Ecotoxicology of polychlorinated biphenyls in fish--a critical review.

http://informahealthcare.com/txc ISSN: 1040-8444 (print), 1547-6898 (electronic) Crit Rev Toxicol, 2015; Early Online: 1–19 © 2015 Informa Healthcare USA, Inc. DOI: 10.3109/10408444.2015.1038498

REVIEW ARTICLE

Ecotoxicology of polychlorinated biphenyls in fish—a critical review T. B. Henry1,2,3

Critical Reviews in Toxicology Downloaded from informahealthcare.com by Fudan University on 05/11/15 For personal use only.

1School of Life Sciences, John Muir Building, Heriot-Watt University, Edinburgh, EH14 4AS, UK, 2The University of Tennessee Center for Environmental Biotechnology, 676 Dabney Hall, Knoxville, Tennessee 37996, USA and 3Department of Forestry, Wildlife and Fisheries, The University of Tennessee, 274 Ellington Plant Sciences Building, Knoxville, Tennessee, 37996, USA

Abstract

Keywords

Polychlorinated biphenyls (PCBs) are widespread persistent anthropogenic contaminants that can accumulate in tissues of fish. The toxicity of PCBs and their transformation products has been investigated for nearly 50 years, but there is a lack of consensus regarding the effects of these environmental contaminants on wild fish populations. The objective of this review is to critically examine these investigations and evaluate publicly available databases for evidence of effects of PCBs in wild fish. Biological activity of PCBs is limited to a small proportion of PCB congeners [e.g., dioxin-like PCBs (DL-PCBs)] and occurs at concentrations that are typically orders of magnitude higher than PCB levels detected in wild fish. Induction of biomarkers consistent with PCB exposure (e.g., induction of cytochrome P450 monooxygenase system) has been evaluated frequently and shown to be induced in fish from some environments, but there does not appear to be consistent reports of damage (i.e., biomarkers of effect) to biomolecules (i.e., oxidative injury) in these fish. Numerous investigations of endocrine system dysfunction or effects on other organ systems have been conducted in wild fish, but collectively there is no consistent evidence of PCB effects on these systems in wild fish. Early life stage toxicity of DL-PCBs does not appear to occur at concentrations reported in wild fish embryos, and results do not support an association between PCBs and decreased survival of early life stages of wild fish. Overall, there appears to be little evidence that PCBs have had any widespread effect on the health or survival of wild fish.

ecotoxicology, endocrine disruption, persistent organic pollutants, toxicology, estradiol, wild fish

Table of Contents Introduction ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... 1 Physicochemical properties of PCB congeners and consequences on environmental fate ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... 2 Concentrations of PCBs in the abiotic environment ... ... ... ... ... ... ... 3 Concentrations of PCBs in wild fish ... ... ... ... ... ... ... ... ... ... ... ... ... 4 Environmental relevance of laboratory studies to predict PCB toxicity in wild fish.. ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... 7 Induction of oxidative stress ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... 8 Effects of PCBs on reproduction in wild fish ... ... ... ... ... ... ... ... ... ...10 Developmental toxicity .. ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...12 Changes in population genetics linked to toxicant exposure in wild fish .. ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...13 Investigations of PCB-induced organ system dysfunction in wild fish 13 Conclusions ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...14 Acknowledgments . ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...14 Declaration of interest ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...14 References . ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ... ...14

Introduction Polychlorinated biphenyls (PCBs) are a group of substances containing two linked benzene rings (biphenyl) in which the remaining five carbons in each ring can be bonded to a Address for correspondence: T. B. Henry, Department of Forestry, Wildlife and Fisheries, The University of Tennessee, 274 Ellington Plant Sciences Building, Knoxville, Tennessee, 37996, USA. Tel: ⫹ 44 (0)131 451 4315. Fax: ⫹ 44 (0)131 451. E-mail: [email protected]

History Received 2 October 2014 Revised 24 March 2015 Accepted 3 April 2015

chlorine molecule. Individual PCB congeners are defined by the configuration of chlorine molecules on the biphenyl, and 209 congeners are possible (Hawker and Connell 1988). PCBs have useful physicochemistry including good chemical stability and high insulating and dielectric properties, and these characteristics led to mass production of PCBs with widespread use in electrical transformers, solvents, and flame retardants (Kutz et al. 1991, National Academy of Sciences 1979). An unexpected consequence of the unique physicochemical properties of PCBs is high environmental persistence and a tendency for some congeners to accumulate in organisms, two characteristics that are particularly important when chemicals are recognized to be of concern as environmental toxicants (UNEP 2012). Because of these characteristics and their suspected toxicity, production of PCBs ended in most countries by the mid-1980s (Kutz et al. 1991). Quantities of PCBs remain in the abiotic and biotic compartments of many aquatic environments throughout the world and remain a concern for human and ecosystem health (UNEP 2012). The presence of PCBs in aquatic environments is a legacy of industrial chemical production and management, and their presence will continue until the slow transformation of these compounds and dilution will reduce amounts to minimal levels of detection. During the long life span of PCBs in the environment, some environmental compartments (e.g., fish tissues)


