by a filter) or ultraviolet light (305–387 nm)) for 1 h and then immediately centrifuged. The catalysts were separated from the suspension and the remaining clear liquid was used for ICP measurements. The influence of the pH value on the removal of Cr(VI) was evaluated under UV light (300 W, 1 h) on the suspension samples containing 8 mg Fe2+ /TiO2 catalyst and 8 mL K2 Cr2 O7 stock solutions (10 mg L−1 ) with different pH values. After ultrasonic dispersed for 1 min, the initial pH values of suspensions were adjusted
by either 5% H2 SO4 or 5% NaOH solution. After photocatalytic reaction the suspensions were centrifuged, and the supernatant liquid was divided into two parts. One was examined directly by ICP and another was adjusted the pH values in the range of 7.5–8 to allow of the transformation of reduced Cr(III) to Cr(OH)3 precipitation before ICP examination [31,32]. The ICP result of the first part represents the total Cr concentration remained in the supernatant liquid and that of the second part is approximately to the concentration of the remained Cr(VI) after photocatalytic reaction. Kinetics of the photocatalytic removal of Cr(VI) was studied with Fe2+ /TiO2 catalyst in a sample containing 25 mg Fe2+ /TiO2 catalyst and 25 mL K2 Cr2 O7 stock solutions (20 mg L−1 ). After ultrasonic dispersed for 1 min, the initial pH value of the suspension was adjusted to 3.5 with 5% H2 SO4 solution. The sample was illuminated with UV light (300 W) with agitating tenderly. Every 3 mL suspension was taken out from the sample after 1, 5, 30, 60, 120, 180, and 240 min time-interval illumination, and then centrifuged immediately for ICP measurement without further treatment. The dependence of the removal capacity on Cr(VI) concentration was evaluated with Fe2+ /TiO2 catalyst under UV illumination (300 W). Six samples contained 8 mg Fe2+ /TiO2 catalyst and 8 mL Cr(VI) ions stock solutions with the concentration of 10, 20, 40, 80, 160 and 200 ppm were prepared, and after ultrasonic dispersed for 1 min, the initial pH value of the suspensions was adjusted to 3.5. After illuminated for 240 min, the suspensions were centrifuged immediately for ICP measurements. The recycle ability of the catalyst was test. The Fe2+ /TiO2 catalyst after photocatalytic reaction was soaked in 10% HNO3 solution for 1 h, then washed with deionized water several times and dried at 70 ◦ C. The reborn Fe2+ /TiO2 catalyst was tested in fresh Cr(VI) solution under the same experimental conditions as mentioned in the section of kinetics study.
3. Results and discussion 3.1. Morphology and structure Fig. 1 shows typical FESEM images and XRD pattern of hierarchical micro/nanostructured TiO2 tubes. The hierarchical architecture is composed of numerous TiO2 nanosheets on the surface of microtubes. The microtubes are about 3–10 m in length, 800–1300 nm in diameter and 200 nm in wall thickness (Fig. 1a and b). A close-up view of Fig. 1c shows that the nanosheets nearly vertically stand on the tubular surface with random distribution. The corresponding XRD analysis (Fig. 1d) reveals that the annealed samples are composed of anatase TiO2 . The three-level hierarchical structure of the TiO2 tubes can be clearly seen in Fig. 2. The TiO2 tube is the main body of the threelevel structure, and its hollow feature can be clearly observed from the different contrasts of the dark edge and pale center (Fig. 2a). The nanosheets formed the second level (Figs. 2b) are composed of numerous TiO2 nanoparticles. The nanoparticles (the third level) with the diameter of 15–20 nm have a special crystal shape of clearcut edges and corners. HRTEM characterizations indicate that the nanoparticles are anatase TiO2 single crystal (see Fig. A1). The TiO2 nanocrystals are connected with each other in nearly a single layer and form a sheet-like structure of about 20 nm in thickness and 250 nm in diameter. The nanosheets stand freely on the surface of the microtube (Fig. 2a) and the nanocrystals consisting of the nanosheet are loosely connected with each other (Fig. 2b), producing numerous mesopores and nanopores in the hierarchical architecture. Such three-level hierarchical structure shows no conspicuous change after doping with Fe nanoparticles, and the Fe (or its oxides) nanoparticles with about several nanometers in diameter situate inside the TiO2 matrix (mainly on the surface of TiO2
S.C. Xu et al. / Journal of Hazardous Materials 196 (2011) 29–35
31
Fig. 1. ESEM images (a–c) and XRD pattern (d) of the TiO2 tubes.
Fig. 2. Typical TEM image of (a) a single hierarchical micro/nanostructured TiO2 tube, (b) nanosheets composed of TiO2 nanocrystals, (c) Fe nanoparticles (inside the circles) doped nanosheet.
nanoparticles), as shown in Fig. 2c. It was found that the reduction step is essential in controlling the size and distribution of Fe nanoparticles inside the TiO2 matrix, and the more generally direct calcination of the TiO2 /FeCl2 mixture results in a large size distribution of Fe oxides. Low-temperature nitrogen sorption–desorption isotherms measurement (Fig. 3) shows that such three-level hierarchical structured TiO2 tubes have a specific surface area of about 140.3 m2 g−1 (evaluated by the Brunauer–Emmett–Teller equation [33]). One can see from the pore size distribution curve (the inset in Fig. 3, calculated from desorption branch of the nitrogen isotherm by the BJH method [34]) that the nanopores have a size distribution from 5.4 to 73 nm with maximum at about 12 nm. The small pore is attributed to the interspace of the nanocrystals and the large one comes from the overlapping of the nanosheets.
Fig. 3. Nitrogen sorption–desorption isotherms and pore size distribution (inset) of the TiO2 tubes.
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Fig. 5. Sketch illustration of the mechanism of photocatalytic removal of Cr(VI) for the Fe2+ /TiO2 catalyst.
Fig. 4. Comparison of removal Cr(VI) capacities of TiO2 , Fe2+ /TiO2 and Fe3+ /TiO2 catalysts in dark and illuminated with visible or UV light for 1 h with the initial Cr(VI) concentration of 20 mg L−1 .
3.2. Mechanism of Fe2+ /TiO2 catalyst for removal of Cr(VI) under UV light The removal capacities of the TiO2 , Fe2+ /TiO2 and Fe3+ /TiO2 catalysts under different illumination conditions are shown in Fig. 4. Without illumination, the TiO2 catalyst shows a slightly preferable removal of Cr(VI) compared with Fe2+ /TiO2 and Fe3+ /TiO2 catalysts, which is considered mainly due to its surface adsorption. As the TiO2 catalyst is synthesized from TiOSO4 , the existed sulfur (as proved by the XPS analysis, see Fig. B1) will increase the surface acidity of TiO2 catalyst in aqueous and make the surface more positively charged, and thus enhancing the adsorption for Cr(VI) in the form of HCrO4 − , CrO4 2− or Cr2 O7 2− [35]. Under visible light illumination, the number of the photoelectron is very limited and cannot enhance the removal capacity but reduce the adsorption as the photoelectrons will neutralize some positive charges on the catalyst surface. When illuminated with UV light, the Cr(VI) will be reduced by enough photoelectrons and the removal capacity will increase accordingly. For Fe2+ /TiO2 and Fe3+ /TiO2 catalysts, the adsorption capacities in dark or under visible light are both lower than the TiO2 catalyst because of the weakened surface acidity by Fe ions (the initial pH values of the solutions are 5.36, 6.53 and 6.05 for TiO2 , Fe2+ /TiO2 and Fe3+ /TiO2 catalysts, respectively). Interestingly, ferrous ions and ferric ions show different roles for removal of Cr(VI) under UV light. Fe2+ /TiO2 catalyst shows an obvious improvement in removal capacity while Fe3+ /TiO2 catalyst show a decreased removal capacity. The mechanisms for removal of Cr(VI) under UV light can be understood by the following equations: − + h+ TiO2 + hv → eCB VB
(1)
− eCB
(2)
+ Cr(VI) → Cr(III)ads
− Fe2+ + eCB → Fe0
(3)
Fe0 + Cr(VI) → Fe2+ /Fe3+ + Cr(III)ads
(4)
− → Fe2+ Fe3+ + eCB
(5)
The photoelectrons generated from TiO2 upon absorbing UV light can reduce Cr(VI) to Cr(III), and the Cr(III) can be removed by adsorbing on the catalyst surface according to Eqs. (1) and (2). For Fe2+ /TiO2 catalyst, the Fe2+ ions will be firstly reduced to Fe0 atoms by the photoelectrons according to Eq. (3). The Fe0 atoms have better reducibility than Fe2+ ions and can easily reduce Cr(VI) to Cr(III) according to Eq. (4). While for Fe3+ /TiO2 catalyst, the Fe3+ ions will capture photoelectrons and transform to Fe2+ ions, leading to the obvious decrease in the removal capacity according to Eq. (5). Here, the hierarchical architecture of the catalyst serves not
only as a reactor for the photocatalytic reaction of Cr(VI) ions but as an effective absorbent for the removal of Cr(III) ions as well. Fig. 5 illustrates the sketch mechanism of the photocatalytic removal of Cr(VI) ions in aqueous for the Fe2+ /TiO2 catalyst. Different from the previous reports, the ferrous ions in our catalyst serve not as the photo-electron trap but as an intermedium of a twostep reduction. The TiO2 photoreduces the Fe2+ ions to Fe atoms firstly, and then the Fe atoms reduce the Cr(VI) to Cr(III). The XPS analysis further confirms our suggestion (Fig. 6). Before the photocatalytic reaction, the high-resolution spectrum shows that the binding energies of Fe 2p are at about 710.9 and 724.5 eV, corresponding respectively to Fe 2p3/2 and Fe 2p1/2 , as shown in Fig. 6a, indicating that the Fe oxide is mostly composed of Fe2+ ions. After the photocatalytic reaction, the binding energies of Fe 2p respectively shift to 713.9 and 726.5 eV, indicating that the main Fe oxide is Fe3+ ions. The binding energies of Cr 2p adsorbed on the Fe2+ /TiO2 catalyst are at about 577.6 and 586.8 eV (Fig. 6b), which are assigned respectively to Cr 2p3/2 and Cr 2p1/2 , indicating that the adsorbed Cr ions are trivalent Cr. Because the reduction of Cr(VI) occurs mainly on the surface Fe nanoparticles, the original pure Fe core in some large Fe oxides does not affect above mentioned two-step reduction mechanism. 3.3. Influence of the pH value on photocatalytic activity Fig. 7 shows the influence of the pH value on the photocatalytic activity. One can see that the Fe2+ /TiO2 catalyst shows an excellent reducibility in the pH range from 1.85 to 6.5. At a lower pH value of 1.85, the concentration of the remained Cr(VI) is less than 0.092 ppm and the concentration of total Cr ions reaches 9.58 ppm, suggesting that most of the Cr(VI) ions are reduced to Cr(III) ions, but the adsorption capacity of the Fe2+ /TiO2 catalyst for Cr(III) at this pH value is very low. With increasing pH value, the reducibility of Fe2+ /TiO2 catalyst further increases (the concentration of the remained Cr(VI) is about 0.057 ppm at pH 3.5) and the adsorption capacity of Cr(III) increases dramatically (the concentration
Fig. 6. XPS spectra of (a) Fe 2p in the Fe2+ /TiO2 catalyst before (curve 1) and after (curve 2) photocatalytic reaction and (b) Cr 2p absorbed on the Fe2+ /TiO2 catalyst after photocatalytic reaction.
S.C. Xu et al. / Journal of Hazardous Materials 196 (2011) 29–35
Fig. 7. Concentration of Cr(VI), total Cr and Fe ions remained in the photocatalytic reaction solution after 1 h illumination with 8 mg Fe2+ /TiO2 catalyst for 8 mL Cr(VI) solution at different initial pH values (at initial concentration of 10 mg L−1 ).
of total Cr ions is about 1.088 ppm at pH 3.5). As the pH value further increases from 3.8 to 6.5, the concentration of Cr(VI) and total Cr ions increases slightly. As the photocatalytic reduction depends on the Fe0 atoms reduced by photoelectrons generated from TiO2 under UV light, the pH value has little influence on the reducibility of the Fe2+ /TiO2 catalyst. Whereas the adsorption of Cr(III) depends strongly on the pH value, the lower the pH value, the heavier the positively charged of the catalyst surface, and the lower the adsorption ability of Cr(III) due to the charge repel. The concentration of the Fe ions dissolved in the solution after the photocatalytic reaction is also depends on strongly on the pH value of the solution, as shown in Fig. 7. The concentration of Fe ions drops to 10.2 from 98.7 ppm as the pH value increases from 1.85 to 3.5. When the pH value is larger than 4.65 the concentration of Fe ions keeps at a nearly constant value of 0.154 ppm, indicating that the consumption of Fe is low at relative high pH value. At lower pH value, Fe atoms will apt to react with H+ and dissolve into the solution, leading to the lost of the loaded Fe in the catalyst. It was found that the optimal pH values of the Fe2+ /TiO2 catalyst for removal of Cr(VI) are in the range of 3.5–3.8. 3.4. Kinetics of the photocatalytic removal of Cr(VI) Time evolution of the photocatalytic reduction of Cr(VI) by the Fe2+ /TiO2 catalyst is characterized by optical absorption measurements, as shown in Fig. 8. At the beginning of the photocatalytic reaction, the concentration of Cr(VI) decreases rapidly and remarkably due to the high adsorption induced by the high specific surface area of the catalyst. The intensity of the absorption peaks of the Cr(VI) gradually decreases with increasing illumination time, and
Fig. 8. Time dependent absorption spectra of Cr(VI) solution with initial concentration of 20 mg L−1 after photocatalytic reaction with 25 mg Fe2+ /TiO2 catalyst for 25 mL Cr(VI) solution. The inset shows corresponding concentrations of Cr(VI) at different reaction times.
33
Fig. 9. Pseudo-first-order kinetic rate plot for removal of Cr(VI) by the Fe2+ /TiO2 catalyst.
almost disappears after 240 min. The final concentration of the total Cr ions remained in the solution drops to 2.03 ppm from the initial 20 ppm (the inset of Fig. 8), indicating the Fe2+ /TiO2 catalyst can effectively remove Cr(VI). The kinetics of photocatalytic reduction for Cr(VI) of the Fe2+ /TiO2 catalyst exhibits a pseudo-first-order model as shown in Fig. 9 (except for the first minute due to the reason mentioned above). The change of ln(Ct /C0 ) with respect to the reaction time has a linear form and obeys the following equation: ln
C t
C0
= A + Kobs × t
(6)
where C0 and Ct are the concentration of Cr (VI) at initial and at different illumination times, respectively, A is a experimental constant associated with the initial adsorption of the Fe2+ /TiO2 catalyst, and Kobs is the observed pseudo-first-order rate constant representing the photocatalytic reduction rate and can be obtained from the slope of the linear plot of ln(Ct /C0 ) vs. time. In our experiments, the absolute value of A and Kobs are found to be about 0.4291 and 0.0069 min−1 , respectively. 3.5. Removal effectiveness and capacity on the initial concentration of Cr(VI) The dependence of the removal effectiveness and capacity on the initial concentration of Cr(VI) was evaluated at pH 3.5 after 4 h illuminations. The removal capacity is determined by the following equation: Qe =
mr (C0 − Ct ) × V = m m
(7)
where Qe is the amount of Cr(VI) ions reduced and removed by a unit mass Fe2+ /TiO2 catalyst (expressed in mg g−1 ), mr the mass of Cr(VI) ions reduced and removed (mg), m the catalyst mass (g), C0 the initial concentration of Cr(VI) ions (ppm or mg L−1 ), Ct the concentration of Cr(VI) (ppm or mg L−1 ) after photocatalytic reaction and V the volume of the solution from which the photocatalytic reaction occurs (L). The removal effectiveness is evaluated by the ratio of the mass of Cr(VI) ions removed over its initial mass. The results are shown in Fig. 10. The removal effectiveness is very prominent at low C0 (with 99.3% for initial Cr(VI) concentration of 10 ppm and 89.0% of 20 ppm), then it gradually decreases with increasing C0 , and at a high C0 , such as at 80 ppm, the removal effectiveness becomes poor. While the removal capacity increases with increasing C0 , and almost has a constant value at high C0 . The reason for this is considered mainly due to the limited generation of Fe0 atoms by the photoelectron, as our further experiment conformed that the removal effectiveness and capacity can be enhanced by extending the illumination time, for example, the removal effectiveness and capacity increases respectively from 25.78 to 34.36%
34
S.C. Xu et al. / Journal of Hazardous Materials 196 (2011) 29–35
4. Conclusion
Fig. 10. Removal effectiveness and capacities after 4 h photocatalytic reaction at different initial Cr(VI) concentrations with 8 mg Fe2+ /TiO2 catalyst for 8 mL Cr(VI) at initial pH value 3.5.
and 20.62 to 27.49 mg g−1 as the illumination time prolonging from 4 to 8 h for the initial Cr(VI) concentration of 80 ppm.
Three-level hierarchical micro/nanostructured Fe2+ /TiO2 catalyst were fabricated by a facile solvothermal approach combined with a reduction route. The Fe2+ /TiO2 catalyst shows an effective removal of Cr(VI) from water under UV light, and the removal effectiveness can reach 99.3% at the Cr(VI) initial concentration of 10 mg L−1 . The mechanism of the Fe2+ /TiO2 catalyst for the removal of Cr(VI) ions is a two-step reduction: the TiO2 catalyst photoreduces the Fe2+ ions to Fe atoms firstly, and then the Fe atoms reduce the Cr(VI) to Cr(III). The hierarchical architecture of the catalyst serves not only as a reactor for the photocatalytic reaction of Cr(IV) ions but as an effective absorbent for the removal of Cr(III) ions as well. The catalyst can be easily magnetically separated from the wastewater after photocatalytic reaction and recycled after acid treatment conveniently and economically. The Fe2+ /TiO2 catalyst may find potential applications as environmental benign photocatalyst for the removal of Cr(VI) from the wastewater, particularly in depth treatment of drinking water.
Acknowledgments 3.6. Recycle and separation of the Fe2+ /TiO2 catalyst The recycle ability of the Fe2+ /TiO2 catalyst was tested after treatment with 10% HNO3 . TEM characterizations show that no observable morphology change occurs after the treatment. The absorption spectra of the Cr(VI) in two cycles are shown in Fig. 11. One can see that the absorption peaks of Cr(VI) only slightly increase in the second cycle (curve 3) as compared with the first cycle (curve 2). The removal effectiveness in the second cycle is about 79.1%, which is about 89% capacity of the first cycle. The decrease in the removal effectiveness is considered mainly due to the consumption of Fe ions in the first cycle. The consumption of Fe ions during the photocatalytic reaction is shown in the inset (a) of Fig. 11, in which the consumption is negligible (below 0.01 ppm) at the first 2 h and increases to about 0.0488 ppm in the following 2 h. Such low consumption of Fe ions also indicates that the Fe2+ /TiO2 catalyst possess favorable durability in the photocatalytic activity. Furthermore, as the Fe2+ ions change partly to Fe3+ ions after photocatalytic reaction, the reacted catalyst can be magnetically separated from the wastewater as shown in the inset (b) of Fig. 11. This result demonstrates that the Fe2+ /TiO2 catalyst can be recycled conveniently and economically.
Fig. 11. Recyclable of the Fe2+ /TiO2 catalyst. Cr(VI) absorption spectra of curve (1) initial, curve (2) after the first photocatalytic cycle and curve (3) after the second cycle. The inset (a) is Fe ion concentration dissolved in the solution at different photocatalytic reaction times in the first cycle and the inset (b) demonstrates the magnetically separate the catalyst from the water.
The authors acknowledge the financial support from the National Natural Science Foundation of China (Grant 50802095), National Basic Research Program of China (Grant 2007CB936601) and the Foundation of President of the Hefei Institute of Physical Sciences of Chinese Academy of Sciences.
Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.08.068.
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Journal of Hazardous Materials 196 (2011) 36–43
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Thiol-functionalization of metal-organic framework by a facile coordination-based postsynthetic strategy and enhanced removal of Hg2+ from water Fei Ke a , Ling-Guang Qiu a,∗ , Yu-Peng Yuan a , Fu-Min Peng a , Xia Jiang a , An-Jian Xie a , Yu-Hua Shen a , Jun-Fa Zhu b a b
Laboratory of Advanced Porous Materials, School of Chemistry and Chemical Engineering, Anhui University, Hefei 230039, China National Synchrotron Radiation Laboratory, University of Science and Technology of China, Hefei 230029, China
a r t i c l e
i n f o
Article history: Received 22 June 2011 Received in revised form 8 August 2011 Accepted 26 August 2011 Available online 3 September 2011 Keywords: Metal-organic frameworks Functionalization Coordinatively unsaturated metal centers Postsynthetic strategy Heavy metal removal
a b s t r a c t The presence of coordinatively unsaturated metal centers in metal-organic frameworks (MOFs) provides an accessible way to selectively functionalize MOFs through coordination bonds. In this work, we describe thiol-functionalization of MOFs by choosing a well known three-dimensional (3D) Cu-based MOF, i.e. [Cu3 (BTC)2 (H2 O)3 ]n (HKUST-1, BTC = benzene-1,3,5-tricarboxylate), by a facile coordinationbased postsynthetic strategy, and demonstrate their application for removal of heavy metal ion from water. A series of [Cu3 (BTC)2 ]n samples stoichiometrically decorated with thiol groups has been prepared through coordination bonding of coordinatively unsaturated metal centers in HKUST-1 with –SH group in dithioglycol. The obtained thiol-functionalized samples were characterized by powder X-ray diffraction, scanning electron microscope, energy dispersive X-ray spectroscopy, infrared spectroscopy, and N2 sorption–desorption isothermal. Significantly, the thiol-functionalized [Cu3 (BTC)2 ]n exhibited remarkably high adsorption affinity (Kd = 4.73 × 105 mL g−1 ) and high adsorption capacity (714.29 mg g−1 ) for Hg2+ adsorption from water, while the unfunctionalized HKUST-1 showed no adsorption of Hg2+ under the same condition. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Heavy metals released into the environment have posed a significant threat to the environment and public health due to the fact that they can accumulate in the human body [1,2]. As a result, the removal and recovery of heavy metal ions from industrial waste water have been a significant concern in most industrial branches. A number of technologies, such as chemical precipitation [3], adsorption [4], membrane systems [5], ionic exchange [6], and liquid extraction [7], have been developed over the years, among them adsorption technology has attracted considerable attention due to its simplicity and low cost. Many kinds of adsorbents, such as activated carbon, coal fly ash and zeolites, have been used for the removal of heavy metal ions from wastewater [8–10]. Recently, mesoporous silicas modified by functional groups such as –NH2 and –SH have been demonstrated to be a new type of high-efficient heavy metal adsorbent. These silica-based
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[email protected] (L.-G. Qiu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.069
adsorbents have unique large specific surface areas and regular pore structures. Significantly, the functional groups modified on the pore surface of adsorbents afford these materials the ability to interact strongly with metallic cations [11]. A variety of amino- or thiol-functionalized mesoporous silica materials have been successfully fabricated. However, the synthesis procedure of these functionalized mesoporous silica is considerably complicated. Large amounts of surfactants have to be used as a structure-directing agent. To remove surfactant molecules from pores formed in the solid, the sample obtained has to be extracted several times using a large amount organic of solvent or calcined in air at a high temperature. Porous metal-organic frameworks (MOFs) [12], a new kind of crystalline porous material formed by metal ions (or clusters) and multidentate organic ligands, have been currently of intensively attraction, due to their great potential applications in catalysis [13,14] separations [15], drug delivery [16–19], gas storage [20], and sensing [21]. Their porous structures provide apparent surface areas of up to 5200 m2 g−1 [22], together with a large variety of pore dimensions and topologies. Remarkably, the ability to chemically modulate the physicochemical properties of MOF structures after
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Fig. 1. Schematic illustration of the thiol-functionalization of MOFs through coordination bonding between one thiol group of dithioglycol and coordinatively unsaturated metal centers (UMCs) in MOFs.
the formation of crystalline MOFs by a postsynthetic modification (PSM) concept allows tuning of the interactions with guest species, and increases the use of MOFs as high performance, tailor-made materials [23]. There are mainly three different strategies to achieve MOF functionalization by PSM strategies. The first relies on noncovalent interactions include guest removal, guest exchange, or ion exchange [24,25], whereas the second approach exploits functional properties by the covalent modification of porous MOFs [26,27]. The presence of coordinatively unsaturated metal centers (UMCs) in MOFs materials also provides functionalization of MOFs, and the resulting functionalized MOFs have clearly demonstrated to exhibit highly selective separation, chemisorptions, and catalytic performance [28–31]. Herein we describe the synthesis of thiol-functionalized MOFs by a facile coordination-based PSM strategy (see Fig. 1) and their application for Hg2+ adsorption. To demonstrate such PSM route to thiol-functionalized MOFs, we chose a well known copper-based MOF with UMCs, [Cu3 (BTC)2 (H2 O)3 ]n (HKUST-1, BTC = benzene1,3,5-tricarboxylate) [32]. HKUST-1 has a three dimensional (3D) ˚ that contains up to 10 square-shaped channel system (9 × 9 A) additional water molecules per formula unit. The lability of axial aqua ligands on the paddle-wheel secondary building units in [Cu3 (BTC)2 (H2 O)3 ]n permits their replacement by other molecules. The coordinated water molecules in the framework as well as excess of water guests in the channels can easily be removed after vacuum treatment at 150 ◦ C for 12 h, thus providing accessible sites for the surface functionalization. By treating the dehydrated MOF with dithioglycol, a series of thiol-functionalized MOF-based materials has been prepared. Their application for removal of heavy metal ions was preliminarily demonstrated by Hg2+ adsorption from water. Significantly, the resulting thiol-functionalized MOF exhibited both high adsorption affinity (Kd = 4.73 × 105 mL g−1 ) and remarkable adsorption capacity (714.29 mg g−1 ) for Hg2+ adsorption from water, clearly demonstrating that thiol-functionalized MOFs could act as useful and effective adsorbents to remove heavy metal ions from the contaminated water.
2. Experimental 2.1. Materials and methods Benzene-1,3,5-tricarboxylic acid (H3 BTC) was purchased from Aldrich, cupric nitrate trihydrate was purchased from Sinopharm (Shanghai) Chemical Reagent Co., Ltd., China, and dithioglycol was
purchased from Tokyo Chemical Industry Co., Ltd., Japan. All other chemicals used in this work were of analytical grade, obtained from commercial suppliers, and used without further purification unless otherwise noted. High concentration (1431.04 mg L−1 ) Hg2+ solution used for adsorption experiments, as well as Hg2+ stock solutions (6.51 mg L−1 ), were prepared by dissolving irradiated HgCl2 in deionized water, while Hg2+ solutions at lower concentrations (0.081–1.30 mg L−1 ) were prepared by dilution of measured volumes of stock solution (6.51 mg L−1 ) with deionized water. The powder X-ray diffraction (PXRD) patterns of the samples were collected using an X-ray diffractometer with Cu target (36 kV, 25 mA) from 5 to 50◦ . Hg contents were determined by atomic fluorescence spectrometry (AFS) using a AFS-3100 spectrometer. Analyses of the morphology and chemical composition of the samples were conducted by a Hitachi S-4800 field emission scanning electron microscope (FE-SEM) equipped with an energy dispersive X-ray (Oxford Instruments INCA EDX) system. Nitrogen sorption–desorption isotherms were obtained at 77 K on a Micromeritics ASAP 2020M+C analyzer.
2.2. Synthesis of thiol-functionalized [Cu3 (BTC)2 ]n samples [Cu3 (BTC)2 (H2 O)3 ]n crystals were prepared by a solvothermal reaction as reported previously [32]. Typically, 1.087 g (4.5 mmol) of Cu(NO3 )2 ·3H2 O was dissolved in 15 mL deionized water and then mixed with 0.525 g (2.5 mmol) of H3 BTC dissolved in 15 mL ethanol. The mixture was stirred for 30 min, and then transferred to a 50 mL Teflon autoclave liner and sealed to heat at 120 ◦ C for 12 h. The obtained blue powder was filtered off, washed several times with deionized water and ethanol, and then dried overnight at 150 ◦ C under air atmosphere. Yield of [Cu3 (BTC)2 (H2 O)3 ]n crystals was calculated to be 65.04% on the basis of Cu. To prepare the thiol-functionalized [Cu3 (BTC)2 ]n samples, 0.1 g as-synthesized [Cu3 (BTC)2 (H2 O)3 ]n sample was dehydrated at 150 ◦ C for 12 h, and then suspended in 10 mL of anhydrous toluene. An appropriate amount of 0.24 mol L−1 dithioglycol solution (0.5 mL for sample A, 1 mL for B, and 1.5 mL for C, respectively) in anhydrous toluene was added to the suspension, and the mixture solution was then stirred magnetically for 24 h at room temperature. The product was recovered by filtration and washed with ethanol (15 mL × 5), and then dried overnight at room temperature in vacuum. The relative contents of S in the functionalized samples A, B and C were determined by EDX spectra. The molar ratios of S to Cu in the framework for samples A, B and C were calculated to be 0.18, 0.92, 1.52, respectively, and thus the obtained
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samples A, B, and C were also labeled as Cu-BTC-DTG-0.18, Cu-BTC-DTG-0.92, and Cu-BTC-DTG-1.52 (DTG = dithioglycol), respectively, in the present work. 2.3. Mercury adsorption studies The capacity of the thiol-modified [Cu3 (BTC)2 ]n samples to adsorb mercury ion from water was determined using a batch of mercury chloride aqueous solutions of known concentration. 10 mg of the thiol-functionalized [Cu3 (BTC)2 ]n crystals were added to 10 mL of the Hg2+ solutions (1431.04, 1252.16, 1073.28, 894.40, 715.52, 143.10, 6.51, 1.30, 0.65, 0.16 and 0.08 mg L−1 , respectively) under constant shaking at room temperature for 24 h. After high-speed centrifugation, both initial and the remaining Hg2+ concentrations were determined with AFS. 3. Results and discussion 3.1. Characterization 3.1.1. PXRD measurements The synthesis of the dithioglycol grafted [Cu3 (BTC)2 ]n through coordination of dithioglycol to the unsaturated metal centers of the dehydrated [Cu3 (BTC)2 ]n framework was performed in anhydrous toluene under magnetic stirring at room temperature. Initial evidence of the as-synthesized thiol-functionalized samples was confirmed by PXRD (Fig. 2). All of the diffraction peaks for thiolfunctionalized samples A and B show that the sketch of the MOF [Cu3 (BTC)2 ]n crystal is well retained even after the modification with dithioglycol, because all diffraction peaks of the functionalized [Cu3 (BTC)2 ]n can readily be indexed to HKUST-1. For sample C, however, modification of [Cu3 (BTC)2 ]n with excessive dithioglycol resulted in loss of crystalline order of the framework, as evidenced by significant decrease in diffraction intensities (see Fig. 2d), which is due to partial decomposition of the crystalline [Cu3 (BTC)2 ]n .
Fig. 2. PXRD patterns of (a) as-synthesized [Cu3 (BTC)2 (H2 O)3 ]n crystals, and (b–d) thiol-modified [Cu3 (BTC)2 ]n samples A (Cu-BTC-DTG-0.18), B (Cu-BTC-DTG-0.92), and C (Cu-BTC-DTG-1.52), respectively.
3.1.2. SEM–EDX measurements To interrogate the surface morphology and chemical composition of the functionalized samples, the thiol-functionalized samples were characterized by SEM (Fig. 3) and EDX (Fig. 4). The crystals of original [Cu3 (BTC)2 ]n sample before the thiol-functionalization are octahedral with a smooth surface and have an average size of 8 m (Fig. 3a). However, surfaces of the thiol-modified [Cu3 (BTC)2 ]n samples A, B and C tend to be rougher after the functionalization (Fig. 3b–d). Moreover, for the thiol-modified sample C, increasing S/Cu molar ratio to 1.52 caused a significant morphological change from regular octahedral [Cu3 (BTC)2 ]n crystals to irregular particles, which is due to partial decomposition of the framework. The EDX spectra of the thiol-functionalized samples reveal that the samples
Fig. 3. SEM images of (a) the as-synthesized [Cu3 (BTC)2 (H2 O)3 ]n , and (b–d) the thiol-modified [Cu3 (BTC)2 ]n samples A (Cu-BTC-DTG-0.18), B (Cu-BTC-DTG-0.92), and C (Cu-BTC-DTG-1.52), respectively.
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Fig. 4. EDX spectra of (a–c) the thiol-modified [Cu3 (BTC)2 ]n samples A (Cu-BTC-DTG-0.18), B (Cu-BTC-DTG-0.92), and C (Cu-BTC-DTG-1.52), respectively.
are composed of C, O, Cu and S, as shown in Fig. 4a–c. The relative contents of S in the functionalized samples A, B and C were determined to be 0.83, 4.24 and 6.98 mmol g−1 of [Cu3 (BTC)2 ]n , which correspond to S/Cu ratio of 0.18, 0.92, 1.52, respectively (Table 1), suggesting that content of –SH group grafted on the framework can easily be tuned by varying molar ratio of dithioglycol to the framework.
3.1.3. IR measurement Fig. 5 shows IR spectra of the thiol-functionalized samples, as well as the bare [Cu3 (BTC)2 (H2 O)3 ]n crystals. The bare [Cu3 (BTC)2 (H2 O)3 ]n shows bands (ArC–H) corresponding to aromatic groups at 3050 cm−1 , and vibrational bands characteristics of the –O–C–O– group around 1550 and 1430 cm−1 (Fig. 5a). The bands around 2900, 2581, and 686 cm−1 observed in thiolfunctionalized samples can be attributed to the presence of (C–H), (S–H), and (C–S) vibrations, respectively (see Fig. 5b–d),
indicating the presence of dithioglycol. Although the peak at 2581 cm−1 corresponding to (S–H) is not very strong, obvious shift of the aliphatic (C–H) stretching vibrations at 2800–3000 cm−1 to larger values was found, as observed when the molecule is coordinated to a Lewis acid center [30,33]. The result clearly reveals that dithioglycol molecules were successfully grated onto the UMCs in channels created in the framework, rather than adsorbed on external surface of [Cu3 (BTC)2 ]n crystals. This result was also confirmed by N2 sorption–desorption isotherms of the samples as shown below.
3.1.4. N2 sorption–desorption isotherms The resulting of thiol-grafting on coordinatively unsaturated copper centers of [Cu3 (BTC)2 ]n crystal is also visible in the N2 sorption–desorption isotherms. The N2 sorption–desorption isotherms of the as-synthesized samples reveal a typical microporous material (Fig. 6). Compared with the pristine
Table 1 Physicochemical properties of [Cu3 (BTC)2 (H2 O)3 ]n and its dithioglycol-functionalized analogs. Sample
SBET a (m2 g−1 )
Vt b (m3 g−1 )
S contentc (mmol g−1 )
S/Cu ratio
Kd (Hg)d (mL g−1 )
[Cu3 (BTC)2 (H2 O)3 ]n A (Cu-BTC-DTG-0.18) B (Cu-BTC-DTG-0.92) C (Cu-BTC-DTG-1.52)
1492.19 800.42 331.39 79.92
0.75 0.38 0.18 0.05
– 0.83 4.24 6.98
– 0.18 0.92 1.52
0
a b c d
SBET represents BET surface areas obtained from N2 sorption–desorption isotherms. Vt represents pore volumes obtained from N2 sorption–desorption isotherms. S content represents the relative contents of S in the functionalized samples per gram of [Cu3 (BTC)2 ]n . Kd (Hg) represents the distribution coefficient of Hg2+ ion.
4.73 × 105
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F. Ke et al. / Journal of Hazardous Materials 196 (2011) 36–43 Table 2 Hg2+ concentrations in aqueous solutions before and after treating with the thiolmodified [Cu3 (BTC)2 ]n sample B (Cu-BTC-DTG-0.92).a Hg2+ concentration (mg L−1 )
Fig. 5. Infrared spectra of (a) the as-synthesized [Cu3 (BTC)2 (H2 O)3 ]n , and (b–d) the thiol-modified [Cu3 (BTC)2 ]n samples A (Cu-BTC-DTG-0.18), B (Cu-BTC-DTG-0.92), and C (Cu-BTC-DTG-1.52), respectively.
Initial
After adsorption
Adsorption capacity (mg g−1 )
1431.04 1252.16 1073.28 894.40 715.52 143.10 6.51 1.30 0.65 0.16 0.081
712.90 542.67 388.75 179.63 1.51 0.13 0.011 0.015 0.019 0.0074 0.0075
718.14 709.49 684.53 714.78 714.01 142.96 6.50 1.29 0.63 0.16 0.074
Removal efficiency (%) 50.18 56.67 63.78 79.92 99.79 99.91 99.83 98.85 97.07 95.38 90.74
a Adsorption conditions: 10 mg of the thiol-functionalized [Cu3 (BTC)2 ]n crystals were added to 10 mL of the Hg2+ solutions under constant shaking at room temperature for 24 h. After high-speed centrifugation, both initial and the remaining Hg2+ concentrations were determined with AFS.
3.2. Adsorption isotherms for Hg2+ removal from water [Cu3 (BTC)2 (H2 O)3 ]n , the thiol-modificated samples exhibit a significant decrease in both surface areas and pore volumes. The Brunauer–Emmett–Teller surface areas (SBET ) of the thiolfunctionalized samples A, B, and C decreased significantly from 1492.19 m2 g−1 to 800.42, 331.39 and 79.92 m2 g−1 , and pore volumes (Vt ) decreased from 0.75 cm3 g−1 to 0.38, 0.18 and 0.05 cm3 g−1 , respectively, after dithioglycol-grafting (Table 1). The result indicates that the pores of the [Cu3 (BTC)2 ]n are partially blocked after the thiol-grafting, leaving limited accessible pore volume for nitrogen molecules. The occupation of partial space in the channels of the [Cu3 (BTC)2 ]n sample by dithioglycol molecules resulted in a significant decrease in both surface area and pore volume of the thiol-functionlized [Cu3 (BTC)2 ]n samples, clearly demonstrating that the grafted thiol groups are present in the channels of [Cu3 (BTC)2 ]n rather than on the surfaces of the [Cu3 (BTC)2 ]n crystals as discussed above.
The ability of such thiol-functionalized MOFs to remove heavy metal ions from contaminated water was preliminarily demonstrated by removal of mercury ion from water using the thiol-functionlized [Cu3 (BTC)2 ]n sample B under a wide range of known mercury concentrations. As shown in Table 2, the treatment of Hg2+ solution with sample B resulted in 99.79% removal of Hg2+ ion with an unprecedented adsorption capacity of 714.01 mg g−1 when initial Hg2+ concentration was as high as 715.5 mg L−1 . More importantly, the sample could still remove 90.74% of Hg2+ ion from the solution even when initial Hg2+ concentration was as low as 81 ppb. Adsorption of Hg2+ from aqueous solution using bare [Cu3 (BTC)2 (H2 O)3 ]n sample was also carried out as a control experiment, and no detectable amount of Hg2+ ion was adsorbed by the unfunctionalized [Cu3 (BTC)2 (H2 O)3 ]n . In addition, compared with other functionalized porous adsorbents such as ion-exchange resin and hybrid silica materials, the thiol-functionalized MOF exhibits a significantly high adsorption capacity [6,11]. Such high capacity of the thiol-functionalized MOFs for Hg2+ adsorption can be attributed to the thiol groups densely populated on the inner surface of porous MOFs with unique large specific surface areas and high density of adsorption sites. To gain a better understanding of the mechanism of the heavy metal removal, the thiol-functionalized sample was characterized using the adsorption isotherm. Fig. 7 shows the adsorption isotherm of Hg2+ onto the thiol-functionalized MOF sample B. Remarkably, the equilibrium adsorption capacities of Hg2+ increase sharply with an increase in Hg2+ initial concentration when its value is lower than 715.52 mg L−1 . However, when concentration of Hg2+ ion is higher than this value, increasing the Hg2+ concentration does not affect adsorption capacity of Hg2+ , which is due to saturated adsorption of Hg2+ onto the thiol-functionalized sample. By fitting the equilibrium adsorption data with Langmuir adsorption model, adsorption capacity of Hg2+ onto the thiolfunctionalized sample was calculated from Eq. (1) [34,35]: Ce 1 Ce = + qe qm KL qm
Fig. 6. Nitrogen sorption–desorption isotherms of (a) the as-synthesized [Cu3 (BTC)2 (H2 O)3 ]n , and (b–d) the thiol-modified [Cu3 (BTC)2 ]n samples A (Cu-BTCDTG-0.18), B (Cu-BTC-DTG-0.92), and C (Cu-BTC-DTG-1.52), respectively.
(1)
where Ce is the equilibrium concentration of remaining Hg2+ ion in the solution (mg L−1 ), qe is the amount of Hg2+ ion adsorbed per mass unit of adsorbent at equilibrium (mg g−1 ), qm is the monolayer adsorption capacity (mg g−1 ), and KL is the Langmuir constant (L mg−1 ).
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Fig. 7. (a) Adsorption curve of Hg2+ at different concentrations by using the thiol-modified [Cu3 (BTC)2 ]n samples B (Cu-BTC-DTG-0.92); inset shows adsorption curve of Hg2+ at low concentrations. (b) The linear regression by fitting the equilibrium adsorption data with Langmuir adsorption model.
The linear regression between Ce /qe and Ce was calculated, and the correlation coefficient of the straight line is 0.9997, indicating that the adsorption of Hg2+ ion conforms Langmuir’s adsorption model. Remarkably, the maximum adsorption capacity of the thiol-functionalized [Cu3 (BTC)2 ]n sample was calculated to be 714.29 mg g−1 for Hg2+ , which corresponds to capture of 0.952 Hg2+ ion per terminal thiol group in the thiol-functionalized [Cu3 (BTC)2 ]n , suggesting that Hg2+ ion were allowed to access almost all free thiol groups in the functionalized [Cu3 (BTC)2 ]n . The maximum adsorption capacity value is higher than many other conventional porous adsorbents [34,35]. The ability of the thiol-functionalized MOF to remove Hg2+ ion from water can be also expressed with a distribution coefficient (Kd ), which is defined as the following equation [36] Kd =
(Ci − Cf ) Cf
×
V m
(2)
where Ci and Cf represent the initial and final solution concentrations, respectively, V represents the volume of solution (mL), and m represents the mass of adsorbent (g). The Kd value of the thiol-functionalized sample B for Hg2+ adsorption is listed in Table 1. Although bare [Cu3 (BTC)2 (H2 O)3 ]n has no binding affinity to Hg2+ , the thiol-functionalization of [Cu3 (BTC)2 (H2 O)3 ]n leads to a significant increase in the binding capacity for Hg2+ . The Kd value for the thiol functionalized sample B was calculated to be 4.73 × 105 mL g−1 for Hg2+ in the single solution, which is higher than some functionalized nanoporous silica reported very recently [36]. Although a few of MOFs and zeolitic imidazolate frameworks (ZIFs), such as MIL-101 and ZIF-8, exhibit exceptional chemical stability, many other organic–inorganic hybrid materials are commonly unstable under strong acidic conditions. As a result, it may be a challenge to recycle such MOF-type absorbents after heavy metal adsorption because efficient regeneration of the absorbents (e.g. activated carbon, ion-exchange resin, and thiol-functionalized mesoporous silica) after heavy metal adsorption has always to be carried out under strong acidic conditions. However, the present work establishes a simple and efficient route to a novel type of thiolfunctionalized adsorbents for heavy metal removal on the basis of MOFs-based materials. Such the thiol-functionalized MOFs exhibit remarkably high adsorption capacity for heavy metal removal, as clearly demonstrated in the present work. In addition, high removal efficiencies were also observed for heavy metal removal from water; removal percent varied from 95.38 to 99.79% depending on various Hg2+ concentrations. More importantly, such absorbent exhibits a significantly high efficiency for Hg2+ removal even when concentration of heavy metal ion is extremely low (81 ppb). Very
recently, we reported a facile and environmentally friendly fabrication of a novel type of magnetic porous MOF-based nanocomposites by incorporation of Fe3 O4 nanorods with MOF nanocrystals. Such MOF-based nanocomposites exhibited both magnetic characteristics and high porosity [19]. These properties may provide the thiol-functionalized MOFs more chances in biomedical applications (for example, decorporation of metal toxins from biological environment and targeted drug delivery). 3.3. Adsorption kinetics In order to evaluate the kinetic mechanism that controls the adsorption process, the effect of contact time on the adsorption of Hg2+ onto the thiol-functionalized [Cu3 (BTC)2 ]n was further investigated by adding 10 mg of sample B to 10 mL of 6.51 mg L−1 Hg2+ solution. As can be seen in Fig. 8, the thiol-functionalized sample attained 99% of the adsorption capacity at equilibrium within 120 min, suggesting that the thiol-functionalized MOF possesses both a high adsorption capacity and a high adsorption efficiency for the removal of heavy metal ions from water. The rate constant for pseudo-second-order adsorption could be obtained from the following equation [37,38]: t 1 t = + qt qe k2 q2e
(3)
where k2 is the rate constant of the pseudo-second-order adsorption (g mg−1 min−1 ), and qt is the amount of Hg2+ adsorbed at time t
Fig. 8. Adsorption curve of Hg2+ versus contact time in aqueous solution by using the thiol-modified [Cu3 (BTC)2 ]n samples B (Cu-BTC-DTG-0.92). Inset shows the pseudosecond-order kinetic plot for the adsorption (Hg2+ concentration: 6.51 mg L−1 ).
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(mg g−1 ). By fitting the experimental date with the pseudo-secondorder kinetic model using Eq. (3), the adsorption rate constant k2 , the calculated qe value and the correlation coefficient were obtained. An extremely high correlation coefficient (0.9999) was obtained. Moreover, the calculated qe value also agrees with the experimental data in the case of pseudo-second-order kinetics, suggesting the adsorption of Hg2+ onto the adsorbent follows a pseudo-second-order kinetic model. The value of the adsorption rate constant k2 under this condition in the present work was determined to be 0.13 g mg−1 min−1 . This value exceeds many other porous absorbents [34,35,37,38], which can also be attributed to the thiol groups densely populated on the inner surface of porous MOFs with unique large specific surface areas and high density of adsorption sites. 4. Conclusions In conclusion, the present work demonstrates a facile strategy to fabricate thiol-functionalized porous MOFs as a novel type of adsorbent for removal of Hg2+ . Thiol-functionalized [Cu3 (BTC)2 ]n samples with different S/Cu molar ratio were prepared by thiol grafting on coordinatively unsaturated copper center in a 3D porous MOF, [Cu3 (BTC)2 ]n , and their structures, morphologies and porosity were characterized. Although the unfunctionalized [Cu3 (BTC)2 ]n showed no adsorption of Hg2+ , the thiol-functionalized samples exhibited a high affinity (distribution coefficient Kd = 4.73 × 105 mL g−1 ) and significant adsorption capacity (714.29 mg g−1 ) for Hg2+ removal from aqueous solution, making them potential for selective removal of heavy metal ions from water. We believe that this work establishes a simple and energy efficient route to a novel type of adsorbents for heavy metal removal on the basis of the functionalization of MOFs-based materials. Also, the present strategy for tailoring MOFs-based porous materials by grafting different organic functional groups using coordination chemistry will offer a plethora of opportunities for generating new functionalized materials for dye adsorption, CO2 capture, sensing, as well as gas purification. Acknowledgments This work was supported by the National Natural Science Foundation of China (NSFC, 20971001), the NSFC-CAS Joint Fund for Research Based on Large-Scale Scientific Facilities (10979014), the Program for New Century Excellent Talent in University, Ministry of Education, China (NCET-08-0617), and the “211 Project” of Anhui University. J.-F. Zhu is grateful for the financial support from NSFC (20873128) and the “Hundred Talents Program” of the Chinese Academy of Sciences. References [1] C. Wang, S.Y. Tao, W. Wei, C.G. Meng, F.Y. Liu, M. Han, Multifunctional mesoporous material for detection, adsorption and removal of Hg2+ in aqueous solution, J. Mater. Chem. 20 (2010) 4635–4641. [2] P. Miretzky, A.F. Cirelli, Hg(II) removal from water by chitosan and chitosan derivatives: a review, J. Hazard. Mater. 167 (2009) 10–23. [3] W.S. Wan Ngah, M.A.K.M. Hanafiah, Removal of heavy metal ions from wastewater by chemically modified plant wastes as adsorbents: a review, Bioresour. Technol. 99 (2008) 3935–3948. [4] V.K. Gupta, P.J.M. Carrott, M.M.L. Ribeiro Carrott, Suhas, Low-cost adsorbents: growing approach to wastewater treatment—a review, Crit. Rev. Environ. Sci. Technol. 39 (2009) 783–842. [5] P. Kumar, V.V. Guliants, Periodic mesoporous organic–inorganic hybrid materials: applications in membrane separations and adsorption, Micropor. Mesopor. Mater. 132 (2010) 1–14. ˛ [6] A. Dabrowski, Z. Hubicki, P. Podko´scielny, E. Robens, Selective removal of the heavy metal ions from waters and industrial wastewaters by ion-exchange method, Chemosphere 56 (2004) 91–106.
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Journal of Hazardous Materials 196 (2011) 44–51
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Bioremoval of trivalent chromium using Bacillus biofilms through continuous flow reactor K. Sundar, I. Mohammed Sadiq, Amitava Mukherjee, N. Chandrasekaran ∗ Centre for Nanobiotechnology, Nano Bio-Medicine Laboratory School of Bio Sciences and Technology VIT University, Vellore – 632014, India
a r t i c l e
i n f o
Article history: Received 27 June 2011 Received in revised form 12 August 2011 Accepted 27 August 2011 Available online 2 September 2011 Keywords: Trivalent chromium Continuous flow reactor Bacillus Confocal laser scanning microscope Biofilms
a b s t r a c t Present study deals with the applicability of bacterial biofilms for the bioremoval of trivalent chromium from tannery effluents. A continuous flow reactor was designed for the development of biofilms on different substrates like glass beads, pebbles and coarse sand. The parameters for the continuous flow reactor were 20 ml/min flow rate at 30 ◦ C, pH4. Biofilm biomass on the substrates was in the following sequence: coarse sand > pebbles > glass beads (4.8 × 107 , 4.5 × 107 and 3.5 × 105 CFU/cm2 ), which was confirmed by CLSM. Biofilms developed using consortium of Bacillus subtilis and Bacillus cereus on coarse sand had more surface area and was able to remove 98% of Cr(III), SEM-EDX proved 92.60% Cr(III) adsorption on biofilms supported by coarse sand. Utilization of Bacillus biofilms for effective bioremoval of Cr(III) from chrome tanning effluent could be a better option for tannery industry, especially during post chrome tanning operation. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Vellore district is mushroomed with many tanneries. There is lot of unorganized tanning industries; which forms the livelihood of Vellore population. Cr(III) is used as a tanning agent in the leather industries; the wastewater resulting from chrome tanning processes contains high amount of chromium metal, which is harmful for the environment and human health [1]. Toxicological studies of Cr(III) compounds reported skeletal and neurological disorders [2]. Cr(III) compounds are cytotoxic and forms DNA adduct [3]. Though there are many methods for effluent treatment like precipitation, chemical oxidation or reduction, lime neutralization, ion exchange, filtration, electrochemical treatment, reverse osmosis, membrane technologies and evaporation recovery. All these methods are expensive and will produce solid sludge containing toxic compounds [4]. And also chromium at low concentrations in the effluent waters cannot be removed by conventional methods [5]. Thus an alternate treatment strategy is required, which would be environment friendly. Indigenous chromium tolerant bacterial strains might be a better choice for the development of biofilms towards chromium removal from tannery effluent.
∗ Corresponding author. Tel.: +91 9442405757. E-mail addresses:
[email protected],
[email protected] (N. Chandrasekaran). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.066
The purpose of the present study was to isolate Cr(III) tolerant bacterial species from tannery effluent polluted sites in Palar river basin and to develop a bacterial biofilm on different substrates like glass beads, pebbles, coarse sand for bioremoval of Cr(III). The study focuses on the bioremoval of Cr(III) using indigenous chromium tolerant bacterial biofilms through continuous flow reactor. The biofilm bioreactor would be a better choice for tanneries to alleviate Cr(III) pollution in the effluent waters. This technology could be adopted in an industrial scale for environmental problems of tanneries.
2. Materials and methods 2.1. Isolation, screening and characterization of effective strains Soil and water samples were collected from chromium polluted sites in the Palar river basin of Vellore district, Tamilnadu, India. Samples were processed for the isolation of Cr(III) tolerant bacterial strains as per APHA [6]. Trivalent chromium tolerant bacteria were isolated using nutrient agar plates amended with Cr(NO3 )3·9H2 O at pH 4 and incubated at 30 ◦ C for 3–5 days. The effective strains were screened based on maximum tolerable concentration (MTC), Cr(III) bioremoval ability and exopolysaccharide (EPS) production [7]. The selected strains were characterized morphologically, biochemically and physiologically following Gerhardt et al. [8]. The taxonomical identifications of the selected bacterial strains were confirmed by 16S rRNA gene sequencing.
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biofilm development. The reactor was packed with sterile glass beads (5 mm)/pebbles10–15 mm)/coarse sand (3–5 mm) up to 2/3 of the column length. The bacterial inoculum was fed through the substrates at different flow rates of 10, 20, 30, 40 and 50 ml/min, up to 72 h at 30 ◦ C. Intermittent samples were collected at 24, 48 and 72 h and were further characterized. 2.4. Characterization of developed biofilms 2.4.1. Biomass estimation The developed biofilms were estimated for total bacterial biomass by viable plate count method. 1 cm2 biofilm samples were scraped from the substrates using a sterile scalpel and dispersed in 10 ml of sterile water. The dispersed biofilm was plated on nutrient agar plates and incubated at 30 ◦ C for 24 h. The NA plates were calculated for colony forming units and there by biomass per cm2 were calculated. 2.4.2. Stability of biofilms Stability of biofilm on the substrates was studied to check the irreversible biofilm formation. After the biofilm development, sterile water was passed through the column at 50 ml/min flow rate for 24 h. Stable biomass was calculated by detecting biomass adhered on the substrate/cm2 after water flow from initial biomass calculated as in Section 2.4.1.
Fig. 1. Schematic diagram of continuous flow biofilm cultivation reactor.
2.2. Biofilm formation by tissue culture plate assay The bacterial biofilm estimation was carried out using standard tissue culture plate (TCP) assay as described by Christensen et al. [9]. In the present study, two bacterial strains (Bacillus subtilis VITSCCr 01 and Bacillus cereus VITSCCr 02) were observed for their ability to form biofilm on tissue culture plate. Overnight cultures of Bacillus strains were diluted to 1 × 106 CFU/mL in fresh Nutrient Broth (NB) with 0.5% glucose after adjusting the OD at 600 nm. Aliquots (200 L) of various concentrations of Cr(III) (10, 25, 50, 75, and 100 mg/l) was transferred onto 96-well flat-bottom tissue culture plates, uninoculated broth served as control. Plates were incubated in static and dynamic conditions for 48 h at 30 ◦ C. After incubation, the content of each well was gently removed by tapping the plates. The wells were washed twice with 200 L of deionized water to remove free-floating ‘planktonic’ bacteria. Biofilms in plates were dried at 60 ◦ C for 1 h and adherent bacteria were stained with 200 L of 0.1% crystal violet for 5 min. The plates were rinsed twice with deionized water to remove excess stain and dried at 37 ◦ C for 2 h. Stained adherent cells were detached from the plates using 200 L of 30% (w/v) glacial acetic acid for 10 min with shaking at 300 rpm. The OD of stained biofilm was determined at 492 nm (Power Wave XS2, Biotek Corp, Microplate reader). The mean OD value obtained from the media control well was deducted from all the test OD values.
2.4.3. Microscopic examination Morphological characterization of biofilms was done by Scanning Electron Microscope (SEM), Atomic force microscope (AFM) and Confocal laser scanning microscopy (CLSM). For SEM the biofilms samples were fixed with 2.5% glutaraldehyde, ethanol (dehydrated) and coated with gold under vacuum in an argon atmosphere. The surface morphology of the gold coated samples was visualized by a Scanning Electron Microscope (Hitachi S4000). SEM allowed the identification of any interesting structural features on the morphology of biofilms. For AFM imaging the biofilms were fixed with 0.4% paraformaldehyde for 5 min and fixed cells were imaged using Atomic Force Microscopy (Non contact mode) (Nanosurf Easy Scan2, Nanosurf Inc.; USA). This mode of AFM enhances the topography of biofilm by scanning the tip above the surface without damaging the biofilm. For CSLM biofilm samples from all substrates were thoroughly washed thrice with 0.05 M phosphate buffer, pH 7.4 and immersed in 1 ml solution of phosphate buffered-saline for 5 min. Samples were stained with acridine orange (0.01%, w/v) for 10 min at room temperature in the dark. Then, samples were washed with PBS and fixed with 4% glutaraldehyde for 1 h. Laser confocal scanning microscope (Leica SP2; Leica Microsystems, Heidelberg, Germany) was used for the imaging of samples. The Leica confocal software was used for analysis of biofilm images, which allowed for collection of z-stacks three-dimensional (3D) reconstruction. Excitation and emission wavelength were set to 460 and 650 nm respectively by adjustable spectrum slit.
2.3. Continuous flow reactor design for biofilm development
2.5. Cr(III) bioremoval studies in continuous flow reactor using Bacillus biofilms
Simplified continuous flow reactor was designed for the cultivation of biofilm and for the treatment of chrome tanning effluent. The continuous flow reactor was fabricated using silicate glass material attached with a two channel peristaltic pump (Ravel Hiteks, RH-P 100 S–100–2H) to enable the uninterrupted flow of culture. The description of the reactor set up is given in Fig. 1. The overnight culture of B. subtilis VITSCCr 01 and B. cereus VITSCCr 02 and their consortium were used as seed inoculum for
The chromium uptake studies were performed with standard chromium solutions (25, 50 and 100 mg/l) and chrome tanning effluents (1:10 and 1:100 diluted with distilled water). After biofilm formation, the biofilm bed was washed out and the Cr(III) solutions or the sterile chrome tanning effluents were passed continuously through the column with a flow rate of 20 ml/min. Cr concentration at the inlet (initial concentration) as well as at the outlet (final concentration) of the column was measured by Atomic
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Absorption Spectroscopy (VARIAN SPECTRAA-240) at wavelengths of 357.9 nm, 425.4 nm and 520.8 nm. Optimizations of temperature, pH, column size and flow rate were performed to achieve maximum bioremoval. Surface morphology of the biofilms used for Cr(III) bioremoval was visualized by Scanning Electron Microscope (Hitachi S4000) with combined Energy Dispersive X-ray Analyser at a voltage of 10 keV. SEM allowed the identification of any interesting structural features on the morphology of biofilms. Elemental composition of the insoluble precipitates in Cr-incubated biofilms was verified using energy dispersive X-ray analysis.
Table 1 Biomass quantification of biofilms of Bacillus subtilis and Bacillus cereus consortium on different substrates. Substrate
After 24 h (CFU/cm2 )
After 48 h (CFU/cm2 )
After 72 h (CFU/cm2 )
Glass beads Pebbles Coarse sand
1.6 × 102 2.4 × 102 2.7 × 102
4.2 × 104 1.3 × 105 2.2 × 105
3.5 × 105 4.3 × 107 4.8 × 107
3.2. Biofilm formation on different substrates by continuous flow reactor
2.6. Adsorption isotherms Two isotherm equations have been tested in this study and details of the equations and parameters are presented below [10]. (A) The general Langmuir sorption model is expressed by Qe =
(Qmax bCe ) (1 + bCe )
(1)
Qe (mg/g) is the amount of metal ion sorbed by the biofilm at equilibrium, Qmax (mg/g) is the maximum metal sorption, Ce (mg/l) is the concentration of metal in solution at the equilibrium and b (1 mg–1 ) is the Langmuir adsorption equilibrium constant.Freundlich isotherm is expressed by 1/n
Qe = Kf Ce
(2)
Qe and Ce are the same as in the Langmuir equation, and Kf and n relate to the capacity and intensity of adsorption, respectively 2.7. Statistical analysis Each set of experiments was carried out at least in duplicate, and in triplicate in some cases. Experiments were repeated separately to ensure reproducibility. In each set of repeated experiments, standard deviations and standard error showed 95% confidence interval. 3. Results and discussion 3.1. Isolation, screening and characterization of Cr(III) tolerant biofilm forming bacteria Forty five indigenous Cr(III) tolerant bacteria were isolated from water and soil samples of tannery polluted sites in Palar river basin. Bacterial strains were screened based on Maximum tolerance concentration to Cr(III), Cr(III) bioremoval ability and EPS production [7]. Among the isolates two bacterial strains were selected and observed for biofilm formation by tissue culture plate assay (TPC), both the bacterial strains showed maximum biofilm formation in microtitre plates, which was confirmed with crystal violet staining. Further, they were characterized as B. subtilis [7] and B. cereus (showing 99% similarity in BLAST search to B. cereus and the sequence size was 1423 bp) by 16S rRNA sequencing (GenBank accession number GQ395343 and GQ395344). Isolation of bacteria from tannery environments represents an appropriate method for chromium removal and bioremediation [11]. It has been already reported, that chromium resistant bacterial strains isolated from chromium polluted environment are capable of reducing chromate [12]. Exopolysaccharide plays a major role in the biofilm formations and also sequestration of metallic ions [13], hence EPS production is one of the important criteria for selection of effective strain.
The B. subtilis VITSCCr01 and B. cereus VITSCCr02 strains were used for biofilm development on glass beads, pebbles and coarse sand by continuous flow reactor. Optimization of reactor for flow rate, substrate quantity and time taken for the biofilm formation was done for each substrate. The optimum flow rate for biofilm formation was 20 ml/min for glass beads and pebbles and it was 30 ml/min for coarse sand. The substrate quantity was fixed as 2/3 of the total column length for all the substrates. Time taken for biofilm formation was varied among the substrates in the following order; coarse sand 24 h, glass beads – 48 h and pebbles – 72 h. Studies by Quintelas et al. [14,15] and Lin et al. [16] developed biofilms on granular activated carbon, kaolin, zeolite and chitosan bead for the removal of hexavalent chromium and other metals. But there are no reports on biofilm development on substrates like glass beads, coarse sand and pebbles for Cr(III) removal studies, using continuous flow reactor design. Moreover, continuous flow biofilm reactor design using indigenous bacterial strains to remove Cr(III) would be an ideal green technology for tanneries. 3.3. Characterization of biofilms The biofilms were analyzed periodically at regular intervals (24, 48 and 72 h). The developed biofilms were observed for their increase in biomass, stability and morphology variations. 3.3.1. Biomass The biomass quantification was estimated by bacterial enumeration method. The biomass of biofilms on different used are: coarse sands > pebbles > glass beads and it was 4.8 × 107 (after 24 h), 4.5 × 107 (after 48 h) and 3.5 × 105 (after 72 h) CFU/cm2 respectively (Table 1). The biofilm formation was more on coarse sand and thereby bacterial biomass was more in coarse sand than pebbles and glass beads. The rough surface and hydrophobic nature may be a probable reason for it. 3.3.2. Stability The biofilm stability studies were carried out on different substrates under force flow condition. The biomass withstanding the force flow was calculated as the stable biofilm. After 24 h of continuous flow glass beads the stable biomass was found to be 2.4 × 104 out of 3.5 × 105 in glass beads, 3.6 × 106 out of 4.8 × 107 in coarse sands and 3.1 × 106 out of 4.5 × 107 CFU/cm2 in pebbles. The more stable biofilm was found on coarse sand and pebbles than glass beads. 3.3.3. Morphology Bacterial biofilms of B. subtilis VITSCCr01 and B. cereus VITSCCr02 were obtained individually and as consortium in all the substrates and were analyzed with and without chromium using SEM, AFM and CLSM. The bacterial biofilms interacted with Cr(III) and uninteracted biofilms were analyzed for SEM imaging. Fig. 2a shows the diversity
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Fig. 2. (a) SEM imaging of Bacillus subtilis and B. cereus consortium biofilms without Cr(III). (b) SEM imaging of B. subtilis and B. cereus consortium biofilms with Cr(III) interaction. (c) AFM imaging of immature biofilms of B. subtilis and B. cereus consortium during biofilm formation. (d) AFM imaging of mature biofilms of B. subtilis and B. cereus consortium. (e1) CLSM imaging of B. subtilis and Bacillus cereus consortium biofilms without Cr(III). (e2) 3D structural imaging of B. subtilis and B. cereus consortium biofilms without Cr(III). (f1) CLSM imaging of B. subtilis and B. cereus consortium biofilms with Cr(III) interaction. (f2) 3D structural imaging of B. subtilis and B. cereus consortium biofilms with Cr(III) interaction.
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Fig. 3. (a) Bioremoval of Cr(III) from synthetic chromium solutions by Bacillus subtilis and Bacillus cereus consortium biofilms supported on coarse sand. (b) Bioremoval of Cr(III) from synthetic chromium solutions by B. subtilis and B. cereus consortium biofilms supported on pebbles. (c) Bioremoval of Cr(III) from synthetic chromium solutions by B. subtilis and B. cereus consortium biofilms supported on glass beads.
Fig. 4. (a) Bioremoval of Cr(III) from chrome tanning effluent by Bacillus subtilis and B. cereus consortium biofilms supported on coarse sand. (b) Bioremoval of Cr(III) from chrome tanning effluent by B. subtilis and B. cereus consortium biofilms supported on pebbles. (c) Bioremoval of Cr(III) from chrome tanning effluent by B. subtilis and B. cereus consortium biofilms supported on glass beads.
3.4. Chromium bioremoval studies using biofilms of the mixed culture of B. subtilis and B. cereus, with the presence of rod shaped bacteria in multilayers. The bacterial cells covered with EPS (Exopolymeric substances) layers were seen during biofilm formation on all the substrates. Extracellular polymeric material released by the bacteria was more in biofilms interacted with chromium. ‘The precipitation of insoluble Cr(III) may be covered by extracellular polymeric materials in the bacterial biofilms. The cells after interaction with Cr(III) showed a clear difference in the cell morphology and size enlargement (pleomorphism) and surface modification (Fig. 2b). There was surface modification changes between biofilms developed after 24 h than that after 48 h (Fig. 2b and c). The Cr(III) interacted biofilms on coarse sand showed more thickness of biomass than pebbles and glass beads. Fig. 2e1 shows average projections of CLSM z-series images from uninteracted biofilms on coarse sand and Fig. 2e2 shows the thickness of biofilm with z-stacks three-dimensional (3D) reconstruction. Thickness of the uninteracted biofilms was calculated as 78 ± 2.55 m. The thickness of biofilm decreased to 65 ± 3.21 m for Cr(III) interacted biofilms (Fig. 2f1 and f2) after 24 h on coarse sand. The biomass were comparatively less when the biofilms developed in the presence of Cr(III) this in turn affect the biofilm thickness.
The bioremoval of Cr(III) in different concentrations (25, 50, 75 and 100 mg/l) of synthetic chromium solution were studied using biofilms of B. subtilis and B. cereus consortium. Biofilms supported on coarse sand showed maximum chromium bioremoval than pebbles and glass beads. The Cr(III) bioremoval percentage for biofilms coated on coarse sand was 100% Cr(III) at 25, 50 and 75 mg/l synthetic chromium solution. In 100 mg/l synthetic chromium solution the bioremoval was 98% after 24 h (Fig. 3a). Biofilms coated on pebbles and glass beads showed 86 and 79% Cr(III) removal at 100 mg/l concentration of synthetic chromium solution (Fig. 3b and c). The chromium bioremoval by biofilms were more effective than the planktonic cells. The Cr(III) bioremoval was also studied with real chrome tanning (1:10 and 1:100 dilution) using biofilms of B. subtilis and B. cereus consortium. The Cr(III) concentration in raw effluent was 3400 mg/l so it was diluted for bioremoval studies. Biofilms supported on coarse sand showed maximum Cr(III) bioremoval from tannery effluent compared to pebbles and glass beads. Biofilms coated on coarse sand were able to remove 92% Cr(III) for 1:10 dilution and 94% Cr(III) at 1:100 dilution after 24 h (Fig. 4a). Biofilms coated on pebbles showed 81% and 75% Cr(III) removal for 1:10 and 1:100 dilution after 24 h respectively. Biofilms on glass beads
K. Sundar et al. / Journal of Hazardous Materials 196 (2011) 44–51
49
Fig. 5. (a) SEM imaging and chromium mapping of Bacillus subtilis VITSCCr01 and Bacillus cereus VITSCCr02 biofilms after chromium interaction by SEM-EDX. (b) EDX spectra of B. subtilis VITSCCr01 and B. cereus VITSCCr02 biofilms after chromium interaction, Si peak due to glass slide and Al peak originates from sample holder.
removed 78% and 72% Cr(III) at 1:100 and 1:10 dilutions respectively (Fig. 4b and c). The physico-chemical properties of chrome tanning effluent before and after biofilm treatment are given in Table 2. There was a
Table 2 Physico-chemical properties of raw and biofilm treated tannery effluent. Parameters
Concentration (mg/l) ± SD before treatment
Concentration (mg/l) ± SD after treatment
Color Odor Temperature (◦ C) pH Conductivity ( ) Total dissolved solids Biochemical oxygen demand Chemical oxygen demand Chloride Sulphite Sulphate Ammonia nitrogen Total nitrogen Total hardness Sodium
Dark green Foul smell 25.50 ± 0.33 2.5 ± 0.61 10.50 ± 0.71 815 ± 30.0 650.00 ± 56.9 1264.00 ± 85.9 680.00 ± 10.0 21.00 ± 1.5 33.00 ± 5.0 142.00 ± 8.0 729.00 ± 1.0 460.00 ± 32.0 44.70 ± 5.2
Colorless No odor 25.50 ± 0.23 5.5 ± 0.85 4.40 ± 0.75 523 ± 18.0 150.00 ± 16.0 460.00 ± 35.0 430.00 ± 11.0 15.00 ± 1.8 12.00 ± 4.4 82.00 ± 6.0 426.00 ± 1.5 180.00 ± 22.0 24.70 ± 3.2
significant difference in the phsico-chemical properties of chrome tanning effluent after biofilm treatment. SEM-EDX showed higher percentage of chromium in the biofilms of B. subtilis VITSCCr01 and B. cereus VITSCCr02 than other elements. The percentage of chromium concentration present was 92.60 wt% and 90.43 at% (Table 3). Fig. 5a shows SEM imaging and chromium mapping of B. subtilis and B. cereus consortium biofilms after synthetic Cr(III) interaction by SEM-EDX. EDX spectra of B. subtilis and B. cereus consortium biofilms after chromium interaction showed, Si peak due to substrate (coarse sand), Ti and Al peak originates from sample holder (Fig. 5b). Biosorption of Cr(VI) by a Bacillus coagulans biofilm supported on granular activated carbon (GAC) removed 46.86% Cr(VI) [17]. Lameiras et al. [18], studied biosorption of Cr(VI) using a bacterial biofilm supported on granular activated carbon and on zeolite and obtain 42% biosorption. Quintelas et al. [10], have performed bioremoval of Cr(VI) in a pilot-scale bioreactor through a biofilm of Arthrobacter viscosus supported on GAC. Smith and Gadd [19] studied Cr(VI) removal using SRB (sulphate reducing bacteria) biofilms. There are no reports on the bioremoval of Cr(III) in a pilot scale bioreactor through bacterial biofilms. This study deals with the bioremoval of trivalent chromium from real tanning effluents using chromium tolerant bacterial biofilms. As these bacterial strains were adapted to higher chromium and salt ion concentrations they
50
K. Sundar et al. / Journal of Hazardous Materials 196 (2011) 44–51
Table 3 EDX elemental analysis of developed biofilms of Bacillus subtilis VITSCCr01 and Bacillus cereus VITSCCr02. Element line
Net counts
Net counts error
wt%
±62 ±31 ±88 ±34 ±18 ±22 ±36 10 ±17
57 856 7137 934 298 306 975 10 24
Na K Al K Si K KK KL Ti K Ti L Cr K Cr L
2.40 6.54 – 4.96 – 2.20 – – 92.60
Total
100
can be used in bioremoval of chromium from tannery effluent, which is used to have high chromium and salt ions. 3.5. Adsorption isotherms The adsorption studies were conducted at a fixed biosorbent dosage, by varying the initial concentration of trivalent chromium in the effluent by 1:10 and 1:100 dilutions for 4, 8, 12, 16, 20 and 24 h. According to the Langmuir adsorption isotherm the adsorbed
Cf/qe
a
30 28 26 24 22 20 18 16 14 12 10 8 6 4 2 0
Coarse sand - 34 mg/l Pebbles - 34 mg/l Glass beads - 34 mg/l
12.54 8.42 – 4.96 – 2.15 – – 90.43
Formula Na Al K Ti
Cr
100
layer will be only one molecule thick and all the sides of the adsorbent will have equal affinity for the metal ions. Therefore the presence of adsorbed molecule at one side will not affect the adsorption of molecules at the adjacent side. The Langmuir correlation coefficient (R2 ) values are 0.97783, 0.95704 and 0.9825 for coarse sand, pebbles and glass beads respectively at 1:100 dilution (Fig. 6a). For 1:10 dilution the Langmuir correlation coefficient (R2 ) values are 0.96196 (coarse sand), 0.93674 (pebbles) and 0.92921 (glass beads) (Fig. 6b). The Freundlich adsorption isotherm assumes that the adsorbent consists of a heterogeneous surface, which is composed of different classes of adsorption sites. The correlation coefficient values (R2 ) and the Freundlich constants Kf and n are calculated and as the values of intensity of adsorption isotherm (1/n) are greater than 1 (1/n > 1), the sorption process is known to be homogeneous. Further the applicability of Langmuir and Freundlich adsorption isotherms is derived on the basis of correlation coefficient (R2 ), as in this case the R2 values for Langmuir plot is higher than the Freundlich plot the sorption obeys Langmuir adsorption isotherm.
4. Conclusion
0
b
at%
2
4
6
8
10
12
14
16
18
20
22
Cf 45
Coarse sand - 340 mg/l Pebbles - 340 mg/l Glass beads - 340 mg/l
40 35
Present study attempted a novel approach towards bioremoval of Cr(III) in a plot scale bioreactor design using indigenous Bacillus biofilms. Since Cr(III) poses a threat to humans and environment, it is pertinent to study a biofilm based chromium remediation strategy. Simplified flow through system was designed for biofilm development on economically feasible substrates for Cr(III) removal. This study gives an insight into in situ remediation of chromium, as these bacterial strains were indigenous for tannery environment. These design and substrates can be recycled for desorption of Cr(III) and could be continuously reused for further Cr(III) bioremoval in tanning industries.
30
Acknowledgement
Cf/qe
25
Authors thank the Department of Science and Technology (DST), govt. of India for funding this project and also the management of VIT University for supporting this research.
20 15 10
References
5 0 20
40
60
80
100 120 140 160 180 200 220 240
Cf Fig. 6. (a) Langmuir adsorption isotherm of 1:100 dilution chrome effluent (34 mg/l Cr) for coarse sand, pebbles and glass beads. (b) Langmuir adsorption isotherm of 1:10 dilution chrome effluent (340 mg/l Cr) for coarse sand, pebbles and glass beads.
[1] A.M. Chaudry, S. Ahmad, M.T. Malik, Supported liquid membrane technique applicability for removal of chromium from tannery wastes, Waste Manag. 17 (1988) 211–218. [2] M.M. Bailey, J.G. Boohaker, R.D. Sawyer, J.E. Behling, J.F. Rasco, J.J. Jernigan, R.D. Hood, J.B. Vincent, Exposure of pregnant mice to chromium picolinate results in skeletal defects in their offspring, Birth Defects Res. B: Dev. Reprod. Toxicol. 77 (2006) 244–249. [3] A. Levina, P.A. Lay, Chemical properties and toxicity of chromium (III) nutritional supplements, Chem. Res. Toxicol. 21 (2008) 563–571.
K. Sundar et al. / Journal of Hazardous Materials 196 (2011) 44–51 [4] I. Sharma, D. Goyal, Adsorption kinetics: bioremoval of trivalent chromium from tannery effluent by Aspergillus sp Biomass, Res. J. Environ. Sci. 4 (2010) 1–12. [5] S.S. Ahluwalia, D. Goyal, Microbial and plant derived biomass for removal of heavy metals from wastewater, Bioresour. Technol. 98 (2007) 2243–2257. [6] L.S. Clesceri, A.E. Greenberg, A.D. Eaton, M.A.H. Franson, Standard methods for the examination of water and wastewater, 20th ed., American Public Health Association (APHA), Washington, DC, USA, 2005. [7] K. Sundar, A. Mukherjee, N. Mohammed Sadiq, Chandrasekaran, Cr(III) bioremoval capacities of indigenous and adapted bacterial strains from Palar river basin, J. Hazard. Mater. 187 (2011) 553–561. [8] P. Gerhardt, R.G.E. Murray, W.A. Wood, Methods for General and Molecular Bacteriology, American Society for Microbiology, Washington, DC, 1994. [9] G.D. Christensen, W.A. Simpson, J.J. Younger, L.M. Baddour, F.F. Barrett, D.M. Melton, E.H. Beachey, Adherence of coagulase-negative staphylococci to plastic tissue culture plates: a quantitative model for the adherence of staphylococci to medical devices, J. Clin. Microbiol. 22 (1995) 96–1006. [10] C. Quintelas, B. Fonseca, B. Silva, H. Figueiredo, T. Tavares, Treatment of chromium(VI) solutions in a pilot-scale bioreactor through a biofilm of Arthrobacter viscosus supported on GAC, Bioresour. Technol. 100 (2009) 220–226. [11] R. Aravindhan, B. Madhan, J.R. Rao, B.U. Nair, T. Ramasami, Bioaccumulation of chromium from tannery wastewater: an approach for chrome recovery and reuse, Environ. Sci. Technol. 38 (2004) 300–306.
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[12] A. Malik, Metal bioremediation through growing cells, Environ. Int. 30 (2004) 261–278. [13] A. Scott, S.J. Palmer, Cadmium bio-sorption by bacterial exopolysaccharide, Biotechnol. Lett. 10 (1998) 21–24. [14] C. Quintelas, Z. Rocha, B. Silva, B. Fonseca, H. Figueiredo, T. Tavares, Biosorptive performance of an Escherichia coli biofilm supported on zeolite NaY for the removal of Cr(VI), Cd(II), Fe(III) and Ni(II), Chem. Eng. J. 152 (2009) 110–115. [15] C. Quintelas, Z. Rocha, B. Silva, B. Fonseca, H. Figueiredo, T. Tavares, Removal of Cd(II), Cr(VI), Fe(III) and Ni(II) from aqueous solutions by an E. coli biofilm supported on kaolin, Chem. Eng. J. 149 (2009) 319–324. [16] Y.H. Lin, C.L. Wu, H.L. Li, C.H. Hsu, Verification of model for adsorption and reduction of chromium(VI) by Escherichia coli 33456 using chitosan bead as a supporting medium, Appl. Math. Modell. 35 (2011) 2736–2751. [17] C. Quintelas, B. Fernandes, J. Castro, H. Figueiredo, T. Tavares, Biosorption of Cr(VI) by a Bacillus coagulans biofilm supported on granular activated carbon (GAC), Chem. Eng. J. 136 (2008) 195–203. [18] S. Lameiras, C. Quintelas, M.T. Tavares, Development of a biosorption system for chromium (VI) using a Arthrobacter viscosus biofilm supported on granular activated carbon and on natural zeolites, Bioresour. Technol. 99 (2008) 206–801. [19] W.L. Smith, G.M. Gadd, Reduction and precipitation of chromate by mixed culture sulphate-reducing bacterial biofilms, J. Appl. Microbiol. 88 (2001) 983–991.
Journal of Hazardous Materials 196 (2011) 52–58
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Improvement of gaseous pollutant photocatalysis with WO3 /TiO2 heterojunctional-electrical layered system Yuan Liu, Changsheng Xie ∗ , Huayao Li, Hao Chen, Tao Zou, Dawen Zeng State Key Laboratory of Material Processing and Die & Mould Technology, Nanomaterials and Smart Sensors Laboratory, Department of Materials Science and Engineering, Huazhong University of Science and Technology, Wuhan 430074, PR China
a r t i c l e
i n f o
Article history: Received 23 June 2011 Received in revised form 27 August 2011 Accepted 29 August 2011 Available online 2 September 2011 Keywords: Holes enrichment Photocatalysis Gas phase Synergistic effect Thermodynamics
a b s t r a c t Since the photogenerated holes play a much more important role than electrons in gas-phase photocatalysis, it is better to enrich the holes in the surface of a material system. Here, a novel [interdigital electrode/WO3 /TiO2 ] heterojunctional-electrical layered (HEL) system is proposed to realize this attempt. The HEL system consists of interdigital electrode, WO3 layer and TiO2 layer, and they are orderly printed onto the alumina substrate from bottom to top using the technology of screen printing. It is surprise that the synergistic effect of layered heterojunction and external low bias can strengthen the separation of electron–hole pairs in both TiO2 and WO3 , and enrich the TiO2 surface layer with photogenerated holes to degrade the gaseous pollutants. In comparison with the pure TiO2 film, a 6-fold enhancement in photocatalytic activity was observed using the HEL system by applying a very low bias of 0.2 V. Furthermore, the results also showed that the remarkable improvement could not be obtained when either the WO3 layer or the low external bias was absent. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Photocatalytic technology is an effective pathway for solving the problems of energy crisis and environmental pollution [1,2]. TiO2 semiconductor has been frequently investigated owing to its high oxidation ability, inexpensive, nontoxicity, and chemical stability. However, as a photocatalyst, the relatively high rate of electron–hole recombination often results in a low quantum yield and poor efficiency of photocatalytic reactions [3–6]. Among the ways of enhancing the separation of photogenerated charge carriers, constructing heterojunction structures [7–9] and applying external biases [10–12] are both effective methods to improve the photocatalytic activity of TiO2 . However, the two methods suffer severe constraints when they are applied in the gas-phase photocatalysis, respectively. With regard to the heterojunction structure, it is generally prepared by the simple physical mixing, such as ball milling and sol–gel method, which produces a random mixture of different nanoparticles [6,13]. Thus far, WO3 coupling has been widely studied to improve the photocatalytic performances of TiO2 , since WO3 has the suitable matching band potential to serve as electron accepting species. However, in the confirmation of degradation of gaseous volatile organic chemicals (VOCs), the photocatalytic activity of TiO2 /WO3 composite is not obviously enhanced than that of pure
∗ Corresponding author. Tel.: +86 27 8755 6544; fax: +86 27 8754 3778. E-mail address:
[email protected] (C. Xie). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.067
TiO2 , and it is even weakened with the increased amount of WO3 [14–16]. Recently, both theoretical and experimental studies indicate that the reduction reactions initiated by the electrons are much less effective than the oxidation reactions conducted by the holes which can induce complete decomposition of gaseous pollutants at room temperature [17–21]. Therefore, in the mixed WO3 /TiO2 composite, WO3 as an electron accepting species not only poorly contributes to the gaseous photocatalysis, but also reduces the surface area between TiO2 and the gaseous pollutants. Hence, the presence of WO3 can only be accounted for enhancing the photocatalytic activity, when its content is significantly less than TiO2 [13,16]. As a result, the interface of the heterojunction for charge separation cannot be sufficiently utilized in the mixed form. It can be seen that the layered structure with the hole accepting species facing the VOCs may be more promising than the mixed form. With regard to the application of an external bias voltage, the applied bias on the catalyst can draw the photogenerated electrons away via the external circuit, leaving holes for mineralization of organic pollutants by their oxidation, which is the main characteristic of photoelectrocatalytic (PEC) process [22,23]. Until now, the PEC degradation researches are mainly carried out in aqueous solution, but rarely in gas phase. This is because that catalyst efficiency of traditional TiO2 was improved with very high bias (∼100 V), which needed special potentiostat or voltage booster and consumed more electric energy [24,25]. In the view of low carbon economy, it is obviously more worthwhile to deal with the improvement by applying a relative low bias (<0.5 V).
Y. Liu et al. / Journal of Hazardous Materials 196 (2011) 52–58
Fig. 1. Schematic diagram of the [interdigital electrode/WO3 /TiO2 ] HEL system (Al2 O3 substrate, Au interdigital electrode layer, WO3 layer and TiO2 layer from bottom to top).
Taking the above into consideration, we firstly introduce a new strategy for sufficiently utilizing the photogenerated holes by combining the two methods in one photocatalytic system for gas-phase application, which is defined as heterojunctional-electrical layered (HEL) system. When a very low bias is needed and the holes can be enriched in the surface layer, this system can well settle the bottlenecks of the two methods applied in the gas-phase photocatalysis, respectively. In this study, the system consists of interdigital electrode, WO3 layer and TiO2 layer, and the three functional layers are laminated from bottom to top using the technique of screen printing (schematically shown in Fig. 1). Based on the experimental results, the charge transportation in terms of the thermodynamics in this system is discussed in detail. We expect that the proposed photocatalytic model for the gas phase application may open a new avenue for the design of high efficient photocatalyst with a powerful oxidative ability, and give a novel sight in the combination of heterojunctional photocatalytic and PEC techniques. 2. Experimental 2.1. Materials TiO2 (Degussa P25, average grain size: 30 nm), WO3 (Tianjin Kermel Chemical Reagent Co. Ltd., China, average grain size: 80 nm) and other chemicals used in the experiment were of analytically pure grade. All of them were used as received without any pretreatment. 2.2. Samples preparation and characterization There are many methods to prepare semiconductor film, e.g., sol–gel, sputter deposition, coprecipitation, etc. However, few works are reported to investigate the photocatalyst film by screen printing technique, which is a well-known process in the ceramic industry [26]. In this study, the preparation of [interdigital electrode/TiO2 /WO3 ] HEL system was based on the screen-printing technique. The designed sample is schematically shown in Fig. 1. Firstly, the chemical powder (TiO2 or WO3 ) and a certain amount of organic carrier (a 55:30:10:4:1 wt% combination of terpineol, butyl carbitol, di-n-butyl phthalate, span 85, and ethyl-cellulose) were mixed in a 7:3 wt ratio. All of them were put into the agate ball milling tank. After ball milling for 4 h at the speed of 300 rpm, the suitable pastes used for the process of screen printing were obtained. Additional details of paste preparation are available in our previous report [27]. Secondly, the WO3 paste was printed onto the Au interdigital electrode which had been preprinted on the alumina substrate (Zhuhai Yueke Jinghua Electronic Ceramics CO. LTD., China). After drying, the TiO2 paste was printed onto the WO3 layer. The thickness of each photocatalyst layer was about 10 m, which could be controlled by the screen printing machine. Then, the samples were preheated at 200 ◦ C for 15 min to eliminate the organic carrier.
53
In the end, the samples were sintered at 550 ◦ C for 2 h. Following the above procedures, the [interdigital electrode/WO3 /TiO2 ] HEL system could be successfully obtained, in which the Al2 O3 substrate, interdigital electrode, WO3 layer and TiO2 layer were laminated from bottom to top. Furthermore, the [interdigital electrode/TiO2 /WO3 ], [interdigital electrode/TiO2 /TiO2 ], [interdigital electrode/WO3 /WO3 ] parallel samples were also prepared by the same method. The prepared samples were characterized by field-emission scanning electron microscopy (FSEM, Sirion 200, FEI), X-ray diffraction (XRD, X’Pert PRO, PANalytical B.V.), UV–vis DRS spectra (Lambda 35, PerkinElmer) and photoluminescence spectra (PL, USB2000-FLG Ocean Optics Spectrometer). 2.3. Evaluation of photocatalytic activity Photocatalytic activities of the prepared samples under UV light were evaluated by the following method. The samples could be installed in a gas reactor system. The total volume of this reactor was 825 ml. The whole printed area of the photocatalytic film was 50 mm × 50 mm, which was irradiated by a flat-type LEDlight (Shenzhen Ti-times Electronics Co. Ltd., China, wavelength: 365 ± 5 nm). The power of the light irradiated to the surface of photocatalytic sample was 5 mW/cm2 . In our experiments, the gaseous toluene was used as a model pollutant, which was a persistent indoor volatile harmful gas. Prior to a test, 250 ppm toluene gas (dry air and toluene mixture) was allowed to enter the reactor until the toluene reached adsorption and desorption equilibrium with the catalyst and the reactor. The initial concentration of toluene was 250 ppm (±5 ppm). Each experiment was lasted for 30 min. The analysis of toluene, carbon dioxide concentration in the reactor was conducted on line with a Photoacoustic IR Multigas Monitor (Model 1412; INNOVA Air Tech Instruments). The photocatalytic activity of different samples can be quantitatively evaluated by comparing the apparent reaction rate constants. In a heterogeneous solid–gas reaction, the photocatalytic degradation of toluene is a pseudo-first-order reaction, and its kinetic equation may be expressed as follows [28,29]: ln
C0 = kt Ct
where k is the apparent rate constant, C0 and Ct are the initial and the reaction concentration of toluene, respectively. 2.4. Measurements of photoelectric response To evaluate the separation ability of photoexcited electron–hole pairs in the semiconductor materials, photoelectric property measurement was performed. Photocurrent characterization was carried out with a flat-type LED-light (365 ± 5 nm) as the UV light source. The light intensity was about 5 mW/cm2 . The platform was developed independently by our laboratory, which had the testing photocurrent range in 10−8 to 10−3 A. In this study, the photocurrent was measured in the dry air and the amplitude illuminated for 270 s was chosen as a parameter to assess the photoelectric response of the material. The detailed descriptions of the platform and the test procedures can refer to our previous work [30,31]. 3. Results and discussion 3.1. Film characteristics The detailed descriptions of the prepared samples are shown in Table 1. Designations are used in the following discussions. [Interdigital electrode/WO3 /TiO2 ], [interdigital
54
Y. Liu et al. / Journal of Hazardous Materials 196 (2011) 52–58
Table 1 Designations and structure descriptions of the as-prepared samples. Sample
[IE/WO3 /TiO2 ]
Designation Substrate Underlayer Toplayer Film thickness
S-W/T S-T/W Al2 O3 substrate printed with Au interdigital electrode WO3 TiO2 TiO2 WO3 20 m in total (toplayer = underlayer = 10 m)
[IE/TiO2 /WO3 ]
[IE/TiO2 /TiO2 ]
[IE/WO3 /WO3 ]
S-T
S-W
TiO2
WO3
20 m
20 m
IE: interdigital electrode.
electrode/TiO2 /WO3 ], [interdigital electrode/TiO2 /TiO2 ] and [interdigital electrode/WO3 /WO3 ] samples are called S-W/T, S-T/W, S-T and S-W, respectively. The surface morphologies and phase components of the S-W/T prepared by the screen-printing technique are shown in Fig. 2. The SEM photograph in Fig. 2a shows the cross-sectional morphology of the WO3 /TiO2 layered structure. The printed film is about 20 m in thickness, in which the TiO2 and WO3 layers are both about 10 m. An apparent interface between two layers is observed, indicating that it is two-layer type heterojunction and is different from the mixture type in a composite system. From the insert of Fig. 2a, we can see that the film has lots of micro-holes, which are formed due to volatilization of organic component during the heat-treatment process and originated from agglomerated structures in the screen printing paste. These micro-holes let the film to have a good absorption property for the light and gaseous chemicals [12]. Moreover,
Fig. 3. X-ray diffraction patterns of as-prepared (a) S-T, (b) S-W and (c) S-T/W parallel samples.
sintering neck as an important structure for composite photocatalyst is formed between two contacted grains. It is considered as the chemically bounded interface caused by ionic thermal diffusion to allow the smooth charge transfer [9,14,32]. The XRD patterns in Fig. 2b illustrate that the phase components of the S-W/T sample are not changed after the heat-treatment at 550 ◦ C. No other phases are found in the sample, suggesting that there is no appreciable chemical reaction between TiO2 and WO3 . The S-T/W, S-T, S-W samples have the similar results, as shown in Fig. 3.
Fig. 2. (a) A typical cross-sectional view SEM image of the as-prepared S-W/T sample, the inset in showing the top surface SEM image of it; (b) X-ray diffraction patterns of the as-prepared S-W/T sample.
Fig. 4. UV–vis diffuses reflectance spectra of the as-prepared S-T, S-W, S-W/T and S-T/W samples.
Y. Liu et al. / Journal of Hazardous Materials 196 (2011) 52–58
Fig. 5. The comparisons of degradation rate among the samples without bias under UV-light (365 nm) irradiation.
Fig. 4 indicates the UV–vis diffuse reflectance spectral of the four samples. The S-T and S-W samples exhibit absorption below 390 and 450 nm, which correspond to their absorption edges (Eg = 3.2 eV for TiO2 and Eg = 2.8 eV for WO3 ), respectively. Furthermore, the SW sample has higher absorption intensity than S-T sample in both UV and visible regions. For coupled films, the S-W/T and S-T/W samples show the similar absorption profiles as their toplayers. In addition, S-W/T sample has higher absorption intensity than S-T sample in the visible region. This suggested that the visible light could penetrate the TiO2 toplayer to excite the WO3 underlayer because of a lot of micro-holes in the film. To simplify the following discussion about the charge separation, the UV (365 nm) light is used for the photocatalytic activity and photocurrent tests, since the S-W/T and S-T samples have the same absorption intensity at this wavelength. 3.2. Photocatalytic activity without bias
55
the photogenerated electron–hole pairs in the WO3 film. In a word, the photoexcited electrons of WO3 cannot be used for one-electron chemical reaction effectively at room temperature. In addition, the recombination of the photogenerated electron–hole pairs is further enhanced. This fact lets us to believe that WO3 is unsuitable for achieving the efficient oxidative decomposition of organic compounds in air. Fig. 6 shows the PL spectra of S-W and S-T samples under excitation at 300 nm. The PL signals of semiconductor materials result from the recombination of photoinduced charge carriers. In general, the lower PL intensity indicates the lower recombination rate of photoinduced electron–hole pairs and the higher photocatalytic activity of semiconductors. As can be seen in Fig. 6, TiO2 shows the lower and narrower PL intensity than WO3 . Therefore, the PL results in Fig. 6 are consistent with the photocatalytic activity of the samples in Fig. 5. It is concluded that the rapid recombination of photogenerated electron–hole pairs is directly harmful to the photocatalytic efficiency. Importantly, it is noted that the S-W/T sample exhibits the best photocatalytic activity, though the S-W/T sample has the same absorption of UV light as the S-T sample. Therefore, the increase in the photocatalytic activity can be attributed to the charge separation between TiO2 and WO3 . A mechanistic scheme of the charge separation for the S-W/T system is shown in Fig. 7. Both TiO2 and WO3 are n-type semiconductor with suitable bandgap energies, strongly absorbing UV light. Upon photon-excitation, electron–hole pairs are generated in each semiconductor film, expressed as the process 1 and process 2, respectively: 1
TiO2 + hv−→eTiO2 − + hTiO2 2
+
WO3 + hv−→eWO3 − + hWO3
+
O2 (g) + e− → O− 2 (ad),
− 4.216 eV versus Vacuum
eTiO2 − −→eWO3 − + 4
hWO3 −→hTiO2
+
(4) (5)
So far, variety of the TiO2 /WO3 coupled systems have been widely reported. As shown in Fig. 7a, only the above four processes
(1)
It is generally considered that the CB level of a semiconductor should be more negative than the potential for the single-electron reduction of oxygen in order to allow efficient consumption of photoexcited electrons and subsequent oxidative decomposition of organic compounds by holes to proceed in air [34]. Otherwise, photoexcited electrons probably recombine with photogenerated holes, resulting in the high probability of the fast recombination of
(3)
The valence band (VB) edges of TiO2 and WO3 are situated at −7.41 and −7.94 eV versus Vacuum level, respectively. The CB edge of WO3 (−5.24 eV) is much lower than that of TiO2 (−4.21 eV) [33]. In terms of the thermodynamics, electrons can flow into the WO3 underlayer, while holes oppositely diffuse into the TiO2 toplayer, expressed as the process 3 and process 4, respectively: 3
The photocatalytic activity of the as-prepared samples is evaluated by the photocatalytic degradation of toluene in air. Blank experiments indicate that the photocatalytic reactions do not take place in the absence of either photocatalysts or UV light. Fig. 5 shows the photocatalytic removal of toluene versus irradiation time and the apparent rate constants are 3.03, 7.39, 13.57 and 20.04 × 10−3 min−1 for the S-W, S-T/W, S-T and S-W/T, respectively. Based on Fig. 5, the order of photocatalytic activity is SW < S-T/W < S-T < S-W/T. The S-W sample has a much lower photocatalytic activity than S-T sample, although it has higher absorption intensity. The possible reason for the low rate constant of S-W sample is associated with the poor photocatalytic activity of WO3 . The WO3 has a narrow band of 2.8 eV and a low potential of the conduction band (CB) bottom (−5.24 eV versus Vacuum) [33]. So that the photo-generated electrons do not have enough reducibility to react with the molecule oxygen at room temperature, expressed as the following equation [34]:
(2)
Fig. 6. Excitation dependent PL of S-W and S-T samples.
56
Y. Liu et al. / Journal of Hazardous Materials 196 (2011) 52–58 Table 2 The apparent rate constants and the amount of evolved CO2 when the S-T and S-W/T samples are unbiased and 0.2 V biased. Samples
S-T S-W/T
Rate constant (10−3 min−1 )
Evolved CO2 (ppm)
Unbiased
0.2 V biased
Unbiased
0.2 V biased
13.57 20.04
13.88 84.73
458 609
469 1186
resulting in the large recombination probability (process 6). Then, the transportation of the holes from the valance of WO3 to that of TiO2 is further suppressed (process 4). As a result, the higher photocatalytic activity of the TiO2 /WO3 coupled films is just due to the lower recombination rate in the TiO2 toplayer (process 5). Thus, the above discussions can deeply explain the higher photocatalytic activity of the WO3 /TiO2 layered film. The detailed production, recombination and transportation processes of charge carriers are schematically illustrated in Fig. 7b. For the reverse coupled TiO2 /WO3 film (S-T/W sample), the WO3 toplayer accumulated with electrons gets in touch with the gaseous toluene. As noticed in Fig. 5, the photocatalytic activity of the ST/W sample is higher than S-W sample but lower than S-T sample. This proves that the contribution of the electrons to the photocatalytic activity does not have so much as the holes. It should be also mentioned that the film prepared by the screen-printing method, as indicated in the insert of Fig. 2a, is micro-porous. Therefore, a portion of gaseous toluene may penetrate the WO3 toplayer to contact with the TiO2 layer. In this case, both WO3 and TiO2 are attributed to the photocatalytic reaction, resulting in a little higher photocatalytic activity than single WO3 film. 3.3. Photocatalytic activity with a low bias
Fig. 7. Schematic diagram for energy band matching and flow of electrons and holes in the S-W/T system. (a) The recombination processes are generally ignored in previous reports. (b) No bias is supplied. (c) A low bias of 0.2 V is supplied. Note that the process is drawn as a dotted line when it is inhibited.
were considered to discuss the charge transfer in the correlative references. Nevertheless, the process of electron–hole recombination that is directly related to the photocatalytic ability was not carefully concerned. The recombination of the photogenerated electron–hole pairs in the TiO2 layer and WO3 layer is expressed as the process 5 and process 6, respectively: + 5
eTiO2 − + hTiO2 −→energy + 6
eWO3 − + hWO3 −→energy
(6) (7)
As is well known, in n-type semiconductor, the mobility of the electrons is much higher than that of the holes [35,36]. When the TiO2 /WO3 coupled system is illuminated by the UV-light, the photogenerated electrons in the CB of TiO2 would be very rapidly and easily transferred to the CB of WO3 (process 3). However, the process 4, that the holes move in the opposite direction from the electrons and are trapped in the TiO2 , should be slow and inefficient. That is to say, the WO3 would be accumulated with more electrons, but the holes of WO3 are not efficiently separated,
The application of an external bias voltage on the catalyst can draw the photo-generated electrons away via the external circuit, leaving the holes for mineralization of organic pollutants by their oxidation [37,38]. Therefore, compared to the unbiased photocatalytic process, the probability of the rapid recombination of electron–hole pairs in the PEC process can be reduced, and the photo-degradation ability can be enhanced. The processes that the electrons in the printed catalyst are driven to the external circuit by the biased interdigital electrode, has been reported in our previous work [25]. The observations in Section 3.2 indicate that the holes generated in the WO3 are not sufficiently utilized for degrading the pollutants because of the suppression of the process 4. We believe that, when the electrons in the WO3 could be further migrated to the external circuit (process 7), the recombination rate in the WO3 layer would be reduced, and process 6 would be inhibited (as shown in Fig. 7c). Then, the process 4 can be smoothly operated, and the TiO2 layer could obtain much more holes from the WO3 layer. As a result, the photocatalytic activity of the S-W/T system could be further enhanced. To prove this hypothesis, a very low bias voltage of 0.2 V was applied on the interdigital electrode to draw the accumulated electrons of the WO3 away via the external circuit. Fig. 8 illustrates the dependence of Ct /C0 with irradiation time over the S-T and S-W/T samples when 0.2 V bias is supplied. Surprisingly, the photocatalytic activity of S-W/T sample was enhanced markedly by applying a low bias voltage of 0.2 V. At this condition, the rate constant of S-W/T sample with 0.2 V bias was 5.2 times larger than that of S-T sample in the initial stage, and the evolved CO2 was improved by 2.6 times (as shown in Table 2). It is also found that the reduced toluene is completely decomposed into CO2 , when the activity is obviously improved. As well, the results are consistent with the explanations in Fig. 7c.
Y. Liu et al. / Journal of Hazardous Materials 196 (2011) 52–58
57
Table 3 The photocurrent responses of different samples by application a bias voltage of 0.2, 0.5, 1, 2 V, respectively. U (V)
S-T I (A)
S-W I (A)
S-T/W I (A)
S-W/T I (A)
0.2 0.5 1 2
4.0 × 10−8 1.1 × 10−7 3.1 × 10−7 9.2 × 10−7
1.6 × 10−6 4.0 × 10−6 7.0 × 10−6 1.3 × 10−5
1.1 × 10−7 3.6 × 10−7 7.4 × 10−7 3.5 × 10−6
2.0 × 10−6 5.4 × 10−6 1.2 × 10−5 2.2 × 10−5
U: bias voltage; I: photocurrent amplitude.
Fig. 8. Comparison of photocatalytic decomposition rates of toluene between S-T and S-W/T with no bias and 0.2 V bias.
Note that the S-T sample within pure TiO2 film was not enhanced by applying a low bias voltage of 0.2 V (as shown in Table 2). Lately, Ye et al. found that the TiO2 was applied with 82.5 V bias voltage, and the degradation rate constant could be raised as 1.26 times as TiO2 without bias [24]. As can be seen, TiO2 needs much higher bias than WO3 /TiO2 coupled film to force the electrons away via external circuit. To explain this phenomenon, the photoelectric response of the S-T and S-W was measured (Fig. 9). Under the same bias voltage of 0.2 V, the photoelectric response of WO3 is two to three orders of magnitude larger than that of TiO2 . In addition, we have noticed that the photoelectric response of TiO2 rises abruptly to a maximum and then falls down to a steady state. However, the photoelectric response of WO3 increases slowly, and the saturation of photocurrent is not attained in the testing process. The effect of lower photoconductivity of TiO2 is attributed to removal of conduction electrons by adsorbed O2 . It is well known that TiO2 has a high potential of CB bottom (−4.21 eV versus Vacuum), so the free electrons have powerful reducibility. Oxygen molecules adsorbed on the TiO2 surface could react with free electrons, creating negatively charged O2 − ions [39]. Thereby, a depletion layer of electrons is created with low conductivity near the grain boundary. The energy supplied by the external bias voltage should be high enough to force the electrons to cross the electrons potential barrier. It can be concluded that the higher
Fig. 9. Photocurrent–time testing curves of S-T and S-W samples during the UV-light illumination process adding 0.2 V bias in dry air.
potential barrier is presented, and the higher bias is needed. In addition, the needed bias has the exponential growth with the height of potential barrier [40]. Thus the photocurrent of TiO2 would fall down after reaching a maximum value. This is why a high bias is needed for pure TiO2 to drive the electrons away and promote the degradation activity. Here, for the S-W/T sample, the photogenerated electrons in the CB of TiO2 could be transferred to the CB of WO3 . Then, the electrons could be much easily driven to the external circuit through the WO3 layer at a low voltage, which is an unimpeded conduction passageway for carrier transporting. For this possibility, the photoelectric responses of different samples were measured by applying a bias voltage of 0.2, 0.5, 1, 2 V, respectively (as shown in Table 3). All the samples had relatively higher response amplitude by applying a higher bias voltage. Furthermore, the photoelectric response of WO3 is much larger than that of TiO2 at any bias voltage. The application of a larger bias voltage is mainly to enable the electron within the semiconductor across the barrier of the grain boundary between two crystals. However, it would hinder the practical application of the photocatalyst when the bias is too high. Therefore, the realization of low bias PEC degradation would promote the rapid development of this technology in practical application and meet the concept of low carbon economy. 3.4. Synergistic effect for charge separation combined with layered heterojunction and external low bias In this study, it is demonstrated that the [interdigital electrode/WO3 /TiO2 ] HEL system applied with a very low bias of 0.2 V can achieve remarkable improvement of gaseous pollutant photocatalysis compared with the single TiO2 film. Both the interdigital electrode and the WO3 layer play important roles in the HEL system. Either of them is indispensable for the multilevel separation of photoinduced electron–hole pairs as illustrated in Fig. 7c. Without the external bias, the holes generated in the WO3 cannot be sufficiently utilized by the TiO2 toplayer. Without the WO3 layer, very high bias is needed for the single TiO2 film to draw the electrons away via external circuit. Therefore, the [interdigital electrode/WO3 /TiO2 ] HEL system combined with layered heterojunction and external low bias can generate a huge synergistic effect, resulting from strengthening the separation of the electron–hole pairs and showing a surface layer enriched with holes to react with the gaseous pollutants. On the basis of the above experimental results, in order to distinctly define the HEL system, three functional layers are concluded as follows: (1) TiO2 as a wide bandgap semiconductor can generate photoinduced holes with a powerful oxidative ability. Under UV-light irradiation, the TiO2 toplayer can enrich with holes from itself and the WO3 layer to degrade the gaseous pollutants. (2) WO3 as a narrow bandgap semiconductor has an excellent electric conductivity and its CB and VB both lie under those of the TiO2 . Under UV and visible light irradiation, the WO3 layer can
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draw the electrons accumulated in its own CB and has potential to supply holes for the TiO2 toplayer. (3) Interdigital electrode is supplied with a low bias, and it can further separate the electrons in the WO3 layer due to its good conductivity. This process can benefit the TiO2 toplayer to obtain much more holes from the WO3 underlayer. 4. Conclusion In the presented work, the hole enriched in the surface of a material system was emphasized for photocatalysis in gas-phase. For this target, an ordered HEL system was proposed to get the holes sufficiently utilized. Furthermore, it was found that layered heterojunction and external low bias in our HEL system showed a huge synergistic effect by the suggested multilevel separation of photoinduced electron–hole pairs. The photocatalytic activity was remarkably improved in degrading the persistent toluene. Compared with the researches of mixed heterojuntion and single phase photoelectrocatalysis with high bias, the HEL system may supply some interesting enlightenments in gas-phase photocatalysis. Acknowledgements This work was supported by the Nature Science Foundation of China (No. 50927201) and the National Basic Research Program of China (Grant Nos. 2009CB939705 and 2009CB939702). The authors are also grateful to Analytical and Testing Center of Huazhong University of Science and Technology. References [1] X.B. Chen, S.H. Shen, L.J. Guo, S.S. Mao, Semiconductor-based photocatalytic hydrogen generation, Chem. Rev. 110 (2010) 6503–6570. [2] U.G. Akpan, B.H. Hameed, Parameters affecting the photocatalytic degradation of dyes using TiO2 -based photocatalysts: a review, J. Hazard. Mater. 170 (2009) 520–529. [3] W. Zhu, X. Liu, H.Q. Liu, D.L. Tong, J.Y. Yang, J.Y. Peng, Coaxial heterogeneous structure of TiO2 nanotube arrays with CdS as a superthin coating synthesized via modified electrochemical atomic layer deposition, J. Am. Chem. Soc. 132 (2010) 12619–12626. [4] M.H. Zhou, J.G. Yu, B. Cheng, Effects of Fe-doping on the photocatalytic activity of mesoporous TiO2 powders prepared by an ultrasonic method, J. Hazard. Mater. 137 (2006) 1838–1847. [5] A. Fujishima, X.T. Zhang, D.A. Tryk, TiO2 photocatalysis and related surface phenomena, Surf. Sci. Rep. 63 (2008) 515–582. [6] H.J. Huang, D.Z. Li, Q. Lin, W.J. Zhang, Y. Shao, Y.B. Chen, M. Sun, X.Z. Fu, Efficient degradation of benzene over LaVO4 /TiO2 nanocrystalline heterojunction photocatalyst under visible light irradiation, Environ. Sci. Technol. 43 (2009) 4164–4168. [7] Y.T. Kwon, K.Y. Song, W.I. Lee, G.J. Choi, Y.R. Do, Photocatalytic behavior of WO3 -loaded TiO2 in an oxidation reaction, J. Catal. 191 (2000) 192–199. [8] Y. Bessekhouad, D. Robert, J. Weber, Bi2 S3 /TiO2 and CdS/TiO2 heterojunctions as an available configuration for photocatalytic degradation of organic pollutant, J. Photochem. Photobiol. A 163 (2004) 569–580. [9] X.P. Lin, J.C. Xing, W.D. Wang, Z.C. Shan, F.F. Xu, F.Q. Huang, Photocatalytic activities of heterojunction semiconductors Bi2 O3 /BaTiO3 : a strategy for the design of efficient combined photocatalysts, J. Phys. Chem. C 111 (2007) 18288–18293. [10] M.V.B. Zanoni, J.J. Sene, H. Selcuk, M.A. Anderson, Photoelectrocatalytic production of active chlorine on nanocrystalline titanium dioxide thin-film electrodes, Environ. Sci. Technol. 38 (2004) 3203–3208. [11] X. Zhao, J.H. Qu, H.J. Liu, C. Hu, Photoelectrocatalytic degradation of triazinecontaining azo dyes at gamma-Bi2 MoO6 film electrode under visible light irradiation lambda >420 Nm, Environ. Sci. Technol. 41 (2007) 6802–6807. [12] X. Zhao, Y.F. Zhu, Synergetic degradation of rhodamine B at a porous ZnWO4 film electrode by combined electro-oxidation and photocatalysis, Environ. Sci. Technol. 40 (2006) 3367–3372. [13] W. Smith, Y.P. Zhao, Enhanced photocatalytic activity by aligned WO3 /TiO2 two-layer nanorod arrays, J. Phys. Chem. C 112 (2008) 19635–19641. [14] V. Keller, P. Bernhardt, F. Garin, Photocatalytic oxidation of butyl acetate in vapor phase on TiO2 , Pt/TiO2 and WO3 /TiO2 catalysts, J. Catal. 215 (2003) 129–138.
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Journal of Hazardous Materials 196 (2011) 59–65
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Electrochemical antimony removal from accumulator acid: Results from removal trials in laboratory cells M.E. Henry Bergmann a,1 , A. Savas Koparal b,∗ a b
Hochschule Anhalt, FB 6/7, Bernburger Str. 55, D-06366 Koethen/Anh., Germany Anadolu University, Iki Eylul Campus, Department of Environmental Engineering, 26555 Eskisehir, Turkey
a r t i c l e
i n f o
Article history: Received 3 November 2010 Received in revised form 28 March 2011 Accepted 30 August 2011 Available online 12 September 2011 Keywords: Antimony Electrochemical deposition Accumulator acid Antimony removal
a b s t r a c t Regeneration of spent accumulator acid could be an alternative process for crystallization, neutralisation and disposal. Therefore, for the first time in a study of the possibilities of electrochemical removal of antimony and accumulator acid regeneration on a laboratory scale, two synthetic and several real systems containing sulfuric acid of concentrations ranging between 28% and 36%, and antimony species were tested. Discontinuous electrochemical reactors with anion exchange membranes were successfully used in these experiments, which were conducted at a temperature of 35 ◦ C. Removal of antimony using cells that were not divided by a separator, however, was not possible. In selected experiments, by varying the electrode material, type of electrolyte, and cell current, the concentration of antimony could be reduced from the range of 5 ppm to 0.15 ppm. This resulted in current efficiencies between 0.00002% and 0.001%, and in specific electroenergy demands between 100 Wh L−1 and 2000 Wh L−1 . In other experiments on substances with antimony contents up to 3500 mg L−1 , the current efficiencies obtained were more than a thousandfold higher. In contrast to the formally high relative energy consumption parameters absolute demand parameters are relatively small and favour the electrochemical method in small scale application. Besides plate electrodes, 3D-cathodes were used. Copper- and graphite cathodes produced the best results. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The US Environmental Protection Agency (USEPA) and the European Union (EU) view antimony (Sb) and its compounds as pollutants whose removal is a priority [1]. Antimony is ubiquitously present in the environment as a result of natural processes and activities of humans [1]. Elevated concentrations of antimony in soils and sediments are either related to the anthropogenic sources of antimony or associated with the high arsenic concentrations present in sulfudic ores. Also, of importance appears to be the subject of airbone pollution by antimony of aquatic and terrestrial systems [1,2]. Antimony is known to be a special element employed in the semiconductor industry and as an additive in antiinflammatory agents. It has been used, in the past, for enhancing the hardness and mechanical stability of lead alloys [3]. Antimony is a constituent of highly concentrated acids, in used accumulator
∗ Corresponding author. E-mail addresses:
[email protected] (M.E.H. Bergmann),
[email protected] (A.S. Koparal). 1 Tel.: +49 03496 67 2313; fax: +49 03496 67 2642. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.073
acids, and hydrometallurgical systems; it is usually encountered in small concentrations [4,5]. Antimony has been the focus of environmental studies for decades. However, the environmental behaviour and effects of this poorly studied metalloid have only recently evoked interest as a consequence of its elevated concentrations in the vicinity of smelters, chemical plants, and mining and mineralised areas [6]. But recent studies on antimony contamination have usually been confined to contaminated areas near smelters [7], and old mining areas [8]. Suggested methods for antimony removal were mostly ion exchange and extraction [9–11]. In electrochemistry, the application of polarography is known even for deposition of antimony [12–15]. Results of kinetic studies using different electrodes in electrochemical processes for antimony removal have been published by the authors earlier [16]. That its removal by this process is possible was clearly demonstrated. However, it was also found that the simulation of real systems in kinetic studies was difficult. As preliminary work showed [16] deposition of antimony by this process depends on the species present in the electrolyte, which may vary in their composition. The deposition process is a multi-step mechanism that involves adsorption effects, formation of intermediates,
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Fig. 1. Crossections of cells used: (a) flow-through cell with anion exchange membrane, (b) box cell with anion exchange membrane without convection, and (c) glass beaker cell with ceramic diaphragm. (1) Anode half cell, (2) cathode half cell, (3) anion exchange membrane, (4) 3D cathode, (5) cathodic feeder or cathode in studies without 3D electrode, (6) plate IrO2 anode in (a), graphite rod anode in (c), (7) ceramic beaker, (8) carbon felt cathode and (9) magnetic stirrer.
and electron transfer. If additives are present, complexation of antimony is possible. Therefore, further detailed studies are necessary to clarify mechanisms. For simplicity, and calculation of energy consumption in the deposition process may be calculated from Eq. (1) using 3 electrons for Sb(III) reduction: Sb(III) + 3e− ↔ Sb
(1)
The complicated chemistry of antimony and the usage of a large number of electrolytes lead to the possibility of using various removal mechanisms. However, only comprehensive and integrated experimental methods can give more insights. There are a few studies considering acid regeneration by several methods mostly not focusing on metal species removal. Ion exchange and electrodialysis are used to obtain 10% H2 SO4 in battery recycling [17]. Recycling industry prefers neutralisation technologies at first stages of recycling. Patent literature also describes thermal decomposition and synthesis processes (US Patents (US Patent 5628976 and 5730950)). The work presented here was carried out against following background: recycling enterprises accumulate spent sulfuric acid from accumulators. Both neutralisation/disposal and replacement are probable more expensive than a selective regeneration, i.e. antimony removal from the system. In addition, a potential user wished to combine a practicable technology with solar energy supply. Therefore, experiments were planned with the aim - to check an electrochemical technology for antimony removal in cathodic processes - to evaluate energy demands focusing on small-scale treatment units, which would be able to treat in m3 per week range. As results, in this study, results obtained from depletion experiments on laboratory scale are presented. Two synthetic electrolyte systems were tested together with other technical systems (spent lead/lead dioxide accumulator acid). Different electrode materials and types of reactors were employed according to their specific requirements to check the feasibility of efficient removal. As mentioned above the study also includes assessment of the energy consumption required for antimony removal in small scale. 2. Experimental Besides preliminary trials with glass beakers and cells using platinum plate electrodes (20 mm × 20 mm) and platinum mesh cathodes, discontinuous experiments were first carried out using a
flow-through cell composed of two cylindrical half cells (Fig. 1a). The chambers for the anolyte and catholyte were cylindrical in form and each had an inner diameter of 30 mm and a depth of 25 mm. Circular plates of nearly same diameter and of 2 mm thickness served as anode (IrO2 mixed oxide electrode from Heraeus) as well as cathode or cathodic current feeder (electrolytic grade copper). In certain experiments, 3D cathodes made of titanium mesh (Heraeus), copper wire wool (wire diameter 0.2 mm), carbon felt (Sigri), and electrode graphite particles with a grain size of 3 mm were used. (Prior to experiments, copper wool was washed in hot 5% Na2 CO3 solution, washed with deionized water, activated by 10% sulfuric acid and washed thrice with deionized water.) The cell was divided into two halves using an anion exchange membrane (SEPRON MA-3475). The electrolyte, stored in glass beakers (0.25–0.5 L), was pumped through the electrode chambers with the help of centrifugal pumps. Due to energy dissipation, the working temperature quickly rose to and stabilized around 35 ◦ C. Regulation of temperature was not resorted to because it remained nearly unchanged over time. Samples were analysed at different intervals. A second laboratory cell using the same separator was designed as an open box (Fig. 1b) having a volume of 75 mL per electrolyte chamber. A third was a stirred divided laboratory cell. An aqueous solution of 100 mL of 36% H2 SO4 containing trivalent antimony with concentrations of 1500 and 3500 mg L−1 was used. In this cylindrical cell, a carbon felt electrode (Sigri) and graphite rod electrode passed through the centre of a ceramic diaphragm (thickness 2 mm) were used (Fig. 1c). The projection of the cathode area was 55 cm2 . The cell which was of diameter 60 mm had a cathodic volume of 200 mL. Anode compartment was filled with Agar. The Agar was prepared by adding 2.5 g Agar powder to 50 mL boiled water. By the addition of Na2 SO4 , its concentration was adjusted to 4 M. Ceramic porous diagram was filled with this gel. Rod electrode was placed in the centre as the anode. Thus electrode bridge was formed by the conductive gel. Gel obtained allows for the electron transfer whilst avoiding the solution in the cathode compartment to pass through the anode compartment. This system is more cost effective than the anode cathode separation by membrane. It has an ease of preparation. When desired it can be removed and replaced after reactor maintenance [18]. The concentration range of H2 SO4 used in the experiments corresponded well with that found in typical Pb/PbO2 accumulators. Precision current supplies from Steiber Mechanic or from Statron were used to provide a constant flow of electric current. This current is given for all experiments. In addition, data for current density were calculated using front area of plate cathode. For
M.E.H. Bergmann, A.S. Koparal / Journal of Hazardous Materials 196 (2011) 59–65
3,5
which does not deposit at the cathode. This is in congruence to electroanalytical conditions [20]. It was concluded that experiments had to continue using undivided (separated) electrochemical cells.
3
Sb concentration, mg L-1
61
2,5
3.2. Experiments in separated cells at relatively low antimony concentration
2 1,5
Pt plate electrode, 100 mA (150 A/m²) Pt net cathode, 500 mA (700 A/m²)
1
Ti fillings, 500 mA (700 A/m²)
0,5 0 0
10
20
30
40
50
60
70
Time, h Fig. 2. Decrease in antimony concentration in discontinuous experiments using electrolysis cell according to Fig. 1a and varying current, and varied electrode material (V = 200 mL). Current density for 3D cathode was calculated with respect to separator area.
experiments using 3D electrode material current density was calculated for comparison with respect to separator area as usual. Commercial accumulator acid (36%) served as the electrolyte, which could be diluted with deionized water when needed. Due to the fact that synthetic antimony solutions cannot completely represent real systems [16] real spent acid was used in many experiments as received by the supplier. In selected experiments, various antimony concentrations were prepared by dissolving Sb2 O3 (Clech/Poland) in hot sulfuric acid. Tartaric acid and King’s water (5 mL nitric acid in 500 mL 36% sulfuric acid) were used to prepare the electrolytes. Addition of tartaric acid stabilizes antimony species by avoiding precipitation. At a first step of preparation, Sb2 O3 was dissolved in 70 ◦ C hot King’s water. Then, sulfuric acid and tartaric acid were added. The second synthetic system was prepared by adding Sb standard (Baker 6968-04) to 36% H2 SO4 . A Nanocolor 100D instrument was used for spectrophotometrical analysis of the antimony [19]. Polarographical analysis was carried out using a Metrohm polarograph, model 646 [19]. Atomic absorption spectroscopy (AAS) was used in the experiments with high antimony concentration (Varian AA+250). The samples used were diluted for this analysis. Experiments were duplicated. 3. Results and discussion 3.1. Experiments in non-divided cells In preliminary experiments, electrochemical deposition was tried in cells with no separator placed between the anode and cathode. This cell type is mostly the cheapest variant and preferred in electrochemical metal winning processes. Fig. 2 shows some chosen results using Pt and Ti as the electrode material. Both materials were chosen because the electrolyte conditions are highly corrosive. Current densities when given are calculated with respect to the cathode area or to the projected separator area (Section 3.2). For comparison, maximum specific charge consumed per liter at 60 h was between 30 and 150 Ah L−1 . Electrolyte losses were compensated by periodic additions of acid. That even platinum is applicable for antimony deposition was demonstrated in an earlier paper [16]. But, the results show that with both cathode materials only a negligible depletion of antimony occurred. (Furthermore, hydrogen evolution starts at much lower cathode potential compared with other materials.) The reason for this could be the partial oxidation at the anode of Sb(III) to Sb(V),
Prediction of deposition behaviour is difficult. Usual factors such as electrode material, surface state, type and concentration of species inside electrolyte, mass transfer conditions and electrode potential (or potential distribution in electrodes of 3D character) have significant influence on deposition quality and efficiency. As it will be explained later the term efficiency can be used in systems, which are poor in reactants, only as a relative characterizing parameter. Parameters used for example in metal winning industry (current efficiency, specific energy demand) must be completed by absolute characteristics considering process and wastewater treatment. This work intended in receiving feasibility information varying and screening promising electrode materials and working conditions as used in plating processes. To prevent the oxidation of the Sb(III) species at the anode, a separator made of SEPRON anion exchange membrane was tested. In control experiments no antimony (detection limit 0.03 mg L−1 ) could be detected in the anolytes. In further experiments, spent accumulator acid (32%) at the same concentration but without the additional antimony dosing was used. Because it is known in electrochemical engineering that 3D electrodes may show better metal removal properties, materials with extended specific surface were used in addition. Fig. 3a shows with one exception some satisfying curves obtained using three-dimensional cathodes. The highest specific charged consumed in these experiments was about 220 Ah L−1 . All cathode materials had again relatively high corrosion resistance against sulfuric acid. The curve relating to spent acid experiments does not, in principle, differ from the one for dissolved Sb2 O3 without admixtures. With the exception of the carbon felt electrode, the other cathodes (of graphite and of copper wool) showed good potential for deposition of antimony. The obviously insufficient current distribution inside the carbon felt was responsible for the reduced activity of the material where a surface block caused by a gas bubble may possibly have occurred [21]. In general, the depositing ability of 3D cathodes depends on additional factors such as electrolyte and bed conductivity, electrochemical kinetics, contact to the feeding plate and hydrodynamics. This variety of conditions makes prediction impossible. However, this material showed better results, when antimony concentration was chosen much higher (see later in Fig. 7). At small antimony concentrations, the removal curves decrease more gradually as was expected. It is also obvious that the removal rate was proportionally higher. This is because the products of antimony covered the cathode, at least in the initial period of experiment, at sufficiently high hydrogen overvoltage [16]. Fig. 3b is related with the same experimental settings for two electrode materials varying current load and initial concentration. The influence of current is marginal because an electrochemical process in lower ppm range of reacting species is usually mass transfer controlled. Additionally, larger amount of gas may hinder deposition inside 3D structures. Both Fig. 3a and b are characterized by high depletion rates in the time period shortly after starting the experiments when deposition occurs still on the initial cathode material. Later, the cathode is coated by the deposition having worse antimony reduction ability. From this behaviour an important conclusion can be drawn: enlarged removal efficiencies could be reached if cathodes with renewing surface are used. There are not so much cells in the field. Famous representatives are the so-called cells of Chemelec type [22] and the VMPB (vertically
62
M.E.H. Bergmann, A.S. Koparal / Journal of Hazardous Materials 196 (2011) 59–65
0,0016
4 Cu wool, 100 mA (150 A/m²) 28% sulphuric acid
3.5 3 2.5
Cu wool, 150 mA (225 A/m²) spent acid
2 1.5
Carbon felt/Sigri, 150 mA (225 A/m²) spent acid
1 0.5 0
Graphite particle bed, 3 mm, 250 mA (375 A/m²) spent acid
0
50
100
150
Differ. current efficiency, %
Sb concentration, mg L-1
a
0,0014
0.25 A (375 A/m²)
0,0012
0.5 A (750 A/m²)
0,001 0,0008
1.5 A (2250 A/m²)
0,0006 0,0004 0,0002 0
200
0
Time, h
0,5
1
1,5
2
2,5
3
-1
b
Sb concentration, mg L
3.5
Sb concentration, mg L -1
Fig. 4. Calculated differential current efficiency versus antimony concentration with varying current density for discontinuous experiments in laboratory cells divided by anion exchange membrane and using graphite cathode (grain size 3 mm, lead dioxide anode, spent 28% accumulator acid, V = 200 mL, conditions of experiments shown in Fig. 3).
Graphite particle bed, 3 mm, 500 mA, spent acid
3
2.5
Graphite particle bed, 3 mm, 1000 mA, spent acid
2
Graphite particle bed, 3 mm, 1500 mA, spent acid
Cu wool, 250 mA, spent acid
1.5
Cu wool, 150 mA, spent acid
1
0.5
0 0
50
100
150
200
250
300
350
Time, h Fig. 3. Decrease in antimony concentration in discontinuous experiments in a laboratory cell separated by anion exchange membrane (Fig. 1a) varying mainly cathode material (a) and cell current (b) (spent acid, for comparison in one experiment dissolved Sb2 O3 was used as electrolyte, V = 200 mL). Current density was calculated with respect to separator area.
moving particle bed) reactor [23,24]. These constructions can keep cathode surface clean. Negative effects are reactions of redissolution and particle formation making filtration necessary. Additional work could be done but this was not in the scope of this paper. Fig. 4 shows the calculated differential current efficiency for experiments using the graphite cathode and with varying current density and initial antimony concentration. Calculations were carried out using Eq. (2): ϕdiff
CSb · 3 · F · V = I · t · MSb · 1000
(2)
where CSb is the decrease of antimony concentration (mg L−1 ) in an incremental time t (h), I is the current (A), MSb is the molar weight of antimony (g mol−1 ), and F is the Faraday constant (Ah mol−1 ), V is the electrolyte volume (L). The parameter ϕdiff is calculated for comparing different experiments. In general, it is expected that for electrolytes with reacting species at mg L−1 level is extremely low. This is the reason that in electrochemical engineering ϕdiff cannot be used as the only characterizing parameter as well as the specific energy demand (Eqs. (3) and (4)) that results in very high amounts for small ϕdiff . Removal seems to be effective at very small concentrations (1–5 ppm) under the concrete experimental conditions used. The
extremely low current efficiency leads to relatively high specific energy consumption which was, however, not a criterion in this study as it is in many applications of environmental technology where only small quantities of metal are removed from solutions. Deposition of metals at very small concentrations under limited current conditions is generally effective when specialised cathode surfaces are used such as 3D electrodes. If, however, very strong gas evolution effects are observed and the surface is blocked, 3D electrodes can lose their advantage and then both types of electrodes (2D and 3D) become comparable. Table 1 shows as an example of comparisons of electrolysis using cathodes made of plates, copper wool, as well as graphite particles (3 mm grain size) with a volume of 0.2 L. The specific electroenergy demand was calculated by Eq. (3): WSb =
U·n·F MSb · ϕfinal
(3)
where WSb is the specific electroenergy consumption (Wh g−1 ), U is the cell voltage (V), F is the Faraday constant (Ah mol−1 ), n is the number of the transferred electrons, ϕfinal is the integral current efficiency taking into account cycle time of electrolysis necessary for reaching the final antimony concentration as indicated in Table 1. When WSb was calculated in Wh L−1 , Eq. (4) was applied for a given depletion C over t, where V is the solution volume (L): WSb =
U · I · t V
(4)
Specific energy demand appears extremely high, but as the amounts of antimony removed are small, the total electricity consumption is expected to be relatively low. Parameters such as current efficiency and specific energy consumption are here given because common in considering deposition processes. But as explained above in electrochemical process water treatment in ppm range of concentration, both parameters may be irritating because of associating high total energy demand. Energy demand or economical efforts per cubic meter of treated solutions are parameters of better choice. A first estimation may explain this: the assumption of 0.150 kWh per treated liter results in electricity costs for electrolysis of 18.75 D m−3 at specific electricity costs of 0.125 D kWh−1 . This is relatively low compared, for example, with 370 D m−3 as only chemical costs in sulfuric bath neutralisation (ISTC Fact Sheet [25]) and demonstrates the economical potential of electrochemical methods even at very low concentrations.
M.E.H. Bergmann, A.S. Koparal / Journal of Hazardous Materials 196 (2011) 59–65
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Table 1 Comparison of experimental parameters and results from discontinuous experiments using 2D- and 3D cathodes and technical- and synthetic accumulator acid in a cell as shown in Fig. 1b.
Current, A Cathode material Initial Sb concentration, mg L−1 Final Sb concentration, mg L−1 Electrolysis time, h Integral current efficiency, % Specific electroenergy consumption Wh g−1 Wh L−1
Fixed bed reactor
Fixed bed reactor
Plate reactor
0.25 Cu 3.08 (technical spent accumulator acid) 0.3 180 0.00081
0.25 Graphite 3.7 (technical spent accumulator acid) 0.4 180 0.00097
0.5 Cu on Pt 3.5 (36%-sulfuric acid) 0.3 72.5 0.00044
81,566 225
33,375 110
453,103 1450
The experiment using Cu and Pt (where Cu was electrochemically deposited prior to commencement of the experiment) is shown for comparison with experiments using 3D electrodes. The concentration decay is comparable but it was reached in quicker time (72.5 h), three times faster at the higher current and without convection. This necessarily affected the energy parameters, which are ten times higher. The experiment shows an extreme case of possible treatment of the electrolyte, applying suitable cathode potential. More results are shown for experiments when 36% sulfuric acid was spiked with antimony standard (Fig. 5). In the separated cells, no antimony was deposited on copper, titanium, and platinum, especially at small current densities. Colouring of the electrolyte was observed at the final stage of experiments indicating anode and/or steel cathode corrosion. Better results were obtained when a copper cathode was used at high current densities. In fact, high current density had a very positive effect on antimony depletion. At the same time, however, the cell potential and the specific electroenergy consumption showed an increase. 3.3. Experiments in separated cells at relatively low antimony concentration For theoretically possible technologies using pre-concentration of antimony one has to expect higher process efficiency in electrochemical deposition due to better mass transfer and current efficiency. To demonstrate this, experiments were carried out at higher antimony concentration. The concentration factor was about
1000. This means that per liter about 2000 kJ would be needed what results in energy demand of approximately 0.5 kWh per liter initial solution. Higher concentration significantly improves energetic parameters of electrochemical deposition. For example, differential current efficiency is dependent on antimony concentration as may be seen from the two experiments shown in Fig. 6. Towards the end of the experiment, the integral current efficiency amounted to 0.000293%, and the specific current consumption to 676773 Wh g−1 or 2300 Wh L−1 . Because both values are relatively high, smaller concentration range or relatively higher final concentrations should be arranged in practical treatment. Under conditions of higher initial Sb concentration, materials such as carbon felt, which showed insufficient behaviour lower ppm range of concentration, were applicable (Fig. 7). Antimony concentration chosen was 3500 mg L−1 . Calculated current efficiency values were more than a thousandfold higher compared with experiments shown in Table 1. Also, energy consumption was a thousandfold lower (Fig. 7). Cheaper ceramic diaphragm could replace the more expensive ion exchange membrane. It can be seen from Fig. 7, that antimony can be removed efficiently. More details are given in [26] but one example is given here to estimate the total costs. If assumed from the studies that both preconcentration and deposition results in 600 kWh/m3 (Fig. 7), 75 D electricity costs under German industrial conditions can be calculated. From this point of view the method may compete with neutralisation technology. (Investment costs are not considered 0,003
5
Cu on Pt, 500 mA (1250 A/m²)
4
Pb, 25 mA (50 A/m²), 250 mL, convection
3
Stainless steel, 75 mA (30 A/m²)
2
Ti plate, 100 mA (40 A/m²)
1
Cu, 50 mA (30 A/m²)
0
0
20
40
60
80
100
120
Time, h
Differ. current efficiency, %
Sb concentration, mg L-1
6
Cu, 500 mA (350 A/m²)
0,0025
0,002
Stainless steel, 75 mA (30 A/m²)
0,0015
0,001
0,0005
0 0
1
2
3
4
5
-1
Sb concentration, mg L Fig. 5. Antimony concentration decay in discontinuous experiments using different plate electrode materials (cell according to Fig. 1b with 75 mL of electrolyte prepared by Sb standard addition to 36% sulfuric acid, no convection). An experiment using lead cathode with convection is shown for comparison with an electrolyte volume of 250 mL at 34–36 ◦ C. Current density was calculated with respect to cathode area.
Fig. 6. Differential current efficiency versus antimony concentration for two discontinuous experiments in a cell according to Fig. 1b without convection (antimony from dissolved Sb2 O3 in 36% sulfuric acid, V = 200 mL). Current density was calculated with respect to cathode area.
64
M.E.H. Bergmann, A.S. Koparal / Journal of Hazardous Materials 196 (2011) 59–65
a
For comparison and better illustration, electrolysis costs were calculated from a curve in Fig. 3b starting at 3.08 mg L−1 antimony in the initial solution. Fig. 8 summarizes specific costs for electrolysis by varying the remaining concentration and electricity costs per kWh between 0.07 and 0.20 D kWh−1 . All results show that treatment costs are still in a reasonable range and that electrolysis can compete with other recuperation or disposal methods.
4000 3500
-1
1500 mg[Sb]/L Sb concentration, mg L
3000
3500 mg[Sb]/L
2500 2000
4. Conclusions 1500
• Removal of antimony by the electrochemical deposition process using separated cells was successfully achieved. • In cells not divided into two halves, antimony was not deposited at the cathode due to the partial oxidation of Sb(III) to Sb(V) at the anode. It can, therefore, be concluded that electrochemical removal of antimony has to be performed in divided cells. • Copper and graphite electrodes were found to be the most suitable electrode materials for deposition of antimony. • Because electrode covering by Sb species lowers deposition efficiency with time periodical replacement or automatic renewing of cathode material is recommended. • Although the calculated specific electroenergy consumption values were relatively high, absolute energy consumption was low because the quantities of antimony removed were small. • Based on the results obtained by laboratory studies a technical anion exchange membrane-divided cell was constructed having a volume of 40 L. This cell is subject of further studies.
1000 8%, 409 kWh/kg
6%, 3116 kWh/kg
500 0 0
1
2
3
4
5
Time, h
Current efficiency, %
b
100 90 80 70 60 50 40 30 20 10
References
0 0
1
2
3
4
5
Time, h Fig. 7. Result of experiments conducted at enhanced antimony concentration. (a) Decrease in antimony concentration in a stirred divided cell (Fig. 1c, 50 A m−2 ). (b) Current efficiency (initial antimony concentration 3500 mg L−1 , 50 A m−2 , V = 200 mL, App. Pot. 3 V).
Electrolysis costs, € m
-3
here because they would have been considered for each treatment technology. Furthermore, in small scale, investment costs are often similar for different technologies.) 20–30 m2 of photovoltaic panels are enough to cover the electricity demand if one week is considered for the whole process.
50
0,07 €/kWh
40
0,15 €/kWh 0,2 €/kWh
30 20 10 0 0
0.5
1
1.5
2
Remaining Sb concentration, mg L-1 Fig. 8. Specific electrolysis cost in dependence on remaining antimony concentration varying specific electricity costs. (Conditions according to one experiment shown in Fig. 3b: starting concentration 3.08 mg L−1 , V = 200 mL, current 0.25 A, voltage 1 V, Cu wool cathode.)
[1] M. Filella, N. Belzile, Y.W. Chen, Antimony in the environment: a review focused on natural waters. I. Occurrence, Earth Sci. Rev. (2002) 57125–57176. [2] E. Merian (Ed.), Metals and Their Compounds in the Environment, VCH, Weinheim, 1991. [3] D.R. Lide (Ed.), CRC Handbook of Chemistry and Physics, 78th ed., CRC Press, NY, 1997–1998. [4] S. Ylasaari, O. Forsen, Einfluss von Nickel und Antimon auf die kathodische Arsenabscheidung bei der Entkupferungselektrolyse, Neue Huette 34 (1989) 181–185. [5] S. Ubaldini, F. Veglio, P. Fornari, C. Abbruzzese, Process flow-sheet for gold and antimony recovery from stibnite, Hydrometallurgy 57 (2000) 187–199. [6] M. Tighe, P. Ashley, P. Lockwood, S. Wilson, Soil, water, and pasture enrichment of antimony and arsenic within a coastal flood plain system, Sci. Total Environ. 347 (2005) 175–186. [7] N. Ainsworth, J.A. Cooke, M.S. Johnson, Distribution of antimony in contaminated grassland. 1. Vegetation and soils, Environ. Pollut. 65 (1990) 65–77. [8] W. Hammel, R. Debus, L. Steubing, Mobility of antimony in soil and its availability to plants, Chemosphere 41 (2000) 1791–1798. [9] K. Ando, N. Tsuchida, Recovering Bi and Sb from electrolyte in copper electrorefining, J. Miner. Met. Mater. Soc. 49 (1997) 49–51. ˇ ´ Antimony availability in highly [10] V. Ettler, M. Mihaljeviˇc, O. Sebek, Z. Nechutny, polluted soils and sediments—a comparison of single extractions, Chemosphere 68 (2007) 455–463. [11] S.M. Saleh, S.A. Said, M.S. El-Shahawi, Extraction and recovery of Au, Sb and Sn from electrorefined solid waste, Anal. Chim. Acta 436 (2001) 69–77. [12] H. Braun, M. Metzger, Umweltanalytische Antimon-Bestimmung durch inverse Wechselstromvoltammetrie mit der Quecksilberfilmelektrode, Fresen. Z. Anal. Chem. 320 (1985) 241–246. [13] H.J. Haase, Elektrochemische Stripping-Analyse, VCH, Weinheim, 1996. [14] G. Jung, C.K. Rhee, Two electrochemical processes for the deposition of Sb on Au(1 0 0 0) and Au(1 1 1): irreversible adsorption and underpotential deposition, J. Electroanal. Chem. 436 (1997) 277–280. [15] R.R. Pradhananga, M. Pradhananga, Electrochemical behavior of electroplated and crystalline antimony electrodes, J. Nep. Chem. Soc. 16 (1997) 30–34. [16] H. Bergmann, S. Koparal, Kinetic studies on electrochemical antimony removal from concentrated sulphuric acid systems, Chem. Eng. Technol. 30 (2007) 242–249. [17] Report: Project 8025 (German Ministry of Environmental Protection), 1997. [18] U.B. Ogutveren, A.T. Pekel, Electrochemical generation of cobalt(III) acetate for the oxidation of alkyl aromatics, Bull. Electrochem. 5 (6) (1989) 452–455.
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Journal of Hazardous Materials 196 (2011) 66–72
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Cadmium phytoextraction potential of different Alyssum species R. Barzanti, I. Colzi, M. Arnetoli, A. Gallo, S. Pignattelli, R. Gabbrielli, C. Gonnelli ∗ Department of Evolutionary Biology, Università di Firenze, via Micheli 1, 50121 Firenze, Italy
a r t i c l e
i n f o
Article history: Received 10 January 2011 Received in revised form 22 March 2011 Accepted 30 August 2011 Available online 10 September 2011 Keywords: Alyssum Cadmium Tolerance Accumulation Phytoextraction
a b s t r a c t This work was planned for providing useful information about the possibility of using serpentine adapted plants for phytoextraction of cadmium, element scarcely represented in such metalliferous environment. To this aim, we investigated variation in cadmium tolerance, accumulation and translocation in three Alyssum plants with different phenotypes: Alyssum bertolonii, that is a serpentine endemic nickel hyperaccumulator, and two populations of Alyssum montanum, one adapted and one not adapted to serpentine soils. Plants were hydroponically cultivated in presence of increasing concentrations of CdSO4 for two weeks. For the metal concentration used in the experiments, the three different Alyssum populations showed variation in cadmium tolerance, accumulation and content. The serpentine adapted population of A. montanum showed statistically higher cadmium tolerance and accumulation than A. bertolonii and the population of A. montanum not adapted to serpentine soil thus deserving to be investigated for phytoextraction purposes. Furthermore, as for the kinetic parameters of the cadmium uptake system, A. montanum serpentine population presented a low apparent Km value, suggesting a high affinity for this metal of its uptake system, whereas the Vmax values were not significantly different among the plants. Present data revealed metallicolous plants are also suitable for the phytoremediation of metals underrepresented in the environment of their initial origin. Nonetheless, field trials on real contaminated soils are essential. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Phytoextraction, together with phytostabilization, phytovolatilization, phytodegradation, and other plant-based technologies, is a subset of phytoremediation used to remediate or stabilize contaminants in the environment [1]. In particular phytoextraction is the technology that uses plants to extract elements from polluted or mineralized soils, and accumulate them in harvestable organs and tissues in order to remove the pollutants/contaminants from the field [2]. Plants suitable for phytoremediation purposes have to show tolerance to trace element/s, fast growing rate on polluted soil, and metal accumulation on harvestable organs [3]. Trace element phytoextraction, as any other technology, has both advantages and limitations. The main advantages are: it has reasonable costs, due to plant’s ability to work as a solar-driven pump, extracting and concentrating particular elements from the environment; there is the possibility of trace element recycling, as the ash of some hyperaccumulators can consist of significant
∗ Corresponding author. Tel.: +39 055 2757384; fax: +39 055 282358. E-mail addresses:
[email protected] (R. Barzanti),
[email protected] (I. Colzi),
[email protected] (M. Arnetoli),
[email protected] (A. Gallo),
[email protected] (S. Pignattelli), gabbrielli@unifi.it (R. Gabbrielli), cristina.gonnelli@unifi.it (C. Gonnelli). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.075
quantities of trace elements and there is no need to pay for safe disposal; it works without disturbing further the site, which is of very great importance for its public acceptance [2]. The main limitations of this technology are: it can only be used for low to moderately contaminated soils; its applicability is limited to surface soils at rooting depth; it is limited to the plant available fraction of the trace elements and it is time-requiring [3]. Hyperaccumulators are effective in phytoextracting and/or phytomining metals or other pollutants from contaminated or mineralized soils [4,5]. For these reasons, metal hyperaccumulators can be used as solar-driven ion-pumps, capable of removing and concentrating metals from the substrate [6] thus extracting the metal from polluted soils through the so-called “green” technologies [7]. From the very beginning of the introduction of the trace element phytoextraction concept, the key question about the preferable use of trace element hyperaccumulator plants or of high biomass producing crop species, is still in debate. Chaney et al. [8] reported that trace element hyperaccumulation is a more important trait than dry biomass yield. In support to this statement, they theoretically calculated that zinc removal by hyperaccumulator and high biomass plants and drew that in any case the use of hyperaccumulators resulted in higher trace element extraction. However, the opposite view also exists [3]. Today, approximately 500 trace element accumulating taxa, belonging to at least 100 plant families, have been identified [9].
R. Barzanti et al. / Journal of Hazardous Materials 196 (2011) 66–72
Most of the hyperaccumulator plant species are able to accumulate just one trace element, but there are also multitrace element accumulators. For example, some populations of T. caerulescens are found to have not only high levels of zinc, but also of cadmium, cobalt, and some other trace elements [10], whereas others do not display this capability [11]. One of the aspects that can be improved to increase phytoextraction yield is certainly the selection of the most suitable plant species. In this context, it is important the selection of ecotypes with high biomass production and the optimization of their cultural condition. At the same time the possibility of these plants to also extract metals not present in their natural environment can offer an opportunity that merit to be studied. In this study we evaluated the potential of the serpentine endemic nickel hyperaccumulating plant A. bertolonii Desv. [12] to tolerate and accumulate cadmium, an important environmental contaminant, extremely toxic to living organisms and human health. This metal is scarcely present in serpentine soils, known for having low Ca to Mg ratio and high content of heavy metals, especially nickel, chromium and cobalt [13]. Furthermore, we compared the response to Cd of A. bertolonii to that of two populations of Alyssum montanum L., harvested from serpentine and normal soil in the same region. In this way, we could compare three distinct phenotypes; serpentine adapted nickel hyperaccumulator, serpentine adapted non-hyperaccumulator and non-serpentine adapted, in relation to cadmium tolerance and accumulation. Beyond giving information on the possible use of these plants in phytoextraction techniques, this model system could also shed light on the still debated existence in nature of the co-tolerance phenomenon. Co-tolerance occurs when metal-tolerant plants can also tolerate high and detrimental concentrations of those metals present at low, non-toxic levels in their environment [14]. Studying cadmium response in serpentine adapted plants could provide useful hints on such intriguing topic.
67
Arnon solution [17] diluted 1:10, pH 5.5 ± 0.1 with the following composition: KNO3 0.06 mmol L−1 , Ca(NO3 )2 4H2 O 0.03 mmol L−1 , NH4 H2 PO4 0.01 mmol L−1 , MgSO4 7H2 O 0.02 mmol L−1 , FeSO4 7H2 O 0.18 mol L−1 , tartaric acid 0.9 mol L−1 , H3 BO3 4.6 mol L−1 , MnCl2 4H2 O 0.92 mol L−1 , CuSO4 5H2 O 0.03 mol L−1 , ZnSO4 7H2 O 0.077 mol L−1 , and H2 MoO4 0.06 mol L−1 . The culture conditions were a 12 h (day) photoperiod, provided by Philips TDL 58W/33 fluorescent tubes (160 mol m−2 s−1 ), at 23 ± 1 ◦ C and a relative humidity of 60–65%. 2.2. Determination of cadmium tolerance After germination, the floating trays with seedlings were placed in the same Arnon fresh hydroponic solutions containing 0, 0.25, 0.5, 1, 2, 5 and 10 M CdSO4 . After 11 days of growth, root and shoot lengths of 30 plantlets for each treatment were measured and chosen as a measurement of metal toxic effects on plants [18]. To compare populations among them, we calculated a tolerance index as the ratio of root or shoot elongation on CdSO4 containing solutions to root or shoot elongation on CdSO4 free solutions. For a quantitative estimation of root cadmium tolerance, experimental data points were fitted to a sigmoid curve to obtain four-parameter logistic function and tolerance parameters were calculated. 2.3. Cadmium accumulation The plantlets, grown in the above-mentioned conditions, were rinsed with milliQ-water and the roots were carefully washed with Pb(NO3 )2 10 mM at 4 ◦ C for 10 min to de-absorb metals adhering to the root cell wall. The plantlets were then divided into shoots and roots and processed as described above for cadmium determination. 2.4. Cadmium accumulation in excised roots
2. Materials and methods 2.1. Plant material and experimental conditions Plants and seeds of Alyssum bertolonii Desv. were collected in a serpentine outcrop at Pieve Santo Stefano, Arezzo, Italy (PSS). Plants and seeds of A. montanum L were harvested in the serpentine outcrop of Monterufoli (MR) and from a normal soil in Monte Prata (MP) respectively. Concentration of nickel in the soil has been measured after nitric and perchloric acid digestion [15] and was respectively of 1228 ± 46 g g−1 (PSS), 1480 ± 47 g g−1 (MR) and 37 ± 0.19 g g−1 (MP). The soil pH, measured following the method of Sparks [16], was around 7.1, 7.5 and 7.3 for PSS, MR and MP, respectively. The soils showed traceable amounts of cadmium, ranging from 1 to 3 g g−1 [15]. Ten adult plants, of the same size for each species and population, were randomly collected over the population entire spatial distribution. Roots and shoots were washed with milliQ-water three times, dried at 70 ◦ C for 24 h and then weighed and mineralised by wet ashing with a mixture of concentrated HNO3 and HClO4 (5:2 v/v) on an electric thermostatic plate (300 ◦ C). Certified materials (grade BCR, Fluka Analitycal, Sigma–Aldrich) were the reference samples used to verify the accuracy of the method. The certified reference materials were digested as described above. Nickel and cadmium concentrations in the digests were determined using ICP-OES. Standard solutions were prepared by using available commercial stock solutions (Fluka Analytical, Sigma–Aldrich). Seeds were germinated for 3 days in the dark on floating trays held in 1 l vessels containing 400 ml of continuously aerated
Roots of two-weeks-old plantlets were excised and incubated in solutions containing different CdSO4 concentrations ranging between 0 and 5 M. After 1 h, samples were washed with milliQwater and desorbed in ice-cold (4 ◦ C) Pb(NO3 )2 10 mM for 10 min. Cadmium was determined as described above. 2.5. Statistical analysis All treatments were performed in triplicate and the significance of differences was analysed by one-way and factorial ANOVA followed by a HSD-Tukey test for post-hoc comparisons between unequal samples performed with Statistica 6 (StatSoft, 2003). For the analysis of toxicity data, the curve fitting and the estimate of kinetic parameters, custom-made worksheets and program files for SigmaPlot 8.0 (SPSS Inc., Chicago, IL) were used. 3. Results 3.1. Metal concentration in field collected plants Table 1 shows nickel and cadmium concentrations in roots and shoots of the field collected plants of A. bertolonii and the two populations of A. montanum. The nickel concentrations were statistically different between the three plants (p < 0.001). The highest values were shown by A. bertolonii and the metal amount was far higher in shoots than in roots. The lowest nickel concentrations were shown by A. montanum Monte Prata (MP) and no significant differences were found between roots and shoots, whereas, A. montanum Monterufoli
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R. Barzanti et al. / Journal of Hazardous Materials 196 (2011) 66–72
Table 1 Metal concentration (mean ± standard error) in field collected plants of A. bertolonii and A. montanum. Metal (g g−1 d.w.)
Population A. montanum Monterufoli (MR)
A. bertolonii Pieve S. Stefano (PSS)
Nickel Cadmium
A. montanum Monte Prata (MP)
Roots
Shoots
Roots
Shoots
Roots
Shoots
396 ± 38 0.30 ± 0.04
7040 ± 1036 n.d.
120 ± 15 0.49 ± 0.01
47 ± 16 0.32 ± 0.06
3.68 ± 1.69 0.19 ± 0.04
2.40 ± 0.17 0.19 ± 0.05
(MR) showed higher nickel concentrations in roots than in shoots (p < 0.001). Regarding cadmium concentrations, only small amounts of the metal were detected in roots and shoots of the three plants. 3.2. Effects of cadmium on growth Root and shoot length variation in A. bertolonii and in the two populations of A. montanum in the presence of increasing cadmium (CdSO4 ) concentrations are reported in Fig. 1. Plants from all the three populations showed a decrease in root and shoot growth with increasing cadmium concentration in the nutrient solution. At the lowest exposure level used in the experimentation, all plants already showed a significant reduction in root and shoot growth in respect to plants grown in absence of cadmium (p < 0.001). Tolerance to cadmium treatments is given as index of tolerance (Fig. 2) which is given by the ratio of root or shoot elongation on
CdSO4 containing solutions to root or shoot elongation on CdSO4 free solutions. Root tolerance index did not show any significant difference at the lowest and at the two highest cadmium concentration used, whereas, A. bertolonii Pieve Santo Stefano (PSS) showed a significantly lower tolerance than A. montanum MR at 0.5 M (p < 0.05) and than A. montanum MR and MP at 1 and 2 M (p < 0.001). Also for shoot tolerance index, no significantly differences between the plants were observed at the lowest and at the two highest cadmium concentrations used. A. bertolonii PSS showed to be less tolerant than A. montanum MR and MP at 0.5 and 1 M cadmium concentration, while at 2 M PSS was less tolerant than MP but no significantly different than MR. The negative effect of cadmium treatment on root growth was analysed with a regression model and in all the three plants, data fitted significantly to a logistic dose-response relationship (Table 2). EC50 was calculated as a tolerance parameter; the highest value was shown by A. montanum MR, followed by A. montanum MP and at last A. bertolonii PSS (Table 2).
1.2
10 A. montanum Monte Prata A. montanum Monterufoli A. bertolonii Pieve S. Stefano
8
1.0
A
A
A
Root Cd tolerance
Root lenght (cm)
A 6
4
AB
0.8
B* 0.6
A
B**
A A
B**
0.4
A A
A
2
0.2
0
0.0
A A
Shoot Cd tolerance
Shoot lenght (cm)
AAA AAA
A
1.0
0.5
0.8
A
B* B**
A
B* 0.6
0.4
A
A
1.0 1.5
0.25 µM 0.5 µM 1 µM 2 µM 5 µM 10 µM
A A B**
A A
A
0.2
0.0
0.0 0
2
4
6
8
10
CdSO4 (µM) Fig. 1. Root and shoot length (mean ± standard error) of A. bertolonii and A. montanum two week old-plantlets treated with 0, 0.25, 0.5, 1, 2, 5, 10 M CdSO4 for eleven days.
Pieve Santo Stefano
Monterufoli
Monte Prata
Fig. 2. Root and shoot cadmium tolerance indexes (mean ± standard error) of A. bertolonii and A. montanum two week old-plantlets treated with 0, 0.25, 0.5, 1, 2, 5, 10 M CdSO4 for eleven days. Letters above the histograms indicate significant differences among plants from the three different populations (* p < 0.05, ** p < 0.001).
R. Barzanti et al. / Journal of Hazardous Materials 196 (2011) 66–72
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Table 2 Cadmium tolerance of A. bertolonii and A. montanum plantlets expressed as EC50 (mean ± standard error). r and p values are reported. Significant differences between the means appear with different letters (* p < 0.05). Population
EC50 (M CdSO4 ) a b c
A. bertolonii Pieve S. Stefano (PSS)a
A. montanum Monterufoli (MR)b
A. montanum Monte Prata (MP)c
0.97 ± 0.19 B*
1.86 ± 0.36 A
1.57 ± 0.33 AB
r = 0.99, p = 0.048. r = 0.99, p = 0.014. r = 0.99, p = 0.018.
3.3. Cadmium accumulation The cadmium concentrations in roots and shoots of all three plants increased with increasing cadmium exposure (Fig. 3) and the metal concentrations showed a significant raise (p < 0.05 for roots and shoots of PSS and roots of MR; p < 0.001 for shoots of MR and roots and shoots of MP) already at the first exposure level (0.25 M). Moreover, values showed a saturating trend in all populations both in roots and in shoots. Regarding cadmium concentrations in roots, significant differences (p < 0.05) between the three plants were found and, except for the lowest concentration used, the highest values were shown by A. bertolonii PSS, followed by A. montanum MP and at last MR. Shoot cadmium accumulation was not significantly different between the three populations, except for the two highest exposure levels (2 and 5 M), where A. bertolonii PSS exhibited the lowest concentrations (p < 0.001) and A. montanum MR the highest ones.
Root Cd concentration (µg g-1 d.w.)
3000
2500
2000
1500
1000
3.4. Cadmium content per plant Total cadmium contents per plant, calculated as the product between mean cadmium concentrations found in tissues of each plant and mean dry weight of plantlets, are reported in Table 4. The three plants showed a significant increase in total cadmium amount already at the first exposure level (0.25 M) compared to the control (p < 0.001), both in roots and in shoots, except for roots of A. bertolonii PSS that showed a significant increase at 0.5 M. The variation of cadmium concentration in roots and shoots of all the plants followed a saturating trend similar to that observed in Fig. 3, although in this case A. bertolonii PSS showed the lowest concentration of metal in roots at 1 and 5 M (p < 0.05). At the other exposure levels there were not significant differences between the populations for root cadmium concentrations. Regarding the total cadmium concentration in shoots, A. bertolonii PSS exhibited the lowest values at 1, 2 and 5 M (p < 0.001), while significant differences between the two populations of A. montanum were at 2 and 5 M where MR showed higher values than MP (p < 0.05).
A. montanum Monte Prata A. montanum Monterufoli A. bertolonii Pieve S. Stefano
500
3.5. Cadmium accumulation in excised roots
0
2500
-1
Shoot Cd concentration (µg g d.w.)
Within each plant, a significant linear correlation between cadmium concentration in root and cadmium concentration in shoot was found. As the shoot:root metal concentration ratio never changed in a exposure-dependent way for any of the three populations at the cadmium concentrations used, the angular coefficient of this regression line was used to estimate the shoot:root metal concentration ratio itself (Table 3). The lowest mean value of this coefficient was found in A. bertolonii PSS, followed by A. montanum MP and at last A. montanum MR. Significant difference (p < 0.05) was found between A. bertolonii PSS and A. montanum MR.
2000
1500
1000
500
0 0
1
2
3
4
5
CdSO4 (µM) Fig. 3. Root and shoot cadmium concentration (mean ± standard error) of A. bertolonii and A. montanum two week old-plantlets treated with 0, 0.25, 0.5, 1, 2, 5, 10 M CdSO4 for eleven days.
Fig. 4 reports cadmium accumulation in excised roots of the three plants treated for 1 h with increasing concentrations of cadmium (CdSO4 ). In all cases, it was observed a significant increase of cadmium concentration in excised roots was observed already at the first exposure level in respect to the control (p < 0.05). For most of the concentrations used, A. bertolonii showed the lowest root cadmium concentrations (p < 0.05). Excised roots accumulation data were elaborated in order to compare uptake kinetics of cadmium by roots of the three plants of Alyssum in the different experimental conditions used. Results showed that the kinetic for cadmium influx into roots followed a hyperbolic pattern that fitted very well to a Michaelis–Menten function (R ranging from 0.92 to 0.98). Table 5 shows the kinetic parameters of the cadmium uptake system, where Km is an approximate measure of the affinity of the substrate for the uptake system and Vmax is the maximum velocity at which the uptake system transports the ion. The Vmax values for roots treated with cadmium were not significantly different between the three plants, whereas for Km a significant difference
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R. Barzanti et al. / Journal of Hazardous Materials 196 (2011) 66–72
Table 3 Shoot:root ratio in cadmium concentrations of A. bertolonii and A. montanum plantlets calculated as the angular coefficient of the linear regression between cadmium shoot concentration and cadmium root concentration. r and p values are reported. Values are mean ± standard error; significant differences between the means appear with different letters (* p < 0.05). Population
Shoot:root ratio a b c
A. bertolonii Pieve S. Stefano (PSS)a
A. montanum Monterufoli (MR)b
A. montanum Monte Prata (MP)c
0.32 ± 0.05 B*
0.81 ± 0.18 A
0.60 ± 0.15 AB
r = 0.95, p = 0.004. r = 0.91, p = 0.011. r = 0.88, p = 0.018.
Table 4 Total cadmium concentrations (ng) in whole roots and shoots of A. bertolonii and A. montanum plantlets treated with CdSO4 (M). Values are mean ± standard error. Roots or shoots intra-populations statistical differences are reported with lowercase letters (at least p < 0.05), roots or shoots inter-population statistical differences are showed with uppercase letters (at least p < 0.05). Treatment (CdSO4 M)
Population A. montanum Monterufoli (MR)
A. bertolonii Pieve S. Stefano (PSS)
A. montanum Monte Prata (MP)
Roots
Shoots
Roots
Shoots
Roots
Shoots
Control
18.97 ± 3.07 aB
9.00 ± 1.96 aB
6.94 ± 1.26 aA
0.32 ± 0.10 aA
14.14 ± 6.47 aAB
14.54 ± 3.18 aB
0.25
42.00 ± 5 bA
134.4 ± 11.20 bA
199.19 ± 23.98 bB
181.69 ± 20.66 bB
174.93 ± 17.68 bB
142.43 ± 29 bAB
0.5
243.41 ± 29 cA
197.44 ± 27.15 cA
287.47 ± 34.60 cA
275.21 ± 36.37 cA
238.63 ± 43.24 bcA
229.03 ± 35.63 bA
1
252.22 ± 37.06 cA
240.88 ± 27.75 cdA
357.74 ± 77.27 cdA
636.82 ± 64.90 dB
349.02 ± 40.52 cA
540.9 ± 94.72 cB
2
217.37 ± 24.92 cA
310.25 ± 52.7 dA
441.50 ± 35.94 dB
1508.53 ± 258.0 eB
346.26 ± 61.70 cAB
1022.69 ± 183.0 dB
5
162.94 ± 31.73 cA
285.76 ± 30.7 dA
314.39 ± 53.93 cdA
1324.25 ± 116.0 eB
279.37 ± 58.88 cA
996.64 ± 123.0 dB
were found between A. montanum MR and A. bertolonii PSS, the latter showing a higher value than the other one (p < 0.001). 4. Discussion
Cd concentration in excised roots (µg g-1 d.w.)
Cadmium concentrations in plants harvested from the field revealed the presence of small amount of this element in roots and shoots of the three plants. Furthermore, in aerial parts cadmium
3500 A. bertolonii Pieve S. Stefano A. montanum Monterufoli A. montanum Monte Prata
3000
concentration was well under the toxic limits for shoots [19], thus excluding any possible plant pre-adaptation to this element. The low cadmium concentration found in plants is likely to depend not only on the low cadmium concentration showed by these soils, but also on their alkaline pH that presumes a very low mobility of cadmium in such environments [19]. Mineral analysis also showed a high nickel concentration in roots and especially in shoots of the nickel hyperaccumulating plant A. bertolonii, whereas, far lower values were detected in the two populations of A. montanum. Between these two latter, the serpentine population showed a nickel concentration higher than the non-serpentine one and the limits of toxicity [19] and underlined its non-hyperaccumulating phenotype, being its nickel concentrations higher in roots than in shoots.
2500 2000
Table 5 Kinetic parameters for cadmium influx into roots of A. bertolonii and A. montanum plantlets. Values are mean ± standard error, r values are reported in the table. Significant differences between the means appear with different letters (** p < 0.001).
1500
Population
1000
A. bertolonii Pieve S. Stefano (PSS)a
A. montanum Monterufoli (MR)b
A. montanum Monte Prata (MP)c
Vmax
185.33 ± 39.09 A
151.34 ± 11.11 A
202.08 ± 26.66 A
Km
4.77 ± 0.86 A
1.84 ± 0.29 B**
2.87 ± 0.71 AB
500 0 0
1
2
3
4
5
6
CdSO4 (µM) Fig. 4. Cadmium concentration (mean ± standard error) in excised roots of A. bertolonii and A. montanum two week old-plantlets treated with 0, 0.25, 0.5, 1, 2, 5, 10 M CdSO4 for 1 h.
a b c
r = 0.95. r = 0.98. r = 0.92.
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We have seen that roots were more susceptible than shoots to cadmium toxicity in all the Alyssum species under evaluation, and, in fact, the reduction in elongation was less significant in shoots than in roots. Although A. bertolonii showed the highest shoot length values, either with or without metal treatment, this plant always displayed the lowest dry biomass values for both roots and shoots in respect to A. montanum plants (data not showed). Many hyperaccumulator species have a decreased growth rate because of the energetic costs for hyperaccumulation mechanisms [20], and unfortunately this feature represents a further disadvantage in their use in phytoextraction techniques. In Alyssum plants, the presence of cadmium in the culture medium decreased root growth in a population dependent way. In fact, analyzing the tolerance indexes, A. bertolonii showed a higher cadmium sensitivity, significant as compared to the serpentine A. montanum, whereas the two A. montanum populations did not show any difference in cadmium tolerance. Neither the serpentine adapted hyperaccumulating phenotype nor the serpentine adapted non-hyperaccumulating one seemed to have any relationship with Cd tolerance thus excluding the occurrence of any kind of Cd cotolerance for these species from a nickel, cobalt, and chromium rich environment. Differently, in some cases there is evidence for cadmium co-tolerance supported by tolerance to other metals. For example, Arnetoli et al. [21] found a very high level of hypertolerance to cadmium in a population of Silene paradoxa from a polimetallic sulphide deposit mainly enriched in copper, arsenic, zinc and lead [22] and with low cadmium levels. Moreover, high cadmium co-tolerance was found in a chromium-tolerant strain in comparison with the wild type one in Scenedesmus acutus [23]. Data on root growth were modeled to calculate the EC50 values. Results confirmed what suggested by tolerance index analysis, showing the serpentine A. montanum as the most tolerant plants, significantly in respect to A. bertolonii. Moreover, interestingly, the calculation of the EC50 values gave useful hints about a remarkable cadmium tolerance of these plants. In fact, considering the M range of the EC50 values found for the three Alyssum, they are far higher than the maximum value of Cd concentration in the solution of non-contaminated soil (about 0.05 M [19]) and about the half of this value for heavy contaminated soil (about 3 M [19]). These plants, especially the A. montanum species, are likely to have a good potential to grow on cadmium-contaminated soil. In fact, a plant, to be utilized in phytoextraction, must have the essential requisite of metal tolerance, to guarantee the defense of the major physiological and metabolic processes. Nonetheless, this feature should be the result of a combination of metal uptake and reduction of harmful effects and not be simply due to metal exclusion [3,24]. Root tolerance means the preservation of the selective property of the cell membrane and so represents the first step in metal uptake and loading into the xylem vessels [25]. Furthermore, the fact that, in our experiments, shoot growth was less affected than root growth, reinforced the previous consideration. The accumulated cadmium concentration in roots and shoots was influenced by the external concentration and metal concentration in the roots was always higher than in the shoots in all the three plants. A preferential root allocation of metals remains a widespread behavior to face metal toxicity in the majority of the plants. As for root cadmium accumulation, A. bertolonii showed the higher concentrations, up to 2500 g g−1 , followed by the other Alyssum species up to 1900–2000 g g−1 . In shoots, significant differences were found only for the two highest medium concentrations used. The pattern of the differences was diverse from roots and interestingly, inverted. In fact, A. montanum plants showed the highest concentrations, up to 1600 and 1200 g g−1 for MR and MP respectively, whereas A. bertolonii showed the lowest, up to only to 900 g g−1 . When the effect of medium cadmium concen-
71
tration began to be stringent, the most tolerant species displayed the lowest root cadmium concentrations and the highest shoot cadmium concentrations, whereas the most sensitive species, A. bertolonii, showed the opposite condition. Probably, the higher cadmium toxicity suffered by the roots of this plant, possibly due to the higher internal cadmium concentration present, impaired all the metabolic processes, cadmium translocation included. In shoots at the lowest concentration used, all the plants showed tissue concentrations higher than 100 g g−1 , value that is considered the limit for cadmium hyperaccumulation [9]. Considering that for that concentration the toxic effect of cadmium was very low on root growth and not present in shoots, or minimum as in the case of A. bertolonii, these plants could be proposed as potential cadmium hyperaccumulators that merit to be studied. After all, also the nickel hyperaccumulator Thlaspi goesingense was demonstrated to accumulate cadmium at very high concentrations in its above ground tissue [11]. Values found in shoots were also higher than those found in some hydroponically cultivated Brassica plants, generally studied for their remarkable and promising cadmium phytoremediation potential (see for example Qadir et al. [26] and Grispen et al. [27]). The very high concentrations found in the shoots of these Alyssum plants may suggest that it is worth studying the behavior of plants also to metals not present in their environment. However, for really generating an hyperaccumulator phenotype, also a high rate of root-to-shoot translocation has to be present and this was not the case for these Alyssum plants for the concentrations used, as root always showed higher cadmium concentrations than shoots. In any case, the translocation coefficients, identified as a fundamental trait for the plant suitable for phytoextraction [3], were very different in the three plants and the serpentine A. montanum showed a value approaching to 1, whereas, the nickel hyperaccumulator showed the lowest. The most cadmium tolerant plants were also those with the highest shoot concentration and the highest translocation coefficient, really suggesting that metal adapted non-hyperaccumulating plants can be useful in exploring the possibility for phytoextraction, rather than too specialized hyperaccumulators. Furthermore, our data also suggest that even non metal adapted plants may be more useful than hyperaccumulators in the extraction of the non hyperaccumulated metals, as is the case of the non-serpentine A. montanum and A. bertolonii. As for cadmium content per plant, this incremented with increasing cadmium concentration in the medium and was higher in shoots than in roots despite the higher root metal concentration. This result depends on the huge difference in both biomass production and metal sensitivity between the two organs. The plantlets showed significant differences in cadmium content. In the case of roots, even if A. bertolonii showed the highest metal concentrations and the lowest biomass production, it had the lowest metal content; in the case of root the low biomass was the factor that determined the differences in metal content. As for shoot cadmium content, the serpentine A. montanum showed the highest concentration and A. bertolonii the lowest. This result was generated by both the lowest cadmium concentration and the lowest dry biomass (data not showed) displayed by A. bertolonii. Therefore, whereas the serpentine adapted nickel hyperaccumulating phenotype did not seem a useful material for cadmium phytoextraction, the serpentine adapted nickel non-hyperaccumulating phenotype presented the advantage of an higher cadmium phytoextraction capability, confirming the idea that studying metal non present in the environment of metal adapted plants can offer the possibility of a useful selection of plants for phytoextraction and phytostabilisation aim. To evaluate if differences in metal concentration may depend on differences in root cadmium uptake systems, short-term uptake
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studies with excised roots were of fundamental importance to exclude the effect of translocation and evaluate the real ability of the root to take the metal up. Our results showed a significantly different cadmium accumulation in the excised roots of the three plants. Experiments showed that the nickel hyperaccumulating plants accumulated cadmium at a significantly lower level than the other ones for most of the cadmium concentrations used. In fact, the calculation of the kinetic parameters of cadmium uptake presented a significant difference in the apparent Km values. A. bertolonii presented the highest apparent Km value, suggesting a lower affinity for this metal of its uptake system, whereas the Vmax values were not significantly different. If there could be a connection between this very low affinity and nickel hyperaccumulation is a question that deserves to be investigated. This low affinity for cadmium is not in contrast with the highest cadmium root accumulation showed by intact plants, as the latter is also the result of shoot translocation that in A. bertolonii was the lowest. This fact, combined also with the lowest dry biomass of this plant, could explain the very low amount of cadmium per plant showed by the nickel hyperaccumulating phenotype. As for A. montanum, the two species did not present any difference in the cadmium kinetic parameters, but, notwithstanding this result, it is interesting to note that the cadmium content per plant was higher in the serpentine adapted plants, probably because of its higher, even not significant, tolerance to this metal in respect to the non-serpentine adapted population. 5. Conclusion The selection of the most efficient plant is indeed a key factor in every phytotechnology for soil reclamation, especially for metal phytoextraction that is still far from reality. In fact, gaining information on the intra-specific and inter-specific variability in plant contaminant accumulation and selecting suitable plants remains a challenge [28]. In this context, the present data demonstrated that evaluating metallicolous plant behavior, even toward metals present at low level in the origin environment, could be a good research field for finding plants suitable for phytoextraction. In fact, in respect to the non-serpentine adapted A. montanum and the serpentine adapted nickel hyperaccumulator A. bertolonii, the serpentine adapted population of A. montanum showed remarkable cadmium tolerance and accumulation that merit to be studied and exploited for the selection of more suitable tools for phytoremediation purposes. Nevertheless, studies on plant cultivation on real contaminated soils are needed. Acknowledgements This study was funded by the MIUR (Cofinanziamento 2008) and the University of Firenze (Fondi di Ateneo 2008–2009). References [1] D.E. Salt, R.D. Smith, I. Raskin, Phytoremediation, Annu. Rev. Plant Physiol. Mol. Biol. 49 (1998) 643–668. [2] R.L. Chaney, J.S. Angle, C.L. Broadhurst, C.A. Peters, R.V. Tappero, D.L. Sparks, Improved understanding of hyperaccumulation yields commercial phytoextraction and phytomining technologies, J. Environ. Qual. 36 (2007) 1429–1443.
[3] J. Vangronsveld, R. Herzig, N. Weyens, J. Boulet, K. Adriaensen, A. Ruttens, T. Thewys, A. Vassilev, E. Meers, E. Nehnevajova, D. van der Lelie, M. Mench, Phytoremediation of contaminated soils and groundwater: lessons from the field, Environ. Sci. Pollut. Res. 16 (2009) 765–794. [4] R.L. Chaney, J.S. Angle, M.S. McIntosh, R.D. Reeves, Y.-M. Li, E.P. Brewer, K.Y. Chen, R.J. Roseberg, H. Perner, E.C. Synkowski, C.L. Broadhurst, S. Wang, A.J.M. Baker, Using hyperaccumulator plants to phytoextract soil Ni and Cd, Z. Naturforsch. 60C (2005) 190–198. [5] S.P. McGrath, F.J. Zhao, E. Lombi, Phytoremediation of metals, metalloids, and radionuclides, Adv. Agron. 75 (2002) 1–56. [6] I. Raskin, R.D. Smith, D.E. Salt, Phytoremediation of metals: using plants to remove pollutants from the environment, Curr. Opin. Biotechnol. 8 (2) (1997) 221–226. [7] D.E. Salt, M. Blaylock, P.B.A.N. Kumar, V. Dushenkov, B.D. Ensley, L. Chet, I. Raskin, Phytoremediation: a novel strategy for the removal of toxic metals from the environment using plants, Biotechnology 13 (2) (1995) 468–474. [8] R.L. Chaney, M. Malik, Y.M. Li, S.L. Brown, E.P. Brewer, J.S. Angle, A.J.M. Baker, Phytoremediation of soil metals, Curr. Opin. Biotechnol. 8 (1997) 279–284. [9] U. Krämer, Metal hyperaccumulation in plants, Annu. Rev. Plant Biol. 61 (2010) 517–534. [10] A.J.M. Baker, R. Reeves, A. Hajar, Heavy metal accumulation and tolerance in British populations of the metallophyte Thlaspi caerulescens J & C Presl (Brassicaceae), New Phytol. 127 (1994) 61–68. [11] E. Lombi, F. Zhao, S. Dunham, S. McGrath, Cadmium accumulation in populations of Thlaspi caerulescens and Thlaspi goesingense, New Phytol. 145 (2000) 11–20. [12] C. Minguzzi, O. Vergnano, Il contenuto di nichel nelle ceneri di Alyssum bertolonii, Atti Soc. Tosc. Sci. Nat. 55 (1948) 49–74. [13] R.R. Brooks, Serpentine and its Vegetation, Dioscorides Press, Portland, OR, 1987. [14] A.J.M. Baker, Metal tolerance, New Phytol. 106 (Suppl) (1987) 93–111. [15] R. Barzanti, Tossicità, tolleranza e batteri endofiti in Alyssum bertolonii Desv.: nuovi aspetti fisiologici, PhD Thesis, Università di Firenze, Italy, 2005. [16] D.L. Sparks, Methods of Soil Analysis. Part 3. Chemical Methods, Soils Science Society of America, Madison, WI, 1996. [17] D.L. Arnon, Microelements in culture solution experiments with higher plants, Am. J. Bot. 25 (1938) 322–325. [18] A.J.M. Baker, P.L. Walker, Physiological responses of plants to heavy metals and the quantification of tolerance and toxicity, Chem. Speciat. Bioavail. 1 (1989) 7–17. [19] A. Kabata-Pendias, H. Pendias, Trace Elements in Soils and Plants, third ed., CRC, Boca Raton, 2001. [20] E. Maestri, M. Marmiroli, G. Visioli, N. Marmiroli, Metal tolerance and hyperaccumulation: costs and trade-offs between traits and environment, Environ. Exp. Bot. 68 (2010) 1–13. [21] M. Arnetoli, R. Vooijs, C. Gonnelli, R. Gabbrielli, J.A.C. Verkleij, H. Schat, Highlevel Zn and Cd tolerance in Silene paradoxa L. from a moderately Cd- and Zncontaminated copper mine tailing, Environ. Pollut. 156 (2008) 380–386. [22] I. Mascaro, M. Benvenuti, F. Corsini, P. Costagliola, P. Lattanzi, P. Parrini, G. Tanelli, Mine wastes at the polymetallic deposit of Fenice Capanne (Southern Tuscany, Italy). Mineralogy, geochemistry and environmental impact, Environ. Geol. 41 (2001) 417–429. [23] E. Torricelli, G. Gorbi, B. Pawlik-Skowronska, L. Sanità di Toppi, M.G. Corradi, Cadmium tolerance, cysteine and thiol peptide levels in wild type and chromium-tolerant strains of Scenedesmus acutus (Chlorophyceae), Aquatic Toxicol. 68 (2004) 315–323. [24] U. Krämer, Phytoremediation: novel approaches to cleaning up polluted soils, Curr. Opin. Biotechnol. 16 (2005) 133–141. [25] M. Zacchini, F. Pietrini, G. Scarascia Mugnozza, V. Iori, L. Pietrosanti, A. Massacci, Metal tolerance, accumulation and trans location in poplar and willow clones treated with cadmium in hydroponics, Water Air Soil Pollut. 197 (2009) 23–34. [26] S. Qadir, M.I. Qureshi, S. Javed, M.Z. Abdin, Genotypic variation in phytoremediation potential of Brassica juncea cultivars exposed to Cd stress, Plant Sci. 167 (2004) 1171–1181. [27] V.M.J. Grispen, H.J.M. Nelissen, J.A.C. Verkleij, Phytoextraction with Brassica napus L.: a tool for sustainable management of heavy metal contaminated soils, Environ. Poll. 144 (2006) 77–83. [28] M. Mench, J.P. Schwitzguébel, P. Schroeder, V. Bert, S. Gawronski, S. Gupta, Assessment of successful experiments and limitations of phytotechnologies: contaminant uptake, detoxification and sequestration, and consequences for food safety, Environ. Sci. Pollut. Res. 16 (2009) 876–900.
Journal of Hazardous Materials 196 (2011) 73–78
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Tentacle carrier for immobilization of potato phenoloxidase and its application for halogenophenols removal from aqueous solutions Nikola Lonˇcar ∗ , Zoran Vujˇcic´ Faculty of Chemistry, University of Belgrade, Studentski trg 12-16, Belgrade, Serbia
a r t i c l e
i n f o
Article history: Received 21 April 2011 Received in revised form 28 July 2011 Accepted 30 August 2011 Available online 3 September 2011 Keywords: Polyphenol oxidase Potato Bromophenol Chlorophenol Immobilization
a b s t r a c t Halogenated compounds represent one of the most dangerous environmental pollutants, due to their widespread usage as biocides, fungicides, disinfectants, solvent and other industrial chemicals. Immobilization of a protein through coordinate bonds formed with divalent metal ions is becoming an attractive method due to its reversible nature, since the protein may be easily removed from the support matrix through interruption of the protein–metal bond hence giving inherently cleaner and cheaper technology for wastewater treatment. We have synthesized novel ‘tentacle’ carrier (TC) and used it for immobilization of partially purified potato polyphenol oxidase (PPO). The obtained biocatalyst TC-PPO showed pH optimum at 7.0–8.0 and temperature optimum at 25 ◦ C. Immobilized PPO shows almost 100% of activity at 0 ◦ C. TC-PPO was more resistant to the denaturation induced by sodium dodecyl sulphate (SDS) detergent as compared to its soluble counterpart and was even slightly activated at SDS concentration of 1%. TC-PPO was tested in the batch reactor for 4-chlorophenol and 4-bromophenol removal. More than 90% removal was achieved for both halogenophenols at concentration of 100 mg/L from aqueous solution. For both halogenophenols TC-PPO works with over 90% removal during first three cycles which decrease to 60% removal efficiency after six cycles each of 8 h duration. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The major obstacle in the commercial application of soluble enzymes for environmental purposes is their limited operational stability, which means that a continuous supply of large amounts of enzyme is required. Enzyme immobilization improves the operational stability and half-life of the enzyme therefore reducing the treatment cost [1]. Enzyme immobilization procedures often require purified enzyme and expensive supports and reagents [2]. The purpose of this procedure is to allow reuse of enzymes for many reaction cycles. However, this implies that immobilized enzyme preparation is more expensive. Novel immobilization protocols are still needed in order to achieve massive implementation of enzymes as catalysts of the complex chemical processes under the benign experimental and environmental conditions. There is always a search for cheaper support and enzymes for preparing an immobilized enzyme preparation for the aforementioned applications. Covalent-type linkages between molecule and matrix are generally considered best, however many of the recommended
∗ Corresponding author. Tel.: +381 11 3282 393; fax: +381 11 2636 061. E-mail addresses:
[email protected] (N. Lonˇcar),
[email protected] ´ (Z. Vujˇcic). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.071
immobilization procedures require pre-derivatisation of the matrix, extended periods for coupling reactions and specialized conditions [3]. In some cases there is considerable loss of activity during the immobilization procedure, in some cases up to 85% as recently reported by Pramparo et al. [4]. We were searching for a method in which matrix derivatisation after preparation could be avoided and where virtually instantaneous coupling could be achieved under mild and easy-to-handle conditions. To accomplish this, we employed the chelating properties of copper – which is often used for purification by immobilized metal-affinity chromatography (IMAC) and immobilization of tyrosinases from different sources [5]. By exploiting this property, high enrichment of enzyme was obtained due to its higher binding affinity to the support relative to other proteins. A similar approach has been applied for direct immobilization of tyrosinase on Celite, d-sorbitol cinnamic ester and copper alginate gel [6–8]. Sepharose and agarose derivatives containing metal chelating groups such as iminodiacetic acid (IDA) or nitriloacetic acid (NTA) are all suitable supports [9–11]. The high cost of commercially available chelating-agarose and cellulose is their biggest limitation. Additionally, short spacer arms restrict the mobility of the bound enzyme and impair the apparent catalytic activity. This opens up the possibility for creating new supports, with longer chelating arms (tentacles) that can offer more fluidity for the enzyme and higher availability, subsequently increasing the enzyme substrate contact.
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Both laboratory and industrial scale separations of proteins rely mostly on ion-exchange resins. These resins may be utilized for a certain number of separation cycles after which their efficiency diminishes. In the present study, this reused ion-exchange (IEX) chromatography matrix – diethylaminoethyl (DEAE) cellulose fibers – were selected as the raw, starting material for preparation of support material for immobilization of PPO. DEAE cellulose is a derivative of the natural polymer cellulose and it possesses a desirable hydrodynamic structure since fibrilar structures have a very large active surface. It was chosen over other materials because of its biodegradability, hydrophilicity and presence of hydroxyl groups, cost and natural origin [12,13]. The tentacle carrier (TC) was synthesized by activation of amino groups of DEAE cellulose with epichlorohydrine followed by introducing IDA groups, which were subsequently charged with copper ions. Binding of enzyme during the immobilization procedure is achieved through coordinative binding of N, O and S atoms of enzyme amino acid residues to copper ions. Horseradish and turnip peroxidases, both soluble and immobilized by different procedures, have been successfully employed in phenol removal [14]. Although those sources of different peroxidases are available in abundance, they are seasonal and rather expensive plants. On the contrary, PPO from potatoes is the enzyme with most potential for biotechnology due to its exceptionally low price since it can be purified from the potato food industry waste [6,15]. It has been shown previously [16] that PPO activity was the greatest at the tuber exterior, including the skin and cortex tissue 1–2 mm beneath the skin, which opens up possibilities for the exploitation of potato peels. This process may be in conjunction with the potato chips industry where excess proteins in effluents are a major problem. It was shown that enzymatic treatment efficiency is independent of the enzyme purity. Moreover a crude or partially purified preparation is protected from inactivation due to the significant quantity of other proteins present [14]. This feature leads to a significant reduction in overall treatment costs; hence we decided to partially purify PPO using only one chromatographic step under oxygen free conditions, with the purpose of enriching phenoloxidase activity. The immobilization procedure consists of an overnight incubation of enzyme preparation with the produced TC to give TC-PPO. The catalyst obtained by this procedure has been tested in batch reactors for removal of 4-chlorophenol (p-CP) and 4-bromophenol (p-BP) from synthetic wastewater.
carried on stirrer for 2 h. Activated DEAE-cellulose was washed and transferred into solution made of 2 g NaOH, 3 g Na2 CO3 and 4.3 g IDA, incubated 2 h at 50 ◦ C and then left overnight. After washing with 500 ml of water, produced tentacle carrier was charged with copper (1 g CuSO4 ·5H2 O in 25 ml H2 O), washed with water again and then with 7% acetic acid. TC-carrier was then washed with 10 volumes of water followed by washing with 10 volumes of 20% ethanol solution. TC-carrier was stored at 4 ◦ C until further use. 2.3. Carrier characterization Water content has been determined in 10 replicative measurements by taking 200 mg of semidry TC-PPO and drying it at 60 ◦ C using Eppendorf Vacuum Concentrator Model 5301 until constant weight. Amount of bound epoxy group prior to introduction of IDA has been determined by method described by Axen et al. [17]. Amount of incorporated Cu(II) ions is determined by inductively coupled plasma-atomic emission spectroscopy (ICP-AES) by iCap 6500Duo, equipped with a CID86 chip detector (Thermo Scientific, UK). Instrumental operating conditions for ICP-AES are RF power: 1150 W; plasma view: axial; nebulizer gas flow: 0.50 L/min; auxiliary gas flow: 0.50 L/min; coolant gas flow: 12 L/min; analysis pump rate: 50 rpm). TC was previously dissolved in 10 ml of 65% HNO3 , 0.2 ml of H2 O2 using ETHOS 1, Advanced Microwave Digestion System, (MILESTONE, Italy). Microwave digestion was done in a two step program, first being heating to 200 ◦ C over 15 min and second being held at 200 ◦ C for 20 min. Sample was then diluted to 25 ml. Blank was prepared same way without sample. 2.4. Partial purification of PPO
2. Materials and methods
Potato (S. tuberosum) tubers were obtained from the local market and were kept at 3 ◦ C for 12 h. Thereafter, whole tubers were homogenized in commercial juicer. The homogenate (1 L) was centrifuged at 3500 rpm at 4 ◦ C. 600 ml of clear supernatant was desalted against 10 mM Na-phosphate buffer pH 7.3 using 6 cm × 60 cm Sephadex G25 Coarse column. 60 g of preswollen QAE Sephadex A-50 was equilibrated with 10 volumes of 10 mM Na-phosphate buffer, pH 7.3. Equilibrated and deaerated ionexchanger was added to extract and mixed with stirrer 30 min in oxygen-free atmosphere. After that matrix was washed with 3 volumes of starting buffer and enzyme was eluted with 0.75 M NaCl in starting buffer in 500 ml. Obtained enzyme preparation was stored at −20 ◦ C until use.
2.1. Reagents
2.5. Polyphenol oxidase activity assay
Potato (Solanum tuberosum) tubers were obtained from the local market. Commercially available DEAE-cellulose (Sigma) was used in this study. However, it was a reused DEAE-cellulose, i.e. not appropriate for protein purification anymore since after many uses it does not provide required resolution in chromatography of proteins. Phosphate buffers, sodium chloride, acetic acid, l-3,4-dihydroxyphenylalanine (l-DOPA), Tris, sodium hydroxide, epichlorohydrin, sodium carbonate, iminodiacetic acid, copper sulphate, SDS, p-CP, p-BP, 4-aminoantipyirine, potassium ferricyanide, sodium hydrogen carbonate and ethanol used were of the highest available purity. They were purchased unless otherwise stated, from Merck (Darmstadt, Germany) and Sigma–Aldrich (St. Louis, MO, USA). Sephadex G-25 Coarse and QAE Sephadex A-50 were purchased from GE Healthcare Life Sciences.
activity was determined using l-3,4PPO dihydroxyphenylalanine (l-DOPA) as a substrate at 25 ◦ C by measuring the initial rate of dopachrome formation [18].
150 mg of dry TC (1 g of semidry) carrier was equilibrated with 25 ml of 0.15 M potassium-phosphate buffer pH 7.0 with 0.5 M NaCl to prevent nonselective ionic adsorption. TC carrier was then resuspended in 30 ml of enzyme preparation (192,000 U). Same carrier was added to 0.5 M NaCl as blank probe. Mixtures were left for 12 h on IKA orbital shaker at 400 rpm. Biocatalysts were removed by centrifugation at 14000 × g and TC was washed six times with 30 ml of 0.9% NaCl (untill total unbound enzyme was eluted).
2.2. Preparation of TC-carrier
2.7. Immobilized PPO activity assay
5 g of preswollen reused DEAE-cellulose was resuspended in 20 ml of 3 M NaOH and 2 ml of epichlorohydrin. Activation was
The activity of immobilized enzyme was assayed using a modified version of the method of Kwon and Kim [18], since
2.6. Immobilization
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immobilizate has to be removed by centrifugation from l-DOPA solution prior to spectrophotometric measurement. Twenty milligrams of air dried biocatalyst was added to 1 ml of 9.3 mM l-DOPA in 50 mM Tris HCl buffer pH 7.0 at 25 ◦ C. The mixture was shaken for 3 min and immobilizate was removed by centrifugation at 14000 × g. Increase in absorbance due to dopaquinone formation in resulting supernatant was measured at 475 nm. The blank sample contained 20 mg of dried blank carrier instead of biocatalyst, besides the other component of activity assay mixture. Specific activity of immobilized PPO is defined as increase in absorbance of 0.001 per min per gram of semidry biocatalyst under the given assay conditions.
using Cintra40 UV–vis Spectrometer in 200–380 nm range. This was done for p-BP and p-CP before and after treatment with TC-PPO as described in Section 2.11. The diminution in absorbance peak of treated sample in UV region was taken as an evidence for the removal of both compounds.
2.8. Determination of pH optimum
2.15. Statistics
To determine the pH optimum against l-DOPA, 10 mg of TC-PPO and 50 l of soluble PPO was incubated with 750 l of appropriate 50 mM buffer (acetate, pH 4.3–5.8; Tris–HCl, pH 7.0–8.0; potassium phosphate, pH 9.0–10.0) for 30 min and 250 l of 9.3 mM l-DOPA was added. Further steps are same as described in Section 2.7.
All experimental results reported in the next sections were based on averaging results of repeated experimental runs (triplicates), with the standard deviation ranging from 2 to 6% of the reported average.
2.14. 2.14. p-CP and p-BP removal reusability of TC-PPO Reusability of TC-immobilizate was tested by repeating the above described batch reactor experiment six times. Between these cycles immobilizate was collected by filtration and washed three times in sodium phosphate buffer pH 7.0.
3. Results and discussion 2.9. Determination of temperature optimum 3.1. Synthesis of ‘tentacle’ carrier To determine the temperature optimum against l-DOPA, 10 mg of TC-PPO and 50 l of soluble PPO were resuspended in preincubated 750 l of 50 mM sodium phosphate buffer pH 7.0 at appropriate temperature (0–95 ◦ C) after which 250 l of 9.3 mM lDOPA preincubated at same temperature was added. Further steps are same as described in Section 2.7. For soluble enzyme above described procedure for PPO activity assay was used. Highest activity was designated as 100%. 2.10. Treatment of soluble and immobilized PPO with detergents Soluble and immobilized enzymes were incubated with SDS detergent (0.25–1.0%) in 50 mM sodium phosphate buffer pH 7.0 at 25 ◦ C for 100 min. Remaining PPO activity was determined as described in Section 2.7. The activity of the untreated preparation was taken as a control (100%). 2.11. Preliminary application study of immobilized PPO using batch reactor
Previously it was described that during synthesis of DEAEcellulose matrix, there is always few percent of residual amino groups besides diethylaminoethyl groups that are primarily derivatised by epichlorohydrin used in this study [20]. This happens because DEAE groups are somewhat sterically hindered and hence less reactive and epichlorohydrin activation leaves some of the DEAE groups unreacted, thus obtaining the mixed IMAC/IEX carrier. Epichlorohydrine spacer arm is introduced in DEAE cellulose via creating bond with nitrogen atom as shown in Fig. 1a. Under these conditions OH groups of cellulose are less reactive than NH2 because most of OH groups are already spent during synthesis of DEAE-cellulose, and those left are less reactive due to steric hindrance. Chelating group, IDA is irreversibly coupled to epoxy-activated DEAE cellulose (Fig. 1a–c.). Copper ions reversibly form a coordinate bond with the chelating group thus creating tentacle Cu-carrier (Fig. 1d). 3.2. Carrier characterization
Removal of phenolic compounds from synthetic wastewater was investigated with 30 mg dry biocatalyst (310,000 U/g) that was added to a 20 ml of 100 mg/L solutions of p-CP and p-BP dissolved in 50 mM sodium phosphate buffer pH 7.0 and incubated on orbital shaker to allow continuous oxygenation. Phenol solution was oxygenated by air sparging for 10 min prior to use.
Water content has been determined to be 85%. Amount of bound epoxy group prior to introduction of IDA has been determined to be 0.144 mmol/g of dry support. Copper content measured by ICP-AES is 7.3110 mg/g of dry support. 3.3. Partial purification of PPO
2.12. Quantitative assay of phenol compounds Treated solutions were sampled at different time scale and filtered through 0.45 m filter and then tested for remaining phenol content using 4-aminoantipyrine (AAP) assay [19]. The concentrations of phenols were measured using a reaction with 2.08 mM AAP and 8.34 mM potassium ferricyanide in 0.25 M sodium bicarbonate solution to form a red quinone-type dye that absorbs light with a peak wavelength of 510 nm. The extent of color generation at 510 nm after a 6 min incubation time was proportional to the concentration of phenols in the assay solution. 2.13. UV spectrometry analysis of TC-PPO treatment products In order to confirm the oxidation of p-BP and p-CP by TC-PPO, spectral analysis was performed. An UV spectrum was recorded
One of the problems for the purification of polyphenoloxidase from plant material is presence of phenolics since these compounds are oxidized to quinones by PPO, which could react covalently with the enzyme, resulting in aggregation of enzyme [18]. Strategy applied in this work for minimizing these unwanted effects is immediate desalting of crude extract using industrial Sephadex G25 (coarse granulation for fast chromatography). Oxygen-free atmosphere in later steps prevents oxidation of enzyme with remaining phenolics, since other substrate – oxygen is not available. QAE Sephadex is a well known industrial chromatographic matrix. Its wide spread use is consequence of its high capacity for protein binding and cheap price. Purification steps were designed to exploit maximum yield from capturing chromatography methodology with respect to time required for this process. Care has been taken to optimize purification conditions so the process can be
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Fig. 2. Effect of the solution pH on the oxidation rate of l-DOPA by soluble and immobilized PPO.
Fig. 1. Synthesis of tentacle carrier. Epichlorohydrin activation of DEAE cellulose is followed by introducing IDA functional groups. Copper ions are bound by coordinative bonds of IDA and three molecules of water replaced during immobilization process by N, O and S atoms of amino acid residues present on surface of protein.
easily scaled-up to industrial level. For this purpose batch chromatography and cheap phosphate buffer have been chosen. This procedure gives a yield of around 70% polyphenoloxidase activity as compared to starting extract. The partially purified PPO showed 8500 U/ml under given assay conditions. 3.4. Immobilization Protein immobilization is achieved by reversible binding of copper and certain amino acid residues of the protein (aa1–aa3 in Fig. 1d). At this stage, the metal ions can be viewed as electronpair acceptors (Lewis acids) and the N, 0, and S atoms of the amino acid side chain ligands as electron-pair donors (Lewis bases). These amino acids must be located on the surface of the protein molecule in order to be accessible to the metal ion chelate. This support offers more fluidity and hence substrate accessibility to bound protein as compared to conventional support of this type with much shorter spacer arm. Using partially purified preparation of PPO under given immobilization conditions biocatalyst with activity of 310,000 U/g of dry biocatalyst was produced.
were kept constant while varying the temperature from 0 to 95 ◦ C and pH from 4.3 to 10. The experimental results for the effect of initial solution pH on the concentration of substrate and oxidation rate of l-DOPA are shown in Fig. 2. These results reveal that TC-PPO is active towards l-DOPA at wider pH range in contrast to soluble PPO. pH optimum for TC-PPO is 7.0–8.0. This is very convenient since pH close to neutrality is desirable for the treatment of halogenophenols effluents. Since the pKa values of p-BP and p-CP at 25 ◦ C are 9.37 and 9.41, respectively [21], the removal efficiency of TC-PPO will decrease at pH close to 9.5 and this is attributed to the formation of the p-BP and p-CP conjugated bases that do not permit the phenol compounds to act as hydrogen donors. A plot of the TC-PPO activity versus temperature is shown in Fig. 3. Determination of temperature optimum showed similar trends for both soluble and immobilized PPO. Clearly, the activity of potato PPO and consequently its ability to oxidize phenolic substrates is optimized at about 25 ◦ C. Higher temperatures seemed to negatively affect the activity of PPO; however, immobilized enzyme has 34% of residual activity at 90 ◦ C. Exposure to lower temperatures is expected to slow down the enzymatic activity, but immobilized PPO has shown almost 100% of activity at 0 ◦ C under assay conditions described in Section 2.9. This is interesting characteristic since it avoids necessity for temperature regulation in large reactors, which further leads to energy saving. 3.6. Effect of SDS detergent Wastewater also gets contaminated with detergents released from industry and municipal sewage so there is a need to investigate biocatalyst efficiency in presence of detergent [22]. The stability of TC-PPO and soluble PPO in presence of anionic detergent
3.5. Effect of pH and temperature Experiments were carried out to assess the effect of temperature and solution pH on the activity of TC-PPO. All other parameters
Fig. 3. Effect of the solution temperature on the oxidation rate of l-DOPA by soluble and immobilized PPO.
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Fig. 4. Effect of SDS on soluble and immobilized PPO.
SDS was examined. Soluble and immobilized potato PPO preparations were exposed to various concentrations of SDS (0.1–1.0%, w/v). Khan et al. [6] conducted investigation of Celite immobilized PPO and showed that immobilized enzyme was more resistant to the denaturation induced by SDS detergent as compared to its soluble counterpart. Our results comply with this as considerable stabilization is obvious as well as slight activation at SDS concentration of 1% as shown in Fig. 4. Activation by submicellar concentrations of SDS is well documented for some plant, fungal and bacterial PPOs [23–25]. On the other side, SDS denatures most proteins by conformational change and thus decreases activity of soluble enzyme. Activation of immobilized PPO may be explained by limited conformational change as a consequence of binding of small amounts of SDS which induce or initiate activation of enzyme. Limited conformational change is direct characteristic of immobilized enzyme. Incubation of soluble enzyme with 1.0% SDS for 1 h resulted in a loss of 77% of its original activity under identical incubation conditions. 3.7. Preliminary application study of TC-PPO As model pollutants we have chosen p-BP and p-CP because they are used as disinfectants in home, hospitals and farms [26] and they had been used previously as model system [2,27,28]. During experimental removal of these two compounds from aqueous solution similar results were obtained for both p-CP and p-BP. In preliminary study for removal of 100 mg/L p-BP from aqueous solution we showed that more than 90% removal was achieved after 8 h as shown in Fig. 5. p-CP was removed by same rate. This experiment has been carried out at constant temperature held at 25 ◦ C. pH of solution remained same during whole treatment experiment. However, color formation was observed to change from transparent at the beginning of the process to lightly brownish after 2–3 h. This is followed by dark brown/black precipitate formation of polymers formed from oxidized halogenophenols. This precipitate partly adsorbs on TC-PPO and partly remains in solution. In order to confirm the oxidation of p-BP and p-CP by TC-PPO, spectral analyses were performed on filtered samples. The diminution in absorbance peak of treated sample in UV region was a clear evidence for the removal of treated halogenophenols (data not shown). By this treatment halogenophenol concentration was lowered below 10 mg/L which is an upper limit value set by water control regulations for phenolic compounds present in wastewater released in open water [29]. Almost equally successful was the process that exploited soluble bitter gourd peroxidase published by Ashraf and Husain [27]. However, some biocatalysts obtained with commercial carriers such as Eupergit, have shown much less effectiveness in phenol compounds removal. It has been shown that PPO
Fig. 5. Removal of 100 mg/L halogenophenols from aqueous solution. Aliquots are taken at different time points and phenol concentration measured.
and HRPO immobilized on Eupergit carriers were able to remove only 45% of p-BP and 50% of phenol [4,30], and this fact is one of the reasons for creating new carriers. 3.8. Reusability of TC-immobilizate The main goal of enzyme immobilization is the industrial reuse of enzymes for many reaction cycles so it was necessary to investigate the reusability of immobilizate for the removal of p-CP and p-BP. For both p-BP and p-CP TC-PPO works with over 90% removal during three cycles, each of 8 h duration. After six times of repeated tests the efficiency of p-CP and p-BP removal by TC-PPO immobilizate decreased to 55% and 60%, respectively (Fig. 6). As described in Section 3.7 accumulation of dark precipitates on immobilizates was observed. Accumulation of dark precipitate gradually increased after each cycle and this can be explanation for a reduction of the removal efficiency in latter cycles since precipitate interacts with bound PPO by hydrophobic interactions thus preventing its movement and possibly inhibits it as pseudo substrate. Further studies should examine application of continuous reactors which will prevent accumulation of reaction product on this biocatalyst. Previously it was reported that colored products generated from phenol compounds by tyrosinase react with amino group containing coagulants such as chitosan
Fig. 6. The effect of repeated use on the activity of TC-PPO.
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and polyethylenimine [31]. In order to decrease time necessary to complete treatment usage of coagulants in batch reactor may be explored because removal of reaction products will provide driving force for reaction to continue at the same rate as at the beginning. Better alternative to coagulants usage is application of TC-PPO in continuous packed bed airlift reactors since this will decrease contact time between catalyst and reactive quinones thus prolonging number of possible reuses. This is a subject of our further study. 4. Conclusions From the industrial point of view simplicity and costeffectiveness are the key demands for immobilized enzymes and these requirements are fulfilled with low cost PPO and low cost TC. Moreover, TC can be reused until enzyme activity decreases and after that carrier can be stripped with EDTA and charged with new dose of enzyme. Presented method of immobilization in overall is very economical procedure for the immobilization of PPO and immobilizate showed good potential for halogenophenol removal by polymerization. Thus, the proposed process can be considered as a first step in further exploration of possible carriers for application of immobilized PPO for phenol and phenol derivatives removal from wastewater. Acknowledgements This work was supported by the Serbian Ministry of Science and Technological Development (grant no. 172048). References [1] N. Duran, E. Esposito, Potential applications of oxidative enzymes and phenoloxidase-like compounds in wastewater and soil treatment: a review, Appl. Catal. B: Environ. 28 (2000) 83–99. [2] I. Levy, G. Ward, Y. Hadar, O. Shoseyov, C.G. Dosoretz, Oxidation of 4-bromophenol by the recombinant fused protein cellulose-binding domainhorseradish peroxidase immobilized on cellulose, Biotechnol. Bioeng. 82 (2) (2003) 223–231. [3] J. Woodward, Immobilized Cells and Enzymes, A Practical Approach, IRL Press, Oxford, Washington, DC, 1985. [4] L. Pramparo, F. Stuber, J. Font, A. Fortuny, A. Fabregat, C. Bengoa, Immobilisation of horseradish peroxidase on Eupergit C for the enzymatic elimination of phenol, J. Hazard. Mater. 177 (2010) 990–1000. [5] F. Richard-Forget, P. Goupy, J. Nicolas, New approaches for separating and purifying apple polyphenol oxidase isoenzymes: hydrophobic metal chelate and affinity chromatography, J. Chrom. A 667 (1994) 141–153. [6] A.A. Khan, S. Akhtar, Q. Husain, Direct immobilization of polyphenol oxidase on Celite 545 from ammonium sulphate fractionated proteins of potato (Solanum tuberosum), J. Mol. Catal. B: Enzym. 40 (2006) 58–63. [7] M.E. Marin-Zamora, F. Rojas-Melgarejo, F. Garcia-Canovas, P.A. Garcia-Ruiz, Direct immobilization of tyrosinase enzyme from natural mushrooms (Agaricus bisporus) on d-sorbitol cinnamic ester, J. Biotechnol. 126 (2006) 295–303.