2

T. B. Henry

can become contaminated at concentrations many times those of the abiotic environment (e.g., sediments) where the bulk of these chemicals reside. The bioaccumulation of PCBs in fish has been particularly well documented and numerous studies have reported elevated concentrations of PCBs in wild fish from various aquatic environments [(freshwater, estuarine, and marine) (e.g., Besselink et al. 1996, Barron et al. 2000, Hinck et al. 2009a, b, c, Bettinetti et al. 2012)]. One consequence of elevated PCBs in fish has been posting of fish consumption advisories to reduce human exposure to these persistent organic chemicals (US EPA 2004, Imm et al. 2005). The issue of PCB-tainted fish and potential human exposure has elevated concern regarding the toxicity of PCBs in humans (Andersson et al. 1998, McElroy et al. 2004), and also regarding the possibility for PCBs to influence fish physiology and survival negatively (Monosson et al. 1994, van der Oost et al. 2003). Linking environmentally relevant cause and effect in wild fish is challenging. Aside from acute toxicity (lethality) following sudden events (e.g., chemical spills), there are few examples in the literature that conclusively link exposure to a specific toxicant with population level effects in wild fish. The toxicological effects of contaminants that bioaccumulate from chronic exposures in the environment are especially difficult to resolve within the background of numerous other environmental stressors that influence fish physiology (Schwarzenbach et al. 2006). The few anthropogenic toxicants for which chronic exposure have been convincingly linked to toxicological effects in fish at population levels include acidification from atmospheric deposition (Larssen et al. 2011), dioxin effects in early life history stages of lake trout in Lake Ontario (Cook et al. 2003), and perhaps deliberate whole-lake exposure to 17α-ethinylestradiol to investigate chronic effects of endocrine disruption in fathead minnow (Kidd et al. 2007). The presence of persistent anthropogenic chemicals in aquatic environments suggests concern for toxicity, but attributing observation of toxic effects in wild fish to specific substances requires more evidence than just the detection of these substances in fish tissues. Within the context of the inherent challenges of establishing causation between PCB accumulation and toxic effects in wild fish, numerous ecotoxicologic studies have been conducted. Investigations have frequently involved collection of fish from locations with PCB contamination, evaluation of the PCB concentrations within tissues, and assessment of one or more biomarkers of toxic effects. These numerous investigations have employed different methods, evaluated different species and effect endpoints, considered different types of aquatic habitats, and drawn different conclusions regarding PCBinduced effects in wild fish. The objective of this review is to examine the literature on the ecotoxicity of PCBs and their transformation products in wild fish, and to critically evaluate the evidence that attributes fish toxicity to PCB exposure. First, the present state of PCB contamination in aquatic environments will be briefly reviewed to establish the nature of exposure, the level of contamination of PCBs in wild fish, and the distribution of PCBs among tissues. The reports of lesions in wild fish that have been observed in association with PCB accumulation will be considered to determine if any consistent relations exist between PCB exposure and toxicity in wild fish. The Web of Science (WOS) was the primary search engine

Crit Rev Toxicol, 2015; Early Online: 1–19

used to identify the evaluated peer-reviewed articles, and the search strategy considered dates from 1965 to 2015. A tiered approach based on a WOS search with “PCB (or polychlorinated biphenyl)” AND “Fish” generated over 14,500 articles and this list was refined by research area (Environmental Sciences AND Toxicology) generating a list of over 3,000 articles with dates from 1971 to 2015. Initial examination of these articles was conducted to determine if the topic of the article was relevant to this review and also to define the selection of the subtopics (i.e., subheadings of this article), and this generated a list of over 500 articles. Evaluation of these articles and consideration of linked citations (i.e., references cited and citing articles) led to the selection of ∼200 articles included in this review.

Physicochemical properties of PCB congeners and consequences on environmental fate The concentration of individual PCB congeners in an environmental sample is the outcome of anthropogenic, physicochemical, abiotic, and biotic processes. Industrial production of PCBs was based on chlorination of biphenyl (e.g., Aroclors, Kutz et al. 1991), a process that led to complex mixtures containing various amounts of the 209 possible congeners that now contaminate environmental compartments (Connolly et al. 2000). The location and number of chlorine atoms on the biphenyl influence water solubility, octanol:water partition coefficients, and fugacity among other properties (Table 1). A complete PCB congener (or homolog group) concentration profile can be used (e.g., Figure 1) to compare differences in relative concentrations of individual congeners among samples, over time, or among environmental compartments (e.g., Howell et al. 2008). When evaluating toxicity to biota, it is important to recognize that PCB congeners have different toxicological properties, and the concentration of a specific congener makes up only a small fraction of the total PCBs present in an environmental sample [e.g., PCB-126 ∼ 0.014% of total PCBs (Figure 2), a proportion consistent with that (0.015–0.036%) reported by Bhavsar et al. (2007)]. Table 1. Properties of selected representative PCB congeners (from Hawker and Connell 1988), and TEFs (from van den Berg et al. 1998). DL PCBs 3,4,4′5-TCB (PCB-81) 3,3′,4,4′-TCB (PCB-77) 3,3′,4,4′,5- (PCB-126) 3,3′4,4′5,5′-HxCB (PCB-169) Mono-ortho PCBs 2,3,3′,4,4′- (PCB-105) 2,3,4,4′,5- (PCB-114) 2,3′,4,4′,5- (PCB-118) 2,3,4,4’,5- (PCB-123) 2,3,3′,4,4′,5- (PCB-156) 2,3,3′,4,4′,5′- (PCB-157) 2,3,3′,4,4′,6- (PCB-158) 2,3′,4,4′,5,5′- (PCB-167) 2,3,3′,4,4′,5,5′- (PCB-189)

Log10 Kow 6.36 6.36 6.89 7.42

TEF 0.0005 0.0001 0.005 0.00005

6.65 6.65 6.74 6.74 7.18 7.18 7.02 7.27 7.71

⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005 ⬍ 0.000005

TEF is based on the recognition that there is a common mechanism of action involving binding to the AhR, and that the toxicity of substances assigned a TEF value is relative to toxicity of 2,3,7,8tetrachlorodibenzo-p-dioxin, which is assigned a TEF value of 1 (van den Berg 1998).

PCBs in wild fish

DOI 10.3109/10408444.2015.1038498

3

Concentration (µg/g)

0.06 0.05 0.04 0.03 0.02 0.01 0 1

9

17

25

33

41

49

57

65

73

81

89

97 105 113 121 129 137 145 153 161 169 177 185 193 201 209

Figure 1. Concentration profile of individual PCB congeners in fillets of largemouth bass, Micropterus salmoides, and smallmouth bass, M. dolomeiu, collected from the Thompson Island Pool (Hudson River, New York, USA) during May–June (2004–2007). Fish had a total length of 400–450 mm (males and females) and total PCB concentration was 2.44 ⫾ 0.35 μg/g (mean ⫾ standard error, N ⫽ 55). The concentration of individual congeners was estimated for congeners that co-eluted within the same peak of the GC by dividing the concentration determined for the PCBs within a GC peak by the number of congeners present within the peak. For example, PCB-126 and -129 co-elute, and the concentration for each is represented by dividing the concentration represented by the peak by two. Fish PCB concentrations were obtained from Hudson River analytical results that were provided to US EPA March 30, 2007, and March 31, 2008 (US EPA 2007, 2008).