[8] G. Palmieri, P. Giardina, B. Desiderio, L. Marzullo, M. Giamberini, G. Sannia, A new immobilization procedure using copper alginate gel: application to a fungal phenol oxidase, Enzyme Microb. Technol. 16 (1994) 151–158. [9] G. Bickerstaff, Immobilization of Enzymes and Cells, Humana Press, 1997. [10] P. Piacquadio, G. De Stefano, M. Sammartino, V. Sciancalepore, Phenols removal from apple juice by laccase immobilized on Cu2+ -chelate regenerable carrier, Biotechnol. Techniq. 11 (7) (1997) 515–517. [11] N. Duran, M.A. Rosa, A. D’Annibal, L. Gianfreda, Applications of laccases and tyrosinases (phenoloxidases) immobilized on different supports: a review, Enzyme Microb. Technol. 31 (2002) 907–931. [12] M.Y. Arica, Immobilization of polyphenol oxidase on carboxymethylcellulose hydrogel beads: preparation and characterization, Polym. Int. 49 (2000) 775–781. [13] F.N. Kok, M.Y. Arica, O. Gencer, K. Abak, V. Hasirci, Controlled Release of aldicarb from carboxymethylcellulose microspheres in vitro and field applications, Pest. Sci. 55 (1999) 1194–1202. [14] V.A. Cooper, J.A. Nicell, Removal of phenols from a foundry wastewater using horseradish peroxidase, Water Res. 30 (4) (1996) 954–964. ´ N. Lonˇcar, B. Dojnov, A. Milovanovic, ´ M. Vujˇcic, ´ NBoˇzic, ´ Characteriza[15] Z. Vujˇcic, tion of leucylaminopeptidase from Solanum tuberosum tuber, Food Chem. 121 (2010) 418–423. [16] P.W. Thygesen, I.B. Dry, S.P. Robinson, Polyphenol oxidase in potato (a multigene family that exhibits differential expression patterns), Plant Physiol. 109 (1995) 525–531. [17] R. Axen, H. Drevin, J. Carlsson, Preparation of modified agarose gels containing thiol groups, Acta Chem. Scan. B 29 (1975) 471–474. [18] D.Y. Kwon, W.Y. Kim, Purification of the glycosylated polyphenol oxidase from potato tuber, J. Biochem. Mol. Biol. 29 (1996) 163–168. [19] I. Alemzadeh, S. Nejati, Phenols removal by immobilized horseradish peroxidase, J. Hazard. Mater. 166 (2009) 1082–1086. [20] M.A. Smith, P.C. Gillespie, Ionization of DEAE-cellulose: dependence of pK on ionic strength, J. Chrom. A 469 (1989) 111–120. [21] D.R. Lideed, CRC Handbook of Chemistry and Physics, CRC Press, BocaRaton, FL, 2005, Internet Version 2005 http://www.hbcpnetbase.com. [22] V. Prigione, V. Tigini, C. Pezzella, A. Anastasi, G. Sannia, G.C. Varese, Decolourisation and detoxification of textile effluents by fungal biosorption, Water Res. 42 (2008) 2911–2920. [23] B.M. Moore, W.H. Flurkey, Sodium dodecyl sulfate activation of a plant polyphenoloxidase, J. Biol. Chem. 265 (9) (1990) 4982–4988. [24] J.C. Espin, H.J. Wichers, Activation of a latent mushroom (Agaricus bisporus) tyrosinase isoform by sodium dodecyl sulfate (SDS). Kinetic properties of the SDS-activated isoform, J. Agric. Food Chem. 47 (9) (1999) 3518–3525. [25] D. Lopez-Serrano, A. Sanhez-Amat, F. Solano, Cloning and molecular characterization of a SDS-activated tyrosinase from Marinomonas mediterranea, Pigment Cell Res. 15 (2002) 104–111. [26] Merck Index, fourteenth ed., Merck & Co. Inc., NJ, USA, 2006. [27] H. Ashraf, Q. Husain, Studies on bitter gourd peroxidase catalyzed removal of p-bromophenol from wastewater, Desalination 262 (2010) 267–272. [28] Z. Tong, Z. Qingxiang, H. Hui, L. Qin, Z. Yi, Removal of toxic phenol and 4chlorophenol from waste water by horseradish peroxidase, Chemosphere 34 (4) (1997) 893–903. [29] K. Yamada, Y. Akiba, T. Shibuya, A. Kashiwada, K. Matsuda, M. Hirata, Water purification through bioconversion of phenol compounds by tyrosinase and chemical adsorption by chitosan beads, Biotechnol. Prog. 21 (2005) 823–829. – ´ I. Andelkovi ´ A. Milovanovic, ´ B. Dojnov, M. Vujˇcic, ´ G. Roglic, ´ [30] N. Lonˇcar, N. Boˇzic, c, ´ Removal of aqueous phenol and phenol derivatives by immobilized Z. Vujˇcic, potato polyphenol oxidase, J. Serb. Chem. Soc. 76 (4) (2011) 513–522. [31] S. Wada, H. Ichikawa, K. Tatsumi, Removal of phenols from wastewater by soluble and immobilized tyrosinase, Biotechnol. Bioeng. 42 (1993) 854–858.
Journal of Hazardous Materials 196 (2011) 79–85
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Partitioning of hexachlorobenzene in a kaolin/humic acid/surfactant/water system: Combined effect of surfactant and soil organic matter Jinzhong Wan a , Lingling Wang a , Xiaohua Lu a,∗ , Yusuo Lin b , Shengtian Zhang b a b
Environmental Science Research Institute, Huazhong University of Science and Technology, Wuhan 430074, PR China Nanjing Institute of Environmental Science, Ministry of Environmental Protection of China, Nanjing 210042, PR China
a r t i c l e
i n f o
Article history: Received 8 June 2011 Received in revised form 30 August 2011 Accepted 30 August 2011 Available online 3 September 2011 Keywords: HOCs Partitioning Triton X-100 Soil organic matter
a b s t r a c t Understanding the combined effect of soil organic matter (SOM) and surfactants on the partitioning of hydrophobic organic compounds in soil/water systems is important to predict the effectiveness of surfactant-enhanced remediation (SER). In the present study we investigate the partitioning of hexachlorobenzene (HCB) within a humic acid (HA)-coated kaolin/Triton X-100 (TX100)/water system, with special emphasis on the interaction between TX100 and HA, and their combined effect on HCB sorption. HA firstly enhanced then suppressed TX100 sorption to kaolin as the amounts of HA increased, while the addition of TX100 led to a consistent reduction in HA sorption. In the HA-coated kaolin/TX100/water system, TX100 played a primary role in enhancing desorption of HCB, while the role could be suppressed and then enhanced as HA coating amounts increased. Only at HA coating above 2.4%, dissolved HA outcompeted clay-bound HA for HCB partitioning, resulting in dissolved HA enhanced desorption. The presence of dissolved HA at these conditions further promoted the effectiveness of TX100 enhanced desorption. Despite a reduced TX100 sorption to clay was achieved due to the presence of dissolved HA, the effect on HCB desorption was comparatively slight. A reliable cumulative influence of HA and TX100 on HCB desorption was observed, although HCB desorption by HA/TX100 mixed was less than the sum of HA and TX100 individually. Our study suggests that for soils of high organic contents, the combined effect of SOM and surfactants on HOCs desorption can be applied to improve the performance of SER. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Surfactant-enhanced remediation (SER) has been proposed as an efficient clean-up technology for soils contaminated with hydrophobic organic compounds (HOCs) [1–3]. Surfactants can hypothetically remediate soils by incorporating HOCs molecules into hydrophobic micelles and enhancing the desorption (i.e., mobility or availability) of HOCs from soils [4]. Soil organic matter (SOM), a ubiquitous substance and also an important part of soil, greatly affects the SER process. SOM not only governs the mobility of HOCs in soils but also contributes to the adsorption of surfactants to soils [5–8]. Therefore, SOM imposes an adverse effect on SER due to its strong sorptive affinity for HOCs and surfactants. The dissolved fraction of SOM (DOM), however, is capable of mobilizing HOCs in soils [9,10]. As a consequence, the interaction between surfactants and SOM/DOM in soil/water systems makes the partitioning behaviors of HOCs more complex in SER process.
∗ Corresponding author. Tel.: +86 27 87792159; fax: +86 27 87792159. E-mail address:
[email protected] (X. Lu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.072
Very recently the combined effect of surfactants and DOM on HOCs dissolution was investigated. Cho et al. [11] found that the apparent solubilities of naphthalene, phenanthrene and pyrene in a Triton X-100 (TX100) and DOM mixed solution were nearly the sum of those in TX100 and DOM solution alone. On the contrary, obvious reduction of HOCs solubilities were recorded when DOM was added to sodium dodecyl sulfate (SDS) solution (with SDS concentration below critical micelle concentration (CMC)) [11,12]. However, present knowledge about the HOCs dissolution behaviors in surfactant-DOM mixed solutions may not necessarily be applicable to predict the mobility of the contaminants in soil/water systems. Cheng and Wong [13] found that the desorption of phenanthrene and pyrene from soils by Tween 80-DOM mixed solutions was more than the sum of Tween 80 and DOM alone, different from above results of HOCs dissolution by Cho et al. [11]. The authors attributed this synergistic effect (i.e., the PAH desorption by DOM and Tween 80 being more than additive) to possible formation of DOM-surfactant complexes that might possess a stronger desorbing capacity for PAHs [13]. However, we suggest that above results may be also relevant to the complex sorptive interactions between Tween 80 and DOM in soil/water systems. To be more specific, SOM may affect the adsorption of
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surfactants to soils dually (i.e., the soil-bound SOM increased the surfactant adsorption while the DOM decreased), and similar for the effect of surfactants on DOM sorption. In other words, other than the interactions between aqueous surfactants and DOM (e.g., the formation of DOM-surfactant complexes), the influence of surfactant-DOM interactions on surfactant and DOM adsorption to soils may also influence the partitioning of HOCs in soil–water systems. Therefore, a deeper insight into this interaction is essential to further verify the combined effect of surfactants, SOM, and DOM on HOCs adsorption in soil/water systems, and to better understand SER. In the present study we investigated the partitioning of hexachlorobenzene (HCB), used here as a representative HOC, in a HA-coated kaolin (HACK)/water system in the presence of TX100. TX100 was chosen as the representative surfactant since it is commonly selected in studies for soil remediation. Purified commercial HA was used as the representative SOM fraction. Our objectives were: (1) to investigate the influence of TX100-SOM interaction on TX100 and DOM sorption to clay, and the resultant effect on HCB partitioning; (2) to study the combined effect of SOM and TX100 on HCB sorption; (3) to illustrate contributions of surfactants, SOM, DOM to HCB sorption in a complicated HACK/surfactant/water system.
2. Materials and methods 2.1. Chemicals, HA and kaolin HCB (99.0%) was purchased from Shanghai General Reagent Factory, China. TX100 (>99.0%) was from Aldrich and used as received. Hexane, sodium chloride, and hydrochloric acid were all of analytical purity. Deionized water (>18.0 m) was used for solution preparation and dilution. Commercial HA powder was obtained from Lemandou Chemicals Co., Ltd., China, which was derived from lignite (Sinkiang, China). The HA was further purified according to the procedures recommended by the International Humic Substances Society (IHSS) before use [14]. The element composition of the purified HA was analyzed on a Vario Micro element analyzer (Elementar, Germany). The HA comprises 48.3% C, 2.6% H, 29.3% O, 1.1% N, and 0.3% S. Kaolin (chemical purity) obtained from Shanghai Qingpu Chemical CO. Ltd. was selected as the model clay. The organic content of kaolin was 0.12%, and the cation exchange capacity (CEC) measured by BaCl2 –H2 SO4 method (ISO 11260-1997) was 16.9 cmol kg−1 . The BET specific surface area measured by a surface area apparatus (Micromeritics Tristar 3000) was 19.4 m2 g−1 . X-ray fluorescence analysis (Genesis, EDAX Inc.) shows that its main mineral elements were Si (59.9% in wt) and Al (36.7%). Other physicochemical properties associated with the kaolin are listed as Table S1 in Supplementary Materials.
2.2. Influence of HA–TX100 interaction on HA and TX100 sorption to kaolin All the sorption experiments were carried out in triplicate using batch equilibrium technique in glass vials sealed with Teflon screw caps. Data processing and fitting (including isotherm model fitting) were performed using Origin v. 8.0. To begin with, the sorption of HA or TX100 in a kaolin/water system was studied respectively. A total of 0.5 g kaolin was mixed with 10 mL HA (0–2500 mg L−1 ) or TX100 solutions (0–10 mmol L−1 ). The vials were agitated in a reciprocating shaker end over end for 48 h (25 ± 1 ◦ C), and centrifuged at 4000 rpm for 10 min. The
supernatant was filtered through a 0.45 m cellulose acetate membrane and subjected for HA or TX100 analysis. The influence of HA–TX100 interaction on TX100 and HA sorption to kaolin was further investigated in two different manners (0.5 g kaolin in 10 mL solutions): with constant HA total concentration of 400 mg L−1 and varying TX100 total concentrations of 0–10 mmol L−1 , or with constant TX100 total concentration of 2 mmol L−1 and varying HA total concentrations of 0–2500 mg L−1 . The pH of the slurries was adjusted to 7.0 ± 0.2 with appropriate concentrations of HCl or NaOH (mostly 1 mol L−1 while 10 mol L−1 HCl or NaOH was used for samples with high HA concentrations). Furthermore, 0.01 mol L−1 NaCl was contained as the background electrolyte. Equilibrium aqueous concentrations of HA and TX100 were both analyzed for each sample. 2.3. Influence of HA or TX100 on HCB sorption to kaolin In this section firstly the influence of HA addition on HCB sorption to kaolin was investigated. A total of 0.5 g kaolin was mixed with 10 mL HA solutions (0–2500 mg L−1 , final pH 7). Then 40 L of HCB acetone solution was added into the vials by a microsyringe before shaking (acetone fraction was below 0.5%). The total concentration of HCB was 0.5 mg L−1 (or 10 mg kg−1 kaolin). After equilibrium and centrifugation, the supernatant was filtered through a 0.45 m cellulose acetate membrane. HCB in the supernatant was then extracted by hexane via liquid-liquid extraction in 11-mL glass vials (with an extraction ratio of 1:2, agitated in a shaker for 2 h) and analyzed by gas chromatography (GC). Meanwhile HA in the filtrate was measured. Secondly, the influence of TX100 on HCB sorption was investigated which followed the same procedure as described above. TX100 concentration used was 0–10 mmol L−1 . Both TX100 and HCB equilibrium concentrations were determined for each sample. 2.4. HCB partitioning in a solid HA/TX100/water system Predetermined volumes of HA and TX100 stock solutions (HA 5 g L−1 and TX100 20 mmol L−1 ) were pipetted into 50 mL glass flasks, and diluted with deionized water to 20 mL. The amounts of HA added were as 0–2.5 g L−1 , and the total TX100 concentration was 2 mmol L−1 . The pH of the mixture was adjusted to 2–3 with HCl to precipitate HA thoroughly. Then 80 L HCB acetone solution was added before shaking. The total concentration of HCB was 0.5 mg L−1 . After equilibrium and centrifugation, the supernatant was filtered through a 0.45 m cellulose acetate membrane and subjected for TX100 and HCB analysis. 2.5. HCB partitioning in a HACK/TX100/water system The sorption of HCB to HACK (detailed preparation procedures are provided in Supplementary Materials) in the presence of TX100 was further investigated. A total of 0.5 g HACK with varied HA coating amounts (0.25–5%) was mixed with 10 mL TX100 solutions (0–10 mmol L−1 ). The pH of the slurry was then adjusted to 7.0 ± 0.2 or 3.0 ± 0.2 (pH 3.0 could avoid HA dissolution from HACK, thereby better reveal the role of solid HA in the partitioning of TX100 and HCB). Then 40 L HCB acetone solution was added into the vials by a micro-syringe before shaking (the final acetone fraction was below 0.5%). Both TX100 and HCB equilibrium concentrations were determined for each sample. 2.6. Chemical analysis Aqueous TX100 was analyzed by a high performance liquid chromatography (Hitachi L7100, Japan) equipped with an
J. Wan et al. / Journal of Hazardous Materials 196 (2011) 79–85
Fig. 1. (a) TX100 sorption isotherm at 0 and 400 mg L−1 HA; (b) influence of HA amount on TX100 (2 mmol L−1 ) sorption to kaolin.
L-7420 ultraviolet–visible (UV–vis) detector and an Agilent ZORBAX Eclipse XDB-C18 column (Agilent, USA). The wavelength was set at 223 nm. The mobile phase was 90% methanol plus 10% water, with a flow rate of 1.0 mL min−1 . HCB in the hexane was determined on a Hewlett–Packard 6890 GC equipped with an electron capture detector and a ZB-5 capillary column (Phenomenex, USA). Detailed information for GC procedure was included in our previous study [15]. The aqueous concentration of HA was measured by a Cary 50 UV–vis spectrophotometer (Varian, USA) at 254 nm (although TOC analysis is frequently used to quantify the humic substances (including HA), herein the co-presence of TX100 especially above 2 mmol L−1 may seriously interfere the TOC analysis of HA. However, a much higher absorptivity of HA under UV-254 than TX100 suggest that the spectrophotometer measurement might be more appropriate). The linear range of the HA working curve was 0–24 mg L−1 (r = 0.9999). The absorbance of HA in the absence of TX100 can be obtained directly. Based on the preliminary observation that the absorbance of TX100 and HA mix at 254 nm was additive, HA content could be deduced by subtracting the absorbance contributed by TX100 from the absorbance of the TX100-HA mixture (the absorbance of TX100 can be calculated from its working curve at 254 nm and the corresponded concentration obtained by HPLC). 3. Results and discussion 3.1. Effects of SOM on TX100 sorption in a kaolin–water system Fig. 1a displays the sorption isotherms of TX100 in the absence and presence of HA (400 mg L−1 ) in a kaolin/water system. The two isotherms can be well fitted by the Langmuir model in the range of 0–10 mmol L−1 , similar to those reported in the literature [7,16]. Comparison of the maximal sorption amounts of TX100 between HA and HA-free systems (24.0 vs 19.1 mmol kg−1 ) indicates that
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the presence of HA at 400 mg L−1 increased the sorption of TX100. Enhanced sorption of surfactant to soil as an effect of SOM has been reported by a number of researchers [17–19]. Zhu et al. [20] revealed that the sorption capacity of HA for TX100 was almost one order larger than that of kaolin. Zhang et al. [21] also reported that much higher sorption capacity and partitioning coefficient (Kd ) for TX100 was exhibited by HA relative to original soil. Furthermore, as shown in Fig. 1b, for the observed HA concentration range, the sorption of TX100 firstly increased and then decreased, with a maximum at HA concentration of about 400 mg L−1 . The initial increase in TX100 sorption was strongly correlated with the partitioning of surfactant to the clay-bound HA. However, as the added HA increased above 400 mg L−1 , the amount and more importantly, the fraction of the dissolved HA increased considerably (as indicated by decreasing Kd values in Table S2). The dissolved HA may compete with the clay-bound HA for the partitioning of TX100, since HA molecule contains both hydrophilic and hydrophobic structure, very similar to surfactants. Another explanation may be that the dissolved HA could reduce the TX100 adsorption by forming HA–TX100 complexes, similar to the complexing of HOCs to DOM [10,22,23]. Lee et al. [22] found that the sorption isotherm of TX100 to a Florida peat was of a skewedGaussian shape, with a sharp decrease in TX100 sorption coefficient at equilibrium concentration above 1000 mg L−1 . It was suggested that at higher TX100 concentrations, more SOM was dissolved, and the DOM was expected to promote the “dissolution” of TX100 and reduce the probability of TX100 sorption [22]. Inspection of Fig. 1b further reveals that even at relatively high concentrations of dissolved HA (100–400 mg L−1 ), the sorption of TX100 was higher than sorption at HA = 0, suggesting a considerable binding of TX100 to the kaolin-bound HA. It can be also estimated from Fig. 1b that the critical HA amount when the dissolved HA outcompeted the clay-bound fraction for TX100 partitioning was approximately 1500 mg L−1 . As a result, it is expected that the effect of SOM on the sorption of nonionic surfactants is dependent on the content of SOM, and only at content high enough as to induce a significant amount of DOM, SOM can reduce the sorption of surfactants, thereby functioning positively to SER.
3.2. Effects of TX100 on HA sorption in a kaolin/water system The partitioning of HA within a kaolin/water system is presented as Fig. 2a. The sorption isotherm of HA in the TX100-free system can be described as a two phase linear relationship: with a steep increase for dissolved HA concentration of 0–7 mg L−1 and a followed slowing increase in the range of 7–360 mg L−1 . Moreover, a linear regression can be applied for HA sorption in the presence of TX100 when the aqueous concentration of HA ranged from 29 to 610 mg L−1 . Inspection of Fig. 2a reveals that the copresence of TX100 at 2 mmol L−1 decreased HA sorption to kaolin. Furthermore, the slope of HA isotherm in the presence of TX100 is lower than the HA isotherm in the absence of TX100 (5.7 vs 7.7), suggesting an increased reduction in HA sorption as the dissolved HA increases. Additionally, Fig. 2b shows a consistent decrease in HA sorption with an increase of TX100 concentration (0–8 mmol L−1 ). In particular, the presence of 8 mmol L−1 TX100 could reduce the sorption of HA by 38% compared with TX100-free system. Similar observations were also recorded by Cheng and Wong [13], therein the presence of Tween 80 at 150 mg L−1 dramatically reduced the sorption of DOM to soil. Preferential binding of Tween 80 molecules to soil with respect to DOM was hypothesized to be relevant to the above results [13]. Despite the lower calculated partitioning coefficients for TX100 (2–34 L kg−1 at 1–10 mmol L−1 , Table S1) in comparison with HA (23.5–80 L kg−1 at 350–1000 mg L−1 , Table S1), it may still be reasonable to suggest that the competitive
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Fig. 2. (a) HA sorption isotherm at 0 and 2 mmol L−1 TX100; (b) influence of TX100 concentration on HA (400 mg L−1 ) sorption to kaolin.
adsorption between the two molecules to the clay surface contributed primarily to the declined sorption of HA as the aqueous TX100 increased. 3.3. HCB partitioning within a HACK/TX100/water system Although the influence of TX100-HA interaction on TX100 and HA sorption to kaolin has been verified, the resultant effects on HCB partitioning in a clay/water system remain unclear. In fact, in a complicated surfactants/SOM/water/clay system, a series of sorption and complexing processes may occur and affect HCB partitioning behavior. To be more specific, these interactions include (1) the formation of DOM-HCB complex that may impede HCB sorption [10,24]; (2) the role of clay-bound SOM as a hydrophobic sorption domain for HCB [23,25,26]; (3) partitioning of HCB to TX100 micelles that can effectively reduce the potential of HCB sorption [3,27]; (4) the influence of TX100-HA interaction on TX100 and HA sorption to clay surface that may further affect the partitioning of HCB; (5) the possible sorption of HCB by the clay-bound TX100 and by the mineral [27–29]. Given the complexity of HOCs partitioning in the co-presence of TX100 and SOM, HCB sorption in SOM and TX100 individual system was investigated first. It can be found from Fig. 3a that HCB sorption correlated negatively with the amount of HA added at HA > 250 mg L−1 . The fraction of sorbed HCB decreased steadily from 1.0 to 0.5 as HA increased to 2500 mg L−1 . With the added HA increasing, the dissolved HA became increasingly abundant, therefore impeding the sorption of HCB. Nevertheless, at a low HA range of 0–150 mg L−1 , HCB sorption increased slightly. It is generally accepted that HA could bind HOCs through hydrophobic interactions to form HA–HOCs complexes [10,24,30], by which means the soil-bound HA sorb HOCs from solution while the dissolved HA desorb HOCs from soils. On the other hand, the formation of micelle-like structure of HA due to molecule aggregation is also proposed as the mechanism for the solubilization/enhanced desorption of HA for HOCs [9,27]. Furthermore, Fig. 3b indicates
Fig. 3. Influence of (a) HA and (b) TX100 on HCB sorption to kaolin individually. HCB 0.5 mg L−1 , pH 7.0.
that the sorption of HCB experienced a steady decline with the increase of TX100 added (0–2 mmol L−1 , Fig. 3b). Particularly, when 2 mmol L−1 of TX100 was added, the fraction of HCB sorbed was about 0.4. It is suggested that only at total surfactant concentration above critical desorption concentration (CDC), i.e., the aqueous surfactant concentration above CMC, reliable enhancement in HCB desorption by surfactants can be obtained [3,31]. The CDC of TX100 herein can be estimated from following equation: CDC = CMC + CsCMC wherein CsCMC is the concentration of TX100 in soil when the corresponding aqueous concentration is at the CMC, which can be calculated from CsCMC = 19.1CMC/(0.085 + CMC) (Fig. 1a). By substituting the measured CMC value of TX100 of 0.1 mmol L−1 into above equations, the estimated CDC is 0.62 mmol L−1 . As a consequence, no apparent desorption of HCB was recorded at total TX100 concentration of 0.5 mmol L−1 , while remarkable HCB desorption was obtained at TX100 dosage above 1 mmol L−1 . The partitioning of HCB in a HACK/TX100/water system, and the effects of clay-bound HA and dissolved HA on the sorption of HCB to HACK are illustrated in Fig. 4a–c. From Fig. 4a it is evident that the sorption of TX100 and HCB to solid HA both correlated positively with the HA content. With the HA concentration increasing from 0 to 2.5 g L−1 , the aqueous TX100 (initially 1 mmol L−1 ) decreased substantially to nearly undetectable, and correspondingly, the sorption of HCB (initial aqueous concentration was 0.5 mg L−1 ) increased steadily to 1.0. The reduced aqueous TX100 concentration was directly associated with the high affinity of HA for TX100, as the Kd values for TX100 herein were estimated as (0.6–8) × 103 L kg−1 , which are 1.8–2.9 orders of magnitude larger than that of TX100 to kaolin (10.9 L kg−1 at TX100 of 2 mmol L−1 ). The reduced aqueous HCB concentration, however, was mainly relevant to the strong affinity of both HA and HA-bound TX100 [27]. The binding constant (Kb ) for the HA–HCB complex was estimated as 5.5 × 104 L kg−1 by using the solubility enhancing method [24,32]
J. Wan et al. / Journal of Hazardous Materials 196 (2011) 79–85
83
Fig. 5. Estimation of factors contributed to HCB sorption, wherein “HA only” and “TX100 only represent HCB sorption as the inherent effect of HA (Fig. 3a) and 2 mmol L−1 TX100 only (Fig. 3b), respectively. “Theoretical effect of TX100 in TX100/HACK system” represents the theoretical HCB sorption by TX100 under the interference of HA coating, calculated from the aqueous TX100 content (Fig. 4c) and the inherent effect of TX100 on HCB sorption (Fig. 3b).
of HA coating amount of 0.5–5% (Fig. S2, the slope of the “SOM content” linear was less than 0.5). As a consequence, the formation of aqueous HA–HCB complex became increasingly pronounced. Furthermore, the reduced sorption of TX100 by the presence of HA may be further responsible for the decreasing sorption of HCB. As indicated in Fig. 4c, the dissolved fraction of TX100 increased from 1.3 to above 1.6 mmol L−1 as the amount of HA coating increased, slightly deviated from the results of Fig. 1b. A plausible explanation for this deviation was that for HACK the clay surface was preliminarily occupied by HA, and the replacement of TX100 for the bound-HA was somewhat harder than the competitive sorption with the dissolved HA (Section 3.1). Finally, the possibly synergistic effect of HA–TX100 complexes on HCB solubilization and reduced adsorption in the HACK/surfactant/water systems cannot be ruled out [13]. 3.4. Contributions of SOM and TX100 to HCB sorption
Fig. 4. Partitioning of TX100 and HCB within (a) solid HA/water, (b) HACK/water (pH 3.0) and (c) HACK/water (pH 7.0) system. Initial concentrations for HCB and TX100 were 0.5 mg L−1 and 2 mmol L−1 , respectively.
(detailed in the notification of Fig. S1 in Supplementary Materials), suggesting a rather strong interaction between HA and HCB. Furthermore, the reduced aqueous TX100 concentration as a function of solid HA could also contribute to the increased sorption of HCB considering the poor inherent solubility of HCB in water. Similar trends for TX100 and HCB sorption were also observed in the HACK/water system at pH 3 (Fig. 4b). Note that much lower aqueous TX100 (about 0.9 mmol L−1 ) and higher HCB sorption (over 0.6) were recorded at a very low HA content of 0.25% (Fig. 4b), which can be mainly attributed to the sorption by kaolin. The important role of mineral components in the partitioning of either surfactants or HOCs especially at low organic carbon contents has been addressed by other researchers [33,34]. Fig. 4c depicts the partitioning of TX100 and HCB at pH 7.0 within the HACK/water system. Contrary to the trends at pH 3.0 (Fig. 4b), the fraction of sorbed HCB decreased from 0.47 to 0.26 with the HA coating ranging from 0.25% to 5%. The declining sorption of HCB was consistent with the aforementioned observations in the HA/kaolin system (Fig. 3a). It is noteworthy to address that although an increase in HA coating may result in a direct increment in SOM content of kaolin, the fraction of dissolved HA was apparently higher than the fraction bound to kaolin all along the range
Based on above results of HCB partitioning behaviors in individual systems, in this section we will further summarize and compare the individual and combined roles of HA and TX100 in HCB sorption. For this purpose, Fig. 5 and Table 1 were constructed. The dashed curve (designated as “HA only”, i.e., Fig. 3a) and dotted curve (designated as “TX100 only”, i.e., Fig. 3b) represent HCB sorption on clay when HA (125–2500 mg L−1 ) and TX100 (2 mmol L−1 ) were added individually. The solid curve (designated as “HA & TX100 both”, i.e., Fig. 4c) displays HCB sorption due to the copresence of TX100 and HA. The curve of dash-dot (designated as “Theoretical effect of TX100 in TX100/HACK system”) is the theoretical HCB sorption caused by 2 mmol L−1 TX100 in the TX100/HACK system, in which the effect of additional aqueous TX100 (resulted from HA coating, Section 3.3 and Fig. 4c) on HCB sorption was involved (see detailed deduction procedure from Notes of Table 1). It should be mentioned that a neutral pH of 7.0 was set for all above systems, thus the influence of pH could be neglected. Inspection of Fig. 5 indicates that TX100 played a dominant role in the reduction of HCB sorption. Other than increasing HCB solubility by the inherent aqueous TX100 (in HA-free system), the reduced sorption of TX100 due to the co-presence of HA may lower the sorption of HCB additionally. This additional fraction of sorbed HCB was expected to be about 0.04, as revealed by comparing the result of “TX100-observed” and “TX100-Theoretical in TX100/HACK system” (Table 1). However, the effect of HA coating on HCB partitioning was dual depending on the coating amount. Although a dramatic decline in HCB sorption was obtained as a function of increasing HA coating in the whole range, the presence of HA was found to promote the HCB sorption
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J. Wan et al. / Journal of Hazardous Materials 196 (2011) 79–85
Table 1 Contributions of TX100 and HA to HCB desorption from HACK. HA coating (%)
0.5 1.0 2.0 3.0 4.0 5.0
Fraction of HCB desorbed in different systems, Ce/C0 HA-observeda
TX100observeda
TX100-theoretical in HACK/TX100 systemb
HA & TX100 both-theoreticalc
Sum-HA & TX100 individuallyd
0.03 0.24 0.26 0.40 0.44 0.50
0.59 0.59 0.59 0.59 0.59 0.59
0.61 0.63 0.63 0.63 0.63 0.63
0.53 0.54 0.58 0.62 0.65 0.74
0.61 0.83 0.84 0.99 1.03 1.10
a
The observed fraction of HCB desorption by HA individually (results of Fig. 3a). Similar for “TX100-observed” (results of Fig. 3b). The theoretical fraction of HCB desorption as an effect of 2 mmol L−1 TX100 in an TX100/HACK system, calculated by substituting the value of the aqueous TX100 (Fig. 3c) to the following equation that correlates HCB sorption to aqueous TX100 concentration: b
FHCB =
0.109 + 1.08 1 + 1.95CeTX100 0.946
wherein FHCB is the fraction of HCB sorbed to HACK, CeTX100 is the aqueous TX100 concentration. c The observed fraction of HCB desorption as the effects of both HA coating and TX100 addition. d The theoretical fraction of HCB desorption by TX100 and HA, i.e., the sum of fraction of HCB desorption by HA and TX100 individually.
at lower coating amounts if taking the result of “TX100 only” (Fig. 5) as the reference. As discussed above, the competition between the aqueous and clay-bound HA determines the effect of HA coating on HCB sorption [10,23], regardless of the interference by TX100. Note that the intersection point of the “HA & TX100 both” and the “TX100 only” curve (P1 in Fig. 5) represents an equilibrium of the facilitating and impeding role of HA coating in HCB sorption. It is speculated that only at HA coating higher than 2.4%, the dissolved HA outcompeted the clay-bound HA for HCB partitioning. Excess HCB desorption from clay-bound HA could be reached due to this dissolved HA, compared to HCB desorptoion by TX100 alone. If we actually take the additional effect of sorption reduction of TX100 into account, the intersection would shift toward the higher HA coating, as depicted in Fig. 5 (P2), which means the positive effects on the reduction of HCB sorption by HA is expected at the coating amount above 3.3%. Inspection of Fig. 5 further reveals that P1 is actually the critical point at which the combined effects of HA and TX100 on HCB sorption equals to that of TX100 alone. That implies that at HA coating higher than P1, the co-presence of TX100 and HA resulted in less HCB sorption than either HA or TX100 alone. Nevertheless, as indicated by Table 1, the combined effect on reduction of HCB sorption was less than the sum of TX100 and HA alone. The results seemingly conflict with previous observations that a synergistic effect on HOCs desorption from soil by co-addition of Tween 80 and DOM was recorded [13]. However, note that the soil texture (one loam sandy soil) and the procedures for agents adding (simultaneous addition of Tween 80 and DOM into the system) therein would probably generate less sorption of DOM as compared with the HACK in our study. Furthermore, since the fractions of HCB “desorption” were 60% and over 40% for TX100 (2 mmol L−1 ) and HA (at HA-coating ≥2%) individually (Table 1), no synergistic or even additive effect can be expected by the combination of TX100 and HA; while as reported by Cheng and Wong [13], even when a synergistic effect was reached, the maximal desorption ratio was as low as 16.2% and 10.9% for phenanthrene and pyrene, respectively. We presumed that similar synergistic or additive effect by the copresence of TX100 and HA can be reached if a proper TX100 and HA concentration was applied (at least, the sum of HCB desorption efficiency by TX100 and HA alone was apparently below 100%).
4. Conclusions A deeper insight into the combined effect of surfactants, SOM, and DOM on HOC adsorption in a soil/water system is required to
better predict the efficacy of SER. Herein we investigated the partitioning of HCB in a HACK/water system in the presence of TX100. Main conclusions can be summed up as follows: (1) Appreciable influence of TX100-HA interaction on TX100 and HA sorption to clay was observed. The addition of HA at lower and higher amount than 1500 mg L−1 was found enhancing and reducing the sorption of TX100 to kaolin, respectively. Furthermore, the presence of TX100 suppressed the sorption of HA to kaolin over the entire observed concentration range of 0.5–2 mmol L−1 . The presented results suggest that for soils of high organic contents, the surfactant-SOM interaction may be beneficial to SER. In addition, a properly high dosage of surfactant may be more preferable to dissolve the SOM and exert such extra desorption. (2) TX100 contributed primarily to HCB desorption in a HACK/TX100/water system. DOM also showed encouraging enhancement in HCB desorption. The coated HA, however, imposed a negative and positive influence on TX100-enhanced desorption of HCB at coating amounts below and above 2.4%, respectively. The combined effect of HA and TX100 on HCB desorption was less than the sum of TX100 and HA alone. (3) It should be noted that HACK may or may not fully represent the whole soil in nature. However, the results we obtained are still of practical significance for SER since HA is often intimately associated with clay minerals in soil, and both HA and clay minerals are primary soil compounds that determine the sorption of surfactants and HOCs, i.e., the performance of SER. (4) Our study suggests that for soils of high organic contents, a combined influence of SOM and surfactants on HOCs removal can be expected, which implies a higher performance of SER, or a smaller dosage of surfactant may be required. However, the critical content of SOM that differentiates the positive role of SOM from the negative one in the SER has been revealed herein, the value may vary from site to site, depending on a series of factors such as characteristics of soil, SOM, surfactants, and HOCs. Furthermore, pH as an important environmental variable may influence the amount of DOM, therefore influencing the SER process. However, for the consideration of system simplification, in present study we only chose pH 7 as a typical circumstance. As a result, further studies are still required to focus on the combined effect of surfactants and SOM on SER for real soils with different organic contents, a broader range of pH conditions and a variety of HOCs.
J. Wan et al. / Journal of Hazardous Materials 196 (2011) 79–85
Acknowledgements This work was supported by the National Natural Science Foundation of China (Grants 20777024), the National High-Tech Research and Development (863) Programme (2009AA063103), and Shanghai Tongji Gao Tingyao Environtal Protection Sci. & Tech. Development Foundation. The Analytical and Testing Center of Huazhong University of Science and Technology is thanked for its help in kaolin and HA characterization. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.08.072. References [1] M. Svab, M. Kubala, M. Muellerova, R. Raschman, Soil flushing by surfactant solution: pilot-scale demonstration of complete technology, J. Hazard. Mater. 163 (2009) 410–417. [2] S.H. Yuan, Z. Shu, J.Z. Wan, X.H. Lu, Enhanced desorption of hexachlorobenzene from kaolin by single and mixed surfactants, J. Colloid Interface Sci. 314 (2007) 167–175. [3] K. Yang, L.Z. Zhu, B.S. Xing, Enhanced soil washing of phenanthrene by mixed solutions of TX100 and SDBS, Environ. Sci. Technol. 40 (2006) 4274–4280. [4] C.N. Mulligan, R.N. Yong, B.F. Gibbs, Surfactant-enhanced remediation of contaminated soil: a review, Eng. Geol. 60 (2001) 371–380. [5] W. Huang, W.J. Weber, A distributed reactivity model for sorption by soils and sediments. 10. Relationships between desorption, hysteresis, and the chemical characteristics of organic domains, Environ. Sci. Technol. 31 (1997) 2562–2569. [6] R. Chefetz, A.P. Deshmukh, P.G. Hatcher, Pyrene sorption by natural organic matter, Environ. Sci. Technol. 34 (2000) 2925–2930. [7] M.J. Salloum, M.J. Dudas, W.B. Mcgill, S.M. Murphy, Surfactant sorption to soil and geologic samples with varying mineralogical and chemical properties, Environ. Toxicol. Chem. 19 (2000) 2436–2442. [8] F.J. Ochoa-Loza, W.H. Noordman, D.B. Jannsen, M.L. Brusseau, R.M. Maier, Effect of clays, metal oxides, and organic matter on rhamnolipid biosurfactant sorption by soil, Chemosphere 66 (2007) 1634–1642. [9] P. Conte, A. Agretto, R. Spaccini, A. Piccolo, Soil remediation: humic acids as natural surfactants in the washings of highly contaminated soils, Environ. Pollut. 135 (2005) 515–522. [10] M. Rebhun, F. De Smedt, J. Rwetabula, Dissolved humic substances for remediation of sites contaminated by organic pollutants. Binding-desorption model predictions, Water Res. 30 (1996) 2027–2038. [11] H.H. Cho, J. Choi, M.N. Goltz, J.W. Park, Combined effect of natural organic matter and surfactants on the apparent solubility of polycyclic aromatic hydrocarbons, J. Environ. Qual. 21 (2002) 275–280. [12] H. Lippold, U. Gottschalch, H. Kupsch, Joint influence of surfactants and humic matter on PAH solubility. Are mixed micelles formed? Chemosphere 70 (2008) 1979–1986. [13] K.Y. Cheng, J.W.C. Wong, Combined effect of nonionic surfactant Tween 80 and DOM on the behaviors of PAHs in soil–water system, Chemosphere 62 (2006) 1907–1916.
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Journal of Hazardous Materials 196 (2011) 86–92
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Geopolymers prepared from DC plasma treated air pollution control (APC) residues glass: Properties and characterisation of the binder phase Ioanna Kourti a , Amutha Rani Devaraj a,b , Ana Guerrero Bustos c , David Deegan d , Aldo R. Boccaccini b,e , Christopher R. Cheeseman a,∗ a
Department of Civil and Environmental Engineering, Imperial College London, London SW7 2AZ, UK Department of Materials, Imperial College London, London SW7 2BP, UK c Institute of Construction Science Eduardo Torroja (CSIC), C/Serrrano Galvache, 4, 28033 Madrid, Spain d Tetronics Ltd., South Marston Business Park, Swindon, Wiltshire SN3 4DE, UK e Institute of Biomaterials, Department of Materials Science and Engineering University of Erlangen-Nuremberg, Cauerstr. 6, 91058 Erlangen, Germany b
a r t i c l e
i n f o
Article history: Received 14 February 2011 Received in revised form 30 August 2011 Accepted 31 August 2011 Available online 6 September 2011 Keywords: Geopolymers Plasma Incineration Energy Waste APC residues
a b s t r a c t Air pollution control (APC) residues have been blended with glass-forming additives and treated using DC plasma technology to produce a high calcium aluminosilicate glass (APC glass). This has been used to form geopolymer–glass composites that exhibit high strength and density, low porosity, low water absorption, low leaching and high acid resistance. The composites have a microstructure consisting of un-reacted residual APC glass particles imbedded in a complex geopolymer and C–S–H gel binder phase, and behave as particle reinforced composites. The work demonstrates that materials prepared from DC plasma treated APC residues have potential to be used to form high quality pre-cast products. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The air pollutions control systems at energy from waste (EfW) plants burning municipal solid waste (MSW) produce granular air pollution control (APC) residues. These are a hazardous waste with an absolute entry in the European Waste Catalogue (19 01 07*), and they contain fly ash, excess lime, carbon, relatively high concentrations of volatile heavy metals and soluble salts, particularly leachable chlorides. They also contain trace levels of organics including dioxins and furans. DC plasma technology provides a sustainable treatment for APC residues that meets the aims of the EU waste policy as it is a recycling/recovery option higher in the waste management hierarchy than alternative options [1,2]. In the DC plasma treatment process APC residues are combined with glass-forming additives and melted to produce inert APC glass [2]. There is increasing interest in developing sustainable construction products which contain recycled materials. Reuse of APC glass would have significant economic and environmental benefits, and
∗ Corresponding author. Tel.: +44 207 594 5971; fax: +44 207 823 9401. E-mail address:
[email protected] (C.R. Cheeseman). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.081
would help to make DC plasma treatment of APC residues commercially attractive. Previous work has investigated the use of APC glass in glass-ceramics [3,4] and sintered ceramic tiles [5]. Geopolymers are synthetic alumino-silicates consisting of silica (SiO4 ) and alumina (AlO4 ) tetrahedra, linked by shared oxygen atoms [6]. Their formation is based on the chemistry of alkaliactivated inorganic binders and involves the chemical reaction of geopolymeric precursors, such as alumino–silicate oxides with alkali poly-silicates, to form polymeric Si–O–Al bonds [7]. The negative charge of Al3+ in four-fold coordination is balanced by the presence of positive ions such as Na+ , K+ and Ca2+ in framework cavities [8]. The empirical formula of geopolymers is therefore: Mn (–(SiO2 )z –AlO2 )n ,wH2 O
(1)
where M is a cation such as Na+ , K+ or Ca2+ , z is 1, 2 or 3, and n is the degree of poly-condensation. Geoploymers are associated with low CO2 emissions compared to Portland cement [9]. Using APC residue glass to form a geopolymer would provide a low-carbon emission reuse option that does not involve thermal processing. The microstructure and properties of geopolymers are determined by the raw materials used, and they can have high early
I. Kourti et al. / Journal of Hazardous Materials 196 (2011) 86–92 Table 1 Chemical composition of APC glass. Oxide
Composition (wt%)
Na2 O MgO Al2 O3 SiO2 P2 O5 K2 O CaO TiO2 Mn3 O4 Cr2 O3 Fe2 O3 Cl
2.88 2.31 14.78 41.10 0.77 0.03 32.59 1.19 0.23 0.06 4.07 2.5
compressive strengths, low shrinkage, rapid or slow setting, good acid resistance and fire resistance, and low thermal conductivity [10–12]. The use of APC glass in geopolymers results in geopolymer–glass composites in which residual APC glass particles act as rigid inclusions in a geopolymer matrix [13]. The presence of calcium in geopolymer systems can result in the formation of calcium silicate hydrate (C–S–H) gel and Al substituted C–S–H gel, and these are reported to decrease porosity and increase the strength of geopolymers [14–25]. This coexistence of geopolymer gel and hydration products has also been observed in alkali activated fly ash-Portland cement blends [26]. Recent research on the effect of alkalis and Al in C–S–H gel has confirmed that the structure is modified by alkali metals and Al [27,28]. The formation of a geopolymer phase or C–S–H gel is determined by the chemical and mineralogical composition, the physical properties of the aluminosilicate and Ca sources, the alkalinity of the activator and the percentage Ca in the system [17–19,22,24]. This work follows from previous research [13] in which novel geopolymers prepared using APC glass were described. The effect of processing parameters on geopolymerisation of APC glass was examined. In the present paper the properties of optimised APC glass geopolymers and the formation of a complex binder phase are investigated in detail. This provides new insight and information on the potential applications of this material and the link between the final geopolymer composite properties and microstructure. 2. Materials and methods 2.1. Materials Glass produced by DC plasma treatment of APC residues was supplied by Tetronics Ltd. (Swindon, UK) in the form of a coarse granular material with <2 mm particles. These were then milled to form a fine powder (TEMA Machinery Ltd.) with a broad particle size range, with all particles <200 m. The chemical composition of APC glass is presented in Table 1. This shows that the composition (wt%) in terms of major oxides was SiO2 (41.1%), CaO (32.6%) and Al2 O3 (14.8%). Sodium silicate solution (VWR, Lutterworth, UK) containing 26.5 wt% SiO2 , 8.5 wt% Na2 O and 65 wt% H2 O and a density of 1310 kg/m3 was used. NaOH (Fisher Scientific, Loughborough, UK) and distilled water were used to form the activating solution. 2.2. Preparation of samples and characterisation Sample preparation involved mixing the APC glass powder with the activating solution. This was prepared by dissolving NaOH pellets in water and allowing the solution to cool to room temperature before adding the required amount of sodium silicate solution. The Si/Al ratio was 2.6, the solid to liquid ratio was 3.4 and the sodium hydroxide concentration in the activating solution was 6 M. The
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activating solution was thoroughly mixed with the APC glass powder for 10 min in a 5 litre mortar mixer. The resulting paste was cast into rectangular moulds (80 mm × 25 mm × 25 mm) on a vibrating table to reduce air voids. The samples were covered with glass slips, de-moulded after 24 h, wrapped in cling film to inhibit water evaporation and cured at room temperature. APC glass geopolymers were characterised for unconfined compressive strength (UCS) at a loading rate of 300 kPa/s on four samples of each composition. The bulk density and porosity of the APC glass geopolymer samples were determined by mercury porosimetry (Micrometrics AutoPore IV 9500). Bar specimens (60 mm × 8 mm × 6 mm) were cut from APC glass geopolymers for three-point flexural strength testing, with one surface of each bar ground and polished and the adjacent edges chamfered prior to testing on a 30 mm span, using a cross-head speed of 0.5 mm/s (Hounsfield H5KS, Salfords, UK). The modulus of elasticity was calculated using the ultrasonic pulse velocity method [29]. Thermal conductivity was measured using a TT-TC Probe (ThermTest Inc.) which provides a non-destructive thermal conductivity test based on the Mathis modified hot-wire technique [30]. Freeze/thaw resistance was determined using the method described in ASTM standard C 1262-05 [31]. Three test samples were subjected to 92 freeze/thaw cycles, with the cumulative mass loss of the samples given as a percentage of the initial mass. The chemical resistance of APC glass geopolymer was determined according to the method described in BS EN 1344: 2002 [32]. APC glass geopolymers were ground and sieved to between 500 and 800 m. The required amount of sample (∼100 g) was then treated with boiling sulphuric and nitric acid (75 ml of 10% sulphuric acid and 25 ml of 10% nitric acid) for 1 h in a round bottom flask, fitted with a reflux condenser. After treatment the residue was rinsed to remove the acid, dried to constant mass, and re-weighed. Results are expressed as the loss in mass of the sample as a percentage of the original mass, with the test performed in duplicate. Leaching of heavy metals from APC glass geopolymer was evaluated using the EU compliance leaching test for granular waste (BS EN 12457-4), on particles less than 10 mm using a liquid to solid ratio (L/S) of 10 L/kg, with water as the leachant [33]. The eluate was vacuum filtered and elemental analysis of leachates completed using inductively coupled plasma atomic emission spectroscopy (ICP-AES, Varian Vista-Pro, Australia). Fractured surfaces were gold coated before examination by scanning electron microscopy (SEM, JEOL JSM 5610LV). Surfaces polished to a 1 m surface finish and carbon coated were also examined (JEOL-JSM-840A). The binder phase in APC glass geopolymers was characterised using the salicylic acid/methanol (SAM) selective extraction method. This test was developed to dissolve alite, belite and free lime from cement clinker. It has previously been used in geopolymer research and the results have shown that salicylic acid–methanol dissolves calcium silicates, while the geopolymer network structure and the un-reacted materials remain unaltered [17,24,26,27,34]. Measuring the amount of SiO2 in the initial material and the solid residue provides an indication of the amount of SiO2 that is fixed either as C–S–H gel or geopolymer gel. Si, Al and Ca in the initial sample and the solid residues were determined by ICP-AES analysis. The initial sample and the solid residue were analysed by FTIR spectroscopy to identify changes in structure of the APC glass geopolymer after SAM selective extraction. FTIR (MagnaIR 560 Spectrometer E.S.P., Nicolet) provides information on the chemical bonds in geopolymers and identifies products formed during geopolymerisation. Samples were ground to a fine powder and pellets prepared with the addition of KBr. The crystalline phases in APC glass geopolymers and solid residues were analysed by XRD (Philips PW1700) using CuK␣ radiation and a secondary graphite crystal as mono-chromator.
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Fig. 1. SEM image of polished surface of APC glass geopolymer.
3. Results 3.1. Properties of APC glass geopolymers A polished surface of APC glass geopolymer is shown in Fig. 1. A significant volume of residual APC glass particles was encapsulated in a geopolymer derived matrix phase. The properties of APC glass geopolymers investigated in this research and various commercial products are presented in Table 2. The term geopolymer cement in Table 2 refers to fly ash based geopolymers or rock based geopolymers which harden at room temperature [8]. The values presented for concrete refer to a typical concrete mix containing 10–15% cement, 60–75% aggregates and 15–20% water [35]. 3.2. Mechanical properties and Young’s modulus of elasticity APC glass geopolymer exhibits high compressive strength (∼110 MPa after 28 days curing), which is much higher than the strength of typical hardened concrete (∼30 MPa) [36], and higher than most geopolymer cements [8]. Flexural strengths of the APC glass geopolymer show similar behaviour to concrete, and are typically 10–15% of the compressive strength [37]. The flexural strength of APC glass geopolymers is 10.5 MPa which is ∼10% of the compressive strength. This value is lower than flexural strength values observed for glass and commercial tiles [5,36]. After 3-point bending testing, the APC glass geopolymer samples developed surface cracking but did not break into separate pieces indicating high fracture toughness. The SEM micrograph of a fractured surface presented in Fig. 2 shows that cracks do not pass through un-reacted APC glass particles but are deflected around them. The geopolymer–glass composite structure results in a high strength material with increased fracture surface energy because of toughening involving crack deflection, typical of particle reinforced ceramics [38]. Data presented in Table 2 show that Young’s modulus is higher in glassy and ceramic materials [5,36] and lower in concrete and geopolymer cements [8,36]. The Young’s modulus of APC glass geopolymer is close to the values reported for concrete [36]. 3.3. Porosity and bulk density by mercury porosimetry APC glass geopolymers have comparable density to concrete and higher density than geopolymers cements, 2.3 g/cm3 , 2.4 g/cm3 and 1–1.9 g/cm3 , respectively [8,36]. They also have lower open porosity (5.5%) compared to concrete (15%) [36] and geopolymer cements (15–30%) [8].
Fig. 2. Fracture surface of APC glass geopolymer sample after 3 point bending strength test.
3.4. Thermal conductivity The thermal conductivity of APC glass geopolymer samples was ∼0.9 W/K m which is higher than the value reported for geopolymer cements of 0.2–0.4 W/K m [8]. The thermal conductivity of APC glass geopolymer is similar to concrete which ranges between 0.8 and 1.3 W/K m [36]. 3.5. Chemical resistance Chemical resistance of APC glass geopolymer was evaluated according to BS EN 1344: 2002 using sulphuric acid and nitric acid. The weight loss of APC glass geopolymer samples was <0.1%, showing that the material has good acid resistance. The value reported for geopolymer cements is similar but is based on a different test method [8]. APC glass geopolymer was boiled in a solution prepared with 75 ml of 10% sulphuric acid and 25 ml of 10% nitric acid while in the test used for geopolymer cements the material was immersed in a 5% sulphuric acid solution [8]. Nevertheless, both materials exhibit very good resistance to acid attack. 3.5.1. Freeze/thaw resistance Freeze/thaw tests were performed on APC glass geopolymer according to ASTM C 1262-05. Samples were subjected to 92 freeze/thaw cycles. The water was changed every ten days and the sample residue in the water was also collected. Residue weight was measured at every collection interval and the weight loss for each sample is presented in Fig. 3. Winitial is the initial weight of the samples which according to ASTM C 1262-05 is calculated using the following equation: Winitial = Wfinal + Wresidue
(2)
After 92 freeze/thaw cycles an average weight loss of 2% was determined. Significant weight loss occurs during the first 60 cycles after which no further weight loss was observed. The highest weight loss was observed in the first 30 cycles and the amount of residue decreased with increasing cycles. Peeling was observed on the surface in contact with water. The freeze/thaw resistance reported for geopolymer cements (<0.1%) [8] is lower than the values obtained for APC glass geopolymers (2%) but a different test has been used. APC glass geopolymers were subjected to 92 freeze/thaw cycles while the standard used for geopolymer cements required just 12 cycles.
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Table 2 Properties of APC glass geopolymers, APC glass materials and other commercial products. Materials
Density (g/cm3 )
Open porosity (%)
Water absorption (%)
Compressive strength (28 days) (MPa)
Flexural strength (MPa)
Young’s modulus (GPa)
Chemical resistance (%)
Freeze/thaw mass loss (%)
Thermal conductivity W/K m
APC glass geopolymer Concrete [36] Bricks [36] Tilesa [5] Glass [36] Geopolymer cements [8]
2.3 2.4 2 2.4b 2.5 1–1.9
5.5 15 – – 0 15–30
11 – – 2–14.5 – –
110 30 100 – 320 >90
10.5 2 10 1.6–53 40 10–15
32 20 50 38 70 >2
<0.1 – 2 – – <0.1
2 – – – – <0.1
0.96 0.8–1.28 – – – 0.2–0.4
a b
Values refer to a range of tiles including floor tiles, wall tiles, vitrified tiles and porcelain tiles. Porcelain tiles.
Table 3 Results of the leaching test from optimum APC glass geopolymer. Element
mg/kg
Element
mg/kg
Element
mg/kg
Al As Ba Ca Cd Cr Cu
13.28
Fe K Mg Mn Na Ni P
Pb Si Ti Zn Zr
Elements in italics are included in Waste Acceptance Criteria. WAC limits: As, 0.5 mg/kg; Ba, 20 mg/kg; Cd, 0.04 mg/kg; Cr, 0.5 mg/kg; Cu, 2 mg/kg; Ni, 0.4 mg/kg; Pb, 0.5 mg/kg, Zn, 4 mg/kg.
3.5.2. Leach testing APC glass which was the starting material for the production of APC glass geopolymers is inert, as previously reported [2]. The results of leaching tests performed on APC glass geopolymers are presented in Table 3. There is negligible heavy metal leaching from APC glass geopolymers. Cr has very limited leaching (0.18 mg/kg), but all values were well below the WAC for inert waste landfill. Significant leaching of Na (1067 mg/kg) occurred, as this was the main component in the activating solution used to prepare the APC glass geopolymers. 3.6. Characterization of the binder phase APC glass geopolymers are not pure geopolymers but a composite material consisting of un-reacted APC glass particles and a binder phase [1]. Data indicate that the binder phase contains geopolymer and amorphous C–S–H gel. The amorphous nature of these phases makes it impossible to identify the presence of
amorphous hydrations products. Salicylic acid and methanol (SAM method) dissolves free lime and C–S–H that exist in the structure, but it does not affect geopolymer network and un-reacted APC glass particles. ICP analysis of the initial geopolymer and the solid residue together with XRD and FTIR analysis are necessary to identify different components in the material. The weight change between the initial sample and the solid residue after the SAM method indicates the percentage of material in the form of calcium containing products such as lime or C–S–H gel. The APC glass geopolymer was tested in duplicate and weight change data are presented in Table 4. The results show that ∼60% of the initial APC glass geopolymer was insoluble and can be attributed to un-reacted APC glass particles and the geopolymer network. The remaining ∼40% passed in the liquid phase is attributed to soluble calcium-containing products such as free lime and C–S–H gel. The APC glass geopolymer and the SAM solid residue were analysed using ICP-AES for Si, Al and Ca content, and the results are presented in Table 5. The Ca percentage in the insoluble residue is lower than the Ca percentage in the initial geopolymer. During the selective chemical attack, phases containing soluble Ca pass into the liquid phase causing the observed decrease in Ca percentage in the solid residue. Moreover, the percentage of Si increased, as the insoluble material is a mixture of un-reacted APC glass particles and geopolymer. ICP-AES results also confirmed the design Si/Al ratio of 2.6 that was used in the preparation of the geopolymers. Based on the composition and quantity of APC glass geopolymer and solid residue from SAM methods, the quantities of Si, Al and Ca were calculated for the initial APC glass geopolymer, the insoluble residue and the soluble material. The results are presented in Table 6. For the soluble material the amounts presented were
Table 4 Weight change of geopolymer samples subjected to salicylic acid/methanol (SAM) selective extraction method. Sample 1 Initial sample weight (g) Insoluble solid residue weight (g) Soluble material in the liquid phase (g) Insoluble material (%) Material in the form of soluble Ca containing products (%)
Sample 2
4.96 2.86 2.10
5.06 3.09 1.97
57.7 42.3
61.1 38.9
Table 5 ICP analysis results for optimum APC glass geopolymer and insoluble solid residue of SAM method.
Fig. 3. Weight loss of three APC glass geopolymer samples subjected to freeze/thaw test.
APC glass geopolymer Insoluble solid residue
Al (wt%)
Ca (wt%)
Si (wt%)
Si/Al
Ca/Si
7.69 6.85
23.2 19.8
19.7 20.2
2.6 2.9
1.2 1.0
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Table 6 Al, Si and Ca quantities in APC glass geopolymer samples subjected to salicylic acid/methanol (SAM) selective extraction method. Samples
Weight (g)
Al (g)
Ca (g)
Si (g)
Al (%)
Ca (%)
Si (%)
Sample 1 APC geopolymer Insoluble residue Soluble material
4.965 2.864 2.101
0.382 0.196 0.186
1.152 0.567 0.585
0.979 0.578 0.401
100 51.3 48.7
100 49.2 50.8
100 59 41
Sample 2 APC geopolymer Insoluble residue Soluble material
5.056 3.088 1.968
0.389 0.212 0.177
1.173 0.611 0.562
0.997 0.623 0.374
100 54.5 45.5
100 52.1 47.9
100 62.5 37.5
calculated from the initial amount on APC glass geopolymer by subtracting the amounts present in the insoluble solid residue: XAPC geo = XInsoluble residue + XLiquid phase
(3)
where X is Si, Al or Ca. Furthermore the percentages of each compound that remained in the solid residue or passed in the liquid phase are included. It can observed from Table 6 that ∼50% of initial Ca has passed into the liquid phase and was therefore bound to soluble compounds such as free lime and C–S–H, although a significant amount remained in the insoluble residue. Part of the Ca in the solid residue can be attributed to the Ca in the un-reacted APC glass particles and the remainder to the non-hydrated alumino–silicate network which also contains some Ca. The solid residue from the SAM test was examined by XRD analysis to identify changes in mineralogy due to the selective chemical extraction. Fig. 4 presents XRD data of the initial APC glass geopolymer and the SAM solid residue. It can be seen that both materials are amorphous, with a broad hump around 30◦ 2 present in both spectra. Fig. 5 presents the results of FTIR analysis of the initial APC glass geopolymer and the SAM solid residue. It also presents the FTIR spectrum of the soluble material generated by subtraction of the spectrum of the insoluble residue from the spectrum of the initial APC glass geopolymer, using computer software. The main peak of the insoluble residue, after the selective chemical attack, shifted to higher frequencies compared to the original APC glass geopolymer. The main band of insoluble residue appears at around 1018 cm−1 , while for APC glass geopolymer it appears at 976 cm−1 . This indicates a higher degree of geopolymerisation in the insoluble residue. The main band of soluble material appears at 968 cm−1 , and this is very similar to the main band of C–S–H gel at 970 cm−1 . 4. Discussion APC glass geopolymers have compressive strengths of 110 MPa after 28 days curing, high density (2.3 g cm−3 ) and low porosity (5.5%). These properties are comparable to the properties of commercial sintered tiles, high strength concrete and geopolymer cements. The APC glass geopolymer samples subjected to flexural strength test developed cracks at the surface but did not fracture into separate pieces. Micro-structural analysis indicates that APC geopolymers have a heterogeneous microstructure containing un-reacted APC glass particles surrounded by a binder phase. The results suggest a toughening mechanism based on crack deflection [39] as previously proposed [13]. Cracks are forced to change direction when they encounter APC glass particles, and this crack deflection mechanism is similar to that reported in particle reinforced ceramic matrix composites [38,40]. The particle reinforcement associated with the un-reacted APC glass particles is considered to be particularly effective because the bond between the APC glass and the complex geopolymer matrix is expected to
be very strong as they both originate from the same material. Thus an effective load transfer mechanism occurs at the interface. The thermal conductivity of APC glass geopolymers is similar to the thermal conductivity of concrete, indicating that APC glass geopolymers are not suitable for refractory applications. Construction materials such as concrete, tiles and geopolymers must be extremely durable. Initial evaluation of the durability of APC glass geopolymer involved testing the chemical resistance and freeze and thaw resistance of samples. The acid resistance test showed that APC glass geopolymers have very good resistance to acid attack as the observed change in weight was <0.1%. This is in accordance with previous research which showed that geopolymers have very good acid resistance [8,41,42], even when Ca is added to the system [25]. It has been reported that geopolymer structures with Si/Al ratio of 1 are more susceptible to acid attack than geopolymers with higher silicon content [43]. APC glass geopolymers have a Si/Al ratio of 2.6 which explains the good acid resistance. Freeze and thaw resistance of the material was evaluated by subjecting APC glass geopolymer samples to 92 freeze and thaw cycles. A 2% weight change and some peeling of the surface in contact with water were observed. Very good freeze and thaw resistance has been reported for geopolymer cements with a weight change <0.1% [8]. The test used required only 12 cycles while the APC glass geopolymers in this study were subjected to 92 cycles, and this makes comparison of the results difficult. The leaching test performed for APC glass geopolymers revealed that APC glass geopolymers are inert and do not leach significant levels of heavy metals. Only low concentrations of Si, Al and Ca were detected, which are the main elements in the material. Significantly higher amounts of Na were observed, which originates from the activating solution. Release of heavy metals and other elements from the original APC glass was not observed as they are effectively encapsulated in the geopolymer micro-structure. This was expected as geopolymers have been used in hazardous waste management to encapsulate heavy metals [44–46] and radioactive wastes [47–50]. The results of the selective chemical attack with salicylic acid and methanol provide valuable information on the nature of APC glass geopolymers. The change in weight of APC glass geopolymer, after SAM method, revealed the presence of C–S–H gel in the APC glass geopolymer. Around 60% of the initial material was insoluble, while the soluble material is associated with calcium containing phases such as lime and C–S–H gel [17,24,26,27]. The results showed that only 50% of the Ca in the initial material was soluble which means that a significant amount remained in the insoluble residue. Part of the Ca in the insoluble residue is due to the Ca in the un-reacted APC glass, while the remainder can be attributed to non-hydrated alumino–silicate network containing Ca. This possibility of formation of silica-rich gel containing Ca has been reported previously in the literature [27,51,52]. The FTIR results confirm this possibility, as the main peak for the insoluble residue is at 1018 cm−1 , where the main peak occurs for
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Fig. 4. XRD analysis of APC glass geopolymer and solid residue after SAM method.
Fig. 5. FTIR analysis of initial APC glass geopolymer, insoluble residue after SAM method and soluble material.
geopolymers [53–55], while the main peak for the soluble material was at 968 cm−1 , where the main peak for C–S–H usually occurs [27,28,51,52,55,56]. The results of the SAM method suggest that the APC glass geopolymer contains both an alumino–silicate network containing calcium and C–S–H gel. The composition of the two phases seems to be quite similar and for that reason they cannot be identified using conventional methods such as XRD or SEM/EDS analysis.
5. Conclusions APC glass geopolymer is not a pure geopolymer, but a composite consisting of a binder phase and un-reacted APC glass particles which act as reinforcement. The binder phase is a three dimensional geopolymeric network that contains C–S–H gel and probably Al modified C–S–H gel. The excellent mechanical properties of APC glass geopolymers can be attributed to these micro-structural characteristics. Geopolymers prepared from the glass derived from DC plasma treatment of APC residues have potential to be used in various pre-cast products such as paving blocks and tiles.
Acknowledgements This work was completed as part of the project ‘Integrated solution for air pollution control residues (APC) using DC plasma technology’ funded by the UK Technology Strategy Board and Defra, through the Business Resource Efficiency and Waste (BREW) programme. The Technology Strategy Board is a business-led executive non-departmental public body, established by the government. Its mission is to promote and support research into, and development and exploitation of, technology and innovation for the benefit of UK business, in order to increase economic growth and improve the quality of life. It is sponsored by the Department for Innovation, Universities and Skills (DIUS). Visit www.innovateuk.org for further information. References [1] I. Kourti, D.A. Rani, D. Deegan, A.R. Boccaccini, C.R. Cheeseman, Production of geopolymers using glass produced from DC plasma treatment of air pollution control (APC) residues, J. Hazard. Mater. 176 (2010) 704–709. [2] D. Amutha Rani, E. Gomez, A.R. Boccaccini, L. Hao, D. Deegan, C.R. Cheeseman, Plasma treatment of air pollution control residues, Waste Manage. 28 (2008) 1254–1262.
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Journal of Hazardous Materials 196 (2011) 93–100
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Feasibility study on the use of soil washing to remediate the As–Hg contamination at an ancient mining and metallurgy area C. Sierra a , J.M. Menéndez-Aguado a , E. Afif b , M. Carrero c , J.R. Gallego a,∗ a b c
Environmental Biotechnology and Geochemistry Group, University of Oviedo, C/Gonzalo Gut, S/N, 33600 Mieres, Asturias, Spain Dpto. de Biología de Organismos y Sistemas, Área de Ingeniería Agroforestal, University of Oviedo, C/Gonzalo Gut, S/N, 33600 Mieres, Asturias, Spain R&D Projects in Civil Construction, R&D&I Directorate, DRAGADOS S.A. Avenida Camino de Santiago 50, 28050 Madrid, Spain
a r t i c l e
i n f o
Article history: Received 14 March 2011 Received in revised form 15 July 2011 Accepted 31 August 2011 Available online 6 September 2011 Keywords: Soil pollution Mercury Arsenic Ore processing Soil washing
a b s t r a c t Soils in abandoned mining sites generally present high concentrations of trace elements, such as As and Hg. Here we assessed the feasibility of washing procedures to physically separate these toxic elements ˜ from soils affected by a considerable amount of mining and metallurgical waste (“La Soterrana”, Asturias, NW Spain). After exhaustive soil sampling and subsequent particle-size separation via wet sieving, chemical and mineralogical analysis revealed that the finer fractions held very high concentrations of As (up to 32,500 ppm) and Hg (up to 1600 ppm). These elements were both associated mainly with Fe/Mn oxides and hydroxides. Textural and geochemical data were correlated with the geological substrate by means of a multivariate statistical analysis. In addition, the Hg liberation size (below 200 m) was determined to be main factor conditioning the selection of suitable soil washing strategies. These studies were finally complemented with a specific-gravity study performed with a C800 Mozley separator together with a grindability test, both novel approaches in soil washing feasibility studies. The results highlighted the ˜ soils. These difficulties are attributed to the presence of contamidifficulties in treating “La Soterrana” nants embedded in the soil and spoil heap aggregates, caused by the meteorization of gangue and ore minerals. As a result of these two characteristics, high concentrations of the contaminants accumulate in all grain-size fractions. Therefore, the soil washing approach proposed here includes the grinding of particles above 125 m. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Abandoned mining-metallurgy sites are sources of environmental pollution. This contamination is a result of mine drainage, waste disposal, subsidence and other phenomena [1,2]. Specifically, Hg mining and processing are frequent causes of environmental concern because of the abundance of Hg and other toxic trace elements, such as As, in the ores exploited [3]. Hg participates in a number of complex environmental cycles. Geochemical studies have shown that, once in the environment, ionic Hg can be converted into organomercury compounds, which are highly toxic to most organisms [4]. Furthermore, As toxicity – specifically As (III) – has triggered severe environmental alarms throughout the world, in particular in relation to groundwater [5]. In this context, soil washing by means of physical and/or chemical procedures is a suitable technique to reduce the concentration of heavy metal contaminants in this matrix [6]. Physical processing technology in particular has been frequently used to remediate
∗ Corresponding author. Tel.: +34 985458064/651043415; fax: +34 985458182. E-mail address:
[email protected] (J.R. Gallego). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.080
heavy metal and semi-metal pollution, including Hg and As [7–11]. These techniques remove contaminants from soil by concentrating them into a minor volume by means of comminution, particle size separation, specific-gravity separation, attrition scrubbing, froth flotation or magnetic separators. Thus, physical processing concentrates contaminants by exploiting differences between the characteristics of metal-bearing particles and soil particles (size, density, hydrophobic surface properties, magnetism), in much the same way as mineral ores can be treated. Given that these technologies are versatile and cost-effective when high amounts of soil are to be treated, they may be highly appropriate for the remediation of former industrial sites and old mine dumps [12]. In contrast, chemical processing usually comprises procedures such as acid or base treatment for solubilisation, or the use of specific solvents, to release pollutants into the liquid fraction [6]. The information reported in feasibility studies usually allows consideration of a range of alternatives for soil washing treatments, the final objective of which is a noteworthy reduction in the volume of contaminated soil (ideally, in a soil washing process the goal pursued for a given contaminant is to achieve a high concentration in a small fraction [6]). This broad view requires a detailed characterisation of the edaphology, mineralogy and geochemical
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behaviour of the grain-size soil fractions. Here we applied this work-plan to soil highly contaminated with As and Hg as a result of the physico-chemical alteration of mining-metallurgic waste [13]. In addition, specific-gravity and liberation degree studies facilitated the evaluation of the viability of applying gravimetric and/or granulometric concentrators. Concretely, the particle-size fractions obtained were used for a specific-gravity study to examine the relationship between particle size, density and contaminant content of the fractions. Milling was also considered by means of a complementary approach consisting of Bond tests to evaluate the grindability of coarse fractions in order to obtain a finer grain-size that is more appropriate for physical separation. Following all of the preceding considerations, the main aims of the current study were as follows: • To integrate grain size distribution, specific-gravity separation and a liberation degree study with edaphological, geochemical and mineralogical information of the study site. • To introduce the “C800 Mozley laboratory mineral separator” and the “Bond Ball Mill Standard test” as effective tools to improve the abovementioned feasibility studies. 2. Materials and methods 2.1. Site description and soil sampling Until the end of the 1970, extensive Hg deposits in the central zone of Asturias (northern Spain) were exploited [14–18]. One of ˜ the main sites was known as “La Soterrana”. There, together with mining activity, ore processing and metallurgy were carried out intermittently from the middle of the XIX century in order to obtain Hg. Regarding ore geology, mining was performed through a lowtemperature hydrothermal ore, which is hosted by highly fractured limestones with dispersion in the flanking sandstones and lutites of Carboniferous age. The paragenesis of this mineral deposit is constituted by cinnabar (HgS), orpiment (As2 S3 ), realgar and pararealgar (AsS), As-enriched pyrite and marcasite (FeS2 ), and arsenopyrite (FeAsS), in a gangue of quartz and calcite [14,19]. These sulphides ˜ mining-metallurgic plant, where the were treated at “La Soterrana” mineral was milled and roasted to obtain Hg vapour, which was then condensed. The emissions of polluting steams and particles and the dumping of mining and smelting waste greatly affected around 80,000 m2 of the surrounding area [14]. Currently, the distribution of the pollutants throughout the site is caused mainly by the mechanical dispersion of the spoil heap waste, together with the oxidation and lixiviation of As–Hg-rich materials, and also the processes of complexation and immobilisation related to soil particles. In order to conduct a soil washing feasibility study, three composite and representative soil samples (labelled S1, S2 and S3, 50 kg each) were collected from the tilled depth (0–35 cm) by means of a stainless steel hand-auger and a shovel. The soil was passed through a 2-cm mesh screen in situ to remove rock fragments, vegetation and other large material; finally, samples were homogenised and stored in inert plastic bags. 2.2. Sample preparation In the laboratory, the three samples were gently dried at room temperature, thoroughly disaggregated, mixed, and subsequently sieved through a 4-mm screen. Materials with a grain size greater than 4 mm were washed and rubbed off to recover fine particles adhered to gravels and pebbles. Once these gravels and pebbles had been cleaned up, they were excluded from the study. Each sample below a grain size of 4 mm was quartered by means of a channel
separator to obtain representative 4-kg batches. These were then oven-dried for 48 h at 45 ◦ C to prevent Hg loss. 2.3. Soil characterisation Regarding pedology, the pH was measured with a glass electrode in a suspension of soil and water (1: 2.5) in H2 O and electrical conductivity was measured in the same extract (diluted 1:5). Organic matter was determined by the ignition method (weight loss at 450 ◦ C). Exchangeable cations (Ca, Mg, K and Na) extracted with 1 M NH4 Cl, and exchangeable Al extracted with 1 M KCl were determined by atomic absorption/emission spectrophotometry [20] in an AA200 Perkin Elmer analyzer; the effective cation exchange capacity (ECEC) was calculated as the sum of the values of the latter two measurements (sum of exchangeable cations and exchangeable Al). 2.4. Wet sieving The representative batches of each sample (S1, S2, S3) were slurried in water and then sieved (cycles of 100 g) into particle-size fractions of <63, 63–125, 125–250, 250–500, 500–1000, 1000–2000 and 2000–4000 m batches were passed through normalised sieves positioned in a shaker (Restch) for 5 min with a water flow of 0.3 l/min (ASTM D-422-63, Standard Test Method for Particle-Size Analysis of Soils). The fractions were recovered with the help of a spray nozzle, and then dried at 50 ◦ C and weighed. To complete the grain size distribution, the silt–clay fraction (<63 m) was studied using a Laser Diffraction Particle Analyser (Beckman Coulter Inc.). Representative samples of the grain size fractions were subjected to chemical analyses by means of ICP-OES (Section 2.5). However, some of these fractions were subdivided to obtain further samples for the mineralogical and specific gravity studies (see Sections 2.7 and 2.8). In order to standardise the conditions used for chemical attack, samples with a grain size over 125 m were ground at 400 rpm for 40 s using a vibratory disc mill (RS 100 Retsch). 2.5. Chemical analyses For chemical analyses, 1-g representative sub-samples of the diverse origins (soils, grain size fractions, light or dense specific gravity fractions, etc.) were leached by means of an ‘Aqua regia’ digestion (HCl + HNO3 ). The digested material was analysed for total concentrations of 19 major and trace elements (Ca, Mg, K, Na, Al, Fe, S, Cu, Pb, Zn, Cd, Ni, Mn, As, Sr, Sb, La, Cr and Hg) by Inductively Coupled Plasma-Optical Emission Spectroscopy (ICPOES) in the accredited (ISO 9002) laboratories Actlabs Int., Ancaster (Ontario, Canada). 2.6. Multivariate statistics Cluster Analysis was undertaken following the Wardalgorithmic method, which maximises the variance between groups and minimises it between members of the same group [21]. A dendrogram obtained with the statistical software SPSS v18.0 was used to show the clustering of results. Groups of elements with a similar geochemical behaviour were identified on the basis of the statistical distance between them (squared-Euclidean distance was selected). 2.7. Mineralogy and liberation degree study The mineralogical composition of the soil was estimated by means of an X-ray diffractometer (DRX, Philips X Pert Pro, incorporating databases of the International Centre for Diffraction Data). In
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Table 1 Parameters used in the specific gravity study performed with a C800 Mozley laboratory separator (recommendations by the manufacturer were adapted to soil properties). Grain-size fractions (m)
Tray
Shake speed (r.p.m.)
Amplitude (mm)
Washwater (l/min)
Feed (g)
Time (min)
<63
‘V’profile
70
2.5
3
50
3
125–63 250–125 500–250
Flat
90
3.5
3
150
3
addition, representative samples of each wet-sieving fraction were used to prepare polished sections to be studied by an Eclipse LV 100 POL Nikon petrographic microscope. The morphology and composition of specific minerals were studied using a SEM-EDX system: Scanning Electron Microscope (Jeol JSM-6100) coupled with Energy Dispersive X-ray analyser (INCA Energy 200). 2.8. Specific gravity study The particle-size fractions obtained were used for a specificgravity study to examine the relationship between particle size, density and contaminant content of the fractions. A C800 Mozley laboratory mineral separator was used for this purpose. This separator, which operates using gravimetric classification, is commonly used to assess mineral processing equipment [22], although in our case it was tested for site remediation purposes. In brief (see [22] for details), this separator comprises a riffleless shaking table; two types of table deck (trays) are available, a “V” profile for materials finer than 1000 m and coarser than 63 m, and a “Flat” profile for the separation of material below 63 m. Physical separation is governed by the flowing film principle [23], in addition to a perpendicular movement to the tray axis, thus favouring the advance of the solids on the tray. Therefore, the separation in this equipment is only partially conditioned by Stokes force (correlated with grain-size); conversely, mass forces related to an asymmetric acceleration are enhanced (specific-gravity separation effect). In our case, 100-g samples of the particle-size fractions of <63, 63–125, 125–250, 250–500 m were shaken under several controlled parameters (shaking speed and amplitude, water flow, and time, Table 1), in order to obtain dense and light fractions in every experiment. However, for fractions above 63 m, the effect of silt/clay particles physically adhered to the coarser ones should be considered; in our case, this adherence was directly linked to the efficiency of the previous wet-sieving performed. To determine the relevance of this effect, we carried out an experiment in which three representative 50-g samples from the 63 to 125 m fraction were directly treated in the C800 separator, while another three samples were pre-treated for 30 min in a solution of dispersing agents (3 g of sodium hexametaphosphate and 0.5 g of anhydrous sodium carbonate dissolved in 250 ml of distilled water) at 400 rpm in a Heidolph RZR 2020 shaker. 2.9. Grindability characterisation In soil washing approaches, milling could be required to free contaminants from the matrix aggregates in which they are embed-
ded, in order to facilitate the ulterior operation in concentrators. Therefore, we estimated the resistance of the soil samples to ball milling in terms of specific power consumption for grinding. This was done by means of the Bond Ball Mill Standard test [24–26]. In brief, with the aim to simulate closed circuit continuous operation with a recirculating load of 250%, the test [27] is carried out in a 12 × 12 lab ball mill, in consecutive cycles of batch operations of grinding and sieving. In our case, the sieves selected to study the variation of the Bond Work index with the milling product size were 250, 180, 125 and 80 m. The test is useful to obtain the internal parameter named “grindability index”, Gbp, in an iterative procedure. The final value of “work index” is calculated using the following equation: wi =
44.5 Pi0.23
0.82
· Gbp
√10 − √10 P80
F80
where Pi (in m): screen size at which the test is performed. Gbp is the Bond’s standard ball mill grindability; net weight of ball mill product passing sieve size Pi produced per mill revolution (g/rev), once the end of the test is reached. F80 and P80 (in m) are the 80% sieve opening through which 80% of the product passes (for Feed and Product respectively). wi is the Bond Work index (in kWh/sht). Once obtained the work index, the specific power consumption W [kWh/t] can be calculated using the Bond Formula:
W = 10 · wi
1
P80
−
1
F80
3. Results and discussion 3.1. Pedology, grain size study and soil geochemistry The pedological characterisation (average of three determinations over the initial bulk samples S1, S2, S3, with a standard error below 5%) revealed a slightly acid pH (6.3), low electrical conductivity (EC = 0.19 dS m−1 ), low content of exchangeable base cations (5.42; 0.31; 0.28 and 0.47 cmolc kg−1 for Ca, Mg, K and Na respectively), low ECEC (sum of exchangeable cations and exchangeable Al = 7.38 cmolc kg−1 ), and a low organic matter content in the upper horizon (0.75%). All of these data are consistent with the geological origin of the soil and its development under particular geochemical conditions as a result of the influence of waste derived from mining activities and the metallurgy industry. Furthermore, ICP-OES analyses (Table 2) revealed very high concentrations of As and Hg, and a lower presence of other contaminants such as Pb. These findings are also consistent with the
Table 2 Element concentration of representative subsamples of the three initial bulk samples (results correspond to the average of three determinations with a standard error <5%). Sample
Trace elements (mg kg−1 ) Hg
S1 S2 S3
805 1600 132
Major elements (%)
As
Pb
Zn
Sb
Al
Ca
Fe
32,500 17,100 6350
98 49 28
98 71 46
211 111 23
2.86 2.08 2.8
2.52 5.6 5.9
3.19 3.09 2.81
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Cumulave passing
100% 80% 60% 40% 20% 0%
1
10
100
1000
10000
Size (microns) Fig. 1. Particle-size distribution of sample S2, obtained by compositing wet-sieving (4000–63 m), and laser dispersion (<63 m) data.
˜ mineralisation-type of the ores treated at “La Soterrana”. In contrast, the high concentrations of Ca, Al and Fe point to a soil matrix composed of carbonates, clay minerals and iron oxides. Regarding the grain-size analysis, the results for the three samples were very similar. For instance, the cumulative passing curve of sample S2 (Fig. 1) indicates that the coarsest (500–4000 m: 39.9%) and finest fractions (<63 m: 38.1%) were predominant, whereas intermediate fractions (250–500 m: 1.8%; 125–250 m: 15.7%; 63–125 m: 4,5%) were less abundant. These data are particularly relevant given that the fineness of the material is one of the main factors impeding acceptable performance of soil washing for remediation purposes [7]. To facilitate the study of the relationship between contaminant contents and grain-size fractions, we then measured the total content of major and trace elements in the abovementioned grain-size fractions. Hg and other trace elements showed higher contents in the fine fraction, while As showed a slightly more homogeneous distribution in all the grain-size fractions (Table 3). Of the major elements, only Fe showed a similar pattern to that observed for the previously mentioned trace elements, especially Hg. Therefore we hypothesised that most of the trace elements remain geochemically associated with former iron-rich sulphur minerals, which are probably oxidised in the present soil conditions. In contrast, a different profile was observed for Al and Ca, thereby suggesting a distinct behaviour of these elements to that of Hg and As. This notion was confirmed by a multivariate approach in which a hierarchical cluster analysis and also a Principal Components Analysis (data not shown) revealed two main groups of elements, as shown in the dendrogram (Fig. 2). This outcome is congruent and complementary to the mineralogical and edaphological studies, thus suggesting the following about the geochemical behaviour of As and Hg: Group AB: comprising mainly chalcophile trace elements and some major elements such as Fe, Al and K. This group is subdivided into two subgroups: ‘A’ contains a clear association between Fe, Mn, Hg, As and other trace elements of concern (Sb, Ni). Consequently, and given the high contents of ultrafine materials in the soil and the low amount
Fig. 2. Dendrogram showing the clustering of elements associated by their geochemical affinity within the samples (irrespective of the original sample and grain-size fraction). Main groups are indicated based on statistical distance between them.
of organic matter, Hg and As behaviour appears to be controlled mostly by their ion binding to iron/manganese oxides-hydroxides. The presence of Al and K in this subgroup suggests a main role of argillaceous minerals; however, the DRX data (see below) and the low ECEC suggest that phyllosilicates have little relevance. ‘B’ includes the well-known Pb–Zn–Cd association, which probably originated from the weathering of sphalerite (ZnS, slightly enriched in Cd) and galena (PbS), both accessory minerals in La ˜ ore. Soterrana Group C: comprising elements linked to the alteration of the gangue rocks (for instance Ca from limestone). In addition, the presence of S in this group is linked to secondary minerals, such as gypsum (CaSO4 ·2H2 O), and others that can be observed in the local paragenesis of the weathered ores. According to the statistical treatment, the elements included in C are distant from the Group AB (negative correlation). 3.2. Mineralogy and liberation degree study DRX data showed that the predominant components in the samples were quartz and secondarily, calcite, hematite, ferrihydrite, goethite and maghemite. In contrast, clay minerals were clearly minority. Mineralogical analyses, particularly microscopy observations, revealed that some of the former sulphides from the ore were undamaged in the coarse fractions, and were usually covered by gangue minerals, which shielded them from the effects of weathering (Fig. 3a and b). Although scarce, free grains of arsenopyrite, (Fig. 3c), pyrite and chalcopyrite were also found. Cinnabar was the
Table 3 Total content in grain-size fractions for major and trace elements of sample S2 (results correspond to the average of three determinations with a standard error <5%). Grain-size fraction (m)
4000–2000 2000–1000 1000–500 500–250 250–125 125–63 <63
Trace elements (mg kg−1 )
Major elements (%)
Hg
As
Pb
Zn
Sb
Al
Ca
Fe
495 720 1000 1920 2030 2520 4810
15,800 18,400 21,900 24,500 27,900 24,300 26,350
36 44 54 65 67 68 76
60 70 80 125 125 133 155
87 101 144 165 195 158 212
2.9 3.25 3.1 2.47 2.95 3.29 3.31
3.31 3.42 3.71 3.62 4.51 3.77 3.42
3.87 3.74 3.75 3.80 3.84 3.97 4.39
C. Sierra et al. / Journal of Hazardous Materials 196 (2011) 93–100
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Fig. 3. (a) Arsenopyrite crystals within a quartz grain. (b) Pyrite inclusion in a quartz grain. (c) Unaltered free idiomorphic grain of arsenopyrite. (d) Quartz grain partially reemplaced by cinnabar (soft component). (e) Free altered cinnabar. (f) Hematite coating surrounding quartz grains. (Horizontal frame is 650 m for Fig. 3a, 260 m for b, c and e and 1.3 mm for d and f.)
most common suphide and also appeared as inclusions in gangue minerals (Fig. 3d) and as free altered small grains (Fig. 3e and g). The texture is typical of a calcine material, the fractions below 250 m being richer in metallic oxides and oxy-hydroxides, especially hematite (usually combined with gangue materials, Fig. 3f) and, to a lesser extent, goethite. These textures and morphologies can be attributed to the roasting process during ore treatment and also weathering. However, the most aggressive oxidising process was the former as it has the capacity to destroy As and Hg sulphides.
Nevertheless, the temperature range used in the ovens (much lower than 700 ◦ C) was not enough to destroy As-rich pyrite, arsenopyrite and chalcopyrite, irrespective of their original grain size (Fig. 4). In this regard, the presence and texture of cinnabar, and other mineral phases containing Hg denotes a deficient roasting procedure. Therefore, the mineralogy of Hg (mainly as cinnabar) and As (mainly linked to Fe sulphides and oxides), which both show an irregular distribution within the grain-size fractions, hinders the design of an effective soil washing treatment. In fact, in the case
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C. Sierra et al. / Journal of Hazardous Materials 196 (2011) 93–100
Table 4 Major and trace elements concentrations measured in the dense fractions after experiments carried out with a C800 separator (125–63 m fraction). Results are an average of those obtained with three samples and show a standard error <5%. Trace elements (mg kg−1 )
Samples
Feed material Dense fraction with dispersant pretreatment Dense fraction without pretreatment % variation
Major elements (%)
Hg
As
Pb
Zn
Sb
Al
Ca
Fe
785 925 897 2.99
29,433 17,100 15,733 7.99
96 70 76 9.05
152 176 165 6.06
180 90 82 8.18
2.38 1.56 1.47 5.98
3.19 2.31 2.26 2.16
4.17 3.76 3.81 −1.51
Table 5 Element content (percentage by weight) measured in the dense fraction for the indicated grain-size fractions. The separation between dense and light fractions was achieved in a C800 separator working as specified in Table 1. Results are an average of those obtained with three samples and show a standard error <5%.
Fig. 4. Scanning electron micrograph of a hexagonal goethite grain that preserves the original morphology of pyrite. The surrounding area contains 22% of Hg. Smaller and more reflecting grains correspond to arsenopyrite.
of cinnabar (HgS), the liberation size calculated under the optical microscope (grain size for which the cinnabar grains are not mineral inclusions in other mineral phases) was below 200 m, thus clearly limiting any physical separation equipment for fractions above this grain size. To obtain these data, we performed a complex liberation size study composed of dozens of microphotographs with the accompanying measurements of grain sizes.
Grain-size fraction (m)
Textural classification
500–250 250–125 125–63 x < 63
Medium sands Fine sands Silt–clay
% in dense fraction
Hg
As
Fe
50.06 39.33 19.73 15.65
42.17 33.42 14.09 6.82
65.47 55.04 29.72 14.06
was poorer in As and Hg than the light fraction (Table 5). In this context, the differences in density between the two fractions were quite low (e.g., 2.5 g/cm3 for the lighter particles and 2.8 g/cm3 for the heavier ones in the case of the 63–125 m fraction, both measured with a water pycnometer). This is an unexpected effect, given that the presence of ‘free’ As and Hg dense minerals, for instance cinnabar and arsenopyrite, arises when grain size decreases. Consequently, the concentrations in the dense fraction should be higher than those indicated in Table 5. The most probable explanation for this observation is that the size ranges used in this study were not narrow enough, and therefore a granulometric separation overlies the densimetric separation, as generally occurs with other types of gravimetric separators [23]. In practical terms, these results imply that physical separation of this soil by means of gravity concentrators (spirals, shaking tables, etc.) would be very difficult. 3.4. Grindability characterisation
3.3. Specific-gravity study The use of dispersants did not imply a clear improvement in the classification obtained in the C800 system (none of the chemical elements notably changed in the dense fraction concentration, Table 4). Thus it can be concluded that the quality of the wet-sieving was sufficient to prevent distortions in the data described below. Although the liberation degree was below 200 m for Hg, a considerable amount of contaminants might be recovered by gravity separation procedures in the grain-size interval between 500 and 125 m. Initially, for the grain-size interval between 500 and 250 m, the C800 separator was not effective as the sizes with which this equipment operates are clearly above the liberation size (Table 5). Alternatively, for all the fractions below 250 m, the results indicate that the dense fraction obtained in the C800
The results described in the previous section indicate that the use of grain size separators or gravity concentrators is strongly hindered in fractions coarser than 200 m; therefore milling is required to improve the liberation degree of Hg and As. This process, which normally accounts for 30% of the total costs in ore processing [23], can be significant enough to make the process unworkable. Thus, the cost of milling must also be taken into consideration in soil remediation procedures. In addition, milling would be an interesting option to address in future research into the immobilisation of trace elements [28]. Following the previous considerations, Bond tests were performed and the main results are presented in Table 6. The specific power consumption ranged from 16.45 kWh/t to 5.82 kWh/t, confirming that the finer the sieve, the greater the energy consumption.
Table 6 Results of Bond ball mill grindability test, corresponding to the average of three determinations with a standard error <10%. F80 and P80 are the 80% sieve opening through which 80% of the product passes (for Feed and Product respectively); Pi is the screen size at which the test is performed; Gbp is the Bond’s standard ball mill grindability, wi is the Bond index (‘work index’), and W is the specific power consumption. F80 (m)
Pi (m)
Gbp (g/rev)
P80 (m)
wi (kWh/sht)
wi (kWh/t)
W (kWh/t)
2957 2973 2917 2951
80 125 180 250
1.1032 1.8917 2.2967 2.8523
72 109 154 194
15.01 11.22 10.97 9.90
16.54 12.37 12.09 10.91
16.45 9.69 7.50 5.82
C. Sierra et al. / Journal of Hazardous Materials 196 (2011) 93–100
Depending on the liberation size, the power consumption in comminution operations can double from one size fraction to another. This issue should be considered when addressing the operating costs of the process. In our case, 125 m can be considered a safe cut-off point in accordance to the results of the mineralogical study; hence the specific power consumption to take into account is 9.69 kWh/t. This value reflects a medium-high level, similar to that required for grinding an average limestone [29]. Along with the previous statements, appropriate selection of the milling machine will yield a well balanced power efficiency process [30]. 3.5. Consequences for soil washing design The feasibility study carried out in this work should be completed by a cost analysis, highly conditioned by the huge amount of soil to be treated (80,000 m2 ) and by other important questions such as cost of milling (as described above) and treatment of washing effluents. At any case, it is possible to propose a first approach to define soil washing stages, in brief: - The grain-size fractions below 125 m containing “free” As and Hg minerals could be treated with hydrocycloning because of the demonstrated increasing concentration of the contaminants with increasing fineness. A factorial design would be required to define optimum conditions (see for example a successful application in [10]). - The design of the strategy for the fractions coarser than 125 m is clearly conditioned by the results of the mineralogical and specific-gravity studies. Therefore in order to improve the Hg and As liberation degree, a milling process should be required. The ground material would be below 125 m, thereby allowing hydrocycloning. Some other supplementary technologies might be helpful. For instance, a considerable amount of minerals such as pyrite, goethite and hematite (enriched/bound to As and Hg) could be separated prior to hydrocycloning by means of magnetic field technologies [31,32]. Furthermore, given that cinnabar is highly hydrophobic [33,34], froth flotation may be an appropriate option to obtain satisfactory Hg recovery yields [35,36]. 4. Conclusions Hg and As are primary soil pollutants, particularly in areas formerly devoted to mining activities and the metallurgy industry. In these circumstances, a proper remediation approach to reduce the volume of contamination is soil washing. This approach has been reported to be a cost-effective physical separation technique. ˜ shows soils with The feasibility study at the site of “La Soterrana” very high As and Hg content. From geochemical, edaphological and mineralogical data, we have demonstrated that Hg is concentrated mainly in fine grain fractions (below 200 m), where it is present in the original sulphide form and bound to Fe–Mn oxyhydroxides. In contrast, As is abundant in fine-medium (below 500 m) fractions and predominantly linked to Fe mineral phases as well. In the context of the feasibility study, the C800 separator proved effective to study narrow grain-size intervals whenever significant differences in specific-gravity within the particles of the soil were detected. However, in our case, the results obtained were not conclusive. In fact, the mineralogical data reflected the low liberation degree of As and Hg in sandy fractions, and thus the yields of gravimetric or grain-size separation of these fractions would be unsatisfactory. Accordingly, an interesting alternative, though possibly expensive, would consist of previous milling of the medium and coarse fractions in order to allow treatment. In this context,
99
the grindability test is a novel approach carried out to indicate the extent of power consumption and its influence on the energy efficiency and potential recovery of pollutants. This approach complemented the information from the feasibility study, and showed that in this case physical soil washing would be clearly conditioned by milling costs.
Acknowledgements This research was funded by the CDTI (‘Centro para el Desarrollo Tecnológico e Industrial’), part of the Spanish Council for Industry, within the research programme named “CLEAM CENIT”. The corporations involved in this project are the final holders of the results presented here. We thank Dr. Rodrigo Álvarez for his help with the mineralogical study. Carlos Sierra obtained a grant of the “Severo Ochoa” programme (Ficyt, Asturias, Spain).
References [1] J.E. Gray, J.G. Crock, D.L. Fey, Environmental geochemistry of abandoned mercury mines in West-Central Nevada, USA, Appl. Geochem. 17 (2002) 1069–1079. [2] J. Martínez, J. Llamas, E. de Miguel, J. Rey, M.C. Hidalgo, A.J. Sáez, Multivariate analysis of contamination in the mining district of Linares (Jaén, Spain), Appl. Geochem. 23 (2008) 2324–2336. [3] P. Higueras, R. Oyarzun, J. Lillo, J.C. Sánchez Hernández, J.A. Molina, J.M. Esbrí, S. Lorenzo, The Almadén district (Spain): anatomy of one of the world’s largest Hg-contaminated sites, Sci. Total Environ. 356 (2005) 112–124. [4] A. Davis, N.S. Bloom, S.S. Que Hee, The environmental geochemistry and bioaccessibility of mercury in soils and sediments: a review, Risk Anal. 17 (1997) 557–569. [5] A.A. Duker, E.J. Carranza, M. Hale, Arsenic geochemistry and health, Environ. Int. 31 (2005) 631–641. [6] G. Dermont, M. Bergeron, G. Mercier, M. Richer-Laflèche, Soil washing for metal removal: a review of physical/chemical technologies and field applications, J. Hazard. Mater. 152 (2008) 1–31. [7] R. Anderson, E. Rasor, Particle size separation via soil washing to obtain volume reduction, J. Hazard. Mater. 6 (1998) 89–98. [8] R. Griffiths, Soil washing technology and practice, J. Hazard. Mater. 40 (1994) 175–189. [9] M.J. Mann, Full scale and pilot scale soil washing, J. Hazard. Mater. 66 (1998) 119–136. [10] C. Sierra, J.R. Gallego, E. Afif, J.M. Menéndez-Aguado, F. González-Coto, Analysis of soil washing effectiveness to remediate a brownfield polluted with pyrite ashes, J. Hazard. Mater. 180 (2010) 602–608. [11] G. Dermont, M. Bergeron, G. Mercier, M. Richer-Laflèche, Metal-contaminated soils: remediation practices and treatment technologies, practice periodical of hazardous, toxic, an, Radioact. Waste Manage. 12 (2008) 188–209. [12] M. Pearl, M. Pruijn, J. Bovendeur, The application of soil washing to the remediation of contaminated soils, Land Contam. Reclam. 14 (2006) 713–726. ˜ [13] J. Loredo, A. Ordónez, R. Álvarez, Environmental impact of toxic metals and ˜ Cimero mercury-mining area (Asturias, Spain), J. metalloids from the Munón Hazard. Mater. 136 (2006) 455–467. [14] C. Luque, Las mineralizaciones de Hg de la Cordillera Cantábrica, Doct. Thesis, Universidad de Oviedo (unpublished), 1985. [15] C. Luque, El mercurio en la Cordillera Cantábrica, in: García Guinea, Martínez ˜ C.S.I.C. Textos universitarios No. 15, Frías (Eds.), Recursos minerales de Espana, Madrid, 1992, pp. 803–826. ˜ [16] C. Baldo, J. Loredo, A. Ordónez, J.R. Gallego, J. García-Iglesias, Geochemical characterization of wastes from a mercury mine in Asturias (northern Spain), J. Geochem. Explor. 67 (1999) 377–390. ˜ Release of toxic metals and metalloids form Los [17] J. Loredo, R. Álvarez, A. Ordónez, Rueldos mercury mine (Asturias, Spain), Sci. Total Environ. 340 (2005) 247–260. ˜ M.J. Rucandio, Physicochemical [18] R. Fernández-Martínez, J. Loredo, A. Ordónez, characterization and mercury speciation of particle-size soil fractions from an abandoned mining area in Mieres, Asturias (Spain), Environ. Pollut. 142 (2006) 217–226. [19] J. Loredo, C. Luque, J. García-Iglesias, Conditions of formation of Hg deposits from the Cantabrian zone (Spain), Bull. Mineral. 111 (1998) 393–400. [20] M. Pansu, J. Gautheyrou, Handbook of Soil Analysis: Mineralogical, Organic and Inorganic Methods, Springer, Berlin, 2006. ˜ J. Loredo, Investigation of trace element sources from [21] J.R. Gallego, A. Ordónez, an industrialized area (Avilés, northern Spain) using multivariate statistical methods, Environ. Int. 27 (2002) 589–596. [22] M.G. Cordingley, M.P. Hallewell, J.W. Turner, Release analysis and its use in the optimisation of the comminution and gravity circuits at the Wheal Jane Tin concentrator, Miner. Eng. 7 (1994) 1517–1526.
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[23] B.A. Wills, T.J. Napier-Munn, Mineral Processing Technology: An Introduction to the Practical Aspects of Ore Treatment and Mineral Recovery, 7th ed., Butterworth-Heinemann, Burlington, MA, 2006. [24] F.C. Bond, Third theory of comminution, Miner. Eng. Trans. AIME 193 (1952) 484–494. [25] N. Magdalinovic, A procedure for rapid determination of the Bond work index, Int. J. Miner. Process. 27 (1989) 125–132. [26] J.M. Menéndez-Aguado, B.R. Dzioba, A.L. Coello-Valazquez, Determination of work index in a common laboratory mill, Miner. Metall. Process. 22 (2005) 173–176. [27] R.J. Deister, How to determine the Bond work index using lab ball mill grindability tests, Eng. Miner. J. 188 (1997) 42–45. [28] S. Montinaro, A. Concas, M. Pisu, G. Cao, Remediation of heavy metals contaminated soils by ball milling, Chemosphere 67 (2007) 631–639. [29] V. Deniz, H. Ozdag, A new approach to Bond grindability and work index: dynamic elastic parameters, Miner. Eng. 16 (2003) 211–217. [30] H.T. Ozkahraman, A meaningful expression between bond work index, grindability index and friability value, Miner. Eng. 18 (2005) 1057–1059.
[31] C.W. Williford, R.M. Bricka, Physical separation of metal-contaminated soils, in: I.K. Iskandar (Ed.), Environmental Restoration of Metals-Contaminated Soils, 1st ed., CRC Press, Boca Raton, FL, 2000. [32] R.A. Rikers, P. Rem, W.L. Dalmijn, Improved method for prediction of heavy metal recoveries from soil using high intensity magnetic separation (HIMS), Int. J. Miner. Process. 54 (1998) 165–182. [33] E.G. Erspamer, R.R. Wells, Selective extraction of mercury and antimony from cinnabar-stibnite ore. Bureau of Mines Report of Investigations 5243, United States Department of the Interior, 1956. [34] P. Cauwenberg, F. Verdonckt, A. Maes, Flotation as a remediation technique for heavily polluted dredged material. A feasibility study, Sci. Total Environ. 209 (1999) 113–119. [35] M. Vanthuyne, A. Maes, P. Cauwenberg, The use of flotation techniques in the remediation of heavy metal contaminated sediments and soils: an overview of controlling factors, Miner. Eng. 16 (2003) 1131–1141. [36] G. Dermont, M. Bergeron, M. Richer-Lafleche, G. Mercier, Remediation of metalcontaminated urban soil using flotation technique, Sci. Total Environ. 408 (2010) 1199–1211.
Journal of Hazardous Materials 196 (2011) 101–108
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Water-soluble organo-building blocks of aminoclay as a soil-flushing agent for heavy metal contaminated soil Young-Chul Lee a , Eun Jung Kim b , Dong Ah Ko a , Ji-Won Yang a,b,∗ a b
Department of Chemical and Biomolecular Engineering (BK21 program), KAIST, 335 Gwahak-ro, Yuseong-gu, Daejeon 305-701, Republic of Korea Advanced Biomass R&D Center, KAIST, 291 Daehakno, Yuseong-gu, Daejeon 305-701, Republic of Korea
a r t i c l e
i n f o
Article history: Received 12 April 2011 Received in revised form 28 August 2011 Accepted 31 August 2011 Available online 6 September 2011 Keywords: Heavy metals Soil flushing Water-soluble clay Phyllosilicate Remediation
a b s t r a c t We demonstrated that water-soluble aminopropyl magnesium functionalized phyllosilicate could be used as a soil-flushing agent for heavy metal contaminated soils. Soil flushing has been an attractive means to remediate heavy metal contamination because it is less disruptive to the soil environment after the treatment was performed. However, development of efficient and non-toxic soil-flushing agents is still required. We have synthesized aminoclays with three different central metal ions such as magnesium, aluminum, and ferric ions and investigated applicability of aminoclays as soil flushing agents. Among them, magnesium (Mg)-centered aminoclay showed the smallest size distribution and superior water solubility, up to 100 mg/mL. Mg aminoclay exhibited cadmium and lead binding capacity of 26.50 and 91.31 mg/g of Mg clay, respectively, at near neutral pH, but it showed negligible binding affinity to metals in acidic conditions. For soil flushing with Mg clay at neutral pH showed cadmium and lead were efficiently extracted from soils by Mg clay, suggesting strong binding ability of Mg clay with cadmium and lead. As the organic matter and clay compositions increased in the soil, the removal efficiency by Mg clay decreased and the operation time increased. Crown Copyright © 2011 Published by Elsevier B.V. All rights reserved.
1. Introduction Soil contamination with toxic heavy metals such as copper, zinc, chromium, cadmium, and lead, has been a big concern in many countries due to their adverse health effects [1]. Thousands of sites contaminated with heavy metals need to be remediated in the United States and other countries, which requires new development of effective soil treatment technologies [2–4]. Currently, various soil treatment technologies, including excavation and landfill, isolation, containment, electrokinetics (EK), biological treatment, and soil flushing (washing) have been applied to remediate heavy metal contaminated sites [3–6]. Among them, soil flushing has shown several advantages such as relatively low cost and less environmentally disruptive consequences compared to the conventional excavation and landfill methods [3,4]. Therefore, it has been widely applied to remove heavy metals from contaminated soils. During soil flushing, acids and other solvents have been applied to enhance the performance, which have caused disruption in the soil environment after treatment [7,8]. Thus, chelating agents such as pyridine-2,6-dicarboxylic acid
∗ Corresponding author at: Department of Chemical and Biomolecular Engineering (BK21 program), KAIST, 335 Gwahak-ro, Yuseong-gu, Daejeon 305-701, Republic of Korea. Tel.: +82 42 350 3924; fax: +82 42 350 3910. E-mail address:
[email protected] (J.-W. Yang).
(PDA), N-(2-acetamido) iminodiacetic acid (AIA), and ethylenediaminetetraacetic acid (EDTA), or surfactants [9–14] have been widely employed in place of acids and other solvent to extract heavy metals from soils. However, these chelating agents also resulted in problems because of their recalcitrance and the difficulties separating them from the heavy metal cations, which required additional processes for disposal treatment. Recently, a new class of nanoscale water-soluble chelants of poly(amidoamine) (PAMAM) dendrimer in environmental applications was explored in aqueous solutions [15,16] and soils [1,17]. However, these chelants are expensive and need an additional nanofiltration process to recover them. Therefore, development of alternative chelating agents is still required for an environmentally non-destructive and cost-effective remediation process. Organic–inorganic hybrid nanomaterials have been getting attractions because of their wide applications. Mann et al. have developed 3-aminopropyl functionalized phyllosilicates (aminoclay) by covalent bonding with a central metal ion and sandwiched organo-functionalities via Si–C bonds [18,19] and reported nanocomposites with biomolecules [20–24]. Unfortunately, they do not have focused on environmental areas. Thus, because aminoclay has unique nano-sized and water-soluble properties with a high density of amino groups, it can lead to applications in environmental studies. In addition, its high chelating capacity with metal ions and less toxicity make the aminoclay as an excellent candidate for a soil flushing agent.
0304-3894/$ – see front matter. Crown Copyright © 2011 Published by Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.077
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Table 1 Properties of Jumunjin sand, Soil A, B, and C used in this study.
Jumunjin sand Soil A Soil B Soil C
Density (g/cm3 )
Bulk density (g/cm3 )
Porosity
pH
Organic content (%)
2.320 2.340 2.286 2.254
1.500 1.521 1.136 0.789
0.582 0.350 0.497 0.650
– 5.85 4.00 6.48
0.11 0.6 3.5 12.1
Soil composition (%) Sand (2–0.02 mm)
Silt (0.02–0.002 mm)
Clay (<0.002 mm)
100 93.5 75.0 55.0
0 2.5 15.0 28.0
0 4.0 10.0 17.0
The objective of this study is to demonstrate that aminoclay can be applied as an alternative soil-flushing agent to remediate heavy metals from soils. We synthesized aminoclays with three different central metal ions and investigated heavy metal binding of aminoclays through batch and column experiments.
day where adsorption equilibrium time was confirmed <1 h. Samples were agitated with magnetic stirrer during the reaction. After reactions, all samples were filtered by an Amicon® centrifugal filter (3 kDa) and then analyzed by atomic absorption spectrometry (AAS, Perkin–Elmer). The removal capacity (q) was calculated by
2. Experimental
(Co − Cfiltered ) × V = q (mg/g) W
2.1. Materials All chemical used in this study were reagent grade or higher grade and used without further purification. 3-Aminopropyltriethoxysilane (APTES, 99%), ferric chloride hexahydrate, and aluminum chloride hexahydrate were purchased from Sigma–Aldrich (St. Louis, MO, USA) and magnesium chloride hexahydrate (98.0%) was obtained from Junsei Chemical Co., Ltd. (Tokyo, Japan). Cadmium and lead sources were used as chloride counterions (>99%, A.C.S. reagent). The 1000 mg/L standard solutions of cadmium, lead, and magnesium were obtained from Fluka. Ethanol (>99.9%) was purchased from Merck KGaA (Darmstadt, Germany). Centrifugal filters (Amicon® Ultra, 3 kDa) were bought from Millipore. HCl or NaOH standard solutions (1 N or 0.1 N) were used from DAE JUNG (Shiheung, Korea) to adjust the pH of the samples. Double-distilled deionized water (>18 m, DI water) was used through all of the experiments. 2.2. Preparation of hydrophilic aminoclays An aminopropyl functionalized magnesium phyllosilicate clay was prepared as previously reported [18–24]. The APTES (1.3 mL, 5.63 mmol) was added in a drop-wise manner to a magnesium chloride (0.84 g, 8.82 mmol) in ethanol (40 mL) at room temperature. The molar ratio of Mg to Si was approximately 0.75. The white slurry was formed after 5 min, which was stirred overnight. The precipitate was isolated by centrifugation, washed with ethanol (50 mL) to remove excess magnesium chloride, and dried at 40 ◦ C. Aminoclays with backbones of Fe and Al were also synthesized using the same procedure described above. For brevity, Mg, Al, and Fe backboned aminoclays will be referred to as Mg clay, Al clay, and Fe clay respectively. 2.3. Cadmium and lead binding of Mg clay in batch system Batch experiments were conducted to determine the maximum binding capacity of the heavy metal cations in Mg clay solution as a function of pH. The Mg clay solution (0.5 g/L) was reacted with 50 mg/L of cadmium or lead solution at various pH values for 1
where C0 and Cfiltered were the initial and filtered concentrations after centrifugation (mg/L), V (L) was the sample volume, and W was the mass (g) of the Mg clay. The experiments were conducted in triplicate to get a reliable result. 2.4. Mg clay flushing for cadmium or lead contaminated soils Column flushing experiments were performed with a 30 mm × 160 mm column (Kontes, USA) containing 110 g of Jumunjin sand (20–30 mesh), and Soil A, B, and C. Jumunjin sand was obtained from seashore of east coast of Korea (Jumunjin, Korea) and Soil A, B, and C were collected on a hill site inside of the university (Daejeon, Korea) and their physical properties were artificially controlled to have different sand, silt, and clay contents. The clay and silt contents increased in the order of Soil A < B< C (Table 2). The elemental compositions observed by X-ray fluorescence (XRF) and physical properties of the soils are summarized in Tables 1 and 2, respectively. X-ray diffraction analysis showed that quartz, illite, and albite are the major mineral composition of Soil A, B, and C. Approximately the same loading of each soil was packed in the column by small incremental actions to obtain homogeneous condition with uniform bulk density [25]. And soil column experiment was The schematic diagram of the column experiment in detail was depicted in Fig. 1. The pore volume of the sand in the column was 30 mL. Initially, 20 pore volumes of DI water were pushed through the column in an upward direction at a flow rate of 2 mL/min. Next, the column was contaminated by circulating an initial concentration of approximately 300 mg/L of divalent cadmium and lead, individually, at a flow rate of 1 mL/min to achieve an equilibrium state for 6 h (at pH 7). Then, more than 30 pore volumes of DI water was introduced in an upward direction in order to remove slightly bounded heavy metals from the soil surfaces. As a result, complete saturation with water for the column was achieved. The preloaded Cd2+ and Pb2+ uptakes for the Jumunjin sand soil were 120 ± 10 and 180 ± 20 mg/kg, respectively. In cases of Soils A, B, and C (order of clay contents: A < B < C), the preloaded concentrations of Pb2+ were 1016, 1216, and 1446 mg/kg, respectively, reflecting sorption of heavy metals on soils were highly affected by the characteristics of the soils. This will be further discussed in Section 3.5. For the flushing of
Table 2 Elemental compositions (%) of Jumunjin sand, Soil A, B, and C observed by X-ray fluorescence (XRF) spectrometer in this study.
Jumunjin Soil A Soil B Soil C
Si
Al
K
Fe
Ca
Ti
Cu
S
Mn
Na
68.4 67.3 42.7 46.4
8.9 13.9 23 19.2
12.7 8.95 9.15 7.47
0.992 5.43 15 18.2
0.36 2.34 3.86 1.2
0.12 0.936 1.39 1.74
0.065 0.11 0.079 0.11
0.23 0.51 0.36 0.66
– 0.11 0.34 0.53
6.6 – – –
Y.-C. Lee et al. / Journal of Hazardous Materials 196 (2011) 101–108
103
Fig. 1. Schematic presentation of the column experiment for soil flushing.
2.5. Characterizations Surface morphology of samples of aminoclays in aqueous media was observed by a transmission electron microscopy (TEM, JEM2100F, JEOL LTD, Tokyo, Japan). In order to observe TEM samples, tiny amount of aminoclay powder was dispersed in DI water for 5 min sonication, and then subsequently was dropped onto a carbon-coated copper grid. For crystallographic structure of aminoclays, Powder X-ray diffraction (PXRD) data were obtained on a Rigaku D/max lllC (3 kW) with a wide angle goniometer equipped with a CuK␣ radiation generator at 40 kV and 45 mA, and the scan range was from 3 to 65◦ at a rate of 1.2◦ × 2/min. Each aminoclay that prepared with mechanical mortar was filled in the holder homogeneously. Zeta potentials of the aminoclay solutions were measured by the Malyern Zetasizer Nano-ZS particle analyzer according to pH. The particle size distributions of aminoclays in aqueous solutions were examined by dynamic light scattering (DLS) method (HELOS/RODOS & SUCELL, Germany). The point of zero charge (PZC) of Soil A, B, and C was measured by titration method [29]. The titration was performed with addition of HCl and KOH (0.1 M), which was corrected by subtraction from the pH of the blank to be required the amounts of acid. To confirm the covalent bonding of central metals and APTES and its organofunctionalities, Fourier transform infrared (FT-IR) spectra of KBr pellets (FT-IR 4100, Jasco, Japan) were collected from 4000 cm−1 to 540 cm−1 . Each spectrum was recorded as the average of 32 scans with a resolution of 4 cm−1 . The pellet disks consisted of 10% aminoclay powder and 90% KBr by weight. The compositions of the Mg clay, Jumunjin sand, and Soil A, B, and C were examined by X-ray fluorescence (XRF, MiniPal2, PANalytical) and an elemental analysis analyzer (EA1108 andNA2000, CEinstrument, USA) was used for aminoclays. Along with that, total amine (TN) and protonated amine (ammonium) in aqueous solution were measured by automatic water analyzer (AACS V, Center for Research Facility in Chungnam National University, Daejeon, Korea). The pH was determined using an Orion pH meter (Thermo Orion, model 710, USA).
3. Results and discussion 3.1. Characterizations of aminoclays Powder X-ray diffraction (PXRD) patterns for the synthesized aminoclays are shown in Fig. 2. The XRD patterns of Mg, Fe, and Al clays showed typical layered organoclay structure, i.e., d0 0 1 = 1.5 − 1.8 nm, which corresponded to the previous results [30]. In addition, 2 = 59◦ was a distinct peak in the Mg clay, which indicated a 2:1 trioctahedral phyllosilicate ratio. Al and Fe clays had a similar 2:1 dioctahedral phyllosilicates ratio or 1:1 phyllosilicate [30,31]. The main functional groups of aminoclays were characterized as –CH2 , –NH2 , Si–O, Si–O–Si, and metal-O by Fourier transform infrared (FT-IR) spectroscopy (Fig. 3). All aminoclays formed covalent bonds with the metal ions and organo-functionalities, further –NH2 , and protonated –NH3 + peaks appeared [18,20,24]. X-ray fluorescence (XRF) study showed that Mg-, Al-, and Fe-clay were composed of metal (Mg: 13.0%, Al: 11.8%, and Fe: 43.5%), Si (24.0%, 31.6%, and 26%), and Cl (63.0%, 56.6%, and 30.9%), respectively, which were calculated by summation of three compositions of each metal, Si, and Cl is 100 wt%. Elemental analysis (EA) determined the compositions of C, H, O, and N of Mg clay were 18.3%, 5.6%, 9.4% and 6.7%, respectively. Furthermore, N (%)
Mg clay Fe clay Al clay
Intensity (a.u.)
contaminants from the column, water-soluble Mg clay solutions of 10 mg/mL were pumped in an upward direction through the column at a flow rate of 1 mL/min. After the flushing procedure, residual amounts of Mg clay on the soil were analyzed. During column experiments, flow rate was controlled by a peristaltic pump (Masterflex® L/S® , Cole-Parmer Instrument, USA) [26–28].
d020,110
d001
d130,200
d002
10
20
30
d060,330
40
50
60
2 theta Fig. 2. Powder X-ray diffraction (PXRD) patterns of aminoclays.
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Transmittance (%)
Table 3 Particle size distributions of aminoclays.
X10 X50 X90 X99 SMD VMD Sv
NH3+
Mg clay Fe clay Al clay
OH
NH2
3000
2000
1000 -1
Wavenumber(cm ) Fig. 3. Fourier transform infrared (FT-IR) spectra of aminoclays. M is denoted as metal.
compositions of Al and Fe clay were 7.91% and 8.20%. From the composition results above, the density of the amino groups of bulk powder of aminoclays was calculated as 4.78 mmol, 5.65 mmol, and 5.86 mmol of nitrogen/g of Mg clay, Al clay, and Fe clay respectively. In order to examine detailed amine function of aminoclays in aqueous solution, total amine (TN) and protonated amine (ammonium form, −NH3 + ) concentrations in aqueous solution (pH 5–6) were measured. The TN concentrations of Mg-, Al-, and Fe-clay were 97.21 × 10−3 , 85.95 × 10−3 , and 86.50 × 10−3 mg/L, respectively. The protonated amine concentrations of each clays were 3.78 × 10−3 , 7.04 × 10−3 , and 10.48 × 10−3 mg/L. Interestingly, among the aminoclays, Mg clay in aqueous solution showed highest amine concentration, whereas it showed lowest concentration of protonated amine (−NH3 + ), which was not proportional to N density (mmol) by EA. Zeta potential of aminoclays showed positive values of +20–25 mV, while talc, which is a parent mineral of Mg clay, had negative potential (Fig. 4a). Even with increased Mg clay loading, no significant increase of zeta potential occurred (data not shown). As illustrated in Fig. 4b, the positive value of the zeta potential decreased to approximately half of the maximum value, i.e., +5–10 mV, near neutral pH, corresponding to reported dendrimer (∼+5 mV) [15,16]. The pH of Al clay in aqueous solution exhibited neutral pH while other clays in aqueous solution showed an approximate pH of 9.8 at 1.5 g/L of clay loading.
a
Al clay
Fe clay
24.73 nm 31.87 nm 41.08 nm 50.49 nm 31.27 nm 32.49 nm 191.89 m2 /cm3
269.69 nm 320.56 nm 370.93 nm 394.09 nm 315.84 nm 320.39 nm 19.00 m2 /cm3
72.90 nm 92.42 nm 117.03 nm 141.67 nm 90.93 nm 94.02 nm 65.99 m2 /cm3
X10 %, X50 %, and X90 % are the cumulative probability sizes at 10%, 50%, and 90%, abbreviations of surface mean diameter and volume mean diameter are SMD and VMD, Sv is the ratio of SMD and VMD.
Si-O-Si Si-C
CH2
4000
M-O
CH2
Mg clay
40
3.2. Screening of soil flushing agents The size distributions of aminoclays observed by dynamic light scattering (DLS) are presented in Table 3. For the Mg clay, the 99% cumulative probability size (X99 ) was 50.49 nm and the ratio of the surface area mean diameter (SMD) and volume mean diameter (VMD), Sv, was 191.89 m2 /m3 . The higher the value of Sv means the more spherical shape or the smaller particle size. The average diameters (X99 ) of Al and Fe clays were 394.09 and 141.67 nm, respectively, while the Sv of them were 19.00 and 65.99 m2 /m3 , respectively. Thus, among the clays, Mg clay had the smallest diameter and more spherical shape, suggesting relatively high probability to exist as discrete particles in aqueous solution. The diameter and the shape of the clays were affected by the metal ions of the clay even though they were synthesized under the identical sol-gel reaction with APTES. The discrete dispersion of Mg clay sheets in aqueous solution was confirmed by TEM images, having particle size of 30–200 nm (Fig. 5a). On the other hand, Al clay (Fig. 5b) and Fe clay (Fig. 5c) sheets showed aggregation by stacking, which may be resulted from sample preparation. The smallest size of Mg clay in aqueous phase suggests that the Mg clay might be the most efficient candidate as a soil-flushing agent among the aminoclays when considering the mass transfer between the pores of the soil media. To indirectly observe the solubility of aminoclays in aqueous solution, UV-vis absorption spectra were measured at pH 6 (Fig. 6). Mg and Al clays were highly soluble in aqueous solution in the visible region (400–800 nm of wavelength), even at 10 mg/mL of Mg clay dosage, while Fe clay was more turbid when compared to Mg and Al clays at the same concentration. Based on the size distribution, solubility, amino density, and the zeta potential of aminoclays, we selected Mg clay as a promising candidate for soil flushing and further studied heavy metal binding capacity and soil flushing efficiency of Mg clay in the following sections.
b
35
Zeta Potential (mV)
Zeta potential (mV)
30 30 20 10 0
25 20 15 10 5
-10
0 Mg Clay
Al Clay
Fe Clay
Talc
2.33
3.26
4.2
5.76
6.98
pH Fig. 4. Zeta potential values of aminoclays (a) at 1.5 g/L of Mg clay as a function of pH (b).
8.83 10.16
10.56
Y.-C. Lee et al. / Journal of Hazardous Materials 196 (2011) 101–108
105
Fig. 5. TEM images of exfoliated dispersions of Mg (a), Al (b), and Fe (c) clay sheets at 1.5 g/L.
a
b
2.0
1.5 1.20 mgml-1 4.99 mgml-1 10.44 mgml-1 15.00 mgml-1 20.10 mgml-1 -1 25.00 mgml
1.0
Absorbance
Absorbance
1.5
1.20 mgml-1 4.99 mgml-1 10.44 mgml-1 15.00 mgml-1 20.10 mgml-1 25.00 mgml-1
1.0
0.5
0.5
0.0 200
300
400
500
600
700
800
0.0 200
300
400
500
600
700
Wavelength (nm)
Wavelength (nm)
c
2.0
2.0
Absorbance
1.5
1.0 1.00 mgml-1 1.50 mgml-1 2.00 mgml-1 2.50 mgml-1 3.18 mgml-1 3.51 mgml-1
0.5
0.0 200
300
400
500
600
700
800
Wavelenght (nm) Fig. 6. UV–Vis absorption spectra of Mg clay (a), Al clay (b), and Fe clay (c) as a function of clay loading in aqueous solution.
800
-1
-1
binding capacity by Mg clay (mgg )
Y.-C. Lee et al. / Journal of Hazardous Materials 196 (2011) 101–108
100
a
80
60
40
2+
2+
20
0 2
4
6
8
10
Pb
Cd binding capacity by Mg clay (mgg )
106
b
100
80
60
40
20
0 2
4
pHfinal
6
8
10
pHfinal
Fig. 7. Binding capacity (q) of cadmium (a) and lead (b) by Mg clay.
Fig. 7 shows a binding capacity of cadmium and lead by Mg clay (0.5 mg/mL) as a function of pH. The cadmium and lead binding capacities of Mg clay were strongly dependent on the solution pH value. The removal of both metals by Mg clay increased as pH increased. However, in case of Cd2+ : >pH 8.4 and Pb2+ : >pH 6.0 to be plotted by chemical equilibrium modeling of Minteq (version 2.0) (data not shown), the removals of metal ions could be also contributed by alkaline precipitates (M(OH)2 , where M = metal ion) as well as binding of Mg clay. Although ionic mobility of Cd2+ is higher than that of Pb2+ ion at neutral pH by consideration of chemical equilibrium modeling, Mg clay showed higher affinity with lead than cadmium in all pH conditions. Those are in line with aminofunctionalized mesoporous material for removal of heavy metals [32–34]. The FT-IR study showed that main functional groups of aminoclays were –CH2 , –NH2 , Si–O, and Si–O-metal, corresponding to the ideal unit structure of aminoclay [35]. Among the functional groups, heavy metal ions are reported to be well complexed with –NH2 groups. This can be explained that in an acidic condition, Cd2+ or Pb2+ forms of heavy metals and –NH3 + forms of Mg clay sheets are predominant, resulting in prevention of heavy metal cations adsorbed on Mg clay sheets, but at near neutral pH, free NH2 of Mg clay sheets can easily chelate on heavy metal ions. Therefore, the uptake of cadmium and lead ions by Mg aminoclay seemed to be mainly contributed by chelating with –NH2 groups of amin-
Cd2+ concentration (mgL-1)
a
200 180 160 140 pH 1.85 of Mg clay solution pH 3.04 of Mg clay solution pH 6.40 of Mg clay solution pH 7.75 of Mg clay solution pH 9.95 of Mg clay solution DI water
120 100 80 60 40 20
oclay. Less protonated NH2 group in neutral pH is known to better complex with heavy metal ions at near neutral pH [36–39], which explained the higher binding of cadmium and lead ions at near neutral pH. The point of zero charge (PZC) of Mg clay sheets become approximately pH 11.0 [40], indicating that dispersions of Mg clay in DI water shows positive surface potential in wide pH range. The zeta potential (mV) to be dealt in pH range presents +5 to +15 mV. While the measured PZC of Soil A, B, and C used in this study indicates pH 5.4 ± 0.2, 5.0 ± 0.2, and 4.6 ± 0.2, respectively. Thus, heavy metal cations could be loosely bound onto soil surface at near neutral pH due to repulsion with soil surface, this result would be result in higher removal efficiency of heavy metal by Mg clay, which was correlated that free NH2 groups of Mg clay increased as increase of pH. However, in acidic conditions, the binding was low because of the competition with proton ions. This indicates that Mg clay can be recovered by acid treatment of Mg clay and metal complexes after soil flushing treatment [41]. 3.4. Soil-flushing with Mg clay: effect of pH Fig. 8 shows the effluent concentrations of cadmium and lead extracted from Jumunjin sand soil during soil flushing column experiments with Mg clay as a function of pH. When only DI water was flushed without pH control, the effluent concentration was almost negligible indicating that DI water only did not extract metals from soil. During soil flushing, the effluent Cd2+ and Pb2+ concentrations increased sharply and at all pH conditions high-
b
300
Pb2+ concentration (mgL-1)
3.3. Cadmium and lead binding efficiency (q) with Mg clay in the batch system
250 pH 1.85 of Mg clay solution pH 3.04 of Mg clay solution pH 6.40 of Mg clay solution pH 7.12 of Mg clay solution pH 9.95 of Mg clay solution DI water
200 150 100 50 0
0 0
50
100
150
200
Effluent volume (mL)
250
300
0
50
100
150
200
Effluent volume (ml)
Fig. 8. Mg clay flushing of cadmium (a) and lead (b) contaminated with Jumunjin sand soil.
250
300
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3.5. Soil-flushing with Mg clay: effect of soil properties In order to investigate the effect of soil properties on soil flushing with Mg clay, soil flushing experiments were conducted with three different types of soils contaminated with lead. We only tested lead-contaminated soil because Mg clay showed better removal efficiency of lead compared to cadmium (Fig. 8). Soil A, B, and C had different physical properties and organic matter contents. The clay and organic matter contents, whose main fraction is humic acid and fulvic mixture, of the soils increased in the order of Soil A, B and C (Table 1). These soils were contaminated with lead and the preloaded lead concentrations were 1016, 1216, and 1446 mg/kg for Soil A, B, and C, respectively. Fig. 9 shows the effluent concentrations of lead extracted from Soil A, B, and C during soil flushing column experiments with Mg clay at pH 6.3, which was the optimum pH of the soil flushing with Mg clay in the previous section. The highest effluent lead concentration was observed for Soil A. About 81% of initially loaded lead on Soil A was removed after soil flushing with Mg clay. The removal efficiencies of Soil B and C were 72.5% and 43.1%, respectively. The preloaded amount of lead on soil increased with the increase of clay and organic matter contents in soil, while the removal efficiency of lead by Mg clay decreased, indicating soil properties affected binding of lead on soil and interactions between lead ions and the removal efficiency of Mg clay. Natural organic matter (humic acid and fulvic acid mixture) and metal oxides such as Mn oxide, Fe oxide, Al oxide, and Ti oxides present in soil have shown to have high affinity for lead [13,42–44]. The XRF study showed that metal oxides as well as sulfates increased in the order of Soil A, B, and C but silicates content did not show a trend (Table 2). The high contents of organic mat-
1800
2+
-1
concentration (mgL )
1600
Pb
est concentrations were observed when effluent water volume was 60 mL, which is the two pore volume of the column. This is related to the pore volume of column, approximately 30 mL. After displacing saturated DI water, Mg clay solution collides with heavy metal ions bound onto soil, following that heavy metal ions are directly released out after binding with Mg clay sheets. The cadmium and lead elution from the soil column with Mg clay solution indicates the strong binding ability of Mg clay with cadmium and lead. The removal efficiency of lead by Mg clay was higher than cadmium, which can be explained by higher binding capacity of Mg clay for lead than cadmium as observed in Section 3.3 (Fig. 7). The highest removal efficiency of cadmium and lead was observed at pH 6.4, where the total maximum removal efficiency of cadmium and lead was 61.95% and 83.83%, respectively. Solution pH affected the removal efficiency of cadmium and lead from soil. The pH could affect both availability of metals in soil and affinity to Mg clay. As observed in the previous section, removal of metals by Mg clay increased as pH increased because of increase in free NH2 of Mg clay. This resulted in increase of the affinity on heavy metal ions with Mg clay where the solubility of Mg clay was negligibly affected. However, metal precipitates could occur at higher pH, which could make hard to elute from the soil. On the other hand, at lower pH the available metal ions increased from the sand soil due to the increased protonation of soil surface, but affinity of metals to Mg clay was also lowered at the same time. In this study, the optimum pH of the soil flushing with Mg clay was observed at pH 6.4 for lead and cadmium ions. In the case of lead, the lowest removal efficiency was observed at the lowest pH, pH 1.85, which suggests that the affinity to Mg clay of lead is more important factor for lead removal efficiency from soil flushing than the available lead in soil. In addition, the shape of the breakthrough curve (BTC) of cadmium elution by Mg clay solution showed symmetry and no tailing without irregularities (Fig. 8a), while lead BTC showed the tailing phenomena. The tailing of the BTC suggests relatively slow mass transfer of lead bounded Mg clay in the soil.
107
1400 1200 1000
Soil A Soil B Soil C
800 600 400 200 0 0
100
200
300
400
500
Effluent volume (ml) Fig. 9. Mg clay flushing of lead contaminated Soils A, B, and C.
ter, metal oxides, and sulfates in Soil C seemed to make lead more strongly bound to the soils as well as Mg clay was predominantly interacted with organic matters than surrounding anionic metals [45], which results decrease of removal efficiency of Mg clay. Furthermore, slow mass transfer of lead bounded Mg clay occurred in Soil C, which was suggested by obvious tailing phenomena of BTC of Soil C. It took twice longer time (up to 10 h) than flushing of Soil A to reach the time when lead leaching into eluent was stopped. The maximum concentration of lead in the eluent was delayed about one pore volume (30 mL for 20 min). These results suggested that soil properties should be considered for the application to real or weathered heavy metal contaminated soil. Thus, pretreatment of organic matter will be needed prior to application of aminoclay flushing to remediate heavy metals or development of a novel aminoclay by controlling functionalities to interact with heavy metals fast and strongly than with organic matters. Finally, recovery of Mg clay was calculated after soil flushing. The amounts of Mg clay remained in the Jumunjin sand, Soil A, B, and C after lead flushing were 86 ± 20 mg, 89 ± 30 mg, 92 ± 30 mg, and 95 ± 20 mg, respectively. This indicated that when the organic matter and clay contents in the soil increased, the sorption of Mg clay on soils also slightly increased as sorption of metals on soils. However, the differences were not statistically significant (S.D. = 4.5). 4. Conclusion In this study, we have synthesized aminoclays with three different central metal ions such as magnesium, aluminum, and ferric ions and investigated applicability of aminoclays as soil flushing agents. Among them, Mg clay showed the smallest size distribution and superior water solubility, which suggests Mg clay as a promising candidate for soil flushing agent due to its efficient mass transfer in soils. In batch experiments, the cadmium and lead binding capacities of Mg clay were best near neutral pH, while at acidic conditions, the binding capacities of them were almost negligible. This indicates that when Mg clay is applied to soil flushing, the pH of the solution should be adjusted to approximately neutral pH to achieve the best performance. In addition, acid treatment process could recover the Mg clay after soil flushing. Due to a high-density of amine groups and water-solubility of Mg clay, it shows a possibility of nanoparticle approaches for the use of soil-flushing agents, instead of molecular (ion) applications such as surfactants, acid, and base agents. Moreover, various aminoclays with a higher aminodensity can be applied to enhance removal efficiency. Especially when consideration of soils with high contents of organic matter,
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a design of novel aminoclays interacted with heavy metals more strongly is in progress.
[22]
Acknowledgements [23]
This subject is supported by Korea Ministry of Environment as “Converging technology project (2010)” and supported by the Advanced Biomass R&D Center (ABC) of Korea Grant funded by the Ministry of Education, Science and Technology (ABC-20100029728).
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Journal of Hazardous Materials 196 (2011) 115–122
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Efficacy of liquid and foam decontamination technologies for chemical warfare agents on indoor surfaces Adam H. Love a,1 , Christopher G. Bailey b,∗,1 , M. Leslie Hanna b , Saphon Hok b , Alex K. Vu b , Dennis J. Reutter b , Ellen Raber c a
Johnson Wright, Inc., 3730 Mt Diablo Blvd, Suite 230, Lafayette, CA 94549, United States L-091, Lawrence Livermore National Laboratory, Livermore, CA, 94550, United States c L-184, Lawrence Livermore National Laboratory, Livermore, CA, 94550, United States b
a r t i c l e
i n f o
Article history: Received 27 April 2011 Received in revised form 1 September 2011 Accepted 1 September 2011 Available online 8 September 2011 Keywords: Chemical warfare agent Decontamination Sarin Soman Sulfur mustard VX
a b s t r a c t Bench-scale testing was used to evaluate the efficacy of four decontamination formulations on typical indoor surfaces following exposure to the liquid chemical warfare agents sarin (GB), soman (GD), sulfur mustard (HD), and VX. Residual surface contamination on coupons was periodically measured for up to 24 h after applying one of four selected decontamination technologies [0.5% bleach solution with trisodium phosphate, Allen Vanguard Surface Decontamination Foam (SDFTM ), U.S. military Decon GreenTM , and Modec Inc. and EnviroFoam Technologies Sandia Decontamination Foam (DF-200)]. All decontamination technologies tested, except for the bleach solution, performed well on nonporous and nonpermeable glass and stainless-steel surfaces. However, chemical agent residual contamination typically remained on porous and permeable surfaces, especially for the more persistent agents, HD and VX. Solvent-based Decon GreenTM performed better than aqueous-based bleach or foams on polymeric surfaces, possibly because the solvent is able to penetrate the polymer matrix. Bleach and foams out-performed Decon Green for penetrating the highly polar concrete surface. Results suggest that the different characteristics needed for an ideal and universal decontamination technology may be incompatible in a single formulation and a strategy for decontaminating a complex facility will require a range of technologies. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Implementing an efficient remediation and recovery process after a civilian facility is contaminated with a chemical warfare agent (CWA) requires understanding the efficacy of a range of decontamination technologies. Many liquid and foam decontaminants have been developed to address the decontamination needs and performance criteria for military operations [1,2]. Far less is known about the performance of such technologies for application to civilian infrastructure [3]. Even trace amounts of residual chemical contamination may prove unacceptable in civilian settings [4,5]. As part of an effort funded by the U.S. Department of Homeland Security to improve the nation’s preparedness for indoor facility restoration after a CWA release, four liquid and foam
∗ Corresponding author. Tel.: +1 925 422 0578; fax: +1 925 423 9014. E-mail addresses:
[email protected] (A.H. Love),
[email protected] (C.G. Bailey),
[email protected] (M.L. Hanna),
[email protected] (S. Hok),
[email protected] (A.K. Vu),
[email protected] (D.J. Reutter),
[email protected] (E. Raber). 1 The first two authors contributed equally to this work. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.005
decontamination technologies for typical indoor surfaces were evaluated experimentally for efficacy against GB, GD, HD, and VX contamination. The U.S. Environmental Protection Agency has evaluated separately the efficacy of other technologies, principally fumigation, for decontaminating CWAs on typical indoor surfaces [6–8]. Although it was anticipated that each of the decontamination technologies tested as part of the present investigation would have some efficacy under the conditions for which it was designed, it is important to compare and quantify the efficacy of each technology as part of an effort to develop an effective overall decontamination strategy for civilian applications. The results of decontamination efficacy are typically reported in two ways: (1) by a commercial vendor through marketing material, which often does not contain the level of experimental detail necessary to evaluate the validity of efficacy claims, or (2) in military reports that describe efficacy in terms of meeting military criteria. Many efficacy evaluations are not performed on target chemicals themselves but, rather, on chemical surrogates [9,10] that may have limited ability to mimic all the important physico-chemical properties of target chemicals. In addition, previous decontamination efficacy testing [11] is often performed on the simplest of substrates, namely nonporous and
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nonpermeable surfaces, where the CWA resides as a free liquid on the surface instead of being bound physically or chemically by the surface. Furthermore, because no experimental data are available that compare the performance of different liquid and foam decontamination technologies using similar testing protocols, the relative performance among existing technologies is not known. Surface decontamination technologies require contact between an active decontamination component and a CWA. If the kinetics of a reaction with a CWA are slow, or there are mass-transport limitations, increasing the contact time may increase the overall decontamination performance. Decontamination foams or gels [12] cling to surfaces and are designed to increase the contact time of the decontamination technology on the surface compared to liquid formulations that rapidly run off some surfaces, such as vertical walls and ceilings. Although numerous decontamination technologies are in development, the current availability of a particular technology is also a critical operational consideration when rapid facility decontamination and restoration are paramount. The four surface decontamination technologies evaluated in this study were chosen to span a range of available technology characteristics. Bleach (sodium hypochlorite) is a strong oxidizing aqueous solution that has widespread availability and a long history of use in CWA decontamination [8]. Liquid bleach does not have good contact time on vertical walls or ceilings, thus typical decontamination operations with bleach employ bleach scrubbing or multiple applications. Bleach is highly corrosive and should not be applied to sensitive electronic equipment intended for reuse. Two commercially available, aqueous-based, oxidizing foams were tested. The Allen Vanguard Surface Decontamination Foam (SDFTM ) formulation, which is a member of the foam family based on the Canadian Aqueous System for Chemical/Biological Agent Decontamination (CASCADTM ), is specifically designed for building decontamination. Sandia Decontamination Foam (DF-200) is available from Modec Inc. and EnviroFoam Technologies. Both of these foam decontamination technologies feature a less aggressive oxidation technology, compared to bleach, and both result in better corrosion prevention. The fourth surface decontamination technology tested is the latest U.S. military liquid formulation, Decon GreenTM , which represents a solvent-based decontamination technology in contrast with aqueous-based technologies. Decon GreenTM is not commercially available, but has been licensed to Strategic Technologies Enterprises, Inc. (STE), a subsidiary of STERIS Corp. Although corrosion concerns for solvent-based liquid decontamination technologies are minimal, potential materialscompatibility issues with plastics and polymers may prevent the reuse of such materials after decontamination. 2. Experimental 2.1. Agent synthesis Neat liquids of four CWAs were synthesized at Lawrence Livermore National Laboratory (LLNL): (1) sarin, (GB, isopropyl methylphosphonofluoridate); (2) soman, (GD, pinacolyl methylphosphonofluoridate); (3) sulfur mustard, (HD, bis (2-chloroethyl) sulfide); and (4) VX, (O-ethyl S-[2(diisopropylamino)ethyl] methylphosphonothioate). The purity of each of the four CWAs was verified to be >97% using gas chromatography–mass spectrometry (GC–MS) analysis. International treaties regulate possession of highly toxic chemical warfare agents, and handling is only permitted in laboratories approved for CWAs under strict scrutiny. All work was performed through the Forensic Science Center and Lawrence Livermore National Laboratory, which has the authority and capability to synthesize and safely handle small quantities of CWAs. All experiments were conducted in triplicate, using standard scientific QA
procedures, including positive and negative controls and routine instrument calibrations. 2.2. Indoor materials Materials chosen for exposure to CWAs were selected from a range of typical indoor surfaces. Common materials that are easily removable (e.g., carpeting, acoustic ceiling tiles, and furniture) were not considered. Materials purchased and used as-is for the evaluation were stainless-steel coupons made from 1/16-in.-thick (∼1.56-mm) sheets of 304 stainless steel; vinyl floor tile [Armstrong commercial flooring, Standard Excelon vinyl composition tiles, Pattern 51858, Imperial Texture, sandrift white, 1/8-in. (3.175-mm) thick]; latex-painted drywall [standard 0.438-in. (11.1-mm) drywall painted with 1 coat of Glidden commercial latex primer and 1 coat of interior eggshell paint]; and glass (Gold Seal Microslides, Becton Dickinson and Co., soda-lime microscope slide glass, precleaned, ground polished edges, plain). Concrete coupons were made at LLNL from a water and Portland cement mass ratio commonly used in construction (0.485:1.0), but made lean in sand (sand to cement ratio = 3, instead of 5–6) to be workable and to avoid extensive entrapped air in the cast coupons (35-mm diam, 17mm thick). Portland cement type I/II (Quikrete brand) was used with a well-graded sand aggregate from U.S. Silica (ASTM 20/30, C-778), with 98% of the particles between 600 and 850 m (20 and 30 mesh). Because the reactivity of newly cured concrete with CWA may not be representative of most concrete in facilities [13], concrete coupons were rapidly aged in a 25% CO2 atmosphere for 2 weeks to reduce the reactivity through carbonation, the same mechanism by which concrete naturally reduces its reactivity, albeit more slowly. Other than the cast concrete coupons, all other materials for chamber exposure were cut into pieces, with top surface areas ranging from approximately 2 to 10 cm2 . 2.3. Preparation of decontamination technologies All decontamination technologies were prepared immediately prior to surface application. The liquid bleach decontamination formulation was prepared by diluting Clorox® regular bleach (5% by mass sodium hypochlorite) with Milli-Q water in a 1:9 ratio to create a 0.5% by mass sodium hypochlorite solution. Trisodium phosphate was added as a surface wetting agent so that the final decontamination solution contained 0.0625% by mass of trisodium phosphate. Decon GreenTM was prepared in 100-mL batches according to the formulation described in U.S. military reports [11], namely 60% propylene carbonate, 10% aqueous H2 O2 , 10% triton X-100, 2.07 g K2 CO3 , and 0.48 g K2 MoO4 . Easy DeconTM DF-200 foam was purchased and prepared in a 2-L beaker by mixing the 3-part formulation in the same ratios as those specified by the manufacturer. Initially, 95 mL each of Part 1 “Penetrator,” containing quaternary ammonium compounds and benzyl-C12-C16 alkyl di-methyl chlorides, and Part 2 “Fortifier,” containing liquid hydrogen peroxide, were mixed well. Then, 4 mL of Part 3 “Fortifier Booster,” containing diacetin, was added, and a Cuisinart hand blender was used to create the decontamination foam. SDFTM Foam was purchased and prepared according to the manufacturer’s directions for the multicomponent formulation. Thus, 1.8 g of GPB-2100 dry-powder buffer component was rinsed using Milli-Q water into a 50-mL cylinder. Then 4.5 mL of GCE-2000 surfactant or foaming agent component was added to the graduated cylinder, and the total volume in the graduated cylinder was increased to 50 mL with Milli-Q water. A micro stir-bar was used to mix until dissolved. Separately, 7.8 g of GP-2100 dry-powder decontaminant component was dissolved in Milli-Q water and
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brought to a volume of 50 mL. The two solutions were combined and brought to a final volume of 150 mL with the addition of Milli-Q water. The 150 mL of solution was then blended in a Waring Blender (TD-WB-000) with special impeller (TD-WB-001) for 2–3 s to create the decontamination foam. 2.4. Liquid agent exposure Each CWA was applied to coupons as 5 separate 200-nL neat drops evenly spaced over the surface using a 2.5-L pipette. The only exception was concrete coupons, which required that a larger amount of HD (5 separate 1.2-L droplets) be applied to the surface to be measurable after surface reactivity occurred. Contaminated coupons were then placed on a vertical stainless-steel rack within a polished stainless-steel chamber to ensure realistic contact time of CWA with the various coupon surfaces before decontamination was initiated. These times were chosen based on the observed penetrating ability and persistence of agents on the surfaces. Clean air flowing into the chamber at 22 ◦ C and 11% relative humidity was controlled using a mass flow controller (150 mL/min) that resulted in an air exchange rate in the chamber of approximately 5 air volumes per hour. Vapor exiting the chamber was bubbled through 5% bleach (sodium hypochlorite) to decontaminate any CWA in the vapor exhaust. At various times for up to 1 week, the chamber was opened, and coupons were removed for decontamination testing. Liquid GB was allowed to remain on the concrete surface for 1 h and on latex-painted wallboard and vinyl floor tile surfaces for 24 h. No quantifiable GB was found on glass or stainless steel after 1 h following droplet deposition. Liquid HD was allowed to remain on the surface of glass, stainless steel, and concrete for 2 h, and it remained on the surface of latex-painted wallboard and vinyl floor tile for 1 week. Liquid GD was allowed to remain on the surface of glass for 30 min, concrete for 2 h, and vinyl tile and wallboard for 6 h. Liquid VX was allowed to remain on the surface of glass, stainless steel, and concrete for 24 h. VX remained on the surface of latex-painted wallboard and vinyl floor tile for 1 week. Given the different quantity of agent applied, persistence, volatility, and aging time, the beginning concentration of agent varies with each experiment. 2.5. Application of decontamination technologies For each combination of CWA and coupon surface, and at each predetermined elapsed time, 3 contaminated coupons were extracted without decontamination for analysis to determine recoverable contamination levels after the allotted CWA contact times. In all, a total of 60 no-treatment controls were evaluated (i.e., 4 CWA types tested × 5 coupon types × 3 coupons for each combination). The remaining contaminated coupons were placed in individual 2-oz jars for application of a selected decontamination technology. Liquid decontamination technologies were applied to horizontal coupons using a spray bottle with an adjustable nozzle that disseminated liquid as a mist until the surface was completely covered. Foam decontamination technologies were applied to horizontal coupons using a spatula to spread a 2–4-mm coating over the entire surface. Immediately after application of liquid or foam technologies, coupons were stored either vertically or horizontally for the range of decontamination times evaluated. The decontamination treatments were applied in a far greater than stoichiometric amount, since this would be consistent with the approach taken in decontaminating a large facility. Three coupons were extracted a few minutes after the initial application and then at 1, 8, and 24 h of decontamination contact time. Data points in the Results section represent the mean value of three replications of coupons tested in both horizontal and vertical orientation. Error bars represent 1 standard deviation of the mean.
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2.6. Solvent extraction Chemical agent was extracted from coupons using 10 mL of dichloromethane:acetone (1:1) and 15 min of sonication. Vinyl floor tile and latex-painted wallboard required a second extraction. Materials that deteriorated during extraction were filtered before analysis. No losses were observed in tests of CWA filtration in the solvent mixture using a 1-m syringe filter. Any liquid or foam that did not remain on the coupon was extracted using 10 mL of dichloromethane:acetone (1:1) and 15 min of sonication. 2.7. Analytical methods Each 5-L sample of solvent extract was analyzed by GC–MS using a Hewlett Packard (HP) 6890 GC coupled to a HP5973 mass selective detector with the injection temperature set at 250 ◦ C. Chromatographic separation was achieved using a DB-5MS column (30-m × 0.25-mm i.d., 0.25-m film thickness), with the GC oven ramped from 40◦ to 250 ◦ C over 35.5 min. The mass spectrometer was operated in full-scan mode (30–550 m/z) to identify possible CWA degradation products that could result from chemical interactions on a coupon surface. 3. Results Data are reported as residual contamination (g or ng of CWA recovered) to link experimental outcomes to eventual health-riskbased cleanup standards for re-occupancy of civilian infrastructure. In addition, Table 1 reports the percentage of CWA removed over 24 h in the presence of decontamination reagent to provide a measure of the efficiency of each technique. None of the captured liquid or foam that dripped off coupons had a detectable concentration of CWA. No toxic degradation products of HD (sulfones) and VX (EA2192) were observed in any solvent extracts of surfaces with or without decontamination. Figs. 1 and 2 show decontamination performance against HD and VX, respectively, on horizontal nonporous or nonpermeable surfaces. GB decontamination was not evaluated on these surfaces because liquid droplets of GB applied to glass and stainless-steel surfaces resulted in persistence times of less than 1 h. Similarly, GD decontamination was not studied on stainless steel because of the short persistence of liquid droplets on that surface (below the detection limit of 1 g after 8 h of natural attenuation). The HD contamination on both glass and stainless steel before application of any decontamination was measured to be ∼700 g. Fig. 1(a) shows the result for HD on glass, where SDF, DF-200, and Decon Green all resulted in rapid and complete decontamination of HD to below detection levels (<100 ng residual contaminant). Decontamination using the bleach solution was slower, such that a measurable residual of 14-g HD remained 24 h after the bleach application. Fig. 1(b) shows the results for HD on stainless steel, where SDF and Decon Green resulted in rapid initial reactivity, similar to that on glass, whereas DF-200 reactivity appeared to be slower than that on glass and similar to the rate of the bleach solution. For stainless steel, both SDF and DF-200 reduced residual HD contamination to nondetectable levels 24 h after decontaminant application. Decon Green and bleach solution resulted in ∼6 g of HD residual remaining on a coupon surface 24 h after application of the decontaminant. The VX contamination on both glass and stainless steel before any decontamination application was measured to be ∼750 g. Fig. 2(a) shows the result for VX on glass, where DF-200 and Decon Green both resulted in rapid and complete decontamination of VX to below detection levels (<2 g residual). Decontamination using SDF was also rapid; however, residual VX (12 g) remained
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Table 1 CWA decontamination percent after 24 h of decontamination treatment on horizontal samples. Chemical agent
Decontamination method
Stainless steel (%)
Glass (%)
VX
None Bleach SDF DF-200 Decon Green
0 67 99+ 98 99
7 60 98 99+ 99+
46 85 98 85 55
0 42 50 20 93
0 50 50 75 90
HD
None Bleach SDF DF-200 Decon Green
42 99 99+ 99+ 99
46 98 99+ 99+ 99+
97 97 95 90 85
14 50 79 43 81
0 93 88 68 99+
GB
None Bleach SDF DF-200 Decon Green
–a – – – –
– – – – –
99+ 99+ 99+ 99+ 99+
0 94 94 99+ 99+
55 75 55 99+ 99+
GD
None Bleach SDF DF-200 Decon Green
– – – – –
99+ 99+ 99+ 99+ 99+
99+ 99+ 99+ 99+ 99+
60 78 85 92 98
50 98 83 95 99
a
Concrete (%)
Vinyl tile (%)
Wallboard (%)
Dash indicates the decontamination technology for the specified agent was not studied on this surface.
24 h after decontaminant application. The bleach solution showed some initial reactivity, but a residual (∼300 g) remained 24 h after its application. Fig. 2(b) shows the result for VX on stainless steel, where SDF resulted in rapid and complete decontamination of
VX to below the detection level. Decon Green and DF-200 showed similar rapid initial reactivity in the first hour, but both resulted in detectable VX residuals (3 g and 15 g for Decon Green and DF200, respectively) 24 h after decontamination application. As with
Fig. 1. HD decontamination on horizontal (a) glass and (b) stainless-steel surfaces.
Fig. 2. VX decontamination on horizontal (a) glass and (b) stainless-steel surfaces.
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Fig. 3. GB decontamination on (a) horizontal and (b) vertical latex-painted wallboard.
Fig. 4. GB decontamination on horizontal (a) floor tile and (b) concrete. Results for vertically oriented coupons (data not shown) were not statistically different from those for horizontally oriented coupons.
glass, bleach solution showed some initial reactivity, but a residual amount (∼250 g) remained 24 h after its application. Figs. 3–10 summarize the decontamination performance for GB, GD, HD, and VX on porous or permeable surfaces. In most cases, efficacy was apparent for each decontamination technology compared to no-treatment controls. Although the decontamination technologies typically demonstrated lower residual contamination levels than those on the no-treatment controls, in many cases some measurable residual contamination remained on a coupon surface even 24 h after decontaminant was applied. Fig. 3(a) shows the results for horizontal latex-painted wallboard contaminated with GB, and Fig. 3(b) shows the results for vertical latex-painted wallboard. The GB contamination on wallboard before any decontamination was ∼150 g. Both Decon Green
and DF-200 reduced the GB residual contamination on latexpainted wallboard to near or below the detection limit (<1 g), whereas SDF and the bleach solution resulted in residual contamination ranging from no difference compared to no-treatment controls (i.e., 80 g) to 20 g. The different foams, SDF and DF200, give different results on wallboard reflecting their different (aqueous vs organic, respectively) chemical make up. The performance of each treatment technology on GBcontaminated wallboard was similar for vertical and horizontal coupons, with most data points for each treatment overlapping one standard deviation of the mean for the 3 replications of the horizontal and vertical variations. The same was true of all remaining tests involving comparisons of coupons in vertical and horizontal orientations, regardless of the CWA tested or substrate evaluated.
Fig. 5. HD decontamination on (a) horizontal and (b) vertical latex-painted wallboard.
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Fig. 6. HD decontamination on horizontal (a) floor tile and (b) concrete. Results for vertically oriented coupons (data not shown) were not statistically different from those for horizontally oriented coupons.
Fig. 7. VX decontamination on (a) horizontal and (b) vertical latex-painted wallboard.
Fig. 8. VX decontamination on horizontal (a) floor tile and (b) concrete. Results for vertically oriented coupons (data not shown) were not statistically different from those for horizontally oriented coupons.
Fig. 9. GD decontamination on (a) horizontal and (b) vertical latex-painted wallboard.
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Fig. 10. GD decontamination on horizontal (a) floor tile and (b) concrete. Results for vertically oriented coupons (data not shown) were not statistically different from those for horizontally oriented coupons.
Because almost all the data points for each decontaminant technology overlapped one standard deviation of the mean for all replications of horizontal versus vertical positions, results for vertical orientation are not shown in some of the remaining figures. Fig. 4(a) shows the results for horizontal vinyl floor tile contaminated with GB, and Fig. 4(b) shows results for horizontal concrete. The GB contamination on vinyl floor tile and concrete before application of any decontaminant was measured to be ∼18 g and 100 g, respectively. Decon Green rapidly reduced the residual GB contamination in vinyl tile to near the detection limit (<0.1 g), and DF-200 reduced GB residual contamination to near the detection limit within 24 h of application. Both SDF and the bleach solution resulted in residual contamination of ∼1 g 24 h after decontamination application. Fig. 4(b) shows that although the no-treatment results indicate residual GB contamination levels are low (∼1 g) on concrete after 24 h, each of the treatment technologies achieved <1 g between the initial application and 1 h of contact time. Each of the technologies tested resulted in nondetectable (<0.1 g) residual contamination 8 h after application of decontaminant. Fig. 5(a) shows the results for horizontal latex-painted wallboard contaminated with HD, and Fig. 5(b) shows results for vertical wallboard. The HD contamination on latex-painted wallboard before any decontamination application was measured to be ∼340 g. Decon Green was the only decontamination technology able to reduce residual HD contamination on latex-painted wallboard to below detection limits (<1 g) within 24 h. SDF and bleach solution reduced the HD residual contamination to 40 g and 25 g, respectively, whereas DF-200 resulted in ∼110 g of residual contamination compared to no treatment (∼340 g). Fig. 6(a) shows the results for horizontal vinyl floor tile contaminated with HD, and Fig. 6(b) shows results for horizontal concrete. The HD contamination on vinyl floor tile and concrete before any decontamination application was measured to be ∼580 g and 400 g, respectively. None of the treatment technologies tested was highly effective at decontaminating HD on vinyl tile. Decon Green and SDF resulted in residual HD contamination of ∼125 g after 24 h of decontaminant application. DF-200 and bleach solution resulted in residual HD contamination of ∼300 g after 24 h of decontaminant application. Fig. 6(b) shows that residual CWA contamination on concrete decreased rapidly for each treatment technology evaluated. However, after 24 h of exposure to the decontamination mixtures, there was little difference between the treated and untreated control coupons, suggesting that the concrete tested has a reactivity that affects the observed degradation of HD more than the treatment technologies themselves. Fig. 7(a) shows the results for horizontal latex-painted wallboard contaminated with VX, and Fig. 7(b) shows results for vertical wallboard. The VX contamination on latex-painted wallboard
before any decontamination was measured to be ∼700 g. Decon Green was the best decontamination technology of those studied for reducing residual VX contamination on latex-painted wallboard within 24 h. Nevertheless, its application resulted in measurable VX residual contamination (∼70 g) after 24 h. Bleach solution demonstrated the next-best performance in this group of treatment technologies, but it resulted in VX residual contamination of ∼150 g after 24 h. SDF and DF-200 left residual contamination of approximately 300 g of VX on latex-painted wallboard. Fig. 8(a) shows the results for horizontal vinyl floor tile contaminated with VX, and Fig. 8(b) shows results for horizontal concrete. The VX contamination on vinyl floor tile and concrete before any decontamination was measured to be ∼700 g and 370 g, respectively. Decon Green was the best decontaminant of those evaluated for reducing residual VX contamination on vinyl tile within 24 h, but it resulted in VX residual contamination of ∼50 g after 24 h. SDF, DF-200, and bleach solution resulted in greater residual VX contamination (ranging from ∼250 to 500 g) 24 h after decontamination application. Fig. 8(b) shows that SDF best reduced the VX residual contamination level (to 6 g) on concrete. DF-200 and bleach solution reduced residual VX contamination to ∼40 g. Application of Decon Green resulted in 110 g of VX residual contamination on treated samples of concrete evaluated after 24 h. Fig. 9(a) shows the results for horizontal latex-painted wallboard contaminated with GD, and Fig. 7(b) shows results for vertical oriented wallboard. The GD contamination on latex-painted wallboard before any decontamination was measured to be ∼260 g. Decon Green and bleach were most effective at removing GD from latex-painted wallboard, leaving approximately 4 g and 6 g of residual, respectively, after 24 h. DF-200 resulted in 12 g of residual, and SDF left 46 g, less than the no-treatment final amount of 125 g. Fig. 10(a) shows the results for horizontal vinyl floor tile contaminated with GD, and Fig. 10(b) shows results for horizontal concrete. The GD contamination on vinyl floor tile and concrete before any decontamination was measured to be ∼270 g and 120 g, respectively. All decontamination reagents and natural attenuation were effective at removing nearly all GD from concrete after 8 h, and the amount of residual agent was below the detection limit after 24 h. GD did persist on vinyl tile in the presence of decontamination reagents. Decon Green was the best decontaminant of those evaluated leaving ∼1 g of GD after 24 h. DF200 was almost as effective, leaving 6 g. Bleach left 16 g and SDF left 11 g of GD, while 24 h of natural attenuation left 28 g. 4. Discussion The results of our efficacy testing reinforce the difficulty associated with developing a universal decontamination technology
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for cleanup of a CWA in a civilian setting. No perfect surface decontamination technology exists, largely because the various decontamination characteristics required to address all potential applications are likely incompatible. Except for bleach, all the decontaminants evaluated demonstrated excellent performance on nonporous and nonpermeable glass and stainless steel, which are surfaces that are the typical design basis of those decontamination technologies. Such surfaces do not pose transport challenges for decontamination because CWA remains a free liquid on the surface. Porous and permeable surfaces pose a more complex, coupled transport and reactivity challenge for decontamination. The leading decontamination technology candidates we tested for CWAs often left substantial residual contamination on porous and permeable surfaces even after 24 h of contact time. Performance appeared to become progressively less effective as the persistence characteristics of the tested CWAs increased. Solventbased Decon Green generally resulted in excellent performance on permeable polymers (latex-painted wallboard and vinyl floor tile) on which it could better penetrate hydrophobic surfaces compared with aqueous-based liquids and foams, whereas aqueous-based liquids and foams appeared to out-perform solvent-based Decon Green in penetrating pores of the polar surface of concrete. Material compatibility also varies as a function of the type of formulation used in a given decontaminant. Aqueous-based products raise potential corrosion concerns, whereas solvent-based decontamination technologies are designed to avoid corrosion while penetrating into polymers and plastics. However, such penetration raises different materials-compatibility issues in that solventbased systems have the potential to irreversibly alter polymer and plastic structures through swelling or weakening, eliminating their potential reuse following decontamination. In effect, the very characteristics that make a given decontamination technology most effective can also result in destructive processes that necessitate disposal of certain decontaminated materials. 5. Conclusions The test results reported here, together with work being performed by the EPA, represent important steps toward improving our understanding of the efficacy of various decontamination strategies for the wide range of materials and surfaces relevant to the restoration of civilian infrastructure. Such understanding requires (1) identifying those materials that are easily decontaminated by the technologies available; (2) identifying those materials that should be removed because decontamination is impractical; (3) determining the most appropriate decontamination approach to apply in light of the contamination scenario and agent–material combinations; and (4) assessing all requirements related to waste disposal. The potential interactions among a given CWA, substrate surface, and decontaminant are complex. Consideration of all potential interactions increases the likelihood that an efficient cleanup strategy will employ a range of decontamination technologies according to site- and incident-specific conditions. In situations where contaminated materials are destined for disposal, the decontamination criteria are much easier to satisfy and more options are available compared to situations in which materials are destined for reuse. Decontamination for reuse mandates not only that residual contamination be reduced to a value equal to or less than an established or recommended health guideline (clearance goal) but also that
the material being decontaminated not be damaged or destroyed, which can be highly problematic for the numerous surface types prevalent in typical civilian infrastructure. Where decontamination is impractical and disposal is unacceptable, at present only natural attenuation, or methods that enhance natural attenuation such as ventilation at elevated temperatures or relative humidity, remain as the most feasible options. Clearly, additional development is needed of decontamination technologies that will maximize material reuse in an effort to facilitate efficient facility restoration and minimize waste generation after a CWA release in a civilian setting. Acknowledgements This work was funded by the U.S. Department of Homeland Security Science & Technology Directorate, Chemical and Biological Research & Development Branch, as part of a larger Facility Restoration Operational Technology Demonstration Project. This work was performed under the auspices of the U.S. Department of Energy by Lawrence Livermore National Laboratory under contract DE-AC5207NA27344. Thanks to Brian Viani for developing the methodology to make and artificially age concrete coupons. We also acknowledge the questions and valuable feedback from other project team members at Sandia National Laboratory, Oak Ridge National Laboratory, and Pacific Northwest National Laboratory. References [1] Department of the Army, Textbook of Military Medicine. Part I. Medical Aspects of Chemical and Biological Warfare, Office of the Surgeon General, Department of the Army, 1997. [2] A. Richardt, M.-M. Blum (Eds.), Decontamination of Warfare Agents, Wiley-VCH Verlag Gmbh & Co., KGaA, Weinheim, Germany, 2008. [3] E. Raber, T. Carlsen, R. Kirvel, Remediation following chemical and biological attacks, in: S.M. Maurer (Ed.), WMD Terrorism: Science and Policy Choices, MIT Press, Cambridge MA, 2009. [4] A. Watson, L. Hall, E. Raber, V.D. Hauschild, F. Dolislager, A.H. Love, M.L. Hanna, Developing health-based pre-planning clearance goals for airport remediation following a chemical terrorist attack: introduction and key assessment considerations, J. Hum. Ecol. Risk Assess. 17 (1) (2011). [5] A. Watson, F. Dolislager, L. Hall, E. Raber, V.D. Hauschild, A.H. Love, Developing health-based pre-planning clearance goals for airport remediation following a chemical terrorist attack: decision criteria for multipathway exposure routes, J. Hum. Ecol. Risk Assess. 17 (1) (2011). [6] Environmental Protection Agency, Decontamination of toxic industrial chemicals and chemical warfare agents on building materials using chlorine dioxide fumigant and liquid oxidant technologies, EPA report 600-R-09-01, 2009. [7] Environmental Protection Agency, Assessment of fumigants for decontamination of surfaces contaminated with chemical warfare agents, EPA report 600-R-10-035, 2010. [8] Environmental Protection Agency, Compilation of available data on building decontamination alternatives, Office of Research and Development, National Homeland Security Research Center, Cincinnati, EPA report/600/R-05/036, 2005. [9] C.M. Reynolds, D.B. Ringelberg, L.B. Perry, Efficacy of DECON Green Against VX Nerve and HD Mustard Simulants at Subfreezing Temperatures, U.S. Army Engineer Research and Development Center, ERDC/CCREL TR-0614, 2006. [10] B. Singer, A. Hodgson, H. Destaillats, T. Hotchi, K. Revzan, R. Sextro, Indoor sorption of surrogates for sarin and related nerve agents, Environ. Sci. Technol. 39 (2005) 3203–3214. [11] G.W. Wagner, L.R. Procell, D.C. Sorrick, G.E. Lawson, C.M. Wells, C.M. Reynolds, D.B. Ringelberg, K.L. Foley, G. Lumetta, D.L. Blanchard Jr., All-weather hydrogen peroxide-based decontamination of CBRN contaminants, Ind. Eng. Chem. Res. 49 (7) (2010) 3099–3105. [12] E. Raber, R.R. McGuire, Oxidative decontamination of chemical and biological warfare agents using L-Gel, J. Hazard. Mater. B93 (2002) 339. [13] G.S. Groenewold, A.D. Appelhans, G.L. Gresham, J.E. Olson, M. Jeffrey, M. Weibel, Characterization of VX on concrete using ion trap secondary ionization mass spectrometry, J. Am. Soc. Mass Spectrom. 11 (1) (2000) 69–77.
Journal of Hazardous Materials 196 (2011) 123–130
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Treatment of oilfield wastewater in moving bed biofilm reactors using a novel suspended ceramic biocarrier Zhiyong Dong a,∗ , Mang Lu b , Wenhui Huang c , Xiaochun Xu d a
State Key Laboratory of Heavy Oil Processing, China University of Petroleum, Beijing 102249, China School of Materials Science and Engineering, Jingdezhen Ceramic Institute, Jingdezhen 333001, Jiangxi Province, China School of Energy Resources, China University of Geosciences, Beijing 100083, China d School of Geosciences and Resources, China University of Geosciences, Beijing 100083, China b c
a r t i c l e
i n f o
Article history: Received 29 April 2011 Received in revised form 31 August 2011 Accepted 1 September 2011 Available online 6 September 2011 Keywords: Produced water Sepiolite Fourier transform ion cyclotron resonance mass spectrometry Biodegradation degree
a b s t r a c t In this study, a novel suspended ceramic carrier was prepared, which has high strength, optimum density (close to water), and high porosity. Two different carriers, unmodified and sepiolite-modified suspended ceramic carriers were used to feed two moving bed biofilm reactors (MBBRs) with a filling fraction of 50% to treat oilfield produced water. The hydraulic retention time (HRT) was varied from 36 to 10 h. The results, during a monitoring period of 190 days, showed that removal efficiency of chemical oxygen demand was the highest in reactor 3 filled with the sepiolite-modified carriers, followed by reactor 2 filled with the unmodified carriers, with the lowest in reactor 1 (activated sludge reactor), at an HRT of 10 h. Similar trends were found in the removal efficiencies of ammonia nitrogen and polycyclic aromatic hydrocarbons. Reactor 3 was more shock resistant than reactors 2 and 1. The results indicate that the suspended ceramic carrier is an excellent MBBR carrier. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Oilfield produced water (OPW), designated the wastewater generated after separation from crude oil during the primary separation process, accounts for the majority of the waste derived from oilfield development. Since OPW contains high concentrations of petroleum hydrocarbons, oilfield chemicals, different salts, suspended solids and heavy metals, it can cause considerable environmental impacts if discharged without effective treatment [1]. Nowadays, the petroleum industry faces a huge challenge in meeting increasingly stringent environmental standards. During the past decades, various technologies such as membrane filtration [2], reverse osmosis [3,4], electrochemical oxidation [5], land disposal [6], and biological treatment [7,8], have been developed for treating OPW. Physical and chemical technologies cannot remove small suspended oil particles and dissolved elements. Besides, chemical treatments may produce hazardous sludge. Compared with physical and chemical processes, biological treatment of oily wastewater can be a cost-effective and environmental friendly method, and more compatible with existing plant facilities and operation [9]. As a prevalent biological wastewater treatment process, however, activated sludge process is unsatisfactory to remediate OPW because of fila-
∗ Corresponding author. Tel.: +86 10 89733070. E-mail address:
[email protected] (Z. Dong). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.001
mentous bulking and foaming, high suspended solid content, and the presence of complex components in the wastewater [10]. Biofilm processes have been successfully used in wastewater treatment for over a century, and have proved to be reliable for contaminant removal without some of the problems encountered in activated sludge processes [11]. The moving bed biofilm reactor (MBBR) process has been employed successfully in treatment of drinking water, municipal and various types of industrial wastewater [12]. The MBBR has been devised to offer the advantages of the biofilm system (compact, stable removal efficiency and simplicity of operation) without its shortcomings (medium channeling and clogging) [13]. The core of MBBR is the supporting media for microbial adhesion. The properties of biocarrier can directly influence the ability to form biofilms, the quantity of biomass and the efficiency of treatment. A wide range of biofilm carriers, including polyethylene, polyurethane (PU), granular activated carbon, sand and diatomaceous earth, have been used in MBBR systems [14]. In general, polymer carriers have low density and excellent processability, and expansion can be obtained easily as the water circulates. However, the poor hydrophilicity and biocompatibility of plastic carriers often lead to some deficiencies in the rate and amount of biofilm culturing, and the adhesion extent of biofilm. In addition, biomass buildup and reactor head-loss may occur due to rapid clogging and jamming of the plastic carriers below the upper grid, inducing the upper retention grid to break [15]. Inorganic carriers such as limestone, zeolite, activated carbon, graystone, slag, and
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Table 1 Characteristics of the oilfield produced water. pH COD, mg/L BOD5 , mg/L TOC, mg/L TPH, mg/L NH3 -N, mg/L Total nitrogen, mg/L Total phosphorus, mg/L Fe2+ /3+ , mg/L Mg2+ , mg/L Ca2+ , mg/L S2− , mg/L SO4 2− , mg/L Total dissolved solids, mg/L
6.4–6.7 343–365 82–95 82–89 24–28 41–48 68–71 0.19–0.22 9.5–9.8 25–26 165–167 1.5–1.6 36–38 4950–5230
coke have good mechanical strength and biocompatibility. However, relatively high densities of inorganic carriers can increase the energy requirement for expansion. Many efforts have been devoted to the development and application of novel biocarriers. Chen et al. [16] used a tube chip type of polymer mixed with nano-size inorganic materials as biocarrier to treat pesticide wastewater by integration of MBBR and Fentoncoagulation pretreatment. The results showed that the MBBR had excellent advantages such as flexibility, easy operation and strong resistance against loading impact. Delnavaz et al. [13] utilized light expanded clay aggregate (density 0.55 g/cm3 ) as carrier for a toxic and hard biodegradable aniline removal, and up to maximum of 90% removal efficiencies were obtained for COD of 2000 mg/L after 3 days of treatment. The PU foam and biodegradable polymer polycaprolactone (PCL) particles were applied by Chu and Wang [15] to treating wastewater with a low C/N ratio using MBBRs. The results exhibited that total organic carbon and ammonium removal efficiencies were 90% and 65%, respectively, in the reactor filled with PU carriers, compared with 72% and 56% in the reactor filled with PCL carriers. With the advance of technology, searching for novel, low-cost, long-life, and easily attached-growth biomass carriers becomes a research subject. In the present study, a novel suspended ceramic biocarrier was developed and employed as MBBR carrier to treat OPW. The performance of the carriers was investigated, and the characteristics of pollutant removal were evaluated. 2. Materials and methods 2.1. Chemicals 2,3,5-Triphenyltetrazolium chloride (TTC) was purchased from Sigma (St. Louis, MO, USA). All other chemicals and solvents used were of analytical or glass-distilled grade and were obtained from commercial supplier (Beijing Chemical Reagent Company, China). 2.2. Wastewater The OPW was collected on 7 April 2010, from a water injection pipe of Henan Oilfield, Central China. The wastewater was allowed to settle for 24 h, and the floating crude oil was discarded. The wastewater was stored at 4 ◦ C until required. Characteristics of the OPW sample are listed in Table 1. 2.3. Preparation and modification of suspended ceramic biocarriers Fly ash, collected from Jingdezhen Power Plant in Jiangxi Province in China, was used as raw materials. The fly ash was crushed in a mortar, dehydrated at 120 ◦ C for 2 h and then heated
at 900 ◦ C for 6 h until the fly ash had constant weight to remove remaining carbon and other decomposable components. Preparation of suspended ceramic body was performed according to Sepulveda et al. [17–20]. Briefly, fly ash powder (16 g) was mixed with deionized water (16 mL), the monomer (acrylamide, 5 g), the cross-linker (N,N -methylene bisacrylamide, 3 g), and the dispersant (2 drops) to produce a slurry. The mixture was then vigorously stirred to produce foam with the assistance of Triton X100 (0.1 mL). Polymerization was triggered by the addition of the initiator (0.52 g/mL ammonium persulfate solution, 2 mL) and the catalyst (tetramethylethylene diamine, 3 mL). The foam was immediately cast into moulds prior to the gelation. The sample was dried at 150 ◦ C, and then sintered in a gas furnace at 1250 ◦ C for 2 h with heating rate of 5 ◦ C/min. In order to obtain higher porosity, the prepared suspended ceramic body was modified with sepiolite. Raw sepiolite powder was received from Leping, Jiangxi Province, China. Sepiolite samples were treated before using in the experiments as follows [21]: the raw sepiolite was sieved through a 20-mesh sieve to remove the impurity; then mixed with deionized water in proportion of 1:20 and stirred at 2000 rpm for 24 h; the suspension was centrifuged, and the supernatant was discarded and the clay particles on the top layer of the sediment was scraped off by using spatula; the solid sample was dried at 105 ◦ C for 24 h, ground then sieved by 50 m sieve; the particles below 50 m were used in further experiments; the clay particles were then suspended in deionized water and acidified with 1 M HCl to pH 3.0 to remove the carbonates and then centrifuged, followed by calcination at 300 ◦ C for 4 h. The ceramic foam body and refined sepiolite particles were mixed with distilled water, and then subjected to ultrasonic agitation (480 W; UR1, Retsch GmbH, Carl Stuart Limited, Germany) for 2 min. Afterwards, the mixture was dried at 120 ◦ C and then mixed with the organic binder (5% aqueous PVA-1750 solution). After drying at 80 ◦ C for 12 h, the granules were sintered between 550 and 650 ◦ C in intervals of 50 ◦ C for 2 h. The Brunauer–Emmett–Teller (BET) specific surface area of biocarrier was determined with ASAP 2020 M + C (Micromeritics, America). Flexure strength was determined by the three-point bending method. The bulk density was computed from the weight-to-volume ratio. 2.4. Set-up and operation of reactors Three identical reactors made of pexiglas with a useful volume of 5-L were continuously fed. The first reactor (R1) contained only activated sludge taken from a wastewater treatment plant at Yanshan Petrochemical Co., Ltd., Beijing. The second reactor (R2) was filled with the suspended ceramic granules as biomedia. In the third reactor (R3) ceramic granules modified with sepiolite was added. The amount of carrier corresponded to a volume fraction of 50%, for both reactors. Fig. 1 shows a schematic representation of one MBBR system and the suspended ceramic biocarriers. The activated sludge reactor has the same configuration as shown in Fig. 1; the only difference is that there is no carrier in the activated sludge reactor. The operation of the three reactors was started by inoculating 0.5 L of fresh activated sludge. After inoculation, OPW was introduced into the reactors to obtain a working volume of 5 L. Air was diffused in order to supply oxygen to the biomass and to mix the carriers. After aeration for 12 h, one-half of the mixed liquid was drained off and then the same volume OPW was introduced for the next 12 h of the operation. This procedure was repeated during the first week of the startup. After that, the feed mode of wastewater was changed into continuous flow mode during a period of 183 days. All laboratory experiments were conducted at room temperature and dissolved oxygen (DO) concentration was always controlled above 3 mg/L during the operation by
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Fig. 1. Schematic diagram of the aerobic MBBR system (above) and photos of suspended ceramic biocarriers (below).
regulating the aeration rate. The variations in operational parameters are listed in Table 2. As shown in Table 1, the content of N in the OPW was sufficient to support the COD removal. However, the amount of phosphorus was far below the conventional treatment standard of 100:5:1. Therefore, K2 HPO4 used as a phosphorus source was added to the wastewater based on a COD:P ratio of 100:1. 2.5. Analytical methods DO, pH, and temperature were monitored in situ daily using portable devices during the experiments. Water samples were centrifuged at 8000 rpm for 5 min, and then analyzed according to standard methods [22]. Briefly, COD was determined by titrimetric method after dichromate closed reflux. BOD5 was measured via the oxygen consumption of bacteria breaking down organic matter in the sample over a 5-day period under standardized conditions. NH3 -N was determined by Nessler’s reagent colorimetry. Water samples that could not be analyzed immediately were frozen at −20 ◦ C prior to analysis. Biofilm thickness was measured in 10 different parts of a carrier and then averaged, using a dissecting microscope (Nikon, Japan) equipped with an ocular micrometer. Table 2 Operational conditions during the experiment.
Morphology of carriers was observed with a field emission scanning electronic microscope (SEM) (JSM-6700 F, JEOL, Japan). 2.6. Biochemical analyses Influent and effluent samples were filtered through 0.22 m membrane filter to obtain wastewater filtrate. Aliquots (20 L) of the bacterial reagent (Photobacterium phosphoreum) were added to a series of dilutions of filtrates. The salinity of the samples was adjusted to 2% with reagent-grade NaCl. The luminescence of bacteria was measured after 15 min of exposure at 15 ◦ C, using a SDI M500 analyzer (SDI, USA). Microtoxicity values were the average of five replicates of each filtrate sample, expressed as EC50 , which was defined as the effective nominal concentration of elutriate (volume percent) that reduces the intensity of light emission by 50%. The activity of microorganisms attached to the biomedia was measured using the TTC dehydrogenase test adopting the method described elsewhere [23]. In brief, one entire ceramic carrier was taken from the reactors at regular intervals and immersed into the reagents. The mixture was incubated and microbial activity was measured. Results were expressed as g triphenylformazan (TPF)/g, based on the mass of the carrier after sintering (550 ◦ C, 24 h). Each measurement was made in triplicate, and the average of three independent measurements was presented. 2.7. Instrumental analyses
Phase
Operational days
HRT (h)
I II III IV
1–7 8–95 96–155 156–190
0 36 18 10
Analyses of saturated and aromatic fractions were carried out using a simplified sample clean-up and a gas chromatography–mass spectrometric (GC–MS) system. Briefly, wastewater samples were extracted three times by liquid–liquid
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Table 3 Characteristics of synthesized suspended ceramic biocarriers with or without sepiolite modification. Item Bulk density, g/cm3 Flexure strength, MPa BET specific surface area, m2 /g
Original 0.92–0.96 2.5–3.0 3.8–4.1
Modified 0.93–0.98 2.5–3.0 5.6–5.9
technique with dichloromethane for pH 2, 7, and 12. The three extract layers were combined and condensed to 1 mL in a rotary evaporator, and then fractionated by silica-gel column chromatography to separate saturate and aromatic fractions. The fractions were analyzed by an Agilent7890-5975c GC–MS equipped with an Agilent HP-5MS fused silica capillary column (60 m × 0.25 mm × 0.25 m). Concentrations were determined for the following 16 priority polycyclic aromatic hydrocarbons (PAHs): naphthalene; acenaphthene; acenaphthylene; fluorene; phenanthrene; anthracene; fluoranthrene; pyrene; benz[a]anthracene; chrysene; benzo[b]fluoranthene; benzo[k]fluoranthene; benzo[a]pyrene; indeno[1,2,3-cd]pyrene; dibenz[ah]anthracene; and benzo[ghi]perylene. An aliquot of the above-obtained extract without fractionation was examined by electrospray ionization Fourier transform ion cyclotron resonance mass spectrometry (ESI FT-ICR MS). Additional details have been published elsewhere [24]. 2.8. Statistical analysis The data are presented in terms of arithmetic averages of three replicates values ± standard deviation. Statistical analysis (arithmetic average and standard deviation) was made using SPSS (Statistical Product and Service Solution) for Windows. The comparisons between treatments were done using the one-way analysis of variance (ANOVA). The statistical significance in this analysis was defined at p < 0.05. 3. Results and discussion 3.1. Characteristics of suspended ceramic biocarriers Biocarriers for MBBR should be provided with large surface area, a certain strength, no blocking and congregation, and good expansion during the operation. The photos of synthesized ceramic biocarrier are shown in Fig. 1. The ceramic granule has a diameter of around 1.25 cm. The bulk density is slightly less than that of water (Table 3). The density of the ceramic granule would be close to that of water after biofilm growth, therefore only gentle agitation or aeration was needed to produce good carrier expansion. As shown in Table 3, the three-point flexure strength of the ceramic granule did not change after modification; however the BET specific surface area was increased significantly. Sepiolite, which has a Si12 Mg8 O30 (OH)4 (OH2 )4 ·8H2 O unit-cell formula, is a magnesium hydrosilicate with a micro-fibrous structure and has a theoretical high surface area (more than 200 m2 /g) and high chemical and mechanical stability [25]. In the region of high temperature (>800 ◦ C), however, the collapse of micro-fibrous structure of sepiolite will occur, and micropores will be destroyed [26]. Therefore the sintering temperature was controlled below 650 ◦ C during the modification process. The mechanical characteristics of the carriers are important in the MBBRs. The three-point flexure strength of the carriers was kept at 2.5–3.0 MPa after 190 days of reactor operation, indicating a good durability of the carriers. In addition, the average weight of the carriers after calcination (500 ◦ C in air) only decreased by
Fig. 2. Surface changes of biocarriers during the operation of R3 (filled with sepiolite-modified ceramic granules): (a) clear carrier; (b) carrier cultivated for 2 days; (c) carrier cultivated for 7 days; and (d) photo of a carrier at the end of the experiment (190 days).
approximately 5% after 190 days of operation, indicating a good abrasion resistance of the carriers. 3.2. Biomass and biofilm characteristics The pH value of the feeding OPW was always kept at 6.4–6.7 and no further pH adjustment was used for the system. During the operation days, the pH stayed at the range 6.1–6.5 for the three reactors. The sludge retention time in the three reactors was maintained at 10 days by sludge withdrawal from the reactors once every day.
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HRT = 36 h
HRT = 18 h
HRT = 10 h
a
127
HRT = 36 h
HRT = 18 h
HRT = 10 h
100
140
R2 R3
80
100
COD removal (%)
Biofilm thickness (µm)
120
80 60 40
60
R1 R2 R3
40
20
20 0
0
0
20
40
60
80
100
120
140
160
180
200
0
20
40
60
Time (days)
To visualize the presence of biofilm on the surface of support moving in the reactor, SEM images of clean and biofilm-containing carriers were prepared. The surface of clean biocarrier (Fig. 2a) was bare. On day 2 only a part of the carrier was covered with biofilms (Fig. 2b). One-week operation of the reactor led to the formation of biofilms with a thickness larger than 45 m (Fig. 2c). After 190 days of cultivation, the structure of the biofilm was rather dense (Fig. 2d). Thickness of biofilm attached on the carrier was measured on the 2nd, 4th, 7th, 10th days and every 5 days afterwards. Fig. 3 shows that the growth curves of biofilms in R2 and R3 varied with time. Biofilms in R2 and R3 vigorously grew at a transient condition from 4th to 10th and 20th days, respectively. Under a steady-state condition, biofilms reached a maximum growth rate and biofilm thickness in R2 was larger than that in R3. This is because that the sepiolite-modified media has a larger specific surface area than the unmodified ones. Under the same conditions (i.e. suspended biomass amount, organic loading), the amount of attached biomass was identical between R2 and R3. Larger the specific surface area, thinner was the biofilm. However, the time required to reach steady state was significantly reduced due to the modification of biocarrier with sepiolite (10 and 20 days for R2 and R3, respectively), which can be attributed to the powerful adsorbent properties of sepiolite. Hrenovic et al. [27] demonstrated that sepiolite was a good carrier of immobilized and metabolically active phosphate-accumulating bacteria, and the number of microorganisms immobilized onto sepiolite was higher in purified form than in original form. It was also observed that biofilm thickness on media increased as the hydraulic retention time (HRT) was decreased from 36 to 18 h. When the HRT was further reduced to 10 h, biofilm thickness on sepiolite-modified media remained the same, whereas significantly decreased in the case of unmodified media. The decrease in biofilm thickness on media at higher HRT may be attributed to the lack of sufficient substrate and DO in inner layers and dominance of anaerobic conditions in the film, which led to the gradual sloughing of biofilm [28]. A steady state biofilm thickness is dependent on the comprehensive effects of biofilm growth rate, mechanical shearing, and subsequent detachment [29]. 3.3. Performance of treatment systems Considering that COD and NH3 -N are the two leading monitoring indices of emission of petrochemical wastewater, these two
100
120
140
160
180
200
Time (days)
b
Ammonia nitrogen removal (%)
Fig. 3. Grow curves of biomass. R2: the MBBR filled with unmodified ceramic media and R3: the MBBR filled with sepiolite-modified ceramic media.
80
HRT = 36 h
HRT = 18 h
HRT = 10 h
100
80
60
R1 R2 R3
40
20
0 0
20
40
60
80
100
120
140
160
180
200
Time (days) Fig. 4. COD (a) and NH3-N (b) removal efficiency variations at HRTs of 36, 18, and 10 h. R1: the activated sludge reactor; R2: the MBBR filled with unmodified ceramic media; and R3: the MBBR filled with sepiolite-modified ceramic media.
parameters were monitored every two days during the treatment process. 3.3.1. COD removal According to HRT variations from 36 to 10 h, organic loading was increased from 1.17 to 4.21 kg COD/m3 d. The average COD removal efficiencies in the three reactors during the operating period are shown in Fig. 4a. The results showed that at an HRT of 36 h, the average efficiencies for COD removal were consistently higher than 73% for the three reactors. It was also interesting to note that there was no significant difference (p < 0.05) in COD removal among the three processes at 36 h HRT. When the HRT decreased from 36 to 18 h, however, the COD removal efficiency decreased significantly from average 75.4% to 61.5% for R1 treatment. After a period of adaptation, the COD removal efficiency recovered to previous levels for both R2 and R3 treatment when the HRT was reduced to 18 h, indicating that the MBBRs had a strong capacity to resist shock loading caused by the change in influent flow rate. When the HRT was further reduced from 18 to 10 h, the COD removal efficiencies decreased from an average value of about 62% to 47% for R1, from 77% to 63% for R2, and from 79% to 74% for R3, respectively. During the experiments, R3 exhibited the advantage over R2 at 10 h HRT. The sepiolite-modified ceramic carrier has a higher specific surface area than the unmodified ones (Table 3). At lower
Z. Dong et al. / Journal of Hazardous Materials 196 (2011) 123–130
substrate loadings, sepiolite-modified surfaces were not fully utilized by microorganisms as suggested above that biofilm thickness in R3 was less than in R2. This also indicates that the increasing extent in biofilm thickness was greater for sepiolite-modified media than for unmodified ones with the increase of substrate loading. The specific surface is an important factor influencing the performance of MBBRs. Andreottola et al. [30] found that MBBR and activated sludge pilot plants showed nearly comparable performances in COD removal during the treatment of municipal wastewater, which was attributed to the insufficient specific surface (160 m2 /m3 ) of the MBBR plastic media used to overcome the activated sludge system in performance. In recent years, a considerable number of studies have been published that purport to treat oilfield wastewater using microorganisms. Freire et al. [31] studied COD removal efficiency of acclimated sewage sludge in a sequencing batch reactor (SBR) with different percentages of produced water and sewage. In 45% and 35% (v/v) mixtures of wastewater, COD removal efficiencies varied from 30% to 50%. Baldoni-Andrey et al. [32] evaluated the feasibility of the biotreatment of saline produced water in SBR. Results showed that the biodegradability of the produced water can be obtained, though the salinity was up to 200 g/L. However, a loss of biomass occurred during continuous SBR operation. Pendashteh et al. [33] investigated the biological pretreatment of oilfield produced water in SBR. It was found that 90% COD removal was obtained at salt concentration of 35,000 mg/L and at an organic load rate of 1.8 kg COD/m3 d. Lu and Wei [24] used the combined process of chemical oxidation and biological degradation to treat oilfield wastewater containing hydrolyzed polyacrylamide (HPAM) in a batch activated sludge reactor. The results showed that under the optimum conditions, the total removal efficiencies of HPAM, total petroleum hydrocarbons, and COD were 96%, 97% and 92%, respectively. 3.3.2. NH3 -N removal The time course of NH3 -N variation is plotted in Fig. 4b. After the lapse of 4 days, the ammonia oxidation efficiencies were 79%, 84% and 86% for R1, R2 and R3, respectively. The NH3 -N removal of the three reactors remained practically constant in the first two runs. However, NH3 -N removal efficiencies dropped when the HRT was reduced to 10 h. The evidence of nitrification process was confirmed through the determination of effluent nitrate content (data not shown). For phase IV, nitrate concentrations in the reactor effluent decreased significantly, revealing the nitrification collapse for the highest range of applied organic loadings. During wastewater treatment processes NH3 -N can be removed in two ways: (a) assimilation into biomass; (b) ammonia volatilization; and (3) biological nitrification under aerobic conditions and denitrification process under depleted oxygen levels or anoxic conditions [34].
500
R2 R3
400
DHA (µg TPF/g)
128
300
200
100
0 60 60
0 8
120 120
180 180
Time (days) Fig. 5. Time course of changes in dehydrogenase activity of biofilm. R2: the MBBR filled with unmodified ceramic media; and R3: the MBBR filled with sepiolitemodified ceramic media. Table 4 Biodegradation extents (%) of saturated hydrocarbons in the produced water as determined on the 180th day at 10 h HRT. R1: the activated sludge reactor; R2: the MBBR filled with unmodified ceramic media; and R3: the MBBR filled with sepiolite-modified ceramic media.
R1 R2 R3
Linear
Branched
Cyclic
58.4 ± 2.9 71.3 ± 3.1 82.6 ± 2.7
34.6 ± 3.5 58.2 ± 3.1 67.1 ± 2.3
15.2 ± 1.8 30.6 ± 1.7 41.2 ± 2.1
also pointed out that the biofilm metabolic activity was affected not only by substrate loading, but also by other physical and biological factors. 3.5. Microtoxicity tests Microtoxicity assessment gives refined information on the changes in wastewater quality undergoing treatment. Microtox analysis was performed on the 8th, 60th, 120th, and 180th days, to monitor toxicity levels of the bioremediated OPW and the results are shown in Fig. 6. It can be found that the acute toxicity to the 100 Influent Effluent in R1 Effluent in R2
80
Effluent in R3
3.4. Biofilm dehydrogenase activity EC50(%)
60
Dehydrogenase activity (DHA) provides a measure of overall microbial activity and consequently indicates whether stimulation or inhibition of the microbial communities is present. The DHA of attached microorganisms was measured on the 8th, 60th, 120th, and 180th days. The DHA measurements over the time period under different HRT conditions are plotted in Fig. 5. On average, the highest biofilm activity occurred on the 120th day (18 h HRT), followed by that on the 180th day (10 h HRT), with the lowest on the 8th day. In addition, it is also observed that the biofilm in R3 had slightly higher DHA than that in R2. It should be noted that the MBBRs had been operated in continuous flow mode since the 8th day. Wijeyekoon et al. [35] demonstrated that the lower loaded biofilms exhibited the highest DHA, followed by the higher loaded biofilms, with the medium loaded biofilms having the lowest activity. It was
40
20
0 0
60
120
180
Time (days) Fig. 6. Time course of changes in microtoxicity of influents and effluents. R1: the activated sludge reactor; R2: the MBBR filled with unmodified ceramic media; and R3: the MBBR filled with sepiolite-modified ceramic media.
Z. Dong et al. / Journal of Hazardous Materials 196 (2011) 123–130
129
Table 5 Concentrations of the 16 priority PAHs in influent and effluent as determined on the 180th day at 10 h HRT. R1: the activated sludge reactor; R2: the MBBR filled with unmodified ceramic media; and R3: the MBBR filled with sepiolite-modified ceramic media. Concentration (g/L) Influent Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthrene Pyrene Benz[a]anthracene Chrysene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[a]pyrene Indeno[1,2,3-cd]pyrene Dibenz[ah]anthracene Benzo[ghi]perylene
125.3 14.2 18.5 78.4 329.3 32.4 46.8 55.6 18.5 216.8 54.3 8.1 11.4 2.2 3.5 7.4
± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±
Effluent of R1 4.9 1.2 1.6 4.2 8.4 2.1 2.5 2.4 1.3 11.3 5.8 0.9 1.3 0.3 0.4 1.1
bacterium P. phosphoreum decreased after the start-up of reactors. However, as the HRT decreased from 18 to 10 h, the reductions in the toxicity of treated wastewater declined. This may be due to the less extent of biodegradation of some toxic intermediates under high HRT conditions. Overall, R3 treatment showed the greatest reductions in microtoxicity, followed by R2 treatment, with R1 treatment having the lowest reductions. In addition, the differences in microtoxicity reduction of treated wastewater between the activated sludge process and the MBBR treatment were significant on the 120th and 180th days (18 and 10 h HRT, respectively) (p < 0.05). Therefore, it is concluded that the MBBR was a quite efficient process for decontaminating the OPW. It has been reported that the attached biomass can be up to 500 times more resistant to antibacterial agents than freely suspended ones [36]. 3.6. Biodegradation of petroleum hydrocarbons GC analyses taken on days 180 showed the degradation extents of linear, branched and cycloalkanes in the OPW by the three processes (Fig. S1, Supporting Information). It can be observed from Table 4 that most of linear alkanes were removed after biotreatment, with relatively low degradation of cycloalkanes. The extent of biodegradation was significantly higher in R2 and R3 than in R1 (p < 0.05). The degradation of PAHs was confirmed by monitoring the disappearance of the 16 priority PAHs in the influent and effluent streams. Table 5 lists the concentration changes of the 16 priority PAHs during treatment. The total content of the 16 PAHs were: 1022 g/L in the influent stream; 493 g/L in the effluent stream of R1 corresponding to a degradation efficiency of 52%; 351 g/L in the effluent stream of R2 corresponding to a degradation efficiency of 65%; and 306 g/L in the effluent stream of R3 corresponding to a degradation efficiency of 70%. Naphthalene, phenanthrene and chrysene were the three major constituents, representing 12%, 32%, and 21%, respectively, of the total 16 PAHs found in the influent oil field wastewater that was the characteristic of Henan Oilfield. Of the minor constituents, indeno[1,2,3-cd]pyrene, dibenz[ah]anthracene, and benzo[ghi]perylene were also degraded with respective efficiencies of 41%, 30%, and 34% for R1, 64%, 57%, and 54% for R2, and 68%, 63%, and 58% for R3. 3.7. Analysis of biodegradation extent Since its introduction in 1974 by Comisarow and Marshall [37], FT-ICR MS has advanced to become an extraordinarily
43.9 5.6 8.3 42.6 116.5 11.9 15.3 37.1 11.6 147.3 32.8 5.4 6.5 1.3 2.4 4.9
± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±
Effluent of R2
1.7 0.4 0.7 2.8 7.4 1.1 1.0 2.0 0.9 8.3 4.5 0.6 0.6 0.2 0.1 0.4
25.6 4.3 6.9 31.5 84.3 8.5 9.6 25.7 8.5 106.5 24.9 3.6 4.8 0.8 1.5 3.4
± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±
1.8 0.5 0.8 2.7 6.5 0.9 0.8 1.6 0.6 7.4 3.6 0.4 0.6 0.1 0.2 0.4
Effluent of R3 26.4 4.6 6.7 30.6 67.4 7.6 7.5 22.5 7.3 93.3 18.6 3.5 4.3 0.7 1.3 3.1
± ± ± ± ± ± ± ± ± ± ± ± ± ± ± ±
1.6 0.6 0.8 3.1 5.3 0.9 0.9 1.4 0.7 7.8 3.3 0.4 0.5 0.1 0.1 0.3
versatile mass spectrometric technique, which affords ultrahigh mass resolving powers m/m50% > 400,000 and mass accuracy < 0.5 ppm, allowing the molecular formula assignment of 10,000+ ions in a single mass spectrum [38]. In this study, the calibrated negative-ion FT-ICR mass spectra of the oils extracted from the influent and effluents taken on days 120 are presented in Fig. S2, Supporting Information. In each spectrum, more than 8000 peaks with a signal to noise ratio of 3 were detected in the range. The broadband mass spectra clearly showed that the molecular weight distribution of polar compounds was greatly altered by biodegradation. The weight-average molecular weight was 483 Da for the influent samples, and decreased with biodegradation to 425, 403 and 392 Da for the effluent samples in R1, R2, and R3, respectively. Kim et al. [39] devised an index to estimate the extent of biodegradation in crude oil (Fig. S3, Supporting Information). This index is based on the ratio of acyclic to cyclic naphthenic acids (A/C ratio) and is calculated as follows: A/C ratio =
O2
O2
z=0
(1)
z=−2,−4,−6
The A/C ratio decreases as the extent of biodegradation increases. For O2 -containing compounds with the molecular formula Cn H2n+z O2 , n is the number of carbon atoms, z is the number of hydrogen atoms that are lost as the structures become more compact. In this study, the oils extracted from the effluents on the 180th day have an Acid Biodegradation Index of 2.6 (moderately degraded), 3.5 (moderately degraded), and 4.1 (severely degraded), for R1, R2, and R3, respectively. It can be concluded, therefore, that the MBBR treatment has a major advantage in terms of biodegradation degree of oils indicated by the Acid Biodegradation Index. In addition, the MBBR filled with sepiolite-modified ceramic biomedia exhibited a higher degree of oil biodegradation than that filled with unmodified ones. This can be attributed to unique properties of sepiolite such as high surface area, and great adsorptive capacity. However, further research is needed to reveal the concrete mechanism of sepiolite-enhanced petroleum biodegradation. 4. Conclusions This investigation demonstrated that MBBR filled with the suspended ceramic biocarrier was an effective and feasible process for removal of COD, petroleum hydrocarbons, and ammonium nitrogen from the OPW in the tested organic loading range of 1.17–4.21 kg COD/m3 d, compared to the conventional activated
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sludge treatment. The modification of ceramic biocarrier with sepiolite produced positive outcomes in the wastewater treatment efficiency. At HRT of 18 h, the concentrations of NH3 -N and COD of effluent in the MBBRs could satisfy the professional emission standard (grade one) (NH3 -N < 15 mg/L, CODCr < 100 mg/L) of petrochemical industry of PR China (GB4287-92). Acknowledgements The authors are grateful to Analysis Center, State Key Laboratory of Heavy Oil Processing, China University of Petroleum, for their technical assistance in FT-ICR MS analysis, and to National Engineering Research Center for Domestic and Building Ceramics, JCU for the assistance in the analytical measurements. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.001. References [1] D.W. Woodall, R.P. Gambrell, N.N. Rabalais, Developing a method to track oil and gas produced water discharges in estuarine systems using salinity as a conservative tracer, Mar. Pollut. Bull. 42 (2001) 1118–1127. [2] M. Ebrahimi, D. Willershausen, K.S. Ashaghi, L. Engel, L. Placido, P. Mund, P. Bolduan, P. Czermak, Investigations on the use of different ceramic membranes for efficient oil-field produced water treatment, Desalination 250 (2010) 991–996. [3] S. Mondal, S.R. Wickramasinghe, Produced water treatment by nanofiltration and reverse osmosis membranes, J. Membrane Sci. 322 (2008) 162–170. [4] M. Melo, H. Schluter, J. Ferreira, R. Magda, A. Júnior, O. de Aquino, Advanced performance evaluation of a reverse osmosis treatment for oilfield produced water aiming reuse, Desalination 250 (2010) 1016–1018. [5] A.M.Z. Ramalho, C.A. Martínez-Huitle, D.R. da Silva, Application of electrochemical technology for removing petroleum hydrocarbons from produced water using a DSA-type anode at different flow rates, Fuel 89 (2010) 531–534. [6] M. Al-Haddabi, M. Ahmed, Land disposal of treated saline oil production water: impacts on soil properties, Desalination 212 (2007) 54–61. [7] X. Zhao, Y. Wang, Z. Ye, A.G.L. Borthwick, J. Ni, Oil field wastewater treatment in biological aerated filter by immobilized microorganisms, Process Biochem. 41 (2006) 1475–1483. [8] A. Fakhru’l-Razi, A. Pendashteha, Z.Z. Abidina, L.C. Abdullaha, D.R.A. Biaka, S.S. Madaeni, Application of membrane-coupled sequencing batch reactor for oilfield produced water recycle and beneficial re-use, Bioresour. Technol. 101 (2010) 6942–6949. [9] A. Fakhru’l-Razi, A. Pendashteha, L.C. Abdullaha, D.R.A. Biaka, S.S. Madaeni, Z.Z. Abidina, Review of technologies for oil and gas produced water treatment, J. Hazard. Mater. 170 (2009) 530–551. [10] M. Lu, Z. Zhang, W. Yu, Z. Wei, Biological treatment of oilfield-produced water: a field pilot study, Int. Biodeter. Biodegr. 63 (2009) 316–321. [11] Y. Rahimi, A. Torabian, N. Mehrdadi, M. Habibi-Rezaie, H. Pezeshk, G.R. NabiBidhendi, Optimizing aeration rates for minimizing membrane fouling and its effect on sludge characteristics in a moving bed membrane bioreactor, J. Hazard. Mater. 186 (2011) 1097–1102. [12] H. Ødegaard, Innovations in wastewater treatment: the moving bed biofilm process, Water Sci. Technol. 53 (2006) 17–33. [13] M. Delnavaz, B. Ayati, H. Ganjidoust, Prediction of moving bed biofilm reactor (MBBR) performance for the treatment of aniline using artificial neural networks (ANN), J. Hazard. Mater. 179 (2010) 769–775. [14] L. Chu, J. Wang, Comparison of polyurethane foam and biodegradable polymer as carriers in moving bed biofilm reactor for treating wastewater with a low C/N ratio, Chemosphere 83 (2011) 63–68.
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Journal of Hazardous Materials 196 (2011) 131–138
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Removal of chromium from Cr(VI) polluted wastewaters by reduction with scrap iron and subsequent precipitation of resulted cations M. Gheju a,∗ , I. Balcu b a b
“Politehnica” University of Timisoara, Faculty of Industrial Chemistry and Environmental Engineering, Bd. V. Parvan Nr. 6, Et. 4, 300223, Timisoara, Romania National Institute for Research and Development in Electrochemistry and Condensed Matter, Str. Dr. Aurel Paunescu Podeanu Nr. 144, 300587, Timisoara, Romania
a r t i c l e
i n f o
Article history: Received 8 March 2011 Received in revised form 28 August 2011 Accepted 2 September 2011 Available online 12 September 2011 Keywords: Hexavalent chromium Scrap iron Packed column Horizontal clarifier Wastewater treatment
a b s t r a c t This work presents investigations on the total removal of chromium from Cr(VI) aqueous solutions by reduction with scrap iron and subsequent precipitation of the resulted cations with NaOH. The process was detrimentally affected by a compactly passivation film occurred at scrap iron surface, mainly composed of Cr(III) and Fe(III). Maximum removal efficiency of the Cr(total) and Fe(total) achieved in the clarifier under circumneutral and alkaline (pH 9.1) conditions was 98.5% and 100%, respectively. The optimum precipitation pH range which resulted from this study is 7.6–8.0. Fe(total) and Cr(total) were almost entirely removed in the clarifier as Fe(III) and Cr(III) species; however, after Cr(VI) breakthrough in column effluent, chromium was partially removed in the clarifier also as Cr(VI), by coprecipitation with cationic species. As long the column effluent was free of Cr(VI), the average Cr(total) removal efficiency of the packed column and clarifier was 10.8% and 78.8%, respectively. Our results clearly indicated that Cr(VI) contaminated wastewater can be successfully treated by combining reduction with scrap iron and chemical precipitation with NaOH. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Chromium compounds are used in a wide variety of industrial processes such as: metallurgy, chemical and refractory industries, textile dying, tanneries, metal electroplating, wood preserving, and preparation of chromate compounds. Therefore, chromium contamination has been often reported in many industrial sites, due to accidental leakages or improper disposals measures [1–4]. In aquatic environments chromium is present mainly as hexavalent and trivalent species, characterized by markedly different chemical behavior and toxicity [5]. While Cr(VI) exists mainly as highly soluble oxyanions [6], Cr(III) is less soluble and readily precipitates as Cr(OH)3 [7]. Cr(III) has a low toxicity, being considered an essential nutrient for many organisms [8]. In contrast, Cr(VI) is up to 1000fold more toxic than Cr(III) [9] and a well-established carcinogen by the inhalation route of exposure [5]. Therefore, Cr(VI) must be removed from wastewaters before their disposal to natural aquatic environments. During last two decades there has been important interest in finding new materials with high removal efficiency or/and low cost, for the removal of Cr(VI) from contaminated waters [10–23]. Reduction to Cr(III) may be considered a satisfactory solution in
∗ Corresponding author. Tel.: +40 256 488441; fax: +40 256 403060. E-mail address:
[email protected] (M. Gheju). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.002
eliminating the toxicity of Cr(VI). Scrap iron is a cheap waste material that has been successfully tested for the removal of Cr(VI) via reduction to Cr(III) according to [24]: 2Cr2 O7 2− (aq) + 6Fe0 (s) + 28H+ (aq) → 4Cr3+ (aq) + 6Fe2+ (aq) + 14H2 O (1) Subsequently, Cr(VI) may be reduced in the solution (homogeneously) by Fe(II): Cr2 O7 2− (aq) + 6Fe2+ (aq) + 14H+ (aq) → 2Cr3+ (aq) + 6Fe3+ (aq) + 7H2 O (2) The two equations can be added together to yield the net reaction for the reduction process: Cr2 O7 2− (aq) + 2Fe0 (s) + 14H+ (aq) → 2Cr3+ (aq) + 2Fe3+ (aq) + 7H2 O (3) Gould [25] reported that 1.33 mol of Fe(0) dissolved for each mol of Cr(VI) reduced. Such a high efficiency suggested that hydrogen generated during iron corrosion acts as a reducing agent for the Cr(VI) (see Eq. (4)). Recent theoretical analysis by Noubactep [26,27] supports this view. In fact, contaminants are demonstrated to be removed by adsorption and co-precipitation, while contaminant reduction, when occurs, mainly results from indirect reducing agents (Fe(II) and H/H2 ). In other words, Fe0 should be regarded as generator of reducing agents [26].
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Cr concentration (mg/L)
25
Fig. 1. Experimental setup: 1 – Cr(VI) storage tank; 2 – NaOH storage tank; 3 and 4 – peristaltic pump; 5 – glass column; 6 – scrap iron filling; 7 – overhead stirrer; 8 – horizontal clarifier; and 9 – treated water.
Cr(VI) at P1 Cr(VI) at P2 Cr(VI) at P3 Cr(III) at P1 Cr(III) at P2 Cr(III) at P3
20
15
10
5
0
Wastewater treatment systems based only on Cr(VI) reduction at pH < 6.0 cannot remove chromium from the aqueous phase because resulted Cr(III) is still soluble [28]. Since the efficiency of Cr(VI) reduction with Fe(0) is very low under circumneutral conditions, the process must be conducted at acidic pH values (2.5–3.0) [29,30]. Therefore, most of the resulted species (Cr(III), Fe(II), and Fe(III)) will remain dissolved. All these species must be removed from the wastewater in a final step, in order to complete the treatment process. To the best of our knowledge, no continuous-flow studies concerning both Cr(VI) reduction and removal of resulted chromium and iron species have been reported. As a continuation of our previous work [30–32], the present study describes the treatment of Cr(VI) polluted wastewater in continuous system, by reduction with scrap iron and subsequent precipitation of the resulted cations. This work will present data regarding the mechanism of Cr(VI) reduction inside the column, and of Cr(total) and Fe(total) removal inside the clarifier. Additionally, the optimum pH for the precipitation of cationic species resulted from the reduction process will also be established. 2. Materials and methods 2.1. Scrap iron iron spirals (5 mm < spiral diameter < 10 mm; Scrap 5 mm < spiral length < 20 mm) used in this study originated from “SPM” metals processing laboratory, at the “Politehnica” University of Timisoara. The scrap iron was washed several times with warm distilled water to assure the complete removal of all impurities, and air dried. 2.2. Background electrolyte The used background electrolyte was made up of: 50 ppm Ca2+ ; 20 ppm Mg2+ ; 128 ppm Cl− ; 104 ppm Na+ ; and 293 ppm HCO3 − . The mixture was chosen to maintain a constant ionic strength. 2.3. Experimental procedure A schematic diagram of the treatment system is shown in Fig. 1. A glass column (inner diameter: 2.5 cm, length: 70 cm) equipped with three lateral sampling ports (P1 , P2 , and P3 ) positioned at distances from the inlet end corresponding to 22.6%, 56.5%, and 100% from the total filling volume, was employed as Cr(VI) reducing reactor. The column was carefully packed with 360 g scrap iron up to a height of 62 cm. An Ismatec IP08 peristaltic pump was used to feed the Cr(VI) solution from a storage tank to the bottom end of the column. The Cr(VI) concentration (25 mg/L), the feed solution pH (2.5), and the pumping rate (1.6 L/h) were held constant throughout the study. The Cr(VI) concentration value was selected because
0
24
48
72
96
120
144
168
192
216
240
264
Time (h) Fig. 2. Cr(VI) and Cr(III) concentration in column pore water vs. time, at P1 , P2 , and P3 sampling ports.
it is within the range of relevant concentrations for electroplating wastewaters [16], while the pH was selected because it was previously reported as optimum value for Cr(VI) reduction with scrap iron in continuous system [30]. Cationic species resulted from Cr(VI) reduction were removed via precipitation with NaOH solution 5 g/L in a rectangular horizontal-flow clarifier having a working volume of 7 L. The column effluent was directed into the mixing chamber of the clarifier, where it was mixed with the NaOH solution using a Heidolph overhead stirrer, at 50 rpm. The NaOH solution was pumped by a second Ismatec IP08 pump. Samples were collected at regular time intervals from column sampling ports and clarifier effluent for pH, Cr(total), Cr(VI), Cr(III), Fe(total), Fe(II) and Fe(III) analysis. 2.4. Analytical method Chromium and iron aqueous species were detected by the 1,5diphenylcarbazide and 1,10-phenanthroline method, respectively [33], using a Jasco V 530 spectrophotometer. The pH of solutions was measured using an Inolab pH-meter, calibrated with pH 4 and 7 standard buffers. All chemicals used were of AR grade. The decanted precipitate was collected from the clarifier as follows: after 48 h (sample N1 ), after 96 h (sample N2 ), after 144 h (sample N3 ), and after 216 h (sample N4 ). After the experiment was completed, scrap iron samples were immediately collected along the packed column at distances corresponding to P1 , P2 , and P3 sampling ports (samples C1 , C2 , and C3 ). Scanning electron microscopy (SEM)–energy dispersive angle X-ray spectrometry (EDAX) was employed to investigate the chemical composition of the scrap iron before (sample C0 ) and after the experiments, as well as the composition of the precipitate. The SEM–EDAX analysis was performed on an Inspect S scanning electron microscope (FEI, Holland) coupled with a GENESIS XM 2i energy dispersive angle X-ray spectrometer. For the speciation of chromium and iron species, samples of secondary minerals (mechanically removed from the surface of exhausted scrap iron), and samples of settled precipitate were dissolved using 3 N HNO3 . The as obtained aqueous solutions where then analyzed using the above mentioned spectrophotometric methods. 3. Results and discussion 3.1. Continuous reduction of Cr(VI) Fig. 2 summarizes the results of Cr(VI) breakthrough. It is shown that, during the first 48 h, Cr(VI) concentration in the column pore
M. Gheju, I. Balcu / Journal of Hazardous Materials 196 (2011) 131–138
4.0
133
3.8
Cr(total) concentration (mg/L)
25
P1 P2 P3
3.6 3.4
pH
3.2 3.0 2.8
24 23 22 21 20 19 18 17 P1 P2 P3
16 15
2.6
0
2.4 0
24
48
72
96
120
144
168
192
216
240
Time (h) Fig. 3. Column pore water pH, at P1 , P2 , and P3 sampling ports.
water decreased from the input value to below the detection limit at the front as the Cr(VI) front passes through the column. High concentrations of Cr(VI) were present in the pore water of P1 from the 6th hour onward. This observation is attributed to two facts: (1) limited extent of Cr(VI) reduction by limited mass of scrap iron available below P1 , and (2) low pH value (<6.0) for which both Cr species are soluble. At P2 , Cr(VI) was totally removed 12 h, while at P3 total removal of Cr(VI) was observed during the first 48 h. At all three sampling ports, Cr(VI) concentration continuously increased after its breakthrough, until a steady-state value was observed. The experiment was stopped at this point. The increase of Cr(VI) concentration proceeds in two stages: high increase rates are observed within the first time interval (approximately 24 h after breakthrough), whereas lower increase rates occurred in the second one. The initial rapid increase of Cr(VI) concentration indicates that the most reactive fraction at the surface of scrap iron filling has been rapidly exhausted. Another explanation is the surface passivation due to progressive formation of an oxide-film. The steady-state Cr(VI) concentration indicates that the extent of iron surface passivation had also reached a steady state. At this stage, a dynamic equilibrium between film grow and film destruction is reached and diffusion processes yielding chromium removal are at equilibrium [34]. With increasing the distance from the inlet end of the column, the Cr(VI) steady state concentrations decreased, as follows: 23.8 mg/L, 18.3 mg/L, and 13.2 mg/L at P1 , P2 , and P3 , respectively.
3.1.1. pH evolution in the column During the first 12 h of the experiment, iron corrosion and Cr(VI) removal were accompanied by an increase in the pore water pH at all sampling ports (Fig. 3). The increase was more significant at P3 and less important at P1 . Subsequently, the pH dropped until it reached a steady-state value of approximately 2.6, 2.55, and 2.5, at P3 , P2 , and P1 , respectively. Iron corrosion and Cr(VI) reduction, accompanied by the formation of Cr/Fe hydroxides, are the main processes responsible for the observed pH change. The first two processes involve consumption of protons (H+ ), while the formation of Cr/Fe hydroxides occurs via consumption of hydroxide ions (HO− ). Prior to the formation of the oxide-film, quantitative Cr(VI) reduction could be expected. The observed pH-decrease was presumably determined by the continuous passivation of scrap iron surface, thereby leading to a higher rate of precipitation than of reduction.
0
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120
144
168
192
216
240
Time (h) Fig. 4. Cr(total) concentration in column pore water vs. time, at P1 , P2 , and P3 sampling ports.
3.1.2. Cr speciation in the column Fig. 2 has already shown that, between P1 and P2 , Cr(III) concentration along the column increased during the entire experiment. This observation may be attributed to an increase in Cr(VI) reduction efficiency as the Cr(VI) front passes from P1 to P2 , reaching scrap iron surfaces less affected by passivation. Between P2 and P3 , Cr(III) concentration decreased during the first 24 h, and continuously increased thereafter until the end of experiment. The initial decrease may be explained by an increase in P3 effluent pH up to 3.8 (Fig. 3), which probably favored the retaining of some Cr(III) inside the top half of the scrap iron filling. Subsequently, the pH at P3 continuously decreased until it reached a steady-state value of approximately 2.6, leading thus to the increase of Cr(III) solubility in pore water. Cr(total) concentrations at P1 , P2 , and P3 , as a function of elapsed time, are presented in Fig. 4. The results show that as the Cr(VI) front passes from P1 to P3 , Cr(total) concentration along the column continuously decreased during the entire experiment, more noticeable at the beginning of the experiment and almost insignificant thereafter. At all three sampling ports, Cr(total) concentration was always less than 25 mg/L, which suggest that chromium was partially retained inside the column during the reduction process. 3.1.3. Fe speciation in the column Aqueous Fe(II) concentrations inside the column increased as the Cr(VI) front passes from P1 to P3 , as shown in Fig. 5. Since Fe(II) occurs in column pore water as a result of iron corrosion and heterogeneous Cr(VI) reduction, the increase of Fe(II) concentration along the column suggests an increase in iron corrosion which is sustained by the presence of a strong oxidizing agent: Cr(VI). Accordingly, Fe(II) concentrations were much higher than it should theoretically be according to the stoichiometry of Eq. (1). This observation is supported by the seminal work of Gould [25] and purchase further experimental data supporting the role of Fe(0) as a generator of reducing agents [26,27,35]. In fact, regardless from the presence of any oxidizing species, protons (H+ ) accept electrons released by Fe(0). This is the primary source of soluble Fe(II), according to Eq. (4) [35]: Fe0 (s) + 2H+ (aq) → Fe2+ (aq) + H2(g)
(4)
H2 bubbles were visualized inside the column at iron–solution interface, especially at the beginning of the experiment. At all three sampling points, Fe(II) concentration continuously decreased in
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80
Fe (III) concentration (mg/L)
Fe(II) concentration (mg/L)
100
P1 P2 P3
80
60
40
20
P1 P2 P3
60
40
20
0
0 0
24
48
72
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216
0
240
24
48
72
3.1.4. Cr removal ˚ Since the ionic radii of Cr3+ and Fe3+ are close (0.63 and 0.64 A, respectively) [36], and Fe(OH)3 and Cr(OH)3 have similar low solubility products (KSP = 1.1 × 10−36 and 5.4 × 10−31 , respectively), the explanations given for Cr(III) behavior inside the column may be applied also for Fe(III). Fe(III) concentrations in column pore water were also much greater than they should theoretically be if all Fe(II) would have been generated only by Eq. (1), confirming thus once again the occurrence of reaction (4) inside the column. Total iron concentrations at P1 , P2 , and P3 , as a function of elapsed time, are presented in Fig. 7. As the Cr(VI) front passes from P1 to P3 , Fe(total) concentration along the column continuously increased during the entire experiment, due to an increase in Cr(VI) reduction efficiency as the solution reaches scrap iron surfaces less affected by passivation. Nevertheless, Fe(total) concentration continuously decreased in time at all three sampling points until the end of the experiment, due to scrap iron passivation. The decrease of Fe concentration is certainly due to precipitation, which is always accompanied by coprecipitation of foreign species (here Cr) [26]. Accordingly, whether Cr(VI) is reduced or not, Cr concentration will decrease along the column. In other words, a fraction of Cr will remain in the column and the fraction passing through should be precipitated (here by NaOH). For the studied wastewater treatment technology, it is important to know the dependence of aqueous chromium and iron species concentration in column effluent (at P3 ) as a function of elapsed time. From Fig. 2 it is apparent that Cr(VI) breakthrough
144
168
192
216
240
Fig. 6. Fe(III) concentration in column pore water vs. time, at P1 , P2 , and P3 sampling ports.
in column effluent occurred immediately after Cr(III) concentration reached its maximum value. Similarly, the maximum peak of Fe(III) concentration at P3 seems also to predict Cr(VI) breakthrough in column effluent, as may be observed from Figs. 2 and 6. The increase of Cr(VI) and Cr(total) concentration, coupled with the decrease of Cr(III) and Fe(total) concentration, noticed after Cr(VI) breakthrough, can be ascribed to passivation of scrap iron surface, process that blocks the access of Cr(VI) to the iron surface and leads to a decrease in Cr(VI) reduction rate. In spite of the low pH, the passivation process was still possible due to formation of secondary solid species on the scrap iron surface, as reported by several previous studies conducted under acidic conditions [28,37–40]. 3.1.5. Solid phase characterization Scrap iron passivation was visually distinguishable through the walls of the column, and the brownish color of the coatings migrated during the experiment from the inlet to the outlet end of the column. EDAX analysis supports this observation, indicating that, while fresh scrap iron contained 81.45% Fe, 15.56% O, and small amounts of Al, Si, and Mn (Fig. 8 and Table 1), after the reaction with Cr(VI) S and Cr were also detected, and the concentration 120
Fe(total) concentration (mg/L)
time until, after 72 h, Fe(II) could not be identified any more in the column pore water. The rapid decrease of Fe(II) concentration is the result of scrap iron surface passivation, which decreases the rate of reactions (1) and (4). However, this rapid decrease of Fe(II) may also suggest that Fe(II) is one important reducing agent for Cr(VI) (Eq. (2)) as previously reported [26,32]. The disappearance of Fe(II) from the column pore water after 72 h indicates that, starting from this point, all Fe(II) resulted from reactions (1) and (4) was either oxidized to Fe(III), or precipitated. Monitoring of the Fe(III) concentration in P1 , P2 , and P3 pore water showed a similar behavior of this parameter with the one observed for Cr(III): between P1 and P2 Fe(III) concentrations increased during the entire experiment, while between P2 and P3 Fe(III) concentrations decreased over the first 36 h, and continuously increased thereafter until the end of experiment (Fig. 6).
120
Time (h)
Time (h) Fig. 5. Fe(II) concentration in column pore water vs. time, at P1 , P2 , and P3 sampling ports.
96
P1 P2 P3
100
80
60
40
20
0 0
24
48
72
96
120
144
168
192
216
240
Time (h) Fig. 7. Fe(total) concentration in column pore water vs. time, at P1 , P2 , and P3 sampling ports.
M. Gheju, I. Balcu / Journal of Hazardous Materials 196 (2011) 131–138
Fig. 8. EDAX pattern and SEM micrograph of un-reacted scrap iron.
Fig. 9. EDAX pattern and SEM micrograph of exhausted scrap iron.
Table 1 Element composition (wt%) of decanted precipitate and scrap iron surface as measured by SEM–EDAX. Element
O Cr Fe Si Al S Ca Mg Na Mn
Sample N1
N2
N3
N4
C0
C1
C2
C3
54.74 7.21 22.98 2.60 0.53 0.53 5.07 5.23 1.11 NA
54.46 8.51 26.18 2.25 0.47 0.73 3.63 2.44 0.54 NA
43.42 14.71 37.96 1.78 NA 1.09 1.05 NA NA NA
37.14 15.74 42.53 2.01 0.58 0.68 1.31 NA NA NA
15.56 NA 81.45 0.73 1.52 NA NA NA NA 0.74
31.08 6.23 60.41 0.34 0.42 1.51 NA NA NA NA
32.36 5.94 59.80 0.66 NA 1.24 NA NA NA NA
38.11 1.24 58.95 0.56 NA 1.13 NA NA NA NA
NA: not available.
of oxygen increased up to 38.11% (Fig. 9 and Table 1). Chromium and iron concentration at the scrap iron surface decreased with increasing the distance from the inlet end of the column, as presented in Table 1; thus, the top half of the scrap iron filling was less affected by the passivation process. Speciation analysis of the secondary mineral phases revealed the existence of chromium mainly as Cr(III) (Table 2). However, low concentrations of Cr(VI) were also detected, probably as a result of Cr(VI) adsorption. The increase of Cr(VI) concentrations along the column may be ascribed to changes in scrap iron surface reactivity during the experiment. These results are in agreement with similar literature reports who also showed the existence of Cr(VI) at the surface of exhausted ZVI, in concentrations accounting up to 20% of the Cr(total) [41–43]. The predominant iron species was Fe(III), consistent with a previous study reporting a mixture of 90% Fe(III) and 10% Fe(II) at the Table 2 Chromium and iron speciation in exhausted scrap iron coatings and decanted precipitate. Sample
N1 N2 N3 N4 C1 C2 C3
135
wt% from Cr(total)
wt% from Fe(total)
Cr(VI)
Cr(III)
Fe(III)
Fe(II)
0 4.25 11.71 7.94 2.53 2.98 3.19
100 95.75 88.29 92.06 97.47 97.02 96.81
99.62 100 100 100 89.67 94.28 95.87
0.38 0 0 0 10.33 5.72 4.13
surface of spent ZVI [44]. From the substantial increase in oxygen concentration and the occurrence of chromium on the surface of the exhausted scrap iron, it can be deduced that there are not only iron, but also chromium oxides/hydroxides formed onto the scrap iron surface as a result of Cr(VI) reduction. SEM micrographs of the scrap iron also confirm the occurrence of secondary mineral phases; while the un-reacted scrap iron was only partially covered by iron oxides (Fig. 8), the exhausted scrap iron was completely covered by secondary phases (Fig. 9). Three types of precipitates were discerned in the micrograph of the exhausted scrap iron. The predominant morphology seems to be as botryoidal clusters, covering the entire scrap iron surface. Euhedral tabular crystalline structures, occurred as thin plates oriented to the surface, are placed between botryoidal phases. Finally, some amorphous structures were also noticed, partially covering the botryoidal clusters. The observed botryoidal morphologies are consistent with the results of a very recent study, carried out at pH values as low as 4, showing that they were composed of FeCr2 O4 [45]. The euhedral structures and amorphous forms are also in good agreement with previous results, in which they were reported to be iron and chromium (oxy)hydroxides [18,36,41]. Thus, the results of SEM, EDAX, and chemical speciation were coincident one another, revealing the formation at the spent scrap iron surface of iron and chromium secondary phases, mainly as Cr(III) and Fe(III) species. 3.2. Continuous precipitation of the resulted Cr and Fe species Our previous batch study [46] carried out using aqueous solutions with pH 2.5, and Cr(III), Fe(II), and Fe(III) concentrations of 50 mg/L, 75 mg/L, and 50 mg/L, respectively, indicated the 500 mg/L value as optimum NaOH dose for cations removal. However, since the first analysis of the column effluent revealed Cr(total) and Fe(total) concentrations lower than the above mentioned, the precipitation process was initialized, starting from the 6th hour onward, with a 400 mg/L NaOH dose. Six hours later, the pH of the clarifier effluent increased up to 9.1, as presented in Fig. 10. Because this pH was considered to alkaline, from the 12th hour onward the NaOH dose was set to 300 mg/L. As a result, the clarifier’s effluent pH subsequently decreased to 7.4 and therefore this dose was maintained until the end of experiment. After another 48 h with relatively constant pH, despite the unchanged NaOH dose, the pH of clarifier effluent slowly increased up to 7.9 at the end of experiment, as a result of decreasing Cr(III) and Fe(total) concentrations in column effluent.
M. Gheju, I. Balcu / Journal of Hazardous Materials 196 (2011) 131–138
9
400
8
pH
7
350
6 300
5 4
250
NaOH dose (mg/L)
pH influent pH effluent NaOH dose
3 2 0
24
48
72
96
120
144
168
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200 240
Time (h) Fig. 10. Variation of clarifier effluent pH, as a function of NaOH dose.
3.2.2. Cr removal The evolution of chromium species removal efficiency vs. NaOH dose is presented in Fig. 12. After the initialization of the precipitation process, Cr(total) removal efficiency markedly increased from 3.2% to 100%. A slight decrease in Cr(total) removal efficiency, up to 94.2%, was noticed afterwards, caused by the pH decrease
400
350 94 300
92 90
Fe(total) Fe(III) Fe(II) NaOH dose
2 0 0
24
48
72
96
120
144
168
192
216
NaOH dose (mg/L)
Iron species removal efficiency (%)
450 100
96
Cr(total) Cr(VI) Cr(III) NaOH dose
80
400
350
60 300
40
250
20
0 0
24
48
72
96
120
144
168
192
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200 240
Time (h)
Fig. 12. Cr species removal efficiency in clarifier, as a function of elapsed time and NaOH dose.
3.2.1. Fe removal From Fig. 11 it can be observed that, after the precipitation was started, Fe(II) removal efficiency noticeably increased from 2.5% to 100% and remained at this value as long Fe(II) was present in the clarifier influent. A similar removal efficiency increase, from 3.1% to 100%, was observed also for Fe(III) after the initialization of the precipitation process. Subsequently, Fe(III) removal efficiency decreased to 95.3%, as a result of pH decrease from 9.1 to 7.4. From the 48th hour onward, Fe(III) removal efficiency slowly increased back, up to 98% at the end of experiment, due to pH increase inside the clarifier. Since the solubility of Fe(OH)2 is much higher than that of Fe(OH)3 (KSP = 1.6 × 10−14 vs. 1.1 × 10−36 ), the 100% removal efficiency of Fe(II) was mainly determined by the oxidation of Fe(II) to Fe(III) in the mixing chamber of the clarifier, process kinetically favored over the pH range of 5–8 [47]. This conclusion was confirmed by the very low Fe(II) concentration in the precipitate decanted during the first 48 h of the experiment (sample N1 , Table 2). Thus, we can reasonable estimate that, inside the clarifier, Fe(total) was almost entirely removed as Fe(III) species.
98
450 100
NaOH dose (mg/L)
450
10
Chromium species removal efficiency (%)
136
250
200 240
Time (h) Fig. 11. Fe species removal efficiency in clarifier, as a function of elapsed time and NaOH dose.
from 9.1 to 7.4. However, it is important to point out that, during the first 48 h, Cr(total) in clarifier influent consisted only in Cr(III), which explains the high removal efficiency of Cr(total). After Cr(VI) breakthrough in column effluent, Cr(total) removal efficiency dropped significantly, up to 48.5% at the end of experiment. This decrease can be attributed to the fact that, after Cr(VI) breakthrough, Cr(total) was comprised from both Cr(III) and Cr(VI); moreover, from the 48th hour onward, Cr(VI) concentration in clarifier influent continuously increased, while Cr(III) concentration continuously decreased. Despite the fact that Cr(VI) exists under circumneutral conditions as highly soluble chromate oxyanions, it was, however, partially removed from the clarifier influent, by coprecipitation with Cr(III) and Fe(III). For example, Cr(VI) may be entrapped in the structure of growing iron hydroxides [48–50]: Cr(VI) + nFe(OH)x(s) → Cr(VI)-[Fe(OH)x ]n(s)
(5)
Cr(VI) removal efficiency was significantly lower than Cr(III) removal efficiency, and continuously decreased in time until the end of experiment; this decrease was probably determined by the decrease in time of the volume of settled precipitate. On the contrary, Cr(III) removal efficiency recorded after 48 h slowly increased up to 98.5% at the end of experiment, as a result of pH increase. Chemical speciation of the decanted precipitate supports these observations, as presented in Table 2. The precipitate collected during the first 48 h was free of Cr(VI), because the clarifier influent did also not contain Cr(VI) over the same period. Nevertheless, Cr(VI) was still detected in the precipitate, but only after its breakthrough in column effluent. Cr(VI) concentration in precipitate increased up to 11.71% after 144 h, and decreased thereafter during the final 72 h of the experiment. The initial increase of Cr(VI) concentration in precipitate can be attributed to the increase of Cr(VI) and decrease of Cr(III) and Fe(III) concentrations in clarifier influent. But, from the 144th hour onward, Cr(VI) concentration in clarifier influent slowly reached to a steady state value; in the same time, the amount of Cr(III)–Fe(III) precipitate settled in the clarifier was much lower than at the beginning of the experiment. Thus, the mass of Cr(VI) removed by coprecipitation during the final 72 h of the experiment was also much lower, determining the decrease of Cr(VI) concentration in the decanted precipitate. Previous studies have shown that dissolved Fe(III) and Cr(III) readily co-precipitate as mixed Fe(III)–Cr(III) (oxy)hydroxides, at pH values greater than 4 [51,52], according to [13]: (1 − x)Fe3+ (aq) + (x)Cr3+ (aq) + 3H2 O → Crx Fe1−x (OH)3(s) + 3H+ (aq) (6)
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Table 3 Chromium and iron mass balance. Column
CCr(VI) (mg/L) CCr(III) (mg/L) CCr(total) (mg/L) CFe(II) (mg/L) CFe(III) (mg/L) CFe(total) (mg/L) MCr(VI) (mg) MCr(III) (mg) MCr(total) (mg) MFe(II) (mg) MFe(III) mg) MFe(total) (mg) M(scrap iron)i (mg) M(scrap iron)f (mg)
Clarifier
In
Out
Retained
In
Out
Retained
25 0 25 0 0 0 1920 0 1920 0 0 0 360000 348406.9
0 22.3 22.3 26.5 53.7 80.2 0 1712.6 1712.6 2035.2 4124.1 6159.3
– – – – – – 6.2 201.2 207.4 330.4 5103.4 5433.8
0 22.3 22.3 26.5 53.7 80.2 0 1712.6 1712.6 2035.2 4124.1 6159.3
0 2.6 2.6 10.8 4.9 15.7 0 199.6 199.6 829.4 376.3 1205.7
– – – – – – 0 1513 1513 1205.8 3747.8 4953.6
4. Conclusions (1 − x)Fe3+ (aq) + (x)Cr3+ (aq) + 2H2 O → Crx Fe1−x (OOH)(s) + 3H+ (aq) (7) EDAX analysis of the settled precipitate (Table 1) indicates that while chromium and iron concentrations increased, oxygen concentration decreased with increasing the elapsed experimental time. Therefore, the degree of precipitates hydration also decreased in time, which could suggest that at the beginning of experiment Crx Fe1−x (OH)3 was the predominant form in precipitate, while at the end of experiment the precipitate consisted mainly from Crx Fe1−x (OOH).
3.3. Chromium and iron mass balance Because the proposed wastewater treatment process is efficient as long Cr(VI) is totally reduced, the mass balance was calculated only up to the moment of Cr(VI) breakthrough in column effluent; the mathematical equations are presented in the supplementary material. The results of first 48 h mass balance, presented in Table 3, indicate an average removal efficiency of 88.3% for Cr(total), and 80.4% for Fe(total). These moderate values were caused by the fact that, before starting the precipitation process, Cr(total) and Fe(total) removal efficiencies were only 3.2% and 2.5%, respectively. However, after the initialization of the precipitation process, the removal efficiencies markedly increased, reaching up to 96.7% under circumneutral conditions. This means that Cr(total) and Fe(total) concentrations in clarifier effluent were as low as 0.8 mg/L and 2.5 mg/L, respectively. As long the column effluent was free of Cr(VI), the average Cr(total) removal efficiency of the packed column and clarifier was 10.8% and 78.8%, respectively; the remaining 10.4% was found in the clarifier effluent, as dissolved Cr(III). The mass balance also shows that, until the moment of Cr(VI) breakthrough, only 3.2% from the initial mass of metallic scrap iron was consumed. Since Cr(VI) breakthrough however occurred, despite the significant mass of unreacted scrap iron, it proves once more that Cr(VI) reduction was significantly retarded by the build-up of a passivating layer on the scrap iron surface. The scrap iron reduction capacity and the total treatment capacity of the wastewater treatment process, calculated up to the moment of Cr(VI) breakthrough, were 5.3 mg Cr(VI)/g scrap iron and 0.2 L/g scrap iron, respectively. This is consistent with the value of 19.2 mg Cr(VI)/g scrap iron, previously reported by a study carried out under same pH conditions, but at a Cr(VI) concentration of only 10 mg/L [30].
Long-term column experiment performed in this work confirmed the possibility of Cr(VI) conversion to Cr(III) by using scrap iron. Although 96.8% from the initial metallic scrap iron still remained unreacted in the column, Cr(VI) breakthrough occurred after 48 h, due to a compactly Cr–Fe composed passivation film formed on the scrap iron. Cr(VI) breakthrough seems to may be predicted by the maximum peak of Fe(III) and Cr(III) concentrations in column effluent. The precipitation of Cr–Fe secondary minerals was more intense at the bottom half of the column, while the top half was less affected. Cr(III) and Fe(III) were the main chromium and iron species detected at the surface of spent scrap iron; however, low concentrations of Cr(VI) and Fe(II) were also identified. After the initialization of the precipitation process, total chromium and iron removal was achieved in the clarifier at pH 9.1, while under circumneutral conditions removal efficiencies as high as 98.5% were observed. The high removal efficiency of Fe(II), in contrast with Fe(OH)2 relative high solubility, was the result of Fe(II) oxidation to Fe(III), kinetically favored over the pH range of 5–8. Although Cr(VI) is highly soluble, it was however partially removed from the clarifier influent by coprecipitation with Cr(III) and Fe(III). This indicates that high Cr(total) removal efficiencies could be attained in the clarifier even after Cr(VI) breakthrough in column effluent, but only for a short period of time. The optimum precipitation pH range (7.6–8.0) was achieved with a 300 mg/L NaOH dose. The experimental results from this study clearly indicate that chromium can be totally removed from Cr(VI) contaminated wastewater by a treatment process combining reduction with scrap iron and chemical precipitation with NaOH. However, in order to assure total reduction of Cr(VI) for a longer period of time, after Cr(VI) breakthrough the scrap iron filling must be re-activated for further reuse. Alternatively, the use of several columns in series should be tested. In this effort, admixing inert materials (e.g. sand, pumice) to Fe(0) in reactive zones could be an efficient tool to save iron costs [53] while possibly increasing sustainability [54].
Acknowledgments This research was conducted under CNCSIS-UEFISCDI PN II IDEI Exploratory Research Project No. 647/19.01.2009 “Innovative technologies for the removal of hexavalent chromium from wastewaters by reuse of scrap iron”, CNCSIS code 1031/2008. The manuscript was improved by the insightful comments of two anonymous reviewers.
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Journal of Hazardous Materials 196 (2011) 109–114
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Removal of some cationic dyes from aqueous solutions using magnetic-modified multi-walled carbon nanotubes Tayyebeh Madrakian a,∗ , Abbas Afkhami a , Mazaher Ahmadi a , Hasan Bagheri b a b
Faculty of Chemistry, Bu-Ali Sina University, Hamedan, Iran Department of Chemistry, Takestan Branch, Islamic Azad University, Takestan, Iran
a r t i c l e
i n f o
Article history: Received 27 April 2011 Received in revised form 30 August 2011 Accepted 31 August 2011 Available online 6 September 2011 Keywords: Cationic dyes Magnetic-modified multi-walled carbon nanotubes Adsorption Removal
a b s t r a c t An adsorbent, magnetic-modified multi-walled carbon nanotubes, was used for removal of cationic dyes crystal violet (CV), thionine (Th), janus green B (JG), and methylene blue (MB) from water samples. Prepared nanoparticles were characterized by SEM, TEM, BET and XRD measurements. The prepared magnetic adsorbent can be well dispersed in the water and easily separated magnetically from the medium after loaded with adsorbate. The influences of parameters including initial pH, dosage of adsorbent and contact time have been investigated in order to find the optimum adsorption conditions. The optimum pH for removing of all the investigated cationic dyes from water solutions was found to be 7.0. The experimental data were analyzed by the Langmuir adsorption model. The maximum predicted adsorption capacities for CV, JG, Th and MB dyes were obtained as 227.7, 250.0, 36.4 and 48.1 mg g−1 , respectively. Desorption process of the adsorbed cationic dyes was also investigated using acetonitrile as the solvent. It was notable that both the adsorption and desorption of dyes were quite fast probably due to the absence of internal diffusion resistance. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Dyes are widely used in the textile and dyestuff industries. Effluents from these industrial facilities are typically of high organic contents and color strength. Thus, dyestuff wastewaters must be treated before discharge, as to minimize the threat to the environment. The strong color in wastewater can decrease the transparency of water and influence photosynthesis activity, which hinders the microbial activities of submerged organisms. However, removal of color from wastewater is of a great challenge. The common dyes include reactive, disperse, acid, and direct dyes. Usually the dyes are low in toxicity, easy to dissolve in water and can be applied for various industrial uses, including inks, cosmetics, soap, and foods [1–4]. The methods used for the removal of organic dyes and pigments from wastewaters are classified into three main categories: physical [5], chemical [6,7] and biological [8,9]. Adsorption methods is the most applied in the removal of organic dyes and pigments from wastewaters, since it can produce high-quality water and also be a process that is economically feasible [10]. Although activated carbon commonly used as adsorbent for color removal [11,12], but the main disadvantage of activated carbon is its high production
∗ Corresponding author. Tel.: +98 811 8257407; fax: +98 811 8257407. E-mail addresses:
[email protected],
[email protected] (T. Madrakian). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.078
and treatment costs [12]. The biological process is difficult to start up and control [13]. Furthermore, the organics in the wastewater cannot be degraded completely by biological processes, and as a result the total treatment cost can increase because of the need of further treatment. Thus, many researchers throughout the world have focused their efforts on optimizing adsorption and developing novel alternative adsorbents with high adsorptive capacity and low cost. In this regard, much attention has recently been paid to nanotechnology. Magnetic nanoparticles as an efficient adsorbent with large specific surface area and small diffusion resistance has been recognized [14,15]. The magnetic separation provides suitable route for online separation, where particles with affinity to target species are mixed with the heterogeneous solution. Upon mixing with the solution, the particles tag the target species. External magnetic fields are then applied to separate the tagged particles from the solution. The synthetic dyes represent a relatively large group of organic chemicals that are met in practically all spheres of our daily life. The cationic dyes such as MB, Th, JG and CV are an important group of organic compounds, which have a variety of scientific and industrial applications [16–22]. So it would be likely that such chemicals have some undesirable effects on humans as well as on environment. In order to minimize the possible damages to humans and the environment arising from the production and applications of cationic dyes, research was carried out around the world.
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Scheme 1. The structure of the investigated dyes.
Removal of cationic dyes methylene blue, neutral red and brilliant cresyl blue from aqueous solution using magnetic multi-wall carbon nanotubes (MMWCNTs) nanocomposite as adsorbent has been reported by Gong et al. [23]. The morphologies of the adsorbent and the synthesized MMWCNT adsorbent were obtained by SEM. It was observed that iron oxide nanoparticles were successfully coated on the surface of multi-wall carbon nanotubes (MWCNTs) to form multi-walled carbon nanotube iron oxide nanocomposites. In the other work, Qu et al. used multi-walled carbon nanotubes filled with Fe2 O3 particles for removal of methylene blue and neutral red from aqueous solutions. The TEM images of this synthesized multi-walled carbon nanotubes demonstrated that the tip of MWCNTs was opened and the inside of the nanotubes were filled with in situ produced Fe2 O3 nanoparticles [24]. Carbon nanofibers can remove cationic and anionic dyes from water samples. The selectivity of adsorption is obtained by adjustment of adsorption pH [25]. But the separation of MWCNTs from solution is very difficult. In this study the multi-walled carbon nanotube that modified with magnetic nanoparticles was synthesized by a simple method and used for removal of cationic dyes from wastewater samples. The technique was found to be very useful and cost-effective for a better removal of dyes. Separation of the dye loaded magnetic-modified multi-walled carbon nanotubes (MMMWCNTs) from the solution was then occurred by an external magnetic field. The adsorption isotherms of the dyes into adsorbent were investigated.
for determination of dye concentration in the solutions. The size, morphology and structure of the nanoparticles were characterized by transmission electronic microscopy (TEM, Philips, CM10, 100 KV) and scanning electron microscope (SEM-EDX, XL30, Philips Netherland). The crystal structure of synthesized materials was determined by an X-ray diffractometer (XRD, 38066 Riva, d/G. Via M. Misone, 11/D (TN) Italy) at ambient temperature. Specific surface area and porosity were defined by N2 adsorption–desorption porosimetry (77 K) using a porosimeter (Bel Japan, Inc.). 2.3. Synthesis of magnetic-modified multi-walled carbon nanotubes
2. Experimental
The synthesis of MMMWCNTs was achieved according to the literature previously reported with some modification [26]. Typically, MWCNTs were first dispersed in concentrated nitric acid at 130 ◦ C for 30 min under stirring to remove the impurities and then washed by DDW until the filtrate is neutral. MWCNTs filled with Fe3 O4 or Fe2 O3 nanoparticles prepared using a simple solution method, in which 0.6 g of ammonium iron (II) sulfate hexahydrate was dissolved in 20 mL of DDW and hydrazine hydrate solution (volume ratio 3:1) to form a grass-green solution, then pretreated MWCNTs of 0.25 g was added. The mixture was sonicated and stirred vigorously. Subsequently, the pH of the mixture was adjusted to 11–13 and kept refluxed at boiling point for 2 h. Finally, the iron oxides nanoparticles-filled MWCNTs were filtered and washed with DDW and anhydrous alcohol for several times, and dried under vacuum for 24 h.
2.1. Reagents and materials
2.4. Removal dyes experiments
All chemicals were of analytical reagent grade or the highest purity available from Merck (Darmstadt, Germany) and double distilled water (DDW) was used throughout the study. In addition, all glassware were soaked in dilute nitric acid for 12 h and finally rinsed for three times with DDW prior to use. Scheme 1 shows the structure of the investigated dyes. Stock solutions of dyes were prepared by dissolving the powder in DDW. Dye solutions of different initial concentrations were prepared by diluting the stock solution in appropriate proportions.
Fifteen milligram of MMMWCNTs were added to 20 mL of 20 mg L−1 of each of cationic dyes MB, Th, JG and CV solutions with predetermined concentration, and the pH of the solution was adjusted at 7.0 with 0.1 mol L−1 HCl or 0.1 mol L−1 NaOH solutions. The mixed solution was then shaken at room temperature for 15 min. Subsequently, the MMMWCNTs with adsorbed dyes were separated from the mixture via a permanent hand-held magnet within 30 s. The residual amounts of dyes in the solution were determined spectrophotometrically at 663, 618, 596, and 601 nm for MB, JG, CV and Th, respectively. The adsorption percentage for each dye, i.e. the dye removal efficiency, was determined using the following expression:
2.2. Instrumentation A Metrohm model 713 pH-meter was used for pH measurements. A single beam UV-mini-WPA spectrophotometer was used
%R =
(C − C ) o t Co
× 100
(1)
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Fig. 1. The SEM image of MMWCNTs. Conditions: voltage: 15 KV, resolution.
where Co and Ct represent the initial and final (after adsorption) concentrations of dyes in mg L−1 , respectively. All tests were performed in triplicate to insure the repeatability of the results; the mean of the three measurements was then reported. All the experiments were performed at room temperature. 3. Results and discussion
111
Fig. 3. X-ray diffraction patterns of (a) MWCNTS, (b) Fe3 O4 , (c) MMWCNTs.
by physical adsorption of a gas on the surface of the solid and by measuring the amount of adsorbed gas corresponding to a monomolecular layer on the surface. The data are treated according to the BET theory [31–34]. The results of the BET method showed that the average specific surface area of MMMWCNTs was 144.68 m2 g−1 , which was higher than that of MWCNTs (44.29 m2 g−1 ) [29] since the MMWCNT adsorbent was a nanocomposite of MWCNTs and iron oxide nanoparticles.
3.1. Characterization of the adsorbent 3.2. Effect of pH Solution pH is an important parameter that affects adsorption process of dye molecules. The solution pH would affect both aqueous chemistry and surface binding-sites of the adsorbent. The effect of the initial pH of the solution in the range 4.0–12.0 with a stirring time of 15 min on the removal of four dyes was investigated using 0.1 mol L−1 HCl or NaOH solutions for pH adjustment, with the initial dye concentration fixed at 20 mg L−1 . The results are shown in Fig. 4. It was observed that, for all the investigated dyes, adsorption quantity increased when the pH of the solution increased from 4 to 7 and remained nearly constant at higher pHs. This can be due to the negative charge of the surface of MMMWCNTs in a wide pH range. The negative charge of the surface of the nanoparticles was also confirmed from the data on the zeta potential [28]. Therefore, the fact that cationic dyes adsorption on MMMWCNTs adsorbent increased with pH values suggested that one of the contributions of MMMWCNTs adsorption toward cationic dyes resulted from electrostatic attraction between the negatively charged MMMWCNTs adsorbent surface and the positively charged cationic dyes [28]. Ideally, wastewater is neutralized and this is an advantage for application of adsorbent in removal cationic dyes from wastewaters.
Removal efficiency, %
The SEM and TEM images of the MMMWCNTs, as shown in Figs. 1 and 2, revealed that the diameter of synthesized MMMWCNTs was around 58 nm. The XRD profile of MWCNTs, MMMWCNTs and Fe3 O4 are shown in Fig. 3. Fig. 3a shows the XRD of MWCNTs and the typical peaks of MWCNTs at 2 = 25.91◦ can be observed. At Fig. 3b the typical peaks of Fe3 O4 and Fe2 O3 , at 2 = 30.2◦ , 35.6◦ , 43.3◦ and 57.2◦ are observed that can be assigned to maghemite or magnetite [27]. Other peaks are also observed at 2 = 53.7◦ and 62.8◦ may be related to the presence of hematite [28,29] and Fig. 3c shows the XRD of MMMWCNTs that includes all the carbon nanotube and iron oxide nanoparticles peaks. The results have a good agreement with other reported [26,30]. The results confirm that iron oxide nanoparticles are encapsulated into the interiors of MWCNTs or adsorbed on the surface of MWCNTs. Specific surface areas are commonly reported as BET surface areas obtained by applying the theory of Brunauer, Emmett, and Teller (BET) to nitrogen adsorption/desorption isotherms measured at 77 K. The specific surface area of the sample is determined
100.0 90.0
MB Th
80.0 70.0
CV JG 0.01
0.02
0.02
0.03
0.03
Weight, g Fig. 2. TEM images of MMWCNTs.
Fig. 4. Percentage of dye removal at different pHs for JG, MB, Th and CV. Conditions: 0.015 g of MMWCNTs, 20 mL of 20 mg L−1 of dye, agitation time of 15 min.
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200.00
90.0
MB Th
80.0
CV
70 0.01
JG 0.02
0.02
0.03
50.00 40.00
150.00
qe, mg g-1
Removal efficiency, %
60.00 100.0
30.00
100.00
20.00 50.00
0.03
Weight, g
10.00
0.00 0.00
Fig. 5. Percentage of dye removal at different amount of MMWCNTs for cationic dyes. Conditions: pH 7.0, 20 mL of 20 mg L−1 of dye, agitation time of 15 min.
50.00
100.00
0.00 150.00
Ce, mg L-1 Fig. 7. Langmuir adsorption isotherm of JG, CV, MB and Th for MMWCNTs.
3.3. Effect of the amount of adsorbent The dependence of the adsorption of dye on the amount of modified nanoparticles was studied at room temperature and at pH 7.0 by varying the adsorbent amount from 0.01 to 0.03 g in contact with 20 mL solution of 20 mg L−1 of each dye. The results are shown in Fig. 5. Apparently, the percentage removal of dyes increased by increasing amount of MMMWCNTs due to the availability of higher adsorption sites. The adsorption reached a maximum with 0.015 g of adsorbent that maximum percentage removal was about 95% for MB and Th and 100% for JG and CV. 3.4. Effect of contact time The effect of contact time on the adsorption of dyes was studied to determine the time taken by MMMCNTs to remove 20 mg L−1 dye solution at pH 7.0. A 0.015 g of adsorbent was added into a 20 mL of dye solution. Absorbance of the solution at related wavelength of each dye with time was determined to monitor the dye concentration. It can be seen that after about 15 min, almost all the dye became adsorbed. The results are shown in Fig. 6. Agitation time of 15 min was selected for further works.
The absorption equilibrium curves of the four dyes to MMMWCNTs were evaluated by adding weighted samples of MMMWCNTs to 50.0 mL solutions of different concentrations of each dye at pH 7.0. The amounts of dyes in the solution were determined after equilibration. The general form of the Langmuir isotherm is: qe KL Ce = qm (1 + KL Ce )
(2)
where KL is a constant and Ce is the equilibrium concentration (mg L−1 ), qe is the amount of dye adsorbed per gram of adsorbent (mg g−1 ) at equilibrium concentration Ce , and qm is the maximum amount of solute adsorbed per gram of surface (mg g−1 ), which depends on the number of adsorption sites. The Langmuir isotherm shows that the amount of dyes adsorbed increases as the concentration increases up to a saturation point. As long as there are available sites, adsorption will increase with increasing dye concentrations, but as soon as all of the sites are occupied, a further increase in concentrations of dyes does not increase the amount of dyes on adsorbents (Fig. 7). After linearization of the Langmuir isotherm, Eq. (2), we obtain:
C 1 e
3.5. Adsorption isotherms
Ce = qe
In order to optimize the use of MMMWCNTs adsorbents, it is important to establish the most appropriate adsorption isotherm. Thus, the correlation of equilibrium data by either theoretical or empirical models is essential to practical operation. Langmuir [35] and Freundlich [36] equations were used to analysis the experimental data of the MMMWCNTs adsorbents for four dyes in our work.
The parameters of this equation for dye were calculated and are given in Table 1. The Freundlich empirical model is represented by:
Removal efficiency, %
100.0 90.0 80.0
+
JG
60.0
MB
50.0
CV
40.0
Th 5.0
10.0
15.0
20.0
25.0
30.0
Time,min Fig. 6. The effect of contact time on the adsorption of cationic dyes on MMWCNTs. Conditions: 20 mL of 20 mg L−1 of dye, pH 7, adsorbent dosage 0.015 g.
KL qm
qe = Kf Ce 1/n
(3)
(4)
where Kf (mmol1−1/n L1/n g−1 ) and 1/n are Freundlich constants depending on the temperature and the given adsorbent–adsorbate couple, n is related to the adsorption energy distribution, and Kf indicates the adsorption capacity. The linearized form of the Freundlich adsorption isotherm equation is ln qe = ln Kf +
70.0
30.0
qm
1 n
ln Ce
(5)
the parameters of Eqs. (3) and (5) for investigated dyes were calculated and are given in Table 1. The results indicate that the experimental data for four dyes do not fit the Freundlich model. Table 2 shows the evaluated parameters of Eqs. (3) and (5) for investigated dyes on to CNTs. The results show capacity factor for the adsorption of JG and CV on the MMMWCNTs is higher than that for MB and Th. Therefore CNTs is the more efficient adsorbent than MMMWCNTs for the adsorption of Th and MB. As the results show, the capacity factor for JG and CV is higher than that for MB and Th. The difference in capacity may be due to the difference in the structure of dyes. JG and CV have quaternary
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113
Table 1 Adsorption isotherms parameters of dyes onto MMMWCNTs. Dyes
JG CV MB Th
Langmuir model
Freundlich model
qmax (mg g−1 )
KL (L mg−1 )
KL qmax −1
r
Kf (mg1−1/n L1/n g−1 )
1/n
r
250.00 227.27 48.08 36.63
51.02 38.46 400.00 166.67
0.204 0.169 8.319 4.550
0.9989 0.9980 0.9982 0.9997
48.23 44.74 32.69 26.26
0.343 0.351 0.134 0.108
0.914 0.879 0.842 0.925
Table 2 Adsorption isotherms parameters of dyes onto CNTs. Langmuir model
Freundlich model
Dyes
qmax (mg g−1 )
KL (L mg−1 )
b (KL qmax −1 )
r
Kf (mg1−1/n L1/n g−1 )
1/n
r
JG CV MB Th
166.67 163.93 76.92 56.49
60.24 55.86 144.92 142.85
0.36 0.34 1.88 2.53
0.9999 0.9996 0.9949 0.9998
59.41 85.42 44.31 33.20
0.23 0.14 0.13 0.12
0.8966 0.8844 0.9156 0.9838
Table 3 Effect of type of eluting agent on recovery (%) for dyes adsorbed on MMMWCNTs (N = 5). Eluent
Recovery (%) Th
CV
MB
JG
Methanol N,N-Dimethylformamide Acetonitrile
85 ± 2 92 ± 4 98 ± 1
93 ± 2 88 ± 5 99 ± 4
82 ± 3 91 ± 3 99 ± 1
94 ± 2 95 ± 4 98 ± 2
Table 4 Comparison the calculated capacity factor for some synthetic adsorbents with proposed method. Sorbent
ammonium group and MB and Th are cationic sulfide dyes. In the JG and CV the positive charge is dispersed on the molecule and in two other dyes the positive charge is located on the heteroatom ring. Therefore the adsorption of CV and JG on the adsorbent is better than that for MB and Th dyes. 3.6. Desorption and reuse study In order to evaluate the possibility of regeneration and reuse of the MMMWCNTs adsorbent, desorption experiments have been performed. Dye desorption from the MMMWCNTs was conducted by washing the dyes loaded on MMMWCNTs using 5.0 mL of methanol, N,N-dimethyl formamide and acetonitrile. For this purpose 5.0 mL of eluent was added to the 0.015 g of dye loaded MMMWCNTs in a beaker. The MMMWCNTs were collected magnetically from the solution. The concentration of dyes in the desorbed solution was measured spectrohotometrically. The results are given in Table 3. As the results show, desorption efficiencies for acetonitrile were higher than other solutions. It was notable that the equilibrium of desorption was achieved within about 2 min, that was fast, similar to the adsorption equilibrium. This was due to the absence of internal diffusion resistance. After elution of the adsorbed dyes, the adsorbent was washed with DDW and vacuum dried at 25 ◦ C overnight and reused for dye removal. The reusability of the sorbent was greater than 5 cycles without any loss in its sorption behavior. Therefore, the MMMWCNTs can be a good reusable and economical sorbent. 4. Conclusion A simple and effective method was presented for removal of cationic dyes from water samples using MMMWCNTs. When iron oxide nanoparticles were adsorbed on the surfaces of MWCNTs, dispersed among the MWCNTs, or encapsulated into the interiors of MWCNTs. The prepared magnetic adsorbent can be well dispersed in the water and can be easily separated magnetically from
Maghemite modified by SDS MMWCNTs MMWCNTs(Fe2 O3 ) MMMWCNTs
Capacity factor (mg g−1 ) MB
Th
JG
CV
– 15.87 42.3 48.1
200 – – 36.4
172 – – 250
– – – 227
Ref.
[29] [23] [24] Proposed method
the medium after adsorption. The rapid adsorption rate is mainly attributed to their unique carbon nanotubes multi-walled structure and carboxylic groups on the carbon nanotube shell providing large surface area and good affinity for the facile and fast adsorption of dye molecules. Table 4 shows the comparison results of our procedure with some methods [23,24,36]. It should be highlighted that the major advantages that the magnetic separation offers is the ability to recover the dye from the nanoparticles and reticulate the particles for further dye separations and easily separated magnetically from the medium after adsorption vs. MWCNTs. Also, the short duration of these experiments have significant practical importance, as it will facilitate smaller reactor volumes ensuring efficiency and economy. References [1] S.Y. Oh, D.K. Cha, P.C. Chiu, B.J. Kim, Conceptual comparison of pink water treatment technologies: granular activated carbon, anaerobic fluidized bed, and zero-valent iron-Fenton process, Water Sci. Technol. 49 (2004) 129–136. [2] S. Papic, N. Koprivanac, A.L. Bozic, A. Metes, Removal of some reactive dyes from synthetic wastewater by combined Al(III) coagulation/carbon adsorption process, Dyes Pigments 62 (2004) 291–298. [3] P. Bayer, M. Finkel, Modeling of sequential groundwater treatment with zero valent iron and granular activated carbon, J. Contam. Hydrol. 78 (2005) 129–146. [4] Y. Lin, C. Weng, F. Chen, Effective removal of AB24 dye by nano/micro-size zero-valent iron, J. Sep. Purif. Technol. 64 (2008) 26–30. [5] A. Afkhami, R. Moosavi, Adsorptive removal of Congo red, a carcinogenic textile dye, from aqueous solutions by maghemite nanoparticles, J. Hazard. Mater. 174 (2010) 398–403. [6] I. Arslan, I.A. Balcioglu, Degradation of remazol black b dye and its simulated dye bath wastewater by advanced oxidation processes in heterogeneous and homogeneous media, Color Technol. 117 (2001) 38–42. [7] M.C. Gutierrez, M. Pepio, M. Crespi, Electrochemical oxidation of reactive dyes method validation and application, Color Technol. 118 (2002) 1–5. [8] A. Stolz, Basic applied aspects in the microbial degradation of azo dyes: a review, Appl. Microbiol. Biotechnol. 56 (2001) 69–80. [9] I.K. Kapdan, R. Ozturk, Effect of operating parameters on color and COD removal performance of SBR, sludge age and initial dyestuff concentration, J. Hazard. Mater. 123 (2005) 217–222. [10] A. Ozcan, A.S. Ozcan, Adsorption of acid red 57 from aqueous solutions onto surfactant-modified sepiolite, J. Hazard. Mater. 125 (2005) 252–259.
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Journal of Hazardous Materials 196 (2011) 139–144
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Preparation of ceramic-corrosion-cell fillers and application for cyclohexanone industry wastewater treatment in electrobath reactor Suqing Wu 1 , Yuanfeng Qi 1 , Yue Gao, Yunyun Xu, Fan Gao, Huan Yu, Yue Lu, Qinyan Yue ∗ , Jinze Li Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental Science and Engineering, Shandong University, 250100 Jinan, China
a r t i c l e
i n f o
Article history: Received 20 March 2011 Received in revised form 29 August 2011 Accepted 2 September 2011 Available online 8 September 2011 Keywords: Ceramic Corrosion cell Sludge Scrap iron Cyclohexanone wastewater
a b s t r a c t As new media, ceramic-corrosion-cell fillers (Cathode Ceramic-corrosion-cell Fillers – CCF, and Anode Ceramic-corrosion-cell Fillers – ACF) employed in electrobath were investigated for cyclohexanone industry wastewater treatment. 60.0 wt% of dried sewage sludge and 40.0 wt% of clay, 40.0 wt% of scrap iron and 60.0 wt% of clay were utilized as raw materials for the preparation of raw CCF and ACF, respectively. The raw CCF and ACF were respectively sintered at 400 ◦ C for 20 min in anoxic conditions. The physical properties (bulk density, grain density and water absorption), structural and morphological characters and toxic metal leaching contents were tested. The influences of pH, hydraulic retention time (HRT) and the media height on removal of CODCr and cyclohexanone were studied. The results showed that the bulk density and grain density of CCF and ACF were 869.0 kg m−3 and 936.3 kg m−3 , 1245.0 kg m−3 and 1420.0 kg m−3 , respectively. The contents of toxic metal (Cu, Zn, Cd, Pb, Cr, Ba, Ni and As) were all below the detection limit. When pH of 3–4, HRT of 6 h and the media height of 60 cm were applied, about 90% of CODcr and cyclohexanone were removed. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Sewage sludge, as a by-product of wastewater treatment, contains pollutants and unstable pathogen substance such as abundant organics, heavy metals and pathogenic bacteria [1–3], thus leading to potential hazards to human health and environment. Traditional treatments such as incineration, landfilling or composting were not economical, safe or sanitary [4,5]. The most widely used final disposal method of sludge is landfilling, with all the risk of soil contamination and degradation of the urban landscape [6,7]. Therefore, in order to prevent secondary pollution and convert sludge into useful resources, an effective and suitable method to treat large amounts of sludge should be found [8]. Scrap iron is a kind of solid waste from production process of machinery plants. The large quantity of scrap iron not only occupies a lot of land, but also causes environmental pollution. How to reduce environmental pollution induced by scrap iron and convert it into useful resources has been widely investigated [9–11]. Ceramics are widely used in construction industry, chemical industry, metallurgy, agriculture and environmental protection [12,13]. Sludge and scrap iron are utilized as additives in ceram-
∗ Corresponding author. Tel.: +86 531 88365258; fax: +86 531 88364513. E-mail addresses:
[email protected],
[email protected] (Q. Yue). 1 These authors equally contributed to this work. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.003
ics production, which can reduce the usage of clay. Thus wastes can be turned into valuable substances [14]. Micro-electrolysis technology was developed on the basis of electrochemistry in Europe during the 1960s [15]. It was believed to operate on the principles very similar to electrochemical methods, and the electrons were supplied from the galvanic corrosion of many micro-scale sacrificial anodes instead of external power supply. Numerous microscopic galvanic cells were formed between the particles of iron and carbon when they were in contact with wastewater (electrolyte solution). The half-cell reactions can be represented as follows [16]: Anode (oxidation) : 2Fe → 2Fe2+ + 4e− , E(Fe2+ /Fe) = −0.44 V Cathode (reduction) : Acidic 2H+ + 2e− → H2 ↑, O2 + 4H+ + 4e− → 2H2 O,
E(H+ /H2 ) = 0 V
E(O2 /H2 O) = +1.23 V
Neutral to alkaline O2 + 2H2 O + 4e− → 4OH− , E(O2 /OH− ) = +0.40 V As a kind of physical–chemical methods, micro-electrolysis technology is widely used to treat refractory wastewater [17], including pesticide wastewater [18], pharmacy wastewater [19], and dye wastewater [20,21]. Micro-electrolysis technology can be used to break the construction of organic pollutants [22] and increase the ratio of BOD5 to COD [23] which can facilitate biological treatment, remove color of wastewater, and reduce energy
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consumption and infrastructure investment. At present, the service life of the electrobath reactor of micro-electrolysis is too short and failure of the electrobath reactor is frequent. CCF and ACF have lower bulk density and grain density, which is beneficial for the backwash process. Therefore, they are suitable to be used as fillers of electrobath reactor for micro-electrolysis. The cyclohexanone wastewater contains hazardous substance such as cyclohexanone and cyclohexane which is harmful to the blood vessel of human body and results in coagulation necrosis of viscera and brain. Therefore, it is significant to find an effective treatment method. At present, the treatment of cyclohexanone mainly includes extraction and electrochemistry. There are few studies on the application of micro-electrolysis for cyclohexanone wastewater treatment. This research investigated the three aspects as follows: (1) The possibility of novel media-ceramic-corrosion-cell fillers (CCF and ACF) used as fillers in electrobath of micro-electrolysis. (2) The electrobath employed in cyclohexanone industry wastewater treatment. (3) The optimum conditions (including pH, HRT and the height of fillers) for the wastewater treatment in terms of the removal efficiency of CODCr and cyclohexanone. 2. Materials and methods
raw pellets were stored in draught cupboard at room temperature (22 ◦ C) for 24 h. Step 2: Sintering treatment. The dried raw pellets were transferred into electric tube furnace (KSY-4D-16, made in China) and sintered at 400 ◦ C for 20 min in anoxic conditions. The pellets were placed in the center of the heated zone. Step 3: Cooling treatment. After the sintering process, the pellets were kept in draught cupboard until they cooled down to room temperature (22 ◦ C). 2.3. Characterization of CCF and ACF Water absorption and bulk density were determined according to GB/T 17431.2-1998 [24]. Before the tests, sintered pellets were kept in an exsiccator (105 ◦ C) for 4 h. Then 100 g of dried pellets were put in a measuring cylinder (500 mL) and leveled completely. The bulk volume of dried pellets was determined. 200 mL of water was added into the measuring cylinder to cover the pellets completely. A dry towel was used to dry the surface of the wet pellets after 1 h, and 1 h saturated wet pellets were weighed. Water absorption and bulk density were calculated by Eqs. (1) and (2), respectively: Water absorption =
2.1. Raw materials Clay and dried sewage sludge (DSS) were utilized as raw materials to prepare CCF, and clay and scrap iron were used as raw materials for ACF production. Clay was obtained from a brickfield (in Zibo, Shandong Province), and dried sewage sludge and scrap iron were obtained from Jinan Wastewater Treatment Plant and Jinan Machinery Plant. Clay, DSS and scrap iron were dried at 105 ◦ C for 4 h, crushed to pass sieve No. 100 (the diameter of mesh was 0.154 mm), and stored until being used in polyethylene vessels to avoid humidification. 2.2. Preparation of CCF and ACF The raw pellets were thermally treated according to the following three steps as shown in Fig. 1: Step 1: Dosage, mixing and drying. Raw pellets were prepared with clay and DSS or clay and scrap iron, which were completely mixed, respectively. Then, the mixture was poured into a pelletizer (DZ20 equipment) to produce pellets (about 7.00 wt% of water was added). Two sieves (the diameters of meshes were 5.00 mm and 6.00 mm, respectively) were used to sift the pellets, which were selected for the following treatment. Before thermal treatment,
mass of 1 h saturated wet pellets − mass of dry pellets ×100% mass of dry pellets (1)
Bulk density (kg m−3 ) =
mass of dry pellets bulk volume of dry pellets
(2)
Grain density was determined by the dry mass (Mdry ) and the volume of the sintered pellets (Vgrain ). Individual grain density was calculated according to the Archimedes’ principle [25]. Structural and morphological analysis was conducted by scanning electron microscopy (Hitachi S-520) both in the surface and in the cross-section (Au coated). 1000.00 g of CCF (or ACF) was soaked into 1.00 L of hydrochloric acid (0.20 mol L−1 ; HCl: = 1.19 g mL−1 Guaranteed Reagent (GR)) for 24 h. 1.00 mL of leach solution obtained from the supernatant was collected for leaching test of the toxic metal elements. Toxic metal concentrations (Cu, Zn, Pb, Cr, Cd, Hg, Ba, Ni, and As) of 1000.00 g of CCF (or ACF) were determined by ICP-AES (IRIS Intrepid II XSP equipment) and were compared with GB 5085.3-2007 [26]. 2.4. Electrobath reactor An electrobath reactor was set up as shown in Fig. 2. The cylindrical reactor made from polymethyl methacrylate had a diameter of 200 mm and an effective volume of 12.5 L with a height of 1.15 m.
Fig. 1. Flow chart for preparation of CCF and ACF.
S. Wu et al. / Journal of Hazardous Materials 196 (2011) 139–144
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Table 1 Properties of CCF and ACF. Ceramics
Bulk density (kg m−3 )
Grain density (kg m−3 )
Water absorption (wt%)
CCF ACF
869.0 936.3
1245.0 1420.0
12.5 14.4
3. Results and discussion 3.1. Properties of CCF and ACF
Fig. 2. Schematic diagram of experiment (dimensioning unit: mm).
CCF (12.5 L) and ACF (12.5 L) were filled in the column and the height of fillers was 80 cm, and the height of graded gravel layer was 20 cm. Tray aerator was installed at the bottom of reactor, aerated sufficiently. The raw wastewater was fed into the column with a feed pump and the treated effluent was collected by an effluent tank. 2.5. Starting and running During the whole experiment period, the electrobath reactor was fed with cyclohexanone industry wastewater which was obtained from a chemical plant (in Dongying, Shandong Province). CODCr of the influent was about 64,000 mg L−1 , and the percentage of cyclohexanone in the influent was about 4.5 wt%. At the start-up, the electrobath was inoculated with the cyclohexanone industry wastewater, the basic operation parameters were as follows: pH of 3, HRT of 12, and the height of fillers of 80 cm. Then pH was varied, HRT was reduced, and the height of fillers was changed, respectively. The optimum conditions were determined in terms of the removal of CODCr and cyclohexanone. 2.6. Analytical methods The concentrations of CODCr in influent and effluent were measured by standard methods [27]. The concentrations of cyclohexanone in influent and effluent were measured by GC–MS (GC/MS-2010QP). Other parameters such as pH, temperature and dissolved oxygen, were monitored regularly. All measurements of each sample were determined in three replicates. The system was operated at room temperature ranging from 25.2 to 30.5 ◦ C.
The physical properties of CCF and ACF were shown in Table 1. It indicated that CCF and ACF had low bulk density and water absorption, and grain density was a little higher than density of water which was beneficial for backwashing and could prevent failure of the electrobath reactor. Table 2 showed the results of the toxic metal leaching test of CCF and ACF. It revealed that all the nine metal (Cu, Zn, Pb, Cr, Cd, Hg, Ba, Ni, and As) concentrations in lixivium were much lower than the limits of the national standards specified by GB 5085.3-2007, China (Hazardous Wastes Distinction Standard-Leaching Toxicity Distinction). It can be seen that CCF and ACF used as fillers for wastewater treatment will not cause secondary pollution. The appearance and microstructure of CCF and ACF ((A) surface of CCF, (B) fracture surface of CCF, (C) surface of ACF, (D) fracture surface of ACF) were shown in Fig. 3. It can be seen from Fig. 3A and C that the surfaces of CCF and ACF were rough and some small apertures contributed to the rough surface, which could increase their specific surface area, meanwhile increase the probability of fillers contacting with wastewater. Fig. 3B and D revealed that there were some large apertures among the frameworks of CCF and ACF, wastewater could easily flow into the fillers, which increased the time for wastewater contacting with fillers. The analysis above indicated that CCF and ACF utilized as fillers in this electrobath for wastewater treatment could increase pollutant removal efficiency. 3.2. Influence of pH on removal of CODCr and cyclohexanone pH was a crucial parameter in micro-electrolysis process. Ten pHs (1–10) were selected and other conditions were as follows: HRT of 12 h and air–liquid ratio (A/L) of 1:1. Fig. 4 showed the effect of pH on removal of CODCr and cyclohexanone. It can be seen that both the removal efficiency of CODCr and cyclohexanone decreased gradually as pH increased. When pH increased from 1 to 4, the removal efficiency of CODCr and cyclohexanone was almost constant, which reached about 90%. When pH increased from 4 to 10, the removal efficiency of CODCr and cyclohexane decreased rapidly, which was lower than 10% as pH was 10. It can be deduced that acidic condition was beneficial for microelectrolysis process, and that the removal efficiency was higher when pH was lower. It may be due to that the electrode voltage in an acidic condition (+1.23 V) was higher than that in a neutral or alkaline condition (+0.40 V), which enhanced micro electrolysis
Table 2 Toxic metal leaching tests of CCF and ACF. Toxic metal
Contents (mg kg−1 of CCF)
Contents (mg kg−1 of ACF)
Threshold (mg kg−1 of hazardous waste)
Toxic metal
Contents (mg kg−1 of CCF)
Contents (mg kg−1 of ACF)
Threshold (mg kg−1 of hazardous waste)
Total Cu Total Zn Total Cd Total Pb Total Cr
0.05 0.03 0.01 0.07 –
0.07 0.01 – 0.03 0.02
100.00 100.00 1.00 5.00 15.00
Total Hg Total Ba Total Ni Total As
– 0.03 – –
– 0.05 – 0.01
0.10 100.00 5.00 5.00
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Fig. 3. The appearance and microstructure of CCF and ACF (SEM): (A) surface of CCF, (B) fracture surface of CCF, (C) surface of ACF, (D) fracture surface of ACF.
reactions and accelerated the dissolving of iron. As a result, more energy was provided which was beneficial for micro-electrolysis process [28]. However, the speed of micro electrolysis reactions and the dissolving of iron decreased rapidly as pH increased, which
led to the decrease of the removal efficiency of CODCr and cyclohexanone. Especially, when pH of wastewater was higher than 7, the removal efficiency of CODCr and cyclohexanone was much lower than that in lower pH, it may be that micro electrolysis reactions were greatly inhibited in basic solution [29]. When pH of wastewater was too low, there would be some disadvantages such as erosion. Overall, in order to ensure the removal efficiency of CODCr and cyclohexanone and to reduce cost of wastewater treatment, pH of 3–4 was selected as the optimum pH. 3.3. Influence of HRT on removal of CODCr and cyclohexanone
Fig. 4. Influence of pH on removal of CODCr and cyclohexanone.
HRT was a key influence factor in micro-electrolysis process. According to the effective volume of the column, ten HRTs (0.5, 1, 2, 3, 4, 5, 6, 8, 10 and 12) were selected and determined. Other conditions were as follows: pH of 3–4 and A/L of 1:1. The flow rate of raw wastewater ranged from 1.0 to 25.0 L h−1 . Fig. 5 showed the influence of HRT on the removal of CODCr and cyclohexanone. It was shown that the removal efficiency of CODCr and cyclohexanone increased gradually as HRT increased. When HRT increased from 0.5 h to 6 h, the removal efficiency increased rapidly, which reached about 90% as HRT was 6 h. When HRT increased from 6 h to 12 h, the removal efficiency was almost
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Fig. 5. Influence of HRT on removal of CODCr and cyclohexanone.
Fig. 6. Influence of height of fillers on removal of CODCr and cyclohexanone.
constant. It can be deduced that the wastewater was still in acidic condition, which could still enhance micro electrolysis reactions and accelerate the dissolving of iron. Moreover, when pH of wastewater was higher than 4, Fe3+ mainly precipitated as Fe(OH)3 , the flocculation precipitation process from ferric hydroxides would result in further removal of organic contaminants [30]. Therefore, micro-electrolysis process was more sufficient as HRT increased, which led to the increase of removal efficiency of CODCr and cyclohexanone as HRT increased from 0.5 h to 6 h. However, the influence of HRT on the removal of CODCr and cyclohexanone was insignificant when HRT was too long. It may be due to that when HRT was longer than 6 h, the wastewater has reached the alkaline condition, where micro electrolysis reactions and the dissolving of iron were greatly restrained, and flocculation precipitation process from ferric hydroxides was also greatly inhibited. Therefore, the removal efficiency of CODCr and cyclohexanone was almost constant as HRT was longer than 6 h. Overall, considering the removal efficiency of CODCr and cyclohexanone and investment in wastewater treatment, HRT of 6 h was a better choice than other HRTs.
was higher than 60 cm. Overall, 60 cm should be selected as the optimum height of fillers. During the whole experiment, attention should be paid to the fact that the removal of CODCr was almost the same as that of cyclohexanone. It can be deduced that cyclohexanone in the wastewater was almost degraded entirely in the electrobath. The results indicated that the removal efficiency of CODCr and cyclohexanone was satisfactory for the wastewater treatment when CCF and ACF were utilized as fillers in the electrobath.
3.4. Influence of height of fillers on removal of CODCr and cyclohexanone The effect of height of fillers on removal of CODCr and cyclohexanone was investigated at pH of 3–4, HRT of 8 h and A/L of 1:1. Four heights of fillers (20 cm, 40 cm, 60 cm, and 80 cm) were selected. The results of the effect of height of fillers on the removal of CODCr and cyclohexanone were shown in Fig. 6. It revealed that the removal efficiency of CODCr and cyclohexanone increased rapidly as the height of fillers increased, which reached about 90% when the height was 60 cm. The removal efficiency was almost constant when the height increased from 60 cm to 80 cm. It can be inferred that increasing the height of fillers could increase HRT, which was beneficial for micro-electrolysis process. When media height was low, the contact time between wastewater and fillers was short, and the wastewater was still in acidic condition, where micro electrolysis reactions and flocculation precipitation process from ferric hydroxides still went on, therefore the removal efficiency of CODCr and cyclohexanone increased as media height increased from 0 to 60 cm. However, when media height reached 60 cm, the contact time between wastewater and fillers was too long, and the wastewater was already in alkaline condition, where micro electrolysis reactions and flocculation precipitation process from ferric hydroxides were inhibited greatly, therefore, the removal efficiency of CODCr and cyclohexanone almost did not increase as media height
3.5. Influence of aeration on removal of CODCr and cyclohexanone Aeration played a key role in micro-electrolysis process. The experiment was performed at pH of 3–4, HRT of 6 h and the height of fillers of 60 cm. The result revealed that the removal efficiency of CODCr and cyclohexanone in aerobic condition reached about 90%, while in the anaerobic condition it was less than 30%. It may be due to that the electrode voltage in aerobic condition (+0.4–1.23 V) was much higher than that in anaerobic condition (0 V), and more energy was provided when oxygen participated in the cathodic reaction, which was beneficial for micro-electrolysis process [31]. Overall, aeration was selected as the experimental condition. 3.6. Backwashing The backwashing operation for the electrobath was water backwashing, and this step was continued for 30 min. According to the effective volume of the column, the superficial velocities of backwash water (40.5 m h−1 ) were determined. And the electrobath was backwashed every three days. As bulk density, grain density and water absorption of CCF and ACF were low, hydraulic resistance of backwash water in backwashing operation was small. It can be deduced that fillers with low bulk density, grain density and water absorption were easily moved, which was beneficial for backwashing. And failure of the electrobath reactor did not occur during the whole experiment. 4. Conclusions As new media, ceramic-corrosion-cell fillers (CCF and ACF) employed in electrobath were investigated for cyclohexanone industry wastewater treatment. The results were as follows: (1) The feasibility of DDS and scrap iron as raw materials of ceramics was verified according to the properties of CCF and ACF.
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(2) CCF and ACF used as fillers in micro-electrolysis process for cyclohexanone industry wastewater treatment were feasible, which could prevent failure of the electrobath reactor. (3) The optimum conditions for wastewater treatment were pH of 3–4, HRT of 6 h, height of fillers of 60 cm and in aerobic condition. Acknowledgements This work was supported by a grant from the Ph.D. Programs Foundation of Ministry of Education of China (No. 20100131110005) and Shandong High-tech Project 2007GG20006003, Shandong Province of China. References [1] J. Werther, T. Ogada, Sewage sludge combustion, Prog. Energ Combust. 25 (1999) 55–116. [2] P.R. Warman, W.C. Termeer, Evaluation of sewage sludge, septic waste and sludge compost applications to corn and forage: yields and N, P and K content of crops and soils, Bioresour. Technol. 96 (2005) 955–961. [3] S. Werle, R.K. Wilk, A review of methods for the thermal utilization of sewage sludge: the Polish perspective, Renew. Energ. 35 (2010) 1914–1919. [4] W.D. Li, W.F. Li, H.F. Liu, Z.H. Yu, Influence of sewage sludge on the slurryability of coal–water slurry, Fuel 88 (2009) 2241–2246. [5] M.J. Wang, Land application of sewage sludge in China, Sci. Total Environ. 197 (1997) 149–160. [6] B.D. Carlos, M.B. Gonzalo, G. Osman, B.M. Lina, A.B. Witold, Processed wastewater sludge for improvement of mechanical properties of concretes, J. Hazard. Mater. 192 (2011) 108–115. [7] A. Hoffman, Competitive Environmental Strategy: A Guide to the Changing Business Landscape, Island Press, Washington, DC, 2000. [8] M.M. Jordan, M.B. Almendro-Candel, M. Romero, J.M. Rincon, Application of sewage sludge in the manufacturing of ceramic tile bodies, Appl. Clay Sci. 30 (2005) 219–224. [9] M. Gheju, A. Iovi, Kinetics of hexavalent chromium reduction by scrap iron, J. Hazard. Mater. B 135 (2006) 66–73. [10] M. Gheju, A. Iovia, I. Balcu, Hexavalent chromium reduction with scrap iron in continuous-flow system. Part 1: Effect of feed solution pH, J. Hazard. Mater. 153 (2008) 655–662. [11] M. Gheju, I. Balcu, Hexavalent chromium reduction with scrap iron in continuous-flow system. Part 2: Effect of scrap iron shape and size, J. Hazard. Mater. 182 (2010) 484–493. [12] S.X. Han, Q.Y. Yue, M. Yue, B.Y. Gao, Y.Q. Zhao, W.J. Cheng, Effect of sludgefly ash ceramic particles (SFCP) on synthetic wastewater treatment in an A/O combined biological aerated filter, Bioresour. Technol. 100 (2009) 1149–1155.
[13] S.Q. Wu, Q.Y. Yue, Y.F. Qi, B.Y. Gao, S.X. Han, M. Yue, Preparation of ultra-lightweight sludge ceramics (ULSC) and application for pharmaceutical advanced wastewater treatment in a biological aerobic filter (BAF), Bioresour. Technol. 102 (2011) 2296–2300. [14] X.R. Wang, Y.Y. Jin, Z.Y. Wang, R.B. Mahar, Y.F. Nie, A research on sintering characteristics and mechanisms of dried sewage sludge, J. Hazard. Mater. 160 (2008) 489–494. [15] X.C. Ruan, M.Y. Liu, Q.F. Zeng, Y.H. Ding, Degradation and decolorization of reactive red X-3B aqueous solution by ozone integrated with internal microelectrolysis, Sep. Purif. Technol. 74 (2010) 195–201. [16] H.F. Cheng, W.P. Xu, J.L. Liu, H.J. Wang, Y.Q. He, G. Chen, Pretreatment of wastewater from triazine manufacturing by coagulation, electrolysis, and internal microelectrolysis, J. Hazard. Mater. 146 (2007) 385–392. [17] X.Y. Yang, Y. Xue, W.N. Wang, Mechanism, kinetics and application studies on enhanced activated sludge by interior microelectrolysis, Bioresour. Technol. 100 (2009) 649–653. [18] Y.P. Wang, L.J. Wang, P.Y. Peng, T.H. Lu, Treatment of naphthalene derivatives with iron–carbon micro-electrolysis, T. Nonferr. Metal. Soc. China 16 (2006) 1442–1447. [19] H.L. Liu, L.P. Zou, W.J. Xie, Treatment of antibiotic wastewater by microelectrolysis-UASB-MBR, Chem. Eng. 116 (2005) 1–3. [20] C. Li, J. Xia, Testing study on electrolysis–catalytic oxidation–SBR process for treating dyeing wastewater, Environ. Preserv. Sci. 117 (2003) 6–8. [21] C.C. Cheng, D.W. Hu, J.X. Zhou, Study on the process of dyestuff wastewater treated by microelectrolysis, Chem. Bioeng. 1 (2005) 29–30. [22] X.L. Yin, W.J. Bian, J.W. Shi, 4-Chlorophenol degradation by pulsed high voltage discharge coupling internal electrolysis, J. Hazard. Mater. 166 (2009) 1474–1479. [23] L. Fan, J.R. Ni, Y.J. Wu, Y.Y. Zhang, Treatment of bromoamine acid wastewater using combined process of micro-electrolysis and biological aerobic filter, J. Hazard. Mater. 162 (2009) 1204–1210. [24] GB/T 17431.2-1998, China, Lightweight aggregates and its test methods. Part 2. Test methods for lightweight aggregate, 1998. [25] Y.F. Qi, Q.Y. Yue, S.X. Han, M. Yue, B.Y. Gao, H. Yu, T. Shao, Preparation and mechanism of ultra-lightweight ceramics produced from sewage sludge, J. Hazard. Mater. 176 (2010) 76–84. [26] GB 5085.3-2007, China, Identification standards for hazardous wastes – identification for extraction toxicity, 2007. [27] State Environmental Protection Administration of China, Monitoring and Analysis Methods of Water and Wastewater, fourth ed., China Environmental Science Press, Beijing, (2002). [28] F. Ju, Y.Y. Hu, Removal of EDTA-chelated copper from aqueous solution by interior microelectrolysis, Sep. Purif. Technol. 78 (2011) 33–41. [29] W. Bian, X. Shen, L. Lei, Degradation of 4-CP in an internal electrolysis system, J. Environ. Sci. 16 (2004) 234–237. [30] P. Gao, X. Chen, F. Shen, G. Chen, Removal of chromium (VI) from wastewater by combined electrocoagulation–electroflotation without a filter, Sep. Purif. Technol. 43 (2005) 117–123. [31] X.Y. Yang, Interior microelectrolysis oxidation of polyester wastewater and its treatment technology, J. Hazard. Mater. 169 (2009) 480–485.
Journal of Hazardous Materials 196 (2011) 145–152
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Pre-irradiation of anatase TiO2 particles with UV enhances their cytotoxic and genotoxic potential in human hepatoma HepG2 cells Jana Petkovic´ a , Tadeja Küzma b , Katja Rade c , Saˇsa Novak c , Metka Filipiˇc a,∗ a b c
Department of Genetic Toxicology and Cancer Biology, National Institute of Biology, SI-1000 Ljubljana, Slovenia Faculty of Chemistry and Chemical Technology, University of Ljubljana, SI-1000 Ljubljana, Slovenia Department for Nanostructured Materials, Joˇzef Stefan Institute, Jamova c. 39, SI-1000 Ljubljana, Slovenia
a r t i c l e
i n f o
Article history: Received 20 July 2011 Received in revised form 23 August 2011 Accepted 2 September 2011 Available online 8 September 2011 Keywords: Anatase TiO2 Cytotoxicity Genotoxicity Photocatalytic HepG2 cells
a b s t r a c t Titanium dioxide (TiO2 ) is active in the UV region of the light spectra and is used as a photocatalyst in numerous applications. Photo-activated anatase TiO2 particles promote increased production of free radicals. This is a desirable property, although the potential toxicity of such photo-activated TiO2 particles on exposure of humans and the environment remains unknown. Therefore, we studied whether preirradiation of TiO2 particles with UV influences their cytotoxic and genotoxic potential. The TiO2 particles, as TiO2 -A (<25 nm) and TiO2 -B (>100 nm), were UV pre-irradiated (24 h) and tested for cytotoxic and genotoxic activities in human hepatoma HepG2 cells. Non-irradiated TiO2 -A/B at 1.0–250 g/ml did not reduce viability of HepG2 cells, nor induce significant increases in DNA strand breaks; only TiO2 -A induced significant increases in oxidative DNA damage. After UV pre-irradiation, both TiO2 -A and TiO2 -B reduced cell viability and induced significant increases in DNA strand breaks and oxidative DNA damage. This is the first study that shows that UV pre-irradiation of anatase TiO2 particles results in increased cytotoxic and genotoxic potential. This warrants further studies as it has important implications for environmental and human health risk assessment and preventive actions to limit human exposure. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Titanium dioxide (TiO2 ) is widely used as a white pigment in the production of paints, paper, ceramics, and as a food additive [1–4]. It is inexpensive, relatively stable chemically, and strongly absorbs UV light. TiO2 is believed to be safe for use: in early studies, submicron-sized TiO2 particles (>100 nm) were classified as harmless [5,6]. However, safe use of TiO2 nanoparticles (NPs) is questionable, as there have been reports from the literature of potential toxicity, with oxidative stress as the main toxicity mechanism [4,7–11]. Under UV light irradiation the physical properties of TiO2 change, as UV activates its photocatalysis. TiO2 is the most investigated photocatalyst system and it has been shown to promote the decomposition of a variety of organic and inorganic compounds, which implies potential applications in sterilisation, sanitation, and pollution remediation [12,13]. Materials coated with TiO2 are already in use, and they show self-cleaning, anti-fogging, and antibacterial properties [14].
∗ Corresponding author at: Department for Genetic Toxicology and Cancer Biology, National Institute of Biology, Veˇcna pot 111, SI-1000 Ljubljana, Slovenia. Tel.: +386 5 9232861; fax: +386 12573847. E-mail address: metka.fi
[email protected] (M. Filipiˇc). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.004
The basic principle of semiconductor photocatalysis involves photon-generated electrons (e− ) and holes (h+ ) that migrate to the surface and serve as redox sources that react with adsorbed reactants, which leads to the destruction of pollutants [15]. The photocatalytical properties of TiO2 depend on the crystalline structure, particle size and specific surface area. Since the anatase crystalline structure has the highest photocatalytic activity, it is the most commonly used form of TiO2 for such purposes [12,13]. Particle size is also an important parameter for catalysis in general, as it directly impacts on the specific surface area of a catalyst. With smaller particle sizes, the numbers of active surface sites increase, as does the surface charge-carrier-transfer rate in photocatalysis [15]. Due to this, TiO2 NPs represent better photocatalyst material than larger, submicron-sized TiO2 particles, so TiO2 NPs are increasingly used nowadays instead of larger TiO2 particles [2,3]. As photocatalytic activation changes the properties of TiO2 towards a greater generation of reactive oxygen species (ROS), it is likely that this will also change its toxicity potential. Many studies have shown that TiO2 is toxic only in the presence of UV irradiation [16,17], and that in the presence of UV irradiation the toxicity of TiO2 is higher than in the dark [18–21]. However, in all of the studies published to date, cells were simultaneously exposed to UV irradiation and TiO2 , and therefore toxic or genotoxic effects of the UV irradiation alone cannot be excluded. Indeed, it is well known that UV irradiation is genotoxic through direct photochemical reactions with DNA, while also
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having indirect genotoxic effects that occur through the production of ROS by photosensitisation of biological molecules other than DNA [22,23], which can all lead to cell mutations or cell death [24]. To exclude the cytotoxic and genotoxic effects of UV irradiation per se, we compared the cytotoxic and genotoxic potential of non-irradiated and UV pre-irradiated anatase TiO2 particles in an experimental model of human hepatoma HepG2 cells. We used two sizes of anatase TiO2 particles, as TiO2 -A (<25 nm) and TiO2 -B (>100 nm). The effects of non-irradiated and UV preirradiated TiO2 particles on cell viability were determined using a 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide (MTT) assay. Induction of DNA damage was determined with a comet assay, and induction of oxidative DNA damage with a modified version of the comet assay with the lesion-specific DNA-repair enzyme formamidopyrimidine-DNA glycosylase (Fpg), which converts oxidised purines to apurinic sites and strand breaks [25].
250 g/ml. These samples were then sonicated for 30 min, to produce non-agglomerated suspensions, before addition to the cells in culture. The time from the end of the UV irradiation and the cell treatment was 1 h maximum. During the experimental work, illumination of the particles was avoided as much as possible (using aluminium foil to wrap the tubes); however, the experiments were not conducted in complete darkness. During the exposure of the cells to the TiO2 -A and TiO2 B, the incubations were kept in complete darkness in a 5% CO2 incubator at 37 ◦ C.
2. Materials and methods
2.5. Determining cytotoxicity: the MTT assay
2.1. Chemicals
Cytotoxicity was determined with the MTT assay according to Mosmann [27], with minor modifications [28]. This assay measures the conversion of MTT to insoluble formazan by dehydrogenase enzymes of intact mitochondria of living cells. The HepG2 cells were seeded into 96-well microplates (Nunc, Naperville, USA) at a density of 40,000 cells/ml, and incubated for 20 h at 37 ◦ C for their attachment. The cell-growth medium was then replaced by fresh medium containing 0, 1, 10, 100 or 250 g/ml TiO2 -A or TiO2 B, and incubated for 4 h, 24 h and 48 h. Each experiment included negative control (cell-growth medium) and positive control (5 M CdCl2 ). The protocol was continued as described previously [28]. Cell survival was determined by comparing the optical density of the wells containing the TiO2 -A/TiO2 -B treated cells with those of the negative control. Five replicates per concentration and three independent experiments were performed. Student t-tests were used to analyse the differences between treated and control cells; p < 0.05 was considered as statistically significant.
MTT, tert-butyl hydroperoxide, benzo(a)pyrene, and dimethyl sulphoxide were all obtained from Sigma–Aldrich (USA). The Fpg enzyme was a gift from Dr. Andrew R. Collins (Department of Nutrition, University of Oslo, Oslo, Norway). 2.2. Characteristics of TiO2 nanoparticles We used two commercial anatase TiO2 particles from Sigma–Aldrich (USA), which we abbreviate as TiO2 -A (Cat. No. 637254: anatase; particle size <25 nm; surface area 200–220 m2 /g) and TiO2 -B (Cat. No. T8141: anatase; no data provided about size and specific surface area). Their sizes, specific surface areas, zeta-potentials and phase compositions were determined experimentally. The sizes and morphologies of TiO2 particles were examined by field-emission-gun scanning electron microscopy (FEG-SEM) with a JEOL 7600 F instrument, and transmission electron microscopy (TEM) with JEM 2010F and JEOL instruments. The phase composition of the particles was verified by X-ray diffraction (Bruker AXS D4 Endeavor) using CuK␣ radiation, and the average crystallite size was calculated according to the Scherrer formula [26]. The specific surface areas were determined by gas adsorption using the BET method (Gemini 2370, Micromeritics). Zeta-potentials were measured before and after UV irradiation, in 3 wt.% suspensions in distilled water and in cell-growth medium, by the electrokinetic sonic amplitude (ESA) technique, using a ZetaProbe device (Colloidal Dynamics, USA). 2.3. TiO2 particles, stock-solution and treatment-media preparation The UV pre-irradiation of the TiO2 -A and TiO2 -B particles (dry, without mixing) was performed by 24 h irradiation in a UV chamber (I-265 CK UV, Kambiˇc Laboratory Equipment) as a simulated sun spectrum with Osram UV bulbs without UVC (ULTRA VITALUX, 300 W, wavelength >290 nm). The stock suspensions were prepared immediately after the end of the UV irradiation of TiO2 particles (within 15 min maximum). Afterwards, the procedure was the same for the non-irradiated and UV-irradiated TiO2 particles. For both non-irradiated and UV-irradiated TiO2 particles, the stock suspensions were prepared at 10 mg/ml in PBS. These were sonicated for 30 min in an ultrasonic bath (Sonorex, Bandelin Electronic, Germany) at a frequency of 60 kHz, to ensure uniform suspension. These stock suspensions were subsequently diluted in cell-growth medium, to final concentrations from 1 g/ml to
2.4. Cell culture The HepG2 cells were obtained from the European Collection of Cell Cultures (UK) and they were grown as described previously [10].
2.6. Determining genotoxicity: the classical and modified comet assays The HepG2 cells were seeded at a density of ca. 60,000 cells/ml into 12-well microtitre plates (Corning Costar Corporation, USA). After incubation at 37 ◦ C in 5% CO2 for 20 h to allow the cells to attach, the cell-growth medium was replaced with fresh medium containing 0, 1, 10, 100 or 250 g/ml TiO2 -A or TiO2 -B, and further incubated for 2 h, 4 h and 24 h. Each experiment included negative control (cell-growth medium) and positive controls (0.3 mM tertbutyl hydroperoxide and 50 M benzo(a)pyrene). At the end of this treatment, the cells were harvested and DNA damage was determined by the protocol of Singh et al. [29], with minor modifications [30]. The level of oxidised purines was determined with a modified comet assay, as described by Collins et al. [25], with minor modifications [30]. Three independent experiments were performed for each of the treatment conditions. Percentages of tail DNA were used to measure the levels of DNA damage. One-way analysis of variance (ANOVA, Kruskal–Wallis) was used to analyse the differences between the treatments within each experiment. Dunnet’s tests were used for comparing median values of percentage tail DNA; p < 0.05 was considered as statistically significant. 3. Results 3.1. Characterisation of TiO2 -A and TiO2 -B The general characteristics of the particles examined are summarised in Table 1, and their morphology is illustrated in Fig. 1. It
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Table 1 Characteristics of the TiO2 particles used in this study.
Supplier information (Sigma–Aldrich) Analysis of particles, as-received (technique used, units) Specific surface area (BET, m2 /g) Crystallite size within the agglomerates/aggregates (XRD, nm) Agglomerates/aggregates size (FEG-SEM, m) Particle shape (TEM) Crystal structure (XRD)
TiO2 -A
TiO2 -B
Nanoparticle, anatase crystalline structure, particle size <25 nm
Particle, anatase crystalline structure
129.3 18 ∼1 Elongated crystallites Anatase
8.6 105 ∼50 Spherical crystallites Anatase
is evident from Fig. 1A, C, and E that TiO2 -A is composed of smaller, but more aggregated (firmly compacted), primary particles (crystallites), compared to TiO2 -B (Fig. 1B, D, and F). As illustrated in the TEM in Fig. 1E and F, the crystallites of TiO2 -A are elongated, while crystallites of TiO2 -B, are spherical. The TiO2 -B particles appear granulated, i.e. compacted into ca. 50 m soft agglomerates (Fig. 1B), which prevents the crystallites from free flowing. Due to the larger size of the TiO2 -B crystallites, the specific surface area of TiO2 -B is significantly lower than that of TiO2 -A. However,
as the specific surface area is measured in gas, it does not necessarily reflect the properties in liquid. In suspension, the granules of the TiO2 -B particles de-agglomerate and disperse, which significantly affects its colloidal behaviour. In both cases, X-ray diffraction confirms the anatase crystal structure (Fig. 2), although for the TiO2 -A particles the peaks are broadened due to the very small crystallite size. The average crystallite sizes calculated according to the Scherrer formula are 18 nm for TiO2 -A and 105 nm for TiO2 -B.
Fig. 1. FEG-SEM images of the TiO2 -A (A and C) and TiO2 -B (B and D) particles, and corresponding TEM images (E and F, respectively) (note: low magnification in B).
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Fig. 2. X-ray diffractogram of the TiO2 -A and TiO2 -B particles (all of the peaks correspond to anatase TiO2 ). Table 2 The properties of the as-received TiO2 particles in suspensions (3 wt.% solids). Property, as-prepared
TiO2 -A
TiO2 -B
pH of aqueous suspension Zeta-potential in water at pH 7 (mV) Electrical conductivity (mS/cm) Iso-electric point in water Zeta-potential in medium (mV) Stability of aqueous suspension
6.5 −20 0.20 6.9 −8.7 Fast sedimentation
7.8 −98 0.10 1.4 −13.5 Stable for days
As stated above, in addition to the characteristics of the particles as-received, their behaviours in liquids are also highly relevant. As given in Table 2, the particles examined revealed significantly different behaviours when mixed into water. The first observation is the difference in the pH of the suspensions at 3 wt.% solids: the natural pH of TiO2 -A was 6.5, while for TiO2 -B it was 7.8, which indicates different chemical compositions of the surface layers of the particles. This is also reflected in the large differences in electrical conductivities and iso-electric points, and in particular in the absolute values of the zeta-potentials at pH 7; the absolute values of the zeta-potentials were higher for TiO2 -B (−98 mV) than for TiO2 -A (−20 mV), which reflected the higher stability of the TiO2 -B suspensions than those of TiO2 -A. It is worth noting that the electrical conductivities increased and the zeta-potentials decreased with time, which indicates a degree of solubility of the particles. As illustrated in Fig. 3, the highly negative surface charge of TiO2 -B in aqueous suspension appears within a wide pH range. In the range of physiological pH values, the TiO2 -B particles are consequently well dispersed and the suspension is stable,
while conversely, TiO2 -A flocculates around pH 7 (near its isoelectric point), which explains the observed poor dispersibility and fast sedimentation of TiO2 -A. It is also evident, that TiO2 -A is positively charged under acidic conditions. The opposite surface charges of these particles imply a tendency for binding to different biomolecules. This difference for TiO2 particles with regard to their surface charges was also reported by Liao et al. [31]. The wide differences in zeta potentials for the TiO2 -A and TiO2 B particles in aqueous suspensions are largely lost in cell-growth medium (Fig. 3). The absolute zeta potential value in cell-growth medium is −8.7 mV and −13.5 mV for TiO2 -A and TiO2 -B, respectively. This will be due to the high ionic strength of the cell-growth medium, as a consequence of the high concentrations of proteins and electrolytes. We also determined whether the properties of the particles change during the 24 h of UV irradiation; here, the changes were insignificant and the morphology remained unchanged. The only change noted was the change in the conductivity and zeta potential of the suspension of the TiO2 -A particles: their conductivity decreased from 0.19 mS/cm before irradiation, to 0.135 mS/cm after irradiation, and their zeta potential increased from 19 mV to 25 mV. These findings suggest the presence, and the UV-induced degradation, of organic molecules on the surfaces of the particles as-received, which affect the behaviour of the suspensions. Hence, although both of these particle types have the same chemistry (TiO2 ) and crystal structures (anatase), they appeared different not only in crystallite size, but also in other chemical and physical properties that probably arise from their different synthetic routes. This can, in turn, affect their bioavailability and toxicity. 3.2. MTT assay Survival of the HepG2 cells treated with non-irradiated and UV pre-irradiated TiO2 -A and TiO2 -B was determined using the MTT assay. The HepG2 cells were exposed to 0, 1, 10, 100 and 250 g/ml non-irradiated or UV pre-irradiated TiO2 -A and TiO2 -B for 4 h, 24 h and 48 h in the dark. Exposure of the HepG2 cells to non-irradiated TiO2 -A and TiO2 -B for up to 48 h did not affect their viability (Fig. 4). On the contrary, UV pre-irradiated TiO2 -A and TiO2 -B significantly (p < 0.05) decreased the viability of the HepG2 cells at the highest two doses (100 and 250 g/ml), which was already evident after 4 h exposure to UV pre-irradiated TiO2 -A and TiO2 -B, and became even more pronounced during the prolonged exposure for 48 h to TiO2 -A (Fig. 4A–C) and TiO2 -B (Fig. 4D–F). 3.3. Induction of DNA strand breaks and oxidative DNA damage: classical and modified comet assays
Fig. 3. Zeta potential of the TiO2 -A and TiO2 -B particles in water and in cell-growth medium (as indicated), as a function of pH.
The extent of DNA strand breaks after exposure of the HepG2 cells to non-irradiated and UV pre-irradiated TiO2 -A and TiO2 -B for 2 h, 4 h and 24 h was determined using the classical comet assay. Exposure of the HepG2 cells to non-irradiated TiO2 -A induced a significant (p < 0.05) increase in DNA strand breaks at 250 g/ml after 2 h, 4 h and 24 h (Fig. 5A–C). Exposure to UV pre-irradiated TiO2 -A for 2 h resulted in a significant (p < 0.05) increase in the level of DNA strand breaks at concentrations above 10 g/ml (Fig. 5A), although after 4 h exposure, the increase was significant (p < 0.05) only at 250 g/ml (Fig. 5B), while after 24 h exposure, significant (p < 0.05) increases were seen for both 100 g/ml and 250 g/ml (Fig. 5C). In cells exposed to non-irradiated TiO2 -B for 2 h, the levels of DNA strand breaks were not significantly increased (Fig. 5D); however, with the longer exposure of 4 h and 24 h, there was a significant (p < 0.05) increase in DNA strand breaks compared to the
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Fig. 4. The effects of non-irradiated and UV pre-irradiated TiO2 -A and TiO2 -B on the viability of HepG2 cells, as indicated: after 4 h (A and D), 24 h (B and E) and 48 h (C and F). Cell viability was measured using the MTT assay, as described in Section 2. PC, positive control of 5 M CdCl2 treatment. Data are means ±SD from three independent experiments (each with five replicates). * p < 0.05, treated versus control cells (Student t-test).
relevant controls (Fig. 5E and F). With the exposure of the cells to UV pre-irradiated TiO2 -B, there were significant (p < 0.05) dosedependent increases in the level of DNA strand breaks at 100 g/ml and 250 g/ml at all times of exposure (Fig. 5D–F).
The induction of oxidative DNA damage was studied with a modified comet assay with the purified Fpg enzyme (see Section 2) that recognises and excises oxidised purines, which are indicative of DNA strand breaks (Fpg-sensitive sites).
Fig. 5. The effects of non-irradiated and UV pre-irradiated TiO2 -A and TiO2 -B on induction of DNA strand breaks in HepG2 cells, as indicated: after 2 h (A and D), 4 h (B and D) and 24 h (C and F). DNA damage was assessed using the comet assay, as described in Section 2. PC, positive control of 0.3 mM (tert-butyl hydroperoxide for 2 h and 4 h treatments, and 50 benzo(a)pyrene for 24 h treatments. Data are means ± SD from three independent experiments, with 50 cells analysed per experimental point. * p < 0.05, treated versus control cells (ANOVA, Kruskal–Wallis with Dunnet’s post test).
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Fig. 6. The effects of non-irradiated and UV pre-irradiated TiO2 -A and TiO2 -B on induction of Fpg-sensitive sites in HepG2 cells, as indicated: after 2 h (A and D), 4 h (B and E) and 24 h (C and F). The modified comet assay was performed as described in Section 2. All further details as for legend to Fig. 5.
In the HepG2 cells exposed to non-irradiated TiO2 -A, there were time-dependent increases in formation of Fpg-sensitive sites, which reached significance (p < 0.05) at 2 h exposure with 250 g/ml (Fig. 6A), at 4 h exposure with 100 g/ml and 250 g/ml (Fig. 6B), and at 24 h exposure with concentrations above 10 g/ml (Fig. 6C). In the cells exposed to UV pre-irradiated TiO2 -A there were significant (p < 0.05) dose-dependent increases in Fpg-sensitive sites at all times of exposure (2 h, 4 h and 24 h) and at all applied concentrations (Fig. 6A–C). From these data, it can also be seen that UV pre-irradiated TiO2 -A induced higher levels of Fpg-sensitive sites than non-irradiated TiO2 -A. Exposure of the HepG2 cells for up to 24 h to non-irradiated TiO2 -B induced significant (p < 0.05) increases in Fpg-sensitive sites only after 2 h of exposure at 250 g/ml (Fig. 6D–F). On the other hand, exposure of the HepG2 cells to UV pre-irradiated TiO2 -B induced significant (p < 0.05) increases in Fpg-sensitive sites after 2 h and 4 h exposure at all of the concentrations tested (Fig. 6D and E), while after 24 h exposure there were significant increases at concentrations above 10 g/ml (Fig. 6F). From these data we can see that UV pre-irradiated TiO2 -B induced much higher levels of Fpg-sensitive sites than non-irradiated TiO2 -B. The levels of oxidative DNA damage induced after 2 h and 4 h exposure by UV pre-irradiated TiO2 -B (Fig. 6D and E) was lower than that induced by UV pre-irradiated TiO2 -A (Fig. 6A and B), while after 24 h exposure, the levels of oxidative DNA damage were comparable (Fig. 6C and F).
4. Discussion Due to the numerous applications of TiO2 photocatalysis, human and environmental exposure to photo-activated TiO2 particles is very likely. However, to our knowledge, to date, no studies have addressed the possibility that photo-activated TiO2 might also have greater reactivity, and consequently greater toxicity, after the UV irradiation has been discontinued. In the present study, we
compared the toxic effects of non-irradiated and UV pre-irradiated anatase TiO2 particles of different average particle sizes, TiO2 -A (18 nm) and TiO2 -B (105 nm), in the experimental model of HepG2 cells. We showed that the toxic potential of these TiO2 particles after UV pre-irradiation is drastically increased not only for the TiO2 -A NPs, but also for submicron-sized TiO2 -B particles. Nonirradiated TiO2 particles did not affect survival of the cells, even at relatively high concentrations and after long times of exposure (48 h), which is in agreement with a number of other studies, for TiO2 NPs [9,16,17,32–35], as well as for submicron-sized TiO2 particles [4–6,35]. Conversely, UV pre-irradiated TiO2 particles of both sizes significantly decreased HepG2 cell viability at all times of exposure, and after UV pre-irradiation their genotoxic potential also increased. Many studies in the literature have suggested that TiO2 particles larger than 100 nm are biologically inert [5,6,36], while TiO2 NPs show greater cytotoxic and genotoxic potential. There is a general hypothesis that the smaller the particle is, the higher the potency it has to induce toxicity in the absence of photo-activation, which has been confirmed in various cell types [4,32,37–39]. However, in the present study, the coarser particles, TiO2 -B, non-irradiated and UV pre-irradiated, induced similarly high levels of DNA strand breaks in HepG2 cells as the TiO2 -A NPs. Although the primary crystallites in TiO2 -B particles are larger than in TiO2 -A, the genotoxic effects of the TiO2 -B particles can be explained by their colloidal behaviour. We showed that the two TiO2 particles used in this study behave significantly differently in water, while the differences observed in cell-growth medium are relatively small. It is likely that not only crystallite size, but especially the form of the particles (strongly aggregated or softly granulated) and its colloidal properties (zeta potential, dispersibility) will influence their biological effects. The zeta potential and also the dispersibility of the TiO2 -B particles are much higher and might increase its bioavailability, and hence its induction of DNA damage. On the other hand, oxidative DNA damage was induced by non-irradiated TiO2 -A, but not TiO2 -B, which is most probably a consequence of the observed
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difference in size of the crystallites (18 nm for TiO2 -A, and 105 nm for TiO2 -B) and the specific surface areas (129.3 m2 /g for TiO2 -A, and 8.6 m2 /g for TiO2 -B). Although TiO2 -A particles are agglomerated and aggregated, some of them are still free, and because of their small size and larger specific surface area, they are more reactive and cause more damage than larger TiO2 -B particles. Gurr et al. [4], who studied the same particles as in the present study in terms of TiO2 -B (Sigma T8141), showed that without UV irradiation, these TiO2 particles did not induce oxidative stress, while anatase TiO2 NPs (10 and 20 nm) did. They also measured the particle sizes in cell-growth medium and found that TiO2 NPs produced aggregations of 1000 nm in diameter, while the >100 nm particles showed no aggregation, which is in agreement with our observations here. Another possible explanation is the different chemical compositions of the surface layers of the particles that were indicated by the difference seen in the natural pH of the two particles. In addition, a decrease in the zeta potential and an increase in conductivity over time indicate a degree of solubility of the particles. However, after UV irradiation, TiO2 -A and TiO2 -B resulted in higher and comparable oxidative DNA damage, because the photocatalytic effects predominate now, such that their toxic potential is not significantly different any more. As already mentioned, a number of studies have shown that in the presence of UV irradiation TiO2 is more cytotoxic and genotoxic than in the absence [18–21,40–43]. Nakagawa et al. [16] studied genotoxicity of anatase and rutile nano-sized and submicron-sized TiO2 in the presence and absence of UV irradiation using several in vitro genotoxicity assays (a comet assay and cell mutation assay with mouse lymphoma cells, a microbial mutation assay with Salmonella typhimurium, and a chromosomal aberration assay with Chinese hamster cells). Without UV irradiation, the nano-sized and submicron-sized TiO2 particles showed little or no genotoxicity, while in combination with UV irradiation, the TiO2 particles showed significant genotoxicity, with nano-sized anatase TiO2 as the most potent. Similar results were obtained in a study by Reeves et al. [17], who in a test system with goldfish skin cells showed that anatase TiO2 NPs alone do not affect cell viability, although they do cause oxidative DNA damage. The combination of TiO2 NPs with UV showed a significant dose-dependent decrease in cell viability and further increases in oxidative DNA damage. They also showed that the toxic/genotoxic effects they observed were most likely due to the formation of hydroxyl radicals. Similar findings were reported by Uchino et al. [19] in a test system with Chinese hamster ovary cells. They also observed that UV irradiation of anatase TiO2 NPs produced larger numbers of hydroxyl radicals in comparison to rutile TiO2 NPs. The conclusion from all of these studies was that photo-activated TiO2 is more toxic than non-photo-activated TiO2 . However, in these studies in which simultaneous exposure of cells to UV and TiO2 was used, it is not possible to exclude any contributions of the UV irradiation per se to the increased toxic effects observed. Our study has several implications that need to be considered in the future evaluation of the potential toxicity of TiO2 particles. The first and most important is related to our finding that irrespective of the particle size, photo-activated anatase TiO2 particles retained elevated reactivity even after the termination of UV exposure. This was seen as greater cytotoxicity and genotoxicity, and in particular, oxidative DNA damage. Among the mechanisms responsible for observed toxic effects, the generation of ROS after UV irradiation is definitely the greatest. Exposure to UV has enough energy to excite electrons from the valence band to conduction band resulting in formation highly reactive electron–hole pairs [13]. In aqueous environments electron (e− ) reduces oxygen to give superoxide anion radicals, which can be dismutated to hydrogen peroxide, and the hole oxidizes water to give hydroxyl radical, respectively. Apparently the UV-irradiated TiO2 particles remained in excited stage also after the irradiation was discontinued, which can explain their
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higher toxic and genotoxic potential compared to non-irradiated TiO2 . From what we have also seen from our data, the toxicity and genotoxicity of photo-activated TiO2 is not dependent on particle size (at least in the investigated range of particle sizes), since both sizes of TiO2 used in our study (TiO2 -A < 100 nm, and TiO2 B >100 nm) showed comparable cytotoxic and genotoxic potential. Also, coarse-sized anatase TiO2 without or with photo-activation cannot be generally considered as safer than nano-sized anatase TiO2 . Our data show that quite different particles can behave relatively similarly in the cell-growth medium, and that not only crystallite size, but also the form of the particles (strongly aggregated or softly granulated) and their colloidal properties (zeta potential, dispersibility) influence their biological effects. This gives us a whole new perspective on the behaviour of TiO2 particles of different sizes in different media, and this will have to be taken into account in any further safety evaluations. 5. Conclusions The results of the present study can be considered relatively alarming and they promote concerns for the safety of all biological systems that might be exposed to photo-activated TiO2 almost anywhere. Considering that 90–99% of the UV light that reaches the surface of the earth shows a UV spectrum with wavelengths of 320–400 nm, almost all of the TiO2 present in the environment that is not specifically protected from this UV can be photoactivated, and become potentially more dangerous in comparison to non-photo-activated TiO2 . Even more worrying, UVA penetrates deep into the skin [44], and as such the users of sunscreens that contain TiO2 can actually be exposed to the more dangerous, photo-activated TiO2 . To our knowledge, this aspect has not been addressed appropriately, as in studies carried out to date only shortterm concurrent exposure to UV and TiO2 has been evaluated. In addition, the possibilities of the application of TiO2 photocatalytic properties are enormous, which will contribute further to increased environmental and human exposure to photo-activated TiO2 . Since the literature contains no similar studies, it is very important to evaluate the toxic potential of photo-activated TiO2 in different experimental models. Our data also change the generally accepted concepts about the toxicity of TiO2 NPs and about the ‘inertness’ of larger TiO2 particles. We have shown that cytotoxity and genotoxicity of TiO2 anatase particles after UV irradiation drastically increases irrespective of particle size. Our new approach that avoids direct UV irradiation of the cells is an appropriate model for toxicological studies of photo-activated TiO2 , and also of other materials that can be photoactivated, and that might provide more reliable evaluation of their potential toxicity. We thus recommend the described experimental approach for inclusion in future studies of the toxicological properties of photo-activated materials. Acknowledgments This study was supported by the Slovenian Research Agency. We thank Ana Gantar and Katja Koenig for the zeta-potential measurements, and Barbara Horvat for the TEM analyses. We thank dr. Cristopher Berrie for critical reading of the manuscript. References [1] H. Nordman, M. Berlin, Titanium, in: L. Friberg, G.F. Nordberg, V.B. Vouk (Eds.), Handbook on the Toxicology of Metals, Elsevier, Amsterdam, 1986, pp. 595–609. [2] M.C. Lomer, R.P. Thompson, J.J. Powell, Fine and ultrafine particles of the diet: influence on the mucosal immune response and association with Crohn’s disease, Proc. Nutr. Soc. 61 (2002) 123–130.
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Journal of Hazardous Materials 196 (2011) 153–159
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Persulfate activation by naturally occurring trace minerals Amy L. Teel, Mushtaque Ahmad, Richard J. Watts ∗ Department of Civil and Environmental Engineering, Box 2910, Washington State University, Pullman, WA 99164-2910, United States
a r t i c l e
i n f o
Article history: Received 11 May 2011 Received in revised form 1 September 2011 Accepted 3 September 2011 Available online 10 September 2011 Keywords: Activated persulfate Minerals In situ chemical oxidation (ISCO) Remediation Hydroxyl radical Sulfate radical
a b s t r a c t The potential for 13 naturally occurring minerals to mediate the decomposition of persulfate and generate a range of reactive oxygen species was investigated to provide fundamental information on activation mechanisms when persulfate is used for in situ chemical oxidation (ISCO). Only four of the minerals (cobaltite, ilmenite, pyrite, and siderite) promoted the decomposition of persulfate more rapidly than persulfate–deionized water control systems. The other nine minerals decomposed persulfate at the same rate or more slowly than the control systems. Mineral-mediated persulfate activation was conducted with the addition of one of three probe compounds to detect the generation of reactive oxygen species: anisole (sulfate + hydroxyl radical), nitrobenzene (hydroxyl radical), and hexachloroethane (reductants and nucleophiles). The reduced mineral pyrite promoted rapid generation of sulfate + hydroxyl radical. However, the remainder of the minerals provided minimal potential for the generation of reactive oxygen species. The results of this research demonstrate that the majority of naturally occurring trace minerals do not activate persulfate to generate reactive oxygen species, and other mechanisms of activation are necessary to promote contaminant destruction in the subsurface during persulfate ISCO. © 2011 Elsevier B.V. All rights reserved.
1. Introduction In situ chemical oxidation (ISCO) has become a dominant technology for the remediation of contaminated soils and groundwater. Of the three common ISCO processes, permanganate, catalyzed H2 O2 propagations (CHP), and activated persulfate, activated persulfate is the least mature and least understood technology. Nonetheless, persulfate appears to have many advantages as an ISCO reagent. Although the persulfate anion (S2 O8 2− ) is a strong oxidant (E0 = +2.01 V), it is usually activated by heat, transition metals, or base to generate sulfate radical (SO4 •− ), a stronger oxidant (E0 = +2.6 V) [1,2]. Sulfate radical can then react with water to form another oxidant, hydroxyl radical (OH• ) [3]. Numerous investigators have focused on the activation of persulfate. Thermal activation of persulfate has been studied for treating methyl tert-butyl ether (MTBE) [4] and trichloroethylene (TCE) [5]. Persulfate activation by iron has been approached from several perspectives. Liang et al. [6] found that iron (II), but not iron (III), activates persulfate, and proposed using thiosulfate to regenerate iron (II) after it was oxidized by persulfate activation. Anipsitakis and Dionysiou [7] demonstrated that out of nine transition metals tested, only three activated persulfate to promote degradation of 2,4-dichlorophenol: silver (I), iron (II), and iron (III),
∗ Corresponding author. Tel.: +1 509 335 3761; fax: +1 509 335 7632. E-mail addresses: amy
[email protected] (A.L. Teel),
[email protected] (M. Ahmad),
[email protected] (R.J. Watts). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.011
with silver (I) providing the most activation and iron (III) the least. Killian et al. [8] and Rastogi et al. [9] subsequently demonstrated that iron chelates are more effective than iron (II) in activating persulfate. Furman et al. [10] recently elucidated the mechanism of base activation of persulfate, which involves the base-catalyzed hydrolysis of persulfate to hydroperoxide anion. The hydroperoxide anion then reduces another persulfate molecule, generating sulfate radical and sulfate anion. High ratios of base to persulfate are often required for effective persulfate activation [11]. Although a modicum of fundamental information is known about persulfate activation in relatively simple aqueous systems, its chemistry in the subsurface has received little attention. Subsurface chemistry has been shown to be a dominant factor in the effectiveness of other ISCO technologies [12]. For example, minerals decompose hydrogen peroxide leading to the formation of reactive oxygen species and resulting in rapid decomposition of hydrogen peroxide in the subsurface [13,14]. Subsurface minerals decompose H2 O2 effectively enough that the addition of soluble iron is usually not required for CHP ISCO [15]. In contrast to research on CHP, minimal attention has been given to the possible activation of persulfate by naturally occurring minerals found in surface and subsurface soils. Liang et al. [16] studied TCE destruction in columns packed with a sandy soil; however, the authors did not investigate the potential for the minerals in the soil to activate persulfate. Costanza et al. [17] studied perchloroethylene (PCE) destruction by thermally activated persulfate in batch reactors containing several solids and soils, but they did not investigate the effect of soil minerals on persulfate activation.
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Similarly, Johnson et al. [18] studied the effect of soil natural oxidant demand, but not soil minerals, on persulfate decomposition. Ahmad et al. [19] investigated the potential for major soil minerals (e.g., goethite, birnessite) found in surface and subsurface soils to activate persulfate. They found that most of the dominant soil minerals, such as goethite and ferrihydrite, did not activate persulfate. Birnessite activated persulfate, but only at concentrations significantly greater than typically found in soils. Many trace minerals are characterized by surface charge couples significantly different from the dominant iron and manganese oxides found in soils. However, the potential for trace minerals, such as ilmenite, cuprite, etc., to activate persulfate has not been investigated to date. The objective of this research was to investigate the potential for soil trace minerals to promote the decomposition of persulfate and activate it to generate reactive oxygen species.
specific to sulfate radical was not identified because all potential sulfate radical probes directly activated persulfate [23]. Nitrobenzene (NB) was used as a hydroxyl radical-specific probe because of its high reactivity with hydroxyl radical (kOH• = 3.9 × 109 M−1 s−1 ) but low reactivity with sulfate radical (kSO4 • − = ≤106 ), allowing a comparison between the two radicals [22,23]. Hexachloroethane, which is unreactive with hydroxyl radical (kOH• ≤ 106 M−1 s−1 ), was used as a reductant probe because it is readily degraded by the reductant superoxide in the presence of electrolytes such as hydrogen peroxide and persulfate [24] and in the presence of mineral surfaces [25]; hexachloroethane is also reduced by alkyl radicals. Hexachloroethane has previously been used as probe for superoxide in CHP and persulfate systems [19,26].
2. Materials and methods
All reactions were conducted in triplicate at 20 ◦ C ± 2 ◦ C in 20 mL borosilicate volatile organic analysis (VOA) vials fitted with polytetrafluoroethylene (PTFE) lined caps. Reactions contained 2 g of mineral and 5 mL of 0.5 M sodium persulfate. This concentration of persulfate was chosen because it is commonly employed in ISCO field applications [10]; the relatively high concentration of trace minerals was used to evaluate the maximum potential effect of the trace mineral on persulfate activity. Reactions investigating total sulfate + hydroxyl radical, hydroxyl radical alone, and reductants contained 1 mM of anisole, nitrobenzene, and hexachloroethane, respectively. The pH of the reactions was allowed to drift downward as it would in a natural groundwater system during ISCO; the pH decreased to a range of 4–5 in all of the reactions. Positive control reactions containing 5 mL of 0.5 M persulfate and no mineral were conducted in parallel to all reactions. In addition, control reactions consisting of 2 g mineral and 5 mL of deionized water in place of persulfate were conducted in parallel to reactions containing probe compounds, to control for any direct effect of the minerals on the probe compounds. At selected time points, a triplicate set of reactors was analyzed for persulfate or probe compound residual concentrations. Reactors analyzed for probe compound residuals were extracted with hexane followed by analysis of the probe compounds by gas chromatography.
2.1. Materials The minerals anatase, bauxite, calcite, cobaltite, cuprite, hematite, ilmenite, magnesite, malachite, pyrite, pyrolusite, siderite, and willemite were purchased from D.J. Minerals (Butte, MT). They were received as cubes approximately 1 cm3 and were crushed to a fine powder using a 150 mL capacity Spex shatter box with hardened steel as a grinder. Mineral surface areas were determined by Brunauer–Emmett–Teller (BET) analysis under liquid nitrogen on a Coulter SA 3100 [20]. The surface areas of the minerals and their chemical formulas are listed in Table 1. The minerals were analyzed for impurities by digesting the samples using EPA Method 3050 followed by inductively coupled plasma/mass spectrometry (ICP/MS); undetectable concentrations were found for metals exclusive of those expected in the minerals. The trace minerals were also evaluated for metals potentially released into persulfate solutions. Aqueous phases of mineral–persulfate slurries were analyzed for 15 elements and metals by ICP/MS. Sodium persulfate (≥98%), anisole (99%), and hexachloroethane (99.9%) were purchased from Sigma–Aldrich (Milwaukee, WI). Nitrobenzene (99%) was obtained from J.T. Baker (Phillipsburg, NJ). HPLC grade hexane was purchased from Fisher Scientific (Fair Lawn, NJ). Double-deionized water (>18 M cm) was purified using a Barnstead Nanopure II Ultrapure system. 2.2. Probe compounds Anisole was used as a probe compound for the total flux of sulfate radical + hydroxyl radical because of its high reactivity with both hydroxyl radical and sulfate radical (kOH• = 5.4 × 109 M−1 s−1 ; kSO4 •− = 4.9 × 109 M−1 s−1 ) [21,22]. A suitable probe compound Table 1 Mineral formulas and surface areas. Mineral
Formula
Surface area (m2 /g)
SEa
Anatase Bauxite Calcite Cobaltite Cuprite Hematite Ilmenite Magnesite Malachite Pyrite Pyrolusite Siderite Willemite
TiO2 Al(OH)3 CaCO3 CoAsS Cu2 O Fe2 O3 FeTiO3 MgCO3 Cu2 (CO3 )(OH)2 FeS2 MnO2 FeCO3 Zn2 SiO4
11.7 28.8 38.0 2.21 49.5 28.2 1.70 38.0 3.65 2.12 1.39 6.80 1.80
0.2 0.2 0.2 0.03 0.08 0.2 0.04 0.3 0.03 0.01 0.04 0.4 0.02
a
Standard error of the mean of three replicates for surface area analysis.
2.3. Experimental procedures
2.4. Analysis Persulfate concentrations were quantified using iodometric titration with 0.01 N sodium thiosulfate [27]. Solution pH was measured using with a Fisher Accumet 900 pH meter. Hexane extracts containing probe compounds were analyzed using Hewlett-Packard 5890 series II gas chromatographs (GC). For anisole and nitrobenzene analysis, the GC was fitted with a 15 m × 0.53 mm (i.d.) SPB-5 capillary column and flame ionization detection (FID) was employed. The injector port temperature was 250 ◦ C, the detector port temperature was 200 ◦ C, the initial oven temperature was 70 ◦ C, the program rate was 20 ◦ C min−1 , and the final temperature was 210 ◦ C. For hexachloroethane analysis the GC was fitted with a 15 m × 0.53 mm (i.d.) Equity-5 capillary column and electron capture detection (ECD) was used. The injector port temperature was 220 ◦ C, the detector port temperature was 270 ◦ C, the initial oven temperature was 100 ◦ C, the program rate was 30 ◦ C min−1 , and the final temperature was 240 ◦ C. ICP/MS analysis of minerals by EPA Method 3050 and the aqueous phase of mineral–persulfate slurries was performed using an Agilent 7500cx ICP/MS system. Triplicate samples were analyzed for each mineral system. Conditions included Rf-power 1450 W; carrier gas flow rate 0.80 L/min; make-up gas flow rate 0.34 L/min; sample uptake 0.5 mL/min; sampling depth 7.5 mm; 1.0 mm Pt sampler; 0.4 mm Pt skimmer.
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Fig. 1. Decomposition of persulfate alone and in the presence of 13 trace minerals. Reactors: 5 mL of 0.5 M persulfate and 2 g of mineral; T = 25 ◦ C ± 1 ◦ C.
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Fig. 2. Pseudo first order rate constants for persulfate decomposition in the presence of thirteen minerals, normalized to mineral surface area. Rate constants for calcite, magnesite, malachite, and willemite were zero.
3. Results and discussion 3.1. Mineral-mediated decomposition of persulfate The decomposition of persulfate in the presence of 13 trace minerals is shown in Fig. 1. The rates of mineral-mediated persulfate decomposition varied substantially; therefore, the minerals were classified into four distinct groups based on their potential to decompose persulfate. The minerals that promoted rapid persulfate decomposition were cobaltite and pyrite with >90% of the persulfate decomposed within 5 h and 30 h, respectively. The minerals classified as slowly decomposing persulfate were ilmenite and siderite, which mediated 48% and 39% persulfate decomposition, respectively, over 60 d. The minerals termed the control group were pyrolusite and hematite, which were characterized by an undetectable difference in persulfate decomposition compared to aqueous controls containing persulfate with no minerals. The scavenging group of minerals (calcite, anatase, bauxite, cuprite, magnesite, malachite, and willemite) promoted persulfate decomposition at rates slower than in aqueous controls with no minerals. Therefore, nine of the 13 minerals either slowed or had minimal effect on the rate of persulfate decomposition, while four minerals (pyrite, cobaltite, ilmenite, and siderite) increased the rate of persulfate decomposition compared to control systems containing persulfate without minerals. These results are consistent with the findings of Johnson et al. [18] and Costanza et al. [17]; they documented that different subsurface solids had wide-ranging effects on persulfate decomposition rates in thermally activated persulfate systems. Possible mechanisms for the decrease in persulfate decomposition rate in the presence of some trace minerals include differences in redox potential or pH in the microenvironment at the surface of the trace mineral [28]. Alternatively, some functional groups on the surfaces of minerals may scavenge reactive oxygen species that might otherwise decompose persulfate [8,29,30]. The metal composition of the minerals did not correlate with their potential to decompose persulfate. For example, ilmenite, pyrite, and siderite are Fe (II) minerals, yet ilmenite and siderite were relatively slow in decomposing persulfate, while pyrite-mediated decomposition of persulfate was very rapid. Furthermore, persulfate decomposition mediated by the Fe (II) minerals was more rapid than that promoted by the Fe (III) mineral hematite, while the copper-based minerals showed no significant difference in persulfate decomposition rates between the Cu (I) of cuprite and the Cu (II) of malachite. The absence of correlation between mineral composition and reactivity with persulfate may
be a function of the nature of persulfate activation; the decomposition of persulfate may be affected by the nature of the mineral surface, or alternatively it may be related to the release of soluble metals into the aqueous phase. Just as hydrogen peroxide is decomposed both by soluble metals and through heterogeneous catalysis by minerals [12], persulfate decomposition in these mineral systems may be due to dissolved metals from the minerals. Iron pyrite dissolves rapidly in mildly acidic water, providing a source of iron (II) for persulfate activation [31,32]. ICP/MS analysis of the aqueous phase of the pyrite–persulfate reactions confirmed the release of soluble iron (II); after 12 h of reaction, the soluble iron (II) concentration in the system was 3.4 mg/L. Pseudo first order rate constants for reactions with minerals were calculated using the data of Fig. 1, and the rate constants were then normalized to the surface area of each mineral (Fig. 2). The trends in persulfate decomposition when normalized to mineral surface area were very similar to the trends seen in Fig. 1, with cobaltite and pyrite promoting the highest rates of persulfate decomposition. However, the normalized pyrolusite-mediated decomposition rate of persulfate increased relative to the normalized rates of the other systems. Manganese oxides are characterized by high surface areas, which likely has a significant influence on their potential to decompose persulfate. A similar trend was noted by Ahmad et al. [19] with the manganese oxide birnessite. Based on the results of Figs. 1 and 2, calcite, ilmenite, and pyrite were used as representative minerals of three groups: (1) those that provided less persulfate decomposition than control systems, (2) those that slowly decomposed persulfate, and (3) those that rapidly decomposed persulfate. The three minerals were used for detailed investigation of reactive oxygen species generation in mineral–persulfate systems. 3.2. Sulfate + hydroxyl radical activity Anisole was used as a probe compound to quantify relative generation rates of sulfate + hydroxyl radical in persulfate systems containing the three representative minerals. The degradation of anisole in persulfate systems containing calcite, ilmenite, and pyrite is shown in Fig. 3(a)–(c). Loss of anisole in the calcite–persulfate system was not significantly different from the persulfate-only control (Fig. 3(a)). These data indicate that calcite does not activate persulfate to generate additional sulfate + hydroxyl radical beyond that generated in persulfate–water systems, which is expected
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Fig. 3. Degradation of the combined sulfate + hydroxyl radical probe anisole by trace mineral–persulfate systems at 25 ◦ C ± 1 ◦ C. (a) Calcite; (b) Ilmenite; (c) Pyrite.
based on the minimal persulfate decomposition promoted by calcite (Fig. 1). In the ilmenite–persulfate system (Fig. 3(b)), anisole degraded 99% within 42 h, demonstrating significant generation of sulfate + hydroxyl radical relative to the persulfate–water control. Anisole degraded even more rapidly in the pyrite–persulfate system (Fig. 3(c)), with >95% degradation over 4 h compared to <20% in the persulfate–water control. These results indicate that rapid decomposition of persulfate in the presence of pyrite (Fig. 1) mediates a corresponding high rate of sulfate + hydroxyl radical generation. 3.3. Hydroxyl radical activity Hydroxyl radical generation, quantified by oxidation of the hydroxyl radical probe nitrobenzene, in persulfate systems
Fig. 4. Degradation of the hydroxyl radical probe nitrobenzene by trace mineral–persulfate systems at 25 ◦ C ± 1 ◦ C. (a) Calcite; (b) Ilmenite; (c) Pyrite.
containing the representative minerals calcite, ilmenite, and pyrite is shown in Fig. 4(a)–(c). Nitrobenzene was oxidized more slowly in the presence of calcite than in the persulfate–water control system (Fig. 4(a)), with 60% nitrobenzene loss in the calcite–persulfate system compared to 90% in the persulfate–water control system over 72 h. These results suggest that calcite inhibited hydroxyl radical generation, and are consistent with the inhibitory effect of calcite on persulfate decomposition. In contrast, a greater flux of hydroxyl radical was promoted in the ilmenite–persulfate system; >99% of the nitrobenzene was oxidized over 36 h, with only 50% of the nitrobenzene lost in the persulfate–water control system over the same time period (Fig. 4(b)). Hydroxyl radical generation was rapid in the pyrite–persulfate system (Fig. 4(c)), with >99% nitrobenzene loss over 14 h in the presence of pyrite while 25% of the nitrobenzene was lost in the persulfate–water control system. The rapid generation of hydroxyl radical in pyrite–persulfate systems may
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have been due to dissolution of the pyrite, providing soluble iron (II) to activate persulfate [31,32]. The results of Fig. 4(a)–(c) demonstrate that the three trace minerals calcite, ilmenite, and pyrite promote highly different rates of hydroxyl radical generation in persulfate systems, consistent with their different potential to promote persulfate decomposition. 3.4. Reductant activity Superoxide is a weak nucleophile and reductant that is generated in CHP systems [24] and in activated persulfate systems [10]. In CHP systems superoxide is responsible for the degradation of highly oxidized organic contaminants (e.g., carbon tetrachloride) [33], for the enhanced desorption of hydrophobic organic contaminants [34], and the enhanced treatment of dense nonaqueous phase liquids (DNAPLs) [35]. HCA was used to investigate the activity of superoxide and other nucleophiles and reductants in persulfate systems containing the representative minerals calcite, ilmenite, and pyrite (Fig. 5(a)–(c)). In calcite–persulfate systems HCA loss was 27% after 72 h (Fig. 5(a)). This loss was not significantly different from the 24% loss in the persulfate–water control, demonstrating that the addition of calcite to persulfate systems did not promote the generation of nucleophiles or reductants. In contrast, reductant activity in ilmenite–persulfate systems was significantly greater than in persulfate–water control systems or in ilmenite systems with no persulfate (Fig. 5(b)) with 63% HCA degradation over 72 h. The data of Fig. 5(a) and (b) demonstrate that while some nucleophiles or reductants are generated in persulfate systems containing ilmenite, a minimal flux of these species is generated by the calcite activation of persulfate. These results are consistent with rates at which ilmenite and calcite promote persulfate decomposition (Fig. 1). HCA loss in the pyrite–persulfate systems was 36% over 72 h (Fig. 5(c)), which was greater than the 24% loss in the persulfate–water control. However, unlike results with the other two minerals (Fig. 3(c) and Fig. 4(c)), systems containing pyrite with no persulfate rapidly degraded HCA, with 77% loss of HCA in the first 24 h and 93% lost over 72 h. Kriegman-King et al. [36] documented that pyrite directly reduces carbon tetrachloride. Because HCA is similar to carbon tetrachloride in degree of chlorination and carbon oxidation state, it was likely also directly reduced by pyrite. Direct reduction of HCA by pyrite may have been minimized in the pyrite–persulfate system due to oxidation of pyrite and Fe (II) by persulfate. 3.5. Screening of reactive oxygen species generation In addition to the detailed analyses of hydroxyl radical, sulfate radical, and superoxide generation by the three representative minerals (calcite, ilmenite, and pyrite) shown in Figs. 3–5, a 24 h evaluation of all the minerals was conducted to provide a simple comparison of the potential for all 13 minerals to generate the three reactive oxygen species. The 24 h results using the probe compounds anisole and nitrobenzene, listed in Table 2, show wideranging potential for mineral activation of persulfate to generate sulfate and hydroxyl radical. Oxidation of anisole and nitrobenzene ranged from less than that of persulfate–water systems for some minerals, to >99.9% for pyrite–persulfate systems. In contrast, the generation of reductants and nucleophiles, quantified by loss of the probe compound HCA, was minimal in all mineral–persulfate systems relative to persulfate–water controls. To better interpret and compare the results of Table 2, the data were normalized to the persulfate–water control systems. The relative degradation of anisole, nitrobenzene, and hexachloroethane in 13 mineral–persulfate systems normalized to corresponding persulfate–water control systems is shown in
Fig. 5. Degradation of the reductant probe hexachloroethane in trace mineral–persulfate systems at 25 ◦ C ± 1 ◦ C. (a) Calcite; (b) Ilmenite; (c) Pyrite.
Fig. 6. The probe compounds degraded at rates similar to or slower than persulfate–water control systems in many of the mineral–persulfate systems. Furthermore, not all minerals were consistent in activation of persulfate for generation of all of the reactive oxygen species, or inhibiting generation of all of the reactive oxygen species. For example, pyrite and malachite promoted the generation of sulfate + hydroxyl radical, but did not result in significant generation of reductants and nucleophiles. Ilmenite, magnesite, and willemite provided a net increase in hydroxyl radical generation, but not sulfate radical generation, relative to the persulfate–water control systems. In contrast, cuprite and malachite promoted some generation of sulfate radical but not hydroxyl radical. Nonetheless, the majority of the 13 minerals evaluated resulted in minimal increases in generation rates of reactive oxygen species relative to persulfate–water systems. The minimal activation of persulfate by most of the trace minerals used in this study is consistent with the results of Kanwartej et al. [37], who found
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Table 2 Percent degradation of the probe compounds anisole, nitrobenzene, and hexachloroethane in mineral–persulfate systems. Anisole
Control (DI Water) Persulfate Alone Anatase Bauxite Calcite Cobaltite Cuprite Hematite Ilmenite Pyrolusite Pyrite Malachite Magnesite Siderite Willemite a
Nitrobenzene
Hexachloroethane
% Loss
SEa
% Loss
SE
% Loss
SE
7.55 45.4 49.1 52.0 40.3 40.4 57.1 66.9 43.3 48.6 >99.9 83.0 51.2 50.9 43.5
1.34 0.321 0.824 0.934 1.73 0.713 1.39 3.14 3.19 0.366 0.0 2.82 1.07 0.381 1.06
15.5 37.2 36.7 29.5 26.4 43.6 36.3 47.6 48.0 23.1 >99.9 43.4 63.0 35.1 50.6
1.84 0.405 2.15 0.962 2.00 0.804 0.848 0.812 0.606 0.378 0.0 0.397 1.04 0.340 0.808
10.9 17.9 22.2 23.9 19.4 29.4 22.1 20.6 23.6 25.1 23.6 19.9 26.0 19.5 19.0
0.587 1.03 0.887 0.764 0.538 0.087 0.751 0.628 1.40 1.09 1.36 0.200 0.893 0.883 2.19
Standard error of the mean of three replicates.
Fig. 6. Degradation of anisole, nitrobenzene, and hexachloroethane in persulfate–mineral systems over 24 h relative to the persulfate–water control.
that in the absence of added activators, the half-life of persulfate in seven aquifer materials ranged from 2 to 600 d. Unlike CHP systems in which most major and trace minerals catalyze the decomposition of hydrogen peroxide to generate a robust mixture of reactive oxygen species, persulfate is not appreciably activated by most trace minerals; furthermore, many of the trace minerals actually decrease rates of persulfate activation and scavenge reactive oxygen species. In addition, those trace minerals that do activate persulfate are usually present in the subsurface at concentrations much lower than those used in this study. These results are consistent with those of Ahmad et al. [19], who found that typical concentrations of goethite, ferrihydrite, and birnessite, the major minerals found in soils, did not activate persulfate to generate reactive oxygen species. Minimal persulfate reactivity has been observed at a number of activated persulfate field applications [38]. Such low reactivity may be due to scavenging of reactive oxygen species by trace minerals present in the subsurface solids. If trace minerals also significantly affect commonly used methods of persulfate activation (e.g., ironchelate activation), then new activators may be required that are
not affected by the scavenging of reactive oxygen species by trace minerals.
4. Conclusions Thirteen trace minerals were evaluated for their potential to promote the decomposition of persulfate and generate reactive oxygen species. Trace minerals were characterized by wideranging potential to promote persulfate decomposition. Seven of the minerals (anatase, bauxite, calcite, cuprite, magnesite, malachite, and willemite) mediated the decomposition of persulfate at rates slower than the persulfate–water control systems containing no minerals. Two minerals (pyrolusite and hematite) promoted persulfate decomposition at rates nearly equal to the persulfate–water control systems. Four minerals (cobaltite, ilmenite, pyrite, and siderite) decomposed persulfate at rates greater than the persulfate–water control systems. Therefore, nine of the 13 minerals either slowed or had minimal effect on the rate of persulfate decomposition.
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The reaction-specific probe compounds anisole, nitrobenzene, and HCA were used to detect total sulfate + hydroxyl radical, hydroxyl radical alone, and reductants and nucleophiles, respectively. The most reactive activator of persulfate for the generation of sulfate radical and hydroxyl radical was pyrite; this result may be due to the reduced nature of pyrite. All of the other minerals provided lower rates of generation of these oxidants than pyrite, and in some cases lower rates of radical generation than the persulfate–water control systems. All of the minerals promoted a small amount of reductant or nucleophile generation relative to the persulfate–water control systems. The majority of the 13 minerals showed minimal potential to generate reactive oxygen species. Acknowledgement Funding for this research was provided by the Strategic Environmental Research and Development Fund through grant no. ER-1489. References [1] D.A. House, Kinetics and mechanism of oxidations by peroxydisulfate, Chem. Rev. 62 (1962) 185–203. [2] A.A. Berlin, Kinetics of radical-chain decomposition of persulfate in aqueous solutions of organic compounds, Kinet. Catal. 27 (1986) 34–39. [3] G.P. Peyton, The free-radical chemistry of persulfate-based total organic carbon analyzers, Marine Chem. 41 (1993) 91–103. [4] K.C. Huang, R.A. Couttenye, G.E. Hoag, Kinetics of heat-assisted persulfate oxidation of methyl tert-butyl ether (MTBE), Chemosphere 49 (2002) 413–420. [5] C.J. Liang, C.J. Bruell, Thermally activated persulfate oxidation of trichloroethylene: experimental investigation of reaction orders, Ind. Eng. Chem. Res. 47 (2008) 2912–2918. [6] C.J. Liang, C.J. Bruell, M.C. Marley, K.L. Sperry, Persulfate oxidation for in situ remediation of TCE. I. Activated by ferrous ion with and without a persulfatethiosulfate redox couple, Chemosphere 55 (2004) 1213–1223. [7] G.P. Anipsitakis, D.D. Dionysiou, Radical generation by the interaction of transition metals with common oxidants, Environ. Sci. Technol. 38 (2004) 3705–3712. [8] P.F. Killian, C.J. Bruell, C.J. Liang, M.C. Marley, Iron (II) activated persulfate oxidation of MGP contaminated soil, Soil Sed. Contam. 16 (2007) 523–537. [9] A. Rastogi, S.R. Al-Abed, D.D. Dionysiou, Effect of inorganic, synthetic and naturally occurring chelating agents on Fe(II) mediated advanced oxidation of chlorophenols, Water Res. 43 (2009) 684–694. [10] O. Furman, A.L. Teel, R.J. Watts, Mechanism of base activation of persulfate, Environ. Sci. Technol. 44 (2010) 6423–6428. [11] O. Furman, A.L. Teel, R.J. Watts, Effect of basicity on persulfate reactivity, J. Environ. Eng. 137 (2011) 241–247. [12] R.J. Watts, A.L. Teel, Chemistry of modified Fenton’s reagent (catalyzed H2 O2 propagations—CHP) for in situ soil and groundwater remediation, J. Environ. Eng. 131 (2005) 612–622. [13] R.J. Watts, Hazardous Wastes: Sources, Pathways, Receptors, John Wiley & Sons, Inc., New York, 1998. [14] A.L. Teel, D.D. Finn, J.T. Schmidt, L.M. Cutler, R.J. Watts, Rates of trace mineralcatalyzed decomposition of hydrogen peroxide, J. Environ. Eng. 133 (2007) 853–858.
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[15] A.J. Khan, R.J. Watts, Mineral-catalyzed peroxidation of tetrachloroethylene, Water Air Soil Pollut. 88 (1996) 247–260. [16] C.J. Liang, I.-L. Lee, I.-Y. Hsu, C.-P. Liang, Y.-L. Lin, Persulfate oxidation of trichloroethylene with and without iron activation in porous media, Chemosphere 70 (2008) 426–435. ˜ J. Callaghan, K.D. Pennell, PCE oxidation by sodium per[17] J. Costanza, G. Otano, sulfate in the presence of solids, Environ. Sci. Technol. 44 (2010) 9445–9450. [18] R.L. Johnson, R.L.P.G. Tratnyek, R.O. Johnson, Persulfate persistence under thermal activation conditions, Environ. Sci. Technol. 42 (2008) 9350–9356. [19] M. Ahmad, A.L. Teel, R.J. Watts, Persulfate activation by subsurface minerals, J. Contam. Hydrol. 115 (2010) 34–45. [20] D.L. Carter, M.M. Mortland, W.D. Kemper, Specific surface, in: A. Klute (Ed.), Methods of Soil Analysis. Part 1. Physical and Mineralogical Methods, American Society of Agronomy, Soil Science Society of America, Madison, WI, 1986, pp. 413–423. [21] P. O’Neill, S. Steenken, D. Schulte-Frohlinde, Formation of radical cations of methoxylated benzenes by reaction with OH radicals, Ti2+ , Ag2+ , and SO4 • − in aqueous solution. An optical and conductometric pulse radiolysis and in situ radiolysis electron spin resonance study, J. Phys. Chem. 79 (1975) 2773–2779. [22] G.V. Buxton, C.L. Greenstock, W.P. Helman, A.B. Ross, Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (• OH/• O− ) in aqueous solution, J. Phys. Chem. Ref. Data 17 (1988) 513–780. [23] P. Neta, V. Madhavan, H. Zemel, R.W. Fessenden, Rate constants and mechanism of reaction of SO4 • − with aromatic compounds, J. Am. Chem. Soc. 99 (1977) 163–164. [24] B.A. Smith, A.L. Teel, R.J. Watts, Identification of the reactive oxygen species responsible for carbon tetrachloride degradation in modified Fenton’s systems, Environ. Sci. Technol. 38 (20) (2004) 5465–5469. [25] O. Furman, D.F. Laine, A. Blumenfeld, A.L. Teel, K. Shimizu, I.F. Cheng, R.J. Watts, Enhanced reactivity of superoxide in water-solid matrices, Environ. Sci. Technol. 43 (2009) 1528–1533. [26] R.J. Watts, D.D. Finn, L.M. Cutler, J.T. Schmidt, A.L. Teel, Enhanced stability of hydrogen peroxide in the presence of subsurface solids, J. Contam. Hydrol. 91 (2007) 312–326. [27] I.M. Kolthoff, R. Belcher, Volumetric Analysis, Volume III, Titration Methods: Oxidation–Reduction Reactions, John Wiley & Sons, Inc., New York, 1957. [28] H.Y. Erbil, Surface Chemistry of Solid and Liquid Interfaces, Wiley-Blackwell, Hoboken, NJ, 2006. [29] N. Kitajima, S. Fukuzumi, Y. Ono, Formation of superoxide ion during the decomposition of hydrogen peroxide on supported metal oxides, J. Phys. Chem. 82 (1978) 1505–1509. [30] R.J. Watts, J. Sarasa, F.J. Loge, A.L. Teel, Oxidative and reductive pathways in manganese-catalyzed Fenton’s reactions, J. Environ. Eng. 131 (2005) 158–164. [31] M. Descostes, P. Vitorgea, C. Beaucaire, Pyrite dissolution in acidic media, Geochim. Cosmochim. Acta 68 (2004) 4559–4569. [32] P. Chirit¸a˘ , M. Descostes, M.L. Schlegel, Oxidation of FeS by oxygen-bearing acidic solutions, J. Colloid Interface Sci. 321 (2008) 84–95. [33] R.J. Watts, J. Howsawkeng, A.L. Teel, Destruction of a carbon tetrachloride DNAPL by modified Fenton’s reagent, J. Environ. Eng. 13 (2005) 1114–1119. [34] J.F. Corbin, A.L. Teel, R.M. Allen-King, R.J. Watts, Reactive oxygen species responsible for the enhanced desorption of dodecane in modified Fenton’s systems, Water Environ. Res. 79 (2007) 37–42. [35] B.A. Smith, A.L. Teel, R.J. Watts, Mechanism for the destruction of carbon tetrachloride and chloroform DNAPLs by modified Fenton’s reagent, J. Contam. Hydrol. 85 (3–4) (2006) 229–246. [36] M.R. Kriegman-King, M. Reinhard, Transformation of carbon tetrachloride by pyrite in aqueous solution, Environ. Sci. Technol. 28 (1994) 692–700. [37] S.S. Kanwartej, N.R. Thomson, J.F. Barker, Persistence of persulfate in uncontaminated aquifer materials, Environ. Sci. Technol. 44 (2010) 3098–3104. [38] S. Borchert, Personal Communication, CH2M-Hill, Freeport, IL, 2010.
Journal of Hazardous Materials 196 (2011) 160–165
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Effect of additives on Hg2+ reduction and precipitation inhibited by sodium dithiocarbamate in simulated flue gas desulfurization solutions Rongjie Lu, Jiaai Hou, Jiang Xu, Tingmei Tang, Xinhua Xu ∗ Department of Environmental Engineering, Zhejiang University, Hangzhou 310058, People’s Republic of China
a r t i c l e
i n f o
Article history: Received 16 March 2011 Received in revised form 1 August 2011 Accepted 3 September 2011 Available online 10 September 2011 Keywords: Hg2+ removal Hg0 re-emission FGD liquors DTCR
a b s t r a c t Mercury (II) (Hg2+ ) ion can be reduced by aqueous S(IV) (sulfite and/or bisulfite) species, which leads to elemental mercury (Hg0 ) emissions in wet flue gas desulfurization (FGD) systems. Numerous reports have demonstrated the high trapping efficiency of sodium dithiocarbamate over heavy metals. In this paper, a novel sodium dithiocarbamate, DTCR, was utilized as a precipitator to control Hg2+ reduction and Hg0 emission against S(IV) in FGD solutions. Results indicated that Hg2+ reduction efficiency decreased dramatically while precipitation rate peaked at around 91.0% in consistence with the increment of DTCR dosage. Initial pH and temperature had great inhibitory effects on Hg2+ reduction: the Hg2+ removal rate gradually increased and reached a plateau along with the increment of temperature and initial pH value. Chloride played a key role in Hg2+ reduction and precipitation reactions. When Cl− concentration increased from 0 to 150 mM, Hg2+ removal rate dropped from 93.84% to 86.05%, and the Hg2+ reduction rate remained at a low level (<7.8%). SO4 2− , NO3 − and other common metal ions would affect the efficiency of Hg2+ reduction and precipitation reactions in the simulated desulfurization solutions: Hg2+ removal rate could always be above 90%, while Hg2+ reduction rate was maintained at below 10%. The predominance of DTCR over aqueous S(IV), indicated by the results above, has wide industrial applications in FGD systems. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Over the past several years, rapid economic growth has created an extremely high energy demand, stimulating a wide expansion of coal combustion industry. However, lacking environmental awareness and adequate technology, China suffered from severe air pollution caused by coal-fired flue gas. Mercury was a major contaminant in the coal-fired flue gas and posed a serious threat to human health and the environment. Wu et al. reported that Hg emissions from coal combustion increased from 202 to 257 t from 1995 to 2003 [1]. In 2005, mercury emissions increased dramatically to a value of 334 t, and the annual average growth rate between 1995 and 2005 was amazingly high at 5.1% [2]. There has been a pressing need to develop effective strategies to cope with mercury emission in the flue gas from coal-fired plants. According to Hg species in the flue gas, some conventional pollution facilities were employed to remove specific mercury species. Particulate-bound mercury (Hgp ) is typically captured in a particulate control device. Compared to Hg0 , oxidized mercury (Hg2+ ) is more soluble in water, less volatile at stack temperatures and more active with mineral matters, so Hg2+ could be easily removed in
∗ Corresponding author. Tel.: +86 571 88982031; fax: +86 571 88982031. E-mail address:
[email protected] (X. Xu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.014
typical air pollution control devices (APCD) [3] such as scrubbers, electrostatic precipitators (ESP) and fabric filters. However, poor capture efficiency of the total mercury was often reported to be coupled with the observation that Hg0 concentration in outlet gas was higher than that in inlet gas [4], which stimulated numerous investigations of novel methods. Yang et al. utilized a series of synthesized nanosized (Fe3−x Mnx )1−ı O4 to oxidize Hg0 to Hg2+ [5]. Yan et al. introduced RuO2 to modify SCR catalyst for Hg2+ oxidation [6]. However, those novel approaches could be prohibitive when taking the economic factors and the stable operation into consideration. In many coal-fired power plants, Hg2+ was removed by FGD systems together with SO2 . Several prior studies suggested that chemical reduction of Hg2+ to Hg0 in the FGD systems was the major source for Hg0 re-emission [7–10]. Thus, it was of/has been attached to great importance to exploring effective methods to inhibit reduction of Hg2+ from wet flue gas desulfurization (FGD). Among various techniques employed, the chelation/precipitation method has been widely adopted due to its convenience and high efficiency. Sodium sulfide, Na2 S, or sodium hydrogen sulfide, NaHS, was employed to form the compound HgS, which could be easily removed from aqueous solution [11]. A synthetic chelating ligand (K2 BDET) was shown to be able to remove mercury from groundwater [12], and the approach demonstrated its capability of removing Hg2+ in groundwater below 0.05 ppb at pH 4.7 and 6.4.
R. Lu et al. / Journal of Hazardous Materials 196 (2011) 160–165
Sodium dithiocarbamates (DTC), which has been widely researched and utilized since 1850s, was obtained through the reaction between a primary or secondary amine and carbon dissulfide in basic media [13]. Due to their high stability, relatively low toxicity to human and low price, numerous studies have demonstrated the versibility and applicability of DTC in medical treatment, industrial production, food sterilization, pollution control, etc. [13,14]. One of sodium dithiocarbamates, DTCR, a new heavy metal macromolecule precipitator, was synthesized and could be coordinated with various heavy metals, such as Hg2+ , Cd2+ , Cu2+ , Pb2+ , Mn2+ , Ni2+ , Zn2+ , Cr3+ , forming chelates with high stability constant. Tang et al. reported [15] that DTCR showed dramatic treatment effect on Hg2+ precipitation in simulated wet FGD solutions, the precipitation efficiency of which was about 87% at the stoichiometric ratio of 1.0. Other similar reports also indicated the unique structure of Hg(DTCR)2 contributing to the high precipitation performance of DTCR over Hg2+ . Ito et al. analyzed the structure of Hg(DTC)2 and found that the two species of Hg–S bonds coexisted as ionic bonds and coordinate bonds, which strengthened the stability of the chelate [16]. The structure of DTCR and Hg(DTCR)2 could be represented as follows [14]. The mechanism of precipitation reaction was illustrated as below: Hg2+ (aq.) + 2DTCR − (aq.) → Hg(DTCR)2 (s)
CH 2
N
CH 2
C
CH 2NCH2CH 2NHCH 3
n
n
S
S
S- Na+
C
C S-
M 2+
S
S-
To better understand the performance of DTCR on inhibiting Hg2+ reduction and Hg0 re-emission in the wet FGD solutions, a lab-scale batch-simulation apparatus was designed. Numerous factors related to the wet FGD properties were explored. The objective of this study was to evaluate the Hg2+ removal efficiency and reduction rate from the wet FGD solutions to determine the inhibition of Hg0 re-emission in the presence of DTCR under different experimental conditions, so as to a theoretical basis for industrial applications.
pH
pH Controller
161
2. Experimental 2.1. Materials Sodium dithiocarbamate (DTCR) (30%) was purchased from the Prode Limited Co., Suzhou, China. Other chemicals including mercury chloride (>99.0%, AR), stannous chloride (98.0%, AR), potassium dichromate (99.95–100.05%, GR), sodium sulfite (>97.0%, AR), potassium permanganate (>99.5%, AR), sodium chloride (>99.5%, AR), calcium chloride (≥96.0%), magnesium chloride(≥98%), nitric acid (65–68%, GR), hydrochloric acid (36–38%, GR), sulfuric acid (95–98%, AR), etc. were all used as received without further purification. The gaseous Ar (99.95%) and N2 (99.9%) were both purchased from Jingong Gas Co. Ltd. 2.2. Experimental apparatus The laboratory-scale (or bench-scale) experimental design shown in Fig. 1 involved two sections: the Hg2+ reduction and precipitation reactor and the Hg0 bubbling absorbing reactor. The former was carried out in a 1-L three-necked flask placed in a water bath at the desired temperature for 2 h. HgCl2 and Na2 SO3 were employed as the sources of Hg2+ and S(IV) respectively. The precipitating agent, DTCR, was added to the reactor. In order to prevent Hg2+ reduction by dissolved oxygen, the reaction mixture was stirred under a nitrogen environment. The bubbling reactors were made of two 100-ml washing-bottles attached to the end of the stem, and contained 10 mL solution composed of 10% (v/v) H2 SO4 –4% (w/w) KMnO4 to absorb and oxidize outlet mercury (mainly Hg0 ) to Hg2+ . This method was a standard protocol adopted by the US Environmental Protection Agency (EPA) [17]. 2.3. Test methods All samples obtained from the experiment were measured using QM201 cold vapor atomic fluorescence spectroscopy (CVAF) coupled with a fluorescence mercury analyzer (Qing’an Instrument Co., Suzhou, China). Prior to analysis, 7% (w/w) SnCl2 solution was prepared as reducing reagent. During the reduction reaction, mercury vapors were flushed out by Ar gas and measured by cold vapor
T
mercury analyzer
H SO /NaOH Feed Tank
Reaction tank N2 Reagent Feed Pump
10%(v/v)H SO 4%(w/v)KMnO
UV-Vis
Stirrer
Fig. 1. Schematic representation of the bench-scale experiment layout for Hg2+ reduction and precipitation reactions.
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generation atomic absorption spectrometry. Before initiating the reaction, all reactant solutions were adjusted to a desired pH value with diluted H2 SO4 and NaOH through a digital pH meter (MettlerToledo Co., Shanghai, China), but no further adjustments were made during the reaction processes. The reaction temperature was controlled by a water bath with a temperature controller. In this study, the reaction efficiency of Hg2+ reduction and precipitation was calculated as follows: Hg2+ rm (%) = Hg2+ rd (%) =
CHg2+ rm CHg2+ in CHg2+ out CHg2+ in
Hg2+ p (%) =
1−
× 100
(1)
× 100
(2)
CHg2+ rm CHg2+ in
−
CHg2+ out CHg2+ in
× 100
(3)
where CHg2+ out (mM) is the Hg2+ concentration in the outlet gas, Hg2+ rd (%) is the Hg2+ reduction rate, Hg2+ p (%) is the Hg2+ precipitation rate, CHg2+ in (mM) is the inlet Hg2+ concentration, and CHg2+ rm (mM) represents the concentration of Hg2+ that remains in solution after the reaction. 3. Results and discussion 3.1. Effect of the precipitator’s dosage The performance under different precipitator dosages was analyzed to study the inhibition of Hg2+ reduction and Hg0 re-emission. Results shown in Fig. 2 indicated that the Hg2+ reduction ratio was 33.62% in the absence of the precipitator in simulated wet FGD solution. When DTCR dosage increased from 0 to the theoretic dosage (Qth = 0.14 mg l−1 ), Hg2+ reduction rate decreased sharply while the precipitation efficiency increased rapidly. When DTCR dosage ranged from 1.0Qth to 3.0Qth , the Hg2+ precipitation rate reached a plateau of around 91.0% while the Hg2+ reduction rate was controlled below 7.0%. Based on the results presented above, the outstanding precipitation efficiency might have a close relationship with its structure. It could be deduced that the precipitation reaction of Hg2+ was dominated by the formation of water-insoluble chelate Hg(DTCR)2 . Apart from the ionic bonds between Hg2+ and DTCR, another
coordinate bonds existed in the complex and formed a bidentate chelate, which ensured the stability of Hg(DTCR)2 [16]. Research also indicated that two –CSS– legends in DTCR could chelate with Hg2+ to form insoluble polymer matrices with a cross-linked network structure [18]. When DTCR dosage was 1.0Qth , 91.1% of Hg2+ was precipitated, and therefore 1.0Qth was chosen as DTCR dosage in the subsequent experiments. 3.2. Effect of the reaction temperature The performance under different reaction temperatures (30–60 ◦ C) was tested to study the temperature effect on Hg2+ reduction and Hg0 re-emission inhibited by DTCR. As shown in Fig. 3, the Hg2+ reduction efficiency varied from 9.14% to 5.53% by adding DTCR when reaction temperature increased to a certain extent. Other experimental conditions such as an initial pH value of 5.0, S(IV) concentration of 5.0 mM, and reaction time of 2 h remained the same as previous study. When the temperature rose to 50 ◦ C, the reduction rate decreased to 6.29% from initial 8.45% while the precipitation efficiency reached 92.72% compared to 86.08% at 30 ◦ C, achieving good inhibition performance. Since the temperature of 50 ◦ C was obtained in the practical desulfurization solution, that solution was chosen in the subsequent experiments. The observation that precipitation rate saw a slight increase along with the temperature rise might be attributed to the thermodynamics of the reaction. High temperature accelerated collisions between Hg2+ and DTCR, which in maro-phenomenon indicated the increment of the reaction rate. Furthermore, the increment of temperature produced more confirmations of DTCR owing to the multiple polymer chains. Thus, the structure of Hg(DTCR)2 was more cross-linked and stable. 3.3. Effect of initial pH values The initial pH value of the reaction solution was a critical factor on Hg2+ reduction and precipitation since the reactivity of precipitators could be influenced by pH value. Binding capacities between Hg2+ and DTCR were weakened due to the protonation of sulfur atoms in functional groups of DTCR at low pH values, which was in accordance with the results shown in Fig. 4. Along with the increase in pH value, the Hg2+ reduction efficiency declined from 11.27% to 2.30% while the Hg2+ precipitation efficiency increased 100
100
80
80
60
2+
Hg (%)
Hg (%)
60
Hg precipitation 2+
Hg remained
40
2+
Hg reduction 2+
Hg remained
40
2+
Hg precipitation
2+
Hg reduction 20
20
0
0 0.0
0.5
1.0
1.5
2.0
2.5
3.0
Dosing quantity (Q th) Fig. 2. Effect of dosing quantity on Hg2+ reduction and Hg0 re-emission inhibited by the introduction of DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, T = 50 ◦ C, and pH value = 5).
30
40
50
60
o
Temperature ( C) Fig. 3. Effect of the reaction temperature on Hg2+ reduction and Hg0 re-emission inhibited by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, DTCR = 0.14 mg l−1 and pH value = 5).
100
100
80
80
60
2+
Hg
2+
Hg remained
40
2+
Hg
163
2+
60
reduction Hg (%)
Hg (%)
R. Lu et al. / Journal of Hazardous Materials 196 (2011) 160–165
Hg reduction 2+
Hg remained
40
2+
precipitation
Hg precipitation
20
20
0 0 3
4
5
6
7
pH
gradually and was ultimately maintained at a relatively high level (about 90%). At an initial pH value of 3.0, the sulfur atoms in functional groups of DTCR were protonated and those positively charged functional group generated Coloumbic repulsion against Hg2+ , which would reduce their binding capacities with Hg2+ [15]. On the contrary, the presence of OH− with increasing pH value would greatly strengthen the transformation of Hg2+ to Hg(OH)+ , and then to Hg(OH)2 , which settled down in the form of a precipitate. Thus the concentration of Hg2+ remaining in solution decreased [15], consequently leading to higher efficiencies in inhibiting Hg2+ reduction. 3.4. Effect of Cl− concentration Effect of Cl− concentration on inhibiting Hg2+ reduction and Hg0 re-emission by DTCR was shown in Fig. 5. Results indicated that the Hg2+ precipitation efficiency gradually declined with increase in Cl− concentration. The results might be due to the reaction between Hg2+ and Cl− and the formation of HgClx 2−x complex, which reduced Hg2+ exposure to organic sulfur and thereby lowered Hg2+ precipitation efficiency by DTCR [19]. The mercury 100
80
Hg (%)
2+
Hg precipitation 2+
Hg remained
40
20
40
60
80
100
CSO2(mM) 4
Fig. 4. Effect of initial pH values on Hg2+ reduction and Hg0 re-emission inhibited by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, DTCR = 0.14 mg l−1 and T = 50 ◦ C).
60
0
Fig. 6. Effect of sulfate concentration on Hg2+ reduction and Hg0 re-emission inhibited by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, Cl− = 100 mM, DTCR = 0.14 mg l−1 , T = 50 ◦ C and pH value = 5).
reduction rate was dependent on Cl− concentrations as shown in Fig. 5. It was worth noting that the Hg2+ reduction rate increased slightly first and then decreased. It was likely that in the absence of Cl− , the mercury precipitation was predominantly controlled by Hg–S complex to form insoluble chelates, while the introduction of Cl− to desulfurization solutions led to the competition against DTCR and generated HgClx 2−x complexes. Subsequently, due to the instability of HgClx 2−x complexes, more and more Hg2+ was released from the complexes, which became available for the reaction with S(IV). However, higher Cl− content decreased the Hg2+ reduction efficiency. According to Wo et al. [9], with increasing Cl− concentration, several intermediates including ClHgSO3 − and Cl2 HgSO3 2− were produced. Since Cl2 HgSO3 2− was very stable and Hg2+ was steadily combined in the complex, further Hg2+ reduction reaction was inhibited from the complex, which led to a decrease in Hg2+ reduction rate at higher Cl− concentrations. 3.5. Effect of sulfate concentration The SO4 2− concentration had a significant effect on inhibiting Hg2+ reduction and precipitation reactions by DTCR (as shown in Fig. 6). When the concentration of SO4 2− increased from 0 to 20 mM, Hg2+ precipitation gradually increased while the Hg2+ reduction ratio decreased from 14.18% to 2.18%. The efficiency of Hg2+ precipitation remained steady when SO4 2− concentration further increased. Consequently, more than 90% of Hg2+ was precipitated by DTCR. Liu et al. had reported the inhibition of Hg2+ reduction by SO4 2− . In the presence of S(IV), SO4 2− would react with Hg2+ to generate HgSO3 SO4 2− and thereby damping the formation of HgSO3 , whose decomposition was assumed to be the critical step of Hg2+ reduction [20].
2+
Hg reduction
3.6. Effect of nitrate concentration
20
0 0
20
40
60
80
100
120
140
160
Ccl-(mM) Fig. 5. Effect of Cl− concentration on Hg2+ reduction and Hg0 re-emission inhibited by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, DTCR = 0.14 mg l−1 , T = 50 ◦ C and pH value = 5).
Results shown in Fig. 7 indicated that NO3 − concentration had an adverse effect on Hg2+ precipitation while Hg2+ reduction rate saw a slight increase in the circumstance of DTCR. More specifically, with NO3 − concentration increasing from 0 to 100 mM, Hg2+ precipitation efficiency decreased from 89.7% to 83.3%. The strong oxidation potential of NO3 − in acid solutions might be a contributing factor for precipitation decrease. When pH value was at 5 and the NO3 − concentration was 100 mM, the electrode potential calculated was 0.7794 V, showing relatively strong oxidation ability. Thus, large
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R. Lu et al. / Journal of Hazardous Materials 196 (2011) 160–165
2+
from 0 to 10 mM, Hg2+ reduction rate increased from 3.42% to 6.15%, indicating that low concentration of Ca2+ promoted the removal of Hg2+ . The hypothesis was that the change in Ca2+ concentration affected numerous factors in the desulfurization solutions (such as SO3 2− , pH value), which in turn affected Hg2+ reduction efficiency. When Ca2+ and SO3 2− coexisted in the desulfurization solutions, some chemical reactions could be expressed as follows:
2+
Ca2+ + SO3 2+ → CaSO3 ↓
100
Hg (%)
80
60
Hg reduction Hg remained
40
2+
Hg
(3-1)
Hg2+ + SO3 2− HgSO3
precipitation
HgSO3 + SO3
20
2−
(3-2)
Hg(SO3 )2
Ca2+
2−
(3-3)
2−
quantities of functional groups in DTCR were prone to be oxidized by NO3 − , which inhibited the precipitation performance over Hg2+ . Therefore, Hg2+ remaining in the FGD solution benefited from the hampered precipitation rate and rose dramatically to over 5 times the initial residual rate, which had a positive effect on the increase of Hg2+ reduction. However, the rate of Hg2+ reduction did not have a similarly tremendous growth with Hg2+ that remained. This phenomenon could be ascribed to the strong oxidation potential of NO3 − which could react with S(IV) to inhibit Hg2+ reduction. Furthermore, from those data obtained above, there existed an urgent need for denitrification during flue gas desulfurization to prohibit further Hg0 re-emission.
reacted with SO3 to form slightly soluble CaSO3 , and the precipitation reaction made the equilibrium (3-3) move towards the left and reduced the chance of generating Hg(SO3 )2 2− . As a result, more HgSO3 was produced to shift the equilibrium (3-2) to the left, which in turn raised Hg2+ reduction rate. In contrast, when Ca2+ concentration was higher than 10 mM, Hg2+ reduction rate gradually decreased and white particles that could be seen with naked eyes were formed. The strengthened sedimentation effect was probably due to the acceleration of formation of less soluble CaSO3 along with increase in Ca2+ concentration. Because of the low solubility of CaSO3 in water, most CaSO3 was suspended in the solution. Hg2+ in the FGD solutions might be adsorbed to the surface of CaSO3 particles [21], which weakened Hg2+ reduction. Meanwhile, Ca2+ could not easily react with coordinated sulfur atoms in DTCR, leading to non-competition against Hg2+ . Furthermore, the chelation between DTCR and Hg2+ could carry a certain amount of negatively charged groups, resulting in mutual exclusion in the formed chelate and a decrease in flocculation. However, Ca2+ played a key role in double-layer compression and thereby accelerated flocculation during Hg2+ reduction and precipitation reaction, which ultimately contributed to the improvement of the Hg2+ precipitation rate [22].
3.7. Effect of Ca2+ concentration
4. Conclusion
As shown in Fig. 8, low concentrations of Ca2+ (<10 mM) had an adverse effect on inhibiting the Hg2+ reduction by DTCR while high concentrations of Ca2+ stimulated the reduction rate. The increase in Ca2+ concentration led to the gradual increase in Hg2+ precipitation rate by DTCR. More specifically, when Ca2+ concentration rose
DTCR was shown to be an effective agent to inhibit Hg2+ reduction and Hg0 re-emission in FGD liquors. More than 80% of Hg2+ was captured by DTCR, and Hg2+ reduction efficiency decreased to less than 10%. The experimental results indicated that increasing the precipitator dosage, the temperature (≤60 ◦ C), the initial pH value, or SO4 2− concentration could contribute to inhibition of Hg2+ reduction in simulated wet FGD solutions and Hg0 re-emission across a wet FGD scrubber by DTCR. Effects of Cl− , NO3 − , and Ca2+ concentration on Hg2+ reduction and Hg0 re-emission in FGD solutions in the presence of DTCR were complicated but still consistent with the law of conservation of mass. The general mechanism and control methods for Hg2+ reduction and Hg0 re-emission was discussed, and such exploration would provide a profound theoretical basis for industrial applications such as mercury restraint in wet FGD systems.
0 0
20
40
60
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CNO- (mM) 3
Fig. 7. Effect of nitrate concentration on Hg2+ reduction and Hg0 reemission restrained by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, Cl− = 100 mM, DTCR = 0.14 mg l−1 , T = 50 ◦ C and pH value = 5).
100
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Hg
Acknowledgments
20
This research is supported by the National High-Tech Research and Development Program of China (No. 2007AA06Z340) and the National Natural Science Foundation of China (No. 20877066).
0 0
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Fig. 8. Effect of Ca concentration on Hg reduction and Hg reemission restrained by DTCR (Hg2+ = 0.1 mg l−1 , S(IV) = 5 mM, Cl− = 100 mM, DTCR = 0.14 mg l−1 , T = 50 ◦ C and pH value = 5).
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[13] V.D. Benedini, P.A. Antunes, É.T.G. Cavalheiro, G.O. Chierice, Thermoanalytical and solution stability studies of hexamethylenedithiocarbamates, J. Braz. Chem. Soc. 17 (2006) 680–688. [14] Y.F. Liu, Y.J. Li, Y.M. Ni, Development in the research on DTC derivatives for heavy metal treatment, Elect. Pollut. Cont. 23 (2003) 1–3. [15] T.M. Tang, J. Xu, R.J. Lu, J.J. Wo, X.H. Xu, Enhanced Hg2+ removal and Hg0 reemission control from wet fuel gas desulfurization liquors with additives, Fuel 89 (2010) 3613–3617. [16] K. Ito, A.T. Ta, D.B. Bishop, A.J. Nelson, J.G. Reynolds, J.C. Andrews, Mercury L3 and sulfur K-edge studies of Hg-bound thiacrowns and back-extracting agents used in mercury remediation, Microchem. J. 81 (2005) 3–11. [17] X.H. Xu, Q.F. Ye, T.M. Tang, D.H. Wang, Hg0 oxidative absorption by K2 S2 O8 solution catalyzed by Ag+ and Cu2+ , J. Hazard. Mater. 158 (2008) 410–416. [18] T.J. Bellos, S.L Louis, Polyvalent metal cations in combination with dithiocarbamic acid compositions as broad spectrum demulsifiers, US Patent 6019912, 2000. [19] C. Acuna-Caro, K. Brechtel, G. Scheffknecht, M. Branb, The effect of chlorine and oxygen concentrations on the removal of mercury at an FGD-batch reactor, Fuel 88 (2008) 2489–2494. [20] Y. Liu, Y.J. Wang, Z.B. Wu, S.Y. Zhou, H.Q. Wang, A mechanism study of chloride and sulfate effects on Hg2+ reduction in sulfite solution, Fuel 90 (2011) 2501–2507. [21] H. Akiho, M. Nunokawa, H. Shiras, The behavior of gaseous mercury in a wet scrubber, in: Proceedings of the IV International Air Quality Conference, Mariott Crystal Gateway Arlington, VA, USA, 2003. [22] F.l. Fu, R.M. Chen, X. Ya, Application of a novel strategy-coordination polymerization precipitation to the treatment of Cu2+ -containing wastewaters, Sep. Purif. Technol. 52 (2006) 388–393.
Journal of Hazardous Materials 196 (2011) 166–172
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Decolorization characteristics and mechanism of Victoria Blue R removal by Acinetobacter calcoaceticus YC210 Chiing-Chang Chen a,1 , Chih-Yu Chen b,1 , Chiu-Yu Cheng c , Pei-Yi Teng c , Ying-Chien Chung c,∗ a b c
Department of Science Application and Dissemination, National Taichung University of Education, Taichung 403, Taiwan Department of Tourism, Hsing Wu Institute of Technology, Taipei 244, Taiwan Department of Biological Science and Technology, China University of Science and Technology, Taipei 115, Taiwan
a r t i c l e
i n f o
Article history: Received 9 January 2011 Received in revised form 21 August 2011 Accepted 4 September 2011 Available online 10 September 2011 Keywords: Acinetobacter calcoaceticus Biodegradation Dealkylation Victoria Blue
a b s t r a c t Acinetobacter calcoaceticus YC210 has been isolated and its ability to remove Victoria Blue R (VBR) from aqueous solution was assessed. The effects of various factors on decolorization efficiency were investigated in a batch system. The decolorization efficiency was found to be optimal within a pH of 5–7 and increased with VBR concentration up to 450 mg/l with high efficiency (94.5%) in a short time. The decolorization efficiency was significantly affected by cell concentrations. The decolorization of VBR by A. calcoaceticus YC210 followed first order kinetics. The apparent kinetic parameters of the Lineweaver–Burk equation, RVBR,max and Km , were calculated as 6.93 mg-VBR/g-cell/h and 175.8 mg/l, respectively. Based on the biodegradation products, VBR degradation by A. calcoaceticus YC210 involves a stepwise demethylation process to yield partially dealkylated VBR species. To our knowledge, this is the first report using microbes to remove VBR. It clearly demonstrates the dealkylation pathway of VBR degradation. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Triphenylmethane dyes are used extensively in the textile and fish farming industries, paper and leather industries, food and cosmetic industries, and in medicines [1]. In the study of Black et al. [2], it was noted that triphenylmethane dyes could be mutagenic and carcinogenic to biota. Some triphenylmethane dyes have been shown to be potent clastogens, possibly responsible for promoting tumor growth in some species of fish [3,4]. The discharge of triphenylmethane dye solution into the biosphere will negatively impact the exposed aquatic organisms. Such dyes are not easily degradable so their discharge will cause serious problems to the environment, and achieving legal purification levels is very difficult. Furthermore, triphenylmethane dyes could cause strong coloration and contribute to the organic load and toxicity of wastewater [5]. As there is a significant health risk to humans and aquatic organisms, it is very important to devise an efficient method to remove this substance from wastewater. A number of studies have been carried out in which in physicochemical methods, such as adsorption, precipitation, flocculation, oxidation, and photodegradation, have been used to treat triphenylmethane dyes [4]. However, none has been found to be completely successful because present methods have proved to
∗ Corresponding author. Tel.: +886 2 89116337; fax: +886 2 89116338. E-mail address:
[email protected] (Y.-C. Chung). 1 Equal contribution by these authors. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.015
be methodologically demanding, relatively inefficient, and timeconsuming [1,6]. Recent focus has therefore turned to biological processes as viable alternatives as such systems are environmentally friendly, cost-competitive, and produce less sludge [7]. The triphenylmethane dye, crystal violet has been successfully degraded by bacteria [7]. Theoretically, Victoria Blue R, another triphenylmethane dye, should be degraded by bacteria. However, there has been no published work on VBR decolorization or biodegradation by this means till now. Acinetobacter is a gram-negative genus of Bacteria belonging to the Gammaproteobacteria. Non-motile, Acinetobacter species are known to be involved in the biodegradation of aromatic pollutants such as biphenyl, chlorobiphenyl, and aniline [8]. Acinetobacter calcoaceticus, which belongs to the genus of Acinetobacter, was able to degrade crude oil such as alkanes and aromatics [9]. Although there are no published data currently available on its efficiency to biodegrade VBR, preliminary experiments carried out in our laboratory revealed that A. calcoaceticus is easily able to degrade VBR. In addition, A. calcoaceticus can be cultured easily in culture medium. The aim of the study reported here was to isolate VBR-degrading bacteria from contaminated soil near a wastewater treatment plant, determine the basic physiological characteristics of the isolated strain, and provide further details on A. calcoaceticus YC210 in terms of its capability to degrade VBR. The effects of initial concentration of VBR, pH, and cell numbers of A. calcoaceticus YC210 on VBR removal was evaluated, and the primary metabolic pathway during the biodegradation process was established.
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2. Materials and methods 2.1. Chemicals All the chemicals used in our experiment were analytical grade. Victoria Blue R was obtained from Sigma–Aldrich, Inc. and confirmed to be a pure organic compound through high-performance liquid chromatography (HPLC) analysis. The formula, molecular weight and solubility in water of Victoria Blue R are C29 H32 ClN3 , 457.5 g/mol, and 0.5%, respectively. Its absorption maximum occurs at 615 nm.
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by A. calcoaceticus YC210 in the culture medium. Organic intermediates were analyzed by means of a coupled HPLC–electrospray ionization mass spectrometry (ESI-MS) after readjustment of the chromatographic conditions to make the mobile phases (solvents A and B) compatible with the working conditions of the MS. Solvent A consisted of 25 mM aqueous ammonium acetate buffer (pH = 6.9) and solvent B, methanol. HPLC was carried out on an AtlantisTM dC18 column (250 mm × 4.6 mm, 5 m). The flow rate of the mobile phase was 1.0 ml/min. A linear gradient was run as follows: t = 0, A = 95, B = 5; t = 20, A = 50, B = 50; t = 35–40, A = 10, B = 90; t = 45, A = 95, B = 5. The column effluent was introduced into the ESI source of the MS.
2.2. Microorganisms and cultivation A. calcoaceticus YC210 was isolated from the soil near a wastewater sewage treatment plant in southern Taiwan. Each soil sample (10 g) was mixed with 200 ml sterile basal mineral medium [BMM (per liter distilled water): 4.8 g K2 HPO4 , 1.2 g KH2 PO4 , 1 g NH4 NO3 , 0.25 g MgSO4 ·7H2 O, 0.04 g CaCl2 , 0.001 g Fe2 (SO4)3 ] in a 350 ml flask and vortexed vigorously for 30 min. A 10 ml aliquot of each suspension solution was pippetted into a 350 ml flask containing 200 ml sterile BMM with 10 mg/l VBR and the solution mixed thoroughly. After 3 days of culture, 10 ml of culture solution was pippetted out and mixed with fresh BMM containing 20 mg/l VBR. The screening cycle was repeated until the major colony could be isolated by the spread plate method in BMM containing agar and VBR. Subsequently, following a 30-day acclimation period, the VBRdegrading bacterium was isolated from BMM containing 100 mg/l VBR. To identify the VBR-degrading bacterium, the cells of the dominant isolate were lysed and the DNA extracted. The 16S rRNA gene sequence of the dominant isolate was compared using BLASTN programs to search for nucleotide sequences in the NCBI website. A phylogenetic tree was constructed using the neighbor-joining method [10], and its topology was evaluated by bootstrap analyses of the neighbor-joining dataset using the SEQBOOT and CONSENSE options from the PHYLIP package. During the culture period, bacterial growth was monitored by spectrophotometry at 600 nm and by the spread plate method. 2.3. Batch decolorization system The batch decolorization experiments were conducted in 350 ml flasks containing BMM, VBR, and A. calcoaceticus cells. To analyze the effects of different environmental factors on the efficiency of A. calcoaceticus cells to decolorize VBR, we tested different initial VBR concentrations (8–450 mg/l), different concentrations of bacterial cells (104 –109 cfu/ml), and various pH values (5–9). The pH was adjusted using diluted solutions of NaOH and HCl. The batch experiments were run as a shaking culture at 30 ◦ C and 250 rpm, with 60 mg/l VBR and 108 cfu/ml of A. calcoaceticus cells, unless stated otherwise. The decolorization efficiency was analyzed after 150 min of reaction time. 2.4. Analysis of VBR degradation by A. calcoaceticus YC210 The amounts of VBR in the culture medium at the end of the various culture periods (=degradation times) were determined by high performance liquid chromatography (HPLC). The samples were concentrated by solid phase extraction (SPE) and the extract identified by HPLC. The HPLC system consisted of an Econosil column (5 m, 4.6 mm × 250 mm) and an isocratic mobile phase containing 70% (v/v) acetonitrile in an ammonium acetate solution (pH 5.5), max = 615 nm. The flow rate was 0.5 ml/min. A Waters ZQ LC/MS system equipped with the 1525 Binary HPLC pump, 2996 Photodiode Array Detector, 717 plus Autosampler and micromass-ZQ4000 Detector were used to identify the intermediates of VBR degraded
3. Results and discussion 3.1. Characterization of VBR-degrading bacteria The PCR amplification and sequencing procedures were according to Weisburg et al. [11] and resulted in the isolate being identified as A. calcoaceticus YC210 with 99.2% similarity (GenBank accession no. GU339280). The phylogenetic tree analysis of the isolate indicates that the isolate YC210 has a high similarity with A. calcoaceticus (GU339280) (Fig. 1): both isolate YC210 and A. calcoaceticus (GU339280) are in the same cluster. A. calcoaceticus was isolated in BMM containing VBR as the sole carbon source. This microorganism comprised non-fermentative aerobic gram-negative rods that are widely dispersed in nature [12]. A. calcoaceticus YC210 was observed to enter the late logarithmic growth phase after 18 h of culture in BMM containing 50 mg/l VBR. The optimal pH range for the growth of A. calcoaceticus was 5–7, as estimated by cell numbers (data not shown), and the optimal temperature range was 20–35 ◦ C (data not shown). These results are similar to those reported by Zhan et al. [13] who used A. calcoaceticus to degrade phenol. Consequently, in accordance with our results and their data, we performed all subsequent experiments at pH 7.0 and 30 ◦ C to optimize the activity of A. calcoaceticus YC210. 3.2. Effect of reaction time and initial VBR concentration on VBR removal To gain an understanding of the adsorption and biodegradation of VBR by A. calcoaceticus in terms of VBR removal, a batch removal study using autoclaved cells and living cells in BMM containing 10 mg/l MG was carried out. During the first 30 min, the efficiency of autoclaved cells and living cells for decoloring VBR was 7.5% and 85.0%, respectively (data not shown). The maximum decolorization efficiency of VBR by autoclaved cells only attained 12.6% (30–150 min), while that with living A. calcoaceticus increased with increasing reaction time up to 150 min, at which time the removal equilibrium was reached at a decolorization efficiency of 82.3 ± 0.5%. This indicated that autoclaved cells could remove dye by an adsorption mechanism but the major dye removal is probably biological oxidation by living A. calcoaceticus cells in the decolorization processes [14,15]. Because the decolorization efficiency of VBR by A. calcoaceticus exhibited an arc increasing curve between 30 and 150 min of the reaction time, we conducted a kinetic analysis. The results indicated that the biodegradation behavior of VBR by A. calcoaceticus YC210 fitted first order kinetics well: ln C = −0.0102t + 0.9243, R2 = 0.983. The decolorization efficiency of A. calcoaceticus YC210 was studied in media containing different concentrations of VBR. Fig. 2 shows the dependence of VBR concentration on the biological decolorization efficiency after 150 min of reaction time. When the initial dye concentration was increased from 8 to 450 mg/l, the decolorization efficiency of VBR increased as well – from 77.4 ± 1.5%
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Fig. 1. Bootstrap N-J tree of isolate YC210 (in this study) illustrating its high similarity to A. calcoaceticus (GU339280) based on 16S rDNA sequence comparisons. P. putida (GU048853) was used as an outgroup in this analysis. The scale bar indicates 1% sequence dissimilarity.
up to 94.5 ± 1.4%. The results indicated that the initial concentration of VBR possesses a strong driving force to overcome the mass transfer resistances between VBR and the A. calcoaceticus cells [16]. However, the decolorization efficiency of malachite green (a triphenylmethane dye, but whose chemical structure is simpler) by P. pulmonicola only achieved 92.7% after 210 min of reaction time [15]. Hence, we suggest that both the structure of the dye and the bacteria species have a significant influence on decolorization [17]. 3.3. Effect of bacterial cell concentration and pH on VBR removal The decolorization efficiency of VBR by A. calcoaceticus YC210 increased with increasing cell concentration of A. calcoaceticus from 104 to 108 cfu/ml (Fig. 3a). The optimal decolorization efficiency, 90.7 ± 0.6%, was found at 108 cfu/ml of A. calcoaceticus. However, an overabundance of bacterial cells (i.e. 109 cfu/ml) had no appre-
Fig. 2. Effect of VBR concentration on the decolorization efficiency of VBR by culture medium with A. calcoaceticus YC210. pH: 7.0; bacterial cells: 108 cfu/ml; 30 ◦ C. Results are means of triplicate experiments (SD is indicated with error bars).
ciable effect on the decolorization efficiency. Prior to degrade dye by microbes, the adsorptive efficiency of A. calcoaceticus was significantly increased in the concentration of 107 cfu/ml in which the higher VBR adsorption was developed and has led to higher decolorization efficiency. That is be regarded the characteristics of A. calcoaceticus. Similar relationships between decolorization efficiency and cell concentration have been reported by Shahvali et al. [14], who treated textile wastewater with Phanerochaete chrysosporium, and by Sani and Banerjee [18], who treated textile wastewater with Kurthia sp. However, Moosvi et al. [6] reported that there was no proportionate increase of decolorization with increase initial concentration of bacteria. Thus, the relationship between decolorization efficiency and cell concentration in a biological system seems to alter from case by case. In this study, 108 cfu/ml of A. calcoaceticus was the optimal cell concentration to decolorize VBR. The decolorization efficiency of VBR in culture medium with/without A. calcoaceticus (control) was determined to understand the decolorization behaviors at different pH values (pH 5–9). The results showed that culture media without A. calcoaceticus had an insignificant decolorization efficiency within the pH range 5–9 (data not shown). In the A. calcoaceticus YC210 system, the decolorization efficiency increased from pH 5 to 6, leveled off between pH 6 and 7 (91.1 ± 0.5%), and sharply decreased to 80.1 ± 1.3% at pH 8 (Fig. 3b). For the genus of Acinetobacter, the isoelectric point was at a pH of 5.8 [19]. At a lower pH (pH < 5.8), the H+ ions compete effectively with dye cations to adsorb on the cell surface, causing a decrease in color removal efficiency. At a higher pH (pH > 5.8), the surface of A. calcoaceticus changes to being negatively charged, which results in an increased electrostatic attractive force between the bacterial cell and the positively charged VBR molecule. This increased electrostatic force accounts for the significant increase in the decolorization efficiency of VBR at pH 6–7. As for the decrease in decolorization efficiency of VBR at pH 8–9 conditions, we suggested the optimal pH range for the growth of A. calcoaceticus YC210 is pH 5–7 that gives an additional explanation. Similar trends were found in the study of the decolorization of triphenylmethane and azo dyes by Pseudomonas sp. [20].
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95
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y = 25.373x + 0.1443 R2 = 0.9962
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Fig. 4. Relationship between VBR degradation 1/R and 1/C in the medium inoculated with A. calcoaceticus YC210. pH: 7.0; bacterial cells: 108 cfu/ml; 30 ◦ C. Results are means of triplicate experiments (SD is indicated with error bars).
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in biomass affinity for the target substrate [15]. These important parameters (RVBR,max and Km ) can be used for comparison if other biological systems are to be utilized in the removal of VBR.
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3.5. Separation and identification of VBR metabolism intermediates
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pH Fig. 3. (a) Effect of bacterial cell concentration on the decolorization efficiency of VBR by A. calcoaceticus YC210. VBR: 60 mg/l; pH: 7.0; 30 ◦ C. (b) Effect of pH value on the decolorization efficiency of VBR by A. calcoaceticus YC210. VBR: 60 mg/l; pH: 5–9; bacterial cells: 108 cfu/ml; 30 ◦ C. Results are means of triplicate experiments (SD is indicated with error bars).
3.4. Apparent kinetic analysis The VBR degradation rate in this system was calculated using the following equation derived from the Lineweaver–Burk plot: Km 1 1 1 = × + RVBR RVBR,max CVBR RVBR,max where RVBR (mg-VBR/g-cell/h) = VBR degradation rate; CVBR (mg/l) = VBR concentration; RVBR,max (mg-VBR/gcell/h) = maximum apparent VBR degradation rate; Km (mg/l) = apparent half-saturation constant. From the linear relationship between 1/CVBR (x) and 1/RVBR (y), RVBR,max and Km was obtained by calculating from the slope and intercept; the relationship is shown in Fig. 4. When the initial VBR concentration was increased from 8 to 498 mg/l, the VBR degradation rate increased significantly between 8 and 450 mg/l and then leveled off at 450–500 mg/l (data not shown). The regression equation in Fig. 4 is expressed as y = 25.373 x + 0.1443 (R2 = 0.9962) for the system inoculated with A. calcoaceticus YC210. The apparent kinetic parameters RVBR,max and Km are calculated to be 6.93 mg-VBR/gcell/h and 175.8 mg/l, respectively. The variations in the Km value for different substrates are mainly due to the affinity between cells and substrates. The decrease in Km suggests an enhancement
We isolated the intermediates of the VBR decolorization process using HPLC and ESI-MS and identified these by comparison with commercially available standards. Fig. 5 shows the VBR biodegradation intermediates formed by A. calcoaceticus YC210 after a 7-day culture period (sampling analysis from culture day-1 (upper), day3, day-4, day-6, and day-7 (down), respectively). Peak A is the parent VBR dye (retention time: 54.6 min), and peaks B–I (retention time: 37.2–52.4 min) are degradation intermediates. Nine components were successfully detected, all with the retention times less than 60 min. The maximum absorption band of each intermediate in the visible spectral region in Table 1 was measured. The intermediates were further identified using the HPLC–ESI mass spectrometric method; the relevant mass spectra are illustrated in Table 1. The intermediates appeared to be in their acidic forms based on their molecular peaks. Other intermediates may have been under the detection limit. From these results, intermediates can be classified into several groups. From the results of mass spectral analysis, we confirmed that the component A, m/z = 422.18, in the liquid chromatogram is VBR; the other components are B, m/z = 408.35; C, m/z = 394.76; D, m/z = 394.76; E, m/z = 394.25; F, m/z = 380.44; G, m/z = 380.16; H, m/z = 366.77; I, m/z = 366.04. The intermediates have the wavelength position of their major absorption bands moved toward the blue region, max , A, 613.0 nm; B, 606.9 nm; C, 603.2 nm; D, 609.4 nm; E, 619.2 nm; F, 598.3 nm; G, 608.1 nm; H, 587.3 nm; I, 545.8 nm. The intermediates may represent the N-de-alkylation of the VBR. 3.6. Product identification and proposed biodegradation pathways A sequential identification of primary and secondary metabolites enabled us to elucidate the metabolic pathway of VBR degradation by A. calcoaceticus YC210. The MS analysis was a powerful tool and enabled us to confirm that the components A–I,
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Fig. 5. Absorption spectra of the intermediates formed during the biodegradation process (7 days) of the VBR dye corresponding to the peaks in the HPLC chromatograph. Spectra were recorded using the photodiode array detector. The sampling analysis was from culture day-1 (upper), day-3, day-4, day-6, and day-7 (down).
respectively (data not shown). Fig. 6 presents a scheme of this degradation pathway. Each of the three side chains of the VBR dye molecules terminates in two dimethyl groups (compound A). B(or E) (compound B(E)) is obtained by the removal of one methyl(or ethyl) group from the VBR molecule; C–E, F–G and H–I (compounds C–I) correspond to three pairs of isomeric molecules with two to three (or one) fewer methyl (or ethyl) groups than the VBR dye molecule. C (compound C) is formed by removal of a methyl group from two different sides of the VBR molecule, whereas the other corresponding isomer in this pair, D (compound D), is produced by the removal of two methyl groups from the same side of the VBR structure. Another corresponding isomer in this pair, E (compound E), is produced by the removal of one ethyl group from the side of the VBR structure. In the second pair of isomers, F–G (compound F–G) is formed by the removal of three methyl groups from each side of the VBR molecule, whereas G (compound G) is produced by the removal of one methyl group from one side of the VBR structure while one ethyl group is removed from the other side of the
VBR structure. In the third pair of isomers, H (compound H) is produced by removal of four methyl groups from the VBR structure, whereas I (compound I) is formed by removal of two methyl and one ethyl groups from the VBR molecule. The results of the HPLC chromatograms, UV-visible spectra and HPLC–ESI mass spectra are summarized in Table 1. All of these analyses clearly reveal that VBR degradation by A. calcoaceticus YC210 is a stepwise dealkylation process in which the dominating mechanism of the initial step of VBR biodegradation is demethylation. A similar degradation mechanism has also been reported for crystal violet degradation by Pseudomonas putida [7]. These reductions may involve different enzymes and chemical reactions. Further studies will focus on the enzyme analyses to determine whether A. calcoaceticus YC210 has potential applications in the industrial degradation of VBR in wastewater systems. Furthermore, the stepwise demethylation process to yield mono-, di-, tri-, tetra-, and penta-dealkylated VBR species is similar to that reported by Mai et al. [21], in which VBR was degraded by TiO2 photocatalysis.
Table 1 Identification of the N-de-alkylated intermediates from the biodegradation of VBR by HPLC–ESI-MS. HPLC peaks
De-alkylation intermediates
[M+H+ ]
Absorption maximum (nm)
A B C D E F G H I
Bis (4-dimethylaminophenyl) (4-ethylaminonaphthenyl) methylium (4-Dimethylaminophenyl) (4-methylaminophenyl) (4-ethylaminonaphthyl) methylium (4-Dimethylaminophenyl) (4-aminophenyl) (4-ethylaminonaphthyl) methylium (4-Methylaminophenyl) (4-methylaminophenyl) (4-ethylaminonaphthyl) methylium (4-Methylaminophenyl) (4-dimethylaminophenyl) (4-aminonaphthyl) methylium (4-Methylaminophenyl) (4-aminophenyl) (4-ethylaminonaphthyl) methylium (4-Dimethylaminophenyl) (4-methylaminophenyl) (4-aminonaphthyl) methylium (4-Aminophenyl) (4-aminophenyl) (4-ethylaminonaphthyl) methylium (4-Dimethylaminophenyl) (4-aminophenyl) (4-aminonaphthyl) methylium
422.18 408.35 394.76 394.76 394.25 380.44 380.16 366.77 366.04
613.0 606.9 603.2 609.4 619.2 598.3 608.1 587.3 545.8
C.-C. Chen et al. / Journal of Hazardous Materials 196 (2011) 166–172
171
Fig. 6. The proposed degradation pathway of VBR by A. calcoaceticus YC210. The intermediates were identified by HPLC–ESI-MS. See Table 1 for the definition of the letters.
4. Conclusions Based on its decolorization efficiency and characteristics A. calcoaceticus YC210 is an efficient degrader of VBR. The decolorization efficiency is dependent on its initial concentration, cell concentration of A. calcoaceticus YC210 and pH. A. calcoaceticus YC210 can rapidly remove high concentrations of VBR and decolorize
it under slightly acid and neutral conditions. The dependence of the VBR degradation rate on VBR concentration is described by the Lineweaver–Burk equation, while HPLC–ESI-MS analysis indicates that the biodegradation of VBR by A. calcoaceticus YC210 is initiated via a demethylation process. The ability of A. calcoaceticus YC210 to decolorize VBR could have biotechnological applications.
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Acknowledgment The work was supported by Grant from the National Science Council. References [1] W. Azmi, R.K. Sani, U.C. Banerjee, Biodegradation of trimethane dyes, Enzyme Microb. Technol. 22 (1998) 185–191. [2] J.J. Black, M. Holmes, P.P. Dymerski, W.F. Zapisek, Fish tumor pathology and aromatic hydrocarbon pollution in a great lakes estuary, in: B.K. Afghan, D. Mackoy (Eds.), Hydrocarbons and Halogenated Hydrocarbon in the Aquatic Environment, Plenum Press, New York, 1980, pp. 559–565. [3] B.P. Cho, T. Yang, L.R. Blankenship, J.D. Moody, M. Churchwell, F.A. Bebland, S.J. Culp, Synthesis and characterization of N-de-methylated metabolites of malachite green and leucomalachite green, Chem. Res. Toxicol. 16 (2003) 285–294. [4] K.T. Chen, C.S. Lu, T.H. Chang, C.W. Wu, C.C. Chen, Comparison of photodegradative efficiencies and mechanisms of Victoria Blue R assisted by nafion-coated and fluorinated TiO2 photocatalysts, J. Hazard. Mater. 174 (2010) 598–609. [5] O. Demirbas, M. Alkan, M. Dogan, The removal of victoria blue from aqueous solution by adsorption on a low-cost material, Adsorption 8 (2002) 341–349. [6] S. Moosvi, H. Keharia, D. Madamwar, Decolourization of textile dye reactive violet 5 by a newly isolated bacterial consortium RVM 11.1, World J. Microbiol. Biotechnol. 21 (2005) 667–672. [7] C.C. Chen, H.J. Liao, C.Y. Cheng, C.Y. Yen, Y.C. Chung, Biodegradation of crystal violet by Pseudomonas putida, Biotechnol. Lett. 29 (2007) 391–396. [8] D. Abdel-El-Haleem, Acinetobacter: environmental and biotechnological applications, Afr. J. Biotechnol. 2 (2003) 71–74. [9] B. Lal, S. Khanna, Degradation of crude oil by Acinetobacter calcoaceticus and Alcaligenes odorans, J. Appl. Bacteriol. 81 (1996) 355–362. [10] N. Saitou, M. Nei, The neighbor-joining method: a new method for reconstructing phylogenetic trees, Mol. Biol. Evol. 4 (1987) 406–425.
[11] W.G. Weisburg, S.M. Barns, D.A. Pelletier, D.J. Lane, 16S ribosomal DNA amplification for phylogenetic study, J. Bacteriol. 173 (1991) 697–703. [12] H.F. Retailliau, A.W. Hightower, R.E. Dixon, J.R. Allen, Acinetobacter calcoaceticus: a nosocomial pathogen with an unusual seasonal pattern, J. Infect. Dis. 139 (1979) 371–375. [13] Y. Zhan, H. Yu, T. Yan, S. Ping, W. Lu, W. Zhang, M. Chen, M. Lin, Benzoate catabolite repression of the phenol degradation in Acinetobacter calcoaceticus PHEA-2, Curr. Microbiol. 59 (2009) 368–373. [14] M. Shahvali, M.M. Assadi, K. Rostami, Effect of environmental parameters on decolorization of textile wastewater using Phanerochaete chrysosporium, Bioprocess Eng. 23 (2000) 721–726. [15] C.Y. Chen, J.T. Kuo, C.Y. Cheng, Y.T. Huang, I.H. Ho, Y.C. Chung, Biological decolorization of dye solution containing malachite green by Pandoraea pulmonicola YC32 using a batch and continuous system, J. Hazard. Mater. 172 (2009) 1439–1445. [16] N. Daneshvar, M. Ayazloo, A.R. Khataee, M. Pourhassan, Biological decolorization of dye solution containing malachite green by microalgae Cosmarium sp., Bioresour. Technol. 98 (2007) 1176–1182. [17] J. Zhang, Y. Zhang, C. Li, Y. Jing, Adsorption of malachite green from aqueous solution onto carbon prepared from Arundo donax root, J. Hazard. Mater. 150 (2008) 774–782. [18] R.K. Sani, U.C. Banerjee, Decolorization of triphenylmethane dyes and textile and dye-stuff effluent by Kurthia sp., Enzyme Microb. Technol. 24 (1999) 433–437. [19] E. Pessione, S. Divari, E. Griva, M. Cavaletto, G.L. Rossi, G. Gilardi, C. Giunta, Phenol hydroxylase from Acinetobacter radioresistens is a multicomponent enzyme, Eur. J. Biochem. 265 (1999) 549–555. [20] P.L. Mali, M.M. Mahajan, D.P. Patil, M.V. Kulkarni, Biodecolourisation of members of triphenylmethane and azo groups of dyes, J. Sci. Ind. Res. 59 (1999) 221–224. [21] F.D. Mai, C.S. Lu, C.W. Wu, C.H. Huang, J.Y. Chen, C.C. Chen, Mechanisms of photocatalytic degradation of Victoria Blue R using nano-TiO2 , Sep. Purif. Technol. 62 (2008) 423–436.
Journal of Hazardous Materials 196 (2011) 173–179
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Effects of chloride ion on degradation of Acid Orange 7 by sulfate radical-based advanced oxidation process: Implications for formation of chlorinated aromatic compounds Ruixia Yuan, Sadiqua N. Ramjaun, Zhaohui Wang ∗ , Jianshe Liu ∗∗ College of Environmental Science and Engineering, Donghua University, Shanghai, 201620, China
a r t i c l e
i n f o
Article history: Received 15 June 2011 Received in revised form 22 August 2011 Accepted 4 September 2011 Available online 19 September 2011 Keywords: Azo dyes Sulfate radicals Decoloration AOX formation Chlorinated aromatic compounds
a b s t r a c t Sodium chloride is a common salt used during textile wet processes. Here a dual effect of chloride (i.e. inhibitory and accelerating effect) on azo dye (Acid Orange 7, AO7) degradation in an emerging cobalt/peroxymonosulfate (Co/PMS) advanced oxidation process (AOP) was reported. Compared to • OH-based AOPs, high concentrations of chloride (>5 mM) can significantly enhance dye decoloration independent of the presence of the Co2+ catalyst, but did greatly inhibit dye mineralization to an extent which was closely dependent upon the chloride content. Both UV–vis absorbance spectra and AOX determination indicated the formation of some refractory byproducts. Some chlorinated aromatic compounds, including 3-chloroisocoumain, 2-chloro-7-hydroxynaphthalene, 1,3,5-trichloro-2-nitrobenzene and tetrachlorohydroquione, were identified by GC–MS measurement in both Co/PMS/Cl− and PMS/Cl− reaction systems. Based on those experimental results, two possible branched (SO4 • − radical-based and non-radical) reaction pathways are proposed. This is one of the very few studies dealing with chlorinated organic intermediates formed via chlorine radical/active chlorine species (HOCl/Cl2 ) attack on dye compounds. Therefore, this finding may have significant technical implications for utilizing Co/PMS regent to detoxify chloride-rich azo dyes wastewater. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Dyeing industrial wastewater contains large amount of chloride ions with high non-biodegradable organic concentration, which can seriously inhibit the effectiveness of aerobic and anaerobic biological treatment of wastewater [1,2]. Recently, advanced oxidation processes (AOPs) are gaining significant importance in detoxification of dyeing wastewaters [3–5]. However, previous investigations have proved that the presence of chloride could make AOPs treatment of saline waters to be inefficient [6–8]. Since • OH (2.8 V) can oxidize Cl− to less reactive chlorine species, Cl2 /2Cl− (1.36 V) and HOCl/Cl− (1.48 V), the involvement of Cl− can inhibit the • OHbased radical chain reactions, thus significantly reducing the overall efficiency of the AOPs. Whereas our very recent study indicated that higher concentrations of chloride (>50 mM) in sulfate radical (SO4 •− ) based AOP could enhance the decoloration rate of an azo dye-AO7, although the underlying reaction mechanism has not been well elucidated [9].