Physicochemical properties influence PCB transport and environmental fate and also the abiotic and biotic transformation of PCBs. Transformation of PCBs has been reported in numerous studies that have analyzed PCB concentrations in sediments (e.g., Brown and Wagner 1990, Lake et al. 1992, Imamoglu et al. 2002), and the influence of various environmental factors (e.g., pH, oxidation/reduction status, organic carbon content, etc. of sediments) have been thoroughly investigated (e.g., Wu et al. 1996, Kim and Rhee 1997, Chang et al. 2001, Furukawa and Fujihara 2008, Reiner 2010, Ho and Liu 2011). The outcome of microbial processes includes de-chlorination of the biphenyl (e.g., Chang et al. 2001), which appears to occur most readily for PCBs with 3 or fewer chlorines (aerobically) and more slowly for more highly chlorinated PCBs (aerobic/anaerobic conditions required) (Vasilyeva and Strijakova 2007, Hughes et al. 2010, Quensen et al. 1998, Abramowicz 1990). The long-term outcome of microbial action could be transformation and the eventual removal of PCBs [e.g., modeled half-lives in sediments: 3 years, PCB 28; 38 years, PCB 180 (Sinkkonen and Paasivirta 2000)]. Chlorine atoms are removed from biphenyls and transformation of higher chlorinated PCBs leads to formation of PCBs with fewer chlorine atoms (Wiegel and Wu 2000). Hydroxylated PCBs (OH-PCBs) can

0.06 Concentration (µg/g)


Congener number

Dioxin-like PCBs

be formed when PCBs are metabolized in vertebrates (see below) and OH-PCBs have been detected in the environment [e.g., surface water maximum total OH-PCB, 130 pg/L; aqueous particulate matter maximum OH-PCB of 990 pg/g particulate matter (Ueno et al. 2007)]. In aquatic environments, OH-PCBs undergo microbial transformation generating other substances, although the biochemical pathways of these processes are not well understood (Wiegel and Wu 2000). Mixtures of PCB congeners, PCB metabolites, and PCB transformation products constitute the PCB contaminant profile in the abiotic/biotic environment (Haranczyk et al. 2010) to which wild fish are exposed.

Concentrations of PCBs in the abiotic environment The persistence and transport of PCBs within the environment have led to global distribution of these substances and their detection is nearly ubiquitous. Historical overall production of PCBs is estimated at millions of metric tons (e.g., as estimated by Breivik et al. 2002a,b) and the scale of the management of this global environmental issue impacts most nations (UNEP 2012). With the ban on further production of PCBs in most industrialized countries in place by the 1980s (Kutz et al. 1991), overall reductions in environmental PCB

Mono-ortho PCBs

0.05 0.04 0.03 0.02 0.01 0 77

81

126

169

105 114 118 123 Congener number

156

157

167

189

Figure 2. Mean (⫾ coefficient of variance) concentration of DL PCBs and mono-ortho PCBs from Micropterus sp. collected from the Thompson Island Pool (Hudson River, New York, USA) during May–June (2004–2007), presented in Figure 1. The concentration of PCB-126 was 0.00035 μg/g, which was ∼0.014% of total PCB concentration.


4

T. B. Henry


concentrations have been indicated (U.S. EPA 1999), and this observation has been documented in numerous studies that have monitored PCB levels over time (e.g., Marvin et al. 2004, Howell et al. 2008, de Boer et al. 2010). There are examples of PCB levels that have increased due to redistribution of PCBs in the environment [e.g., re-suspension of contaminated sediments or changes in community structure (e.g., invasion by zebra mussels) (French et al. 2011)]; however, as these localized conditions stabilize, PCB concentrations at these sites are also expected to decrease with time. Interpretation of PCB concentrations among studies that have used different approaches for analysis can be difficult because of limitations in comparability between analytical methods. Differences in methods used to measure PCB concentrations reflect evolution of analytical methods over time, differences in available instrumentation, cost of analyses, and different objectives for undertaking the analysis (e.g., screening, targeting specific congeners, and low detection limits). Since PCB concentrations in environmental matrices can be extremely low [e.g., low pg/L in seawater samples (Schulzbull et al. 1995)], various methods for extracting and concentrating PCBs from matrices have been developed (methods reviewed in Muir and Sverko 2006). Following extraction, methods are applied to “clean up” the extracts to enhance separation of PCBs from co-extracting substances that are also present in samples. Further separation of PCBs from remaining coextracting substances (e.g., lipids from tissue samples) can be completed by size-exclusion gel permeation chromatography or high-performance liquid chromatography. The ultimate outputs of these procedures are the PCBs obtained from the original sample dissolved in an organic solvent, which can be evaporated off several times to concentrate PCBs for subsequent analysis. The preferred method for PCB analysis for regulatory and litigation purposes is gas chromatography (GC)–high-resolution mass spectrometry (HRMS), which is recognized to provide higher selectivity and sensitivity compared with other methods, although GC–HRMS does require a longer sample processing time (Reiner 2010). Other GC-based methods for analysis of PCBs have been developed, and the application, capabilities, and limitations of these techniques have been extensively reviewed (e.g., Muir and Sverko 2006, Reiner et al. 2006, Reiner 2010). Comparability among analytical methods and results is highly important and required for international implementation of the Stockholm Convention on Persistent Organic Pollutants (UNEP 2012). In addressing issues in fish ecotoxicology, a variety of these methods for PCB quantification have been applied (Table 2). The analytical technique used to determine PCB concentrations, type of standards used for quantification, and presentation of results influence the ability to make comparisons among studies that have investigated ecotoxicity of PCBs in fish. For

each recognized analytical procedure [e.g., U.S. Environmental Protection Agency (EPA) test methods including those indicated in Table 2] that has been appropriately conducted (internal controls, standard reference materials, detection limits, etc.), the results are suitable for interpretation within the context of the test method, but limitations exist when making comparisons to results obtained by different techniques (e.g., Martin et al. 2003). Many studies that report total PCB concentrations (e.g., Baldigo et al. 2006, Hinck et al. 2007a, b, 2008, 2009a, b) have used various methods including the sum of specific Aroclors (e.g., Baldigo et al. 2006), computation based on correlations between representative congener concentrations and total PCBs (e.g., National Oceanic and Atmospheric Administration (NOAA) EPA method, NOAA National Seabird Program method, or Babut et al. 2009), or the sum of concentrations of each PCB congener assessed individually (e.g., Howell et al. 2008). As these methods for determination of total PCB concentrations are based on different assumptions (e.g., relative congener profile in environmental sample is the same as that of a specific Aroclor), each procedure can lead to different total PCB concentrations if they are applied to analyze the same sample (Martin et al. 2003, Bhavsar et al. 2007). Howell et al. (2008) compared different methods for determining total PCB concentrations, and indicated that the limitations of incomplete congener analyses can lead to incorrect conclusions, an aspect that has not been adequately considered in the literature (Bhavsar et al. 2007). Total PCBs implies the sum of the concentrations of all the individual PCB congeners in a sample, but the analytical approach used to obtain this value must be considered when comparing study results (Martin et al. 2003).

Concentrations of PCBs in wild fish The concentrations of PCBs in wild fish are a consequence of congener physicochemistry, biotic and abiotic factors that influence exposure, and physiological processes of fish that effect absorption, distribution, metabolism, and excretion of these substances. Perhaps the most extensive and comparable dataset on the concentrations of total PCBs in fish is that of the United States Geological Service (USGS)’s Biomonitoring of the Environmental Status and Trends (BEST) program, which evaluated fish health and contaminant levels in large rivers systems of the U.S. from 1995 to 2002. Results of the USGS BEST program are publicly available (http://www. cerc.usgs.gov/data/best/search/index.htm), and indicate total PCB levels in fish over a large geographical scale (Figure 3). Numerous separate studies have evaluated fish collected from environments known to have especially elevated PCBs concentrations; among these, total PCBs levels as high as 1200 μg of total PCBs/g (lipid weight of fillet) have been reported

Table 2. Examples of concentrations of total PCBs in sediments from various locations and concentrations of individual PCB congeners and the analytical technique used for quantification. Location Rhone River (France) Hudson River (NY, USA) Houston Ship Channel (TX, USA)

Analysis target Σ209 PCB congeners ΣAroclors 1242, 1254, 1260 Σ209 PCB congeners

Concentration (ng/g dry weight) 0.25–131.5 7.5–2480 4.18–4601

Analytical technique GC–HRMS GC–MS-ECD GC–HRMS

See Babut et al. (2009) for reports of sediment PCB concentrations in other locations reported by other studies.

Method (US EPA) 1668 1668A

Reference Babut et al. (2009) Baldigo et al. (2006) Howell et al. (2008)

PCBs in wild fish

DOI 10.3109/10408444.2015.1038498

5

Total PCB concentration (µg/g)

5 4

Mississippi

RioGrande

Southeastern

Yukon

Colorado

Columbia

3 2 1 0 0

20

40

60

80

100

120

140

160

180


Sample number (in both rank order and count)

Figure 3. The PCB concentration (total PCBs) in whole fish composite samples (wet weight) from all fish collected and analyzed within the USGS BEST program. The USGS BEST program investigated fish health and contaminant levels in large river systems from 1995 to 2002: Mississippi River Basin, Columbia River Basin, Rio Grande Basin, Yukon River Basin, and Southeast Rivers. Specifics of the sampling sites and study details are available on the website indicated above. In this figure, all samples analyzed in each river basin are assigned numbers in increasing order of concentration (rank order identification) numbers. The total number of samples indicates the level of effort for each basin. In the Rio Grande Basin, total PCBs were below the reported detection level (⬍ 0.0500 μg/g) in all samples. In the Yukon River Basin, total PCBs in all (31) reported samples were below 0.09 μg/g.

(Table 3). Detection of individual congeners in fish from PCBcontaminated sites have been reported with concentrations as high as 2.9 ng/g (liver wet weight) for PCB-126 (Yuan et al. 2001). Concentrations of PCBs in the abiotic environment (e.g., sediments) are frequently correlated with fish tissue PCB burdens (e.g., Baldigo et al. 2006, Meador et al. 2010), and reductions in tissue PCB levels in wild fish have been documented in areas where abiotic PCB concentrations have decreased over time (e.g., El-Shaarawi et al. 2011, Fu and Wu 2005, 2006, Sadraddini et al. 2011a, b). However, the observed reductions in tissue PCB levels over time in some Canadian Great Lakes fish may not continue at the same rate and low

concentrations of PCBs may persist in fish for long periods of time (Bhavsar et al. 2007). PCB congener concentration profiles within fish tissues (e.g., Figure 1) have been related to congener concentration profiles of the environment in which the fish inhabit (e.g., Zlokovitz and Secor 1999, Ashley et al. 2000, Fu and Wu 2005, 2006), although independent variables such as fish species, age, lipid content, length, sex, weight, growth, habitat use, migration, and season can confound this relation (Ashley et al. 2003, Lopes et al. 2011). There is strong evidence that fish accumulate some PCB congeners more rapidly than others, and likewise some congeners are more rapidly excreted (Brown 1992, Stapleton et al. 2001,

Table 3. PCB concentrations reported in selected species of fish. Species Largemouth bass Smallmouth bass Common carp Bullhead Creek chub Longear sunfish Atlantic killifish Atlantic killifish Atlantic killifish Chinook salmon (eggs) Chinook salmon (tissue) Atlantic Tomcod (2 years) Common carp Common carp Gudgeon Mullet Bass Roach White croaker Halibut Shiner surfperch Sacramento Splittail aTotal

Location Hudson R Hudson R Hudson R Hudson R Logan, KY Logan, KY New Bedford H New Bedford H New Bedford H Lake Michigan Lake Michigan St Lawrence estuary Kalamazoo R Kalamazoo R Rhone R Er-Jen estuary Housatonic R Seine R San Fran Bay San Fran Bay San Fran Bay Sac-San Joaquin R

PCB (μg/g) a1,200total PCB a315total PCB a350total PCB a400total PCB b13.4Total PCB b5.1total PCB c0.534PCB126

d34.7non⫹ mono ortho c0.323PCB126 e3.8total PCB e1.5total PCB f1.1total PCB f15.4total PCB f10total PCB g1.012