SO4 •− as a major oxidizing specie with high standard redox potential (2.5–3.1 V), may react selectively via electron transfer with organics over a wide pH range (2–8), and completely destroy the pollutants present in wastewater or convert them into simple harmless compounds [10]. This emerging AOP has attracted great scientific and technological interest in environmental application. Many transition metals such as Co(II) and Fe(II) can efficiently activate peroxymonosulfate (PMS) to produce sulfate radicals as followings (Eqs. (1)–(8)) [11–14]: Co2+ + H2 O ↔ CoOH+ + H+ −
+
CoOH + HSO5 → CoO + SO4 +
+
CoO + 2H ↔ Co
0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.007
3+
•−
+ H2 O
+ H2 O
(4)
Co3+ + HSO5 − → Co2+ + SO5 •− + H+ 2SO5
•−
−
−
↔ O3 SOOOOSO3 →
{SO4 •− O2 SO4 •− }
→ O2 + S2 O8
(2) (3)
Co2+ + SO4 •− → Co3+ + SO4 2−
{SO4 •− O2 SO4 •− } ∗ Corresponding author. ∗∗ Co-corresponding author. E-mail addresses:
[email protected] (Z. Wang),
[email protected] (J. Liu).
(1)
+
(5)
{SO4 •− O2 SO4 •− }
(6)
2−
(7)
•−
(8)
→ O2 + 2SO4
From a thermodynamic point of view, both PMS and sulfate radical also can oxidize chloride ions to active chlorine species HOCl/Cl2
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or chlorine radicals. Chloride ions may be involved in PMS decomposition reaction via either non-radical pathways (Eqs. (9) and (10)) or sulfate radical-based pathways (Eqs. (11)–(16)) [15,16]. Cl− + HSO5 − → SO4 2− + HOCl −
−
+
2Cl + HSO5 + H → SO4
2−
SO4 •− + Cl− ↔ SO4 2− + Cl•
(9)
+ Cl2 + H2 O
(10)
kf = (3.2 ± 0.2) × 108 M−1 s−1 ;
kr = (2.1 ± 0.1) × 108 M−1 s−1
Cl• + Cl− ↔ Cl2 •−
(11)
Cl2
+ Cl2
•−
→ Cl2 + 2Cl
−
(12)
9
k = (2.1 ± 0.05) × 10 M
−1 −1
s
(13) Cl2 •− + H2 O → ClOH•− + H+ + Cl−
ClOH•− + H+ ↔ Cl• + H2 O
k[H2 O] < 100 s−1
(14)
kf = (2.6 ± 0.6) × 1010 M−1 s−1 ;
kr [H2 O] = (1.6 ± 0.2) × 105 s−1
ClOH•− ↔ • OH + Cl−
2.1. Materials Orange 7 (C16 H11 N2 O4 SNa) and Oxone® Acid ([2KHSO5 ·KHSO4 ·K2 SO4 ] salt, 95%) were purchased from SigmaAldrich. NaCl and CoSO4 ·7H2 O were of laboratory reagent grade and used without further purification. All sample solutions were prepared using deionized water from Barnstead UltraPure instrument. The salinity was varied by addition of sodium chloride (NaCl). Stock solutions of all chemicals were freshly prepared. Prior to each experiment, certain aliquots were transferred to the reactor vessel to obtain the specific concentrations.
kf = (7.8 ± 0.8) × 109 M−1 s−1 ;
kr = (5.7 ± 0.4) × 104 s−1
•−
2. Experimental
(15)
kf = (6.1 ± 0.8) × 109 s−1 ;
kr = (4.3 ± 0.4) × 109 M−1 s−1
(16)
These radical-based chain reactions and elementary reactions with unknown rate constants (e.g. Eqs. (1)–(8)) indeed complicate kinetic analysis in the Co/PMS/Cl− systems. In this context, identification of reaction intermediates may be an alternative choice for clarifying the underlying reaction mechanism. All reactive chlorine species, including chlorine atom, dichloride radical and free chlorine, can add to some unsaturated bonds of the compounds present during oxidation processes, generating chlorinated hydrocarbons, therefore leading to an undesirable increase of the parameter AOX (halogenated organic compounds adsorbable on activated carbon) [17]. In the Co/PMS/Cl− systems, it was observed that chlorine radicals may participate in phenol and 2,4-dichlorophenol transformation reactions [18]. Similarly, some chlorinated organic compounds or AOX are expected if Co/PMS is applied to decolor dye wastewater containing high concentration of Cl− . Kiwi et al. measured amount of AOX during Acid Orange 7 (AO7) oxidation in photoassisted and dark Fenton processes in the presence of Cl− anions, but did not give detailed information of reaction intermediates formed by mass spectral identification [8]. Also, information of both decoloration and mineralization of the dye wastewater in a wide salinity range is lacking. To improve the understanding of chloride effects on SO4 •− based treatment, AO7 was selected as a model nonbiodegradable azo dye to monitor its degradation and resulting intermediates, because it is a common dye used in industrial applications and laboratory studies and its degradation products in varied AOPs have been extensively characterized. The main objectives of this paper are (1) to test the effect of different concentration of chloride ions on the extent of decoloration and mineralization of AO7; (2) to study whether AOX can be formed during dye oxidation processes by chloride addition; and (3) to propose and compare the different dye degradation mechanisms under salt-free and hypersaline environment by detecting the intermediate products.
2.2. Experimental procedures All reactions were initiated by mixing appropriate concentrations of AO7, cobalt salt, chloride ion and Oxone® in this order immediately without the pH being controlled. The final experimental conditions, including the concentrations of PMS, Co2+ and AO7, have been set based on our preliminary experiments (please see Figs. SI-1 and SI-2) to obtain optimum conditions. The time scan mode of a Hitachi U-2900 spectrophotometer was used to in situ monitor the dye decoloration at 485 nm. A Shimazu TOC-VCPH analyzer was employed for TOC (total organic carbon) measurement. At the given reaction time intervals, samples taken from the solution for TOC analysis were immediately quenched with sodium nitrite. The AOX determinations were carried out by instrumental analysis (AOX, multi X® 2000, Jena, Germany) after enrichment on activated carbon (European Standard EN 1485 H14, 1996). GC/MS analysis was performed to identify the intermediate products formed during the AO7 degradation process. The experiments were carried out at two different chloride ion concentrations (without Cl− for 1# and in the presence of 200 mM Cl− for 2# and 3#). To determine whether the cobalt is involved in the degradation at extremely high concentration of chloride, the initial concentration of cobalt was different (in the presence of 0.2 mM Co2+ for 1# and 2#, without Co2+ for 3#). Prior to GC–MS analysis, the unbuffered solutions (100 mL) were quenched by 50 mL of NaNO2 solution (0.1 M) and filtered with 0.22 m membrane. Samples were pretreated using SPE (solidphase extraction), SPME (solid-phase microextraction) methods (details in Table SI-1) to extract and concentrate compounds of different polarity and volatility. Spectra were obtained with a gas chromatograph (Agilent 7890A), equipped with DB-5 MS capillary column (30 mm × 320 m × 0.5 m film thickness), interfaced directly to the mass spectrometer (5975A inert XL MSD with TripleAxis Detector) used as a detector. Electron impact (EI) mode at 70 eV was used and the mass range scanned was 30–400 m/z. The substance analysis was undertaken with reference to the NIST08 mass spectral library database. 3. Results and discussion 3.1. Effects of Cl− concentration on dye degradation rate Fig. 1 showed that in PMS/Cl− system AO7 was still rapidly degraded even in the absence of Co2+ ion, although the decoloration rate is lower than that with the catalyst. This Co-independent PMS decomposition suggests that PMS would react with chloride to produce active chlorine species HOCl/Cl2 (Eqs. (9) and (10)), which are able to decolor dyes rapidly. Since the amount of HOCl/Cl2 based on PMS decomposition by chloride was dependent upon chloride and PMS concentration, dye bleaching rate consequently increased exponentially with chloride and PMS content (Fig. SI-3).
R. Yuan et al. / Journal of Hazardous Materials 196 (2011) 173–179
175
3.2. Spectral change of AO7
Fig. 1. Effects of chloride ions on AO7 bleaching rate constants (k) in the solution. Experimental conditions: [AO7]0 = 0.1 mM; [PMS]0 = 0.5 mM; [Co2+ ]0 = 0.02 mM. Inset: plots of combined effect of chloride ions and catalyst Co2+ on the decoloration rate (k) with the initial concentration of AO7 0.1 mM and PMS 0.5 mM. All rate constants are pseudo-first order rate constants.
While in Co/PMS/Cl− system decoloration of AO7 declined significantly with increasing concentration of sodium chloride from 0.05 to 5 mM. Similar inhibitory effects have been reported in Ozonation [19], UV/H2 O2 [20] and UV/TiO2 processes [21]. However, further addition of Cl− (>5 mM) apparently accelerated dye degradation and the bleaching rate in presence of 500 mM Cl− was even larger than that without chloride ions. This dual effect of chloride on dyes bleaching has not been observed in other AOPs systems which are mainly based on • OH-radicals. In Co/PMS/Cl− system, the dye degradation can be attributed to three kinds of active species: (1) SO4 •− radicals (Eqs. (1)–(8)); (2) Cl• , Cl2 •− and ClOH•− (Eqs. (11), (12), (14) and (15)) and (3) HOCl and Cl2 (Eqs. (9) and (10)). The variation tendency of dye degradation rate (k) is not only related to the percentage of each oxidant species, but also to their reaction kinetic constants with dye. Consumption of SO4 •− radicals by Cl− and formation of less reactive chlorine radicals should be responsible for the adverse effect of chloride on Co/PMS performance. However, the formation of HOCl and Cl2 (Eqs. (9) and (10)) would make the dye bleaching rate higher with the addition of large amount of chloride ions similar to PMS/Cl− system. Therefore the dye degradation can be enhanced at higher chloride concentration. However, this dual effect of chloride has been not observed in H2 O2 -based AOPs such as Fenton and photo-Fenton systems [8], possibly because chloride alone could not activate H2 O2 to oxidize organics.
Representative UV/Vis spectra obtained from the dye solution as a function of reaction time are shown in Fig. 2. The absorption spectrum of AO7 in water is characterized by two bands in the visible region, with their maxima located at 485 nm and 430 nm corresponding to the hydrazone form and the azo form of the dye, respectively, and by two bands in the ultraviolet region located at 230 nm and 310 nm due to the benzene and naphthalene rings of AO7, respectively. It is observed that the absorbance at 485 nm almost completely disappears regardless of chloride ion, indicating the destruction of azo chromophore. In Fig. 2A, a new absorbance peak ( = 251 nm) appears at 2 min, but subsequently disappears after 30 min. However, in the presence of chloride, the intensities of the absorbance peaks at 230 nm and 310 nm decrease much more slowly with respect to that of the azo bond within 2 h. Previous reports have indicated that the ultraviolet bands located at 228 and 310 nm correspond to –* transitions in the benzoic and naphthalene rings of AO7, respectively [22]. It implies that some intermediates generated from chromophore cleavage may still contain benzoic- and naphthalene-type rings. Besides, there were some new peaks observed during the decomposition of the dye and their absorbance maxima increased continuously from 249 nm to 260 nm, which were not observed in the absence of chloride. These differences indicate the possible formation of new breakdown products like chlorinated aromatic compounds. Repeta et al. reported that chlorinated aromatic acids display strong absorption bands (240–400 nm) in the ultraviolet and near-ultraviolet regions [23]. Our following GC–MS analysis also proved that at least nineteen chlorinated compounds were produced during AO7 degradation, in accordance with spectral results. 3.3. Identification of reaction intermediates 3.3.1. AO7 intermediates (Co/PMS, No Cl− ) The reaction intermediates of AO7 without any chloride present in the reacting solution (Figs. SI-4 and SI-5) can be classified into six groups: (I) two naphthalene-type compound, 1-nitro-2-naphthalenol (1) and coumarin (2), which are primary degradation intermediate accompanying AO7 cleavage in the vicinity of the azo bond; (II) four fused heterocyclic compounds, such as 1(3H)-isobenzofuranone (3), 1,3-indanedione (4), phthalimide (5) and phthalic anhydride (6); (III) two aromatic compounds, such as phthalic acid (7), benzoic acid (8), which might be the degradation products of the partial oxidation of the aforementioned naphthalene-type and heterocyclic compounds, and their formation has been also reported by several previous studies concerning AO7 photocatalytic degradation [24,25]; (IV) two
Fig. 2. UV–vis spectral changes of AO7 in solutions containing PMS (5 mM) and Co2+ (0.1 mM) in the presence of different concentrations of chloride as a function of time. Experimental conditions: (A) [AO7]0 = 0.08 mM; [NaCl]0 = 0 mM; (B) [AO7]0 = 0.08 mM; [NaCl]0 = 200 mM, initial pH 6.5 (natural). Reaction times: (a) 0 min, (b) 2 min, (c) 30 min, and (d) 120 min.
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mono-heterocyclic compounds, including 1-nitrosopiperidine (9) and butyrolactone (10); (V) two alkane compounds, such as hexane (11) and 3-methylpentane (12); and (VI) one small molecule organic acid compound, formic acid (13), whose further oxidation yielded CO2 , H2 O. The major intermediates detected from the sulfate radical attack on AO7 were similar with those identified in hydroxyl radical oxidation processes such as the Fenton reagent [25], sonolytic oxidation [26] and photocatalysis [22,24]. Other reaction intermediates, such as 2-naphthnol and p-benzoquinone proposed by Velegraki et al. [26] and other investigators [25], were not detected in our study and this might have been due to the different degradation mechanisms involving in different radicals as well as to the different experimental conditions and analytical techniques employed in various studies. (Co/PMS/Cl−
PMS/Cl− )
3.3.2. AO7 intermediates & The intermediates detected under a chloride-rich environment not only included the oxidation products measured in Co/PMS system, but also included several chlorinated organic products, regardless of the presence of Co2+ (Figs. SI-6–SI-11). These chlorinated intermediates can be grouped into four kinds: (I) five naphthalenetype compounds, including 4-chloro-1,2-naphthalenedione (14), 2,4-dichloro-1-naphthalenol (15), 6-chlorochromone (16), 3chloroisocoumarin (17), and 3,4-dichloroisocoumain (18); (II) two heterocyclic compounds, including 5-chloroisobenzofuran-1,3dione (19) and 5-chlorobenzofuran (20); (III) twelve aromatic compounds, such as 4-chlorophthalic acid (21), 1,2-benzenedicarbonyl dichloride (22), 1,2-dichlorobenzene (23), 1,4-dichlorobenzene (24), 1,3,5-trichlorobenzene (25), 1,2-dichloro-3-nitrobenzene (26), 1,3,5-trichloro-2-nitrobenzene (27), 2,4-dichlorophenol (28), 2,5-dichlorohydroquinone (29), 2,4,6-trichlorophenol (30), 2,3,5,6tetrachlorophenol (31) and tetrachlorohydroquione (32); and (IV) one low molecule chlorinated product: tetrachloromethane (33). Some of these intermediates mentioned above such as 2,4-dichlorophenol, 2,4,6-trichlorophenol, tetrachlorohydroquione and tetrachloromethane have been identified as the phenol degradation products in Co/PMS/Cl− system [18]. To our best knowledge, none of previous studies dealing with dye degradation gives information on these chlorinated by-products. Such information is only available for several experiments with much simpler molecules. To quantify total chlorinated by-products, AOX value was tested. 3.4. AOX formation AOX is an important parameter for wastewaters from cleaning agents and disinfectants. Photochemical reactions are capable of transforming Cl− into Cl• radical, leading to the chlorination of organic compounds [8]. Chlorine atoms are electrophilic (the element is electronegative, and Cl• will readily take up an electron to complete its octet) and thus readily add to the double bond of organic compounds. Chlorination by HOCl/Cl2 can also generate AOX [17,19]. To determine the effect of chloride concentration on AOX formation, experiments were carried out at different chloride ion concentrations (0 mM Cl− , 0.2 mM Cl− and 100 mM Cl− ). Fig. 3 presents results of AOX measurement under different experimental conditions. Prior to oxidative treatment, no AOX was detected as expected, due to the absence of chlorine atom in the AO7 molecular structure. AOX type products were formed once the dye in aqueous solution was mixed with 0.1 mM Co2+ , 10 mM PMS and 0.2 mM Cl− . The amount of AOX formed was dependent upon reaction time and the chloride content in solution. The increase of Cl− concentration led to an increase in the yields of AOX as seen in Fig. 3. The maximum AOX production was 3.5 mg/L in the presence of 100 mM
Fig. 3. AOX concentration of the dye wastewater during the Co/PMS treatment ([AO7]0 = 0.2 mM; [PMS]0 = 10 mM; [Co2+ ]0 = 0.1 mM) containing different concentrations of chloride.
Cl− , comparable to those detected in • OH-based Fenton oxidation processes of AO7 [8]. AOX formation in the presence of Cl− can be explained by reactions (17)–(19): SO4 •− + AO7 → R •+ + SO4 2− R •+
+ Cl2
•−
→ RCl + Cl
−
Cl2 /HClO + AO7 → RCl + H2 O
(17) (18) (19)
Sulfate/chlorine radicals attack on the aromatic ring of dye itself or its byproducts, or direct chlorination by Cl2 /HClO yields RCl type organohalogens. The formation of AOX has already been observed in other AOPs systems such as Fenton reaction [8] and UV/H2 O2 oxidation [17]. Kiwi et al. [8] reported that Fenton and photo-Fenton degradation of AO7 may produce chlorinated byproducts due to the radical chain reactions of chloride ion with hydroxyl radicals, although their molecular structures have not been well identified. Baycan et al. [17] observed a de novo synthesis of AOX in the presence of 1000 mg/L or 10000 mg/L Cl− during UV/H2 O2 oxidation of acetone and sodium dodecyl sulfate (ABS). It should be pointed out that AOX only represents a fraction of the chlorinated organic compounds. They are considered recalcitrant and expected to be completely mineralized with longer oxidative treatment. 3.5. AO7 mineralization All experimental results obtained from UV–vis spectroscopy, GC–MS and AOX measurements show that chlorinated organic intermediates were generated when certain concentrations of chloride were present in Co/PMS systems. Therefore, using TOC values rather than decoloration rates seems more reasonable to evaluate the degradation efficiencies of Co/PMS/Cl− system. The approximate stoichiometry of AO7 mineralization in the presence of Oxone® can be expressed as: 2C16 H11 N2 NaO4 S + 37(2KHSO5 ·KHSO4 ·K2 SO4 ) → 32CO2 + 10H2 O + 2N2 + NaHSO4 + 111KHSO4 + 37K2 SO4
(20)
The stoichiometric amount of Oxone® needed to mineralize 0.2 mmol of AO7 is 3.7 mmol (less than Oxone® dosage used in this study). Therefore, complete mineralization of AO7 should be expected under the present experimental conditions (AO7, 0.2 mM; PMS, 10 mM; Co2+ , 0.1 mM). As shown in Fig. 4, 81.8% of AO7 could be mineralized without the addition of NaCl after 6 h oxidation treatment. However, under similar conditions but in the presence of salt (0.2–200 mM of NaCl), TOC removal was reduced with increasing chloride concentrations. For example, the decrease of TOC was only 6.65% with a chloride concentration of 10 mM.
R. Yuan et al. / Journal of Hazardous Materials 196 (2011) 173–179
177
less reactive inorganic radicals (e.g. Cl• /Cl2 •− ); and (2) generation of more refractory organohalogen compounds as identified with GC–MS and AOX measurements. 3.6. Proposed reaction mechanism Based on the kinetic results and degradation products identified by GC/MS methods and previous researches [25,27], two possible branched pathways for AO7 degradation by sulfate radicals under a chloride-rich environment are proposed.
Fig. 4. TOC removal of dye wastewater containing different chloride concentrations after 6 h oxidation treatment. Inset: influence of chloride ions on mineralization of dyes. Experimental conditions: [AO7]0 = 0.2 mM; [Co2+ ]0 = 0.1 mM; [PMS]0 = 10 mM.
AO7 was hardly mineralized after 6 h oxidation when the concentration of NaCl was up to 200 mM, although complete decoloration could be reached within 1 h. The inhibitory effects of chloride on dye mineralization would be attributed to (1) the consumption of the oxidant PMS and SO4 •− radicals by Cl− ions, resulting to the
3.6.1. SO4 •− radical-based reaction pathway The sulfate radical is a primary oxidizing species via Co2+ mediated PMS reduction, and may react with dye molecule via addition to the aromatic ring followed by hydrolysis. The hydroxylation intermediates are the major products formed by sulfate radical attack on aromatics [28]. The overall decomposition of AO7 dye via sulfate radical-based pathway is illustrated in Fig. 5. The oxidation decomposition of AO7 can be described by a series of consecutive degradation steps. AO7 was firstly decomposed to aromatic intermediates, further oxidized to ring opening products and organic acids, and finally mineralized to CO2 , H2 O and inorganic salts as pointed out in the literature [25,27]. Sulfonated molecules are non-volatile and highly soluble in water therefore cannot be analyzed by GC/MS. After 120 min of oxidation, the reaction
Fig. 5. SO4 • − radical-based reaction pathway of AO7 oxidative degradation using Co/PMS system in the presence of chloride ions. The identified byproducts by GC–MS were numbered in parentheses (see details in Supporting Information).
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Fig. 6. Sulfate/chlorine radicals attack on degradation intermediates of AO7.
mixture nearly exclusively consists of aliphatic compounds presumably due to further oxidation of AO7 and its ring intermediates. In the presence of chloride, the secondary chlorine radicals such as Cl• , Cl2 •− and ClOH•− , were generated by scavenging SO4 •− radical by Cl− ion [29]. These chlorine radicals are generally less reactive than the primary SO4 •− radical, leading to slower degradation of AO7 dye. Cl• /Cl2 •− may react with dye and its intermediates by one-electron oxidation, H-abstraction and addition to unsaturated C–C bonds [30]. A termination reaction of chlorine radical with AO7 oxidation intermediates radicals, such as quinine radical, leads to the formation of the chlorinated derivatives as detected from the transformation of phthalic anhydride provided in Fig. 6. The first step in all cases is the addition of a sulfate radical to the aromatic ring of byproducts of dye decomposition [18]. The formation of the corresponding cation radicals take place via elimination of the sulfate group, followed by chlorination by Cl2 •− , as shown in more details in Fig. 6.
3.6.2. Non-radical reaction pathway The halide ions (e.g. Cl− , Br− and I− ) are directly oxidized through two-electron transfer producing reactive halogens (e.g. Cl2 /HClO) (Eqs. (9) and (10)) [31]. Chlorine in the form of a liquid or gas has been effectively used for the color removal processes of many dyes [32]. The first step of aquatic chlorination is electrophilic substitution on the aromatic ring, depending on the combined resonance and inductive effects of the ring substituents. HClO has higher instability due to pronounced ionic nature and thus more reactivity towards the aromatic nucleus [33]. Chlorination of anisole, o-methoxybenzoic acid, ethylbenzene [34] and other phenolic compounds [35] led to formation of notable array of organochlorines. Narender et al. [33] reported the selective oxychlorination of aromatic compounds using KCl as a chlorine source and Oxone® as an oxidant. They found that introduction of an electron donating group on the aromatic ring substantially increased the rate of ring chlorination while on electron-withdrawing group decreases it. It is noteworthy that the identified chloride-containing intermediates by Cl2 /HClO chlorination were very similar to those detected in Co/PMS/Cl− systems (see Figs. SI-6 and SI-7). It should be noted that many chlorinated aromatic pollutants generated during the dye bleaching process are recalcitrant and have long half-lives. Some of them show a tendency to bioaccumulate due to their hydrophobicity and excellent ability to penetrate cell membranes while some are proven carcinogens and mutagens [36]. It is unlikely that the risk of yielding more toxic compounds
would be deemed worth the effectiveness that Co/PMS reagent represents. If these noncolored azo dye byproducts are released into water bodies as a discharge effluent, it will pose a high toxicological risk to exposed humans and aquatic biota [37].
4. Conclusions Effects of chloride ion on degradation of an azo dye (Acid Orange 7) by the emerging cobalt/peroxymonosulfate (Co/PMS) advanced oxidation process were investigated in this study. A dual effect of chloride (i.e. inhibitory and accelerating effect) on dye bleaching kinetics was firstly observed. In contrast to the Fenton-based AOPs, chloride ions with high concentrations (>5 mM) can greatly increase the decoloration efficacy of AO7, regardless of the presence of the Co2+ catalyst. However, high dosage of chloride did significantly inhibit dye mineralization to an extent which was closely dependent upon the initial chloride concentrations. According to the results obtained from UV–vis spectroscopy, AOX determination and GC–MS measurements, some undesirable aromatic chlorinated compounds were generated in the presence of 200 mM NaCl, independent of the addition of cobalt catalyst. These interesting results above can be attributed to two possible branched (SO4 •− radicalbased and non-radical) pathways for AO7 degradation by sulfate radicals under a chloride-rich environment. The present study indicates that the appreciable levels of salts in wastewater from the textile industries may reduce the level of dye mineralization and even lead to the formation of more toxic chlorinated compounds during AOPs treatment. Therefore, attempts to develop strategies for circumventing the adverse effects of chloride in Co/PMS system should be urgently taken prior to the large scale application of this emerging advanced oxidation technology.
Acknowledgments The authors gratefully acknowledged the financial support from the Fundamental Research Funds for the Central Universities (2010B04-04-1), the Shanghai Leading Academic Discipline Project (B604) and Shanghai Tongji Gao TingYao Environmental Science & Technology Development Fund. This work was also partially supported by the priming scientific research foundation for the junior teachers of Donghua University (No. 113-10-0044049) and the PhD thesis innovation foundation of the Donghua University (No. 11D11311).
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Journal of Hazardous Materials 196 (2011) 180–186
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Alkaline modified oil shale fly ash: Optimal synthesis conditions and preliminary tests on CO2 adsorption Janek Reinik a,∗ , Ivo Heinmaa a , Uuve Kirso a , Toivo Kallaste b , Johannes Ritamäki e,g , Dan Boström c , Eva Pongrácz d,e , Mika Huuhtanen e , William Larsson g , Riitta Keiski e , Krisztián Kordás f,g , Jyri-Pekka Mikkola g,h a
National Institute of Chemical Physics and Biophysics, Akadeemia tee 23, EE-12618, Tallinn, Estonia Institute of Geology at Tallinn University of Technology, Ehitajate tee 5, EE-19086, Tallinn, Estonia c Energy Technology and Thermal Process Chemistry ETPC, Department of Chemistry and the Department of Applied Physics and Electronics, Chemical-Biological Center, Umeå University, SE-90187, Umeå, Sweden d Thule Institute, NorTech Oulu, University of Oulu, P.O. Box 8000, FI-90014 University of Oulu, Finland e Mass and Heat Transfer Process Laboratory, P.O. Box 4300, FI-90014 University of Oulu, Finland f Microelectronics and Materials Physics Laboratories, EMPART Research Group of Infotech Oulu, University of Oulu, P.O. Box 4500, FI-90014 Oulu, Finland g Technical Chemistry, Department of Chemistry, Chemical-Biological Center, Umeå University, SE-90187, Umeå, Sweden h Industrial Chemistry and Reaction Engineering, Process Chemistry Center, Åbo Akademi University, FI-20500, Åbo-Turku, Finland b
a r t i c l e
i n f o
Article history: Received 18 July 2011 Received in revised form 1 September 2011 Accepted 4 September 2011 Available online 10 September 2011 Keywords: Oil shale fly ash Tobermorite synthesis CO2 thermo-gravimetric analysis
a b s t r a c t Environmentally friendly product, calcium–silica–aluminum hydrate, was synthesized from oil shale fly ash, which is rendered so far partly as an industrial waste. Reaction conditions were: temperature 130 and 160 ◦ C, NaOH concentrations 1, 3, 5 and 8 M and synthesis time 24 h. Optimal conditions were found to be 5 M at 130 ◦ C at given parameter range. Original and activated ash samples were characterized by XRD, XRF, SEM, EFTEM, 29 Si MAS-NMR, BET and TGA. Semi-quantitative XRD and MAS-NMR showed that mainly tobermorites and katoite are formed during alkaline hydrothermal treatment. Physical adsorption of CO2 on the surface of the original and activated ash samples was measured with thermo-gravimetric analysis. TGA showed that the physical adsorption of CO2 on the oil shale fly ash sample increases from 0.06 to 3–4 mass% after alkaline hydrothermal activation with NaOH. The activated product has a potential to be used in industrial processes for physical adsorption of CO2 emissions. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Oil shale processing residues pose a problem for the oil shale industry. Only in Estonia the power plants based on firing of oil shale produce 5–7 million tons of ash annually [1]. Oil shale ash is rendered so far mainly as a hazardous industrial waste because of high alkalinity and content of trace elements [2,3]. Less than 10 per cent of oil shale ash produced in Estonian power plants is utilized at the moment, mainly oil shale fly ash in cement industry [4]. However, the concentrations of heavy metals (e.g. Pb, Cd, Cr, Zn) in fly ash are higher than those in bottom ash and utilizing fly ash for construction materials can cause secondary pollution [5]. The focus of current work was to investigate possibilities for new industrial applications to oil shale fly ash. Oil shale fly ash from Estonian power plants has a very high content of CaO (around 50%) and relatively high content of SiO2 (ca.
∗ Corresponding author. Tel.: +372 639 8356; fax: +372 670 3662. E-mail addresses: jreinik@kbfi.ee, janek.reinik@kbfi.ee (J. Reinik). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.006
22%) and Al2 O3 (ca. 5%) by mass weight [4]. However, it is important to remember that the composition of oil shale fly ash vary, depending on combusted oil shale. In the 1980s, German scientists invented a method for producing zeolites from coal fly ash [6]. Using the same hydrothermal activation method for Ca-rich oil shale ash, calcium–silica–aluminum (CSA) hydrates containing mainly tobermorites have been synthesized [7,8]. Tobermorites are core-binding phase in concrete [9] and can also be used as non-asbestos fire resistant materials [10]. Other uses of tobermorites include cation exchangers for cleaning low-level nuclear waste and adsorbents [11,12]. In the structure of tobermorites, the Al3+ ions are isomorphously substituted for Si4+ ions in tetrahedral sites, which in turn give to this structure a surplus negative charge. Na+ ions are placed in the interlayer and compensate the deficient charge. Na+ ions are mobile and easily pass into solution when substituted by heavy metal ions (Co2+ , Ni2+ , Zn2+ , etc.) [13,14]. Taking account the immobilizing properties of tobermorite structure in activated oil shale ash we assume that the release of heavy metals should diminish compared to original ash, witch in turn would make the product environmentally safer.
J. Reinik et al. / Journal of Hazardous Materials 196 (2011) 180–186
Tobermorite occurs naturally, e.g. in Tobermory, Scotland [15] and tobermorites have been synthesized from a range of parent materials and industrial by-products [16–18]. Part of the work focused on finding optimal conditions for the synthesis of tobermorites from oil shale fly ash. The CSA hydrates from oil shale fly ash were synthesized at temperatures of 130 and 160 ◦ C and NaOH concentrations of 1, 3, 5 and 8 M, respectively. The solid/solution ratio (350 g/dm3 ) and reaction time (24 h) were constant throughout the experiments. The second objective of the work was to find a new application to synthesized product. We investigated the possibility of using the product for physical adsorption of CO2 . Fossil CO2 emissions are considered a problem encountered in many industrial processes [19]. Adsorption is one of the more promising technologies for capturing CO2 from exhaust gases, potentially avoiding the shortcomings of aqueous amine systems [20,21]. Several research studies have been made to evaluate the performance of various sorbents for CO2 adsorption from flue gases, e.g. zeolites [22], carbon based materials (activated carbon and molecular sieve carbon) [23], hydrotalcite-like compounds and metal oxides (e.g. CaO and MgO) [24]; polymer based materials [25]; magnesium double salts, e.g. K2 Mg(CO3 )2 [26] and also ionic liquids [27]. There are several ways to enhance the performance of an adsorbent, e.g. by means of chemical modification by addition of alkalis [28]. Moreover, several different CO2 adsorption technologies are available, differing by the regeneration technology applied, such as pressure swing, vacuum swing, temperature swing, electric swing adsorption or combination of these regeneration technologies [29]. In the present work, laboratory scale tests of cyclic CO2 adsorption on activated oil shale ash were conducted via thermal desorption to evaluate their repeated availability for CO2 capture.
2. Materials and methods Oil shale fly ash was collected from the 1st unit of electric precipitators of Estonian Power Plant’s boiler (Narva Power Plants Ltd.) operating on circulating fluidized bed (CFB) principle. For the activation of oil shale ash the same reactor set-up and procedure was used as described in authors’ previous work [8], except the activated product was washed only one time in distilled water. X-ray diffraction patterns of original and activated ash samples were recorded with a HZG4 diffractometer (Freiberger Präzisionsmechanik, former DDR) with scintillation detector. Diffraction patterns of Co Ka were registered in the range of 5–65◦ 2-theta. Semi-quantitative XRD analyze of one activated oil shale ash sample (activated in 8 M NaOH solution at 160 ◦ C for 24 h) was performed with Rietveld technique using DIFFRACplus TOPAS R 2.1 (Bruker AXS GmnH, Karlsruhe, Germany, 2003). Structures from Inorganic Crystal Structure Database (Fachinformationszentrum, Karlsruhe, Germany, 2009) were used as starting models for the phases identified in the sample. Also, semi-quantitative XRD analysis was made using a Bruker d8 Advance instrument in − mode, with an optical configuration consisting of a primary Göbel mirror and a Våntec-1 detector. Continuous scans were applied on the sample. By adding repeated scans, the total data collection time lasted for 6 h. The PDF2 databank (ICDD, Newtown Square, PA, 2004) together with Bruker software was used to analyze the diffraction patterns. The chemical composition of major elements in original and activated ash samples was analyzed by the wavelength dispersive X-ray fluorescence (XRF) spectrometer S4 Pioneer (Bruker AXS GmbH) using integrated standardless evaluation technique (SPECTRAplus ). 29 Si MAS-NMR (Magic Angle Spinning-Nuclear Magnetic Resonance) spectra of original and activated samples was recorded on Bruker AMX-360 spectrometer at 8.5 T external magnetic field,
181
using a bespoke MAS probe and 10 mm od zirconia rotors (rotation speed 5 kHz, simple 90◦ pulse excitation). About 400 accumulations with recycle time of 200 s were used to get reasonable signal to noise ratio. Specific surface area was measured with Micrometrics TriStar 3000 sorptometer (Micrometrics Instrument Corp. USA). The samples (ca. 0.3 g) were degassed at temperature 120 ◦ C for 5 h and before installing into sorptometer. Analysis adsorptive was N2 and bath temp. −196 ◦ C. Isotherm data, BET surface area, t-plot, pore volumes were processed with TriStar 3000 v. 6.07 software. SEM and EFTEM imaging for original ash and a characteristic sample activated in 4 M NaOH solution at temp. 130 ◦ C were conducted with Zeiss FE-SEM Ultra Plus and Zeiss EFTEM (energy filtering transmission electron microscope, Carl Zeiss 201), respectively. Thermo-gravimetric analysis for CO2 adsorption was conducted with TGA Q 5000 (TA Instruments-Waters LLC, USA) supplied with Advantage for Q Series and Thermal Advantage (release 4.6.9) software. The method details were as follows: 1. The sample was dried at 120 ◦ C for 100 min in N2 environment. 2. The sample was equilibrated at 40 ◦ C. 3. The gas flow was switched to a 70% CO2 at flow rate 40 mL/min for 120 min. 4. The gas flow was reversed to pure N2 and the sample was equilibrated at 115 ◦ C and kept isothermal for 90 min. 5. The whole sequence was repeated. In the thermo-gravimetric cyclic tests, the adsorption took place at the temperature of 40 ◦ C (60 min) and desorption at the temperature of 115 ◦ C (30 min). The cycle was repeated for 2 times. 3. Results 3.1. Nitrogen physisorption For characterizing the product, the specific surface area was calculated from the N2 absorption–desorption isotherm (P/P0 = 0.025–0.999) using the B.E.T. equation [30]. The estimated constant C in BET equation was ranging from −6200 to 1000 in many sorption analyses, which means that in addition to mesopores the material also contained micropores. Therefore the t-plots and BJH cumulative pore specific surface area; cumulative pore volume and average pore diameter were also estimated (Table 1). 3.2. X-ray fluorescence (XRF) analysis The chemical composition of original and activated ash (8 M NaOH solution at 160 ◦ C) samples is presented in Table 2. The XRF analysis shows the increased concentration of Na+ ions in the product. Excess of Na+ can be removed by thorough washing of the activated material [31], which was not the objective of the work. 3.3. X-ray Roentgen diffraction (XRD) analysis The X-Ray Roentgen diffraction patterns of the original and activated ash samples are presented in Fig. 1 and Fig. 2, respectively. The XRD analyses clearly demonstrate that the quartz in original ash is gradually converted into calcium–aluminum–silicate hydrates, depending on the NaOH concentration. Tobermorite 1 in Figs. 1 and 2 refers to reference pattern of Ca2.25 [Si3 O7.5 (OH)1.5 ]·H2 O orthorhombic – a = 5.5860 | b = 3.6960 | c = 22.7790 [32] and tobermorite 2 refers to reference pattern Ca4.5 Si6 O16 (OH)·5H2 O monoclinic – a = 6.7320 | b = 7.3690 | c = 22.6800 | gamma = 123.1800 [33].
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J. Reinik et al. / Journal of Hazardous Materials 196 (2011) 180–186
Table 1 Results of nitrogen physisorption analysis. Surface area (m2 /g)
Samplea
Original ash 1M130 3M130 5M130 8M130 1M160 3M160 5M160 8M160 a
Pore volume (mm3 /g)
Average pore size (nm)
BET
t-plot micro-pore
BJH cumulative of pores between 1.7 and 100 nm
t-plot micro-pore
BJH cumulative of pores between 1.7 and 300 nm
BET
BJH
6.3 38 59 61 70 58 50 47 58
0.77 5 14 14 18 11 11 10 14
7.3 38 58 61 67 59 50 45 57
0.35 2 7 7 9 5 5 5 7
36 143 264 269 360 236 196 250 310
23 15 18 18 21 16 16 22 21
20 15 18 18 21 16 16 22 22
Sample name refers to activation conditions (1M130 means: ash sample activated in 1 M NaOH solution at the temperature of 130 ◦ C, etc.).
Table 2 Chemical composition of major compounds in original fly ash and ash sample activated in 8 M NaOH solution at 160 ◦ C. Component
Original ash (wt%)
Activated ash, 8 M, 160 ◦ C (wt%)
CaO SiO2 Al2 O3 Fe2 O3 MgO Na2 O
24 31 9.3 3.9 3.2 0.1
27 38 9.0 4.5 4.6 2.5
Cps 1800 Si Si
1300
0 1M
800 3M 5M
The orthorhombic tobermorite suited best for all activated ash samples except one activated at temp 130 ◦ C and in 8 M NaOH solution were monoclinic tobermorite achieved the best fit and also unidentified crystal phase was formed (see Fig. 1). The Si-external standard was inserted into sample for technical reason and gap in pattern 26–27◦ (2) of the original ash sample is a technical error. The results of semi-quantitative analysis of mineral content of one activated oil shale fly ash sample are presented in Table 3. The activation conditions were: 8 M NaOH, at 160 ◦ C for 24 h. The results show that the main mineral synthesized during the activation is tobermorite, whilst part of silica was also converted to katoite. Calcite and cancrinite are minerals originating from the native oil shale ash [34].
8M
300
quartz
-200
tobermorite 1 tobermorite 2 calcite hematite katoite
-700 5
15
25
35
45
55 °2θ, Co Kα
Fig. 2. XRD patterns of the original and activated ash samples at different NaOH concentrations with reference patterns of major minerals. Temperature 160 ◦ C; Si refers to internal silicon standard; tobermorites 1 and 2 refer to orthorhombic and monoclinic polytypes of 1.1 nm tobermorite, respectively.
3.4. Magic angle spinning nuclear magnetic resonance (MAS-NMR) analysis Cps 1800 Si Si
1300 0 1M
800
3M 5M 8M
300
29 Si MAS-NMR spectra of the original and activated ash samples (at the temperatures of 130 and 160 ◦ C) as well as the deconvolution of the spectra are presented in Fig. 3 and Fig. 4, respectively. The spectra of the original ash exhibit broad resonance in the chemical shift range from −80 to −110 ppm. This resonance can be assigned to the variety of silicon sites in the amorphous fly ash glass. The resonance lines at −71.5 and −107.3 ppm arise from belite (-Ca2 SiO4 ) [35] and quartz [36], respectively. The chemical shift values of the 29 Si resonance of activated oil shale ash samples are typical to the silicon sites in silicate chains of tobermorites [37–40] or that of silicate hydrate gels [41–43]. The spectrums of activated oil shale ash samples present five resonance
quartz quatz
-200
tobermorite 1 tobermorite 2
Table 3 Semi-quantitative XRD mineral analysis of the oil shale sample activated in 8 M NaOH solution at the temperature of 160 ◦ C.
calcite hematite katoite
-700 5
15
25
35
45
55 °2θ, CoKα
Fig. 1. XRD patterns of the original and activated ash samples, at different NaOH concentrations with reference patterns of major minerals. Temperature 130 ◦ C; Si refers to internal silicon standard; tobermorite 1 and 2 refer to orthorhombic and monoclinic polytypes of 1.1 nm tobermorite, respectively.
Mineral Tobermorite Calcite Katoite Cancrinite
wt% 75 12 9 4
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183 2
Q
2
Q (1Al) 8M 2
Q
Q (1Al)
1
2
Q
3
Q (1Al)
8M
Q
1
3
Q (1Al) 3 Q
5M
5M
3M
2
2
Q (1Al)
Q
1M
3
Q (1Al) 3 Q
1
-50
-60
Q
3M
original ash
-70
-80
-90
-100
-110
-120
ppm 1M Fig. 3. MAS-NMR analyses of the original ash and samples activated at 130 ◦ C in 1–8 M NaOH solution. Deconvolution with Gaussian curves.
original ash lines at −80, −82.7, −85.8, −92 and −96.7 that can be assigned to the silicon sites Q1 , Q2 (1Al), Q2 (0Al), Q3 (1Al) and Q3 of 1.1 nm tobermorite, respectively [37]. The site Q1 is the silicon site at the end of the polymeric silicate chain in the SiO4 /AlO4 tetrahedron chains (“dreierkette”), Q2 is the silicon site in the chain middle groups, Q2 (1Al) is the chain middle group site where one of the two neighboring tetrahedra contains aluminum ion, Q3 is the silicon site with three nearest neighbor SiO4 tetrahedra, the branching tetrahedron which links together two chains, and Q3 (1Al) is the site, where one of the three nearest neighbor tetrahedra contains an Al ion [8]. The best fit was achieved when we added underneath of the well-resolved lines a broad Gaussian background line (Qx ) with center position of the peak at −89.5 ppm, which arises from silicon configurations of amorphous calcium–silicate hydrate phase. Some quantitative characteristics of ash samples can be obtained from the deconvolution of the 29 Si MAS-NMR spectra by Gaussian lines using relative intensities of the 29 Si lines in Table 4. The mean length n of SiO4 /AlO4 chains and the ratio of Al/Si were evaluated from the formulas presented in Refs. [8] and [43].
-50
-60
-70
-80
-90
-100
-110
-120
ppm Fig. 4. MAS-NMR analyses of the original ash and samples activated at 160 ◦ C in 1–8 M NaOH solution. Deconvolution with Gaussian curves.
The data reveals that the average chain length of activated oil shale ash remains constant at the temperature of 130 ◦ C and increases with increasing temperature, except for sample activated in 8 M NaOH solution. Ratio of aluminum to silicon is not dependent on reaction conditions and remains in between 0.12 and 0.19. 3.5. Scanning electron and transmission electron microscopy (SEM-TEM) analysis Imaging results of scanning electron and transmission electron microscopes for the original ash vs. a sample activated in 4 M NaOH solution at 130 ◦ C are presented in Figs. 5 and 6, respectively.
Table 4 Relative intensities of Gaussian lines in 29 Si MAS-NMR spectras with center position of the peak in ppm. I1 –I3 are the intensities of lines assigned to silicon sites Q1 –Q3 , respectively. Ix is the intensity of the broad background line arising from amorphous silicon sites. Ash Sample
I1 −80
I2 (1Al) −82.7
I2 85.8
I3 (1Al) −92
I3 −96.7
Ix −89.5
n
Al/Si
1M130 3M130 5M130 8M130 1M160 3M160 5M160 8M160
0.06 0.09 0.11 0.06 0.04 0.04 0.04 0.11
0.15 0.30 0.24 0.20 0.15 0.31 0.27 0.21
0.27 0.34 0.43 0.23 0.32 0.41 0.41 0.34
0.03 0.02 0.17 0.03 0.05 0.10 0.14 0.10
0.02 0.02 0.02 0 0.06 0.04 0 0
1.90 1.94 0.90 0.90 2.36 2.92 2.95 3.22
21 21 21 21 36 55 53 17
0.14 0.19 0.12 0.19 0.12 0.17 0.16 0.14
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Fig. 5. SEM (left) and TEM (right) images of the native oil shale fly ash.
Fig. 6. SEM (left) and TEM (right) images of the oil shale fly ash sample activated in 4 M NaOH solution at 130 ◦ C.
SEM analysis demonstrated that the original ash had smooth surface structure and particles had spherical shapes (Fig. 5). The texture of the activated ash is more obscure and TEM image confirms that crystal morphologies are formed on the surface (Fig. 6). 3.6. Thermo-gravimetric analysis (TGA) The results of CO2 adsorption on the surface of original and some samples of activated fly ash are presented in Fig. 7 and Fig. 8. The original oil shale ash adsorbs 0.03–0.06 mass% of CO2 at 40 ◦ C during 120 min period. The CO2 adsorption capacity of alkaline treated oil shale ash (3 M NaOH solution, at 160 ◦ C) is up to 4.4 mass% under similar conditions.
Fig. 8. Thermo-gravimetric analysis of CO2 adsorption onto ash sample activated in 3 M NaOH solution at 160 ◦ C.
Results of CO2 adsorption onto original and selected ash samples at temperature 40 ◦ C and 120 min period are presented in Table 5. According to Table 5, the samples activated at temperature 160 ◦ C adsorb 2–3 times more CO2 than those activated at temperature 130 ◦ C. Table 5 Maximum physical adsorption of CO2 onto selected ash samples at temperature 40 ◦ C during 120 min.
Fig. 7. Thermo-gravimetric analysis of CO2 adsorption onto original oil shale fly ash sample.
Sample
CO2 adsorption wt%
Original ash 1M130 8M130 5M160 3M160
0.06 1.8 1.3 3.7 4.4
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Fig. 9. Cyclic thermo-gravimetric analysis of CO2 adsorption onto ash sample activated in 5 M NaOH solution at 160 ◦ C.
According to cyclic adsorption test with ash sample activated in 5 M NaOH solution at 160 ◦ C, the adsorption capacity of the activated ash sample diminishes less than 5% after two adsorption/thermal desorption cycles (Fig. 9). 4. Discussions and conclusions Synthetic calcium–silica–aluminum hydrates, mainly 1.1 nm tobermorite and katoite, were produced from oil shale fly ash at a laboratory scale. Reaction parameters were alkali concentration and temperature whilst solid/liquid ratio and reaction time were kept constant. XRD and MAS-NMR analysis demonstrated that formation of synthetic hydrates and dissolution of silica depends positively on both alkali concentration and reaction temperature, at given range (1–8 M NaOH solution concentration; 130 ◦ C and 160 ◦ C). The highest yield of 1.1 tobermorite was obtained with 5 M NaOH solution at the temperature of 160 ◦ C (Fig. 10). During the activation of oil shale ash specific surface area (m2 /g) increased over 10 times and the cumulative volume of pores (mm3 /g) 100 times, which gives us indication of the formation of micropores in tobermorite structure. Some dependence of specific surface increase on alkali concentration was observed in the samples activated at 130 ◦ C. Semi-quantitative XRD analysis illustrated that minerals synthesized during oil shale activation were tobermorite (75 mass%) and katoite (9 mass%). According to MAS-NMR analysis, the tobermorite can be assigned to 1.1 nm tobermorite.
Intensity, Cps at 9.08 2theta
300
200 160 oC 130 oC 100
0 0
2
4
6
8
10
Alcaline concentration, M Fig. 10. XRD peak intensities of tobermorite at 9.08, at different reaction conditions.
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The results of TGA show that the alkaline treatment of oil shale ash increases the CO2 adsorption capacity and is dependent on reaction temperature. The CO2 adsorption tests were carried out in dry conditions. The residual NaOH in activated materials plays role in CO2 adsorption under humid conditions. Therefore we assume that the main cause of increase in CO2 adsorption capacity is the formation of synthetic minerals, which leads to increased specific area and pore volumes in product. The adsorption capacity of alkali-treated (3–4 mass%) ash is coherent with the results obtained from other CO2 adsorption research [44]. Furthermore, the cyclic test for CO2 adsorption/thermal desorption demonstrated that the activated oil shale ash can be regenerated for multiply adsorption–desorption cycles for CO2 capture. Future study will address CO2 adsorption onto activated oil shale ash in a pilot scale industrial process. Also for other applications, the leaching test of original and activated product will be made in order to assess the trend on bioavailability of heavy metals from the native and treated oil shale fly ash. Acknowledgments This work was supported by grant SF 0690001s09 from the Estonian Ministry of Education and Research. In Sweden, the Bio4Energy program, Knut and Alice Wallenberg Foundation as well as Kempe Foundations are acknowledged. Also, the Micre (Micro Energy to Rural Enterprise) project under the auspices of the Northern Periphery Program by the EU is acknowledged. References [1] R. Mõtlep, T. Sild, E. Puura, K. Kirsimäe, Composition, diagenetic transformation and alkalinity potential of oil shale ash sediments, J. Hazard. Mater. 184 (2010) 567–573. [2] E. Häsänen, L. Aunela-Tapola, V. Kinnunen, K. Larjava, A. Mehtonen, T. Salmikangas, J. Leskelä, J. Loosaar, Emission factors and annual emissions of bulk and trace elements from oil shale fueled power plants, Sci. Total Environ. 198 (1997) 1–12. [3] M. Laja, G. Urb, N. Irha, J. Reinik, U. Kirso, Leaching behavior of ash fractions from oil shale combustion by fluidized bed and pulverized fuel processes, Oil Shale 22 (2005) 453–465. [4] A. Ots, Oil Shale Fuel Combustion, Tallinna Raamatutrükikoda, Tallinn, 2006. [5] J. Luan, A. Li, T. Su, X. Cui, Synthesis of nucleated glass-ceramics using oil shale fly ash, J. Hazard. Mater. 173 (2010) 427–432. [6] H. Höller, U. Wirsching, Zeolites formation from fly ash, Fortschr. Miner. 63 (1985) 21–43. [7] R.A. Shawabkeh, Synthesis and characterization of activated carboaluminosilicate material from oil shale, Microporous Mesoporous Mater. 75 (2004) 107–114. [8] J. Reinik, I. Heinmaa, J.P. Mikkola, U. Kirso, Hydrothermal alkaline treatment of oil shale ash for synthesis of tobermorites, Fuel 86 (2007) 669–676. [9] I.G. Richardson, The calcium silicate hydrates, Cem. Concr. Res. 38 (2008) 137–158. ˜ A. Alastuey, E. Hernández, A. López-Soler, F. [10] X. Querol, N. Moreno, J.C. Umana, Plana, Synthesis of zeolites from coal fly ash: an overview, Int. J. Coal Geol. 50 (2002) 413–423. [11] N.J. Coleman, Interactions of Cd(II) with waste-derived 11 A˚ tobermorites, Sep. Purif. Technol. 48 (2006) 62–70. [12] R.A. Shawabkeh, Equilibrium study and kinetics of Cu2+ removal from water by zeolite prepared from oil shale ash, Process Saf. Environ. Prot. 87 (2009) 261–266. [13] R. Siauciunas, V. Janickis, D. Palubinskaite, R. Ivanauskas, The sorption properties of tobermorite modified with Na+ and Al3+ ions, Ceramics – Silikáty 48 (2004) 76–82. [14] M. Miyake, S. Komarneni, R. Roy, Kinetics equilibria and thermodynamics of ion exchange in substituted tobermorites, Mater. Res. Bull. 24 (1989) 311–320. [15] M.F. Heddle, Preliminary notices on substances which may prove to be new minerals, Mineral. Mag. 4 (1882) 117–123. [16] N.J. Coleman, Synthesis, structure and ion exchange properties of 11 A˚ tobermorites from newsprint recycling residue, Mater. Res. Bull. 40 (2005) 2000–2013. [17] Z. Yao, C. Tamura, M. Matsuda, M. Miyake, Resource recovery of waste incineration fly ash: synthesis of tobermorite as ion exchanger, J. Mater. Res. 14 (1999) 4437–4442. [18] N.J. Coleman, C.J. Trice, J.W. Nicholson, 11 A˚ tobermorite from cement bypass dust and waste container glass: a feasibility study, Int. J. Miner. Process. 93 (2009) 73–78.
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Journal of Hazardous Materials 196 (2011) 187–193
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
A nanofiber functionalized with dithizone by co-electrospinning for lead (II) adsorption from aqueous media Jianjun Deng a , Xuejun Kang a,c,∗ , Liqin Chen a , Yu Wang a , Zhongze Gu b,c , Zuhong Lu a a b c
Key Laboratory of Child Development and Learning Science (Ministry of Education), Research Center for Learning Science, Southeast University, Nanjing 210096, China State Key Laboratory of Molecular and Biomolecular Electronics, Southeast University, Nanjing 210096, China Suzhou Key Laboratory of Environment and Biosafety, Suzhou 215123, China
a r t i c l e
i n f o
Article history: Received 24 January 2011 Received in revised form 5 September 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: Electrospun nanofiber Packed fiber solid phase extraction (PFSPE) Lead (II)
a b s t r a c t An electrospun nanofiber was utilized as a sorbent in packed fiber solid phase extraction (PFSPE) for selective separation and preconcentration of lead (II). The nanofiber had a polystyrene (PS) backbone, which was functionalized with dithizone (DZ) by co-electrospinning of a PS solution containing DZ. The nanofiber exhibited its performance in a cartridge prepared by packing 5 mg of nanofiber. The nanofiber was characterized by a scanning electron microscope and IR spectra. The diameter of the nanofiber was less than 400 nm. After being activated by 2.0 mol L−1 NaOH aqueous solution, the nanofiber quantitatively sorbed lead (II) at pH 8.5, and the metal ion could be desorbed from it by three times of elution with a small volume of 0.1 mol L−1 HNO3 aqueous solution. The breakthrough capacity was 16 g mg−1 . The nanofiber could be used for concentration of lead (II) from water and other aqueous media, such as plasma with stable recovery in a simple and convenient manner. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Heavy metals, which are the main pollutants not only in the industrial sector, but also in our living environment, are causing severe public health problems. One typical example of it is lead. According to the Guidelines of the United States Center for Disease Control, a blood lead level below 10 g dL−1 is interpreted as “safe”, while medical evaluation and treatment are recommended for blood leads above 20 g dL−1 [1]. However, many works have found that even low-level lead exposure is harmful to our health, especially for children who are more susceptible [2–5]. Therefore, the determination of heavy metal ions at trace level is very important in environmental protection and disease prevention. The direct determination of heavy metal ions in complex matrices is limited due to their low concentrations and matrix interferences. Therefore, a preconcentration and separation procedure, such as liquid–liquid extraction, coprecipitation, and cloud point extraction [6–9], is necessary to improve the sensitivity and selectivity of the determination of heavy metal ions. Recently, solid phase extraction (SPE) has been the most common technique used for preconcentration of heavy metal ions because of its advantages
∗ Corresponding author at: Key Laboratory of Child Development and Learning Science (Ministry of Education), Research Center for Learning Science, Southeast University, Nanjing 210096, China. Tel.: +86 25 83795664; fax: +86 25 83795929. E-mail address:
[email protected] (X. Kang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.016
of a high enrichment factor, high recovery, rapid phase separation and low consumption of organic solvents [10–13]. Among various types of SPE techniques, there is a new technique called packed fiber solid phase extraction (PFSPE) [14], which is based on the use of electrospun nanofiber as the sorbent. Although the electrospinning technique was invented in the early 1900s, electrospun nanofiber was reported as early as 1971 [15], and various types of electrospun nanofiber and various applications have been reported after that [16–18]; there are few reports related to the extraction. Compared to the conventional SPE technique, the nanofiber sorbent possesses a large surface area which facilitates the attachment of target molecules, so that less amounts of sorbent, and less volume of sample and eluent are required [19]. The PFSPE technique can perform perfectly in both environmental and biological sample pretreatment. Some successful applications of PFSPE have been reported in the extraction of target compounds, such as the cortisol in the saliva [19], vitamins in the plasma [20] and in the beverages [21], drugs [22] in the plasma, and aromatic pollutants in the environmental water [23]. The target molecules which have been reported were organic matters. However, the pretreatment of metal ions by PFSPE is rarely found to be reported. In this work, a selective sorbent, polystyrene–dithizone composite electrospun nanofiber (PS–DZ nanofiber), was fabricated, characterized and applied to extraction of lead (II) in aqueous samples. Dithizone, as a well-known reagent, has been used for the determination of many metal ions [24]. It can be immobilized on the solid phase by synthetic reaction. The chelating resin containing
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s-bonded dithizone was reported to have good performance on SPE of metal ions [25], but the procedure of fabrication was not straightforward enough. The modification was also achieved by coating the solid phase with dithizone [26], however, the dithizone on the solid phase was possibly not stable enough. In this work, the immobilization of DZ on the solid phase was finished by co-electrospinning of a mixture solution of PS polymer and DZ, and adsorbing performance of the composite nanofiber was investigated. 2. Experimental 2.1. Reagents and solutions All chemicals used in this work were of analytical reagent grade and were used without further purification. Doubly distilled deionized water was used throughout. The polystyrene (PS), the polyacrylonitrile (PAN) and the acrylic resin (AR) were obtained from Shanghai Chemical Agents Institute. The standard labware and glassware used were cleaned with HNO3 and rinsed with double distilled water, according to Tuzen et al. [27]. Standard stock solutions (1 mg mL−1 ) of lead (II) were prepared by dissolving spectral pure grade chemicals Pb(NO3 )2 in double distilled water. Buffer solutions (NH3 /NH4 Cl) were prepared by mixing appropriate volumes of 1 mol L−1 ammonium chloride and ammonia for pH 8–9. Phosphate buffer solutions for pH 6–7 were prepared from disodium hydrogen phosphate and citric acid, and the concentration of Na2 HPO4 was approximately 1 mol L−1 . The acetic buffer solutions (HAc/NaAc) were prepared by mixing 1 mol L−1 sodium acetate and glacial acetic acid for pH 3–5. The sodium hydroxide solutions were prepared by dissolving sodium hydroxide in distilled water.
Fig. 1. The system of electrospinning.
The PAN–DZ nanofiber and the AR–DZ nanofiber were prepared in the following procedures: 8% (w/v) polyacrylonitrile (PAN) was dissolved in the dimethylformamide by magnetic stirring after ultrasonic processing for 30 min; and 8% (w/v) acrylic resin (AR) was dissolved in the ethanol by magnetic stirring. Then dithizone (5%, w/w of polymer) was dissolved in the polymer solution. The solutions continued to be stirred at room temperature for more than 10 h before electrospinning. Other procedures were similar to those of the PS–DZ nanofiber, except for a voltage of 14 kV for the PAN–DZ nanofiber and 10 kV for the AR–DZ nanofiber. A novel cartridge, as shown in Fig. 2, was designed for pretreatment of the aqueous sample. 3–5 mg of nanofiber were packed into a column with an inside diameter of 1.5 mm. A conical liquid storage cartridge was attached to the column. The pressurizer was available by using a syringe with a modified tip which was fitted to the liquid storage cartridge.
2.2. Instruments and apparatus A high performance liquid chromatography, consisting of a LC-20A pump and a PDA detector (SHIMADZU, Japan) was used. A C18, 5 m, 150 mm × 4.6 mm Rad-Pak reversed-phase column (SHIMADZU, Japan) was used to achieve fast separation and analysis of metal ions. A HPLC software package (SHIMADZU, Japan) was used for the data analysis. A pH meter (Shanghai, China) with a glass electrode was used for all pH measurements. A high-voltage power supply (model DWP403-1AC, Tianjin, China), and a syringe pump were used for the electrospinning. The nanofiber was examined using a Hitachi S3000N scanning electron microscope (SEM, Tokyo, Japan). The IR spectra were carried out on a NICOLET 5700 FT-IR Spectrophotometer (Nicolet, US).
2.4. Procedures 2.4.1. Adsorption and desorption procedures The nanofiber should be activated before being used. Before activation, the pretreatment device consisted of the column and the liquid storage cartridge was washed with 0.1 mol L−1 nitric acid and water, so as to clean up the metal ion remaining in the device.
2.3. Fabrication of electrospun nanofiber and preparation of cartridge The electrospun solution was prepared by dissolving an appropriate amount of PS (10%, w/v) and dithizone (5%, w/w of PS) in a mixture of dimethylformamide and tetrahydrofuran (4:6, v/v). At the beginning, PS was dissolved in the mixed solvent by magnetic stirring. The dithizone was added into the polymer solution after PS was dissolved completely. The solution continued to be stirred at room temperature for more than 10 h before electrospinning. This solution was loaded into a glass syringe which was fitted to a steel needle with a tip diameter of 0.5 mm whose tip was filed flat. Then electrospinning performance was shown in Fig. 1 in this condition: an anodic voltage of 17 kV, 25 cm from the tip of the needle to the collecting equipment, the feed rate of 1 mL h−1 for precursor solution. The nanofiber was collected on the collector which was covered by a piece of gauze pretreated by dilute HNO3 (1 + 9) and rinsed with distilled water.
Fig. 2. The component of the pretreatment device.
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2 mol L−1 sodium hydroxide aqueous solution was used to activate the nanofiber. After had been washed with water, the nanofiber was ready for sample pretreatment. Test solutions containing lead (II) were adjusted to the desired pH before being loaded into the cartridge. After loading and elution, the column was washed with buffer solutions (pH 8.5–9.0), and then desorbed with the one and the same 0.1 mol L−1 nitric acid solutions in a small volume (0.1 mL) throughout 3 times. The flow rate was carefully controlled in a slow dropwise manner in the adsorption and desorption procedures. 2.4.2. Chromatographic procedures Analysis of metal ions by high performance liquid chromatography had been extensively reported before [28–30]. Based on the works of Dilli et al. [31] and Wang and Wai [32], we used the sodium dimethylaminaocarbodithioate as chromogenic agent. The mobile phase was composed of methanol, acetonitrile and water (40:35:25, v/v/v). The HPLC flow rate was 1.5 mL min−1 . The detection wavelength was 260 nm. 2.5. Application The tap water, the lake water and the plasma provided by The Nanjing Blood Donor Service (Nanjing, China), which was donated from a healthy volunteer, were collected in the polytetrafluoroethylene centrifugal tube. 10 mL of water samples were adjusted to the desired pH with the buffer solution. A 0.5 mL nitric acid solution (5%, v/v) was added into 1 mL of the plasma sample. After agitation for 1 min and ultrasonic processing for 10 min, the mixture was centrifuged at 12,000 rpm for 5 min. The supernatant then was transferred into another centrifugal tube. The precipitate was washed with a 0.5 mL nitric acid solution twice. Then the supernatant which was collected into the same centrifugal tube was adjusted to the desired pH with sodium hydroxide solutions and buffer solutions. The pretreatment and the detection procedures given above were applied to the samples. 3. Results and discussion 3.1. Characterization Identified by scanning electron microscope, illustrated in Fig. 3a and b, the diameter of the PS–DZ nanofiber was 200–400 nm, and the nanofiber was dense with network structure. Though PS was modified with dithizone, no morphological change was observed in the view of the SEM images shown in Fig. 3c. PS–DZ nanofiber was also identified by the IR spectra, as shown in Fig. 4. Compared to the IR spectra of the PS nanofiber (Fig. 4a), many new peaks appeared in the IR spectra of PS–DZ nanofiber (Fig. 4b) from 1000 to 2000 cm−1 , which were similar to the IR spectra of dithizone. According to the literature [33], the new peaks at 1257 cm−1 , 1172 cm−1 and 1522 cm−1 were respectively due to C–N stretching vibration, C S stretching vibration and N N
Fig. 4. IR spectra of PS nanofiber (a) and PS–DZ nanofiber (b).
stretching vibration donated by dithizone. It seems that dithizone was successfully impregnated in the PS nanofiber. 3.2. Activation of nanofiber The PS–DZ nanofiber performed poorly without activation. The recovery of lead (II) was less than 20%. The commonly used activating solvent for PFSPE, such as methanol [14], did not improve the recovery very much. The sodium hydroxide aqueous solutions were found to activate the PS–DZ nanofiber effectively. Furthermore, the absorption was influenced by the concentration of sodium hydroxide aqueous solutions (Fig. 5). With 0.1–5.0 mol L−1 of sodium hydroxide aqueous solutions, the recovery improved correspondingly when the concentration of the sodium hydroxide aqueous solution increased, and was stable when the concentration of sodium hydroxide aqueous solutions was higher than 1 mol L−1 . In addition, the activation efficiency decreased when the methanol was added into the sodium hydroxide solutions. Finally, the 2.0 mol L−1 of sodium hydroxide aqueous solution was selected as the activating solution throughout the following work. Dithizone is an S, N-donating ligand as shown in Fig. 12a. It can easily react with many heavy metal ions at particular pH [34]. Primary dithizonate is formed when dithizone reacts with the metal ion as an anion of monobasic acid (HDz− ) [35]. And dithizone dissolves in the alkaline aqueous medium in the form of the dithizonate whose polarity is stronger than that of the dithizone, but is undissolved in the neutral aqueous medium [36]. When
Fig. 3. Scanning electron microscope images of nanofiber: (a) PS–DZ nanofiber magnified 2k times, (b) PS–DZ nanofiber magnified 10k times, and (c) PS nanofiber magnified 10k times.
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Fig. 5. Effect of the concentration of NaOH in the activation solution. Other conditions: sample pH = 8.5. eluent: 100 L of 0.1 mol L−1 HNO3 aqueous solution eluted 3 times with the same eluent. N = 3.
the dithizone was impregnated in the fiber, it was expected that the chelation between dithizone and lead (II) became more difficult in the neutral aqueous medium because of phase interfacial resistance, which caused the poor adsorption efficiency without activating the PS–DZ nanofiber. When the PS–DZ nanofiber was treated by strong alkaline aqueous solutions, dithizone on the nanofiber transformed into dithizonate (NaHDz), which was possibly easier to react with lead (II). That may be the reason that treating with methanol did not activate the PS–DZ nanofiber. 3.3. Effect of pH The pH of the sample solutions is usually the most critical parameter on the SPE studies of metal ions [37,38]. Especially for the SPE of lead (II) in this trial based on the chelation, the pH of the sample solution is one of the decisive parameters for quantitative recovery of the analytes [39]. The effect of pH was investigated in the pH range of 3–9. As shown in Fig. 6, stable recoveries were obtained in the pH range of 7–9. For dithizone chelated respectively with zinc at the neutral pH, and cadmium in the strong alkaline solution [40], the investigation was carried out at pH 8.5. 3.4. Effect of organic solvent in the medium The effect of the organic solvent was also investigated with methanol as a model solvent. The sample solutions containing 0–20% (v/v) of methanol were carried out to find out how the
Fig. 7. Effect of organic solvent in the medium on the retention of lead (II).
organic solvent influenced the quantitative recovery. The result was given in Fig. 7. The existence of methanol impacted the absorption efficiency of the PS–DZ nanofiber notably when the concentration of methanol was higher than 7% (v/v). To obtain a good recovery, the ratio of organic solvent in the sample should be as less as possible. 3.5. Eluent Chelation of dithizonate is pH sensitive. Plumbous dithizonate is instable in an acidic solution. For selection of the best eluent, various acidic solutions were studied for desorption of lead (II). The result revealed that nitric acid performed better on desorption of lead (II). Moreover, the addition of methanol was favorable for desorption (Table 1). However, a small amount of dithizone was swilled out from the nanofiber by absolute methanol which was an environmental pollutant. Therefore, the nitric acid aqueous solution was considered as the most suitable eluent. The concentration of nitric acid was an influential factor in the desorption procedure. As shown in Fig. 8, recovery was lower than 95% while the concentration of nitric acid was lower than 0.08 mol L−1 . Hence, 0.1 mol L−1 nitric acid was selected as a suitable desorption solution. A decreased volume of eluent was good for detection sensitivity, but a small volume of eluent, for example, a 100 L of 0.1 mol L−1 nitric acid aqueous solutions could not desorb the adsorbate entirely. To solve this problem, multi-washing processes which repeatedly used the one and the same acid solution were carried out. As presented in Table 2, the frequency of eluting should be at least twice, but in order to ensure the entire recovery, three times was more suitable. 3.6. Effect of the sample volume In order to find out the effect of the sample volume on the sorption behavior of lead (II) on the PS–DZ nanofiber, 1–20 mL of lead (II) aqueous solutions containing 500 ng mL−1 of lead (II) in the desired Table 1 Effect of eluent on the recoveries of lead (II). Eluent volume: 100 L, eluting for 3 times. N = 3.
Fig. 6. Effect of pH on the retention of lead (II).
Type of eluent
Recovery (%)
0.05 M HNO3 0.1 M HNO3 0.1 M HCl 0.1 M H2 SO4 0.05 M HNO3 with 20% (v/v) MeOH 0.05 M HNO3 with 20% (v/v) MeOH MeOH
91.36 98.99 81.46 85.52 98.74 99.10 8.28
± ± ± ± ± ± ±
2.31 2.34 1.90 2.53 2.68 2.46 3.67
J. Deng et al. / Journal of Hazardous Materials 196 (2011) 187–193
Fig. 10. Breakthrough volume of lead (II).
Fig. 8. Eluting effect of HNO3 aqueous solution of different concentrations. Table 2 Effect of the frequency of eluting. N = 3.
191
Table 3 Reutilization of PS–DZ nanofiber. N = 3.
Times of eluting
Recovery (%)
1 2 3 4 5
91.35 98.67 98.99 98.52 98.61
± ± ± ± ±
2.73 1.79 2.34 3.13 3.04
pH were pretreated by the PS–DZ nanofiber. The result illustrated in Fig. 9 revealed that the recoveries of lead (II) from different sample volumes (1–20 mL) were quantitative. 3.7. Breakthrough capacity To evaluate the amount of lead (II) sorbed per milligram on PS–DZ nanofiber under the operating conditions, the breakthrough capacity was calculated with the assumption that breakthrough occurs at Ce /Ci = 0.01; Ce is the concentration of lead (II) in the effluent, and Ci is the concentration of lead (II) in the influent [41]. 10 g mL−1 of lead (II) solutions at the desired pH passed through the column packed with 5 mg of PS–DZ nanofiber. The concentration of lead (II) in each milliliter of the effluent was determined. The breakthrough capacity presented in Fig. 10 was calculated to be 16 g mg−1 for the PS–DZ nanofiber. 3.8. Reutilization
Times of reutilization
Recovery (%)
0 1 2 3
98.99 98.75 99.03 98.26
± ± ± ±
2.34 2.16 3.72 2.58
before reutilization. Then the trial followed the optimum activation, adsorption and desorption procedures given above. The result was given in Table 3. The recoveries in three times of reutilization were quantitative. Although the PS–DZ nanofiber was reutilizable, it is recommended to be throwaway when used for determination of the trace of lead (II).
3.9. Effect of amounts of dithizone The effect of amounts of dithizone in electrospun solutions was investigated in 0–20% mass ratio of PS. The result was given in Fig. 11. The nanofiber rarely sorbed lead (II) without dithizone, in contrast, the nanofiber modified with dithizone performed well. The recovery was over 98% in a mass ratio range of 5–20%. However, a little dithizone visibly brushed off from PS–DZ nanofiber whose contents of dithizone were over 10% when the nanofiber was being activated.
The reutilization of PS–DZ nanofiber was also investigated. Sufficient nitric acid solution and water were used to clean the device
Fig. 9. Effect of the sample volume.
Fig. 11. Effect of amounts of dithizone in PS–DZ nanofiber. Each electrospun solution was electrospinned in the same conditions.
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Fig. 12. The molecular structure of dithizone (a) and the representative schema of PS–DZ nanofiber (b).
Table 5 The determination of lead (II) in some real sample. N = 3.
3.10. Effect of foreign ions In order to evaluate the possibility of selective recovery of analyte ions, the effect of coexisting ions needs to be considered. Various amounts of metal ions were added to an aqueous solution containing 500 ng mL−1 of lead (II), and the optimum procedure was followed. The result including the tolerance limit and the relevant recovery was given in Table 4. The metal ions normally present in natural water and biological samples did not interfere under the optimized experimental conditions, which implied the method was desired in view of applications on real samples.
Sample
Added (ng mL−1 )
Found (ng mL−1 )
Recovery (%)
Lake water
– 50 100
10.26 ± 1.73 59.37 ± 2.18 108.95 ± 1.85
– 98.22 98.69
Tap water
– 50 100
ND 49.63 ± 2.35 98.71 ± 1.89
– 99.26 98.71
Plasma
– 50 100
ND 39.68 ± 2.81 80.54 ± 2.52
– 79.36 80.54
3.11. Difference of polymers backbone
ND: not detected.
Dithizone was impregnated in different sorts of polymer nanofiber, and the performance on the adsorption of lead (II) was investigated. AR–DZ nanofiber was destroyed when activated by a sodium hydroxide solution for high solubility of acrylic resin in alkaline solution. PAN–DZ nanofiber performed almost as well as PS–DZ nanofiber in the adsorption and desorption procedure. However, a part of the dithizone from the PAN–DZ nanofiber was brushed off in the activation procedure. It was suggested that dithizone in PAN–DZ nanofiber was just inlayed on the surface of the nanofiber, in contrast, dithizone in PS–DZ was not only inlayed on the surface of the nanofiber but also immobilized by the conjugation of the benzene ring of dithizone and polystyrene (Fig. 12b). Therefore, the PS–DZ nanofiber was more stable than PAN–DZ nanofiber, and it was selected as the absorbent in our work.
4. Conclusions
3.12. Applications The real samples, including lake water in a tourism area, tap water and plasma donated from a healthy volunteer, were used to investigate how PS–DZ nanofiber performed in the real samples. Various amounts of lead (II) were added to the real samples to examine recovery of lead (II). As illustrated in Table 5, PS–DZ nanofiber performed well in the water samples with quantitative recoveries. However, the performance in the plasma was not as good as in the water samples. Probably, a part of lead (II) coprecipitated with the protein, which could not be released by the nitric acid solution without digestion.
Table 4 Effect of coexisting ions on the recoveries of lead (II). N = 3.
The PS–DZ nanofiber was used as the sorbent for solid phase extraction of lead (II). The most important characteristic of the PS–DZ nanofiber was its excellent selectivity towards lead (II). A novel modification for sorbent by co-electrospinning of the polymer solution containing functional molecules was developed, which was simple and rapid. The performance of the functional PS–DZ nanofiber in retaining lead (II) was investigated by packing it into a novel sample pretreatment device. The conditions relevant to the performance of PS–DZ nanofiber were optimized for the quantitative recovery of lead (II). The pretreatment device was practicable in the analysis of water samples. The enrichment factor of PS–DZ nanofiber for lead (II) was higher than that of other SPE materials [42–44] when used in pretreatment of the same volume of liquid samples. Though the application in plasma without digestion was not as efficacious as in the water samples, this device with the PS–DZ nanofiber could be applicable in screening blood lead or lead in urine, for the simple and rapid procedures and small sample volumes it required. Acknowledgements This work was supported by Jiangsu Province Science and Technology Department Foundation (Grant No. BE2010088), National Natural Science Foundation of China (Grant No. 81172720), Suzhou Science and Technology Department Foundation (Grant Nos. SYJG0912, SYN201006), National Basic Research Program of China (Grant No. 2007CB936300). References
Ions
Added as
Concentration (mg mL−1 )
Recovery (%)
Na+ K+ Mn2+ Zn2+ Hg2+ Cu2+ Ca2+ Al3+ Fe3+
NaCl KCl MnSO4 ZnNO3 HgAc2 CuSO4 CaAc2 AlCl3 FeCl3
5 3 0.2 0.2 0.5 0.5 0.3 0.4 0.4
93.53 96.61 93.24 91.79 98.85 97.12 93.41 98.83 98.03
± ± ± ± ± ± ± ± ±
2.58 2.40 3.36 3.53 2.39 2.44 3.13 2.96 3.38
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Journal of Hazardous Materials 196 (2011) 194–200
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Adsorption and desorption performance of benzene over hierarchically structured carbon–silica aerogel composites Baojuan Dou, Jinjun Li, Yufei Wang, Hailin Wang, Chunyan Ma ∗ , Zhengping Hao ∗∗ Department of Environmental Nano-Materials, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences, Beijing 100085, PR China
a r t i c l e
i n f o
Article history: Received 9 March 2011 Received in revised form 2 September 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: Carbon–silica aerogel composites VOCs Adsorption and desorption Micro/mesoporous structure
a b s t r a c t Hierarchically structured carbon–silica aerogel (CSA) composites were synthesized from cheap water glass precursors and granulated activated carbon via a post-synthesis surface modification with trimethylchlorosilane (TMCS) and a low-cost ambient pressure drying procedure. The resultant CSA composites possess micro/mesoporous structure and hydrophobic surface. The adsorption and desorption performance of benzene on carbon–silica aerogel composite (CSA-2) under static and dynamic conditions were investigated, comparing with pure silica aerogel (CSA-0) and microporous activated carbon (AC). It was found that CSA-2 has high affinity towards aromatic molecules and fast adsorption kinetics. Excellent performance of dynamic adsorption and desorption observed on CSA-2 is related to its higher adsorption capacity than CSA-0 and less mass transfer resistance than AC, arising from the well-developed microporosity and open foam mesostructure in the CSA composites. © 2011 Elsevier B.V. All rights reserved.
1. Introduction From an environmental point of view, it is necessary to control volatile organic compounds (VOCs) emissions, because such compounds are hazardous to human health, contributing to seriously environmental problems such as the destruction of the ozone layer, photochemical smog and global warming [1,2]. Adsorption technologies for controlling VOC emissions have been recognized as preferred strategies, especially in cases where the captured organic pollutants have alternative uses [3]. Among various adsorbents, activated carbon is the most widely used for VOC adsorption due to its low cost and well-developed microporosity which ensures excellent adsorption capacity [4]. However, serious diffusion restrictions imposed by micropores in activated carbon tend to inhibit its ability to adsorb large VOC molecules. Besides, molecules adsorbed in the micropores of activated carbon are strongly held by adsorption forces [5], hence, it is hard for the adsorbent to be regenerated. It is obvious that the regeneration of saturated adsorbents is a critical factor that should be considered in the adsorbent selection process [6]. Moreover, when it is used under various operational conditions, activated carbon frequently encounters problems, such as fire risk, pore clogging, hygroscopicity, lower selectivity and limited modification flexibility [7,8]. Silica aerogels, prepared by sol–gel processing and subsequent solvent extraction, have large surface area, high porosity and
∗ Corresponding author. Tel.: +86 10 62849194; fax: +86 10 62849194. ∗∗ Corresponding author. Tel.: +86 10 62923564; fax: +86 10 62923564. E-mail addresses:
[email protected] (C. Ma),
[email protected] (Z. Hao). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.019
extremely low densities due to the open pore structure of the mesopores [9]. The unique pore structure permits easy regeneration of the silica aerogels [10,11]. In addition, silica aerogel surface can be chemically modified to improve the hydrophobicity and selectivity for target specific compounds, like VOCs [12]. El Rassy et al. [13] showed that the hydrophobicity of silica aerogel surface was enhanced through the modification with methyltrimethoxylsilane. Standeker et al. [11] found that silica aerogels modified with methyltrimethoxysilane or trimethylethoxysilane exhibited high adsorption capacities for toxic organic compounds from water, which were from 15 to 400 times higher as compared with granulated active carbon. Traditional silica aerogels preparation approaches involve supercritical drying process to avoid capillary stress and related drying shrinkage, whereas the supercritical drying process is energy intensive, expensive and dangerous [14]. To overcome these disadvantages of traditional supercritical aerogel process, a new ambient pressure drying process technique has been developed, making the manufacture and application of silica aerogels in large scale possible [15]. Moreover, the fragility and brittleness of silica aerogels could limit their applications to some extent in various fields [9,16]. Therefore, it is expected that the composite materials derived from the combination of activated carbons and silica aerogels, namely, carbon–silica aerogel composites, may improve the adsorption and desorption performance of the adsorbent and thereby reduce the cost. Carbon–silica aerogel composites have potential applications in various fields, such as catalyst supports [17,18], adsorbents for groundwater treatment [12] and electroconductive materials, etc. [19]. However, there is limited information in literature related to adsorption and desorption properties of carbon–silica aerogel
B. Dou et al. / Journal of Hazardous Materials 196 (2011) 194–200
composites. Coleman et al. [12] found that a cost-effective granulated activated carbon (GAC)–aerogel composite was superior in removing uranium from a stock solution compared with GAC alone. To the best of our knowledge, there has been no report on VOC emission controlling by using carbon–silica aerogel composites as adsorbents, especially in terms of adsorption and desorption kinetics which might play fundamental roles in the application of adsorbent for real situations. In this work, hierarchically structured carbon–silica aerogel composites were synthesized by adding powered activated carbon to the silica sol just before gelation. In order to reduce the CSA composites cost, cheap water glass was used as precursors to synthesize silica aerogels. After solvent exchange and surface modification, the wet gels of composites were subsequently dried by a low-cost technique of ambient pressure drying. Adsorption and desorption performance as well as the corresponding kinetics of VOCs for carbon–silica aerogel composites were intensively studied, comparing with pure silica aerogels and microporous activated carbon. The investigations of adsorption and desorption for VOCs were carried out under both static and dynamic conditions by a digital microbalance and a flow adsorption measurement. We believe that the obtained CSA composites could be effective adsorbents for VOC emission controlling. 2. Experimental 2.1. Material synthesis Silica sols were prepared by using cheap water glass (Be0 = 40, Na2 :SiO2 molar ratio = 1:3.3, Beijing Chemical Works) as precursors. In a typical synthesis, 10 mL of water glass was diluted by deionized water with volume ratio of 1:4 and then mixed with strongly acidic type ion exchange resin so as to remove Na+ ions. After stirring the mixture for 10 min, pH of the solution was decreased in the range of 2–3, resulting in the formation of silicic acid. In the subsequent step, the base catalyst (NH4 OH, 1 M) was added drop by drop to raise pH to 5 with constant stirring. Waiting until the silica sol is just about to gel, 1 g of powered activated carbon (BN-09, <60 mesh, Ningxia BENNIU Activated Carbon Works) was added, followed by stirring for some minutes and keeping the colloidal mixture for gelation at room temperature. After aging for 3 h, water in the composite gels was exchanged with ethanol for three times in 36 h. To prevent the reverse reaction of surface modification [20], ethanol was exchanged with n-hexane for three times in 36 h. The gels were modified with 20% trimethylchlorosilane (TMCS) in n-hexane for 24 h and then the unreacted TMCS was washed with n-hexane for two times in 24 h. Finally, the wet gels of composites were dried under ambient pressure at 60 ◦ C, 80 ◦ C, 120 ◦ C and 180 ◦ C for 6 h, respectively. The obtained CSA composites were denoted as CSA-x, where x is referred to the mass fraction of activated carbon in wet gels. 2.2. Characterizations and measurements Nitrogen adsorption and desorption measurements on CSA composites were conducted at liquid nitrogen temperature (−196 ◦ C) by using NOVA 1200 gas sorption analyzer. Each sample was degassed under vacuum condition at 120 ◦ C for 18 h prior to the measurement. The BET specific surface area was calculated by using adsorption data at relative pressure (P/P0 ) of 0.05–0.25, and the total pore volume was estimated from the amount adsorbed at a relative pressure of about 0.99. Pore size distributions (PSD) of CSA composites were calculated from the analysis of desorption branch of the isotherm, by using Barrett–Joyner–Halenda (BJH) algorithm. Fourier transform infrared (FT-IR) spectroscopy was measured by
195
the KBr method recorded on a Bruker Tensor 27, scanned from 4000 to 600 cm−1 . The thermogravimetric analysis (TGA) of adsorbents was performed on a TG/DTA analyzer (Setaram, Labsys). The heating rate was 10 ◦ C min−1 from 40 to 700 ◦ C under an airflow of 30 mL min−1 . 2.3. Static adsorption and desorption measurements The static adsorption and desorption equilibrium measurements for CSA composites were recorded by using an Intelligent Gravimetric Analyzer (Model IGA-002, Hiden Isochema Instrument) with a sensitivity of 0.1 g. The apparatus had an ultrahigh vacuum system allowing isotherms and the corresponding kinetics to be determined by setting pressure steps. For each pressure increment or decrement, weight changes due to adsorption or desorption were used to calculate kinetic parameters. Before the measurement, each sample was degassed at 110 ◦ C over night. 2.4. Dynamic adsorption and desorption measurements The investigations of dynamic adsorption and desorption of VOCs for CSA composites were carried out by a flow method on an experimental set-up [21]. After degassing at 110 ◦ C for 24 h, about 1 g of adsorbent (40–60 mesh) was packed into the adsorption bed. The adsorption measurement was performed under both dry and wet conditions. The concentration of benzene was controlled at 700 ppm and the total flow rate was 100 mL min−1 . To observe the effects of water vapor on adsorption behaviors of CSA composites, test gases consisted of nitrogen, 700 ppm of benzene under the relative humidity (RH) of 18%, were passed through the adsorption bed. After samples were saturated with benzene vapor, the benzene saturator was bypassed to start purging samples with nitrogen flow at room temperature. In addition, dynamic desorption properties of adsorbents were investigated by temperature programmed desorption (TPD) technique. The concentration of benzene before and after the adsorption process was tested by using gas chromatograph (GC) equipped with a flame ionization detector. 3. Results and discussion 3.1. Characterization of carbon–silica aerogel composites Fig. 1 shows the nitrogen adsorption and desorption isotherms of CSA composites at −196 ◦ C and the corresponding pore size distributions obtained by the BJH method. For comparison, pure silica aerogel CSA-0 and commercial AC were studied. As can be seen in Fig. 1a, the shape of adsorption isotherms for all CSA samples can be considered as a combination of type I and type IV, according to the IUPAC classification [22]. The steep increase of adsorbed amount at low relative pressure (P/P0 < 0.1) indicates the presence of micropores in these materials. At intermediate P/P0 , a capillary condensation step and an H2-type hysteresis loop was observed, suggesting the formation of worm-like mesopores during the synthesis process of CSA composites [23]. However, the adsorption isotherm for AC is of type I, which is typical for microporous activated carbon. With increasing mass fraction of activated carbon, the microporosity of CSA composites increases while the total pore volume decreases (see Table 1). The nitrogen adsorption and desorption results provide evidence for the existence of both micropores and mesopores in CSA composites. The FT-IR spectrum of modified silica aerogel CSA-0 and carbon–silica aerogel composite CSA-2 is shown in Fig. 2. The FT-IR spectrum represents a broad band at around 3500 cm−1 and a peak at around 1600 cm−1 , which can be attributed to O–H groups [24]. The peaks at around 1100 and 800 cm−1 are due to asymmetric and symmetric modes of SiO2 [25]. The peaks at 2980 and 1450 cm−1
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B. Dou et al. / Journal of Hazardous Materials 196 (2011) 194–200
a
b
CSA-0
1200
CSA-2
1000
CSA-0
CSA-4 dV/dD
3 -1
Volume Adsorbed ( cm g , STP)
1400
800
CSA-6
CSA-2
600 400
CSA-4
AC
CSA-6
200
AC
0 0.0
0.2
0.4
0.6
0.8
1.0
1
2
Reletive pressure (P/Po )
3
4
5
6
7
8
9 10
Pore dimeter (nm)
Fig. 1. Nitrogen adsorption and desorption isotherms (a) and BJH pore size distributions (b) for CSA composites. The isotherms for CSA-6, CSA-4, CSA-2 and CSA-0 materials are shifted by 300, 500, 700 and 900 cm3 g−1 STP, respectively.
Table 1 Textural properties of carbon–silica aerogel composites. Sample
BET surface area (m2 g−1 )
Average pore diametera (nm)
Total pore volume (cm3 g−1 )
Micropore volume (cm3 g−1 )
Micropore area (m2 g−1 )
CSA-0 CSA-2 CSA-4 CSA-6 AC
726 727 710 758 936
3.9 3.6 3.6 2.8 2.2
0.71 0.66 0.64 0.53 0.52
– 0.05 0.05 0.18 0.33
– 102 110 393 715
a
Calculated using the Barrett–Joyner–Halenda (BJH) model based on the desorption branch of the isotherm.
are assigned to C–H bonds, while the peak at 840 cm−1 is attributed to Si–C bond. FT-IR results clearly confirm that methyl groups are covalently anchored onto the CSA composite surface. The thermal stability of carbon–silica aerogel composites was observed by TGA–DTA analysis, as shown in Fig. S1 (Supporting Information). The weight loss at temperature lower than 150 ◦ C is attributed to the loss of adsorbed water and the evaporation of residual organic compounds. The weight loss at temperature higher than 400 ◦ C is due to the oxidation of the –CH3 on the surface of adsorbents developed by surface modification with TMCS (Fig. S1a),
Absorbance (a.u.)
CSA-2
4000
CSA-0 Si-OH C- H
3000
Si-OH
C- H Si-O-Si
2000
Wavenumber
Si-C
1000
(cm-1)
Fig. 2. FT-IR spectra of patterns for CSA-0 and CSA-2 materials.
implying that the retention of surface hydrophobicity can be up to 400 ◦ C.
3.2. Static adsorption and desorption behaviors The carbon–silica aerogel composite CSA-2 with developed microporosity and open mesostructure was adopted here as a representative sample to estimate its adsorption and desorption properties for VOCs, comparing with pure silica aerogel CSA-0 and microporous AC. Fig. 3 shows the adsorption and desorption isotherms of benzene on CSA-0 (Fig. 3a), CSA-2 (Fig. 3b) and AC (Fig. 3c) at temperatures of 25, 35, and 45 ◦ C, respectively. The shapes of adsorption and desorption isotherms of benzene for all the samples are similar to the nitrogen isotherms. It is found that the amount of benzene adsorbed at low pressure increases in the order of AC > CSA-2 > CSA-0, with the augmentation in microporosities of the materials. The adsorption in micropores at low pressure is a direct consequence of the overlap in the adsorption field from adjacent walls of micropores. In comparison with AC (Fig. 3c), the adsorption isotherms of benzene for both CSA-0 (Fig. 3a) and CSA-2 (Fig. 3b) exhibit capillary condensation steps which is attributed to the presence of mesopores in these materials. Therefore, the adsorption isotherms of benzene on CSA-0 and CSA-2 exhibit a combination of microporous and mesoporous adsorptive behavior. As shown in Table 2, the equilibrium adsorption capacity of benzene for CSA-2 (5.06 mmol g−1 ) is higher than that for microporous AC (4.37 mmol g−1 ) at relatively high temperature of 45 ◦ C, and the equilibrium adsorption capacity changes in the order of CSA-2 > CSA-0 > AC. This result can be explained by the fact that
B. Dou et al. / Journal of Hazardous Materials 196 (2011) 194–200
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Table 2 Equilibrium adsorption capacities Qe (mmol g−1 ) and Henry constants K (10−5 mol g−1 Pa−1 ) of benzene on CSA-0, CSA-2 and AC at different temperatures. Equilibrium adsorption capacities Qe (mmol g−1 )
Samples
◦
CSA-0 CSA-2 AC
◦
25 C
35 C
45 C
25 ◦ C
35 ◦ C
45 ◦ C
7.42 5.18 4.82
7.17 5.20 4.74
7.29 5.06 4.37
0.32 0.40 0.94
0.28 0.37 0.77
0.20 0.22 0.29
the static VOC adsorption capacity is proportional to the total pore volume, in accordance with the literature [26]. The Henry constants, which reflect adsorption affinity in the line region of the adsorption isotherm, can be estimated from the adsorption date at very low pressure (P < 3 mbar), where the interaction between adsorbed molecules may be neglected and only interactions between adsorbed molecules and the surface remain [27]. The Henry constants of benzene adsorption at different temperatures are listed in Table 2. As expected, Henry constants of all the samples decrease with the increase of temperature. At the same temperature, Henry constants increase in the order of AC > CSA-2 > CSA-0, with the augmentation in microporosities of the CSA composites. This result can be explained by strong adsorption energy in micropores [21]. In the micropores, due to dispersion interactions, field superposition from the close separated opposing pore walls enhances the physical adsorption behavior. In addition, the isosteric heat of adsorption represents the strength of the interaction between adsorbent and adsorbate, hence it can be used to gauge the compatibility between them. Additionally, the isosteric
8
a
6 4
25 C o 35 C o 45 C
2
120
160
200
240
-1
Amount adsorbed ( mmol g )
80
b
5 4 3
o
25 C o 35 C o 45 C
2 1 0 0
40
80
120
160
200
240
c
5 4 3
o
25 C o 35 C o 45 C
2 1 0 0
40
80
120
160
3.3. Adsorption and desorption kinetics Adsorption and desorption kinetics are of fundamental importance in actual applications [30]. The mass relaxation curve cursed by pressure increment or decrement can be used to calculate the adsorption and desorption kinetic parameters. Studies found that the linear driving force (LDF) model was followed in most cases of the adsorption kinetics of various gases/vapors on activated carbons [30], metal organic framework materials [31], carbon molecular sieves [32] and silica mesoporous materials [33]. The LDF model for adsorption is described by the following equation:
200
(1)
where Mt is the uptake at time t, Me is the equilibrium uptake for the given pressure increment, and k is the rate constant. The corresponding LDF model for desorption is described by the equation:
0 40
heat of adsorption (Hads ) can be calculated from the measured adsorption equilibrium data at different temperatures by using the Clausius–Clapeyron equation [28]. The adsorption isosteric heat for adsorption of benzene versus the loading on CSA-0, CSA-2 and AC is shown in Fig. S2 (Supporting Information). For all the samples, the isosteric heat of adsorption reaches a constant value after an initial sharp decrease. This decrease may be attributed to the heterogeneity of adsorption sites [29]. Afterwards, owing to the confinement effect, the adsorption heat decreases gradually with the increase in adsorbate loading and then gets close to the heat of liquefaction in bulk liquid phase.
Mt = 1 − e−kt Me
o
0
Henry constants K (10−5 mol g−1 Pa−1 ) ◦
240
P (mbar) Fig. 3. Adsorption and desorption isotherms of benzene on CSA-0 (a), CSA-2 (b) and AC (c) materials at different temperatures.
Mt = e−kt − 1 Me
(2)
where Me is the amount desorbed at equilibrium for the given pressure decrement and Mt is the amount desorbed at time t [30]. Fig. 4 shows typical graphs of Mt /Me versus time for adsorption and desorption of benzene and the corresponding fits for the LDF model on CSA-0, CSA-2 and AC, respectively. It is evident that adsorption kinetics for the three adsorbents obeys the LDF model, which indicates that the diffusion through a barrier is the rate determining step. The barrier is either due to constrictions in the porosity with similar size to that of benzene or a surface diffusion barrier [33]. Obviously, the desorption kinetics also obeys the LDF model, similar to the adsorption ones. A comparison of the variation of adsorption and desorption rate constants of benzene with relative pressure at 35 ◦ C on CSA-0, CSA-2 and AC is shown in Fig. 5. In the initial uptake region, the adsorption rate constants for CSA-0 (Fig. 5a) and CSA-2 (Fig. 5b) are significantly faster than those for AC (Fig. 5c). This observation can be attributed to the fact that the presence of open mesoporous structure in CSA-2 and CSA-0 allows high accessibility of adsorptives. For all the samples, the adsorption rate constants initially decrease to a plateau (observed at P/P0 ∼ 0.4) before increasing with relative pressure. The decrease in rate constants can be attributed to the development of benzene molecular clusters in micropores of adsorbents. On the other hand, the increase in rate constants when the relative pressure is greater than 0.6 is probably related to the growth of the benzene molecular clusters filling the
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1.0
0.0
a
-0.5
0.5
-1.0
0.0 0
500
1000
1500
0
2000
1.0
500
1000
1500
2000
0.0
b
e Mt / Me
Mt / M e
d
0.5
-0.5
-1.0
0.0 0 1.0
500
1000
0
1500
500
1000
0.0
c
f
-0.5
0.5
-1.0
0.0 0
200
400
Time (s)
0
200
400
600
800
Time (s)
Fig. 4. Adsorption kinetic profiles of benzene on (a) CSA-0 (P/P0 = 0.576–0.614), (b) CSA-2 (P/P0 = 0.577–0.614) and (c) AC (P/P0 = 0.577–0.615), and desorption kinetic profiles on (d) CSA-0 (P/P0 = 0.355–0.316), (e) CSA-2 (P/P0 = 0.435–0.395) and (f) AC (P/P0 = 0.395–0.356) at 35 ◦ C, line is the LDF model.
12
mesopores. In the case of desorption, the rate constants are similar to the adsorption for given pressure increments or decrements. Particularly, the rate constant for the final desorption step is much slower than that for the corresponding adsorption increment. This result is attributed to the difficulty in removing adsorbate from ultramicropores in adsorbents, which is caused by the overlap of potential fields from pore walls [34].
a
10 8 6 4 2 0 -3
8
Rate Constant ( s , 10 )
10
-1
0.0
0.2
0.4
0.6
0.8
1.0
b
6 4 2 0 0.0
0.2
0.4
0.6
0.8
10
1.0
c
8 6 4 2 0 0.0
0.2
0.4
0.6
0.8
1.0
Relative Pressure, P/P0 Fig. 5. Variation of rate constants for adsorption (filled symbols) and desorption (open symbols) of benzene with relative pressure on CSA-0, CSA-2, and AC at 35 ◦ C.
3.4. Dynamic adsorption and desorption behaviors The investigations of dynamic adsorption and desorption were conducted for the removal of typical VOCs, such as benzene, by representative samples of CSA-0, CSA-2 and AC. A breakthrough measurement is a direct method designed to explore the dynamic performance of VOC adsorption at low concentrations [26]. The breakthrough curves of benzene adsorption under dry and wet conditions and the corresponding purge curves with nitrogen on CSA-0, CSA-2 and AC samples are shown in Fig. 6. Generally, longer breakthrough time results in higher dynamic adsorption capacity. The dynamic adsorption capacity (Qa ) was obtained by measuring the area between the maximum baseline and experimental curves, while the desorption capacity (Qd ) was obtained from the area between minimum baseline and experimental curve profiles [35]. The dynamic adsorption and desorption capacities of benzene on the CSA-0, CSA-2 and AC are listed in Table 3. In the adsorption region (Fig. 6a), all the samples exhibit slightly longer breakthrough time for benzene under dry condition as compared with that under wet condition (18% RH). Correspondingly, a /Q a ) for all Table 3 shows that the adsorption efficiency (Qwet dry the samples is relatively high and the dynamic adsorption capacity under wet condition is about 85% of that under dry condition, suggesting the hydrophobic surface of adsorbents. It is obvious that AC represents the longest breakthrough time of benzene. However, the post-breakthrough curves of AC for benzene adsorption
B. Dou et al. / Journal of Hazardous Materials 196 (2011) 194–200
a
1.0
199
b
CSA-0 CSA-2 AC
C/Co
0.8 0.6 0.4 0.2 0.0 0
100
200
300
400
500
600
700
t (min) Fig. 6. The breakthrough curves for benzene adsorption (a) and the purge curves with nitrogen (b) on CSA-0, CSA-2, and AC materials. Curves with filled symbols represent results obtained under dry conditions and curves with empty symbols represent results obtained under wet conditions (18% RH).
Table 3 Dynamic adsorption and desorption capacities of CSA-0, CSA-2 and AC.
CSA-0 CSA-2 AC
Dynamic adsorption capacity Qa (mmol g−1 )
Dynamic desorption capacity Qd (mmol g−1 )
a Qdry
a Qwet
a a Qwet /Qdry (%)
Desorption by d purge Qpurge
Desorption by d TPD QTPD
0.88 1.58 3.99
0.76 1.34 3.41
86 85 85
0.68 1.11 1.86
0.20 0.46 2.12
increase more gradually with time compared with the other two samples, implying significantly large mass transfer resistance in AC [26]. The dynamic adsorption capacity of benzene increases in the order of AC > CSA-2 > CSA-0 (Table 3), corresponding to the micropore volume of the adsorbents, which is consistent with the literature [26]. This result confirms that the presence of micropores in the adsorbents is an essential factor in determining the dynamic adsorption capacity. In contrast, the post-breakthrough sharpness of the increase in benzene concentration for CSA-0 and CSA-2 is more rapid, implying less diffusion resistance in the adsorbents during the adsorption process. The dynamic adsorption results indicate that the carbon–silica aerogel composite CSA-2 exhibits an excellent adsorption performance with higher dynamic adsorption capacity than silica aerogel CSA-0 and less mass transfer resistance than microporous AC. In the case of the purge region from Fig. 6b, the amount of benzene adsorbed on the external surface or relatively large pores of adsorbents can be desorbed during the nitrogen purge. The desorption breakthrough curve of AC for benzene decreases more slowly than those of CSA-0 and CSA-2. The VOC molecules weakly adsorbed on the external surface or the open mesopores of CSA-0 and CSA-2 adsorbents can be easily desorbed by purge gases. However, VOC molecules adsorbed in the micropores of AC are strongly held by adsorption forces and the strongly adsorbed molecules can resist the effects of purge gases. Fig. 7 shows the TPD curves of benzene on CSA-0, CSA-2, and AC materials. The desorption peaks of benzene for TPD process at around 75 ◦ C for CSA-0, 80 ◦ C for CSA-2 and 105 ◦ C for AC were observed. Furthermore, the desorption peaks for CSA-0 and CSA-2 are significantly sharp, suggesting that the desorption of benzene molecules from CSA-0 and CSA-2 is fast. In contrast, the desorption peak for AC is broad, demonstrating that the desorption of VOCs from micropores in AC is slow. The temperature for complete elimination of residual VOCs from CSA-0 and CSA-2 is significantly lower than that from AC, and the decrease of temperature is very important from the practical point of energy saving. To test the reusability of the CSA composites, CSA-2 was regenerated by thermal treatment at 120 ◦ C for 4 h, and the regenerated adsorbent was subjected to adsorption/desorption cycles at least 7 times with-
1.0
CSA-0 CSA-2 AC
0.8
C/Co
Samples
0.6 0.4 0.2 0.0
50
100
150
200
250
300
350
T (ºC) Fig. 7. TPD curves of benzene on CSA-0, CSA-2, and AC materials.
out observing any appreciable loss in its adsorption capacity (see Fig. S3 in the Supporting Information). 4. Conclusions In this study, carbon–silica aerogel composites were synthesized from cheap water glass precursors and granulated activated carbon. The hydrophobic surface of the CSA composites was obtained by post-synthesis modification with TMCS and subsequent ambient pressure drying procedure. The resultant CSA composites exhibit both microporous and mesoporous structures. As indicated by the static adsorption–desorption results, CSA composites have good affinity towards aromatic molecules and fast adsorption kinetics. The equilibrium adsorption capacity of benzene on CSA-2 (5.06 mmol g−1 ) is higher than that on AC (4.37 mmol g−1 ). The carbon–silica aerogel composite CSA-2 exhibits the best dynamic adsorption and desorption performance, because CSA-2 shows higher adsorption capacity than silica aerogel CAS-0 and less mass transfer resistance than AC. Furthermore, owing to the open mesostructure of CSA-2, the desorption of VOCs from CSA-2 by purge or TPD process is significantly faster than AC. With excellent adsorption and desorption performance, as potential adsorbents, the cost-effective CSA composites with hierarchical micro/mesoscopic structure could have a promising future in VOC emission controlling.
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Acknowledgments This study was financially supported by the Special CoConstruction Project of Beijing Municipal Commission of Education, National Natural Science Foundation of China (20725723, 20807050), National Basic Research Program of China (2010CB732300). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.019. References [1] T. Yamamoto, S. Kataoka, T. Ohmori, Characterization of carbon cryogel microspheres as adsorbents for VOC, J. Hazard. Mater. 177 (2010) 331–335. [2] M.E. Ramos, P.R. Bonelli, A.L. Cukierman, M.M.L. Ribeiro Carrott, P.J.M. Carrott, Adsorption of volatile organic compounds onto activated carbon cloths derived from a novel regenerated cellulosic precursor, J. Hazard. Mater. 177 (2010) 175–182. [3] R. Serna-Guerrero, A. Sayari, Applications of pore-expanded mesoporous silica. 7. Adsorption of volatile organic compounds, Environ. Sci. Technol. 41 (2007) 4761–4766. [4] N. Mohan, G.K. Kannan, S. Upendra, R. Subha, N.S. Kumar, Breakthrough of toluene vapours in granular activated carbon filled packed bed reactor, J. Hazard. Mater. 168 (2009) 777–781. [5] K.J. Kim, C.S. Kang, Y.J. You, M.C. Chung, M.W. Woo, W.J. Jeong, N.C. Park, H.G. Ahn, Adsorption–desorption characteristics of VOCs over impregnated activated carbons, Catal. Today 111 (2006) 223–228. [6] D. Das, V. Gaur, N. Verma, Removal of volatile organic compound by activated carbon fiber, Carbon 42 (2004) 2949–2962. [7] W.G. Shim, J.W. Lee, H. Moon, Adsorption equilibrium and column dynamics of VOCs on MCM-48 depending on pelletizing pressure, Micropor. Mesopor. Mater. 88 (2006) 112–125. [8] P. Liu, C. Long, Q. Li, H. Qian, A. Li, Q. Zhang, Adsorption of trichloroethylene and benzene vapors onto hypercrosslinked polymeric resin, J. Hazard. Mater. 166 (2009) 46–51. [9] M. Moner-Girona, E. Martinez, J. Esteve, A. Roig, R. Solanas, E. Molins, Micromechanical properties of carbon–silica aerogel composites, Appl. Phys. A: Mater. 74 (2002) 119–122. [10] A. Venkateswara Rao, N.D. Hegde, H. Hirashima, Absorption and desorption of organic liquids in elastic superhydrophobic silica aerogels, J. Colloid Interface Sci. 305 (2007) 124–132. [11] S. Standeker, Z. Novak, Z. Knez, Adsorption of toxic organic compounds from water with hydrophobic silica aerogels, J. Colloid Interface Sci. 310 (2007) 362–368. [12] S.J. Coleman, P.R. Coronado, R.S. Maxwell, J.G. Reynolds, Granulated activated carbon modified with hydrophobic silica aerogel-potential composite materials for the removal of uranium from aqueous solutions, Environ. Sci. Technol. 37 (2003) 2286–2290. [13] H. El Rassy, P. Buisson, B. Bouali, A. Perrard, A.C. Pierre, Surface characterization of silica aerogels with different proportions of hydrophobic groups, dried by the CO2 supercritical method, Langmuir 19 (2003) 358–363. [14] P.R. Aravind, P. Mukundan, P. Krishna Pillai, K.G.K. Warrier, Mesoporous silica–alumina aerogels with high thermal pore stability through hybrid sol–gel route followed by subcritical drying, Micropor. Mesopor. Mater. 96 (2006) 14–20. [15] F. Shi, L. Wang, J. Liu, Synthesis and characterization of silica aerogels by a novel fast ambient pressure drying process, Mater. Lett. 60 (2006) 3718–3722.
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Journal of Hazardous Materials 196 (2011) 201–209
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Preparation and characterization of immobilized [A336][MTBA] in PVA–alginate gel beads as novel solid-phase extractants for an efficient recovery of Hg (II) from aqueous solutions Yun Zhang a,∗ , Daniel Kogelnig b,c , Cornelia Morgenbesser b , Anja Stojanovic b,d , Franz Jirsa b , Irene Lichtscheidl-Schultz e , Regina Krachler b , Yanfeng Li a , Bernhard K. Keppler b a State Key Laboratory of Applied Organic Chemistry, College of Chemistry and Chemical Engineering, College of Resources and Environment, Lanzhou University, Lanzhou 730000, PR China b University of Vienna, Institute of Inorganic Chemistry, Waehringer Strasse 42, 1090 Vienna, Austria c Cytec Surface Specialties S.A./N.V., Anderlechtstraat 33, 1620 Drogenbos, Belgium d Martin-Luther-University Halle-Wittenberg, Institute of Chemistry, Division of Technical and Macromolecular Chemistry, Faculty of Natural Sciences II, Von-Danckelmann-Platz 4, D-06120 Halle, Germany e University of Vienna, Core Facility of Cell Imaging and Ultrastructure Research, Althanstrasse 14, A-1090 Vienna, Austria
a r t i c l e
i n f o
Article history: Received 12 May 2011 Received in revised form 12 August 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: PVA–alginate–boric acid Hg (II) [A336][MTBA] PVA/IL
a b s t r a c t The coarse PVA–alginate matrix gel beads entrapping the micro-droplets of the ionic liquid tricaprylylmethylammonium 2-(methylthio) benzoate ([A336][MTBA]) as novel solid-phase extractants were prepared for the removal of mercury (II) from aqueous media. The ionic liquid [A336][MTBA] immobilized PVA–alginate beads (PVA/IL) have been characterized by FTIR, SEM and TGA. The influence of the uptake conditions was investigated including aqueous pH, PVA/IL dosage, the content of [A336][MTBA] and initial Hg (II) concentration; maximum Hg (II) ion adsorption capacity obtained was 49.89 (±0.11) mg g−1 at pH 5.8 with adsorptive removal of approximately 99.98%. The selectivity of the PVA/IL beads towards Hg (II), Pb (II) and Cu (II) ions tested was Hg > Pb > Cu. The rate kinetic study was found to follow second-order and the applicability of Langmuir, Freundlich and Tempkin adsorption isotherm model were tested as well. The results of the study showed that PVA/IL beads could be efficiently used as novel extractants for the removal of divalent mercury from aqueous solutions under comparatively easy operation conditions. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Mercury and its compounds are one of the most toxic chemicals present in the environment, which can be easily accumulated in humans and animals mainly causing neurological and renal disturbances as well as impairment of pulmonary function [1]. The recommended permissible limit of Hg (II) in potable water is 0.001 mg L−1 [2]. Therefore, industry is facing many constraints because of the increasingly drastic regulations concerning
Abbreviations: PVA, poly(vinyl alcohol); IL, ionic liquid); FTIR, Fourier transform infrared spectroscopy); SEM, scanning electron microscopy); ATR, attenuated internal total reflection); TGA, thermal gravimetric analysis); SDTA, scalable decision tree algorithm); EDAX, energy dispersive X-ray analysis); [A336][MTBA], tricaprylylmethylammonium 2- (methylthio) benzoate); ACS, American chemical society); AAS, atomic absorption spectroscopy); FIMS, flow injection mercury system); CaCl2 , calcium chloride); DOC, dissolved organic carbon); NPOC, non purgeable organic carbon); TOC, total organic carbon); LOD, limit of determination. ∗ Corresponding author. Tel.: +86 931 8912528; fax: +86 931 8912113. E-mail address:
[email protected] (Y. Zhang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.018
wastewater discharge. Adsorption has been considered as an effective alternative among conventional processes especially for metal ions recovery from dilute solutions. The development of suitable sorbents characterized by a high efficiency and selectivity has been a problem for many years. Recently, the solid-phase adsorbents combining the efficiency of solvent extraction systems (liquid/liquid extraction) and the stability of resins (sorption process) have found the widest application because of easy separation and simplicity of their preparation [3]. Different techniques have been used for manufacturing such solid-phase extractants like impregnation and encapsulation or immobilization of solvents with support materials [4]. The application of ionic liquids (ILs) opens a new route to prepare novel solid-phase adsorbents because of some unique physicochemical properties: a very low (often negligible) vapor pressure, low melting points and good thermal stability [5]. A number of papers have been devoted to the description of synthesis and properties of various ILs, including imidazoliumbased ILs and phosphonium-based ILs focusing on their potential for the extraction of organic and inorganic compounds [6]. The
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Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209 _ O
O
C8H17 N
+
C8H17
S
C8H17
Tricaprylmethylammonium, [A336]
2-(Methylthio)benzoate, [MTBA]
Fig. 1. The structure of the ionic liquid [A336][MTBA].
ability of ILs to be retained on solid surfaces allows one to prepare sorption materials by means of immobilization of ILs on different matrices. Guibal et al. [7] tested adsorption performance of Hg (II) by capsules prepared by immobilization of Cyphos IL-101(trihexyl(tetradecyl)phosphonium chloride) in alginate particles. However, most of those supports reported for the immobilization of ILs focused on alginate, chitosan and mineral materials [6,8]. Poly(vinyl alcohol) (PVA) is a water-soluble material containing large amounts of hydroxide groups that has been developed for biomedical applications since it is biocompatible and exhibits minimal cell adhesion. It has been reported that PVA gel has no degradation by microorganisms and a higher mechanical strength and larger durability in high acid solutions than alginate gel, which has been mostly employed as polymer for microcapsules [9]. So combining PVA gel with ILs would be an interesting and promising solution for developing novel solid-phase sorbents leading to easy possible design for extracting heavy metals industrially. The present work proposes the coarse PVA–alginate matrix gel beads entrapping the microdroplets of the IL [A336][MTBA], which contains electron donor atoms: S and O. At first, the immobilized [A336][MTBA] PVA–alginate gel beads were prepared and characterized. Then the performance of a Hg (II) extraction process was evaluated in terms of the Hg (II) removal efficiency, the adsorption mechanism, optimization of operation and the effect of coexisting ions. This alternative solid-phase extraction system was expected to be simple, competitive, efficient and environmentally friendly for treating wastewater containing Hg (II) ions. 2. Experimental 2.1. Materials and characterization of the PVA/IL beads [A336][MTBA] was synthesized in the lab as reported before (shown in Fig. 1) [10]. Polyvinyl alcohol (99+% hydrolyzed, average Mw 146,000–186,000) and sodium alginate with a viscosity of 3500 cps (2% solution, 25 ◦ C) were purchased from Sigma–Aldrich, Germany. All the other chemicals in this study were of ACS reagent grade and used as received. Double deionised water (Milli-Q Millipore 18.2 M cm−1 conductivity) was used for all dilutions. A ProLab2000 pH meter was used for the measurement of pH values in the aqueous phase. The concentration of mercury remnant in the solution was determined by a FIMS 400 Mercury Analysis System (PerkinElmer). Lead and copper concentrations were determined using flame-AAS with an AAanalyst 200 by PerkinElmer. FT-IR spectra of dried PVA beads, ILs, PVA/IL and Hg (II) loaded PVA/IL beads were recorded using a Bruker Vertex 70 FT-IR spectrometer with ATR device (4000–400 cm−1 ). Thermogravimetric measurements were conducted on a Mettler-Toledo TGA/SDTA 851e analyzer at a scan rate of 10 ◦ C min−1 over a temperature window of 25–700 ◦ C. Nitrogen was used as the purge gas. SEM–X-ray microanalyses: dried beads were split after cooling them in fluid nitrogen and sputtered with carbon prior to the examination in a scanning electron microscope (SEM) XL20 by Philips. Element determination was done by an X-ray microanalysis system (EDAX). DOC was measured
as NPOC with a TOC-VCPH total organic carbon analyser in combination with a total nitrogen measuring unit TMN-1 by Shimadzu, bringing the samples to pH ≈ 2 with HCl (p.a. by Fluka) and sparging them with carrier gas for 5 min prior to combustion. Element concentrations were calculated from the corresponding regression lines (correlation factor >0.9995) using five different dilutions of a potassium hydrogen phthalate solution. LOD was determined as 0.5 mg L−1 C. 2.2. Preparation of PVA/IL beads The PVA–alginate–boric acid immobilization method was used. It was reported that monodiol-type is the cross-linking mechanism between PVA and boric acid [9]. The recommended percentage of PVA in the beads was kept in the range of 8–12.5% (w/v) to get high bead strength as PVA contributed strength and durability to the beads, whereas calcium alginate improved the surface properties, reducing the tendency to agglomerate [11]. In this work, 4 g PVA and 0.675 g sodium alginate were dissolved in 50 mL of Millipore water with boiled water bath for 1 h. The PVA (8%, w/v)–alginate solution and a required amount of [A336][MTBA] were then blended together with stirring at 500 rpm at 30 ◦ C for 6 h to obtain a homogeneous gel blend which was extruded into a gently stirred saturated 3% (w/v) CaCl2 –boric acid solution by a syringe with a diameter of approximately 1 mm and immersed for 24 h to form spherical beads. The PVA/IL beads were washed with Millipore water to remove residual reagents. The wet spherical PVA/IL beads exhibit an elastic property with an average diameter of 1.6 (±0.02) mm and swelling ratio of approximately 56.51 (±0.13)% studied in 100 ml Millipore water using 2 g wet beads. The term wet means the state of the beads immediately after the preparation. After 24 h, the beads were separated from the water. Immediately, the beads were wiped gently with paper and weighed. The swelling ratio was calculated according to the formula: Swelling ratio = (Ws − Wi )/Wi × 100%, where Ws is the weight of the adsorbent in the swollen state and Wi is the initial weight of the adsorbent. Finally, the beads stability was tested as follows: 2 g of plain PVA–alginate beads and wet PVA/IL beads were immersed into 50 mL of Millipore water respectively for 72 h. Then DOC in the solutions was measured for each sample. 2.3. Batch sorption procedure A 50 mg L−1 solution of Hg (II) was prepared by dissolving 0.0677 g of mercury (II) chloride in 1 L solutions. The solutions (50 mL) including different amounts of the PVA/IL beads (0.01, 0.05, 0.1, 0.47 and 0.94 weight) at initial pH 6.0 were shaken in an electrically thermostatic reciprocating shaker at 120 rpm for 24 h. The batch adsorption experiments were performed to investigate Hg (II) removal as a function of initial metal concentration (10–50 mg L−1 ), contact time (5 min–24 h), aqueous pH (2–7.0), PVA/IL beads concentration (0.2–20 g L−1 ), the content of [A336][MTBA] in PVA/IL beads and coexisting ions. The equilibrium time was estimated by drawing samples at regular intervals of time till equilibrium was reached. The adsorption capacity of the PVA/IL beads and the percentage removal of Hg (II) are calculated using the following Eqs. (1) and (2) respectively: Adsorption capacity(mg/g) = Hg(II) ions removal(%) =
(Ci − Cf )V/1000 W
(Ci − Cf ) × 100 Ci
(1) (2)
where Ci and Cf are the initial and final concentration of Hg (II) in mg L−1 , V is the volume of Hg (II) solution in mL and W is the total amount of PVA/IL beads in g.
Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209
O
203
O
O Hg S
+
Hg2+
S
Fig. 3. Proposed main schematic mechanism for Hg (II) adsorption by PVA/IL.
Fig. 2. FTIR spectra of [A336][MTBA], PVA, PVA/IL and Hg (II) loaded PVA/IL.
Each determination was replicated two times. For each set of data present, standard statistical methods were used to determine the mean values and stand deviations. Confidence intervals of 95% were calculated for each set of samples to determine the margin of error. 3. Results and discussion 3.1. Characterization of the PVA/IL beads The FT-IR spectra of PVA, [A336][MTBA], PVA/IL and Hg (II) loaded PVA/IL are shown in Fig. 2. Bands at about 1554 and 1464 cm−1 could be assigned to stretching and bending vibrations
of aromatic C–H groups of [A336][MTBA]. The bands at 2855 and 2925 cm−1 correspond to stretching vibrations of CH3 and CH2 from [A336][MTBA] as well. COO– group stretching vibrations were observed at 1598 cm−1 for PVA/IL which may contribute to the Hg (II) adsorption (Fig. 3) since it was reported that PVA show little affinity to the heavy metal ions [12]. These observations indicated that [A336][MTBA] was immobilized in PVA/IL beads successfully. A scanning electron microscope was used for obtaining microscopic images of the immobilized ILs in PVA–alginate gel beads. The shape of the PVA/IL beads and their morphologies of the cross section as well as the porous structure can be seen clearly from Fig. 4(A and B), respectively. From the element analysis by EDAX, it can be clearly seen that the peak of S emerges for PVA/IL beads, which suggests that [A336][MTBA] was wrapped up by porous PVA–alginate matrix after immobilization. And Hg (II) could be found in the PVA/IL beads after adsorption as well (Fig. 4). Fig. 5 shows the thermal degradation behavior of the IL [A336][MTBA], PVA and PVA/IL beads assessed by TGA. The curve of PVA/IL showed that the degradation process occurred between 155 and 520 ◦ C. There were three distinct weight loss stages: 147–248 ◦ C; 248–351 ◦ C; 351–516 ◦ C. The first weight loss stage (about 17%) could be assigned to the decomposition of [A336][MTBA] immobilized onto the PVA–alginate beads; the second and third stages were probably due to the thermal degradation of [A336][MTBA] immobilized into PVA–alginate beads (248–300 ◦ C) and PVA. Thus, it was assumed that the required amount of [A336][MTBA] was successfully immobilized in the PVA–alginate beads by the above mentioned encapsulation method. And the results from the stability test showed only 2.2% of the ionic liquid [A336][MTBA] was released from PVA/IL beads prepared with this method after 72 h.
Fig. 4. Element determination of PVA, PVA/IL and Hg (II) loaded PVA/IL by EDAX ((A): digital photo of PVA/IL; (B): SEM picture of the cross section of PVA/IL).
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they cause precipitation of metal compounds, which then does not allow one to distinguish between the mechanisms of sorption and/or precipitation. According to the speciation diagram of mercury (5 × 10−3 mol L−1 )[13], HgCl2 is the main species in solution below a pH of 7.0. In this study the initial mercury ion concentration was in the range of 10–50 mg L−1 which was much less than the concentration of 5 × 10−3 mol L−1 (1357.5 mg L−1 ), so a pH range of 2–7.0 was chosen. As shown in Fig. 6, the uptake of Hg (II) onto PVA/IL beads is slightly dependent on the pH value. In aqueous solutions with a mercury ion concentration below 20 mg L−1 , the pH hardly affected the adsorption. However, when the Hg (II) initial concentration increased to 30 mg L−1 or above, the pH of the solution showed a visible effect on Hg (II) adsorption by PVA/IL beads. This result agrees with that reported previously by Robichaud et al. [14]. This may be mainly owing to the fact that Hg (II) is a ‘soft’ metal, reported to adsorb independently of the pH value [15] and/or a strong affinity of [A336][MTBA] towards mercury ions. It may be concluded that these PVA/IL beads can be used for an efficient removal of mercury ions from solutions of a wide pH range, which might be of great interest for complex wastewater system.
Fig. 5. Thermogram of the IL [A336][MTBA], PVA and PVA/IL.
10 ppm 40 ppm
Adsorption capacity (mg/g)
50
20 ppm 50 ppm
30 ppm
3.2.2. Effects of [A336][MTBA] content in preparation of PVA/IL beads and the beads dosage To study the effect of [A336][MTBA] content on Hg (II) adsorption, a series of PVA/IL beads with different volume of [A336][MTBA] (0, 0.5 mL, 1 mL, 2 mL, 4 mL) were prepared and the result of the adsorption test is shown in Fig. 7(a): the adsorption capacity of Hg (II) increased with the volume increase of [A336][MBTA]. When the ratio of [A336][MTBA] volume (mL) to PVA mass (g) was 0.25, the removal reached 87.8%. Nearly 100% of Hg (II) was removed when [A336][MBTA]/PVA reached 0.5. This suggests that effective adsorption of mercury ions could be mainly attributed to the IL [A336][MTBA] and a ratio in the range of 0.25–0.5 may be the most economical. The optimization of PVA/IL synthesis should thus take into account the rational use of the IL resource. So PVA/IL beads with the ratio of volume of [A336][MTBA] (mL) to the mass of PVA (g) 0.5 were used in all following experiments. Another important parameter for an economical design is the dosage of PVA/IL beads. Therefore, the adsorption efficiency and capacity for Hg (II) ions as a function of PVA/IL beads dosage was investigated (Fig. 7(b)). The percentage of adsorption removal steeply increased with the PVA/IL loading up to 1 g L−1 . This may be due to the increase in active sites on the adsorbent and thus allowing an easier penetration of the metal ions into the sorption sites.
40
30
20
10 2
3
4
5
6
7
pH Fig. 6. Hg (II) adsorption onto PVA/IL beads as a function of solution pH (dosage: 2 g wet beads L−1 ; room temperature; initial Hg (II) concentration: 50 mg L−1 ; solution volume: 50 mL; contact time: 24 h).
3.2. Hg (II) adsorption 3.2.1. Effect of initial pH value Generally, the solution pH value has a significant effect on the metal binding capacity. Too high pH values should be avoided as
100
100
a
b
50
Hg(II) ions removal (%)
Hg(II) ions removal (%)
80
60
40
20
40
Hg(II) ions removal Adsorption capacity
80
30 70 20
60
10
50
0
40
0 0.0
0.2
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0.6
[A336][MTBA]/PVA(mL/g)
0.8
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Adsorption capacity (mg/g)
90
0
2
4
6
8
10
12
14
16
18
20
PVA/IL beads dosage (g/L)
Fig. 7. Effect of [A336][MTBA] content in preparation of PVA/IL beads (a) and the wet beads dosage (b) on Hg (II) adsorption(room temperature; pH: 5.8; initial Hg (II) concentration: 50 mg L−1 ; solution volume: 50 mL; contact time: 24 h).
Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209
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Adsorption capacity (mg/g)
50
40
Hg in Cu - Hg system Hg in Pb - Hg system Pb in Pb - Hg system Cu in Cu - Hg system
30
20
10 20
30
40
Co-exsiting ion concentration (mg/L)
50
Fig. 8. Selective adsorption of Hg (II) onto the PVA/IL beads in binary system (Pb2+ –Hg2+ , Cu2+ –Hg2+ ) (initial concentration of metals: Pb(II) = 10–50 ppm, Cu(II) = 10–50 ppm, Hg (II) = 50 ppm; pH = 5.78; PVA/IL dosage = 1 g L−1 ; temperature = 298 K, contact time = 24 h).
The maximum adsorptive removal was found to be 99.3% as PVA/IL concentration was 2 g L−1 . The percent removal was almost the same at higher dosages of 9.4 and 18.8 g L−1 . A further increase in PVA/IL dosage over 1 g L−1 did not lead to a significant improvement in adsorptive removal due to the saturation of Hg (II) with active sites of PVA/IL beads. Therefore, the optimal wet PVA/IL dosage was selected as 1 g L−1 for further experiments. 3.2.3. Effect of coexisting ions Adsorption of coexisting metal ions onto the PVA/IL beads was examined in both binary and ternary system. As shown in Fig. 8, generally, the Hg (II) adsorption amount decreased slightly with the increases in the concentration of Pb (II) and Cu (II) varying from 10 to 50 mg L−1 . However, absorbed Hg (II) content in the binary system Pb (II)–Hg (II) was higher than that in the system of Hg (II)–Cu (II). Moreover, the adsorption capacity of PVA/IL beads for Pb (II) was higher than that for Cu (II). The same trend can be seen from the ternary system Pb (II)– Hg (II)–Cu (II). The highest adsorption of Hg (II) may be due to [A336][MTBA], which can combine Hg (II) strongly, while higher adsorption of Pb (II) onto PVA/IL beads could be owing to PVA–alginate matrix, which shows a better affinity to Pb (II) than Cu (II) [16]. It may be concluded that the adsorption affinity of metals onto the PVA/IL beads is in the following order: Hg (II) > Pb (II) > Cu (II). As shown in Fig. 9, Hg (II) ions were dominantly adsorbed attaining equilibrium after 20 h. The results were likely due to high adsorption affinity of Hg (II), which led to substitution of already adsorbed metals on the adsorption sites. Rengaraj [17] also reported that favored adsorption of a metal onto the adsorbent can result in desorption of other metals into solution. It was concluded that even higher concentrations of Cu (II) and Pb (II) ions did not affect the efficient removal of Hg (II) dramatically, which is of significance in practical treatment of Hg (II) from aqueous solutions contaminated with other heavy metals such as Pb (II) and Cu (II). 3.3. Kinetics of mercury ions adsorption and effect of contact time The effect of contact time on the adsorption capacity of the PVA/IL beads was investigated in the time ranges of 5 min–24 h (Fig. 10). It was observed that initial adsorption of Hg (II) was rapid on the PVA/IL beads. The adsorption sites on the PVA/IL beads were quickly covered by Hg (II) and the adsorption rate became dependent on the rate at which the metal ions were transported from the bulk liquid phase to the actual adsorption sites [18]. Then the adsorption rate slowed down after initial 2 h of contact time. The
Fig. 9. Competitive adsorption for Cu(II), Pb(II) and Hg (II) on PVA/IL beads in multimetals solutions (initial concentration of metals: Hg (II) = 50 ppm, Cu (II) = 50 ppm, Pb (II) = 50 ppm; pH 5.78; PVA/ILs dosage = 1 g L−1 ; temperature = 298 K, contact time = 24 h).
maximum capacity was 49.89 (±0.11) mg g−1 at around 20 h. Thus the contact time of 24 h was used in the following sections to ensure the adsorption equilibrium. The kinetics of Hg (II) removal could be explained using pseudo first-order, second-order and Elovich kinetic models to examine the rate controlling mechanisms of the adsorption process. Lagergren showed that the rate of adsorption of solute on the adsorbent is based on the adsorption capacity and followed a pseudo-first-order equation [19]. The linear form of the pseudo first-order equation is described by Eq. (3): log(Qe − Qt ) = log Qe −
kad t 2.303
(3)
where Qe and Qt are the amounts of Hg (II) adsorbed (mg g−1 ) at equilibrium time and at any instant of time, t, respectively, and kad (L min−1 ) is the rate constant of the pseudo-first-order sorption. The values of first order rate constants kad and Qe for the initial Hg2+ concentration of 50 ppm by keeping the adsorbent amount constant (1 g L−1 ) at 298 K are calculated and listed in Table 1. The coefficient of determination (R2 ) is found to be 0.919 and the true value of Qe obtained from experiment, 49. 89 (±0.11) mg g−1 , is not 50
Adsorption capacity (mg/g)
10
40
30
20
10
0 0
5
10
15
20
25
Time (h) Fig. 10. Effect of contact time on adsorption of Hg (II) onto the PVA/IL beads (Hg2+ concentration: 50 mg L−1 ; pH 5.8, PVA/ILs dosage = 1 g L−1 ; temperature: 298 K, contact time: 24 h; volume of solution: 100 mL).
206
Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209
Table 1 Comparison of pseudo-first-order, pseudo-second-order and Elovich kinetic model for Hg (II) adsorption by PVA/IL beads.
a 35
Parameters
Pseudo-first-order
kad (min−1 ) 2.303 × 10−3
Qe1,cal (mg g−1 ) 43.9
R2 0.919
Pseudo-second-order
k2 (g mg−1 min−1 ) 4.82 × 10−4
Qe2,cal (mg g−1 ) 50.7
R2 0.999
˛ 10.2
ˇ 1.09
R2 0.988
Elovich model
30 25
t/Qt
Kinetic model
40
15 10
in agreement with the predicted values as given in Table 1. So the plot of log (Qe − Qt ) versus t for the pseudo-first-order adsorption kinetics was not given here. Ho [20] developed a pseudo-second-order kinetic expression for the sorption system of divalent metal ions using sphagnum moss peat, which is described in the following linear form: t t 1 = + Qt Qe h
5 0 0
50
where k2 is the second-order rate constant, h = k2 Qe 2 can be regarded as the initial sorption rate as t approaches 0. The application of the second-order kinetics by plotting t/Qt versus t as shown in Fig. 11(a) yielded the second-order rate constant, k2 , Qe and the coefficient of determination (R2 ). As can be seen from Table 1, the calculated Qe value shows a good agreement with the experimental value and the obtained value for the coefficient of determination (R2 ) is more than 0.995, which indicates that the second-order kinetic model describes the removal of Hg2+ by PVA/IL beads as solid extractants well. Elovich equation is a rate equation based on the adsorption capacity describing adsorption on highly heterogeneous adsorbents which is commonly expressed as Eq. (5) [21]:
45
400
600
800
Q(mg/g) t
35 30
2
R = 0.9882 25 20 15 10 2
3
4
5
6
7
ln t Fig. 11. Pseudo-second-order (a) and Elovich (b) kinetic model plot for the adsorption of Hg2+ using PVA/IL beads (pH 5.8, PVA/IL dosage = 1 g L−1 , temperature: 298 K, contact time: 24 h).
55
100
50
Hg(II) ions removal (%)
90 Hg(II) ions removal Qe
80
45 40 35
70
30 60 25 50
20 15
40
The initial Hg (II) concentration was adjusted in the range of 10–50 mg L−1 for Hg (II) adsorption under a pH of 5.8 as shown in Fig. 12. The removal amount of Hg (II) increased rapidly with increase in the Hg (II) concentration. As the mercury ions concentration was higher than 30 mg L−1 it became slow. When the Hg (II) concentration was 30 mg L−1 , the amount of Hg (II) absorbed
1400
Qe (mg/g)
3.4. Adsorption equilibrium isotherms and effect of initial Hg2+ concentration
1200
40
(5)
where ˛ (mg g−1 min−1 ) is the initial adsorption rate and ˇ (g mg−1 ) is the desorption constant. Fig. 11(b) represents the application of the linear form of Elovich kinetic equation. The constants ˛ and ˇ are obtained from the intercept and the slope, respectively. The coefficient of determination (R2 ) obtained is 0.9882, which is found to be less than the value calculated using second-order kinetic model as shown in Table 1. Thus it may lead to the conclusion that there were three stages in the adsorption process: an external surface adsorption stage; the gradual adsorption stage where intraparticle diffusion was ratecontrolled; the final equilibrium stage where intraparticle diffusion started to slow down due to the extremely low adsorbate concentrations in solution. Overall, these observations suggest that metal sorption by PVA/IL beads followed the second-order reaction indicating that the process controlling the rate may be a chemical sorption and the correlation coefficients of Elovich equation (above 0.98) at pH 5.8 also may confirm the predominant chemical nature of Hg (II) adsorption on PVA/IL beads.
1000
b
min−1 )
1 1 Qt = ln(˛ˇ) + ln t ˇ ˇ
200
Time (min)
(4)
(g mg−1
R2=0.999
20
10
20
30
40
50
Initial mercury ions concentration (mg/L) Fig. 12. Effect of initial Hg2+ concentration on adsorption of Hg2+ using PVA/IL beads (pH 5.8; PVA/ILs dosage = 0.5 g L−1 , temperature: 298 K, contact time: 24 h).
Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209
0.4
0.3
1 1 + bC0
(7)
The value of RL lies between 0 and 1 for a favorable adsorption, while RL > 1 represents an unfavorable adsorption, and RL = 1 represents the linear adsorption, while the adsorption operation is irreversible if RL = 0. The isotherm data has been linearized using Eq. (6) shown in Fig. 13(a). The Langmuir constant Qm , 50.17 mg g−1 , was close to the experimental data (49.89 (±0.11) mg g−1 ) which indicated a good adsorption on the PVA/IL beads. The Qm of PVA/IL beads was higher than some Qm values of other sorbents reported by literatures (Table 2). Another constant denoting adsorption energy, b, is found to be 0.0025 L mg−1 . The high value of the related coefficient (R2 = 0.9998) obtained indicates a good agreement between the experimental values and isotherm parameters and also confirms the monolayer adsorption of Hg2+ onto PVA/IL beads surface because of the macroporous structure of the PVA/IL beads. The dimensionless parameter RL , a measure of adsorption favorability, is found in the range of 0.0025–0.0126 (0 < RL < 1) when C0 is varying from 10 to 50 ppm. The results confirm the favorable adsorption process for Hg2+ removal using PVA/IL beads, especially when the initial Hg2+ concentration is lower than 50 ppm. 3.4.2. Freundlich isotherm The Freundlich isotherm theory says that the ratio of the amount of solute adsorbed onto a given mass of sorbent to the concentration of the solute in the solution is not constant at different concentrations [30]. The linear Freundlich isotherm is commonly expressed as follows: log Qe = log Kf +
0
1 log Ce n
(8)
where Kf (mg1−1/n L1/n g−1 ) and n (g L−1 ) are the Freundlich constants characteristic of the system. The Freundlich constants Kf and n are obtained by plotting the graph between log Qe versus log Ce . The values of Kf and n are 38.18 and 9.387, respectively. It is found that the coefficient of determination obtained from the Freundlich constants for PVA/IL is 0.892, which is lower than that for Langmuir isotherm model as given in
5
10
15
20
25
Ce (mg/g) 4.0
b
3.8
3.6 2
ln Q e
RL =
R2=0.9998
0.0
(6)
where Qm is the quantity of adsorbate on unit mass of adsorbent (mg g−1 ) and Qe is the amount adsorbed on unit mass of the adsorbent (mg g−1 ) when the equilibrium concentration is Ce (mg L−1 ) and b (L mg−1 ) is the Langmuir constant. A further analysis of the Langmuir equation can be made on the basis of a dimensionless equilibrium parameter, RL , given by Eq. (7):
0.2
0.1
3.4.1. Langmuir isotherm Langmuir model has been widely applied to many metal ions sorption process [22]. The model takes the following linear form: 1 Ce Ce = + Qe Qm bQm
a
0.5
Ce/Qe
by the PVA/IL beads was 49.80 (±0.1) mg g−1 compared with 19.61 (±0.04) mg g−1 when 10 mg L−1 while the adsorption effective of Hg (II) kept constant at nearly 100%. At higher concentrations, more Hg (II) was left un-adsorbed in solution due to the saturation of binding sites. The adsorption capacity on the beads in 50 mg L−1 Hg (II) solution reached the maximum while the Hg (II) ions removal (49.88 (±0.06)%) is much lower than that in 30 mg L−1 (99.92 (±0.12)%). This indicates that PVA/IL beads are effective for being used to treat wastewater with Hg (II) of concentrations lower than 50 mg L−1 . Adsorption isotherms are important to describe the adsorption mechanism for the interaction of Hg2+ on the adsorbent surface for the design of an adsorption process.
207
R = 0.998 3.4
3.2
3.0 0.00E+000 2.00E+007 4.00E+007 6.00E+007 8.00E+007 1.00E+008 2
[ln (1+1/Ce )]
Fig. 13. Isotherm model for the Hg2+ adsorption onto PVA/IL beads (a: Langmuir isotherm model, b: Dubinin–Radushkevich isotherm model).
Table 3. The results indicate that the equilibrium data is not fitted well with the Freundlich isotherm model. 3.4.3. Dubinin–Radushkevich (D–R) isotherm Dubinin and Radushkevich [31] have proposed another isotherm applied to estimate the mean free energy of adsorption (E). D–R equation is represented in a linear form by Eq. (9): ln Qe = ln Qm − Kε2
(9)
where K (mol2 kJ−2 ) is a constant related to mean adsorption energy; and ε is the Polanyi potential, which can be calculated from Eq. (10).
ε = RT ln 1 +
1 Ce
(10)
The sorption energy can also be worked out using the following relationship 1 E= √ −2K
(11)
Fig. 13(b) shows the plot ln Qe versus ε2 at 298 K. The constants, Qm and K obtained are 50.88 mg g−1 and 9.67 × 10−4 mol2 kJ−2 respectively. It has been reported [32] that physisorption processes have adsorption energies less than 40 kJ mol−1 and relatively high values for the heat of adsorption indicate a strong interaction between sorbate and sorbent. The mean free energy of adsorption,
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Y. Zhang et al. / Journal of Hazardous Materials 196 (2011) 201–209
Table 2 Comparison of Hg (II) sorption capacity of PVA/IL beads with other typical sorbents. Adsorbent
Operating conditions
Guava bark Hydrogel sorbents A cation exchanger Carbonized newsprint fiber Commercial sulfur-impregnated AC Natural zeolites PEI-attached PHEMA gel beads PVA/IL ([A336][MTBA]) beads
Qm
pH
T (K)
C0 (mg L−1 )
(mg g−1 )
9.0 5.0 6.0 7.0 4.8 8.5 5.0 5.8
303 293 NA 293 298 293 293 298
50 50 75 100 100 NA NA 50
3.364 13.46 ± 1.15 32.9 48.3 41 57.5 335.2 49.89 ± 0.11
References
Lohani et al. [23] Yetimoglu et al. [24] Anirudhan et al. [25] Aoyama et al. [26] Cai et al. [27] Chojnacki et al. [28] Denizli et al. [29] In this study
NA: not available. Table 3 Isotherm constants and regression data for various adsorption isotherms for adsorption of Hg (II) on PVA/IL beads. R2
Adsorption isotherm
Isotherm parameters
Langmuir
Qm b
50.17 mg g−1 0.0025 L mg−1
0.9998
Freundlich
Kf n
38.18 mg1−1/n L1/n g−1 9.387 g L−1
0.892
Dubinin–Radushkevich
Qm K × 10−4 E
50.88 mg g−1 9.67 439.94 kJ mol−1
0.998
E, is calculated as 439.94 kJ mol−1 (Table 3) indicating that the sorption process could be chemisorption because of [A336][MTBA] on the surface and inner of PVA/IL. The value of coefficient of determination (R2 = 0.998) indicates that the Dubinin–Radushkevich model fits well with the equilibrium data. Overall, the experimental data are found to be fitted well with the Langmuir and Dubinin–Radushkevich models, which confirm favorable adsorption process for Hg2+ removal using PVA/IL beads. This suggests that it may include both physisorption and chemisorption because of the macroporous structure and [A336][MTBA] on the surface and inner of PVA/IL beads. 4. Conclusions The present study dealt with the removal of mercury from aqueous solutions using [A336][MTBA]-immobilized on PVA–alginate beads (PVA/IL). The work was primarily based on the principle of modification of the solid sorbents by introducing liquid extractants in inorganic materials to increase the efficiency and environmentally friendly performance. So far, a few research work on this topic has been carried out using alginate, silica and zeolite as supported materials and phosphonium-based ILs as liquid extractants. The use of PVA and [A336][MTBA] with the same purpose has not been investigated. The present study showed that PVA–alginate gel beads could entrap the ionic liquids successfully characterized by FT-IR, EDAX and TGA. It is found that the prepared PVA/IL beads can be used to remove mercury with an efficiency of more than 99% from a 50 mg L−1 Hg (II) solution with nearly no leakage of the ionic liquid. Immobilized [A336][MTBA] samples with an initial [A336][MTBA] volume of 2 ml and 4 g of PVA in a 50 mL alginate solution showed the most economic adsorption for 50 mg L−1 of Hg (II) solutions. Experiments to characterize the PVA/IL beads including the effect of pH, the initial Hg (II) ions concentration, PVA/IL beads dosage and the contact time revealed that a pH of 5.8 and an adsorbent dosage of 1 g L−1 are the optimum conditions for its operation in batch mode. More than 99% of mercury ions can be removed after 20 h with the maximum adsorption capacity
of 49.89 (±0.11) mg g−1 . Modelling of the sorption equilibrium performed well using the Langmuir and Dubinin–Radushkevich equations, which confirm favorable adsorption process for Hg2+ removal using PVA/IL beads. This suggests that it may include both physisorption and chemisorption because of the porous structure and [A336][MTBA] on the surface and inner of PVA/IL beads. The pseudo-second-order chemical reaction kinetic provided the best correlation with the experimental data as well. Moreover, the competitive experiments showed that the selectivity of the PVA/IL beads towards Hg (II), Pb (II) and Cu (II) ions was Hg > Pb > Cu and even in the solutions including high concentration of Cu (II) and Pb (II) ions, PVA/IL beads kept almost the same adsorption removal efficiency of Hg (II). Thus, it can be concluded that this alternative solid-phase extractant based on [A336][MTBA] is promising to be exploited for applications in the treatment of Hg (II) polluted water. However, the repeated use of the PVA/IL beads and the column study will be the main work in future. Acknowledgements The authors gratefully acknowledge financial supports from the Austrian Science Fund and China Scholarship Council. References [1] D.W. Boening, Ecological effects, transport, and fate of mercury: a general review, Chemosphere 40 (2000) 1335–1351. [2] World Health Organization, International Standards of Drinking Water, WHO, Geneva, 1971. ˜ [3] R. Navarro, I. Saucedo, A. Núnez, M. Ávila, E. Guibal, Cadmium extraction from hydrochloric acid solutions using Amberlite XAD-7 impregnated with Cyanex 921 (tri-octyl phosphine oxide), React. Funct. Polym. 68 (2008) 557–571. [4] A.F. Ngomsik, A. Bee, J.M. Siaugue, V. Cabuil, G. Cote, Nickel adsorption by magnetic alginate microcapsules containing an extractant, Water Res. 40 (2006) 1848–1856. [5] T.P.T. Pham, C.-W. Cho, Y.-S. Yun, Environmental fate and toxicity of ionic liquids: a review, Water Res. 44 (2010) 352–372. [6] M. Regel-Rosocka, Extractive removal of zinc(II) from chloride liquors with phosphonium ionic liquids/toluene mixtures as novel extractants, Sep. Purif. Technol. 66 (2009) 19–24. [7] E. Guibal, K. Campos, P. Bunio, T. Vincent, A. Trochimczuk, Cyphos IL 101 (tetradecyl (trihexyl) phosphonium chloride) immobilized in biopolymer capsules for Hg (II) recovery from HCl solutions, Sep. Sci. Technol. 43 (2008) 2406–2433. [8] Y. Kume, K. Qiao, D. Tomida, C. Yokoyama, Selective hydrogenation of cinnamaldehyde catalyzed by palladium nanoparticles immobilized on ionic liquids modified-silica gel, Catal. Commun. 9 (2008) 369–375. [9] T. Kobayashi, M. Yoshimoto, K. Nakao, Preparation and characterization of immobilized chelate extractant in PVA gel beads for an efficient recovery of copper(II) in aqueous solution, Ind. Eng. Chem. Res. 49 (2010) 11652–11660. [10] A. Stojanovic, D. Kogelnig, L. Fischer, S. Hann, M. Galanski, M. Groessl, R. Krachler, B.K. Keppler, Phosphonium and ammonium ionicliquids with aromatic anions: synthesis, properties and platinum extraction, Aust. J. Chem. 63 (2010) 511–524. [11] S. Hashimoto, K. Furakawa, Immobilization of activated sludge by PVA–boric acid method, Biotechnol. Bioeng. 30 (1987) 52–59. [12] L. Lebrun, F. Vallée, B. Alexandre, Q.T. Nguyen, Preparation of chelating membranes to remove metal cations from aqueous solutions, Desalination 207 (2007) 9–23.
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Journal of Hazardous Materials 196 (2011) 210–219
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Occurrence of metolachlor and trifluralin losses in the Save river agricultural catchment during floods Laurie Boithias a,b,∗ , Sabine Sauvage a,b , Lobat Taghavi a,b,1 , Georges Merlina a,b , Jean-Luc Probst a,b , José Miguel Sánchez Pérez a,b,∗ a b
University of Toulouse; INP, UPS; Laboratoire d’Ecologie Fonctionnelle et Environnement (EcoLab); ENSAT, Avenue de l’Agrobiopole, 31326 Castanet Tolosan Cedex, France CNRS, EcoLab, 31326 Castanet Tolosan Cedex, France
a r t i c l e
i n f o
Article history: Received 12 May 2011 Received in revised form 18 August 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: Floods Metolachlor Trifluralin SWAT model Save river
a b s t r a c t Rising pesticide levels in streams draining intensively managed agricultural land have a detrimental effect on aquatic ecosystems and render water unfit for human consumption. The Soil and Water Assessment Tool (SWAT) was applied to simulate daily pesticide transfer at the outlet from an agriculturally intensive catchment of 1110 km2 (Save river, south-western France). SWAT reliably simulated both dissolved and sorbed metolachlor and trifluralin loads and concentrations at the catchment outlet from 1998 to 2009. On average, 17 kg of metolachlor and 1 kg of trifluralin were exported at outlet each year, with annual rainfall variations considered. Surface runoff was identified as the preferred pathway for pesticide transfer, related to the good correlation between suspended sediment exportation and pesticide, in both soluble and sorbed phases. Pesticide exportation rates at catchment outlet were less than 0.1% of the applied amount. At outlet, SWAT hindcasted that (i) 61% of metolachlor and 52% of trifluralin were exported during high flows and (ii) metolachlor and trifluralin concentrations exceeded European drinking water standards of 0.1 g L−1 for individual pesticides during 149 (3.6%) and 17 (0.4%) days of the 1998–2009 period respectively. SWAT was shown to be a promising tool for assessing large catchment river network pesticide contamination in the event of floods but further useful developments of pesticide transfers and partition coefficient processes would need to be investigated. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Rising pesticide levels in stream waters draining intensively managed agricultural land have become a widespread problem throughout Europe in recent decades. Intensive agriculture is known to have a detrimental effect on soils, surface water and groundwater quality, leading to acute problems such as soil erosion and water contamination [e.g. 1–4]. Excessive loading of pesticides, transferred into the environment through various pathways (e.g. surface runoff, subsurface and groundwater flows) either in solution or sorbed onto particles, may be harmful to terrestrial and aquatic ecosystems [5–8], rendering stream water and groundwater unfit for human consumption.
∗ Corresponding authors at: University of Toulouse, INPT, UPS, Laboratoire Ecologie Fonctionnelle et Environnement (EcoLab), ENSAT, Avenue de l’Agrobiopole, 31326 Castanet Tolosan Cedex, France. Tel.: +33 5 34 32 39 20; fax: +33 5 34 32 39 01. E-mail addresses:
[email protected] (L. Boithias),
[email protected] (J.M. Sánchez Pérez). 1 Present address: Department of Environment and Energy, Science and Research Branch, Islamic Azad University, Tehran, Iran. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.012
In Europe, pesticides are considered hazardous substances in accordance with current directives regarding water [9,10]. Drinking water quality standard should not exceed 0.1 g L−1 for an individual pesticide concentration and 0.5 g L−1 for all pesticide concentration [11]. River basins were adopted as territorial management units and the scientific community was asked to provide reliable modelling tools to evaluate pesticide source contribution to water pollution and to quantify pesticide river loads. To model pesticide fate at catchment scale, spatially variable land management and landscape characteristics, temporally variable climatology and hydrology as well as dissipation processes in the river need to be taken into account. Therefore, the combination of watershed models and river water quality models is needed to calculate pesticide fluxes to the river and transformation processes in the river channel [12]. Various models describe the pesticide fate, allowing a better understanding of the processes involved. Among the very first, the Chemical Migration and Risk Assessment (CMRA) methodology [13] included the Agricultural Runoff Management (ARM) [14] and Chemicals, Runoff and Erosion from Agricultural Management Systems (CREAMS) [15] models. It simulates the transport and fate of both dissolved and sediment-sorbed contaminants and by predicting acute and chronic impacts, provides risk assessment on aquatic
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biota. Many one-dimension, river and catchment scale models simulating pesticide fate have been then developed [e.g. 16–19]. The Soil and Water Assessment Tool (SWAT, [20]) is a semi-distributed model that provides long-term continuous predictions, including hydrology, plant growth, nutrients and suspended sediments from the field to the catchment outlet at daily time-step. Two main processes describe the pesticide transfer in both the soluble and the sorbed phases: the pesticide load generated in the hydrological unit and the fate of the load in the river. Few works have been published so far on pesticide fate modelling using the SWAT model. Molecules of a wide range of solubility were simulated (e.g. atrazine, metolachlor, trifluralin, diazinon and chlorpyrifos) in catchments ranging from 30 to 15,000 km2 [21–24]. To our knowledge, no work has been published on pesticide modelling in both the dissolved and sorbed phases at flood-event scale, i.e. during a few days of high flow. This study had four objectives: (i) to assess the performance of the SWAT model in the Save catchment (1110 km2 ) in the Gascogne region, an agriculturally intensive area of south-western France, in predicting daily pesticide river loads and concentrations at the catchment outlet; (ii) to test the sensitivity of SWAT long-term response, in terms of pesticide exportations, to interannual hydrological constraint by using an 11-year constant pesticide supply; (iii) to hindcast earlier pesticide data in order to make the model reliable for predicting river network contamination (e.g. exceeding drinking water standards) depending on the climatic context and possible flood events; (iv) to identify factors controlling exportations and preferred pathways.
2. Material and methods 2.1. Study area The River Save drains an area of 1110 km2 which is mostly farmed with intensive agriculture. It is located in the Coteaux de Gascogne region (south-western France) near Toulouse (Fig. 1). The River Save has its source in the Pyrenees piedmont. It joins the River Garonne after a 140 km course at a 0.4% average slope. Altitudes range from 663 m in the piedmont to 92 m at the Garonne confluence. The Larra gauging station elevation is 114 m (Fig. 1). The climate is oceanic. The Save river hydrological regime is mainly pluvial with a maximum discharge in May and low flows during the summer (July–September). Annual precipitation is 600–900 mm and annual evaporation is 500–600 mm (1998–2008). The hydrology is complex and subject to large climatic variations: annual average rainfall is 721 mm with a 99 mm standard deviation. The catchment lies on detrital sediments. Calcic soils represent over 90% of the whole catchment with a clay content ranging from 40% to 50%. Non-calcic silty soils represent less than 10% of the soil in this area (50–60% silt) [25]. Because of its high clay content, the catchment substratum is relatively impermeable. River discharge is consequently supplied mainly by surface and subsurface runoff and groundwater is limited to alluvial and colluvial phreatic aquifers. The mean interannual discharge (1965–2006) for the River Save is 6.1 m3 s−1 . The annual discharge of ‘dry’ years is approximately 4.1 m3 s−1 whereas for ‘wet’ years it is about 8.1 m3 s−1 . Low water discharge is about 1.3 m3 s−1 (data from the Compagnie d’Aménagement des Coteaux de Gascogne (CACG) at the Larra gauging station). During low flows, river flow is sustained upstream by the Neste canal (about 1 m3 s−1 ). 90% of the catchment area is used for agriculture. The upstream part of the catchment is a hilly agricultural area mainly covered with pasture – a 5-year rotation including one year of corn and 4 years of grazed fescue – and sometimes forest. The lower part is devoted to intensive agriculture with mainly a 2-year crop rotation
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Table 1 Spatial and temporal 3-year survey averaged management practices for sunflower grown on the Save catchment including metolachlor and trifluralin spreading. Type of operation
Date of operation
Quantity (kg ha−1 )
Pesticides spreading: trifluralin Fertilizer: 15-15-15 Sowing Pesticides spreading: metolachlor Fertilizer: 15-15-15 Harvest
05 April 05 April 10 April 15 April 16 May 01 October
0.874 193.3 – 1.12 193.3 –
of sunflower and winter wheat. Fertilizers are generally applied from late winter to spring. Average nitrogenous supply throughout the catchment is approximately 72 kg N ha−1 , i.e. 320 kg ha−1 nitrate-equivalents. About 150 mm of water is supplied by irrigation of corn. Various pesticides are applied in the catchment throughout the year depending on the crops. Our study focuses on the most applied pesticides: each year, 23 tons of metolachlor, a highly soluble chemical (Sw = 488 mg L−1 , log Kow = 2.9), and 18 tons of trifluralin, a poorly soluble chemical (Sw = 0.221 mg L−1 , log Kow = 4.83), are applied on the catchment. Both pesticides are herbicides. They are applied each year on sunflower in early April (Table 1). On average, sunflower fields cover 18.4% of the catchment (20,600 ha). 2.2. Observed discharge data The River Save has been monitored for discharge since 1965. At the Larra hydrometric station, hourly discharges (Q) were obtained from CACG. The hourly discharge was plotted by the rating curve H(Q) in which the water level (H) was measured continuously and then averaged for each day. 2.3. Observed water quality data 2.3.1. Nitrate and suspended sediment monitoring Nitrate loads and suspended sediment concentrations were monitored continuously from January 2007 to March 2009 at the Larra gauging station, both manually and automatically, as described previously in Oeurng et al. [26–28]: an automatic water sampler, connected to a probe, was programmed to activate pumping water on the basis of water level variations ranging from 10 cm (during low flows) to 30 cm (during high flows) for the rising and falling stages. Grab sampling was also undertaken near the probe position at weekly intervals. 2.3.2. Pesticide monitoring Pesticides were monitored from March 2008 to March 2009 at the Larra station with weekly grab sampling during low flow and daily grab sampling during flood events. Laboratory analyses were performed as described in Taghavi et al. [29,30]. Additional data from Agence de l’Eau Adour-Garonne (AEAG) were used for longterm total pesticide concentration comparison (Source: Système d’Information sur l’Eau du Bassin Adour-Garonne, data exported in 2009). 2.3.3. Load calculation Based on the high frequency of data collection, a linear interpolation method was applied between two neighbouring sampling points to construct the continuous nitrate, suspended sediment and pesticide concentration series and thus calculate continuous daily loads through the product of concentration and water volume. Yearly loads were calculated by totalling daily loads.
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Fig. 1. Localisation of Save catchment, Larra gauging station and meteorological stations.
2.4. Modelling approach 2.4.1. The SWAT model SWAT is a physically based agro-hydrological model [20]. It operates at a daily time-step and was designed to predict the impact of management practices on water quality in ungauged catchments. It allows the addition of flows by including measured data from point sources. SWAT discretises catchments into sub-basins. Sub-basins are then subdivided into Hydrological Response Units (HRUs). HRUs are areas of homogenous land use, soil type and slope. HRUs outputs are inputs for the connected stream network. One sub-basin is drained by one reach. Authors refer to Neitsch et al. [31] for detailed description of the model. 2.4.2. The pesticide component in SWAT Pesticide processes in SWAT are divided into three components: (i) pesticide processes in land areas, (ii) transport of pesticides from land areas to the stream network, and (iii) instream pesticide processes. SWAT uses algorithms from GLEAMS (Groundwater Loading Effects on Agricultural Management Systems) [16] to model pesticide movement and fate in land areas. The partitioning of a pesticide between the dissolved and sorbed phases is defined by a soil adsorption coefficient. Algorithms governing movement of soluble and sorbed forms of pesticide from land areas to the stream network were taken from the EPIC (Erosion-Productivity Impact Calculator) model [32]. The SWAT model incorporates a simple mass-balance method [33] to model the transformation and transport of pesticides in streams. Only one pesticide can be routed through the stream network in a given simulation. The fraction of pesticide in each phase is a function of the pesticide’s partition coefficient and the reach segment’s suspended solid concentration. Degradation is based on half-life. Authors refer to Neitsch et al. [31] for further details. 2.5. SWAT data inputs Spatialised data used in this study were: - Digital Elevation Model with a resolution of 25 m × 25 m from Institut Géographique National (IGN) France (BD TOPO R).
- Soil data on the scale of 1:80,000 from CACG and digitised by Cemagref de Bordeaux [34] and soil properties for the SWAT soil database [35]. - Land use data [34] from Landsat 2005 with associated management practices: spatial and temporal average of planting/seedling dates, amounts, type and date of fertilisation, pesticide application and irrigation, grazing, tillage and harvest operations dates from a 3 year survey (2003–2005) with catchment farmers, applied for each year of simulation. - Meteorological data from 5 stations (Fig. 1) with daily precipitation from Météo-France. Missing data were generated by linear regression equation from data from the nearest stations with complete measurements. Two stations in the upstream section had a complete set of measurements of daily minimum and maximum air temperature, wind speed, solar radiation and relative humidity that were used to simulate the reference evapotranspiration in the model by Penman–Monteith [36,37] method. - Point source data: the Save river network is connected upstream to the Neste canal. Daily discharge was given by CACG. Since it is water from a mountainous agricultural extensive area, concentrations of nitrate and pesticides were set constant and equal to 2 and 0 mg L−1 respectively. Since water is derived by a dam where sediments are trapped, suspended sediment concentration was set constant and equal to 10 mg L−1 . In this study, version 2009.93.3 of ArcSWAT was used. The catchment was discretised into 73 sub-basins with a minimal area of 500 ha. 1642 HRUs were generated integrating 8 land uses classes, 23 soil classes and 5 slope classes (%: 0–1, 1–3, 3–5, 5–8 and 8 and over). Whole simulation was carried out daily from January 1998 to March 2009 (excluding 4 years’ warm-up from 1994 to 1997). The sensitivity of 15 parameters governing discharge, nitrate and suspended sediment dynamic was tested using the ArcSWAT2009 sensitivity analysis tool [38]. Calibration of discharge, nitrate, suspended sediment and pesticide at daily time-step was performed manually. Pesticide input values are given in Table 2. 2.6. Model evaluation The performance of the model was evaluated using the Nash–Sutcliffe efficiency (ENS ) index [39] and the coefficient of
L. Boithias et al. / Journal of Hazardous Materials 196 (2011) 210–219 Table 2 Manually calibrated values of pesticides parameters: half-lifes, Koc and CHPST KOC and degradation rates in the channel water and in the sediment bed (respectively CHPST REA and SEDPST REA). Parameters
Input file
Metolachlor
Aclonifen
Soil half-life Met/Tri (days) Koc Met/Tri (mg kg−1 /mg L−1 ) CHPST KOC Met/Tri (m3 g−1 ) CHPST/SEDPST REA (days−1 )
pest.dat pest.dat .swq .swq
90 667 0.1 0.025
60 13,196 2.6 0.025
determination (R2 ). ENS ranges from negative infinity to 1 whereas R2 ranges between 0 and 1. We refer to Krause et al. [40] for further discussion on these evaluation criteria. Daily ENS and R2 were applied on daily discharge for low flows (below the 6.1 m3 s−1 mean annual discharge), high flow (over 6.1 m3 s−1 ) and total (1998–2006 for calibration and 2007–2009 for validation). They were also calculated for nitrate loads and suspended sediment concentrations for the 2007–2009 period. Daily and monthly R2 were calculated for dissolved, sorbed and total pesticides concentrations for calibration period (2008–2009). Due to data limitation, ENS was not calculated on pesticide concentration. In this study we deemed ENS satisfactory when higher than 0.36 [41] and R2 satisfactory when higher than 0.5 [42]. 2.7. Water quality simulation 2.7.1. Discharge, nitrate and suspended sediment simulation Daily SWAT interpolated rainfall and simulated water yield (i.e. the amount of water flowing down the outlet) were totalled for each year. Daily simulated nitrate and suspended sediment loads at the Larra outlet were totalled for each year and for low flow and high flow (using the 6.1 m3 s−1 threshold). Incoming suspended sediment in a given reach is the total of the amount of suspended sediments coming from the HRUs related to the reach, added to the amount entering from the upstream reach. 2.7.2. Pesticide simulation Daily pesticide loads were totalled for each year, for April flood (11/04/08–30/04/08) and June flood (14/05/08–18/06/08), and for low flow and high flow (using the 6.1 m3 s−1 threshold). Total pesticide concentration is the concentration of pesticide as measured in unfiltered water. Simulated total concentration is calculated as the total of dissolved and sorbed pesticide concentration. The particulate fraction is the fraction of pesticide in the sorbed phase and is calculated as the sorbed pesticide load divided by the total pesticide load. The exportation rate is calculated as the ratio from pesticide load exported at outlet and the amount of pesticide applied on the catchment. As pesticide toxicity is more relevant as concentration than as load, long-term simulation concentrations (1998–2009) were compared to European drinking water quality standards of 0.1 g L−1 for a single pesticide and of 0.5 g L−1 for both pesticides being modelled. Correlation analyses were performed on 1998–2008 interannual-averaged metolachlor and trifluralin (both dissolved and sorbed) loads from reach and HRUs to relate them to reach and HRUs variables such as slope, soil classes, rainfall, water, suspended sediment and nitrate yields. 3. Results 3.1. Discharge, nitrate and suspended sediment simulation According to sensitivity analysis, parameters governing discharge and nitrate were mostly parameters governing runoff and groundwater transfer (CN2, RCHRG DP, GWQMN). Parameters governing suspended sediment were
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Table 3 Goodness-of-fit indices for daily discharge, nitrate load and suspended sediment simulation (p < 0.05).
Discharge – calibration Discharge – validation Low flow High flow Nitrate load Suspended sediments concentration
Periods
R2
ENS
1998–2006 2007–2009 1998–2009 1998–2009 2007–2009 2007–2009
0.52 0.58 0.02 0.51 0.46 0.36
0.50 0.56 0.02 0.54 0.37 0.27
mostly parameters governing runoff (CN2) and in-stream processes (CH N2, SPCON, SPEXP). Fig. 2 focuses on discharge simulated daily from January 2007 to March 2009. The goodness-of-fit indices for daily discharge were satisfactory during both calibration and validation period (Table 3). They were also satisfactory for high flow but unsatisfactory for low flow (Table 3). Nitrate daily load predictions (Fig. 3(a)) were correlated to observations for the 2007–2009 period (Table 3). Observed and simulated cumulated nitrate loads in 2007 were 2514 and 2388 tons respectively, they were 3047 and 3018 tons respectively in 2008 (Fig. 3(b) and (c)). Considering daily and annual loads prediction on 2007–2009, the model was considered to hindcast past daily and annual nitrate loads back to 1998 with little error. Annual nitrate loads were correlated to annual water yield (R2 = 0.84, p < 0.05). Daily simulated suspended sediment concentrations fitted observations (Fig. 4(a)) although ENS and R2 were unsatisfactory (Table 3). Observed and simulated annual suspended sediment loads were 9000 and 15,000 tons respectively in 2007 and 58,000 and 64,000 tons respectively in 2008 (Fig. 4(b) and (c)). Considering concentration and load prediction on 2007–2009, the model was considered to reconstruct past annual loads back to 1998 with little error. Annual suspended sediment loads were correlated to annual water yield (R2 = 0.59, p < 0.05). On average across the 73 reaches, the annual ratio of deposited/incoming suspended sediment in the reach is of 54%. 3.2. Pesticide simulation Simulation results were shown to be poorly sensitive to application date change for both molecules (results not shown) although pesticide losses are known to be determined mainly by the period of time between application and the first rainfall event and by the application dose [43,44]. SWAT pesticide component parameters, including the soil adsorption coefficient (Koc ), were also poorly sensitive (results not shown) except the channel partition coefficient between water and suspended sediment (CHPST KOC). 3.2.1. SWAT performances The range of daily simulated concentrations of metolachlor and trifluralin followed the range of respective measurements during both the calibration 2008–2009 period (Fig. 5) and the long-term validation 1998–2009 period (Fig. 6). During the calibration period, simulated dissolved and sorbed molecule concentrations during low flow matched respective observations. Flood concentration peaks of dissolved metolachlor were predicted although overestimated during April and June flood events. Sorbed pesticide concentration peaks during the same period were underestimated. Trifluralin concentration peaks in May were skipped. The model did not simulate the trifluralin concentration peak in early 2009. Simulated partition between soluble and sorbed phases of both molecules roughly followed the observed partition. In terms of loads, average simulated annual loads at outlet were in the range of observed annual loads (Table 4). The simulated
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Fig. 2. Observed and simulated daily discharge (m3 s−1 ) at the Larra gauging station (January 2007–March 2009).
Fig. 3. Observed and simulated daily (a) nitrate load (tons) at the Larra gauging station (January 2007–March 2009), (b) 2007 accumulation (tons) and (c) 2008 accumulation (tons).
metolachlor and trifluralin particulate fractions at catchment outlet fitted the observed particulate fractions and were consistent with the fractions simulated in surface runoff out of the HRUs (Table 4). At flood scale, simulation followed the measured range of
values (Table 5). However, metolachlor loads were underestimated whereas trifluralin loads were overestimated during the April flood and underestimated during the June flood. R2 for monthly concentrations of metolachlor and trifluralin were over 0.38 except
Fig. 4. Observed and simulated daily (a) suspended sediment concentration (mg L−1 ) at the Larra gauging station (January 2007–March 2009), (b) 2007 accumulation (tons) and (c) 2008 accumulation (tons).
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215
Fig. 5. Observed and simulated daily pesticide concentrations (g L−1 ) at the Larra gauging station (2008–2009): (a) metolachlor and (b) trifluralin.
Table 4 Observed (2008–2009) and simulated (1998–2008) average annual loads of metolachlor and trifluralin in each catchment compartment (mg ha−1 yr−1 ) and particulate fraction out of HRUs and at catchment outlet. Metolachlor (mg ha−1 yr−1 ) Observed HRUs Runoff
Lateral flow Outlet Flow water
Trifluralin (mg ha−1 yr−1 ) Simulated
Observed
Dissolved: 475 Sorbed: 1237 Partition: 0.72 Dissolved: 10
Dissolved: 1653 Sorbed: 317 Partition: 0.16 Dissolved: 200 Dissolved: 204 Sorbed: 27 Partition: 0.12
Simulated
Dissolved: 135 Sorbed: 14 Partition: 0.09
Dissolved: 60 Sorbed: 195 Partition: 0.77
Dissolved: 2 Sorbed: 5 Partition: 0.72
Table 5 Total and dissolved metolachlor and trifluralin loads (g) during April and June 2008 floods at Larra outlet. Metolachlor
Trifluralin
2008 floods
April (11/04–30/04)
June (14/05–18/06)
April (11/04–30/04)
June (14/05–18/06)
Measured total load (g) Simulated total load (g) Measured dissolved load (g) Simulated dissolved load (g)
5015 3588 4591 3262
16,692 11,283 16,178 10,257
66 92 13 26
3285 422 186 117
Table 6 Daily and monthly R2 for metolachlor and trifluralin concentration (dissolved, sorbed and total) at Larra outlet (2008–2009). Metolachlor
Trifluralin
Daily
Monthly
Daily
Monthly
Dissolved Sorbed
0.25 0.01
0.43 0.38
0.21 0.01
0.60 0.15
Total
0.26
0.45
0.02
0.16
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Fig. 6. Observed and simulated daily pesticide total concentrations (g L−1 ) at the Larra gauging station (1998–2009): (a) metolachlor and (b) trifluralin.
for sorbed and total trifluralin. R2 for daily concentrations did not exceed 0.26 (Table 6).
3.2.2. Long-term pesticide exportation balances at outlet During the 1998–2008 period, high flows represented 17% of the time considering the 6.1 m3 s−1 threshold. 50% of nitrate load and 57% of suspended sediment load were exported during high flow. Annual pesticide loads are shown in Fig. 7. The total metolachlor load varied between 0.1 kg yr−1 (2003) and 80 kg yr−1 (2000), whereas trifluralin varied between 0.01 kg yr−1 (2003) and 2.3 kg yr−1 (2000). Average total metolachlor and trifluralin annual loads were 16.7 kg (SD = 23 kg) and 0.8 kg (SD = 1 kg) respectively. The total metolachlor and trifluralin exported were around 0.072% and 0.005% of the applied amount respectively (exportation rate was 1% out of the HRUs for both molecules). At outlet 61% of the total metolachlor and 52% of the total trifluralin were exported during high flows (Table 7). Out of the 4108 simulated days (1998–2009), metolachlor at Larra outlet exceeded the European standard threshold of 0.1 g L−1 for 149 days and trifluralin exceeded this same threshold for 17 days (3.6 and 0.4% of the time period respectively). Maximum metolachlor concentration was 5.4 mg L−1 , predicted in July 2001 whereas maximum trifluralin concentration was 0.2 mg L−1 , predicted in April 1998 (Fig. 6). Considering the sum of total metolachlor and total trifluralin concentrations, the threshold of 0.5 g L−1 was exceeded during 24 days at catchment outlet.
3.2.3. Pesticide transfer controlling factors At catchment scale, the simulated preferred pathway of pesticide transfer was surface runoff (Table 4) with associated nitrate and suspended sediment exportations. At sub-basin scale, trifluralin and metolachlor in the sorbed phase were correlated to suspended sediment loads (R2 = 0.69 and 0.64, respectively). Trifluralin and metolachlor in the dissolved phase were poorly correlated to nitrate loads (R2 = 0.21 and 0.16, respectively). Eventually, trifluralin and metolachlor in the dissolved phase were better correlated to suspended sediment loads (R2 = 0.33 and 0.67 respectively). At HRU scale, the correlation analysis did not show any correlation between suspended sediment loads, metolachlor and trifluralin loads in the dissolved phase and catchment variables. Metolachlor and trifluralin loads in the sorbed phase correlated weakly to suspended sediment yields (R2 = 0.24 and 0.43 respectively). 4. Discussion 4.1. Discharge, nitrate and suspended sediment simulation Overall ENS and R2 for daily discharge were over the satisfactory threshold. However, gaps between observed and simulated values are explained by errors in observed and simulated values. Errors in observed values can stem from the precision of the sensor and from the use of a rating curve. Errors in simulated values can be attributed to (i) actual local rainfall storms that were not well represented by the SWAT rainfall data interpolation and (ii) the flow
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217
Fig. 7. Simulated annual dissolved and sorbed pesticide loads (kg yr−1 ) at Larra outlet (1998–2008): (a) metolachlor and (b) trifluralin.
uncertainty in the dense network of canals diverted from the river network to bring part of river flow to many watermills. ENS for low flow was below the satisfactory threshold. Low ENS during low flow has to be related to its generally poor performance in periods of low flow: with only minor simulation errors the denominator of the equation tends towards zero and ENS approaches negative infinity. Low ENS is however of minor concern since pesticides were shown to be mostly exported during floods. Save river discharge simulation quality is however comparable to Oeurng et al. [45]. Daily nitrate prediction is related to daily discharge prediction, as it is a very soluble nutrient. Bias is therefore linked to the inaccuracy of discharge predictions mentioned above. Uncertainty in point-source nutrient input also explains bias. In addition, averaged land use and associated management practice inputs may not reflect well enough the actual and local land use and management practices. They also can evolve over the modelled period depending on agricultural policy trends. ENS and R2 for daily suspended sediments concentration were below the satisfactory threshold. Calibration of sediment is difficult in the Save catchment considering the dense network of canals diverted from the river network. The succession of dams and gates traps sediments till their random emptying back to river network. Uncertainty in point-source suspended sediments input may also
explain bias. Suspended sediment results were however consistent with Oeurng et al. [45]. 4.2. Pesticide simulation 4.2.1. SWAT performances R2 of daily pesticide concentration was below the satisfactory threshold. As highlighted by Luo et al. [24] possible time shifts in the precipitation, agricultural activity, and measurements for flow and water quality data render daily pesticide statistics poor. Errors in pesticide concentration and load predictions may be related to (i) an inadequate calibration of the parameters governing flow and soluble (e.g. nitrate) and particulate (e.g. suspended sediment) phases transport and to (ii) an inadequate calibration of pesticide component parameters. Sorbed pesticide underestimation has to be related to the high deposition rate of suspended sediment in the channel. Also, as Koc was shown to be poorly sensitive and as CHPST KOC is set as a constant value, both partition coefficients modelled in HRUs and in reaches vary depending on suspended sediment concentration. They may not reflect actual variations: metolachlor and trifluralin measurements at outlet show various inversions (i.e. [soluble] < [sorbed] for metolachlor and [soluble] > [sorbed] for tri-
Table 7 Average metolachlor and trifluralin daily loads (g d−1 ) during high flow and low flow in dissolved and sorbed phases and percentage of pesticide load exported during high flow at Larra outlet (1998–2008). Dissolved
Sorbed −1
Low flow (g d Metolachlor Trifluralin
19 0.4
)
−1
High flow (g d 137,118 1805
)
Flood losses (%)
Low flow (g d−1 )
61 52
2 1
High flow (g d−1 ) 13,712 4694
Flood losses (%) 61 52
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fluralin) of the partition coefficient Kd, as defined by Taghavi et al. [30] at catchment outlet, that were not modelled. Regarding agricultural activity, i.e. land use and associated management practices, the model did not simulate any trifluralin concentration rise in early 2009. The land use input map was based on a 2005 land cover satellite image that may not reflect long-term actual operations: e.g. canola, representing 1% of the catchment land use in 2005, has been grown increasingly over the past five years in the northern part of the catchment. Up to 2009, canola was managed with an average spread of 1 kg ha−1 of trifluralin in August. This was not taken into account in the model because SWAT land use approximation was set to skip land use representing less than 10% of the sub-basin area. Reliable pesticide supply input data are therefore a necessary condition to achieve satisfactory simulation outputs. A last source of error is the modelled transfer pathways: SWAT simulates pesticide transfers through surface runoff and subsurface lateral flow but not through groundwater flow, drainflow nor atmospheric deposition [31]. This may lead to additional errors in soluble pesticide simulation although leaching of pesticides into deep groundwater and a possible input of pesticides into surface waters by outflowing groundwater is known to be negligible [46,47]. In addition, no point-source, such as the cleaning of the equipment, was modelled in this study [12,48]. Finally, daily pesticide total concentrations R2 were in the lower range of the values mentioned by Neitsch et al. [21]. They reported R2 ranging from 0.41 to 0.28 for daily metolachlor total concentration and R2 ranging from 0.51 to 0.02 for daily trifluralin total concentration. 4.2.2. Long-term pesticide exportation balance at outlet Simulation showed that pesticides were exported mainly during flood events. Role of floods in pesticide exportation was previously shown at the Save catchment outlet [29]. As interannual average, exportation rates of both pesticides have to be related to their application time (April) and the month of maximum discharge (May). However, simulated values of exportation during a similar high flow period were less than measured values in a small catchment or at field scale [49,50]. The size of the drained area, but also the soil and the land use may modulate the exportation. Pesticide exportations from land to outlet were less than 1% of applied amount. Such a value was reported by various studies on metolachlor in France, Switzerland and Québec [51–53] and on other pesticides [50,53,54]. About 93% and 99% of metolachlor and trifluralin respectively entering the stream network does not reach the outlet, suggesting high deposition (discussed above) and degradation in stream water. The latter would be consistent with the carbon consumption by river biota shown by Sánchez-Pérez et al. [55] on a similar catchment of Gascogne. Trifluralin concentration exceeded the 0.1 g L−1 maximum permissible level less than metolachlor concentration. In the model, trifluralin was less applied than metolachlor among the catchment, its soil half-life was 60 days (instead of 90 days for metolachlor) and its transport was more likely depending on land and river bed erosion. It is worth noting that the European standard of 0.5 mg L−1 was exceeded by a pool of only two molecules during 0.6% of the simulation time. 4.2.3. Pesticide transfer controlling factors Runoff was shown to be the preferred simulated pathway for pesticide exportation. Luo et al. [24,56] already reported the control of runoff on pesticide transfer simulation. In the environment, pesticides may be sorbed onto mineral suspended matter, Particulate Organic Carbon (POC) and complexed by Dissolved Organic Carbon (DOC) [29]. Better correlation was found between dissolved phases and suspended sediment than between dissolved phases
and nitrate. Although pesticide transfer through groundwater is not yet modelled in SWAT, contrarily to nitrate, the relationships highlighted a control of both surface and sub-surface runoffs during floods on dissolved phase exportation. This is consistent with the pesticides’ actual ability to sorb onto DOC, i.e. smaller than the 0.45 m mesh filter, and its transfer through sub-surface flow [29]. 5. Conclusions The SWAT model was applied to simulate pesticide transfer at the outlet of a large intensive agricultural catchment. Simulation results were deemed to be satisfactory, taking into account that the modelled transfer processes were simplified and do not represent all actual transfers. Further improvements of the SWAT model may be investigated. The transfer of pesticide in the dissolved phase from land to river through groundwater could be tested to assess possible water-table effect. Also, investigating the variations of the partition coefficient between observed dissolved and sorbed phases during floods in various points in the river (areas of rapid, deep, etc.) would help to assess the accuracy of the SWAT modelled partition at flood-event scale. However, extrapolation to other chemicals is conceivable and SWAT was shown to be promising for providing a robust decision tool for water quality managers. Floods are quick events. Such a tool would help (i) to target ‘what and when’ to monitor, (ii) to highlight pesticide concentration peaks without cost-intensive field measurements and predict future peaks and (iii) to evaluate exported loads as contamination indicator. Suggestion of localised mitigation practices to reach water policy objectives such as the European Water Framework Directive is made possible. Acknowledgements This work was performed as part of the EU Interreg SUDOE IVB program (SOE1/P2/F146 AguaFlash project, http://www.aguaflashsudoe.eu) and funded by ERDF and the Midi-Pyrénées Region. We sincerely thank the CACG for discharge data, Météo-France for meteorological data, AEAG for long-term pesticide concentration measurements, Arnaud Mansat and Erwan Motte for their data-processing expertise, and Hugues Alexandre for IT support. References [1] C. Soulsby, A.F. Youngson, H.J. Moir, I.A. Malcolm, Fine sediment influence on salmonid spawning habitat in a lowland agricultural stream: a preliminary assessment, Sci. Total Environ. 265 (2001) 295–307. [2] G. Zalidis, S. Stamatiadis, V. Takavakoglou, K. Eskridge, N. Misopolinos, Impacts of agricultural practices on soil and water quality in the Mediterranean region and proposed assessment methodology, Agric. Ecosyst. Environ. 88 (2002) 137–146. [3] M. Probst, N. Berenzen, A. Lentzen-Godding, R. Schulz, M. Liess, Linking land use variables and invertebrate taxon richness in small and medium-sized agricultural streams on a landscape level, Ecotoxicol. Environ. Safe. 60 (2005) 140–146. [4] M. Zeiger, N. Fohrer, Impact of organic farming systems on runoff formation processes—a long-term sequential rainfall experiment, Soil Tillage Res. 102 (2009) 45–54. [5] J.G.M. Cuppen, P.J. Van den Brink, E. Camps, K.F. Uil, T. Brock, Impact of the fungicide carbendazim in freshwater microcosms. I. Water quality, breakdown of particulate organic matter and responses of macroinvertebrates, Aquat. Toxicol. 48 (2000) 233–250. [6] P.J. Van den Brink, J. Hattink, F. Bransen, E. Van Donk, T. Brock, Impact of the fungicide carbendazim in freshwater microcosms. II. Zooplankton, primary producers and final conclusions, Aquat. Toxicol. 48 (2000) 251–264. [7] R.M. Niemi, I. Heiskanen, J.H. Ahtiainen, A. Rahkonen, K. Mäntykoski, L. Welling, et al., Microbial toxicity and impacts on soil enzyme activities of pesticides used in potato cultivation, Appl. Soil Ecol. 41 (2009) 293–304. [8] A.G. Becker, B.S. Moraes, C.C. Menezes, V.L. Loro, D.R. Santos, J.M. Reichert, et al., Pesticide contamination of water alters the metabolism of juvenile silvercatfish, Rhamdia quelen, Ecotoxicol. Environ. Safe. 72 (2009) 1734–1739. [9] EC, Directive 2006/11/EC of the European Parliament and of the Council of 15 February 2006 on pollution caused by certain dangerous substances discharged into the aquatic environment of the Community, 2006.
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[34] F. Macary, E. Lavie, G. Lucas, O. Riglos, Méthode de changement d’échelle pourl’estimation du potentiel de contamination des eaux de surface par l’azote, Ingénieries - EAT 46 (2006) 35–49. [35] J.M. Lescot, P. Bordenave, A decision support to choose between changes of agricultural practices: a spatially distributed cost-effectiveness assessment framework, in: Integrated Assessment of Agriculture and Sustainable Development, Setting the Agenda for Science and Policy, Egmond aan Zee, NLD, 2009, pp. 452–453. [36] H.L. Penman, Natural evaporation from open water, bare soil, and grass, Proc. Roy. Soc. Lond. A 193 (1948) 120–146. [37] J.L. Monteith, Evaporation and environment, in: The State and Movement of Water in Living Organisms, Cambridge University Press, Swansea, 1965, pp. 205–234. [38] A. Van Griensven, T. Meixner, S. Grunwald, T. Bishop, M. Diluzio, R. Srinivasan, A global sensitivity analysis tool for the parameters of multi-variable catchment models, J. Hydrol. 324 (2006) 10–23. [39] J.E. Nash, V. Sutcliffe, River flow forecasting through conceptual models. Part I. A discussion of principles, J. Hydrol. 10 (1970) 282–290. [40] P. Krause, D.P. Boyle, F. Bäse, Comparison of different efficiency criteria for hydrological model assessment, Adv. Geosci. 5 (2005) 89–97. [41] M.W. Van Liew, J. Garbrecht, Hydrologic simulation of the little Washita river experimental watershed using SWAT, J. Am. Water Resour. Assoc. 39 (2003) 413–426. [42] C.H. Green, M.D. Tomer, M. Di Luzio, J.G. Arnold, Hydrologic evaluation of the soil and water assessment tool for a large tile-drained watershed in Iowa, Trans. ASABE 49 (2006) 413–422. [43] J.S. Martins, L.B. Owens, Atrazine, deethylatrazine, and deisopropylatrazine in surface runoff form conservation tilled watersheds, Environ. Sci. Technol. 37 (2003) 944–950. [44] K. Müller, M. Deurer, H. Hartmann, M. Bach, M. Spiteller, H.G. Frede, Hydrological characterisation of pesticide loads using hydrograph separation at different scales in a German catchment, J. Hydrol. 272 (2003) 1–17. [45] C. Oeurng, S. Sauvage, J.M. Sanchez-Perez, Assessment of hydrology, sediment and particulate organic carbon yield in a large agricultural catchment using the SWAT model, J. Hydrol. 401 (2011) 145–153. [46] B. Röpke, M. Bach, H.G. Frede, DRIPS—a DSS for estimating the input quantity of pesticides for German river basins, Environ. Model. Soft. 19 (2004) 1021–1028. [47] R.P. Richards, D.B. Baker, Pesticide concentration patterns in agricultural drainage networks in the Lake Erie Basin, Environ. Toxicol. Chem. 12 (1993) 13–26. [48] S. Reichenberger, M. Bach, A. Skitschak, H.G. Frede, Mitigation strategies to reduce pesticide inputs into ground- and surface water and their effectiveness: a review, Sci. Total Environ. 384 (2007) 1–35. [49] X. Louchart, M. Voltz, P. Andrieux, R. Moussa, Herbicide transport to surface waters at field and watershed scales in a Mediterranean vineyard area, J. Environ. Qual. 30 (2001) 982–991. [50] M. Rabiet, C. Margoum, V. Gouy, N. Carluer, M. Coquery, Assessing pesticide concentrations and fluxes in the stream of a small vineyard catchment—effect of sampling frequency, Environ. Pollut. 158 (2010) 737–748. [51] S.M. Novak, J.M. Portal, M. Schiavon, Effects of soil type upon metolachlor losses in subsurface drainage, Chemosphere 42 (2001) 235–244. [52] C. Leu, H. Singer, C. Stamm, S.R. Muller, R.P. Schwarzenbach, Variability of herbicide losses from 13 fields to surface water within a small catchment after a controlled herbicide application, Environ. Sci. Technol. 38 (2004) 3835–3841. [53] L. Poissant, C. Beauvais, P. Lafrance, C. Deblois, Pesticides in fluvial wetlands catchments under intensive agricultural activities, Sci. Total Environ. 404 (2008) 182–195. [54] N. Berenzen, A. Lentzen-Godding, M. Probst, H. Schulz, R. Schulz, M. Liess, A comparison of predicted and measured levels of runoff-related pesticide concentrations in small lowland streams on a landscape level, Chemosphere 58 (2005) 683–691. [55] J.M. Sánchez-Pérez, M. Gerino, S. Sauvage, P. Dumas, E. Maneux, F. Julien, et al., Effects of wastewater treatment plant pollution on in-stream ecosystems functions in an agricultural watershed, Int. J. Lim. 45 (2009) 79–92. [56] Y. Luo, M. Zhang, Management-oriented sensitivity analysis for pesticide transport in watershed-scale water quality modeling using SWAT, Environ. Pollut. 157 (2009) 3370–3378.
Journal of Hazardous Materials 196 (2011) 220–227
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Magnetic loading of TiO2 /SiO2 /Fe3 O4 nanoparticles on electrode surface for photoelectrocatalytic degradation of diclofenac Xinyue Hu, Juan Yang, Jingdong Zhang ∗ College of Chemistry and Chemical Engineering, Huazhong University of Science and Technology, Luoyu Road 1037, Wuhan 430074, PR China
a r t i c l e
i n f o
Article history: Received 26 May 2011 Received in revised form 29 August 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: TiO2 /SiO2 /Fe3 O4 Magnetic loading Photoelectrocatalysis Diclofenac
a b s t r a c t A novel magnetic nanomaterials-loaded electrode developed for photoelectrocatalytic (PEC) treatment of pollutants was described. Prior to electrode fabrication, magnetic TiO2 /SiO2 /Fe3 O4 (TSF) nanoparticles were synthesized and characterized by X-ray diffraction (XRD), transmission electron microscopy (TEM) and FT-IR measurements. The nanoparticles were dispersed in ethanol and then immobilized on a graphite electrode surface with aid of magnet to obtain a TSF-loaded electrode with high photoelectrochemical activity. The performance of the TSF-loaded electrode was tested by comparing the PEC degradation of methylene blue in the presence and absence of magnet. The magnetically attached TSF electrode showed higher PEC degradation efficiency with desirable stability. Such a TSF-loaded electrode was applied to PEC degradation of diclofenac. After 45 min PEC treatment, 95.3% of diclofenac was degraded on the magnetically attached TSF electrode. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Over past decades, TiO2 has been extensively utilized as an efficient photocatalyst for the removal of pollutants from wastewater [1,2]. However, practical applications of slurry photocatalyst to wastewater treatment are usually limited because a costly and difficult post-treatment process is required to separate slurry TiO2 from treated water. To overcome this problem, magnetically separable composite photocatalysts such as TiO2 /Fe3 O4 and TiO2 /SiO2 /Fe3 O4 (TSF) have been prepared and applied to the degradation of pollutants [3–7]. Alternatively, preparation of TiO2 film by immobilizing photocatalyst onto various solid supports has also been considered as an effective method to overcome the separation problem of slurry TiO2 [8–13]. Unfortunately, the immobilized TiO2 film reduces the reaction surface area of catalyst and thus decreasing the degradation efficiency, as compared with the slurry catalyst. To improve the photocatalytic efficiency of TiO2 film, photoelectrocatalytic (PEC) technique by applying an external bias potential on TiO2 film-coated electrode has been developed [14]. For PEC electrode, the catalyst must be firmly adhered to the solid support to keep high stability of electrode and allow facile electron transfer between catalyst and electrode. Traditional PEC electrodes prepared from TiO2 sol–gel are generally treated
∗ Corresponding author. Tel.: +86 27 87792154; fax: +86 27 87543632. E-mail address:
[email protected] (J. Zhang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.009
with high temperature to reinforce the adherence of catalyst to the electrode surface [15–18]. Recently, magnetic loading has been developed for constructing electrochemical sensors [19,20]. To prepare magnetic electrode, magnetic Fe3 O4 –SiO2 core–shell bio-nanoparticles have been prepared and immobilized on the electrode surface with the help of magnetic force [21]. A novel electrochemical-sensing platform based on the magnetic loading of CNT/Fe3 O4 composite on electrodes for the detection of hydrogen peroxide has been demonstrated [22]. As a group of emerging contaminants, pharmaceuticals and personal care products (PPCPs) have received considerable attention in recent years. PPCPs are regarded as potentially hazardous compounds since many of them are ubiquitous, persistent and biologically active compounds with endocrine disruption functions [23]. Diclofenac is a widely used non-steroidal anti-inflammatory drug. It has been one of the most frequently detected PPCPs in water environment at concentrations up to 1.2 g/L [24]. Advanced oxidation processes such as photocatalysis [24], UV [25], UV/H2 O2 [26], photo-Fenton [27,28], ozonation [26,29,30], and sonolysis [31,32] have been applied to the removal of diclofenac from polluted water. In this work we reported a novel magnetic composite nanomaterials-loaded electrode with photoelectrochemical activity, which was fabricated by immobilizing TSF nanoparticles on the electrode surface with aid of magnet. The magnetic force between external magnet and TSF nanoparticles allowed a desirable adherence of catalyst to the electrode. Such a TSF-loaded electrode was successfully applied to the PEC degradation of diclofenac.
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2. Experimental 2.1. Chemicals FeCl3 ·6H2 O, FeSO4 ·7H2 O, tetraethoxysilane (TEOS), tetrabutyl orthotitante (TBOT), sodium dodecyl sulfate (SDS), methylene blue (MB) and Na2 SO4 were obtained from Sinopharm chemical Reagent Co. (Shanghai, China). Diclofenac sodium was obtained from Sigma Chemical Reagent Co., USA. Multi-walled CNTs (diameter 20–30 nm, length 1–10 m) provided by Nanjing University (Nanjing, China) were pretreated in a mixture of concentrated nitric acid/sulfuric acid (1:3, v/v) with ultrasonication for 8 h, swilled to neutral and then dried at 60 ◦ C. Other chemicals such as HCl, H2 SO4 , NH3 ·H2 O (25 wt%), NaAc, HAc, Na2 HPO4 , NaH2 PO4 , dimethylformamide (DMF), cetyltrimethylammonium bromide (CTAB) and ethanol were of analytical grade. Doubly distilled water was used in all experiments. 2.2. Preparation of Fe3 O4 , SiO2 /Fe3 O4 and TSF nanoparticles The magnetite nanoparticles were prepared by the coprecipitation method. Typically, 60 mL mixture of iron salts with a molar ratio (FeCl3 : FeSO4 ) of 1:2 was prepared in the presence of 0.5% SDS. A concentrated solution of 25 wt% NH3 ·H2 O was added drop-by-drop with stirring until pH 10. The reaction mixture was heated at 40 ◦ C for 30 min, then separated with magnet and swilled to neutral. The obtained Fe3 O4 particles were washed with ethanol for three times followed by drying at 80 ◦ C under vacuum for 11 h. SiO2 /Fe3 O4 (SF) was prepared using a modified sol–gel process. In this process, a suspension of the synthesized Fe3 O4 nanoparticles (0.4 g) was diluted by a mixture of the precursor of TEOS (11 mL) and ethanol (60 mL), followed by the drop-by-drop addition of a mixture of 25 wt% NH3 ·H2 O (8 mL) and ethanol (40 mL) with stirring at room temperature for 6 h. The product was separated with magnet, washed with ethanol and dried at 60 ◦ C for 8 h. Coating SF with TiO2 was also carried out using the sol–gel technique. Briefly, the suspension with SF particles (0.5 g), TBOT (5 mL) and ethanol (20 mL) was prepared and ultrasonic for 10 min. Then, a mixture of 1.5 mL H2 O, 0.1 mL hydrochloric acid and 20 mL ethanol was added drop-wise to the SF suspension under vigorous stirring at 40 ◦ C until gel was formed. The product was dried at 80 ◦ C to evaporate ethanol, and then ground into powder and calcined at 450 ◦ C for 1 h. 2.3. Preparation of magnetically attached TSF electrode Typically 50 mg of TSF was dispersed in 1 mL of ethanol to give a 50 mg/mL suspension with the aid of ultrasonic agitation. Prior to modification, the graphite electrode surface (2 cm × 4 cm) was polished with emery papers, ultrasonically washed in ethanol and distilled water for several minutes and then dried with nitrogen gas. A Nd–Fe–B magnet (3500 Gauss, 2 cm × 4 cm × 0.5 cm, Yingshen Magnet Co., China) enwrapped with a layer of PVC thin film was glued to the back of the working surface of graphite and then 120 L of 50 mg/mL TSF suspension was coated on the working surface of electrode and dried in air. The PVC film could avoid the contact of magnet with solution when the magnetically attached TSF electrode was utilized in electrolyte solution. 2.4. Apparatus and procedures XRD patterns were measured using a diffractometer (X’ pert PRO, PANalytical B.V., the Netherlands) with radiation of a Cu target (Ka, = 0.15406 nm). TEM observation was performed on a JEM100CXII TEM instrument (JOEL Ltd., Japan) operated at a 100 kV
Fig. 1. Diagram of PEC degradation experimental setup using magnetically attached TSF electrode.
accelerating voltage. The FT-IR spectra were measured with an Equinox 55 FTIR spectrophotometer (Bruker Co., Germany). The concentration of MB was measured by monitoring the absorbance at the maximum absorption wavelength at 660 nm with a UV2000 UV-visible spectrometer (Shanghai Unico Instruments Co., China). The PEC degradation experiments were performed in a quartz photoreactor containing 100 mL sample solution. The magnetically attached TSF electrode was placed in the photoreactor as the working electrode (W.E.). A platinum wire and a saturated calomel electrode (SCE) were used as the counter electrode (C.E.) and reference electrode (R.E.), respectively. All the potentials were versus SCE. The electrochemical measurements were carried out using a CHI 660A potentiostat (Shanghai Chenhua Instrument Co., China). A 15 W low pressure Hg lamp with a major emission wavelength of 253.7 nm was used for UV irradiation. The experimental setup for PEC degradation of pollutants was illustrated in Fig. 1. The voltammetric measurements were performed in a 10 mL conventional three-electrode electrolytic cell. A CNT-modified glassy carbon electrode described as our previous report [33], a platinum wire and a SCE served as the working, auxiliary and reference electrodes, respectively. For voltammetric analysis of degraded diclofenac, 1.0 mL sample was taken out of the photoreactor, and diluted with 8.0 mL doubly distilled water and 1.0 mL of 1.0 mol/L H2 SO4, and then analyzed in the electrolytic cell using the CNT-modified electrode. The chemical oxygen demand (COD) value was determined using the traditional dichromate method [33]. LC/MS analysis was carried out in a Finnigan LTQ XL linear ion trap mass spectrometer (Thermo Fisher, USA). The degradation product of diclofenac was separated by Ultimate 3000 (Donex, USA) liquid chromatography using an Acclaim 120 C18 column, 250 mm × 4.6 mm, with 60–80% CH3 CN (0–7 min), 80% CH3 CN (7–25 min) as eluent. The flow rate was 2.5 mL/min. The detection wavelength was set at 365 nm. The mass spectrometer was equipped with an electrospray ionization (ESI) source. The source voltage was set at the 3.5 kV value. The tuning parameters adopted for ESI source were the following: source current 100.00 A, capillary voltage −35.00 V, capillary temperature 250.00 ◦ C, tube lens offset −200.00 V, sheath gas flow 20.00 L/min, auxiliary gas flow 6.00 L/min.
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Fig. 2. (A) XRD patterns, (B) TEM images and (C) FT-IR spectra for Fe3 O4 (a), SF (b) and TSF (c) nanoparticles (bar: 140 nm).
3. Results and discussion 3.1. Surface characterization of TSF nanoparticles Fig. 2A shows the XRD patterns for pure Fe3 O4 , SF and TSF nanoparticles. As can be seen, the characteristic peaks at 2 = 30.4◦ , 35.6◦ , 43.2◦ , 53.8◦ , 57.2◦ and 62.9◦ for pure Fe3 O4 nanoparticles agree well with those XRD patterns of Fe3 O4 nanoparticles in the literature [34–36], indicating a cubic spinel structure of magnetite. The XRD pattern of SF particles is similar to the pattern of Fe3 O4 but shows an obvious diffusion peak at 2 = 15–25◦ , generally considered as the diffusion peak of amorphous silica. For the sample of TSF particles, several diffraction peaks at 25.4◦ , 37.8◦ , 48.1◦ , 54.1◦ , 55.2◦ , 68.8◦ , 70.3◦ , 75.2◦ reveal the formation of anatase TiO2 . At the
same time, the TEM images of three types of prepared particles are shown in Fig. 2B, which indicate the diameter of Fe3 O4 nanoparticles are increased after it is coated with SiO2 layer. While the second layer of TiO2 is coated on SF, the size of nanoparticles is further enlarged up to 120 nm. Fig. 2C compares the FT-IR spectra of these particles. For Fe3 O4 , the absorption band at 568 cm−1 corresponds to the Fe–O vibration from the magnetite phase [37]. While SiO2 is coated on Fe3 O4 , the absorption bands at 1093, 807, 466 cm−1 are observed due to the stretching and deformation vibrations of Si–O–Si. For TSF, the absorption bands in the range from 500 to 900 cm−1 relating to the stretching vibration of Ti–O–Ti bond are clearly observed while the band at 1093 cm−1 corresponding to the asymmetric stretching vibration of Si–O–Si disappears due to the coat of TiO2 in the outlayer.
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3
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Fig. 3. (A) Effect of different dispersants including (a) ethanol, (b) DMF, (c) 0.5% CTAB, (d) 0.5% SDS and (e) water on the photocurrent response of magnetically attached TSF electrode in 0.1 mol/L Na2 SO4 at a bias potential of +0.8 V. (B) Comparison of MB degradation efficiency by different processes. PEC: UV + TSF + 0.8 V; PC: UV + TSF; EC: TSF + 0.8 V; P: UV. (C) Degradation curves of magnetically attached TSF electrode without (a) and with (b) magnet. (D) 100 min degradation efficiencies for MB in successive degradation cycles using the magnetically attached TSF electrode. Error bars were derived from three repeated measurements.
3.2. PEC activity of magnetically attached TSF electrode To obtain a high-performance TSF electrode, the influence of different dispersants such as 0.5% SDS, 0.5% CTAB, DMF, water and ethanol on the photocurrent response of magnetically attached TSF electrodes was measured in an aqueous solution of 0.1 mol/L Na2 SO4 by applying +0.8 V anodic bias potential. As shown in Fig. 3A, the result clearly indicates that all these magnetically attached TSF electrodes sensitively respond to the UV illumination. However, the response is dispersant-dependent. The photocurrent is only ca. 15 A on the electrode coated with water-dispersed TSF. While DMF, 0.5% SDS, or 0.5% CTAB was used as dispersant instead of pure water, the photocurrent is improved to ca. 30 A. The photocurrent response is dramatically improved to ca. 85 A on the TSF-loaded electrode prepared using ethanol as dispersant. This difference arises from the dispersion of TSF nanoparticles. In aqueous suspension, TSF nanoparticles tend to aggregate due to their high surface energy [38]. The aggregation of TSF nanoparticles reduces the surface area of catalyst to absorb UV illumination, leading to low photocurrent response. When surfactants such as SDS and CTAB are present in suspension, surfactant molecules adsorbed to TSF nanoparticles diminish the aggregation of nanoparticles by modifying their electrostatic, hydrophobic and steric interactions [39]. While TSF is dispersed in DMF, this organic solvent with high dielectric constant forms an electrical double-layer structure on TiO2 , which decreases the interfacial surface energy and improves the dispersibility of TSF nanoparticles [40]. In comparison, ethanol is the most efficient dispersant for ultrafine TSF nanoparticles due to the formation of a solvent layer on the particle surface which markedly reduces the repulsive potentials between
particles [41,42]. Accordingly, ethanol was selected as the dispersant to prepare TSF electrodes in the following degradation experiments. MB is an effective organic dye for evaluating the photocatalytic (PC) or PEC activity of semiconductor materials. Thus, the PEC degradation performance of TSF-loaded electrode was investigated using MB as the model pollutant. The degradation experiments were carried out in a solution containing 0.1 mol/L Na2 SO4 and 5 mg/L MB. Fig. 3B compares the PEC removal of MB on the magnetically attached TSF electrode with other degradation techniques such as photocatalysis (PC), electrochemical oxidation (EC), and direct photolysis (P). The results are described as the variation of the removal efficiency (C0 − C)/C0 with degradation time t, where C0 is the initial concentration of MB, and C is the concentration of MB at time t. It is found that the 100 min removal efficiencies for MB by PC, EC and P processes are 18.1%, 11.7% and 11.9%, respectively. In comparison, the PEC process provides the most efficient way to the degradation of MB. By PEC, 33.4% of MB is degraded after 100 min treatment. The role of magnetic loading was clarified by comparing the degradation efficiencies for MB on TSF electrodes with and without an external magnet (Fig. 3C). It is observed that at the initial degradation time less than 20 min, the difference between two degradation curves is not noticeable. However, with further increasing the degradation time, the magnetically attached TSF electrode shows obviously higher degradation efficiency. For example, the 100 min degradation efficiency of MB on the TSF electrode without magnet is only 27.3%; whereas the efficiency is improved to 33.4% on the magnetically attached TSF electrode. This difference can be attributed to the good adherence of TSF nanoparticles
X. Hu et al. / Journal of Hazardous Materials 196 (2011) 220–227
to the electrode surface by the attraction of magnet. While in the absence of magnet, many TSF particles peel off from the electrode surface during the long-time PEC degradation process, leading to the reduced degradation efficiency. To check the stability of the magnetically attached TSF electrode, several successive PEC degradation experiments were carried out. As can be seen from Fig. 3D, the degradation efficiency only shows a slight decrease in several PEC degradation cycles, demonstrating the desired stability of such a magnetically attached TSF electrode. Moreover, the graphite electrode was also stable and we did not observe dissolution of graphite during the long-time degradation process. Thus, graphite electrode was selected as support for TSF loading.