PCBs

a60total PCB b556total PCB h16total PCB b0.34total PCB b0.055total PCB b0.16total PCB b0.16total PCB

PCBs relative to lipid weight of fillets. PCBs relative to wet weight of edible portion. cPCBs relative to dry weight of liver. dSum of non-ortho and mono-ortho PCBs relative to dry weight of liver. eTotal PCBs relative to wet weight. fTotal PCBs relative to wet weight of liver. gSum of 12 PCB congeners relative to muscle lipid. hTotal PCBs relative to dry weight of fillet. bTotal

Author (year) Baldigo et al. (2006) Baldigo et al. (2006) Baldigo et al. (2006) Baldigo et al. (2006) Brammell et al., 2004 Brammell et al. (2004) Nacci et al. (2009) Black et al. (1998) Black et al. (1998a) Ankley et al. (1991) Ankley et al. (1991) Couillard et al. (2004) Fisher et al. (2006) Fisher et al. (2008) Flammarion et al. (1999) Fu and Wu (2005) Reiser et al. (2004) Chevreuil et al. (1995) Fairey et al. (1997) Fairey et al. (1997) Fairey et al. (1997) Greenfield et al. (2008)


6

T. B. Henry

Fernandez et al. 2004, Antunes et al. 2007a, b). The PCB congeners found most frequently in fillets (excluding skin, kidney, and fins) of wild and farmed salmon collected from four regions of coastal British Columbia were congeners 77, 105, 114, 118, 123, 156, 157, 167, and 189 [based on those congeners that have a toxicity equivalent factor (TEF, see Table 1) (Yunker et al. 2011)]. Dietary exposure is considered most environmentally relevant in fish (e.g., Daouk et al. 2011), and absorption appears consistent with PCB congener physicochemistry (e.g., hydrophobicity) with no evidence of active or facilitated transport of PCBs across epithelial membranes (Buckman et al. 2006). Laboratory experiments with fish fed mixtures of PCB congeners demonstrated that the PCB half-life (time for concentration to be reduced by half after cessation of exposure) is a curvilinear response related to congener log10Kow value, and congeners with log10Kow values between 6.5 and 7.0 (Table 1) had the longest half-lives (Buckman et al. 2006). The observation of relations between PCB bioaccumulation, trophic magnification, and PCB log10Kow values has also been supported in wild fish collected from various aquatic ecosystems (Walters et al. 2011). In yellow eels, Anguilla anguilla, collected from PCB-contaminated locations and transferred to waters with low PCB levels, half-lives for some congeners was years; for highly chlorinated PCBs, no elimination was reported at the end of the eight-year study (de Boer et al. 2010). Once PCBs are absorbed into fish, concentrations in tissues are dependent on continued PCB exposure, metabolism and excretion of PCBs and metabolites, and growth of the fish [i.e., as fish grow PCB concentration can decrease (Rypel and Bayne 2010)]. Metabolism of PCBs occurs by enzymatic processes, and the most-studied metabolites are hydroxylated derivatives [OH-PCBs (Buckman et al. 2006, 2007a)] and methyl sulfone PCBs [CH3SO2-PCBs (Maervoet et al. 2004)], which are also of toxicological concern (Gerpe et al. 2000, Carlson and Williams 2001). Generally, the enzyme systems that biotransform PCBs into metabolites are considered to have limited activity in fish (Boon et al. 1989, Buckman et al. 2006); however, OH-PCBs have been detected in tissues of a variety of fish species (Campbell et al. 2003, Li et al. 2003), and the ability of fish to metabolize PCBs into OH-PCBs has been demonstrated experimentally in fish fed PCB mixtures in the laboratory (Buckman et al. 2006). Conversely, the metabolism of PCBs into CH3SO2PCBs, which occurs in some vertebrate species (e.g., birds and mammals), has only been reported in deepwater sculpin, Myoxocephalus thompsoni, and investigators speculated that this metabolite was generated by a novel biochemical pathway (Stapleton et al. 2001). While fish appear to have at least some capability to metabolize PCBs, and can be exposed to PCB metabolites that are present in the environment [e.g., OH-PCBs generated by metabolic processes of other organisms including humans and released in sewage effluents (Ueno et al. 2007)], the concentrations of PCB metabolites in wild fish are typically extremely low (e.g., ⬍ 175 pg/g plasma for 16 OH-PCBs detected in Great Lakes salmonids; Campbell et al. 2003). Nonetheless, PCB congener profile signatures in fish (e.g., Figure 1) at any point in time are driven in part by the metabolic processes of the fish that


generate OH-PCBs (and perhaps CH3SO2-PCBs) and the relative excretion rates for these metabolites, and this complex exposure scenario must be interpreted to understand any relation between exposure and toxic effects. Toxicity of substances is dependent on the presence and interaction of substances at target sites, and factors that influence the distribution and accumulation of substances within tissues will affect toxicological outcome. Toxicity is the degree to which a substance can damage an organism, and toxicant damage to or interference with biological molecules can manifest consequences at tissue, whole organism, and population levels of biological organization (Di Giulio and Hinton 2008). Because of PCB physicochemistry, tissues with higher lipid levels are the tissues where PCBs accumulate most in fish (Henshel et al. 2006, Antunes et al. 2007a, b); for this reason, PCB concentrations in tissues are frequently reported in terms of lipid concentration. In fish, lipid levels within tissues can vary considerably among species, sexes, and seasons (Rypel and Bayne 2010); and unequal mobilization or deposition of lipid stores among tissues within an individual fish can redistribute and influence tissue PCB concentrations (Bodiguel et al. 2009). Further, if fish condition changes over time (e.g., as a consequence to alterations in habitat), alterations in tissue lipid levels [(or lipid composition (van der Heijden and Jonker 2011)] can lead to changes in PCB tissue concentrations that are not a consequence of differences in PCB exposure (Rypel and Bayne 2010). For example, in adult European hake, Merluccius merluccius, concentrations of lipids (% of tissue) were 5–7% in muscle and ∼70% in liver and concentrations of PCBs in muscle were ∼6% of the concentrations reported in liver based on dry tissue weight [Bodiguel et al. (2009); (Figure 4)]. Fish species differ in relative lipid content among tissue types and this may contribute to differences in relative levels of PCBs among tissues in different species. Redistribution of PCBs from tissues (e.g., skeletal muscle) and deposition in oocytes during oogenesis were attributed to reductions in PCB body burdens in female after spawning compared with male fish (Bodiguel et al. 2009). Similar relative tissue distributions of PCBs have been reported for adult Atlantic cod, Gadus morhua, with largest PCB concentrations in liver followed by skeletal muscle and gonad (Dabrowska et al. 2009). Since distribution of lipids varies among tissues and lipids can be stored or metabolized in different amounts according to fish life history stage or season, redistribution of PCBs can occur within fish tissues. For example, utilization of lipids for metabolism during winter fasting and emaciation in arctic char, Salvelinus alpinus, was reported to induce mechanisms involved in PCB metabolism (e.g., cytochrome P450 enzyme systems, discussed below) and alteration of PCB dose–response relations (Jorgensen et al. 2002, 2006). PCB persistence and bioaccumulation [and perhaps biomagnification of some PCB congeners (Bettinetti et al. 2012)] in fish have been exploited in numerous studies to investigate and model trophic dynamics in aquatic ecosystems (e.g., Campfens and Mackay 1997). The issue of PCB bioaccumulation and biomagnification has been critically examined and reviewed in detail (e.g., Borga et al. 2005a, b, c, Walters et al. 2011), and the movement of PCBs within biota can provide information on food web dynamics (e.g., De Laender et al.

PCBs in wild fish

DOI 10.3109/10408444.2015.1038498

7

1

Concentration (µg/g)

0.8

0.6

0.4

0.2

0


110

149

118

153

132

105

138

187

128

156

180

170

194

Congener number

Figure 4. Concentration (μg/g dry tissue weight) of individual PCB congeners in liver (open bars) and muscle (solid bars) in European hake, M. merluccius, collected from the NW Mediterranean Sea and reported in Bodiguel et al., (2009). Data are means (⫹/⫺ standard deviation) of values reported in Bodiguel et al., (2009) for adult male and female fish. Congener concentration profiles were similar among tissues; however, lipid concentrations were considerably higher in liver (∼70% lipid) than in muscle (∼6% lipid). For each congener, the concentration in muscle was ∼6% of the concentration in the liver.

2010); however, the presence of PCBs in tissue was assessed as an indication of trophic level rather than as a contaminant of toxicological concern in these situations (e.g., Borga et al. 2005c). Perhaps if the PCBs were negatively affecting fish physiology then some of the conclusions on the food web dynamics obtained from these studies would not be valid (e.g., Sobek et al. 2010). Based on a large dataset (N ⫽ 984 fish), correlations between PCB concentrations (0.01–33.4 mg/kg lipid corrected) and growth metrics in six species of wild fish in riverine and lacustrine environments in Alabama (USA) supported differences among ecosystems related to diversity and abundance of prey items and no toxicological effects were reported (Rypel and Bayne 2010). Evaluation of the changes in PCB accumulation patterns in biota might provide information on alterations to food webs and ecosystems that could occur in response to changes in global climate (Borga et al. 2005c), and the behavior of PCBs in organisms provides an additional tool for long-term comparisons of food web dynamics and ecosystem processes. It is also possible (and perhaps likely) that fish with elevated PCB tissue levels live in habitats where other anthropogenic substances [e.g., polyaromatic hydrocarbons (PAHs)] are present, and the detection of PCBs in fish could/should initiate investigations of the presence and toxicity of these other substances at these locations.

Environmental relevance of laboratory studies to predict PCB toxicity in wild fish Laboratory exposures of fish to toxicants can be useful to investigate mechanisms of action and to reduce the number of variables that confound understanding effects of toxicant exposure, but results of laboratory experiments must be applied carefully to interpret toxicity in wild fish. Numerous laboratory experiments have exposed fish to PCBs and these studies have reported effects (Table 4) including induction of oxidative stress (Duffy et al. 2002, Grinwis et al. 2000), changes in reproductive system physiology (e.g., Monosson et al. 1994, Olsson et al. 1999, Daouk et al. 2011), thyroid axis dysfunction (Besselink et al. 1996, Adams et al. 2000, Buckman et al. 2007b), compromised immune system function (Iwanowicz et al. 2005, 2009, Grinwis et al. 2001), and histological lesions in tissues (Grinwis et al. 2001, Olsson et al. 1999, Daouk et al. 2011). Laboratory studies (e.g., zebrafish—Danio rerio) have also reported PCBassociated changes in biochemical processes including alterations in global gene expression pathways (e.g., Williams et al. 2008) and in the proteome (e.g., Sanchez et al. 2009, Berg et al. 2011). However, while results within the context of these studies contribute to understanding PCB toxicity in general, there are important limitations when results of laboratory studies

Table 4. Representative studies that have investigated toxicity of PCBs in fish exposed in the laboratory and the associated endpoints that have been assessed. Species Oryzias latipes Platichthys flesus Danio rerio Danio rerio Oryzias latipes Morone americana

PCB congener or mixture PCB-126 PCB-126 PCB-126 PCB-126 PCBs TCB

Platichthys flesus Hippoglossoides platessoides

Clophen A50 PCB-77, PCB-126

0.01–1.0 μg/g 0.5–50 mg/kg bw 7.5 μg/L 100nM 1–125 μg/g 0.2–5 mg/kg

Concentration

Exposure route I.p. injection Diet (8 weeks) Aqueous Aqueous Dietary I.p. injection

20–500 mg/kg 5–500 ng/g

I.p. injection I.p. injection

Reported endpoint observations Oxidative stress, immunosuppression Oxidative stress, decrease in thymus size Abnormal embryonic development Abnormal embryo development/mortality Abnormal swimming behavior Impaired female maturation, decreased larval survival Oxidative stress, thyroid hormones Increased deiodination activity of thyroid hormones in some tissues

Author Duffy et al. (2002) Grinwis et al (2000) Grimes et al. (2008) Na et al. (2009) Nakayama et al. (2004) Monosson et al. (1994) Besselink et al. (1996) Adams et al. (2000)