(A)
18
0.6
0.4
0.0 0
10
20
30
40
50
60
Time / min
(B) 100
Diclofenac degradation COD removal
80
Percentage (%)
(A)
P PC PEC
0.2
3.3. Application of magnetically attached TSF electrode to the PEC degradation of diclofenac The magnetically attached TSF electrode was applied to study the PEC degradation of diclofenac. Considering that diclofenac was an electroactive molecule and CNT-based voltammetry could provide a cheap and useful analytical tool for studying the pollution control [33,43], we developed a voltammetric method using a CNT-modified electrode to monitor the degradation of diclofenac during the PEC treatment. Fig. 4A shows the voltammetric response recorded on the CNT-modified electrode, which
1.0
0.8
(C0-C) / C0
224
60
40
20 15 0
a
12
0
10
20
30
40
50
60
i / µA
Time / min 9
Fig. 5. (A) Degradation efficiency curves for 1.0 × 10−3 mol/L diclofenac in 0.1 mol/L Na2 SO4 with different techniques. (B) Comparison between COD removal efficiency and diclofenac degradation efficiency treated by PEC technique. Error bars were derived from three repeated measurements.
g 6 3 0 0.5
0.6
0.7
0.8
0.9
1.0
1.1
E / V vs. SCE
(B) 15
ip / µ A
12
9
6
3
0 0.00
0.02
0.04
0.06
0.08
0.10
C / mmol/L Fig. 4. (A) Linear sweep voltammograms of 0.1 mol/L H2 SO4 solution containing (a) 1.0 × 10−4 mol/L, (b) 8.0 × 10−5 mol/L, (c) 5.0 × 10−5 mol/L, (d) 3.0 × 10−5 mol/L, (e) 1.0 × 10−5 mol/L, (f) 5.0 × 10−6 mol/L and (g) 1.0 × 10−6 mol/L diclofenac on CNTmodified electrode at 20 mV/s. (B) Linear relationship between the concentration of diclofenac and peak current. Error bars were derived from three repeated measurements.
is linearly increased with the concentration of diclofenac from 1.0 × 10−6 to 1.0 × 10−4 mol/L (Fig. 4B). The linear regression equation is expressed as ip (A) = 1.479 + 138.6c (mmol/L) (correlation coefficient r = 0.9963). Such a calibration curve was employed to evaluate the degradation efficiency of diclofenac. Typically, diclofenac was degraded on the TSF electrode for appropriate time, and then 1.0 mL of sample was taken out of the reactor and added in electrolytic cell containing suitable supporting electrolyte to analyze the concentration of diclofenac by voltammetric measurement on the CNT-modified electrode. Fig. 5A illustrates the degradation efficiency curves for 1.0 × 10−3 mol/L diclofenac with different techniques. The result indicates that 59.1% of diclofenac is degraded after 60 min UV irradiation. When the magnetically attached TSF electrode is employed to degrade diclofenac, the degradation efficiency reaches 77.3% after 60 min UV irradiation, meaning the photocatalysis of TSF film towards diclofenac. While +0.8 V and UV irradiation are simultaneously applied on the magnetically attached TSF electrode, the degradation efficiency of diclofenac is improved to 95.3% after 45 min treatment, indicating the effective PEC degradation of diclofenac on this magnetic electrode. However, the COD removal efficiency is only 47.8% after 45 min PEC treatment (Fig. 5B), almost half of the degradation efficiency. The difference between the degradation efficiency determined by voltammetry and COD analysis indicates that mineralization of diclofenac is slow and many diclofenac molecules are quickly degraded to form intermediates during the PEC process. Thus, prolonging the PEC treatment time is necessary for further oxidation of diclofenac and its degradation intermediates to achieve the complete mineralization.
X. Hu et al. / Journal of Hazardous Materials 196 (2011) 220–227
(A)
225
data03_110323205147 #2056 RT: 22.10 AV: 1 NL: 5.05E5 F: ITMS - c ESI Full ms [75.00-400.00] 294.05
100 250.16
90
Relative Abundance
80 70 296.06
60 252.15
50 40 30 20 297.98
10 83.11 97.17
0
80
100
119.15
120
141.16 157.24
140
178.14
196.32
180
200
160
214.13 216.26
258.10
241.51
220
240
260
281.41
280
316.03
300
320
376.05 393.18
341.08 353.47
340
360
380
400
m/z
(B)
data03_110323205147 #1891 RT: 20.36 AV: 1 NL: 4.75E5 F: ITMS - c ESI Full ms [75.00-400.00] 214.10
100 90
257.99
Relative Abundance
80 70 60 50 40 216.07
30
260.02
20 10
178.26 83.13
106.11 119.07
141.24
80
100
140
0
120
196.28
166.27
160
180
253.31
227.29
200
220
240
260
281.19 296.05
323.11 340.07
280
320
300
340
359.17
378.14 393.10
360
380
400
m/z Fig. 6. LC-MS spectra of (A) diclofenac and (B) its PEC degradation intermediate.
To understand the PEC degradation mechanism of diclofenac on this magnetic electrode, the main degradation intermediate after 60 min PEC treatment was separated by HPLC and identified with ESI-MS. Fig. 6 compares the LC/MS spectra of diclofenac and its PEC degradation intermediate. The mass spectrum of diclofenac produces a molecular ion at m/z 294 and a fragment ion at m/z 250 (Fig. 6A). While for the degradation intermediate of diclofenac, the spectrum shows a molecular ion at m/z 258 and a fragment ion at m/z 214 (Fig. 6B). The difference of 36 between the degradation intermediate and diclofenac is due to the loss of chlorine and hydrogen atoms, consistent with the proposal by Cavalheiro [44]. Based on this result, the mechanism of PEC oxidation of diclofenac by hydroxyl radical can be proposed. In the first step, photogenerated electrons and holes over TiO2 are obtained when TSF is irradiated with UV light [45]. hv
TiO2 −→e− + h+ Then, holes react with H2 O or h+ + H2 O → H+ + • OH
h+ + OH− → • OH
While +0.8 V potential is applied, diclofenac is electro-oxidized on the TSF-loaded electrode to form a radical cation [46].
CO2H Cl
to form hydroxyl radicals. (2)
NH
Cl
Cl
CO2H +. + eNH Cl (4)
At the same time, the applied anodic bias potential drives photogenerated electrons to the counter electrode through the external electric circuit, which reduces the recombination of electrons and holes. Thus, more hydroxyl radicals according to reaction (2) or (3) are obtained to oxidize diclofenac radical cation.
(1) OH−
(3)
Cl
CO2H +. NH Cl
+
.OH
CO2H + H2O + Cl-
NH Cl
(5)
226
X. Hu et al. / Journal of Hazardous Materials 196 (2011) 220–227
Such an oxidation intermediate of diclofenac can undergo further reactions and ring cleavage to form small molecular derivatives under hydroxyl radical attack, which is eventually mineralized to CO2 , Cl− , H2 O and NH3 . 4. Conclusions Magnetic TSF core–shell nanoparticles were synthesized and loaded on the electrode surface with the aid of an external magnet. A series of experiments were carried out to study the PEC activity of magnetically attached TSF electrode. It was found that dispersant had a prominent effect on the performance of the prepared TSF electrode and the ethanol-dispersed TSF electrode showed high photoelectrochemical activity. For such a magnetic nanoparticles-loaded electrode, the external magnetic force kept TSF nanoparticles firmly adhered to the electrode and allowed efficient electron transfer between catalyst and electrode, leading to high PEC degradation efficiency and desirable stability during long-time degradation process. Based on this TSF-loaded electrode, effective PEC degradation of diclofenac was obtained. Our work demonstrates that magnetic loading of TSF nanoparticles on electrode surface provides a new strategy for preparing photoelectrochemical electrode with high stability, which could be applied to the PEC degradation of organic pollutants. Acknowledgements This work was supported by the National Natural Science Foundation of China (grant no. 20977037). The authors thank the Analytical and Testing Center of Huazhong University of Science and Technology for the help in surface analysis. References [1] K. Pirkanneime, M. Sillanpaa, Heterogeneous water phase catalysis as an environmental application: a review, Chemosphere 48 (2002) 1047–1060. [2] A. Fujishima, T.N. Rao, D.S. Truk, Titanium dioxide photocatalysis, J. Photochem. Photobiol. C 1 (2000) 1–21. [3] D. Beydoun, R. Amal, G. Low, S. Mcevoy, Occurrence and prevention of photodissolution at the phase junction of magnetite and titanium dioxide, J. Mol. Catal. A: Chem. 180 (2002) 193–200. [4] T.A. Gad-Allah, K. Fujimura, S. Kato, S. Satokawa, T. Kojima, Preparation and characterization of magnetically separable photocatalyst (TiO2 /SiO2 /Fe3 O4 ): effect of carbon coating and calcination temperature, J. Hazard. Mater. 154 (2008) 572–577. [5] S. Rana, R.S. Srivastava, M.M. Sorensson, R.D.K. Misra, Synthesis and characterization of nanoparticles with magnetic core and photocatalytic shell: anatase TiO2 –NiFe2 O4 system, Mater. Sci. Eng. B 119 (2005) 144–151. [6] F. Chen, Y. Xie, J. Zhao, G. Lu, Photocatalytic degradation of dyes on a magnetically separated photocatalyst under visible and UV irradiation, Chemosphere 44 (2001) 1159–1168. [7] Y. Gao, B. Chen, H. Li, Y. Ma, Preparation and characterization of a magnetically separated photocatalyst and its catalytic properties, Mater. Chem. Phys. 80 (2003) 348–355. [8] D.-K. Lee, I.-C. Cho, Characterization of TiO2 thin film immobilized on glass tube and its application to PCE photocatalytic destruction, Microchem. J. 68 (2001) 215–223. [9] J.-M. Herrmann, H. Tahin, Y. Ait-Ichou, G. Lassaletta, A.R. González-Elipe, A. Fernández, Characterization and photocatalytic activity in aqueous medium of TiO2 and Ag–TiO2 coatings on quartz, Appl. Catal. B: Environ. 13 (1997) 219–228. [10] X. Cao, Y. Oda, F. Shiraishi, Photocatalytic and adsorptive treatment of 2,4dinitrophenol using a TiO2 film covering activated carbon surface, Chem. Eng. J. 156 (2010) 98–105. [11] M. Huang, C. Xu, Z. Wu, Y. Huang, J. Lin, J. Wu, Photocatalytic discolorization of methyl orange solution by Pt modified TiO2 loaded on natural zeolite, Dyes Pigm. 77 (2008) 327–334. [12] N.J. Peill, M.R. Hoffmann, Mathematical model of a photocatalytic fiber-optic cable reactor for heterogeneous photocatalysis, Environ. Sci. Technol. 32 (1998) 398–404. [13] S. Horikoshi, N. Watanabe, H. Onishi, H. Hidaka, N. Serpone, Photodecomposition of a nonylphenol polyethoxylate surfactant in a cylindrical photoreactor with TiO2 immobilized fiberglass cloth, Appl. Catal. B: Environ. 37 (2002) 117–129.
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Journal of Hazardous Materials 196 (2011) 228–233
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Treatment of radioactive liquid waste (Co-60) by sorption on Zeolite Na-A prepared from Iraqi kaolin Yasmen A. Mustafa a,∗ , Maysoon J. Zaiter b a b
Department of Environmental Engineering, University of Baghdad, P.O. Box 47121, Jadria, Baghdad, Iraq Ministry of Science and Technology, Iraq
a r t i c l e
i n f o
Article history: Received 14 July 2011 Received in revised form 3 September 2011 Accepted 5 September 2011 Available online 10 September 2011 Keywords: Ion exchange Zeolite Cobalt-60 Fixed bed Thomas model
a b s t r a c t Iraqi synthetic zeolite type Na-A has been suggested as ion exchange material to treat cobalt-60 in radioactive liquid waste which came from neutron activation for corrosion products. Batch experiments were conducted to find out the equilibrium isotherm for source sample .The equilibrium isotherm for radioactive cobalt in the source sample showed unfavorable type, while the equilibrium isotherm for the total cobalt (the radioactive and nonradioactive cobalt) in the source sample showed a favorable type. The ability of Na-A zeolite to remove cobalt from wastewater was checked for high cobalt concentration (822 mg/L) in addition to low cobalt concentration in the source sample (0.093 mg/L). A good fitting for the experimental data with Langmuir equilibrium model was observed. Langmuir constant qm which is related to monolayer adsorption capacity for low and high cobalt concentration was determined to be 0.021 and 140 mg/gzeolite . The effects of important design variables on the zeolite column performance were studied these include initial concentration, flow rate, and bed depth. The experimental results have shown that high sorption capacity can be obtained at high influent concentration, low flow rate, and high bed depth. Higher column performance was obtained at higher bed depth. Thomas model was employed to predict the breakthrough carves for the above variables. A good fitting was observed with correlation coefficients between 0.915 and 0.985. © 2011 Elsevier B.V. All rights reserved.
1. Introduction There are many sources of low- and intermediate-level radioactive liquid waste in Altwatha site in Iraq, those include radiochemical, radiomedical and reactors wastes. All these activities were stopped after the war at 1991. 14-Tammuze reactor, material testing reactor, MTR (swimming pool reactor type), its containment building was destroyed completely and the reactor pool was opened to the surrounding environment. Emergency and health physics teams had been emptied the radioactive liquid waste from the pool and filled with other clean water in order to reduce exposure to the surrounding environment. Radioactive liquid waste used in this study came from reactor pool. The pollutant cobalt60 in the liquid waste came from neutron activation for corrosion products. Cobalt is a major contributor toward the radiation buildup problem because of its rather long half-life (t1/2 = 5.27 y) and high energy (sum peak of 2.5 MeV) gamma emission. Radiations can produce harmful effects on living organisms. Partly as a result of ignorance and partly due to accidental circumstances, a number of cases of injury, ranging from minor early
∗ Corresponding author. Tel.: +964 7705094551. E-mail address:
[email protected] (Y.A. Mustafa). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.013
skin lesions to delayed bone cancer and leukemia, where reported among radiologists and others who were exposed to excessive amount of radiations. Ion exchange process is considered as one of the more effective methods for the removal of radioisotopes from liquid waste. Ion exchange is a chemical treatment process used to remove unwanted ionic species from wastewater. It is basically a simple process based on reversible interchange of ions between liquid and solid with no permanent changes in the structure of the solid. Compared with other separation methods ion exchange has a number of advantages and limitations. The immediate reason for their use in the nuclear industry is the comparative simplicity, of operation and the associated equipment involved, together with their high efficiency and the possible volume reduction. A limitation of ion exchange is its unsuitability for use with substances without charge as in the case in many complexes and for particles of the colloidal type. Many researchers are utilized different materials to remove metal radionuclides from aqueous radioactive wastes by ion exchange process in preparation for their final disposal [1–5]. The low cost of natural and synthetic zeolite makes their use attractive in water treatment applications. The use of zeolites towards decontamination of low- and intermediate-level radioactive liquid waste and as carriers of target elements in a nuclear reactor has received considerable attention [6–9]. The principle
Y.A. Mustafa, M.J. Zaiter / Journal of Hazardous Materials 196 (2011) 228–233
229
Table 1 Zeolite composition. Element
Percentage (%)
SiO2 Al2 O3 L.O.I. Na2 O CaO TiO2 Fe2 O3 MgO K2 O
34.48 29.94 15.05 13.4 2.52 1.70 0.95 0.08 0.03
reason for using zeolite in radioactive waste treatment is their high selectivity and resistance to degradation from radiation. Synthetic type Na-A zeolite as crystalline with molecular formula Na2 O:0.79, Al2 O3 :1, SiO2 :1.95, powder and pellets are prepared in Iraqi Atomic Energy Commission (IAEC) Laboratories was used as material under investigation in this study. After the zeolite has been exhausted in the treatment of radioactive liquid waste, there is no regeneration and it will be disposed as radioactive solid waste. The zeolite with retained radioactivity must be immobilized by cementation as final process for disposal [10,11]. Zeolite amounts in weight up to 10% of cement amount did not cause any weakness on the strength of cemented matrix [12,13]. The aim of this research is to investigate the performance of Iraqi Na-A zeolite in the treatment of radioactive liquid waste contained Cobalt-60 by batch and fixed bed ion exchange experiments. Different variables in fixed bed experiments (concentration, flow rate and bed depth) were studied. Thomas model was applied to predict the breakthrough curves for ion exchange, and to determine the characteristic parameters useful in design using nonlinear regression. 2. Materials and methods 2.1. Adsorbent Zeolite was used as sorbent, it is of type A in the sodium form (Na-A). Zeolite in crystalline powder form was prepared at Iraqi Atomic Energy Commission (IAEC) laboratories [14]. This material was prepared from Iraqi Kaolin (from Kaara region). It treated with 2 N sodium hydroxide solution at boiling point under agitation for 2 h. Pellets were formed by mixing the product with natural kaolin in specific proportion as cross linking material. These pellets were calcined in the furnace at 600 ◦ C for 3 h for solidification. Zeolite in crystalline powder form was used in batch system experiments, whereas the pellet form was used in continues system experiments. Elemental analysis was accomplished in the Geological Survey and Mining Company’s Laboratories. The composition of zeolite is shown in Table 1. The physical properties of zeolite pellets are listed in Table 2. X-ray diffraction and infrared scanning were performed in Geological Survey and Mining Company’s Laboratories (Figs. 1 and 2). The results were found to be coincided with type Na-A zeolite which was used throughout the research.
Fig. 1. X ray diffraction for Na-A zeolite.
2.2. Adsorbate Two types of liquid waste were used in the present research: • Radioactive liquid waste (the source sample) collected with the aid of health physics team from the reactor pool. The samples contained Co-60 from neutron activation of corrosion product. This type of liquid waste was used in the batch experiments. • Simulated waste solutions were prepared from cobaltous oxide (CoO carrier free of Co-60) supplied by BDH chemical Ltd. Pool/England (Mwt. equal to 74.93 with 99.9% purity). Deionized water with electrical conductivity of less than 3 siemens/cm was used to prepare the solutions. This type of liquid waste was used in the column experiments. 2.3. Equipments Gamma spectrometry (Canberra-USA) with sodium iodide scintillation detector was used to analyze the radioactivity concentration of the samples, and the atomic adsorption spectrophotometry (A.A-6200 Shemadzu, Japan), flame emission type, was used to analyze the total cobalt concentration of samples (radioactive and nonradioactive). 2.4. Batch experiments Batch experiments were used to study the sorption rate and sorption isotherm of cobalt-60 on synthetic type Na-A zeolite. To determine the sorption rate, 250 mL of radioactive liquid waste (the source sample with radioactivity of 19.684 Bq/L) was placed into five beakers of 1 L size. Accurately weighted amount of zeolite powder (0.5 g) was added to these beakers. Each beaker was sealed with parafilm and placed in an oscillating shaker (90 rpm).
Table 2 Physical properties of zeolite pellets. Dimension
2 mm × 6 mm
Bulk density Particle density Internal porosity Void fraction of bed Surface area
0.58 g/cm3 2.38 g/cm3 0.3 0.756 7.82 × 104 cm2 /g
Fig. 2. Infrared scanning for Na-A.
Y.A. Mustafa, M.J. Zaiter / Journal of Hazardous Materials 196 (2011) 228–233
Activity of cobalt (Bq/L)
230
20 18 16 14 12 10
0
50
100
150
200
Time (min) Fig. 3. Sorption rate for radioactive cobalt on zeolite.
2.5. Column experiments For the design of industrial scale fixed-bed sorption systems, column operations are an absolute must. Column experiments were carried out at different variables, concentration, flow rate, and bed depth to study the performance of synthetic type Na-A zeolite pellets in removing cobalt from liquid waste. Glass column of 1 cm (I.D.) and 30 cm height was used. The zeolite was confined in the column by fine stainless steel screen at the bottom. Two containers were used; the first one as feed container while the second one as receiver. After preparing the desired concentration in the feed container the solution was pumped through the bed. The concentration of the cobalt in the samples was determined by means of atomic adsorption spectrophotometer. 3. Results and discussion 3.1. Source sample analysis Radioactive liquid wastes (source samples) were analyzed in gamma spectrometry scintillation detector (NaI) sodium iodide type. The analytical graph is presented in Fig. 4. It was observed that the radioactivity of the sample came from cobalt-60 presented in the reactor coolant by neutron activation of corrosion products. Two source samples were used for batch experiments. The analysis of the source sample which used for sorption rate experiments was found to be of concentration 19.684 Bq/L (4.79 × 10−10 mg/L radioactive cobalt). The analysis of the source sample which used for sorption isotherm experiments
0.8 0.6 0.4 0.2 0
0
1
2
3
4
Cex1010(mg/L) Fig. 5. Sorption isotherm for radioactive cobalt in source sample on zeolite.
was found to be of concentration 17.872 Bq/L (4.35 × 10−10 mg/L radioactive cobalt), this sample was analyzed by atomic adsorption spectrophotometry .The results have shown that total cobalt concentration (radioactive and nonradioactive) was 0.093 mg/L. 3.2. Sorption isotherms Fig. 5 shows the adsorption isotherm for radioactive cobalt in the source sample, it is of unfavorable type, while Fig. 6 shows the adsorption isotherm for total cobalt (radioactive and nonradioactive cobalt) in the source sample, it represents the favorable type of adsorption isotherm. The ability of zeolite to remove cobalt from wastewater was checked for high cobalt concentration (822 mg/L), in addition to the low concentration (0.093 mg/L in the source sample). Fig. 7 shows the adsorption isotherm for high cobalt concentration. A favorable type of adsorption isotherm can be observed from this figure. Langmuir equation was applied to the sorption isotherms for low and high concentration of cobalt: Ce Ce 1 + = qe qm qm b
(1)
where Ce is the equilibrium concentration (mg/L), qe is the amount of adsorbate adsorbed per unit mass of the adsorbent at Ce (mg/g), qm is Langmuir constant related to monolayer adsorption capacity
qe x102 (mg/g zeolite lit )
Each beaker was shacked for a various periods of time (10, 30, 60, 120, and 180 min). After shaking, the contents were distributed into six closed plastic ampoules, and then placed in a centrifuge (10,000 rpm) for 30 min. The supernatant of all ampoules were collected into 1 L beaker and then transferred to special container to analyze in gamma spectrometry scintillation detector .The results have shown that 1 h is sufficient for reaching equilibrium (Fig. 3). To determine the sorption isotherm for low cobalt concentration, 250 mL of radioactive liquid waste (the source sample with radioactivity of 17.872 Bq/L) was placed into five 1 L beakers. Accurately weighted amounts of zeolite powder (0.25, 0.5, 0.73, 1.02, and 2.0 g) were added to these beakers. Each beaker after sealing by parafilm was placed in an oscillating shaker (90 rpm) for 4.5 h to reach equilibrium. The content of each beaker was distributed into six closed ampoules, and then placed in a centrifuge (10,000 rpm) for 30 min. The supernatant was collected into (1 L) beaker, after that it transferred to special container to be analyzed using Gamma Spectrometry Scintillation Detector. Concerning high cobalt concentration simulated waste solution was used. The zeolite powder weights that used in the experiments were (0.5, 1, 2, and 3 g). The same steps for the above procedure was applied, using atomic adsorption spectrophotometry for analysis.
qe x1010 (mg/g zeolite)
Fig. 4. Analytical graph for source.
2 1.5 1 0.5 0
0
2
4
6
Cex102 (mg/L) Fig. 6. Sorption isotherm for total cobalt sample on zeolite.
8
150
8
100
6 C (mg/L)
qe (mg/g)
Y.A. Mustafa, M.J. Zaiter / Journal of Hazardous Materials 196 (2011) 228–233
50 0
0
200
400
4
0
Fig. 7. Sorption isotherm for high cobalt concentration on zaeolite.
1000
2000
3000
Fig. 8. The breakthrough curves at different concentrations.
qm (mg/gZeolite )
b (L/g)
R2
0.021 140
6 × 10 9.5
0.97 0.98
4
12
Capacity (q) mg/gzeolite
0.093 822
0
Time (min)
Table 3 Langmuir constants for low and high concentration.
Low High
3 mg/L 5 mg/L 7 mg/L Thomas
2
600
Ce(mg/L)
Concentration (mg/L)
231
(mg/g), and b is Langmuir constant related to the energy of adsorption (L/mg). Langmuir constants were determined from the slop and the intercept of the straight line when plotting Ce /qe vs. Ce . The Langmuir constants were listed in Table 3, the correlation coefficients are 0.97 and 0.98. From Table 3, it can be observed that high monolayer adsorption capacity (140 mg/gzeolite ) for high cobalt concentration is so high compared with monolayer adsorption capacity (0.021 mg/gzeolite ) for low cobalt concentration. The increase in uptake capacity of Zeolite material with the increase of initial cobalt concentration may be due to higher probability of collision between cobalt ion and Zeolite. Also the driving force for sorption process was increased due to high concentration gradient at high cobalt concentration.
10 8 6 4
3 mg/L 5 mg/L 7 mg/L
2 0
0
1000
2000
3000
Time (min) Fig. 9. Rate of sorption at different concentrations.
6 5
C (mg/L)
3.3. Fixed bed experiments Fixed bed sorption experiments were carried out to study the sorption dynamics. The effects of process variables on the zeolite column performance were studied. These variables include initial concentration, flow rate, and bed depth (Table 4). The breakthrough curves (the concentrations of cobalt in the effluent C vs. t), and the rate of sorption curves (cobalt adsorbed per unit mass of zeolite q vs. t) for the above variables were plotted (Figs. 8–13).
4 3
1.5 mL/min
2
2.5 mL/min 5 mL/min
1 0
Thomas
0
2000
4000
6000
Time (min) 3.4. Effect of initial concentration Fig. 10. The breakthrough curves at different flow rates.
Experiments with different initial concentrations (3, 5, and 7 mg/L) were carried out at fixed flow rate 2.5 mL/min and bed depth 3.5 cm. Fig. 8 shows the breakthrough curve for the above concentrations. It can be observed that the breakthrough point (which is a point on the breakthrough curve where the effluent concentration reaches the maximum allowable concentration) appears earlier for highest initial concentration. This happens because the binding
sites and the ion exchange sites become quickly saturated for high concentrations and the primary adsorption zone mobilized rapidly across the bed. Fig. 9 shows the rate of sorption for the above concentrations. Apparent increase in the sorption capacity was observed for high concentration.
Table 4 Fixed bed experimental data and Thomas model parameters. Co (mg/L)
Q (mL/min)
Z (cm)
Bed capacity (mg/gzeolite )
Column performance (%)
KTh (mL/mg/min)
qo (mg/g)
R2
3 5 7 5 5 5 5
2.5 2.5 2.5 1.5 5 2.5 2.5
3.5 3.5 3.5 3.5 3.5 4.5 5.5
5.05 8.64 9.64 8.94 8.2 8.94 9.2
46 47.3 38 50 32 52.2 59
1.2 0.72 0.43 0.36 0.84 0.62 0.6
5.6 10.1 10.4 11.4 11.5 10.2 10.4
0.985 0.976 0.915 0.982 0.955 0.976 0.979
232
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Capacity (q) mg/gzeolite
10 8 6 4
1.5 mL/min
2
5 mL/min
0
2.5 mL/min
0
2000
4000
3.6. Effect of bed height
6000
Time (min) Fig. 11. Rate of sorption at different flow rates.
6
C( mg/L)
5 4 3 2
3.5 cm 4.5 cm 5.5 cm Thomas
1 0
1000
0
2000
3000
4. Modeling the break through curve
Fig. 12. The breakthrough curves at different bed depths.
3.5. Effect of flow rate Experiments with different flow rates (1.5, 2.5, and 5 mL/min) were carried out at fixed initial concentration 5 mg/L and bed depth of 3.5 cm. The breakthrough and the rate of sorption curves were shown in Figs. 10 and 11. It can be noticed that the breakthrough point appears earlier for the highest flow rate (Fig. 10). At earlier period of experiment the uptake of cobalt is high for the highest flow rate, but earlier saturation leads to lower capacity .The uptake at low flow rate continues to longer period of time leads to high capacity (Fig. 11). For low flow rate the mobile phase spent greater time with the solid surface of stationary phase, the primary adsorption zone mobilized slowly across the bed in the column, hence later saturation of the bed can be achieved. A linear portion exists for each curve at early period of experiments (Fig. 11). The linear portion of the curve indicate that the film diffusion controls as rate limiting step to a point at which the external surface area becomes
10
Capacity (q) mg/gzeolite
Experiments with different bed depths (3.5, 4.5, and 5.5 cm) were carried out at fixed initial concentration 5 mg/L and flow rate of 2.5 mL/min. The total bed depth must be larger than the primary adsorption zone otherwise the breakthrough point cannot be identified at the breakthrough curve. From Fig. 12 the breakthrough point appears later for highest bed depth, since increasing bed depth provide extra surface area for sorption process to carry on, allowing more than one primary sorption zone to be appeared. Fig. 13 shows nearly the same uptake at early period of experiment then sequential saturation was attained beginning from the lower bed depth. Referring to Table 4, it can be noticed that high percent of column performance ((the amount of cobalt retained in the column)/(amount of cobalt inter the column) × 100) was obtained at high bed depth 5.5 cm. While the maximum sorption capacity obtained at high cobalt concentration 7 mg/L.
4000
Time (min)
The most important criterion in the design of fixed bed sorption systems is the prediction of breakthrough curve, which determines the operation life span of the bed. Thomas model is one of the most general and widely used methods in column performance theory. It gives a general analytical solution for the differential equations that describe the transport of fluid in fixed bed sorption process with nonlinear equilibrium relationship. Thomas model is based on second order reaction kinetics .The equilibrium relationship is assumed to be of Langmuir (favorable) isotherm model. Thomas excludes the effect of axial dispersion. When the sorption isotherm is highly favorable, the actual Thomas model is reduced to the Bohart-Adams model [15]: C 1 = Co 1 + exp((KTh qo M/Q )KTh Co t)
6 4
3.5 cm 4.5 cm
2
5.5 cm
0
1000
2000
3000
Time (min) Fig. 13. Rate of sorption at different bed.
4000
(2)
where KTh is the kinetic coefficient (mL/mg/min), qo is the maximum (equilibrium) sorption capacity of the bed (mg/g), M is the mass of adsorbent packed in the column (g), Q is the flow rate (mL/min), C and Co are the concentrations of cobalt in the influent and in the effluent at any time t. The linearized form of the above equation is as follows: ln
8
0
essentially saturated, and deviation from linearity occur because of the increasing influence of intraparticle transport on the overall rate of mass transfer as each run progresses. The linear segment of the curve extended over a shorter period of time for high flow rate and steeper curve was observed leading to fast saturation of bed. Insufficient residence time of the solute in the column may limit the diffusion through the pores.
C
o
C
−1
=
KTh qo M − KTh Co t Q
(3)
The kinetic coefficient KTh and the equilibrium sorption capacity of the bed qo can be determined from a plot of ln ((Co /C) − 1) against t at a given conditions for the fixed bed. Figs. 8, 10 and 12 show that the Thomas model gives a good fit of the experimental data. The value of KTh and qo are given in Table 4. It can be observed that qo increases as cobalt concentration increases. High value of qo was obtained at low flow rate 1.5 mL/min. In addition high value of qo was obtained at high flow rate 5 mL/min, this may be due to high uptake in early period of experiment. No apparent difference in qo for different bed depths. KTh decreases with the increase in cobalt concentration and bed depths while it increases with the increase of flow rate.
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5. Conclusion Based on the experimental results, the following conclusions can be drawn: Synthetic Iraqi Na-A zeolite can be used to removal cobalt-60 from radioactive liquid waste. Equilibrium isotherm for radioactive cobalt in the source sample shows unfavorable type, while the total cobalt in the source sample shows a favorable type. Langmuir model were successfully modeled the sorption data .The Langmuir constant qm for low and high cobalt concentration was determined to be 0.021 and 140 mg/gzeolite , respectively, which represent the mono layer capacity. The fixed bed column sorption experiments show that high sorption capacity can be attained at high influent concentration, low flow rate, and high bed depth. In the present research, the highest sorption capacity was obtained at high initial cobalt concentration while the highest percent of column performance was obtained at high bed depth. Good fitting of the experimental data and Thomas model was obtained with correlation coefficient of 0.915–0.985. References [1] A. Bhaskarapillai, N.V. Sevilimedu, B. Sellergren, Synthesis and characterization of imprinted polymers for radioactive waste reduction, Ind. Eng. Chem. Res. 48 (2009) 3730–3737. [2] C.R. Preetha, J.M. Gladis, T.P. Rao, Removal of Toxic Uranium from synthetic nuclear power reactor effluents using uranyl ion imprinted polymer particles, Environ. Sci. Technol. 40 (2006) 3070–3074.
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[3] A. Dyer, M. Pillinger, J. Newton, R. Harjula, T. Moller, S. Amin, Sorption behavior of radionuclides on crystalline synthetic tunnel manganese oxides, Chem. Mater. 12 (2000) 3798–3804. [4] I.M. Latheef, M.E. Hukman, R.G. Anthony, Modelling cesium ion exchange on fixed-bed columns of crystalline silicotitanate granules, Ind. Eng. Chem.Res. 39 (2000) 1356–1363. [5] D. Gu, L. Nguyen, C.V. Philip, M.E. Huckman, R.G. Anthony, Cs ion exchange kinetics in complex electrolyte solution using hydrous crystalline silicotitanates, Ind. Eng. Chem. Res. 36 (1997) 5377–5383. [6] Z. Weihua, Z. Lei, H. Runping, Removal of uranium (VI) by fixed bed ionexchange column using natural zeolite coated with manganese oxide, Chin. J. Chem. Eng. 17 (2009) 585–593. [7] A.M. El-Kamash, Evaluation of zeolite A for the sorptive removal of Cs+ and Sr2+ ions from aqueous solutions using batch and fixed bed column operations, J. Hazard. Mater. 151 (2008) 432–445. [8] A.E. Osmanlioglu, Treatment of radioactive liquid waste by sorption on natural zeolite in Turkey, J. Hazard. Mater. B 137 (2006) 332–335. [9] N.V. Elizondo, E. Ballesteros, B.I. Kharisov, Cleaning of liquid radioactive wastes using naturalzeolites, Appl. Radiat. Isotopes 52 (2000) 27–30. [10] R.O. Abdel Rahman, A.A. Zaki, A.M. El-Kamash, Modeling the long-term leaching behavior of 137 Cs, 60 Co, and 152,154 Eu radionuclides from cement–clay matrices, J. Hazard. Mater. 145 (2007) 372–380. [11] J. Li, J. Wang, Advances in cement solidification technology for waste radioactive ion exchange resins: a review, J. Hazard. Mater. B 135 (2006) 443– 448. [12] A.E. Osmanlioglu, Progress in cementation of reactor resins, Prog. Nucl. Energy 49 (2007) 20–26. [13] L. Junfeng, Z. Gang, W. Jianlong, Solidification of low-level-radioactive resins in ASC-zeolite blends, Nucl. Eng. Des. 235 (2005) 817–820. [14] F. Al-Mashta, N. Al-Daghistani, Formation of zeolite type A from Iraqi kaolinite, Petrolum Res. 8 (1989) 83–92. [15] K.H. Chu, Fixed bed sorption: setting the record straight on the Bohart-Adams and Thomas models, J. Hazard. Mater. 177 (2010) 1006–1012.
Journal of Hazardous Materials 196 (2011) 278–286
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Effects of gene-augmentation on the formation, characteristics and microbial community of 2,4-dichlorophenoxyacetic acid degrading aerobic microbial granules Xiang-chun Quan ∗ , Jing-yun Ma, Wei-cong Xiong, Zhi-feng Yang Key Laboratory of Water and Sediment Sciences of Ministry of Education/State Key Laboratory of Water Environment Simulation, School of Environment, Beijing Normal University, Beijing 100875, PR China
a r t i c l e
i n f o
Article history: Received 2 July 2011 Received in revised form 6 September 2011 Accepted 7 September 2011 Available online 12 September 2011 Keywords: Aerobic granule 2,4-dichlorophenoxyacetic acid Bioaugmentation Microbial community
a b s t r a c t Development of 2,4-dichlorophenoxyacetic acid (2,4-D) degrading aerobic granular sludge was conducted in two sequencing batch reactors (SBR) with one bioaugmented with a plasmid pJP4 donor strain Pseudomonas putida SM1443 and the other as a control. Half-matured aerobic granules pre-grown on glucose were used as the starting seeds and a two-stage operation strategy was applied. Granules capable of utilizing 2,4-D (about 500 mg/L) as the sole carbon source was successfully cultivated in both reactors. Gene-augmentation resulted in the enhancement of 2,4-D degradation rates by the percentage of 65–135% for the granules on Day 18, and 6–24% for the granules on Day 105. Transconjugants receiving plasmid pJP4 were established in the granule microbial community after bioaugmentation and persisted till the end of operation. Compared with the control granules, the granules in the bioaugmented reactor demonstrated a better settling ability, larger size, more abundant microbial diversity and stronger tolerance to 2,4-D. The finally obtained granules in the bioaugmented and control reactor had a granule size of around 600 m and 500 m, a Shannon–Weaver diversity index (H) of 0.96 and 0.55, respectively. A shift in microbial community was found during the granulation process. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Aerobic sludge granulation is an innovative cell immobilization technology in biological wastewater treatment. Compared with conventional activated sludge, aerobic granular sludge demonstrates the advantages like good settling ability, high biomass retention, and high tolerance to medium toxicity. Aerobic granules are formed through the process of microbial self-aggregation under specific selective pressures such as hydraulic shearing forces, short settling time and rich-famine alternative nutrition status. Aerobic granules are generally easily formed in high strength wastewater comprising easily biodegradable substrates like glucose, acetate, sucrose and ethanol [1–4]. In recent years, more efforts have been directed at cultivation of sludge granules for recalcitrant pollutants removal. Several researches have explored the cultivation of aerobic granules with toxic organic compounds such as chlorophenol, phenol, chloroanilines and tert-butyl alcohol [5–9]. For most of these studies, the toxic compounds degrading aerobic granules were cultivated with a mixture of the target compounds and some easily degradable compounds. Up to date, only a few studies have
∗ Corresponding author. Tel.: +86 10 58802374; fax: +86 10 58802374. E-mail addresses:
[email protected],
[email protected] (X.-c. Quan). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.023
reported the successful cultivation of aerobic granules with hazardous compounds as the sole carbon source [6]. Gene-augmentation, as an important mean of bioaugmentation, has often been used at polluted sites or in biotreatment systems aimed to achieve a microbial community with desired functions. Gene-augmentation works through mobile genetic elements (MGEs) horizontal transfer from inoculated donor strains to recipient microorganisms. This technology has been well investigated in biofilm reactors because gene horizontal transfer generally occurs with a high frequency at the site having a dense microbial population [10–14]. Aerobic granules, as an important form of microbial aggregation, have the potential to become another “hot spot” of gene horizontal transfer. A few studies have explored the feasibility of cultivation of aerobic granules through gene-augmentation. Nancharaiah et al. [15] first studied the bioaugmentation of aerobic granules with a strain Pseudomonas putida carrying a TOL plasmid, and found a significant increase in the degradation of benzyl alcohol in the bioaugmented system. Our previous research explored the possibility of cultivating 2,4-dichlorophenoxyacetic acid (2,4D) degrading aerobic granules through inoculating a plasmid pJP4 donor strain. Results showed that bioaugmentation accelerated the establishment of 2,4-D degradation ability and the finally obtained granules were able to degrade 2,4-D in the presence of glucose [16]. Although the aforementioned researches have demonstrated
X.-c. Quan et al. / Journal of Hazardous Materials 196 (2011) 278–286
the possibility of developing aerobic granules for toxic compounds degradation through MGEs based gene-augmentation, the effects of gene-augmentation on the formation, characteristics and microbial community of microbial granules under long term operation has not been fully investigated. 2,4-D, as one of the most commonly used phenoxy acid herbicides in agriculture and gardening, has created potential risks on public health and ecosystem. In this study, 2,4-D was chosen as the target compound. A genetically engineered microorganism P. putida SM1443, which carries a conjugative plasmid pJP4 responsible for 2,4-D degradation, was used as the donor strain. Development of 2,4-D degrading granules was investigated in two sequencing batch reactors (SBR) with one bioaugmented with the pJP4 donor strain P. putida SM1443 and the other not bioaugmented as control. Half-matured granules pre-grown on glucose were used as the starting seeds. This research aimed to reveal the effects of gene-augmentation on aerobic granules formation, pollutants removal, 2,4-D biodegradation kinetics, granule morphology, microbial communities and diversities during the granulation process. This work will be helpful for further understanding the effects of gene-augmentation and promoting its application in the development of specific aerobic granules. 2. Materials and methods 2.1. Reactor operation Two column type reactors each having a working volume of 1.7 L (5 cm internal diameter and 87 cm total height) were established. Both reactors were seeded by half-matured aerobic granular sludge, which was previously cultivated with glucose as the main substrate and had a mean diameter of approximately 500 m. A culture of the strain P. putida SM1443 carrying conjugative plasmid pJP4 was inoculated to one reactor at an inoculation ratio of 10%. The other reactor was not bioaugmented as a control. As the strain P. putida SM1443 carries a dsRed tagged pJP4 plasmid and a chromosomally labeled gfp gene (donated by Prof. Stephan Bathe), it expresses a constitutive gfp fluorescence but no dsRed fluorescence due to repression by a chromosomally encoded lac-repressor [17]. The transconjugants receiving plasmids could express dsRed fluorescence. The donor strain was able to grow in mineral salt medium containing 5 mM 2,4-D and 1 mM NH4 Cl [17]. Both reactors had an initial biomass concentration of 9 g mixed liquor suspended solid (MLSS)/L. The whole granulation process could be divided into two stages (Stage I and Stage II). Stage I (Days 1–107) was operated with mixed substrates of glucose and 2,4-D, with 2,4D stepwise increased from 14 to 496 mg/L and Chemical Oxygen Demand (COD) ranged 597–1171 mg/L. Stage II (Days 108–220) was operated with wastewater containing the sole carbon source of 2,4D (approximately 500 mg/L). The two reactors were operated in a sequencing batch mode with a cycle duration of 6 h (Days 1–121) and 4 h (Days 122–220). Each cycle consisted of 5 min fill, 345 or 225 min aeration, 5 min settle and 5 min draw. Fine air bubbles for aeration were supplied through a dispenser installed at the reactor bottom at an air-flow rate of 3 L/min. The volume exchange ratio was set at 50%. Temperature was maintained at about 20 ◦ C and pH was controlled at 7.0–8.0. The other components of the synthetic wastewater were described by Quan et al. [16]. 2.2. Biodegradation kinetics of 2,4-D by the aerobic granules Some aerobic granule samples were withdrawn from the two reactors on Days 18 and 105. Degradation to 2,4-D by those granules at different initial 2,4-D concentrations (41–713 mg/L) were investigated in a small reaction volume of 100 mL.
279
Reactions were performed at 25 ◦ C and 180 rpm. The specific 2,4-D degradation rates were calculated from 2,4-D degradation curves. Kinetic analysis of the degradation data was performed on the basis of Haldane kinetics model for an inhibitory substrate, V = Vmax S/[Ks + S + (S2 /Ki )], where V and Vmax are the specific and maximum specific substrate degradation rates (mg 2,4-D/(g of volatile suspended solid (VSS)·h), respectively; and S, Ks and Ki are the substrate concentration, half-saturation constant and inhibition constant (mg of 2,4-D/L), respectively. 2.3. Microbial community analysis by denaturing gradient gel electrophoresis (DGGE) Genomic DNA was extracted from the granule samples withdrawn at different operation times using the EZ-10 Spin Column Bacterial Genomic DNA MiniPreps Kits (Bio Basic Inc., Canada). Bacterial 16S rRNA fragments of the granule samples were amplified by Polymerase Chain Reaction (PCR) using the primers 341F-GC and 907 R [18,19]. One PCR reaction (50 L) contained: Taq-DNApolymerase (5 U/L), 0.25 L; GC buffer I (Mg2+ Plus), 25 L; dNTP mixture (each 2.5 mM), 4 L; DNA template, 1 L; primer 341FGC (10 M), 2 L; primer 907R (10 M), 2 L; and sterilized MilliQ water, 15.75 L. Amplification was performed with touchdown PCR under the following conditions: an initial denaturing step at 95 ◦ C for 7 min; then 8 cycles of denaturing at 94 ◦ C for 30 s, annealing at 63–56 ◦ C for 1 min (decreasing by 1 ◦ C each cycle) and extension at 72 ◦ C for 90 s; followed by 25 cycles of 94 ◦ C for 30 s, 56 ◦ C for 1 min, 72 ◦ C for 90 s; final extension at 72 ◦ C for 7 min and then kept at 4 ◦ C. The DGGE was performed using a DCode universal mutation detection system (Bio-Rad, USA). PCR products (15 L) were run on 6% acrylamide gels with a denaturing gradient of 35–55%. Electrophoresis was performed at 120 V for 10 h at 60 ◦ C. Gels were stained with SYBR Green I and photographed using a GEL imaging system (VILBER INFINITY 3000, France). Bands of interest were excised from the gels and DNA was recovered from the target bands for sequence determination. The isolated sequences were compared with 16S rRNA sequences obtained via BLAST searches of the National Center for Biotechnology Information database (http://blast.ncbi.nlm.nih.gov). The scanned gels containing DNA band profiles were analyzed using quantity One 1-D analysis software. The Shannon–Weaver index of species diversity (H) was calculated to evaluate the microbial diversity according to the following equation: S
H=−
pi log(pi )
(1)
i=1
where pi is the proportion of band i in the DGGE profile and s is the total number of the bands. 2.4. Analytical methods The effluent water samples withdrawn from the reactors were filtered with 0.22 m filter prior to determining the concentrations of 2,4-D and COD. 2,4-D was analyzed with a high performance liquid chromatograph (HPLC, Waters 1525, USA) equipped with a UV–vis detector and a C18 reverse-phase column (250 × 4.6 mm). The detection wavelength used was 285 nm and the mobile phase was a mixture of methanol, water and acetic acid (85:13:2). The detection limit of 2,4-D was 0.01 mg/L. COD, MLSS, volatile suspended solids (VSS) and the sludge volume index (SVI) were measured according to Standard Methods [20]. Granule size was measured by a laser particle size analysis system with a range of 0.02–2000 m (MasterSizer 2000, Malvern Instruments, UK). Granule morphology was observed by a microscope and the microbial composition was observed with a scanning
100
1200
80
1000
Influent COD Effluent COD (Bioaug) 60 Effluent COD (Control)
800 600
40
Removal(Bioaug) Removal (Control)
400
20
200 0
0
20
40
60
80
100
120
140
160
180
2,4-D influent Conc. (mg/L)
600
0 200 220 100 80
500 400
60
300
Influent Efluent (Bioaug) Effluent (Control)
200 100
40 20 0
0 0
20
40
60
80
100
120
140
160 180
200
220
Time (d) Fig. 1. Removal of COD and 2,4-D in the granule reactors during the whole operation.
electron microscope (Quanta200, FEI, Netherland). The granule samples were gently fixed with 5% glutaraldehyde and 1% OsO4 , and then dehydrated by a graded series of ethanol solutions (50%, 70%, 80%, 90%, 95% and 100%). The dehydrated granules were dried with a liquid CO2 critical point dryer and observed in SEM. The distribution of transconjugants in the granules was examined by a confocal laser scanning microscope (CLSM) (Carl Zeiss LSM 510, Jena, Germany). A 488 nm laser line with a 505 nm long-pass emission filter and a 543 nm laser line with a 560–615 band-pass emission filter were used to detect fluorescence emitted by gfp and dsRed, respectively.
16
A
14
Day 18
12 10 8 6 4 2 0
0
50
Specific degradation rate (mg2,4-D/gVSS.h)
Stage II
Stage I
COD Removal percentage(%)
COD Conc. (mg/L)
1400
Specific degradation rate (mg 2,4-D/gVSS.h)
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2,4-D effluent Conc. (mg/L)
280
100
B
200
300
Day 105
400
500
600
700
800
Granule of bioa ugmented reactor Granule of control reactor Fitted line
40
30
20
10
0
0
100
200
300
400
500
600
2,4-D Conc.(mg/L) Fig. 2. Specific 2,4-D degradation rates for the granules from the bioaugmented reactor and its control (a) Day 18; (b) Day 105.
showed that nearly 100% removal of 2,4-D and about 90% removal of COD (data not shown) could be achieved at the end of a cycle time, indicating that 2,4-D could be mineralized by the microbial granules.
3. Results 3.1. Reactor performance The performance of the two reactors based on the removal of 2,4-D and COD is presented in Fig. 1. When a mixture of glucose and 2,4-D was fed to the reactor during Stage I, 2,4-D at the influent concentration of 14 mg/L was degraded to below detection limit within 5 days for both reactors, and a nearly complete removal of 2,4-D and more than 90% removal of COD were maintained till the end of this operation stage even under the enhancement of 2,4-D loading. During Stage II, when glucose was not added and 2,4-D (at about 500 mg/L) served as the sole carbon source in influent, both reactors maintained a stable removal of pollutants with 2,4-D and COD average removal percentages at 99% and 87%, respectively. On the whole, the bioaugmented reactor and the control demonstrated a similar performance on pollutants removal during the granulation process (Fig. 1). Pollutants removal profiles for a typical cycle period
3.2. Biodegradation kinetics of 2,4-D Biodegradation kinetics of 2,4-D by the granules sampled on Days 18 and 105 were investigated through batch degradation experiments. Specific degradation rates achieved at different 2,4D initial concentrations are presented in Fig. 2. The granules in the bioaugmented reactor demonstrated higher 2,4-D degradation rates than that in the control reactor. The difference in degradation rates was more significant for the granules sampled on Day 18 than that on Day 105, indicating that bioaugmentation exhibited more benefits for the granules at the initial operation stage. The maximum specific degradation rates demonstrated by the aerobic granules in the bioaugmented reactor and the control were 13.52 and 6.63 mg 2,4-D/gVSS·d for Day 18, 36.49 and 30.46 mg 2,4-
Table 1 Degradation kinetic parameters obtained through fitting data to Haldane equation. Granule type
Day
Vmax (mg2,4-D/gVSS·h)
Ks (mg2,4-D/L)
Ki (mg2,4-D/L)
R2
Bioaugmented reactor
18 105
29.9 35.9
256.7 13.2
600.5 13462.9
0.9784 0.9841
Control reactor
18 105
10.4 33.6
213.3 11.9
2986.6 4612.5
0.9993 0.9977
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60 MLSS SVI
SVI (mL/g)
50 40
8
30
6
20
4
10
2
0
B
0
20
40
60
80
100
120
140
160
180
200
0
800 Granule size
700
Granule size (µm)
10
MLSS (g/L)
A
600 500 400 300 100
0
20
40
60
80
100
120
SVI (mL/g)
180
200
40
4
20
2
20
40
60
80
100
120
140
160
3.4. Granule morphology
8 6
0
180
200
0
900 Granule size
800 700 600 500 400 300 0
20
40
60
80
100 120 Time (d)
140
160
180
at 20–30 mL/g and biomass reached around 6 g MLSS/L at end of operation; meanwhile, granule size also increased linearly to a peak value of 721 m on Day 167 and then decreased slightly and finally stabilized at about 600 m. For the granules in the control reactor, granules settling ability remained relatively stable with SVI at about 30 mL/g during Stage I except an ephemeral deterioration at the beginning of operation (Days 1–10); meanwhile, biomass concentration varied in the range of 4–8 g MLSS/L. Different to the granules in the bioaugmented reactor, granule size for the granules in the control reactor was not significantly influenced by the feeding of 2,4-D during Stage I, as it first increased to 770–820 m during Days 10–60 and then decreased and stabilized at 620–660 m during Days 70–105. However, at the beginning of Stage II, sludge settling ability deteriorated with a sharp increase of SVI to 73 mL/g and decrease of biomass to below 2 g MLSS/L on Day 110; sludge granule size also dropped down correspondingly to 300 m on Day 145. After a period of adaptation, the granules recovered settling ability and granule size began to increase. The matured granules had a granule size of about 500 m and a SVI value of about 30 mL/g.
10
60
0
Granule size (mm)
160
MLSS SVI
80
D
140
MLSS (g/L)
C
281
200
Fig. 3. Variation of biomass, SVI and granule size during the granulation process for the bioaugmented reactor (A and B) and its control (C and D).
D/gVSS·d for Day 105, respectively. Gene-augmentation resulted in the enhancement of 2,4-D degradation rates by the percentage of 65–135% for the granules on Day 18, and 6–24% for the granules on Day 105. The Haldane equation was used to model the degradation data and the obtained kinetic parameters are presented in Table 1. 3.3. Evolution of biomass and formation of granules The evolution of biomass, granule size and sludge settling ability during the granulation process are presented in Fig. 3. For the granules in the bioaugmented reactor, sludge settling ability deteriorated due to the feeding of 2,4-D at the beginning of operation, with the SVI values sharply increased from 25 mL/g to 55 mL/g accompanied by a fast drop of biomass from 11 g MLSS/L to 4 g MLSS/L. Granule size increased first from 560 m to 700 m, and then declined gradually to 300 m on Day 105. When the influent switched to the 2,4-D sole carbon source wastewater, the granules recovered settling ability gradually with SVI finally stabilized
Microscopic observation revealed the dynamic changes of granule morphology in the bioaugmented reactor during the whole operation (Fig. 4). The half-matured granular sludge used as the starting seeds had an irregular shape and loose structure with mean diameter of about 550 m (Fig. 4(a)). Some black dots or areas appeared within the granules at the beginning of operation due to the toxicity of 2,4-D (Fig. 4(b)). Those sludge granules further broke into small particles of about 300 m on Day 105 (Fig. 4(c)). When the reactor was operated with the wastewater of 2,4-D as sole carbon source (Stage II), small granular particles began to aggregate and grew; meanwhile, some “super” sludge flakes with a diameter of 3–4 mm and an obvious granule color change from yellow to black was observed (Fig. 4(d and e)). Those “super” sludge flakes involved several sludge particles embedded by a dense layer of extracellular polymeric substances (EPS), and finally disintegrated under the function of hydraulic washing and formed matured aerobic granules (Fig. 4(f)). Detailed microstructures of the granules were examined using a SEM (Fig. 5). Both types of aerobic granules showed a compact structure and the domination of non-filamentous bacteria that were tightly linked and covered with extrapolysaccarides. The granules in the bioaugmented reactor were composed of rod, coccus bacteria and a few filamentous bacteria, while that in the control reactor showed the predominance of coccus. Filamentous bacteria are generally regarded as the backbone of sludge granules [21], whereas, only a few filamentous bacteria was found in the granules of this study, which indicated that filamentous structure in not necessary for the formation of the 2,4-D degrading granule. The strong toxicity of 2,4-D and a high DO condition applied in this study may inhibit the growth of the growth filamentous bacteria. The aerobic granules in the bioaugmented reactor were also examined with a CLSM to investigate the survival status of the transconjugants produced through receiving the pJP4 plasmid. A large quantity of bacteria emitting red fluorescence was found in the granules throughout the experiment (data not shown), which indicated that pJP4 successfully transferred to granule microbes and transconjugants had become an important member of the granule microbial community through subsequent growth and proliferation. As the fluorescent cells existed both on surface and inner part of the granules,
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Fig. 4. Images of the sludge in the bioaugmented reactor at different operation times: (a) seeded half-matured granular sludge on Day 0, (b) Day 44, (c) Day 105, (d) Day 134, (e) Day 167, (f) Day 220.
their number was hard to measure accurately. Other molecular technologies like Real-Time PCR are required for further investigation the dynamic changes of transconjugants number. 3.5. Dynamic changes of microbial community Well-resolved DGGE band profiles were obtained for both types of granules during the granulation process (Fig. 6). A gradual succession in granule microbial community was observed for both types of granules. The granule seeds showed four dominant bands (bands 1–4), with two belong to the member of Actinobacteria and two belong to ˇ-proteobacteria. With the feeding a mixture of glucose and 2,4-D (Days 1–107), bands 1–4 disappeared, some new bands like bands 5–9 appeared. When the reactors were operated with the 2,4-D sole carbon source wastewater, bands 5, 6, 7, 9 disappeared and bands 10–15 appeared. Both reactors showed an altered microbial community during different operation stages. Some dominant genes bands from the DGGE profile were sequenced and their Blast results were presented in Table 2. The finally obtained granules in the bioaugmented reactor and the control shared four dominant species of
Pseudoxanthomonas taiwanensis (band10), Uncultured Sphingobium sp. (band 11), Novosphingobium sp. TYA-1(band 14) and uncultured ˇ-proteobacteria bacterium QEDR3DA12 (band 13). Besides the four common strains, the granules in the bioaugmented reactor also showed the presence of Aquincola tertiaricarbonis L108 (band 8) and Sphingobium sp. JW16.2a (band 12), and the granules in the control reactor found the existence of uncultured Xanthomonas sp. (band 15). The Shannon–Weaver index of species diversity (H) value was calculated based on DGGE band patterns to evaluate the apparent diversity of a microbial community (Fig. 7). The H value is influenced by both the number and abundance of species. A high H value signifies high species diversity [22]. For the granules in the bioaugmented reactor, the H index increased slightly from 0.75 to 0.91 during Days 1–105, and then fluctuated in the range of 0.78–0.96 for the subsequent operation days. For the granules in the control reactor, index H increased from 0.78 to 0.88 first and then declined gradually and finally stabilized at about 0.55. These data indicated that the granules in the bioaugmented reactor showed more microbial species than that in the control.
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283
Fig. 5. SEM images of granules in the bioaugmented reactor (a and b) and the control reactor(c and d) on Day 160 (a × 300, b × 5000, c × 250, d × 5000).
4. Discussion Chlorophenoxy herbicides are reported to have toxic and metabolic uncoupling effects on microbial cell growth [23,24]. Therefore, cultivation of sludge granules degrading these compounds is much more difficult than cultivation of sludge granules degrading common pollutants. To promote the fast formation of 2,4-D degrading sludge granules, half-matured sludge granules pre-cultured with glucose were used as the starting seeds, and a specific strain P. putida SM1443 carrying a conjugative plasmid
pJP4 was bioaugmented. Half-matured granules were selected as the starting seeds because of its strong tolerance to inhibitory compounds. In addition, half-matured granules are more suitable than completely matured granules to incorporate bioaugmented strains because of their less compact structure. To promote the steadily transform of glucose-fed granules to 2,4-D degrading granules, glucose was added as a benign substrate and 2,4-D was increased gradually at the initial operation stage. Bioaugmentation is often applied in biotreatment systems to promote system start-up. In this study, it took 5 days for both the
Table 2 Sequence analysis and species identification of selected DGGE bands for the aerobic granules (the band numbers are shown in the schematic DGGE profiles in Fig. 7). Band no. 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15
Closest relatives
Ref. accession no.
Leifsonia sp. Microbacterium pumilum Uncultured Alicycliphilus sp. Uncultured Comamonas sp. Uncultured Alphaproteobacteria bacterium 16S rRNA gene from clone QEDN1CD03 Hydrogenophaga sp. Gsoil 1545 Uncultured bacterium partial 16S rRNA gene, clone H2SRC126X Aquincola tertiaricarbonis L108 Lysobacter sp. T-15 Pseudoxanthomonas taiwanensis Uncultured Sphingobium sp. Sphingobium sp. JW16.2a Uncultured Betaproteobacteria bacterium from clone QEDR3DA12 Novosphingobium sp. TYA-1 Uncultured Xanthomonas sp.
HQ222274.1 AB234027.1 GQ891858.1 FJ439050.1 CU927486.1 AB271047.1 FM174360.1 DQ656489.1 AB490175.1 AB210278.1 HM438584.1 FN556564.1 CU922449.1 AB491194.1 EU381114.1
Identity
Phylogenetic division
Accession no.
97% 99% 99% 98% 99%
Actinobacteria Actinobacteria -proteobacteria -proteobacteria ␣-proteobacteria
JF804658 JF804659 JF804660 JF804661 JF804662
99% 99% 99% 99% 98% 96% 97% 96% 98% 98%
-proteobacteria unknown -proteobacteria ␥-proteobacteria ␥-proteobacteria ␣-proteobacteria ␣-proteobacteria -proteobacteria ␣-proteobacteria ␥-proteobacteria
JF804663 JF804664 JF804673 JF804665 JF804666 JF804672 JF804667 JF804669 JF804670 JF804674
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Fig. 6. DGGE fingerprints for the granule samples in the bioaugmented reactor and the control. Lanes are labeled with letter SG (seed granule), GEM (genetically engineered microorganism Pseudomonas putida SM1443) and the number indicates sampling period in days.
bioaugmented and control reactors to achieve a stable removal of 2,4-D and COD. Bioaugmentation failed to reduce the start-up period. This may be attributed to the low concentration of 2,4D (14 mg/L) applied at the beginning of operation and a strong tolerance of sludge granules to toxic organic compounds [7,25]. When influent 2,4-D increased to 375 mg/L on Day 25, both granule reactors maintained a nearly complete removal of 2,4-D and 90% removal of COD. The start-up time for the two granule reactors was much less than that for activated sludge based reactors reported previously. Mangat and Elefsiniotis [26] investigated the degradation of 2,4-D in three SBR reactors seeded with phenol degrading
1.0 0.9 0.8
Diversity index H
0.7 0.6 0.5
Bioaugmented reactor Control reactor
0.4 0.3 0.2 0.1 0.0
0
50
100
150
200
250
Time (d) Fig. 7. Bacterial species diversity using the Shannon–weaver diversity (H) calculated from the DGGE band profiles for the granules from the bioaugmented reactor and its control.
microorganisms, activated sludge and their mixture. A long acclimation period (about 4 months) was required for those reactors to establish a stable degradation ability for 2,4-D at the concentration of about 100 mg/L. Orhon et al. [27] indicated that activated sludge needed a acclimation period of 35–45 days to degrade 2,4-D at about 100–400 mg/L. Although no obvious effects of bioaugmentation were found in pollutants removal during the granulation process, batch degradation experiment conducted on Days 18 and 105 revealed a considerable difference in 2,4-D degradation rates between the two types of granules. The granules in the bioaugmented reactor demonstrated much higher 2,4-D degradation rates than that in the control. This may be due to the production of transconjugants through plasmid horizontal transfer and subsequently cell proliferation, as a large number of transconjugants emitting red fluorescence were found in the granules from the bioaugmented reactor but not found in the control during the whole operation period. The number of transconjugants in the matured 2,4-D degrading granules was approximately 2 × 105 cell/L. The donor strain P. putida SM1443 disappeared 8 days after bioaugmentation as no green fluorescence cells could be detected thereafter. The methodology employed in this study failed to exactly discriminate the relative contribution of the donor cells, transconjugant cells and indigenous bacteria in degradation, so further researches based on advanced molecular biotechnology such as real-time PCR are required. Biodegradation kinetics of 2,4-D by granules showed that both types granules demonstrated high degradation rates over a wide range of 2,4-D concentration (100–700 mg/L). This strong resistance to high concentrations of 2,4-D could be attributed to the compact granule structure which created a diffusion barrier and made the 2,4-D concentration inside granules lower than that in the bulk liquid. Similar phenomenon was observed for granules degrading other recalcitrant pollutants like phenol, p-Nitrophenol
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and 2,4-dichlorophenol [7,8,28]. The 2,4-D specific degradation rates achieved by the granules in this study were higher than that by activated sludge. For example, Mangat and Elefsiniotis [26] investigated the biodegradation of 2,4-D in a SBR seeded with activated sludge and obtained the specific removal rates of 1.46–1.75 mg 2,4D/g VSS·h. Selective pressure is an important factor influencing the effectiveness of bioaugmentation. It has been reported that a strong selective pressure of target recalcitrant compounds existing alone and persistently was more beneficial for exhibiting effectiveness of gene augmentation [14,15]. For example, Quan et al. [14] conducted two gene augmentation experiments in a biofilm reactor under different substrates conditions aimed to enhance the removal of 2,4-D. No enhancement in 2,4-D removal was observed during start-up period for the bioreactor operated with mixed carbon sources of glucose and 2,4-D, but a significant enhancement was found for the reactor operated with the sole carbon source of 2,4-D. Similar phenomena were also found in this study. The granule size for the bioaugmented reactor decreased gradually during Stage I (mixed substrates of 2,4-D and glucose) but step increase during Stage II (2,4-D as the sole carbon source), whereas the granule size for the control reactor showed an opposite variation trend. The granules in the two reactors demonstrated a distinct trend in the changes of granule size. During Stage I, the granule size for the granules in the bioaugmented reactor declined gradually, while that in the control reactor were relatively stable, which indicated that the granules in the bioaugmented reactor was less stable than that in the control when operated with mixed substrates of glucose and 2,4-D. During Stage II, the granule size increased steadily for the bioaugmented reactor while that for the control decreased first and then increased slightly, indicating that the granules in the bioaugmented demonstrated obvious advantages in treating the wastewater with 2,4-D as the sole carbon source. This finding was also corroborated by other researches. Another important feature for the granules in the bioaugmented reactor was more microbial diversity than that in the control. This may be due to the fact that many indigenous bacteria originally having no 2,4-D degradation ability established this ability through receiving the conjugative plasmid pJP4 and thus survived in the 2,4-D containing wastewater. The dominant bacteria in the granules cultivated through bioaugmentation were found to mainly belong to the member of Aquincola tertiaricarbonis, Pseudoxanthomonas taiwanensis, Sphingobium sp. and Novosphingobium sp. Pseudoxanthomonas taiwanensis has been reported to be a thermophilic bacteria isolated from the thermophilic aerobic granular biomass and paper mill slime [29,30]. Novosphingobium sp. TYA-1 was able to degrade bisphenol A and bisphenol F in the rhizosphere sediment [31]. Sphingobium sp. was isolated from soils, sediment and activated sludge and exhibited degradation ability to many aromatic hydrocarbons including herbicides, polyaromatic hydrocarbons and estrogen [32–35]. A. tertiaricarbonis L108 was reported to grow on the fuel oxygenates methyl tert-butyl ether (MTBE), ethyl tert-butyl ether (ETBE) and tert-amyl methyl ether (TAME) [36]. Many 2,4-D-degrading microorganisms have reported to be isolated from agricultural, urban, and industrial soils and sediments, which mainly belong to the subdivision of Proteobacteria such as Rhodoferax, Bulkholderia, Ralstonia, Alcaligenes, Halomonas, Variovarax and Pseudomonas, Bradyrhizobium sp. and Sphingomonas [37–41]. All the identified strains in the finally obtained granules have not been reported to be 2,4-D degrading bacteria before except Sphingobium sp. Above research results indicates that gene-augmentation is a promising strategy for the fast establishment granular sludge possessing specific functions. Compared with the traditional bioaugmentation method, cell augmentation, it shows the advantage of lower requirement for the survival of the added strains in a
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bioaugmented system, and therefore can be regarded as an in situ method of genetic modification [16,42]. In summary, aerobic granules with a mean diameter of 500–600 m and capable of utilizing 2,4-D of about 500 mg/L as the sole carbon source were successfully established in both the bioaugmented reactor and the control. The granules cultivated in the bioaugmented reactor exhibited larger sizes, better settling ability, stronger 2,4-D degradation ability and richer microbial species than that in the control. More researches are required to understand the dynamic changes of transconjugants number and their contribution in 2,4-D degradation in the complex microbial communities. Acknowledgements This research was supported by “National Natural Science Foundation of China” (nos. 50878024 and 21077012). The authors wish to express their gratitude to Prof. Stephan Bathe for the donation of Pseudomonas putida SM1443::gfp2x carrying the plasmid pJP4::dsRed. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.023. References [1] H.P. Fang, H. Liu, T. Zhang, Characterization of a hydrogen-producing granular sludge, Biotechnol. Bioeng. 78 (1) (2002) 44–52. [2] Y. Liu, J.H. Tay, B.Y.P. Moy, Characteristics of aerobic granular sludge in a sequencing batch reactor with variable aeration, Appl. Microbiol. Biotechnol. 71 (2006) 761–766. [3] Y.M. Zheng, H.Q. Yu, S.J. Liu, X.Z. Liu, Formation and instability of aerobic granules under high organic loading conditions, Chemosphere 63 (2006) 1791–1800. [4] J. Wan, M. Sperandio, Possible role of denitrification on aerobic granular sludge formation in sequencing batch reactor, Chemosphere 75 (2009) 220–227. [5] S.T.L. Tay, B.Y.P. Moy, H.L. Jiang, J.H. Tay, Rapid cultivation of stable aerobic phenol-degrading granules using acetate-fed granules as microbial seed, J. biotechnol. 115 (4) (2005) 387–395. [6] S.T.L. Tay, W.Q. Zhuang, J.H. Tay, Start-up, microbial community analysis and formation of aerobic granules in a tert-butyl alcohol degrading sequencing batch reactor, Environ. Sci. Technol. 39 (2005) 5774–5780. [7] S. Yi, W.Q. Zhuang, B. Wu, S.T. Tay, J.H. Tay, Biodegradation of p-Nitrophenol by aerobic granules in a sequencing batch reactor, Environ. Sci. Technol. 40 (2006) 2396–2401. [8] S.G. Wang, X.W. Liu, H.Y. Zhang, W.X. Gong, Aerobic granulation for 2,4dichlorophenol biodegradation in a sequencing batch reactor, Chemosphere 69 (5) (2007) 769–775. [9] L. Zhu, X. Xu, W. Luo, Z. Tan, H. Lin, N. Zhang, A comparative study on the formation and characterization of aerobic 4-chloroaniline-degrading granules in SBR and SABR, Appl. Microbiol. Biotechnol. 79 (2008) 867–874. [10] L.J. Ehlers, E.J. Bouwer, RP4 plasmid transfer among species of Pseudomonas in a biofilm reactor, Water Sci. Technol. 39 (7) (1999) 163–171. [11] M. Hausner, S. Wuertz, High rates of conjugation in bacterial biofilms as determined by quantitative in situ analysis, Appl. Environ. Microbiol. 65 (8) (1999) 3710–3713. [12] A.K. Lilley, M.J. Bailey, The transfer dynamics of Pseudomonas sp. plasmid pQBR11 in biofilms, FEMS Microbiol. Ecol. 42 (2002) 243–250. [13] A.R. Johnsen, N. Kroer, Effects of stress and other environmental factors on horizontal plasmid transfer assessed by direct quantification of discrete transfer events, FEMS Microbiol. Ecol. 59 (2007) 718–728. [14] X.C. Quan, H. Tang, J.Y. Ma, Effects of gene augmentation on the removal of 2,4-dichlorophenoxyacetic acid in a biofilm reactor under different scales and substrate conditions, J. Hazard. Mater. 185 (2–3) (2011) 689–695. [15] V. Nancharaiah, H.M. Joshi, M. Hausner, V.P. Venugopalan, Bioaugmentation of aerobic microbial granules with Pseudomonas putida carrying TOL plasmid, Chemosphere 71 (2008) 30–35. [16] X.C. Quan, H. Tang, W.C. Xiong, Z.F. Yang, Bioaugmentation of aerobic sludge granules with a plasmid donor strain for enhanced degradation of 2,4dichlorophenoxyacetic acid, J. Hazard. Mater. 179 (2010) 136–1142. [17] S. Bathe, T.V.K. Mohan, S. Wuertz, M. Hausner, Bioaugmentation of a sequencing batch biofilm reactor by horizontal gene transfer, Water Sci. Technol. 49 (11–12) (2004) 337–344. [18] G. Muyzer, E.C. De Waal, A.G. Uitterlinden, Profiling of complex microbial populations by denaturing gradient gel electrophoresis analysis of polymerase chain
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[19]
[20]
[21] [22] [23]
[24]
[25] [26]
[27] [28]
[29]
[30] [31]
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Journal of Hazardous Materials 196 (2011) 234–241
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Synthesis of poly(aminopropyl/methyl)silsesquioxane particles as effective Cu(II) and Pb(II) adsorbents Xin Lu, Qiangfeng Yin, Zhong Xin ∗ , Yang Li, Ting Han State Key Laboratory of Chemical Engineering, School of Chemical Engineering, East China University of Science and Technology, P.O. Box 545, Meilong Road 130, Shanghai 200237, People’s Republic of China
a r t i c l e
i n f o
Article history: Received 2 March 2011 Received in revised form 6 September 2011 Accepted 6 September 2011 Available online 10 September 2011 Keywords: Poly(aminopropyl/methyl)silsesquioxane Particle Adsorption Copper Lead
a b s t r a c t Poly(aminopropyl/methyl)silsesquioxane (PAMSQ) particles have been synthesized by a onestep hydrolytic co-condensation process using 3-aminopropyltriethoxysilane (APTES) and methyltrimethoxysilane (MTMS) as precursors in the presence of base catalyst in aqueous medium. The amino functionalities of the particles could be controlled by adjusting the organosilanes feed ratio. The compositions of the amino-functionalized polysilsesquioxanes were confirmed by FT-IR spectroscopy, solid-state 29 Si NMR spectroscopy, and elemental analysis. The strong adsorbability of Cu(II) and Pb(II) ions onto PAMSQ particles was systematically examined. The effect of adsorption time, initial metal ions concentration and pH of solutions was studied to optimize the metal ions adsorbability of PAMSQ particles. The kinetic studies indicated that the adsorption process well fits the pseudo-second-order kinetics. Adsorption phenomena appeared to follow Langmuir isotherm. The PAMSQ particles demonstrate the highest Cu(II) and Pb(II) adsorption capacity of 2.29 mmol/g and 1.31 mmol/g at an initial metal ions concentration of 20 mM, respectively. The PAMSQ particles demonstrate a promising application in the removal of Cu(II) and Pb(II) ions from aqueous solutions. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The pollution of heavy-metal ions has already become a worldwide serious problem that endangers the environment and health of human beings [1,2]. There are a variety of methods that can be used for removing of heavy-metal ions including liquid–liquid extraction, chemical precipitation, metal replacement, ion exchange [3,4], electrolysis [5], and membrane separation [6]. Adsorption using suitable adsorbents has also been widely used for the concentration and retrieval of metal ions. A variety of new adsorbents are currently being explored, including activated carbon [7], chelating resins [8], biosorbents [9], and mesoporous silica [10]. And nowadays, seeking for novel adsorbents for heavy-metal ions with high efficiency and low cost has become the main direction of the research on the adsorption and removal of heavy-metal ions. Modified matrices with amino groups have attracted a great attention for metal uptake. Pearson’s hard–soft, acid–base (HSAB) principle states that hard (Lewis) acids prefer to bind to hard (Lewis) bases and soft (Lewis) acids prefer to bind to soft (Lewis) bases [11]. The RNH2 are considered to be hard Lewis base
∗ Corresponding author. Tel.: +86 2164252972; fax: +86 2164240862. E-mail address:
[email protected] (Z. Xin). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.020
according to Pearson’s HSAB principle and selecting RNH2 as functional groups should create an adsorbent that displays affinity for adsorption of hard Lewis acid, such as copper. Lam and coworkers [12] demonstrated that Cu(II) was selectively removed by NH2 -MCM-41 containing NH2 adsorption sites from binary AgNO3 /Cu(NO3 )2 solutions. Yantasee et al. [13] found aminefunctionalized activated carbon had an affinity for metal ions in decreasing order of Cu2+ > Pb2+ > Ni2+ > Cd2+ and with a saturation loading capacity of 0.86 mmol of Cu/g. I.M. El-Nahhal and co-workers have prepared a series of amino group functionalized polysiloxane-immobilized ligand systems via the sol–gel process and found their application in separation of heavy metal ions from aqueous solution [14–18]. In recent years, there has been intense interest in the preparation and application of functional polysilsesquioxane particles [19–21]. We have reported the strong adsorbability of Ag(I) ions onto poly(3-mercaptopropylsilsesquioxane) (PMPSQ) microspheres recently [22]. Beari et al. [23] have studied the hydrolytic condensation of 3-aminopropyltriethoxysilane (APTES) in aqueous solutions and found the hydrolysis and condensation products of APTES was not precipitated from the solution even after several weeks due to their excellent water solubility. In this paper, we report a one-step synthetic process for synthesizing amino-functionalized polysilsesquioxane having high content of amino groups to develop an efficient adsorbent of
X. Lu et al. / Journal of Hazardous Materials 196 (2011) 234–241
Nomenclature APTES b C0 Ce h k1 k2
3-aminopropyltriethoxysilane Langmuir constant (L/mmol) initial metal ions concentration (mmol/L) equilibrium metal ions concentration (mmol/L) initial adsorption rate (mmol g−1 min−1 ) pseudo-first-order rate constant (min−1 ) pseudo-second-order rate constant (g mmol−1 min−1 ) m mass of the adsorbent (g) MTMS methyltrimethoxysilane PAMSQ poly(aminopropyl/methyl)silsesquioxane PMSQ poly(methylsilsesquioxane) adsorption capacity (mmol/g) qe qm theoretical saturation adsorption capacity (mmol/g) qt adsorption capacity at t (mmol/g) regression coefficient R2 t time (min) the solution volume (L) V
heavy metals. Poly(aminopropyl/methyl)silsesquioxane (PAMSQ) particles were obtained by hydrolytic co-condensation of 3aminopropyltriethoxysilane (APTES) with methyltrimethoxysilane (MTMS) in aqueous medium. The PAMSQ particles have the ability to effectively remove the Cu(II) and Pb(II) ions from the aqueous solution. The effect of adsorption time, initial metal ions concentration, and solution pH was studied by a static adsorption method to optimize the Cu(II) and Pb(II) adsorbability of PAMSQ particles. 2. Experimental 2.1. Materials 3-Aminopropyltriethoxysilane (APTES, ≥98.0%) was purchased from Diamond Advanced Material of Chemical Inc. Methyltrimethoxysilane (MTMS, ≥98.0%) was purchased from Jiangsu Danyang Organosilicon Material Industrial Corporation. Ammonium hydroxide solution (NH4 OH, 25%), copper sulfate pentahydrate and lead nitrate of analytical reagent grade were commercially obtained and used as received.
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constant temperature of 20 ◦ C at 100 rpm. After a desired period of adsorption, the particles were filtered from the solution with millipore filter membrane (0.22 m). The final concentrations of the metal ions in the solution were analyzed by inductively coupled plasma, atomic emission spectrometry (ICP-AES, IRIS 1000, Thermo Elemental). The equilibrium adsorption capacity was calculated from Eq. (1). qe =
(C0 − Ce )V m
(1)
where qe (mmol/g) is the adsorption capacity and C0 (mmol/L) and Ce (mmol/L) are respectively the initial and equilibrium metal concentrations. V (L) is the solution volume, and m (g) is the weight of the adsorbent. 2.4. Measurements Elemental analyses were performed with a Vario EL III elemental analyzer (Elementar Analysen systeme GmbH, Germany). FT-IR spectra were recorded on a iS10 FT-IR spectrophotometer (Nicolet, USA). The samples were mixed with potassium bromide and pressed to a disk to measure the absorption spectrum. High-resolution solid-state 29 Si NMR spectra were measured at room temperature on a Bruker Avance 400 MHz spectrometer (silicon frequency 99.36 MHz) equipped with a Bruker solid-state accessory. Spectra were obtained using a broadband probehead with a 4 mm double air bearing magic-angle spinning assembly. Chemical shifts of silicon atoms in silsesquioxane compounds are referred to using the traditional terminology Tn , where the superscript corresponds to the number of oxygen bridges to other silicon atoms. Thus, an uncondensed monomer was designated T0 and a fully condensed polymer with no residual silanols was assigned as T3 silicon atoms [24]. A field emission SEM (JSM-7401F FE-SEM, JEOL Ltd., Japan) was utilized to study the morphology of the PAMSQ particles. The particles were sputter-coated with gold for SEM observations. Specific surface area data were performed on a Micromeritics ASAP 2020 surface area and porosity analyzer with BET method. TGA was performed on a SDT Q600 (TA Instruments, USA) thermal analyzer at a heating rate of 10 ◦ C min−1 in air. 3. Results and discussion 3.1. Synthesis and properties of PAMSQ particles
2.2. Synthesis of PAMSQ particles by hydrolytic co-condensation process PAMSQ particles were prepared using base catalyzed sol–gel process in aqueous medium. The mixture of APTES and MTMS at different molar ratios was added to 100 mL of water, maintaining 10% weight percentage. Ammonium hydroxide solution (0.16 mL) was added into the above solution. The reaction was continued overnight at room temperature. Then the resulting precipitate was filtrated with millipore filter membrane (0.22 m) and rinsed thoroughly with distilled water and ethanol several times to remove the residual NH4 OH as well as unreacted monomers or oligomers. Finally, the products were dried in vacuum and the PAMSQ particles were attained. 2.3. Adsorption of Cu(II) and Pb(II) onto the PAMSQ particles Batch adsorption experiments were conducted using PAMSQ particles as a adsorbent to adsorb Cu(II) or Pb(II) ions from aqueous single metal ion solutions. The sample pH was adjusted to the desired value with HNO3 or ammonia solution. The batch adsorption experiments were conducted in a shaker bath kept at
Synthesis of PAMSQ particles with controllable amount of aminopropyl functional groups using APTES and MTMS as precursors by hydrolytic co-condensation process were conducted (Scheme 1). Generally, hydrolytic condensation of the organotrimethoxysilanes in water or ethanol–water was quite rapid under basic conditions [23–25]. Initial hydrolysis of the APTES and MTMS resulted in silanol oligomers. Silanol (Si–OH) was very reactive and then condensed to form polysilsesquioxanes in the presence of base catalyst. The amino groups of APTES could increase the pH of solution and accelerates the hydrolytic co-condensation process. Simultaneous hydrolysis of APTES and MTMS led to cocondensation, but the product state was different with the variation of APTES/MTMS molar ratios. If the APTES molar ratio in the precursors was less than 40%, white precipitate appeared. However, the co-condensation products were not precipitated from the solution when the APTES molar ratio in the precursors was above 50% due to the excellent solubility of co-condensed PAMSQ in water. SEM images show morphology of polysilsesquioxanes particles (see Fig. 1). The PMSQ particles prepared from MTMS alone are spherical with a medium size of 2.0 m. The particle aggregation was quite evident for the copolymerized PAMSQ particles
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Scheme 1. Schematic illustration of the synthetic strategy for PAMSQ particles. Note that “OR” can present either silanol (OH) or other silane units
and the particle size decreased with increasing APTES amount in the APTES/MTMS mixtures, as seen in Fig. 1b–d. The results could be due to the excellent solubility of APTES in water, which makes the copolymerized PAMSQ particles to be more hydrophilic. Additionally, the amino groups of APTES raised pH of reaction solution and catalyze the hydrolytic co-condensation reactions of APTES and MTMS. Increase in hydrolysis rates with increasing pH led to a higher nucleation rate, which also resulted in a larger number of particles but a smaller final particle size [24]. Table 1 shows the properties of polysilsesquioxanes particles. Amino group contents of PAMSQ particles determined by element analysis were lower than the theoretical values. The results could be due to the excellent solubility of APTES and its hydrolysis and condensation products in water [23]. These results are consistent with those reported by Liu et al. [20] that copolymerized aminopropyl/phenylsilsesquioxane microparticles synthesized from the hydrolytic co-condensation of APTES and phenyltriethoxysilane (PTES). The specific surface areas of the polysilsesquioxanes particles were evaluated using the BET method as shown in Table 1.
The relatively small BET surface area values were due to complete condensation of polysilsesquioxanes. The influence of amino groups content on Cu(II) and Pb(II) adsorption onto the PAMSQ particles was investigated. As shown in Table 1, the aminopropyl functionalized PAMSQ samples revealed a high affinity towards Cu(II) and Pb(II), and the adsorption capacity of metal ions onto the PAMSQ particles increase with increasing amino groups content. However, the unmodified PMSQ without amino groups adsorbed only small amount of Cu(II) and Pb(II) ions. The results of Table 1 suggest that the mechanism of adsorption involves primarily metal ion complexation by the amino groups. High amino group contents make PAMSQ3 fit as a representative copolymerized product to be used for further studies in adsorption experiment. Solid-state 29 Si NMR spectra are powerful methods for characterizing the chemical structure of polysilsesquioxane frameworks. According to Arkhireeva et al. [26], in the case of the polysilsesquioxane derived from MTMS, resonances at −65.9, −57.1, −48.5 ppm can be assigned to T3 , T2 and T1 species, respectively. And Caravajal et al. [27] reported the 29 Si NMR spectra
Fig. 1. SEM images of polysilsesquioxanes particles: (a) PMSQ, (b) PAMSQ1, (c) PAMSQ2 and (d) PAMSQ3.
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Table 1 Properties of polysilsesquioxanes particles. Samples
APTES in comonomers (mol%)
a
Amino content in PAMSQ (mol%)
BET surface area (m2 /g)
b Cu(II) adsorption capacity (mmol/g)
b Pb(II) adsorption capacity (mmol/g)
PMSQ PAMSQ1 PAMSQ2 PAMSQ3
0 20 30 40
0 15 22 27
4.3 4.8 6.7 5.0
0.13 0.80 1.62 2.25
0.17 0.58 0.90 1.14
a b
The amino content in PAMSQ is determined by elemental analysis of nitrogen content. Initial metal ions concentration: 10 mM, adsorption time: 5 h, adsorbent dose: 2 g/L.
1482 cm−1 (ıN–H ), 3376 cm−1 (N–H ), 2934 and 2970 cm−1 (C–H ). A broad band around 3400 cm−1 could be attributed both to the adsorbed water and to the Si–OH group [29]. The thermal stability of the PMSQ and PAMSQ3 in air was investigated using TGA (Fig. 4). Weight loss in the 100–250 ◦ C range for polysilsesquioxane is probably due to the residual reaction of alkoxysilyl groups [26]. The thermal reduction of polysilsesquioxane in the 250–700 ◦ C range appeared to be mainly due to the decomposition of organic moieties groups [24]. The thermal decomposition temperatures of the PAMSQ3 particles in air, at 5% weight loss and at 10% weight loss are 257 ◦ C and 370 ◦ C, respectively. The TGA result shows that both the PMSQ and PAMSQ3 particles have good thermal stability.
Fig. 2. Solid-state
29
Si NMR spectrum of PAMSQ3 particles.
3.2. Adsorption kinetics of Cu(II) and Pb(II) onto the PAMSQ particles
exhibited major peaks in the regions of −66, −58 and −49 ppm, due to the silicons of the attached CH2 CH2 CH2 NH2 moiety of APTES-modified silica. Fig. 2 shows solid-state 29 Si NMR spectrum of PAMSQ3 particles. There is one large peak at −62.1 ppm and a weak shoulder peak at about −53 ppm assigned to fully condensed T3 and linear T2 species, respectively. The formation of T1 and T0 species is insignificant, suggesting that the co-condensation is quite completed [28]. Fig. 3 shows IR spectra of PMSQ and PAMSQ3particles. PMSQ and PAMSQ3 exhibit well-defined methyl group and Si–O–Si absorption bands at: 2970, ca. 2921–2934 cm−1 (C–H ), 1410 cm−1 (ıC–H in Si–R), 1272 cm−1 (ıC–H in Si–R), ca. 1119–1127, ca. 1032–1036 cm−1 (Si–O–Si ), and 778 cm−1 (Si–C ) [29,30]. Si–O–Si stretching peaks at 1119–1127 cm−1 indicate the presence of cagestructure while adsorption at 1032–1036 cm−1 show that the ordered structure is probable ladderlike or layered [20,30]. The spectrum of PAMSQ3 showed bands due to aminopropyl groups at
The kinetics of adsorption is one of the important characteristics that define the efficiency of adsorption. The effect of contact time on the adsorption of Cu(II) and Pb(II) onto the PAMSQ3 particles is shown in Fig. 5. The kinetic curve showed that the adsorption was rapid for the first 10 min, when the adsorption capacity reaches up to 1.49 mmol/g and 0.58 mmol/g for Cu(II) and Pb(II), respectively, and then slowed gradually. The initial rapid step of metal ions adsorption may be attributed to the physical and reactive adsorption between metal ions and the amino groups on the surface of the PAMSQ particles. However, the subsequent slow step is attributable to the adsorption inside the particles, representing the diffusion of Cu(II) and Pb(II) ions into the inner of the particles over a long period. Experimental results suggest that the amount of metal ions adsorbed increased with increasing adsorption time and reached equilibrium at 300 min for Cu(II) and Pb(II). Hence, in the present study, we used 300 min contact time for further experiments. Pseudo-first-order and pseudo-second-order models [31] were used to test the experimental data and thus elucidate the
Fig. 3. IR spectra of PMSQ and PAMSQ3 particles.
Fig. 4. TGA curves of PMSQ and PAMSQ3 in air.
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Fig. 5. Effect of adsorption time on Cu(II) and Pb(II) adsorption onto PAMSQ3 particles (initial metal ions concentration: 10 mM; adsorbent dose: 2 g/L).
adsorption kinetic process. The Lagergren pseudo-first-order kinetic model, represented as: k1 log(qe − qt ) = log qe − t 2.303
(2)
where qe and qt are the amounts of metal ions adsorbed (mmol/g) at equilibrium and time t, respectively, and k1 (min−1 ) is the pseudo-first-order rate constant. The qe and rate constant k1 were calculated by plotting the log (qe − qt ) vs. t. As seen from Fig. 6, the pseudo-first-order model does not fit the data well. The experimental and calculated qe values, pseudo-first-order rate constants and regression coefficient (R2 ) values are presented in Table 2. The calculated qe values in the pseudo-first-order model were not in agreement with the experimental qe values, suggesting that the adsorption of Cu(II) and Pb(II) does not follow pseudo-firstorder kinetics. In order to find a more reliable description of the adsorption kinetics of Cu(II) and Pb(II) ions onto the particles, a pseudo-second-order kinetic model was applied to the experimental data. The pseudo-second-order equation can be written as: t 1 1 = + t qt qe k2 q2e
(3)
where qe and qt are defined as in the pseudo-first-order kinetic model; k2 is the pseudo-second-order rate constant. The slope and intercept of the linear plot t/qt vs. t in Fig. 7 yielded the values
Fig. 6. Pseudo-first-order kinetic plots for the adsorption of Cu(II) and Pb(II) onto PAMSQ3 particles.
Fig. 7. Pseudo-second-order kinetic plots for the adsorption of Cu(II) and Pb(II) onto PAMSQ3 particles.
of qe and k2 . Additionally, the initial adsorption rate (h) can be determined from k2 and qe values using h = k2 q2e . The regression coefficients (R2 ) and several parameters obtained from the pseudosecond-order kinetic model are also shown in Table 2. As seen from Table 2, the calculated qe values are in good agreement with experimental qe values. Moreover, the obtained R2 values for Cu(II) and Pb(II) adsorption both are above 0.99. Hence, the adsorption kinetics could well be approximated more favorably by pseudosecond-order kinetic model for Cu(II) and Pb(II) onto the PAMSQ particles. The pseudo-second-order model was developed based on the assumption that the determining rate step may be chemisorption promoted by covalent forces through the electron exchange, or valency forces through electrons sharing between adsorbent and adsorbate [31], indicating that the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles is mainly the chemically reactive adsorption. 3.3. Effect of initial metal ions concentration and adsorption isotherm The effect of the initial metal ions concentration on adsorption of Cu(II) and Pb(II) onto the PAMSQ3 particles is shown in Fig. 8. At a lower initial metal ions concentration, abundant aminopropyl groups on the surface of the PAMSQ particles can react with metal ions, resulting in a significantly increased adsorption of Cu(II) and Pb(II). Then the adsorption process gradually becomes slow with increasing initial metal ions concentration.
Fig. 8. Effect of initial metal ions concentration on adsorption of Cu(II) and Pb(II) on PAMSQ3 particles (initial metal ions concentration: 1–20 mM; adsorption time: 5 h; adsorbent dose: 2 g/L).
X. Lu et al. / Journal of Hazardous Materials 196 (2011) 234–241
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Table 2 Kinetic model equations for Cu(II) and Pb(II) adsorption onto the PAMSQ3 particles. Metal ions
Cu(II) Pb(II)
qe (exp.) (mmol/g)
2.25 1.14
Pseudo-first-order
Pseudo-second-order
k1 (min−1 )
qe (cal.) (mmol/g)
R2
k2 (g mmol−1 min−1 )
h (mmol g−1 min−1 )
qe (cal.) (mmol/g)
R2
0.01073 0.01262
1.37 0.85
0.8695 0.8830
0.021 0.028
0.11 0.04
2.30 1.22
0.9915 0.9908
facile method under moderate condition would have a promising application as a cost-effective adsorbent. 3.4. Effect of solution pH on adsorption
Fig. 9. Langmuir plots for Cu(II) and Pb(II) adsorption on the PAMSQ3 particles (initial metal ions concentration: 1–20 mM; adsorption time: 5 h; adsorbent dose: 2 g/L).
Fig. 10 shows the effect of the solution pH on the adsorptions of Cu(II) and Pb(II) by PAMSQ3 particles, respectively. The pH in a range of 2.0–5.0 was chosen to avoid the precipitations of Cu(OH)2 and Pb(OH)2 . The adsorption capacities increased with an increase in solution pH in the pH range of 2.0–5.0 and no adsorption was observed at pH 2.0. This could be attributable to a competitive adsorption between metal ions and H+ ions on PAMSQ3 particles. At low pH value, the adsorption of metal ions is decreased because high concentrations of competitive H+ ions occupy the adsorption sites, whereas the protonated amino groups are deprotonated with increasing pH value, enhancing metal ions adsorbability [15,44]. Therefore, the solution pH around 5.0 could be optimal for the application of the PAMSQ3 particles as efficient Cu(II) and Pb(II) adsorbent. 3.5. Adsorption mechanism of metal ions onto PAMSQ particles
Fig. 9 shows the adsorption isotherms of Cu(II) and Pb(II) by the PAMSQ3 particles at 20 ◦ C. The adsorption data were plotted according to Langmuir equation: Ce 1 Ce = + qe qm qm b
(4)
where qm and b are the characteristic Langmuir parameters. qm is the theoretical saturation adsorption capacity of the monolayer (mmol/g) and b is a constant related to the intensity of adsorption. Plotting Ce /qe against Ce gives straight lines as shown in Fig. 9. Table 3 displays the coefficients of the Langmuir model along with regression coefficients (R2 ). As seen from Table 3, the R2 values for the Langmuir isotherm models were both above 0.99, suggesting that the Langmuir model closely fits the experimental results. The calculated qm values are in good agreement with those experimentally found. PAMSQ3 particles possess a strong capability to adsorb Cu(II) and Pb(II) ions from aqueous solutions, indicating a great potential as a high efficiency absorbent. The variations of metal ions uptake on various adsorbents are associated with adsorbent properties such as structure, functional groups, and specific surface area. Table 4 shows the comparison of the maximum adsorption capacity of PAMSQ for Cu(II) and Pb(II) onto various adsorbents reported in the literature. The results demonstrate that the adsorption capacities of PAMSQ3 particles for Cu(II) and Pb(II) were high when compared to several other adsorbents. Therefore, it could be believed that the PAMSQ3 particles synthesized from common silane coupling agent through a
Sorption is broadly defined as the transferring of ions from the solution phase to the solid phase via various mechanisms such as physical and chemical adsorption, surface precipitation, or solidstate diffusion or fixation [45]. According to hard and soft acids and bases theory of Pearson [11], aminopropyl group functionalized PAMSQ has bonding ability with heavy metal ions such as Cu(II) and Pb(II). The FTIR spectra of PAMSQ3 particles before and after adsorption of Cu(II) ion are shown in Fig. 11. Appearance of a sharp peak at 619 cm−1 after adsorption of Cu(II) on PAMSQ3 is assigned to the stretching vibration of N–Cu bond formed during complexation process [37]. The FTIR results confirm that nitrogen of PAMSQ3 particles are actively participated during the adsorption process through complexation with Cu(II) ion. So the adsorption mechanism of metal ions onto PAMSQ3 particles involves primarily metal ions complexation by the amino groups.
Table 3 Coefficients of Langmuir isotherms. Metal ions
qm (exp.) (mmol/g)
qm (cal.) (mmol/g)
b (L/mmol)
R2
Cu(II) Pb(II)
2.29 1.31
2.26 1.47
11.82 0.48
0.9986 0.9925
Fig. 10. Effect of the pH on adsorption of Cu(II) and Pb(II) on PAMSQ3 particles (initial metal ions concentration: 2.5 mM; adsorption time: 5 h; adsorbent dose: 2 g/L).
240
X. Lu et al. / Journal of Hazardous Materials 196 (2011) 234–241
Table 4 Comparison of maximum adsorption capacity of PAMSQ for Cu(II) and Pb(II) onto various adsorbents reported in the literature. Adsorbents
Adsorption capacity (mg/g)
SBA-15 mesoporous silica with melamine-based dendrimer amines Silica gel chemically modified by triethylenetetraminomethylenephosphonic acid Silica gel modified with 5-amino-1,3,4-thiadiazole-2-thiol Ethylenediaminetriacetic acid functionalized silica–gel 4-Amine-2-mercaptopyrimidine modified silica gel 2-Aminophenylaminopropylpolysiloxane Epichlorohydrin cross-linked xanthate chitosan Aminated polyacrylonitrile nanofibers Porous chitosan monoliths PS-EDTA resin 2-((2-Aminoethylamino)methyl)phenol-functionalized activated carbon Iron oxide coated sewage sludge Ulva lactuca algae Potassium hydroxide treated pine cone powder PAMSQ particles
Cu(II)
Pb(II)
Ref.
126 19.8 1.21 99.4
130 16.8 1.54
[32] [33] [34] [35] [36] [16] [37] [38] [39] [40] [41] [42] [9] [43] This work
80.19 130.91 43.47 116.52 141.8 42.1 12.1 17.3 112 19.22 146
32.1 16.2 42.4 230 26.27 272
Fundamental Research Funds for the Central Universities (Project no. WA1013012), and East China University of Science and Technology (ECUST) through fostering the Undergraduates Innovating Experimentation Project (No. X0807).
References
Fig. 11. FTIR spectra of PAMSQ3 particles (a) before adsorption and (b) after adsorption of Cu(II). Inset: scheme of copper ions binding.
4. Conclusions Amino-functionalized polysilsesquioxane particles have been synthesized by hydrolytic co-condensation using APTES and MTMS as precursors in the presence of base catalyst in aqueous medium. The process is a one-step co-condensation synthetic route where the functionalities of the particles can be easily controlled by changing the organosilanes feed ratio. The results of solid-state NMR spectroscopy, FT-IR analysis, and elemental analysis confirmed the co-condensation between organosilanes. The PAMSQ particles had shown as an efficient adsorbent for the removal of Cu(II) and Pb(II). The adsorption behavior of Cu(II) and Pb(II) onto PAMSQ particles is influenced by the adsorption time, initial concentration of metal ions, and solution pH. The kinetic studies indicated that the adsorption process well fits the pseudo-second-order kinetics with a rapid initial adsorption rate. The experimental data was well fit by the Langmuir isotherm model. The PAMSQ particles demonstrate the highest Cu(II) and Pb(II) adsorption capacity of 2.29 mmol/g and 1.31 mmol/g at an initial metal ions concentration of 20 mM, respectively. Therefore, there are good prospects for the PAMSQ particles in practical applications for the removal of Cu(II) and Pb(II) ions from their aqueous solutions. Acknowledgments This work was financially supported by the National Natural Science Foundation of China (Project no. 21006025), the
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Journal of Hazardous Materials 196 (2011) 242–247
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Decontamination of waters polluted with simazine by sorption on mesoporous metal oxides Veria Addorisio a , Domenico Pirozzi b , Serena Esposito c , Filomena Sannino a,∗ a Dipartimento di Scienze del Suolo, della Pianta, dell’Ambiente e delle Produzioni Animali, Facoltà di Scienze Biotecnologiche, Università di Napoli “Federico II”, Via Università 100, 80055 Portici, Napoli, Italy b Dipartimento di Ingegneria Chimica, Facoltà di Ingegneria, Università degli Studi di Napoli “Federico II”, P.le Tecchio, 80, 80125 Napoli, Italy c Laboratorio Materiali del Dipartimento di Meccanica, Strutture, Ambiente e Territorio, Facoltà di Ingegneria dell’Università di Cassino, Via G. di Biasio 43, I-03043, Cassino (Fr), Italy
a r t i c l e
i n f o
Article history: Received 19 April 2011 Received in revised form 6 September 2011 Accepted 6 September 2011 Available online 12 September 2011 Keywords: Decontamination Sorption Simazine Mesoporous metal oxides Regeneration
a b s t r a c t Two mesoporous metal oxides, Al2 O3 and Fe2 O3 , were evaluated as regards their ability to remove simazine, a highly persistent herbicide of s-triazines, using a batch equilibrium method. The effect of several experimental parameters such as pH, contact time, initial concentration and sorbent dosage on the sorption of the herbicide was investigated. The maximum sorption of simazine on Al2 O3 and Fe2 O3 was observed at pH 6.5 and 3.5, respectively. The different sorption capacities of the two oxides were explained considering a set of factors affecting the sorption process such as the surface area and the porosity. The kinetics of sorption on both oxides was described using a pseudo second-order model. The sorption of simazine on Fe2 O3 was faster in comparison to that observed on Al2 O3 . It was shown that aluminum oxide can be regenerated by incineration, and consequently can be considered for industrial treatment systems designed to mitigate the pesticide pollution in the aquatic environments. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Water pollution by pesticides has been recognized in agricultural areas of the world for many years, and considerable evidences suggest that many water resources are contaminated by organic pesticides. Common agricultural practices, accidental spillage or uncontrolled release of contaminated waters due to washing of pesticide containers or industrial effluents in the environment have resulted in the contamination of air, soils, surface and ground waters, as well as of living organisms. Thus, in order to protect the environment and the human health, it is important to develop new remediation technologies. Currently, sorption is believed to be a simple and effective technique for water and wastewater treatment and its success largely depends on the development of efficient sorbents. Activated carbon [1], clay minerals [2], biomaterials [3], zeolites [4], and some industrial solid wastes [5] have been widely used with varying efficiency. In a wastewater treatment process that utilizes sorption,
∗ Corresponding author at: Dipartimento di Scienze del Suolo, della Pianta, dell’Ambiente e delle Produzioni Animali, Università degli Studi di Napoli Federico II, Via Università 100, 80055 Portici (NA), Italy. Tel.: +39 081 2539187/2539183; fax: +39 081 2539186. E-mail address:
[email protected] (F. Sannino). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.022
the regeneration of the sorbent is crucially important. However, the high costs associated with the regeneration of the sorbents or the necessity of an extraction achieved by acid or alkali solutions represent a serious problem. New sorbents are required to remove organic pollutants in water decontamination processes. An ideal sorbent should have high surface area (i.e. high density of sorption sites), uniformly accessible pores and physical and/or chemical stability [6]. It is believed that the sorption capacity of a sorbent is largely determined by its surface area, which usually increases with decreasing the particle size, although the pore size distribution is also decisive for an optimal sorption process. Therefore, thanks to the introduction of nano-structured oxide materials, the pollutant removal efficiency can be increased dramatically. Mesoporous materials, a class of nanoporous materials, have attracted a lot of attention in both the scientific and the industrial communities since the introduction of well-ordered mesoporous silicas which possess large surface areas and uniform and tunable pore sizes (2–50 nm) [7,8]. The great interest of these materials as adsorbents for environmental remediation is due not only to their high surface area but also to their fast contaminant sorption kinetics. Recent works [9–11] have shown that mesoporous materials can have large adsorption capacity, good selectivity and improved recoverability for the removal of toxic compounds from aqueous solutions.
V. Addorisio et al. / Journal of Hazardous Materials 196 (2011) 242–247
The encouraging results obtained from these studies prompted us to investigate the sorption of simazine (2-chloro-4,6bis(ethylamino)-s-triazine), a basic herbicide belonging to striazine family, on the mesoporous metal oxides. The s-triazines are selective persistent herbicides, widely investigated due to their still large application in forestry and pre- and post-emergence in agricultural soils [12]. Even though these herbicides are now forbidden in some countries, the recalcitrance of s-triazines against chemical and biological degradation has led to their accumulation in the environment [12]. In Italy, the annex number 5 included in the Legislative Declaration 152/2006 on the environment, states the safe limit of atrazine (s-triazine herbicide) in soil that varies from 0.01 to 1.0 mg kg−1 , whereas the limit in waters is 0.3 g L−1 . Simazine is a synthetic s-triazine herbicide used for pre-emergence control of broad-leaf weeds and annual grasses in agricultural and non-crop fields [13,14]. It is the second most commonly detected pesticide in surface and groundwater in the United States, Australia and Europe [15]. Due to the carcinogenic potential of s-triazines, simazine presence in water is of increasing concern [16]. A significant research on the removal of s-triazines by sorption on soils and different organic and inorganic sorbents has been performed [17,18]. Nevertheless, as far as we know, no papers have been published on the sorption capacity of mesoporous oxides towards triazine. Therefore, the objective of this work was to evaluate two commercial metal oxides with mesoporous structure (Al2 O3 and Fe2 O3 ) as regards their ability to remove simazine from aqueous solutions. In view of future applications, the regeneration of these materials is also discussed.
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2.3. Sorption experiments A stock solution of herbicide was prepared by dissolving 2 mg of simazine in 500 mL of KCl 0.03 M (final concentration 20 mol L−1 ). This solution was then kept refrigerated. All sorption experiments were carried out by adding 10 mg of sorbent to 20 mL of simazine solution in glass vials with Teflon caps at a temperature of 20 ◦ C. The samples, after incubation for 24 h in a rotatory shaker (35 rpm), were centrifuged at 7000 rpm for 20 min. The supernatant was analyzed to evaluate the herbicide concentration using high pressure liquid chromatography (HPLC) technique as described below. The amount of simazine sorbed on the oxides was calculated as the difference between the initial quantity of herbicide added and that present at the equilibrium. Blanks of simazine in KCl 0.03 M were analyzed in order to check the pesticide stability and the sorption to vials. Several experiments were carried out to study the effect of different factors affecting the sorption of simazine on Al2 O3 and Fe2 O3 , as summarized below:
2-Chloro-4,6-bis(ethylamino)-1,3,5-triazine (simazine) (Fig. S1 of Supporting Information) was purchased from Sigma–Aldrich Chemical Company (Poole, Dorset, UK; 99.0% purity). All solvents were of HPLC grade (Carlo Erba, Milan, Italy) and were used without further purification. The water used in the preparation of all solutions was obtained from a Millipore Waters Milli-Q water purification system. All other chemicals were obtained from Sigma–Aldrich unless otherwise specified. ␥-Aluminum (Al2 O3 ) and iron(III) (Fe2 O3 ) oxides were purchased from IoliTec Nanomaterials (Denzlingen, Germany; 99.9 and 99.5% purity for Al2 O3 and Fe2 O3 , respectively).
(a) Effect of pH: Experiments were carried out by adding pesticide solutions at fixed concentration (10 mol L−1 ) and different pH values from 3.0 to 7.0. The pH was controlled by the addition of a 0.10 or 0.01 mmol L−1 solution of HCl or KOH. The samples were shaken for 24 h and subsequently, after centrifugation, analyzed as described below. (b) Effect of sorbent amount: The experiments were carried out by adding pesticide solutions at two concentrations (5 and 10 mol L−1 ), at different solid/liquid ratios. Ratios of 0.1, 0.5, 1.0 and 2.0 were obtained by adding 2.0, 10, 20 and 40 mg, respectively, of Al2 O3 or Fe2 O3 to a final volume of 20 mL, at 20 ◦ C. The samples were incubated at pH values of 6.5 (tests with Al2 O3 ) and 3.5 (tests with Fe2 O3 ), for 24 h. (c) Effect of incubation time: Kinetic studies were performed using 10 mol L−1 solutions of simazine at pH 6.5 (tests with Al2 O3 ) and pH 3.5 (tests with Fe2 O3 ). The solutions were stirred for 2.0, 5.0, 20, 40, 60, 90, 120, 320, 960 and 1800 min. (d) Sorption isotherm: Different volumes of a stock solution of herbicide (20 mol L−1 ) were added to each oxide to give an initial simazine concentration ranging from 0.50 to 10.69 mol L−1 . The pH of each solution was kept constant at pH 6.5 (tests with Al2 O3 ) and 3.5 (tests with Fe2 O3 ), by the addition of 0.10 or 0.01 mmol L−1 solutions of HCl or KOH. The samples were incubated for 20 min (tests with Al2 O3 ) and 180 min (tests with Fe2 O3 ); then, after centrifugation, the supernatants were analyzed as described below.
2.2. Chemical and physical analysis of Al2 O3 and Fe2 O3
2.4. Analytical determination
The determination of point of zero charge (pzc) of Al2 O3 and Fe2 O3 was performed according to the methods described by Addorisio et al. [11]. The specific surface area (SSA) of Al2 O3 and Fe2 O3 was calculated by the Brunauer–Emmett–Teller (BET) method [19]. N2 adsorption–desorption isotherms at 77 K were obtained by a Micromeritics Gemini II 2370 apparatus. Before each measurement the sample was degassed at 250 ◦ C for 2 h under N2 flow. Pore volumes were determined from the amounts of adsorbed N2 at P/P◦ = 0.98 (desorption curve), assuming the presence of liquid N2 (density = 0.807 g cm−3 ) in the pores under these conditions. The average values of the pore diameter dp were calculated from the relation: dp = 4V/ABET , where V is total pore volume. The Barrett–Joyner–Halenda (BJH) approach [19] was used to calculate the pore size distribution of the sample using the desorption data.
Simazine was analyzed with an Agilent 1200 Series HPLC apparatus (Wilmington, U.S.A.), equipped with a DAD array and a ChemStation Agilent Software. The procedure of analysis is described in detail in Supporting Information.
2. Materials and methods 2.1. Materials
2.5. Diffuse Reflectance Infrared Fourier Transform Spectroscopy (DRIFTS) analysis The procedure of sample preparation for DRIFTS determinations is reported in detail in Supporting Information. 2.6. Scanning Electron Microscopy (SEM) analysis The SEM analysis of Al2 O3 samples at pH 4.0 and 6.5 was carried out by a FEI Quanta 200 ESEM.
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-1
Simazine sorbed (µmol kg )
Al2O3
Fe2O3
6000 5000 4000 3000 2000 1000 0
3.0
4.0
5.0
pH
6.0
7.0
8.0
Fig. 1. Effect of pH on the sorption of simazine by Al2 O3 and Fe2 O3 at solid/liquid ratio of 0.5.
2.7. Analysis of the data All the experiments were performed in triplicate and the relative standard deviation was in all cases lower than 3%. 3. Results and discussion 3.1. Effect of pH In order to evaluate the optimum pH to be used in the subsequent experiments, sorption tests were carried out to study the effect of pH at a 0.5 solid/liquid ratio, preliminarily found as the optimal value for sorption. The results reported in Fig. 1 show that the greatest sorbed amount of simazine was observed at pH 6.5 when using Al2 O3 , and at pH 3.5 in the tests with Fe2 O3 . In aqueous solution, triazines such as simazine exist in either neutral or protonated form, depending on the pKa of the compound (the pKa of simazine is 1.70) and on the pH of the system. The ring nitrogen atom, located in the 3-position between the electron-rich alkyl-side chains, is the most basic and hence the most likely site of protonation. At low pH values (e.g. 3.0–3.5) the surfaces of Feoxide and of soluble species are strongly protonated [20], so that the most basic triazinic nitrogen (N-3) could easily form coordination bonds with Fe2 O3 deriving from the overlap of nitrogen lone pair electrons and partially filled metal d-orbitals on iron, being this latter a transition metal (d-block element). The adsorption of simazine onto Al2 O3 as a function of pH is quite different. A possible explanation of our findings is that the key role in the sorption of the herbicide is played by the aggregation state of the oxide, which is greatly influenced by the pH of the medium. In particular, as reported in Fig. 2, large aggregates and small particles were observed at pH 4.0 and pH 6.5, respectively. Consequently, the textural properties of Al2 O3 are expected to be modified by the pH. To support the latter hypothesis, a physical analysis on Al2 O3 sample at pH 4.0 and pH 6.5 was performed through the analysis of the relative N2 adsorption–desorption isotherms. As shown in Table 1, the surface area of Al2 O3 sample at pH 4.0 (157 m2 g−1 ) results comparable to that at pH 6.5 (150 m2 g−1 ). Table 1 Physical properties of Al2 O3 at pH 4.0 and pH 6.5. Sample
ABET (m2 g−1 )
Pore volume (cm3 g−1 )
Average dp (nm)
Al2 O3 (pH 4.0) Al2 O3 (pH 6.5)
157 150
0.352 0.643
8.9 17.3
Fig. 2. SEM image of Al2 O3 at 500× magnification at pH 4.0 (a) and at pH 6.5 (b), respectively.
However, more interesting indications can be obtained from a comparison between the pore size distributions (Table 1), obtained by the elaboration of the desorption data by the BJH method. Clearly, the contact with solutions at different pH strongly affects the intrinsic organization of Al2 O3 particles, generating a porosity made of smaller cavities in the case of Al2 O3 sample at pH 4.0. Moreover, the pore volume of the samples at pH 6.5 is much greater than that observed at pH 4.0 (Table 1). These observations are confirmed by the SEM analysis (Fig. 2). Finally, at pH 6.5, the herbicide could give an acid–base reaction with Al2 O3 , that at this pH is present as Al[(H2 O)6 ]3+ . Alternatively, being simazine more nucleophilic than water molecules, it can replace some water molecules in the hexacoordinate complex. Chappell et al. [21] showed the interaction of atrazine with smectite surfaces through hydrogen bonding, and at the same time demonstrated that alkyl tails of the herbicide may interact with hydrophobic nanosites on the smectite basal surfaces. Other studies demonstrated that noncovalent binding forces, cation-, may occur between s-triazine and metallic cation [22]. In the literature, the information about the binding mechanism of triazine herbicides on the oxides is very scarce. Consequently, the explanation of the observed behaviour reported above deserves a close attention.
V. Addorisio et al. / Journal of Hazardous Materials 196 (2011) 242–247
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-1
Simazine sorbed (µmol kg )
Al O 2
1.2 104 1 10
Fe O
3
2
3
4
8000 6000 4000 2000 0
0
1
2
3
4
5
6
7
8
Equilibrium concentration of Simazine ( µmol L ) -1
Fig. 3. Effect of solid/liquid ratio on the sorption of two different concentrations of simazine by Al2 O3 and Fe2 O3 at pH 3.5 and 6.5, respectively.
Alternatively, the sorption kinetic curves were analyzed adopting the pseudo second-order kinetic model:
3.2. Effect of solid/liquid ratio Sorption studies of simazine were carried out using Al2 O3 (at pH 6.5) and Fe2 O3 (at pH 3.5), varying the amount of sorbent and adding two different concentrations of herbicide. The results reported in Fig. 3 show, for both the oxides and regardless of the herbicide concentration, a higher sorption at solid/liquid ratio 0.5. In particular, the amount of sorbed herbicide on Al2 O3 was already significant at the lowest solid/liquid ratio (0.1) and was greatly increased by increasing the amount of oxide. However, in the presence of 20 and 40 mg of oxide, no sorption of simazine was observed. Evidently, the greater the amount of oxide, the greater the resistance to the diffusion in the mesoporous structure, which results in a lower sorption of the herbicide.
3.3. Effect of incubation time The kinetic data were analyzed adopting a pseudo first-order kinetic equation [23]: dq = k1 · (qe − q) dt
(E1)
where qe and q are the amounts of herbicide sorbed (mol kg−1 ) at equilibrium and at time t, respectively, k1 is the rate constant of sorption (min−1 ) and t is the time (min). Integrating with the boundary condition q|t=0 = 0 Eq. (E1), the following expression is obtained: log(qe − q) = log qe −
Ka t 2.303
0.20
(E2)
Al O 2
3
Fe O 2
3
-1
t/q (min kg µmol )
Fig. 5. Sorption isotherm of simazine by Al2 O3 and Fe2 O3 .
0.15
dq = k2 · (qe − q)2 dt
(E3)
where k2 is the rate constant of sorption (kg mol−1 min−1 ). Upon integration with the boundary condition q|t=0 = 0, Eq. (E3) yields the following expression: t 1 t − = q qe k2 · q2e
(E4)
The best model to describe the sorption kinetics data was the pseudo second-order model (i.e. Eq. (E3)), as shown by the linear behaviour of the (t/q) versus time plot (Fig. 4). The corresponding model parameters (qe and k2 ) were estimated with reference to the simazine sorption on Al2 O3 (qe = 6098 mol kg−1 , k2 = 2.85 × 10−5 kg mol−1 min−1 , r2 = 0.99), and on Fe2 O3 (qe = 1695 mol kg−1 , k2 = 9.03 × 10−4 kg mol−1 min−1 , r2 = 0.99). The sorption on Fe2 O3 , reaching the equilibrium after 5 min, was faster in comparison to that pertaining Al2 O3 , showing an equilibrium time of 120 min. Therefore, all the equilibrium determinations were carried out adopting an incubation period of 20 min for Fe2 O3 and 180 min for Al2 O3 . 3.4. Sorption isotherm The sorption isotherms of simazine on Al2 O3 and Fe2 O3 are displayed in Fig. 5. The obtained data were analyzed according to the Freundlich equation: x = Kc 1/N
(E5)
where x is the amount of pesticide sorbed (mol kg−1 ), c is the equilibrium concentration of pesticide (mol L−1 ), K [(mol kg−1 )/(mol L−1 )1/N ] and N (dimensionless) are constants that give estimates of the sorptive capacity and intensity, respectively, according to Giles et al. [24]. The sorption isotherms of simazine on Al2 O3 and Fe2 O3 , shown in Fig. 5, were well-fitted by the linearized form of Freundlich equation (r2 > 0.99) (Table 2). According to the classification of Giles et al.
0.10 Table 2 Freundlich parameters for the sorption of simazine on Al2 O3 and Fe2 O3 .
0.05
Freundlich parameters
0
0
50
100
150
Time (min) Fig. 4. Effect of time on the sorption of simazine by Al2 O3 and Fe2 O3 .
200
Al2 O3 Fe2 O3 a
K (mol kg−1 )/(mol L−1 )1/N
N (dimensionless)
r2 a
168.11 156
0.44 0.56
0.99 0.99
Correlation coefficient.
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Table 3 Comparison of surface area, pore volume and average pore diameter of Al2 O3 and Fe2 O3 samples. Sample Al2 O3 Al2 O3 500 Fe2 O3 Fe2 O3 500
ABET (m2 g−1 ) 195 200 106 33
Pore volume (cm3 g−1 )
Average dp (nm)
0.723 0.770 0.239 0.0650
14.8 14.7 9.2 7.9
[24], the experimental sorption isotherms were of S-type for Al2 O3 and C-type for Fe2 O3 . In particular, at low equilibrium concentrations the sorbed amount of simazine on Fe2 O3 was similar to that detected on Al2 O3 , whereas at concentrations greater than 3 mol L−1 there was a marked difference in the behaviour of the two mesoporous oxides. As a matter of facts, at a 6.0 mol L−1 equilibrium concentration of simazine, the amount of herbicide sorbed on Fe2 O3 was 4000 mol kg−1 , whereas that sorbed on Al2 O3 was ∼8000 mol kg−1 . The S-type isotherm of simazine on Al2 O3 indicates that the presence of molecules of herbicide already sorbed on the surface favours the sorption process by a cooperative effect. This effect can be explained assuming that the molecules already sorbed modify the affinity of the sorption sites towards the molecules present in solution. On the contrary, the C-type isotherm of simazine on Fe2 O3 was characterized by a straight line trend, indicative of a constant partition of the herbicide between solution and sorbent until reaching saturation. The Freundlich constants (K and N) (Table 2) showed that Al2 O3 sorbed the herbicide with a higher sorptive capacity and a lower affinity in comparison to Fe2 O3 . The presence of secondary small pores at the boundary of micropores region in Al2 O3 may affect positively the sorption of small organic molecules such as simazine (0.784 nm), as it is possible that the sorption energy increases in those pores whose dimensions approach the herbicide dimensions (0.7–0.9 nm). In fact, as reported in Fig. S2 of Supporting Information, the pore size distribution of Al2 O3 appears to be bimodal, characterized by two maxima at about 3 nm and at about 15 nm. On the contrary, Fe2 O3 shows a unimodal distribution and most of the N2 volume is adsorbed in the pore size range 6–10 nm [11]. A combination of factors such as the surface area and the porosity concurs significantly to influence the highest sorption capacity of Al2 O3 than Fe2 O3 (see Table 3). Finally, DRIFT analyses were carried out on each metal oxide after the sorption of the herbicide. The DRIFT spectra of Al2 O3 and Fe2 O3 -simazine complexes were recorded and compared with those of simazine and untreated oxides (Fig. S3 of Supporting Information). In particular, in Fig. S3a, the characteristic sorption bands of simazine corresponding to NH stretching (3260 cm−1 ) and C N stretching (1637, 1565 and 1406 cm−1 ) were observed [27]. Fig. S3b and c shows that, after the sorption of the herbicide, the sorption band at 3440 cm−1 , corresponding to –OH stretching of each oxide, was reduced more or less strongly due to a possible coordination interactions simazine-Fe2 O3 and acid–base reactions or replacement of the herbicide with water molecules in the acid hexacoordinate complex [Al(H2 O)6 ]3+ , respectively. In a sorption–desorption study of atrazine and simazine by model soil colloidal components, Celis et al. [25] demonstrated that ferrihydrite does not adsorb triazine herbicides. Enhanced sorption of these herbicides on montmorillonite was measured after increasing the surface acidity of the clay. On the contrary, a carbonrich product (biochar) generated from biomass through pyrolysis sorbed an amount of simazine of ∼2480 mol kg−1 [26].
3.5. Regeneration of Al2 O3 and Fe2 O3 In a wastewater treatment involving a sorption process, the regeneration of the sorbent is crucially important. Nowadays, in many applications, the reuse of the sorbent through regeneration of its sorption properties is an economic necessity. Desorption agents (e.g. sodium hydroxide solution) are commonly used to recover sorbents such as Fe- and Al-based supports [28]. However, the utilization of a desorption agent has some disadvantages because it increases the operating cost, and the waste solution containing NaOH discarded from the regeneration of the sorbent causes environmental pollution. The incineration method could be considered as an alternative way for the regeneration of the sorbents, as it avoids the use of hazardous desorption agents. To assess the feasibility of this option, Al2 O3 and Fe2 O3 were annealed at 500 ◦ C for 1 h; subsequently, to ascertain whether the textural properties were retained, a physical characterization on heat treated oxides was performed through the analysis of the relative N2 adsorption–desorption isotherms. The porosities of Al2 O3 and Fe2 O3 have been previously analyzed by the authors [11]; herein we report a comparison with the physical properties of the heat treated samples. The notation used for the samples is referred to the chemical formula followed by a number indicating the temperature of the heat treatment, i.e. Al2 O3 500 and Fe2 O3 500; the samples before the heat treatment are simply denoted with the chemical formula, as in the text. The perfect correspondence between the isotherms obtained with Al2 O3 500 and Al2 O3 [11], indicated that the mesoporous structure was not damaged by the annealing. The adsorption isotherms were elaborated using the BET method, to obtain the corresponding surface areas reported in Table 3 together with the total pore volume and the estimated average pore diameter. As clearly shown by the data in Table 3, all the textural properties of Al2 O3 are well preserved after the heat treatment. These results drive us to consider the incineration method as an effective option for the regeneration of aluminum oxide. On the contrary, the thermal stability of Fe2 O3 was not comparable to that of Al2 O3 as regards the pore structure. As a matter of facts, the heat treatment strongly altered the textural properties of the sample and a drastic collapse of the surface area was observed (Table 3). To observe to which extent the pore size distribution of iron oxide was modified by the annealing procedure, the desorption data were elaborated by the BJH method. The comparison between the pore size distributions of Fe2 O3 and Fe2 O3 500 samples (Fig. S4 of Supporting Information) indicates beyond doubt that the mesoporous structure was completely destroyed by the heat treatment at 500 ◦ C, making the iron oxide not recoverable by incineration.
4. Conclusions In this study, two metal oxides with mesoporous structure, Al2 O3 and Fe2 O3 , showed different capacities to adsorb simazine, a highly persistent herbicide. In particular, the optimum pH for sorption was found to be 6.5 for Al2 O3 and 3.5 for Fe2 O3 . The different sorption capacities of the two oxides were explained by considering a set of factors significantly concurring to influence the sorption process, such as the surface area and the porosity. The kinetics of sorption was described by a pseudo secondorder model, demonstrating that Fe2 O3 adsorbs simazine faster than Al2 O3 . Finally, we demonstrated that Al2 O3 can be regenerated by incineration and could be considered for industrial treatment systems, to remove effectively simazine from the aquatic environments and eventually to mitigate the pesticide pollution.
V. Addorisio et al. / Journal of Hazardous Materials 196 (2011) 242–247
Supplementary data The chemical formula of simazine and its analytical determination, the Diffuse Reflectance Infrared Fourier Transform Spectroscopy (DRIFTS) analysis, the pore size distribution of Al2 O3 and Fe2 O3 , and the pore size distribution of Fe2 O3 at 500 ◦ C are reported. Acknowledgement This manuscript is contribution DiSSPAPA 248. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.022. References [1] S. Baup, C. Jaffre, D. Wolbert, A. Laplanche, Adsorption of pesticides onto granular activated carbon: determination of surface diffusivities using simple batch experiments, Adsorption 6 (2000) 219–228. [2] F. Bruna, I. Pavlovic, C. Barriga, J. Cornejo, M.A. Ulibarri, Adsorption of pesticides carbetamide and amitron on organohydrotalcite, Appl. Clay Sci. 33 (2006) 116–124. [3] G. Crini, Recent developments in polysaccharide-based materials used as adsorbents in wastewater treatment, Prog. Polym. Sci. 30 (2005) 38–70. [4] S. Wang, Y. Peng, Natural zeolites as effective adsorbents in water and wastewater treatment, Chem. Eng. J. 156 (2010) 11–24. [5] N. Ratola, C. Botelho, A. Alves, The use of pine bark as a natural adsorbent for persistent organic pollutants—study of lindane and heptachlor adsorption, J. Chem. Technol. Biotechnol. 78 (2003) 347–351. [6] H. Yoshitake, T. Yokoi, T. Tatsumi, Adsorption of chromate and arsenate by amino functionalized MCM-41 and SBA-1, Chem. Mater. 14 (2002) 4603–4610. [7] C. Lee (Ed.), Adsorption Science and Technology, World Scientific, Singapore, 2003, pp. 605–609. [8] Y. Kim, C. Kim, I. Choi, S. Rengaraj, J. Yi, Arsenic removal using mesoporous alumina prepared via a templating method, Environ. Sci. Technol. 38 (2004) 924–931. [9] M. Anbia, N. Mohammadi, K. Mohammadi, Fast and efficient mesoporous adsorbents for the separation of toxic compounds from aqueous media, J. Hazard. Mater. 176 (2010) 965–972. [10] P. Wang, I.M.C. Lo, Synthesis of mesoporous magnetic ␥-Fe2 O3 and its application to Cr(VI) removal from contaminated water, Water Res. 43 (2009) 3727–3734.
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Journal of Hazardous Materials 196 (2011) 287–294
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Valorisation of electric arc furnace steel slag as raw material for low energy belite cements R.I. Iacobescu a , D. Koumpouri b , Y. Pontikes c , R. Saban a , G.N. Angelopoulos b,∗ a b c
Department of Materials Science and Engineering, Politehnica University of Bucharest, Splaiul Independentei 313, 060032 Bucharest, Romania Laboratory of Materials and Metallurgy, Department of Chemical Engineering, University of Patras, 26500 Rio, Greece Department of Metallurgy and Materials Engineering, Katholieke Universiteit Leuven, Kasteelpark Arenberg 44 bus 2450, B-3001 Heverlee (Leuven), Belgium
a r t i c l e
i n f o
Article history: Received 5 July 2011 Received in revised form 7 September 2011 Accepted 7 September 2011 Available online 12 September 2011 Keywords: Electric arc furnace slag Belite cement
a b s t r a c t In this paper, the valorisation of electric arc furnace steel slag (EAFS) in the production of low energy belite cements is studied. Three types of clinkers were prepared with 0 wt.% (BC), 5 wt.% (BC5) and 10 wt.% (BC10) EAFS, respectively. The design of the raw mixes was based on the compositional indices lime saturation factor (LSF), alumina ratio (AR) and silica ratio (SR). The clinkering temperature was studied for the range 1280–1400 ◦ C; firing was performed at 1380 ◦ C based on the results regarding free lime and the evolution of microstructure. In order to activate the belite, clinkers were cooled fast by blown air and concurrent crushing. The results demonstrate that the microstructure of the produced clinkers is dominated by belite and alite crystals, with tricalcium aluminate and tetracalcium-alumino-ferrite present as micro-crystalline interstitial phases. The prepared cements presented low early strength development as expected for belite-rich compositions; however the 28-day results were 47.5 MPa, 46.6 MPa and 42.8 MPa for BC, BC5 and BC10, respectively. These values are comparable with OPC CEMI 32.5 N (32.5–52.5 MPa) according to EN 197-1. A fast setting behaviour was also observed, particularly in the case of BC10, whereas soundness did not exceed 1 mm. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Recent years have seen the cement industry growing dynamically with most of the activity taking place in emerging economies. Despite financial turbulence, population growth and the resulting need for housing along with state investments in infrastructure are strong drivers to offset the downturn in cement markets. Globally, cement production increased from 2.568 Mt in 2006 to 3.294 Mt in 2010 [1]. Unavoidably, as with any industrial activity, cement production has its own environmental footprint. Estimations suggest that cement production is responsible for 5–7% of the worldwide CO2 emission [2,3]. If all the greenhouse gases emitted by anthropogenic activities are considered, the cement manufacturing industry contributes about 3% of the total anthropogenic greenhouse gases emissions [2]. This is predominantly the result of the fuels used to generate the required energy, estimated at 0.37 kg/kg clinker, and of the de-carbonation of limestone (CaCO3 ) which takes place during cement production, estimated at 0.53 kg/kg clinker CO2 [2]. Consequently, reducing the limestone in the raw meal and thus changing its chemistry, could lead to lower CO2 emissions. This potential has resulted in increased scientific interest in innovative
∗ Corresponding author. Tel.: +30 2610969530; fax: +30 2610990917. E-mail address:
[email protected] (G.N. Angelopoulos). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.024
types of cement [4], and more specifically, in belite-rich cements over the last 20 years [5,6]. This type of cement, unlike conventional OPC, contains a higher percentage of belite (C2 S) and a lower percentage of alite (C3 S). In order to reach the desirable percentages of C2 S and C3 S, the lime saturation factor (LSF) must be between 78% and 83% [7]. The environmental benefits of the belite type cements over OPC can be summarized as follows: energy saving could rise up to 16% [8], burning temperatures could be reduced by 6–10% and levels of emitted CO2 and NOx could fall [9,10]. However, the early strength of such cements is lower and milling energy might be increased due to the hardness of C2 S. By combining the production of belite cements with alternative raw materials as a substitute for limestone, such as metallurgical slag’s, additional benefits may be obtained [11]. During the production of iron and steel, several types of slag are generated. These include blast furnace (BF), basic oxygen furnace (BOF), electric arc furnace (EAF) and stainless steel (SS-EAF and SS-AOD) slag. Nearly 50 Mt/y of steel slag is produced globally and 12 Mt/y is produced in Europe [12]. About 65% of this is used in qualified fields of applications, mainly construction, while the remainder is stored or used for other small purposes [13]. Around 37% of the steel slag produced in Europe in 2010 was used for cement production [14]. Nowadays more than 40% of global steel production takes place in EAFs [15] and is associated with 20 Mt/y slag generation. Greece has a cement production capacity of approximately 18 Mt/y [16],
288
R.I. Iacobescu et al. / Journal of Hazardous Materials 196 (2011) 287–294
and a steel production capacity of 3.5 Mt/y. The annual EAF slag (EAFS) production varies from 300.000 t/y to 400.000 t/y. Of the total amount of EAFS processed annually, around 55% is used in the production of coarse aggregates for road construction. The main environmental problems associated with the disposal of the wasted EAFS are the “dusting” of the slag and the release of leachates. Thus, quite apart from the needs of the cement industries, steel producers have their own motivations for finding a use for their slag. In principle, there are two methods of incorporating slag in cement production: either in the raw meal or in a later stage, as a (latent) hydraulic or pozzolanic material [17]. Prior studies indicate that an EAFS addition of up to 10 wt.% in the raw meal is effective without any detrimental effect on the technical properties of the resultant cement [17]. Other authors reached similar results, in terms of sintering, microstructure, as well as, hydrating properties of the final clinker with 10.5 wt.% EAFS addition. In addition to the above, other authors suggest that the addition of EAFS in clinker production will reduce the sintering temperature of raw meal and the theoretic heat consumption [18]. Moreover, cement containing steel slag as BOF and/or EAF can also have improved corrosion resistance than conventional Portland cement [19]. Finally, combinations of EAFS, BOF and AOD slags were tested for the production of sulfo-aluminate belite cements with encouraging results [20]. Despite the work in the field, no studies were focused on the use of EAFS for belite-rich cements to the best of our knowledge. In terms of disadvantages, the heavy metals content (Cr, V, etc.) in steel slag is an issue of concern. In the European Union (EU), a directive regarding Cr(VI) came into effect in 2005 and prohibits the use or supply of cements containing more than 2 ppm water-soluble chromium by mass of cement [21]. Typically, Cr(VI) compounds are more water-soluble (although insoluble do exist as well), thus more likely to participate in leaching. A number of adverse health effects have been associated with Cr(VI) exposure, ranging in severity. According to NIOSH [22] all Cr(VI) compounds are considered potential occupational carcinogens. Reducing agents, such as ferrous sulphate, either the monohydrate (FeSO4 ·H2 O) or heptahydrate (FeSO4 ·7H2 O) form, or stannous sulphate (SnSO4 ) are added to control the oxidation state of chromium [23]. The present work explores how EAFS can be exploited as a raw material in the production of low energy belite cements. The clinkers produced were characterised by SEM/EDS and Rietveld QXRD. Water demand, initial setting time, soundness and compressive strength were measured on both cement and cement paste. The hydration behaviour of these cements as well as their leaching potential is addressed in a separate work.
2. Materials and methods The raw materials used in the preparation of the raw meals were limestone, flysch and EAFS. The chemical analysis was performed by X-ray fluorescence spectrometry (XRF, Philips PW 2400). The crystalline phases of the raw materials were identified by Xray diffraction analysis (D5000 Siemens). Qualitative analysis was performed by the DIFFRACplus EVA® software (Bruker-AXS) based on the ICDD Powder Diffraction File. The mineral phases were quantified using a Rietveld-based quantification routine with the TOPAS® software (Bruker-AXS). This routine is based on the calculation of a single mineral-phase pattern and the refinement of the pattern using a non-linear least squares routine [24]. A number of corrections, including adjustments to the instrument’s geometry, background, sample displacement, detector type and mass absorption coefficients of the refined phases, were applied in order to achieve the optimum pattern fitting. Diffraction patterns were
measured in 2 range of 5–70◦ using CuK␣ radiation of 40 kV and 30 mA, with a 0.01◦ step size and step time of 1 deg/min. The design of the raw meals was based on the predictions of Bogue equations. In order to produce high belite cement, the lime saturation factor (LSF) was adjusted between 78% and 83% [7] whereas the alumina (AR) and silica ratios (SR) varied from 1.00% to 1.87% and 1.96% to 3.29%, respectively, similar to those adopted in the production of OPC. The quality indices LSF, AR and SR were calculated according to Eqs. (1)–(3) [5,23]:
LSF =
%CaO 2.8 ∗ %SiO2 + 1.2 ∗ %Al2 O3 + 0.65 ∗ %Fe2 O3
(1)
SR =
%SiO2 %Al2 O3 +%Fe2 O3
(2)
AR =
%Al2 O3 %Fe2 O3
(3)
Based on the above and on the chemical analysis of the raw materials, an MS Excel© worksheet was used in order to derive the syntheses of the raw meals. Three types of clinker were prepared: one as a reference (named BC), a second with the addition of 5 wt.% EALS (BC5) and third with the addition of 10 wt.% EAFS (BC10). The obtained meal contents in limestone/flysch/EAFS were in wt.%: 84.0/16.0/0.0, 80.5/14.5/5.0 and 77.0/13.0/10.0 for BC, BC5 and BC10, respectively. The results of the quality indices are presented in Table 3. BC10 presents a maximum in terms of EAFS addition (10 wt.%), while keeping LSF within the desired limits. The mineralogical phases of the clinkers were calculated also by the Rietveld method, besides the estimations derived from the Bogue equations (Table 4). For the preparation of the clinkers, raw materials were individually milled in a Siebtechnik® planetary mill at a particle size below 90 m. After mixing and homogenizing, pellets of approximately 15–20 mm diameter were formed by hand with a minimum water addition. The pellets were dried for 24 h at 110 ◦ C, followed by calcination at 1000 ◦ C for 4 h. Firing of the clinker was performed in a Nabertherm® type Super Kanthal resistance furnace at 1380 ◦ C. Optimum clinkering temperature was determined by burnability tests at 1280 ◦ C, 1300 ◦ C, 1320 ◦ C, 1350 ◦ C, 1380 ◦ C and 1400 ◦ C, with 40 min of soaking time to determine the free lime content according to ASTM C114-03, as well as by SEM observations which evaluated the quality of the clinker. For the stabilization of ␣ - and - C2 S polymorphic forms, fast cooling was applied by simultaneously applying blown air and crushing by means of a hammer. Clinkers were characterised by QXRD and SEM/EDS microanalysis (Jeol JSM 6300 and LINK PentaFET 6699, Oxford Instruments). Carbon coated samples, fractured, polished as well as etched with 1% Nital, were used. All EDS analyses were undertaken well away from phase boundaries. In the case of belite and alite crystals, spot analyses were performed. In the case of the interstitial phase, due to the micro-crystalline texture of the individual phases finely distributed within the amorphous one, analysis of an approximately 4 m × 4 m area away from alite and belite boundaries was performed. The default standards of LINK ISIS have been used. For the preparation of the cement, clinker was milled by means of the aforementioned planetary mill to finesse in the range of 4000–4100 cm2 /g. After milling, 5 wt.% gypsum with grain size lower than 90 m was added. Specific surface (Blaine method) was measured according to EN 196-6 [25], setting time and soundness according to EN 196-3 [26] and compressive strength according to EN 196-1 [27].
R.I. Iacobescu et al. / Journal of Hazardous Materials 196 (2011) 287–294 Table 1 Chemical composition of the raw materials (wt.%). Oxides
EAFS
Limestone
Flysch
CaO FeOtotal SiO2 Al2 O3 MnO MgO Cr2 O3 P2 O5 TiO2 SO3 Cl BaO Na2 O K2 O V2 O5 LOI
32.50 26.30 18.10 13.30 3.94 2.53 1.38 0.48 0.47 0.44 0.14 0.14 0.13 n.d. 0.06 0.00
48.90 1.00 9.00 1.36 n.d. 0.65 n.d. n.d. n.d. n.d. n.d. n.d. 0.10 0.15 n.d. 38.00
5.55 5.90 58.25 13.75 n.d. 2.86 n.d. n.d. n.d. 0.05 n.d. n.d. 1.10 2.50 n.d. 9.80
Total
99.91
99.16
99.76
LOI, loss on ignitions; n.d., not determined.
3. Results and discussion 3.1. Characterisation of raw materials XRF chemical analyses of the raw materials, limestone, flysch and EAFS are given in Table 1. It is observed that EAFS contains elements such as Cr, P, Ti, S and Ba that are considered as dopants for belite activation. The introduction of such ions into the crystal lattice of C2 S, can stabilise ␣ - and - polymorphs; ␣ -C2 S being more active than -C2 S [7,28]. XRD analyses of the raw materials are shown in Fig. 1 whereas the results for the semi-quantitative mineralogical analysis are presented in Table 2. The main mineralogical phases identified are calcite and quartz for limestone, quartz, illite, dolomite, albite and clinochlore for flysh. EAFS contains significant amounts of larnite (-belite), gehlenite, wüstite, magnetite, and brownmillerite. The Rietveld analysis results are 41.0 wt.%, 14.7 wt.%, 12 wt.%, 10.0 wt.% and 9.4 wt.% for the above phases, respectively.
289
Table 2 Mineralogical composition of the raw materials, wt.%, according to Rietveld analysis, normalised. Limestone Calcite Quartz Illite Microcline Muscovite Kaolinite Hematite Clinochlore Cristobalite Total
Flysch 90.5 5.7 1.1 1.1 0.6 0.5 0.2 0.2 0.1
Illite Quartz Kaolinite Dolomite Albite Calcite Microcline Muscovite Clinochlore Hematite
100
EAFS 34.1 28.2 8.5 6.8 6.7 5.9 5.7 2.2 1.3 0.6
Larnite Gehlenite Wüstite Magnetite Brownmillerite Mayenite Merwinite Spinel
100
41.0 14.7 12.0 10.0 9.4 7.2 3.7 2.0
100
3.2. Clinker quality as a function of firing temperature The composition of the produced clinkers, as well as the quality indices are presented in Table 3. In the firing tests, the maximum free lime content was 1.6 wt.% for both BC and BC10 at 1280 ◦ C. For the temperature range 1300–1400 ◦ C, free lime varied from 0.4 wt.% to 0.2 wt.% for all mixtures tested. For firing higher than 1300 ◦ C, therefore, free lime values are well below the commonly defined threshold of 1 wt.% for OPC clinker. In Figs. 2 and 3, backscattered electron images revealing the development of clinker microstructure at different firing temperatures for BC and BC10 are presented. In both cases for temperatures up to 1320 ◦ C, clinker microstructure is poorly developed. An extended interstitial phase is also observed. At temperatures exceeding 1350 ◦ C, the dissolution and transport phenomena through the melt are enhanced: the precipitation of stable, rounded belite is apparent in conjunction with the formation, at a lesser extent, of angular alite. The resulting microstructures are characterised by uniform distribution and development of the phases. Firing at 1400 ◦ C has no detectable difference in terms of phase morphology and growth compared to 1380 ◦ C. Comparing BC and BC10 microstructures, slag addition disfavours alite formation and promotes the formation of the interstitial phase as well as that of Table 3 Composition of raw metals, resulted chemical composition of the produced clinkers and quality indexes results.
Raw material EAFS Limestone Flysch Oxides SiO2 Al2 O3 Fe2 O3 CaO MgO K2 O Na2 O SO3 MnO Cr2 O3 P2 O5 TiO2 Cl BaO V2 O5 Fig. 1. XRD patterns of the raw materials. The main minerals identified are: 1, calcite (CaCO3 ); 2, quartz (SiO2 ); 3, illite ((K,H3 O)(Al,Mg,Fe)2 (Si,Al)4 O10 ((OH)2 ,(H2 O))); dolomite (CaMg(CO3 )2 ); 5, albite (NaAlSi3 O8 ); 6, clinochlore 4, (Mg2.5 Fe1.65 Al1.5 Si2.2 Al1.8 O10 (OH)8 ); 7, larnite (-Ca2 SiO4 ); 8, gehlenite (Ca2 Al(AlSi)O7 ); 9, wüstite (FeO); 10, magnetite (Fe3 O4 ); 11, brownmillerite (Ca2 (AlFe3 )2 O5 ); 12, mayenite (Ca12 Al14 O33 ).
Total Quality indexes LSF AR SR
0 wt.% BC
5 wt.% BC5
10 wt.% BC10
0.0 84.0 16.0
5.0 80.5 14.5
10.0 77.0 13.0
25.38 5.03 2.68 63.09 1.51 0.79 0.39 0.01 0.00 0.00 0.00 0.00 0.00 0.00 0.00
24.41 5.52 4.38 61.47 1.57 0.71 0.36 0.04 0.29 0.10 0.04 0.03 0.01 0.01 0.00
23.48 6.00 6.00 59.92 1.62 0.63 0.34 0.07 0.57 0.20 0.07 0.07 0.02 0.02 0.01
98.88
98.95
99.01
80.13 1.87 3.29
79.00 1.26 2.47
78.10 1.00 1.96
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Fig. 2. Backscattered images of preliminary firings. BC clinker formed at different firing temperatures: (a) 1280 ◦ C, (b) 1300 ◦ C, (c) 1320 ◦ C, (d) 1350 ◦ C, (e) 1380 ◦ C and (f) 1400 ◦ C.
belite. The size of crystals varies from 5 m to 50 m. These values are typical for a good quality clinker. According to the above results, it was decided that the firing of the clinker should be performed at 1380 ◦ C. 3.3. Clinker characterisation The XRD patterns of the prepared clinkers are depicted in Fig. 4. Table 4 presents the mineralogical compositions calculated by Rietveld and the estimations obtained by Bogue equations. As expected, discrepancies exist between the results obtained by Bogue and Rietveld, the most significant being Bogue’s underestimation of C3 S and overestimation of C2 S. These are attributed
to the fact that Bogue’s method is based on ideal stoichiometries for the clinker phases, without taking into consideration solid solutions, and also implies a certain reaction and solidification path. On the other hand, Rietveld analysis can reflect the changing conditions induced by the different raw materials as well as the non-equilibrium conditions during firing and cooling. Nonetheless, the qualitative trend is soundly predicted by Bogue. Based on the X-ray diffraction results, the main mineralogical phases identified were: alite (C3 S), belite (C2 S), tricalcium aluminate (C3 A), and tetracalcium-alumino-ferrite (C4 AF). The specific polymorphism is important as affects the hydration and, as a consequence, the development of microstructure and mechanical properties. The main polymorphic form of stabilised alite is tri-
Fig. 3. Backscattered images of preliminary firings. BC10 clinker formed at different firing temperatures: (a) 1280 ◦ C, (b) 1300 ◦ C, (c) 1320 ◦ C, (d) 1350 ◦ C, (e) 1380 ◦ C and (f) 1400 ◦ C.
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291
Fig. 4. X-ray patterns of the prepared clinkers. The main minerals identified: 1, C3 S; 2, C2 S; 3, C3 A; 4, C4 AF.
clinic (T), which decreases as the slag content increases towards the monoclinic (M) one. This is expected to affect early strength development towards lower values [29]. The rhombohedral (R) alite is reacting slightly more rapidly compared with T and M polymorphic forms [29,30] but without a significant effect in the present case due to the low content. Regarding belite, mainly the ␣ polymorph was identified. The ␥- and -C2 S polymorphs were also detected and increased as the slag content rose. Beta, - and ␣ -C2 S play an important role during late days of hydration. On the contrary, ␥C2 S does not present notable hydraulic properties. C3 A was formed as cubic and orthorhombic in BC, whereas only the orthorhombic form was found in BC5 and BC10. The predominance of orthorhombic C3 A in the last two mentioned clinkers was most probably due to the higher sulphate content in the EAFS which inhibits the formation of cubic C3 A [31]. The identified orthorhombic C3 A polymorph will react faster in the presence of gypsum than its cubic counterpart [32]. In general, the increase in EAFS content results in a decrease in C3 S and C3 A and an increase in C4 AF; this is attributed to the high iron content of the slag.
Table 4 Estimated (Bogue) and calculated (Rietveld) mineralogical composition of the prepared clinkers. Phases
BC Rietveld
BC5 Bogue
Rietveld
BC10 Bogue
Bogue
1.0 0.3 33.3 0.7
Total C2 S-␣’ C2 S- C2 S-␥
35.30 45.5 1.9 0.0
27.43
28.90 41.6 4.0 1.5
22.39
21.80 42.2 5.8 0.8
17.59
Total C3 A – Cubic C3 A – Orth.
47.40 2.7 6.0
54.96
47.10 0.0 4.9
56.16
48.80 0.0 4.3
57.32
8.70 7.50 0.4 0.7 100
9.13 8.49 0.0 0.0 100
4.90 18.50 0.0 0.6 100
7.54 13.90 0.0 0.0 100
4.30 24.10 0.0 1.0 100
6.01 19.08 0.0 0.0 100
Total C4 AF Lime MgO Total
3.4 1.3 20.8 3.4
Rietveld
C3 S (M1) C3 S (M3) C3 S-T C3 S-R
4.0 1.5 15.4 0.9
Fig. 5. Backscattered images of clinkers in polished section: (a) BC, (b) BC5 and (c) BC10.
Backscattered electron images (BEI) are presented in Fig. 5a, b and c for BC, BC5 and BC10, respectively. In all cases the microstructures consist predominantly of well-developed crystals of type I belite according to Insley’s classification [33]. Their diameter varies from 5 m to 50 m. Most of the belite crystals display complex twin lamellae. The formed striations arise as a consequence of phase transformation during cooling. In type I belite, it has been reported that this is a skeleton structure rather than polysynthetic twinning, consisting of beta and alpha forms of belite [33]. Crystals with parallel striations are less noticeable. Angular, euhedral and subhedral alite crystals are also present. In the case of BC10, alite crystals
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they undergo inversions on cooling which cause twinning resulting in strain accumulations [34] and (b) they were caused by the etching [33]. Representative EDS micro-analyses of C3 S and C2 S crystals and interstitial phase for each clinker prepared are presented in Table 5. Ferrite and aluminate phases are not presented individually due to the difficulties imposed by their micro-crystalline texture, but analysis of the interstitial phase is denoted as C3 A + C4 AF in the table. As a general remark, it was observed that alite and belite become enriched with iron as the slag content rises, whereas Ba, Cr, Ti and P are more likely to be found in the belite crystals. According to the present QXRD and SEM/EDS results and under the adopted experimental conditions, no clear conclusions can be drawn concerning the influence of slag addition on the belite crystals, although it is a host of elements such as P, S, Cr which are known as belite stabilisers. However, this will be the subject of a forthcoming communication dealing with the composition and substitutions in the individual phases. 3.4. Specific surface, setting time and soundness measurements In Table 6 the results of water demand, initial setting time and soundness of the cement pastes are presented. To prepare the cement paste with the standard consistency [26] the water demand was 27.6 wt.% for all cases. Initial setting times obtained for BC, BC5 and BC10 were 240 min, 170 min and 20 min, respectively according to EN 197-1 standard in the first 2 cases. It is observed that the use of slag decreases the setting time, with the BC10 behaving as fast-setting cement. This is attributed to the increase of the molten phase, forming upon cooling higher amounts of C4 AF, and the interaction with C3 A and gypsum during hydration. In more detail, as early hydration of cement is principally controlled by the amount and activity of C3 A, setting is balanced by the amount and type of sulphate interground with the cement. Tetracalcium-alumino-ferrite (C4 AF) reacts much like C3 A, i.e., forming ettringite in the presence of gypsum. The higher amounts of C4 AF in the clinkers with slag (more than two-fold and three-fold increase for BC5 and BC10, respectively compared to BC) consume also higher amounts of gypsum. Hence, it is likely that setting occurs due to the uncontrolled reaction of the C3 A, after depletion of the sulphate by reaction with the C4 AF. Addition of higher amounts of gypsum or of retarders, possibly of organic nature, could control the setting behaviour. Expansion was 1 mm for all prepared cements. In order to obtain Blaine of 4000 cm2 /g, 4080 cm2 /g and 4057 cm2 /g, the required milling times were 60 sec, 80 sec and 103 sec for BC, BC5 and BC10, respectively. Notably, increasing the slag addition results in an increased milling time for comparable finesse. 3.5. Compressive strength of the BC, BC5 and BC10 cements Fig. 6. Backscattered images of clinkers in fractured surface: (a) BC, (b) BC5 and (c) BC10.
contain inclusions of belite in some cases. The crystallised-duringcooling interstitial phase presents a micro-crystalline texture and mainly consists of a mixture of ferrite and C3 A. It partially separates the primary alite and belite crystals. The addition of slag favours the formation of belite and ferrite phases as it disfavours the formation of alite. Ferrite crystals pipework exposed on a broken surface are presented in Fig. 6b and c for BC5 and BC10, respectively. Voids in their structure are presumably occupied by aluminate. The lamellar belite structure is slightly visible. In Fig 6b micro-cracks that developed on the belite crystals are clearly visible. Plausible hypotheses for the formation of these micro-cracks are: (a) as dicalcium silicate crystals are considered to be complex, containing point defects,
In Fig. 7 the compressive strength results are presented. As expected for belite cements, the early days strength development is significantly lower than that of the OPC. For BC, the results for 2 days were 6.5 MPa whereas for BC5 and BC10 they were even lower, at 2.5 MPa and 1.6 MPa, respectively. However the 28-day results for BC, BC5 and BC10 were 47.5 MPa, 46.6 MPa and 42.8 MPa, respectively, which are comparable to OPC CEMI 32.5N (32.5–52.5 MPa) according to EN 197-1 [35]. These results concur with previously reported results [36–39]. The low early strength development observed in the cements with slag addition is attributed, as in cases of fast setting behaviour, to the extended formation of ferrite. The higher results for BC at early days are attributed to the higher C3 S and C3 A content, compared to BC5 and BC10. Compared to other belite cements incorporating wastes, belite cement with EAFS has
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293
Table 5 Typical EDS micro-analyses for clinkers: (a) BC, (b) BC5 and (c) BC10, in wt.%. Phase
BC
BC5
BC10
C3 S
C2 S
C4 AF + C3 A
C3 S
C2 S
C4 AF + C3 A
C3 S
C2 S
C4 AF + C3 A
NaO MgO Al2 O3 SiO2 P2 O5 SO3 K2 O CaO TiO2 MnO FeOtotal V2 O5 BaO Cr2 O3
0.10 0.40 1.23 24.05 0.00 0.00 1.43 71.25 0.00 0.00 1.54 0.00 0.00 0.00
0.20 0.36 1.73 30.65 0.00 0.20 1.44 64.10 0.00 0.00 1.32 0.00 0.00 0.00
0.67 1.45 20.74 5.35 0.00 0.08 0.50 59.14 0.00 0.00 12.07 0.00 0.00 0.00
0.11 0.30 1.66 23.66 0.00 0.00 1.36 69.97 0.04 0.36 2.19 0.00 0.00 0.35
0.01 0.36 0.84 30.88 0.23 0.03 1.35 61.38 0.10 0.39 3.22 0.00 0.09 0.22
0.28 1.53 20.04 4.16 0.00 0.14 0.50 56.41 0.06 1.35 15.43 0.00 0.05 0.05
0.09 0.81 2.25 22.49 0.01 0.10 1.50 68.01 0.08 0.94 3.34 0.00 0.03 0.35
0.12 0.97 2.82 30.25 0.43 0.00 0.81 59.03 0.04 0.38 4.52 0.00 0.13 0.50
0.19 1.35 18.99 4.78 0.00 0.13 0.54 53.59 0.09 1.38 18.86 0.02 0.06 0.02
Total
100
100
100
100
100
100
100
100
100
• Early days compressive strength results are low, as is expected for belite cements, however the 28-day results for BC, BC5 and BC10 were 47.5 MPa, 46.6 MPa and 42.8 MPa, respectively which are comparable to EN 197-1, OPC CEMI 32.5N ones. • The addition of slag did not affect the water demand of the cements and soundness did not exceed 1 mm, although setting time was decreased for BC10, which behaved like “flash set” cement. • The fast setting that occurred for the cements with slag addition is attributed to the extended formation of the molten phase which forms ferrite upon cooling and the interaction with C3 A and gypsum during hydration. Acknowledgements
Fig. 7. Compressive strength results of the BC, BC5 and BC10 cements.
Table 6 Physical properties of the cement and cement pastes. Cement types Specific surface (cm2 /g) Initial setting time (min) Water demand (wt.%) Soundness (mm)
BC 4000 240 27.6 1
BC5
BC10
4080 170 27.6 1
4057 20 27.6 1
R.I. Iacobescu and R. Saban acknowledge the support of the Sectoral Operational Programme for Human Resources Development 2007–2013 of the Romanian Ministry of Labour, Family and Social Protection through the Financial Agreement POSDRU/6/1.5/S/16. D. Koumpouri and G.N. Angelopoulos acknowledge the support of University of Patras through the “Karatheodoris” 2011 research program. Y. Pontikes is thankful to the Research Foundation – Flanders for the post-doctoral fellowship. TITAN Cement Company S.A. and SOVEL S.A. metallurgy industries are gratefully acknowledged for providing raw materials as well as their technical assistance. References
lower early and late compressive strength than belite with Bayer’s process red mud [38] (5.0 MPa at 1 day and 53.7 MPa at 28 days) and higher early and late compressive strength than belite cement with boron waste [39] (1.5 MPa at 1 day and 32.3 MPa at 28 days). 4. Conclusion The production of belite cements with EAFS is feasible and can offer significant environmental advantages. Specifically the characteristics of the cements are as follows: • Clinkers predominantly contain well-formed belite crystals. Alite crystals are also present. • The interstitial phase is a mixture of C4 AF and C3 A, and partially separates the primary alite and belite crystals. • Slag addition favours the formation of the belite and ferrite phase and disfavours the formation of alite, in accordance with Bogue’s predictions under the requirement of comparable quality indices.
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[27] European Committee for Standardization, EN 196-1, Methods of testing cement. Part 1. Determination of Strength, 1994. [28] S.N. Ghosh, P.B. Rao, A.K. Paul, K. Raina, Review. The chemistry of dicalcium silicate mineral, J. Mater. Sci. 14 (1979) 1554–1566. ´ The influence of the alite polymorphism on the strength [29] T. Stanek, P. Sulovsky, of the Portland cement, Cem. Concr. Res. 32 (2002) 1169–1175. [30] R.T.H. Aldous, The hydraulic behaviour of rhombohedral alite, Cem. Concr. Res. 13 (1983) 89–96. [31] L. Gobbo, L. Sant’ Agostino, L. Garcez, C3 A polymorphs related to industrial clinker alkalies content, Cem. Concr. Res. 34 (2004) 657–664. [32] A. Kirchheim, V. Fernàndez-Altable, P. Monteiro, D. Dal Molin, I. Casanova, Analysis of cubic and orthorhombic C3 A hydration in presence of gypsum and lime, J. Mater. Sci. 44 (2009) 2038–2045. [33] D.H. Campbell, Microscopical Examination and Interpretation of Portland Cement and Clinker, Portland Cement Association, Skokie, IL 60077-1083, USA, 1999. [34] S.N. Ghosh, Advances in Cement Technology: Critical Reviews and Case Studies on Manufacturing, Quality Control, Optimisation and Use, New Delhi, India, 1983. [35] European Committee for Standardization, EN 197-1, Cement. Part 1. Composition, Specifications and Conformity Criteria for Common Cements, 2000. [36] J. Stark, A. Müller, R. Schrader, K.R. Rumpler, Existence Conditions of Hydraulically Active Belite, Zement-Kalk-Gips, Bauverlag GmbH, Wiesebaden, Germany, 1981. [37] I. Vangelatos, Valorisation of Red Rud in the Cement Industry, Department of Chemical Engineering, University of Patras, Patras, 2008. [38] I. Vangelatos, Y. Pontikes, G.N. Angelopoulos, Ferroalumina as a raw material for the production of “Green” belite type cements, in: SERES’ 09. I. International Ceramic, Glass, Porcelain Enamel, Glaze and Pigment Congress, Eskisehir, Turkey, 2009. [39] T. Kavas, I. Vangelatos, S. Koyas, Y. Tabak, G.N. Angelopoulos, Wastes from alumina and boron production as raw materials for belite cement, in: J. Heinrich, C. Aneziris (Eds.), 10th Conference and Exhibition of the European Ceramic Society, Berlin, Germany, 2007, pp. 1799–1803.
Journal of Hazardous Materials 196 (2011) 248–254
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The influence of preparation method, nitrogen source, and post-treatment on the photocatalytic activity and stability of N-doped TiO2 nanopowder Shaozheng Hu ∗ , Fayun Li, Zhiping Fan Institute of Eco-environmental Sciences, Liaoning Shihua University, Fushun 113001, PR China
a r t i c l e
i n f o
Article history: Received 19 May 2011 Received in revised form 6 September 2011 Accepted 6 September 2011 Available online 10 September 2011 Keywords: Lattice-nitrogen TiO2 Photocatalysis Stability Post-treatment
a b s t r a c t NH3 plasma, N2 plasma, and annealing in flowing NH3 were used to prepare N doped TiO2 , respectively. XRD, UV–vis spectroscopy, N2 adsorption, FT-IR, Zeta-potential measurement, and XP spectra were used to characterize the prepared TiO2 samples. The nitridation procedure did not change the phase composition and particle sizes of TiO2 samples, but extended its absorption edges to the visible light region. The photocatalytic activities were tested in the degradation of an aqueous solution of a reactive dyestuff, methylene blue, under visible light. The photocatalytic activity and stability of TiO2 prepared by NH3 plasma were much higher than that of samples prepared by other nitridation procedures. The visible light activity of the prepared N doped TiO2 was improved by increasing the lattice-nitrogen content and decreasing adsorbed NH3 on catalyst surface. The lattice-nitrogen stability of N-doped TiO2 samples improved after HCl solution washing. The possible mechanism for the photocatalysis was proposed. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Nanocrystalline TiO2 has great potential for many applications such as photocatalysis, solar energy conversion, and gas sensor [1,2]. However, with a wide band gap energy of 3.0–3.2 eV, TiO2 cannot be activated to generate photoexcited electrons and holes to promote redox reaction unless it is irradiated by ultraviolet. This hinders the application of TiO2 as a photocatalyst with response to solar light or even indoor light. Therefore, it is highly desirable to shift the absorption edge of TiO2 to the visible light region. In 2001, Asahi et al. [3] prepared nitrogen doped TiO2 films by sputtering TiO2 in a N2 /Ar gas mixture, and concluded that the doped N atoms narrowed the band gap of TiO2 by mixing N 2p and O 2p states, therefore demonstrating the activity for the decomposition of acetone and methylene blue. Since then, N-doping has become a hot topic and been widely investigated. Heating TiO2 powders in N2 and/or NH3 at elevated temperatures is the conventional method to prepare nitrogen-doped TiO2 [3]. Besides the energy waste, the treatment at such high temperature usually results in the low surface area due to grain growth, which would decrease the number of photoactive sites. Therefore, new strategies for preparing nitrogen-doped TiO2 , such as sputtering [4], sol–gel [5], ion implantation [6], pulsed laser deposition [7], hydrothermal synthesis [8], and plasma treatment [9] have been proposed more recently.
∗ Corresponding author. Tel.: +86 24 23847473. E-mail address:
[email protected] (S. Hu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.021
Non-thermal plasma is composed of atoms, ions and electrons, which are much more reactive than their molecule precursors. Plasma is able to initiate a lot of reactions, which take place efficiently only at elevated temperatures and high pressures, under mild conditions. So far, some literatures on preparation of N doped TiO2 by plasma treatment have been reported [9–13]. Yamada et al. [9–11] investigated the photocatalytic activity of TiO2 thin films prepared by plasma treatment using N2 as nitrogen source. They suggested that the substitutional N-doping contributed to the band gap narrowing, therefore absorbing visible light and demonstrating the photocatalytic activity. Abe et al. [12] prepared N doped TiO2 by NH3 (10%)/Ar plasma. The influence of the NH3 /Ar gas pressures (50, 300 and 1000 Pa) on the physical and photocatalytic property of the powder was investigated. Miao et al. [13] reported the structural and compositional properties of TiO2 thin films prepared by N2 –H2 plasma treatment. HRTEM results indicated that the primitive lattice cells of anatase TiO2 films are distorted after plasma treatment in comparison with that of bulk TiO2 , which confirmed the N doping by N2 –H2 plasma. It is shown from the above literatures that N2 and NH3 were usually used as nitrogen source to prepare N doped TiO2 under plasma treatment. However, few literature on the comparison of N2 and NH3 plasma treated TiO2 were reported. In this work, NH3 plasma, N2 plasma, and annealing in flowing NH3 were used to prepare N doped TiO2 , respectively. The structural and optical properties of prepared N doped TiO2 were compared. The photocatalytic performance was evaluated in the degradation of methylene blue under visible light. The possible mechanism for the photocatalysis was proposed.
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2. Experimental 2.1. Preparation and characterization
2.2. Photocatalytic reaction Methylene blue (MB) was selected as model compound to evaluate the photocatalytic performance of the prepared TiO2 particles in an aqueous solution under visible light irradiation. 0.1 g TiO2 powders were dispersed in 100 ml aqueous solution of MB (initial concentration C0 = 50 ppm, pH 6.8) in an ultrasound generator for 10 min. The suspension was transferred into a self-designed glass reactor, and stirred for 30 min in darkness to achieve the adsorption equilibrium. The concentration of MB at this point was considered as the absorption equilibrium concentration C0 . The adsorption capacity of a catalyst to MB was defined by the adsorption amount of MB on the photocatalyst (C0 − C0 ). In the photoreaction under visible light irradiation, the suspension was exposed to a 110 W high-pressure sodium lamp with main emission in the range of 400–800 nm, and air was bubbled at 130 ml min−1 through the solution. The UV light portion of sodium lamp was filtered by 0.5 M NaNO2 solution [14]. The light intensity is 130 mW cm−2 . All runs were conducted at ambient pressure and 30 ◦ C. At given time intervals, 4 ml suspension was taken and immediately centrifuged to separate the liquid samples from the solid catalyst. The concentrations of MB before and after reaction were measured by means of a UV–vis spectrophotometer at a wavelength of 665 nm. It is the linear relationship between absorbance and concentration of liquid sample in the experimental concentration range. Therefore, the percentage of degradation D% was determined as follows: D% =
A0 − A × 100% A0
TO-PN2
TO-CNH3
P25
20
30
40
50
60
2θ / deg. Fig. 1. XRD patterns of P25 and prepared N-doped TiO2 samples.
where A0 and A are the absorbances of the liquid sample before and after degradation, respectively. 3. Results and discussion It is reported that the phase composition and particle size of TiO2 have significant influence on its photocatalytic activity [2]. The XRD patterns of P25 and prepared N-doped samples (Fig. 1) indicate that all TiO2 samples were mixtures of anatase and rutile phases. The phase contents and the particle sizes of the samples were calculated by their XRD patterns according to the method of Spurr [15] and Debye–Scherrer equation [16], respectively. The results (Table 1) indicate that there were no remarkable changes in phase composition and particle sizes. Up to date, the mechanism of the enhancement by N-doping is still controversial. Asahi et al. [3] concluded that the doped N atoms narrowed the band gap of TiO2 by mixing N 2p and O 2p states, therefore demonstrating the activity. Irie et al. [17] argued that the isolated narrow band located above the valence band is responsible for the visible light response. Lee et al. [18] suggested that substitutional N-doping would narrow the band gap by the coupling of the O 2p and N 2p orbitals, while interstitial N-doping would create an isolated defect state between the conduction band and valence band. Fig. 2 shows the UV–vis spectra of P25 and prepared N-doped TiO2 samples. Compared with the spectra of P25, obvious red-shifts of the absorption bands were observed for prepared N-doped TiO2 . 1.0
TO-PNH3
0.8
TO-CNH3 TO-PN2
0.6
Abs.
The doping of TiO2 was conducted in a dielectric barrier discharge (DBD) reactor, consisting of a quartz tube and two electrodes. The high voltage electrode was a stainless-steel rod (2.5 mm), which was installed in the axis of the quartz tube and connected to a high voltage supply. The grounding electrode was an aluminum foil, which was wrapped around the quartz tube. For each run, 0.4 g commercial TiO2 powder (P25) was charged into the quartz tube. At a constant NH3 flow (40 ml min−1 ), a high voltage of 9–11 kV was supplied by a plasma generator at an overall power input of 50 V × 0.4 A. The discharge frequency was fixed at 10 kHz, and the discharge was kept for 15 min. After discharge, the reactor was cooled down to room temperature. The obtained TiO2 sample was denoted as TO–PNH3 . When N2 was used to replace NH3 following the same procedure in the preparation of TO–PNH3 , the product is denoted as TO–PN2 . For comparison, P25 was calcined under NH3 flow (40 ml min−1 ) for 15 min at 500 ◦ C. The obtained sample was denoted as TO–CNH3 . XRD patterns of the prepared TiO2 samples were recorded on a ˚ Rigaku D/max-2400 instrument using Cu K␣ radiation ( = 1.54 A). UV–vis spectroscopy measurement was carried out on a Jasco V-550 spectrophotometer, using BaSO4 as the reference sample. FT-IR spectra were obtained on a Nicolet 20DXB FT-IR spectrometer in the range of 400–2300 cm−1 . The zeta-potential of the catalyst was measured at room temperature on Zetasizer Nano S90 (Malvern Instruments). The pH was adjusted by dropwise addition of dilute HCl or NaOH solution. Photoluminescence (PL) spectra were measured at room temperature with a fluorospectrophotometer (FP-6300) using a Xe lamp as excitation source. XPS measurements were conducted on a Thermo Escalab 250 XPS system with Al K␣ radiation as the exciting source. The binding energies were calibrated by referencing the C1s peak (284.6 eV) to reduce the sample charge effect.
Intensity / a.u.
TO-PNH3
P25 0.4
0.2
0.0 300
400
500
600
700
Wavelength / nm (1) Fig. 2. UV–vis spectra of P25 and prepared N-doped TiO2 samples.
800
250
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Table 1 Summary of physical properties of P25 and prepared N-doped TiO2 samples. Sample
Size (nm)
XA (%)a
SBET (m2 g−1 )
Pore volume (cm3 g−1 )
Central pore size (nm)
Eg (eV)
Nfresh (at.%)b
Nused (at.%)c
P25 TO–PN2 TO–CNH3 TO–PNH3
28.2 28.5 29.3 28.1
74.6 74.7 74.4 75.1
43 40 36 41
0.07 0.06 0.05 0.06
3.6 3.2 3.1 3.4
3.10 2.92 2.75 2.67
0 1.32 1.64 1.95
0 0.76 1.17 1.91
c
XA represents the phase composition of anatase. Nfresh represents the lattice-nitrogen content before photocatalytic reaction Nused represents the lattice-nitrogen content after photocatalytic reaction.
The band gap energies of TiO2 samples which calculated according to the method of Oregan and Gratzel [19] (Table 1) indicate that the prepared N-doped TiO2 samples exhibited much narrowed band gap energies. According to the previous result [3,18], this indicated that substitutional N-doping existed in the prepared N-doped TiO2 samples. It is shown that the band gap energy decreased in the order: TO–PN2 > TO–CNH3 > TO–PNH3 , which is probably due to the different doping N content in prepared N-doped TiO2 samples. Xu et al. [20] prepared N doped TiO2 by pulsed laser deposition and suggested that more absorption edge red-shift indicated higher nitrogen concentration. Besides, the distinct differences in visible light absorption are observed between calcination and plasma treated samples. In the spectrum of TO–CNH3 , the obvious absorption in 400–550 nm is observed, which is a typical absorption region for N doped TiO2 materials. This typical absorption is due to the electronic transition from the isolated N 2p level, which is formed by incorporation of nitrogen atoms into the TiO2 lattice, to the conduction band [21]. However, the spectra of plasma treated samples are obvious different. The broad absorptions in the whole visible light region are observed in the spectra of TO–PN2 and TO–PNH3 . Huang et al. [22] prepared the visible light responsive TiO2 by nitrogen-plasma surface treatment, and found the similar broad absorption in visible light region. Abe et al. [12] prepared N doped TiO2 by NH3 /Ar plasma, and suggested that such broad absorption is attributed to the presence of Ti3+ , which might be formed by plasma treatment. It is noted that TO–PNH3 showed much stronger broad absorption in visible light region than TO–PN2 . This is probably due to that NH3 plasma consists of not only various active nitrogen species but excited hydrogen, leading to Ti4+ reduced easily, thus more Ti3+ were formed. Therefore, according to the conclusion of Lee et al. [18], it is proposed that substitutional and interstitial Ndoping existed simultaneously in TO–CNH3 which caused the band gap narrowing and remarkable absorption in 400–550 nm, whereas only substitutional N-doping existed in TO–PN2 and TO–PNH3 . The
N 1s
TO-PN2
Intensity / a.u.
a b
TO-PNH3
TO-CNH3
396
398
400
Fig. 3. XP spectra of prepared N-doped TiO2 samples in the region of N1s.
broad absorptions of TO–PN2 and TO–PNH3 in the whole visible light region were due to the presence of Ti3+ caused by the N-doping. XPS is an effective surface test technique to characterize elemental composition and chemical states. According to the previous literatures [10,11], the peaks around 396 and 400 eV are attributed to the formation of lattice-nitrogen and other surface N species such as N–N and N–O bond. The XP spectra in the region of N1s (Fig. 3) indicated that most N species in prepared TiO2 using NH3 as nitrogen source existed in lattice-nitrogen, whereas other surface N species such as N–N and N–O bond were dominant in TO–PN2 which using N2 as nitrogen source. The lattice-nitrogen content calculated by XPS data were shown in Table 1. The Nfresh content decreased in the order: TO–PNH3 > TO–CNH3 > TO–PN2 , which indicated that NH3 plasma treatment is more effective than another two methods
(B)
(A)
O 1s
Intensity / a.u.
Intensity / a.u.
Ti 2p
TO-PNH3 TO-CNH3
TO-PNH3 TO-PN2 TO-CNH3
TO-PN2
P25
P25
456
458
460
462
Binding Energy / eV
464
466
402
Binding Energy / eV
529
530
531
532
Binding Energy / eV
Fig. 4. XP spectra of P25 and prepared N-doped TiO2 samples in the region of Ti 2p (A) and O1s (B).
533
534
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P25 TO-CNH3
20
TO-P HN3
TO-PN2
1070 cm
TO-C HN3
-1
1224 cm
-1
P25 1402 cm -1 1620 cm-1
655 cm-1
Zeta potential / mV
Absobance / a.u.
10
TO-PNH3 0 2
4
6
8
10
12
-10
-20
-30
pH 400
600
800
1000
1200
1400
1600
Wavenumber / cm
1800
2000
2200
-1
Fig. 6. Plots of the zeta-potential as a function of pH for P25 and prepared N-doped TiO2 suspensions in the presence of NaCl (10−3 M).
Fig. 5. FT-IR spectra of P25, TO–CNH3 and TO–PNH3 .
to form lattice-nitrogen. This is probably due to that NH3 plasma consists of various nitriding species, leading to the formation of lattice-nitrogen easier [23]. Besides, trace N species, which located at 395.3 eV were present in TO–PNH3 and TO–CNH3 . Li et al. [24] prepared N doped TiO2 in NH3 /ethanol fluid under supercritical condition, and suggested that the N species with binding energy at 395.3 eV was attributed to the surface adsorbed NH3 molecules. In this investigation, the N species located at 395.3 eV only existed in TO–PNH3 and TO–CNH3 , which prepared using NH3 as nitrogen source. This confirmed the result of Li et al. Furthermore, the peak intensity of TO–CNH3 at 395.3 eV was obvious higher than that of TO–PNH3 , indicating more NH3 molecules adsorbed on TO–CNH3 surface. Fig. 4 shows the XP spectra of P25 and prepared N-doped TiO2 samples in the region of Ti 2p and O1s. Compared with the spectra of P25, obvious shifts to lower binding energies were observed for N-doped TiO2 samples in the Ti 2p region (458.4 eV) as well as the O1s region (529.7 eV). This is probably attributed to the change of chemical environment after N doping [24]. It is known that the binding energy of the element is influenced by its electron density. A decrease in binding energy implies an increase of the electron density. The electrons of N atoms may be partially transferred from N to Ti and O, due to the higher electronegativity of oxygen, leading to increased electron densities on both Ti and O. The peaks around 530 and 532 eV in the O1s region are attributed to crystal lattice oxygen (Ti–O) and surface hydroxyl group (O–H) of TiO2 . The ratio of these two peak areas (SO–H /STi–O ) represents the abundance of surface hydroxyl groups. The calculated results indicated that SO–H /STi–O ratio for TO–CNH3 was 0.06, much lower than that of P25 (0.14). Whereas, SO–H /STi–O ratios for TO–PN2 and TO–PNH3 were 0.13 and 0.11, which were slight lower than P25. This indicated the content of surface hydroxyl groups decreased more drastically after the calcination procedure under NH3 flow compared with plasma treatment. Those surface hydroxyl groups are known to play an important role in photocatalysis. They react with photogenerated holes, producing active hydroxyl radicals, which are responsible for the photo-degradation [25]. The FT-IR spectra of P25 and TO–PNH3 were shown in Fig. 5. The absorption peak at 1620 cm−1 is attributed to bending vibration of hydroxyl group. The band at around 655 cm−1 belongs to O–Ti–O structure of TiO2 . There are three bands at 1402, 1224, and 1070 cm−1 which were observed in the spectra of TO–CNH3 and TO–PNH3 , but not in that of P25. The band at 1402 cm−1 is attributed to the surface adsorbed NH3 species on Brönsted acid sites (–OH)
[26]. It is known that NH3 can adsorb on Brönsted acid sites (–OH) located at 1400 cm−1 and Lewis acid sites (Ti4+ ) located at 1225 and 1190 cm−1 [26,27]. However, in Fig. 5, no NH3 adsorbed on the Lewis acid sites was observed. There are many previous literatures, which report the FT-IR results of NH3 adsorbed on TiO2 materials. Some of them reported that NH3 adsorbed on both Brönsted acid and Lewis acid sites [26,27]. Other results showed that only adsorption on Brönsted acid sites was obtained which is consistent with the result of Fig. 5 [24,28]. Therefore, it is proposed that the preparation methods and conditions probably affect the adsorption state, leading to the adsorption site different from different literatures. It is known that the TiO2 surface is hydrophilic. In this investigation, TiO2 materials were treated under calcination and plasma condition only for 15 min, leading to most of the H2 O adsorbed on Ti4+ still existed on TiO2 surface. It is reported that when Ti4+ sites are saturated by hydroxyl groups, NH3 will adsorb mainly on Brönsted acid sites by the formation of an N· · ·HO bond [29]. Therefore, only adsorption on Brönsted acid sites was obtained. In Fig. 5, the peaks at 1224 and 1070 cm−1 could be attributed to the nitrogen atoms embedded in the TiO2 network, which is consistent with XPS result [28]. These results confirmed the formation of doping N species in the TiO2 lattice. Fig. 6 shows the plots of the zeta-potential as a function of pH for P25 and prepared N-doped TiO2 suspensions in the presence of NaCl (10−3 M). It is known that the point of zero charge (PZC) of TiO2 is around 3–6, indicating the surface of TiO2 particles is positively charged. Compared with P25, the distinct shifts to lower value of the PZC were observed for all the N doped TiO2 , indicating the positive charge on TiO2 surface decreased. The PZC value decreased in the order: P25 > TO–PN2 > TO–PNH3 > TO–CNH3 . It is possible that the lone electron pair of doping N counteract a few of positive charge. Besides, NH3 readily adsorbed on the catalyst surface during the nitridation process, due to the numerous acidic hydroxyl groups on the TiO2 surface. The presence of these surface-adsorbed NH3 decreased the number of acidic hydroxyl groups, resulting in lower PZC value of TO–PNH3 and TO–CNH3 . Furthermore, plasma treatment caused the NH3 decomposition more drastically, leading to less NH3 adsorbed on TO–PNH3 surface compared with TO–CNH3 . Therefore, the PZC value of TO–PNH3 is higher than TO–CNH3 . The adsorption of MB on TiO2 -based catalysts was measured by the equilibrium adsorption capacity. The adsorption capacities of all the N doped TiO2 samples were lower than that of P25 (Fig. 7). The BET specific surface area (SBET ), pore volume, and central pore size are listed in Table 1. Compared with P25, the SBET , pore volume, and central pore size of prepared samples decreased.
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6
100
TO-PNH3
5
TO-PNH3(HCl)
TO-PN2 4
D/%
TO-PNH3 3
TO-CNH3
60
40
2 20
1
0
0
0
Fig. 7. Adsorption capacity of MB on P25 and N-doped TiO2 samples.
This probably caused the decreased equilibrium adsorption capacity shown in Fig. 7. Besides, it is possibly that such decreased adsorption of MB is attributed to the coverage of TiO2 surface by excess surface N species. It is noted that the equilibrium adsorption capacity decreased in the order: P25 > TO–PN2 > TO–PNH3 > TO–CNH3 , which is completely consistent with the order of surface hydroxyl groups content. This indicated that the content of surface hydroxyl groups influenced significantly on the equilibrium adsorption capacity. It is shown that equilibrium adsorption capacities of TO–PNH3 and TO–CNH3 were lower than TO–PN2 , which using N2 as nitrogen source. Besides the lower surface hydroxyl groups content than that of TO–PN2 , large numbers of NH3 molecules adsorbed on hydroxyl groups of TO–PNH3 and TO–CNH3 , caused the reduced surface sites for adsorbing MB, leading to lower equilibrium adsorption capacity than TO–PN2 . The photocatalytic performances under visible light shown in Fig. 8 indicate that prepared N-doped TiO2 samples exhibited much higher activities than that of P25. Since no obvious change were observed in phase compositions and particle sizes between P25 and prepared N-doped TiO2 samples, the enhanced photocatalytic activity must result from the doping of nitrogen in TiO2 , which gave rise to the narrowed band gap and thus to the enhanced absorption in the visible region. Moreover, it is shown that the photocatalytic activity increased in the order: TO–PN2 > TO–CNH3 > TO–PNH3 , which is in agreement with the order of lattice-nitrogen content (Nfresh ). This proved that the lattice-nitrogen significantly 100
P25 TO-PN2
80
TO-CNH3 TO-PNH3
D/%
TO-PNH3(H2O)
80
-6
Adsorption amount / 10 g/g
P25
60
40
20
0 0
1
2
3
4
t/h Fig. 8. Photocatalytic performances of P25 and prepared N-doped TiO2 samples in the degradation MB under visible light irradiation.
1
2
3
4
t/h Fig. 9. Photocatalytic performances of TO–PNH3 , TO–PNH3 (H2 O), and TO–PNH3 (HCl) in the degradation MB under visible light irradiation.
influenced the visible light activity, which is consistent with the earlier results of Yamada [10]. On the other hand, the stronger absorption in visible light region of TO–PNH3 caused the visible light utilization more effectively, thus leading to the much higher activity than that of TO–PN2 and TO–CNH3 . Nused is calculated and shown in Table 1. Obviously, Nused of TO–CNH3 and TO–PN2 are much lower than that of Nfresh , whereas lattice-nitrogen of TO–PNH3 is relatively stable. It is reported that the lattice-nitrogen was oxidated by photogenerated holes during the degradation reaction, leading to the decrease of lattice-nitrogen content [30]. Therefore, it is deduced that the oxidation of lattice-nitrogen of TO–PNH3 is more difficult than that of another two samples. This difference in lattice-nitrogen stability is probably due to the different preparation method among three samples. Besides, Chen et al. [30] prepared N doped TiO2 by heating TiO2 powders in NH3 flow and found that the presence of surface-adsorbed NH3 decreased the number of surface sites accessible for reactants, resulting in low photocatalytic activity. In this investigation, compared with TO–PNH3 , more NH3 adsorbed on TO–CNH3 surface, thus leading to the lower adsorption capacity and photocatalytic activity of TO–CNH3 . To confirm the detrimental effect of NH3 , Chen et al. [30] washed the prepared N doped TiO2 with pure water for several times to remove adsorbed NH3 . The photocatalytic activity of obtained sample was improved after washing, but still much lower than that of postcalcination sample (NT400). This is probably due to that NH3 was not removed completely by washing with pure water. In this investigation, TO–PNH3 was washed with HCl (0.1 M) to remove the adsorbed NH3 , and then cleaned with deionized water. The obtained sample was denoted as TO–PNH3 (HCl). For comparison, TO–PNH3 (H2 O) was obtained by washing TO–PNH3 with deionized water directly. The FT-IR results (not shown) indicated that adsorbed NH3 were removed completely after HCl washing, whereas residual NH3 still existed on TO–PNH3 (H2 O) surface. The photocatalytic performances (Fig. 9) show that the activity increased in the order: TO–PNH3 < TO–PNH3 (H2 O) < TO–PNH3 (HCl), which confirmed NH3 detrimental effect on photocatalytic activity. When TO–PN2 and TO–CNH3 were used to replace TO–PNH3 following the same procedure as in the preparation of TO–PNH3 (HCl), the product were denoted as TO–PN2 (HCl) and TO–CNH3 (HCl), respectively. The photocatalytic performances of TO–PN2 (HCl), TO–CNH3 (HCl), and TO–PNH3 (HCl) were investigated in three cycles to check the photocatalytic stability (Fig. 10). It is shown that the activity of TO–PNH3 (HCl) decreased slightly in 1st reuse and kept stable in the next two cycles. However, for TO–PN2 (HCl) and
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120 110
TO-PNH3(HCl)
TO-CNH3(HCl)
+
N 1s
TO-PN2(HCl)
after Ar etching
100 90
Intensity / a.u.
D/%
80 70 60 50 40
TO-PNH3(HCl)
30
TO-CNH3(HCl)
20 10 0
TO-PN2(HCl) fresh
1st reuse
2nd reuse
3rd reuse
Fig. 10. Photocatalytic stability of prepared N-doped TiO2 samples in the degradation of MB.
TO–CNH3 (HCl), the activities decreased gradually, from 41.8% and 55.1% for fresh catalyst to 30.9% and 49.6% for 3rd reused catalyst. This hinted that the photocatalytic stability of TO–PNH3 (HCl) was much better than TO–PN2 (HCl) and TO–CNH3 (HCl). It is proposed that this difference in photocatalytic stability is attributed to the different lattice-nitrogen stability among the three samples. Therefore, the lattice-nitrogen contents of fresh and reused TO–PN2 (HCl), TO–CNH3 (HCl), and TO–PNH3 (HCl) were calculated according to the relevant XPS data (Table 2). The lattice-nitrogen contents of TO–PN2 (HCl) and TO–CNH3 (HCl) decreased gradually from 1.28 at.% and 1.62 at.% to 0.46 at.% and 1.04 at.% after three cycles which confirmed that the lattice-nitrogen significantly influenced the visible light activity. However, lattice-nitrogen content of TO–PNH3 (HCl) decreased slightly from 1.94 at.% to 1.78 at.% for the 1st reuse, and then kept stable in the next two cycles. This indicated that the lattice-nitrogen atoms in TO–PNH3 (HCl) remained relatively stable. As mentioned above, Nused of TO–CNH3 and TO–PN2 are much lower than that of Nfresh , whereas lattice-nitrogen of TO–PNH3 is stable (Table 1). This difference in lattice-nitrogen stability is probably due to the different preparation method among three samples. In order to elucidate why the photocatalytic stability is different among the samples, the XP spectra of fresh TO–PN2 (HCl), TO–CNH3 (HCl), and TO–PNH3 (HCl) in N1s region after Ar+ ion etching were measured and shown in Fig. 11. Apparently, the surface adsorbed NH3 species located at 395.3 eV and N–N (N–O) species located at 400 eV were removed after Ar+ ion etching to get rid of the surface layer. Only one peak around 396 eV, which attributed to lattice-nitrogen was observed for all the three samples. The calculation according to the relevant XPS data revealed that the lattice-nitrogen content of TO–PN2 (HCl), TO–CNH3 (HCl), and TO–PNH3 (HCl) were 0.32, 0.74, and 1.58 at.%, respectively. Compared with the data of fresh catalyst in Table 2, more than 50% and 75% lattice-nitrogen in TO–CNH3 (HCl) and TO–PN2 (HCl) was eliminated after Ar+ ion etching. This indicated that a great number of N atoms doped only into the surface layer of TO–CNH3 (HCl) and TO–PN2 (HCl), which were oxidated easily by photogenerated holes Table 2 Lattice-nitrogen content of fresh and reused TO–PNH3 (HCl), TO–CNH3 (HCl), and TO–PN2 (HCl) determined by XPS data. Sample
Fresh catalyst (at.%)
1st recycle (at.%)
2nd recycle (at.%)
3rd recycle (at.%)
TO–PNH3 (HCl) TO–CNH3 (HCl) TO–PN2 (HCl)
1.94 1.62 1.28
1.78 1.31 1.02
1.78 1.22 0.75
1.75 1.04 0.46
390
395
400
405
Binding Energy / a.u. Fig. 11. XP spectra of TO–PN2 (HCl), TO–CNH3 (HCl), and TO–PNH3 (HCl) in the region of N1s after Ar+ ion etching.
Table 3 Comparison of lattice-nitrogen stability of N-doped TiO2 samples before and after HCl solution washing. Sample
Fresh catalyst (at.%)
1st reuse (at.%)
Retention ratea
TO–PNH3 (HCl) TO–CNH3 (HCl) TO–PN2 (HCl) TO–PNH3 TO–CNH3 TO–PN2
1.94 1.62 1.28 1.95 1.64 1.32
1.78 1.31 1.02 1.91 1.17 0.76
0.92 0.81 0.80 0.98 0.71 0.58
a Retention rate is equal to the ratio of lattice-nitrogen content in 1st reused catalyst to that of fresh catalyst.
during the degradation reaction, leading to the decrease of latticenitrogen content. Therefore, TO–PN2 and TO–CNH3 exhibited the poor photocatalytic stability. On the contrary, compared with the data of fresh TO–PNH3 (HCl) in Table 2, less than 20% lattice-nitrogen of TO–PNH3 (HCl) was removed after Ar+ ion etching. This is probably due to that the excited hydrogen species produced by NH3 plasma made the N atoms doped into crystal lattice of deeper layer, thus caused it oxidated difficulty by photogenerated holes. Therefore, the photocatalytic stability of TO–PNH3 (HCl) was much higher than that of TO–PN2 and TO–CNH3 . The retention rate of lattice-nitrogen, which represents the lattice-nitrogen stability is calculated and shown in Table 3. Compared with the sample before HCl washing, more than 10% and 20% enhancement of retention rate are observed in TO–CNH3 (HCl) and TO–PN2 (HCl), whereas only slight decrease of retention rate is shown in TO–PNH3 (HCl). This indicated the lattice-nitrogen stability of N-doped TiO2 samples improved after HCl solution washing. This is probably due to that the surface N species absorbed on Brönsted acid sites (–OH) were removed by HCl solution, leading to more surface hydroxy groups are available to trap the photogenerated holes, thus restrain the oxidation of lattice-nitrogen by photo-generated holes. 4. Conclusion NH3 plasma, N2 plasma, and annealing in flowing NH3 were used to prepare N doped TiO2 respectively to investigate the influence of preparation method, nitrogen source, and post-treatment on the photocatalytic activity and stability. The photocatalytic activity increased in the order: TO–PN2 < TO–CNH3 < TO–PNH3 , indicating NH3 plasma is most effective among the three
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methods. The lattice-nitrogen significantly influenced the visible light activity. NH3 which adsorbed on catalyst surface led to the low adsorption capacity for reactant MB, thus decreased photocatalytic activity. After removal NH3 by HCl washing, the obtained catalysts exhibited much higher activities under visible light, which confirmed the detrimental effect of NH3 . The photocatalytic stability of N doped TiO2 prepared by NH3 plasma was much higher than that of samples prepared by other nitridation procedures. This is proposed that the excited hydrogen species produced by NH3 plasma made the N atoms doped into crystal lattice of deeper layer, which caused it oxidated more difficult by photogenerated holes than other catalysts, thus leading to the stable lattice-nitrogen. Besides, the lattice-nitrogen stability of N-doped TiO2 samples improved after HCl solution washing. This is probably due to that the surface N species absorbed on Brönsted acid sites (–OH) were removed by HCl solution, leading to more surface hydroxy group are available to trap the photo-generated holes, thus restrain the oxidation of lattice-nitrogen. This stable lattice-nitrogen during the degradation reaction caused the high photocatalytic stability. Acknowledgments This work was supported by National Natural Science Foundation of China (no. 41071317, 30972418), National Key Technology R & D Programme of China (no. 2007BAC16B07), the Natural Science Foundation of Liaoning Province (no. 20092080). The authors would like to thank Prof. Anjie Wang, Dalian University of Technology, for the contribution to the manuscript. References [1] A. Fujishima, T.N. Rao, D.A. Tryk, Titanium dioxide photocatalysis, J. Photochem. Photobiol. C 1 (2000) 1–21. [2] M.R. Hoffmann, S.T. Martin, W. Choi, D.W. Bahnemann, Environmental applications of semiconductor photocatalysis, Chem. Rev. 95 (1995) 69–96. [3] R. Asahi, T. Morikawa, T. Ohwaki, A. Aoki, Y. Taga, Visible-light photocatalysis in nitrogen-doped titanium oxides, Science 293 (2001) 269–271. [4] T. Lindgren, J.M. Mwabora, E. Avendano, J. Jonsson, A. Hoel, C.G. Granqvist, S.E. Lindquist, Photoelectrochemical and optical properties of nitrogen doped titanium dioxide films prepared by reactive DC magnetron sputtering, J. Phys. Chem. B 107 (2003) 5709–5716. [5] M. Qiao, S.S. Wu, Q. Chen, J. Shen, Novel triethanolamine assisted sol–gel synthesis of N-doped TiO2 hollow spheres, Mater. Lett. 12 (2010) 1398–1400. [6] H. Shen, L. Mi, P. Xu, W.D. Shen, P.N. Wang, Visible-light photocatalysis of nitrogen-doped TiO2 nanoparticulate films prepared by low-energy ion implantation, Appl. Surf. Sci. 17 (2007) 7024–7028. [7] L. Zhao, Q. Jiang, J.S. Lian, Visible-light photocatalytic activity of nitrogen-doped TiO2 thin film prepared by pulsed laser deposition, Appl. Surf. Sci. 15 (2008) 4620–4625. [8] S.Z. Hu, A.J. Wang, X. Li, H. Löwe, Hydrothermal synthesis of well-dispersed ultrafine N-doped TiO2 nanoparticles with enhanced photocatalytic activity under visible light, J. Phys. Chem. Solid 71 (2010) 156–162.
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Journal of Hazardous Materials 196 (2011) 255–262
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Enhanced photocatalytic activity of Bi2 WO6 with oxygen vacancies by zirconium doping Zhijie Zhang, Wenzhong Wang ∗ , Erping Gao, Meng Shang, Jiehui Xu State Key Laboratory of High Performance Ceramics and Superfine Microstructures, Shanghai Institute of Ceramics, Chinese Academy of Sciences, 1295 Dingxi Road, Shanghai 200050, PR China
a r t i c l e
i n f o
Article history: Received 3 July 2011 Received in revised form 5 September 2011 Accepted 6 September 2011 Available online 10 September 2011 Keywords: Zr4+ -doped Bi2 WO6 Oxygen vacancy Photocatalysis RhB Phenol
a b s t r a c t To overcome the drawback of low photocatalytic efficiency brought by electron–hole recombination, Bi2 WO6 photocatalysts with oxygen vacancies were synthesized by zirconium doping. The oxygen vacancies as the positive charge centers can trap the electron easily, thus inhibiting the recombination of charge carriers and prolonging the lifetime of electron. Moreover, the formation of oxygen vacancies favors the adsorption of O2 on the semiconductor surface, thus facilitating the reduction of O2 by the trapped electrons to generate superoxide radicals, which play a key role in the oxidation of organics. Visible-light-induced photodegradation of rhodamine B (RhB) and phenol were carried out to evaluate the photoactivity of the products. The results showed that oxygen-deficient Bi2 WO6 exhibited much enhanced photoactivity than the Bi2 WO6 photocatalyst free of oxygen deficiency. This work provided a new concept for rational design and development of high-performance photocatalysts. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Semiconductor-based photocatalysis has been attracting a great deal of attention due to its potential applications in renewable energy and environment fields such as dye-sensitized solar cells, hydrogen generation from water splitting and photocatalytic water/air purification [1–5]. Because these applications are based on the photogeneration of charge carriers such as electrons and holes, success in the applications relies on the transfer efficiency of electron or hole, which is closely related to the recombination rates of the photogenerated charge carriers. Unfortunately, due to the much faster recombination rate (nanoseconds) than the interfacial transfer rate (microseconds to milliseconds), many charge carriers recombine and dissipate the input energy as heat, which seriously limits the overall quantum efficiency for photocatalysis [6]. Therefore, to improve the photocatalytic activity of the semiconductors, it is important to control the recombination dynamics of the photogenerated charge carriers. If a suitable scavenger or surface defect state is available to trap the electron or hole, recombination is inhibited and ensuing redox reaction may occur. It was reported that oxygen vacancies may act as electron capture centers, and thus play an important role in retarding the recombination of charge carriers, which can lead to an enhanced photocatalytic activity of the photocatalysts [7–9]. Moreover, the existing oxygen vacancies
can act importantly as specific reaction sites for reactant molecules in heterogeneous reactions [10]. Therefore, introducing oxygen vacancies into the photocatalysts can be a feasible approach for developing highly active photocatalysts. As one of the simplest Aurivillius oxides with layered structure, Bi2 WO6 has recently attracted considerable attention for its good photocatalytic performance in water splitting and organic contaminant decomposing under visible light irradiation [11–15]. Up to now, much work has been done to facilitate the electron–hole separation and enhance the photocatalytic activity of Bi2 WO6 , including surface modification [16,17], anion doping [18], and coupling with other semiconductors [19,20]. However, to the best of our knowledge, the effect of oxygen vacancies on the photocatalytic activity of Bi2 WO6 has seldom been reported. Here for the first time we introduce oxygen vacancies into Bi2 WO6 through zirconium doping and the relationship between oxygen vacancies and the photocatalytic activity of Bi2 WO6 has been investigated. Bi2 WO6 does not contain oxygen vacancies and it was reported that substitution of W by appropriate cations with lower valence states could lead to an extrinsic oxygen deficiency by charge compensation [21,22]. Defect calculations show that the low solution energy (0.05 eV) is favorable for the substitution of ZrIV at WVI site with the creation of extrinsic oxygen vacancies [22], which may be described by defect reactions written as: WO
3 ZrO2 −→Zr
∗ Corresponding author. Tel.: +86 21 5241 5295; fax: +86 21 5241 3122. E-mail address:
[email protected] (W. Wang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.017
W
+ VO •• + 2OO
In this study, we succeeded in preparing oxygen deficient Bi2 WO6 phases by substitution for WVI with ZrIV . The photoactivity
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Fig. 1. (A) The XRD patterns of the as-synthesized products; (B) diffraction peak positions of the (1 3 1) plane in the range of 2 = 27.5–29◦ .
evaluation, via the photocatalytic degradation of RhB and phenol under visible light, demonstrated that the photocatalytic activity is dependent on the concentration of oxygen vacancy, and the Zr4+ -doped Bi2 WO6 exhibit much better photocatalytic performance than undoped Bi2 WO6 sample. Moreover, the role of oxygen vacancy in promoting the separation of charge carriers and enhancing the photocatalytic activities is elucidated in detail.
2. Experimental 2.1. Preparation of Zr4+ -doped Bi2 WO6 photocatalysts The Zr4+ -doped Bi2 WO6 photocatalysts were prepared by a hydrothermal method. In a typical process, 2 mmol of Bi(NO3 )3 ·5H2 O and 1 mmol of Na2 WO4 ·2H2 O were dissolved in 2 mL of 2 M nitric acid and 30 mL of deionized water, respectively. After that, these two solutions were mixed together and stirred for 30 min. Then aqueous solution containing desired amounts of ZrOCl2 ·8H2 O was added for Zr4+ -doped Bi2 WO6 . The molar ratios of Zr to Bi2 WO6 were set as 0, 2.0%, 3.0% and 4.0%, respectively, and the corresponding products were named as Zr-0, Zr-0.02, Zr0.03 and Zr-0.04. The pH value of the final suspension was adjusted to about 7 and the mixture was stirred for several hours at room temperature. Afterward, the suspensions were added into a 50 mL Teflon-lined autoclave up to 80% of the total volume. The autoclave was sealed in a stainless steel tank and heated at 160 ◦ C for 24 h. Subsequently, the autoclave was cooled to room temperature naturally. The products were collected by filtration, washed with distilled water for several times, and then dried at 60 ◦ C in air for 12 h.
2.2. Characterization The phase and composition of the as-prepared samples were measured by X-ray diffraction (XRD) studies using an X-ray diffractometer with Cu K␣ radiation under 40 kV and 100 mA and with the 2 ranging from 20◦ to 60◦ (Rigaku, Japan). The morphologies and microstructures of the as-prepared samples were investigated by transmission electron microscopy (TEM, JEOL JEM-2100F). UV–vis diffuse reflectance spectra (DRS) of the samples were recorded with an UV–vis spectrophotometer (Hitachi U-3010) using BaSO4 as reference. Chemical compositions of the derived products were analyzed using X-ray photoelectron spectroscopy (XPS) analysis (Thermo Scientific Escalab 250). All binding energies were referenced to the C 1s peak (284.8 eV) arising from adventitious carbon. The photoluminescence (PL) spectra of the samples were recorded
with a Perkin Elmer LS55. Total organic carbon (TOC) analysis was carried out with an elementar liqui TOC II analyzer. 2.3. Photocurrent measurement Photocurrent measurements were carried out by using a CHI 660C electrochemical workstation. 25 mg of photocatalyst was suspended in de-ionized water (50 mL) containing acetate (0.1 M) and Fe3+ (0.1 mM) as an electron donor and acceptor, respectively. A Pt plate (both sides exposed to solution), a saturated calomel electrode (SCE), and a Pt gauze were immersed in the reactor as working (collector), reference, and counter electrodes, respectively. Photocurrents were measured by applying a potential (+1 V vs SCE) to the Pt electrode using a potentiostat (EG&G). 2.4. Measurement of photocatalytic activities Photocatalytic activities of the Zr4+ -doped Bi2 WO6 photocatalysts were measured by monitoring photo-degradation of rhodamine B (RhB) and phenol in aqueous solution. 100 mg of the photocatalysts were dispersed in a 100 mL solution of RhB (10−5 mol/L) or phenol (20 mg/L). Before illumination, the suspensions were magnetically stirred in the dark for 1 h to ensure adsorption/desorption equilibrium of RhB or phenol with the photocatalyst powders, and then exposed to visible light from a 500 W Xe lamp with a 420 nm cutoff filter. After a certain period of irradiation, 3 mL suspension was sampled and centrifuged to remove the photocatalysts. After that, the supernatant was taken out to measure the absorption spectral change of RhB or phenol through a UV–vis spectrophotometer (Hitachi U-3010) to monitor the photodegradation rate. The concentration change of rhodamine B and phenol were determined by monitoring the optical intensity of absorption spectra at 553 nm and 270 nm, respectively. 3. Results and discussion 3.1. Crystal structure and morphology of the products The XRD diffraction patterns of the pure Bi2 WO6 and Zr4+ doped Bi2 WO6 samples are shown in Fig. 1(A). All of the diffraction peaks match the standard data for a Russellite Bi2 WO6 structure (JCPDS 39-0256), and no characteristic peaks of any impurities are detected in the patterns, which demonstrates that doping with zirconium does not result in the development of new phases. However, a careful comparison of the (1 3 1) diffraction peaks in the range of 2 = 27.5–29◦ (Fig. 1(B)) shows that the peak position of Bi2 WO6 shifts slightly toward a lower 2 value with the increase
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Fig. 2. TEM micrograph of (A) Zr-0 and (B) Zr-0.03.
of zirconium contents. The same results are also presented in other diffraction peaks. According to Bragg’s law, d(h k l) = /(2 sin ), where d(h k l) is the distance between crystal planes of (h k l), is the X-ray wavelength, and is the diffraction angle of the crystal plane (h k l) [23], the decrease in 2 value should result from the increase in lattice parameters (d(1 3 1) value). Because the ionic radius of Zr4+ (0.080 nm) is smaller than that of Bi3+ (0.108 nm) but larger than that of W6+ (0.062 nm), the observed shift of diffraction peak toward lower angles should be due to the larger lattice parameter expected for substitution of W6+ by Zr4+ . In order to obtain detailed information about the microstructure and morphology of the as-synthesized samples, TEM observations
are carried out. Fig. 2(A) and (B) shows the representative TEM images of pure Bi2 WO6 sample and Bi2 WO6 sample doped with a zirconium content of 3.0 mol%, respectively. Both samples exhibit sheet-like morphology, which indicates that zirconium doping has no obvious influence on the morphology of Bi2 WO6 . 3.2. X-ray photoelectron spectroscopic (XPS) analysis The surface composition and elementary oxidation states of the as-prepared sample with a zirconium content of 3.0 mol% is investigated using XPS analysis, and the corresponding experiment results are shown in Fig. 3. The overall XPS spectra shown in Fig. 3(A)
Fig. 3. XPS spectra of Zr-0.03. (A) The overall XPS spectra of the sample; (B) Bi 4f spectrum; (C) W 4f spectrum and (D) Zr 3d spectrum.
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Fig. 4. High-resolution XPS spectra of O 1s for (A) Zr-0 and (B) Zr-0.03.
indicates that all of the peaks on the curve are ascribed to Bi, W, O, Zr and C elements and no peaks of other elements are observed. The presence of C comes mainly from carbon tape used for XPS measurement. Parts B–D of Fig. 3 display the high-resolution spectrum for Bi, W and Zr species, respectively. According to Fig. 3(B), the binding energies of Bi 4f7/2 and Bi 4f5/2 are 159.2 eV and 164.4 eV, respectively, which correspond to the characteristic peak of Bi3+ . The W 4f orbital is clearly resolved into W 4f5/2 and W 4f7/2 contributions, centered upon 37.3 eV and 35.2 eV, respectively (Fig. 3(C)), which are very close to previously reported values [19], suggesting that the tungsten in the Zr4+ -doped Bi2 WO6 sample exists as W6+ . The Zr 3d spectra shown in Fig. 3(D) consists of Zr 3d3/2 and Zr 3d5/2 main peaks with a peak separation of 2.4 eV, which is in agreement with the literature data of Zr4+ [24]. Moreover, we investigated the presence of oxygen vacancies by the XPS spectra. The high-resolution O 1s XPS spectra of undoped Bi2 WO6 and Zr4+ -doped Bi2 WO6 samples were presented in Fig. 4(A) and (B), respectively. Both profiles are asymmetric and can be fitted to two Gaussian features, which are normally assigned as the low binding energy component (LBEC) and the high binding energy component (HBEC), indicating two different kinds of O species in the sample. The LBEC and HBEC can be attributed to the lattice oxygen and chemisorbed oxygen caused by the surface chemisorbed species such as hydroxyl and H2 O, respectively [25]. It has been previously reported that the HBEC component develops with the increase of oxygen vacancies [26], which can lead to the asymmetry of the main peak. The high-resolution O 1s XPS spectra indicate that the peak area of HBEC is obviously larger in the zirconium doped sample as compared to the undoped one. Moreover, the calculated ratios of the adsorbed oxygen to the lattice oxygen are 1.14 and 0.43 for the zirconium doped sample and undoped sample, respectively, which strongly suggests the presence of oxygen deficiencies in the zirconium doped sample.
associated with oxygen vacancies just below the conduction band minimum [27]. The oxygen vacancies are positive charges centers, which bound electrons easily. Excitation of the electrons from such local states to the conduction band can lead to better visible light absorbance. Therefore, with more zirconium dopant concentration, more oxygen vacancies are created, and optical absorption properties of the samples become stronger. 3.4. Photoluminescence spectra and photoelectrochemical measurements Since photoluminescence (PL) emission mainly results from the recombination of free carriers, PL spectra is useful in determining the migration, transfer, and recombination processes of the photogenerated electron–hole pairs in a semiconductor. A weaker PL intensity implies a low recombination rate of the electron–hole under light irradiation [28]. Fig. 6(A) shows the PL spectra of undoped Bi2 WO6 and Zr4+ -doped Bi2 WO6 (Zr-0.03) when the excitation wavelength was 300 nm. There was a significant decrease in the intensity of PL spectra of Zr4+ -doped Bi2 WO6 , which confirmed that zirconium doping could effectively inhibit the recombination of photogenerated charge carriers. Photocurrent reflects indirectly the semiconductor’s ability to generate and transfer the photogenerated charge carriers, which correlates with the photocatalytic activity [29]. To investigate the photo-induced charges separation efficiency of undoped and zirconium doped samples, the photocurrent measurement was carried out under visible light irradiation. As shown in Fig. 6(B), the Zr0.03 sample generates higher photocurrent than undoped Bi2 WO6 ,
3.3. Optical properties of the products The UV–vis diffuse reflectance spectra (DRS) of Zr4+ -doped Bi2 WO6 samples in comparison with pure Bi2 WO6 are shown in Fig. 5. Pure Bi2 WO6 sample presented the photoabsorption ability from the UV light region to the visible light with the wavelength shorter than 450 nm. It was noteworthy that the absorption onset of Zr4+ -doped Bi2 WO6 samples was red-shifted apparently. With the increasing zirconium doping amount, the visible light absorption intensity of the samples became stronger, which may be attributed to excitations of trapped electrons in localized states
Fig. 5. UV–vis diffuse reflectance spectra of the as-prepared samples.
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Fig. 6. (A) The room temperature photoluminescence (PL) spectrum of Zr-0.03 and Zr-0 (Ex = 300 nm); (B) photocurrent generated with visible light irradiation time over Zr-0.03 and Zr-0 suspended with acetate and Fe3+ .
which indicates that zirconium doping can effectively promote the charge carrier transfer and reduce the electron–hole recombination. 3.5. Photocatalytic performance of Zr4+ -doped Bi2 WO6 samples RhB, a hazardous compound as well as a common model pollutant, was chosen as a representative pollutant to evaluate the photocatalytic performance of the photocatalysts. The RhB concentration variation versus the reaction time in the presence of Zr4+ -doped Bi2 WO6 samples compared with pure Bi2 WO6 is plotted in Fig. 7(A). The results demonstrate that the photoactivity of the samples was strongly dependent on the zirconium doping concentration. With zirconium concentration increasing from 0 mol% to 3.0 mol%, the photoactivity of the samples for the RhB photodegradation was enhanced. When the zirconium concentration increased to 4.0 mol%, the photoactivity decreased as compared to that of 3.0 mol%, but still higher than that of undoped Bi2 WO6 . The maximum photoactivity was observed for Zr-0.03, which can degrade RhB completely in 20 min, while only 65.4% of RhB was degraded in the presence of pure Bi2 WO6 within the same time period. Moreover, the comparison of the apparent rate constant k in Fig. 7(B) demonstrated that Zr-0.03 had the highest k value in the photodegradation of RhB, while that of Zr-0.04 decreased compared with that of Zr-0.03. The reason can be interpreted as follows: appropriate amount of oxygen vacancies can trap the electrons, resulting in the holes free to diffuse to the semiconductor surface where oxidation of organic species can occur.
Therefore, appropriate content of oxygen vacancies will improve the photocatalytic process by separating the electron–hole pairs effectively. If it exceeded the optimum value, however, the oxygen vacancies would act as the recombination centers for the photoinduced electrons and holes, which is unfavourable to the photocatalytic performance [27,30]. When the Zr doping concentration was 3.0 mol%, appropriate content of oxygen vacancies were generated and this photocatalyst exhibited the highest photocatalytic activity. When the doping concentration of Zr was increased further, the excess oxygen vacancies generated led to a poor photocatalytic performance. Therefore, appropriate zirconium doping amount can significantly enhance the photocatalytic activity of Bi2 WO6 . Photocatalytic activities of the above-mentioned photocatalysts can be further tested by the degradation of some other organic compound, such as phenol that has no light absorption property in the visible light region and no photosensitization, as shown in Fig. 8. Obviously, upon visible-light irradiation, phenol is degraded more efficiently by Zr-0.03 than by pure Bi2 WO6 (Fig. 8(A)). About 62.5% and 14.2% degradation efficiency was reached within 120 min by Zr-0.03 and pure Bi2 WO6 , respectively. In addition, due to pseudofirst-order kinetics of phenol photodegradation on Bi2 WO6 , the apparent rate constant k is calculated to be 0.0012 min−1 and 0.0083 min−1 for pure Bi2 WO6 and Zr-0.03, respectively (Fig. 8(B)). In other words, the photocatalytic activity of Zr-0.03 is about 7 times that of pure Bi2 WO6 . In order to further investigate the photodegradation of phenol, total organic carbon (TOC), which has been widely used to evaluate
Fig. 7. (A) Photocatalytic degradation of RhB under visible light ( > 420 nm) as a function of irradiation time by the as-prepared samples; (B) the comparison of rate constant k.
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Fig. 8. (A) Photocatalytic degradation of phenol under visible-light irradiation by Zr-0.03 and Zr-0, respectively; (B) the comparison of rate constant k; (C) TOC removal efficiency during the course of photocatalytic degradation of phenol in the presence of Zr-0.03 and Zr-0, respectively; (D) cycling runs in the photocatalytic degradation of phenol under visible-light irradiation.
the degree of mineralization of organic species, was measured in the photodegradation process by the as-prepared samples under visible light, as shown in Fig. 8(C). The results confirm that phenol is steadily mineralized by the as-prepared samples. Moreover, the TOC removal efficiency in the presence of Zr-0.03 reaches a value of 29.2% after 120 min of irradiation, while that of Bi2 WO6 is only 5.3%. Based on the above results, it can be deduced that Zr-0.03 is a much superior photocatalyst to pure Bi2 WO6 . To check the stability of the Zr4+ -doped Bi2 WO6 photocatalyst, the circulating runs in the photocatalytic degradation of phenol were performed under visible light. As shown in Fig. 8(D), after five recycles for the photodegradation of phenol, the catalyst did not exhibit any significant loss of activity, confirming the Zr4+ -doped Bi2 WO6 is not photocorroded during the photocatalytic oxidation of the pollutant molecules, which is especially important for its application.
groups or H2 O to form surface-bound hydroxyl radicals (• OH) and the conduction band electrons can interact with adsorbed O2 to form superoxide radicals (O2 •− ), which are both strong oxidative species and play crucial roles in the oxidative degradation of organics [32,33]. However, for the Bi2 WO6 system, the holes could not react with OH− /H2 O to form • OH due to the more negative redox potential of BiV /BiIII (+1.59 V) than that of • OH/OH− (+1.99 V) [34]. In order to ascertain the active species in the degradation process, holes and hydroxyl radicals scavengers were added into the degradation system. Fig. 9 showed that the addition of isopropanol (IPA) as hydroxyl radicals scavenger [34] caused a minor change in the photocatalytic degradation of phenol, indicating that • OH is not the major oxidation species in this process. However, when holes scavenger, EDTA [35] was introduced, the degradation rates of phenol
3.6. Mechanism of enhanced photoactivities Photocatalysis generally involves four processes [31]: (i) lightinduced generation of conduction band electrons and valence band holes; (ii) transfer of the photogenerated charge carriers to the photocatalyst surface; (iii) subsequent reduction/oxidization of the adsorbed reactants directly by electrons/holes or indirectly by reactive oxygen species; and (iv) recombination of the photogenerated electron–hole pairs. The photocatalysis efficiency is determined by the competition between the charge separation process and the charge recombination process. Desired photocatalysts are expected to promote the charge transfer processes while suppressing recombination process. Typically, for photocatalysts in an aqueous solution, the photoinduced valence band holes can react with chemisorbed hydroxyl
Fig. 9. Photocatalytic degradation of phenol with the addition of hole and hydroxyl radical scavengers and in N2 -saturated solutions under visible light irradiation ( > 420 nm).
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more superoxide anions on the photocatalyst surface. Generation of superoxide anions was a process of trapping the photoinduced electrons, which facilitated the charge separation and resulted in a lower electron–hole recombination rate. Our work suggests that the idea of oxygen vacancies introduction can be a plausible strategy to develop efficient visible-light-driven photocatalysts for environmental remediation. Acknowledgements
Fig. 10. Proposed photocatalytic mechanism of oxygen deficient Bi2 WO6 . OP: organic pollutant; DP: degradation product.
were depressed to a large extent. Therefore, holes play an important part in Bi2 WO6 photocatalysis. The superoxide radical is another important intermediate for oxidative degradation of organics [36,37]. In order to examine the role of superoxide radical in the photocatalysis, photocatalytic degradation of phenol was carried out under N2 -saturated conditions. The result shown in Fig. 9 indicated that under the anoxic condition, the photodegraded rate of phenol was largely suppressed, suggesting that superoxide radical is an important oxidation species in the photocatalytic process. Oxygen molecules as the electron scavengers play a crucial role in photocatalysis by reacting with electrons to generate superoxide radicals. However, when the rate of O2 reduction by electrons is not sufficiently fast to match the rate of reaction of holes, an excess of electrons will accumulate on the photocatalyst particles, and the electron–hole recombination rate will increase consequently. In this case, electrons transfer to O2 may be the rate limiting step in photocatalysis [38,39]. This, however, can be overcome by the introduction of oxygen vacancies into the photocatalyst. In order to facilitate the reaction between oxygen and electrons, strong oxygen adsorption on the photocatalyst surface and longer lifetime of electrons are indispensable. It was reported that adsorption of O2 molecules is mainly mediated by oxygen vacancies and oxygen physisorbs on defect-free oxide surfaces but interacts strongly with oxygen vacancies [40,41]. So the adsorption of oxygen can be promoted by the formation of oxygen vacancies. On the other hand, if an electron is freely mobile in the semiconductor particle, it is hardly possible for it to escape destiny of recombination. However, this can be prevented if the electrons are transiently but efficiently trapped in the particles. The positively charged VO •• defects can work as electron acceptors and can trap the photogenerated electrons temporarily to reduce the surface recombination of electrons and holes [8,9]. Therefore, the creation of oxygen vacancies can not only favor oxygen adsorption but also retard the recombination of charge carriers, which facilitates the O2 reduction rate to generate more superoxide anions (O2 •− ) on the photocatalyst surface (Fig. 10), and thus lead to an enhanced photocatalytic activity. 4. Conclusions Oxygen-deficient Bi2 WO6 photocatalysts were synthesized by zirconium doping, and the relationship between oxygen vacancies and photocatalytic activities of Bi2 WO6 was investigated. The visible-light-induced photo-degradation of RhB and phenol demonstrated that zirconium doping could significantly enhance the photocatalytic performance of Bi2 WO6 . The higher photocatalytic activity was attributed to the formation of oxygen vacancies, which promote the O2 adsorption and O2 reduction rate, leading to
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Journal of Hazardous Materials 196 (2011) 295–301
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Monitoring of PCBs at facilities related with PCB-containing products and wastes in South Korea Guang-Zhu Jin a,1 , Ming-Liang Fang a , Jung-Ho Kang a , Hyokeun Park a , Sang-Hyup Lee b , Yoon-Seok Chang a,b,∗ a b
School of Environmental Science and Engineering, Pohang University of Science and Technology (POSTECH), San 31, Hyoja-dong, Nam-gu, Pohang 790-784, Republic of Korea Water Research Center, Korea Institute of Science and Technology (KIST), Hwarangno, 14-Gil 5, Seongbuk-gu, Seoul 136-791, Republic of Korea
a r t i c l e
i n f o
Article history: Received 1 March 2011 Received in revised form 8 September 2011 Accepted 8 September 2011 Available online 28 September 2011 Keywords: PCBs Inventory Emission factor South Korea
a b s t r a c t Polychlorinated biphenyl (PCB) contents were analyzed in samples collected from facilities related to PCB-containing products or wastes in South Korea. Average concentrations of the atmospheric 209 PCBs were 7420 (37.0–104,048) pg m−3 and 16.8 (ND–34.2) fg WHO-TEQ m−3 in indoor air samples; and 1670 (106–13,382) pg m−3 and 5.64 (ND–36.0) fg WHO-TEQ m−3 in outdoor air samples. The highest levels were observed in indoor air samples from disposal facilities (7336–104,048 pg m−3 ), followed by production (330–25,057 pg m−3 ), recycling, and storage facilities, indicating that PCB emissions from PCB-containing products and wastes remains very high and the facilities related with those may be an important source to atmospheric PCBs. Principal component analysis of PCB profiles showed that the homologue patterns of PCBs in outdoor and indoor air samples collected from the facilities were similar to those of boundary air samples and PCB commercial products, e.g. Aroclor 1016, 1221, 1232 and 1242. Evaluation of the PCB mass balance in a facility, dismantling and solvent-washing PCB-contaminated transformers, showed that of the total PCBs treated in this facility, approximately 0.0022% was emitted to the atmosphere, and most was transferred to waste oil for disposal by incineration or chemical methods. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Polychlorinated biphenyls (PCBs) are classified and regulated as one of the 12 persistent organic pollutants (POPs) under the Stockholm Convention on POPs [1]. The sources of PCBs can be divided into two major categories: intentional chemicals produced in the chemical industries, and unintentionally de novo synthesized by-products during thermal processes [2,3]. The production and consumption of global PCBs for industrial purposes are relatively well established. PCBs were mostly produced commercially from 1929 to the early 1970s. During this period, total global production of PCBs was estimated approximately 1.3 million tons [4]. The commercial PCBs are known by a variety of trade names, such as Aroclor (USA, UK), Kanechlor (Japan), Sovol (Russia), Chlophen (Germany, Poland), and Phenoclor (France) [5,6].
∗ Corresponding author at: School of Environmental Science and Engineering, Pohang University of Science and Technology (POSTECH), San 31, Hyoja-dong, Namgu, Pohang 790-784, Republic of Korea. Tel.: +82 54 279 2281; fax: +82 54 279 8299. E-mail address:
[email protected] (Y.-S. Chang). 1 Present address: Key Laboratory of Nature Resource of the Changbai Mountain and Functional Molecular (Yanbian University), Ministry of Education, 133002, Park Road 977, Yanji, Jilin Province, China. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.030
In South Korea, industrial PCB mixtures have never been produced and their use in electronic equipments has been banned since 1979, and their import and use was completely banned in 1996. Kim et al. [7] reported that the ambient air in Korea was more influenced by combustion processes than that in Japan and also the contribution of PCB commercial products was relatively small. PCB levels in iron and steel complexes in South Korea have been reported to be higher than those in residential areas, indicating that iron and steel complexes are probably an important source of PCBs [8]. However, the emission of PCBs caused by de novo synthesis is not believed to contribute significantly to the global historical PCB mass balance [9]. The relative importance of atmospheric emissions from various source categories is not well known with considerable uncertainty [10]. Jamshidi et al. [11] reported that the principal contemporary source of PCBs in UK conurbation was ventilation of indoor air and not volatilization from soil. According to the Korean Law, wastes that contain PCBs (>0.0001 mg kg−1 in solids or >0.01 mg kg−1 in liquids) are considered as “PCB-containing wastes” which must be treated by specialized methods [12]. Recycling of PCB-containing wastes only limits for wastes which contain less than 2 mg kg−1 PCBs. In 2007, the amount of PCB-containing wastes generated in South Korea that were contaminated with >2 mg kg−1 PCBs was 2543 tons [13]. Therefore, the emission of PCBs from PCB-containing products and
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Fig. 1. PCB levels in air samples collected using HVAS from facilities related to PCB-containing products or wastes. Cross bars are mean values and vertical represent maximum and minimum concentrations.
wastes remains very high, even 30 years after PCB production ceased. Emission inventories are essential for identifying, evaluating, and prioritizing sensible control strategies on a regional or a global scale [1,14]. Also, to understand and predict the long range transport features and environmental fates of these substances, quantitative information on their atmospheric releases is deemed essential [15,16]. Hosomi et al. have estimated volatilization of PCBs from PCB-containing ballast in a fluorescent lamp [17]. However, few studies have investigated PCB contamination in facilities related to PCB-containing products and wastes rather than unintentional sources such as incinerators; no such evaluation has been performed in South Korea. In this study, we investigated atmospheric levels and distribution of PCBs in facilities related to PCB-containing products and wastes. These facilities include production, in-use, recycling, storage and disposal facilities across South Korea. We also evaluated PCB emission factors and the mass balance in a PCB disposal facility. The emission of PCBs caused by de novo synthesis was not considered in this study. This is the first study to investigate PCB emissions from facilities related to PCB-containing products and wastes in South Korea, providing valuable data in planning for comprehensive management and final elimination of PCB-containing products and wastes. 2. Experiment and method 2.1. Sampling Air samples were collected from 44 sites (9 production, 8 inuse, 14 storage, 10 recycling, 1 disposal, and 2 boundary) related to PCB-containing products or wastes across South Korea from October 2007 to July 2008 (Fig. S1). The specific information of sampling was shown in the Table S1. Samples were collected using high volume air sampling (HVAS, DHA-1000S, SIBATA). A glass fiber filter (GFF) and two consecutive polyurethane foam plugs (PUF) were used to collect airborne particles and vaporphase PCBs, respectively. Before sampling, the GFFs were baked at 450 ◦ C for 12 h, and the PUF disks were Soxhlet extracted for 16 h with acetone, then for 16 h with dichloromethane, then dried in a desiccator under vacuum for 24 h. A total of 20 outdoor air samples and 39 indoor air samples were collected for 24 h and at a flow rate of 700 L/min. It is important to note that
room sizes of a few in-use facilities were smaller than collected air volumes (1000 m3 ). Therefore, PCB concentrations could be underestimated by dilution effects in those small facilities. Outdoor air samples were collected within 5 m from the facilities (or rooms). Boundary PCB concentrations were measured at sites situated at the boundaries (500–800 m) of facilities. An additional 89 bottom samples were collected at the 37 facilities by wiping floor dust with hexane rinsed glass wool. At each site, 1–3 bottom samples were collected according to its facility size. For mass balance case study, several final product samples, such as copper, silicon steel plate, waste paper, waste oil etc., were collected from a dismantling and cleaning facility of PCB-containing wastes. 2.2. Analytical methods In the laboratory, samples were treated, extracted and analyzed according to the methods established at the US EPA’s method 1668A [18]. Briefly, the samples were spiked with the internal standard containing 27 13 C-labled PCB congeners (1, 3, 4, 15, 19, 37, 54, 77, 81, 104, 105, 114, 118, 123, 126, 155, 156, 157, 167, 169, 188, 189, 202, 205, 206, 208, and 209) (Wellington, 1668-LCS), then Soxhlet-extracted for 24 h using toluene. The extracts were then washed with concentrated H2 SO4 followed by hexane-saturated H2 O. Sample cleanup was performed using multi-layer silica and florisil columns. The eluent was reduced to 0.5 mL by rotary evaporation and a gentle stream of N2 gas. Finally, the extracts were transferred to GC vials, and 13 C-labled PCBs (9, 28, 52, 101, 111, 138, 178, and 194) were added as recovery standards. PCB contents were analyzed using an Agilent Hewlett-Packard 6890 gas chromatograph/Jeol JMS-700T high resolution mass spectrometer (GC/HRMS) with a DB-5MS column (J&W Scientific, 60 m length, 0.25 mm ID, 0.25 m film thickness). The instrument was operated using He as the carrier gas with a constant flow of 1 mL min−1 . The temperature program of the GC oven was as follows: the temperature was held at the initial value of 110 ◦ C for 2 min, then raised at 40–200 ◦ C min−1 and held for 3 min, then raised at 2–230 ◦ C min−1 , then raised at 7–300 ◦ C min−1 and held for 7 min. 1 L sample was injected at a temperature of between 280 and 300 ◦ C for the analysis of PCB contents. The GC/HRMS was operated under positive EI conditions (38 eV) with a resolution of 10,000. Data were obtained in the selected ion monitoring (SIM) mode.
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Fig. 2. PCB levels in bottom samples. Cross bars are mean values and vertical represent maximum and minimum concentrations.
Peak assignment was conducted to quantify 209 PCB congeners, but typically only 120 PCB congeners were detected. Bottom samples were analyzed using a GC with an electron capture detector (HP 6890, Agilent) following the Korean waste official method [19].
3. Results and discussion 3.1. PCB levels Detected levels of PCBs were generally lower in outdoor air samples than in indoor air samples, although the range was very large (Table 1, Fig. 1). The mean PCB concentration (209 PCBs) in outdoor air samples was 1670 pg m−3 (5.64 fg WHO-TEQ m−3 ) and ranged from 106 pg m−3 at a PCB-containing waste storage site to 13,400 pg m−3 at a PCB disposal (dismantling and cleaning) facility. These PCB concentrations in outdoor air samples were consistent with the PCB levels in the ambient air of South Korea in a previous study [7] and comparable with those in global urban sites (mean: 1700 pg m−3 ) [20]. The mean PCB concentrations of indoor air samples were 7420 pg m−3 (16.8 fg WHO-TEQ m−3 ) and ranged from 37 pg m−3 at an indoor transformer site (containing 46.6 tons of transformer oil contaminated with 0.15 mg kg−1 PCBs) to 104,048 pg m−3 at a PCB disposal facility. PCB levels in both indoor and outdoor air samples were highest at the disposal facility followed by the production facility, the recycling facility, the storage site and the in-use site. The high concentration at the PCB disposal facility might be due to the volatilization of
2.3. Quality assurance/quality control Several steps were taken to obtain data that would allow an assessment of the accuracy and reliability of the data. Analytical blanks were included at a rate of one per 10 samples. All data have been blank corrected. The average recoveries of 27 13 C-labled PCB congeners ranged from 25 to 93% (Table S2), which satisfied the criteria (25–150%) recommended by US EPA method 1668A. Recovery statistics are given in Table S6. The method detection limit was calculated as 3 times the standard deviation of seven blank replicate samples (Table S3). The criteria for the quantification of analytes were as follows: retention time within 2 s of that of the standard, isotope ratio within 20% of that of the standard, and signal-to-noise ratio ≥3.
Table 1 PCB levels (209 PCBs) in indoor and outdoor air samples from different sites using HVAS. Site type
Sample type
Mean PCB concentration 3
Boundary Production In use (indoor) In use (indoor) In use (outdoor) Storage (indoor) Storage (indoor) Storage (outdoor) Recycling Recycling Disposal Disposal Background Industrial Residential In use condenser (containing PCBs) Industrial area Urban area Background
Outdoor (n = 2) Indoor (n = 9) Indoor (n = 5) Outdoor (n = 3) Outdoor (n = 3) Indoor (n = 11) Outdoor (n = 9) Outdoor (n = 1) Indoor (n = 10) Outdoor (n = 1) Indoor (n = 4) Outdoor (n = 2) Outdoor
Indoor Outdoor (n = 3) Outdoor (n = 3) Outdoor
References −3
(pg/m )
(fg WHO-TEQ m
153 (146–161) 8722 (330–25,057) 788 (37–2273) 815 (199–1159) 1017(671–1561) 1469 (353–5404) 790 (106–2527) 2539 (2539–2539) 5731 (160–17,710) 1667 (1667–1667) 33,692 (7336–104,048) 8257 (3131–13,382) 180–280 2080–5820 240–28,000 26,000–110,000
1.44 (0.029–2.85) 11.50 (0.073–31.0) 4.79 (0.157–12.3) 8.20 (ND–24.6) 8.76 (ND–14.9) 0.72 (ND–3.74) 6.81 (ND–36.0) 0.033 6.05 (0.018–41.3) ND 114 (ND–342) 2.31 (1.70–2.91)
21–27 19–46 0.6 (1–1.9)
) This study This study This study This study This study This study This study This study This study This study This study This study Kim et al. [7]
Hosomi [17] Martínez et al. [29] Menichini et al. [30]
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more-volatile PCB congeners during transformer dismantling and solvent washing. PCB levels were also investigated in 89 bottom samples collected from 36 facilities (Fig. 2). PCBs were detected in about 80% of bottom samples. PCB concentrations ranged from ND to 342 ng cm−2 and the highest level was found in samples from a transformer oil production site (ND–342 ng cm−2 ), followed by a disposal facility (4–152 ng cm−2 ). PCB concentrations of 26,000–110,000 pg m−3 have been observed in the indoor air of an office where fluorescent lamps with PCB-containing ballast had been used [17]; the PCB volatilization rates from this ballast were temperature-dependent and the PCB composition of the emission gas was similar to that observed in the ballast samples collected. Our samples were not collected to quantify the effect of temperature on PCB levels in indoor air samples; however, the highest PCB concentrations were observed in samples collected in a disposal facility during summer (104,000 pg m−3 in July vs. 7340 pg m−3 in October), followed by a production facility (25,100 pg m−3 ). In a major UK conurbation, the principal contemporary source of PCBs has been reported to be not the volatilization from soil but the ventilation of indoor air; existing structures, especially older buildings in which PCBs had been used in the past, were the major source of PCBs in outdoor air [11]. Generally, urban areas are more polluted by PCBs than rural areas [21]. Based on our data and previous studies, it seems that PCBs volatilized from the PCBcontaining products or wastes are important sources of PCBs in the ambient air in South Korea. Future reductions in PCB concentrations in the outdoor air and ultimately in human exposure may be best achieved by actions to these remaining sources of PCBs from PCB-containing products and wastes. In our air samples, the average contribution of gas phase PCBs to total PCBs was about 96%, which was consistent with the previous study of 24 PCB congeners in South Korea [22]. Gas-particle partitioning of PCBs in air samples from each type of facilities related to PCB-containing products or wastes showed similar patterns (Fig. S5). Gas phase contribution to total PCBs decreased from 89 (mono-CBs) to 24% (deca-CB) in indoor air samples, and from 92 (mono-CBs) to 22% (deca-CB) in outdoor air samples (Fig. S6). These results suggest that PCBs in the air exist predominantly in the gas phase and that the contribution of PCB congeners to the gas phase decreases as congeners become more-highly chlorinated (i.e., less volatile). 3.2. Homologue patterns The homologue patterns of PCBs in air samples were similar at all sampling sites (Fig. S2). In all air samples, the dominant PCB homologues found were low chlorinated PCBs such as mono-, diand tri-PCBs which accounted for about 14%, 35%, and 33% of total PCBs in indoor air samples and about 15%, 41%, and 30% of total PCBs in outdoor air samples, on average, respectively (Fig. S3). Many sources of PCBs can influence atmospheric PCB levels, including incinerators, industrial thermal processes, and PCBcontaining products and wastes [23]; homologue patterns can provide clues to where and how these substances originated [24]. PCBs in the outdoor air in this study were apparently influenced by the indoor air PCBs due to the higher levels of PCBs in the indoor environment. For further source identification, principal component analysis (PCA) of the data was conducted using SPSS 12.0 software (SPSS, Inc.) and homologue patterns of air samples from this study were compared to those of commercial mixtures of Aroclor 1016, 1221, 1232, 1242, 1248, 1254, 1260, 1262 and 1268 from other studies [6,25]. Total concentrations of each homologue PCBs (i.e. 1 Cl, 2 Cls, . . ., 10 Cls) were used for PCA analysis. PCB data were normalized by dividing by the total PCB concentrations for each sample, producing data ranging from 0 to 1. Finally, these normalized PCB compositions were used as input data for PCA. As a result,
PC1 and PC2 accounted for 60% of the total variance (Fig. 3). In the loading plot, the variables are well grouped by the number of chlorine. The homologue patterns of PCBs in air samples from various sites in this study were similar to commercial mixtures such as Aroclor 1016, 1221, 1232, and 1242, suggesting that the homologue patterns of many air samples were simultaneously influenced by these commercial mixtures. It is consistent with another previous study that the homologue patterns of PCBs found in sediments in South Korea indicated that their sources were commercial mixtures such as Aroclor 1016, 1242, 1254, and 1260 or corresponding Kanechlor products [26]. PCA was also used to compare homologue patterns of air samples in this study to those of ambient soil [27], incineration flue gas, and cement plant flue gas samples from other researches [6,28]. As a result, PC1 and PC2 accounted for 90% of the total variance (Fig. 4). The homologue patterns of PCBs in our air samples from various sites in South Korea were different from those of ambient soil, incineration flue gas, and cement plant flue gas samples, suggesting that there are other significant sources. However, all our air samples including boundary air samples had similar homologue patterns with the general ambient air samples (Korea, n = 15; Japan, n = 11) [7], indoor air samples from a disposal facility (Japan, n = 5), and indoor air samples of a room where PCB-containing sealant was used (Japan, n = 3) [6] (Fig. S4). Kim et al. [7] reported that the PCB levels in the ambient air of South Korea were more influenced by combustion processes than that in Japan, and also that the contribution of commercial PCB products was relatively small. However, our results strongly suggest that the ambient air in South Korea is contaminated by mixtures of commercial Aroclor products with various chlorine contents, particularly lowly chlorinated mixtures. 3.3. A case study of PCB mass balance To evaluate the PCB mass balance, a PCB disposal facility was selected, where PCBs in the indoor air showed the highest level. A series of air, bottom, and final product samples were collected from each process of dismantling and cleaning. This facility mainly treats waste transformers which contain PCB-contaminated transformer oil (>2 mg kg−1 ). The main processes are removal of transformer oil, dismantling of outer transformer cases, first extraction with toluene, dismantling of transformer inner assemblies, and second extraction with toluene. The final products are recycled (metals) or passed on for further disposal (waste oil). The mean PCB concentrations were measured as 14,560 pg m−3 in indoor air samples and 8130 pg m−3 in outdoor air samples (Fig. 5). Since all PCB sampling was conducted only in autumn season, no seasonal variation of PCBs in outdoor air was reflected. This is one of the limitations in the present study. Limited field monitoring data need further study in the future. If the concentrations of PCBs in the indoor air are relatively constant over the operation period, air PCB equilibrium between indoor and outdoor air can be described by the following Eq. (1): V
dC1 = Ea Q − VR (C1 − C0) = 0 dt
(1)
and the PCB air emission factor Ea in this facility can be calculated using Eq. (2), Ea =
(C1 − C0)VR Q
(2)
where V is the volume of the room (m3 ), C1 is the concentration of PCBs in the indoor air (ng m−3 ), C0 is the concentration of PCBs in the outdoor air (ng m−3 ), t is time, Q is the quantity of PCBs treated in the facility, and R is the natural air exchange rate at which outdoor air replaces the whole indoor air. Generally, natural air exchange rate of the concrete building is 7–24 day−1 [17],
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Fig. 3. Comparison of homologue patterns of air samples from this study with commercial mixtures of PCBs.
Fig. 4. Comparison of homologue patterns of air samples from this study with samples from other sources.
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PCB-containing products and wastes in South Korea. The total PCB concentrations ranged from 37.0 to 104,048 pg m−3 in indoor air samples and from 106 to 13,382 pg m−3 in outdoor air samples. The homologue patterns of PCBs in outdoor and indoor air samples collected from various facilities were similar to those of boundary air samples and the PCB commercial mixtures of Aroclor 1016, 1221, 1232 and 1242. These results suggest that PCB emissions during the production, recycling, in-use and disposal of PCB-containing products and wastes can be an important source of atmospheric PCBs. Therefore, it provides valuable data in planning for comprehensive management and final elimination of PCB-containing products and wastes. Fig. 5. PCB mass balance in a PCB disposal facility.
Acknowledgements which was used in this study. Using the data observed in this facility (Table S4), Ea was estimated to be 9.8 × 10−4 to 3.4 × 10−3 g-PCB gPCB−1 yr−1 . This value is comparable or slightly higher than that reported in a previous study [4]; Ea ranged from 1.58 × 10−5 to 2.56 × 10−2 g-PCB g-PCB−1 yr−1 for open in-use and storage sites, and from 3.38 × 10−9 to 5.22 × 10−4 g-PCB g-PCB−1 yr−1 for closed in-use and storage sites [9]. In the previous study, Breivik et al. [4] has reported emission factors of only 22 PCB congeners, which have high uncertainties. The specific amounts of 22 congeners in the Aroclor mixtures in South Korea are not available. However, direct comparison of PCB air emission factors reported in this study with those above might be reasonable, because the 22 congeners were dominant congeners of PCBs and both air emission factors had the same unit. The uncertainty of air emission factor calculation in this study mainly comes from natural air exchange rate, variation of PCB concentrations in indoor and outdoor air samples, temperature etc. PCB bottom emission factor Eb in this facility can be calculated using Eq. (3): Eb =
Cb A Q
(3)
where Cb is the average concentration of PCBs in bottom samples, and A is the area of the facility. Using the data measured in this facility, Eb was estimated for bottom samples to be 3.0 × 10−4 gPCB g-PCB−1 yr−1 . The PCB mass balance in this disposal facility was calculated (Fig. 5). There are major uncertainties, like Breivik et al. [9], which mainly come from natural air exchange rate, variation of PCB concentrations in air samples and temperature etc. Of the total PCBs disposed in this facility, approximately 0.0022% was emitted to the atmosphere and 0.03% was deposited to the indoor bottom as dust particles or transformer oil leakage. Meanwhile, most PCBs (98.7%) were transferred as waste transformer oil for later disposal by incineration or chemical treatment. If this facility were to operate at its maximum capacity of 100 tons/week, the estimated maximum 209 PCB emission to the air would be 13 g/yr. This is much smaller than the previous estimation, where estimated PCB emissions to air in South Korea was 199 kg (for 22 PCB congeners, mid scenario, maximum is hundreds-fold of minimum scenario) in the reference year 2008 [4]. Although the production, import and use of PCBs have been banned in South Korea since 1999, however, the amounts of in-use PCB-containing products are still huge. In South Korea, the amount of PCB-containing waste, which is contaminated with >2 mg kg−1 PCBs, was 2543 tons in 2007 [13]. Therefore, atmospheric emission of PCBs from PCB-containing products and wastes still can be a significant source for some period by the time of their complete elimination. 4. Conclusion In this study, we investigated PCBs from various types of facilities and calculated the PCB mass balance in a facility related to
This work was supported by the National Research Foundation of Korea (NRF) grant funded by the Korea government (MEST) (No. 2011-1128723), and partially supported by the Korea Institute of Science and Technology (KIST) as the Institutional Program (2E22173). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.030. References [1] UNEP, The Stockholm convention on persistent organic pollutants (POPs), Chemicals, United Nations Environment Programme, 2001. [2] S. Sakai, M. Hiraoka, N. Takeda, K. Shiozaki, Formation and emission of nonortho CBs and mono-ortho CBs in municipal waste incineration, Chemosphere 29 (1994) 1979–1986. [3] B. Wyrzykowska, N. Hanari, A. Orlikowska, N. Yamashita, J. Falandysz, Dioxinlike compound compositional profiles of furnace bottom ashes from household combustion in Poland and their possible associations with contamination status of agricultural soil and pine needles, Chemosphere 76 (2009) 255–263. [4] K. Breivik, A. Sweetman, J.M. Pacyna, K.C. Jones, Towards a global historical emission inventory for selected PCB congeners – a mass balance approach. 3. An update, Sci. Total Environ. 377 (2007) 296–307. [5] S.K. Shin, T.S. Kim, Levels of polychlorinated biphenyls (PCBs) in transformer oils from Korea, J. Hazard. Mater. 137 (2006) 1514–1522. [6] Y. Ishikawa, Y. Noma, Y. Mori, S.-i. Sakai, Congener profiles of PCB and a proposed new set of indicator congeners, Chemosphere 67 (2007) 1838–1851. [7] K.S. Kim, B.-J. Song, J.-G. Kim, K.-K. Kim, A study on pollution levels and source of polychlorinated biphenyl (PCB) in the ambient air of Korea and Japan, J. KSEE 27 (2005) 170–176. [8] S.-D. Choi, S.-Y. Baek, Y.-S. Chang, Passive air sampling of persistent organic pollutants in Korea, Toxicol. Environ. Health Sci. 1 (2009) 75–82. [9] K. Breivik, A. Sweetman, J.M. Pacyna, K.C. Jones, Towards a global historical emission inventory for selected PCB congeners – a mass balance approach: 2. Emissions, Sci. Total Environ. 290 (2002) 199–224. [10] K. Breivik, R. Alcock, Y.F. Li, R.E. Bailey, H. Fiedler, J.M. Pacyna, Primary sources of selected POPs: regional and global scale emission inventories, Environ. Pollut. 128 (2004) 3–16. [11] A. Jamshidi, S. Hunter, S. Hazrati, S. Harrad, Concentrations and chiral signatures of polychlorinated biphenyls in outdoor and indoor air and soil in a major UK conurbation, Environ. Sci. Technol. 41 (2007) 2153–2158. [12] Korea, Persistent Organic Pollutant Special Management Law, Ministry of Environment, 2008. [13] Korea, Generation and Disposal Status of Designated Wastes, Ministry of Environment, 2008. [14] UN/ECE, The 1998 Aarhus protocol on POPs, United Nations/Economic Council for Europe, 1998 www.unece.org/env/lrtap/pops h1.htm. [15] K. Breivik, B. Bjerkeng, F. Wania, A. Helland, J. Magnusson, Modeling the fate of polychlorinated biphenyls in the Inner Oslofjord, Norway, Environ. Toxicol. Chem. 23 (2004) 2386–2395. [16] H. Hung, C.L. Sum, F. Wania, P. Blanchard, K. Brice, Measuring and simulating atmospheric concentration trends of polychlorinated biphenyls in the Northern Hemisphere, Atmos. Environ. 39 (2005) 6502–6512. [17] M. Hosomi, Volatilization of PCBs from PCB-containing Ballast in Fluorescent Lamp and Indoor PCB pollution: odor of PCBs, J. Japan Assoc. Odor Environ. 36 (2005) 323–330. [18] USEPA, Method 1668, Revision A: Chlorinated Biphenyl Congeners in Water, Soil, Sediment, and Tissue by HRGC/HRMS, 1999. [19] Korea, Official Analysis Method of PCBs in Wastes Samples, Ministry of Environment, 2005.
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Journal of Hazardous Materials 196 (2011) 263–269
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Rhizodegradation gradients of phenanthrene and pyrene in sediment of mangrove (Kandelia candel (L.) Druce) Haoliang Lu a,b , Yong Zhang a,∗ , Beibei Liu a , Jingchun Liu b , Juan Ye b , Chongling Yan b a b
State Key Laboratory of Marine Environmental Science (Xiamen University), Environmental Science Research Center, Xiamen University, Xiamen 361005, Fujian Province, PR China Key Laboratory of Ministry of Education for Coastal and Wetland Ecosystems, and School of Life Sciences, Xiamen University, Xiamen 361005, Fujian Province, PR China
a r t i c l e
i n f o
Article history: Received 21 January 2011 Received in revised form 7 September 2011 Accepted 7 September 2011 Available online 14 September 2011 Keywords: Rhizodegradation Phenanthrene Pyrene Mangrove Kandelia candel (L.) Druce
a b s t r a c t A greenhouse experiment was conducted to evaluate degradation gradient of spiked phenanthrene (Ph, 10 mg kg−1 ) and pyrene (Py, 10 mg kg−1 ) in rhizosphere of mangrove Kandelia candel (L.) Druce. Rhizosphere model system was set up using a self-design laminar rhizoboxes which divided into eight separate compartments at various distances from the root surface. After 60 days of plant growth, presence of the plant significantly enhanced the dissipation of Ph (47.7%) and Py (37.6%) from contaminated sediment. Higher degradation rates of the PAHs were observed at 3 mm from the root zone (56.8% Ph and 47.7% Py). The degradation gradient followed the order: near rhizosphere > root compartment > far-rhizosphere soil zones for both contaminants where mangrove was grown. Contribution of direct plant uptake and accumulation of Ph and Py were very low compared to the plant enhanced dissipation. By contrast, plant-promoted biodegradation was the predominant contribution to the remediation enhancement. The correlation analysis indicates a negative relation between biological activities (microbial biomass carbon, dehydrogenase, urease, and phosphatase activity) and residual concentrations of Ph and Py in planted soils. Our results suggested that mangrove rhizosphere was effective in promoting the depletion of aromatic hydrocarbons in contaminated sediments. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous pollutants persisting in the environment. Anthropogenic inputs of PAHs from oil spills, ship traffic, urban runoff and emission from combustion and industrial processes have caused significant accumulation of PAHs in coastal mangrove wetlands especially those near urban centers and industrial cities [1,2]. Phytoremediation of PAHs is a promising alternative approach to sediment remediation due to its cost effectiveness, convenience and environmental acceptability [1,3]. There are several branches of phytoremediation identified by the USEPA (2000), including phytoextraction, rhizofiltration, phytovolatization, rhizodegradation and phytodegradation, and phytostabilization. Rhizodegradation refers to the microbial breakdown of organic contaminants in the root zone (rhizosphere) soil and sediment. This process uses the natural ability of plants to manipulate the biological, chemical and physical characteristics of the rhizosphere for reducing organic contaminant concentrations in soil and sediment [4,5]. In sediment, rhizoremediation was suggested to be the primary mechanism responsible for PAH
∗ Corresponding author. Tel.: +86 592 2188685; fax: +86 592 2184977. E-mail address:
[email protected] (Y. Zhang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.031
degradation in plant-assisted remediation efforts [6,7]. In this case, roots contribute to the dissipation of hydrocarbon contaminants through an increase in the number of microbes, improvement of physical and chemical soil conditions, increased root exudates and humification, and adsorption of pollutants in the rhizosphere was investigated. Mangrove ecosystems, important inter-tidal estuarine wetlands along coastlines of tropical and subtropical regions, are closely tied to industrial activities and are subject to contamination [8,9]. Mangrove may contribute to the dissipation of organic contaminants through an increase in the number of microbes, improvement of physical and chemical soil conditions, increased humification and adsorption of pollutants in the rhizosphere, but the impact of each process has not been clearly elucidated. A number of bacterial strains able to degrade PAHs have been isolated from surface mangrove sediment and the degradation of PAHs by these consortia and isolates in culture medium and in sediment slurry have been studied [10,11]. Nevertheless, the question of how far a mangrove rhizosphere effect on degradation of PAHs may extend has, however, never been approached, but the preferential use of plants with fibrous root systems for rhizodegradation indicates that it was rather narrow. The production of protons, exudates and metabolites is released by plant roots in rhizosphere soil which led to significant
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differences between rhizosphere and non-rhizosphere in soil properties have been reviewed in previous reports [12,13]. The rhizosphere, a layer of soil surrounding plant roots, was difficult to physically sample and manipulate with precision. Rhizosphere soil was commonly separated from plant root by gentle shaking. In an attempt to overcome some of these problems, a rhizobox was designed where soil in close proximity to roots which allowed for the harvesting of thinner consecutive sections (1–5 mm, and >5 mm) of rhizosphere soil in the lab [14]. Recently, phytodegradation of PAH contaminated sediments using mangrove plants has been the subject of several studies [2,8]. However, limited by sampling techniques of rhizosphere soil in the proximity to roots, the distance-dependent microscale depletion of PAHs in root–soil interface along the rhizosphere gradient has thus far been seldom studied. Therefore, we hypothesized that the differences with distance in rhizosphere effects would coincide with the degradation gradients of PAHs. The objective of this study was therefore to investigate the rhizosphere effect on the removal process of PAHs in a specially designed rhizobox which permitted the separation of rhizosphere soil (root compartment) and soil affected by root exudation (root-free compartment). The study also attempts to reveal the effect of the increasing distance on PAH removal. 3-Ring PAHs Ph, and 4-ring PAHs Py were used as target PAHs. Kandelia candel (L.) Druce (K. candel), a common red mangrove species in China, was chosen as the model plant. 2. Materials and methods 2.1. Chemicals Ph and Py with a purity of 99.9% were obtained from Sigma–Aldrich Co. Ltd., UK. All the other chemicals used in the study were of analytical purity. 2.2. Preparation of PAH-spiked sediment Bulk samples of surface sediments were collected from Jiulong Estuary mangrove wetland, PR China, and sieved through a 0.5 cm sieve to remove coarse debris, homogenized, and then stored at 4 ◦ C until use. The physical and chemical properties of the sediments were measured in the laboratory as follows: pH 6.63, moisture content 49.5%, total organic content 2.1% and total nitrogen amount 0.90 g kg−1 dry sediment, total phosphate amount 0.62 g kg−1 dry sediment and cation exchange capacity 15.8 cmol kg−1 . PAHs were detected in the sediment samples with concentration of 19.3 g kg−1 Ph and 24.56 g kg−1 Py, respectively. Sediment was taken and spiked with PAHs as follows: a portion of the sediment was accurately weighed in the vessel. Then, a volume of the PAHs dissolved in acetone is added and allowed to equilibrate with the matrix, stored in the dark and allowed to dry. Mass balance is used to determine the evaporation of acetone. The acetone was evaporated 12 h and the portion of spiked sediment was first mixed with near 25% of total sediment, and then to mix with the remaining 75% of wet sediments followed by mechanical mixing. After aging for 7 days, the sediment was used for rhizobox experiment. The detected concentration of Ph and Py was 10 ± 0.5 and 10 ± 0.4 mg kg−1 , respectively in 7 days aged sediment. No nutrient amendments were added to the soil during the experiment. 2.3. Experimental design A laboratory rhizobox modified from our previous study [15] (Fig. 1) was used to plant K. candel. The dimension of the rhizobox (Fig. 1) was 150 mm × 300 mm × 200 mm
Fig. 1. Sketch diagram of rhizobox (modified from Lu et al. [15]). S0: sediment for seedling growth; S1: rhizosphere; S2: near rhizosphere; S3: near bulk soil; and S4: bulk soil.
(length × width × height). The rhizobox was divided into five sections from central to left or right boundary of rhizobox which were surrounded by nylon cloth (400 mesh), viz. a central zone for plant growth (20 mm in width), rhizosphere zones (1 mm in width), near rhizosphere zones (2 mm in width), near bulk soil zones (10 mm in width) and bulk soil zones (52 mm in width). In the rhizobox soil for seedlings growth, rhizosphere, near rhizosphere, near bulk soil and bulk soil zones were designated as S0, S1, S2, S3 and S4, respectively. The design successfully prevents root hairs from entering the adjacent soil zones as well as keeping the soil zones separated, while permitting the transfer of soil microfauna and root exudates between the compartments. About 12 kg of the treated sediment was added to each rhizobox, each treatment had three replicates. Five K. candel seedlings were planted in the central zone of the box. The plants were grown under greenhouse conditions with natural illumination and the relative humidity of 85%, the temperature ranging from 26 to 32 ◦ C for 60 days. Sediment moisture content was adjusted to 100% of the water holding capacity by watering with fresh water to minimize drainage and simulates anoxic, waterlogged conditions. Rhizobox without plant was used as control. Rhizoboxes were arranged in a randomized design within the glasshouse and their position was rotated regularly to ensure uniform conditions. Before harvesting, rhizoboxes were withheld from watering for 2 days. Harvesting involved the sequential dismantling of each rhizobox, separating the layers of each soil zone of the rhizobox and removing the plants from the root compartment. Roots and shoots were manually separated from soils washed with deionized water, and then blotted dry with filter paper. The soil samples from different soil zones of each rhizobox were homogenized separately before analysis. 2.4. Analyses 2.4.1. Measurements of enzymatic activities Soil microbial biomass carbon (Cmic ) was determined by the chloroform-fumigation–extraction method [16,17]. Sediment dehydrogenase activity was measured by the reduction of 2,3,5triphenyl tetrazolium chloride (TTC) to 1,3,5-triphenyl formazan (TPF). Briefly, 5.0 g of freeze-dried sediment sample was incubated for 24 h at 37 ◦ C in 5.0 mL of TTC solution (5.0 g L−1 in 0.2 mol/L Tris–HCl buffer, pH 7.4). Two drops of concentrated H2 SO4 were immediately added after incubation to stop the reaction. The sample was then blended with 5.0 mL of toluene to extract TPF and shaken for 30 min at 250 rpm (25 ◦ C), followed by centrifugation
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at 5000 rpm for 5 min, and absorbance in the extract was measured at 492 nm. Finally, soil dehydrogenase activity was calculated as 1.00 g TPF g−1 dry sediment [18]. Colorimetric method was utilized to determine the urease and phosphatase activities [19]. Enzyme activities expressed as mg NH4 –N released kg−1 dry sediment at 37 ◦ C and mg P released kg−1 dry sediment at 37 ◦ C each for urease and phosphatase, respectively. 2.4.2. PAH analysis Sediment samples were freeze-dried, meshed, and extracted with an accelerated microwave extraction system modified from Zhang et al. [20]. Briefly, 10.00 g of freeze-dried sediment was extracted with 50 mL mix solvent (n-hexane/acetone 1:1, v/v) using microwave extraction system (CEM Co., Matthews, NC, USA). The surrogates Ph-d10 and chrysene-d12 (Chy-d12 ) (Sigma–Aldrich, UK) were added to the samples prior to extraction. Activated copper (stirring copper with 5% of iodide/acetone solution for about 10 min) was added into the extract for desulphurization, and then pre-concentrated to 2 mL by a rotary evaporator (Buchi Vac V-800, Switzerland). Concentrated extracts were fractionated with alumina/silica gel (100–200 mesh) column chromatography (40 cm × 1.5 cm i.d.) packed from the bottom with glass wool, 10.00 g neutral aluminum oxide (100–200 mesh, dried at 440 ◦ C for 4 h), 18.00 g silica gel (100–200 mesh, dried at 170 ◦ C for 4 h) and 2.00 g anhydrous sodium sulphate. Target analytes were eluted from the column with 150 mL of mix solvent n-hexane/methylene chloride (1:1, v/v). This fraction was then concentrated to 2 mL by rotary vacuum evaporation in a water bath at 60 ◦ C and solventexchanged to n-hexane. The PAH fraction was finally concentrated to 1 mL under a gentle stream of nitrogen before GC/MS analysis. Plant samples were ground and homogenized, and extracted using the same method as to sediment. The concentrations of the PAHs in the extracts were determined by a Hewlett-Packard 6890 gas chromatography equipped with a mass spectroscopy detector (HP5975B). The HP-5MS column (Agilent Co., USA) was 30 m in length, with an internal diameter of 0.25 mm and a film thickness of 0.25 m. The temperature was raised from 60 ◦ C to 150 ◦ C at a rate of 15 ◦ C min−1 , increased to 220 ◦ C at 5 ◦ C min−1 , and increased to 300 ◦ C at 10 ◦ C min−1 , then held at 300 ◦ C for 5 min. Helium was used as the carrier gas. The injector and detector temperatures were 280 ◦ C and 300 ◦ C, respectively. The electron-impact energy was 70 eV and the mass to charge ratio scan (m/z) was from 50 to 400 amu. The selected ion mode (SIM) was chosen. Detection limits derived from replicate and procedural blanks were 2.2 and 1.6 g kg−1 dry weight for Ph and Py, respectively. All data were subject to strict quality control procedures. Matrix spikes, laboratory sample duplicates, and laboratory blanks were processed with each batch of samples (10 samples per batch) as part of the laboratory internal quality control. The mean recoveries of deuterated surrogate were 87.2 ± 2.1% for Ph-d10 and 90.3 ± 1.8% for Chy-d12 , respectively (n = 3). Spiked samples in each batch were analyzed with mean recoveries of 86.7 ± 2.6% for Ph and 89.6 ± 1.4% for Py, respectively (n = 3). Each extract was analyzed in duplicate form and relative standard deviations were less than 20%. Any analyses not meeting quality assurance requirements were re-analyzed. 2.5. Statistical analyses All of these experiments were performed in triplicates and the results presented were average values of the three replicates. Data were analyzed statistically using analysis of variance (ANOVA) and the Duncan’s multiple range tests was employed to determine the significance of the differences between the parameters. The
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Table 1 Removal percentages of phenanthrene and pyrene in various sampling zones in planted and unplanted treatments after 60 days of K. candel growth. Zones
Treatment Phenanthrene (%) Unplanted
S0 S1 S2 S3 S4
27.1 26.5 25.6 26.1 25.5
± ± ± ± ±
3.5Cb 3.4Cb 3.1Cb 2.9Cb 3.2Cb
Pyrene (%) Planted 47.5 53.6 56.8 43.2 37.3
± ± ± ± ±
Unplanted 4.1Aa 6.4Aa 6.2Aa 3.4Aa 3.2Ba
23.5 24.4 21.9 23.2 20.9
± ± ± ± ±
2.9Cb 2.6Cb 2.3Cb 2.8Cb 2.6Ca
Planted 32.4 46.2 47.7 38.1 23.5
± ± ± ± ±
3.7Ba 5.1Aa 4.9Aa 4.2Ba 2.6BCa
Note: Values in each column followed with different capital letters (A, B, C and D) indicated significant (p < 0.05) differences among different distances (0, 1, 2, 4 and 6 mm) from roots, and in each row followed with different lowercase letters (a and b) indicated significant difference between planted and unplanted soils by statistically using Duncan’s multiple range tests. Values represent means ± standard deviation. S0–S4 represented the distance of 0, 1, 2, 4 and 6 mm far from the root surface.
statistical package used was SPSS statistical software package (Version 11.0) and the confidence limit was 95%. 3. Results and discussion 3.1. Dissipation gradients of Ph and Py in sediment At the beginning of the experiment, 10 ± 0.5 and 10 ± 0.4 mg kg−1 of the added Ph and Py in the sediment slurry were adsorbed onto the sediments, respectively. This indicates that evaporation of acetone did not cause any significant loss of the spiked PAHs in sediment slurries. At the end of 60 days experiment, the results showed initial Ph (10.0 mg kg−1 ) and Py (10.0 mg kg−1 ) concentrations significantly decreased in the planted sediment as well as in unplanted control, but a more marked rate of disappearance was evident when plants were presented. The removal percentages for Ph and Py were 37.3–56.8% and 23.5–47.7%, respectively, in different gradient zones of planted sediment, which were significantly higher compared to unplanted treatments (25.5–27.1% for Ph and 20.9–24.4% for Py) (Table 1). The PAH concentration in the sediment after growing mangrove was affected by proximity to the roots. At both spiked Ph and Py sediment with planted treatments, the general trend in the degradation of the PAHs was typically rhizosphere > compartment > far-rhizosphere apart from a slight difference in root zones (Table 1). The mass balance results suggest that although the spiked PAHs were adsorbed tightly onto the sediments at the beginning of the experiment, PAH-degraders had the ability to utilize and degrade the sorbed PAHs efficiently (Table 2). The loss of the PAHs from mangrove sediment could be due to biotransformation, Table 2 Mass balance of phenanthrene and pyrene in rhizobox sediment after 60 days of K. candel growth. PAH (mg)
Non-vegetation
Plant-treatment
Ph
Input Leachate Plant uptake Remain in sediment Losses
100.0 ± 5.0 ND NA 74.2 ± 4.5 25.8 ± 1.7
100.0 ± 5.0 ND NA 59.8 ± 3.7 40.2 ± 2.3
Py
Input Leachate Plant uptake Remain in sediment Losses
100.0 ± 4.0 ND NA 78.4 ± 5.2 21.6 ± 1.2
100.0 ± 4.0 ND NA 72.3 ± 4.4 27.7 ± 1.8
Note: ND, not detected; NA, not applicable because no plants were grown in the rhizobox or PAHs uptake was negligible. Values represent means ± standard deviation.
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Phenanthrene Planted
16
Cmic(mg kg-1)
biodegradation, plant uptake, or abiotic dissipation, including leaching and volatilization [21,22]. In this study, abiotic losses by leaching were insignificant because non leachate was produced during the experiment. The losses of Ph and Py via volatilization from sediment are also unlikely to occur due to whole water covered condition and low vapor pressure of the PAHs (10−1.00 and 10−2.05 L atm mmol−1 for Ph and Py, respectively). Our data showed that mangrove plant only accumulate a little PAHs (detailed in Section 3.4), thus, the loss of the PAHs from soil by plant uptake/accumulation can be assumed to be negligible. In rhizosphere, Reilley’s results suggested that abiotic dissipation (chemical degradation and irreversible sorption) was not a possible pathway of loss of anthracene and Py [23]. Therefore, our results indicated that the enhanced dissipation of the PAHs might be caused by increased rhizosphere microbial density and activity compared to unplanted soil, since the root exudates and plant litter could enhance the bioavailability of the contaminant, provide more substrate for co-metabolic degradation, and modify the soil environment to be more suitable for microbial transformation. The roots are known to release some organic compounds, such as amino acids, organic acids, sugars, enzyme and complex carbohydrates, providing carbon source and energy for the growth of rhizosphere microorganisms [13,24]. The increased dissipation of PAHs in the rhizosphere may also be due to the decreased extractability of the PAHs with the formation of bound residues. The rhizosphere could stabilize pollutants by polymerization reactions such as humification [12,25]. The PAH degradation gradients observed in the rhizobox showed that the dissipation of Ph and Py was higher in the sediment only receiving root exudates than the soil with root exudates and plant roots. This is not consistent with the gradients in root exudates and plant enzymes or the depletion zones of the most diffusion limited mineral nutrients in lots of reports [26,27]. This is an important but interesting conclusion. It might be the result of the competition between plant roots and soil microbes for soil nutrients influencing the activities of soil microbes, especially in low organic matter soil such as sediment used in this study. In addition, the Ph of suberization induced by root senescence might render the PAHs more hydrophobic and potentially interfere with their availability through adsorption [28]. While PAHs accumulated in plants only accounted for a small amount of removed PAHs, whether this plant K. candel itself was able to produce enzymes for PAHs degradation is unknown. However, the synergism mechanism still need to be confirmed by further studies, in which more plants should be involved and different plant species should be studied.
Unplanted
12 8 4 0 S0
S1
S2
S3
S4
Pyrene
16
Cmic(mg kg-1)
266
Planted
Unplanted
12 8 4 0 S0
S1
S2
S3
S4
Fig. 2. The amount of microbial biomass carbon (Cmic ) in various distances proximity to K. candel roots grown in the sediment treated with phenanthrene and pyrene. Bars are the standard error of means of three replicates.
sediment with plants compared to unplanted treatments. In the PAHs-treated sediment, enzyme activities were largest in the rhizosphere or root compartment, and then decreased with increased distance from the root surface (Figs. 2–5). However, enzyme activities did not decrease with distance in unplanted soils. This matches well the PAH degradation data (Table 1). The most probable number of PAH degraders was influenced by planting regime. Our microbial biomass data support the hypothesis that micro-organisms were responsible for the observed PAHs degradation. Planted treat-
3.2. Microbial biomass and enzyme activities of the sediment The content of soil microbial biomass carbon and activities of soil dehydrogenase, urease and phosphatase were measured to evaluate the gradient effect of the rhizosphere on the PAH degradation. The different gradient sediments from roots displayed different responses to the presence of the PAHs in the rhizobox. Overall, in the unplanted sediment, microbial biomass measured as total Cmic was the same in various compartments but was lower than the planted sediment (Fig. 2). Likewise, Cmic was 16–234% greater with, than without, plants. The largest Cmic concentration (12.65 mg kg−1 for Ph and 10.68 mg kg−1 for Py, respectively) at either gradient zone was found in the rhizosphere (S1). The activities of soil dehydrogenase, urease and phosphatase in the planted soils were higher than those of unplanted treatments over the process of 60 days biodegradation (Figs. 3–5). In our study, it was shown that the relative lower concentrations of Ph and Py (10.0 mg kg−1 ) had a stimulatory effect on enzyme activity in sediment. However, other researchers found that higher concentration of PAHs could inhibit the enzyme in the soil [10]. Our data indicated that rhizosphere effects caused the increased response characteristics in the
Fig. 3. Urease activities in various distances proximity to K. candel roots grown in the sediment treated with phenanthrene and pyrene. Bars are the standard error of means of three replicates.
H. Lu et al. / Journal of Hazardous Materials 196 (2011) 263–269
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Table 3 The linear regression between residual PAH concentrations (Y) in sediment and different biological parameters in rhizosphere (S1) (Y = ax + b). Spiked-PAHs
Indexes (x)
a
b
r2
Phenanthrene
Cmic Urease Dehydrogenase Phosphatase
−0.36 −0.15 −5.76 −0.03
8.79 13.43 10.01 8.50
0.792* 0.693* 0.749* 0.508*
Pyrene
Cmic Urease Dehydrogenase Phosphatase
−0.41 −0.17 −7.22 −0.04
9.37 12.69 10.05 9.77
0.645* 0.745* 0.761* 0.459*
* Significant (p < 0.05) difference between residual PAHs concentrations and biological parameters.
of the soil biological status [29]. In our experiment, although the responses of activities of urease and phosphatase were different and inconsistent to some extent with the degradation of PAHs, it could still be concluded that increased urease and phosphatase activity occurred in the planted soil, especially in the rhizosphere, compared to the unplanted soil. The reason for different enzyme activities in different soil zones might relate to the gradient impact of root exudates. Fig. 4. Dehydrogenase activities in various distances proximity to K. candel roots grown in the sediment treated with phenanthrene and pyrene. Bars are the standard error of means of three replicates.
3.3. Correlation between microbial activities and PAH dissipation in rhizosphere
ments, especially rhizosphere, contained a significantly increased and large microbial biomass that could mediate the enhanced degradation of PAHs. The differences observed between soil with and without plants, as well as among various sampling zones in proximity to roots of the planted soils, were expected on the basis of microbial growth and community structure modified by both PAHs and root exudates. Overall microbial activity, as determined by dehydrogenase, urease and phosphatase activities is indicator
The successful application of rhizoremediation is largely dependent on the capacity of contaminant degraders or plant growth promoting microbes to efficiently colonize growing roots. Table 3 lists the relationships between microbial activities and the PAH dissipation in rhizosphere after 60 days of cultivation. A significant negative correlation was found between residual contaminant concentrations and soil enzymes in the rhizosphere. Statistical correlations (r2 ) of both spiked PAHs, especially for Cmic and dehydrogenase two indexes, had better values (rC2 = 0.792 and mic
2 0.645, p < 0.001; rdehydrogenase = 0.749 and 0.761, p < 0.001). For phosphatase, it showed a relatively poorer correlation. Some plant species appear to increase the numbers of degradative microbes in a large volume of soil that extends beyond the rhizosphere. The release of compounds or enzymes from roots is presumed to be associated with rhizosphere biodegradation and plant types vary in the nature and quantity of compounds released, it follows that the plant species used could be a significant factor influencing the efficacy of phytoremediation. Parrish et al. [30] reported that after 12 months of plant growth, the PAH degrading microbial populations in vegetated treatments were more than 100 times greater than those in unvegetated controls. This microbial consortia can provide various benefits to plants, including the synthesis of compounds that protect the plants by decreasing plant stress hormone levels; chelators for delivering key plant nutrients; protection against plant pathogens; and degradation of contaminants before they can negatively impact the plants [31]. Therefore, differences between rhizosphere soils and nonrhizosphere soils could be explained by the rhizosphere effect.
3.4. Accumulative potential of Ph and Py in plant tissues
Fig. 5. Phosphatase activities in various distances proximity to K. candel roots grown in the sediment treated with phenanthrene and pyrene. Bars are the standard error of means of three replicates.
When using spiked sediments for remediation experiments, the focus has often been on the ability of a given plant to accumulate a specific compound and can be removed along with the biomass for sequestration or incineration. In order to acquire a comprehensive understanding about the mechanisms of the PAH degradation, the
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Table 4 Phenanthrene and pyrene concentrations (mg kg−1 ) and concentration factors (CFs) in plant after 60 days of plant growth. Treatments
Root
RCFs
Stem
SCFs
Leaf
LCFs
Phenanthrene Pyrene
0.83 ± 0.11 1.56 ± 0.25
0.18 0.29
0.55 ± 0.07 1.83 ± 0.24
0.12 0.34
0.32 ± 0.04 0.59 ± 0.07
0.07 0.11
uptake of Ph and Py by K. candel was measured (Table 4). The concentrations of Ph and Py in root were higher than those in the stem and leaf, and the concentrations of Ph in leaf were lowest among plant tissues (Table 4). Plant concentration factors (CFs) were calculated as the ratio of the PAH concentrations in plant tissues (root, shoot and leaf) and in sediments on a dry weight basis. The results also indicated that root concentration factors (RCFs) of Ph (0.18) were much lower than those of Py (0.29) treatment. It might be explained by the higher Kow (octanol–water partition coefficient) value of Py than Ph [32]. It was demonstrated that hydrophobic compounds with log Kow > 4 are not readily taken up by plants through transpiration due to their hydrophobicity; log Kow for Ph and Py was 4.17 and 5.13, respectively [33]. Our data indicated that K. candel was not a hyper accumulation plant for PAHs (CFs from 0.07 to 0.34). There were no significant correlations between concentration of Ph and Py in roots and the dissipation of Ph and Py from rhizosphere sediments (S1) as well as other gradient sediments (S2–S4) were found. This indicated that accumulation of Ph and Py by roots was not the major factor contributing to the removal of PAHs from soil. This result is similar to previous report [34] which indicated contribution of K. candel accumulation and plant uptake to the removal of Py from contaminated sediments was insignificant. 4. Conclusion We investigated the rhizodegradation gradient of mangrove plant K. candel for the PAH contaminated sediment. The presence of mangrove plant significantly increased the dissipation of Ph and Py in contaminated sediment. Effect of root proximity was important in the removal process of Ph and Py, which was depended on the distance from the root surface. Enhanced dissipation rate in different gradients of planted versus unplanted sediment was 11.8–29.9% for Ph and 2.9–25.8% for Py. Accumulation of the PAHs in plant parts showed negligible contributions to the total remediation. Plant root-promoted dissipation was the predominant contribution to the remediation enhancement for sediment Ph and Py in the presence of K. candel. Our results suggested that the enhancement of Ph and Py disappearance is caused by an increase in the rhizosphere biological activity compared to root free sediment. Moreover, there is a scope for future work particularly regarding underlying mechanisms responsible for observed rhizoremediation outcomes. These future directions include elucidation of the complex processes at the interface of soil, microorganisms and roots. More effort also should be made to investigate of root exudates being deposited into the rhizosphere and involved microbe activities during the remediation process, and the achievable outcome using the mangrove under field trials. Acknowledgements The present study was supported by the China Postdoctoral Science Foundation Found Project and National Natural Science Foundation of China (10805036, 20777062, and 30710103908). The authors wish to thank Bosen Weng and Yong Huang for assistance in the sampling work, and Dr. Youwei Hong for the PAH analysis.
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Journal of Hazardous Materials 196 (2011) 270–277
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Amphiphilic hollow carbonaceous microspheres for the sorption of phenol from water Zhengrong Guan, Li Liu, Lilu He, Sen Yang ∗ College of Resources and Environmental Sciences, Biomass Engineering Center, China Agricultural University, Beijing 100193, People’s Republic of China
a r t i c l e
i n f o
Article history: Received 29 June 2011 Received in revised form 19 August 2011 Accepted 7 September 2011 Available online 14 September 2011 Keywords: Carbonaceous spheres Removal Partition Phenol
a b s t r a c t Amphiphilic porous hollow carbonaceous spheres (PHCSs) were synthesized via mild hydrothermal treatment of yeast cells and further pyrolyzing post treatment. The morphology, chemical composition, porosity, and structure of the carbonaceous materials were investigated. It is evident that the carbonaceous materials were composed of the carbonized organic matter (COM) and the noncarbonized organic matter (NOM), and the relative COM and NOM fractions could be adjusted through changing the temperature of hydrothermal and/or pyrolyzing treatment. The phenol sorption properties of the carbonaceous materials had been investigated and the sorption isotherms fit well to the modified Freundlich equation. It was found that the sorption isotherm of phenol onto PHCSs was practically linear even at extreme high concentrations, which was fewer reported for activated carbon or other inorganic materials. This type of sorption isothermals was assigned to a partition mechanism, and the largest value of the partition coefficient (Kf ) and carbon-normalized Kf (Koc ) is 56.7 and 91.5 mL g−1 , respectively. Moreover, PHCSs exhibit fast sorption kinetic and facile regeneration property. The results indicate PHCSs are potential effective sorbents for removal of undesirable organic chemicals in wastewater, especially at high concentrations. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Phenol-like compounds have attracted global concerns because of their toxicity and the frequency and quantity of their presence in wastewaters of various industries, such as refineries (6–500 mg L−1 ), coking operations (28–3900 mg L−1 ), coal processing (9–6800 mg L−1 ), and manufacture of petrochemicals (2.8–1220 mg L−1 ) [1–4]. Phenol-containing wastewater usually involves multiple, different contaminants and the concentration is different from low-concentration of several mg L−1 to highconcentration of several thousands mg L−1 . Those meanings the best abatement technologies for phenol from wastewaters to be applied strongly depends on single cases, in particular on the concentration of phenol in the stream, the co-presence of other contaminants, the nature of the plant where this problem is found [2]. Now, a number of strategies such as oxidation with ozone/hydrogen peroxide [5], biological methods [6,7], membrane filtration [8], ion exchange [9], electrochemical oxidation [10], photocatalytic degradation [11], and adsorption [12] have been used for the removal of phenol. Review on available technologies for phenol removal from fluid streams has been recently published providing comparison of the experimental conditions and the performances of different techniques [2]. Adsorption is generally considered an
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operationally simple, effective and widely used process for the removal and recovery of the phenol. Among vast number of different adsorbents, activated carbon (AC) is the most commonly used adsorbent in industrial scale and experimental research [1]. However, AC is not an ideal adsorbent for practical applications at high phenol concentrations, since the adsorption amount of AC will soon saturate. The demanding regeneration and poor mechanical rigidity of AC are also problems for its wider application [13]. Consequently, a large variety of non-conventional adsorbents have been studied for removal of phenol and phenolic pollutants including organoclay [14], polymer [13], carbon nanotube [15], sewage sludge-based adsorbents [16] and activated carbon cloth [17]. However, development of novel adsorbent materials for removal and recovery of the phenol and phenolic pollutants with high concentrations still remains an important challenge. Yeast, a by-product of the brewing industry, is considered as an industrial organic waste that causes a great deal of concern [18]. In a previous study, we reported a facile method for fabricating porous hollow carbonaceous spheres (PHCSs) with controlled shell porosity from Saccharomyces cerevisiae (S. cerevisiae) cells [19,20]. Through mild hydrothermal treatment of these tiny unicellular organisms, hollow microspheres with controllable mesoand macroporous shells were synthesized. Most interestingly, the surfaces of these hollow spheres were found to be covered with both hydrophobic and hydrophilic functional groups, endowing the as-obtained microspheres with amphiphilic property. Actually, we found that PHCSs could be well dispersed not only in water but also
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in nonpolar solvents such as toluene and chloroform. This inspired us to speculate that PHCSs may be promising sorbent for organic pollutant from aqueous solution. Herein, four kinds of PHCSs were synthesized via mild hydrothermal treatment of yeast cells and further pyrolyzing post treatment. And the morphology, chemical composition, porosity, and structure of the carbonaceous materials were investigated. The sorption behavior (i.e., sorption isotherms, kinetics models, and the influence of pH and temperature) of PHCSs toward phenol was examined. The relationship existing between the specific surface area and the chemical composition of PHCSs and their phenol sorption capability was elucidated. Lastly, principal sorption mechanisms were clarified. 2. Experimental 2.1. Synthesis of PHCS PHCSs were synthesized via mild hydrothermal treatment of yeast cells using modified methods described in our previous studies [19]. Typically, S. cerevisiae cells (3–4 g, purchased from Angel Yeast Co., Ltd., China) pre-washed with acetone were dispersed in 2–3% (v/v) glutaraldehyde and diluted (less than 0.01 mol L−1 ) nitric acid aqueous solution (40 mL), which was then placed in a 50 mL Teflon-sealed autoclave and maintained at 180 or 230 ◦ C for 8 h. The puce solid products were centrifugal separated, then washed by three cycles of centrifugation/washing/redispersion in deionized water and alcohol, and oven-dried at 80 ◦ C for 4 h. The PHCSs samples hydrothermally treated at 180 and 230 ◦ C were denoted as P180 and P230, respectively. The samples denoted as P350 and P700 were prepared via further pyrolyzing P180 at temperature of 350 ◦ C and 700 ◦ C for 1 h, respectively, by a tubular reactor in a flow of nitrogen (15 mL min−1 ). 2.2. Characterization of PHCS The morphology of the materials was inspected with a field emission scanning electron microscope (SEM, JEOL, JSM-6700F, Japan). Surface area and pore volume of the materials were measured by N2 adsorption/desorption isotherms at 77 K with a Physisorption Analyzer (Micromeritics, ASAP 2020, U.S.A.). Fourier-transform infrared (FTIR) analysis was performed by a Micro FTIR spectrometer (Nicolet, Magna 750 Nic-Plan FTIR Microscope, U.S.A.) in the spectral region of 4000–650 cm−1 . Solid-state cross-polarization magic angle-spinning and totalsideband-suppression 13 Carbon nuclear magnetic resonance (13 C NMR) spectrum (CPMAS-TOSS) were obtained by a Bruker Avance 400 MHz spectrometer (Karlsruhe, Germany). The C, H, N, O contents of the samples were determined using elemental analyzer (Elementar, Vario EL, Germany). The atomic ratio of (O + N)/C, O/C, H/C were calculated with the element content. 2.3. Sorption experiments Batch experiments were carried out using a series of 15 mL screw cap centrifuge tube covered with Teflon sheets to prevent the introduction of any foreign particle contamination. In typical batch experiment, 20 mg of the sorbent was added to 10 mL phenol solution at various concentrations (0–10,000 mg L−1 ) taken in sealed tubes, which were placed in the thermostat shaking assembly. The solutions were shaken at 150 rpm and constant temperatures for 24 h to achieve equilibration. After equilibrium, the mixtures were filtrated through 0.45 m nitrocellulose syringe filters. The phenol concentrations in the filtrates were determined by a UV-visible spectrometer (Persee, TU-1810, China) at the maximum adsorption wavelength of phenol (270 nm) and pH 6 (pH adjusted with 0.5 M
271
HCl or NaOH). Isotherms were performed by taking different concentrations of phenol at designed temperatures and pH values. All the experiments were carried in triplicate and the averaged data were reported. Standard deviations were found to be within 2.0%. Furthermore, the error bars for the figures were smaller than the symbols used to plot the graphs and hence are not shown in the figures. 2.4. Sorption models and statistical analysis Freundlich model was applied for sorption data fitting in this work due mainly to its advantages in isotherm nonlinearity investigation. To facilitate direct comparisons of sorption affinities among the samples tested, and investigate the effect of temperature on sorption, the modified Freundlich equation was applied for sorption data fitting in this work: log qe = log KF + n log Cr Cr =
Ce Sw
(1) (2)
where qe is the solid-phase concentration (mg g−1 ), Ce is the liquidphase equilibrium concentration (mg L−1 ); Sw is the solubility of phenol for a given temperature (mg L−1 ) and Cr is dimensionless since the value of Sw is constant for a given temperature and is expressed in the same unit as Ce ; K F and n are the modified Freundlich adsorption parameters, K F (mg g−1 ) is the sorption capacity coefficient, which represents the mass of phenol sorbed per unit mass of sorbent when the Ce concentration approaches saturation, and n (dimensionless) is an indicator of isotherm nonlinearity related to the heterogeneity of sorption sites[21]. The partition–adsorption model for describing sorption from aqueous solutions on heterogeneous solids was also analyzed [14,22]: QT = QA + QP
(3)
where QT is the total amount of phenol sorbed onto the sorbent; QA and QP are the amounts contributed by adsorption and partition, respectively. According to the partition–adsorption model, the partition effect is favored progressively by increasing the solute concentration, whereas the adsorption contribution reaches saturation more rapidly with the solute concentration. The isotherm at high concentrations should approach linearity. Therefore, at the high solute concentration range the adsorption becomes saturation and the linear partition remains. Thus, Eq. (3) can be transformed to: QT = QAmax + QP = QAmax + Kf Ce Koc =
Kf foc
(4) (5)
where QAmax is the saturated adsorption capacity estimated from the high concentration data; Kf Ce is the partition contribution at high concentration with Kf being the partition coefficient; Ce is the solute equilibrium concentration (mg L−1 ); Koc is carbon-normalized Kf , and foc is the percentage of carbon contents of sorbent. Linear regression between QT and Ce were conducted at high solute concentration range, and the QAmax corresponds to the y-axis intercept of the line, and Kf to the slope. 2.5. Sorption kinetics For kinetic studies the batch technique was used. Typically, 20 mg of PHCSs were added 10 mL phenol solution (2000 mg L−1 ) into series of screw cap centrifuge tubes, and then shaken in constant temperature rotary shaker (150 rpm) at 25 ± 0.5 ◦ C. The
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Fig. 1. SEM images of (a) P180, (b) P230, (c) P350 and (d) P700.
concentrations of phenol in solution were sampled and analyzed with different time intervals, and the averaged data were reported. Linear pseudo-first and pseudo-second order models were given in the following equations [23,24]: ln(qe − qt ) = ln qe − k1 t
(6)
t 1 t = + qt qe k2 q2e
(7)
t1/2 =
1 k2 qe
(8)
h = k2 q2e
(9) mg g−1 )
where qe and qt (both in are the amount of phenol sorbed on per unit mass of sorbent at equilibrium and at time t (h), respectively; k1 (h−1 ) and k2 (g mg−1 h−1 ) are the pseudo-first-order and pseudo-second-order rate constants, respectively. Moreover, for the pseudo-second-order kinetic model, the half sorption time (t1/2 ) and the initial sorption rate (h) are given (Eqs. (8) and (9)). The value of t1/2 is the time required to uptake half of the maximal sorbed amount of sorbate at equilibrium. 3. Results and discussion 3.1. Characterization of PHCSs SEM images and optical microscopy photos of the carbonaceous products are shown in Figs. 1 and S1, respectively. It was found that the hydrothermal products (P180 and P230) were porous hollow microspheres in the size range 2.0–4.0 m, which was consistent
with our previous results [19]. The thermal stability of the microspheres was satisfactory. After post pyrolyzing treatment of P180 at 350 ◦ C, the microsphere structure was still well preserved (Fig. 1c for P350). Further increasing the pyrolyzing treatment temperature to 700 ◦ C, however, the morphology of materials changed from microspheres into larger carbon blocks (Fig. 1d for P700). Surface area and pore volume of the materials were measured by N2 adsorption/desorption isotherms at 77 K and all the BET surface areas of the materials are below 10 m2 g−1 (Table 1). N2 adsorption/desorption isotherm of the materials at 77 K with corresponding pore size distributions are shown in Fig. S2. Fig. 2 shows the solid-state 13 C NMR spectra of the carbonaceous products. Peaks at ı = 26 and 31 ppm can be attributed to methyl and methylene, respectively, and those in the ı = 120–150 ppm region can be assigned to long-range conjugated C C bonds and oxygen-substituted C C bonds, revealing the existence of aromatic furan ring compounds [25,26]. Moreover, oxygenated functional groups, including carbonyl, carboxy, hydroxy, ether, and ester groups were also detected. The content of aromatic species was enhanced remarkably by pyrolyzing treatment of the hydrothermal products, indicating high carbonization degree for P350 and especially P700. Moreover, the average chemical shift decreases with the sequence of the carbonization, resembling to some extent the resonance pattern of graphite. Corresponding Fourier transform infrared (FTIR) spectra of the carbonaceous materials are shown in Fig. 3. Both aliphatic and aromatic species were revealed for P180, P230 and P350. The bands at 2925, 2861, 1442, and 1372 cm−1 are assigned mainly to CH2 units [27]. Those at 1693 and 1160 cm−1 are assigned to C O and C–O stretching vibrations of ester bonds, respectively. And the band at 1602 cm−1 is assigned to C C and C O
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Table 1 Elemental compositions, atomic ratios, BET surface area (SA), and total pore volume (TPV) of the carbonaceous materials. Sample
Treatment temperature (◦ C)
C (wt.%)
H (wt.%)
O (wt.%)
N (wt.%)
(O + N)/C ratio
O/C ratio
H/C ratio
SA (m2 g−1 )
TPV (mL g−1 )
P180 P230 P350 P700
180 230 350 700
62.00 72.96 73.22 77.17
6.25 6.99 4.44 1.94
21.16 12.15 10.61 5.71
6.08 4.88 6.27 5.06
0.34 0.18 0.18 0.11
0.26 0.13 0.11 0.06
1.21 1.14 0.72 0.30
9.504 2.635 1.960 2.830
0.0498 0.0095 0.0052 0.0080
H/C: atomic ratio of hydrogen to carbon. O/C: atomic ratio of oxygen to carbon. (O + N)/C: atomic ratio of sum of nitrogen and oxygen to carbon.
Fig. 2. Solid-state 13 C NMR spectrum of (a) P180, (b) P230, (c) P350 and (d) P700.
stretching in the aromatic ring [22]. The peak of 786, 706 cm−1 can be assigned to the aromatic CH out-of-plane deformation. In the case of P700, almost all of the band intensities mentioned above are dramatically decreased or disappeared, indicating destruction of
the surface function groups and chemical structure after pyrolyzing treatment at high temperatures. The elemental compositions shown in Table 1 agree well with the observations of FTIR. O content decreased from 21.16% for
Fig. 3. FTIR spectrum of (a) P180, (b) P230, (c) P350 and (d) P700.
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Fig. 4. Sorption isotherms of phenol on carbonaceous products in aqueous solution at 25 ◦ C and pH 6, inset: sorption isotherms on P230 at high phenol concentrations.
P180 to 12.15% for P230 and further declined slightly to 10.61% for P350. When the pyrolyzing treatment temperature was increased to 700 ◦ C, the O content declined drastically to 5.71% (P700). The H content of P350 and P700 declined dramatically from about 6.25% (P180) to 4.44% and 1.94%, respectively. These results mean that the degree of carbonation of samples increased with higher hydrothermal and pyrolyzing temperature. The H/C ratio and O/C ratio decreases with deep carbonization, indicating the surfaces of the products become less hydrophilic [28]. The decrease of the polarity index [(O + N)/C] with the degree of carbonization reveals a reduction of the surface polar functional groups [22]. All the above results reveal that both hydrothermal and pyrolyzing treatment changed the chemical composition of PHCSs to a large degree. Due to the partial loss of the hydrophilic groups and aromatization of the molecule networks, wettability characteristics of PHCSs were supposed to be very different from that of the original hydrophilic yeast cells. We found all samples except P700 could be well dispersed not only in water but also in nonpolar solvents (Fig. S3), suggesting that PHCSs possess amphiphilic surfaces, as we reported previously [19,20]. Moreover, there are two types carbon species in the PHCSs, namely alkyl carbon and aromatic carbon. The alkyl carbon is mainly derived form the noncarbonized organic matter (NOM) of the yeast cells, and the aromatic carbon is the products of carbonization, i.e., the carbonized organic matter (COM). It is evident that the relative COM and NOM fractions could be adjusted through changing the temperature of hydrothermal and/or pyrolyzing treatment. While there are large amounts of NOM in the samples of P180, P230 and P350, COM is dominant in P700. The amphiphilic property and the specific chemical composition of the PHCSs make them may be excellent potential sorbents for organic compounds removal from wastewater. Subsequently, the sorption isotherms of phenol were carried out. 3.2. Sorption isotherms Sorption of phenol onto the carbonaceous materials at 25 ◦ C was investigated and the results are presented in Fig. 4. The sorption of P700 was saturated at very low phenol concentration (86 mg L−1 ) and the sorption amount was only 5.5 mg g−1 , which may be due to the very low surface area (2.8 m2 g−1 ). Because P700 was produced by pyrolyzing P180 at 700 ◦ C for 1 h, the high carbonization degree and few surface function groups (Figs. 2 and 3) suggested that the sorption mechanism of P700 is physical adsorption.
The result means P700 cannot be used as sorbents for organic compounds, and so we did not discuss about P700 any more. The surface area of P180, P230 and P350 was all less than 10 m2 g−1 , however, they exhibited quite different sorption behaviors: surface area-independent sorption amount and the almost linear isotherm without a saturated sorption. This sorption characteristic was fewer reported for AC [29,30] or other inorganic adsorbent materials [31], which suggests that the sorption mechanism of PHCSs is not simple physical or chemical adsorption. Modified Freundlich model was applied to investigate the isotherm nonlinearity of PHCSs. The sorption isotherms fit well to the modified Freundlich equation and the calculated parameters are listed in Table 2. The isotherm of phenol onto P180 is practically linear, with Freundlich n = 1.016 ± 0.025, and the isotherms for P230 and P350 display different nonlinearity of a concavedownward curvature at low solute concentrations but exhibit a practically linear shape at moderate to high concentrations. The nonlinear effects are relatively more visible for the P350 than for the P230. Similar results were observed for sorption of organic solutes from water over a wide range of Ce /Sw by black carbons (BCs) or biochars, derived mainly from the incomplete combustion of biomass and fossil fuel [22,32,33]. The unique isotherm shape, i.e., nonlinear at low Ce /Sw but virtually linear at other Ce /Sw , suggests that more than a single mechanism be operative over the entire concentration range [34]. To explain the nonlinearity of sorption isotherms, dual-mode sorption models (DMSM) or dual-reactive domain models (DRDM) were suggested. According to these models [14,22,32,33,35], the sorbent was considered to be a heterogeneous substance, and a concept of NOM (or “soft carbon”, expanded, rubbery state) vs. COM (or “hard carbon”, condensed, glassy state) was been invoked to operationally delineate chemical heterogeneity of sorbent and to elucidate the mechanisms for sorption by soils, sediments, biochars and charcoal. The COM is expected to behave as an physical adsorbent, producing isotherm nonlinearity, and the NOM can uptake pollutants via a partition (sorption) mechanism [22,32,33]. One unique feature of the partition process is that the ratio of solid phase to aqueous-phase concentrations remains unchanged with the variation of the solute concentration. In this sense, the sorptive uptakes are determined by the relative carbonized and noncarbonized fractions and their surface and bulk properties. Thus, the linearity of P180 can be assigned to sorption of phenol to NOM via partition mechanism, since NOM is the mainly carbon species in this sample. The nonlinearity of P230 and P350 at low Ce /Sw is attributed to a combined physical adsorption on a small amount of COM and a partition effect of NOM. At moderate to high Ce /Sw , the physical adsorption becomes largely saturated and the partition in NOM predominates to give an essentially linear isotherm. The larger curvature at low Ce /Sw of P350 is presumably due to the higher content of aromatic moieties with increased carbonization degree. The adsorption behavior of P700 with highest carbonization degree, validate this speculation. High isotherm nonlinearity was also observed for high-temperature chars [33]. From the partition–adsorption model, QAmax , Kf and Koc were calculated (Table 2), The Kf and Koc increase in the order of P350 < P230 < P180, which is in inverse with their carbonization degree. The physicochemical nature of the organic carbon has been suggested as a major factor controlling sorption of organic compounds to natural or modified organic sorbent. According to FTIR and NMR data (Figs. 2 and 3), the major partition phase in P180, P230 and P350 is a polymeric aliphatic fraction preserved during the hydrothermal process, and the content of aliphatic fraction increased in the order of P350< P230 < P180, which is in the line with the increased partition effect. The similar conclusions were also presented by other researchers [22,36]. Chefetz et al. [36] tested the sorption of pyrene to a series of sorbents comprised of different levels of aromaticity and aliphaticity. In that study, a
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Table 2 Modified Freundlich model parameters, partition coefficients and saturated adsorption of phenol to PHCs. Sample
Modified Freundlich model parameters log KF a
P180 P230d P230 P230e P350
3.733 3.102 3.434 3.230 2.543
± ± ± ± ±
nb 0.049 0.024d 0.058 0.018e 0.041
1.016 0.661 0.822 0.700 0.378
± ± ± ± ±
0.025 0.013d 0.025 0.009e 0.019
Partition–adsorption model R2
Kf (mL g−1 )c
QAmax (mg g−1 )c
0.996 0.997d 0.987 0.991e 0.986
56.7 25.8d 38.2 39.4e 17.3
0 67.3d 45.8 34.3e 45.15
Koc (mL g−1 )
max QA,SA (mg g−1 m−2 )
91.5 35.4d 52.5 54.0e 23.6
0 25.5 17.4 13.0 23.0
max Koc is carbon-normalized Kf . QA,SA is the SA-normalized QAmax . The solubility of phenol in water at 15, 25 and 35 ◦ C is 8200 mg/100 mg, 8660 mg/100 mg and 9910 mg/100 mg, respectively. a 95% confidence interval of log KF . b 95% confidence interval on n. c The slope and y-axis intercept of the linear equation were used to calculate partition coefficient (Kf ) and the maximum adsorption capacity (QAmax , mg g−1 , Chen et al. [22]), respectively. d Tested at 15 ◦ C. e Tested at 35 ◦ C.
positive trend was observed between the Koc level and the aliphaticity of the set of samples. Chen et al. [22] performed sorption experiments with biochars, produced by pyrolysis of pine needles at different temperatures. This study clearly demonstrated that a greater sorption affinity of naphthalene, nitrobenzene, and m-dinitrobenzene with aliphatic-rich biochars than with aromaticrich biochars. max of P230 and P350 was listed in Table 2, which exceeds The QA,SA greatly the amount accountable by the small surface area of the sorbent. Some researchers also reported the higher adsorption of polar solutes compared with the little sorbent surface area. To explain the higher sorption of polar pesticides at low (relative) concentrations, Spurlock and Biggar [37] suggested the specific-interaction model. The model postulates that the specific interaction of polar solutes with highly active sites of organic carbon phase. That implies that more highly active site, more specific interaction, and max . P350 showed less oxygen content and highly active higher QA,SA max site because of pyrolyzed dehydration at 350 ◦ C, however, the QA,SA of P350 is 23.0 mg g−1 m−2 , higher than the 17.4 mg g−1 m−2 of P230. We speculated the possible phenol–sorbent interaction may be hydrophobic effect since P350 showed higher hydrophobicity compared with P230.
3.3. Sorption kinetics Fig. 5 shows the sorption kinetics of phenol onto P180 and P230. Apparently, the sorption rates were very fast. For example, given the tested conditions, approximately 86 and 85% of sorption was accomplished within 0.5 h for P180 and P230, respectively. For quantitative comparison of apparent sorption kinetics between P180 and P230, the data were fitted to the pseudo-first order and pseudo-second order models, respectively. The sorption kinetic constants were listed in Table 3. The regression coefficient (R2 ) for the pseudo-first order model varied from 0.9830 to 0.7022 and the Qe values calculated from the model deviated tremendously from the experimental values, together indicating the invalidity of the model. However, the pseudo-second-order model provided the best fitting for the all experimental data. The plots show regression coefficients higher than 0.9997 for P180 and P230. The value of the constant k2 of P180 is higher than that of P230, which inversely correlates with the Kf . Thus, it could be speculated that kinetics of phenol following the pseudo-second-order model are controlled by an adsorption process and adsorption was the main sorption rate-limiting step. 3.4. Effect of temperature on sorption To investigate the effect of temperature on the sorption of phenols, sorption experiments were conducted at 15, 25 and 35 ◦ C for P230. The results are shown in Fig. 6. The parameters of
Fig. 5. Sorption kinetics plotted as sorbed amount of phenol vs. time on P180 and P230 with a 2000 mg L−1 initial phenol concentration and 25 ◦ C.
Fig. 6. Sorption isotherms of phenol on P230 at 15, 25 and 35 ◦ C at pH 6.
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Table 3 Adsorption rate constants for two kinetic models at 25 ◦ C and pH of 6. Sample
P180 P230 a b
qe (exp)a (mg g−1 )
98.63 107.85
First-order model
Second-order model
K1 (h−1 )
qe (cal)b (mg g−1 )
R2
K2 (g mg−1 h−1 )
qe (cal)b (mg g−1 )
t1/2 (h)
h (mg g−1 h−1 )
R2
3.17 0.12
82.92 19.86
0.9830 0.7022
0.1028 0.0660
110.31 106.80
0.099 0.14
1000 769
0.9997 0.9998
Experimental data. Calculated data from models.
modified Freundlich adsorption model and partition–adsorption model are incorporated in Table 2. A comparison of modified Freundlich parameter n shows that the biggest value was got at 25 ◦ C and the smallest value at 15 ◦ C. This means that the effect of temperature on the adsorption and partition is different. Comparing with QAmax and Kf of P230 at 15, 25 and 35 ◦ C, we can find that partition increase with an increase in temperature, meaning that temperature may play an important role for the partition coefficient Kf . To date, fewer Kf values have been reported for hydrophobic organic contaminants (HOCs) at temperatures other than 25 ◦ C, and the reported conclusion is different [38]. Chen and Pawliszyn [39] evaluated the effect of temperature on Kf for BTEX compounds and found that Kf should not be greatly dependent on temperature. Muijs and Jonker [40] determined Kf values for several PAHs and found that temperature can play a significant role. We speculated that the increasing Kf values of P230 for phenol with increasing temperature maybe due to the polarity of phenol and the structure of P230. However, the adsorption decreases with an increase in temperature indicate that the process is apparently exothermic.
3.5. Effect of pH value on sorption The effect of pH on the sorption ability of P230 was investigated and the results are shown in Fig. 7. According to the model proposed by Deryło-Marczewska and Marczewski [41], the adsorption of non-ionized compound does not depend on the pH and the surface charge. The sorption isotherms at the pH of 3 and 6 coincided. This may be due to the dissociation degree of phenol is very low (pKa = 9.89) in acidic conditions. However, the sorption at pH of 11 obviously decreased since phenol is highly dissociated at pH 11 and dissolved into water.
3.6. Sorbent regeneration Important goals in the development of sorbent materials include simple regeneration and sorbate isolation [42]. Regeneration allows for the repeated use of the sorbent material and decreasing costs. It was found that desorption of phenol from the loaded P230 using 0.01 M NaOH solution was rapid. For example, after placing P230 that had sorbed 141 mg g−1 phenol in a 0.01 M NaOH solution for 10 min, the regenerated P230 showed greater than 98% of the phenol was removed from the sorbent. The facile regeneration is due to the high solubility of the sodium salt of phenol in water. 4. Conclusions Amphiphilic PHCSs were synthesized via mild hydrothermal treatment of yeast cells and further pyrolyzing post treatment. The sorption properties of PHCSs for phenol in aqueous solutions were investigated, and then reached the following conclusions: (1) PHCSs were composed of COM and NOM, and the relative COM and NOM fractions could be adjusted through changing the temperature of hydrothermal and/or pyrolyzing treatment. (2) The sorption isotherm of phenol onto PHCSs was practically linear even at extreme high concentrations. This type of sorption isothermals was assigned to a partition mechanism, and the largest value of the partition coefficient (Kf ) and carbonnormalized Kf (Koc ) is 56.7 and 91.5 mL g−1 , respectively. (3) PHCSs exhibit fast sorption kinetic and facile regeneration property. The results indicate they are potential effective sorbents for removal and recovery of undesirable organic chemicals in water treatment, especially at high concentrations. Acknowledgments This work was supported by the National Natural Science Foundation of China (20703065, 20877097, and 20806089), the Ministry of Science and Technology of China (2008AA06Z324) and Chinese Universities Scientific Fund (2011JS160). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.025. References
Fig. 7. Sorption isotherms of phenol on P230 at pH 3, 6 and 11 at 25 ◦ C.
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Journal of Hazardous Materials 196 (2011) 302–310
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Partitioning behavior and stabilization of hydrophobically coated HfO2 , ZrO2 and Hfx Zr1−x O2 nanoparticles with natural organic matter reveal differences dependent on crystal structure Divina A. Navarro 1 , Sean W. Depner, David F. Watson, Diana S. Aga ∗ , Sarbajit Banerjee ∗∗ Department of Chemistry, University at Buffalo, State University of New York, Buffalo, NY 14260-3000, USA
a r t i c l e
i n f o
Article history: Received 11 April 2011 Received in revised form 7 September 2011 Accepted 8 September 2011 Available online 14 September 2011 Keywords: Twin-metal oxides Natural organic matter Phase transfer Nanoparticle structure Environmental mobility Colloidal interactions
a b s t r a c t The interactions of engineered nanomaterials with natural organic matter (NOM) exert a profound influence on the mobilities of the former in the environment. However, the influence of specific nanomaterial structural characteristics on the partitioning and colloidal stabilization of engineered nanomaterials in various ecological compartments remains underexplored. Herein, we present a systematic study of the interactions of humic acid (HA, as a model for NOM) with monodisperse, well-characterized, ligandpassivated HfO2 , ZrO2 , and solid-solution Hfx Zr1−x O2 nanoparticles (NPs). We note that mixing with HA induces the almost complete phase transfer of hydrophobically coated monoclinic metal oxide (MO) NPs from hexane to water. Furthermore, HA is seen to impart appreciable colloidal stabilization to the NPs in the aqueous phase. In contrast, phase transfer and aqueous-phase colloidal stabilization has not been observed for tetragonal MO-NPs. A mechanistic model for the phase transfer and aqueous dispersal of MO-NPs is proposed on the basis of evidence from transmission electron microscopy, -potential measurements, dynamic light scattering, Raman and infrared spectroscopies, elemental analysis, and systematic experiments on a closely related set of MO-NPs with varying composition and crystal structure. The data indicate the synergistic role of over-coating (micellar), ligand substitution (coordinative), and electrostatic processes wherein HA acts both as an amphiphilic molecule and a charged chelating ligand. The strong observed preference for the phase transfer of monoclinic instead of tetragonal NPs indicates the importance of the preferential binding of HA to specific crystallographic facets and suggests the possibility of being able to design NPs to minimize their mobilities in the aquatic environment. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The imminent large-scale commercialization of engineered nanomaterials (ENMs) has raised concerns regarding their potential environmental impact [1–4]. Some preliminary data are starting to become available regarding the toxicity of ENMs at the sub-cellular, cellular, and organism levels [5–8]. Among these materials, transition metal oxide (MO)-based nanoparticles (NPs) are finding various applications as nanoceramic fillers within composite materials, magnetic recording media, catalyst supports, and sensing elements [9]. Most notably, HfO2 and ZrO2 NPs have found widespread applications in optical and protective coating
∗ Corresponding author. Tel.: +1 716 645 4220; fax: +1 716 645 6963. ∗∗ Corresponding author. Tel.: +1 716 645 4140; fax: +1 716 645 6963. E-mail addresses:
[email protected] (D.S. Aga),
[email protected] (S. Banerjee). 1 Current address: CSIRO Land and Water, Advanced Materials Transformational Capability Platform, Nanosafety, Biogeochemistry Program, Waite Campus, Waite Rd, SA 5155, Australia. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.028
technologies due to their thermal stability and high dielectric constants [10]. In particular, these materials are promising alternatives to SiO2 as gate dielectric layers for flexible electronics [11–13]. The underlying premise of flexible electronics has been affordability and ubiquitous availability on standard media such as paper and cloth. Consequently, with increasing commercial production, consumer use, and end-user disposal, the release of these NPs to different environmental compartments (especially water and soil) is inevitable. In particular, waste generated during manufacturing processes will have a high concentration of NPs. Although, the likelihood of the release of MO NPs affixed within device structures is not high, environmental discharge may occur over a protracted period of time, especially as a result of material abuse and towards the end of product life. Given the low proposed cost of flexible electronic devices and lack of specifications regarding disposal outlined by manufacturers, several of these materials may eventually enter landfills and run-off streams through household waste disposal. A systematic understanding of partitioning behavior, potential mobilities, and persistence of MO NPs is thus necessary for
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evaluating potential ecological hazards and framing informed policy [1]. Up till now, the environmental fate and transport of ENMs have been characterized under different environmental conditions (i.e., pH, ionic strength, organic colloids, etc.). The influence of natural organic matter (NOM) on NP behavior in the environment has been emphasized in many studies because of the ubiquity of the former in aquatic and soil environments. Indeed, the nature and amount of NOM in water have been demonstrated (both theoretically and experimentally) to affect the stability and bioavailability of a variety of NPs [14–20]. In many cases, the stability of NPs in aqueous suspension is often attributed to the adsorption of NOM. However, these reports only tend to deem adsorption as a primary mechanism for interaction based on indirect measures such as transmission electron microscopy (TEM), -potential, and light scattering, which are somewhat limited in characterizing surfacerelated interactions. An evaluation of the chemical structure of NOM and colloidal MO-NPs suggests several distinctive processes that can facilitate the stabilization of NPs in the aqueous phase. NOM has a highly complex molecular structure, which includes a skeleton of alkyl and aromatic units with pendant functional groups including carboxylic acid, phenolic hydroxyl, and quinone moieties [21–23]. MO-NPs, on the other hand, comprise an inorganic core passivated by a layer of organic ligands [11,24]. The crystalline core can adopt different crystal structures depending on the specifics of composition and stoichiometry (HfO2 and ZrO2 NPs adopt monoclinic and tetragonal crystal structures, respectively) [25,11]. The intrinsic morphological, energetic, and surface chemical properties of NOM enable them to interact with and stabilize different species via amphiphilic and metal-chelating processes. In addition to the organic ligands that surround the crystalline core, MO-NPs also have highly reactive surfaces (edge and corner sites) that greatly influence their behavior and reactivity [26–28]. While these structural and surface characteristics are well known to be important in surface science, these details have typically been overlooked in many fate and transport studies. Very few studies thus far have focused on the transformations of ligand-passivated NPs prepared by hot colloidal chemistry methods upon interactions with NOM. Such ligand-capped NPs are indeed likely to be the mainstay of most nanoscience-enabled technologies [29–32]. Herein, we describe systematic studies on the interaction of ligand-passivated HfO2 , ZrO2 , and Hfx Zr1−x O2 NPs with Suwannee River humic acid (HA) as a model for NOM. The following topics have been addressed in this work: (1) the partitioning of hydrophobically coated MO-NPs, with or without HA, in the aqueous phase; (2) examination of interactions that enable HA to colloidally stabilize MO-NPs (i.e., electrostatic, coordinative, and dispersive interactions); and (3) the importance of both NP surface (crystal structure: monoclinic or tetragonal) and passivating ligands (surface coating: tri-n-octylphosphine oxide (TOPO)) on the stabilization of NPs by HA in water; TOPO is a ubiquitous ligand in nanoscience and is commonly used in hot colloidal synthesis. This study provides a mechanistic report of the aqueous-phase stabilization of different types of hydrophobically coated MO-NPs with and without HA. In particular, we have attempted to directly characterize processes that govern the adsorption and agglomeration of MO-NPs with HA.
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were synthesized by the non-hydrolytic sol–gel condensation of metal alkoxides with metal halides using TOPO (Strem Chemicals, MA, USA) as the coordinating ligand [11]. This synthetic approach provides monodispersity and excellent control over crystal structure, size, and stoichiometry. The TOPO ligand coordinates to surficial atoms, completing the coordination shell for undercoordinated metal sites, and thereby serving as a passivating coating. The use of monodisperse systems provides standardization required for careful mechanistic studies and precludes obscuration from polydispersities in particle size, surface capping, and crystal structure, which would substantially complicate studies of MO-NPs with NOM. All MO-NP powders were readily dispersible in hexane. Approximately 500–1000 mg/L NP suspensions were prepared and used for the phase transfer experiments; these concentrations were chosen for ease in detection of NPs. No consensus has yet emerged on what constitutes an environmentally realistic concentration of MO NPs. The use of the said concentrations allows us to deploy standard microscopy and spectroscopy tools for elucidation of the nature of NP–NOM interactions and indeed such interactions are likely to persist even at low concentrations. For aqueous solubility tests, NP powders were dispersed in water. Suwannee River HA (SRHA-II) standard was purchased from International Humic Substances Society (St. Paul, MN, USA). The use of well-characterized SRHA-II also enables standardization, articulated at various international workshops as an urgent goal for establishing generalizable means of evaluation of ENM fate and transport. Given our primary goal of elucidating the mechanistic basis for ENM phase transfer, the use of well-characterized NOM acquires paramount importance. For Hf/Zr analysis, metal standards (fluoride-soluble metals) and Aristar Ultra grade concentrated HF and HNO3 from BDH Chemicals (West Chester, PA, USA) were used in standard preparation and acid digestions, respectively. Deionized (DI) water from a Barnstead NANOpure (USA) water system was used to prepare all aqueous solutions (resistivity = 18.2 M/cm). 2.2. Phase transfer experiments A 5-mL aliquot of the MO-NP suspension (in hexane) was mixed with 5 mL of 20 mg/L HA in DI water (pH ∼ 4.4) in a clear vial. This experimental construct is referred to as a “phase transfer set-up”. A 20 mg/L HA solution contains 12.5 mg/L dissolved organic carbon that is within levels typical of natural waters (0.1–200 mg/L) [33]. The low natural pH used here is also representative of the low pH of the Suwanee River where the HA was sampled. Similar conditions were used in our previous work on CdSe QDs [30,31]. Phase transfer set-ups were also prepared in DI water (no HA) to serve as controls. Our intention was to study the interactions of MO-NPs with analogs of actual environmental samples, containing controlled concentrations of HA. In between measurements, each set-up was protected from light using Al foil, and was stirred continuously at room temperature using a rocking platform shaker to stimulate natural mixing and fluid diffusion processes. Mixing was performed for 15 days (∼2 weeks). For metal analyses, individual phase transfer set-ups (0, 1, 3, 5, 10, and 15 days) were prepared; this avoids sampling errors due to interfacial aggregation (i.e., removal of flocculated aggregates as the volume of solution is reduced).
2. Experimental
2.3. Analysis and instrumentation
2.1. Materials
Qualitative and quantitative analyses were performed using -potential measurements, dynamic light scattering (DLS), TEM, Raman and Fourier transform infrared (FTIR) spectroscopies and inductively coupled plasma mass spectrometry (ICP-MS).
The monoclinic (m-) HfO2 and Hf0.37 Zr0.63 O2 , and tetragonal (t) ZrO2 and Hf0.37 Zr0.63 O2 NP powders used in these experiments
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Fig. 1. HRTEM images of (A1–2) the m-HfO2 and t-ZrO2 NPs and (B3–4) the HA-transferred MO-NPs in aqueous solution. Insets (i–ii) highlight the predominant crystal planes of the NPs. Lattice spacings were assigned based on monoclinic (JCPDS# 780050) and tetragonal (JCPDS# 881997) structures.
-Potential and DLS data were acquired using a Zetasizer Nano ZS90 instrument (Malvern Instruments, Malvern Hills, UK). TEM measurements were performed using a JEOL JEM-2010 (Tokyo, Japan) operating at an accelerating voltage of 200 kV. Raman spectra were acquired at room temperature using a Horiba JobinYvon (Villeneuve d’Ascq, France) Labram HR Raman spectrometer using 784.51 nm laser excitation from a diode laser. FTIR spectra were collected using a Nicolet-Magna (USA) 550 spectrometer purged with dry air with a spectral resolution of 4 cm−1 . Total Hf and Zr concentrations present in the organic and aqueous phase were quantitatively determined by ICP-MS. ICP-MS measurements were conducted using a Thermo Scientific (Germany) X-Series 2 instrument. Concentrations of Hf and/or Zr in the samples were determined using an external calibration curve. Details on sampling and the acid digestion protocol are described in the Supporting information (SI).
3. Results and discussion 3.1. Characteristics of the MO-NPs Figs. 1A and S1A depicts TEM and lattice-resolved high-resolution TEM (HRTEM) images of m-HfO2 , t-ZrO2 , mHf0.37 Zr0.67 O2 , and t-Hf0.37 Zr0.67 O2 NPs. All NP surfaces are passivated with TOPO and phosphonate ligands. The m-HfO2 particles adopt an elongated rice-grain-type morphology with aspect ratios ranging from 3 to 4. In contrast, the t-ZrO2 NPs adopt a quasi-spherical morphology with an average diameter of 3.3 nm. Solid-solution m-Hf0.37 Zr0.67 O2 NPs are slightly elongated, whereas t-Hf0.37 Zr0.67 O2 NPs are quasi-spherical [11]. Representative XRD patterns of the MO-NPs used in the phase transfer experiments are shown in Fig. S2. Figs. 1A and S1A demonstrate the monodispersity of the NPs used. The HRTEM images indicate
D.A. Navarro et al. / Journal of Hazardous Materials 196 (2011) 302–310
the exposed crystal facets for each set of particles. The monoclinic NPs consistently show a preference for {1 0 0} and {1 1 1} crystal facets, whereas the tetragonal NPs predominantly exhibit {1 0 1} surface-terminating planes. These assignments are consistent with surface energy calculations for m-HfO2 and t-ZrO2 , which indicate preferential energetic stabilization for these planes in the monoclinic and tetragonal crystal structures [34,35]. Given the hydrophobic nature of TOPO, it would be reasonable to expect to first approximation that such NPs will have insignificant dispersibilities or mobilities in the aquatic environment. Indeed, these NPs exhibit very low aqueous solubilities: <0.01 g/L of Hf and <0.02 g/L of Zr for the m-HfO2 and t-ZrO2 NPs, respectively. 3.2. Phase transfer of MO-NPs 3.2.1. Visual examination In this study, to probe the interactions of NOM with well-defined MO-NPs, we have vigorously stirred hexane suspensions of the four systems noted above with aqueous solutions of SRHA. As evidenced by the TEM images (Figs. 1B and S1B), some phase transfer of MONPs into the aqueous phase is noted for all the systems tested here, although there are some key differences in the magnitude of phase transfer depending on the crystal structure (vide infra). Comparison of TEM images (Figs. 1B/S1B to1A/S1A) clearly indicates the formation of MO-HA agglomerates upon phase transfer to the aqueous phase. The images are consistent with the formation of extended amphiphilic humic structures through associations that are stabilized by dispersive hydrophobic interactions, hydrogen bonding, and intramolecular rearrangements [36]. Despite extensive agglomeration of the MO and HA units, the lattice planes of crystalline MO-NPs enable their identification within the amorphous HA matrix. To first approximation, the sizes of the crystalline cores of the NPs do not appear to be affected but considerable agglomeration of NPs within each colloidal humic entity is evidenced. Figs. 2 and S3 are digital photographs summarizing the results of the mixing experiments. The top layer is the hexane (organic) phase and the bottom layer is the aqueous phase. Agglomeration of the NPs results in significant visible light scattering, observed as cloudiness/turbidity of the solutions. Clearly, at the start of
305
the experiments, the lower aqueous phase is optically transparent. After 5 days of mixing, appreciable turbidity develops in the aqueous phase for the m-HfO2 and m-Hf0.37 Zr0.63 O2 NPs, whereas significantly less turbidity is observed for the tetragonal samples; interfacial accumulation of NPs between the organic and aqueous phases is observed to varying extents for the different NPs. After 15 days of mixing, phase transfer and uniform dispersion of NPs in the aqueous phase are evidently most pronounced for the m-HfO2 and m-Hf0.37 Zr0.63 O2 samples with the top organic layer almost completely clear. Some interfacial accumulation and phase transfer are observed for the m-HfO2 and m-Hf0.37 Zr0.63 O2 NPs even without the addition of HA. However, in the presence of HA, the coated MO-NPs exhibit greater colloidal dispersion and stability with respect to agglomeration upon phase transfer, whereas in the absence of HA, interfacial accumulation appears to dominate and even the particles that are phase transferred to the aqueous phase eventually flocculate from suspension. In stark contrast, the t-ZrO2 and t-Hf0.37 Zr0.63 O2 NPs show a significantly lower tendency for phase transfer. 3.2.2. Quantitation by measurement of Hf/Zr concentrations ICP-MS measurements provide a more quantitative perspective of the colloidal stabilization of the MO-NPs achieved in the aqueous phase. Herein, our goal has been to demonstrate the gradual increase in the concentration of NPs (based on Hf/Zr concentrations) dispersed in the aqueous phase and to compare the extent of phase transfer for monoclinic vs. tetragonal NPs. Hf and Zr concentrations in the organic and aqueous phases (Table 1 and Figs. 3 and S4) show the transfer characteristics of the different MO-NPs over a 15-day period. As also suggested by Figs. 2 and S3, the data show pronounced differences between the phase transfer of monoclinic and tetragonal NPs and similarities between phase transfer in H2 O and in HA. After mixing for 15 days with or without HA, the organic phase appears to be almost completely depleted of m-HfO2 NPs. The high HfO2(aq.) /HfO2(org.) ratio (Table 1) and the high levels of Hf found in the aqueous phase (∼260 mg/L in H2 O and ∼280 mg/L in HA) indicate that the m-HfO2 NPs are significantly partitioned into the aqueous phase. In contrast, very little phase transfer appears to have occurred for the t-ZrO2 NPs with or without HA, which is also manifested in the ZrO2(aq.) /ZrO2(org.) ratio and low levels of Zr
Fig. 2. Digital photographs illustrating phase transfer of the m-HfO2 and t-ZrO2 NPs in 20 mg/L HA compared to the control set-ups in H2 O. The top layer is the hexane phase and the bottom layer is the water phase. Similar images for Hf0.37 Zr0.63 O2 NPs are in Fig. S3.
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Table 1 Quantitative information on the phase transfer of the different NPs in 20 mg/L HA and H2 O (control) after 15 days. Ratio of NP concentration in the aqueous and organic phase is determined from the concentration of Hf/Zr in each phase, as measured by ICP-MS.
HfO2 , ZrO2 Monoclinic Hf0.37 Zr0.63 O2 b Tetragonal Hf0.37 Zr0.63 O2 b
H2 O
HA
61 0.10 17, 16 0.02, 0.02
67 0.13 5.4, 5.5 0.02, 0.02
149 131
236 248 266 319
Intensity (a.u.)
Distribution ratio (Diao )a
MO sample
499
639
379
679
578 HfO2
HfO2-HA
HfO2-H2O
a
Diao = [MO nanocrystal]aqueous /[MO nanocrystal]organic = [Hf or Zr]aqueous /[Hf or Zr]organic . b Diao are reported as ratios calculated from Hf, Zr.
found in the aqueous phase (∼5 mg/L Zr in H2 O and HA). To determine whether the observed selectivity of phase transfer arises from differences in chemical composition (m-HfO2 vs. t-ZrO2 ) or crystal structure (monoclinic vs. tetragonal), the same experiments have been performed for solid-solution Hf0.37 Zr0.63 O2 NPs with identical composition but different crystal structures. It is evident from the results summarized in Table 1, Figs. 3 and S4 that the crystal structure (and not composition) has the predominant effect on the observed phase transfer: significant phase transfer is observed for m-Hf0.37 Zr0.63 O2 but not t-Hf0.37 Zr0.63 O2 NPs (analogous to the differences noted above for m-HfO2 and t-ZrO2 NPs). After the 15-day period, m-Hf0.37 Zr0.63 O2 NPs had higher distribution ratios than the t-Hf0.37 Zr0.63 O2 NPs (Table 1). Levels of Hf (and Zr) were also higher in the aqueous phases of the m-Hf0.37 Zr0.63 O2 NPs (∼150 mg/L Hf in H2 O and ∼130 mg/L Hf in HA) than the t-Hf0.37 Zr0.63 O2 NPs (∼4 mg/L Hf in H2 O and ∼1 mg/L Hf in HA). It is also interesting that the Diao calculated using concentrations from Hf and Zr correlated well with each other (Table 1) suggesting that intact NPs were transferred. Some differences in the phase-transfer behavior of m-HfO2 and m-Hf0.37 Zr0.63 O2 are also apparent in Fig. 3. The m-HfO2 NPs reached equilibrium within 1 day of mixing, whereas the m-Hf0.37 Zr0.63 O2 NPs reached equilibrium only after 10 days of mixing; the value of Diao is also higher for m-HfO2 NPs. Taken together, this set of data implies the primary importance of crystal structure but also shows some distinctions based on the specific Hf content in the NPs.
200
400
600
800
Raman shift (cm-1) Fig. 4. Raman spectra for as-prepared TOPO-coated m-HfO2 NPs and phasetransferred HfO2 NPs transferred in the presence or absence of HA showing the 100–800 cm−1 region highlighting the different Raman modes characteristic of monoclinic HfO2 NPs.
close similarity to the spectrum of the pure NPs. The spectra show characteristic signatures of the monoclinic phase with 14 of the 18 predicted Raman phonon modes for m-HfO2 clearly identifiable in the 100–800 cm−1 range [25]. In other words, the Raman spectra suggest retention of the monoclinic phase even after phase transfer; this data further corroborates TEM evidence that there is no discernible surface reconstruction or change in crystal structure. While some leaching of Hf/Zr ions cannot be ruled out, from the Raman data, there is no evidence for the speciation of other solid phases such as metal hydroxides and phosphates. 3.3.2. Surface charge A comparison of the -potential distributions for the MO-NPs, HA solution and phase-transferred samples is shown in Figs. 5 and S5. In general, the as-prepared NPs in hexane have broad potential distributions (positive to negative), whereas HA in water has an overall negative -potential (centered at −20.4 mV). The presence of both positive and negative particles is consistent with hot colloidal syntheses where coordinatively unsaturated cationic and anionic sites are remnant on the NP surface (based on sterics alone, TOPO can not passivate every cationic and anionic site) [41].
3.3. Characterization of the phase-transferred MO-NPs HfO 2
Intensity (a.u.)
3.3.1. Crystal structure Fig. 4 shows Raman spectra of the as-prepared colloidal m-HfO2 NPs along with spectra for solid samples freeze-dried after transfer to the aqueous phase with or without the presence of HA. Notably, the spectrum for the phase-transferred m-HfO2 -HA adducts bears
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Time (d) Fig. 3. Concentration vs time plots for the phase transfer of m-HfO2 and t-ZrO2 NPs in 20 mg/L HA and in H2 O. Similar plots for Hf0.37 Zr0.63 O2 NPs are in Fig. S4. Error bars correspond to standard deviation of concentrations determined from different sample aliquots (n = 2).
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Zeta Potential (mV) Fig. 5. -Potential distribution of (A) all starting materials, as-prepared TOPOcoated MO-NPs and 20 mg/L HA, and (B) the aqueous phases of the different set-ups after 1 day of continuous mixing. All samples were measured in water except for the NP suspensions which were in hexane. -Potential distribution is from the average of n = 3 measurements.
*HfO2 in HA
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100000
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Time (µs) Fig. 6. DLS correlation data (aqueous phase) collected for the phase transfer of mHfO2 NPs in 20 mg/L HA. Insets show raw correlation data for the control and asprepared HA. Correlation data for other MO NPs are in Fig. S4.
Nonetheless, the resulting NPs are still considered hydrophobic because of the capping TOPO ligands with their pendant hydrocarbon chains, which are strongly hydrophobic and preclude ready access of other species to charged sites [37]. When the NPs are mixed with HA for 1 day, the -potential values in the aqueous phase shifted to more negative values; from −20.4 mV to −36.8 mV (for m-HfO2 in HA) and −43.1 mV (for t-ZrO2 in HA), consistent with other studies [16,17,19]. These values are consistent with the formation of larger, more stable (more negative) HA agglomerates in suspension, as also seen in Fig. 1B. For the MO-HA adducts observed in the TEM, the negative -potential also confirms that the negatively charged humics are present at the surface and envelope the NPs. The shift to more negative -potential values in the presence of HA and the loss of positively charged NPs suggest electrostatic interactions between the two species. Although both m-HfO2 and tZrO2 experienced the same changes in average -potential, ICP-MS results still show that monoclinic NPs have a stronger predilection to phase transfer compared to tetragonal NPs. 3.3.3. Particle agglomeration Agglomeration of the NPs during phase transfer has been further monitored by DLS, as shown in Figs. 6, S6 and S7. Fig. 6 compares the DLS correlation curves acquired at different time intervals (0–50 h). As we have noted in a previous research article, the validity of fitting correlograms measured from fixed-angle DLS measurements to extract exponential decays and an “average” size for systems as heterogeneous and polydisperse as MO-HA agglomerates is questionable [30,31]. Consequently, a more reliable and realistic approach involves qualitatively comparing the directly measured correlograms. The shift of the correlograms to longer times implies that the particles stay correlated over longer periods while undergoing Brownian motion—suggesting the formation of larger more sluggish agglomerates. The measured correlation curves for the aqueous phase after mixing m-HfO2 NPs in hexane with HA for different periods of time are shown in Fig. 6. It is apparent that as the m-HfO2 NPs transfer from hexane to the aqueous phase, larger species (likely MO-HA agglomerates) are formed. With progressively increased transfer of NPs to the aqueous phase, the MO-HA agglomerates tend to expand in size. The MO-HA adducts first appear in the aqueous phase as early as 6 h after initiation of mixing and subsequently increase in size after 22 and 50 h, as evidenced by the correlation curves; the considerable breadth of the decay traces is indicative of substantial polydispersity in the size of the agglomerates, as also evidenced by the TEM images in Fig. 1B. For the control sample (m-HfO2 NPs in H2 O), no evidence
4000
3000
2000
1000
Wavenumbers (cm-1) Fig. 7. FTIR spectra of as-prepared TOPO-coated m-HfO2 NPs, phase-transferred MO-NPs and HA showing the 4000–400 cm−1 region.
of adducts is observed before 22 h. Hence, HA considerably accelerates phase transfer in addition to better dispersing the m-HfO2 NPs in the aqueous phase. DLS data presented in the inset also demonstrate that substantial agglomeration of HA does not occur under these conditions confirming that the observed increase in the times for the correlation curves to decay to the baseline must arise from the transfer of m-HfO2 NPs to the aqueous phase. Similar results are obtained for the m-Hf0.37 Zr0.63 O2 NPs (Fig. S6). For the tetragonal NPs exhibiting minimal phase transfer, the DLS correlation curves are shown in Fig. S7. DLS data are unable to pinpoint the specific interactions between HA and MO-NPs that induce phase transfer but clearly establish the role of the HA in this process and are consistent with the formation of MO-HA adducts. Note that although HA provides some extent of colloidal stabilization to the monoclinic NPs, it is a much bulkier molecule that is prone to crosslinking and does not have the steric or electrostatic properties required to completely preclude NP agglomeration over an extended period of time (>2 weeks). 3.3.4. Surface chemistry To further investigate the nature of the interactions between HA and the MO-NPs, FTIR spectroscopy data for the phase-transferred MO-HA adducts are compared to the data for the individual constituents (Fig. 7). The infrared spectrum of the as-prepared m-HfO2 NPs is dominated by absorptions that can be attributed to the TOPO ligands. In particular, a strong P O stretch is observed at 1100 cm−1 and asymmetric and symmetric C–H stretches arising from the alkyl groups are observed at ∼2850–2930 cm−1 [11]. In contrast, HA shows a characteristic O–H band at ∼3450 cm−1 and asymmetric COO stretches at 1720 cm−1 and 1623 cm−1 [18,38,39]. FTIR spectra of m-HfO2 -HA adducts recovered from the aqueous phase clearly indicate some peaks characteristic of TOPO, specifically C–H stretching bands at 2930 cm−1 and 2850 cm−1 as well as the P O stretching band at ∼1100 cm−1 , though these are significantly diminished relative to the corresponding peaks for TOPO-coated NPs. The presence of these bands indicates that the TOPO ligands on the NP surfaces are not completely displaced by the functional groups on HA (or by H2 O molecules). Since TOPO has a very low aqueous solubility (0.058 mg/L in H2 O at 25 ◦ C [40]), our results suggest that HA has enabled the phase transfer and aqueous stabilization of the TOPO-coated MO-NPs. Some peaks characteristic of HA are also apparent in the FTIR spectra for the m-HfO2 -HA adducts, specifically the asymmetric COO stretching bands. In the spectrum acquired for pure HA, prominent asymmetric COO stretches, corresponding to carboxylic acid (as (CO2 H)) and carboxylate (as (CO2 − )) moieties [30,31,39], are present at relatively equal intensities. In the spectrum of the
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m-HfO2 -HA adducts, there is a significant diminution in the intensity of the (as (CO2 H)) band at 1720 cm−1 and instead a single prominent as (CO2 − ) band is observed at 1613 cm−1 . This shift to lower wavenumbers and dimunition of the as (CO2 H) peak can be attributed to the formation of metal-carboxylate linkages. The precise coordination mode (i.e., monodentate, bidentate) however, cannot be clearly determined since the symmetric carboxylate stretching bands are not adequately well resolved. Although studies of metal-humate complexes with higher valent metal ions are limited, the binding of HA to hafnium has indeed been demonstrated [44]. 3.4. Proposed phase transfer mechanism Taken together, the data presented above provide a clear picture of multiple interfacial interactions that enable the phase transfer and aqueous dispersal of monoclinic MO-NPs. As noted previously for the NOM-induced phase transfer of CdSe quantum dots [30,31], two distinct stages can be demarcated: (1) the flocculation of the MO-NPs at the hexane/water interface wherein a turbid interfacial layer is formed and a large density of surface sites are primed for interactions with H2 O or HA, and (2) dispersion of the NPs from the interface into the aqueous layer where they form a colloidal dispersion that is stable over >2 weeks. In the absence of HA, phase transfer occurs at a relatively slower rate and the transferred MONPs tend to flocculate and eventually settle at the bottom of the vial. Indeed, the TEM images, -potential, DLS, and FTIR spectral data all suggest direct interaction between the MO-NPs and HA. Three distinctive interactions of HA with the MO-NPs leading to aqueous phase dispersal and stabilization can be envisaged [30,31,45]. The first process involves overcoating of the MO-NPs with HA through non-specific adsorption of humics onto the TOPO ligand shell. The amphiphilic characteristics of HA enable the formation of pseudomicellar agglomerates that overcoat the NPs, wherein hydrophobic aromatic and heteroaliphatic regions form a hydrophobic interior cavity, whereas pendant hydrophilic carboxylic acid, phenolic, and amine moieties are directed outwards imparting colloidal stability in water via electrostatic stabilization [38,46]. Provided that the NP surface ligands are engaged in this interaction, this mechanism is not expected to show any discrimination between monoclinic and tetragonal crystal structures. The second process involves dative interactions between HA and the MO surfaces, where the carboxylic acid (estimated to be 10% in SRHA-II, [47]), phenolic, and amine moieties that are abundant in the HA structure serve as versatile polydentate chelating ligands [38,48–51]. Interaction involving the NP surface is conceivable given the highly reactive surfaces of nanoscale materials, wherein most of the constituent atoms reside at or near the surface [27,28,52]. This mechanism is predicated on the availability and accessibility of metal sites on the MO-NP surface that can participate in coordinative interactions. Different crystal structures have distinctive planes, corners, and edges exposed at the surface (Fig. 1). Hence, in contrast to the overcoating mechanism, this interaction can possibly provide some discrimination between monoclinic and tetragonal crystal structures. The third process involves electrostatic interactions between the coordinatively unsaturated surface sites on the MO-NPs (positively and negatively charged) and the carboxylic and phenolic moieties (negatively charged) in the HA structure. these processes (micellar/overcoating, coordinaAll tive/substitutional, and electrostatic interactions) appear to work synergistically to facilitate the phase transfer and subsequent aqueous-phase dispersal of monoclinic MO-NPs passivated with TOPO but not the tetragonal NPs. The differences in the phase transfer behavior observed between the two different polymorphs suggest that specific rather than non-specific interactions mediate
phase transfer; in this case, the surface-structure-dependent coordinative/subsitutional interactions are the likely genesis of the distinctive reactivity. Upon mixing with H2 O/HA, some of the TOPO ligands are likely displaced; FTIR spectra of phasetransferred MO-HA agglomerates suggest that the amount of TOPO is diminished compared to the TOPO-coated NPs. Removal of TOPO provides access to oxophilic, coordinatively unsaturated cationic sites on the MO-NP surfaces that can be accessed by the carboxylic acid moieties of HA (electrostatic interactions likely induce the initial approach of the HA moieties and NPs). Apart from ligand displacement, incomplete coverage/passivation of the initial NPs by TOPO would also make Hf/Zr surface sites available for binding to HA. Impurities in technical grade TOPO (90%), particularly alkylphosphonic and alkylphosphinic acids that have been shown to play an active role in passivating surfaces of CdSe quantum dots, rods, and wires [53–56], could also influence the interactions between HA and the NP surface [31]. The fractional surface coverage of the coating groups and the precise nature of the ligand passivation shell are beyond the scope of this study (ligand passivation shells remain to be adequately characterized even for CdSe quantum dots that are possibly the most mature of this class of materials). Nonetheless, as suggested by FTIR, the formation of metal–humate linkages tethers the MO-NPs to the humic colloids, which likely draws the NPs to the hexane/water interface such that both the hydrophobic NPs and the hydrophilic HA colloids can be adequately solvated. Subsequently, as described in the literature [46,49,57], the flexible humic colloid can undergo molecular rearrangement and cross-linking with proximal HA moieties at the hexane/water interface to optimize hydrophobic interactions with the pendant aliphatic chains on the TOPO ligand [41–43,58]. In addition, our results indicate that H2 O by itself enables some phase transfer of NPs. The oxophilicity of early transition metal oxide surfaces may also allow for facile ligand substitution by H2 O, which can eventually result in appreciable hydroxylation of the MO surfaces. The affinities of different crystallographic facets for HA may reasonably be assumed to parallel the likelihood of ligand substitution by H2 O, which may explain the significant phase transfer observed for the monoclinic NPs even in the absence of HA. Phase transfer only shows differences between H2 O and HA on shorter timescales, when the MO-H2 O interaction is likely limited by ligand substitution. The interfacial turbidity and shorter-lived phase transfer noted in control samples in the absence of HA likely arise from the displacement of some TOPO ligands by H2 O. Consistent with the proposed mechanism, since surface-coordinated H2 O molecules lack the amphiphilic characteristics of HA, they are not able to adequately stabilize the TOPO-coated MO-NPs in the aqueous phase. In other words, ligand substitution of TOPO by H2 O can induce sedimentation at the hexane/water interface but does not permit colloidal stabilization in the aqueous phase in the absence of HA. Preferential binding of HA or H2 O to monoclinic instead of tetragonal surfaces may form the basis for the observed selectivity of phase transfer such as between m-HfO2 and t-ZrO2 NPs, and between m-Hf0.37 Zr0.63 O2 and t-Hf0.37 Zr0.63 O2 . As shown in Fig. 1A, the m-HfO2 (and m-Hf0.37 Zr0.63 O2 ) NPs preferentially expose {1 0 0} and {1 1 1} surfaces, whereas for t-ZrO2 (and tHf0.37 Zr0.63 O2 ) NPs {1 0 1} surfaces are energetically preferred. While the difference between m-HfO2 and t-ZrO2 NPs with regard to phase transfer could be related to the extent/strength of the Hf-HA bonds vs. the Zr-HA bonds, results from m-Hf0.37 Zr0.63 O2 and t-Hf0.37 Zr0.63 O2 NPs (same chemical composition) suggest that phase transfer is more responsive to changes in crystal structure. Consequently, despite both having potentially reactive surfaces, we speculate that the differences in phase transfer of the NPs may originate from (a) the relative binding affinities of the different surfaces in monoclinic and tetragonal NPs for TOPO and HA; (b) the degree of
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coordinative unsaturation and steric hindrance for transition metal sites in the different surface planes; and (c) the density of exposed transition metal cation sites. A combination of these factors could make metal sites less available and the displacement of TOPO ligands by HA functionalities more difficult for tetragonal NPs. In the absence of metal–humate linkages, phase transfer may not be as readily initiated resulting in the poor phase transfer efficiencies observed for t-ZrO2 and t-Hf0.37 Zr0.63 O2 NPs. An analogous preference for binding different surfaces is also expected for coordinative interactions with H2 O molecules. Calculation of binding affinities, extent of surface unsaturation, and density of cation and anions on the NP surface is beyond the scope of this study; these measurements have indeed not been experimentally validated for ligand-passivated colloidal NP systems with any degree of accuracy. Nonetheless, our extensive characterization data is adequate to draw some conclusions with regard to the mechanisms that dictate phase transfer and aqueous phase stabilization of these NPs. 4. Conclusions The interactions between HA and MO-NPs indicate the role of HA as both a coordinating ligand and an amphiphilic surfactant. Our results further provide experimental validation of theoretical predictions [14] and experimental observations [15–17] of modifications to the colloidal stability of MO-NPs upon the acquisition of NOM coatings. In this study, HA and H2 O both exhibit a distinct preference for monoclinic rather than tetragonal MONPs, possibly because of stronger binding affinities to monoclinic surfaces (coordinative/substitutional interactions) and the ease of formation of cylindrical pseudo-micellar structures (overcoating/micellar interaction). The extent of phase transfer and degree of colloidal stabilization observed for well-defined MO-NPs with hydrophobic coatings in the presence of HA also underline the importance of developing a detailed understanding of the potential environmental transformations of ENPs. The distinctive selectivity in the HA-induced phase transfer and dispersion of different polymorphs suggests that interactions of different inorganic ENMs with NOM, which dictate the extent of NP transport and stabilization, are not necessarily generalizable and that it may be possible to design NMs to minimize their residence time in the aquatic environment. In this regard, further research is required to determine the actual binding affinities of HA onto monoclinic and tetragonal surfaces and to investigate the influence of different surface coatings. Acknowledgements This work was primarily supported by the US Environmental Protection Agency (Grant# R833861). SB acknowledges partial support of this work from National Science Foundation under DMR 0847169. We acknowledge the NSF MRI Program CHE 0959565 for acquisition of the ICP-MS instrument. Both CSIRO and the US EPA have not subjected this manuscript to internal peer and policy review. Therefore, no official endorsement should be inferred. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.028. References [1] P.J.J. Alvarez, V.L. Colvin, J. Lead, V. Stone, Research priorities to advance ecoresponsible nanotechnology, ACS Nano 3 (2009) 1616–1619. [2] A. Maynard, R.J. Aitken, T. Butz, V. Colvin, K. Donaldson, G. Oberdorster, M.A. Philbert, J. Ryan, A. Seaton, V. Stone, S.S. Tinkle, L. Tran, N.J. Walker, D.B. Warheit, Safe handling of nanotechnology, Nature 444 (2006) 267–269.
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Journal of Hazardous Materials 196 (2011) 311–317
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Reduction of As(V) to As(III) by commercial ZVI or As(0) with acid-treated ZVI Fenglong Sun a , Kwadwo A. Osseo-Asare b , Yongsheng Chen c , Brian A. Dempsey a,∗ a
Department of Civil and Environmental Engineering, Penn State University, UP, United States Department of Materials Science and Engineering, Penn State University, UP, United States c Department of Energy and Mineral Engineering, Penn State University, UP, United States b
a r t i c l e
i n f o
Article history: Received 24 April 2011 Received in revised form 7 September 2011 Accepted 8 September 2011 Available online 14 September 2011 Keywords: Arsenic Zero-valent iron Reduction X-ray absorption spectroscopy (XAS)
a b s t r a c t Zero-valent iron (ZVI) consists of an elemental iron core surrounded by a shell of corrosion products, especially magnetite. ZVI is used for in situ removal or immobilization of a variety of contaminants but the mechanisms for removal of arsenic remain controversial and the mobility of arsenic after reaction with ZVI is uncertain. These issues were addressed by separately studying reactions of As(V) with magnetite, commercial ZVI, and acid-treated ZVI. Strictly anoxic conditions were used. Adsorption of As(V) on magnetite was fast with pH dependence similar to previous reports using oxic conditions. As(V) was not reduced by magnetite and Fe(II) although the reaction is thermodynamically spontaneous. As(V) reactions with ZVI were also fast and no lag phase was observed which was contrary to previous reports. Commercial ZVI reduced As(V) to As(III) only when As(V) was adsorbed, i.e., for pH < 7. As(III) was not released to solution. Acid-treated ZVI reduced As(V) to As(0), shown using wet chemical analyses and XANES/EXAFS. Comparisons were drawn between reactivity of acid-treated ZVI and nano-ZVI; if true then acid-treated ZVI could provide similar reactive benefits at lower cost. © 2011 Elsevier B.V. All rights reserved.
1. Introduction: High arsenic concentrations in groundwater have been reported throughout the world, notably in Bangladesh and Taiwan [1]. Arsenic in aquatic environments usually exists in inorganic forms as arsenate (As(V)) and arsenite (As(III)). As(III) is usually more mobile and toxic than As(V) although As(III) can also become immobilized in the presence of sulfide. Elevated arsenic concentrations in groundwater can occur due to reductive dissolution of ferric oxide sorbents and consequent reduction and mobilization of As(III), desorption of As(V) under alkaline pH conditions especially in the presence of phosphate or other competing adsorbates, or oxidation of sulfidic materials [1,2]. Zero-valent iron (ZVI) has been used to remove organic and inorganic contaminants including chlorinated solvents, nitrate, uranyl ion, chromate, lead, and arsenic [3,4]. ZVI can be incorporated into permeable reactive barriers or nano-ZVI (nZVI) can be injected into contaminated soils [5]. ZVI is also found in some point-ofuse potable water treatment systems [6]. ZVI is usually reported to have a core–shell structure. The shell contains oxidized iron that is mostly magnetite and often with maghemite (␥-Fe2 O3 ) or lepidocrocite (␥-FeOOH) [3,7–10].
∗ Corresponding author. Tel.: +1 814 865 1226; fax: +1 814 863 7304. E-mail address:
[email protected] (B.A. Dempsey). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.029
ZVI immobilizes arsenic by adsorption of As(V) or As(III) onto iron corrosion products in the shell surrounding the elemental iron core and this is sometimes accompanied by reduction of As(V) to As(III) [11–14]. Detailed mechanisms have been reported for removal of many contaminants by ZVI. Fe(0), dissolved Fe(II), solidbound Fe(II), and H2 have been proposed as elementary reductants [3,9,15]. However there is still controversy about the mechanisms for removal of arsenic especially regarding redox reactions. The rate and extent of As(V) reduction by ZVI may depend on the experimental conditions. In different studies, 25% of initial As(V) was reduced to As(III) by nano-ZVI at neutral pH after 90 days [13,16], As(V) was partially reduced to As(III) by commercial ZVI at slightly basic pH after 60 days [17], and there was no As(V) reduction using iron wires [18]. It was also reported that As(III) was reduced to As(0) with acid-treated iron filings [19]. Magnetite is often observed to be a dominant component in the corroded ZVI shell. In an effort to identify mechanisms by which ZVI immobilized contaminants, Lago and co-workers [20,21] used mechanical grinding to produce a magnetite/ZVI reactant that reduced methylene blue, H2 O2 , and Cr(VI). Other iron oxides (␣Fe2 O3 , FeOOH, or ␥-Fe2 O3 ) mixed with ZVI were much less reactive. It was suggested that the semi-conductor behavior of magnetite was important for effective reduction of contaminants. The reactivity of magnetite may depend on whether the ZVI has been in contact with air. In this context White and Peterson [22] showed that magnetite reduced Cr(VI) at a much faster rate under anoxic conditions than under oxic conditions. It has also been shown that
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“stoichiometric magnetite” that is produced and maintained under strictly anoxic conditions has a Fe(II)/Fe(III) ratio close to 0.5 and is a stronger reducing agent than “non-stoichiometric magnetite” that has been produced or stored in the presence of O2 resulting in a Fe(II)/Fe(III) ratio <0.5 [23]. Magnetite is a strong adsorbent for arsenate [24–26]. Magnetite in the presence of Fe(II) is theoretically capable of reducing arsenate to arsenite (discussed later in this paper) but to our knowledge there are no reports of As(V) reduction using magnetite. There are several previous studies dealing with the reduction of arsenate by ZVI but these were all conducted under oxic conditions that can result in the production of “nonstoichiometric magnetite”. The objectives of this study were to identify the reaction mechanisms and reaction products when As(V) is remediated using ZVI. Experiments were conducted using three solid phases: (1) stoichiometric magnetite, (2) commercial ZVI, and (3) ZVI that had been acid-treated to disrupt the adherent layer of magnetite. Experiments were run under rigorously anoxic conditions [27]. The rate and extent of reactions were followed by analyzing for As(V) and As(III) in both dissolved and solid phases. The nature of the reduced arsenic species was also investigated using X-ray absorption near-edge structure (XANES) and extended X-ray absorption fine structure (EXAFS).
As(V)/magnetite experiments with fixed pH were performed in 250 mL bottles. After adding 100 mL of 0.65 mM as Fe3 O4 (0.15 g/L) the bottles were connected to the O2 -trap in the anaerobic chamber and were continuously shaken (VWR Model 3750 500 rpm). After 8 h, As(V) stock solution was injected to achieve 100 M. Suspension samples (1.5 mL) were extracted by syringe and needle through the cap for subsequent measurement of dissolved and adsorbed As(V)/As(III) concentrations. As(V)/magnetite experiments with variable pH were conducted in five polypropylene vials (30 mL). Twenty millilitre 0.65 mM magnetite suspension and As(V) stock solution to obtain 50 M were added to each vial. The pH was adjusted with concentrated NaOH or HCl. The vials were placed in a dark box on a shaker in the anaerobic chamber for 48 h. Suspension samples were taken at intervals for determination of dissolved and adsorbed As(V) and As(III). Reactions of As(V) and commercial or acid-treated ZVI were conducted in 30 mL polypropylene vials, with sequential additions of 2.0 g ZVI (100 g/L), 20 mL deoxygenated DI water, As(V) stock solution, and pH adjustment. Samples were handled as in the As(V)/magnetite experiments. Dissolved arsenic was determined after centrifugation (Eppendorf MiniSpin Microcentrifuge) of 1 mL samples. Adsorbed As(V) and As(III) were measured after NaOH extraction. 2.3. Characterization of solids
2. Materials and methods 2.1. Materials and chemicals Commercial ZVI (Alfa-Aesar, 99.2% Fe and ∼0.3 mm diameter) was used as received and also after acid pre-treatment. Acid-treated ZVI was prepared by adding 50 mL 1 M HCl to 50 g commercial ZVI to dissolve the oxidized film [28]. The slurry was aerated for 2 h, washed 3 times with DI water, dried overnight at 80 ◦ C, and stored in the anaerobic chamber. Commercial ZVI materials were also stored in the anaerobic chamber. Magnetite was prepared in the anaerobic chamber by mixing equal volumes of 4.0 mM FeCl3 (1.08 g/L as FeCl3 ·6H2 O) and 2.0 mM FeCl2 (0.40 g/L as FeCl2 ·4H2 O) and pH adjustment to 11 [24]. Settled precipitates were washed with deoxygenated water three times. XRD analysis showed that the sample was magnetite. No other solid phases were identified. Samples were completely dissolved in concentrated HCl and then analyzed for Fe(II) and Fe(III). The Fe(II)/Fe(III) molar ratio was 0.47, thus the magnetite was stoichiometric Fe3 O4 . All stock solutions were prepared in deoxygenated Milli-Q water in an anaerobic chamber (Coy Laboratory Products, Inc., 5% H2 in N2 with Pd catalyst). Deoxygenated water was prepared by bubbling 99.9% nitrogen gas while heating to 70–80 ◦ C for at least five hours before transfer and storage in the anaerobic chamber.
2.2. Wet chemistry All experiments were conducted in reactors that were connected to an O2 -trap (in the anaerobic chamber) to maintain <7.5 × 10−9 atm of O2 [27,29] unless specified. As(V) and the sum of As(V) + As(III) (after oxidation with 0.2 mM potassium iodate at pH 1) were determined by molybdenum blue colorimetric analysis [30]. Adsorbed arsenic was extracted with 5 M NaOH in the anaerobic chamber for 24 h. Separate experiments demonstrated that >98% of adsorbed As(V) or As(III) was recovered without change in oxidation state. The detection limit of As(V) by molybdenum blue analysis was ∼1 M. Total Fe(II) plus Fe(III) was measured by adding 100 mM ascorbic acid to 1 g ZVI at pH 3 for 48 h, reducing iron oxides to Fe(II) [31], which was measured with ferrozine [32].
Magnetite was filtered using 0.2 m Teflon membranes and the filter and deposits were dried in the anaerobic chamber at room temperature. Commercial and acid-treated ZVI were characterized without further treatment. XRD patterns were measured with a PANalytical Theta-2-Theta Powder Diffractometer using Cu K␣ radiation at 60 kV and a PIXcel solid detector. SEM was performed using a FEI Quanta 200 scanning electron microscope. A WITec CRM200 confocal Raman spectroscopy configured with 514 nm laser excitation was used to record the Raman spectra. The laser power was kept below 0.5 mW to avoid sample degradation [33]. The laser was focused into a round spot with ∼0.5 m in diameter with a resolution of 1 cm−1 . Samples for XANES and EXAFS analysis were prepared by reacting commercial ZVI (pH 8.5 ± 0.5) and acid-treated ZVI (pH 7) with 100 M As(V) for 5 days. The samples were rinsed and dried, with all operations in the anaerobic chamber. Na2 HAsO4 ·7H2 O and NaAsO2 powders were used as As(V) and As(III) standards. Arsenic K-edge (11.868 keV) EXAFS spectra were collected on the bending magnet beamline (9-BM) at the advanced photon source (APS), Argonne National Laboratory using Si (1 1 1) monochromator crystals. Harmonics were rejected by use of an Rh-coated flat mirror in the experimental station. Energy was calibrated using a Ga foil by setting its K-edge energy to 10,367 eV. The beam was focused to a 1 mm diameter spot size using an Rh-coated toroidal mirror. The XAFS spectra were collected in fluorescence mode using a SII Vortex 4-element Silicon Drift Detector. An aluminum filter was placed before the fluorescence detector to suppress Fe fluorescence. Ten scans were collected and averaged. XANES and EXAFS spectra were processed using Iffefit software (Athena and Artemis) [34]. A straight line to the pre-edge and a spline function to the post-edge region were used for background removal. XANES spectra were normalized to unity edge height. Fitting EXAFS data to retrieve structural information was performed using Artemis [34]. Data processing and fitted parameters are described elsewhere [35]. The As EXAFS oscillations (k) were weighted by k2 and windowed between 2.5 < k < 10 A˚ −1 using a Hanning window with dk = 1.0 A˚ −1 . The fits were to both real ˚ The ˜ (R) in the region of 1.0 < R < 2.0 A. and imaginary parts of amplitude-reduction factor S02 was determined by fitting two standards Na2 HAsO4 ·7H2 O and NaAsO2 [36,37]. Fitting included only
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Fig. 2. X-ray diffraction spectra of (a) commercial ZVI and (b) acid-treated ZVI. Fig. 1. Raman spectra of commercial ZVI and acid-treated ZVI: (a) commercial ZVI; (b) acid-treated ZVI.
3. Results and discussion
tified in the XRD pattern for commercial ZVI because it was below the detection limit (∼3% by weight) [7,38]. Others have reported that commercial ZVI has an oxide surface film [3,7,38,39] with an inner magnetite layer and an outer ␥-Fe2 O3 (maghemite) layer [9]. SEM images in Fig. 3 show that commercial ZVI (a and b) had a smooth surface and acid-treated ZVI (c and d) had a more porous surface that was covered with ∼0.5 m particles.
3.1. Characteristics of commercial ZVI and acid-treated ZVI
3.2. Adsorption of As(V) on magnetite
Iron oxides (from ascorbic acid dissolution) constituted <0.2% (w/w) of commercial ZVI and >10% of acid-treated ZVI. Fig. 1 shows that magnetite was the only iron oxide phase detected by Raman spectroscopy (542 cm−1 and 671 cm−1 ). Fig. 2 shows XRD results, confirming magnetite in acid-treated ZVI. Magnetite was not iden-
The first experiments were conducted with stoichiometric magnetite. Fig. 4a shows that removal of As(V) at pH 6.5 was fast, with half-time (t1/2 ) for adsorption <1 h. The adsorption capacity was ∼0.08 As(V)/Fe3 O4 (mol/mol), similar to adsorption of As(V) onto magnetite under oxic conditions [24]. Desorption with 5 M
one As–O path based on bond distance (R), coordination number (N), and the Debye–Waller factor ( 2 ).
Fig. 3. SEM images of commercial ZVI and acid-treated ZVI at two magnifications: (a and b) commercial ZVI; (c and d) acid-treated ZVI.
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Fig. 4. As(V) adsorption on stoichiometric magnetite in an anoxic environment: (a) As(V) removal versus time at pH 6.5; (b) pH dependence of As(V) removal (As(V) = 50 M, Fe3 O4 = 0.65 mM, 25 days).
NaOH recovered ∼100% recovery of arsenic from magnetite, all in the +5 valence. Fig. 4b shows the pH dependence of As(V) adsorption on magnetite. Adsorption of As(V) decreased at pH above the pHzpc = 6.6 of pristine magnetite [40]. The pKa1 and pKa2 of H3 AsO4 are 2.2 and 7.0. The anions H2 AsO4 − and HAsO4 2− are major dissolved species over the entire pH range for Fig. 4b. An electrostatic model was used by Dixit and Hering [41] to explain the adsorption of As(V)/As(III) on hydrous ferric oxides (HFO) and goethite. Surface species FeH2 AsO4 , FeHAsO4 − and FeAsO4 2− (in the order of increasing pH) were proposed as the surface complexes formed from As(V) adsorption on HFO and goethite. The formation constants of the surface species have this order: k( FeH2 AsO4 ) > k( FeHAsO4 − ) > k( FeAsO4 2− ), which is consistent with decreased As(V) adsorption with increasing pH. As(V) adsorption on magnetite formed similar inner-sphere complexes as on HFO and goethite [42] and similar adsorption fonts have been reported [24–26,42]. As(V) adsorption on magnetite under anoxic conditions in this study showed similar pH dependence as previously reported for oxic conditions. All of the dissolved and adsorbed arsenic in the magnetite experiments was recovered as As(V), even when the experiment was conducted in the presence of excess dissolved Fe(II). Information about the arsenic oxidation state is shown in Fig. S1. Although we used stoichiometric magnetite (a stronger reducing agent than non-stoichiometric magnetite [22,23]), a rigorous anoxic
Fig. 5. As(V) adsorption and reduction to As(III) using commercial ZVI (2.0 g/0.02 L) in an anoxic environment: (a) pH = 6.5 ± 0.5; (b) pH = 8.5 ± 0.5; (c) pH = 10.
environment, and a stoichiometric excess of dissolved Fe(II), magnetite did not reduce As(V) at pH 5–10. 3.3. Reactions of As(V) with commercial ZVI The results of experiments with As(V) and commercial ZVI at pH 6.5 ± 0.5, 8.5 ± 0.5 and 10 are shown in Fig. 5 in which concentrations of dissolved and adsorbed As(V) and As(III) are plotted against time. There was partial reduction of As(V) to As(III) at the two lower pH ranges and some loss of As(V) + As(III) at pH 8.5 ± 0.5.
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Fig. 6. As(V) adsorption and reduction to As(0) using acid-treated ZVI (2 g in 0.02 L) in an anoxic environment at pH 7. No As(III) was produced.
The t1/2 for adsorption and for reduction to As(III) were <10 h at pH 6.5 ± 0.5 and at pH 8.5 ± 0.5. Adsorption was very fast at pH 10 (t1/2 <1 h) even though the extent of adsorption was less than for lower pH values. None of the As(III) was released back into solution, suggesting that the reduction occurred within the bulk magnetite or at the Fe(0)/magnetite interface. Previous investigators reported slow reduction of As(V) with a 60–90 days lag time, which was probably due to less rigorous exclusion of O2 in their experimental systems [13,17]. Dissolved oxygen (DO) could be an important factor that changes the reduction kinetics. It has been reported that As(III) could be oxidized to As(V) on iron oxides in the presence of O2 through a Fenton-like reaction mechanism [43]. In order to study the effect of dissolved oxygen on As(V) reduction by commercial ZVI, additional experiments were conducted in which the closed reactors were removed from the anaerobic chamber after 5 days and placed on a shaker. As(III) was completely re-oxidized to As(V) within a few hours at pH 6.5, 8.5 and 10 by O2 that diffused through the tightly attached screw caps. The results indicated that As(III) oxidation even with low DO was an important back-reaction that inhibits early observation of As(V) reduction by ZVI. The long incubation time observed by previous researchers was likely the time required for complete removal of DO due to direct or indirect reaction with ZVI. Care was taken in the current study to eliminate any O2 in the reaction vials and consequently fast As(V) reduction kinetics were observed.
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Fig. 7. p–pH redox ladder of arsenic-ZVI system. Thermodynamic data were taken from V-minteq (v2.51): Fe2+ = 1 mM, As = 0.1 mM. The redox ladder indicates Fe(0) can reduce As(V) or As(III) to As(0), and Fe(II) is not a dominant species for any p conditions at pH 10.
with the previously observed 20% conversion of As(V) to As(III) at pH 8.5 ± 0.5 (see Fig. 5b). The XANES spectrum for the reaction products of As(V) with acid-treated ZVI contained a shoulder near 11,865 eV. The binding energy for As(0) K electrons is 11,866.7 eV [44]. This result was consistent with reduction of As(V) to As(0) by acid-treated ZVI. To further evaluate this hypothesis, the first shell of the radial structure function (RSF) from Fourier-transformed EXAFS data was fitted with a theoretical model. Experimental data (solid line) and fitting results (dotted line) are compared in Fig. S2. The amplitude value was set to 0.91 based on the results from the As(V) standard. The data were not phase corrected. The fitting showed that there were 2.3 ± 0.5 oxygen atoms at the As–O first-shell inter-atomic dis˚ The corresponding Debye–Waller parameter ( 2 ) is tance of 1.69 A. −1 ˚ 0.0003 A . Since the As–O coordination number is four for As(V), three for As(III), and zero for As(0) [45], the experimental result of 2.3 oxygen atoms is consistent with a mixture of As(V) and As(0). Thus the EXAFS data, the wet chemical analyses, and the XANES results all support the hypothesis that As(0) was produced due to reaction between As(V) and acid-treated ZVI.
3.4. Reactions of As(V) with acid-treated ZVI Fig. 6 shows the removal of As(V) using acid-treated ZVI at pH 7. Adsorption was very fast (t1/2 <1 h) and there was no measurable dissolved arsenic after ∼ 5 h. The concentration of As(V) decreased with time even though the colorimetric procedures demonstrated that there was no accumulation of As(III). These results led to the hypothesis that As(V) was reduced to As(0), which is a thermodynamically spontaneous reaction as shown in Fig. 7. Reduction of As(V) to As(0) in the presence of acid-treated ZVI was confirmed by XANES. Fig. 8 shows the XANES spectra of the As(V) reaction products using commercial ZVI or acid-treated ZVI. The As(V) white line (first most intense absorption peak) was at 11,872.8 eV and the As(III) white line was at 11,869.5 eV. Both As(V)/ZVI samples had a white line at 11,872 eV, confirming the presence of As(V) species. Reaction between As(V) and commercial ZVI produced an As(III) shoulder at 11,870 eV. Although the XANES results are not quantitative, the spectrum was consistent
Fig. 8. XANES spectra at near-K␣ adsorption edge of As(V) (Na2 HAsO4 ), As(III) (NaAsO2 ), As(V) + commercial ZVI at pH 8.5 ± 0.5, As(V) + acid-treated ZVI at pH 7.
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Fig. 9. Schematic mechanisms of As(V) adsorption/reduction on (a) commercial ZVI, and (b) acid-treated ZVI.
3.5. Mechanisms The experimental findings demonstrated the following differences among magnetite, commercial ZVI, and acid-treated ZVI: (1) stoichiometric magnetite adsorbed As(V) but did not reduce As(V) even with a stoichiometric excess of dissolved Fe(II) although magnetite/Fe(II) is theoretically capable of reducing As(V) to As(III) at pH 6 and to As(0) at pH 8 and 10 as shown in Fig. 7; (2) commercial ZVI adsorbed As(V) and reduced As(V) to As(III) rapidly and without a lag phase although Fig. 7 shows that Fe(0) can reduce As(V) to As(0); and, (3) acid-treated ZVI adsorbed As(V) and partially reduced As(V) to As(0). The physical structures of commercial ZVI and acid-treated ZVI and proposed mechanisms for As(V) removal are illustrated in Fig. 9. Commercial ZVI is covered with a continuous adherent film of magnetite which inhibits direct contact of As(V) with the Fe(0) core. Reducing species such as e− , Fe2+ , H2 , or H• [3] are produced by corrosion of the core and migrate through the magnetite shell to react with As(V). Magnetite is a semiconductor with conductivity 106 higher than other iron oxides [46] and it is possible that there is a potential gradient through the magnetite shell [20,21,47]. Using commercial ZVI, As(V) was partially reduced to As(III) at pH 6.5, reduced to As(III) and perhaps partially to As(0) at pH 8.5. As(V) was not reduced by commercial ZVI at pH 10 because the reaction is not thermodynamically spontaneous at that pH. Acid-treatment of ZVI disrupted the adherent magnetite film, producing a massive and porous oxidized iron shell that could have allowed transport of As(V) to the Fe(0) core where the redox potential was low enough to produce As(0), as shown in Fig. 9. These observations were consistent with a study [48] in which nZVI reduced Ag(I), Cu(II), and Hg(II) to the zero valence but Zn(II) and Cd(II) were adsorbed without being reduced. The first group of metals have higher redox potentials and the second group have lower redox potentials than Fe(II) → Fe(0). The reaction products were consistent with predictions based on direct reaction of the metal cations with the ZVI core. Acid-treated and nZVI have relatively massive and flocculated shells due to rapid precipitation [13] resulting in a porous oxidized shell. The oxidized shell on commercial ZVI is formed during slow corrosion in the presence of air and this produces a thin adherent passivation layer that could inhibit direct contact of contaminants with Fe(0). Both nZVI and acid-treated ZVI used in these studies are produced and stored under anoxic conditions and this could further
inhibit formation of a thin adherent passivation layer. Perhaps acidtreated ZVI can be produced in situ by introducing commercial ZVI followed by an acid flush. The acid-flushing procedure might be analogous to renovation of sampling or other wells during which acids and other agents are used to reductively dissolve oxidized precipitates. This treatment could result in inexpensive, rapid, nolag removal of arsenic and other contaminants from contaminated soils and ground waters. Reduction to As(0) could have significant effects on mobility and toxicity of arsenic in ground water environments. As(V) and As(III) can be released from solids due to change in pH or addition of competing adsorbates [41,49]. Pump and treat remediation efforts can require years to decades for remediation due to rebound effects. To the contrary As(0) does not dissolve and is relatively inert against O2 or even O3 [50]. Partly for these reasons As(0) is less toxic than As(III) or As(V). These properties could make As(0) an attractive reaction product when remediating contaminated water or soil. 4. Conclusions We conducted experiments using strictly anoxic conditions and found that both adsorption and reduction of As(V) were fast, with no lag phase, when reacted with either commercial ZVI or acidtreated ZVI. Previous studies have reported a significant lag phase before commencement of the reduction of As(V) by ZVI. These experiments showed that the Fe(0) core of ZVI is necessary for reduction of As(V), even though magnetite/Fe(II) is theoretically capable of reducing As(V). Further, although commercial ZVI only reduced As(V) to As(III) (with the possibility of slow reduction of As(V) to As(0) at slightly alkaline conditions), disruption of the adherent magnetite shell by acid-treatment resulted in reduction of As(V) to As(0). Thus the physical nature of ZVI can have profound effects on reactivity and acid-treatment of ZVI should be considered for field applications. Reduction to As(0) could result in reduced toxicity and increased stability after remediation of arsenic-contaminated soil and water. In addition As(V) removal capacity on acid-treated ZVI was ∼10 times higher than commercial ZVI of similar size, probably due to both reduction to As(0) and also increased adsorption due to the thicker iron oxide shell. We have drawn comparisons between the reactivity of acidtreated ZVI and nZVI, which was previously observed to reduce some metal cations to their zero valence [48]. Production of nZVI
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also results in the formation of a relatively massive flocculated shell of iron oxides, similar to the shell formed on acid-treated ZVI. Thus it is possible that reduction to As(0) could also occur with nZVI. If acid-treated ZVI has similar reactivity to nZVI and can be produced in situ by acidification of commercial ZVI, then it will be possible to produce a material with reactivity of nZVI but at a much lower cost. Acknowledgement This research was partially supported by the Center for Environmental Kinetics Analysis at Penn State University (NSF Grant No. Che-0431328). The authors thank Trudy Bolin and the U.S. DOE Office of Basic Energy Sciences (Contract No. W-31-109-Eng-38) for XAS support. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.029. References [1] M. Bissen, F.H. Frimmel, Arsenic-a review. Part I: occurence, toxicity, speciation, mobility, Acta Hydrochim. Hydrobiol. 31 (2003) 9–18. [2] S. Stauder, B. Raue, F. Sacher, Thioarsenates in sulfidic waters, Environ. Sci. Technol. 39 (2005) 5933–5939. [3] C. Noubactep, A critical review on the process of contaminant removal in Fe0 –H2 O systems, Environ. Technol. 29 (2008) 909–920. [4] S.H. Joo, A.J. Feitz, T.D. Waite, Oxidative degradation of the carbothioate herbicide, molinate, using nanoscale zero-valent iron, Environ. Sci. Technol. 38 (2004) 2242–2247. [5] X.Q. Li, D.W. Elliott, W.X. Zhang, Zero-valent iron nanoparticles for abatement of environmental pollutants: materials and engineering aspects, Crit. Rev. Solid State Mater. Sci. 31 (2006) 111–122. [6] T.K.K. Ngai, R.R. Shrestha, B. Dangol, M. Maharjan, S.E. Murcott, Design for sustainable development – household drinking water filter for arsenic and pathogen treatment in Nepal, J. Environ. Sci. Health A – Toxic/Hazard. Subst. Environ. Eng. 42 (2007) 1879–1888. [7] I. Diez-Perez, F. Sanz, P. Gorostiza, In situ studies of metal passive films, Curr. Opin. Solid State Mater. Sci. 10 (2006) 144–152. [8] J. Farrell, M. Kason, N. Melitas, T. Li, Investigation of the long-term performance of zero-valent iron for reductive dechlorination of trichloroethylene, Environ. Sci. Technol. 34 (2000) 514–521. [9] B.A. Balko, P.G. Tratnyek, Photoeffects on the reduction of carbon tetrachloride by zero-valent iron, J. Phys. Chem. B 102 (1998) 1459– 1465. [10] B.C. Reinsch, B. Forsberg, R.L. Penn, C.S. Kim, G.V. Lowry, Chemical transformations during aging of zerovalent iron nanoparticles in the presence of common groundwater dissolved constituents, Environ. Sci. Technol. 44 (2010) 3455–3461. [11] P. Rao, M.S.H. Mak, T. Liu, K.C.K. Lai, I.M.C. Lo, Effects of humic acid on arsenic(V) removal by zero-valent iron from groundwater with special references to corrosion products analyses, Chemosphere 75 (2009) 156–162. [12] B.A. Manning, M.L. Hunt, C. Amrhein, J.A. Yarmoff, Arsenic(III) and arsenic(V) reactions with zerovalent iron corrosion products, Environ. Sci. Technol. 36 (2002) 5455–5461. [13] S.R. Kanel, J.-M. Greneche, H. Choi, Arsenic(V) removal from groundwater using nano scale zero-valent iron as a colloidal reactive barrier material, Environ. Sci. Technol. 40 (2006) 2045–2050. [14] S. Bang, G.P. Korfiatis, X. Meng, Removal of arsenic from water by zero-valent iron, J. Hazard. Mater. 121 (2005) 61–67. [15] Z. Ai, Y. Cheng, L. Zhang, J. Qiu, Efficient removal of Cr(VI) from aqueous solution with Fe@Fe2 O3 core–shell nanowires, Environ. Sci. Technol. 42 (2008) 6955–6960. [16] S.R. Kanel, B. Manning, L. Charlet, H. Choi, Removal of arsenic(III) from groundwater by nanoscale zero-valent iron, Environ. Sci. Technol. 39 (2005) 1291–1298. [17] C. Su, R.W. Puls, Arsenate and arsenite removal by zerovalent iron: kinetics, redox transformation, and implications for in situ groundwater remediation, Environ. Sci. Technol. 35 (2001) 1487–1492. [18] J. Farrell, J. Wang, P. O’day, M. Conklin, Electrochemical and spectroscopic study of arsenate removal from water using zero-valent iron media, Environ. Sci. Technol. 35 (2001) 2026–2032.
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Riley, A modified single solution method for the determination of phosphate in natural waters, Anal. Chim. Acta 27 (1962) 31–36. [31] O. Larsen, D. Postma, R. Jakobsen, The reactivity of iron oxides towards reductive dissolution with ascorbic acid in a shallow sandy aquifer – (Romo, Denmark), Geochim. Cosmochim. Acta 70 (2006) 4827–4835. [32] E. Viollier, P.W. Inglett, K. Hunter, A.N. Roychoudhury, P.V. Cappellen, The ferrozine method revisited: Fe(II)/Fe(III) determination in natural waters, Appl. Geochem. 15 (2000) 785–790. [33] D.L.A.d. Faria, S.V. Silva, M.T.d. Oliveira, Raman microspectroscopy of some iron oxides and oxyhydroxides, J. Raman Spectrosc. 28 (1997) 873–878. [34] B. Ravel, M. Newville, Athena, Artemis, Hephaestus: data analysis for Xray absorption spectroscopy using IFEFFIT, J. Synchrotron Radiat. 12 (2005) 537–541. [35] Y.S. Chen, J.L. Fulton, W. Partenheimer, A XANES and EXAFS study of hydration and ion pairing in ambient aqueous MnBr2 solutions, J. 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Journal of Hazardous Materials 196 (2011) 318–326
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Co2+ -exchange mechanism of birnessite and its application for the removal of Pb2+ and As(III) Hui Yin, Fan Liu, Xionghan Feng, Mingming Liu, Wenfeng Tan, Guohong Qiu ∗ Key Laboratory of Subtropical Agricultural Resources and Environment, Ministry of Agriculture, College of Resources and Environment, Huazhong Agricultural University, Wuhan 430070, PR China
a r t i c l e
i n f o
Article history: Received 28 May 2011 Received in revised form 30 August 2011 Accepted 8 September 2011 Available online 14 September 2011 Keywords: Birnessite Ion exchange Octahedron vacancy Cobalt oxidation Lead adsorption Arsenite oxidation
a b s t r a c t Co-containing birnessites were obtained by ion exchange at different initial concentrations of Co2+ . Ion exchange of Co2+ had little effect on birnessite crystal structure and micromorphology, but resulted in an increase in specific surface areas from 19.26 to 33.35 m2 g−1 , and a decrease in both crystallinity and manganese average oxidation state. It was due to that Mn(IV) in the layer structure was reduced to Mn(III) during the oxidation process of Co2+ to Co(III). The hydroxyl groups on the surface of Co-containing birnessites gradually decreased with an increase of Co/Mn molar ratio owing to the occupance of Co(III) into vacancies and the location of large amounts of Co2+/3+ and Mn2+/3+ above/below the vacant sites. This greatly accounted for the monotonous reduction in Pb2+ adsorption capacity, from 2538 mmol kg−1 for the unmodified birnessite to 1500 mmol kg−1 for the Co2+ ion-exchanged birnessite with a Co/Mn molar ratio of 0.16. The amount of As(III) oxidized by birnessite was enhanced after ion exchange, but the apparent initial reaction rate was greatly decreased. The present work demonstrates that Co2+ ion exchange has great influence on the adsorption and oxidation behavior of inorganic toxic metal ions by birnessite in water envrionments. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Birnessite is a kind of ubiquitous hydrous-layered manganese oxide (phyllomanganate) in geological environments [1]. Usually occurring as fine-grained particles, it exhibits high reactivity and is widely involved in a series of geochemical processes, particularly plays a pivotal role in the fate of heavy metal(loids) (Pb2+ , Cd2+ , As(III), Cr(III) and so on) and other pollutants in contaminated soil and water systems, due to mixed valences, large surface area and low point of zero charge [2–9]. Cobalt is one of the trace elements enriched in deep-sea manganese nodules and crusts [10]. It is also found that manganese minerals present in soils contain relatively large amount of Co [11]. Manganese oxides remarkably affect the geochemistry behavior of Co likely due to the electron transfer between adsorbed Co2+ and high-valence Mn [12–17]. A direct evidence for the oxidation of Co2+ to Co3+ on the surface of manganese oxides was first reported by Murray and Dillard [18] using XPS. Previous literatures have suggested that surface-adsorbed Co2+ on birnessite could be oxidized by certain Mn(IV) in the vicinity of vacancies [18,19]. During the reactive transport of Co(II)EDTA2− by pyrolusite (MnO2 ), Mn(IV) was reduced to Mn(III) rather than Mn(II) to form a
∗ Corresponding author. Tel.: +86 27 87280271; fax: +86 27 87280271. E-mail address:
[email protected] (G. Qiu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.027
stable trivalent manganese solid (␣-Mn2 O3 ), which passivated the surface after an initial reaction period and ultimately limited the yield of Co(III)EDTA− [20]. However, X-ray powder diffraction and polarized EXAFS spectroscopy were combined to determine the Co-sorbed birnessite and indicated that Mn3+ is the more likely electron sink for the oxidation of Co2+ [21]. These results indicated that, during the adsorption and oxidation of Co2+ by manganese oxides, the electron sink, whether the Mn(IV) or the Mn(III), is mostly likely dependent on the manganese oxide structures, surface chemical characteristics and reaction conditions. However, the mechanism of Co2+ oxidation by acid birnessite is not clear yet. Furthermore, as for the first transition-metal series, Co(III), Mn(III), and Mn(IV) have similar ionic radius and charges, and all can stably exist in layered structures composed of edge-sharing octahedra [22]. However, the electronegativity of cobalt is different from that of manganese in similar crystallographic structures, and the Co3+ /Co2+ redox conjugate pair has a higher standard reduction potential than those of MnO2 /Mn3+ /Mn2+ [23]. Thus, incorporation of cobalt may have some influence on the crystal structure, morphology, manganese average oxidation state (Mn AOS), and hydroxyl content of birnessite, subsequently on the removal of toxic metal(loids) from wastewater. The adsorption, oxidation, and ion exchange are common characteristics of elemental geochemistry. Our previous work demonstrated that the adsorption performance and removal capacity for Pb2+ and As(III) from aquatic systems by acid birnessite could be improved remarkably
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after doping with cobalt during the synthesis process [24]. In this paper, Co-containing birnessites obtained by ion exchange with Co2+ at different concentrations were synthesized, characterized, and their adsorption and oxidation performance towards Pb2+ and As(III) were investigated, respectively. The ion-exchange process and underlying mechanism was further studied.
corresponding to alkyl type carbon (C–C, C–H), alcohol (C–OH) and/or ester (C–O–C) functionalities, and C O/ O–C O, respectively. All samples have been charge corrected to give the adventitious C (1s) spectral component (C–C, C–H) a binding energy of 284.80 eV. This process has an associated error of ±0.1–0.2 eV [27,28].
2. Materials and methods
2.4. Pb2+ adsorption experiments
2.1. Sample preparation
Pb2+ adsorption experiments were performed and details on the experimental procedure are reported in our previous work [24]. A solid/solution ratio of 1.67 g L−1 , initial Pb2+ concentration ranges from 0 to10 mmol L−1 , and a constant supporting electrolyte concentration (NaNO3 , Ic = 0.1 mol L−1 ) were used. The mixtures were shaken at a rate of 250 r min−1 for 24 h at 25 ◦ C, and the pH of the reaction system was maintained at 5.00 using a pH-stat technique in the process. The metal ions (Pb2+ , Mn2+ , Co2+ , and K+ ) in solution were analyzed by AAS and flame spectrometry.
Acid birnessite was prepared according to McKenzie’s method [25]. Co2+ adsorption was conducted by adding Co(NO3 )2 solution to a birnessite suspension (1.67 g L−1 ) with pH 5 at 25 ◦ C. The pH of the mixtures was maintained by adding 0.1 mol L−1 HNO3 and NaOH solution. The Co concentration in solution was adjusted to obtain Co/Mn molar ratios, such as 0, 0.02, 0.05, 0.10, and 0.20. After allowing 24 h for equilibration, the samples were filtered, rinsed, and dried at 40 ◦ C for several days, and then ground and sieved (100 mesh). The prepared Co-containing birnessite samples were designated as HB, HC2, HC5, HC10, and HC20 according to the initial Co coverages, respectively. 2.2. Sample characterization All samples were characterized by powder XRD, FTIR, SEM, N2 adsorption, and chemical analysis. XRD analyses were carried out on a Bruker D8 Advance diffractometer equipped with a LynxEye detector using Ni-filtered Cu K␣ radiation ( = 0.15418 nm). The diffractometer was operated at a tube voltage of 40 kV and a tube current of 40 mA with 1.2 s counting time per 0.02◦ 2 step. FTIR analyses were conducted on a Bruker Equinox 55 model spectrophotometer using KBr pellets with a spectral range of 4000–400 cm−1 . The final spectrum was the average of 64 scans at a nominal resolution of 4 cm−1 . The crystallite morphologies of the samples were probed by scanning electron microscopy using a JSM-6390LV microscope after being coated with a gold evaporated film. The specific surface area was obtained by nitrogen adsorption using an Autosorb-1 standard physical adsorption analyzer (Quantachrome Autosorb-1). The samples were degassed at 110 ◦ C for 3 h under vacuum prior to adsorption measurement. The chemical compositions of the samples were determined using atomic absorption spectrometry (AAS, Varian AAS 240FS) and flame spectrometry (Sherwood Model 410). 0.1000 g of sample was dissolved in 25 mL solution of 0.25 mol L−1 NH2 OH·HCl and 1 mol L−1 H2 SO4 . The Mn AOS was obtained by a titration method: a mass of 0.2000 g sample was completely reduced to Mn2+ in 5 mL of 0.5000 mol L−1 H2 C2 O4 and 10 mL of 1 mol L−1 H2 SO4 . Excess C2 O4 2− was determined by back-titration using a KMnO4 standard solution at 75 ◦ C [26].
2.5. As(III) oxidation experiments The procedure of As(III) oxidation was also the same as that documented in reference [24]. A solid/solution ratio of 0.5 g L−1 , 0.08 mmol L−1 of initial As(III) concentration, and a constant supporting electrolyte concentration (NaNO3 , Ic = 0.1 mol L−1 ) were used. The reaction was carried out at pH 7 with stirring at 25 ◦ C. As(V) was measured using the colorimetric method [29]. Release amount of Co2+ , Mn2+ and K+ was monitored by AAS and flame spectrometry. 3. Results and discussion 3.1. Characterization 3.1.1. Crystal structure of the Co-containing birnessites Fig. 1 shows the powder XRD patterns of the obtained samples. The patterns could be indexed corresponding to Hexagonal, R-3m (JCPDS 86-0666). Birnessite was characterized by four detectable peaks: 0.723, 0.361, 0.244, and 0.142 nm. Peaks at 0.723 nm and 0.361 nm were both symmetrical, belonging to (0 0 l) reflections, whereas the other two at higher angles were broad. The 0.244 nm peak is convoluted by (1 0 1) and (0 1 2) reflections while the 0.142nm one is convoluted by (1 1 0), (1 1 3) and (10, 13) reflections (JCPDS 86-0666). The intensity of the 0.361-nm peak was higher than that of the 0.244-nm peak, which suggested that the diffracting crystallites contained more layers stacked coherently along the c axis than that of “acid birnessite” reported by Villalobos et al. [30,31].
2.3. XPS analysis X-ray photoelectron spectra were collected using a VG Multilab2000 X-ray photoelectron spectrometer with an Al K␣ X-ray source (1486 eV) and a base pressure of 3 × 10−9 Torr in the analytical chamber. The scans were recorded using the large area mode. The survey scans were collected using a fixed pass energy of 100 eV and an energy step size of 1.0 eV, whereas the narrow scans had a pass energy of 25 eV and an energy step size of 0.1 eV. The spectra were analyzed using the Avantage software. The Shirley-type background was subtracted before deconvolution and fitting. A ratio of 30:70 of the Lorentzian: Gaussian mix-sum function was used for all the fittings. A broad asymmetric peak was observed for C (1s) (Fig. S1). Three peaks were used to fit C (1s) spectra (Table S1 and Fig. S1)
Fig. 1. Powder XRD patterns of birnessite and Co-containing birnessites.
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Table 1 Physicochemical properties of Co-containing birnessites. Sample
HB HC2 HC5 HC10 HC20
Element content (%) Mn
Co
K
52.74 50.43 49.57 48.25 46.96
0 1.16 2.77 5.90 7.88
8.36 6.65 5.61 4.39 3.00
Co/Mn ratio
Mn AOS
SSA (m2 g−1 )
0 0.02 0.05 0.11 0.16
3.78 3.68 3.65 3.59 3.56
19.26 19.14 18.43 28.70 33.35
Co-containing birnessites exhibited the same XRD characteristics as the original sample. There were four peaks with similar symmetry and relative intensity. As the initial molar ratios of Co/Mn increased, no extra peak was detected, suggesting that the import of cobalt ions into the lamellar birnessite did not affect the crystal structure, and no second phase was introduced. However, the peak intensity of Co-containing birnessites slightly decreased, and their FWHM slightly increased, indicating a decrease in particle size and crystallinity. This was further confirmed by the monotonous decrease in the relative intensity of the 0.361-nm peak to the 0.244nm peak. 3.1.2. Elemental analyses and Mn AOS Table 1 shows the content of Co, Mn, and K in birnessite before and after adsorption reaction. The initial Co/Mn molar ratio was controlled as 0.02, 0.05, 0.10, and 0.20 in the adsorption experiment, and it was changed to be 0.02, 0.05, 0.11, and 0.16 in birnessite, respectively. When the initial concentration was low, all Co2+ ions were adsorbed. However, when the initial molar ratio of Co/Mn increased to 0.20, only 80% Co2+ ions could be retained. The final molar ratio of Co/Mn was lower than that obtained by ion exchange of HBir with Co2+ by Manceau et al. [21]. This was ascribed to that vacant sites (16.7%) in HBir were higher than those in acid birnessite (12%) [31]. The content of potassium was reduced by 20.5%, 32.9%, 47.5%, and 64.1%, respectively, compared to HB, due to the replacement of quite a relatively large amount of K+ by Co2+ . The Mn AOS of Co-free birnessite was 3.78, confirming the mixed valence states of Mn in the form of Mn3+ and Mn4+ with Mn4+ being the predominant species. The Mn AOS gradually decreased with an increase in cobalt content (Table 1). 3.1.3. FTIR analysis Fig. 2 presents the FTIR spectra of Co-containing birnessites. Five well-resolved peaks were observed at 3412, 1619, 920, 506 and 436 cm−1 . The first two bands were ascribed to the stretching and bending vibrations of molecular H2 O [32]. Vibrations of Mn–O bond in the structure of birnessite were reflected at 506 and 436 cm−1 , which resembled the natural analogs [33,34]. Due to
Fig. 2. FTIR spectra of Co-containing birnessites: (a) HB, (b) HC2, (c) HC5, (d) HC10, and (e) HC20.
that the absorption band at 423 cm−1 could indicate the crystalline order [34], the decrease of the adsorption band intensity suggested the lower degree of crystallinity for Co-containing composites. It was noteworthy that the intensity of the adsorption peak at 920 cm−1 gradually decreased with an increase in Co content. As [MnO6 ] octahedral layer in birnessite containing vacancies was similar to that of [AlO6 ] layer in the structure of dioctahedral clay minerals. As for the latter, 2/3 octahedra were occupied by cations. These cations were coordinated to OH groups, which were parallel to the (0 0 1) planes and pointing to vacant sites. If two Al3+ cations were located there, vibration frequencies of Al–Al–OH were 911–917 cm−1 [35]. For the similar electronegativity of Mn3+ (HS, = 1.675), Mn4+ (HS, = 1.923) in the [MnO6 ] octahedron of birnessite with Al3+ ( = 1.515), and coordination radius of Mn3+ ˚ Mn4+ (r = 0.670 A) ˚ with Al3+ (r = 0.675 A), ˚ it could be (r = 0.785 A), inferred that the vibration of Mn–Mn–OH in the vicinity of vacancy was close to 911–917 cm−1 [36–38]. Therefore, the decrease in the peak intensity at 920 cm−1 was possibly due to that the content of hydroxyl groups around vacancies gradually decreased after Co2+ ion exchange reaction. 3.1.4. Micromorphology Fig. 3 shows the SEM images of birnessite formed by three-dimensional (3D) hierarchical microspheres consisting of two-dimensional (2D) nanoplates. The radius of the 3D microspheres was approximately 200 nm in diameter. Cobalt induction had no obvious effect on the micromorphologies of the products but slightly reduced the particle dimensions. 3.1.5. Specific surface area The surface areas of the samples are listed in Table 1. The specific surface area of HB was 19.26 m2 g−1 , which was similar to that reported by Mckenzie [13]. After ion-exchange reaction, the surface area values of HC2, HC5, HC10, and HC20 were determined to be 19.14, 18.43, 28.70, and 33.35 m2 g−1 , respectively. As indicated by XRD patterns, FTIR, and SEM images, the crystallinity of Co-containing samples gradually decreased. The specific surface area increased with a decrease in crystallinity of birnessite. 3.2. Surface analysis using XPS X-ray photoelectron spectroscopy is recently being used to investigate the abundance and chemical state of elements in the uppermost few atomic layers of solid surfaces. Fig. S2 shows the broad scans for all samples. The peak at a BE of 780 eV was the photoelectron line of cobalt, and its intensity gradually increased with an increase in Co content. The photoelectron line of K (2p) was at the left of C (1s). Its relative intensity greatly decreased for Co-containing birnessites compared to HB, suggesting a lower abundance of potassium in samples after Co2+ ion exchange, which coincided with the results of the chemical analysis. In order to determine the chemical states of Mn, Co and O on the surface of Co-containing samples, narrow scans were also performed. To obtain the relative quantity of Mn2+ , Mn3+ , and Mn4+ on birnessite surface, the Mn (2p3/2 ) spectrum was fitted [28,39]. Table S2, Fig. 4 and Table 2 show the parameters used and the fitting results, respectively. The content of Mn4+ gradually decreased from 79.83% in HB to 66.90% in HC20 while Mn3+ increased from 13.72% to 23.70%, and Mn2+ content slightly increased. According to the fitting results, the calculated Mn AOS were 3.73, 3.71, 3.68, 3.62, and 3.58 for HB, HC2, HC5, HC10 and HC20, respectively. These results agreed well with the titration data (Table 1). Narrow scans for Co (2p) are plotted in Fig. 5. The measured BE values for Co (2p1/2 ) and Co (2p3/2 ) were 795 and 780 eV, respectively, similar to those of Co(III)OOH reported by Crowther et al. [19]. Furthermore, the Co(2p1/2 )–Co(2p3/2 ) splittings were identical
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Fig. 3. SEM of birnessites before and after ion exchange with Co2+ at different concentrations: (a) HB, (b) HC2, (c) HC5, and (d) HC20.
at 15.0 eV(±0.20 eV) which was a characteristic of the spectrum of Co(III) compounds [19]. These results suggested that Co(III) mainly existed in the samples. After being adsorbed on the surface of birnessite, Co2+ was oxidized to Co(III) by high valence Mn in the structure. As illustrated by elemental analysis and multi-component fitting of Mn (2p3/2 ) spectra of Co2+ ion-exchanging hexagonal birnessites, the Mn AOS are generally lower than +4, indicating the coexistence of Mn4+ , Mn3+ , and Mn2+ in the structure. Both Mn4+ and Mn3+ can accept electron from Co2+ , but acid birnessite with hexagonal layer symmetry has the most intricate and polytropic structure characteristics. It consists of edge-sharing [Mn(IV)O6 ] octahedra, [Mn(III)O6 ] octahedra, and vacancies. Large quantities of cations, such as H+ , K+ , and Mn2+/3+ , exist in the interlayer space of adjacent layers to balance the high negative permanent structural charge originated from the substitution of Mn(IV) by Mn(III) and vacancies within the octahedral sheet [31,40–42]. It is difficult to investigate the oxidation mechanism of Co2+ due to the complex structure of birnessite. The element composition in HB changed after ion exchange with Co2+ . Variations () in the quantity of Mn(IV), Mn(III), Mn2+ , and Co(III) were calculated and summarized in Table 3 based on the results of chemical composition (Table 1) and XPS (Table 2). As for HC20, the content of Mn(IV) was reduced by 2353 mmol kg−1
Table 2 Near-surface compositions of Mn and O species derived from fittings of Mn (2p3/2 ) and O (1s) spectra. Sample
HB HC2 HC5 HC10 HC20
Mn (at.%)
O (at.%)
Mn, whereas Mn(III) and Mn(II) were increased by 1817 and 535 mmol kg−1 , respectively. The amount of Co(III) imported was 2847 mmol kg−1 . Given that Mn2+ was adsorbed almost totally on the surface of birnessite as Mn2+ was not detected during the ion exchange process, (Mn2+ ) should be equal to (Co3+ ) if the layer and/or interlayer Mn(III) was the oxidant as proposed by Manceau et al. [21]. However, (Mn2+ ) was only about 1/5 of (Co3+ ). HC2, HC5 and HC10 all generally abided to this rule. In the present work, the Mn(IV) was most likely the electron acceptor. Two possible pathways for the Co2+ oxidation by Mn(IV) were proposed in this study. In the first pathway, two moles of Co2+ reacted with one mole of Mn(IV), producing one mole of Mn2+ and two moles of Co(III) as Eq. (1): Mn(IV) + 2Co2+ → Mn2+ + 2Co(III)
(1)
As for the above equation, (Mn2+ ) = −(Mn4+ ) = (1/2)(Co3+ ), and (Mn3+ ) should be no more than zero (for disproportionation). The other pathway involved one mole of Mn(IV) participating in the oxidation of one mole of Co2+ to Co(III), and Mn(IV) was reduced to Mn(III): Mn(IV) + Co2+ → Mn(III) + Co(III)
(2)
Table 3 Variances in the content of Mn(IV), Mn(III), Mn2+ and Co(III) in Co-containing birnessites.
Mn4+
Mn3+
Mn2+
O2-
OH-
H2 O
Sample
(Mn4+ )a
(Mn3+ )
(Mn2+ )
(Co3+ )
79.83 77.87 75.71 70.76 66.90
13.72 15.32 16.90 20.76 23.70
6.46 6.81 7.39 8.48 9.40
54.78 60.26 56.98 59.02 57.74
22.45 22.03 21.42 18.88 17.00
22.77 17.71 21.60 22.10 25.26
HC2 HC5 HC10 HC20
−357 −750 −1651 −2353
291 579 1281 1817
64 169 368 535
390 948 2075 2847
a
“−” means that the content of Mn4+ decreased.
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Fig. 4. Mn (2p3/2 ) spectra of Co-containing birnessites: (a) HB, (b) HC2, (c) HC5, (d) HC10, and (e) HC20 (The upper circles represent the observed data, and the thick, solid curve is the best fit of the data. The dash-dot curves represent Mn(IV) multiplet peaks, while the thin, solid curves are Mn(III) and the dotted lines are Mn(II).).
If Eq. (2) occurred, (Mn3+ ) = −(Mn4+ ) = (Co3+ ) should be correct. However, Mn(III) are unstable and disproportionate as follows [43]: 2Mn(III) → Mn(IV) + Mn2+
(3)
Therefore, the change trend of various species should increase in the order: (Mn2+ ) < (Mn3+ ) < −(Mn4+ ) < (Co3+ ), in agreement with the results as listed in Table 3. Consequently, Mn(III) was the primary reduction product of Mn(IV).
The O (1s) spectra of as-prepared birnessites had main characteristic peaks at 529.94, 529.93, 529.88, 529.61 and 529.75 eV. The electronegativity () values of Mn3+ (HS, = 1.675) and Co3+ (LS, = 1.791) in the [Mn(Co)O6 ] octahedron of birnessites were lower than that of Mn4+ (HS, = 1.923) [36,37]. Larger amounts of Mn3+ and Co3+ in the layer led to an increase in the ionic bond character of the Mn(Co)–O bond. Therefore, the electron cloud density of cores of O2− and OH− increased, and their BEs shifted to the lower.
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Fig. 6. Isothermal curves of Pb2+ uptake by Co-containing birnessites. Fig. 5. Co (2p) spectra of Co-containing samples.
Lattice oxygen (O2− ), hydroxide oxygen (OH− ), and oxygen in molecular water (physisorbed, chemisorbed, and structural H2 O and water in poor electrical contact with the mineral surface) [44] were the three oxygen species in birnessite materials corresponding to the main peak, the broad shoulder, and the pronounced tail, respectively. The peak parameters used in fitting O (1s) spectra are listed in Table S3 and the corresponding results are presented in Fig. S3 and Table 2, respectively. The content of hydroxide group gradually decreased with an increase of cobalt content, which was consistent with the FTIR results. 3.3. Adsorption of Pb2+ by Co-containing birnessites Lead is one of the widespread contaminant in water systems. Its concentration in many industrial water bodies was reported to be as high as 200–500 mg L−1 . This concentration is very high in relation to the current water quality standard of 0.05–0.1 mg L−1 [45,46]. Consequently, it is essential to reduce the lead levels in wastewater before it was discharged. In order to investigate the effect of Co2+ ion exchange on the Pb2+ adsorption behavior by birnessite, as-prepared composites were used for removal of Pb2+ in laboratory artificial waste water at the concentration level of 0–10 mmol L−1 at 25 ◦ C. The adsorption of Pb2+ on birnessites conformed to the L-type isotherm [47]. The removal efficiency of Pb2+ increased sharply when the equilibrium concentration of Pb2+ was increased from the lowest value, indicating a high affinity isotherm with strong preference for solid phase partitioning in these systems. Then the removal efficiency increased slightly, and last tended to remain stable and approached a maximum as the equilibrium Pb2+ concentration increased (Fig. 6). The Langmuir equation was shown as Eq. (4): Q =
Amax KC 1 + KC
(4)
where Q is the amount of Pb2+ adsorbed (mmol kg−1 mineral), Amax the maximum adsorption capacity (mmol kg−1 ), C the equilibrium concentration of the adsorbate (mmol L−1 ), and K a constant related
to the adsorption energy as function of temperature and adsorption enthalpy [48]. The adsorption capacities are listed in Table 4. HB had a capacity of 2538 mmol Pb2+ per kilogram birnessite. Co2+ ion exchange lessened the removal of Pb2+ from solution by birnessite. The maximum capacity of Pb2+ removed by HC2, HC5, HC10 and HC20 were 2352, 2045, 1879 and 1500 mmol kg−1 , respectively. These as-obtained birnessites exhibited more effective adsorbability for Pb2+ in aqueous solution than other reports, such as single manganese oxides [5], and manganese oxide coated sand, zeolite and bentonite [3,49,50]. Birnessite has the highest affinity for Pb2+ among various heavy metal ions [5]. Pb(II) forms strong inner-sphere surface complexes mainly at two sites on hexagonal birnessite nanoparticles: triple corner-sharing (TCS) surface complexes and triple edge-sharing (TES) surface complexes on vacancies, and double corner-sharing (DCS) and double edge-sharing (DES) complexes on lateral edge surfaces [51]. Regardless of whatever Pb(II) complexes formed, the content of vacancies in birnessite structure positively determined the adsorption capacity for Pb2+ [52]. But some other literatures reported that, the external edge surface played a key role in Pb2+ adsorption. A positive linear relationship between BET surface area and maximum Pb2+ capacity had been established [53]. Nevertheless, the surface areas of birnessites treated with different concentrations of Mn2+ greatly increased from 9.84 m2 g−1 to 67.0 m2 g−1 , but their Pb2+ adsorption capacity gradually decreased [52]. Therefore, the total Pb2+ adsorption capacity was determined not only by the content of vacant sites and the surface areas, but also by their relative abundance. In the Pb(II) uptake experiments of biogenic manganese oxides, the surface areas of the minerals were much high for their finite crystal size, and the contribution for Pb uptake by external sites was higher than the internal sites. However, as for the Co-containing birnessites here, although the surface areas of Co-containing composites was increased from 19.26 to 33.35 m2 g−1 with an increase in cobalt content (Table 1), Pb2+ adsorption capacity gradually decreased (Fig. 6; Table 4). This decrease was probably attributed to the decrease of the content of vacant sites. As demonstrated by Manceau et al. [21], when Co2+ was adsorbed on vacancies and then oxidized to Co3+ , most of the Co3+ directly incorporated into the vacant sites, and
Table 4 Langmuir parameters for the adsorption of Pb2+ and maximum concentrations of Mn2+ , Co2+ , H+ , K+ released. Sample
Ions released (mmol kg−1 )
Parameters −1
Amax (mmol kg HB HC2 HC5 HC10 HC20
2538 2352 2045 1879 1500
)
2
K
R
Co2+
Mn2+
H+
K+
17.7 4484 1222 165.3 349.6
0.998 0.999 0.995 0.987 0.998
0 18.9 34.6 78.9 174.9
0 0.5 0.5 9.1 34.7
2778 2629 2297 2304 2199
1283 978 730 546 372
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Fig. 7. Speciation of lead(II) as a function of pH. [Pb] = 10 mmol L−1 in 0.1 mol L−1 NaNO3 (calculated using ECOSAT4.9 [54]).
Fig. 8. Kinetics curves of As(III) oxidation by Co-containing birnessites.
Pb(II) adsorbed and the release of ions as follows: partial Co3+/2+ was located above or below vacancies, resulting in the decrease of the number of vacant sites and the content of hydroxyl groups. The decrease of vacant sites loudly accounted for the decrease of Pb2+ adsorption capacity. Although the location of Co3+/2+ on the vacancies did not affect the amount of vacant sites, but these cations would compete with Pb(II) species for binding sites. Therefore, the maximum lead adsorption capacity of Co2+ exchange birnessites was greatly reduced. This was different from the case of Pb2+ adsorption by Co-doped birnessites. The enhanced Pb2+ removal of Co-doped birnessites compared with the undoped one was attributed to the gain of negative charge of the octahedral layer by the substitution of layer Mn(IV) by Co(III) and the induction of vacancies in the structure by the insertion of Co(III) because of the heterogeneity between the exotic ion and Mn4+/3+ [24]. During Pb2+ adsorption, Mn2+ , Co2+ , H+ , and K+ were simultaneously released into the solution in the order: H+ > K+ > Co2+ > Mn2+ (Table 4). This implied that the negative charge of birnessite layer was mainly balanced by H+ and K+ . Mn2+ was not detected during Pb2+ adsorption by HB, indicating the presence of small amounts of Mn2+/3+ in the interlayer, which was consistent with the XPS results. The release amount of H+ and K+ was positively related to the Pb2+ adsorbing capacity, however, that of Co2+ and Mn2+ increased with the increase of cobalt content. With an increase in cobalt content, more Mn2+ ions were released into the solution. These low valence Mn ions were formed from the reduction of Mn3+/4+ by Co2+ . About 9.6%, 7.4%, 7.9% and 13.1% of the total cobalt in HC2, HC5, HC10 and HC20 was released, respectively. It indicated that there might be some Co2+ , existing in the interlayer or on the surface of birnessite besides Co3+ when the initial Co2+ concentration was high [21]. The Co2+ can be replaced by Pb2+ during the adsorption. Furthermore, the Co3+ in the interlayer might also be driven by Pb2+ into the solution coupled with redox reactions, but the underlying mechanisms are not clear yet. The aqueous speciation of Pb2+ is shown in Fig. 7 as a function of pH in the range of 3–11 (calculated using ECOSAT4.9 [54]). When pH was below 5.4, Pb2+ and [Pb(NO3 )]+ were the main species. As it increased to 5.4, the formation of Pb(OH)2 (s) began to limit the concentrations of aqueous species. At ∼pH 5.5, Pb(OH)2 (s) was the major form present in the system. When pH was above ∼6.5, Pb(II) occurred predominantly as Pb(OH)2 (s). At pH 5.0, the species of Pb(II) in the adsorption systems existed as Pb2+ (61.21%), [Pb(NO3 )]+ (34.84%), and Pb(NO3 )2 (3.95%). According to the charge conservation law in the process of ion exchange, two moles of H+ and/or K+ or a mole of divalent cation (Me2+ ) would be released after the adsorption of one mole of Pb2+ whereas adsorption of [Pb(NO3 )]+ would drive only one mole of H+ or K+ or 1/2 mol of Me2+ away from the surface of birnessite. This was confirmed by the algebraic relationship of the amounts of
n(Pb(II)) × (0.6121 × 2 + 0.3484) ≈ n(H+ ) + n(K+ ) + (n(Mn2+) + n(Co2+ )) × 2 where n denotes the amount of adsorbed/released ions, and was listed in Table 4. 3.4. Effects of Co2+ -exchange on the oxidation of As(III) Arsenic contamination is an issue of great concern. Depending on its source, arsenic concentrations in natural waters may range up to several hundred milligrams per liter. Due to its acute toxicity to humans, a maximum contaminant level (MCL) should be less than 10 g L−1 for arsenic in drinking water [55]. Manganese oxides are reactive oxidants for the transformation of As(III) to As(V) under natural conditions. As-prepared birnessites were used to study the oxidative transformation of sodium arsenite at the interface of minerals and water. Arsenite oxidation first occurred quickly and the reaction rate then decreased to keep a balance after 1–2 h. HC2 and HC5 had the same shape as that of HB. However, in the case of HC10 and HC20, the oxidation amount of As(III) gradually increased as the reaction progressed (Fig. 8). An apparent oxidation capacity of As(III) to As(V) was calculated to evaluate the oxidation ability of birnessite, due to the fact that adsorption and fixation of As(III) and As(V) occurred simultaneously during the oxidation process [8]. The calculated apparent oxidation capacity of As(III) by HB is 77.3% at equilibrium. After reaction for 7 h, the conversion of As(III) to As(V) by HC2, HC5, HC10 and HC20, were 91.0%, 93.2%, 88.4% and 60.2%, respectively. High oxidation ability towards As(III) was ascribed to the participation of Co(III) in the reaction, since the standard reduction potential for Co3+ /Co2+ (E◦ = 1.92 V) is higher than the MnO2 /Mn2+ (E◦ = 1.224 V) and Mn3+ /Mn2+ (E◦ = 1.5415 V) half reactions [23]. This was already confirmed in our previous work using XPS analysis [24]. The relationship of As(III) concentration with time was analyzed by fitting a first-order rate equation to the 0–0.33 h portions of all the five systems (Fig. 9). The apparent reaction rate constants (kobs ) of HB, HC2, HC5, HC10, and HC20 were calculated to be 0.0226, 0.0175, 0.0161, 0.0123, and 0.0035 min−1 , respectively. The higher initial reaction rate of As(III) oxidation for HB than for Co-containing ones can be ascribed to several reasons. Firstly, in Cocontaining birnessites, oxygen atoms bound to Co3+ will be more strongly held than those bound to Mn3+/4+ due to the high crystal field stabilization energy (CFSE) of the low-spin Co3+ ion [15]. This would increase the activation energy at these sites, resulting in a slower reaction rate [56]. Secondly, the As(III) oxidation by birnessites is a complex process. Investigation on the arsenite oxidation by a poorly crystalline manganese-oxide exhibited that As(III) oxidation and As(V) sorption is greatly affected by Mn AOS in
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Fig. 9. Linear regression analysis of normalized As(III) uptake by as-obtained birnessite.
the ␦-MnO2 structure [57]. The Mn(III) reactive sites on Mn-oxide surfaces were expected to be less reactive than Mn(IV) reactive sites in terms of As(III) oxidation [39,58,59]. Moreover, analysis of XPS and the release of ions during Pb2+ adsorption revealed that certain amounts of Mn2+/3+ and Co2+/3+ were adsorbed on the surface of birnessites. These low valence ions also blocked reactive sites on the mineral surface. Release of Mn2+ , Co2+ and K+ into the solution was also monitored. The concentrations of K+ released by HB, HC2, HC5, HC10 and HC20 during the arsenite oxidation process were 1998, 1523, 1247, 1068 and 669 mmol kg−1 , respectively. However, Mn2+ and Co2+ were not detected when HB, HC2, and HC5 were used as oxidants. For HC10 and HC20, 13 and 0, 36 and 8 mmol kg−1 of Mn2+ and Co2+ were determined, respectively. There were two reasons for this low Mn2+ /Co2+ release: (i) MnO2 particle surface contained negatively charged surface functional groups (≡Mn–O− ), thus soluble Mn2+/3+ and Co2+/3+ formed by reductive dissolution of Co-containing birnessites were adsorbed on the surface at pH 7.00 in the present study, and (ii) the precipitate of Co3 (AsO4 )2 (pKsp = 28.17) [60] and the possible Krautite was formed [61]. The oxidation of As(III) by manganese oxide was an important reaction in both the natural cycling of As and in developing remediation technology for lowering the As(III) concentration in drinking water. In the presence of Co-containing birnessite, As(III) in wastewater or underground water would be oxidized to As(V). As(III) has higher mobility and weaker adsorption, and thereby is more poisonous than As(V) [8]. Because As(V) exists as deprotonated oxyanions in broad pH ranges [23] and has high affinity of mineral surfaces, oxidation of As(III) to As(V) not only reduces its toxicity, but also facilitates the removal of As species. Metal compounds (Fe/Al oxides, hydroxides, etc.) are the most widely used adsorbents for As, for their higher removal efficiency at lower cost versus many other adsorbents [4]. The maximum amount of As(V) produced during the oxidation of As(III) by birnessite was greatly enhanced in the presence of goethite. The combined effects of the oxidation (by birnessite)-adsorption (by goethite) led to rapid oxidation and immobilization of As and alleviation of the As toxicity in the environments [62]. Hence, the powerful oxidation of arsenite by manganese oxides, followed by adsorption of arsenate by Fe/Al compounds as adsorbents, is an applicable approach for the treatment of As(III) contaminated water systems, and is worth further investigation. 4. Conclusions Co2+ ion exchange with birnessite was conducted at different Co2+ concentrations. Co2+ ions were totally retained by birnessite at low concentrations. However, when initial Co/Mn molar ratio was increased to 0.2, only 80% of Co2+ could be located in the
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structure. Induction of Co2+ had no effect on the crystal structure and morphology of birnessite. The crystallinity of Cocontaining birnessites gradually decreased and specific surface areas increased. The valence of Co was exclusively +3 and Mn AOS of Co-containing birnessites gradually decreased. The content of hydroxyl groups in the structure of Co-containing birnessites gradually decreased, which accounted for the reduced Pb2+ adsorption capacities. As(III) oxidation by Co-containing birnessite was enhanced. Conversely, with an increase in cobalt concentration, the initial reaction rate constant was greatly reduced. During the process of Co2+ oxidation, Mn(IV) was more likely the electron sink, and subsequently reduced to Mn(III). The present work provides a new insight into the environmental chemical behavior and interaction mechanism of cobalt and manganese oxides. Further, these modified materials have higher adsorption capacity for Pb2+ than many other adsorbents. Simultaneously, their enhanced oxidation ability for As(III) to As(V) can greatly reduce the toxicity of As(III) in the environment. These as-obtained birnessites have great potential applications in the remediation of heavy metal-contaminated soil and water. Acknowledgements The authors gratefully thank the National Natural Science Foundation of China (Grant numbers: 40830527, 41171375) and the Fundamental Research Funds for the Central Universities (Program number: 2011PY015) for financial support. The authors also acknowledge research assistant Homer Genuino at University of Connecticut for improving English writing in the paper. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.027. References [1] Y.N. Vodyanitskii, Mineralogy and geochemistry of manganese: a review of publications, Eurasian Soil Sci. 42 (2009) 1170–1178. [2] F. Liu, C. Colombo, P. Adamo, J.Z. He, A. Violante, Trace elements in manganese–iron nodules from a Chinese Alfisol, Soil Sci. Soc. Am. J. 66 (2002) 661–670. [3] R.P. Han, W.H. Zou, Z.P. Zhang, J. Shi, J.J. Yang, Removal of copper(II) and lead(II) from aqueous solution by manganese oxide coated sand. I. Characterization and kinetic study, J. Hazard. Mater. B137 (2006) 384–395. [4] D. Mohan, C.U. Pittman Jr., Arsenic removal from water/wastewater using adsorbents—A critical review, J. Hazard. Mater. 142 (2007) 1–53. [5] X.H. Feng, L.M. Zhai, W.F. Tan, F. Liu, J.Z. He, Adsorption and redox reactions of heavy metals on synthesized Mn oxide minerals, Environ. Pollut. 147 (2007) 366–373. [6] R.N. Dai, J. Liu, C.Y. Yu, R. Sun, Y.Q. Lan, J.D. Mao, A comparative study of oxidation of Cr(III) in aqueous ions, complex ions and insoluble compounds by manganese-bearing mineral (birnessite), Chemosphere 76 (2009) 536–541. [7] Y.T Meng, Y.M. Zheng, L.M. Zhang, J.Z. He, Biogenic Mn oxides for effective adsorption of Cd from aquatic environment, Environ. Pollut. 157 (2009) 2577–2583. [8] X.J. Li, C.S. Liu, F.B. Li, Y.T. Li, L.J. Zhang, C.P. Liu, Y.Z. Zhou, The oxidative transformation of sodium arsenite at the interface of ␦-MnO2 and water, J. Hazard. Mater. 173 (2010) 675–681. [9] S.B. Lee, J.S. An, Y.J. Kim, K. Nam, Binding strength-associated toxicity reduction by birnessite and hydroxyapatite in Pb and Cd contaminated sediments, J. Hazard. Mater. 186 (2011) 2117–2122. [10] R.G. Burns, V.M. Burns, The mineralogy and crystal chemistry of deep-sea manganese nodules—a polymetallic resource of the twenty-first century, Philos. Trans. R. Soc. London Ser. A286 (1977) 283–301. [11] R.M. Taylor, R.M. Mckenzie, The association of trace elements with manganese minerals in Australian soils, Aust. J. Soil Res. 4 (1966) 29–39. [12] R.M. Mckenzie, The reaction of cobalt with manganese dioxide minerals, Aust. J. Soil Res. 8 (1970) 97–106. [13] R.M. Mckenzie, The adsorption of lead and other heavy metals on oxides of manganese and iron, Aust. J. Soil Res. 18 (1980) 61–73. [14] R.G. Burns, The uptake of cobalt into ferromanganese nodules, soils, and synthetic manganese(IV) oxides, Geochim, Cosmochim. Acta 40 (1976) 95–102.