8

T. B. Henry

are applied to explaining potential effects of PCBs in wild fish (Reiser et al. 2004, Letcher et al. 2010, Deshpande et al. 2013). Laboratory exposure of fish to PCBs must be able to be related to environmentally relevant PCB exposure levels, exposure routes, and exposure time courses. Relations between PCB exposure and toxic effects in fish (i.e., dose–response curves) are not established. In the context of the environmental relevance of laboratory tests that investigate PCB toxicity in fish, exposure scenarios (e.g., intraperitoneal (i.p.) injection of PCBs or bolus ingestion) that deliver a sudden quantity of PCBs that do not reproduce natural exposure conditions (e.g., Iwanowicz et al. 2009) must be interpreted cautiously to be useful in understanding the toxicity of similar amounts of PCBs that accumulate in wild fish over longer periods of time. Further, when PCBs are added to water and effects on fish early life history stages are investigated (e.g., Li et al. 2014), dose–effect relations are almost impossible to determine because PCBs rapidly partition out of the aqueous phase and the dose to which fish are actually exposed is unknown [an exception is the study of Chambers et al. (2012), in which the PCB concentrations in exposed fish embryos were carefully determined using radiolabeled PCBs]. Investigations of the effects of PCBs in wild fish frequently cite laboratory fish studies to support field observations; however, the limitations of laboratory studies may not be adequately acknowledged (Barnthouse et al. 2003).

Induction of oxidative stress Cellular metabolism of some endogenous and exogenous substances including PCBs occurs by enzyme-mediated biochemical reactions that tend to generate substances of higher water solubility that are easier to excrete (Tillitt et al. 2008). Many organic substances including some PCBs can bind to receptors within cells [e.g., aryl hydrocarbon receptor (AhR)], and the receptor–ligand complex initiates transcription of genes that code for the enzymes [e.g., phase I metabolism, cytochrome P450 monooxygenases (e.g., CYP1A)] that facilitate initial stages of metabolism (i.e., phase I metabolism) (Nebert et al. 1993). Enzymes of the CYP1A (and other CYP enzymes) monooxygenase system generate reactive oxygen species (ROS) that react with the organic substance (Isin and Guengerich 2007) to add a functional group [e.g., a hydroxyl group OH-PCB as discussed above, (Buckman et al. 2006)], and the complex may be metabolized further in phase II metabolic processes (Tillitt et al. 2008). The induction of CYP1A in cells is a normal physiological response for metabolism of some endogenous and exogenous substances; however, substance-induced de-coupling of the reduction of oxygen and the oxidation of substrate in the catalytic cycle can lead to inappropriate production of ROS that subsequently react with biological molecules and cause cellular toxicity (Guengerich 2006, Schlezinger et al. 2006). Effects of ROS can cause oxidative stress that includes lipid and protein peroxidation, damage to DNA, and inactivation of cellular enzymes (Guengerich 2006; 2008). The consequences of oxidative stress can include disruption of organism physiology, lesions, and death. However, it is important to recognize that ROS are generated during normal cellular physiology (e.g., during metabolism of endogenous and exogenous substances) and that various mechanisms exist to protect cells from ROS damage [e.g., enzyme systems


including superoxide dismutase (SOD) and glutathione peroxidase (GPX)]. Biomarkers of exposure (e.g., induction of CYP1A activity) to exogenous substances that induce ROSmediated metabolic pathways and biomarkers of induced ROS defense mechanisms (e.g., SOD or GPX activity) are not by themselves indicators of toxicological effects (i.e., ROS-induced damage to biomolecules). Investigations of the biochemistry of AhR regulation, cytochrome P450 induction, and toxicity through ROS induction (and perhaps other mechanisms) are rapidly progressing in the scientific literature (Doering et al. 2013) as refined techniques evolve and more thorough understanding of these processes occur. Numerous substances can generate ROS via induction of the CYP1A enzyme system, and the toxic potency of these substances is closely related to the binding efficiency of the substance to the AhR (Ma 2001). Dioxins [notably tetrachlorodibenzo-p-dioxin (TCDD)] are substances with high affinity to the AhR and consequently generate more ROS and are of higher relative toxicity (Hahn et al. 1997, Clark et al. 2010). Other substances with DL structure can also bind to the AhR and their affinity to the receptor relative to TCDD is used to rank their relative toxicity [TEF (van den Berg et al. 2013)]. Numerous other toxicants found in contaminated waters also have DL activity [e.g., chlorinated dibenzofurans and some PCBs], and summation of the TEF values for the individual contaminants in mixtures is recommended (van den Berg et al. 2013) and frequently used to estimate overall TCDD toxic equivalent (TEQ) (Babut et al. 2009). TEF values of PCBs have been revised as methods for assessment of their DL activity have been improved, and recently TEF values have been reduced (indicating lower toxicity) for PCBs in humans and rodents based on in silico and in vitro screening (Larsson et al. 2015). Dioxin-like PCBs (DL-PCBs) represent approximately 5% of all PCB congeners, concentrations of DL-PCBs are only a small fraction of the total PCBs present (e.g., Figure 1 and 2), and the highest TEF value (0.005 for PCB126) is orders of magnitude lower than the toxicity of TCDD (Table 1). Because of differences in physiology of AhR and CYP1A induction between mammals and fish, TEF values for PCBs in fish are considerably lower than those reported for mammals (van den Berg et al. 1998), which further reduces the importance of PCBs in the computation of TEQ in fish. Consequently, PCBs generally contribute only a small fraction to the total TEQ in fish from contaminated environments [e.g., Newark Bay, NJ, USA; lower Hudson River, USA (Yuan et al. 2001)]. For example, in Atlantic tomcod, Microgadus tomcod, that inhabit some PCB-contaminated estuarine environments on the Atlantic coast of North America (Courtenay et al. 1999, Roy et al. 2001), the highest reported concentration of PCB126 (Hackensack River, NJ, USA) was 2.9 ng/g wet weight of liver corresponding to a TEQ value of 15 pg/g [TEQ for all PCBs present was 28 pg/g (Yuan et al. 2001)]. Assessment of toxicant-induced oxidative stress in wild fish has frequently included measurement of the induction of genes that code for components of the cytochrome P450 enzyme system (e.g., CYP1A), activity of gene products [e.g., ethoxyresorufin-O-deethylase (EROD) activity], or evaluation of oxidative damage in tissues [e.g., formation of lipid peroxidation products and thiobarbituric-acid-reactive substances, protein oxidation, or DNA damage]. In fish, co-planar (DL) PCBs can