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Journal of Hazardous Materials 196 (2011) 327–334
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Adsorption behavior of some aromatic compounds on hydrophobic magnetite for magnetic separation Takahiro Sasaki ∗ , Shunitz Tanaka Division of Environmental Material Science, Graduate School of Environmental Science, Hokkaido University, Sapporo, Hokkaido, 060-0810, Japan
a r t i c l e
i n f o
Article history: Received 4 March 2011 Received in revised form 8 September 2011 Accepted 9 September 2011 Available online 16 September 2011 Keywords: Adsorption behavior Aromatic compounds Hydrophobic magnetite Hydrophobic interaction -electron interaction
a b s t r a c t In this study, a hydrophobic magnetite coated with an alkyl chain or a phenyl group on the surface was prepared and used as an adsorbent to investigate the adsorption behavior of aromatic compounds having various values of log Pow (phenol 1.46, benzonitrile 1.56, nitrobenzene 1.86, benzene 2.13, toluene 2.73, chlorobenzene 2.84 and o-dichlorobenzene 3.38) onto hydrophobic magnetite. The hydrophobic magnetites were modified with stearic acid and phenyltrimethoxysilane, and the modification amounts were 9.84 × 10−3 and 4.17 × 10−2 mmol/g, respectively. The aromatic compounds used in this study were divided into 3 groups depending on the log Pow : 1 < log Pow < 2, 2 < log Pow < 3 and 3 < log Pow . The adsorption amounts of above each group on the magnetite at an initial concentration of 100 ppm were 3.62 × 10−3 (nitrobenzene), 1.92 × 10−2 (phenol), 1.13 × 10−1 (chlorobenzene), 2.42 × 10−1 (benzene), and 3.10 × 10−1 mmol/g (dichlorobenzene), respectively. This indicates that the adsorption behaviors depend on the strength of hydrophobicity of aromatic compounds. The adsorption mechanism for 2 < log Pow < 3 and 3 < log Pow is hydrophobic interaction and that for 1 < log Pow < 2 is -electron interaction. The quantitative relationship between the amount of adsorbed compounds and modified functional groups and the fitting for adsorption isotherm models suggested that this adsorption might form a multi-layer adsorption in the most cases. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Global attention to water pollution from harmful chemicals has increased in recent decades. Many kinds of pollutants have been discovered in aquatic environments such as rivers, ponds and seas. These contaminants originated from industrial and domestic wastewater, sometimes from accidental spills. Various mono and polycyclic aromatic compounds have been found in aquatic environments. These aromatic compounds must be removed before the water is discharged or consumed. In order to treat harmful organic compounds in effluent, two types of technology are currently available. The first is a decomposition technology where hazardous organic compounds are converted to more environmentally friendly compounds. Technologies such as chemical oxidation [1], electrolysis [2], photo oxidation [3], and ozonation [4] are included in this category. The second type is a separating technology, where harmful organic pollutants are separated from the effluent by various methods.
∗ Corresponding author at: N10W5 kita-ku, Sapporo, Hokkaido 060-0810, Japan. Tel.: +81 11 706 2219, fax: +81 11 706 2219. E-mail addresses:
[email protected] (T. Sasaki),
[email protected] (S. Tanaka). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.033
Technologies such as membrane separation [5], adsorption [6], and solvent extraction [7] belong to the second category. Among these types of technologies, adsorption is one of the simplest and most effective processes. Adsorption is fast, economic and widely applicable techniques. Using adsorption is applicable for various pollutants such as organic compounds and heavy metals by selecting the type of adsorbent and adsorption conditions. In addition, recently treatment methods for wastewater using lowcost adsorbents such as by-products or waste materials have been reported [8,9]. Gupta and co-workers have reviewed the details of treatment methods for various pollutants in water using low-cost adsorbents [10,11]. Activated carbon is the most widely used adsorbent in various cases because of a large capacity and a wide variety of adsorbates. However, there are some limitations, particularly in regeneration [12]. There is poor mechanical rigidity and low selectivity when activated carbon is applied to real environmental pollution. Furthermore, it is difficult to collect activated carbon powder that has been widely diffused into the environment. If it is not collected, the adsorbent that is used to adsorb harmful pollution could become a secondary source of pollution. Magnetic separation has been applied recently in various fields such as analytical biochemistry [13], medical science [14] and biotechnology [15]. From an environmental point of view, magnetic separation offers advantages due to the easy recovery of the
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1 < log Pow < 2 OH
1.46 a)
1.56 a)
Phenol
Benzonitrile
NO2
NO2
CN
1.62 d)
1.86 a)
Nitrocyclohexane
Nitrobenzene
2 < log Pow < 3 Cl
CH3
2.13 a)
2.73 b)
Benzene
Toluene
2.84 a) Chlorobenzene
3 < log Pow Cl Cl
3.38 c)
o-Dichlorobenzene Fig. 1. Structures and log Pow s of aromatic compounds and nitrocyclohexane. a) the values of log Pow was referred to Ref. [25] b) the values of log Pow was referred to Ref. [26] c) the values of log Pow was referred to Ref. [27] d) the values of log Pow was referred to Ref. [28].
adsorbent without filtration or centrifugation. Several studies have reported magnetic separation using modified magnetite (Fe3 O4 ) as an environmentally friendly approach to remove heavy metal ions [16,17] and organic pollutants [18,19]. The removal of harmful organic compounds with adsorbents is in most cases based on the hydrophobic interaction between an adsorbent and its target compounds. This hydrophobic interaction has been applied to not only the removal of harmful substances by adsorbents but also to the preconcentration of analytes by using solvents [7] and solid extraction [20]. A hydrophobic adsorbent can be prepared by using a hydrophobizing agent on the surface of materials such as magnetite and silica beads [18,21]. One of the most popular techniques for hydrophobizing involves the use of silane coupling agents. Silane coupling agents are known as surface modifiers that can add an organic property to the surface of an inorganic material [22,23]. Therefore, a silane coupling agent is important when forming a hybrid inorganic-organic material. Another technique is the use of an ionic surfactant. An ionic group of the surfactant will turn toward the surface of mineral oxides such as alumina, silica, titanium dioxide and ferric oxide, and then the alkyl chain, the hydrophobic group of the surfactant, will orient to the outside. As a result, the surface of the material becomes hydrophobic [24]. The strength of the hydrophobic interaction depends on the degree of hydrophobicity of both the adsorbent and the adsorbate. The hydrophobicity of an organic compound can be varied by the structure and functional group of the compound. The octanol–water partition coefficient (Pow ) is a well-known indicator of the hydrophobicity of an organic compound. A higher hydrophobicity compound will have a larger Pow . Therefore, the hydrophobicity, or Pow , of an organic compound is very important in the prediction of the adsorptive behavior of some organic compounds in water. The aim of the present study was to clarify the adsorption behaviors of organic compounds on hydrophobic magnetite and to evaluate the possibility of magnetic separation for the removal of organic compounds dissolved in water. Organic compounds
with low Pow s, or relatively weak hydrophobicities, were used in this study. These organic compounds were selected on the basis of the value of log Pow . The selected compounds were divided into 3 groups according to log Pow : 1 < log Pow < 2, 2 < log Pow < 3, and 3 < log Pow . The Pow s of phenol, benzonitrile and nitrobenzene fell into the 1 < log Pow < 2 group. Those of benzene, toluene and chlorobenzene were in the 2 < log Pow < 3 group. o-Dichlorobenzene was in the 3 < log Pow group [25–29]. The individual values of log Pow s of aromatic compounds used in the adsorption experiments are summarized in Fig. 1. The hydrophobic magnetite was prepared by hydrophobizing the surface of a magnetite particle. Stearic acid and phenyltrimethoxysilane were used to hydrophobize the surface of magnetite. By using two different types of hydrophobic magnetite, the difference in adsorption behaviors of the various aromatic compounds was investigated. 2. Experiments 2.1. Materials Magnetite where average size was 0.3 m (the data was provided from Kishida Chemical.), was purchased from Kishida Chemical Co., Ltd. (Osaka, Japan) and used as the adsorbent carrier. The modifying reagents, stearic acid and phenyltrimethoxysilane were purchased from MP biomedicals Japan K. K. (Tokyo, Japan) and Tokyo Chemical Industry Co., Ltd. (Tokyo, Japan), respectively. The organic compounds used as adsorbates, phenol, nitrobenzene and benzene were purchased from Wako Pure Chemical Industries, Ltd. (Osaka, Japan). Toluene, chlorobenzene and o-dichlorobenzene were purchased from Nacalai Tesque, Inc. (Kyoto, Japan). Benzonitrile was purchased from Kanto Chemical Co., Inc. (Tokyo, Japan). Nitrocyclohexane was purchased from Tokyo Chemical Industry Co., Ltd. In analysis for fatty acid, a boron trifluoride solution was used as an esterification agent, anhydrous sodium sulfate was used as a dehydration agent, and methylene chloride was used as a solvent, and all were purchased from Wako Pure Chemical Industries, Ltd. The reagents used in Si analysis, (hydrochloric acid, nitric acid,
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sulfuric acid, hydrofluoric acid, boric acid, hexaammonium heptamolybdate tetrahydrate and sodium chloride) were purchased from Wako Pure Chemical Industries, Ltd. All regents were of analytical grade and used without further purification. 2.2. Preparation of hydrophobic magnetite Two kinds of hydrophobic adsorbents were prepared: stearic acid modified magnetite (SA-mag) and phenyl group modified magnetite (Ph-mag) [29,30]. SA-mag was prepared in the following way. 1.0 g of magnetite was added in 50 ml of methanol for distribution. Then 10 mg of stearic acid was added to the suspension with stirring to dry out the methanol. The residue was washed twice with methanol and dried in an oven at 50 ◦ C. Ph-mag was prepared in the following way. 1.0 g of magnetite was added in 40 ml of ethanol and distributed with stirring. Then, 0.1 ml of phenyltrimethoxysilane was added to the suspension. After phenyltrimethoxysilane was sufficiently dissolved in the solvent, 0.058 ml of H2 O and 0.025 ml of 1 M HCl were added to the suspension. The suspension was then heated at 50 ◦ C with continuous stirring until the solvent was dry. The obtained magnetite was heated in a muffle furnace at 120 ◦ C for 1 h. After heating, the modified magnetite was washed twice with ethanol and dried in an oven at 50 ◦ C. 2.3. Characterization of hydrophobic magnetite 2.3.1. Surface areas of hydrophobic magnetite The surface areas of SA-mag and Ph-mag were determined using a N2 BET analysis (AutoSorb6 YUASA, Japan). The samples were pretreated by degassing at 80 ◦ C for 6 h. 2.3.2. Amount of modified stearic acid on magnetite For desorption of modified stearic acid from SA-mag, 30 mg of SA-mag was washed with 20 ml of ethanol for 1 h with sonication. The washing procedure was repeated 3 times. And then, the collected ethanol solution that included stearic acid was evaporated under vacuum. The obtained sample was dissolved in 3 ml of methanol and 1 ml of methanolic solution containing lauric acid (1 mg/l) as the surrogate standard. The mixture was transferred to a test tube and 2 ml of 14% BF3 methanolic solution was added. The sample was placed in a water bath at 80 ◦ C for 3 min and then 1 ml of water was added to stop the reaction. The fatty acid esters were extracted twice by 1 ml of methylene chloride each time. The collected organic phase was dehydrated with anhydrous sodium sulfate and placed in a 10 ml volumetric flask after filtration. The adjusted 10 ml of sample solution was measured using a GC-17A (Shimadzu, Japan) equipped with a GCMS-QP5050A (Shimadzu, Japan). A DB-5 ms (30 m × 0.25 mm × 0.25 m) column (Agilent, USA) was used. 1 l of sample solution was injected. The injections were performed in splitless mode. The carrier gass was helium (Air Water, Japan) at a constant flow of 1.5 ml/min. The injection port was heated to 200 ◦ C. The oven temperature was set at 75 ◦ C for 2 min, then increased 30 ◦ C/min to 270 ◦ C, and the final temperature was held for 1 min. The temperature at the detector was 280 ◦ C. All mass spectra were acquired in the electron impact (EI) mode as the ionization source with a quadrupole mass filter. The analysis was carried out in SIM mode, and the selected ions of the compound were m/z 55, 74 and 87. The concentrations of the fatty acids were calculated using the internal standard method [31]. 2.3.3. Amount of modified phenyl group on magnetite The amount of modified phenyltrimethoxysilane on the magnetite surface was determined from the measurement of silica dioxide as a decomposition component of Ph-mag in the following way [32]. First, 1 g of Ph-mag was decomposed using 40 ml of the mixed acid contained 6 M hydrochloric acid and 6 M nitric acid
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covering a watch glass with heating at 80 ◦ C until the black precipitate disappeared. The obtained precipitated silicate was filtrated by a membrane filter and washed with water several times. Second, the filter paper carrying the precipitation was moved into a PTFE beaker and then 5 ml of 4% (w/v) NaCl solution and 3 ml of 60% hydrofluoric acid were added to the beaker. The mixture was heated until dry here in a water bath. After dissolving the obtained residue with 15 ml of water, 10 ml of saturated boric acid solution was added to the mixture and then was heated almost to the boiling point (<100 ◦ C). Finally, 2 ml of 5N sulfuric acid was added to the obtained mixture followed by cooling to about 40 ◦ C to an acidic condition. After filtration of the obtained mixture, the filtrate and the hot water used for washing were collected in 50 ml of volumetric flask and cooled to room temperature. Then, 5 ml of 2% (w/v) ammonium molybdate solution was added to the solution and the volume was adjusted to 50 ml with water. After 10 min, the absorbance of the mixed solution was measured by UV–vis at 425 nm. 2.4. Adsorption experiments All adsorption experiments were carried out by the following batch test. To 10 ml of organic compound solution was added 20 mg of adsorbent that was mixed with a shaker (V BR-36 TAITEC Co., Ltd., Saitama, Japan). After adsorption, the magnetite was separated from the solution using a Nd–Fe–B magnet (50 mm × 10 mm × 5 mm, 0.4 T). Then, the supernatant was measured by UV–vis (V-550 Jasco Co., Tokyo, Japan). Nitrocyclohexane measurement was performed by HPLC (Waters 2695 separation module and 2487 dual absorbance detector Waters, Milford, MA) with a C18 ODS column (CAPCELL PAK C18 MG II S5 (5 m, 4.6 mm id × 150 mm length) Shiseido Oo., Ltd. Tokyo, Japan). The mobile phase was acetonitrile–water (80:20), the flow rate was 1 ml/min and the wavelength was 210 nm. 3. Resuts and discussion 3.1. Preparation of hydrophobic magnetite The modification processes of the surface modifiers on the magnetite surface are shown in Fig. 2. Stearic acid had a carboxyl group. However, the magnetite had a hydroxyl group on the surface. Therefore, the carboxyl group of the stearic acid and the hydroxyl group on the magnetite interacted by hydrogen bond. As a result, the carboxyl group of the stearic acid faced inward to contact the magnetite, and the alkyl chain of the stearic acid was oriented toward the outside (Fig. 2A) [24].
Fig. 2. Hydrophobization of magnetite surface (A) is modification with stearic acid and (B) is with phenyl group on magnetite surface.
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Fig. 3. Photographs of the distribution behaviors of (A) SA-mag, (B) Ph-mag and (C) plane-mag in two-phase system. The upper phase was water and the bottom phase was chloroform.
Phenyltrimethoxysilane had 3 methoxyl groups in its structure. These methoxy groups were substituted for hydroxyl groups in a solution of HCl. The hydroxyl groups interacted with a hydroxyl group on the magnetite surface through a hydrogen bond, then a covalent bond was formed by dehydration and condensation followed by heating for 1 h at 120 ◦ C (Fig. 2B) [30]. This produced magnetite with a modified phenyl group. To confirm that the prepared modified magnetite had a hydrophobic surface, the powder of the modified magnetite was added and shuffled into a two-phase system composed of water in a neutral pH and chloroform. As a comparative control, plane magnetite was also added to the same system. Fig. 3 shows the distribution behaviors of 3 types of magnetite in a two-phase system. SA-mag and Ph-mag were selectively dispersed in the chloroform phase at the bottom of a vial container. These results showed that the surfaces of magnetite modified with stearic acid and phenyltrimethoxysilane were hydrophobic. However, the plane magnetite was dispersed in both phases. This result indicated that the plane magnetite had a middle property between hydrophilic and hydrophobic.
3.2. Characterization of hydrophobic magnetite 3.2.1. Surface areas of hydrophobic magnetite The N2 adsorption/desorption isotherms of the SA-mag and Ph-mag are shown in Fig. 4. Both isotherms scarcely revealed a hysteresis loop, therefore, SA-mag and Ph-mag were nonporous
20 Ph-mag Adsorption
18
Ph-mag Desorption
16
SA-mag Adsorption SA-mag Desorption
3
Volume (cm /g)
14 12 10 8 6
materials. The surface areas of SA-mag and Ph-mag were 3.65 m2 /g and 4.88 m2 /g, respectively. 3.2.2. Amount of modified stearic acid on magnetite Ethanol was used in the extraction of fatty acid because of a higher dissolving ability for stearic acid and a lower volatility than other organic solvents. A high sensitivity with no interference was obtained in the SIM analysis. The retention times of the peaks of lauric acid methyl ester and stearic acid methyl ester were about 7 and 9.5 min, respectively. A calculation result showed that the modified amount of stearic acid on the magnetite was 9.84 × 10−3 mmol/g. After the extraction of stearic acid, a hydrophobic property on the magnetite surface could not be observed. 3.2.3. Amount of modified phenyl group on magnetite The amount of the modified phenyl group on the magnetite surface was determined by measurement of the silicate included in the modified phenyltrimethoxysilane. The first treatment using hydrochloric acid and nitric acid with heating was performed for the decomposition of magnetite. Simultaneously, the modified silane coupling agent, an organosilicon compound, was oxidized with nitric acid and precipitated as an insoluble inorganic silicate. Next, the treatment by hydrofluoric acid can dissolve the insoluble silicate by formation of sodium hexafluorosilicate with sodium ion and hydrofluoric acid. The excess hydrofluoric acid after the above reaction was removed from the mixture with further heating. The dissolved mixture of boric acid solution was heated almost to the boiling point, which led to the release of soluble silicic acid with a reaction between the fluoric ion in the hexafluoric silicate and boric acid. Molybdosilicate acid, which is measurable by UV–vis, was formed with a reaction between the obtained soluble silicic acid and ammonium molybdate in an acidic condition. According to the analysis results of the molybdosilicate acid, the amount of modified phenyl group on the magnetite surface was 4.17 × 10−2 mmol/g. These results of modifier amounts on magnetite were used in the following discussion about the adsorption behaviors of aromatic compounds. 3.3. Adsorption experiments for aromatic compounds
4 2 0 0.0
0.1
0.2
0.3
0.4
0.5
0.6
0.7
0.8
0.9
P/P0 Fig. 4. N2 adsorption/desorption isotherm of SA-mag and Ph-mag.
1.0
The adsorption behaviors of several aromatic compounds were examined using two kinds of hydrophobic magnetite. Fig. 5 shows the results of adsorption experiments in various initial concentrations (10–100 ppm of these compounds). The large slope of adsorption isotherms in Fig. 5 represents a high adsorption ability. Both SA-mag and Ph-mag showed a higher ability for
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0.30 0.20 0.10 0
case. In addition, Fig. 6 shows that the adsorption amounts of o-dichlorobenzene onto SA-mag or Ph-mag under all initial concentration were larger than the modified amounts of stearic acid and phenyl groups on the magnetite surface. Thus, the adsorption behavior of o-dichlorobenzene was multi-layer adsorption rather than mono-layer adsorption for an adsorption site. The adsorption behavior of o-dichlorobenzene did not depend on the type of modifiers (SA-mag and Ph-mag). The hydrophobic interaction between the modified surface of magnetite and o-dichlorobenzene may be too strong to determine the dependency of the adsorption amount on the functional group on the surface.
o-Dichlorobenzene Benzene Toluene Chlorobenzene Phenol Benzonitrile Nitrobenzene 0
20
40
60 Ce (ppm)
80
100
qe (mmol/g)
B 0.40 0.30 0.20 0.10 0
0
20
40
60 80 Ce (ppm)
100
Fig. 5. The adsorption amounts of aromatic compounds onto (A) SA-mag and (B) Ph-mag in various initial concentrations (10–100 ppm).
o-dichlorobenzene compared with the other compounds. The slopes for benzene, toluene and chlorobenzene were almost equal, and were lower than that of o-dichlorobenzene. The slopes for phenol, benzonitrile and nitrobenzene were the smallest of the compounds tested here. These results show that the adsorption ability of the modified magnetite seemed to depend on the values of log Pow for these compounds. The adsorption behaviors could be divided into 3 groups according to the following Pow order: 1 < log Pow < 2, 2 < log Pow < 3, and 3 < log Pow . 3.3.1. 3 < log Pow o-Dichlorobenzene has the highest hydrophobicity among the compounds used in this study. The results by adsorption experiments with SA-mag and Ph-mag are represented in Fig. 6. Fig. 6 shows that the adsorption amount of o-dichlorobenzene continued to increase as the concentration increased. When the initial concentration of o-dichlorobenzene was 100 ppm, the adsorption ratio (C/C0 ) of o-dichlorobenzene on both SA-mag and Ph-mag reached about 90%. If this adsorption was due to the mono-layer adsorption, the adsorption amount of adsorbate would be expected to level off. However, a leveling off was not observed in this
3.3.2. 2 < log Pow < 3 Benzene, toluene and chlorobenzene had medium hydrophobicity (2 < log Pow < 3) in this study. The results of adsorption experiments using these compounds with hydrophobic magnetite are shown in Fig. 7. The adsorption ratios of all compounds were about 50% of the initial concentration of 100 ppm. In this group, the dependency of the adsorption behavior on the type of the modifiers could not be confirmed, because of the relatively high hydrophobicity of these compounds. As shown in Fig. 7, the adsorption amounts of all compounds under most conditions were larger than the modified amounts of stearic acid and phenyl groups, so the adsorption behavior of this group also seemed to be that of multi-layer adsorption. The order of the adsorbed amounts was as follows: benzene > toluene > chlorobenzene. This result was not consistent with the order of log Pow . That is, the order of the adsorption ability did not always depend on the log Pow of the compounds in this group. The adsorption mechanism in this group was also that of multi-layer adsorption based on the hydrophobic interaction. Therefore, the adsorption amount depended on the strength of the interaction and the size of the molecule forming the multi-layer. Benzene was adsorbed most because there was no substituent group and it had the smallest molecular size [33]. On the other hand, toluene and chlorobenzene are more bulky than benzene because of a substituent group. The methyl group, a substituent group of toluene, supplies electrons to the benzene ring because
A 0.25 qe (mmol/g)
qe (mmol/g)
A 0.40
331
Benzene Toluene Chlorobenzene
0.20 0.15 0.10 0.05 0
0
20
40
60
40
60
Ce (ppm)
0.40
B 0.25
Ph-mag
0.30
qe (mmol/g)
qe (mmol/g)
SA-mag
0.20 0.10 0
0.20 0.15 0.10 0.05
0
5
10
15
Ce (ppm) Fig. 6. The adsorption amounts of o-dichlorobenzene onto SA-mag and Ph-mag in various initial concentrations (10–100 ppm).
0
0
20
Ce (ppm)
Fig. 7. The adsorption amounts of aromatic compounds in 2 < log Pow < 3 onto (A) SA-mag and (B) Ph-mag in various initial concentrations (10–100 ppm).
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the functional group is the electron-donating group. By contrast, chlorine, a substituent group of chlorobenzene, is an electronwithdrawing group and withdraws electrons from the benzene ring. The strength of the -electron interaction depends on the richness of the -electron in the benzene ring, that is, the interaction works well in a -electron-rich state. The -electron interaction between aromatic compounds seems to be important for the formation of a multi-layer in these cases. Thus, toluene adsorbed more than chlorobenzene. 3.3.3. 1 < log Pow < 2 Phenol, benzonitrile, nitrobenzene and nitrocyclohexane had the lowest hydrophobicity (1 < log Pow < 2) in this study. Fig. 8 shows the results of these adsorption experiments. The adsorption amounts of these compounds in this group were small by comparison with other groups. This happened because these compounds have strong polar groups such as hydroxyl, cyano and nitro groups and can be hydrated with water molecules around their polar groups. Therefore, the hydrophobic interaction between these compounds and the modified magnetite becomes weak. In the same way, the interaction between their compounds also becomes weak. According to Fig. 8, phenol, benzonitrile and nitrobenzene were more selectively adsorbed onto Ph-mag than onto SA-mag. These results indicate that the functional group on the magnetite surface significantly affects adsorption behavior and -electron interactions between the compounds and the phenyl group on Ph-mag works stronger rather than hydrophobic interaction in these cases. When nitrocyclohexane, having no aromatic ring, was scarcely adsorbed onto Ph-mag, this suggested that the adsorption mechanism of this group was based on -electron interaction. The order of the adsorption amount of these aromatic compounds was as follows: phenol > benzonitrile > nitrobenzene. These results did not agree with the order of log Pow (phenol 1.46, benzonitrile 1.56, nitrobenzene 1.86). Phenol has a hydroxyl group, an electrondonating group, and the -electron interaction between phenol
qe (mmol/g)
A
0.03
Phenol Benzonitrile Nitrobenzene Nitrocyclohexane
0.02
and Ph-mag was strongest among them. However, the substituent groups of benzonitrile and nitrobenzene are electron-withdrawing groups that make their aromatic rings relatively electron-poor. The electron-withdrawing property of benzonitrile might be weaker than that of nitrobenzene because the dipole moment of benzonitrile, 4.05, is smaller than that of nitrobenzene, 4.22–4.91 [34,35]. The aromatic ring of benzonitrile is more electron-rich than that of nitrobenzene. The -electron interaction between the compound and the phenyl group on the magnetite surface was dominant in this adsorption mechanism and so the adsorption amount of benzonitrile on Ph-mag was more than that of nitrobenzene. According to Fig. 8A, phenol was adsorbed on SA-mag, and the amount of adsorbed phenol was close to the modified amount of stearic acid on magnetite. However, this does not mean that adsorption is a one-to-one relationship between phenol and stearic acid like a mono-layer adsorption because of the shape of the adsorption isotherm. A feature of the adsorption isotherm is that the adsorption amount immediately increases with a high concentration of adsorbate during multi-layer adsorption. Moreover, the dipole moment of phenol, 1.53 [36], was the lowest and hydration for phenol was the weakest. Thus, a small amount of phenol could be adsorbed onto SA-mag by hydrophobic interaction. However, the adsorption behaviors of this group were not clear in the initial concentration range (10–100 ppm). Fig. 8 shows that the adsorption amounts of the compounds are smaller compared with the amounts of the modified functional groups, particularly in the case of a phenyl group with an initial concentration of 100 ppm. Therefore, the adsorption experiment for this group should be carried out with a higher initial concentration. 3.4. Adsorption isotherm models This section describes an investigation into the adsorption isotherms of these aromatic compounds onto two adsorbents (in Fig. 5) by application of the adsorption isotherm models. The multilayer adsorption model in the gas phase is known as the BET model. Many researchers have reported the use of the-multi layer adsorption model in the liquid phase based on the BET model, but this model has not yet been established [37]. Thus, in this study, adsorption behaviors were studied by fitting these adsorption data into the Langmuir and Freundlich models. The Langmuir and Freundlich models are as follows: 1 1 1 = + qe KL qm Ce qm
0.01
1/n
qe = Kf Ce
0.00
B
0.03
qe (mmol/g)
0
0.02
20
40
60 Ce (ppm)
80
100
120
20
40
60 80 Ce (ppm)
100
120
0.01
0.00 0
Fig. 8. The adsorption amounts of aromatic compounds and nitrocyclohexane in 1 < log Pow < 2 onto (A) SA-mag and (B) Ph-mag in various initial concentrations (10–100 ppm).
(1) (2)
where qe is the equilibrium adsorption capacity, Ce represents the solute concentration in equilibrium, qm (mg/g) is the maximal sorption capacity, KL (L/g) is a binding constant, and Kf and n are the Freundlich constants to be determined. Table 1 shows the fitting results for the Langmuir and Freundlich models. When applying the Langmuir model, most correlation coefficients are not improved. The Langmuir model is well known as a mono-layer adsorption model and most of the adsorption behaviors seem not to be of the mono-layer adsorption type, because of the low correlation coefficients. On the other hand, when applying the Freundlich model, most correlation coefficients are better than that of the Langmuir model. However, most of the correlation coefficients for the Freundlich model were not so good. The Freundlich model does not provide insight into adsorption behavior, because of an empirical formula. However, the Freundlich model is well known to be a better fit for adsorption into a porous material such as activated carbon. Thus, these adsorption behaviors seemed not to be porous adsorption. This result also agrees with Fig. 4. These fitting results for the models were not definitive as to whether the
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333
Table 1 Isotherm parameters of aromatic compounds onto SA-mag and Ph-mag. Adsorbent
Adsorbate
Freundlich model
Langmuir model
Kf
n
R2
qm (mmol/L)
KL (L/mg)
R2
SA-mag
Phenol Benzonitrile Nitrobenzene Benzene Toluene Chlorobenzene o-Dichlorobenzene
3.46e−5 2.70e−4 – 3.83 1.76 2.80 16.4
0.453 0.647 – 2.52 1.80 1.45 2.97
0.807 0.988 – 0.984 0.962 0.930 0.883
1.17e−3 9.52e−4 – 0.234 0.203 0.0690 0.248
0.0101 9.86e−3 – 0.135 0.0543 0.933 1.90
0.213 0.991 – 0.960 0.911 0.423 0.979
Ph-mag
Phenol Benzonitrile Nitrobenzene Benzene Toluene Chlorobenzene o-Dichlorobenzene
0.528 0.0383 0.115 3.78 1.53 1.56 16.1
3.63 1.26 1.83 3.07 1.70 0.885 2.82
0.932 0.978 0.995 0.951 0.935 0.986 0.853
0.0189 0.0185 0.0134 0.135 0.199 0.0907 0.292
0.135 0.0151 0.0310 0.508 0.0534 0.224 1.19
0.819 0.937 0.908 0.812 0.927 0.724 0.944
0.05 Ph-mag
qe (mmol/g)
0.04
SA-mag
0.03 0.02 0.01 0
0
2
4
6
8
10
Conc. of Salt (w/v%) Fig. 9. Influence of salt in adsorption of nitrobenzene onto SA-mag and Ph-mag.
adsorption behaviors in 1 < log Pow < 2 were mono or multi-layer adsorption. 3.5. Nitrobenzene adsorption in various salt concentrations Shown in Fig. 9 are nitrobenzene adsorption isotherms for various concentrations of salt on SA-mag and Ph-mag. This experiment was performed to validate whether hydration with water inhibits the adsorption of aromatic compounds in 1 < log Pow < 2 on hydrophobic magnetite. The dehydration effects of the addition of salt were investigated. The addition of salt deprives hydrated adsorbate of water molecules. If the adsorption amount of nitrobenzene is increased when the adsorption experiment is carried out under dehydration conditions, then hydration would be an inhibiting factor in its adsorption onto hydrophobic magnetite. As shown in Fig. 9, the adsorption amounts of nitrobenzene on SA-mag and Ph-mag increased with increasing salt concentration. Furthermore, nitrobenzene was hardly adsorbed onto SA-mag without the addition of salt, but was adsorbed onto SA-mag in the presence of salt. The nitrobenzene adsorption onto SA-mag was by hydrophobic interaction, which seems to indicate that the hydrophobicity of nitrobenzene was enhanced by dehydration. This result indicates that hydration is one of the inhibitors of the adsorption of nitrobenzene onto hydrophobic magnetite. 4. Conclusion Aromatic compounds with various log Pow s were investigated for adsorption onto hydrophobic magnetite coated alkyl chains and phenyl groups. The adsorption behaviors of each compound were divided into 3 groups depending on the log Pow : 1 < log Pow < 2, 2 < log Pow < 3 and 3 < log Pow . The adsorption amounts generally
increased as the log Pow of each group increased. However, the adsorption amounts for the compounds in each group did not depend on the values of log Pow . In the 2 < log Pow < 3 and 3 < log Pow groups, the adsorption mechanism was mainly a hydrophobic interaction between aromatic compounds and the surface of hydrophobic magnetite. The adsorption behaviors did not depend on the difference in modified functional groups on the magnetite surface. The adsorption behavior of hydrophobic magnetite in these groups seemed to form a multi-layer on the hydrophobic magnetite because of the quantitative relationship between the amount of adsorbed aromatic compound and the amount of modified functional groups on the magnetite, and there was a poor fitting for the adsorption isotherm models. However, the adsorption behavior of the 1 < log Pow < 2 group was sensitive to the modified functional group on the hydrophobic magnetite. The main adsorption mechanism for this group was the -electron interaction between the compounds and the phenyl group on Ph-mag rather than the hydrophobic interaction. The adsorption behavior of this group could not be demonstrated for either mono or multilayer adsorption under these adsorption conditions. The results of adsorption experiments under dehydration conditions indicated that an inhibiting factor for nitrobenzene adsorption is hydration from water molecules. The simple system of the present study, which used magnetite without porosity, enabled clarification of the factors that determine adsorption behavior. Thus, the results were significant with regard to the selection of an optimal surface modifier for adsorption or solid-phase extraction as well as magnetite separation, which could be valuable in the design of a novel high-performance adsorbent. Acknowledgements The N2 BET analysis in this work was carried out with Autsorb6 at the OPEN FACILITY, Hokkaido University Sousei Hall. References [1] E. Ferrarese, G. Andreottola, I.A. Oprea, Remediation of PAH-contaminated sediments by chemical oxidation, J. Hazard. Mater. 152 (2008) 128–139. [2] Z.M. Shen, D. Wu, J. Yang, T. Yuan, W.H. Wang, J.P. Jia, Methods to improve electrochemical treatment effect of dye wastewater, J. Hazard. Mater. 131 (2006) 90–97. [3] X. Shen, L. Zhu, G. Liu, H. Yu, H. Tang, Enhanced photocatalytic degradation and selective removal of nitrophenols by using surface molecular imprinted titania, Environ. Sci. Techol. 42 (2008) 1687–1692. [4] F.J. Beltrán, G. Ovejero, J.M. Encinar, J. Rivas, Oxidation of polynuclear aromatic hydrocarbons in water. 1. Ozonation, Ind. Eng. Chem. Res. 34 (1995) 1596–1606. [5] U.K. Ghosh, N.C. Pradhan, B. Adhikari, Separation of water and o-chlorophenol by pervaporation using HTPB-based polyurethaneurea membranes and application of modified Maxwell-Stefan equation, J. Membr. Sci. 272 (2006) 93–102.
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Journal of Hazardous Materials 196 (2011) 335–341
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Zinc induces chemokine and inflammatory cytokine release from human promonocytes Tsui-Chun Tsou a,∗ , How-Ran Chao b , Szu-Ching Yeh a , Feng-Yuan Tsai a , Ho-Jane Lin a a b
Division of Environmental Health and Occupational Medicine, National Health Research Institutes, Zhunan, Miaoli 350, Taiwan Department of Environmental Science and Engineering, National Pingtung University of Science and Technology, Neipu, Pingtung 912, Taiwan
a r t i c l e
i n f o
Article history: Received 18 May 2011 Received in revised form 8 September 2011 Accepted 9 September 2011 Available online 19 September 2011 Keywords: Zinc Promonocytes Chemokines Inflammatory cytokines
a b s t r a c t Our previous studies found that zinc oxide (ZnO) particles induced expression of intercellular adhesion molecule-1 (ICAM-1) protein in vascular endothelial cells via NF-B and that zinc ions dissolved from ZnO particles might play the major role in the process. This study aimed to determine if zinc ions could cause inflammatory responses in a human promonocytic leukemia cell line HL-CZ. Conditioned media from the zinc-treated HL-CZ cells induced ICAM-1 protein expression in human umbilical vein endothelial cells (HUVEC). Zinc treatment induced chemokine and inflammatory cytokine release from HL-CZ cells. Inhibition of NFB activity by over-expression of IB␣ in HL-CZ cells did not block the conditioned medium-induced ICAM-1 protein expression in HUVEC cells. Zinc treatment induced activation of multiple immune response-related transcription factors in HL-CZ cells. These results clearly show that zinc ions induce chemokine and inflammatory cytokine release from human promonocytes, accompanied with activation of multiple immune response-related transcription factors. Our in vitro evidence in the zinc-induced inflammatory responses of vascular cells provides a critical linkage between zinc exposure and pathogenesis of those inflammatory vascular diseases. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Epidemiologic studies indicate that exposure to fine particulate matter (PM) in air pollution is associated with systemic inflammatory markers [1] as well as incidence of cardiovascular morbidity and mortality [2–4]. However, mechanisms behind this correlation remain largely unknown. When inhaled into the respiratory tract, small particles tend to rest in the deeper part of the lungs. Ultrafine particles [5–7] and, of particular relevance to the present study, their dissolved chemicals such as metal ions can penetrate the deepest part of the lungs and cross the pulmonary epithelial barrier into the bloodstream, directly exposing the vascular cells, such as monocytes and endothelial cells, to pollutants. Analysis of zinc levels in suspended PM in air revealed that 82–93% of zinc was in the small PM10 particles [8]. Studies in mothers subjected to cigarette smoking or air pollution showed that both cigarette smoking and air pollution contributed to the increased levels of placental zinc [9]. A previous study in ambient air zinc levels and health care utilization for asthma revealed the association between elevated ambient air zinc and increased
∗ Corresponding author. Tel.: +886 37 246 166x36511; fax: +886 37 587 406. E-mail address:
[email protected] (T.-C. Tsou). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.035
pediatric asthma morbidity [10]. ZnO particles induce cytotoxicity and apoptosis in mammalian cells [11,12] and the dissolved zinc ions seem to play the critical role in toxic effect of ZnO particles [13]. In our previous studies, ZnO particles induced ICAM-1 expression in vascular endothelial cells via an NFB dependent pathway [14] and zinc ions alone were sufficient to induce similar levels of ICAM-1 expression as ZnO particles, suggesting that dissolved zinc ions might play the major role in inflammatory effect of ZnO particles on vascular endothelial cells [15]. These studies suggest that metal particle composition, or its dissolved metal ions, may determine the capability of metal oxide nanoparticles to induce inflammation in vascular endothelial cells. Increasing evidence indicates that chronic obstructive pulmonary disease, asthma, and atherosclerosis are associated with systemic inflammatory cytokine changes. Various pathophysiological stimulators induce cytokine release, including modified LDL [16,17], free radicals [18], hemodynamic stress [19,20], and hypertension [21]. On the basis of our previous findings in vascular endothelial cells using ZnO particles, the present study aimed to determine if zinc could cause inflammatory responses in other vascular cells. We found that zinc induces chemokine and inflammatory cytokine release from human promonocytes possibly via activation of multiple immune response-related transcription factors.
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2. Materials and methods 2.1. Materials Zn(CH3 COO)2 (#370080250) was obtained from ACROS Organics (Geel, Belgium). Zinc preparation in this study was tested to be endotoxin-free by using an endotoxin inhibitor, polymyxin B, as previously described [22]. RayBio Human Cytokine Antibody Array 3 (#AAH-CYT-3) was obtained from RayBiotech, Inc. (Norcross, GA, USA). Rabbit polyclonal antibodies against ICAM-1 (sc-7891) and IB␣ (sc-847) and a goat polyclonal antibody against p-IB␣ (sc-7977) were purchased from Santa Cruz Biotechnology (Santa Cruz, CA, USA). A mouse monoclonal antibody against actin (MAB1501) was purchased from Chemicon Int. Inc. (Temecula, CA, USA). Endothelial cell growth supplement (ECGS) was obtained from Sigma–Aldrich (St. Louis, MO, USA). Dulbecco’s phosphatebuffered saline (D-PBS), M199 medium, and RPMI 1640 medium were obtained from Life Technologies (Grand Island, NY, USA). Fetal bovine serum (FBS) was obtained from HyClone (Logan, UT, USA). Penicillin (10,000 units/ml)/streptomycin (10,000 g/ml) solution was obtained from Invitrogen Corp. (Carlsbad, CA, USA). Gentamycin sulfate was purchased from Biological Industries (Kibbutz Beit Haemek, Israel). 2.2. Construction of recombinant adenoviruses Construction of recombinant AdEasy-GFP, AdEasy-IB␣, and AdV-NFB-Luc has been previously described [14]. For construction of AdV-AP-1-Luc, a 2360-bp DNA fragment containing seven copies of the AP-1 response element, a TATA box, and a firefly luciferase gene in pAP-1-Luc (Stratagene, La Jolla, CA, USA), was amplified by PCR using pfu DNA polymerase. The amplified DNA fragment was digested by KpnI and SalI. After separation by agarose gel electrophoresis, the purified KpnI/SalI-digested DNA fragment was cloned into the KpnI/SalI-digested pACCMV.pLpA vector [23]. The function of the CMV promoter of the pACCMV.pLpA vector used here had been abolished. The recombinant adenovirus AdVAP-1-Luc was generated by homologous recombination between the pJM17 plasmid [24] and the pACCMV.pLpA vector in 293 human embryo kidney cells. Construction of other recombinant adenoviruses (AdEasy-C/EBP-Luc, AdEasy-CRE-Luc, AdEasy-NFATLuc, AdEasy-SRE-Luc, and AdEasy-STAT-Luc) was generated using the AdEasyTM Adenoviral Vector System (Stratagene, La Jolla, CA, USA) (see Supplementary material in detail). The recombinant adenoviruses were purified and concentrated according to the manufacturer’s instructions. General information of these immune response-related transcription factor-mediated luciferase reporter adenoviruses is summarized in Table 1. 2.3. Cells and treatments Human promonocytic leukemia cell line HL-CZ (BCRC-60043), originally established by Dr. Wu-Tse Liu (National Yang-Ming University, Taipei, Taiwan) [25], were purchased from Bioresource Collection and Research Center (BCRC, Hsinchu, Taiwan) and were routinely cultured in RPMI 1640 medium. HUVEC cells were obtained by using collagenase digestion of umbilical veins [26] and were routinely cultured in M199 medium as previously described [27]. HL-CZ cells (8.5 × 106 cells per 100-mm dish) were left untreated or treated with Zn(CH3 COO)2 as indicated for 6 h. Following treatments, conditioned media were dialyzed against D-PBS with stirring at 4 ◦ C for 42 h, sterilized with a 0.45-m syringe filter, and then was added with M199 medium (with 20% FBS and 30 g/ml ECGS) at 1/1 ratio (v/v). HUVEC cells were treated with this conditioned medium/M199 mixtures for different time periods
as indicated. Following the treatments, cell lysates were collected for immunoblot analyses. In some cases requiring adenovirus infection, HL-CZ cells (3 × 106 cells per 100-mm dish) were first infected with AdEasyGFP or AdEasy-IB␣ at a multiplicity of infection (MOI) of 50 pfu/cell for 24 h. The infected cells were replaced with fresh RPMI 1640 medium and cultured for another 24 h for recovery. Hereafter, the cells were ready for the zinc treatments as just described. 2.4. Immunoblot analysis Following treatments, cells were lysed in ice-cold RIPA buffer (50 mM Tris–HCl, pH 7.5, 5 mM EDTA, 1 mM EGTA, 1% Triton X-100, 0.25% sodium deoxycholate) containing PMSF, (2 mM), aprotinin (2 g/ml), leupeptin (2 g/ml), NaF (2 mM), Na3 VO4 (2 mM), and -glycerophosphate (0.2 mM). The cell lysates were subjected to SDS–PAGE and immunoblot analysis, as described previously [28]. The blots were probed with a primary antibody against ICAM-1, phosphor-IB␣ (p-IB␣), IB␣, or actin. HRP-conjugated secondary antibodies. Protein bands in the membrane were visualized in an Xray film by using Western Lightning Chemiluminescence Reagent Plus (PerkinElmer Life Sciences, Boston, MA, USA). The protein band intensity was quantified by densitometry scanning of X-ray films. 2.5. Analysis of cytokines in conditioned media from HL-CZ cells Following zinc treatments of HL-CZ cells, conditioned media were dialyzed against D-PBS with stirring at 4 ◦ C for 42 h and sterilized with a 0.45-m syringe filter. Cytokines in conditioned medium were analyzed with the RayBio Human Cytokine Antibody Array 3 according to the manufacturer’s instructions (see Supplementary material in detail). The kit provides a simple array format, and highly sensitive approach to simultaneously detect 42 cytokine expression levels from conditioned media. 2.6. Immune response-related transcription factor-mediated luciferase reporter assay To determine the activation of those immune response-related transcription factors, HL-CZ cells (1.0 × 104 cells per well in 96-well plates) were infected with one of the recombinant adenoviruses (AdV-AP-1-Luc, AdV-NFB-Luc, AdEasy-SRE-Luc, AdEasy-NFATLuc, AdEasy-CRE-Luc, AdEasy-C/EBP-Luc, and AdEasy-STAT-Luc) (Table 1) at a MOI of 1 pfu/cell for 24 h. Following the adenovirus infection, the infected cells were replaced with fresh RPMI 1640 medium and cultured for another 24 h for recovery. Then, the infected cells were left untreated or treated with 150 M Zn(CH3 COO)2 for 6 h. Luciferase activity of each sample was determined using the Luciferase Assay System (Promega, Madison, WI), according to the manufacturer’s instructions. 2.7. Statistics Each experiment was performed independently at least three times. The statistical analysis was expressed using the mean ± standard deviation (SD) from each independent experiment. Induction of ICAM-1 protein expression in HUVEC cells by conditioned medium were examined by Student’s t-tests with 2000 bootstrap samples. One-sample t-tests were used to determine the significant differences in induction of chemokine and inflammatory cytokine release from HL-CZ cells between the zinc-treated and untreated groups (test value = 1). Differences were considered statistically significant when p < 0.05. Analyses were carried out using the Statistical Package for Social Sciences (SPSS) version 12.0 (SPSS Inc., Chicago, IL, USA).
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Table 1 Immune response-related transcription factor-mediated luciferase reporter adenoviruses. Recombinant adenoviruses
Transcription factors
Response element sequences (RES) a direction (5 → 3 )
AdV-AP-1-Luc
AP-1
TGACTAATGACTAATGACTAATGACTAA TGACTAATGACTAATGACTAA
AdEasy-C/EBP-Luc
C/EBP
ATTGCGCAATATTGCGCAATATTGCGCAAT
AdEasy-CRE-Luc
CREBP
AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAG
AdEasy-NFAT-Luc
NFAT
ACGCCTTCTGTATGAAACAGTTTTTCCTCC ACGCCTTCTGTATGAAACAGTTTTTCCTCC ACGCCTTCTGTATGAAACAGTTTTTCCTCC
AdV-NFB-Luc
NFB
GGGGACTTTCCGCTTGGGGACTTTCCGCT GGGGACTTTCCGCTGGGGACTTTCCGCT GGGGACTTTCCGC
AdEasy-SRE-Luc
Elk1/SRF
CCATATTAGGACATCTAGGATGT CCATATTAGGACATCTAGGATGTCCATATTAGG AGCTAGCCCATATTAGGACATGCTAGGATGT CCATATTAGGAGCATCTAGGATGTCCCATATTAGG AC
AdEasy-STAT-Luc
STAT
GGTTCCCGTAAATGCATCAGGTTCCCGTAAA TGCATCAGGTTCCCGTAAATGCATCAGG TTCCCGTAAATG
a
RES-driven luciferase constructs
The response elements are marked by shading or underlining.
3. Results 3.1. Conditioned media from the zinc-treated HL-CZ cells induce ICAM-1 protein expression in HUVEC cells Our previous study showed that zinc ions alone are sufficient to induce similar levels of ICAM-1 expression as ZnO particles, suggesting that the dissolved zinc ions play the major role in inflammatory effect of ZnO particles on vascular endothelial cells. In this study, we used HL-CZ cells, a human promonocytic leukemia cell line, and a soluble zinc compound Zn(CH3 COO)2 to determine if zinc ions could cause inflammatory responses in this human promonocyte cell line. HL-CZ cells were treated with different concentrations of Zn(CH3 COO)2 (0, 30, 50, 100, and 150 M) for 6 h and then the conditioned media were collected. Then, HUVEC cells were treated with the conditioned media for 24 h. Following treatments, HUVEC cell lysates were collected for analysis of ICAM-1 protein expression with immunoblot. Results in Fig. 1A showed that levels of ICAM-1 protein expression in HUVEC cells were positively correlated with the Zn(CH3 COO)2 concentrations used in HL-CZ treatments. The time-dependent ICAM-1 induction in HUVEC cells by the conditioned media from HL-CZ cells treated with 150 M Zn(CH3 COO)2 was also observed (Fig. 1B). The ICAM-1 induction could be up to 6–7 folds. These results suggested that zinc treatments might cause release of inflammatory cytokines from HL-CZ cells into culture medium and the released inflammatory cytokines were able to activate ICAM-1 expression in HUVEC cells. 3.2. Zinc treatments induce chemokine and inflammatory cytokine release from HL-CZ cells Because a large number of cytokines have been characterized, it was complicated that how to effectively identify the expression profiles of multiple cytokines in conditioned medium. By using the RayBio Human Cytokine Antibody Array 3 for detection of secreted/active cytokines, we were able to simultaneously detect 42 cytokine levels in conditioned media. Results in Fig. 1 showed
that conditioned media from HL-CZ cells treated with 150 M Zn(CH3 COO)2 caused the maximum level of ICAM-1 expression in HUVEC cells. Therefore, conditioned media by such zinc treatments were collected for cytokine analysis with the cytokine antibody array. As shown in Table 2, the zinc treatment induced significant releases of GRO-␣, IL-6, IL-7, IL-8, and IL-10 by 3.98, 1.92, 1.72, 1.34, and 1.46 folds, respectively. Although, the array detects only 42 cytokines, the results clearly show that the zinc treatment causes significant releases of chemokines (e.g., GRO-␣ and IL-8), pro-inflammatory cytokines (e.g., IL-6 and IL-7), and antiinflammatory cytokines (e.g., IL-10) from HL-CZ cells. 3.3. Inhibition of NFB activity by over-expression of IB˛ in HL-CZ cells does not block the conditioned medium-induced ICAM-1 expression in HUVEC cells Because NFB plays the major role in regulating the zincinduced ICAM-1 expression in HUVEC cells [14], it was of importance to further ask if NFB also mediates the zincinduced inflammatory cytokine release from HL-CZ cells. By over-expression of IB␣ in HL-CZ cells using an adenovirusmediated expression system, we investigated whether the zinc treatment was able to induce IB␣ phosphorylation in HL-CZ cells and whether overexpression of IB␣ in HL-CZ cells could block the conditioned medium-induced ICAM-1 expression in HUVEC cells. As shown in Fig. 2, in the un-infected and the AdEasyGFP-infected controls, treatment of HL-CZ cells with 150 M Zn(CH3 COO)2 for 6 h induced degradation of endogenous IB␣ by 63%; IB␣ phosphorylation was barely detectable most likely due to the rapid polyubiquitination and subsequent degradation of phosphorylated IB␣ by the 26S proteasome [29]. In the adenovirus-mediated overexpression of IB␣ experiments, results indicated that inhibition of NFB activity by over-expression of IB␣ in HL-CZ cells did not block the conditioned medium-induced ICAM-1 protein expression in HUVEC cells; meanwhile, the zinc treatments did enhance phosphorylation of exogenous IB␣ in HLCZ cells. Because of the abundant IB␣ expression by adenovirus
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Table 2 Analysis of zinc effect on cytokine release from HL-CZ cells with RayBio Human Cytokine Antibody Array 3. Cytokines
ENA-78 GCSF GM-CSF GRO GRO-␣ I-309 IL-1␣ IL-1 IL-2 IL-3 IL-4 IL-5 IL-6 IL-7 IL-8 IL-10 IL-12 p40p70 IL-13 IL-15 INF-␥ MCP-1 MCP-2 MCP-3 MCSF MDC MIG MIP-1␦ RANTES SCF SDF-1 TARC TGF-1 TNF-␣ TNF- EGF IGF-1 Angiogenin Oncostatin M Thrombopoietin VEGF PDGF BB Leptin * **
Induction fold (treated/untreated)
1st
2nd
3rd
0.540 1.027 0.978 1.854 3.454 1.054 0.799 0.992 1.018 0.550 0.764 0.812 1.930 1.427 1.239 1.631 0.613 0.892 0.708 0.904 1.449 1.005 0.540 0.962 1.176 0.699 0.716 1.031 0.951 0.678 0.724 0.813 1.008 0.781 0.827 0.533 1.014 0.886 1.012 1.361 0.890 0.998
1.059 1.009 1.484 2.479 5.212 1.186 1.271 1.052 1.447 0.987 2.004 1.735 2.028 1.868 1.348 1.313 0.987 1.060 1.198 1.001 1.223 1.461 1.503 1.168 0.906 1.099 1.032 0.997 0.864 0.918 0.988 1.017 0.847 1.009 0.920 1.249 0.965 0.976 1.249 1.817 0.837 1.026
0.914 0.797 0.712 1.634 3.276 0.871 0.836 0.975 1.009 1.046 1.014 0.795 1.787 1.849 1.431 1.434 0.913 0.902 0.998 0.979 1.013 1.121 0.749 1.058 0.872 0.863 0.820 0.862 0.794 0.844 1.011 0.970 0.900 1.028 0.894 0.698 0.690 0.761 0.984 1.243 0.725 0.922
Mean
SD
p value for one-sample t-test (test value = 1)
0.838 0.944 1.058 1.989 3.981 1.037 0.969 1.006 1.158 0.861 1.261 1.114 1.915 1.715 1.339 1.459 0.838 0.951 0.968 0.961 1.228 1.196 0.931 1.063 0.985 0.887 0.856 0.963 0.870 0.813 0.908 0.933 0.918 0.939 0.880 0.827 0.890 0.874 1.082 1.474 0.817 0.982
0.268 0.128 0.392 0.438 1.070 0.158 0.262 0.040 0.250 0.271 0.656 0.538 0.121 0.249 0.096 0.161 0.198 0.094 0.246 0.051 0.218 0.237 0.507 0.103 0.167 0.201 0.161 0.089 0.079 0.123 0.159 0.107 0.082 0.137 0.048 0.375 0.175 0.108 0.146 0.303 0.084 0.054
0.404 0.530 0.822 0.060 0.040* 0.725 0.855 0.812 0.388 0.468 0.562 0.749 0.006** 0.038* 0.026* 0.038* 0.292 0.465 0.843 0.319 0.211 0.289 0.835 0.403 0.888 0.433 0.262 0.551 0.103 0.119 0.422 0.393 0.227 0.524 0.050 0.507 0.388 0.181 0.434 0.114 0.064 0.621
p < 0.05. p < 0.01.
system, we were able to detect the IB␣ phosphorylation. These results suggest that inhibition of NFB alone is not sufficient to completely block the inflammatory cytokine release from HL-CZ cells. 3.4. Zinc treatment induces activation of multiple immune response-related transcription factors in HL-CZ cells On the basis of our present results, it was suggested that, in addition to NFB, multiple immune response-related transcription factors might be involved in the zinc-induced cytokine release from HL-CZ cells. To verify this hypothesis, seven recombinant adenoviruses carrying a response element-driven luciferase reporter gene were established (Table 1). These immune response-related transcription factors include AP-1, C/EBP, CREBP, NFAT, NFB, SRF, and STAT [30–32]; the activated transcription factors mediate luciferase expression via binding to their respective response elements. HL-CZ cells were infected with one of the recombinant adenoviruses and then were treated with 150 M Zn(CH3 COO)2 for 6 h. As shown in Fig. 3, the zinc treatment induced activation of AP1, C/EBP, CREBP, NFAT, NFB, SRF, and STAT by 1.18, 2.90, 2.46, 1.64, 4.27, 1.42, and 1.42 folds, respectively, in HL-CZ cells. Among them,
C/EBP, CREBP, NFAT, NFB, and SRF were significantly activated by the zinc treatment. 4. Discussion Zinc is an essential trace element for animals and play important roles in regulation of immune function in humans [33]. However, excess zinc may also deregulate the homeostasis of immune system. Epidemiological studies revealed the association between elevated ambient air zinc and increased pediatric asthma morbidity [10]. Animal studies indicated that the ambient PM2.5 samples with higher levels of metals, such as zinc, caused increases in the allergic respiratory disease in mice [34]. Our previous in vitro evidence revealed an important role for ZnO particle, or its dissolved zinc ions, in modulating inflammatory responses of vascular endothelial cells [14,15]. The present study further demonstrates that zinc ions induce chemokine and inflammatory cytokine release from vascular promonocytes, possibly, via activating multiple immune response-related transcription factors. In the previous study, we evaluated vascular endothelial dysfunction by using ICAM-1 expression, an indicator for inflammatory response. ICAM-1, continuously present in low concentrations in
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Fig. 2. Inhibition of NFB activity by over-expression of IB␣ in HL-CZ cells does not block the conditioned medium-induced ICAM-1 protein expression in HUVEC cells. HL-CZ cells were left uninfected (−) or infected with a recombinant adenovirus (ADV), AdEasy-IB␣ (IB␣) or AdEasy-GFP (GFP). Cells were treated with 150 M of Zn(CH3 COO)2 for 6 h. Following treatments, both HL-CZ cell lysates and the conditioned media were collected. Then, HUVEC cells were treated with the conditioned media for 24 h and HUVEC cell lysates were collected. By using immunoblot analysis, IB␣ expression (IB␣) and phosphorylation (p-IB␣) in HL-CZ cells and protein levels of ICAM-1 and actin in HUVEC cells were determined.
Fig. 1. Conditioned medium (CM) from the zinc-treated HL-CZ cells induces ICAM1 protein expression in HUVEC cells. (A) HL-CZ cells were treated with different concentrations of Zn(CH3 COO)2 (0, 30, 50, 100, and 150 M) for 6 h and then the conditioned medium was collected. HUVEC cells were treated with the conditioned medium for 24 h. (B) HUVEC cells were also treated with the conditioned medium for 6, 12, and 24 h; the conditioned medium was collected from HL-CZ cells treated with 150 M Zn(CH3 COO)2 for 6 h. For C2 controls, HUVEC cells were treated with the conditioned medium for 24 h; the 6-h HL-CZ cultured medium with no zinc treatment were used as the conditioned medium. For C1 controls, HUVEC cells were cultured in M199 medium for 24 h. Following treatments, HUVEC cell lysates were analyzed for ICAM-1 and actin protein levels by immunoblot analysis. Representative immunoblots are shown (inserts). The protein levels of ICAM-1 and actin were quantified. The experiment was repeated three times. The data are expressed as relative ICAM-1 protein levels compared to that of the untreated control (C1) and are presented as means ± SD. Differences in ICAM-1 expression were examined by using Student’s t-tests with 2000 bootstrap samples (for A, 0 vs. C1 and the other CM-treated groups vs. 0; for B, the other CM-treated groups vs. C2). *p < 0.05.
the membranes of leukocytes and endothelial cells, is a ligand for LFA-1 (integrin), a receptor found on leukocytes [35]. Here, we showed that the conditioned medium from HL-CZ cells treated with 150 M Zn(CH3 COO)2 for 6 h caused a maximum level of ICAM1 induction (Fig. 1). Therefore, we collected these conditioned media for analysis of cytokine secretion using the RayBio Human Cytokine Antibody Array 3. The kit detects secreted cytokines in conditioned medium that presents a more accurate reflection of active cytokine levels. Cytokines or other proinflammatory mediators induce adhesion molecule expression and facilitate leukocyte attachment to vascular endothelium via ICAM-1/LFA-1 binding [36,37]. The adhesion of monocytes to the arterial wall and their subsequent infiltration and differentiation into macrophages is a crucial step in the development of atherosclerosis. Analysis of cytokines in conditioned media clearly showed that the zinc treatment induced chemokine and inflammatory cytokine release from HL-CZ cells. With the Cytokine Antibody Array, the GRO antibody detects CXCL1, CXCL2, and CXCL3; the
GRO-␣ antibody detects only CXCL1. The zinc treatment induced marked releases of GRO (CXCL1, CXCL2, and CXCL3) and GRO-␣ (CXCL1) by 1.99 and 3.98 folds, respectively (Table 2). Because GRO activity involves three CXC chemokines, it is suggested that GRO␣ could be the major CXCL chemokine secreted by HL-CZ cells. GRO-␣, initially isolated and characterized by its growth stimulatory activity on malignant melanoma cells, is a chemoattractant for neutrophils [38,39]. Recently, many new functions of GRO-␣ have been discovered and associated with atherosclerosis, angiogenesis, and many inflammatory conditions [40]. Moreover, the zinc treatment also induced significant releases of inflammatory cytokines, including IL-6, IL-7, IL-8, and IL-10. In addition to its known role in mediating the systemic acutephase response, IL-6 plays multiple roles in initiating and sustaining vascular inflammation [41]. IL-7 has many roles in T cells, dendritic cells, and bone biology in humans and is involved in chronic inflammation linking stroma and adaptive immunity [42]. IL-8, or CXCL8, is also a CXC chemokine and bears the primary responsibility for the recruitment of monocytes and neutrophils, the signature
Fig. 3. Zinc treatment induces activation of multiple immune response-related transcription factors in HL-CZ cells. HL-CZ cells were infected with one recombinant virus carrying a transcription factor-mediated luciferase reporter gene. The infected cells were treated with 150 M Zn(CH3 COO)2 for 6 h. Following the treatments, luciferase activity of each sample was determined. The experiment was repeated three times. The data are expressed as relative reporter activity as compared to that of the untreated control and are presented as means ± SD. Differences in reporter activity (the zinc-treated vs. the zinc-untreated) of each transcription factor were examined by using one-sample t-tests. *p < 0.05, **p < 0.01, and ***p < 0.001.
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cells of acute inflammatory response [43]. IL-10, an important cytokine with anti-inflammatory properties, is produced by activated immune cells, in particular monocytes/macrophages and T cell subsets including Tr1, Treg, and Th1 cells [44]. On the basis of this information, the present results demonstrate the potential impacts of zinc exposure on disturbing homeostasis of inflammation via these inflammatory cytokines. The major role of NFB in regulation of zinc-induced ICAM-1 expression in vascular endothelial cells has been demonstrated [14]. However, results in Fig. 2 suggest that, in addition to NFB, other immune response-related transcription factors may be also involved in zinc-induced chemokine/inflammatory cytokine release from HL-CZ cells. Indeed, among the seven immune response-related transcription factors tested, C/EBP, CREBP, NFAT, NFB, and SRF were significantly activated (Fig. 3). Therefore, the zinc-induced inflammatory responses in promonocytes involve a more sophisticated signaling regulation than that in vascular endothelial cells. Moreover, the 6-h zinc treatment conditions were used in cytokine analyses and immune response-related transcription factor-mediated luciferase reporter assay; this short-term treatment was designed to avoid the potential secondary inflammatory responses. However, we could not rule out the possibility in the meantime. Thus, deciphering the time-course activation of inflammation-related signaling molecules and transcription factors in detail is needed in the following study. The present study mainly dealt with the potential impacts on homeostasis of vascular immune system by those excess zinc exposures, especially from ambient particulate pollutants. In addition to vascular endothelial cells [14], here we further demonstrated that zinc induces chemokine and inflammatory cytokine release from human promonocytes. On the basis of this and our previous studies [14,15], the possible scenario of inflammatory responses induced by zinc from ambient particulate pollutants is described. First, fine particles tend to be trapped in the deeper pulmonary alveoli through inhalation. Second, in situ decomposition of the trapped particles results in a local increase of metal ions (such as zinc and nickel) and thus may activate pulmonary inflammation. Third, ultrafine particles and dissolved metal ions are able to cross the pulmonary epithelial barrier and then into the bloodstream. Finally, direct exposure of vascular cells, including endothelial cells and circulating blood cells, to those proinflammatory metals ions, such as zinc ions in this study, elicits vascular inflammation. These studies provide new insight for understanding the mechanisms of those inflammatory diseases induced by ambient particulate pollutants. 5. Conclusions Using human HL-CZ promonocytes as an in vitro system, this study reveals two important findings. Zinc treatment induces chemokine and inflammatory cytokine release from HL-CZ cells. The process involves activation of multiple immune responserelated transcription factors, including C/EBP, CREBP, NFAT, NFB, and SRF. Conflict of interest The authors declare that there are no conflicts of interest. Acknowledgements This work was supported by grants from the National Health Research Institutes (EO-099-PP-03, EO-100-PP-03) and the National Science Council (NSC97-2314-B-400–003-MY3) in Taiwan. We are grateful to Dr. Shu-Ching Hsu (Vaccine Research and
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Journal of Hazardous Materials 196 (2011) 342–349
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Preparation of nanocrystalline Fe3−x Lax O4 ferrite and their adsorption capability for Congo red Lixia Wang a,b , Jianchen Li a , Yingqi Wang a , Lijun Zhao a,∗ a b
Key Laboratory of Automobile Materials (Jilin University), Ministry of Education and School of Materials Science and Engineering, Jilin University, Changchun 130022, China School of Mechanical Science and Engineering, Northeast Petroleum University, Daqing 163318, China
a r t i c l e
i n f o
Article history: Received 2 June 2011 Received in revised form 8 September 2011 Accepted 9 September 2011 Available online 16 September 2011 Keywords: La3+ -doped magnetite Adsorption Desorption Wastewater treatment
a b s t r a c t This investigation was to increase the adsorption capacity of magnetite for Congo red (CR) by adulterating a small quantity of La3+ ions into it. The adsorption capability of nanocrystalline Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10) ferrite to remove CR from aqueous solution was evaluated carefully. Compared with undoped magnetite, the adsorption values were increased from 37.4 to 79.1 mg g−1 . The experimental results prove that it is effectual to increase the adsorption capacity of magnetite by doped La3+ ions. Among the La3+ -doped magnetite, Fe2.95 La0.05 O4 nanoparticles exhibit the highest saturation magnetization and the maximum adsorption capability. The desorption ability of La3+ -doped magnetite nanoparticles loaded by CR can reach 92% after the treatment of acetone. Furthermore, the Fe3−x Lax O4 nanoparticles exhibited a clearly ferromagnetic behavior under applied magnetic field, which allowed their high-efficient magnetic separation from wastewater. It is found that high magnetism facilitates to improve their adsorption capacity for the similar products. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Dyes and pigments are widely used as the coloring agents. Colored organic effluent is produced in industries such as textiles, paper, plastics, leather, food and cosmetic, etc. The total dye consumption in textile industry worldwide is more than 10,000 tonnes/year and approximately 100 tonnes of dyes were discharged into waste streams by the textile industry every year [1]. It was reported that nearly 40,000 dyes and pigments are listed, which consist of more than 7000 different chemical structures [2]. Such colored effluent can affect photosynthetic processes of aquatic plants, reducing oxygen levels in water and, in severe cases, resulting in the suffocation of aquatic flora and fauna [3]. Dye effluents are the pollutants that contain chemicals that exhibit toxic effect towards microbial populations and can be toxic and carcinogenic to organisms and human beings. Congo red (CR) (sodium salt of benzidinediazobis-1naphthylamine-4-sulfonic acid) is metabolized to benzidine, a known human carcinogen and exposure to this dye can cause some allergic responses [4]. The treatment of contaminated CR in wastewater is difficult because the dye is generally present in sodium salt form giving it very good water solubility. Due to their chemical structures, dyes resist fade when exposed to light, water and many chemicals and therefore it was difficult to be
∗ Corresponding author. Tel.: +86 431 85095878; fax: +86 431 85095876. E-mail address:
[email protected] (L. Zhao). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.032
decolorized once dyes were released into the aquatic environment. Synthetic dyes are difficult to biodegrade due to their complex aromatic structures, which provide them physico-chemical, thermal and optical stability. Also, the high stability of its structure makes it difficult to biodegrade and photodegrade [5]. To remove dyes and other colored contaminants from wastewaters, several physical, chemical, physico-chemical and biological methods (e.g., adsorption, coagulation-flocculation [6], biodegradation, ion-exchange, chemical oxidation [7], ozonation [8], nanofiltration [9], micellar enhanced ultrafiltration [10] and electrochemical methods have been developed. A number of adsorbents, such as activated carbon [11], orange peel [12], sawdust [13], montmorillonite [14], wheat bran and rice bran [15], and mesoporous Fe2 O3 [16], have been used for the removal of CR from aqueous solutions. But the adsorption capacity of these adsorbents is not large. Adsorbent-grade activated carbon is cost-prohibitive and both regeneration and disposal of the used carbon are often very difficult [17]. Widespread application of some of these adsorbents is restricted due to high cost, difficult disposal and regeneration. One of the new developments for removing dyes from water or wastewater in recent years is to use ferrite as adsorbents [18,19]. However, there are still some practical problems to be solved, such as the incompatible relation between the magnetic properties and the sizes. As far as we know, the decrease of the particle size will increase the surface disorder of nanoparticles. Thus, the surface energy will increase with the decreasing particle sizes. However, the saturation magnetization of magnetic powders is decreased with the decrease of the particle sizes, which is a disadvantage
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high-resolution transmission electron microscope (HRTEM, JEOL3010, 300 kV) and Energy dispersive X-ray spectrum (EDX, Oxford Instruments INCA Energy TEM 200, 300 kV) were used to characterized the microstructure. The hysteresis loops were measured on a VSM-7300 vibrating sample magnetometer (VSM) (Lakeshore, USA) in room temperature. IR spectra of the samples were characterized using a FTIR spectrophotometer (NEXUS, 670) in KBr pellets. A UV–vis spectrophotometer was used for determination of CR concentration in the solutions.
Scheme 1. Structure of Congo red molecule.
2.4. Adsorption experiments for the magnetic separation after the wastewater treatment. To solve this problem, in this contribution, the high surface activity is trying to be obtained by deforming the crystal structure which can be proved by the change of the lattice constant after substitution. Moreover, we have done an investigation on the relation between the adsorption activity and magnetic properties of Fe3−x Lax O4 nanoparticles. It is found that the higher magnetic properties facilitate to improve the adsorption capacity. 2. Materials and methods
The stock solution of CR (1 g L−1 ) was prepared in deionized water and desired concentrations of the dye were obtained by diluting the same with water. The calibration curve of CR was prepared by measuring the absorbance of different predetermined concentrations of the samples at max = 497 nm using UV–vis spectrophotometer (CR has a maximum absorbency at wavelength 497 nm on a UV–vis spectrophotometer). The amount of adsorbed CR (mg g−1 ) was calculated based on a mass balance equation as given below:
2.1. Adsorbate
qe =
Congo red [CR, chemical formula = C32 H22 N6 Na2 O6 S2 , FW = 696.68, max = 497 nm] is a benzidine-based anionic disazo dye, i.e., a dye with two azo groups. The structure is as illustrated in Scheme 1. An accurately weighed quantity of the dye was dissolved in double-distilled water to prepare stock solution (1 g L−1 ).
where qe is the equilibrium adsorption capacity per gram dry weight of the adsorbent, mg g−1 ; C0 is the initial concentration of CR in the solution, mg dm−3 ; Ce is the final or equilibrium concentration of CR in the solution, mg dm−3 ; V is the volume of the solution, dm3 ; and W is the dry weight of the hydrogel beads, g. Take one adsorption of CR for example. Standard solution with initial concentrations of 30 mg L−1 was prepared. Then, 15 mg of Fe3−x Lax O4 nanoparticles was added to 50 mL of the above solution under stirring. After a specified time, the solid and liquid were separated by magnet and UV–vis adsorption spectra was used to measure the CR concentration in the remaining solutions. A standard curve, which was used to convert absorbance data into concentrations for kinetic and equilibrium studies, was drawn to calculate the concentration of each experiment.
2.2. Synthesis of nanocrystalline Fe3−x Lax O4 ferrite In a typical experiment, FeSO4 ·7H2 O and LaCl3 ·7H2 O were dissolved in 20 mL of ethylene glycol (EG) by intensive stirring, accordingly a homogeneous solution was obtained, and then 1.5 g of NaOH was added to the solution at room temperature with simultaneous vigorous agitation. The mixtures were stirred vigorously for 30 min, and then sealed in a Teflon-lined stainless-steel autoclave and maintained at 200 ◦ C for 8 h. After the completion of the reaction, the solid product was collected by magnetic filtration and washed several times with deionized water and absolute ethanol respectively. The final product was dried in a vacuum oven at 100 ◦ C for 6 h. Black powders were obtained and characterized as Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10). Detailed experimental parameters are listed in Table 1 (from S1 to S4). Furthermore, the experimental works were carried out in winter, so the room temperature was lower about 13 ◦ C.
(C0 − Ce ) × V W
3. Results and discussion 3.1. Characterization of Fe3−x Lax O4 Fig. 1 shows the XRD patterns of S1 and S3. All the diffraction peaks in Fig. 1a can be indexed to the face-centered cubic structure of magnetite according to JCPDS card no. 19-0629, and the lattice ˚ The diffraction peaks of S3 show the constant of S1 is 8.40014 A. same structure with S1, and no impurities can be detected from
2.3. Characterization The phases were identified by means of X-ray diffraction (XRD) with a Rigaku D/max 2500pc X-ray diffractometer with Cu K␣ radiation () 1.54156 (Å) at a scan rate of 0.02◦ /1(s), morphologies were characterized by a JEOL JSM-6700F field emission scanning electron microscopy (FESEM) operated at an acceleration voltage of 8.0 kV. Transmission electron microscope (TEM, Philips Tecnai 20, 200 kV), Table 1 Summary of the experimental parameters. Samples
FeSO4 ·7H2 O (g)
S1 (Fe3 O4 ) S2 (Fe2.99 La0.01 O4 ) S3 (Fe2.95 La0.05 O4 ) S4 (Fe2.90 La0.10 O4 )
0.8341 0.8313 0.8202 0.8063
± ± ± ±
0.0002 0.0002 0.0002 0.0002
LaCl3 ·7H2 O (g) 0.0000 0.0037 0.0185 0.0371
± ± ± ±
0.0002 0.0002 0.0002 0.0002
NaOH (g) 1.5000 1.5000 1.5000 1.5000
± ± ± ±
0.0002 0.0002 0.0002 0.0002
(1)
Fig. 1. XRD patterns of: (a) S1 and (b) S3.
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adsorption capacity (mg g-1 )
80
d c
60
b 40
a
20
0 0
20
40
60
80
100
time (min) Fig. 2. Adsorption capacity of: (a) S1; (b) S2; (c) S4 and (d) S3. (Adsorption conditions for CR: 50 mL of 100 mg L−1 of dye, adsorbent dosage 0.015 g, natural pH, temperature: 13 ◦ C.)
˚ The Fig. 1b. Furthermore, the lattice constant of S3 is 8.39244 A. incorporation of La ions may reside on the boundaries of the magnetite, which make the shortening of Fe–O bond length, therefore, the lattice constant of S3 is smaller than S1. Similar results have been reported in work done by Zhao and El-Bahy [20,21]. The strong and sharp peaks indicate that the S1 and S3 are well crystallized. By comparison of the lattice constants between S1 and S3, we can confirm that the crystal structure of magnetite was deformed little by the doping of La3+ ions. 3.2. Effect of La3+ -doped amount on the adsorption capacity of magnetite After we ascertain the pure phase Fe3−x Lax O4 ferrites, a series adsorption experiments were carried out. The adsorption capacity of S1 to S4 for CR was shown in Fig. 2. Their adsorption values
for CR are 37.4,48.6,79.1 and 63.1 mg g−1 , respectively. An exciting experimental result is obtained that the doping of La3+ ions favors increasing the adsorption capacity of magnetite for CR. Especially, S3 exhibits the maximum adsorption capacity. Furthermore, at the beginning of the contact time about 5 min, a rapid removal of CR was observed. After 90 min, the adsorption for CR almost reaches saturation. In order to study the effect of morphologies or particle sizes of the Fe3−x Lax O4 on the adsorption capacity of CR in the aqueous solution, SEM photos were shown in Fig. 3a. S1 is composed by octahedral nanoparticles with edge length about 10–30 nm. Uniform nanoparticles with particle sizes around 20 nm are observed from Fig. 3b. However, irregular shapes and broad size distribution are appeared with the increasing doped contents of La ions for S3 and S4 (Fig. 3c and d). Their particle sizes are in the range of 60–200 nm and 80–300 nm, respectively. To our best knowledge, the adsorption capacity of nanopowders increased with the decrease of particle sizes (namely the increase of surface areas). However, in this experiment, it is found that the adsorption capacity could be improved by doped La3+ ions, without accompanying the decrease of the particle sizes. Further insight into the nanostructure of samples was gained using TEM and HRTEM. Take S3 as example. Fig. 4a shows the TEM image of S3, which are in agreement with the above SEM findings. The HRTEM image (Fig. 4b) and the corresponding fastFourier-transform (FFT) pattern-which is framed in Fig. 4b with a square-is shown in Fig. 4c, it represents a face-centered cubic diffraction spots pattern. The clear lattice fringes can prove the high crystallinity of the as-prepared S3. Further, the dominantly exposed planes of S3 are {1 1 1}. The lattice spacing between two adjacent fringes we can observe is corresponding to the set of (1 1 1) planes with a lattice space of 0.5 nm. The lattice fringes are parallel throughout, which prove the single-crystalline nature of S3. Besides, EDX analysis (Fig. 4d) exhibited that S3 was essentially composed of Fe, La and O elements, indicating that La ions was introduced into the magnetite.
Fig. 3. SEM images of (a) S1, (b) S2, (c) S3 and (d) S4.
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Fig. 4. (a) SEM image, (b) HRTEM image, (c) FFT pattern framed in (b), and (d) EDX pattern of S3.
By comparing the S1 with other samples, it is certain that the proper doped La3+ ions can effectively increase the adsorption capacity. Among the S2, S3 and S4, it can be concluded that the concentration of the doped La3+ ions and the particle sizes exhibit a combination influence on the adsorption capacity of magnetite. The proper doped amount and particle sizes lead to the maximum adsorption capacity of magnetite. The ionic radius (r) of La3+ is ˚ rFe 3+ = 0.64 Å). Such much larger than that of Fe3+ (rLa 3+ = 1.06 A, a substitution makes the change of lattice constant. Compared S1 with S3, the lattice constant of S3 is smaller than that of S1. The lattice constant is corresponding to the distortion height of octahedral site ([MeO6 ]) for magnetite with face-center cubic structure. The decrease of lattice constant makes the increase of distortion degree, hence, the magnetite substituted by La3+ ions is more unstable. Meanwhile, the dopant would lead the imperfect coordination and produce surface defects of magnetite. Finally, the unstable state might lead to the increase of surface energy. To decrease the surface energy of magnetite, it is prone to the adsorption of CR on its surface. Moreover, the structural mismatch caused by doped La3+ ions should change the surface charge of Fe3 O4 [23], which may also conduce to the adsorption of CR on the surface based on the electrostatic adherence principle. 3.4. Effect of initial dye concentrations and contact time on adsorption In order to know the effect of initial dye concentration and contact time on the removal of CR, four different concentrations (0.030, 0.050, 0.080, and 0.100 g L−1 ) are selected to investigate the adsorption of CR on the surface of Fe2.95 La0.05 O4 . With the increase of initial CR concentrations from 0.030 to 0.100 g L−1 , the amount of
CR removal was increased from 29.2 to 79.11 mg g−1 as shown in Fig. 5. Very rapid adsorption is observed at previous 2–5 min, and thereafter a gradual increase occurs with increasing contact time up to 20–30 min depending on the initial dye concentration. Then, the adsorption keeps a weak increase during the following time. Therefore, the adsorption equilibrium almost happen at 40 min. Similar results have been reported for the adsorption of CR on calcium-rich fly ash in work done by Acemio˘glu [24]. 3.5. Adsorption isotherms Analysis of adsorption isotherm is of fundamental importance to describe how adsorbate molecules interact with the adsorbent surface. To simulate the adsorption isotherm, two commonly used 80
adsorption capacity (mg g-1 )
3.3. Adsorption mechanism
60
40
100mg/l 80mg/l 50mg/l 30mg/l
20
0
0
20
40
60
80
100
time (min) Fig. 5. Effect of initial dye concentration on CR removal from S3. (Conditions: 50 mL of CR, adsorbent dosage 0.015 g, natural pH, temperature: 13 ◦ C.)
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Table 2 Adsorption parameter obtained from adsorption isotherm for S3. Freundich
Langmuir
KF
n
rF 2
qmax (mg g−1 )
KL
rL 2
RL
9.8084 ± 0.0002
2.14 ± 0.01
0.9898 ± 0.0002
107.64 ± 0.01
0.0279 ± 0.002
0.9976 ± 0.0001
0.5444 0.4175 0.3094 0.2639
models, the Freundlich [25] and Langmuir [26] isotherms, were selected to explicate dye–ferrite interaction. The Freundlich adsorption isotherm can be expressed as: log qe = log KF +
1 log Ce n
where KF and n are the Freundlich adsorption isotherm constants, being indicative of the extent of the adsorption and the degree of nonlinearity between solution concentration and adsorption, respectively. KF and 1/n values can be calculated from intercept and slope of the linear plot between log Ce and log qe . The Langmuir isotherm is expressed as:
log(q1 − qt ) = log q1e −
where KL and C0 are the same as defined before. The value of RL calculated from the above expression. The nature of the adsorption process to be either unfavorable (RL > 1), linear (RL = 1), favorable (0 < RL < 1) or irreversible (RL = 0). The Freundlich isotherm was employed to describe heterogeneous systems and reversible adsorption, which does not restrict to the monolayer formations. Unlike the Freundlich isotherm, the Langmuir isotherm is based on the assumption that a structure of adsorbent is homogeneous, where all sorption sites are identical and energetically equivalent. Fig. 6 represents the plot of the experimental data based on Freundlich and Langmuir isotherms model, respectively. Table 2 shows the calculated values of Freundlich and Langmuir model’s parameters. The comparison of correlation coefficients (r2 ) of the linearized
(7)
b y=0.3334x+0.0093
0.025
-1
1/qe (g mg )
logqe(mgg-1)
(6)
The h, q2e and K2 can be obtained by linear plot of t/qt versus t. Fig. 7 is the plots of the pseudo-first order and second order kinetics of CR adsorption on S3. The calculated kinetic parameters are given in Table 3. The correlation coefficient for the pseudo-first-order model is relatively lower (r1 2 = 0.8527), the calculated qe value (q1e ) obtained
2
r =0.9976
0.020
1.6
0.015
1.5 1.2 1.3 1.4 1.5 1.6 1.7 1.8 1.9 2.0
0.010 0.01
logCe(mgL-1)
(5)
h = K2 q2e 2
0.030
1.7
0.1 0.1 0.1 0.1
where K2 is the pseudo-second order rate constant (g mg−1 min−1 ). The initial adsorption rate, h (mg g−1 min−1 ) at t → 0 is defined as:
a
1.8
K1 t 2.303
t 1 t = + qt q2e K2 q2e 2
2.0 1.9
± ± ± ±
where qt is the amount of dye adsorbed per unit of adsorbent (mg g−1 ) at time t, K1 is the pseudo-first order rate constant (min−1 ). The adsorption rate constant (K1 ) were calculated from the plot of log(q1e − qt ) against t. Ho and McKay [28] presented the pseudo-second order kinetic as:
(4)
y=0.4668x+0.9916 r2=0.9898
30.0 50.0 80.0 100.0
The adsorption kinetic models were applied to interpret the experimental data to determine the controlling mechanism of dye adsorptions from aqueous solution. Here, Pseudo-first-order, pseudo-second-order and the intraparticle diffusion model were used to test dynamical experimental data. The pseudo-first order kinetic model of Lagergren [27] is given by:
(3)
1 (1 + KL C0 )
0.0002 0.0002 0.0002 0.0002
3.6. Adsorption kinetics
where qmax is the maximum amount of adsorption with complete monolayer coverage on the adsorbent surface (mg g−1 ), and KL is the Langmuir constant related to the energy of adsorption (L mg−1 ). The Langmuir constants KL and qmax can be determined from the linear plot of 1/Ce versus 1/qe . The essential characteristics of Langmuir isotherm can be expressed by a dimensionless constant called equilibrium parameter RL that is defined by the following equation: RL =
± ± ± ±
form of both equations indicates that the Langmuir model yields a better fit for the experimental equilibrium adsorption data than the Freundlich model. This suggests the monolayer coverage of the surface of S3 by CR molecules. The maximum adsorption capacity (qmax ) of the S3 beads for CR was 107.64 mg g−1 (Table 2). Here, RL -values obtained are listed in Table 2. All the RL -values for the adsorption of CR onto S3 are in the range of 0.5444–0.2639, indicating that the adsorption process is favorable.
(2)
1 1 1 1 = + qe qmax KL qmax Ce
C0 (mg L−1 )
0.02
0.03
0.04
0.05
0.06
-1
1/ce (L mg )
Fig. 6. Adsorption isotherms for adsorption of CR on S3 (15 mg of adsorbent) (a) Freundlich and (b) Langmuir.
L. Wang et al. / Journal of Hazardous Materials 196 (2011) 342–349
347
Fig. 7. Adsorption kinetic for adsorption of CR on S3 (15 mg of adsorbent, initial dye concentration 100 mg L−1 , natural pH, test-temperature: 13 ◦ C) (a) pseudo-first order and (b) pseudo-second order. Inset in (b) in turn is CR solution, mixing with the magnetic absorbents and separation of the adsorbent from solution with a magnet after reaction (2) and 30 min, respectively.
Table 3 Adsorption parameters obtained from Fig. 5.
79.11 ± 0.01
Pseudo-first-order
Pseudo-second-order
K1 (min−1 )
q1e (mg g−1 )
r1 2
K2 (g mg−1 min−1 )
q2e (mg g−1 )
h (mg g−1 min−1 )
r2 2
0.0246 ± 0.0002
12.22 ± 0.01
0.8527 ± 0.0002
0.0085 ± 0.0002
79.43 ± 0.01
53.62 ± 0.01
0.9994 ± 0.0001
from this equation does not give reasonable value (Table 3), which is much lower than experimental data (qe,exp ). This result suggests that the adsorption process does not follow the pseudo-first-order kinetic model, which is similar to the result reported for adsorption of CR onto Australian clay materials [29]. On the contrary, the results present an ideal fit to the second order kinetic for adsorbent with the extremely high r2 2 = 0.9994 (Fig. 7b). A good agreement with this adsorption model is confirmed by the similar values of calculated q2e and the experimental ones for adsorbent. The best fit to the pseudo-second order kinetics indicates that the adsorption mechanism depends on the adsorbate and adsorbent. CR is an acidic dye with negative charge because of the existence of sulphonated group (–SO3 − Na+ ). Here, the higher adsorption capacity of the CR for S3 is probably because of the dopant of La ions, which may increase the surface positive charges of magnetite, so we speculate that an electrostatic attraction may the main adsorption mechanism. Inset in Fig. 7b represented the photograph of adsorption and magnetic separation behavior. A light pink solution was observed after 2 min of adsorption. Further prolonging the adsorption time to 30 min, a colorless solution was gained. More importantly, simple and rapid separation of CR-loaded magnetite adsorbent from treated water can be achieved via an external magnetic field.
at 1050 cm−1 corresponding to C–N bond only appears in Fig. 9b, which discloses that CR was loaded on the surface of S3. This also serves as another evidence of physical adsorption because of in a physical adsorption at the mineral–water interface; an oxyanion 100 80
Desorption(%)
qe,exp (mg g−1 )
60 40 20 0 0
20
40
60
80
100
time(min) Fig. 8. Desorption ratio of loaded magnetite nanoparticles with time.
3.7. Desorption Desorption is also a key role for the practical application of magnetic powders to water treatment. A facile desorption method and high-efficient desorption can facilitate to reduce cost, because the spent adsorbent and the CR can obtain recycling chance. Desorption process was conducted by mixing 5 mg of CR-loaded modified S3 with 30 mL of acetone solutions and shaking for different time. Fig. 8 is the desorption percentage with the time. The desorption efficiency calculated as Eq. (8) was 92%. Therefore, the CR could be desorbed from the loaded nanoparticles by acetone solutions. Desorption ratio (%) =
Amount of desorbed CR × 100 Amount of adsorbed CR
(8)
FT-IR analysis was also performed to reveal the surface nature of S3, as shown in Fig. 9. The spectra display a broad band at 580 cm−1 , which is believed to be associated with the stretching vibrations of the tetrahedral groups (Fe3+ –O2− ) for S3. However, the band
Fig. 9. FT-IR spectra of (a) as-prepared, (b) CR-absorbed and (c) CR-desorbed S3.
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Table 4 Adsorption capacities of CR dye on various adsorbents.
S3 [3] [15] [15] [16] [29] [30] [31] [32] [32] [32] [33] [34] [35] [36] [37] [38] Present study
will retain its hydration shell and will not form a direct chemical bond with the oxide surface [22]. Moreover, it is an evident proof that CR is removed sufficiently from the surface of S3 by acetone, because Fig. 9a and c exhibit the same FT-IR spectra. 3.8. Performance evaluation The maximum adsorption capacity (qmax ) for S3 nanoparticles to CR calculated from the Langmuir isotherm model is listed in Table 4 with literature values of qmax of other adsorbents for CR adsorption [3,15,16,29–38]. All of the adsorbents used for CR adsorption have considerably lower qmax values than S3 used in this study, except chitosan hydrogel beads impregnated with carbon nanotubes CS/CNT [3], maghemite nanoparticles [16] and CTAB modified chitosan beads [37,38]. However, the simplicity of the preparation method and magnetic separation of S3 nanoparticles makes them better adsorbent than the others for CR adsorption. 3.9. Magnetic properties The magnetic properties of magnetic absorbents directly influence the callback efficiency. Hence, excellent magnetic performance is also a key role for the magnetic material as magnetic absorbent. Here, the magnetism of S1 to S4 is evaluated. It is exciting to find that the magnetic properties of magnetic absorbents do have influence on their adsorption ability. The room-temperature hysteresis loops of S1 to S4 were shown in Fig. 10. Furthermore, the magnetic parameters of samples obtained from hysteresis loops were listed in Table 5. The test results show that the doped-La3+ ions make the Ms values of magnetite decreased to some extent. It must acknowledge that magnetite still keep the high Ms values after La3+ ions were doped into them, even if the lowest Ms value is 81.4 emu/g for S2. We amazingly find that the adsorption abilities of La3+ -doped magnetite are proportional to their Ms values and independent of their particle sizes. This is an important proof that excellent magnetism facilitates to increase the adsorption capacity for the similar mag-
S4
40
S1
0 Magnetization (emu/g)
Chitosan hydrogel beads impregnated with carbon nanotubesCS/CNT 450.40 Wheat bran 22.73 Rice bran 14.63 208.33 Maghemite nanoparticles Cattail root 38.79 4.43 Sugar cane bagasse 35.70 Jute stick powder Bentonite 19.90 Kaolin 5.60 Zeolite 4.30 66.23 Palm kernel seed coat 7.08 Activated red mud 22.62 Anilinepropylsilica xerogel 71.46 Marine alga 352.50 CTAB modified chitosan beads 373.29 CS/CTAB beads 107.64 Fe2.95 La0.05 O4
S2
80
Reference
Magnetization (emu/g)
qmax (mg g−1 )
Type of adsorbent
-40 -80 -10000
-5000
40 20 0 -20 -40 -200
0
-100 0 100 Field (Oe)
5000
200
10000
Field (Oe) Fig. 10. Magnetization curve measured at room temperature for the Fe3−x Lax O4 ferrites.
netic products. Both the heavy-metal ions and organic matters show feeble paramagnetism or antiferromagnetism, so the magnetic powders which possess high Ms will be beneficial to the adsorption of heavy-metal ions and organic matters. Moreover, the magnetic materials with high Ms are help to the finally magnetic separation. Therefore, it is very meaningful to both keep the magnetic properties almost constant and increase the adsorption capacity. 4. Conclusions Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10) ferrite nanoparticles were successfully synthesized by a facile one-step solvothermal synthesis. Compared with the pure magnetite, La3+ -doped magnetite exhibit more excellent adsorption ability. Furthermore, among the La3+ -doped products, the sample (Fe2.95 La0.05 O4 ) possessing the biggest Ms value owes the strongest adsorption capacity. The adsorption capacity of magnetite for CR is improved not by increasing the specific area but by deforming the crystal structure via doped- La3+ ions. By comparison with many other adsorbents, Fe3−x Lax O4 nanoparticles have higher adsorption capacities for CR. It would be a good method to increase adsorption efficiency of magnetite for the CR removal in a wastewater treatment process by doping La3+ ions. Analysis of adsorption isotherm shows that our adsorption experiment accord with Langmuir model. Again, adsorption kinetic model indicates that the adsorption mechanism depends on the adsorbate and adsorbent. In a word, the Fe3−x Lax O4 nanoparticles were a kind of excellent absorbent because of their high adsorption, desorption and recovery efficiency. Acknowledgements The financial supports from the Natural Science Foundation of Jilin Province (20101542) of China and the National Foundation of Doctoral Station (grant No. 20100061110019) are acknowledged. References
Table 5 Magnetic parameters obtained from hysteresis loops. Samples
Ms (emu/g)
S1 S2 S3 S4
90.5 81.4 86.2 82.2
± ± ± ±
0.1 0.1 0.1 0.1
Mr (emu/g)
Hc (Oe)
± ± ± ±
156.6 116.5 115.9 105.9
15.6 11.2 14.8 16.4
0.1 0.1 0.1 0.1
± ± ± ±
0.1 0.1 0.1 0.1
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Journal of Hazardous Materials 196 (2011) 350–359
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Organo/layered double hydroxide nanohybrids used to remove non ionic pesticides D. Chaara a,1 , F. Bruna b , M.A. Ulibarri a , K. Draoui c , C. Barriga a,∗ , I. Pavlovic a a Dpto de Química Inorgánica e Ingeniería Química, Instituto Universitario de Química Fina y Nanoquímica (IUQFN),Universidad de Córdoba, Campus de Rabanales, Campus de Excelencia Internacional Agroalimentario, Ceia3, Edificio Marie Curie, 14071 Córdoba, Spain b Instituto de Recursos Naturales y Agrobiología de Sevilla (IRNAS), CSIC, Avenida Reina Mercedes 10, Apartado 1052, 41080, Sevilla, Spain c Department de Chimie, Laboratoire LPCIE, Faculté des Sciences, BP 2121, Tetouan, Morocco
a r t i c l e
i n f o
Article history: Received 15 July 2011 Received in revised form 7 September 2011 Accepted 9 September 2011 Available online 16 September 2011 Keywords: Organo/layered double hydroxide Nanohybrid Pesticide Adsorption Controlled release
a b s t r a c t The preparation and characterization of organo/layered double hydroxide nanohybrids with dodecylsulfate and sebacate as interlayer anion were studied in detail. The aim of the modification of the layered double hydroxides (LDHs) was to change the hydrophilic character of the interlayer to hydrophobic to improve the ability of the nanohybrids to adsorb non-ionic pesticides such as alachlor and metolachlor from water. Adsorption tests were conducted on organo/LDHs using variable pH values, contact times and initial pesticide concentrations (adsorption isotherms) in order to identify the optimum conditions for the intended purpose. Adsorbents and adsorption products were characterized several physicochemical techniques. The adsorption test showed that a noticeable increase of the adsorption of the non-ionic herbicides was produced. Based on the results, the organo/LDHs could be good adsorbents to remove alachlor and metolachlor from water. Different organo/LDHs complexes were prepared by a mechanical mixture and by adsorption. The results show that HTSEB-based complex displays controlled release properties that reduce metolachlor leaching in soil columns compared to a technical product and the other formulations. The release was dependent on the nature of the adsorbent used to prepare the complexes. Thus, it can be concluded that organo/LDHs might act as suitable supports for the design of pesticide slow release formulations with the aim of reducing the adverse effects derived from rapid transport losses of the chemical once applied to soils. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Agricultural pesticides are often detected in natural waters, and this has raised concerns regarding the protection of health and the environment. They are an important group of organic pollutants which production and uses are still increasing but must be controlled to minimize contamination problems [1,2] The research of new adsorbent is a strategy to remediate the contamination of water produced by an increase in the use of pesticides to improve agriculture production. Layered double hydroxides (LDHs), also known as hydrotalcitelike compounds, consist of brucite-like layers which contain the hydroxides of divalent (MII ) and trivalent (MIII ) metal ions and have an overall positive charge balanced by hydrated anion
∗ Corresponding author. Tel.: +34 957 218648; fax: +34 957 218621. E-mail address:
[email protected] (C. Barriga). 1 Permanent address: Department de Chimie, Laboratoire LPCIE, Faculté des Sciences, BP 2121, Tetouan, Morocco. 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.034
between layers. These compounds present a general formula [MII (1−x) MIII x (OH)2 ]x+ (An− )x/n ·mH2 O where An− is the intercalated anion. Because of the strong hydration of these inorganic ions the interlayer spaces have hydrophilic nature they resemble clay minerals. As a result, the natural clay minerals as well as LDHs show rather weak affinity to most of the non ionic organic compounds and are seldom by use as sorbents for organic compounds [3,4] Under suitable conditions, the inorganic ions on clay minerals and layered double hydroxides can be replaced by organic ions which make the interlayer spaces become hydrophobic, [5–9]. Therefore, the adsorption ability of clay minerals and LDHs for organic pollutants can be significantly improved by modifying the interlayer. These organoclays and organo/LDHs have applications in a wide range of organic pollution control fields [10–14] To date, there are few studies of LDHs modified with organic anions and in particular of their adsorption properties. The number of publications of organoclay (more than 2000) is much higher than organo/LDH (approx. 50) over the last 10 years (source: Citation Report of Web of Science in June 2011). However, recently the scientific interest in organo/LDH is increasing. This gives rise
D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359
to an emerging research field of the hydrotalcites since the results obtained regarding its application as adsorbents of pesticides are very promising. The organo/LDHs with suitable anion can provide interlayer spaces between 2 and 4 nm and might be considered as nanohybrids with hydrophobic characteristics in the interlayer space and external surface. The aim of this work was to prepare different nanohybrid intercalated hydrotalcites with two organic anions: dodecylsulfate (DDS) and sebacate (SEB) and with different ratio Mg/Al = 3 and 2, to increase the charge density of the layers and consequently to increase the content of interlayer sebacate anions. These nanohybrids were used to assess the removal of alachlor and metolachlor from water, two widely used herbicides with hydrophobic characteristic. Additionally, the metolachlor has been selected to explore the release behavior from the formulation with organo/LDH nanohybrids. 2. Materials and methods 2.1. Organic anions, pesticides and soil The organic anions used for the preparation of the organo/LDH nanohybrids containing DDS and SEB, were supplied as soluble sodium salt and acid form, respectively, by Sigma–Aldrich. The log Kow values are 5.4 and 1.86 for dodecylsulfate and sebacate acid respectively. (2-chloro-2 ,6 -diethyl-N-(methoxymethyl)) Alachlor and metolachlor (2-chloro-N-(6-ethyl-o-tolyl)-N-[(1RS)-2-methoxy-1methylethyl]acetamide) are selective pre-emergence herbicides which belong to aniline herbicides. Analytical standard alachlor and metolachlor was purchased from Sigma–Aldrich. The water solubility values of alachlor and metolachlor at 25 ◦ C are 0.110 g/L and 0.120 g/L and their log Kow values are 2.9 and 3.45 respectively (data obtained from Scifinder Scholar). The molecular structures of the herbicides and organic anions used are shown in Fig. 1. The soil used in the leaching experiments was fluvisol from the terraces of the Guadalquivir River, Córdoba (Spain). The soil was sampled (0–20 cm), air-dried, and sieved (2 mm) prior to use. It had 660 mg/kg sand, 150 mg/kg silt, 190 mg/kg clay, and 3.6 mg/kg organic matter. Soil pH was 8.7 in a 1:2 (w:w) soil:deionized water mixture.
351
for alachlor and metolachlor respectively. The amount of herbicide adsorbed (Cs ) was calculated from the difference between the initial (C0 ) and equilibrium (Ce ) solution concentrations. Desorption was realized immediately after adsorption from the highest equilibrium point of the adsorption isotherm and was repeated three times. Adsorption–desorption data were fitted to the Langmuir equation: Ce = Cs
C 1 e Cm
+
Cm L
(1)
and the logarithmic form of Freundlich equation: logCs = logKf + nf logCe
(2)
where Cm is the maximum adsorption capacity at the monolayer coverage (mmol/g), L (L/mmol) is a constant related to the adsorption energy and Kf (mmol 1−nf Lnf g−1 ) and nf are the Freundlich constants. 2.4. Characterization of the adsorbents and adsorption products The adsorbents HTSDS1, HTSDS2 and HTSEB and the adsorption products were characterized by different physical chemical techniques. Powder X-ray diffraction (PXRD) patterns were recorded on powder samples at room temperature under air conditions, using a Siemens D-5000 instrument with Cu K␣ radiation. FT-IR spectra were recorded by using the KBr disk method on a Perkin Elmer Spectrum One spectrophotometer and the ATR-FT-IR method was used for alachlor. Elemental chemical analyses for Mg and Al were carried out by atomic absorption spectrometry on a Perkin Elmer AA-3100 instrument. DDS and SEB amounts were calculated from elemental analysis of S and C respectively, carried out on Elemental Analyse Eurovector EA 3000 instrument. The interlayer water amount was obtained from TG-curves recorded on a Setaram Setsys Evolution 16/18 apparatus, in air at the heating rate of 5 ◦ C/min. Scanning electron microscopy (SEM) micrographs were obtained using a JEOL JSM 6300 instrument; the samples were prepared by deposition of a drop of sample suspension on a Cu sample holder and covered with an Au layer by sputtering in a Baltec SCD005 apparatus.
2.2. Synthesis of the organo/LDH nanohybrids
2.5. Preparation of organo/LDH– metolachlor complexes
The organo/LDHs containing DDS and SEB anions and Mg/Al = 3 and 2 respectively, were obtained by the coprecipitation method [15] using N2 atmosphere and CO2 -free water. The samples were figured as HTDDS1 and HTSEB. For comparison purposes another organo/LDH with DDS was obtained under the same experimental conditions but without washing the precipitate. After removing the supernatant the tube containing the solid was dried. This organo/LDH was named HTDDS2. A carbonate–Mg/Al hydrotalcite (HTCO3 ) was also prepared by the coprecipitation method [16] for the same purpose.
Four complexes were prepared with HTDDS1 and HTSEB and metolachlor. Two of them were based on the adsorption isotherms, figured as HTDDS1–MetoAds and HTSEB–MetoAds , and loaded with a 3% of the herbicides. The other two were prepared by mechanical mixture of the components by soft grinding of the adsorbents and the herbicides (in the same proportion as the adsorption complexes) dissolved in acetone and then were let the solvent evaporate. These figured as HTDDS1–MetoM and HTSEB–MetoM . 2.6. Bath release experiments
2.3. Adsorption and desorption experiments Alachlor and metolachlor adsorption isotherms on HTDDS1, HTDDS2 and HTSEB were obtained by the batch equilibration procedure. Triplicate 20 mg adsorbent samples were equilibrated through shaking for 24 h at room temperature with 30 mL of herbicide solutions with initial herbicide concentrations (C0 ) ranging between 0.1 and 0.35 mmol/L. After equilibration, the supernatants were centrifuged and separated to determine the concentration of herbicides by UV–visible spectrophotometry at 265 nm and 220 nm
The release of metolachlor into water from the organo/LDH–herbicide complexes was compared with the release of the herbicides as a free (technical) product. For this purpose, 0.18 mg of metolachlor as organo/LDHs–herbicide complexes or as a technical product was added to 500 mL of distilled water. The experiments were conducted as described by Bruna et al. [17]. The herbicide concentration was determined by HPLC using a Waters 1525 chromatograph coupled to a Waters 2996 diode-array detector and UV detection at 220 nm for metolachlor.
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Fig. 1. The molecular structure of organic anions and herbicides.
2.7. Soil-column leaching experiments Leaching experiments were conducted as described by Bruna et al. [17]. The calculated pore volume of the columns after saturation was 68 ± 2 mL. The leaching experiment was conducted in triplicate and the concentrations of metolachlor were analysed by HPLC. 3. Results and discussion
spacing d(0 0 3) for the organo/LDH adsorbents are also included in Table 1. As expected, the positions of the basal reflections of all the prepared organo/LDHs are shifted to smaller 2 reflection angles regarding carbonate hydrotalcite, which reveals expansion in the interlayer distances, moreover several harmonic are also observed indicating well ordered structure. The van der Waals end-to-end length of dodecylsulfate anion estimated was 2.08 nm taking into account that the LDH layer thickness is 0.48 nm so the dodecylsulfate chains can fit perfectly in the perpendicular direction and
3.1. Characterization of organo/LDH nanohybrid sorbents 3.1.1. Elemental analysis The results of elemental analysis for the nanohybrides used as adsorbents are shown in Table 1 together with other characteristics. The organic anion was determined from the ratio S/Al for the samples HTDDS1 and HTDDS2, and from C/Al for the HTSEB sample. The content of S indicates that the layer charge in HTDDS1 is not balanced only by the organic anion. The exchange percentage in the product was 92% based upon anion exchange capacity (AEC). However, in HTDDS2 it was 100% and the amount of S was higher than required to compensate the layer charge suggesting the precipitation of sodium dodecylsulfate, as it is indicated in Table 1. The elemental analysis of C for the HTSEB sample indicated a small excess of sebacate to compensate the layer charge which could be considered as precipitated salt as has been shown in Table 1. The proposed formulae were obtained from the elemental analysis, assuming that all the positive charge is compensated by the maximum amount possible of dodecylsulfate and sebacate anions for HTDDS and HTSEB respectively. The amount of water was attained from TG data (not included) and metal content analysis. 3.1.2. X-ray diffraction The PXRD patterns of the organo/LDHs included in Fig. 2 together with HTCO3 show that they are typical hydrotalcitelike compounds. The reflections for HTCO3 were indexed on the basis of a hexagonal unit cell with a and c parameters 0.304 nm and 2.34 nm respectively. The corresponding values of the basal
Fig. 2. PXRD patterns for the samples HTDDS1, HTDDS2, HTSEB and HTCO3 .
D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359
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Table 1 Chemical composition of adsorbents, structural data and proposed formulae. Sample
HTDDS1 HTDDS2 HTSEB HTCO3
wt.%
Atomic ratio
Mg
Al
N
S
C
Mg/Al
S/Al
C/Al
11.1 9.0 13.5
3.9 3.3 18.1
– 0.35 – –
4.2 5.9 –‘ –
27.1 31.3 18.9
3.2 3.1 1.9 2.7
0.9 1.5 – –
15.7 21.7 5.3 –
d0 0 3 (nm)
d1 1 0 (nm)
Proposed formulae
2.58 3.84 1.48 0.78
0.152 0.152 0.150 0.152
[Mg0.76 Al0.24 (OH)2 ](C12 H25 SO4 )0.22 (CO3 )0.01 ·0.83H2 O [Mg0.76 Al0.24 (OH)2 ](C12 H25 SO4 )0.24 ·0.79H2 Oa [Mg0.65 Al0.35 (OH)2 ](C10 H16 O4 )0.175 ·1.36H2 Ob [Mg0.73 Al0.27 (OH)2 ](CO3 )0.135 ·0.64H2 O
The excess of S corresponds to a 0.122NaC12 H25 SO4 per FW and the wt.% of N to 0.05NaNO3 . b The same for the sample HTSEB.
the chains could be in an all-trans conformation [6] in the HTDDS1 sample with a basal spacing of 2.58 nm (Table 1). However, for the HTDDS2 sample a basal spacing of 3.84 nm was observed, similar to those obtained in previous studies [18,19], which could suggest that there is a bilayer arrangement with a slant angle ˛ = 60.4◦ between the chain of DDS and the surface of the layer [20,21]. A schematic arrangement of the anions on modified hydrotalcites (nanohybrids) is included in Fig. 3. For the HTSEB sample a basal spacing 1.48 nm was observed. This lead to a gallery height of 1.0 nm so the anions could be considered as monolayer in a vertical position and titled towards the brucite-like layers according to the length of sebacate (≈1.3 nm) [17], the slant angle required is sin−1 (1.0/1.3) = 50.3◦ . The basal spacing was lower than that obtained for an organohydrotalcite with a ratio Mg/Al = 3.2, d0 0 3 = 1.58 nm by Bruna et al. [17]. This can be due to the difference of the layer charge and/or the dried treatment in each case [18]. The value of spacing corresponding to (1 1 0) reflection, included in Table 1, agrees with the metal ratio, which increases slightly as charge density decreases. 3.1.3. FT-IR spectroscopy Fig. 4 shows the FT-IR spectra of the organo/LDH samples (HTDDS1 is not included because it is almost equal to HTDDS2). The data reveal the LDH-like structure of all adsorbents with the corresponding interlayer anions (DDS and SEB) as reported previously [22–25] 3.1.4. Scanning electron microscopy The SEM images, included in Fig. 5, showed that the HTCO3 sample consisted of thin plate-like crystal with a irregular shape and size <10 m (Fig. 5a) and that the HTDDS1 and HTSEB (Fig. 5b and d) are very similar. In contrast, the HTDDS2 (Fig. 5c) particles formed aggregated via hydrophobic interactions and the surfaces are more diffused. This suggests that an excess of DDS could coat the outer surfaces and gives rise to crystallized NaDDS (not observed by PXRD) when the sample were separated by centrifugation and dried as has been noted by other authors [18,26].
3.2. Study of alachlor and metolachlor removal with HTDDS and HTSEB 3.2.1. Adsorption of herbicides by HTDDS1 and HTDDS2 Alachlor and metolachlor adsorption tests were carried out with the HTDDS1 and HTDDS2 sorbents and HTCO3 for comparison purposes. The herbicides adsorption on the inorganic hydrotalcite, HTCO3 was negligible (data not included) as has been observed for carbetamide [15]. This confirmed that the transformation of the hydrotalcite interlayer from hydrophilic to hydrophobic increases its affinity for non-ionic compounds such as alachlor and metolachlor. The influence of pH on alachlor and metolachlor adsorption was examined at an initial value of 3, 11 and the pH value of the corresponding herbicide solution (∼6 for alachlor and 7 for metolachlor). The final pH was tested for each experiment, and the maximum value obtained was 9.5, it was consistent with the buffering properties of hydrotalcites. The results (not shown) indicated that adsorption was not significantly affected by the initial pH. This led us to adopt an initial pH 6 and 7 for alachlor and metolachlor respectively for all adsorption tests. The kinetic results are included in Fig. 6a for adsorption of herbicides in HTDDS1 and HTDDS2 showing that the adsorption process was quite fast in both cases and the equilibrium can be reached at 4 h for alachlor and 6 h for metolachlor, because of the higher hydrophobic character of metolachlor which has slower adsorptions rates in water and the bulky molecules of the metolachlor makes its diffusion more difficult. Moreover, an increase of the herbicides concentration gives rise to an increase on the amount of adsorbed herbicides in all cases. The Cs values obtained by adsorption on HTDDS1 were slightly higher than those obtained for HTDDS2, in spite of that the interlayer space being lower for HTDDS1. However, different aggregated species can be formed according to the amount or concentration of DDS organic anion in the synthesis process of nanohybrid (admicelles, hemicelles or mixed hemicelles) [27]. As has been noted by Moyo et al. [28], intercalation of organic compounds creates diverse types of
Fig. 3. Schematic arrangement of DDS and SEB anions on the prepared adsorbents.
D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359
3000
2000
1412
1665 1665
Meto
1000
500
4000
3500
3000
2500
2000
1500
1242
1354
1671
1458
1500
HTSEB-Meto
HTDDS2-Meto
Ala
Wavenumbers (cm -1)
1084
1219
1566 1655
HTDDS2
1461 1387
2962 2918 2850 2957 2937 2838
1655
1684
2500
1589
1689
3500
HTSEB
2925 2853
1084 1219
1566 1655
1461 1387
1412
HTSEB-Ala
HTDDS2-Ala 2960 2937 2838
4000
b
HTDDS2
1698
2926 2852
2962 2918 2850
% Transmittance
2922 2846
HTSEB
2922 2846
a
% Transmittance
354
1000
500
-1
Wavenumbers (cm )
Fig. 4. FT-IR spectra of the adsorption products of (a) alachlor and (b) metolachlor on the HTDDS2 and the HTSEB adsorbents. The sodium dodecylsulfate, sebacic acid and the herbicides spectra are also included.
supramolecular structures in the clay interlayer. The intercalation of surfactants such as DDS from aqueous solution can be considered as a change in the nature of the micelle structure from spherical to lamellar [29]. Under suitable experimental conditions the
formation of double layers over the surface of the solid is favoured. The formation of double layer of surfactant in the interlayer space of HTDDS2 adsorbent, as has been confirmed by X-ray diffraction, gives rise to negative charges on the surface of organic/LDH [18,30]
Fig. 5. SEM images of the (a) HTCO3 , (b) HTDDS1, (c) HTDDS2 and (d) HTSEB samples.
D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359 0,50
a
0,45 0,40
Cs(mmol/g)
0,35 0,30 HTDDS1-Meto0,35 HTDDS1-Meto0,1 HTDDS2-Meto0,35 HTDDS2-Meto0,1
0,25 0,20
HTDDS1-Ala0,4 HTDDS1-Ala0,1 HTDDS2-Ala0,4 HTDDS2-Ala0,1
0,15 0,10 0,05 0,00 0
5
10
15
20
25
Time (h) 0,50
b
0,45 0,40
Cs (mmol/g)
0,35 0,30 0,25 0,20 0,15 HTSEB-Ala0,4 HTSEB-Ala0,25 HTSEB-Meto0,35 HTSEB-Meto0,1
0,10 0,05 0,00 0
5
10
15
20
25
355
HTDDS1 were slightly higher than on HTDDS2 (Cs = 0.31mmol/L for alachlor and Cs = 0.36 mmol/L for metolachlor), which is in accordance with the orientation of the DDS molecules in the solid water interphase as was previously mentioned. The HTDDS2 adsorbent, with the bilayers of DDS anions, presents its negative charges towards the solution, so the hydrophobic interaction is less favoured in this case while for HTDDS1 the aliphatic chains oriented towards the solution enable its interaction with the non-ionic molecules of the herbicides (alachlor and metolachlor). 3.2.2. Adsorption of herbicides by HTSEB Alachlor and metolachlor adsorption tests were carried out with the HTSEB nanohybrid. The influence of pH was studied in the same way as in the case of HTDDS in the previous section and similar results were obtained. The adsorption process was carried out at initial pH 6 and 7 for alachlor and metolachlor respectively. The kinetic results showed that the adsorption was very fast in both cases (Fig. 6b); there was no appreciable influence of the size of the herbicide molecules to reach equilibrium. An increase of the herbicides concentration gives rise to an increase in the amount of adsorbed pesticides. The isotherms of alachlor and metolachlor on HTSEB (Fig. 7c) correspond to L-type, this indicates that the adsorption occurs at the specific sites of the adsorbents. However, it could be observed that in the case of HTSEB–Meto the adsorbed amount (Cs = 0.28 mmol/L) was the lowest, but the shape of the adsorption isotherm according to Giles et al. [32] was the H-type (special L-type) indicating high affinity between the adsorptive and adsorbent, thus the metolachlor at low concentrations (0.1 and 0.15) was totally adsorbed. The maximum amount of adsorbed alachlor (Cs = 0.43 mmol/L) was higher than for metolachlor (Cs = 0.28 mmol/L) on HTSEB. That can be due to the bulky molecules of metolachlor herbicide in respect to the lower interlayer space of adsorbent HTSEB makes its diffusion more difficult.
Time (h) Fig. 6. The kinetic results for the adsorption process for (a) alachlor and (b) metolachlor on HTDDS1, HTDDS2 and HTSEB adsorbents.
and consequently a decrease of the adsorption of the low polar herbicides will be observed. On the other hand, the formation of monolayer hemicelles favours the herbicide adsorption on the alkyl chain, which is oriented towards the interphase solid–water as in the HTDDS1. The results of Fig. 6a show that the amount of herbicide adsorbed was higher for metolachlor than alachlor by HTDDS. This is likely due to the higher similarity of hydrophobic character of metolachlor and the organic anion DDS, which favour the hydrophobic interactions. The adsorption isotherms are shown in Fig. 7 and the experimental adsorption data were fitted to the Freunlich and Langmuir adsorption models (Table 2). The correlation coefficients (R2 ) for the Freunlich equation ranged from 0.91 to 0.98 and for Langmuir from 0.97 to 0.99. Therefore the adsorption isotherm will be discussed in function of the Langmuir equation which best described the observed data. The isotherms for alachlor on HTDDS1 and HTDDS2 included in Fig. 7a correspond to the L-type [31]. The adsorption processes for metolachlor on HTDSD1 and HTDDS2 (Fig. 7b) are similar and the isotherms correspond to C-type indicating a partition mechanism between the aqueous solution and the organo/LDH. This type of isotherm is very common when the increase of the concentration is limited by the low solubility of the herbicide. A similar behavior was observed for adsorption of metolachlor on berberinemontmorillonite by Rytwo et al. [32]. The amounts adsorbed of alachlor (Cs = 0.39 mmol/L) and metolachlor (Cs = 0.38 mmol/L) on
3.2.3. Desorption experiments of herbicides by the organo/LDH nanohybrids Desorption processes for both herbicides by HTDDS1 and HTDDS2 are included in Fig. 7a and b, which show that they are reversible for metolachlor according to similar hydrophobic character of pesticide and organic anions DDS. However, the desorption processes were highly reversible for HTSEB, according to the isotherm included in Fig. 7c, the negative hysteresis observed could be explained like an experimental artefact [33]. These results agreed with relative weak hydrophobic-type interaction of the herbicides molecules on the interlayer organic phase of the adsorbents and a similar behavior has been observed by terbutylazine in adsorbents alike and by modified montmorillonites [34]. 3.3. Characterization of the adsorption products The PXRD patterns of the adsorption products are included in Fig. 8. The interaction with the herbicides did not modify significantly the interlayer space for HTDDS1 (PXRD pattern not included) however for HTDDS2 the interlayer decreased to the value (d0 0 3 = 2.53 nm) close to a phase with the monolayer of the DDS anion. The contact time (24 h) between the herbicide solutions and the solid adsorbents in the adsorption test is enough to reorganize the interlayer space together with a decrease of the surfactant concentration. The decrease of the surfactant concentration produces the transformation of the bilayer into a monolayer of surfactant on the external surface of LDH (as hemicelles) [26,27] and the presence of monolayer of this surfactant in the interlayer space was confirmed by PXRD. These conditions acted as a washing process, part of dodecylsulfate was dragged by the aqueous
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D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359 0,5
0,5
0,5
Cs (mmol/g)
a
b
c
0,4
0,4
0,4
0,3
0,3
0,3
0,2
0,2
0,2
0,1
0,1
0,1 Ads HTSEB-Ala Ads HTSEB-Meto Des HTSEB-Ala Des HTSEB-Meto
Ads HTDDS1-Meto Ads HTDDS2-Meto Des HTDDS1-Meto Des HTDDS2-Meto
Ads HTDDS1-Ala Ads HTDDS2-Ala Des HTDDS1-Ala Des HTDDS2-Ala
0,0
0,0 0,0
0,1
0,2
0,0 0,0
0,1
0,2
0,0
0,1
Ce(mmol/l)
Ce(mmol/l)
0,2
Ce (mmol/l)
Fig. 7. The isotherms of alachlor and metolachlor on HTDSD1, HTDDS2 and HTSEB adsorbents.
solution and meant that both adsorption products became similar for the two sorbents (HTDDS1 and HTDDS2). The adsorption of alachlor and metolachlor does not cause an important change in the interlayer spacing of the HTDDS. The interlayer spacing of HTDDS1 and HTDDS2 is considerably larger than the pesticide molecules (see Fig. 1), therefore their incorporation can be considered in the edge of the particle tied to the alkyl chain of dodecylsulfate [19]. This has been observed in the adsorption of other non-ionic herbicides [15,17,35] The PXRD patterns, included in Fig. 8, of the adsorption complexes with HTSEB showed a basal spacing of 1.87 nm and 1.84 nm for alachlor and metolachlor respectively, indicating that the adsorption of herbicides produced an increase in the interlayer space. This can be due to accommodation of the pesticide molecules between the HTSEB. Furthermore, the bigger interlayer space of HTDDS and/or the free hydrophobic chain enabled an easier interaction with the herbicides, without the modification of the interlayer, than in the case of HTSEB with a lower value of interlayer spacing. Instead, the HTSEB adsorbent has more free space in the interlayer between the organic anions. The number of mole per formula weight requests to balance the positive charges of the hydroxide layers is half for sebacate than dodecylsulfate due to the double charge of sebacate anion, considering the same layer charge density of the LDH. Therefore, the intercalation of the herbicide molecules in the free space between the sebacate anion [35] is favoured and gives rise to an increase of the interlayer space.
However, a high packing of the organic anion in HTDDS prevents the incorporation of the herbicide molecules into the interlayer space. The FT-IR spectroscopy (Fig. 4) data show that characteristic bands corresponding to DDS and SEB anions are present in the adsorption products and that there is no noticeable change in respect to the adsorbents. Moreover, the herbicides are identified on the adsorbents in agreement with the presence of its characteristic bands as can be seen in Fig. 4. A light shift of the C O vibration band from 1689 cm−1 and 1671 cm−1 in pure alachlor and metolachlor respectively to 1684 and 1665 cm−1 for the adsorption complex was observed on HTDDS-herbicide adsorption products. This could be due to the interactions between the pesticides and the organo/LDH similarly to the pesticides bounded to organoclay [34]. The intensity of the band corresponding to vibration mode C O decreased for alachlor and metolachlor adsorbed by HTSEB and simultaneously the position of the band due to stretching vibrations of C–H bonds of aliphatic chains of sebacate is slightly shifted (from 2922 and 2846 cm−1 to 2926 and 2852 cm−1 for alachlor and 2925 and 2853 cm−1 for metolachlor which is indicative of the interactions between them. These results confirm that the alachlor and metolachlor molecules has been adsorbed in the organo/LDHs studied, although the interlayer arrangement differed in the HTDDS and HTSEB adsorbents
Table 2 The parameters of Freundlich and Langmuir. Freundlich 1/nf HTDDS1–Alachlor HTDDS2–Alachlor HTDDS1–Metolachlor HTDDS2–Metolachlor HTSEB–Alachlor HTSEB–Metolachlor
0.40 0.52 0.79 0.86 0.13 0.17
Langmuir Kf
± ± ± ± ± ±
0.08 0.06 0.06 0.09 0.02 0.02
0.85–1.10 0.85–0.99 1.33–1.58 1.26–1.59 0.76–0.81 0.65–0.71
2
R
Cm (mmol/L)
0.91 0.92 0.98 0.97 0.94 0.92
0.48 0.43 0.8 0.9 0.4500 0.2700
± ± ± ± ± ±
0.02 0.01 0.3 0.3 0.0003 0.0001
R2
L 36 14.5 8 4.5 126 218
± ± ± ± ± ±
5 1.2 5 2.1 11 15
0.98 0.99 0.98 0.97 0.97 0.98
D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359
0,151
0,376
0,493
1,87
0,94
HTDDS2-Meto
40
50
60
0,25 a
0,20 HTDDS-MetoM
0,15
HTDDS-MetoAds
0,10
HTSEB-MetoM HTSEB-MetoAds Metolachlor (Technical)
50
100
150
200
250
Time (h) 1,8
70
80
2θ (degrees) Fig. 8. PXRD patterns of the adsorption products of alachlor and metolachlor on HTDDS2 and HTSEB. (*) Al sample holder.
Metolachlor in leachetes (ppm)
30
0,30
0
0,152
0,421
1,26
2,53
20
0,35
0,00
HTDDS2-Ala
10
0,40
0,05
0,152
2,53 1,27
Intensity (a.u)
HTSEB-Ala
Metolachlor in solution (ppm)
0,151
0,495 0,379
0,92
1,84
*
* *
a
0,45
HTSEB-Meto * *
357
b
HTDDS-MetoAds
1,6
HTSEB-MetoAds
1,4
Metolachlor (Technical)
1,2 1,0 0,8 0,6 0,4 0,2
3.4. Desorption experiments in water and lixiviation in soil
0,0 0
1
2
3
4
5
6
water added /( pore volumen ) 0,20
c
0,18
Metolachlor leached (mg)
The kinetic of release of metolachlor into water under static conditions from the complexes prepared HTDDS1–MetoAds and HTSEB–MetoAds is shown in Fig. 9a. Additionally, a free (technical) product and the mechanical mixture HTDDS1–MetoM and HTSEB–MetoM , which were used for comparison purpose, are also included in Fig. 9a. The release from free (technical) product was instantaneous but the evolution of the release was slower from the mechanical mixed complexes. The concentration of the herbicide for complex HTDDS1–MetoM and HTSEB–MetoM was similar for both and reaches 90% of the total load of metolachlor. The most prominent difference in the release kinetic was between HTDDS1–MetoAds and HTSEB–MetoAds , despite the same preparation method on these complexes. The HTSEB–MetoAds (8.5%) presents a slower release of the herbicide than HTDDS1–MetoAds (26%) in the first 12 h (see insert in Fig. 9a). 72% of metolachlor was released after 8 days from HTSEB–MetoAds against 99% from HTDDS1–MetoAds . This difference in the behavior of the herbicide release may be explained by the fact that the metolachlor molecules are trapped between the sebacate chains for the complex HTSEB–MetoAds as was indicated before while for the complex HTDDS1–MetoAds there is an adsolubilization [36] of the metolachlor on the external surface of the particles. These results are in agreement with the PXRD results that indicated a modification of the interlayer spacing for HTSEB–MetoAds but not for HTDDS1–MetoAds . Fig. 9b and c show the metolachlor breakthrough curves (BTCs) for HTDDS1–MetoAds , HTSEB–MetoAds and the free (technical) product. These formulations have been selected because they appeared to be those that released the herbicide more slowly in the batch release study, Fig. 9a. Metolachlor applied as organo/LDH complexes resulted in lower maximum concentrations in the leachates with respect to the application of the herbicide as free product (Fig. 9b). The maximum concentration of metolachlor in
0,16 0,14 0,12 0,10 0,08 0,06
HTDDS-MetoAds
0,04
HTSEB-MetoAds
0,02
Metolachlor (Technical)
0,00 0
1
2
3
4
5
6
water added /( pore volumen ) Fig. 9. (a) Metolachlor release kinetics into water from its complexes with HTDDS1 and HTSEB. (b) Metolachlor breakthrough curves (BTGs) after application to soil columns as a free (technical) product and as organo/LDH-metolachlor complex (c) cumulative BTCs.
leachates was reduced from 1.76 ppm, for the technical, to 1.56 ppm or 1.12 ppm for the HTDDS1-metolachlor and HTSEB-metolachlor complexes, respectively. Fig. 9b also shows that the maximum of the BTCs of the complexes is shifted towards larger volumes of added water, at 120 mL of added water in the commercial formulation, 165 mL in the HTDDS1–MetoAds and 195 mL in the HTSEB–MetoAds . This shift of the BTCs is an important feature,
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D. Chaara et al. / Journal of Hazardous Materials 196 (2011) 350–359
because it indicates that the leaching of metolachlor was retarded by the application of the herbicide as organo/LDH complexes. Cumulative BTCs (Fig. 9c) show that the amount of metolachlor accumulated in leachates was 99% of the amount applied to the soil as free (technical) product. This amount was a little less (95%) for HTDDS1–MetoAds , but reached only 66% when the complex applied was HTSEB–MetoAds . The amount leached was greater for HTDDS1–MetoAds than for HTSEB–MetoAds . This might be due to the type of interaction between the herbicide and the organo/LDH complex in each case. The metolachlor molecules could be trapped between the sebacate ions and a greater fraction of irreversibly adsorbed herbicide in HTSEB–MetoAds complex was retained compared to the HTDDS1–MetoAds . The amounts of metolachlor extracted from the soil columns at the end of the leaching experiments were less than 3% so the amount not recovered probably corresponded to the degradation of the herbicide within the soil column and/or the formation of strongly bound herbicide residues. The results of the columns leaching test confirm that the sustained release of metolachlor from the formulation based on HTSEB–MetoAds retards the vertical movement of the herbicide through soil columns and reduce the total leaching losses of the herbicide. 4. Conclusions The organo/LDH nanohybrids containing dodecylsulfate and sebacate were prepared and its structure confirmed by different characterization techniques. The interlayer spacing of organo/LDH nanohybrids range 1.5–3.8 nm presents suitable characteristics to remove alachlor and metolachlor. The adsorption tests conducted showed that a noticeable increase of the adsorption of the non-ionic herbicides by organo/LDH was produced with respect to inorganic LDHs. Two different arrangements of DDS ions on HTDDS1 and HTDDS2 were proposed (monolayer and bilayer) but finally after the adsorption process the products had the same arrangement of the interlayer anions. The amount of herbicide adsorbed was higher for metolachlor than alachlor on HTDDS. This is likely due to the similar hydrophobic character of metolachlor and the organic DDS anions, which favour interactions between the pesticide and the organo/LDH. However, the adsorbed amount of metolachlor was lower than alachlor by HTSEB. The adsorbed amount of herbicides and the incorporation in the interlayer strongly depended on the herbicide structures and the properties of the organic anion which modifies the LDH, as well as the arrangement of the DDS and SEB in the organo/LDHs hybrid interlayers.On the other hand, the results obtained in the slow release experiments show that the HTSEB–MetoAds formulation has a controlled release behavior indicating potential applications as prolonged release vehicles for pesticide targeting The design and optimization of the chemical structure of organo/LDH nanohybrids could be considered very interesting as nanocarriers of non-ionic pesticides. Therefore, this research line is an open way to study the non-ionic pesticide interaction with the organo/LDH and design of formulations to improve the use of pesticide regarding agronomical and environmental issues. Acknowledgements This work was partially funded by MCI (AGL2008-04031CO2-02) and J. Andalucía through Research Groups FQM-214 and RNM-124. D. Chaara acknowledges MAEC (Spain) a grant from AECID to research in Universidad de Córdoba (Spain). We appreciate the technical assistance received in the SCAI (Universidad de Córdoba) for Electron Microscopy and Elemental Analysis Units.
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Journal of Hazardous Materials 196 (2011) 360–369
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Synthesis and characterization of activated carbon from sawdust of Algarroba wood. 1. Physical activation and pyrolysis Juan Matos ∗ , Carol Nahas, Laura Rojas, Maibelín Rosales Engineering of Materials and Nanotechnology Centre, Venezuelan Institute for Scientific Research (IVIC) 20632, Caracas 1020-A, Venezuela
a r t i c l e
i n f o
Article history: Received 7 April 2011 Received in revised form 2 September 2011 Accepted 10 September 2011 Available online 16 September 2011 Keywords: Activated carbon Physical activation Pyrolysis Algarroba wood
a b s t r a c t Synthesis of activated carbon (AC) from sawdust of Algarroba wood was performed as a function of the temperature under CO2 and N2 flow. Characterization was performed by adsorption–desorption N2 isotherms, FTIR, XPS and SEM. Functional acid or basic groups were detected on the surface of AC. For both studied atmospheres, the maximum value of surface area was obtained at 800 ◦ C. A monotonic correlation between temperature and mean pore diameter was detected being the higher the activation temperature the lower the mean pore width of AC. Ultramicroporous AC with pore diameters of 6.7 A˚ and 5.3 A˚ were obtained at 900 ◦ C under CO2 and N2 flow, respectively. It can be concluded that pore diameter and the functionalization of the AC surface can be controlled easily controlling the temperature of activation, independently of the gas atmosphere. The present results suggest that waste biomass is a potential source for the synthesis of carbon materials with potential novel applications. © 2011 Elsevier B.V. All rights reserved.
1. Introduction
2. Experimental
The lignocelluloses are the most common precursor materials to obtain activated carbon (AC) of low cost [1–8]. AC has been used extensively as adsorbent and catalytic support mainly due to the high surface area up to 3000 m2 /g and a wide range of pore sizes [1]. AC has been studied in catalytic heterogeneous reactions such as hydrogasification [9,10], hydrodesulphurization [11,12] and photocatalytic detoxification of waste waters [13–15]. The efficiency of AC is associated directly with texture and surface properties [16,17]. An increasing interest of wood-derived carbonsupported catalysts has been performed, mainly upon the kinetics of methane conversion reactions [18–22] to obtain syngas, bio-oil and as feedstock for chemical production. These works showed that physical and chemical characteristics of carbon supports influence the catalytic activity. Therefore, the kinetics parameters related with the synthesis of AC from biomass such as sawdust from wood [8,23–26] has received much attention. As part of a project aimed to obtain up-grading materials from biomass, the principal aim of this work treats with the synthesis of carbon materials from the sawdust of Algarroba wood. Discussion about the influence of the gaseous atmosphere and of the temperature of reaction on the texture and the chemical surface properties of the carbon are also presented.
2.1. Raw material and activated carbon characterization
∗ Corresponding author. Tel.: +58 212 5041166; fax: +58 212 5041166. E-mail addresses:
[email protected],
[email protected] (J. Matos). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.046
Sawdust from Algarroba (Hymenaea Courbaril) wood was employed. It was firstly crashed and sieved before characterization. The mean size of particles was 350 m. Tests of moisture at 120 ◦ C by 2 h, ash at 550 ◦ C by 14 h, and volatiles at 850 ◦ C by 7 min were also performed. Bulk density was estimated by geometrical filled method with sieved microparticles of 350 m. Atomic absorption (AA) was performed to analyze the chemical composition of ashes. Textural characterization was performed by adsorption–desorption N2 isotherms at 77 K. The full isotherms in the range of 4 × 10−3 to 84 kPa were measured in a ASAP-2020 equipment (Micromeritics). Equivalent surface area, micropore area and volume, total pore volume and pore diameters were obtained by Brunauer–Emmet–Teller (BET), Harkins–Jura (HJ) and Horvath–Kawazoe (HK) methods, respectively. These methods were employed because carbon materials can contain slits and spherical pores [26]. Fourier-transform infrared spectroscopy (FTIR), X-ray photoelectronic spectroscopy (XPS), and scanning electron microscopy (SEM) were also performed to characterize the samples. FTIR spectrums were obtained on a spectrophotometer Magna-IR 560 from Nicolet. The powders were mixed with KBr in a 5% (w/w) mixture and pressed to tablets of 1 cm diameter at 10 t for 1 min. Spectra were recorded from 4000 to 400 cm−1 with a resolution on 5 cm−1 . XPS was carried out in an ESCALAB 220i-XL spectrometer (VG scientific) equipped with a hemispherical electron analyzer and a double anode Mg–Al
J. Matos et al. / Journal of Hazardous Materials 196 (2011) 360–369
361
Fig. 1. Scheme of activation set-up.
non-monochromatic X-ray source. The pressure in the analysis chamber was kept below 10−9 Torr. The data were corrected to the C1s level at 285.0 eV. Scanning electron microscopy (SEM) images were obtained from palladium–gold-coated samples in a XL-30 microscopy from Phillips. 2.2. Physical activation and pyrolysis of sawdust Carbon samples were prepared in one-step procedure by two different ways, physical activation under CO2 flow and by pyrolysis under inert N2 flow. All samples were thermally treated by 0, 1 or 2 h at the final temperature (200 ◦ C up to 900 ◦ C) and denoted ACCO2−T and ACN2−T , with T the final temperature. In a typical synthesis [16], 3 g of sieved wood was treated thermally under CO2 or N2 flows (100 mL min−1 ) at constant heating rate (10 ◦ C/min) from ambient to final temperature. Activation was carried out at nearly constant atmospheric pressure (100 kPa) inside a tubular kiln (Heraeus). Fig. 1 shows a schematic representation of the equipment used in this work. Tubular reactor was made of stainless steel or Pyrex glass depending of final temperature. Different reaction time (0, 1 and 2 h) was followed and then the oven was left cooled to temperature environment under flow of inert gas. The yield (%) of carbon was verified after each experiment. 2.3. Phenol photodetoxification Phenol was purchased from Aldrich. The photocatalyst was TiO2 Degussa P25, mainly anatase (ca. 70%) under the shape of nonporous polyhedral particles of ca. 30 nm mean size with a surface area of 50 m2 /g. Activated carbons prepared under CO2 and N2
flow at 800 ◦ C and 1 h were selected for the study because they showed the highest surface areas (discussed below). These samples were denoted: TiO2 –ACCO2 and TiO2 –ACN2 . The experimental set-up has been described before [13,14] but it can be summarized as follows. The batch photoreactor was a cylindrical flask made of Pyrex of ca. 100 mL with a bottom optical window of ca. 4 cm diameter and was open to air. Irradiation was provided by a high pressure mercury lamp (Phillips HPK 125 W) and was filtered by a circulating-water cell (thickness 2.2 cm) equipped with a 340 nm cut-off filter (Corning 0.52). The water cell was used to remove all the IR beams, thus preventing any heating of the suspension, especially in the presence of black activated carbon. The cut-off filter, although decreasing the overall UV-light power available, enables one to eliminate any photochemical side reaction. Millipore disks (0.45 m) were used to remove particulate matter before HPLC analysis. The HPLC system comprised a LDC/Milton Roy Constametric 3200 isocratic pump and a Waters 486 tunable absorbance detector (Millipore) adjusted at 270 nm for the detection of phenol. The quantity of 50 mg of Titania was chosen since in our conditions there is a full absorption of the UV light entering the photoreactor and because it has shown a optimum of composition in the photodegradation of phenol [13]. The quantity of 10 mg AC was chosen to ensure a good adsorption of phenol related to the high surface area of AC without disturbing the UV absorption by titania nor phenol adsorption on it [13]. Samples of the suspension were removed at regular intervals for analysis and from the linear regression of the kinetic data of phenol disappearance as a function of reaction time Ln(Co /Ct ) = f(t), the first-order apparent rate-constants (kapp-phenol ) were obtained to compare the photoefficiency of the TiO2 –AC against TiO2 alone. Further details can be verified elsewhere [13].
J. Matos et al. / Journal of Hazardous Materials 196 (2011) 360–369
Table 1 Characterization of Algarroba (Hymenaea Courbaril) wood. Moisture (wt%) Ash (wt%) Volatiles (wt%) Fixed carbon (wt%) Bulk density (g/cm3 )
9.8 1.30 72.1 16.8 0.89
± ± ± ± ±
A 0.4 0.04 0.7 1,3 0.03
3. Results and discussion
3.2. Thermal degradation of wood The yield (%) after thermal degradation (followed gravimetrically) under CO2 or N2 flow is show in Fig. 2A and B, respectively. The moisture corresponding to the heating up to 120 ◦ C is also observed in Fig. 2. It can be noticed that the thermal decomposition begins on 200 ◦ C in both gaseous atmospheres with burn-off of about 12% and 10% under CO2 and N2 flow, respectively. This is in agreement with previous works of thermal decomposition Table 2 Atomic absorption analysis of ashes from Algarroba wood.
Al Ba B Ca Cr Cu Fe Mg Mn Ni P K Si Na Sr Zn Total
1h 2h
60 40
0
Proportion (%)
18.31 0.51 1.83 538.46 0.17 1.58 86.46 112.05 0.46 0.15 934.57 42.58 0.82 2.61 2.69 4.66
1.05 0.03 0.10 30.80 0.01 0.09 4.95 6.41 0.03 0.01 53.47 2.44 0.05 0.15 0.15 0.27
1747.91
100.00
Wood composition (wt%) 0.0136 0.0004 0.0014 0.4066 0.0001 0.0012 0.0653 0.0846 0.0003 0.0001 0.7057 0.0322 0.0006 0.0020 0.0020 0.0035 1.30
0
200
400
600
800
1000
Temperature (ºC)
B
100
0h 1h
80
Yield (%)
Results of wood characterization are shown in Table 1. It is wellreported [1,2] that carbon yield is affected by many factors such as the precursor material, the activation method, the activation agent, and other activation parameters such as pressure and heating rate. In addition, the high volatile content (about 72%) would suggest low yields during the synthesis in agreement with the low fixed coal of about 17% (Table 1). A low content of ash of about 1.3% was detected in the composition of the wood, suggesting that the carbon materials will have a high purity. Table 2 shows a summary of the results of atomic absorption (AA) obtained from ashes of Algarroba. Alkaline and alkaline-earth metals were detected being Ca the most important (ca. 0.4%). Also, low content of amphoteric elements (Al and Si) and first row transition metals (Cr, Cu, Fe, Ni, Zn and Mn) were detected. However, this composition is low enough (less than 0.1%) that any catalytic influence of transition metals on the thermal degradation of the wood can be rejected. It should be pointed out an important quantity of phosphor (ca. 0.7%) in the composition of wood. Laine et al. [27] have reported that P found in AC obtained from coconut shell is structured commonly in the shape of cyclic or linear polyphosphates [(Pn O3n )]n and [(Pn O3n+1 )](n+2) -, respectively. These phosphates can be intercalated in the borders of the carbon sheets with a strong interaction between the anions phosphates and the carbon atoms.
Concentration in ashes (mg/L)
0h
20
3.1. Raw material characterization
Element
100 80
Yield (%)
362
2h
60 40 20 0
0
200
400
600
800
1000
Temperature (ºC) Fig. 2. Yields of Algarroba wood as function of temperature and reaction time. (A) Activation under CO2 flow and (B) activation under N2 flow.
of lignocelluloses materials [23,24,28]. Fig. 2 shows three thermal different degradation zones as a function of temperature. These zones can be identified at low, moderate and high temperatures as follows. The low temperature zone between 200 and 300 ◦ C, the moderate one between 350 and 600 ◦ C, and the high temperature zone between 600 and 900 ◦ C. For the case of the thermal degradation under CO2 atmosphere, Fig. 2A shows that the higher the activation temperature and reaction time the lower the yield and concomitantly the highest burn-off of about 94% and 70% under CO2 and N2 , respectively. In short, up to 600 ◦ C, the burn-off follows a very similar trend under both atmosphere, but at temperature higher than 600 ◦ C, the burn-off is clearly larger under CO2 because the carbonized wood reacts in an efficient way with CO2 to major temperatures that 690 ◦ C according to Boudouard’s reaction [29]: C + CO2 → 2CO, which is a bimolecular process. In addition, the pyrolysis under N2 flow (Fig. 2B) shows a similar behavior up to 700 ◦ C but then being a constant to temperatures higher than 700 ◦ C. By contrast, under the atmosphere of N2 the unimolecular thermal decomposition of the wood happens from temperatures of 200 ◦ C up to 700 ◦ C. At higher temperatures than 700 ◦ C and under inert atmospheres, the yields remain practically constant because of the aromatization of carbon atoms into graphite sheets [30–33]. It should be pointed out that the samples activated by 1 h were selected for the characterization in order to compare the present results mainly against textural results obtained in a previous work related with the synthesis of AC from the Apamate wood [16] as we discuss below. 3.3. Characterization of carbons 3.3.1. FTIR and XPS FTIR is mainly used as a qualitative technique for the study of the surface chemical functional groups on activated carbon. Since
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363
Fig. 3. FTIR spectra of Algarroba wood.
carbon materials are black materials, they absorb almost all of the radiation in the visible spectrum and the peaks obtained in FTIR are commonly a sum of the interactions of different types of groups. However, some interesting features can be obtained from this technique. The FTIR spectra of Algarroba is show in Fig. 3 and in supplementary material the FTIR spectra of some selected carbon materials are included for comparative purposes. A summary of the functional organic groups detected in this wood is compiled in Table 3. A strong absorption peak at about 3424 cm−1 and another peak at 1638 cm−1 can be assigned to O–H vibrations suggesting the presence of phenol groups. Other important peaks were detected about 2926 cm−1 , 1731 cm−1 and between 1510 and 1600 cm−1 are these peaks are assigned to aliphatic C–H vibrations, carbonyl of lactones, and aromatic rings from lignin, respectively [34,35]. Tables 4 and 5 compile the results of the functional groups detected by FTIR on the surface of carbons prepared under CO2 and N2 flows, respectively. Comparison of these tables shows that the predominant groups detected were the carboxylic acids,
phenols and ethers groups. Phenol was the main acid surface group detected at low temperatures up to 450 ◦ C under both atmospheres. This result was expected because the main peak detected in the FTIR spectra of the Algarroba wood was phenol group (Fig. 3). It should be remarked that at higher temperatures than 600 ◦ C, carboxylic acids and phenol were not detected indicating a less hydrophilic behavior on the carbon samples in agreement with previous results obtained in the synthesis of other carbon materials [16,36,37]. Fig. 4 shows the XPS spectra in the region O1s of the AC prepared under CO2 flow at 450 ◦ C (Fig. 4A) and 900 ◦ C (Fig. 4B). Both spectra show clear differences. A summary of the functional groups detected from the XPS analysis in the region O1s for the AC prepared under CO2 and N2 flows are presented in Tables 6 and
Table 4 Functional groups detected by FTIR on AC prepared after 1 h reaction under CO2 flow. Temperature (◦ C)
Table 3 Summary of FTIR results of Algarroba wood. Functional organic groups
Assignment of absorption bands (cm−1 )
Carboxylic acid Lactone Phenol Ether Aliphatic groups Aromatic rings
3200; 1460; 1400; 1117 1742; 1158 3406; 1400 1276 2920 1629; 892; 774
200 250 300 350 450 600 700 800 900
Carboxylic acid √ √ √ × × × × × ×
Phenol
Ether
√ √ √ √ √
√ √ √ √ √ √ √ √ √
× × × ×
364
J. Matos et al. / Journal of Hazardous Materials 196 (2011) 360–369
Table 5 Functional groups detected by FTIR on AC prepared after 1 h reaction under N2 flow. Temperature (◦ C) 200 250 300 350 450 600 700 800 900
Carboxylic acid √ √ √ √ × × × × ×
Phenol
Ethers
√ √ √ √ √
√ √ √ √ √ √ √ √ √
× × × ×
Fig. 4. XPS spectra of 1sO region of some selected carbons activated under CO2 flow. (A) 450 ◦ C, (B) 700 ◦ C and (C) 900 ◦ C.
7, respectively. For example, AC prepared under CO2 flow developed carbonyl groups (531.5–532.6 eV) in any of the temperatures while for the case of AC prepared under N2 flow, the carbonyl was detected at temperatures higher than 700 ◦ C. This result suggests that the CO2 reacts with the wood even at low temperature to form carbonyl groups on the surface of carbon materials. Phenol and ether groups (533.0–534.1 eV) [36] were also detected in both cases (Tables 6 and 7) in agreement with the FTIR spectra (Tables 4 and 5). The XPS spectrum in Fig. 4A is highly symmetric while the XPS spectrum in Fig. 4B shows a broad shoulder at high binding energy. Fig. 4 also shows the resolved analysis of XPS spectra in the region O1s into four individual peaks. These peaks correspond to the different functional groups detected on the carbon surface. It should be pointed out that at temperatures higher than 600 ◦ C, a broad peak at binding energies between 534.3 and 537.0 eV (Fig. 4B) suggest the presence of chemisorbed water and/or oxygen [38] on carbons. The formation of molecular oxygen radical (• O2 ) has been proposed by Barr [39]. We also suggest that the peak about 536.7 eV (Fig. 4B) corresponds to this oxygen radical in agreeing with the fact that carbons prepared at high temperatures by physical activation or by pyrolysis commonly are characterized by basic surface pH [16,36,37]. In addition, it is difficult to differentiate between carbons prepared by physical activation or pyrolysed carbons when high temperatures are employed. XPS spectra in the O1s region (Fig. 4) showed important peaks about 531 eV. Beside the conventional C O functional group, some authors [37,39] attribute this peak to metal oxides when the chemical composition of the precursor wood is important as in the present case (>1 wt%). It can be suggested that under N2 flow the particles of metallic oxide need higher temperatures to spread up to the carbon surface in the comparison of CO2 , which can react from temperatures moderated to remove hetero atoms (O and H) of the wood. Finally, it can be seen from Table 8, that the higher the temperature of activation or pyrolysis the more basic is the surface of the carbon. It can be inferred from the pHPZC that surface of carbons evolutes from a soft acid surface to a basic one in agreement with the changes in the oxygenated functional groups detected by FTIR and XPS discussed above. In other words, the higher the temperature of preparation the more hydrophobic behavior of carbons is expected. 3.3.2. SEM Fig. 5 shows the SEM images of selected AC prepared under CO2 and N2 flow. SEM images show some white dots that have been
Table 6 Summary of functional carbon groups detected from XPS analysis in the O1s region for the AC prepared under CO2 flow by 1 h. Temperature (◦ C)
Metallic oxides
350 450 600 700 800 900
× × √ √ √ √
Carbonyl √ √ √ √ √ √
Phenol and ethers √ √ √ √ √ √
Quimisorbed oxygen or water × × × × √ √
Table 7 Summary of oxide carbon groups detected from XPS analysis in the O1s region for the AC prepared under N2 flow by 1 h. Temperature (◦ C)
Metallic oxides
Carbonyl
350 450 600 700 800 900
× × × × √ √
× × × √ √ √
Phenol and ethers √ √ √ √ √ √
Quimisorbed oxygen or water × × × × √ √
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365
Table 8 BET surface area (SBET ) and pHPZC of carbons prepared from Algarroba (Alg) under CO2 and N2 flow by 1 h and a comparison against those obtained from Apamate (Apa). T (◦ C)
SBET CO2 (m2 /g)
Alg-350 Alg-450 Alg-600 Alg-700 Alg-800 Alg-900 Apa-450a Apa-600a Apa-700a Apa-800a Apa-900a
92 350 870 1038 1167 752 352 426 570 770 548
a
± ± ± ± ± ± ± ± ± ± ±
1 7 17 28 31 20 5 13 14 16 21
pHPZC ACCO2
SBET N2 (m2 /g)
5.9 6.1 7.0 7.8 8.3 8.9 6.3 7.2 8.0 8.5 9.1
34 220 497 527 549 471 31 360 388 519 590
± ± ± ± ± ± ± ± ± ± ±
1 6 1 2 2 12 5 12 13 15 13
Results obtained from the Apamate wood at the same experimental conditions [16].
Fig. 5. SEM images of selected activated carbons. (A) CO2 , 350 ◦ C, (B) N2 , 350 ◦ C, (C) CO2 , 600 ◦ C, (D) N2 , 600 ◦ C, (E) CO2 , 800 ◦ C and (F) N2 , 800 ◦ C.
pHPZC ACN2 5.8 5.9 6.9 7.7 7.8 8.3 6.1 7.1 7.9 8.5 8.9
366
J. Matos et al. / Journal of Hazardous Materials 196 (2011) 360–369
Fig. 6. Adsorption–desorption N2 isotherm of AC prepared at 800 ◦ C by 1 h. (A) Under CO2 flow and (B) under N2 flow.
associated with the inorganic composition of wood. AC particles are in a micrometer scale in agreement with sieving performed. Independently of the gas and activation temperature, AC showed a cellular fibrous morphology. SEM images suggest that the present activated carbons are constituted by an interconnected channel framework in concordance with the fact that the precursor is constituted by a fibrous structure. For example, Fig. 5C indicates that under CO2 flow, even at moderate temperatures (600 ◦ C), an incipient activation occur in spite of this temperature is lower than that commonly considered as the critical temperature for the spontaneous activation under CO2 flow (about 690 ◦ C) [29]. Similar tendencies of an interconnected porous system as a function of temperature have been reported for the case of activated fibers obtained from rayon fibers [40] and for activated carbons obtained from almond shells [41]. 3.3.3. Texture of AC For the materials prepared at temperatures higher than 600 ◦ C, the adsorption–desorption N2 isotherms of carbon prepared under CO2 and N2 flow, showed very similar trends characteristic of a micropore framework as suggest Fig. 6 and figures in supplementary material. Fig. 6 shows the analysis for two selected carbons prepared at 800 ◦ C by 1 h under CO2 and N2 flow, denoted ACCO2 and ACN2 , respectively. Both isotherms correspond to a type I indicating that the framework is mainly composed by micropores. A summary of BET surface areas (SBET ) of AC prepared from Algarroba (hard wood) is compiled in Table 8 and results obtained from Apamate (soft wood) at the same experimental conditions [16] are also shown in Table 8 for comparative purposes. In general, results obtained from Algarroba and Apamate followed very similar trends.
As expected, BET surface area of AC prepared under CO2 flow was higher than under N2 . It must be noted that a maxima in the BET surface area was obtained at 800 ◦ C, both under CO2 and N2 flow. This temperature is the same than that we have found before [16] for the preparation of AC from the sawdust of Apamate [16] and for the activation of carbon foams obtained from the controlled pyrolysis of saccharose under CO2 or N2 flow [30]. Porosimetry parameters such as micropore area (porearea ), micropore volume (porevolume ), total volume or pore (Vtot ) and pore diameter (Wpore ) are showed in Tables 9 and 10 for the AC obtained under CO2 and N2 flow, respectively. It can be seen that the higher the activation temperature the higher the micropores volume (porevolume ) and the higher the total volume of pore of AC. For the micropore area (porearea ), a maximum is reached at 800 ◦ C in agreement with the BET surface area. In most of cases, the microporous area contributes with about 90% of the total surface area. In addition, it can be seen from Tables 9 and 10 that the higher the final temperature of activation the lower the mean width of pore (Wpore ). For temperatures between 350 and 450 ◦ C macroporous and mesoporous carbons were obtained, respectively; whereas between 600 and 800 ◦ C microporous were obtained. A carefully analysis of mesopore volume compiled in Tables 9 and 10 showed that in spite of the framework of the carbon materials is mainly micropore, the samples prepared by gasification with CO2 showed higher contribution in the mesopore range than samples prepared under N2 flow. This was expected because as indicated above, CO2 reacts effectively with the carbon at temperatures higher than 600 ◦ C. In general, the mean pore diameter decreases monotonically with the increase of the activation or pyrolysis temperature. This could be the consequence of a pseudo-graphitization of the graphene sheets, in spite the present maxima temperature is 900 ◦ C, clearly lower than that require for this phenomena. The mean pore diameter was lower for the AC obtained under N2 flow than those obtained under CO2 atmosphere in any of the temperatures studied. For example, at 350 ◦ C, the mean macropore diameter on N2 flow was about the half ˚ This trend is the same than that on CO2 flow (922 against 1815 A). for the temperature where mesoporous were obtained (450 ◦ C), and also in the range (600–900 ◦ C) where micropore AC were obtained. It should be pointed out that at 900 ◦ C the lowest mean width of pore of about 6.7 A˚ (Table 9) and 5.3 A˚ (Table 10) were obtained under CO2 and N2 flows, respectively. These AC can be classify as an ultramicroporous AC which have shown several potential applications such as a double layer capacitor and electrode material [42], as a separation membrane [43], as catalytic support for the hydrogen production from the dry methane reforming [20,21,44,45], and as an efficient adsorbent for the hydrogen uptake and storage [46]. 3.4. General discussion We present here a first part of a major research showing some insights about the design of both textural and functional groups on the surface of AC. It can be summarized that in this work, a hydrophobic and ultramicroporous AC can be prepared by controlling the temperature and atmosphere of the thermal degradation of waste biomass. The potential of these AC, mainly in the industry and in environmental green chemistry applications by cataand photocatalytic heterogeneous reactions will be presented in the next two works. It should be pointed out that carbon derived materials obtained from sawdust of wood contain lower ash content that those prepared from other lignocellulosic materials such as agroindustrial bio-wastes and clearly much lower than those obtained from petroleum precursors. This is one of the reasons why activated carbon materials prepared from sawdust of wood has been employed successfully in catalytic heterogeneous reactions [1]. For example, Laine et al. [11,12] have showed that pore volume of activated carbon supports play a synergistic role upon
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367
Table 9 Micropore area (porearea ), micropore volume (porevolume ), mesopore volume (mporevolume ), total pore volume (Vtot ) and mean width of pore (Wpore ) of AC prepared by activation under CO2 flow by 1 h. T (◦ C) 350 450 600 700 800 900
– 326.7 624.1 892.0 1003 645.7
porevolume (cm3 /g)b
mporevolume (cm3 /g)b
Vtot (cm3 /g)c
˚ c Wpore (A)
d
d
0.030 0.107 0.336 0.478 0.538 0.667
1815 54.5 20.6 20.0 18.6 6.72
– 0.100 0.241 0.411 0.462 0.573
– 0.007 0.095 0.067 0.076 0.094
Obtained by HJ method. Obtained by HJ method. Obtained by HK method. Not estimated.
the activity and selectivity of NiMo catalysts the in thiophene hydrodesulphurization. Also, our group has showed that pore size distribution clearly influence the catalytic activity of Ni and NiMo catalysts in the ethylene hydrogenation [9] and the kinetics of coke deposition [10]. Also, our group have showed in different works about synthesis of activated carbons by physical activation or by pyrolysis [16], and by chemical activation [17] that the pore size distribution and surface area of activated carbons remarkably affects the photoactivity of TiO2 in the photocatalytic detoxification of 4-chlorophenol. In addition, we have showed that textural properties of activated carbon clearly influence the selectivity of main intermediate products detected during the aromatic molecules as phenol and 2,4-dichlorophenoxiacetic acid [13,14] and more recently on 4-chlorophenol [47,48] and 2-propanol [49] photooxidations. In this sense, Fig. 7 shows the influence of the two carbons prepared at 800 ◦ C by 1 h upon the phenol adsorption and on the photo efficiency of TiO2 in the phenol photo detoxification under UV-irradiation. It can be seen from Fig. 7 that any of two TiO2 –AC binary materials adsorbed higher phenol (after 15 min adsorption in the dark). This enhancement in phenol molecules around photoactive TiO2 enhances the photo efficiency of the semiconductor as can also be seen in Fig. 7. This enhances has been attributed to the presence of a common contact interface between TiO2 and AC that make possible a continuous transfer of the species from the AC to the TiO2 surface [49]. Our present enforces are aimed to prepare hierarchically macro–meso–micro porous carbon materials to study the influence of pore size distribution on the selectivity of NiMo catalysts in hydrocracking reactions and to verify the presence of confinement effects on the selectivity of hydrocracking consequence of specific pore size an pore volume of the support such as in the case of zeolites [50,51]. We do believe that the present results regarding the pyrolysis of sawdust of a hard wood as Algarroba consists of essentially 3 different zones is a very important finding and deserve to be studied carefully. A better explanation for the present results where the reaction rates rather decreased with the increasing temperature from 350 to 600 ◦ C could be due to the
7
7
Phads 6
kapp
6
5
5
4
4
3
3
2
2
1
1
0
kapp (min-1)x10-3
c d
d
Phads (micromols)
a b
porearea (m2 /g)a
0
TiO2
TiO2 + AC-CO2
TiO2 + AC-N2
Photocatalysts Fig. 7. Summary of kinetic results of phenol adsorption in the dark (Phads ) and first-order apparent rate-constants (kapp ) of phenol photodegradation under UVirradiation.
presence of different chemical structures of the starting materials at zero holding time after heating up to the set temperatures (i.e., 350, 400 and 600 ◦ C). Therefore, the estimation of kinetic parameters from data of modulated thermogravimetric analysis is required to better understand the correlation of the influence of each pyrolysis zone on the pore size distribution and pore volume of carbon materials. In addition, the influence of ashes of lignocellulosic carbon precursors and the influence of additives (chemical activators) that can play the role of catalysts to improve the textural properties are necessary to clarify the significance of the present results. In this way, our groups have already reported preliminary studies regarding the influence of the pyrolysis atmosphere [52] and the effect of
Table 10 Micropore area (porearea ), micropore volume (porevolume ), mesopore volume (mporevolume ), total pore volume (Vtot ) and mean width of pore (Wpore ) of AC prepared by activation under N2 flow by 1 h. T (◦ C) 350 450 600 700 800 900 a b c d
porearea (m2 /g)a d
– 201.7 449.1 494.0 515.4 454.8
Obtained by HJ method. Obtained by HJ method. Obtained by HK method. Not estimated.
porevolume (cm3 /g)b
mporevolume (cm3 /g)b
Vtot (cm3 /g)c
˚ c Wpore (A)
d
d
0.016 0.111 0.190 0.201 0.208 0.231
922 43.3 16.3 15.8 15.8 5.33
– 0.101 0.172 0.188 0.195 0.223
– 0.010 0.018 0.013 0.013 0.008
368
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chemical additives [53] on the topological organization of carbon materials obtained from the controlled pyrolysis of saccharose.
[18]
4. Conclusions [19]
AC was prepared from the sawdust of wood by physical activation and pyrolysis under CO2 and N2 flow, respectively. Maxima BET surface area was obtained at 800 ◦ C, both under CO2 and N2 atmospheres to then decrease at higher temperatures of activation. IR and XPS suggest that the higher the activation temperature the more basic is the functional groups on surface of carbons. Porosimetry showed that the higher the activation temperature the higher the micropores volume and the higher the micropore surface of AC. The higher the activation temperature the lower the pore diameter obtaining an ultramicroporous activated carbon at 900 ◦ C. It can be concluded that mean pore width and functionalization on the surface of AC can be easily controlled and this feature permits to think in waste biomass as a potential source for the synthesis of carbon materials with different and potential modern applications.
[26] [27]
Acknowledgement
[28]
J. Matos thanks to the Ministry of Science and Technology for financial support.
[29]
[20]
[21]
[22] [23] [24]
[25]
[30]
Appendix A. Supplementary data [31]
Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.046.
[32]
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Journal of Hazardous Materials 196 (2011) 370–379
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Crystallization evolution, microstructure and properties of sewage sludge-based glass–ceramics prepared by microwave heating Yu Tian a,b,∗ , Wei Zuo a , Dongdong Chen a a b
School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology (SKLUWRE, HIT), Harbin 150090, China
a r t i c l e
i n f o
Article history: Received 19 April 2011 Received in revised form 7 September 2011 Accepted 10 September 2011 Available online 16 September 2011 Keywords: Sewage sludge Glass–ceramics Microwave Double-layer reactor
a b s t r a c t A Microwave Melting Reactor (MMR) was designed in this study which improved the microwave adsorption of sewage sludge to prepare glass–ceramics. Differential scanning calorimetry (DSC), X-ray diffraction (XRD), and scanning electron microscopy (SEM) were used for the study of crystallization behavior and microstructure of the developed glass–ceramics. DSC and XRD analysis revealed that crystallization of the nucleated specimen in the region of 900–1000 ◦ C resulted in the formation of two crystalline phases: anorthite and wollastonite. When the crystallization temperature increased from 900 to 1000 ◦ C, the tetragonal wollastonite grains were subjected to tensile microstresses, causing the cracking of crystal. Al ions substituted partially Si ions and occupied tetrahedral sites, giving rise to the formation of anorthite. The relationship between microwave irradiation and crystal growth was studied and the result indicated that the microwave selective heating suppressed the crystal growth, giving apparent improvements in the properties of the glass–ceramics. The glass–ceramics products exhibited bending strength of 86.5–93.4 MPa, Vickers microhardness of 6.12–6.54 GPa and thermal expansion coefficient of 5.29–5.75 × 10−6 /◦ C. The best chemical durability in acid and alkali solutions was 1.32–1.61 and 0.41–0.58 mg/cm2 , respectively, showing excellent durability in alkali solution. © 2011 Elsevier B.V. All rights reserved.
1. Introduction One of the main environmental problems is the safe disposal of the huge amount of sewage sludge that is produced every day in wastewater treatment plants [1]. Among the methods of the treatment of sewage sludge, glass–ceramics preparation seems to be a promising one for converting sewage sludge into novel materials that possess attractive mechanical and chemical properties [2]. Sewage sludge containing large amounts of CaO, SiO2 , and Al2 O3 can be a good raw material for glass–ceramics production. By controlling the initial composition and by suitable heat treatment, a variety of crystalline phases will be obtained [3]. They exhibit bending strength, Vickers microhardness, fracture toughness, chemical durability and thermal shock resistance superior to those of glass, and in some cases traditional ceramics [4,5]. It should be noted that the chemical energy of the organic components in sewage sludge could be recovered during the prepared procedure of glass–ceramics as an auxiliary energy source. The
∗ Corresponding author at: School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China. Tel.: +86 451 8608 3077/13804589869; fax: +86 451 8628 3077. E-mail address:
[email protected] (Y. Tian). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.045
chemical energy in sewage sludge can be recovered during the prepared procedure of glass–ceramics as an auxiliary energy source, reducing the emission of CO2 which is favorable to Kyoto Protocol [6]. Other advantages of this technology are the possibility of immobilizing heavy metal ions (held in the framework of glass or encapsulated into the crystallization phase) [7], the large reduction of volume (vary between 40 and 90%), and the flexibility of treatment procedure (which may accept different types of sewage sludge, either municipal or industrial) [8]. Preparing glass–ceramics by the conventional technology is an energy-intensive process, with the process temperature as high as about 1300 ◦ C and the process time required as several hours [9]. Economic analysis of a glass–ceramics preparation system which can process 0.5–1.0 ton of sewage sludge per hour showed that the operating costs of this unit ranged from US$100–420 per ton, including labor, fuel and maintenance [10]. Another critical point in glass–ceramics preparation is the difficulty in controlling the size and the type distributions of the crystals due to the thermal inertia of the conventional heating [11]. To overcome these drawbacks, microwave heating has been developed as an alternative technology for the preparation of dense structural glass–ceramics, which is characterized by shorter reaction time, reduced energy consumption, and suppressed crystal size. It was found that the treatment temperature was decreased
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371
Fig. 1. A schematic of the microwave preparation reactor assembly: (1) microwave cavity; (2) Microwave Melting Reactor (MMR); (3) waveguide; (4) magnetron; (5) PC with fuzzy logic algorithm; (6) power governor; (7) infrared radiation thermometer.
from 1300 ◦ C to 1000 ◦ C when a glass–ceramics was sintered from barium aluminosilicate glass in microwave field [12]. It was also reported that an abrasion resistant glass–ceramics was developed from the MgO–Al2 O3 –TiO2 system in 20 min by microwave heating [13]. Moreover, more uniform and strong bonding was observed in the glass–ceramics prepared by microwave, indicating that microwave energy suppressed the grain growth in crystal phase due to a fast heating rate and apparent low-temperature crystallization [8]. Sewage sludge is a poor receptor of microwave energy to achieve the temperature necessary for preparing glass–ceramics. It has been proved that microwave-induced preparation is possible, if an effective receptor is added into the raw sludge. The temperature of sewage sludge can achieve 1200 ◦ C in microwave field when it was homogeneously blended with microwave receptor, such as graphite and char [9]. However, there are fundamental disadvantages of this method when it is applied in glass–ceramics preparation. The chemical composition of the samples shows uncontrollable changes in virtue of adding microwave receptor, leading to the poor properties of the products. In addition, the microwave receptor could not be recovered due to the encapsulation of silicate matrix in the glass–ceramics, increasing the operating cost of the procedure. Attempts termed as “hybrid microwave sintering” were also made to set around the sample directly to initially heat the material at room temperature [14]. However, the temperature of sewage sludge could not reach high enough owing to the significant reflection loss on the interface between microwave receptor layer and the air surrounding it. To solve these problems, a new Microwave Melting Reactor (MMR) was designed in this study for preparing glass–ceramics
from sewage sludge. In MMR, microwave absorption of sewage sludge can be improved by the double-layer structure and the required temperature can be achieved in a very short of time, usually in a few minutes. A wave-transparent layer was introduced into the MMR system to decrease the reflection coefficient of the interface between the air and the MMR. Another important property of the powder was the low thermal conductivity which could give the sample a good heat insulation quality. The double-layer structure in MMR provides the even distribution of temperature and electromagnetic field in the samples, favoring the production of glass–ceramics with desired qualities. Further researches presented in this paper were focused on: (1) investigating the influence of heat-treatment schedule on the crystallization behavior and microstructure of the microwave-prepared glass–ceramics, (2) defining the evolution of crystallization in microwave field which were hardly found by applying the conventional procedures, and (3) gaining an insight into the chemical and physical properties of the glass–ceramics prepared by microwave in comparison to that obtained from conventional process. 2. Experimental 2.1. The design of MMR The 2.45 GHz microwave furnace, which consisted of a rectangular multimode cavity, a continually adjustable power supply (0.50–2.7 kW), a temperature controlling system, and a Microwave Melting Reactor, was used for microwave heating experiment. As shown in Fig. 1, the Microwave Melting Reactor (MMR) consisted of a wave-absorbing layer and a wave-transparent layer. The
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wave-transparent layer is the surface layer which plays an important role in avoiding the reflection loss of the incident wave on the front surface between the reactor and air. The wave-absorbing layer beneath it absorbed the incident wave transmitted through the wave-transparent layer and transformed the electromagnetic energy to thermal energy. In terms of optimizing the MMR performance, the material properties and the thickness of each layer are most important parameters to design the reactor structure. Active carbon, a well known microwave receptor, was filled in the microwave-absorption layer. The material filled in microwavetransparent layer was selected according to the expression for reflection coefficient as given in Eq. (1) [15]:
2 − 1 +
R=
2
(1)
Proximate analysis (wt.%) Aa
Va
Caa , b
Oc
Ha , b
Na , b
Sa , b
24.50
75.50
39.40
24.43
5.71
4.75
1.18
Heavy metal content in dry sewage sludge (ppm) Cr
Cd
Cu
Pb
Zn
Fe
Ni
143
4.75
138
59.9
700
10,200
38.7
A: ash content; V: volatile matter content. a Dry base. b Ash free basis. c Calculated by difference.
1
where R is the reflection coefficient, 1 and 2 are the characteristic impedances of the air and the material filled in microwavetransparent layer, respectively. It is clear that the value of 2 should be close to the value of 1 for decreasing the reflection coefficient of the incident wave. Based on the results of our preliminary experiments, ferric oxides mixed with aluminum oxides (Fe2 O3 /Al2 O3 = 1:1) were adopted as the materials filled in the microwave-transparent layer. Before filled into the microwavetransparent layer, the Fe2 O3 and Al2 O3 grains were ground using a mill to obtain a mixed powder with particle size ≤75.0 m. The mixture powder had the properties of both lower characteristic impedance and higher microwave transmission rate, decreasing the reflection coefficient of the interface between the air and the MMR. Another important property of the powder is the low thermal conductivity which gives the sample a good heat insulation quality. The thicknesses of microwave-transmission layer and microwave-absorption layer were determined according to the penetration depth (DE , depth of the microwave energy penetrates into a material). DE can be calculated by the Fresnel formula [8]: DE =
Table 1 Chemical characteristics of sewage sludge.
0 √ εr tgı
(2)
where 0 is the length of electromagnetic wave in vacuum, εr is the material dielectric constant, and tgı is the dielectric loss tangent. According to the calculation of Eq. (2), both the optimal thicknesses of microwave-transmission layer and microwave-absorption layer were determined as 2 mm. The glass preparation from sewage sludge has been carried out to test the behavior of MMR in microwave heating. It was observed that the temperature of the specimen required for preparing glass was reached (1300 ◦ C) and parent glass was prepared successfully in this reactor. 2.2. Parent glass production Sewage sludge used in experiments was collected from urban wastewater treatment plants in Harbin, China. Selected chemical characteristics and the heavy metal contents of this sludge are given in Table 1. The dehydrated sewage sludge (moisture content was 79.8%) which contained small hard particles were crushed in a mortar and then heated at 1000 ◦ C until the sludge samples reached constant weight to remove the volatile components. Sewage sludge must be mixed with additives to lower the melting temperature from its melting around 1500 ◦ C. In our experiments, CaO and waste glass were used as effective additives. The chemical compositions of the sewage sludge, waste glass and raw materials were examined by X-ray fluorescence spectroscopy (XRF) and the results are shown in Table 2, indicating that the formed glass should be in the SiO2 –CaO–Al2 O3 ternary phase system. Fig. 2 shows the phase diagram of the CaO–Al2 O3 –SiO2 system. The chemical compositions of raw materials for preparing glass–ceramics could be located in the
wollastonite–anorthite subsystem (the region marked by hatching in Fig. 2). The batch composition, prepared by mixing 52.0 wt.% of the raw sludge with 21.0 wt.% of CaO and 21.0 wt.% of waste glass, was chosen on the basis of the eutectic composition (CaO 38.0, Al2 O3 20.0 and SiO2 42.0 mass%) [16]. Additionally, 6.0 wt.% TiO2 were added to the base glass composition as nucleating agents. Mixtures obtained above were melted by microwave processing and conventional processing respectively. In the microwave process, glasses were prepared by melting sewage sludge in a corundum crucible at 2000 W microwave power for 10 min and then cooled naturally to room temperature. In the conventional process, glasses were prepared by melting the mixture in an alumina crucible at 1450 ◦ C for 2 h after which melts were preheated at 600 ◦ C to reduce thermal shock. The results of XRD analysis for sewage sludge, parent glasses obtained from these two different heating processes are shown in Fig. 3. 2.3. Glass–ceramics production It is important to determine nucleation and crystal growth temperatures precisely for effective conversion of glasses to glass–ceramics. Differential scanning calorimetry analysis (DSC) is performed using a calorimeter (STA449C, NETZSCH) with ␣Al2 O3 as standard sample. The glass powders are heated from room temperature to 1100 ◦ C with the rate of 10 ◦ C/min in order to detect the nucleation and crystallization temperatures. According to the results of DSC, heat-treatment schedule for the microwaveproduced parent glass should include a nucleation stage at 760 ◦ C for 30 min followed by a crystal growth stage at different temperatures (900 ◦ C, 950 ◦ C and 1000 ◦ C) for 60 min in microwave irradiation. For conventional glass, sample should be held at nucleation temperature (820 ◦ C) for 90 min and then heated to crystallization temperature (1000 ◦ C) for 120 min in electric furnace. Fig. 4 shows the processes of glass–ceramics preparation by microwave and conventional methods. The types of the crystalline phases were characterized by X-ray with Cu K␣ radiation (XRD: P|max-␥, Rigaku, Japan). The step length was 0.02◦ with scanning speed of 5◦ /min in the range of 10–90◦ (Cu Ka = 1.5418 A). The schemes of glass–ceramics preparation by microwave and conventional heating are shown in Fig. 4. 2.4. Methods to evaluate glass–ceramics properties Several techniques were used to evaluate the properties of glasses and glass–ceramics. The morphology of the crystalline phases made in thermal glass treated was investigated using a scanning electron microscope (SEM, S-4700, HITACHI). Archimedes’ method was employed to measure the apparent density of the glass–ceramics. Hardness and fracture toughness were measured by an indentation method using the Vickers indenter. Vickers hardness was measured with loads of 100–1000 g with loading
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373
Fig. 2. Phase diagram of the CaO–Al2 O3 –SiO2 system. The chemical compositions of sewage sludge correspond to the eutectic point marked by hatching (CS, wollastonite; CAS2 anorthite).
times of 10 s. Bending strength was obtained from a four-point method with spans of 20 and 40 mm at a cross-head speed of 100 mm/min, as designated by American Society of Testing Materials (ASTM) E855-90 [17]. The thermal expansion coefficient (20–400 ◦ C) was measured by TMA with a heating rate of 10 ◦ C/mm in atmosphere. Chemical durability was measured following the designation of American Society of Testing Materials (ASTM) C27988 [18]. First, powdered specimens were prepared in particle sizes of 4.75–6.75 mm. 20.0 g specimen powder was then immersed into 100 ml of 1 mass% H2 SO4 (about 0.10 mol/l) or 1 mass% NaOH (0.25 mol/l) and boiled on a hot plate for 48 h. The specimens were dehydrated and acid/alkali durability was estimated by measuring the weight loss of powders. 2.5. Methods of heavy metal leaching tests Leaching tests of sewage sludge and glass–ceramics were subjected to the toxicity characteristic leaching procedure (TCLP) method according to the US Environmental Protection Agency [19]. The sludge and glass–ceramics samples were manually crushed (<9.00 mm) and placed in a flask. The extraction solution consisted of acetic acid diluted in distilled water with a pH value of 4.93. Extraction fluid was added into the flask to keep a liquid-to-solid ratio of 20. The flask is tightly closed and stored at 25 ◦ C for 18 h. The resultant solutions were filtered through 0.6 m filters and the concentration of heavy metals in the solution was determined by ICP. A Perkin Elmer A Analyst 200 ICP operated at 13.56 MHz (using Ar and N2 gases) was used for the measurements. 3. Results and discussion 3.1. Thermal analysis of the parent glass The glass transition temperature (Tg) and crystallization temperature (Tc) of the parent glasses prepared by microwave process
and conventional process were determined from the DSC traces as shown in Fig. 5. The DSC profiles from the different processes showed different individual properties with regard to peak position and intensity. For the glass prepared by microwave process, an intense exothermic peak (Tc) was observed at 969.5 ◦ C which was attributed to crystallization from the parent glass. In the case of conventional heating, Tc peak was clearly evident at about 985.1 ◦ C. Apparently, microwave irradiation gave rise to the lower temperature of Tc. It was also noted that the exothermic peak of microwave prepared glass had higher intensity than that of conventionally prepared glass, demonstrating that microwave irradiation played an important role in enhancing bulk heterogeneous crystallization. The important differences between the DSC curves of the two glass samples may be attributed to the less required energy for crystallizing when microwave was used in the preparation of the glass–ceramics. Crystallization was a process to consolidate partial glass into strong crystal phases by supplying sufficient thermal energy to overcome the energy barrier between the glass and the crystal. The principle of microwave heating was to apply an electromagnetic field to the sludge samples, causing violent agitation in atoms thus raising the product’s temperature. The atoms with continual rapid motion decreased strength of the bonds within the glass network or directly precipitated from glass as nuclei, resulting in the reduction of energy barrier for crystallization. It must be noted that the glass transition temperatures Tg were difficult to determined due to their weak peaks in the DSC curves of both microwave and conventional prepared samples. Weaker intensity of Tg peak for microwave-prepared glass probably associated with its higher crystallization capability. When glassy phases transformed to crystalline phase(s), the molecular rearrangement phenomenon occurred preceding glass crystallization i.e. precrystallization stage [20]. Regarding the kinetics, heating absorption was necessary in this stage, which was called the activation energy of crystallization (E). The heating absorption was recorded in the DSC curve as endothermic peak and the corresponding temperature
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Fig. 3. XRD patterns for sewage sludge, parent glass obtained from conventional and microwave heating.
was Tg [21]. The increased heat absorption during precrystallization stage resulted in the higher intensity of the Tg peak, indicating the lower crystallization capability of the parent glass. On the contrary, decreased heat absorption caused insignificant Tg
peak, suggesting the higher crystallization capability of the parent glass.
3.2. XRD analysis of the glass–ceramics
Fig. 4. The scheme of glass–ceramics preparation by microwave and conventional heating in the present study.
X-ray diffraction analysis (XRD) was carried out to identify the crystalline phases in both the conventional and microwaveprocessed samples crystallized at 1000 ◦ C. For the conventional samples, there was only one major crystalline phase, anorthite, which had intensive XRD lines at 23.4, 24.3 and 28.6 A˚ (Fig. 6(a)). For the microwave samples, XRD measurement showed the precipitation of anorthite as a major crystalline phase and the precipitation of a small amount of -spodumene solid solution in all the glass–ceramics after the heat treatment (Fig. 6(b)–(d)). The disappearance of wollastonite phase in conventional sample might be explained by thermal inertia of the conventional process which prolonged the duration at high temperature for molten glass, leading to early phase separation and crystallization in glass phase. Furthermore, the X-ray diffraction lines for microwave sample were progressively sharper and more distinct. This behavior might be related to the microwave irradiation, which was believed to decrease the viscosity of the glassy matrix phase and enhance crystallization as discussed above. To design the optimized heat-treatment schedule of microwave process, XRD patterns of microwave glass–ceramics obtained at
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375
Fig. 5. DSC curves (heating rate 10 ◦ C/min) corresponding: (1) parent glass prepared by microwave heating; (2) parent glass prepared by conventional heating, respectively.
Fig. 6. XRD patterns of the glass–ceramics obtained from (a) conventional process and microwave process at (b) 900 ◦ C, (c) 950 ◦ C and (d) 1000 ◦ C.
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Fig. 7. SEM micrographs of the glass–ceramics obtained from conventional process and microwave process (a: conventional-prepared sample; b: microwave-prepared sample at 900 ◦ C; c: microwave-prepared sample at 950 ◦ C; microwave-prepared sample at 1000 ◦ C).
different crystallization temperature were also studied. The XRD results of microwave-prepared samples at 900, 950 and 1000 ◦ C suggested a consecutive transformation of the parent glass into crystalline phases. The X-ray pattern of the sample obtained at 900 ◦ C is basically composed of low intensity peaks of wollastonite (CaSiO3 ). A further temperature increase to 950 ◦ C enhanced the intensity of the wollastonite peaks and weakly new phase of anorthite (CaAl2 Si2 O8 ) was observed. When the glasses were crystallized at 1000 ◦ C, the prepared sample was predominantly composed of anorthite with residual glassy phase, implying that the main crystalline phase had changed from wollastonite to anorthite. From the figure, it was obvious that the anorthite peak intensity was a function of crystallization temperature. The peak intensity increased while crystallization temperature increased, indicating a temperature dependence of anorthite crystallization. According to the theory of stable energy of glass structure unit [22], the crystallization process of CaO–Al2 O3 –SiO2 system was assumed in this paper. At 900 ◦ C, the free Ca2+ was prone to unite [SiO4 ] in order that wollastonite was first formed in glass–ceramics [23]. When the temperature reached 1000 ◦ C, the structure unit Ca[SiO4 ] was forced to rearrange and unite [AlO4 ] to form anorthite as the main crystalline phase in glass–ceramics.
3.3. SEM analysis of the glass–ceramics The microstructural features of both conventionally and microwave processed glass–ceramics crystallized at 1000 ◦ C were analyzed by SEM as shown in Fig. 7. To determine the influence of crystallization temperature on microwave-processed glass–ceramics, the SEM micrographs of microwave processed samples crystallized at 900 ◦ C and 950 ◦ C are also exhibited in Fig. 7. It appeared that microwave process produced finer sized crystallites in the glass–ceramics than that evolved by the conventional heat treatment process. The results revealed that rapidity of microwave method avoided undesirable grain growth and provided a finer and uniform microstructure, which was an attractive feature for the processing of glass–ceramics. The similar results might be found in other literatures [12,14]. Mahajan et al. [24] reported that high-density ceramic materials were obtained in 4 h of cycle time for microwave sintered sample whereas it took 22 h for conventional method. It should also be noticed that the heat-treatment schedules had significant effect on the microstructure of glass–ceramics obtained from microwave process. For sample crystallized at 900 ◦ C, needle-like wollastonite-2 M crystal grains were observed. When the crystallization temperature increased to 950 ◦ C, not only
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377
Table 2 Mineral chemical composition of sewage sludge, waste glass and resulting glass–ceramics (wt.%). Sample
SiO2
Al2 O3
MgO
CaO
K2 O
Na2 O
TiO2
Fe2 O3
P2 O5
Others
Sewage sludge Waste glass Glass–ceramics
47.62 70.63 39.62
18.34 3.25 10.21
2.50 – 1.30
7.91 10.66 27.34
2.74 – 1.42
1.33 15.44 3.92
0.81 – 6.42
8.29 0.02 4.31
7.16 – 3.72
3.29 – 1.74
the amount of crystal grains increased but also the length of the grain became short, suggesting that wollastonite-TC appeared. The conversion process of the wollastonite phase was also described by Toya et al. [25]. They considered that the crystalline phase formed at 900 ◦ C was wollastonite-2 M (the low temperature phase of CaSiO3 ), gradually transforming to wollastonite-TC (the high temperature phase of CaSiO3 ) at ≥950 ◦ C, the vicinity of the phase transition temperature. When the crystallization temperature increased to 1000 ◦ C, the glass–ceramics reached an almost stationary microstructure consisting of flake-like crystals. According to the results of XRD, these crystals should attribute to anorthite. Based on the SEM results, a particular process model of glass–ceramics preparation from sewage sludge by microwave heating was developed. During the fabrication of parent glass, the interaction between air and hot glass might induce the occurrence of active radicals (for example Si–OH) and hetero-matters (for example silica gels) on the surface of glass [12]. It has been reported that the active radicals and hetero-matters could act as the nuclei and bring the nucleation energy lower [26–28]. Accordingly, under the synergetic actions of the above factors, the surface of glass would be homogeneously nucleated and then preferentially deposited wollastonite-2 M as initial crystals. During microwave process, the microwave power was absorbed by parent glass itself and hence, the rise of heat content in the bulk of glass took place at a faster rate compared to that of the near surface region. Thus, the temperature of the bulk was much higher than that of the surface, leading to a latent heat release from bulk to the surface. Under the effect of latent heat release during crystal growth, the initial crystals did not grow evenly in all directions. Instead, the growth would be orientated along directions with lower latent heat release impact. When primary dendrite crystals grew along the surface orientation, the impact of latent heat release was so weak that crystallization grew faster and the crystallite size became larger. Larger crystallite size led to decrease in their bonding area and thereby, the extent of reinforcement to the glass matrix decreased, causing the decrease of the measured hardness. Therefore, the glass–ceramics obtained at 900 ◦ C had slight poor mechanical property. In fact, the wollastonite-2 M phase was thermodynamically unstable but kinetically favorable and as a result, this phase appeared first. When the crystallization temperature reached 950 ◦ C, wollastonite-2 M was replaced by the slower growing, but thermodynamically favored wollastonite-TC phase. Since the wollastonite-TC crystals grew perpendicularly to interface between the surface and bulk, the effect of latent heat release became stronger, and then, the crystallization proceeded slower and the crystalline size became finer. The area fraction enhancement of fine crystallites provided more rigidity and localized strong bonding which could cause an apparent increase in the hardness value of the glass–ceramics. When the crystallization temperature increased to 1000 ◦ C, hydrostatic compressive microstresses developed in crystal phase owing to the much higher latent heat release from the bulk to the surface. The tetragonal wollastonite grains were subjected to tensile microstresses, causing the cracking of the Si–O bond. Al ions substituted partially Si ions in the structure of wollastonite and occupied sites, giving rise to the formation of tetrahedral anorthite.
3.4. Physical and chemical properties The various properties of the resultant glass–ceramics obtained from conventional process and microwave process with different crystallization temperatures are listed in Table 3. The bending strengths of the microwave processed glass–ceramics ranged from 86.5 to 93.3 MPa, showing no apparent relationship with the crystallization temperature. These bending strengths were relatively higher compared with those prepared using conventional method (70.2 MPa), which might be attribute to the differences in the microstructures among the specimens. Undesirable grain growth was avoided in microwave process for its high heating rate and negligible thermal inertia which provided a finer and uniform microstructure [29]. Therefore, crystalline distribution in a microwave-processed sample was roughly uniform, and the impressed pressure value on every microscopic part was almost the same. In addition, the fine, compact and homogeneous distribution of crystals in microwave samples also led to remarkably higher values of Vickers’ microhardness than that of conventional ones. The highest one, which was generated at 950 ◦ C and did not increase in number at higher temperature, was close to 6.54 GPa, thus being comparable to that of commercial glass–ceramics for applications in the building industry [30]. The chemical durability of glass–ceramics was important if the materials were to be considered as a potential building material. The weight losses of the microwave processed glass–ceramics after leaching in acid and alkali solutions were in the range of 1.32–1.61 and 0.41–0.58 mg/cm2 , respectively. The weight losses in acid were higher than those in alkali, indicating higher chemical durability of the microwave processed glass–ceramics to alkali than to acid. The weight losses of the conventionally processed glass–ceramics after leaching in acid and alkali solutions were 1.20 and 0.91 mg/cm2 for acid and alkali, respectively. The glass–ceramics obtained from microwave process thus had slightly lower chemical durability to acid but significantly higher durability to alkali than the conventional ones. This difference might be due to the differences in crystalline amount of the glass–ceramics samples from different processes. The weight loss in acid solution was well known to be generated by selective leaching of the crystalline CaSiO3 phases which were more soluble in acid than the glassy matrix [22]. Since the crystalline phase in microwave samples consisted of wollastonite (␣-CaSiO3 and -CaSiO3 ) with lower durability in acid, the degree of leaching increased and resulted in a slight lower chemical durability to acid [25]. By contrast to the slightly lower chemical durability to acid, microwave processed glass–ceramics exhibited excellent durability to alkali leaching. This might be due to the low glass content in the microwave sample, which was more soluble in alkali than in acid. The thermal expansion coefficients of the microwave glass–ceramics ranged from 5.29 × 10−6 /◦ C to 5.75 × 10−6 /◦ C, a little lower than that of the sample obtained from conventional process (6.21 × 10−6 /◦ C). The present thermal expansion was also lower compared to the values reported for various glass–ceramics, giving the present materials an enhanced ability to suppress thermal stress in applications involving building materials.
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Table 3 Physical and chemical properties of the glass–ceramics obtained from conventional process and microwave process at different crystallization temperatures. Property
Microwave-prepared glass–ceramicsa ◦
◦
900 C Bending strengths (MPa) Vickers’ hardness (GPa) Acid weight loss (mg/cm2 ) Alkali weight loss (mg/cm2 ) Thermal coefficient (×10−6 /◦ C) a b
86.5 6.12 1.61 0.41 5.75
950 C
± ± ± ± ±
0.2 0.05 0.02 0.01 0.03
93.4 6.54 1.53 0.53 5.31
± ± ± ± ±
Conventional glass–ceramicsa
Commercial glass–ceramicsb
◦
1000 C 0.2 0.03 0.03 0.01 0.01
91.6 6.41 1.32 0.58 5.29
± ± ± ± ±
0.4 0.02 0.02 0.01 0.04
70.2 6.28 1.20 0.91 6.21
± ± ± ± ±
0.2 0.03 0.01 0.02 0.03
82.0 5.80 1.32 1.00 6.70
The presented value (mean value ± standard deviation)is average of five results. Data from Ref. [30].
Table 4 TCLP results of the glass–ceramics samples. TCLP results (mg/l)
Cu
Pb
Zn
Cd
Cr
As
Sewage sludge Glass–ceramics (900 ◦ C) Glass–ceramics (950 ◦ C) Glass–ceramics (1000 ◦ C) Regulatory standard of EPA
8.21 – – – 50.0
2.11 0.00537 – – 3.00
22.21 1.21 – – 50.0
0.451 0.00213 – – 0.300
1.54 – – – 1.50
– – – – 1.50
3.5. Leaching characteristics of heavy metals
Acknowledgements
TCLP results of the developed glass–ceramics are given in Table 4. Any heavy metal concentration could not be detected in the extraction solutions of the samples prepared at 950 and 1000 ◦ C, while small concentrations of Pb, Zn and Cd ions which were lower than the limits suggested by EPA [19] were detected in the extraction solutions of samples prepared at 900 ◦ C. Consequently, heavy metals were sufficiently stabilized by preparing glass–ceramics, according to the US EPA standards. It seemed that high bonding strength of anorthite was responsible for this behavior. During the crystallizing process, the nuclei formed at the preferred sites and absorbed ions from the silicate matrix. Heavy metal ions, such as Zn2+ and Pb4+ , replaced Si4+ or Al3+ ions and held in the crystalline phase through the bonding strength of crystal systems. Correspondingly, the stabilities of heavy metals strongly depended on the type of the crystal which had different bonding strength. The bonding strength of anorthite systems could reach 189.45–208.62 × 104 kJ/mol [31], while the value of wollastonite was only 2.75–5.75 × 104 kJ/mol [32]. Anorthite which consisted in the 950 and 1000 ◦ C samples was an ideal crystalline matrix for the immobilization of heavy metals in the production of glass–ceramics materials.
This study was supported by the National High-tech R&D Program (863 Program) of China (Nos. 2009AA064704 and 2007AA06z348), the National Natural Science Fund of China (No. 50978071) and the State Key Laboratory of Urban Water Resource and Environment, Harbin Institute of Technology (No. 2011TS01). The authors also appreciate the National Innovation Team Supported by the National Science Foundation of China (No. 50821002).
4. Conclusion The presented MMR with double-layer structure was used for preparing glass–ceramics from sewage sludge. Glass–ceramics based on CaO–Al2 O3 –SiO2 system was developed successfully. Attractive physical and chemical properties of the microwaveprocessed glass–ceramics were observed, such as higher bending strengths (86.5–93.4 MPa) and lower thermal expansion coefficient (5.29 × 10−6 /◦ C). The leaching tests of heavy metals in the glass–ceramics showed that 950 and 1000 ◦ C samples which contained anorthite as main crystal immobilized the heavy metal ions effectively. A model of crystal growth in the microwave field was developed in this study and the results indicated that microwave heating generated a latent heat release, which suppressed the crystals growth, causing an improvement in the glass–ceramics properties.
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The roles of various plasma species in the plasma and plasma-catalytic removal of low-concentration formaldehyde in air Xing Fan, Tianle Zhu ∗ , Yifei Sun, Xiao Yan School of Chemistry and Environment, Beihang University, No. 37 Xueyuan Road, Haidian District, Beijing 100191, China
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Article history: Received 5 May 2011 Received in revised form 9 September 2011 Accepted 10 September 2011 Available online 16 September 2011 Keywords: Formaldehyde Discharge plasma MnOx /Al2 O3 catalyst Plasma-catalysis Ozone
a b s t r a c t The contributions of various plasma species to the removal of low-concentration formaldehyde (HCHO) in air by DC corona discharge plasma in the presence and absence of downstream MnOx /Al2 O3 catalyst were systematically investigated in this study. Experimental results show that HCHO can be removed not only by short-living active species in the discharge zone, but also by long-living species except O3 downstream the plasma reactor. O3 on its own is incapable of removing HCHO in the gas phase but when combined with the MnOx /Al2 O3 catalyst, considerable HCHO conversion is seen, well explaining the greatly enhanced HCHO removal by combining plasma with catalysis. The plasma-catalysis hybrid process where HCHO is introduced through the discharge zone and then the catalyst bed exhibits the highest energy efficiency concerning HCHO conversion, due to the best use of plasma-generated active species in a two-stage HCHO destruction process. Moreover, the presence of downstream MnOx /Al2 O3 catalyst significantly reduced the emission of discharge byproducts (O3 ) and organic intermediates (HCOOH). © 2011 Elsevier B.V. All rights reserved.
1. Introduction Low-concentration formaldehyde (HCHO), mainly originating from furnishing and building materials, widely exists in indoor environments, and has been correlated to adverse health effects such as eye, nose and throat irritation, allergic asthma, pulmonary function damage and even cancer [1–3]. Conventional technologies, such as adsorption, catalytic and photo-catalytic oxidation, have been widely investigated for the removal of HCHO [4–9]. Complete oxidation of 100-ppm HCHO was achieved over a Pt/TiO2 catalyst even at room temperature [6]. However, the limited removal capacities of adsorbent materials and the expensive cost of noble metals still limit their widespread applications. As an alternative approach, non-thermal plasma (NTP) technology is likely to be more appropriate for indoor air purification because it is capable of removing various indoor pollutants such as particulate matters, bacteria and volatile organic compounds (VOCs) simultaneously under ambient conditions. Atmospheric plasma discharges generate high-energy electrons, while the background gas remains close to room temperature. The energetic electrons excite, dissociate and ionize gas molecules, producing chemically active species for removal of VOCs [10]. NTP process alone, however, has disadvantages such as high energy consumption, poor selectivity towards total oxidation and
∗ Corresponding author. Tel.: +86 10 82314215; fax: +86 10 82314215. E-mail address:
[email protected] (T. Zhu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.044
undesired byproduct formation [11–13]. An attempt to overcome these disadvantages is to combine plasma technology with heterogeneous catalysis [10,14–18]. Fan et al. [17] reported that the BTX (mixture of benzene, toluene and p-xylene) conversion can be greatly enhanced by introducing MnOx /Al2 O3 catalyst after the discharge zone, at the same time harmful O3 can be removed from the exit gas stream. On the other hand, despite the excellent performance of the plasma-catalysis hybrid process [19], the underlying mechanisms involved in the plasma-catalytic destruction of VOCs are not well understood yet. Since various active species, such as energetic electrons, radicals, ions, stable and metastable molecules/atoms, coexist in the discharge plasma, the study of the contributions of these species to the removal of VOCs can bring valuable information about the destruction processes. Harling et al. [20] investigated the role of O3 in the destruction of toluene and cyclohexane using the combination of NTP with either MnO2 or MnO2 –CuO catalyst. It was deduced that the decomposition of O3 to active oxygen species over the catalyst surface initiated heterogeneous destruction reactions, leading to complete removal of toluene and cyclohexane. In the present study, different plasma and plasma–catalyst hybrid systems were designed for the removal of lowconcentration HCHO in air. A link tooth wheel-cylinder energized by positive DC high voltage was used as the plasma reactor while MnOx /Al2 O3 was selected as the post-plasma catalyst. The main objective of this work is to investigate the roles of different plasma species in the conversion of HCHO, both in the gas phase and over the catalyst surface. Besides, the energy efficiency concerning
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Fig. 1. Schematic diagram of the experimental set-up.
HCHO conversion, and the behavior of byproduct formation, including O3 and organic byproducts, were also investigated for the plasma and plasma-catalysis hybrid processes. 2. Experimental 2.1. Experimental set-up A schematic diagram of the experimental system is shown in Fig. 1. It consists of a tandem plasma–catalyst reactor system, a 25 kV/5 mA positive DC high voltage power supply, reaction gas supply and analytical instrumentation. A stainless steel cylinder with an inner diameter of 42 mm was used as the ground electrode of the plasma reactor, while a stainless steel rod (o.d. 6 mm) through which 9 discharge teeth wheels were linked with a space interval of 10 mm was used as the high voltage electrode. The effective discharge length and discharge gap were 89 and 16 mm, respectively. A stainless steel cylinder with an inner diameter of 27 mm was connected to the link tooth wheel-cylinder plasma reactor in series, in order to construct a plasma–catalyst hybrid system or a plasma alone system by introducing the MnOx /Al2 O3 catalyst or not. In order to investigate the roles of different plasma species in the conversion of HCHO, six treatment systems with different gasfeeding methods or configurations were designed, as shown in Fig. 2. Treatment systems A–C were constructed for homogeneous reaction of HCHO while systems D–F also included heterogeneous reactions over the MnOx /Al2 O3 catalyst. HCHO was introduced before the plasma in systems A and D, while in other systems HCHO was directly introduced into the post-plasma cylinder. In systems C and F, a buffer flask with capacity of 7 L (gas residence time of 70 s) was inserted between the two cylinders in order to extinguish active species from the plasma except O3 (O3 has a lifetime of several days at room temperature). 2.2. Experimental methods Gaseous HCHO and water vapor were introduced by passing dry air through two temperature-controlled bubble towers, containing 36 wt.% HCHO solution and deionized water, respectively. The water vapor was first mixed with the dilution air and then with the gaseous HCHO before or after the plasma reactor (Fig. 2). The flow rates of air were controlled by a set of mass flow controllers and the total flow rate of the reaction gas was controlled at 6.0 L/min. The
initial HCHO concentration and relative humidity (RH) of the reaction gas were 2.2 ± 0.1 ppm and 30 ± 3%, respectively. All the tests were carried out at room temperature and atmospheric pressure. MnOx /Al2 O3 (5 wt.% Mn) prepared by the impregnation method, with manganese acetate as the precursor and -Al2 O3 of 2.5–3.5 mm in diameter as the support, was used as the catalyst in this study [17]. Loading amount of the catalyst was 15.0 g, with the residence time in the catalyst bed being around 0.21 s. When the catalyst was introduced, the reaction was started by energizing the plasma reactor with DC only when the HCHO concentration at the exit of the catalytic reactor reached a steady state, meaning that initial adsorption–desorption equilibrium of HCHO over the catalyst surface was achieved. HCHO, HCOOH and O3 were sampled at the exit of the catalytic reactor and then analyzed by acetylacetone spectrophotometric method [21], ion chromatography (ICS200; Dionex Corporation, USA), and indigo disulphonate spectrophotometric method [17], respectively. HCHO and HCOOH were absorbed by distilled water while O3 by the solution of indigo disulphonate. The conversion of HCHO and the yield of HCOOH are defined as follows. HCHO (%) =
[HCHO]off − [HCHO]on × 100 [HCHO]off
(1)
[HCOOH] × 100 [HCHO]off − [HCHO]on
(2)
YHCOOH (%) =
where [HCHO]on and [HCHO]off denote the concentrations of HCHO (ppm) at the exit of the catalytic reactor with the power supply for the plasma reactor being on and off; and [HCOOH] is the concentration of HCOOH (ppm) measured at the exit of the catalytic reactor. The energy efficiency concerning HCHO conversion (E , the removed amount of HCHO per kWh of energy consumption, g/kWh) was also evaluated in this study to determine the capability of the treatment processes for practical applications. Since the specific input energy (SIE, the discharge energy deposited into 1 L of reaction gas, J/L) directly reflects the energy consumption of the plasma discharge, all the experimental results were compared on the basis of SIE in this investigation. SIE and E are calculated as follows. SIE = E =
UI × 60 Q [HCHO]off − [HCHO]on 30 × 3.6 × SIE 24.4
(3) (4)
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Fig. 2. Schematic diagram of the treatment systems: (A) in-plasma; (B) post-plasma; (C) ozone; (D) in-plasma + catalyst; (E) post-plasma + catalyst; (F) ozone + catalyst.
where U is the applied voltage (kV) measured with a high voltage probe (NISSIN EP-50K); I is the discharge current (mA) calculated by measuring the voltage across a 10 resistor with a multimeter; Q is the flow rate of the reaction gas (L/min); 30 is the molar mass of HCHO (g/mol); 24.4 is the molar volume of gas (L/mol) under the ambient condition (around 298 K); and 60 and 3.6 are the unit conversion coefficients. 3. Results and discussion 3.1. HCHO conversion Fig. 3 summarizes the conversion of HCHO as a function of SIE for treatment systems A, B, and D–F. The conversion was measured after the reaction reached the steady state. No HCHO conversion was observed in system C up to an SIE of 100 J/L in this study, although comparable O3 was detected at the outlet of system C to those of systems A and B. This result validates that the reaction
of HCHO with O3 in gas phase is negligible [22]. As seen from Fig. 3, no HCHO conversion occurred without discharge power (SIE = 0) whether the catalyst was introduced or not, showing that discharge plasma is indispensable for the removal of HCHO. The MnOx /Al2 O3 catalyst cannot be activated for HCHO oxidation under ambient conditions. For all the five treatment systems, the HCHO conversion increased with the increase of SIE. And compared with plasma alone (systems A and B), the presence of post-plasma MnOx /Al2 O3 catalyst (systems D–F) significantly enhances the removal of HCHO. The HCHO conversion for different systems is in the order of D > E > F A > B under the same SIE. 3.1.1. Plasma conversion of HCHO Generally, VOCs can be removed by discharge plasma via three pathways, i.e., direct electron impacts, gas-phase radical attacks and ion collisions. Results presented in Fig. 3 show that HCHO can be removed not only in the plasma zone (system A), but also in the post-plasma cylinder (system B). For an SIE of 20 J/L, the conversion of HCHO was 36% and 29% for systems A and B, respectively. Considering that unstable plasma species, such as energetic electrons and some gas-phase radicals, cannot reach the postplasma reactor because of their millisecond lifetimes [23,24], the removal of HCHO in system B can only be attributed to relatively stable (not including O3 ) and/or metastable plasma species (e.g. N2 metastable states). In fact, it has been reported that for plasma removal of HCHO, N2 metastable states may be more important than electrons owing to their longer lifetime [25]. These excited states of N2 contribute to the removal of HCHO via two possible pathways: direct attacks towards HCHO molecules and indirect reactions through the O2 dissociation processes, as shown in Eqs. (5)–(9) [25,26]. N2 (A3 + u ) + HCHO → H + HCO + N2
Fig. 3. Effects of specific input energy on the HCHO conversion.
(5)
N2 (a ) + HCHO → H + HCO + N2
(6)
3 3 N2 (A3 + u ) + O2 → O( P) + O( P) + N2
(7)
N2 (a ) + O2 → O(3 P) + O(3 P) + N2
(8)
O + HCHO → OH + HCO
(9)
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where N2 (a ) denotes the three states of a1 , a1 and ω1 , with a mean energy of 8.4 eV; while the energy of N2 (A3 + u ) is 6.2 eV for the = 0 level. In the case of system A in this study, besides the efficient removal of HCHO in the discharge zone, further removal of the unreacted HCHO can be expected in the downstream cylinder too, due to the residual long-living plasma-excited species.
3.1.2. Plasma-catalytic conversion of HCHO As shown in Fig. 3, HCHO can be more efficiently removed in the plasma–catalyst hybrid systems (D–F) when compared with the plasma alone processes (systems A and B). For an SIE of 20 J/L, the conversion of HCHO was 87%, 76% and 72% for systems D, E and F, respectively. This result suggests that the downstream MnOx /Al2 O3 catalyst can be effectively activated by long-living plasma species for HCHO conversion at room temperature and the heterogeneous oxidation reactions over the catalyst surface are much more important for HCHO removal than the homogeneous reactions in the discharge zone. As stated previously, not only short-living unstable reactive species are produced in plasma discharges, a fraction recombines to form more stable species such as O3 [21,27]. Comparison of the HCHO conversion in systems C and F shows that although O3 does not oxidize HCHO in the gas phase, it does initiate the removal of HCHO over the MnOx /Al2 O3 catalyst. O3 catalytic decomposition mechanism research shows that O3 decomposition over manganese oxide catalyst produces atomic oxygen and peroxide as the intermediate species [28,29]. These highly active oxygen species should be mainly responsible for the catalytic oxidation of HCHO in system F. Besides O3 , other long-living plasma-excited species, which account for the HCHO removal in system B, can also exist in the second cylinder of system E. These species may trigger both homogeneous and heterogeneous reactions of HCHO, resulting in higher HCHO conversion in system E than in system F under the same conditions. The test results in Fig. 3 also show that the treatment system D behaves the best in terms of HCHO conversion in this study, which can be easily attributed to the best use of the plasma-generated active species in a two-stage HCHO destruction process: firstly, HCHO was attacked by energetic electrons and reactive species in the discharge zone; secondly, unreacted HCHO from the discharge zone was further removed in the post-plasma stage mainly via catalytic processes initiated by O3 and also other long-living active species. Comparison of the HCHO conversion in systems D, E and F shows that the O3 initiated catalytic oxidation reactions play a significant role in the plasma-catalytic removal of HCHO. Moreover, it should be noticed that for the three hybrid systems, the difference in HCHO conversion is not significant for an SIE lower than 3 J/L. We may have to consider that this phenomenon occurs for the following reasons. On the one hand, the production of high energy long-living species, such as N2 (A3 + u ) and N2 (a ), is very limited in low-energy discharge plasma. Therefore, the conversion of HCHO in system E may only result from the catalytic ozonation process just as that in system F. On the other hand, it can be seen from Fig. 3 that the heterogeneous reactions are much more important than the homogeneous reactions towards HCHO conversion, especially in the case of low SIE. This probably explains the small difference in HCHO conversion between systems D and E. Nevertheless, the difference in HCHO conversion among the three hybrid systems becomes remarkable at higher SIE, with the increasing production of high-energy long-living species and also higher HCHO conversion in the discharge zone.
Fig. 4. Effects of specific input energy on the energy efficiency concerning HCHO conversion.
In summary, HCHO can be removed not only by short-living active species in the discharge zone, but also by long-living species except O3 downstream the plasma reactor. Compared with the plasma alone processes, the tandem plasma–catalyst hybrid systems perform much better in HCHO conversion, mainly arising from the O3 initiated heterogeneous destruction of HCHO over the post-plasma MnOx /Al2 O3 catalyst. 3.2. Energy efficiency Fig. 4 presents the energy efficiency concerning HCHO conversion as a function of SIE for treatment systems A, B, and D–F. It can be seen that the energy efficiency decreased with the increase of SIE for all the five treatment systems. Obviously, the concentration of HCHO in the gas stream decreased with the increase of HCHO conversion, resulting in lower collision probability between HCHO molecules and active species and hence lower removed amount of HCHO per kWh of energy consumption at higher SIE. Meanwhile, more of the energy in plasma was converted into heat, photons, and used for byproduct formation (as shown in Fig. 5) with the increase of SIE. Higher SIE favors the complete removal of HCHO (Fig. 3) but causes serious energy inefficiency of the process. Therefore, maximum available value for input power in the plasma reactor will be determined not only by HCHO conversion but also by the
Fig. 5. Effects of specific input energy on the emission of O3 .
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energy efficiency. Nevertheless, compared with the plasma alone processes (systems A and B), the energy efficiency was greatly improved by introducing post-plasma MnOx /Al2 O3 catalyst (systems D–F), indicating that the plasma-catalysis hybrid processes have higher HCHO removal capability and are more promising for practical applications. The maximum energy efficiency was 3.1, 3.1 and 2.5 g/kWh for systems D, E and F, respectively, compared to 0.9 g/kWh for system A and 0.3 g/kWh for system B. 3.3. Byproduct formation 3.3.1. Ozone As long as the NTP process is operated in air-like mixtures, the formation of O3 , a hazardous discharge byproduct, is unavoidable. Fig. 5 shows the O3 outlet concentration as a function of SIE for treatment systems A and D. In fact, O3 outlet concentrations were measured in both the presence and absence of HCHO in air in this study for all treatment systems. Results prove that low levels of HCHO in the gas stream as well as the introduction of a buffer flask hardly influence the O3 outlet concentration. As seen from Fig. 5, the O3 outlet concentration increased with the increase of SIE for both plasma alone and plasma-catalysis hybrid processes. Compared with plasma alone, however, the presence of post-plasma MnOx /Al2 O3 catalyst significantly reduced the O3 emission. For an SIE of 20 J/L, the O3 outlet concentration decreased from 57.2 ppm for system A to 13.9 ppm for system D. It is clear that mainly O3 induced by gas discharge was decomposed catalytically over the MnOx /Al2 O3 surface, producing highly active oxygen species which play a key role in the enhanced removal of HCHO in the plasma-catalysis hybrid processes (Fig. 3). Nevertheless, it should be noticed that even in the presence of MnOx /Al2 O3 catalyst, the O3 emission (13.9 ppm for an SIE of 20 J/L) is still high. In future research, it will be tested if the simultaneous catalytic removal of HCHO and O3 can be further improved by introducing catalysts that are more reactive towards O3 decomposition. 3.3.2. Formic acid Formic acid (HCOOH) is a common intermediate produced during the HCHO oxidation process [5,22]. Outlet concentrations of HCOOH were measured for treatment systems A and D in this study to investigate the influence of downstream MnOx /Al2 O3 catalyst on the formation of decomposition byproducts. In order to obtain measurable concentrations of HCOOH, a higher initial HCHO concentration of 40.9 ± 0.5 ppm was used. Fig. 6a and b shows the outlet concentration and the yield of HCOOH as functions of SIE, respectively. As seen from Fig. 6a, the HCOOH concentration at the outlet of system A increased with the increase of SIE, indicating that HCOOH was indeed produced as a byproduct in the plasma decomposition of HCHO and the absolute production of HCOOH increased with the increase of HCHO removed at higher SIE. On the contrary, the HCOOH outlet concentration of system D linearly decreased with the increase of SIE and was much lower than that of system A under the same SIE. For an SIE of 80 J/L, the HCOOH outlet concentration was 2.0 and 0.1 ppm for systems A and D, respectively. The difference in HCOOH production between the two systems indicates that HCOOH produced in the discharge zone can be effectively removed over the downstream MnOx /Al2 O3 catalyst, especially at higher SIE. The presence of post-plasma MnOx /Al2 O3 catalyst does not only significantly enhance the conversion of HCHO (Fig. 3), but also favors the elimination of organic byproducts. In addition, results presented in Fig. 6b show that the HCOOH yield decreased with the increase of SIE for both systems A and D. Since the absolute removal of HCHO increased with the increase of SIE, the decreasing HCOOH yield in system D can be easily attributed to its decreasing production of HCOOH, as shown in
Fig. 6. Effects of specific input energy on the production of HCOOH: (a) HCOOH concentration and (b) HCOOH yield.
Fig. 6a. On the other hand, although the production of HCOOH increased with the increase of SIE in system A (Fig. 6a), the HCOOH yield decreased monotonously, suggesting that more of the removed HCHO tends to undergo further oxidation reactions in higher-energy discharge plasma to form end products such as CO2 and CO.
4. Conclusions The roles of various plasma species in the plasma and plasma-catalytic removal of low-concentration HCHO in air were experimentally studied in this work. The main findings can be summarized as follows:
(1) Both short- and long-living plasma species (other than O3 ) contribute to HCHO removal in the gas phase. (2) O3 does not initiate HCHO removal in the gas phase but does trigger heterogeneous destruction of HCHO over the MnOx /Al2 O3 catalyst, well explaining the greatly enhanced HCHO conversion by combining plasma with the MnOx /Al2 O3 catalyst in series. (3) The best use of plasma-generated active species for HCHO destruction can be achieved in a plasma–catalyst hybrid system where HCHO is introduced through the discharge zone and then the catalyst bed, leading to the highest energy efficiency concerning HCHO conversion. (4) The introduction of MnOx /Al2 O3 catalyst after the plasma reactor significantly reduces the emission of discharge byproducts (O3 ) and organic intermediates (HCOOH), showing great potential for indoor VOCs’ purification.
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Journal of Hazardous Materials 196 (2011) 386–394
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Biodegradation of chlorobenzoic acids by ligninolytic fungi ˇ Milan Muzikáˇr a,b , Zdena Kˇresinová a,c , Kateˇrina Svobodová a , Alena Filipová a , Monika Cvanˇ carová a,c , a a,∗ Kamila Cajthamlová , Tomáˇs Cajthaml a b c
Institute of Microbiology, Academy of Sciences of the Czech Republic, v.v.i., Vídenská 1083, CZ-142 20 Prague 4, Czech Republic ˇ Institute of Chemical Technology Prague, Faculty of Food and Biochemical Technology, Technická 5, CZ-160 28 Prague 6, Czech Republic Institute of Environmental Studies, Faculty of Science, Charles University, Benátská 2, CZ-128 01 Prague 2, Czech Republic
a r t i c l e
i n f o
Article history: Received 27 May 2011 Received in revised form 23 August 2011 Accepted 10 September 2011 Available online 16 September 2011 Keywords: Chlorobenzoic acid Polychlorinated biphenyls Biodegradation White rot fungi Irpex lacteus
a b s t r a c t We investigated the abilities of several perspective ligninolytic fungal strains to degrade 12 mono-, diand trichloro representatives of chlorobenzoic acids (CBAs) under model liquid conditions and in contaminated soil. Attention was also paid to toxicity changes during the degradation, estimated using two luminescent assay variations with Vibrio fischeri. The results show that almost all the fungi were able to efficiently degrade CBAs in liquid media, where Irpex lacteus, Pycnoporus cinnabarinus and Dichomitus squalens appeared to be the most effective in the main factors: degradation and toxicity removal. Analysis of the degradation products revealed that methoxy and hydroxy derivatives were produced together with reduced forms of the original acids. The findings suggest that probably more than one mechanism is involved in the process. Generally, the tested fungal strains were able to degrade CBAs in soil in the 85–99% range within 60 days. Analysis of ergosterol showed that active colonization is an important factor for degradation of CBAs by fungi. The most efficient strains in terms of degradation were I. lacteus, Pleurotus ostreatus, Bjerkandera adusta in soil, which were also able to actively colonize the soil. However, in contrast to P. ostreatus and I. lacteus, B. adusta was not able to significantly reduce the measured toxicity. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Chlorinated organic pollutants are a class of serious environmental contaminants because of their environmental persistence and ecotoxicity. Chlorinated benzoic acids (CBAs) are widespread environmental pollutants resulting primarily from microbial biodegradation of polychlorinated biphenyls (PCBs), reviewed, e.g., in Field and Alvarez [1], and some herbicides [2]. CBAs are significantly more soluble than their parent compounds and can therefore enter into the aqueous phase from the contaminated soil of polluted sites. Some mono-, di-, and tri-CBAs have been shown to cause genomic damage to tobacco plants [3], and to be toxic to aquatic organisms such as ciliate, Daphnia, algae and fish [4–6]. Several mono, di and trichlorinated isomers were also found to possess estrogenic-disrupting activity [7]. CBAs represent crucial recalcitrant metabolites on the biphenyl pathway during bacterial PCB transformation. Although it was found that CBAs are not very toxic toward bacteria, substantial negative effects of their presence on the bacterial transformation of PCBs have been reported
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[8,9]. Moreover, soil bacteria that co-metabolize PCBs via the main biphenyl upper pathway tend to accumulate CBAs as dead-end products because they are generally unable to further transform these substrates [10]. Another great limitation of organopollutant bacterial biodegradation is the fact that bacterial degrading enzymes are usually intracellular and the transfer of the pollutant into the bacterial cell represents an important limiting step. On the other hand, ligninolytic fungi, with their extracellular low-substrate-specificity enzymes, represent a promising alternative for biodegradation of various aromatic pollutants [11]. The ligninolytic system consists of three major peroxidases: lignin peroxidase (LiP), manganese peroxidase (MnP), versatile peroxidases and laccase, which belong among phenoloxidases. Their degradative abilities have been documented e.g., for chlorophenols, polycyclic aromatic hydrocarbons, PCBs, dioxins, furans, endocrine disrupters and others [12–15]. Moreover, the fungi were shown to be capable of splitting the aromatic rings of various persistent pollutants [14,16]. In contrast to a number of articles dealing with bacterial CBA degradation, only few papers have been published describing potential degradation of these compounds by fungi. Kamei et al. identified 4-CBA acid after degradation of 4,4 dichlorobiphenyl by Phanerochaete sp. MZ 142 and suggested its further transformation via a reductive pathway [14]. Other authors showed that ortho and meta mono-CBAs and benzoic acid (BA)
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significantly induced the activities of cytochrom P-450 from Phanerochaete chrysosporium [17]. BA was then proven to be transformed by microsomes containing P-450 from the same fungus. The role of the monooxygenases of P-450 was clarified earlier by other authors, when the enzyme was heterologically expressed and hydroxyl benzoate and protocatechuic acid were detected as the degradation products of benzoate [18]. Generally, the data published in the literature document was efficiency of fungi to degrade PCBs and suggest possible transformation of PCBs to CBAs. CBAs are critical metabolites on the bacterial degradation pathway mainly, due to high specificity of individual bacterial enzymes. Therefore it is reasonable to investigate also CBA degradation abilities of ligninolytic fungi, especially when these organisms represent a promising alternative to bacterial PCB degradation applications. The aim of this work was to investigate the abilities of several promising ligninolytic fungal strains to transform 12 representatives of CBAs with various degree of chlorination (mono-, di-, tri-CBAs). The degradation performance was tested in model liquid nutrient media, where the production of CBA degradation products, the activities of ligninolytic enzymes and changes in the acute toxicity were also monitored. Moreover, the applicability of the fungi was also tested in an artificially contaminated soil, where toxicity was also monitored. 2. Materials and methods 2.1. Materials Standards and chemicals. 2-CBA; 2,3-CBA; 3,4-CBA; 3,5CBA; 2,3,5-CBA; 2,4,6-CBA and HPLC internal standard 2,3dichlorophenol were obtained from Sigma–Aldrich (Steinheim, Germany). 3-CBA; 4-CBA; 2,4-CBA; 2,5-CBA and 2,6-CBA were from Merck (Darmstadt, Germany). 2,3,6-CBA was purchased from Supelco (Steinheim, Germany). All the compounds were employed without further purification to prepare stock solutions in dimethyl formamide as described below. All the solvents were purchased from Merck, Germany or Chromservis (Prague, Czech Republic) and were of p.a. quality, trace analysis quality or gradient grade. All the chemicals used for the biochemical studies were from Sigma–Aldrich (Steinheim, Germany). 2.2. Microorganisms, inocula preparation and enzyme activities measurement Fungal cultures, inocula preparation and degradation experiments. All of the ligninolytic fungal strains used in this study (Irpex lacteus 617/93, Bjerkandera adusta 606/93, Phanerochaete chrysosporium ME 446, Phanerochaete magnoliae CCBAS 134/I, Pleurotus ostreatus 3004 CCBAS 278, Trametes versicolor 167/93, Pycnoporus cinnabarinus CCBAS 595, Dichomitus squalens CCBAS 750) were obtained from the Culture Collection of Basidiomycetes of the Academy of Science, Prague. Fungal inocula were grown under stationary conditions for 7 d at 28 ◦ C in 250 mL Erlenmeyer flasks containing 20 mL of either complex malt extract-glucose (MEG) medium or low-nitrogen mineral medium (LNMM). MEG medium (pH 5.5) contained 5 g malt extract broth (Oxoid, UK) and 10 g glucose per liter of distillated water and LNMM contained 2.4 mM diammonium tartrate [19]. The cultures were then homogenized with the Ultraturrax-T25 (IKA-Labortechnik, Staufen, Germany) and this suspension was used for inoculation in the degradation experiments. Enzyme determination. LiP (E.C. 1.11.1.14) was assayed with veratryl alcohol as the substrate [20] and MnP (E.C. 1.11.1.13) was determined with 2,6-dimethoxyphenol [21]. Laccase (Lac, E.C. 1.10.3.2) was estimated with
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2,2-azinobis-3-ethylbenzo-thiazoline-6-sulfonic acid as the substrate [22]. Manganese-independent peroxidase (MIP) was calculated from the peroxidase activity of MnP assay detected in the absence of Mn2+ ions. One unit of enzyme produced 1 mol of the reaction product per minute under the assay conditions at room temperature. 2.3. Degradation of CBAs in liquid media The degradation experiments in the liquid media were performed as static cultures, incubated in 250 mL Erlenmeyer flasks in five parallel experiments at 28 ◦ C. Twenty milliliters of the respective medium (MEG or LNMM) was inoculated with a 5% suspension of homogenized pre-inocula (1 mL) of the respective fungal strain. The cultures were immediately spiked with a solution of the CBAs in dimethyl formamide (100 L). The final amount of each CBA was 200 g per flask. The heat-killed controls were performed with oneweek growth of fungal cultures, which were killed in an autoclave before addition of the CBA solution. All of the cultures were incubated in the darkness at 28 ◦ C and harvested after 7, 14 and 21 days. 2.4. Fungal treatment of contaminated soil For preparation of the soil degradation experiment, 1.0 mL aliquots of a mycelia suspension of each fungal strain were added to 16 cm × 3.5 cm test-tubes containing 10 g of commercial straw pellets (ATEA Praha, Prague, Czech Republic), the moisture contents of which had been previously adjusted to 70% (w/w) and subsequently sterilized by autoclaving (121 ◦ C, 45 min). After inoculation, the cultures were closed with cotton–wool stoppers and then grown for 14 d at 28 ◦ C [23] The colonized substrate was then covered with a layer of soil (20 g), which had been previously artificially spiked with a mixture of CBAs in acetone. The relevant controls were prepared in the same way, however, without fungal inoculation. Main properties of the used sandy–loamy soil were as follows: total organic carbon 0.8%, total organics 1.4%, pH 5.3, water-holding capacity 31% and the granulometric composition was: sand 50.9%, fine sand 31.2%, silt 10.8%, clay 7.1%. The soil was air-dried and sieved through a 2-mm mesh before contamination and the final concentration of each CBA in the soil after contamination was 10 g/g. The soil samples were then moistened to 15% humidity. The tubes were incubated at 28 ◦ C and the samples were harvested after 30 and 60 d. All the respective controls and samples were performed in five replicates. 2.5. Extraction and quantitative analyses of CBAs The whole content of each liquid culture was homogenized with Ultraturrax and acidified to approximately pH 2. It was then extracted with five 10 mL portions of ethyl acetate, the extracts were dried with sodium sulfate and concentrated using a rotary evaporator to a final volume of 10 mL. The extraction recoveries of all the CBAs were better than 95%. To enable HPLC analysis, an aliquot of the ethyl acetate extract was mixed with acetonitrile in a ratio of 1:10 (v/v), and the mixture was used for injection [16]. The soil samples were submitted to extraction using a Dionex 200 ASE extractor (Palaiseau, France). The soil samples (3 g) were mixed with sodium sulfate before extraction (v/v) and the extraction conditions were: 3 cycles; 150 ◦ C; 10.34 MPa; solvent system hexane–acetone, 1% acetic acid [24]. To avoid CBA volatilization, 500 L of DMSO was added to the extracts as a solvent stopper and the extracts were concentrated using a vacuum rotary evaporator (60 kPa, 40 ◦ C) to approximately 1.5 mL. 50 L of internal standard (IS, 2,3-dichlorophenol 0.9 mg/mL in ACN) was added to
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each sample and the IS was used to calculate the volume extracts. The mixture was then directly injected into the HPLC. The quantitative analyses were performed using the Alliance Waters system (Prague, Czech Republic) equipped with a PDA detector and Empower software was used for data processing. Separation of the CBA mixture was performed on an XBridge C18 column (250 mm × 4.6 mm I.D., particle size 3.5 m) from Waters (Prague, Czech Republic). The separation was carried out with a gradient (v/v) of acetonitrile (B) and water solution (A) of 0.1% TFA. The gradient program was as follows (min/%B): 0/17; 30/17; 60/34; 70/50. The applied flow rate was 0.8 mL min−1 and temperature was 35 ◦ C [24]. 2.6. Qualitative analyses of CBA degradation products Qualitative analysis of the CBA degradation products was performed in that the degradation intermediates were separated and characterized or identified by gas chromatography–mass spectrometry (GC–MS; 450-GC, 240-MS ion trap detector, Varian, Walnut Creek, CA). The ethyl acetate extracts were injected both directly without any derivatization and also after trimethylsilylation with N,O-bis(trimethylsilyl)trifluoroacetamide (BSTFA, Merck, Germany) and methylation with diazomethane [25]. The GC instrument was equipped with a split/splitless injector maintained at 240 ◦ C. DB-5MS column (Agilent, Prague, Czech Republic) was used for the separations (30 m, 0.25 mm I.D., 0.25 mm film thickness). The temperature program was started at 60 ◦ C and was held for 1 min in the splitless mode. Then the splitter was opened with ratio 1:50. The oven was heated to 120 ◦ C at a rate of 25 ◦ C/min with a subsequent temperature ramp up to 240 ◦ C at a rate of 2.5 ◦ C/min, where this temperature was maintained for 20 min. The solvent delay time was set at 5 min and the transfer line temperature was set at 240 ◦ C. The mass spectra were recorded at 3 scans s−1 under electron impact at 70 eV and mass range 50–450 amu. The excitation potential for the MS/MS product ion mode employed was 0.2 V and was increased to 0.8 V for more stable ions. Acetonitrile was used as the medium for chemical ionization (CI), where the ionization maximum time was 2000 and 40 s for the reaction. 2.7. Analyses of ergosterol The total ergosterol was extracted and analyzed as described previously [23]. Briefly, samples (0.5 g) were sonicated with 3 mL of 10% KOH in methanol at 70 ◦ C for 90 min. Distilled water (1 mL) was added and the samples were extracted three times with 2 mL of cyclohexane, evaporated under nitrogen, redissolved in methanol and analyzed isocratically using a Waters Alliance HPLC system (Waters Milford, MA) equipped with a LiChroCart column filled with LiChrosphere® 100 RP-18e (250 × 4.0 mm; particle size 5 m; ˚ equilibrated with 100% methanol at a flow rate pore size 100 A) of 1 mL min−1 . Ergosterol was detected at 282 nm and quantified with a 5-point calibration curve over a linear range from 0.5 to 50.0 g/mL. 2.8. Toxicity assay The luminescent bacteria Vibrio fischeri (strain NRRL-B-11177), which was used for all the toxicity tests, were purchased freeze-dried from the supplier Ing. Musial (Czech Republic). The freeze-dried bacteria were rehydrated and stabilized in 2% (w/v) NaCl solution at 15 ◦ C for 1 h according to the standard procedure ISO 2007 [26]. An acute toxicity test of samples after degradation in liquid media was performed using the corresponding ethyl acetate extracts. Aliquots of the extracts (0.5 mL) were evaporated to dryness and dissolved again in dimethyl sulfoxide, which was directly applied to the test (2% of DMSO in the reaction mixture).
The amount of dimethyl sulfoxide varied between media, due to different sensitivities of the test toward the media matrix (see below). Three replicates for each sample were used to carry out the ecotoxicity test. The luminescence readings were obtained with a Lumino M90a luminometer (ZD Dolní Újezd, Czech Republic) at a temperature of 15 ± 0.2 ◦ C. The inhibition of bioluminescence was recorded after 15-min exposure. The toxicity of soil samples was measured by a kinetic Flash assay using the luminescent bacterium [27,28]. The samples were prepared by weighing 1.5 g dried soil and 6 mL 2% (w/v) NaCl solution. The sample suspension was mixed continuously and 0.5 mL was placed into the measuring cuvette. The contents of the measuring cuvette were mixed continuously by adapted luminometer LUMINO M90a and 0.5 mL of the bacterial solution was dispensed into the sample. The signal was recorded permanently for 60 s. The light inhibition was calculated as the difference between the height of the peak that was observed immediately after addition of the bacteria to the sample and the luminescence intensity after a contact time of 60 s.
3. Results and discussion 3.1. Degradation of CBAs in liquid cultures The representatives of mono, di and tri-CBAs that were tested in this study were employed at a relatively high concentration of 10 g/mL. The fact that the compounds are partially soluble in water and their acute toxic properties were confirmed by the observation of fungal biomass development. Generally, the fungal strains were partly affected by CBAs and their biomass reached about 50–70% compared to non toxic controls (data not shown). This finding is in agreement with the observation of Dittmann et al. who tested the development of the mycelia of fungal strains in two liquid nutrient media after the addition of various concentrations of 3-CBA [29]. In contrast to this observation, the fungal strains in our study were very efficient in degradation of CBAs in the liquid media. The time course of the individual CBA degradation in both media is presented in Tables 1 and 2. The results clearly demonstrate that all of the tested strains were at least partially able to transform CBAs. I. lacteus, P. cinnabarinus and D. squalens were found to be the most efficient degraders in complex MEG media. P. cinnabarinus and D. squalens were able to degrade about 78% and 73% of total CBA, respectively, while I. lacteus degraded 92% of total CBAs in complex media compared to the heat-killed controls. Particularly I. lacteus removed all of the CBAs from the media except 2,6-CBA and 2,3,6-CBA, which were degraded to approximately 50% of original amount, while P. cinnabarinus did not significantly transform 2,6-CBA, 2,3,6-CBA, 2,4,6-CBA and D. squalens did not significantly degrade 2,3,6-CBA (ANOVA, P = 0.05). B. adusta exhibited the poorest degradation ability in both the liquid media. The most efficient strains in the LNNM media were once again found to be I. lacteus, P. cinnabarinus and D. squalens. All of the three most efficient strains were able to transform CBAs with percent removals ranging from 76% to 77%. 2,6-CBA, 2,3,6-CBA and 2,4,6-CBA again appeared to be the most recalcitrant compounds where D. squalens did not significantly degrade 2,6-CBA and 2,4,6-CBA; P. cinnabarinus – 2,6-CBA, 2,3,6-CBA and 2,4,6-CBA; I. lacteus – 2,6-CBA and 2,3,6-CBA. These results suggest a possible connection between the substituted ortho and para positions and persistency toward the fungal degradation mechanism. Only a few publications in journals deal with transformation of CBAs by fungi. The above-mentioned work of Dittmann et al. also included the degradation of 3-CBA by P. chrysosporium, P. ostreatus, Heterobasidion annosum and two other ectomycorrhizal fungi [29]. However, in contrast to our results, the authors observed only limited degradation of the compound in the range of several
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Table 1 Residual CBA amounts in heat-killed controls and after incubation of the tested fungal strains in LNNM media (ND: not detected). Amount of CBA (g per flask) LNNM – 7 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor LNNM – 14 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor LNNM – 21 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor
2-CBA
3-CBA
4-CBA
2,3-CBA
2,4-CBA
2,5-CBA
2,6-CBA
3,4-CBA
3,5-CBA
2,3,5-CBA
2,3,6-CBA
2.4.6-CBA
196 ± 4 156 ± 5 ND ND ND 27 ± 2 156 ± 9 ND 73 ± 39
207 ± 4 53 ± 12 33 ± 4 19 ± 6 ND ND 28 ± 3 ND 25 ± 1
207 ± 2 41 ± 34 ND ND ND ND ND ND ND
195 ± 3 174 ± 5 ND 45 ± 7 94 ± 14 133 ± 2 162 ± 8 ND 25 ± 3
207 ± 2 156 ± 31 ND 66 ± 2 ND 74 ± 4 105 ± 17 ND 104 ± 8
193 ± 3 141 ± 11 ND 85 ± 5 92 ± 15 126 ± 6 167 ± 8 ND 46 ± 3
205 ± 2 197 ± 8 211 ± 15 179 ± 18 196 ± 5 189 ± 3 195 ± 5 190 ± 8 242 ± 11
182 ± 8 64 ± 19 ND ND ND ND 87 ± 60 ND ND
196 ± 1 77 ± 28 ND ND ND 165 ± 2 55 ± 1 ND ND
165 ± 2 161 ± 2 ND 153 ± 2 136 ± 10 ND 124 ± 46 ND ND
199 ± 1 185 ± 7 ND 176 ± 21 177 ± 7 184 ± 8 172 ± 14 176 ± 12 206 ± 14
185 ± 0 176 ± 8 207 ± 6 128 ± 11 171 ± 6 180 ± 4 163 ± 13 173 ± 7 140 ± 4
170 ± 15 124 ± 5 ND ND ND ND 105 ± 1 ND 66 ± 13
173 ± 14 45 ± 2 56 ± 4 ND ND ND ND ND 47 ± 4
184 ± 21 ND ND ND ND ND ND ND ND
168 ± 19 157 ± 22 126 ± 8 13 ± 2 25 ± 1 100 ± 11 132 ± 1 ND ND
208 ± 17 85 ± 5 ND ND ND 60 ± 19 15 ± 2 ND ND
167 ± 18 114 ± 7 ND ND 40 ± 1 77 ± 13 159 ± 3 ND ND
181 ± 16 198 ± 14 189 ± 11 173 ± 7 188 ± 12 182 ± 15 193 ± 11 184 ± 6 186 ± 34
156 ± 17 46 ± 0 ND ND ND ND ND ND ND
165 ± 19 ND ND ND ND 136 ± 23 ND ND ND
138 ± 19 145 ± 5 ND 103 ± 3 92 ± 0 ND 114 ± 6 ND ND
169 ± 17 187 ± 7 ND 171 ± 9 175 ± 12 153 ± 32 188 ± 5 163 ± 8 183 ± 3
159 ± 21 179 ± 10 175 ± 11 95 ± 4 165 ± 8 151 ± 27 161 ± 13 136 ± 41 136 ± 3
196 ± 5 112 ± 10 ND ND ND ND 70 ± 13 ND 76 ± 5
195 ± 9 54 ± 5 71 ± 1 ND ND ND ND ND 32 ± 15
204 ± 8 ND ND ND ND ND ND ND ND
189 ± 3 167 ± 8 105 ± 3 19 ± 2 20 ± 2 24 ± 5 101 ± 19 ND ND
208 ± 6 89 ± 2 ND 22 ± 0 ND 35 ± 5 ND ND ND
191 ± 5 106 ± 5 ND ND 40 ± 6 20 ± 2 130 ± 25 ND ND
203 ± 6 201 ± 14 186 ± 5 170 ± 1 185 ± 24 176 ± 9 182 ± 36 196 ± 7 195 ± 13
180 ± 15 ND ND ND ND ND ND ND ND
188 ± 5 ND ND ND ND ND ND ND ND
170 ± 7 130 ± 8 ND 67 ± 7 98 ± 4 129 ± 7 105 ± 18 ND ND
195 ± 7 190 ± 13 ND 183 ± 1 180 ± 18 152 ± 12 179 ± 39 175 ± 9 194 ± 8
183 ± 7 182 ± 12 191 ± 4 71 ± 6 168 ± 13 153 ± 11 157 ± 30 176 ± 8 134 ± 11
percent, even though, in one case, the authors employed a similar concentration (15.6 mg/L) to that used in our study (20 mg/L). In order to employ the toxicity test, we diluted the samples from the two media in different ways. The theoretical (original) concentrations of the individual CBAs in the reaction mixture for MEG and
LNNM media samples were 0.5 and 0.25 g/mL, respectively. Since we detected only a decrease in the toxicity in preliminary tests, the dilution of the samples was set to reach about 90% inhibition for the controls. The evaluation of the acute toxicity test with V. fischeri was performed by comparison of the inhibition of the sample
Table 2 Residual CBA amounts in heat-killed controls and after incubation of the tested fungal strains in MEG media (ND: not detected). Amount of CBA (g per flask) MEG – 7 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor MEG – 14 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor MEG – 21 days Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor
2-CBA
3-CBA
4-CBA
2,3-CBA
2,4-CBA
2,5-CBA
2,6-CBA
3,4-CBA
3,5-CBA
2,3,5-CBA
2,3,6-CBA
2,4,6-CBA
211 ± 4 201 ± 20 ND ND ND ND 71 ± 9 ND 90 ± 22
195 ± 4 138 ± 64 97 ± 6 ND ND 44 ± 4 ND ND ND
209 ± 2 162 ± 69 ND ND ND ND ND ND ND
197 ± 2 200 ± 8 ND ND 52 ± 5 56 ± 7 127 ± 7 ND 42 ± 2
207 ± 6 192 ± 25 131 ± 11 46 ± 2 ND ND 66 ± 5 ND 89 ± 2
195 ± 6 189 ± 17 ND 70 ± 27 ND 47 ± 10 145 ± 9 ND 39 ± 1
206 ± 4 200 ± 9 205 ± 20 204 ± 4 209 ± 7 213 ± 7 214 ± 14 201 ± 12 202 ± 7
195 ± 5 154 ± 39 ND ND ND ND ND ND ND
199 ± 4 175 ± 32 ND ND ND ND ND ND ND
175 ± 2 166 ± 13 ND 86 ± 5 ND 76 ± 6 132 ± 10 ND ND
172 ± 6 177 ± 8 151 ± 1 151 ± 4 167 ± 8 172 ± 7 156 ± 11 167 ± 18 154 ± 2
193 ± 5 193 ± 7 160 ± 13 114 ± 11 185 ± 6 188 ± 5 150 ± 14 190 ± 15 174 ± 3
210 ± 30 208 ± 3 ND ND ND ND 37 ± 9 ND 84 ± 16
186 ± 32 191 ± 2 ND ND ND 34 ± 1 ND ND ND
197 ± 24 211 ± 1 213 ± 3 ND ND ND ND ND ND
199 ± 27 162 ± 51 101 ± 6 ND ND 40 ± 8 91 ± 19 ND ND
206 ± 26 203 ± 7 124 ± 6 ND ND 39 ± 5 58 ± 3 ND ND
195 ± 21 194 ± 1 ND ND ND 60 ± 15 117 ± 19 ND ND
212 ± 25 202 ± 12 ND 98 ± 10 226 ± 4 221 ± 2 231 ± 11 188 ± 17 197 ± 5
176 ± 26 175 ± 3 ND ND ND ND ND ND ND
181 ± 25 186 ± 1 ND ND ND ND ND ND ND
164 ± 28 151 ± 10 ND ND ND 86 ± 17 115 ± 15 ND ND
178 ± 26 148 ± 28 145 ± 3 107 ± 12 161 ± 5 162 ± 4 170 ± 7 151 ± 25 150 ± 7
193 ± 24 192 ± 1 166 ± 1 ND 181 ± 2 189 ± 11 150 ± 4 170 ± 23 169 ± 6
209 ± 15 140 ± 53 ND ND ND ND 22 ± 1 ND 121 ± 30
175 ± 17 47 ± 2 ND ND ND ND ND ND ND
188 ± 11 42 ± 9 197 ± 5 ND 93 ± 19 29 ± 8 ND ND 136 ± 22
186 ± 13 173 ± 26 ND ND ND 55 ± 9 86 ± 6 ND ND
193 ± 9 133 ± 46 74 ± 14 ND ND 46 ± 1 61 ± 4 ND ND
182 ± 8 138 ± 54 54 ± 9 ND ND 52 ± 3 126 ± 14 ND ND
213 ± 9 200 ± 6 ND 88 ± 20 194 ± 23 236 ± 3 200 ± 13 187 ± 27 206 ± 11
172 ± 10 80 ± 13 ND ND 59 ± 5 56 ± 0 ND ND ND
177 ± 9 105 ± 10 ND ND 75 ± 4 ND ND ND ND
157 ± 6 137 ± 15 ND ND ND 82 ± 1 119 ± 12 ND ND
166 ± 12 170 ± 4 140 ± 2 96 ± 9 142 ± 19 174 ± 3 187 ± 23 150 ± 25 160 ± 5
181 ± 8 183 ± 7 151 ± 14 ND 167 ± 14 195 ± 2 171 ± 18= 162 ± 34 172 ± 6
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Table 3 Luminescence inhibition in heat-killed controls and after incubation of the tested fungal strains in MEG and LNNM media after 21 days of incubation. Luminescence inhibition (%)
LNMM
Control B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor
89.4 72.1 11.9 27.2 44.9 41.9 27.5 11.3 14.5
± ± ± ± ± ± ± ± ±
MEG 8.2 4.3 8.5 13.9 15.4 10.1 8.3 2.7 4.0
94.5 85.9 35.0 7.4 38.2 52.8 34 .8 20.8 36.3
COOCH3 COOCH3
COOH ± ± ± ± ± ± ± ± ±
1.0 12.5 9.4 3.0 30.7 17.7 2.8 5.6 7.5
Clx COOH
Clx
CHO HO luminescence with their respective controls (Table 3). As mentioned above, the toxicity test revealed that the tested fungi were generally able to decrease the measured acute toxicity, suggesting that the degradation products of CBAs were either not accumulated or they were less toxic than the original CBAs.
HO
Clx COOCH3
CH3O
Clx
Clx
Clx
CHO
CH2OH 3.2. Detection of CBA degradation products The CBA metabolites were analyzed with GC–MS and their structures were suggested by comparing the mass spectra with the data in the NIST 08 library and independently by interpreting the fragmentation pattern. Additionally, unknown structures of metabolites were explored using MS/MS (product ion scan) to clarify the fragmentation sequence. The mass spectral characteristics of the detected CBA degradation products are listed in Table 4. Some of the metabolites were detected after trimethylsilylation (e.g., chlorobenzyl alcohols) and several of them were confirmed by comparison with the available chemical standards. All of the intermediates were detected at only trace levels, suggesting that none of them were accumulated during degradation. The group of the detected intermediates includes chlorobenzaldehydes, chlorobenzyl alcohols, chlorobenzoic acid methyl esters and the methoxy or hydroxy derivatives of these structures. The metabolites were found in various fungal strain cultures when representatives of monochloro and dichloro benzaldehydes and alcohols were found in all cultures, as well as methyl esters of di-CBA. Methyl ester representative of tri-CBA was detected only in the culture of I. lacteus, however, methoxy derivatives of tri-CBA and di-CBA methyl esters were found in all fungi. Trichlorinated hydroxybenzyl alcohols were detected in all fungal cultures too. A possible scheme of CBA fungal degradation pathway constructed of the detected metabolites is shown in Fig. 1. The results generally correspond to the results of Kamei et al. [14], who studied the transformation of 4,4 -dichlorobiphenyl by Phanerochaete sp. MZ 142, where these authors detected the formation of 4-CBA, the methyl ester of 4-CBA and further reduced transformation products: 4-chlorobenzyl alcohol and 4-chlorobenzaldehyde. Such a reduction mechanism could be explained by the action of an intracellular aryl alcohol oxidase system [30]. Matsuzaki et al. showed that the enzymes that are probably involved in the transformation, i.e. aryl alcohol dehydrogenase, aryl aldehyde dehydrogenase and also cytochrome P-450 of P. chrysosporium were up-regulated after the addition of BA to the fungal culture [31]. The other types of transformation products, i.e. hydroxyl and methoxy derivatives, which were found in our study, have already been described by Matsuzaki and Wariishi following transformation of BA by P. chrysosporium [18]. The detected metabolites include 4-hydroxy, 2-hydroxy and 4-hydroxy-2methoxy derivatives. In another work, the authors demonstrated that heterologously expressed P-450 cytochromes from the CYP53 family of P. chrysosporium, Aspergilus niger and Rhodotorula minuta were able to hydroxylate BA at the 4-position [32]. P-450-mediated
CH3O
Clx
CH2OH
Clx
CH2OH HO
Clx
CH3O
Clx
Fig. 1. Proposed pathway of CBA degradation by ligninolytic fungi.
hydroxylation of BA at other positions in fungi has not been reported to date. Moreover, the authors employed quantitative PCR to demonstrate that the expression of the cytochrome is regulated by the presence of BA at the transcription level. The induction of cytochrome P-450 by BA and also by 3 and 4-CBA has been published elsewhere [17]. The measurement of fungal extracellular ligninolytic activities in this study demonstrated that most of the activities were suppressed or the maxima activity peaks were delayed during cultivation by the presence of CBAs. Only rare cases when the situation was different were recorded for the activities of MnP and laccase of T. versicolor, which were significantly induced in MEG and LNNM media, respectively. Particularly the laccase activity increased from 20 U/L to 230 U/L. These findings indirectly confirm that ligninolytic enzymes need not play a key role in the degradation of CBAs. 3.3. Degradation of CBAs in soil The soil degradation experiment was monitored after 30 and 60 days and the residual concentrations after the application of the fungal strains are depicted in Fig. 2. The results show that, except for strains P. cinnabarinus and T. versicolor, which degraded only 30% and 39% of the total CBA, respectively, within 60 days of incubation, all the other strains under study were able to substantially remove CBAs from soil in the range of 85–99% of total CBA. The results are partially contrary to the experiments in liquid cultures, because P. cinnabarinus belonged among the most degrading strains in both the liquid media. On the other hand, B. adusta appeared to be effective in soil while this strain belonged among the less degrading in the liquid media. I. lacteus was found to be the most efficient
Table 4 Retention data and electron impact mass spectral characteristics of CBA metabolites. MW according to CI
m/z of fragment ions (relative intensity)
Structural suggestion
5.431 5.494 5.582 7.603 7.724 7.913 8.329 8.532 8.875 9.445 9.685 10.883 11.203 11.729 11.894 12.308
140 140 140 174 174 174 174 174 214 214 214 204 204 204 204 204
o-Chlorobenzaldehyde m-Chlorobenzaldehyde p-Chlorobenzaldehyde 3,5-Dichlorobenzaldehyde 2,4-Dichlorobenzaldehyde 2,5-Dichlorobenzaldehyde 2,3-Dichlorobenzaldehyde 3,4-Dichlorobenzaldehyde TMS p-chlorobenzyl alcohol TMS m-chlorobenzyl alcohol TMS o-chlorobenzyl alcohol 2,6-Dichlorobenzoic acid methyl ester 3,5-Dichlorobenzoic acid methyl ester 2,4-Dichlorobenzoic acid methyl ester 2,5-Dichlorobenzoic acid methyl ester 3,4-Dichlorobenzoic acid methyl ester
12.651
204
12.945 13.139 13.27 13.604 14.091
170 248 248 248 238
14.288 15 15.801
248 248 238
16.391
238
18.625
310
18.699
298
18.741
310
19.695
234
19.82 20.46 20.592
186 298 310
20.688 21.057 21.526 22.693 23.481
298 298 234 298 234
27.313 29.207
268 312
142 (23.8), 141 (36.6), 140 (73.7), 139 (99.9), 111 (55.0), 75 (32.0), 51 (19.6), 50 (29.8) 142 (20.6), 141 (36.1), 140 (66.6), 139 (99.9), 113 (18.4), 77 (22.7), 75 (33.5), 74 (19.1) 142 (16.0), 141 (37.1), 140 (49.4), 139 (99.9), 113 (16.8), 111 (49.6), 77 (15.1), 74 (16.8) 176 (62.4), 174 (70.3), 173 (99.9), 145 (47.0), 139 (54.2), 111 (50.1), 75 (61.1), 74 (52.6) 176 (39.5), 175 (70.2), 174 (61.4), 173 (99.9), 147 (16.9), 145 (25.9), 75 (18.0), 74 (15.0) 176 (38.2), 175 (68.4), 174 (61.3), 173 (99.9), 111 (25.0), 75 (61.0), 74 (45.9), 50 (25.3) 176 (37.5), 175 (69.0), 174 (62.8), 173 (99.9), 147 (21.9), 145 (37.4), 75 (36.4), 74 (26.3) 176 (38.9), 175 (69.5), 174 (64.9), 173 (99.9), 147 (28.1), 145 (43.0), 75 (29.1), 74 (24.9) 201 (34.0), 199 (99.9), 179 (18.5), 163 (30.7), 127 (25.4), 125 (82.5), 89 (25.1), 73 (18.6) 201 (31.8), 199 (90.9), 179 (33.4), 171 (19.9), 169 (60.1), 127 (30.2), 125 (99.9), 89 (32.7) 201 (20.8), 199 (58.7), 179 (24.2), 169 (20.5), 127 (31.9), 125 (99.9), 89 (24.2), 73 (12.5) 206 (32.5), 204 (43.4), 177 (9.9), 175 (63.4), 173 (100), 147 (9.6), 145 (7.9), 109 (7.8), 75 (33.9) 208 (2.0), 206 (14.1), 204 (20.5), 177 (8.9), 175 (63.1), 173 (100), 147 (21.1), 145 (33.2), 109 (17.8), 75 (16.0) 208 (5.4), 206 (12.9), 204 (20.7), 177 (9.9), 175 (61.1), 173 (100), 147 (15.5), 145 (29.0), 109 (16.2), 75 (95.6) 208 (3.7), 206 (18.3), 204 (25.7), 177 (11.9), 175 (62.1), 173 (93.3), 147 (6.4), 145 (19.3), 109 (17.1), 75 (100) 208 (2.7), 206 (19.8), 204 (32.4), 177 (10.3), 175 (64.6), 173 (100), 147 (19.7), 145 (35.1), 109 (22.9), 74 (27.6) 208 (1.3), 206 (16.3), 204 (19.0), 177 (19.0), 175 (56.9), 173 (100), 149 (29.4), 147 (47.7), 145 (19.0), 109 (14.4), 75 (97.4) 172 (19.4), 171 (35.0), 170 (56.3), 169 (100), 141 (6.8), 126 (13.6), 111 (11.7), 77 (15.5) 235 (67.6), 233 (99.9), 205 (18.5), 203 (25.6), 161 (41.4), 159 (64.2), 123 (18.4), 103 (18.9) 235 (58.0), 233 (84.2), 161 (61.1), 159 (99.9), 123 (13.8), 103 (29.1), 73 (12.7) 235 (70.9), 233 (99.9), 205 (32.3), 203 (45.8), 161 (58.5), 159 (89.7), 147 (27.7), 123 (21.1) 242 (6.4), 240 (19.3), 238 (19.7), 211 (29.3), 209 (95.0), 207 (100), 183 (3.4), 181 (14.1), 179 (14.5), 143 (11.0), 109 (12.8), 74 (12.9) 235 (69.6), 233 (99.9), 205 (17.8), 203 (25.3), 161 (46.3), 159 (71.1), 123 (17.5), 103 (22.8) 235 (68.5), 233 (94.6), 203 (16.7), 161 (67.3), 159 (99.9), 75 (27.5), 73 (28.5), 59 (34.2) 242 (7.1), 240 (25.7), 238 (25.1), 211 (31.5), 209 (98.1), 207 (100), 183 (5.7), 181 (12.6), 179 (14.5), 143 (10.5), 109 (11.7), 74 (14.1) 242 (9.3), 240 (29.4), 238 (29.5), 211 (29.7), 209 (95.4), 207 (100), 183 (6.8), 181 (21.6), 179 (20.7), 143 (14.2), 109 (15.6), 74 (14.5) 271 (32.4), 269 (100), 267 (76.8), 241 (15.8), 239 (45.5), 237 (42.8), 197 (13.7), 195 (46.6), 193 (44.2), 157 (14.2), 125 (5.5), 123 (15.8), 93 (15.9) 285 (33.2), 283 (100), 281 (99.5),239 (49.1), 237 (50.6), 209 (80.6), 207 (83.3), 205 (2.8), 165 (13.3), 167 (33.3) 271 (32.8), 269 (100), 267 (87.2), 241 (13.0), 239 (38.3), 237 (37.2), 197 (22.0), 195 (68.5), 193 (65.9), 157 (14.6), 125 (5.6), 123 (15.7), 93 (28.1) 238 (4.3), 236 (41.1), 234 (67.2), 207 (15.4), 205 (84.6), 203 (100), 162 (14.1), 160 (18.6), 111 (15.0), 97 (31.9) 188 (7.2), 186 (25.8), 157 (31.6), 155 (100), 127 (17.8), 99 (13.7) 285 (31.5), 283 (100), 281 (98.0), 239 (73.8), 237 (36.1), 209 (68.2), 207 (68.5), 167 (12.6),165 (33.8) 271 (35.5), 269 (100), 267 (99.1), 241 (10.0), 239 (31.1), 237 (31.1), 197 (26.8), 195 (81.5), 193 (82.1), 157 (16.7), 125 (6.4), 123 (18.0), 93 (39.2) 285 (31.6), 283 (100), 281 (95.3), 239 (73.8), 237 (34.6), 209 (47.7), 207 (43.3), 167 (14.8),165 (33.0) 285 (34.4), 283 (100), 281 (98.5), 239 (36.8), 237 (32.4), 209 (51.5), 207 (54.4), 167 (13.2),165 (35.6) 238 (3.0), 236 (22.3), 234 (30.4), 207 (10.2), 205 (68.4), 203 (100), 162 (5.3), 160 (8.3), 111 (11.2), 97 (17.0) 285 (26.1), 283 (100), 281 (90.9), 239 (52.1), 237 (49.7), 209 (74.2), 207 (64.2), 167 (19.7),165 (48.1) 238 (5.1), 236 (21.3), 234 (29.5), 207 (11.5), 205 (70.5), 203 (100), 162 (10.9), 160 (11.9), 111 (15.4), 97 (19.9) 270 (15.1), 268 (15.3), 241 (28.3), 239 (95.8), 237 (100) 312 (12.3), 314 (13.6), 301 (42.6), 299 (100), 297 (97,3), 271 (24.3), 269 (76.7), 267 (73.4), 227 (31.8), 225 (75.6), 223 (88.5)
Type of derivatization
Trimethylsilylation Trimethylsilylation Trimethylsilylation Trimethylsilylation
2,3-Dichlorobenzoic acid methyl ester ?-Chloro-?-methoxybenzaldehyde TMS 2,5-dichlorobenzyl alcohol TMS 2,4-dichlorobenzyl alcohol TMS 3,5-dichlorobenzyl alcohol 2,4,6-Trichlorobenzoic acid methyl ester
Trimethylsilylation Trimethylsilylation Trimethylsilylation Trimethylsilylation
TMS 2,3-dichlorobenzyl alcohol TMS 3,4-dichlorobenzyl alcohol 2,3,6-Trichlorobenzoic acid methyl ester
Trimethylsilylation Trimethylsilylation Trimethylsilylation
2,3,5-Trichlorobenzoic acid methyl ester TMS ?,?,?-trichlorobenzyl alcohol
Trimethylsilylation
TMS ?,?,?-trichloro-?-hydroxybenzyl alcohol
Trimethylsilylation
TMS ?,?,?-trichlorobenzyl alcohol
Trimethylsilylation
M. Muzikáˇr et al. / Journal of Hazardous Materials 196 (2011) 386–394
tR (min)
?,?-Dichloro-?-methoxybenzoic acid methyl ester ?-Chloro-?-hydroxybenzoic acid methyl ester TMS ?,?,?-trichloro-?-hydroxybenzyl alcohol TMS ?,?,?-trichlorobenzyl alcohol
Methylation Trimethylsilylation Trimethylsilylation
TMS ?,?,?-trichloro-?-hydroxybenzyl alcohol TMS ?,?,?-trichloro-?-hydroxybenzyl alcohol ?,?-Dichloro-?-methoxybenzoic acid methyl ester TMS?,?,?-trichloro-?-hydroxybenzyl alcohol ?,?-Dichloro-?-methoxybenzoic acid methyl ester
Trimethylsilylation Trimethylsilylation
?,?,?-Trichloro-?-methoxybenzoic acid methyl ester TMS ?,?,?-trichloro-?-methoxybenzyl alcohol
Trimethylsilylation
Trimethylsilylation
391
392
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Fig. 2. Residual concentrations of CBAs in contaminated soil after incubation with the tested fungal strains: A – 30 days; B – 60 days.
strain in soil, where this fungus had already depleted 98% of the CBA within 30 days. One probable explanation for the discrepancy between these results and model conditions in the liquid media and the soil degradation experiment lies in the different abilities of fungi to penetrate into contaminated soil [33]. Therefore, we tried to estimate the relative amount of fungal biomass using the analysis of ergosterol in soil samples with CBAs and also without the addition of pollutants (Fig. 3). Despite the high variability of the data, the results indicate that the fungi that showed the highest CBA depletion (I. lacteus, P. ostreatus, and B. adusta), were also the strains that were able to significantly colonize the contaminated soil. The only exception was P. magnolia, where we detected a
significantly lower amount of ergosterol despite the high removal of CBAs (99% within 60 days). The parameters of kinetic Flash toxicity assay were adjusted to also include recording of a possible increase in toxicity. The data obtained from this test are depicted in Fig. 4. The best results, in terms of inhibition reduction, were obtained with strains I. lacteus and P. ostreatus, corresponding to their CBA degradation efficiencies. On the other hand, unexpected results were observed with strains B. adusta and P. magnolia where, in spite of their high degradation rate, the detected residual toxicity was not significantly different from the controls (t-test, P < 0.05). These results of Flash assay are in agreement with the results from toxicity estimation in
Fig. 3. Ergosterol concentrations in non-contaminated soil and in soil contaminated by CBAs after 30 and 60 days of incubation.
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393
Fig. 4. Luminescence inhibition of the Flash assay in contaminated soil (control) and in soil with the tested fungal strains after 60 days of incubation.
the liquid cultures, where a residual toxicity in these fungal cultures was also detected. This could possibly be explained by the formation and accumulation of toxic metabolites and, probably for the same reason, significantly elevated toxicity was observed for T. versicolor. 4. Conclusion To the best of our knowledge, this is the first paper providing a general description of the ability of ligninolytic fungi to biodegrade CBAs that represent crucial toxic and highly persistent metabolites on bacterial biodegradation pathways of polychlorinated biphenyls. The ability of the fungi has been examined under liquid conditions and also verified in contaminated soil. The tested fungal strains were able to degrade CBAs in soil in the 85–99% range within 60 days when I. lacteus was found to be the most efficient degrading strain under both of the tested conditions. Several new degradation products have been identified when mainly methoxy and hydroxy derivatives were produced together with reduced forms of the original acids. The results show that the fungi are probably able to transform CBAs via several pathways with significant reduction of toxicity during the process. The promising degradation results from this study emphasize the need for further research, especially to identify the participation of different enzymatic machineries, in order to improve the understanding of the degradation mechanisms. The results for the liquid media and from the consequent soil experiment show that the presence of a bioremediative organism is of key importance; however, in soil, i.e. under conditions with limited pollutant bioavailability, active colonization of the soil is of equal importance. Acknowledgments This work was supported by grants no. 2B06156 of the Ministry of Education, Youth and Sports of the Czech Republic and no. 525/09/1058 of the Science Foundation of the Czech Republic and by Institutional Research Concept No. AV0Z50200510. References [1] J.A. Field, R.S. Alvarez, Microbial transformation and degradation of polychlorinated biphenyls, Environ. Pollut. 155 (2008) 1–12.
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Journal of Hazardous Materials 196 (2011) 395–401
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Comparison of o-toluidine degradation by Fenton, electro-Fenton and photoelectro-Fenton processes Jin Anotai a , Somporn Singhadech b , Chia-Chi Su c , Ming-Chun Lu c,∗ a National Center of Excellence for Environmental and Hazardous Waste Management, Department of Environmental Engineering, Faculty of Engineering, King Mongkut’s University of Technology Thonburi, Bangkok 10140,Thailand b Department of International Postgraduate Programs in Environmental Management (Hazardous Waste Management), Chulalongkorn University, Bangkok 10330, Thailand c Department of Environmental Resources Management, Chia-Nan University of Pharmacy and Science, Tainan 717, Taiwan
a r t i c l e
i n f o
Article history: Received 27 May 2011 Received in revised form 5 September 2011 Accepted 11 September 2011 Available online 22 September 2011 Keywords: Box–Behnken design Electro-Fenton process Hydroxyl radicals Photoelectro-Fenton process o-Toluidine
a b s t r a c t A Box–Behnken design (BBD) statistical experimental design was used to investigate the degradation of o-toluidine by the electro-Fenton process. This method can be used to determine the optimal conditions in multivariable systems. Fe2+ concentration (0.2–1.0 mM), H2 O2 concentration (1–5 mM), pH (2–4), and current (1–4 A) were selected as independent variables. The removal efficiencies for o-toluidine and chemical oxygen demand (COD) were represented by the response function. Result by 2-level factorial design show that the pH and the Fe2+ and H2 O2 concentrations were the principal parameters. Among the main parameters, the removal efficiencies for o-toluidine and COD were significantly affected by pH and Fe2+ concentration. From the Box–Behnken design predictions, the optimal conditions in the electro-Fenton process for removing 90.8% of o-toluidine and 40.9% of COD were found to be 1 mM of Fe2+ and 4.85 mM of H2 O2 at pH 2. Under these optimal conditions, the experimental data showed that the removal efficiencies for o-toluidine and COD in the electro-Fenton process and the photoelectro-Fenton process were more than 91% and 43%, respectively, after 60 min of reaction. The removal efficiencies for o-toluidine and COD in the Fenton process are 56% and 27%, respectively. © 2011 Elsevier B.V. All rights reserved.
1. Introduction o-Toluidine is an important aromatic amine that is used in the dyestuffs and rubber industry. However, short-term exposure to otoluidine may induce methaemoglobinaemia, whereas long-term or repeated exposure to o-toluidine could be possibly carcinogenic to humans [1]; it may cause bladder cancer [2]. It is difficult to completely treat wastewater containing o-toluidine because of its resistance to biodegradation. Presently, advanced oxidation processes (AOPs) have been used for wastewater treatment, particularly in cases where the contaminant species are difficult to remove by biological or physicochemical processes [3–7]. AOPs are based on the generation of a powerful oxidant, the hydroxyl radical (• OH), which can react with most organic pollutants and then degrade them [8,9]. The Fenton process is one of the most widely used AOPs because of its low investment cost [10]. The Fenton reaction is shown below: Fe2+ + H2 O2 → • OH + Fe3+ + OH−
∗ Corresponding author. Tel.: +886 6 266 4911; fax: +886 6 266 3411. E-mail address:
[email protected] (M.-C. Lu). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.043
(1)
However, the iron sludge produced from the Fenton process requires further treatment and disposal, and this is a major disadvantage of this process. This major drawback can be resolved by coupling the Fenton process with electric discharge, the so-called “electro-Fenton (EF) process”. The advantage of the electrochemical Fenton process is that it produces much less iron sludge than the traditional Fenton process. In this process, ferric ions (Fe3+ ) are effectively electroregenerated to ferrous ions (Fe2+ ), as shown in Eq. (2); this can be expressed in terms of current efficiency. Fe3+ + e− → Fe2+
(2)
The capability of the electro-Fenton process has been confirmed by Harrington and Pletcher [11], with more than 90% chemical oxygen demand (COD) removal with current efficiencies higher than 50% and acceptable energy consumptions. The efficiency of the electro-Fenton process can be improved by using UV or visible light illumination in a process known as the photoelectro-Fenton (PEF) process. This improvement is due to the higher production rate of • OH from the photoreduction of Fe(OH)2+ (Eq. (3)) and the photodecomposition of complexes from Fe3+ reactions (Eq. (4)) [12–15] Fe(OH)2+ + hv → Fe2+ + • OH R(CO2 )-Fe
3+
+ hv →
R(• CO
2)
(3) + Fe(II) →
•R
+ CO2
(4)
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Normally, Fenton-type processes [16–18] are affected by the pH and initial Fe2+ and H2 O2 concentrations. To determine the optimal conditions for o-toluidine degradation and the effect of variables on the electro-Fenton process, the Box–Behnken design (BBD) was used in this investigation. The BBD can be used to find the optimal conditions in multivariable systems [19]. The statistical design of an experiment reduces the number of experiments that must be performed and corresponding time spent and can be used to optimize the operating parameters in multivariable systems. Few studies have used BBD for the degradation of azo-dyes and organic contaminants by the photo-Fenton process [19,20]. However, there have been no studies published on the use of BBD for the degradation of o-toluidine by the electro-Fenton process. In this study, the optimal conditions for o-toluidine degradation and the effect of four variables (pH, Fe2+ concentration, H2 O2 concentration and current) on the electron-Fenton process were investigated using BBD. As only several significant factors were involved in optimization, response surface methodology (RSM) was applied. Additionally, the o-toluidine degradation performances of ordinary Fenton, electroFenton and photoelectro-Fenton processes were also compared. 2. Materials and methods 2.1. Material and reactor o-Toluidine (99.5%, Merck), hydrogen peroxide (H2 O2 , 35%, Merck), and ferrous sulfate hepta-hydrated (FeSO4 ·7H2 O, Merck) were reagent grade and used without further purification. Fig. 1 shows the three kinds of reactors. The Fenton reactor was cylindrical stainless steel (diameter: 13 cm; height: 35 cm). The total volume of the reactor was 3.5 L. The electro-Fenton reactor, a cylindrical reactor, was operated in constant current mode. The anode was titanium net coated with RuO2 /IrO2 (DSA), and the cathode was stainless steel. The DSA anode with an inside diameter of 7 cm and height of 35 cm, and the cathode had an inside diameter of 2 cm and height of 35 cm. The electrodes were connected with direct current (DC) power. In the photoelectro-Fenton reactor, a set of 6 UV lamps fixed inside a cylindrical Pyrex tube (allowing wavelengths > 320 nm to penetrate) were used as the irradiation source. The UV lamps were connected to the power supply E-safe, 2003, Switching Power Supply (Max. 300 W), Model: LC-B300AT. 2.2. Analysis method In the photoelectro-Fenton experiment, synthetic wastewater containing 1 mM o-toluidine was prepared and then initial pH was adjusted with perchloric acid (HClO4 ). After pH adjustment, a predetermined amount of catalytic ferrous sulfate was added into the solution and then the UV lights were turned on. H2 O2 was also added in the same time to start the reaction. Additionally, in the electro-Fenton experiment, solution with 1 mM o-toluidine was prepared and then ferrous ions were added after the pH was adjusted to the desired value. In the meantime, the power supply was turned on, and hydrogen peroxide was added to initiate the reaction. Samples (1 mL) were taken at predetermined time intervals and were immediately injected into a tube containing sodium hydroxide solution to quench the Fenton reaction by increasing the pH to 11. The sample was then filtered (0.45 m filter) to remove precipitates and kept for 12 h before COD analysis. This process was used to avoid quantifying the effect of the H2 O2 concentration on the COD value. COD was analyzed by a closed reflux titrimetric method based on the standard methods [21]. The Fe2+ concentration was determined using the 1,10-phenanthroline method [22]. Total organic carbon was measured with an Elementar liquid TOC analyzer. The concentration of o-toluidine
was determined using high performance liquid chromatography (HPLC) with a Spectra system model SN4000 pump and Asahipak ODP-506D column (150 mm × 6 mm × 5 m). The detection limit of o-toluidine was 0.005 mM or 0.535 ppm. Organic acids were analyzed using a Dionex DX-120 ion chromatograph with an Ion Pac AS11 anion column at 30 ◦ C. 2.3. Experimental design BBD are a class of rotatable or nearly rotatable second-order designs based on three-level incomplete factorial designs. Among all the RSM designs, BBD requires fewer runs [23]. The DesignExpert software version 7.0 (Stat-Ease, Inc., Minneapolis, USA) was used to find the optimal conditions of o-toluidine degradation by the electro-Fenton process. The effects of the significant factors were determined by BBD. The significant factors and the appropriate studied ranges were pH: 2–4, Fe2+ concentration: 0.2–1.0 mM, and H2 O2 concentration: 1–5 mM. The concentration of o-toluidine was fixed at 1 mM for all experiments. 3. Results and discussion 3.1. Effect of various parameters on o-toluidine removal efficiency The Fe2+ concentration, H2 O2 concentration, pH and current were selected as experimental conditions for the BBD. The removal efficiencies for o-toluidine and COD were represented by a response function. Table 1 shows the two levels of the four factors on BBD. The values of variables, the experimental data and the results are presented in Table 2. The maximum removal rate of o-toluidine was 100% and the minimum was 23% (Table 2). When 1.0 mM of Fe2+ , 5.0 mM of H2 O2 and 1.0 A of current at pH 2 were applied, the otoluidine removal was 94.4% (run 2). However, as current increased from 1.0 A to 4.0 A, the removal of o-toluidine slightly increased to 100% (run 6). It was found that the amount of current was not sensitive in the applied range and therefore could be neglected. The correlation of o-toluidine and COD removal efficiencies obtained from BBD is shown in Table 1, where a higher correlation means that the parameter has a higher effect on o-toluidine and COD. The correlation can be as high as 1 or low as −1. The result indicates that the current has a slight effect on o-toluidine, with a correlation of only 0.098 (Table 1). The degradation of o-toluidine depended on the initial concentration of Fe2+ and H2 O2 , showing a high correlation in o-toluidine removal efficiency about 0.617 and 0.278 for Fe2+ and H2 O2 concentration, respectively. The same trend was also observed in COD removal efficiency. This correlation indicates that the Fe2+ and H2 O2 concentrations have a positive effect on the removal efficiencies for o-toluidine and COD, indicating that increasing Fe2+ and H2 O2 concentrations increased the removal efficiencies for o-toluidine and COD. The pH has a negative effect on o-toluidine and COD removal, and so the removal efficiencies for otoluidine and COD decreased with increasing pH of solution. From the correlation values, initial pH and the Fe2+ and H2 O2 concentrations were the most significant factors that affected o-toluidine and COD removal. Table 3 shows the levels of significant factors of o-toluidine and COD removal efficiencies. Results from the experiment revealed that the maximum removal of o-toluidine was 91.4% and that of COD was 42% (run 1) (Table 4). The correlation values indicate that pH had the most pronounced effect on o-toluidine and COD removal (−0.725 for o-toluidine and −0.593 for COD) (Table 3). The Fe2+ concentration had a comparable effect on these responses, and H2 O2 concentration had a greater effect on COD removal than the degradation of o-toluidine. Fig. 2 shows the response surface plot of the effect of pH and Fe2+ concentration on the o-toluidine and COD removal efficiencies. This
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397
Fig. 1. The experimental reactors.
Table 1 The two levels of variables and the value of correlation on o-toluidine and COD removal efficiency from Box–Behnken statistical design. Variables
Symbol
pH Fe2+ (mM) H2 O2 (mM) Current (A)
A B C D
Variable level
Correlation
Low
High
o-Toluidine
COD
2 0.2 1 1
4 1 5 4
−0.628 0.617 0.278 0.098
−0.517 0.633 0.274 0.259
Table 2 o-Toluidine and COD removal from the two levels of variables in electro-Fenton process with 1 mM of o-toluidine designed by the BBD. Run number
pH
Fe2+ (mM)
H2 O2 (mM)
Current (A)
o-Toluidine removal (%)
COD removal (%)
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16
4.0 2.0 4.0 4.0 2.0 2.0 2.0 2.0 4.0 4.0 2.0 2.0 4.0 4.0 2.0 4.0
0.2 1.0 0.2 1.0 0.2 1.0 1.0 1.0 1.0 0.2 0.2 0.2 0.2 1.0 0.2 1.0
5.0 5.0 5.0 5.0 5.0 5.0 1.0 1.0 1.0 1.0 5.0 1.0 1.0 5.0 1.0 1.0
4.0 1.0 1.0 4.0 1.0 4.0 1.0 4.0 4.0 1.0 4.0 1.0 4.0 1.0 4.0 1.0
17.0 94.4 32.0 56.0 42.0 100 66.4 68.5 45.0 23.0 74.0 40.0 26.5 61.4 56.4 48.0
23.0 57.0 22.0 33.02 8.0 59.5 33.0 46.0 34.0 19.0 38.4 22.5 27.0 36.0 34.6 31.0
plot shows the negative effect of pH on the removal efficiencies. The o-toluidine and COD removals decreased as the initial pH of the solution increased from 2.0 to 4.0 because the oxidation potential of hydroxyl radicals (• OH) and the dissolved fraction of iron species decreased [24,25]. The results also show that increasing the Fe2+ concentration can enhance o-toluidine and COD removal efficiencies because more Fe2+ reacts H2 O2 producing more • OH. Analysis of variance (ANOVA) tests for o-toluidine and COD removal were conducted to determine the suitability of the
response function and the significance of the effects of independent variables on the response function (Table 5). ANOVA indicates that the predictability of the model is at the 95% confidence level. Values of “Prob > F” less than 0.05 indicates a significant effect of the corresponding variable on the response. The result shows that the F-values of o-toluidine and COD removal were 11.10 and 8.06, respectively; imply that the model is significant. There are only 0.15% and 0.43% chances for o-toluidine and COD removal, respectively; that the model’s F-values this large could occur due to noise.
Table 3 The levels of significant factors and the value of correlation on o-toluidine and COD removal efficiency from BBD. Significant factor
Symbol
Variable level
Correlation
Low
Center
High
o-Toluidine
COD
pH Fe2+ (mM) H2 O2 (mM)
A B C
2 0.2 1
3 0.6 3
4 1 5
−0.725 0.484 0.294
−0.593 0.455 0.408
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Table 4 o-Toluidine and COD removal from the levels of significant factors in electro-Fenton process with 1 mM of o-toluidine and 1 A designed by the BBD. Run number
pH
Fe2+ (mM)
H2 O2 (mM)
o-Toluidine removal (%)
COD removal (%)
1 2 3 4 5 6 7 8 9 10 11 12 13
2.0 4.0 2.0 2.0 2.0 4.0 3.0 40 3.0 3.0 3.0 4.0 3.0
0.6 0.6 1.0 0.2 0.6 0.6 0.2 0.2 0.6 1.0 0.2 1.0 1.0
5.0 5.0 3.0 3.0 1.0 1.0 5.0 3.0 3.0 1.0 1.0 3.0 5.0
91.4 31.2 75.0 47.4 63.4 34.0 51.0 15.0 60.0 52.0 42.4 49.4 78.0
42.0 17.6 36.0 16.4 30.0 16.0 24.0 17.0 32.0 21.0 16.4 21.2 36.0
In this case, pH and Fe2+ were significant model terms affecting percent o-toluidine and COD removal. 3.2. The prediction of optimal conditions of o-toluidine degradation by BBD The goal of this part was to establish the optimal conditions for maximum removal of o-toluidine and COD by the electro-Fenton process. The BBD can provide an empirical relationship between the response function and the variables. The mathematical relationship between the removal of o-toluidine and the three significant variables can be approximated by quadratic polynomial equation, and the equations for the removal of o-toluidine and COD by the electro-Fenton process are shown below: o − toluidine removal (%) = 64.22 − 11.02A + 12.36B + 3.25C + 2.73AB − 1.02AC + 8.58BC − 1.17ABC
(5)
COD removal (%) = 30.16 − 2.31A + 7.47B + 0.84C + 0.91AB − 2.81AC + 8.03BC − 1.81ABC
of o-toluidine and COD at each value of pH, Fe2+ concentration and H2 O2 concentration. On the basis of the coefficients in Eqs. (5) and (6), it indicates that pH (A) and Fe2+ (B) concentration have negative and positive effects on o-toluidine and COD removal efficiencies, respectively. In other words, removal of o-toluidine and COD decreased with the pH (A) while increasing with Fe2+ (B) and H2 O2 (C) doses. Fe2+ dose had a more profound effect on o-toluidine and COD removal as compared to H2 O2 . The electro-Fenton process utilizes electrochemical generation of ferrous ions from ferric ions and ferric complexes. Ferrous ions were continuously recycled electrochemically, and therefore they were not depleted during the degradation of o-toluidine. Fe2+ concentration has the greatest effect on removal of o-toluidine with the largest coefficient (12.36). In this study, the removal of o-toluidine and COD were selected as “maximize” and then, Fe2+ and H2 O2 concentrations and pH were used as “within the range.” Consequently, these individual goals were combined into an overall desirability function by the software to find the best optimal conditions. The final equation relationship between the response function (o-toluidine and COD removal) and the key parameters can be determined by Eqs. (7) and (8).
(6)
Fe2+
where A, B and C are pH, concentration and H2 O2 concentration, respectively. The equations are used to calculate the removal
o − toluidine removal (%) = 78.74 − 18.45 × pH + 30.81 × [Fe2+ ] + 3.74 × [H2 O2 ]
(7)
Table 5 ANOVA tests for o-toluidine and COD removal by BBD. Source
Sum of squares
df
Mean squares
F-value
p-value Prob > F
o-Toluidine removal Model A (pH) B (Fe2+ ) C (H2 O2 ) AB AC BC ABC Residual Cor Total
7790.65 3387.24 3271.84 663.06 1.56 190.44 262.44 14.06 802.07 8592.72
7 1 1 1 1 1 1 1 8 15
112.95 3387.24 3271.84 663.06 1.56 190.44 262.44 14.06 100.26
11.10 33.78 32.63 6.61 0.016 1.90 2.62 0.14
0.0015 0.0004 0.0004 0.0330 0.9037 0.2055 0.1443 0.7178
Significant Significant Significant
COD removal Model A (pH) B (Fe2+ ) C (H2 O2 ) AB AC BC ABC Residual Cor Total
1808.81 552.25 826.56 155.00 52.56 119.90 68.89 33.64 256.41 2065.22
7 1 1 1 1 1 1 1 8 15
258.40 552.25 816.56 155.00 52.56 119.90 68.89 33.64 32.05
8.06 17.23 25.79 4.84 1.64 3.74 2.15 1.05
0.0043 0.0032 0.0010 0.0591 0.2362 0.0891 0.1808 0.3356
Significant Significant Significant
J. Anotai et al. / Journal of Hazardous Materials 196 (2011) 395–401
399
Fig. 2. Three-dimensional representation of the response surface plot of the effect of pH and Fe2+ concentration on (a) o-toluidine and (b) COD removal efficiency.
COD removal (%) = 30.41 − 6.58 × pH + 12.63 × [Fe2+ ] + 2.27 × [H2 O2 ]
(8)
According to Eqs. (7) and (8), at the optimal conditions of pH 2, 1 mM of Fe2+ and 4.85 mM of H2 O2 , the maximum removals of o-toluidine and COD were 90.8% and 40.9%, respectively. 3.3. Comparison between various processes The optimal conditions were used to investigate the removal efficiencies for o-toluidine and COD in the Fenton process, electroFenton process and photoelectro-Fenton process. The results are shown in Fig. 3. Fig. 3(a) shows that the removal efficiency for otoluidine in the three processes was almost the same in the first 2 min. After 2 min, the removal of o-toluidine in the Fenton process was slightly increased. The removal efficiency for o-toluidine was approximately 56% after 60 min of reaction. However, the removal efficiencies for o-toluidine in the electro-Fenton process and the photoelectro-Fenton process were 91% and 99%, respectively, after 60 min of reaction. The removal of o-toluidine was due to the formation of • OH via Eq. (1). Moreover, Fe3+ in the solution was able to regenerate inside the reactor when electric discharge and UV irradiation were used, allowing numerous Fe2+ react with H2 O2 to generate • OH. Ferrous ions are not depleted during the oxidation reaction, as shown in Eqs. (2)–(4). Therefore, the electro-Fenton process and the photoelectro-Fenton process can enhance the oxidation rate of o-toluidine. The same result was found for COD removal efficiency, as shown in Fig. 3(b). The removal efficiencies for COD in the three processes were similar in the first 2 min. However, the removal efficiency of COD was significantly different between the various processes after two min of reaction. The removal efficiencies for COD were 27% for the Fenton process, 45% for the electro-Fenton process and 43% for the photoelectro-Fenton process after 60 min of reaction. Fig. 4 shows that maleic and oxalic acids were identified as the intermediates from the oxidation of o-toluidine. Maleic and oxalic acid were found in the electro-Fenton and the photoelectroFenton processes after 1 min of reaction and the concentrations increased with time. The decrease in maleic acid occurred in the electro-Fenton and the photoelectro-Fenton processes after 45 and 10 min, respectively (Fig. 4(a)). The decrease in oxalic acid occurred in the electro-Fenton and the photoelectro-Fenton processes after 30 and 45 min, respectively (Fig. 4(b)). However, maleic and oxalic acids were found after 10 min of reaction in the Fenton process and their concentrations increased with time until the end of the reaction. The concentration of maleic and oxalic acid were generated
quickly because of the increased concentration of • OH and degradation of o-toluidine. This result indicates that the electro-Fenton and the photoelectro-Fenton processes have higher efficiencies to degrade o-toluidine than the traditional Fenton process. The accumulation of intermediates in the photoelectro-Fenton process was lower than in the electro-Fenton process (Fig. 4(a) and (b)). Moreover, the removal efficiency for TOC in the electro-Fenton and the photoelectro-Fenton processes were 12% and 31%, respectively (Fig. 4(c)). These phenomena show that the intermediates were efficiently mineralized by the action of UV light in the photoelectroFenton process.
Fig. 3. Comparison between various processes on (a) o-toluidine removal and (b) COD removal efficiency. Experimental conditions: 1 mM of o-toluidine, 1 mM of Fe2+ and 4.85 mM of H2 O2 at pH 2. Each data has twice samplings.
400
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results with 91% of o-toluidine removal and of 45% COD removal in the electro-Fenton process, indicating the reliability of the methodology used. Both electric discharge and UV irradiation could significantly enhance the degradation of o-toluidine. Therefore, more intermediates, such as maleic and oxalic acids, were derived during the beginning of the electro-Fenton and the photoelectroFenton processes than in the traditional Fenton process. Acknowledgements The authors would like to thank the National Science Council of Taiwan, for financially supporting this research under Contract No. NSC 96-2628-E-041-001-MY3. References
Fig. 4. The concentrations of (a) maleic acid, (b) oxalic acid and (c) TOC during the degradation of o-toluidine. Each data has twice samplings.
4. Conclusions This study investigated the optimization of o-toluidine treatment by the electro-Fenton process applying the Box–Behnken experimental design methodology. The results showed that pH and Fe2+ concentrations were important factors in the removal efficiencies for both o-toluidine and COD. The removal efficiencies for o-toluidine and COD increase with decreasing pH and increasing Fe2+ concentration. The optimal conditions for the maximum removal of o-toluidine and COD (90.8% and 40.9%, respectively, from prediction) were 1 mM of Fe2+ and 4.85 mM of H2 O2 at pH 2. Obviously, the calculating o-toluidine and COD removals applying the predicted conditions approaches the experimental
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Reconstruction of metal pollution and recent sedimentation processes in Havana Bay (Cuba): A tool for coastal ecosystem management M. Díaz-Asencio a,∗ , J.A. Corcho Alvarado b , C. Alonso-Hernández a , A. Quejido-Cabezas c , A.C. Ruiz-Fernández d, M. Sanchez-Sanchez c, M.B. Gómez-Mancebo c, P. Froidevaux b, J.A. Sanchez-Cabeza e a
Centro de Estudios Ambientales de Cienfuegos, Carretera Castillo de Jagua, Cienfuegos, CITMA-Cienfuegos, Cuba Institute of Radiation Physics (IRA), University Hospital and University of Lausanne, Rue du Grand-Pré 1, 1007 Lausanne, Switzerland c Centro de Investigaciones Energéticas, Medioambientales y Tecnológicas (CIEMAT), Madrid, Spain d Universidad Nacional Autónoma de México. ICMyL, Mazatlán, Mexico e Institute of Environmental Science and Technology, and Physics Department, Universitat Autónoma de Barcelona, 08193 Bellaterra, Barcelona, Spain b
a r t i c l e
i n f o
Article history: Received 20 June 2011 Received in revised form 12 September 2011 Accepted 12 September 2011 Available online 16 September 2011 Keywords: Pollutants 210 Pb 239,240 Pu 137 Cs Sediment dating Havana Bay
a b s t r a c t Since 1998 the highly polluted Havana Bay ecosystem has been the subject of a mitigation program. In order to determine whether pollution-reduction strategies were effective, we have evaluated the historical trends of pollution recorded in sediments of the Bay. A sediment core was dated radiometrically using natural and artificial fallout radionuclides. An irregularity in the 210 Pb record was caused by an episode of accelerated sedimentation. This episode was dated to occur in 1982, a year coincident with the heaviest rains reported in Havana over the XX century. Peaks of mass accumulation rates (MAR) were associated with hurricanes and intensive rains. In the past 60 years, these maxima are related to ˜ periods, which are known to increase rainfall in the north Caribbean region. We observed a strong El Nino steady increase of pollution (mainly Pb, Zn, Sn, and Hg) since the beginning of the century to the mid 90s, with enrichment factors as high as 6. MAR and pollution decreased rapidly after the mid 90s, although some trace metal levels remain high. This reduction was due to the integrated coastal zone management program introduced in the late 90s, which dismissed catchment erosion and pollution. © 2011 Elsevier B.V. All rights reserved.
1. Introduction Havana Bay is one of the largest and most important estuaries in the Cuban Island. The economic, commercial and recreational values of the bay have been, however, threatened by pollution and the reduction of water depth due to infilling [1]. The environmental degradation of the bay ecosystem has been intensified in the past decades due to the fast economical growth of Havana City. This one has become the most contaminated bay in the island [2]. In order to rehabilitate this marine ecosystem, several pollution-reduction measures have been implemented over the past decade. Reliable information about the input of pollutants to Havana Bay is however required for evaluating the impact of the environmental management practices. In the absence of long-term monitoring data, sedimentary records can provide retrospective information about the past inputs of pollutants into aquatic environments. Pollutants such as heavy metals often have a strong affinity for particle
∗ Corresponding author. E-mail address:
[email protected] (M. Díaz-Asencio). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.037
surfaces and, therefore, accumulate in the sediments. Hence, dated sediment profiles of major and trace elements can be used to obtain reliable information about the extent and history of pollution and sedimentary conditions [3–7]. The quantitative reconstruction of a contaminant input into an aquatic system requires a good sediment chronology. The most widely used method for dating recent marine and lacustrine sediments is based on the examination of 210 Pb profiles. The natural radionuclide 210 Pb (T1/2 = 22.23 yr) enters the aquatic environment mainly by atmospheric deposition; however it can be produced in situ, in the water column and the sediments, by decay of its precursor radionuclide 226 Ra (T1/2 = 1600 yr). The 210 Pb dating methods are based on the radioactive disequilibrium between the 210 Pb and 226 Ra [8,9]. 210 Pb has shown to be an ideal tracer for dating aquatic sediments deposited during the last 100–150 years, a period of time with significant environmental changes due to industrialization and population growth. 210 Pb dating should be always corroborated by an additional chronostratigraphic marker in the same sediment core [10,11]. Among the most commonly used time markers we find anthropogenic fallout radionuclides such as 137 Cs, 239,240 Pu or 241 Am
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the ships arriving to the country. Agriculture and intensive forest exploitation in the bay catchment increased soil erosion and, therefore, sediment input to the Bay. Industrial activities began in the 1850s with the construction of oil refineries, electric power plants and gas production [2,16]. The area of Havana experienced a fast economical growth during the 20th century, with a high diversity of industries and commercial activities, and a large population growth that required massive urbanization (from 250,000 inhabitants in 1899 to 2.2 million inhabitants in 2001). The anarchic growth of many activities over the past 400 years has caused severe damages to the natural resources and facilities of the Havana Bay. The damages have been intensified by the lack of waste treatment facilities [2,17,18]. The bay receives contaminants from numerous sources such as an oil refinery, power stations, urban and industrial wastewaters, three shipyards, riverine and stream discharges, and atmospheric fallout [1]. 3. Sampling and laboratory methods 3.1. Sampling
Fig. 1. Map of Havana Bay, with sampling station.
[13]. The onset of anthropogenic radionuclides, originating from the atmospheric nuclear weapon tests (early 1950s) and their peak value in 1963 [12] have been widely used as time markers in numerous marine and lacustrine studies [13–15]. In this work we reconstructed the historical trends of pollutants entering the Havana Bay by analyzing their sediment concentration profiles. The chronology of the sediment core was based on the 210 Pb dating method. Due to the low levels of 137 Cs found in the sediments, the chronology was further validated with 239,240 Pu and 241 Am fallout radionuclides. Enrichment factors and fluxes of pollutants were used to describe the history of pollutions in this aquatic ecosystem. A full geochemical analysis was undertaken to look at possible impacts of changing catchment sources, diagenesis, and atmospheric contamination. The potential origin of the most important pollutants and the impact of the pollution-reduction measures taken to protect the Havana Bay ecosystem are also discussed. Despite the large number of studies of pollution in coastal environments, only a few have been conducted in the Caribbean region. Hence, this study provides important information about pollution trends in a coastal ecosystem of this tropical region.
In February 2008, sediment cores were collected with an UWITEC corer avoiding dredged areas of the bay (Fig. 1). In order to optimize analytical time and resources, we chose the cores with the best likelihood to show good temporal record (section 1 of Supporting information). We sampled three sediment cores from the station B (23◦ 08.107 N 82◦ 20.043 ) at a water depth of 8 m (Fig. 1): one core for radionuclides, metals and grain size analysis; one for organic pollutants (not reported here) and one was kept frozen for future analysis. The sediment core was vertically extruded and sliced into 1 cm sections. Each section was dried at 45 ◦ C, sieved through a 1 cm sieve and homogenized. The mud content in the sediments showed a slight decreasing trend from 15 cm depth to the surface. The sediments also showed a strong change in color at about 15 cm depth. 3.2. Laboratory analyses Grain size was determined by standard methods of sieving and pipetting analysis [19]. Content of organic matter for each section was estimated by the loss on ignition method (LOI: 450 ◦ C, for 8 h). The content of total carbon and nitrogen was measured by using a CHN analyzer (LECO TRUSPEC). Total carbon was measured as CO2 with an infrared detector. N2 was measured by using a thermal conductivity detector. Inorganic carbon was quantified by using an infrared detector (SSM-500, Shimadzu) after sample acidification with phosphoric acid and heating (200◦ C). Major and trace elements were measured by Wavelength Dispersion X-Ray Fluorescence Spectrometry (WDXRF) using a Panalytical system (AXIOS) with Rhodium tube. Total mercury concentrations were determined by using an Advanced Mercury Analyser (LECO AMA254, detection limit of 0.01 ng Hg).
2. Site description
3.3. Sediment dating
Havana Bay (NW Cuba) is located aside Havana City, is a typical enclosed bay with a catchment area of about 68 km2 . It is characterized by a mean depth of 10 m, an area of 5.2 km2 and a water mean residence time of 7–9 days [1]. The bay is an estuary with deltaic systems in the fluvial discharge zones of the Luyano and Martin Perez rivers, and the Tadeo, Matadero, Agua Dulce and San Nicolas streams (Fig. 1). The population density, commercial and harbors activities in Havana City increased significantly since 1850. The city became a key transshipment point between Europe and America in the 19th century. Nowadays, the port of Havana receives about 50% of
The chronology of the sediment cores was determined by the method (section 2.3 of Supporting information). Sediment samples were placed in sealed plastic containers and stored for at least three weeks in order to allow 226 Ra to reach equilibrium with its daughter nuclides. 226 Ra was then analysed by high-resolution gamma spectrometry, using a low-background intrinsic Ge coaxial detector ORTEC model GX10022. 226 Ra was determined via the 352 keV emitted by its daughter nuclide 214 Pb in equilibrium. Supported 210 Pbsup was derived from the assumption of equilibrium with 226 Ra. The total 210 Pb activity was determined by high-resolution ␣ spectrometry of its decay product 210 Po, assumed 210 Pb
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Fig. 2. Profiles in core B of (a) particle size distribution, (b) carbonates and aluminisilicates distribution, and (c) organic (Corg ) and inorganic (Cinorg ) carbon, phosphorous (P) and nitrogen (N).
to be at equilibrium. Aliquots (0.5 g) of dry sediment were spiked with 209 Po as a yield tracer and dissolved by adding a mixture of 1:1:0.5 HNO3 + HCl + HF using an analytical microwave system [20]. 210 Po was electrodeposited onto silver discs [21,22] and counting was done in an integrated Camberra alpha spectrometer with ion implanted planar silicon detectors (active area of 450 mm2 ; and 18 keV of nominal resolution). The 210 Pb in excess (210 Pbex ) to the 210 Pb supported by 226 Ra (210 Pb sup ) was determined by subtracting the 210 Pbsup from the total activity of 210 Pb measured in each layer. 210 Pb ex was then introduced in the models in order to obtain the sedimentation rate (section 2.3 of Supporting information). Measurements of 137 Cs, 239,240 Pu and 241 Am were used to validate the 210 Pb dating models. 137 Cs was measured via its emission at 662 keV by high-resolution gamma spectrometry. The sediment samples were then crushed and ashed at 550 ◦ C for 48 h prior to the radiochemical analyses of 239,240 Pu and 241 Am. Composite samples were prepared by mixing layers. The method combines high-pressure microwave digestion for the dissolution of the sample and the highly selective extraction chromatographic resins TEVA and DGA (Triskem International, France) for the separation and purification of Pu and Am [23]. The alpha sources were
prepared by electrodeposition onto stainless steel discs [24]. Highresolution ␣-spectrometry was performed on a ␣-spectrometer with PIPS detectors (Alpha Analyst, Canberra Electronic).
4. Results 4.1. Characteristic of the sediments The sediments are predominantly fine and displayed small variations in grain size in the overall samples (Fig. 2a). The sediments consisted mainly of clay (<4 m, 58–85%), and silt and very fine sand (>4 m, 15–42%) sized particles. In the upper 5 cm, the percent of silt and very fine to coarse sand sized particles increased slightly (Fig. 2a). The sediments were mainly composed of carbonates (11–45%) and aluminosilicates (40–80%). The mineral composition of the sediments did not change significantly from the bottom of the core up to 20 cm depth (Fig. 2b). However, from 20 cm depth up to the surface large variations were observed with the amount of carbonates correlating negatively to the amount of alumino-silicates (Fig. 2b).
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Fig. 3. Profiles of Si, Ca, Fe, Al and Mn in core B. Uncertainties were below 2% for all elements. The dashed line indicates the slump.
The amount of carbonates showed a general trend to decrease towards the surface, but it increased rapidly in the top 4 cm (Fig. 2b). The content of inorganic carbon (Cinorg ) in the sediments was almost constant along the whole core; but the organic carbon (Corg ) showed a surface maximum and then decreased with depth depicting two zones of rapid change (at 2–3 cm, and at 16–17 cm depth; Fig. 2c). The large percentage of Corg in the top 2–3 cm may be related to a change in the sources of organic matter, more complex and less biodegradable, typical of industrial organic wastes. However, the increase of organic matter may be also related to the reduction of particles observed in the Bay over the past decades. In the segment 3–16 cm, Corg was nearly constant around 4%; then, below 16 cm depth, Corg decayed to about 1.5% (Fig. 2c). Similarly to the pattern of Corg , nitrogen (N) and phosphorous (P) profiles showed increasing trends towards the surface, with nearly constant concentrations between 3 and 16 cm depth (Fig. 2c). 4.2. Major and tracer elements in the sediment core The concentration profiles of Al, Fe, Si, and Mn showed a slight decrease in the two uppermost layers, but no significant changes below 3 cm depth (Fig. 3). The Ca content showed an opposite trend
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to Al (Fe, Si), with slightly higher concentrations at the surface (Fig. 3). In fact, Ca, which is probably in the form of biogenic carbonates, acts as a dilute of the trace metal concentrations in the bulk sediments. A slight change in the profiles is observed at 15–20 cm, which probably indicates the input of sediments with high content of Ca (Fig. 3). The concentration profiles of Pb, Zn, S, Hg, Sn and Cr in the sediment core showed similar increasing trends towards surface (Fig. 4). Maximum concentrations for Zn (450 mg kg−1 ), Pb (123 mg kg−1 ) and S (1.4 mg kg−1 ) were observed at the core surface, while for Hg (1.4 mg kg−1 ), Sn (18.6 mg kg−1 ) and Cr (365 mg kg−1 ) the maximum concentrations were found at depths between 5 and 15 cm. The maximum concentrations are comparable to the concentrations found in other polluted coastal sediments such as Porto Marghera in Italy [25], Halifax Harbor in Nova Scotia [26] and Barcelona in Spain [27]. The similarities in the Pb, Zn and Sn profiles (linear correlation R2 > 0.9 and p < 0.01 for each combination) suggest that these elements possibly originated from the same source and/or that the geochemical affinities to the sediment particles are similar. Pb, Zn, Hg, Sn, Cr and S profiles showed a plateau between 3 and 15 cm depth suggesting a similar time of deposition for the whole segment (e.g. an episodic event).
4.3. Radionuclide profiles and sediment chronology The 210 Pbex profile has non-monotonic features suggesting irregularities in the process of sediment accumulation (Fig. 5a). The surface 210 Pbex activities were around to 230 Bq kg−1 (Fig. 5a), relatively high compared to activities found in other studies from Havana Bay [28] and from other Cuban coastal locations [4,29]. A detailed analysis of the 210 Pbex distribution with depth suggests that the record can be divided into three distinct segments. At the top (0–3 cm) and bottom (15–35 cm) sections of the sediment core, 210 Pb ex declined exponentially with depth, indicating regular sedimentation. However, 210 Pbex was almost constant throughout the 3–15 cm segment of the core. A flattening of 210 Pbex indicates either a dilution of the 210 Pb atmospheric flux by sediment mixing, accelerating sedimentation and/or the occurrence of slumps due to, for example, heavy rains (typical in this tropical region). The trends observed in the profiles of 210 Pbex (Fig. 5a), Corg , P, N (Fig. 2c) and some major and trace elements (Figs. 3 and 4) suggest that most probably the section from 3 to 15 cm was instantly deposited as a result of an episodic event or slump. This is also supported by the color change of the sediment core at about 15 cm
Fig. 4. Profiles of (a) Pb, Zn, Sn and Cr; and b) S and Hg in core B. The uncertainties were: Pb (<1%); Zn (<1%); S (<2%), Hg (<0.5%), Sn (<4%), Cr (<1%). The dashed line indicates the slump.
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Fig. 5. Profiles of the activities of (a) 210 Pbex , (b) 241 Am and 239,240 Pu in core B. The dashed line indicates the slump.
depth (Table 1 of Supporting information). Sediment packages (slumps) are commonly observed in coastal areas, especially where riverine inputs are important. We conclude that about 87.3 kg m−2 were rapidly deposited (slump) in this zone of the Havana Bay. While the 137 Cs signal in the sediments was very low (<1.8 Bq kg−1 ), close to the detection limit; 239,240 Pu and 241 Am were detected in the sediment core at levels up to 0.50 Bq kg−1 and 0.16 Bq kg−1 , respectively (Fig. 5b). The first appearance of Pu and Am radionuclides in the sediment core (fallout onset) is found at 23 cm, which is a composite of the 22–23 and 23–24 cm layers (Fig. 5b). Despite the low 137 Cs activities in the sediments, its first appearance was also found at about 22 cm. These results indicate that the first appearance of the anthropogenic fallout radionuclides (1952) is indeed associated with the 22–24 cm depth horizon. The vertical distributions of 239,240 Pu and 241 Am did not show a well developed peak corresponding to the expected radionuclide fallout maximum in 1963 (Fig. 5b). However, both profiles depicted slightly higher activities in the middle segment (3–15 cm) associated to an episodic deposition event. This episodic event was responsible for a large input of radionuclides from the catchment area. The significant input of radionuclides from the catchment, compared to the direct atmospheric input, is confirmed by the large 239,240 Pu inventory of in the core (60 Bq m−2 ) compared to the expected Pu fallout inventory in Havana City (15–30 Bq m−2 ; section 2.2 of Supporting information). The 239,240 Pu inventory in the core is 2–4 times higher than the expected Pu fallout inventory in Havana City and in other nearby regions in USA [30–33]. 239,240 Pu inventories higher than the expected fallout inventory have been commonly found in marine sediments from near shore or coastal zones, which are characterized by high particle fluxes [33,34]. Our results show that the site received a substantial sedimentassociated radionuclide input from its catchment area in addition to the direct atmospheric input. We used the Constant Flux (CF) model [8,9] after removing the segment 3–15 cm (slump), and obtained an average mass
accumulation rate (MAR) of 1.6 kg m−2 y−1 and a 210 Pb flux of 137 Bq m−2 y−1 . These parameters are similar to those reported for other coastal regions of Cuba [4,29]. The CF MAR ranged from 0.4 to 3.8 kg m−2 y−1 (Fig. 6a and b), with a significant decrease in the past two decades. The CF ages agree well with the onset fallout determined from Pu and Am radionuclides (Fig. 6c). The sediment chronology was also constructed by the Constant Flux Constant Sedimentation (CFCS) model [35] in the same conditions as the CF model (section 2.3. of Supporting information). The results of both CFCS dating model in core B agree relatively well with the CF model (Fig. 6c). Therefore, the ages and MAR of the CF model, after removing the slump event, were used for the reconstruction of the pollution history of Habana Bay.
5. Discussion 5.1. Origin of major changes in sediment accumulation As shown in Fig. 6b, the MAR peaks most likely correspond to the major hurricanes and intensives rains that affected the Havana area between 1910 and 1982 [36–38]. The slump was dated in 1982, a ˜ period in Cuba (1982–1983; [36,39]). The rainfall in severe El Nino 1982 was reported as the strongest in Havana over 60 years [36]. As a result of the heavy rains in 1982, coastal flooding affected the coastal areas of Havana with intensities not witnessed in the region since the Great Hurricane of 1926 [39]. A closer look to the MAR record shows that three of the recent peaks could be associated ˜ with years (1957 or 1965, 1972 and 1982) with a strong El Nino influence (Fig. 6b). Indeed, some studies have shown that during El ˜ years there is a significant increase in the number of days with Nino heavy precipitations in Cuba [39–42], and therefore an increase in the sedimentation rate due to watershed erosion. Studies in nearby regions have found similar patterns associated to the ENSO warm phases. For example, Jamaica (SE of Cuba) is more likely to
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Fig. 6. Results of the application of the CF and CFCS dating models. (a and b) Temporal evolution of the CF MAR and the CFCS MAR. The date of occurrence of important ˜ events are also shown. (c) 210 Pb ages obtained with the CFCS and the CF models versus the massic climatic events such as cyclones, hurricanes and the years of strong El Nino depth (kg m−2 ), without considering the slump from 3 to 15 cm depth. The location of the onset fallout is also indicated.
experience floods during May–July of years with warm phases [43]. In the region between Florida and the north Caribbean, an excess of rainfall also occurs during the early dry season for ENSO warm phases [44]. This rainfall excess during ENSO warm phases has been associated, for example, with a greater frequency of extra-tropical cyclone passages across the southern continental USA. In the nineties, an ambitious coastal zone management program was implemented to restore the Havana Bay by acting on watershed erosion and pollution sources [16]. As a result, a significant reduction of solid suspended matter was observed in the Bay [45]. The low MARs observed in the upper section of core B (Fig. 6a) is similar to the ones observed at the end of the XIX century for other cores taken within the frame of this study (Table 4 of Supporting information). We concluded that the significant reduction of MAR over the past two decades is due to a combination of the effectiveness of the coastal zone management program implemented since 1998 and the economical contraction of Cuba in the early 90s due to the economic collapse of the socialist block.
5.2. Pollution history in the last 100 years Identifying anthropogenic from natural sources of pollutants is essential to understand the pollution history recorded in the sediments. For this purpose, the comparison of the element content in the sediments with the pre-industrial background concentration is often used. The enrichment factor (EF) and pollutant fluxes were
calculated in order to evaluate the pollution record of the Havana Bay ecosystem (section 2.1. of Supporting information). Enrichments of Pb, Zn, Sn, S, Cr and Hg are already observed since 1900 (Fig. 7a and b), indicating that anthropogenic pollution sources were active during the whole XX century. Indeed, pollution was likely associated to old pre-industrial activities in the bay, such as forges and metal manipulation in shipyards (which provided services to a large fleet), leather manufacture, and sugar and derivatives production. The use of some natural heavy hydrocarbons such as tar (with high levels of S, Pb and Hg) to waterproof ships is reported to occur before 1700 [1]. The EFs increased significantly after 1950s, which corresponds to the period of economical development in Cuba and of fast population growth in Havana City. In the past two decades, the EFs of Pb, Zn and Sn were higher than 5 (500% increase) than the core bottom levels (Fig. 7a). These high EFs are likely a result of an increase of the pollutant specific concentrations due to the significant reduction of solid matter in Habana Bay over the past two decades [45]. For Sn and Hg, despite the significant increase of their EFs, the absolute fluxes have been decreasing since the 80s. This behavior can also be explained by the reduction of solid matter in the Bay over the past two decades as changes in their fluxes (Fig. 8) seem to be slightly modulated by changes in the MAR (Fig. 6a). The anthropogenic fluxes of Pb, Zn, Sn, S, Hg and Cr into the sediments increased rapidly from the beginning of the century until the 1980s (Fig. 8), indicating an increasing impact from the urban and industrial activities around Havana Bay. However, these fluxes
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Fig. 7. Temporal record of the Pb, Zn, Sn, S, Hg, Cr, Cu, Mn, Co and Ni enrichment factors. The uncertainties were: Pb (<1%); Zn (<1%); Sn (<8%); S (<3%); Hg (<1%); Cr (<1%); Cu (<1%); Mn (<2%); Co (<5%); Ni (<1%). The slump event (1982) is not included.
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Fig. 8. Reconstructed natural and anthropogenic fluxes of Pb, Zn, Sn, S, Hg and Cr in the last century. The flux due to the slump event (1982) is not included.
are drastically reduced since the 1990s, due to a combination of the coastal zone management program of the Bay and the economic contraction of Cuba. The EFs of Mn, Co and Ni are close to or below 1 (Fig. 7c), indicating that these elements are depleted towards the surface. The slight decrease of the Mn enrichment factor values towards the core surface (Fig. 7c) could be related to Mn remobilization owed to diagenesis, since it is known that Mn3+ reduction to Mn2+ in anoxic sediments might promote the dissolution of Mn compounds and the upward diffusion of soluble species into the water–sediment interface [46,47]. This same process could also explain the EF profiles obtained for Ni and Co, which showed a slight deficit at sediment core surface (Fig. 7c.) and whose diagenetic behavior in coastal sediments have been reported to be strongly associated
with that of Mn oxides [48]. However, no other trace element profile showed to be influenced by the Mn behavior in the sedimentary column, and Mn, Fe as well as Ni and Co showed significant correlations to Al, indicating that diagenesis influence is not significant on the trace elements depth profiles and that changes observed are mostly due to variations in the trace elements terrestrial input to Havana Bay. 5.3. Change in sediment sources The PCA analysis showed that two factors explained more than 76% of the total variance (Table 1). Factor 1 explained 41% of the total variance and grouped the typical pollutants (e.g. Pb, Zn, Hg, S, Sn, etc.) with negative values and other elements, like Mn, with
Table 1 Principal component (PC) loadings of PCA based on 25 variables of the core from station B. The largest coefficients in each PC are in bold.
Factor 1 Factor 2
Factor 1 Factor 2
MAR
Ca
Si
Al
Fe
S
Mn
Pb
Cr
Zn
Cu
As
Co
−0.34 0.69
0.62 −0.73
0.13 0.96
0.24 0.94
−0.15 0.94
−0.97 0.16
0.94 0.25
−0.97 0.09
−0.49 0.69
−0.97 −0.12
0.55 −0.38
−0.09 −0.52
0.64 0.73
Hg
Ni
Sn
Mg
K
Ti
Br
I
Sr
Ba
Mo
U
−0.84 0.44
0.57 0.78
−0.93 0.23
−0.85 0.35
−0,34 0.76
−0.01 0.93
−0.71 −0.63
−0.63 −0.72
−0.36 0.06
−0.20 0.32
−0.92 −0.28
−0.66 −0.08
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Fig. 9. (a) Projection of 25 variables on the two principal components plane. (b) Temporal evolution of the two principal components.
positive values (Fig. 9a). Factor 2, which explained 35% of the total variance, included the major components of the sediments, and allowed to identify the two main sediment sources: Ca from marine sediments with negative value; Si, Al, Ti and Fe from terrigenous inputs, with positive value (Fig. 9a). The section loadings for the two main factors showed significant time trends (Fig. 9b). From 1890 to approximately 1930, the sediments showed a constant value of major components (Factor 2) and rather constant pollution levels (Factor 1). From 1930 to 1961, pollution levels steadily increased (decrease of Factor 1), caused by the socio-economic development of the Havana City [2,16], and terrigenous inputs increased due to catchment erosion. In 1957–61 and 1982 we observed a peak of terrigenous components, coincident with high MAR and due to extreme rainfall in Havana. From 1961 to approximately 1990, the pollution levels steadily increased. However, after 1990 the terrigenous input steady decreased while pollution levels steadily increased. These opposite trends reflect the reduction of terrigenous input to the bay due to the socioeconomic contraction from 1993 to 1998 and the coastal zone management actions taken in the Bay and its watershed after 1998. The reduction of the sediment load in the bay had the negative effect of increasing the pollutant concentrations in suspended matter, reflected as enhanced sediment concentration levels, but the positive effect of reducing the total flux of pollutants to the sediment. 6. Conclusions A sediment core from Havana Bay allowed us to reconstruct the history of anthropogenic impacts since the beginning of the XX century. We observed a steady increase of pollutants (e.g. Pb, Zn, Sn, and Hg) since the beginning of the last century to the mid 90s, with enrichment factors as high as 6. Relative fast decreases of MAR and pollutants were observed after mid 90s, although some concentrations remain high. We concluded that the economic contraction of Cuba and the integrated coastal zone management program introduced in the 90s are responsible for the reduction of sedimentation and pollution fluxes in the Bay. This observation confirmed the positive results of the management program to reduce the suspended matter supply to the bay. However, it is still necessary to implement new actions oriented to reduce the concentration levels of organic matter and metals associated to particulate matter entering the system by further acting on the pollution sources themselves. This work shows the relevance of dated environmental archives to reconstruct the sedimentation and pollution trends
and its usefulness to evaluate the impact of environmental management practices. Peaks of the mass accumulation rates were found to be associated with the time of occurrence of a strong hurricane or a period of intensive rains. Moreover, over the past 60 years, most of the peaks of the MAR occurred in years (1957, 1972 and 1982) characterized ˜ events. This study shows the impact of the ENSO by strong El Nino oscillation in coastal estuaries like the Havana Bay. This historical reconstruction of the environmental changes in Havana Bay provided an excellent example of the use of Nuclear Techniques as tools in the Management Problems of Coastal Zones in the Caribbean Region. Acknowledgements This work was supported by the IAEA (Regional Project RLA/7/012) and the CIEMAT International Cooperation program. We also acknowledge the Swiss Federal Office for Public Health (OFSP) for their financial support. This work has involved a large number of people, and it is impossible to list all of them here. We would like to give special thanks to CEAC staff (Anabel Pulido, Hector Cartas, Rosario Rodríguez, Miguel Gómez), CIEMAT staff (Lourdes Romero – in memoriam, Manuel Barrera, Dolores M. Sánchez, Manuel Fernández, Ramón Morante, M. Isabel Rucandio, Rodolfo Fernández), CIMAB staff (Reinaldo Fernández, Jesus Beltran) and IRA staff. We are also grateful to IAEA Technical Cooperation Department (Latin America Section, very specially the RLA/7/012 Project Manager, Jane Gerardo-Abaya), CIEMAT Office of International Cooperation (Félix Barrio) and CIEMAT Director General Juan Antonio Rubio (in memoriam). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.037. References [1] ITT, Informe Final Proyecto CUB/80/001, Investigación y Control de la Contaminación Marina en la Bahía de la Habana, PNUMA, UNESCO (1985) (in Spanish). [2] A. Valdes-Mujica, La Bahia y la ciudad, Pasado, presente y futuro de un ecosistema, El pelicano (2008) 4–11. [3] T.M. Church, C.K. Sommerfield, D.J. Velinsky, D. Point, C. Benoit, D. Amouroux, D. Plaa, O.F.X. Donard, Marsh sediments as records of sedimentation, eutrophication and metal pollution in the urban Delaware Estuary, Mar. Chem. 102 (2006) 72–95.
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Journal of Hazardous Materials 196 (2011) 412–419
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Characterization of the propanil biodegradation pathway in Sphingomonas sp. Y57 and cloning of the propanil hydrolase gene prpH Ji Zhang, Ji-Quan Sun, Qiao-Yun Yuan, Chao Li, Xin Yan ∗ , Qing Hong ∗ , Shun-Peng Li Key Lab of Microbiological Engineering of Agricultural Environment, Ministry of Agriculture, College of Life Science, Nanjing Agriculture University, 210095, Nanjing, Jiangsu Province, People’s Republic of China
a r t i c l e
i n f o
Article history: Received 29 June 2011 Received in revised form 9 September 2011 Accepted 12 September 2011 Available online 16 September 2011 Keywords: Sphingomonas sp. Y57 Biodegradation Propanil 3,4-DCA prpH
a b s t r a c t In our previous study, the isoproturon-degrading strain Sphingomonas sp. Y57 was isolated from the wastewater treatment system of an herbicide factory. Interestingly, this strain also showed the ability to degrade propanil (3,4-dichloropropionamilide). The present work reveals that Y57 degrades propanil via the following pathway: propanil was initially hydrolyzed to 3,4-dichloroaniline (3,4-DCA) and then converted to 4,5-dichlorocatechol, which was then subjected to aromatic ring cleavage and further processing. N-acylation and N-deacylation of 3,4-DCA also occurred, and among N-acylation products, 3,4-dichloropropionanilide was found for the first time. The gene encoding the propanil hydrolase responsible for transforming propanil into 3,4-DCA was cloned from Y57 and was designated as prpH. PrpH was expressed in Escherichia coli BL21 and purified using Ni-nitrilotriacetic acid affinity chromatography. PrpH displayed the highest activity against propanil at 40 ◦ C and at pH 7.0. The effect of metal ions on the propanil-degrading activity of PrpH was also determined. To our knowledge, this is the first report of a strain that can degrade both propanil and 3,4-DCA and the first identification of a gene encoding a propanil hydrolase in bacteria. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The intensive use of pesticides in agriculture is a matter of worldwide concern. These substances constitute a substantial source of contamination of non-target systems due to their overuse and application techniques. Contamination of different environmental niches via spray drift, volatilization, run-off and/or leaching has been widely reported in the literature [1–4]. Many pesticides persist in edaphic and aquatic environments either unaltered or only partially degraded. Most of these compounds, as well as their metabolites, represent human health and ecotoxicity threats and should therefore be removed from contaminated environments. Propanil, a highly selective post-emergent contact herbicide, is one of the most extensively used herbicides for rice production worldwide and is ranked within the top 20 pesticides used in agriculture in the United States [5]. Propanil and its major metabolite 3,4-DCA are biologically active pollutants that cause acute toxicity in a wide range of aquatic species [5–8]. Propanil is also an important cause of death from acute pesticide poisoning, of which methemoglobinemia is an important manifestation [5]. In
∗ Corresponding authors. Tel.: +86 025 84396314; fax: +86 025 84396314. E-mail addresses:
[email protected] (X. Yan),
[email protected] (Q. Hong). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.040
soil, biodegradation of propanil generates 3,4-DCA, which is also a product of the microbial transformation of other herbicides, such as diuron and linuron [8–11]. Compared with propanil, 3,4-DCA shows lower toxicity against fish and mammals [8,12]. However, 3,4-DCA can be converted by microbial peroxidases to 3,3 ,4,4 tetrachloroazobenzene (TCAB) and other azo products. There is great concern about TCAB because it is a carcinogen and a potential genotoxin [8]. Microbes play an important role in removing agrochemical residues and their metabolites from contaminated environments [13,14]. In previous studies, several bacteria and fungi capable of hydrolyzing propanil to 3,4-DCA were isolated from different environments [8,15–19]. Strains capable of degrading 3,4-DCA have also been isolated and characterized [10,20–25]. However, to the best of our knowledge, a single strain that can hydrolyze propanil to 3,4-DCA and further degrade 3,4-DCA has never been reported. Furthermore, while several hydrolases that transform propanil to 3,4-DCA have been purified from bacteria or fungi [26,27], no propanil hydrolase gene has been identified. In our previous study, an herbicide-degrading strain Sphingomonas sp. Y57 [28], which can degrade isoproturon, diuron and propanil, was isolated from the wastewater treatment system of an herbicide factory. In this work, it was demonstrated that Y57 first hydrolyzed propanil to 3,4-DCA and subsequently degraded 3,4DCA. The degradation pathway of propanil in Y57 was studied. prpH,
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Table 1 Strains and plasmids used in this study. Strains and plasmids
Characteristic(s)
Source or reference
Strains Sphingomonas sp.Y57 E.coli DH5␣ E.coli BL21 (DE3)
Wild type; Propanil and 3,4-DCA degrader Host strain for cloning vectors Expression host
Lab stock Lab stock Lab stock
Plasmids pUC118 BamH I/BAP pUC118-T1T2 pT-prpH pET-29a(+) pET-prpH
BamH I digested DNA fragment cloning vector pUC118 Derivative carrying the Sau3A I digested DNA fragment containing prpH. pMD18-T Derivative carrying the prpH Expression vector pET-29a(+) Derivative carrying the gene prpH
TaKaRa This study This study Novagen This study
encoding a propanil hydrolase responsible for converting propanil to 3,4-DCA, was cloned from Y57 and was expressed in Escherichia coli; the target protein was then purified and characterized. 2. Materials and methods
on a rotary shaker at 30 ◦ C, and samples of the suspension were taken at regular intervals. The optical densities (OD600 ) of the samples were measured with a Shimadzu UV-2450 UV–vis spectrophotometer (Shimadzu Corporation, Japan). All samples were immediately extracted with an equal volume of dichloromethane and analyzed by HPLC.
2.1. Chemicals and media Propanil (97%) and 3,4-DCA (98%) were purchased from Nanjing Trust Chem. Co., Ltd. and Alfa Aesar A Johnson Matthey Co., respectively. Methanol used for liquid chromatography was purchased from Jiangsu Hanbon Science & Technology Co., Ltd. All other chemicals used in this study were of analytical grade. Minimal salt medium (MSM), prepared at a pH of 7.0, contained (g L−1 ) NaCl 1.0, NH4 NO3 1.0, K2 HPO4 1.5, KH2 PO4 0.5, MgSO4 ·7H2 O 0.2. Concentrated stock solutions of propanil and 3,4-DCA were prepared in methanol at a concentration of 10 g L−1 . Luria-Bertani (LB) medium, prepared at a pH of 7.0, contained (g L−1 ) tryptone 10.0, yeast extract 5.0 and NaCl 5.0. To make selective media, ampicillin and kanamycin were added at concentrations of 100 and 50 mg L−1 , respectively. 2.2. Strains, plasmids and DNA-manipulation techniques Strains and plasmids used in this study are listed in Table 1. Oligonucleotide synthesis and DNA sequencing reactions were performed by Invitrogen Biotechnology Co., Ltd. DNA was gel-purified using the AxyPrep DNA Gel Extraction Kit (Axygen). All enzymes were used as specified by the supplier (TaKaRa Biotechnology (Dalian) Co., Ltd.). 2.3. Degradation of propanil and 3,4-DCA by Y57 In order to analyze biodegradation of propanil and 3,4-DCA, one loop of strain Y57 was inoculated into LB medium and incubated overnight at 30 ◦ C. After harvest by centrifugation (4000 × g, 5 min), the cell pellet was washed twice with sterilized MSM, and suspended in MSM to give an OD600 of 1.0. Propanil and 3,4-DCA degradation tests were performed by adding propanil or 3,4-DCA to the cell suspension at a concentration of 30 mg L−1 . The cell suspension was shaken in a rotary shaker at 180 rpm at 30 ◦ C, and samples of the suspension were taken at regular intervals. All samples were immediately extracted with an equal volume of dichloromethane and analyzed by high performance liquid chromatography (HPLC) as described below. 2.4. Degradation of 4,5-dichlorocatechol by Y57 4,5-Dichlorocatechol (4,5-DCCAT) degradation tests were performed in MSM. Y57 cells were washed in MSM (as above) and diluted 1:100 into 100 mL MSM containing 50 mg L−1 4,5-DCCAT as the sole carbon source. The cell suspension was shaken at 180 rpm
2.5. Identification of propanil and 3,4-DCA degradation products Samples (10 mL) were taken at regular intervals during propanil and 3,4-DCA degradation by Y57 as described above. Cell suspensions lacking propanil and 3,4-DCA were used as negative controls. Propanil and 3,4-DCA (in MSM) controls were included as well. All samples were extracted with an equal volume of dichloromethane; the organic layer was dried and re-dissolved in methanol. To identify the metabolites, samples were analyzed by HPLC and gas chromatography–mass spectrometry (GC–MS). For the HPLC analysis, the separation column (internal diameter, 4.6 mm; length, 25 cm) was filled with Kromasil 100-5 C18. The mobile phase was methanol:water (80:20, v:v), and the flow rate was 0.8 mL min−1 . The detection wavelength was 246 nm. GC–MS analyzes were performed on a Thermo Trace DSQ mass spectrometer. Helium was used as the carrier gas at a flow rate of 1.2 mL min−1 . Gas chromatography was conducted using a RTX5MS column (15 m × 0.25 mm × 0.25 mm, Restek Corp., US). The column temperature profile was programmed as follows: hold at 50 ◦ C for 1.5 min; increase to 150 ◦ C at 40 ◦ C min−1 and hold for 1 min; increase to 200 ◦ C at 10 ◦ C min−1 and hold for 6 min; finally, increase to 260 ◦ C at 50 ◦ C min−1 and hold at 260 ◦ C for 6 min. The injector temperature was set at 220 ◦ C with a split ratio of 20:1. The interface temperature and ion source temperature were both set to 250 ◦ C. The column outlet was inserted directly into the electron ionization source block, operating at 70 eV. 2.6. Cloning and sequence analysis of a propanil hydrolase gene The genomic DNA of Y57 was randomly digested with Sau3A I. 4–6 kb length DNA fragments were recovered by gel purification, ligated into BamH I/BAP-treated pUC118 and transformed into E. coli DH5␣ cells. Transformants were spread onto LB agar containing 100 mg mL−1 ampicillin and incubated at 37 ◦ C overnight. Colonies were transferred onto LB agar plates supplemented with 100 mg L−1 ampicillin and 400 mg L−1 propanil. The solubility of 3,4-DCA is higher than that of propanil in water, so a clear transparent halo forms around colonies that can hydrolyze propanil to 3,4-DCA. Such colonies were selected and further tested by HPLC. Nucleotide and amino acid sequence analyzes were performed using BioEdit software. BlastN and BlastP were used for the nucleotide and protein sequence searches (www.ncbi. nlm.nih.gov/Blast). Enzyme MWs and PIs were predicted using the ExPASy Proteomic Server (http://www.expasy.org/).
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2.7. Expression and purification of PrpH To express PrpH in E. coli using the pET29a expression system (Novagen), prpH was amplified from Y57 genomic DNA. Forward and reverse primers, prpH-F (5 -CATATGTCCAACACCGGCTTCTACACGC-3 , forward) and prpH-R (5 -CTCGAGGCCCAGCAGCACGGCCGCCATC-3 , reverse) incorporated Nde I and Xho I sites (underlined), respectively, into the amplified product. The PCR product was cloned into the pMD18-T vector (TaKaRa) and sequenced. The prpH DNA fragment was then digested by Nde I and Xho I, gel-purified and ligated into the corresponding site of pET29a, generating pET-prpH, which was transformed into E. coli BL21 (DE3). Transformants were subcultured into 50 mL LB medium and were allowed to grow until reaching a culture density of 0.5 (OD600 ). Then, prpH expression was induced by adding isopropyl--d-thiogalactopyranoside (IPTG) to a final concentration of 1 mM, and after a 12 h incubation at 18 ◦ C, cells were harvested by centrifugation, washed twice and suspended in binding buffer (20 mM Na3 PO3 , 500 mM NaCl, 20 mM imidazole; pH 7.8). The cell suspension was disrupted by sonication. Cell debris was removed by centrifugation. The supernatant was loaded onto a His-Bind resin (Novagen). After elution of the nontarget proteins with elution buffer (20 mM Na3 PO3 , 500 mM NaCl; pH 6.0) containing 50 mM imidazole, the target protein was eluted with elution buffer containing 200 mM imidazole. The enzyme was dialyzed against 20 mM Tris–HCl (pH 8.0) containing 200 mM NaCl for 24 h and was concentrated with self-indicating silica gel. The purified enzyme was analyzed using sodium dodecyl sulfate-polyacrylamide gel electrophoresis (SDS-PAGE). 2.8. Enzyme assays Enzymatic activities toward propanil were determined in 20 mM Tris–HCl (pH 8.0) unless described otherwise. Then, 5 l of propanil stock solution was added to 0.5 mL of Tris–HCl buffer containing 25 l enzyme solution. The mixture was incubated for 3 h at the temperatures given. Reactions were stopped by adding 1.5 mL dichloromethane, and the substrate was extracted and analyzed by HPLC. To study the effect of incubation temperature, samples were incubated at temperatures ranging from 20 ◦ C to 60 ◦ C. The effect of pH was studied at 40 ◦ C in Na2 HPO4 -citric acid buffer. Metal ions were added to final concentrations of 1 mM and 10 mM to study their effects on PrpH activity. Three replicates were used for each sample. One activity unit in this study was defined as the amount of enzyme required to catalyze the formation or hydrolysis of 1 mol product or substrate per min. 2.9. GenBank accession number The GenBank accession numbers of gene prpH and the 16S rRNA gene sequence of strain Sphingomonas sp. Y57 are JN024622 and DQ092868, respectively.
Fig. 1. Identification of propanil and its putative degradation intermediates by comparing the HPLC retention times with standard chemicals. (A) propanil standard. (B) 3,4-DCA standard. (C) Propanil degraded by Y57. (D) Propanil hydrolyzed by PrpH.
that Y57 could not only transform propanil into 3,4-DCA but could also degrade 3,4-DCA. This was further demonstrated by 3,4-DCA degradation test (Fig. 2B). Many propanil-degrading strains have been reported [8,15–19], but none are known to degrade propanil beyond 3,4-DCA. 3.2. Identification of the metabolites during 3,4-DCA degradation by Y57
3. Results and discussion 3.1. Propanil degradation in MSM by Sphingomonas sp. Y57 As shown in Fig. 1C, during the degradation of propanil, one metabolite with the retention time of 5.53 min was detected. This product had nearly the same retention time as authentic 3,4-DCA (Fig. 1B and C) and was identified as 3,4-DCA by GC–MS (data not shown). The kinetics of degradation of propanil and production of 3,4-DCA were then concurrently investigated. Over 98% of propanil (30 mg L−1 ) was degraded by cell suspensions of strain Y57 within 16 h (Fig. 2A). During the first 8 h, 3,4-DCA was accumulated. After this point, 3,4-DCA concentration began to decrease, indicating
As shown in Fig. 3, four metabolites appeared during the degradation of 3,4-DCA. The predicted chemical structures, retention times and characteristic ions of the mass spectra are listed in Table 2. The detection of three 3,4-DCA N-acylation products indicated that 3,4-DCA underwent N-acylation and N-deacylation events during the degradation. Among these N-acylation products, 3,4-dichloroformylanilide and 3,4-dichloroacetanilide were reported during the degradation of 3,4-DCA in previous studies [25,29–31], but 3,4-dichloropropionanilide (propanil) was here detected for the first time. Metabolite B was identified as 4,5dichlorocatechol (Fig. 3a and c and Table 2), indicating that 3,4-DCA was converted to 4,5-dichlorocatechol. As shown in Fig. 2C, Y57
J. Zhang et al. / Journal of Hazardous Materials 196 (2011) 412–419
35
A
30
B
30
Propanil 25
25
34DCA 20
20
15
15
10
10
5
5
0
0
2
4
6
8
10
12
14
16
0
34DCA
0
2
4
6
Concentration of 4,5-DCCAT (mgL-1)
Time (h) 60
8
10
12
14
16
Time (h)
0.12
C 4,5-DCCAT OD600
50
0.1
40
0.08
30
0.06
20
0.04
10
0.02
OD600
Concentration of propanil and 3,4DCA (mgL-1)
35
415
0
0 0
2
4
6
8 10 12 14 16 18 20 22 24
Time (h) Fig. 2. (A) Degradation dynamics of propanil in MSM by Y57. () The degradation of propanil; () the accumulation of 3,4-DCA. (B) Degradation dynamics of 3,4-DCA in MSM by Y57. (C) Degradation dynamics of 4,5-DCCAT in MSM by Y57. () The degradation of 4,5-DCCAT; () the growth of Y57.
Table 2 3,4-DCA and its metabolites identified by GC–MS. Chemical structural formula in NIST library
Name
Megtabolite
RT (min)
Characteristic ions in GC–MS (m/z)
A
4.57
160.91, 133.91, 125.96, 98.90, 89.95, 80.44, 72.92, 62.93,
3,4-Dichloroaniline (3,4-DCA)
B
6.39
177.89, 159.94, 148.98, 131.91, 114.91, 96.98, 84.98, 70.89, 61.95
4,5-Dichlorocatechol
C
7.06
188.93, 177.85, 160.95, 134.06, 124.72, 113.09, 108.80, 98.98, 89.97, 71.00, 62.95, 56.93
3,4-Dichloroformylanilide
D
7.47
202.93, 160.91, 144.89, 132.87, 125.97, 108.89, 98.89, 89.97, 72.90, 62.92
3,4-Dichloroacetanilide
E
8.15
216.96, 160.90, 144.89, 132.94, 125.97, 108.91, 100.01, 89.95, 73.94, 62.93,
3,4-Dichloropropionanilide (propanil)
416
J. Zhang et al. / Journal of Hazardous Materials 196 (2011) 412–419
Fig. 3. GC–MS mass spectrum analysis of the 3,4-DCA intermediates transformed by Y57: (a) the GC profiles for 3,4-DCA and its intermediates A, B, C, D, E. The retention times of each compounds was 4.57, 6.39, 7.06, 7.47 and 8.15, respectively. (b–f) The characteristic ions of compounds A–E in GC–MS. They were identified as 3,4-DCA, 4,5-dichlorocatechol, 3,4-dichloroformylanilide, 3,4-dichloroacetanilide and 3,4-dichloropropionanilide (propanil), respectively.
could degrade 4,5-dichlorocatechol and also utilize it as its sole carbon source for growth. Based on the above results, we proposed a degradation pathway of propanil in Y57 (Fig. 4). Y57 was originally isolated, using isoproturon as the sole carbon source, from the wastewater treatment system of an herbicide factory that produced several kinds of herbicides including isoproturon, diuron and propanil. Notably, these herbicides are all aniline
derivatives. Y57 has the ability to degrade a series of aniline derivatives such as 4-IA (4-isopropylaniline), 3-CA (3-chloroaniline), 4-CA (4-chloroaniline) and 3,4-DCA (data not shown). Clearly its acquisition of the ability to degrade aniline derivatives, whether by horizontal gene transfer or mutation, gave strain Y57 advantages over other strains for survival the wastewater treatment system.
J. Zhang et al. / Journal of Hazardous Materials 196 (2011) 412–419
417
H N
Cl
O Cl
3,4-Dichloropropionanilide (Propanil)
Cl
H N
Cl CH
NH 2
Cl
H N O
O Cl
Cl
Cl
3,4-Dichloroformylamilide
3,4-Dichloroaniline (3,4-DCA)
Cl
OH
Cl
OH
3,4-Dichloroacetanilide
4,5-Dichlorocatechol
CO2 and biomass Fig. 4. Proposed pathway for degradation of propanil and 3,4-DCA in Sphingomonas sp.Y57.
3.3. Cloning and sequence analysis of the propanil hydrolase gene Although propanil is known to be hydrolyzed to 3,4-DCA by microbes [8], and several propanil hydrolases responsible for transforming propanil to 3,4-DCA have been purified from bacteria or fungi [26,27], there are no reports of the cloning of a propanil hydrolase gene. To clone the propanil hydrolase gene from Y57, a gene library was constructed from Y57 genomic DNA. Three positive clones were obtained from approximately 6000 transformants. Their propanil-hydrolyzing abilities were verified by HPLC. One positive clone, 3250 bp in length, was selected for further study. Putative ORFs were then subcloned into pMD18T and transformed into E. coli DH5␣ to determine their ability to hydrolyze propanil. The ORF encoding the propanil hydrolase was designated as prpH. Sequence analysis indicated that the prpH ORF was 1110 bp long and that it encoded a protein of 369 amino acids. Proteomic predictions (ExPASy) showed that the molecular weight of PrpH is 40146.6 Da and the theoretical pI of PrpH is 5.20. The ORF started with ATG and ended with TAG. A putative ribosomal binding site (AGGAAG) was located 10 bp upstream of the start codon. The G + C content was 69.28%. The prpH nucleotide sequence and predicted protein sequence were compared with those in the GenBank database by an online alignment search. Y57 prpH shares a high level of similarity with a series of putative histone deacetylases (HDACs) at both the nucleotide and protein sequence level. PrpH showed the highest protein sequence identity (72%) with a putative histone deacetylase from Sphingomonas wittichii RW1 (GenBank Accession Number YP 001263732.1) and 46% identity with a verified bacterial histone
deacetylase-like amidohydrolase (HDAH) from Alcaligenes strain FB188 (DSM11172) [32,33]. We speculated that PrpH was a bacterial histone deacetylase homologue that had the capability to remove the N-propyl group from propanil. S. wittichii RW1 cannot convert propanil to 3,4-DCA (data not shown), indicating that PrpH is different from the histone deacetylase RW1, although it may have obtained this function through gene mutation during long-term exposure to propanil. 3.4. Expression, purification and characterization of PrpH PrpH was expressed in E. coli BL21 (DE3) and purified using Ninitrilotriacetic acid affinity chromatography. The purified enzyme gave a single band on SDS-PAGE (Fig. 5A). The molecular mass of the denatured enzyme was approximately 40 kDa, in agreement with the molecular mass deduced from the amino acid sequence (40146.6 Da) (GenBank accession number JN024622). The effects of temperature, pH and metal ions on the propanildegrading activity of PrpH were assayed with the purified enzyme (Fig. 1D). As shown in Fig. 5B and C, the optimal temperatures and pH for PrpH were 40 ◦ C and 7.0, respectively. The results shown in Fig. 5D indicated that Li+ , Mg2+ , Ga2+ and Mn2+ did not remarkably influence enzyme activity at concentrations of 1 mM or 10 mM but that Ag+ , Cd2+ , Zn2+ , Cu2+ and Hg2+ were strong inhibitors of PrpH. Approximately 40% to 60% of the PrpH activity was inhibited by Ni2+ and Co2+ at the tested concentrations. At a low concentration (1 mM), Cr3+ , Fe2+ , Fe3+ and Al3+ slightly influenced the enzyme activity, but at a high concentration (10 mM), they all strongly inhibited enzyme activity. Notably, Zn2+ was a strong inhibitor of PrpH at a concentration of 1 mM, which is in agreement with
418
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Fig. 5. Characterization of the purified recombinant PrpH produced from the recombinant E. coli BL21(DE3) strain harboring pET29a-prpH: (A) SDS-PAGE analysis of the purified recombinant PrpH, (lane M) low weight protein marker; (lane PrpH) purified recombinant PrpH. (B) Effect of temperature on enzyme activity of PrpH. (C) Effect of pH on enzyme activity of PrpH. (D) Effect of different metal ions on enzyme activity of PrpH.
published propanil degradation tests using Y57 cells [28]. However, HDAH from Alcaligenes strain FB188 contains a Zn2+ ion in the active site, which contributes significantly to catalytic activity, and Zn2+ was able to increase the catalytic activity of HDAH (2.2-fold) at a concentration of 1 mM [32]. This activity may indicate that PrpH is different from HDAH. Additionally, besides propanil, stain Y57 was also able to degrade other types of herbicides, such as isoproturon and diuron, but PrpH could not degrade any other herbicides except propanil (data not shown), indicating the existence of other herbicide-degrading enzymes in strain Y57. 4. Conclusion The isoproturon-degrading strain Y57 degrades both propanil and 3,4-DCA. It first hydrolyzes propanil to 3,4-DCA, which it converts to 4,5-dichlorocatechol before metabolizing it further. Nacylation and N-deacylation of 3,4-DCA also occurred in strain Y57, and N-propionylation was found for the first time. The gene encoding a propanil hydrolase responsible for transforming propanil to 3,4-DCA was cloned from strain Y57 and was designated as prpH. Its protein product was expressed in E. coli, purified and characterized. To the best of our knowledge, strain Sphingomonas sp. Y57 is the first bacterium reported to degrade both propanil and 3,4DCA, and the gene prpH is the first propanil hydrolase gene to be identified in bacteria.
Acknowledgements This work was supported by grants from Chinese National Natural Science Foundation (31070099, 31000060 and 30830001), Natural Science Foundation of Jiangsu Province, China (BK2009312), and the Key Technology R&D Program of Jiang Su Province (BE2009670). References [1] P.C. Wilson, J.F. Foos, Survey of carbamate and organophosphorous pesticide export from a South Florida (USA) agricultural watershed: implications of sampling frequency on ecological risk estimation, Environ. Toxicol. Chem. 25 (2006) 2847–2852. [2] M.I. Tariq, S. Afzal, I. Hussain, N. Sultana, Pesticide exposure in Pakistan: a review, Environ. Int. 33 (2007) 1107–1122. [3] M.J. Cerejeira, P. Viana, S. Batista, T. Pereira, E. Silva, M.J. Valério, A. Silva, M. Ferreira, A.M. Silva-Fernandes, Pesticides in Portuguese surface and groundwaters, Water Res. 37 (2003) 1055–1063. [4] J.L. Pereira, S.C. Antunes, B.B. Castro, C.R. Marques, A.M.M. Gon alves, F. Gonalves, R. Pereira, Toxicity evaluation of three pesticides on non-target aquatic and soil organisms: commercial formulation versus active ingredient, Ecotoxicology 18 (2009) 455–463. [5] R. Darren, H. Renate, B. Nick, D. Andrew, F. Mohamed, E. Michael, E. Peter, Clinical outcomes and kinetics of propanil following acute self-poisoning: a prospective case series, BMC Clin. Pharmacol. 9 (2009). [6] N. Crossland, A review of the fate and toxicity of 3, 4-dichloroaniline in aquatic environments, Chemosphere 21 (1990) 1489–1497. [7] K. Mitsou, A. Koulianou, D. Lambropoulou, P. Pappas, T. Albanis, M. Lekka, Growth rate effects, responses of antioxidant enzymes and metabolic fate of the
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Journal of Hazardous Materials 196 (2011) 420–425
Contents lists available at SciVerse ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Degradation of crystal violet by an FeGAC/H2 O2 process Chiing-Chang Chen a , Wen-Ching Chen b , Mei-Rung Chiou b , Sheng-Wei Chen b , Yao Yin Chen b , Huan-Jung Fan b,∗ a b
Department of Science Application and Dissemination, National Taichung University, of Education, Taichung, Taiwan, ROC Department of Safety, Health and Environmental Engineering, Hungkuang University, Taichung, Taiwan, ROC
a r t i c l e
i n f o
Article history: Received 4 July 2011 Received in revised form 2 September 2011 Accepted 12 September 2011 Available online 16 September 2011 Keywords: Crystal violet Degradation FeGAC FeGAC/H2 O2
a b s t r a c t Because of the growing concern over highly contaminated crystal violet (CV) wastewater, an FeGAC/H2 O2 process was employed in this research to treat CV-contaminated wastewater. The experimental results indicated that the presence of iron oxide-coated granular activated carbon (FeGAC) greatly improved the oxidative ability of H2 O2 for the removal of CV. For instance, the removal efficiencies of H2 O2 , GAC, FeGAC, GAC/H2 O2 and FeGAC/H2 O2 processes were 10%, 44%, 40%, 43% and 71%, respectively, at test conditions of pH 3 and 7.4 mM H2 O2 . FeGAC/H2 O2 combined both the advantages of FeGAC and H2 O2 . FeGAC had a good CV adsorption ability and could effectively catalyze the hydrogen peroxide oxidation reaction. Factors (including pH, FeGAC dosage and H2 O2 dosage) affecting the removal of CV by FeGAC/H2 O2 were investigated in this research as well. In addition, the reaction intermediates were separated and identified using HPLC-ESI–MS. The N-demethylation step might be the main reaction pathway for the removal of CV. The reaction mechanisms for the process proposed in this research might be useful for future application of this technology to the removal of triphenylmethane (TPM) dyes. © 2011 Elsevier B.V. All rights reserved.
1. Introduction More than 10,000 different synthetic dyes and pigments are produced annually worldwide. The annual production of dyes and pigments is more than 7 × 105 tons, and it has been estimated that approximately 5–15% is lost in industrial effluents [1–3]. The large amount of wastewater generated in textile manufacturing processes represents an increasing environmental danger because of the refractory carcinogenic nature of the dyes [4–6]. For example, triphenylmethane (TPM) dyes are suitable for a large variety of technological applications. They are used extensively in the textile industry for dyeing nylon, wool, cotton, and silk, as well as for coloring oil, fats, waxes, varnish, and plastics. The paper, leather, cosmetic, and food industries are major consumers of various TPM dyes as well. A great deal of concern has arisen regarding the thyroid peroxidase-catalyzed oxidation of the TPM class of dyes because of the generation of N-dealkylated primary and secondary aromatic amines, which have structures similar to hazardous aromatic amine carcinogens [7,8]. Crystal violet (CV) is a typical triphenylmethane dye; it is toxic to mammalian cells and is a mutagen and a mitotic poison [9]. Therefore, CV was chosen as the model pollutant in this research.
∗ Corresponding author. Tel.: +886 4 26318652; fax: +886 4 26525245. E-mail address:
[email protected] (H.-J. Fan). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.042
A chemical coagulation process could remove dyes reasonably well, but it would produce a large amount of sludge, which causes another waste problem. Activated carbon processes can remove dyes in aqueous solution effectively, but activated carbon and the subsequent treatment of the spent carbon are costly [2,10–13]. Therefore, more efficient and economical technologies are needed for the treatment of dye-contaminated wastewater. Consequently, there are many researchers trying to explore processes that are more efficient to remove dye from wastewater. For example, both photolytic oxidation (TiO2 /UV or VIS [14–16], ZnO/UV [17], and H2 O2 /UV [18]), and Fenton’s regents (Fenton [12], Fe2+ /H2 O2 /UV [18], and Fe3+ /UV [19]) processes have been reported in literatures with promising results. In addition, Alshamsi et al. reported that Fenton’s process had better CV degradation efficiency than that of photolytic process (H2 O2 /UV) [20]. However, the research also reported that the Fenton process was dramatically inhibited by the presence of hydrogen phosphate ions. In addition, one of the major limitations of the Fenton’s process was that the process would produce a large amount of sludge, which caused another waste problem. Therefore, the purpose of this research is to develop an alternative oxidation process using FeGAC/H2 O2 for the removal of CV in aqueous solutions. With the iron oxides coated on the surface of the GAC (granular activated carbon), the production of sludge can be minimized. The FeGAC/H2 O2 processes combine both advantages of iron oxide coated granular activated carbon (FeGAC) and H2 O2 . FeGAC has high adsorption capacity of dyes due to GAC surface and the coating of iron oxides. In addition, the oxidation ability
C.-C. Chen et al. / Journal of Hazardous Materials 196 (2011) 420–425
CH 3
CH 3
N + CH 3
N H3C
-
Cl
H 3C
N
CH 3
Fig. 1. Chemical structure of crystal violet (CV).
of H2 O2 is enhanced by the catalytic properties of FeGAC. In previous study, the FeGAC/H2 O2 process could effectively remove acid black 24 [10] and humic acids [21], respectively. However, both studies did not identify the reaction intermediates and the reaction pathways. To obtain a better understanding of the mechanistic details of FeGAC/H2 O2 process for degradation of CV dye, the reaction intermediates of the process are identified in this research as well. 2. Materials and methods CV was obtained from Tokyo Kasei Kogyo Co. and confirmed as a pure organic compound using HPLC analysis. The chemical structure of the CV is shown in Fig. 1. Stock solutions containing 2.69 mM (1000 mg/L) of CV in aqueous solution were prepared, protected from light, and stored at 4 ◦ C. Hydrogen peroxide (35%) was purchased from Sigma–Aldrich Chemical Co. The granular activated carbon (GAC) employed herein was offered by Calgon Carbon Corporation (F400) and was provided courtesy of the manufacturer. All of the other chemicals used in this study were reagent grade. FeGAC (iron oxide-coated granular activated carbon) was prepared as follows. A known amount of FeSO4 solid was dissolved in deionized water and then mixed with GAC. The applied iron dosage in this research was approximately 40 mg Fe (II)/g GAC. This suspension was mixed for 12 h before being dried in an oven at 90 ◦ C for 3 d. The resulting mixture was cooled to room temperature and washed several times with double-distilled water to remove detachable iron oxide. The resulting FeGAC composite adsorbent was dried in the oven at 105 ◦ C for another 3 d and stored at room temperature in a covered glass container until needed. The amount of Fe oxide coating on the FeGAC surface was measured by extracting the composite FeGAC in a boiling, concentrated
421
(10%) HNO3 solution for 12 h. The total iron in the extraction solution was measured by flame atomic absorption spectrophotometry. The BET specific surface areas of the samples (FeGAC and GAC) were measured with an automatic system (Micromeritics Gemini 2370C) with nitrogen gas as the adsorbate. The surface morphology of samples was examined by the scanning electron microscopy (SEM). The SEM analysis was carried out on the JSM-7401F model. A Waters ZQ LC/MS system equipped with a Waters 1525 Binary HPLC pump, a Waters 2998 Photodiode Array Detector, a Waters 717 plus Auto sampler, and a Waters micromass-ZQ3100 Detector was used to identify CV and the reaction intermediates. Solvents A and B were used as the mobile phase. Solvent A was 25 mM aqueous ammonium acetate buffer, and solvent B was methanol. The mobile phase flow rate was 1.0 mL/min. An AtlantisTM dC18 column (250 mm × 4.6 mm i.d., dp = 5 m) was used, and the column effluent was introduced into the ESI source of the mass spectrometer. The quadrupole mass spectrometer was equipped with an ESI interface with a heated nebulizer probe at 350 ◦ C and an ion source temperature of 80 ◦ C. The removal efficiencies of five treatment processes (GAC, FeGAC, H2 O2 , GAC/H2 O2 , and FeGAC/H2 O2 ) were studied in this research in a batch reactor. A known amount of adsorbent (GAC or FeGAC) and/or H2 O2 was added into a sealed 1 L reactor filled with 10 mg/L CV and the pH adjusted to the desired value. Samples were collected and analyzed at predetermined time intervals. Blanks containing no GAC or FeGAC were used for each series of experiments as controls. All samples were filtered through filter paper prior to analysis and treatment. The samples were analyzed in triplicate within the accepted analytical error (±5%). 3. Results and discussion 3.1. Characteristics of FeGAC SEM images of GAC and FeGAC are shown in Fig. 2. As shown in Fig. 2b, there were apparently some iron oxides coated on the surface of the GAC. The amount of iron oxides on FeGAC surfaces was approximately 38 mg Fe/g GAC. The BET surface areas for GAC and FeGAC were approximately 810 and 745 m2 /g, respectively. The lower surface area of FeGAC was another indicator of the coating of iron oxides on the GAC surface. The loss of iron oxide on the FeGAC after the treatment of CV was tested. The loss of iron oxide was less than 5%. 3.2. Treatment efficiency of FeGAC/H2 O2 One of the primary aims of this investigation was to study the removal efficiency of CV by the FeGAC/H2 O2 process. Five treatment processes (GAC, FeGAC, H2 O2 , GAC/H2 O2 , and FeGAC/H2 O2 ) were
Fig. 2. SEM images of (a) GAC and (b) FeGAC.
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C.-C. Chen et al. / Journal of Hazardous Materials 196 (2011) 420–425
a
100
100 0.5 g 1g 1.5 g
3.5 mM 80
Removal (%)
Removal (%)
80
60
60 40
40
20 20
0
Fig. 3. CV removal efficiencies by various processes (CV = 10 mg/L, adsorbent 1.5 g/L, 7.4 mM H2 O2 , pH 3, 30 min).
conducted in this research to evaluate their CV removal efficiency. The removal efficiencies of the H2 O2 , GAC, FeGAC, GAC/H2 O2 and FeGAC/H2 O2 processes were approximately 10%, 44%, 40%, 43% and 71%, respectively (Fig. 3). H2 O2 alone is not an effective process for the treatment of CV; FeGAC/H2 O2 had the highest CV removal efficiency. The removal efficiency of FeGAC/H2 O2 (71%) was 20% higher than that of the FeGAC (40%) and H2 O2 (10%) processes combined. This result indicated that FeGAC could enhance the oxidation ability of H2 O2 . On the other hand, the removal efficiency of GAC/H2 O2 was approximately the same as that of the GAC process. Therefore, the better removal efficiency of the FeGAC/H2 O2 process was probably due to the catalytic reactions that occurred between coated iron oxides and H2 O2 . This result indicates that the presence of FeGAC greatly improved the oxidative ability of H2 O2 for the removal of CV.
b
100 7.4 mM 80
Removal (%)
GAC GAC/H2O2 FeGAC FeGAC/H2O2 H2O2
60 40 20 0
c
100 11.6 mM 80
Removal (%)
0
60
40
3.3. Factors affecting removal efficiency 20
The treatment efficiencies of these five treatment processes at initial pH values of 3 and 6 are shown in Figs. 4 and 5, respectively. In general, the sequence of the removal efficiencies of these five treatment processes was FeGAC/H2 O2 > GAC/H2 O2 , FeGAC, GAC > H2 O2 . For instance, at pH 3, the removal efficiencies of the FeGAC/H2 O2 , GAC/H2 O2 , FeGAC, GAC and H2 O2 processes were approximately 71%, 42%, 41%, 42% and 9%, respectively (Fig. 4a) at 1.5 g adsorbent/L and 11.4 mM H2 O2 . FeGAC/H2 O2 had a better removal efficiency at pH 3 (Fig. 4) than at pH 6 (Fig. 5). In addition, a higher dosage of FeGAC or GAC had a higher CV removal rate as well. The removal efficiencies for FeGAC/H2 O2 significantly increased from 37% to 70% as the FeGAC increased from 1.0 to 1.5 g/L (Fig. 4a). In contrast, for the FeGAC system, the removal efficiencies only increased from 31% to 44% as the FeGAC dosage increased from 1.0 to 1.5 g/L. For the FeGAC/H2 O2 system, it seems that there was a minimum dosage (critical dosage) requirement for FeGAC for the catalytic reaction to be dominant or effective. This result is further illustrated in Fig. 6. As shown in Fig. 6a (H2 O2 dosage 3.5 mM), at dosages of FeGAC lower than 1.3 g/L, the higher removal efficiencies occurred at high pH values and high FeGAC dosages. However, at higher FeGAC dosages (>1.3 g/L), the better removal efficiencies occurred at low pH values and high FeGAC dosages. When the H2 O2 dosage increased from 3.5 to 7.4 to 11.6 mM, the critical FeGAC dosages decreased from 1.3 to 1.22 to 0.78 g/L, as shown in Fig. 6a–c, respectively. These figures also indicate that if the FeGAC dosages were higher than the corresponding critical values, the use of H2 O2 would be much more efficient.
0 GAC
GAC/H2O2
FeGAC
FeGAC/H2O2
H2O2
Fig. 4. CV removal efficiencies at pH 3 with various dosages of H2 O2 : (a) 3.5 mM, (b) 7.4 mM, and (c) 11.6 mM (CV = 10 mg/L, GAC or FeGAC dosage if added: 䊉 0.5 g/L, 1.0 g/L, 1.5 g/L, 30 min).
It seems that there are two main different removal mechanisms. When the FeGAC dosage is lower than the critical value, the CV removal is mainly contributed by the adsorption process. The increased adsorption at the higher pH is due to the electrostatic attraction between the negatively charged sites of the adsorbents and the positively charged dye molecules. When the pH increased, the electrostatic attraction force of CV with FeGAC surface might increase as well. An increase of pH of the solution decreases the charge density, so the electrostatic repulsion between the positively charged dye and the surface of the adsorbent is lowered, which results in an increase in the extent of adsorption of CV. In addition, the decreased adsorption at the lower pH is probably due to the presence of excess H+ ions competing with the cation groups on the dye for adsorption sites on FeGAC surface. This result is in agreement with the findings of Monash and Pugazhenthi, Saeed et al. and Mittal et al., who have reported that MCM-41 [22], grapefruit peel [23] and bottom ash (BA) [24] had greater percentage adsorption of CV at the higher solution pH, respectively. One the other hand, when the FeGAC dosage is higher than the critical value,
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a
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0.5 g 1g 1.5 g
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0 GAC
GAC/H2O2
FeGAC
FeGAC/H2O2
3.5
4.0
4.5
H2O2
Fig. 5. CV removal efficiencies at pH 6 with various dosages of H2 O2 : (a) 3.5 mM, (b) 7.4 mM, and (c) 11.6 mM (CV = 10 mg/L, GAC or FeGAC dosage if added: 䊉 0.5 g/L, 1.0 g/L, 1.5 g/L, 30 min)).
5.0
5.5
6.0
pH 11.6 mM H 2O2
c
65 60
1.4 65
55 60
3.4. Separation and identification of the intermediates The experimental results indicated that CV could be removed efficiently by the FeGAC/H2 O2 process, as shown in Fig. 7. The reaction intermediates were examined by HPLC using a photodiode array detector and ESI mass spectrometry. Six components (A–F) were observed and identified. The results of HPLC chromatograms, UV–vis spectra, and HPLC-ESI mass spectra are summarized in Table 1. Typical reaction intermediates observed by HPLC are shown in Fig. 8. Peak A is CV, and peaks B–F are reaction intermediates. The distributions of the intermediates are illustrated in Fig. 9. To minimize errors, the relative intensities of each intermediate were recorded at the maximum absorption wavelength of each intermediate.
55
60
1.2
FeGAC (g/L)
the CV removal is mainly contributed by the catalytic reactions between FeGAC and H2 O2 . The better removal efficiencies occur at the lower solution pH. This result is in agreement with the removal of CV by Fenton and Fenton like processes [8].
50 55 50 55
1.0
45
50 50
45
45
0.8
45
40
40
35
40
35
0.6 30
40
35
3.0
3.5
4.0
4.5
5.0
5.5
6.0
pH Fig. 6. CV removal efficiency contour map of pH and FeGAC dosages at (a) 3.5 mM, (b) 7.4 mM, and (c) 11.6 mM of H2 O2 .
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Table 1 Reaction intermediates of CV by HPLC-ESI–MS. HPLC peaks
De-methylation intermediates
ESI-MS spectrum ions (m/z)
Absorption maximum (nm)
A B C D E F
N,N,N ,N ,N ,N -hexaethylpararosaniline N,N-dimethyl-N ,N -dimethyl-N -methyl pararosaniline N,N-dimethyl-N -methyl-N -methylpararosaniline N,N-dimethyl-N ,N -dimethyl pararosaniline N-methyl-N -methyl-N -methyl pararosaniline N,N-dimethyl-N -methylpararosaniline
372.18 358.14 344.10 334. 09 336.36 336.10
588.5 581.2 570.2 579.8 566.3 570.0
reaction pathways for N-demethylation as well: (1) reaction with N,N-dimethyl groups (such as compounds D, F and H) and (2) reaction with N-methyl groups (compounds C, E and G) [6]. Considering that the N,N-dimethyl groups of compounds D, F and H are bulkier than those of the N-methyl groups of compounds C, E and G molecules, respectively, a nucleophilic attack by free radicals on the N-methyl groups might be favored over that on the N,N-dimethyl groups. However, considering that the attack probability on two N, N-dimethyl groups of compound D is higher than that on one N-methyl group in compound C molecules, the attack by free radicals on the N, N-dimethyl group might be favored over that on the N-methyl group. As shown in Fig. 9, compound C had a higher concentration and reached the maximum concentration sooner than that of compound D. This result indicates that the attack on the Nmethyl groups (compound C) is significantly stronger than that on the N,N-dimethyl group (compound D). This result suggests that the removal of CV was N-demethylated in a stepwise manner (i.e., methyl groups were removed one by one, as confirmed by the gradual peak wavelength shifts). This result is similar with Fenton and Fenton-like [8], and TiO2 /UV [9] processes for the removal of CV. Therefore, N-demethylation might be the dominating mechanism for the removal of CV by the FeGAC/H2 O2 process. The cleavage of the CV chromophore ring structure is not the significant step. Based on these experimental results, reaction
Fig. 7. The absorption spectra changes of CV by FeGAC/H2 O2 as a function of time. Spectra from top to bottom correspond to a time of 0, 10, 20, 30, 60, 120, and 180 min, respectively.
3.5. Reaction pathways There are two possible major competitive reaction pathways for degradation of TPM dyes: (1) N-demethylation and (2) cleavage of the CV chromophore ring structure [3,15,20]. Because compounds B–F were observed in this study, the N-demethylation might be the main degradation pathway for the FeGAC/H2 O2 system, as shown in Fig. 10. In addition, there are two possible major competitive
Fig. 8. HPLC chromatogram of the intermediates of the FeGAC/H2 O2 reaction at pH 6, recorded at 580 nm.
1.2
0.025
1.0
A B C D E F
0.020
C/Co
C/Co
0.8 0.6
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0.4
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0.015
0.000 0
10
20
30
Time (min)
40
50
60
0
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20
30
40
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Fig. 9. Variations in the relative distribution of the intermediates for the FeGAC/H2 O2 process.
50
60
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Fig. 10. Proposed pathways for the removal of CV by the FeGAC/H2 O2 process.
pathways for the N-demethylation of CV are proposed and depicted in Fig. 10. This result is similar to the findings reported in literatures [8,9]. 4. Conclusions The FeGAC/H2 O2 process could effectively remove CV in aqueous solutions. This process combined the advantages of both FeGAC and H2 O2 . In general, the removal efficiency sequence among the five processes tested was FeGAC/H2 O2 > GAC/H2 O2 , FeGAC, and GAC > H2 O2 . The factors affecting the CV removal efficiencies included solution pH, FeGAC dosage, and H2 O2 dosage. A minimum FeGAC dosage might be needed to sufficiently speed up the reaction process for the FeGAC/H2 O2 processes. The reaction mechanisms of CV were proposed and discussed in this research. N-demethylation might be the main reaction pathway for the removal of CV from aqueous solutions. Acknowledgements The authors gratefully acknowledge partial funding from National Science Council, Taiwan, ROC (NSC 97-2221-E-241-006MY3). References [1] F.A. Alshamsi, A.S. Albadwawi, M.M. Alnuaimi, M.A. Rauf, S.S. Ashraf, Comparative efficiencies of the degradation of Crystal Violet using UV/hydrogen peroxide and Fenton‘s reagent, Dyes Pigm. 74 (2007) 283–287. [2] W. Azmi, R.K. Sani, U.C. Banerjee, Biodegradation of triphenylmethane dyes, Enzyme Microb. Technol. 22 (1998) 185–191. [3] C.C. Chen, H.-J. Fan, C.Y. Jang, J.L. Jan, H.D. Lin, C.S. Lu, Photooxidative N-demethylation of crystal violet dye in aqueous nano-TiO2 dispersions under visible light irradiation, J. Photochem. Photobiol. A 184 (2006) 147–154. [4] C.C. Chen, C.S. Lu, H.-J. Fan, W.H. Chung, J.L. Jan, H.D. Lin, W.Y. Lin, Photocatalyzed N-de-ethylation and degradation of Brilliant Green in TiO2 dispersions under UV irradiation, Desalination 219 (2008) 89–100. [5] C.C. Chen, F.D. Mai, K.T. Chen, C.W. Wu, C.S. Lu, Photocatalyzed N-demethylation and degradation of crystal violet in titania dispersions under UV irradiation, Dyes Pigm. 75 (2007) 434–442. [6] F. Chen, J. He, J. Zhao, J.C. Yu, Photo-Fenton degradation of malachite green catalyzed by aromatic compounds under visible light irradiation, New J. Chem. 26 (2002) 336–341. [7] B.P. Cho, T. Yang, L.R. Blankenship, J.D. Moody, M. Churchwell, F.A. Beland, S.J. Culp, Synthesis and characterization of N-demethylated metabolites of
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Journal of Hazardous Materials 196 (2011) 426–430
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Short communication
Electrospun nanofibers of Bi-doped TiO2 with high photocatalytic activity under visible light irradiation Jie Xu, Wenzhong Wang ∗ , Meng Shang, Erping Gao, Zhijie Zhang, Jia Ren State Key Laboratory of High Performance Ceramics and Superfine Microstructures, Shanghai Institute of Ceramics, Chinese Academy of Sciences, 1295 Dingxi Road, Shanghai 200050, China
a r t i c l e
i n f o
Article history: Received 3 June 2011 Received in revised form 25 August 2011 Accepted 4 September 2011 Available online 10 September 2011 Keywords: Nanofibers Photocatalysis TiO2 Bismuth Visible light
a b s t r a c t Bi-doped TiO2 nanofibers with different Bi content were firstly prepared by an electrospinning method. The as-prepared nanofibers were characterized by scanning electron microscopy (SEM), X-ray diffraction (XRD), X-ray photoelectron spectroscopy (XPS), photoluminescence spectra (PL), and UV–vis diffuse reflectance spectroscopy (DRS). The results indicated that Bi3+ ions were successfully incorporated into TiO2 and extended the absorption of TiO2 into visible light region. The photocatalytic experiments showed that Bi-doped TiO2 nanofibers exhibited higher activities than sole TiO2 in the degradation of rhodamine B (RhB) and phenol under visible light irradiation ( > 420 nm), and 3% Bi:TiO2 samples showed the highest photocatalytic activities. © 2011 Elsevier B.V. All rights reserved.
1. Introduction In view of increasingly serious environmental pollution and energy crisis, photocatalysis, a “green” technology, plays an important role in solar energy conversion and degradation of organic pollutants [1]. As a most promising photocatalyst, TiO2 possesses many advantages such as suitable redox potentials of conduction band and valence band, chemical inertness, stability against photocorrosion, low cost, and nontoxicity [2]. However, the wide band gap (3.2 eV for anatase) and the low quantum efficiency limit the practical application of TiO2 photocatalyst [3]. It is well known that doping transitional metal ions is one of the most effective methods for synthesizing visible light active TiO2 photocatalysts with high photocatalytic activities [4]. Some researchers have reported that the incorporation of Bi into TiO2 lattice not only widens the absorption range of TiO2 but also acts as an electron acceptor, which is beneficial to the effective separation of photogenerated electrons and holes, thus facilitating photocatalytic reactions [5,6]. Moreover, it should be noted that the morphology of the photocatalysts is another important factor influencing the photocatalytic efficiencies [7,8]. However, such research is still not satisfying considering the practical application of photocatalysts. More effort is needed to obtain Bi-doped TiO2 photocatalyst with desirable separability and excellent photoactivity.
Fibrous nanostructures are favorable for industrial applications in environmental remediation. On one hand, conventional film photocatalysts can be fixed and reclaimed easily, but their low surface area decreased the photocatalytic activities. On the other hand, the application of particulate photocatalysts is limited owing to the difficulties in separation, which may re-pollute the treated water. Compared to film and particulate counterparts, nanofibrous photocatalysts not only possess large specific surface area, which allow for their surface active sites to be accessible for reactants more efficiently, but also owe high length-to-diameter aspect ratio, which makes the separation of photocatalyst more easily [9]. Electrospining is an effective, straightforward, and convenient way to fabricate polymer, polymer/inorganic hybrid, and inorganic fibers [10]. The electrospun nanofibrous photocatalysts have been found to exhibit excellent performance in terms of photocatalytic activity and recycling [11]. In this work, for the first time, Bi-doped TiO2 nanofibers were successfully synthesized by electrospinning. We investigated the effects of Bi doping on photocatalytic activities of as-prepared photocatalysts, and 3% Bi:TiO2 nanofibers showed the highest photocatalytic activity in the degradation of rhodamine B (RhB) and phenol under visible light irradiation ( > 420 nm). 2. Experimental 2.1. Synthesis
∗ Corresponding author. Tel.: +86 21 5241 5295; fax: +86 21 5241 3122. E-mail address:
[email protected] (W. Wang). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.010
All reagents used in our experiments were of analytical purity and were used as received without further purification. Certain
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amounts of Bi(NO3 )3 ·5H2 O were dissolved in the 1.0 g of N,Ndimethylformamide (DMF) in a capped bottle, then 1.0 g of absolute ethanol and 1.87 g of Ti(OC4 H9 )4 were added into the mixture (the atomic ratio of Bi:Ti ranged from 1% to 3%). The mixture was stirred for several minutes, and 0.28 g of poly(vinyl pyrrolidone) with a molecular weight of 1.3 × 106 was slowly added. After being stirred for 1 h, the solution was loaded in a syringe. The voltage applied to the needle of the syringe was 15 kV and the distance between the tip of needle and aluminum foil was about 20 cm (The setup for electrospinning showed in Fig. S1). The as-collected fibers were calcined at 500 ◦ C for 3 h in air. The as-synthesized samples were denoted as x% Bi:TiO2 , where x refers to the molar ratio of Bi:Ti. The TiO2 nanofibers were prepared by the same method as described above without adding of Bi(NO3 )3 ·5H2 O. 2.2. Characterization The phase and composition of as-prepared samples were recorded by X-ray diffraction (XRD) studies using an X-ray diffractometer (Rigaku, Japan). The morphologies and microstructures were performed on a JEOL JSM-6700F field emission scanning electron microscope (SEM). UV–vis diffuse reflectance spectra (DRS) of the sample were recorded with a UV–vis spectrophotometer (Hitachi U-3010). The chemical states of Bi in TiO2 were analyzed using the X-ray photoelectron spectroscopy (XPS) attachment of a Microlab310F auger microprobe. The photoluminescence spectra (PL) were obtained on a spectrofluorophotometer (Shimadzu RF-5301PC) with a Xe lamp as the excitation source at room temperature. 2.3. Photocatalytic test Photocatalytic activities of the Bi-doped TiO2 nanofibers were evaluated by the degradation of rhodamine B (RhB) and phenol under visible light irradiation of a 500 W Xe lamp with a 420 nm cut-off filter. 0.05 g of the photocatalyst was added into 50 mL of RhB (1 × 10−5 M) or 50 mL of phenol (20 mg L−1 ). Before illumination, the solution was stirred for 30 min in the dark in order to reach adsorption–desorption equilibrium between organic substrates and the photocatalyst. At given irradiation time intervals, a 4 mL solution was sampled. Then, the absorption UV–vis spectrum of the centrifuged solution was recorded with use of a Hitachi U3010 UV–vis spectrophotometer. In order to prove the visible light photoresponse in actual photodegradation process, 3 W light emitting diodes (LEDs) with different wavelengths were used to excite as-prepared photocatalysts. The photodegradation was carried out in a 20 mL suspension of phenol (20 mg L−1 ) with 2% Bi:TiO2 photocatalyst (1 g/L). After 4 h irradiation, the corresponding ratio of final concentration to initial concentration (C/C0 ) was recorded. 3. Results and discussion 3.1. Characterization of the photocatalysts The crystal phases of the different materials were investigated by performing XRD analysis. Fig. 1a shows the XRD patterns of sole TiO2 and Bi-doped TiO2 samples with different molar ratio of Bi:Ti, demonstrating that all the samples were composed of anatase TiO2 (JCPDS No.99-0008) and rutile TiO2 (JCPDS No.99-0090). The introduction of Bi into TiO2 caused the broadening of diffraction peaks, suggesting lower crystallization degree of Bi-doped TiO2 samples than that of sole TiO2 , and inhibited the transformation of TiO2 from anatase-to-rutile, especially for 1% Bi:TiO2 . No diffraction peaks indicative of the Bi species could be observed even at Bi:Ti molar ratio up to 3.0% due to low content of Bi. In order to reveal the presence of Bi and its electronic state, X-ray photoelectron spectroscopy
Fig. 1. (a) XRD patterns of TiO2 and Bi-doped TiO2 samples and (b) Bi 4f XPS profile of 3% Bi:TiO2 sample.
(XPS) with high resolution is used in our study. Fig. 1b presents the Bi 4f XPS spectra of 3% Bi:TiO2 sample. The peaks of 160.15 eV and 165.35 eV were close to the 4f7/2 level and 4f5/2 level of Bi3+ [12], indicating that the incorporated Bi existed in the form of Bi3+ . The morphologies and microstructures of TiO2 and 3% Bi:TiO2 samples prepared by electrospining are revealed by the SEM images. As shown in Fig. 2a and c, both TiO2 and 3% Bi:TiO2 exhibited nanofibrous morphologies, and the lengths of these randomly oriented nanofibers could reach several micrometers. The close-up view of TiO2 nanofibers (Fig. 2b) demonstrated that the diameter of TiO2 nanofibers ranged from 100 nm to 200 nm, whereas the diameter of 3% Bi:TiO2 nanofibers was between 150 nm and 300 nm as shown in Fig. 2d. It could be observed that the electrospun nanofibers possessed large length-to-diameter aspect ratios, which were beneficial for separation of the photocatalysts from reaction solutions, thus avoiding the possible pollution introduced by unseparated catalysts. In addition, the nanofibers had a rough surface (as shown in Fig. 2b and d) due to the decomposition of PVP during the calcination process, and such rough surface may result in larger surface areas allowing for more efficient contact between reactant molecules and photocatalysts, leading to a higher photocatalytic reaction rate. The optical absorption property, which is relevant to the electronic structure feature, is considered as a key factor in determining photocatalytic behavior. Fig. 3 presents the UV–vis diffuse reflectance spectroscopy (DRS) of as-prepared samples. Sole TiO2 displayed no obvious absorbance in visible light region due to its large band gap (3.2 eV for anatase, and 3.0 eV for rutile). All Bi-doped TiO2 samples exhibited spectral response in the visible region (from
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Fig. 2. SEM images of TiO2 and 3% Bi:TiO2 : (a) low magnification and (b) high magnification of TiO2 ; (c) low magnification and (d) high magnification of 3% Bi:TiO2 .
400 nm to 600 nm), and the absorbance became stronger with the increase of Bi content from 0% to 2.0% because of narrower band gap induced by the energy state of Bi3+ 6s lone pairs lying above the valence band of sole TiO2 [13]. However, further increase of Bi content could not enhance the visible absorbance, which was possibly attributed to the lower dispersion of Bi species in TiO2 . In order to prove the effective utilization of absorbed visible light in the photocatalytic process, the photodegradation of phenol was carried out under 3 W LEDs with different wavelengths and Fig. S2 suggests the visible light photoresponse of Bi-doped TiO2 in the actual photodegradation process. Photocatalytic activity is closely related to the efficiency of the photogenerated electron–hole separation and the transfer from the
inner regions to the outer surfaces. PL emission spectra have been widely used to investigate the efficiency of charge carrier trapping, immigration, transfer, and to understand the fate of photogenerated electrons and holes in the semiconductor [14]. As shown in Fig. 4, TiO2 and 3% Bi:TiO2 exhibited emission in the range of 350–550 nm. As for sole TiO2 , the peaks at 435 nm, 449 nm, 451 nm, 467 nm, 491 nm were originated from the surface traps of TiO2 [15,16]. Compared to sole TiO2 , 3% Bi:TiO2 exhibited a decrease in the PL intensity, which indicated that 3% Bi:TiO2 possessed a lower recombination rate of photogenerated electrons and holes. This was possibly ascribed to the fact that Bi species reduced the number of trap states on the surface of TiO2 , thus improved the separation of photogenerated charge carriers [17]. The decreased recombination
Fig. 3. UV–vis diffuse reflectance spectroscopy (DRS) of sole and Bi-doped TiO2 samples.
Fig. 4. PL emission spectra of TiO2 and 3% Bi:TiO2 samples (excitation wavelength is 300 nm).
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Scheme 1. Proposed mechanism of photocatalytic reaction on Bi-doped TiO2 photocatalyst.
Fig. 5. Photocatalytic degradation profiles of (a) RhB and (b) phenol by TiO2 and Bi-doped TiO2 under visible irradiation ( > 420 nm).
of photogenerated charges will provide more opportunities for photoinduced charge carriers to participate in the photocatalytic reactions occurring on the semiconductor surface. 3.2. Photocatalytic activity and proposed mechanism It is well known that the colored rhodamine B (RhB) and uncolored phenol are two kinds of chemicals used in industrial production, which often cause water pollution. Therefore, they are selected as model pollutants to evaluate the photocatalytic activity of the as-prepared TiO2 samples. Fig. 5a shows the photodegradation of RhB under visible irradiation. Since RhB exhibits absorption band around 553 nm, sole TiO2 degraded 88.8% of RhB due to the dye sensitization effect [2]. When the Bi:Ti ratio was increased from 0% to 3%, the photocatalytic activities became higher, and RhB was completely degraded in 90 min by 3% Bi:TiO2 . Clearly, doping TiO2 with Bi3+ enhanced the photocatalytic activity, which may be attributed to two reasons. First, the energy band is assumed between the top of the (lone pair) Bi3+ 6s band and the bottom of the Ti4+ 3d band, giving rise to narrower energy band gap compared to that of TiO2 [13], which is in accordance with the observation of UV–vis DRS. Hence, the absorption spectra shift into visible light region so that lower energy photons can be absorbed for photocatalytic reaction. Second, Bi species doped in TiO2 can inhibit the electron and hole recombination by capturing the photoinduced charge carriers as illustrated in PL spectra. Also, P25 (TiO2 ), as a well known standard, was used in our work and possessed higher activity for degrading RhB compared to as-prepared samples. On the other hand, Fig. 5b represents the variation of phenol
concentrations (C/C0 ) with the irradiation time over TiO2 and Bi:TiO2 samples under visible light illumination. Sole TiO2 degraded very little phenol, only about 5% of phenol in 4 h, which was attributed to the poor ability of TiO2 in harvesting visible lights. TiO2 (P25) exhibited very low photocatalytic activity with about 13% of phenol degraded. However, doping TiO2 with a small amount of Bi species significantly increased the degradation rate, and 3% Bi:TiO2 showed the highest photocatalytic activity with 56.1% of phenol degraded. In addition, compared to the photodegradation of RhB, the photodegradation of colorless phenol was almost ascribed to the photocatalytic process rather than the dye photosensitization effect. Based on the above discussions, the level of Bi 6s [8,13,18], which is located above the valence band of TiO2 , plays an important role in reducing the energy band gap of TiO2 , thus enhancing the photocatalytic activity under visible irradiation. Scheme 1 shows the energy band diagram and possible mechanism of photocatalytic reaction on Bi-doped TiO2 , and the detailed photocatalytic processes are proposed as follows (S represents the organic pollutants): Bi : TiO2 → Bi6s (h+ ) + TiO2 (e− ) OH− + h+ → HO• O2 + e− → O2 •− O2 •− + H+ → HO2 • HO2 • + e− → HO2 − HO2 − + H+ → H2 O2 H2 O2 + e− → OH− +HO• S + HO• → CO2 + H2 O When the Bi:TiO2 photocatalyst is radiated by visible light with a photon energy higher or equal to the band gap between the top of the (lone pair) Bi3+ 6s band and the bottom of the Ti4+ 3d band, electrons (e− ) in the 6s level of Bi3+ can be excited into the VB of TiO2 , leaving same amount of holes (h+ ) in the 6s level of Bi3+ . Then holes are captured by surface hydroxyl groups (OH− ) at the photocatalyst surface to yield hydroxyl radicals (HO• ) [19], and the electrons are trapped by the dissolved oxygen molecules (O2 ), producing superoxide anions (O2 •− ) [20]. The formed superoxide anions (O2 •− ) may either attack the organic molecules directly or generate hydroxyl radicals (HO• ) by reacting with hydrion (H+ ) and photogenerated
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electrons [21,22]. Afterward, hydroxyl radicals (HO• ), as strong oxidizing agents, degrade the organic molecules [23]. 4. Conclusion Bi-doped TiO2 nanofibers with variable Bi:Ti ratios have been successfully fabricated by an electrospinning method. Doping Bi species improved the separation of photogenerated electron and hole and extended the absorption of TiO2 into visible light region due to the new energy states introduced by Bi, and 3% Bi:TiO2 samples showed the highest photocatalytic activities. It is expected that such nanofibrous structures would provide great impetus to the industrialization of photocatalysts in the wastewater remediation. Acknowledgement We acknowledge the financial support from the National Natural Science Foundation of China (50972155, 50902144, 50732004), National Basic Research Program of China (2010CB933503) and the Nanotechnology Programs of Science and Technology Commission of Shanghai (0952nm00400). Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at doi:10.1016/j.jhazmat.2011.09.010. References [1] M.R. Hoffmann, S.T. Martin, W.Y. Choi, D.W. Bahnemann, Environmental applications of semiconductor photocatalysis, Chem. Rev. 95 (1995) 69–96. [2] C.C. Chen, W.H. Ma, J.C. Zhao, Semiconductor-mediated photodegradation of pollutants under visible-light irradiation, Chem. Soc. Rev. 39 (2010) 4206–4219. [3] X. Chen, S.S. Mao, Titanium dioxide nanomaterials: synthesis, properties, modifications, and applications, Chem. Rev. 107 (2007) 2891–2959. [4] J. Choi, H. Park, M.R. Hoffmann, Effects of single metal-ion doping on the visiblelight photoreactivity of TiO2 , J. Phys. Chem. C 114 (2010) 783–792. [5] H.Y. Li, D.J. Wang, P. Wang, H.M. Fan, T.F. Xie, Synthesis and studies of the visible-light photocatalytic properties of near-monodisperse Bi-doped TiO2 nanospheres, Chem. Eur. J. 15 (2009) 12521–12527. [6] Y.Q. Wu, G.X. Lu, S.B. Li, The doping effect of Bi on TiO2 for photocatalytic hydrogen generation and photodecolorization of Rhodamine B, J. Phys. Chem. C 113 (2009) 9950–9955.
[7] S. Sajjad, S.A.K. Leghari, F. Chen, J.L. Zhang, Bismuth-doped ordered mesoporous TiO2 : visible-light catalyst for simultaneous degradation of phenol and chromium, Chem. Eur. J. 16 (2010) 13795–13804. [8] Z.F. Bian, J. Ren, J. Zhu, S.H. Wang, Y.F. Lu, H.X. Li, Self-assembly of Bix Ti1−x O2 visible photocatalyst with core-shell structure and enhanced activity, Appl. Catal. B 89 (2009) 577–582. [9] Y. Yang, C.C. Zhang, Y. Xu, H.Y. Wang, X. Li, C. Wang, Electrospun Er:TiO2 nanofibrous films as efficient photocatalysts under solar simulated light, Mater. Lett. 64 (2010) 147–150. [10] A. Greiner, J.H. Wendorff, Electrospinning: a fascinating method for the preparation of ultrathin fibres, Angew. Chem. Int. Ed. 46 (2007) 5670–5703. [11] M. Shang, W.Z. Wang, J. Ren, S.M. Sun, L. Wang, L. Zhang, A practical visiblelight-driven Bi2WO6 nanofibrous mat prepared by electrospinning, J. Mater. Chem. 19 (2009) 6213–6218. [12] C.H. Wang, C.L. Shao, L.J. Wang, L. Zhang, X.H. Li, Y.C. Liu, Electrospinning preparation, characterization and photocatalytic properties of Bi2 O3 nanofibers, J. Colloid Interface Sci. 333 (2009) 242–248. [13] W.F. Yao, H. Wang, X.H. Xu, X.F. Cheng, J. Huang, S.X. Shang, X.N. Yang, M. Wang, Photocatalytic property of bismuth titanate Bi12 TiO20 crystals, Appl. Catal. A 243 (2003) 185–190. [14] H. Yamashita, Y. Ichihashi, S.G. Zhang, Y. Matsumura, Y. Souma, T. Tatsumi, M. Anpo, Photocatalytic decomposition of NO at 275 K on titanium oxide catalysts anchored within zeolite cavities and framework, Appl. Surf. Sci. 121 (1997) 305–309. [15] B.S. Liu, X.J. Zhao, L.P. Wen, The structural and photoluminescence studies related to the surface of the TiO2 sol prepared by wet chemical method, Mater. Sci. Eng. B 134 (2006) 27–31. [16] T. Toyoda, T. Hayakawa, K. Abe, T. Shigenari, Q. Shen, Photoacoustic and photoluminescence characterization of highly porous, polycrystalline TiO2 electrodes made by chemical synthesis, J. Lumin. (2000) 1237–1239, 87–9. [17] S. Shamaila, A.K.L. Sajjad, F. Chen, J.L. Zhang, Study on highly visible light active Bi2 O3 loaded ordered mesoporous titania, Appl. Catal. B 94 (2010) 272– 280. [18] H. Mizoguchi, K. Ueda, H. Kawazoe, H. Hosono, T. Omata, S. Fujitsu, New mixedvalence oxides of bismuth: Bi1 − xYxO1.5 + delta (x = 0.4), J. Mater. Chem. 7 (1997) 943–946. [19] N. Zhang, S.Q. Liu, X.Z. Fu, Y.J. Xu, Synthesis of M@TiO2 (M = Au, Pd, Pt) coreshell nanoconnposites with tunable photoreactivity, J. Phys. Chem. C 115 (2011) 9136–9145. [20] J. Yang, C.C. Chen, H.W. Ji, W.H. Ma, J.C. Zhao, Mechanism of TiO2 -assisted photocatalytic degradation of dyes under visible irradiation: photoelectrocatalytic study by TiO2 -film electrodes, J. Phys. Chem. B 109 (2005) 21900– 21907. [21] L.R. Zheng, Y.H. Zheng, C.Q. Chen, Y.Y. Zhan, X.Y. Lin, Q. Zheng, K.M. Wei, J.F. Zhu, Network structured SnO2 /ZnO heterojunction nanocatalyst with high photocatalytic activity, Inorg. Chem. 48 (2009) 1819–1825. [22] H.C. Yatmaz, A. Akyol, M. Bayramoglu, Kinetics of the photocatalytic decolorization of an Azo reactive dye in aqueous ZnO suspensions, Ind. Eng. Chem. Res. 43 (2004) 6035–6039. [23] O. Legrini, E. Oliveros, A.M. Braun, Photochemical processes for watertreatment, Chem. Rev. 93 (1993) 671–698.
Journal of Hazardous Materials 196 (2011) 431–435
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Short communication
Procedure to use phosphogypsum industrial waste for mineral CO2 sequestration C. Cárdenas-Escudero a,b , V. Morales-Flórez b,∗ , R. Pérez-López c,d , A. Santos e , L. Esquivias a,b a
Departamento de Física de la Materia Condensada, Facultad de Física, Universidad de Sevilla, Av. Reina Mercedes s/n, 41012 Seville, Spain Instituto de Ciencia de Materiales de Sevilla (CSIC-US), Av. Américo Vespucio, 49, 41092 Seville, Spain c Departamento de Geología, Facultad de Ciencias Experimentales, Universidad de Huelva, Campus Universitario Campus del Carmen, Avenida de las Fuerzas Armadas, 21071 Huelva, Spain d Instituto de Diagnóstico Ambiental y Estudios del Agua (IDÆA-CSIC), Jordi Girona 18, 08034 Barcelona, Spain e Departamento de Ciencias de la Tierra, Universidad de Cádiz, Campus del Río San Pedro, Av. República Saharaui s/n, 11510 Puerto Real, Spain b
a r t i c l e
i n f o
Article history: Received 5 June 2011 Received in revised form 7 September 2011 Accepted 12 September 2011 Available online 16 September 2011 Keywords: Phosphogypsum CO2 capture Separation impurities Industrial waste Alkaline solution Mineral sequestration
a b s t r a c t Industrial wet phosphoric acid production in Huelva (SW Spain) has led to the controversial stockpiling of waste phosphogypsum by-products, resulting in the release of significant quantities of toxic impurities in salt marshes in the Tinto river estuary. In the framework of the fight against global climate change and the effort to reduce carbon dioxide emissions, a simple and efficient procedure for CO2 mineral sequestration is presented in this work, using phosphogypsum waste as a calcium source. Our results demonstrate the high efficiency of portlandite precipitation by phosphogypsum dissolution using an alkaline soda solution. Carbonation experiments performed at ambient pressure and temperature resulted in total conversion of the portlandite into carbonate. The fate of trace elements present in the phosphogypsum waste was also investigated, and trace impurities were found to be completely transferred to the final calcite. We believe that the procedure proposed here should be considered not only as a solution for reducing old stockpiles of phosphogypsum wastes, but also for future phosphoric acid and other gypsum-producing industrial processes, resulting in more sustainable production. © 2011 Elsevier B.V. All rights reserved.
1. Introduction The wet-acid process for manufacture of phosphoric acid (i.e. H3 PO4 ) for fertilizers involves the chemical attack of phosphate rock ore (mainly apatite, Ca5 (PO4 )3 OH) with sulphuric acid (H2 SO4 ), generating a gypsum by-product known as phosphogypsum (CaSO4 ·nH2 O). The phosphogypsum waste is usually slurried with water and then pumped out of the fertilizer industrial plant to nearby settling/disposal area, using a system of pipes. Recycling of phosphogypsum is limited by the high content of metallic impurities and radionuclides [1]. In Spain, the phosphoric acid production began in 1968 in an industrial complex located at the estuary formed by the union of the Tinto and Odiel river mouths (Huelva, SW Spain). The phosphogypsum has been stockpiled over an area of 1200 ha containing about 120 million tonnes on the salt marshes associated with the right margin of the Tinto river, less than 1 km away from the city centre. In fact, the area covered by the phosphogypsum stack is roughly similar to the surface area covered by the city of Huelva itself, with a population of 149,000. The proximity of the waste to
∗ Corresponding author. E-mail address:
[email protected] (V. Morales-Flórez). 0304-3894/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.039
the city has also aroused considerable controversy for its alleged implications for the health of the local population; however, most studies have concluded that populations living close to stacks are not exposed to any significant health risk (see for example [2]). Moreover, salt marshes on the Tinto–Odiel estuarine system hold an important part of Europe’s ornithological biodiversity and they were declared UNESCO Biosphere Reserve in 1983 and RAMSARNATURA wetlands sites in 1989. Sudden changes in land-use and direct dumping of phosphogypsum on these salt marshes dramatically altered the visual landscape and degraded the marshland occupied by the stack. The high content of metals and U–Th series radionuclides in phosphogypsum and the impact of these wastes on the quality of sediments and waters of surrounding environmental receptors have been widely described [3–6]. After two decades of looking for sustainable solutions and alternatives to the stockpiling of wastes, the Huelva’s fertilizer industrial complex ceased dumping of phosphogypsum over salt marshes in December 2010 following a decision of the Spanish Major National Court. Currently, the growing interest for the environmental restoration encourages the search of possible low-cost applications to phosphogypsum waste. Few investigations have been so far reported in the literature for this purpose; and up to now, the only reported use of this phosphogypsum waste is as an additive, in six doses of 20–25 t/ha, to improve fertility and reduce sodium saturation in an area of 140 km2 of agricultural soils [7].
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reach a OH− /Ca molar ratio of 2. The mixture was kept at room temperature for 3 h under constant stirring. The phosphogypsum dissolution resulted in the precipitation of a whitish solid phase and a supernatant liquid. Subsequent evaporation of the supernatant to dryness on a hot plate at 80 ◦ C yielded transparent salts, which were later characterized as a second solid phase. The carbonation capacity and efficiency of the solid phase precipitated in the dissolution experiments were assessed by a simple carbonation experimental set-up: 2 g of this precipitate were dispersed in 40 ml of high-purity water under magnetic stirring into a reactor, and a CO2 flux (1 bar, 20 cm3 /s) was bubbled through the suspension for 15 min at room pressure and temperature. Afterwards, the sample was left to rest overnight into the CO2 -rich water. The resulting solid phase was separated by centrifugation and dried in air at 80 ◦ C, and the supernatant discarded. In Fig. 1, the entire experimental procedure is sketched. Chemical characterization of the raw phosphogypsum and the solid products from dissolution and carbonation experiments was performed by X-ray fluorescence (XRF; AXIOS Panalytical instrument) for major elements and pseudo-total acid-digestion followed by analysis with inductively coupled plasma mass-spectrometry (ICP-MS; HP-4500 instrument) for trace elements. Crystalline phases of the samples were identified by X-ray diffraction (XRD) in a diffractometer (Philips X’Pert) with Cu K␣ radiation, from 5.00◦ to 70.00◦ with a step of 0.05◦ and counting time of 80 s. The carbonation degree of the samples was studied by thermogravimetric analyses (TGA; STD Q600) carried out under a nitrogen flux of 100.0 ml/min, starting from ambient temperature and increasing by 10 ◦ C/min up to 1000 ◦ C.
However, this practice was halted in 2001 due to public concern about safety. Thus, under the framework of the integrated waste management, it is crucial to find the best alternative use for the 120 million tonnes of discarded gypsum by-products without damaging the natural environment. The proposal discussed in this work concerns the utilization of this waste as CO2 sequester agent. Several carbon dioxide sequestration strategies are being studied worldwide to reduce anthropogenic greenhouse gas emissions, and hence, mitigate global warming. Mineral sequestration [8] is a promising approach to the problem of managing and capturing carbon dioxide permanently and safely, and has the potential to sequester CO2 emissions directly from localised sources, mobile sources (e.g. transport), and even past CO2 emissions. The mineral sequestration process involves a reaction where aqueous ions (mainly Ca and Mg from silicates [9] or hydroxides) react with CO2 to form stable carbonate minerals and it has controlled the CO2 content on the atmosphere for millennia. The costs associated with industrial scale mineral sequestration are a major drawback of this technology, but they could be significantly reduced by using industrial alkaline wastes as aqueous Ca and Mg sources [10–14]. In light of the above, the aim of the present study is to evaluate the use of phosphogypsum waste as a Ca source for carbon dioxide mineral sequestration. We believe that this new methodology is especially attractive and ecologically clean, since it has the potential to reduce two environmental problems simultaneously: (1) management of hazardous industrial waste; and (2) greenhouse gasses emissions. 2. Experimental procedure Several samples (approx. 2 kg) were collected from the phosphogypsum stack in November of 2009 at different depths from bore-holes carried out using a soil sampling auger. In the laboratory, samples were oven-dried (40 ◦ C) until complete dryness, ground and homogenised. The data relative to one representative sample of phosphogypsum are discussed in this paper. The proposed methodology starts by the dissolution of the raw sample in an alkaline solution. The dissolution experiments were conducted in 20 ml of high-purity water where 5 g of phosphogypsum was dispersed by magnetic stirring at room pressure and temperature. Immediately after, 2.34 g of NaOH was added to
3. Results and discussion 3.1. Analysis of the samples prior to CO2 sequestration The XRD of the raw phosphogypsum sample is shown in Fig. 2. The diffraction pattern indicates that the sample is composed mainly of gypsum (CaSO4 ·2H2 O, Powder Diffraction File, PDF number: 99-101-0394), as expected [15]. After the dispersion of this sample in water and the addition of NaOH, the resulting whitish precipitate correspond mostly to portlandite (Calcium hydroxide, Ca(OH)2 , PDF number: 99-100-0115), as
Phosphogypsum + Water
Magnetic stirring
NaOH
Mix
Magnetic stirring (3h)
Filter
Solid
CaCO3
Bubble CO2
Ca(OH)2
Liquid
Disolution
Fig. 1. Flowchart of the experimental methodology.
Evaporate
Na2SO4
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Table 2 Contents of trace elements measured by analysis with inductively coupled plasma mass-spectrometry.
CaSO4·2H2 O + 2NaOH ↔ Ca(OH)2 + Na2SO4 + 2H2O
Intensity (counts)
Phosphogypsum Trace elements (mg kg−1 ) Cr 3.70 As 3.58 3.48 U 2.49 Ni 2.10 V Se 1.97 Cd 1.34 1.20 Pb 0.92 Zn Th 0.69
Portlandite
Thenardite
Gypsum
5
10
15
20
25
30
35
40
45
50
55
60
65
Fig. 2. Diffractograms of the raw phosphogypsum (CaSO4 ·2H2 O) from Huelva (Spain), of the solid filtered where portlandite (Ca(OH)2 ) can be clearly observed, and the solid phase obtained from supernatant evaporation where the diffraction pattern of thenardite (Na2 SO4 ) is clearly matched. Diffractograms were shifted for of clarity.
shown in the diffractogram of Fig. 2 and as expected by reaction (␣): CaSO4 ·2H2 O + 2NaOH ↔ Ca(OH)2 + Na2 SO4 + 2H2 O
(␣)
and considering the large difference between gypsum and portlandite’s solubility constants, being the solubility constant of gypsum K = 3.14 × 10−5 and that of the portlandite K = 5.02 × 10−6 [16]. Finally, the diffractogram of the transparent salts precipitated by evaporation of the supernatant liquid (Fig. 2) indicates that is mostly composed of thenardite (Na2 SO4 , PDF number: 99100-4889), again as expected regarding reaction (␣). These results confirm the efficiency of the reaction (␣) as a procedure to deal the raw phosphogypsum waste and obtain a roughly pure portlandite. Major element concentrations determined by XRF of the different solid phases from phosphogypsum dissolution were in agreement with the mineralogical composition estimated by XRD (Table 1). Raw phosphogypsum composition is clearly dominated by S (50.2 wt% as SO3 ) and Ca (44.7 wt% as CaO). However, the calcium content of the gypsum was slightly higher than expected stoichiometrically, revealing the existence of other different calcium compounds poorly crystallized, not easily observable by XRD Table 1 Major element contents of the samples analyzed by X-ray fluorescence. Weight percentages are normalized to the mass without loss-of-ignition (LOI) values.
Major elements (%) 50.2 SO3 44.7 CaO 1.56 F Na2 O 1.16 0.72 Cl 0.67 P2 O5 SiO2 0.43 0.24 Al2 O3 0.14 MgO 0.07 Fe2 O3 SrO 0.07 Y2 O3 0.02 21.2 LOIb a b
n.d., not detected. LOI, loss-of-ignition.
Na2 SO4
Portlandite
Calcite
7.59 8.57 7.84 4.76 4.76 5.04 2.34 2.47 2.02 1.62
8.17 9.32 8.05 5.21 4.79 6.04 2.51 2.56 2.98 1.71
l.d., limit of detection.
2 (degrees)
Phosphogypsum
433
Portlandite
Na2 SO4
4.28 89.4 1.61 2.12 0.04 1.21 0.64 0.27 0.13 0.20 0.06 0.04 14.3
55.8 1.04 n.d.a 42.7 0.23 n.d. 0.16 n.d. n.d. n.d. 0.02 n.d. 27.9
(Fig. 2). Portlandite is manly composed of Ca (89 wt% as CaO); again, calcium content remains slightly higher than expected, so minor calcium compounds are carried in this portlandite phase as impurities, possibly calcium phosphate from original rock. Finally, the solid obtained from the supernatant evaporation is characterized by high contents of S (55.8 wt% as SO3 ) and Na (42.7 wt% as Na2 O). Other major impurities of the starting phosphogypsum are F, Na, Cl, P, Si, Al and Mg, and to a lesser extent Fe, Sr and Y. Excepting some of Cl, Si and Sr present in the sodium sulphate, the most of these impurities were also found in the resulting portlandite. The presence of potentially toxic trace elements in the raw phosphogypsum was analyzed by acid-digestion and ICP-MS (Table 2). The main trace elements identified were Cr, As, U, Ni, V, Se, Cd, Pb, Zn and Th, in order of abundance. The contents of these elements in the products from phosphogypsum dissolution, i.e. portlandite and sodium sulphate, are also shown in Table 2. All trace elements were present at ≤3.70 mg/kg within the phosphogypsum sample. These values are close to the average values analyzed in surface samples from the pile [6], supporting this sample as representative of the stack. The contents of trace elements are slightly inferior to those of phosphogypsum samples produced from other phosphate rock sources around the world [17]. Most of these trace elements were below or near to the detection limit in the precipitated sodium sulphate. Based on reaction (␣), a mass balance or transfer factor was calculated in order to determine the partitioning of trace elements during phosphogypsum dissolution and portlandite precipitation. Accordingly, approx. 100% for all trace elements are transferred from phosphogypsum to the resulting portlandite. Hence, no metals remained in the sodium sulphate solution resulting from the phosphogypsum dissolution. 3.2. Carbon dioxide capture efficiency The XRD diffractogram of the carbonated portlandite resulting from the aqueous carbonation procedure is shown in Fig. 3. It clearly indicates that total conversion of the original portlandite to calcite (CaCO3 , PDF number: 99-101-2108) was achieved. The carbonation reaction is as follow: Ca(OH)2 + CO2 ↔ CaCO3 + H2 O
()
This point was confirmed by thermogravimetric analyses of the original and calcite samples. The analysis of the reference portlandite prior to the carbonation procedure confirmed that the purity of the portlandite was close to 90%. The measured weight loss due to dehydration was 22%, as estimated by XRD (see Table 1). And the carbonated portlandite showed a 38.4% weight loss due to the release of the CO2 at 700 ◦ C (almost 40%, the maximum theoretical weight loss of pure calcite), indicating that all the calcium present in the sample, whether from portlandite or from other minor calcium
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Intensity(counts)
Ca(OH)2 + CO2 ↔ CaCO3 + H2O
Calcite
Portlandite
5
10
15
20
25
30
35
40
45
50
55
60
65
2 (degrees) Fig. 3. Diffractogram of the carbonated porlandite. No other phase than calcium carbonate is observed. Portlandite diffractogram has been repeated for comparison purposes only. Diffractograms were shifted for clarity.
compounds, was carbonated and revealing a extremely high carbonation efficiency of these wastes. During the aqueous carbonation process, toxic trace elements initially contained in the portlandite are released into solution. Based on reaction () stoichiometry and the molecular weights, the concentration of trace metals in the final calcite implies a high transfer factor of around 100% for all elements (Table 2). Again, no metals were kept in the resulting water after the carbonation reaction, and all of them were transferred to the solid precipitate. Metal uptake by calcite precipitation is responsible for trace element removal from solution through co-precipitation or adsorption on its surface. Divalent metals such as Ni, Cd, Pb and Zn can be easily removed from solution and incorporated into the structure of the calcite by co-precipitation [18]. Albeit, adsorption on the calcite surface could be also a significant mechanism for retention of metals as long as their concentrations in solution are lower than the limit of 10−5 M established by Zachara et al. [19], as is the case in our experiments of portlandite dissolution and carbonation. Nevertheless, whatever the scavenging process, calcite precipitation is a successful mechanism for reducing metal mobility in natural aqueous systems [20]. A sustainable and environmental-oriented procedure to manage of these wastes is inferred from these results. In a realistic and numerical approximation to quantities of phosphogypsum that are generated annually by the fertilizer industry of Huelva, the reagents to develop the proposed methodology are summarized as follows: to treat 2 Mt of phosphogypsum, 0.92 Mt of NaOH will be needed, and 0.4 Mt of H2 O from phosphogypsum, 1.64 Mt of Na2 SO4 and 0.84 Mt of Ca(OH)2 will be produced. This amount of Ca(OH)2 could be used to subsequently capture 0.50 Mt of CO2 , yielding 1.16 Mt of CaCO3 . Based on these results, an estimation of the carbon capture capacity of the entire stockpiled phosphogypsum waste was made. 55.2 Mt of NaOH will be needed to treat 120 Mt of phosphogypsum, enabling capture 30 Mt of CO2 . 4. Conclusions In this work, total conversion of phosphogypsum industrial waste into portlandite (calcium hydroxide) and sodium sulphate was confirmed, and the high carbon dioxide capture efficiency of the resulting portlandite was demonstrated. Portlandite was produced by dissolution of dihydrated calcium sulphate from the phosphogypsum and reaction with sodium hydroxide. The main
impurities contained in the phosphogypsum were transferred to the portlandite and, subsequently, to the final calcite after bubbling CO2 , with transfer factors close to 100%. The total and rapid conversion of portlandite into calcite demonstrated in this study makes the proposed methodology an attractive and ecological solution to two environmental problems: (1) the high amount of phosphogypsum waste generated by the fertilizer industry; and (2) the CO2 emissions generated by the same industry. Furthermore, mineral CO2 sequestration using phosphogypsum by-products from the fertilizer industry of Huelva (Spain) is especially attractive since the waste disposal stockpiles are located about 60 km from the Iberian Pyrite Belt, the largest sulphide metallogenic province in the world. The intense mining activity in this region has produced a huge volume of sulphide-rich mining wastes [21]. The oxidation of these mining-wastes releases solutions with abnormally high acidity and metal concentrations [22]. Some restoration strategies using calcite have been tested in the field with satisfactory results to deal this environmental problem [23]. However, the calcite used in the treatment systems is associated with a high economic and environmental cost, since calcite is typically a resource and not a residue. Using the final calcite from dissolution and carbonation of the phosphogypsum would significantly reduce the costs of future remediation plans. The toxic metals contained in the final calcite may be released into solution during the treatment of acid mine waters, but this additional amount of toxic elements is negligible compared to that released by sulphide-rich wastes in acid discharges and the precipitation of metal hydroxides also depletes these elements by sorption process. Finally, the purity of the sodium sulphate by-product allows it to be commercialized, for example to detergent or paper industries. In summary, our results support the development of an economically viable technology of carbon dioxide sequestration based on the reuse of these phosphogypsum industrial wastes. Acknowledgements The authors would like to acknowledge the research services of the ICMS (CSIC-US) and the CITIUS (US). The assistance of Dr. Jesús de la Rosa and Dra. Ana María Sanchez de la Campa for ICP-MS analyses is gratefully acknowledged. The authors are also grateful to the Consejería de Innovación Ciencia y Empresa of the Junta de Andalucía (Spain) for supporting this work with the annual Grant TEP115 and to the Ministerio de Ciencia e Innovación of the Spanish Government for the Grant PIA42008-31. V.M.F. thanks the CSIC for financial support through the JAE programme. References [1] P.M. Rutherford, M.J. Dudas, R.A. Samek, Environmental impacts of phosphogypsum, Sci. Total Environ. 99 (1994) 1–38. ˜ ˜ M.C. Fernández, S. Canete, M. Pérez, Radiological impacts of natural [2] C. Duenas, radioactivity from phosphogypsum piles in Huelva (Spain), Radiat. Meas. 45 (2010) 242–246. ˜ A. Martínez-Aguirre, M. García-León, U- and Th-isotopes in an estu[3] R. Periánez, arine system in Southwest Spain: tidal and seasonal variations, Appl. Radiat. Isotopes 47 (1996) 1121–1125. [4] J.P. Bolívar, R. García-Tenorio, J.L. Más, F. Vaca, Radioactive impact in sediments from an estuarine system affected by industrial waste releases, Environ. Int. 27 (2002) 639–645. ˜ F. Vaca, [5] M. Villa, F. Mosqueda, S. Hurtado, J. Mantero, G. Manjón, R. Perianez, R. García-Tenorio, Contamination and restoration of an estuary affected by phosphogypsum releases, Sci. Total Environ. 408 (2009) 69–77. [6] R. Pérez-López, J.M. Nieto, I. López-Coto, J.L. Aguado, J.P. Bolívar, M. Santisteban, Dynamics of contaminants in phosphogypsum of the fertilizer industry of Huelva (SW Spain): from phosphate rock ore to the environment, Appl. Geochem. 25 (2010) 705–715. ˜ A. Delgado, Effect of soil properties [7] R. Domínguez, M.C. Del Campillo, F. Pena, and reclamation practices on phosphorus dynamics in reclaimed calcareous marsh soils from the Guadalquivir Valley, SW Spain, Arid Land Res. Manage. 16 (2001) 203–221.
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