PCBs in wild fish

bind to AhRs; however, mono-ortho-substituted congeners have considerably lower AhR activity in fish compared with that in mammals, and this has implications on the expression of CYP1A (Hestermann et al. 2000). The inducibility of fish CYP1A by substances that bind to the AhR appears to vary considerably among tissues (Yuan et al. 2006b), life history stages (e.g., Roy et al. 2011), species (Sarasquete and Segner 2000), and among different populations of the same species (e.g., Courtenay et al. 1999, Yuan et al. 2006a). Higher doses (i.p. injection) of PCB-126 related to induction of CYP1A gene expression were reported in juvenile shortnose sturgeon, Acipenser brevirostrum, and an injection of at least 50 ppb was required before CYP1A induction was observed in liver (Roy et al. 2011). Positive relations between CYP1A induction, i.p. exposure concentration (PCB77), and time after injection have also been reported in Atlantic tomcod (Courtenay et al. 1999). The induction of CYP1A gene expression has been reported in wild fish collected from contaminated environments compared with those from control sites in numerous studies; however, interpretation of the substances responsible for induction and the toxicological consequences of CYP1A expression are difficult to ascertain (Yuan et al. 2001). In a separate study with Atlantic tomcod sampled from various contaminated environments, Yuan et al. (2001) reported that levels of CYP1A expression did not correlate with hepatic burdens of known CYP1A inducers, and suggested that differences in induction of CYP1A among wild fish may be more related to inducers that do not persist in the tissues [e.g., polyaromatic hydrocarbons (PAHs), which are metabolized much more quickly than DL-PCBs]. Complex mixtures of substances that can bind to the AhRs and lead to antagonistic or synergistic interactions with the receptor and further complicate the interpretation of CYP1A expression (Suh et al. 2003) as has been observed both in laboratory fish experiments (Gunawickrama et al. 2008) and in wild fish (Yuan et al. 2001). Further, while induction of CYP1A transcription is consistent with a substance binding to the AhR, the outcome of the induction (i.e., translation of mRNA and activity of gene products) must be established along with evidence of toxic effects resulting from ROS generation (i.e., damage to biomolecules). While the linkage between CYP1A gene expression and EROD is well under-

9

stood, the connection with the incidence of lesions is not well established (Whyte et al. 2000). Gene transcripts of CYP1A are translated into enzymes that catalyze reactions with organic substances (as indicated above), and measurement of the activity of these enzymes has frequently been based on the EROD assay. However, the linkage between CYP1A gene expression and EROD activity is poorly understood, and neither of these measurements has been well linked to toxic effects (Whyte et al. 2000). In English sole, Pleuronectes vetulus, collected from sites with different contaminant (including PCBs) levels, positive relations between hepatic CYP1A (both mRNA expression and protein level) and EROD activity was reported and the relation was also consistent with sediment contaminants for most (but not all) sites (Foster et al. 2003). Similarly, livers of cultured rainbow trout with elevated expression of CYP1A mRNA also had elevated EROD activity [although no link with the presence of contaminant was reported (Valdehita et al. 2012)]. However, laboratory investigations with scup, Stenotomus chrysops, and PCB77 indicated both a dose-dependent induction of CYP1A expression and also suppression of CYP1A gene product activity (Schlezinger and Stegeman 2001), which has also been reported after exposure to PCB-126 (White et al. 1994, 1997). Binding of PCB-126 to the CYP1A protein with consequent production of ROS appeared to cause loss of enzyme activity, perhaps because of oxidative damage to the CYP1A protein (Schlezinger et al. 2006). Higher CYP1A transcription was reported in rainbow trout compared with that in white sturgeon A. transmontanus exposed (i.p. injection) to 50–500 mg/kg body weight of β-naphthoflavone, and no consistent dose relation was observed between CYP1A expression and EROD activity (Doering et al. 2012). Collectively, results indicate that wild and laboratory fish respond to substances that bind to the AhR, CYP1A transcription is induced, and EROD activity is affected (Di Giulio et al. 1989, Whyte et al. 2000); however, alteration in transcription or EROD activity is not by itself a toxicological effect (i.e., does not indicate damage to biomolecules). Based on data from the USGS BEST program that evaluated EROD activity in wild fish collected from large river basins, there is no evidence of a relation between total PCBs concentrations in fish and EROD activity (Figure 5).

300 Hepatic EROD activity (pmol/min/mg of protein)


DOI 10.3109/10408444.2015.1038498

250 200 150 100 50 0 0

0.5

1 1.5 2 2.5 Total PCB concentration (µg/g) based on wet weight of whole fish composite

3

3.5

Figure 5. The hepatic EROD activity in 749 common carp, Cyprinus carpio, and 569 largemouth bass, M. salmoides, collected from sites in the Mississippi and Southeastern River Basins (USGS BEST). The USGS BEST program investigated fish health and contaminant levels in these river systems from 1995 to 2002. Specifics of the sampling sites and study details are available on the website indicated above. In this figure, the maximum concentration of total PCBs reported for a whole fish composite sample at each site was used to relate EROD activity to total PCB concentration among all sites from both river basins. No relation is evident between total PCB concentration and EROD activity.


10

T. B. Henry

Contaminant-induced oxidative damage in organisms is the outcome of ROS generation and the inability of cellular mechanisms to defend against ROS-induced damage (Di Giulio et al. 1989). Studies aimed to improve understanding of linkages between biomarkers of oxidative stressor exposure (e.g., induction of CYP1A, EROD, etc.) and indications of induction of oxidative defense mechanisms [e.g., SOD, glutathione-S-transferase (GST), or catalase (CAT)] with observations of oxidative damage (e.g., lipid or protein oxidation) have been conducted in wild fish. Mullet, Mugil cephalus, and flounder, Platichthys flesus, from contaminated (including PCBs and numerous other substances) sites in the Douro estuary (Iberian Peninsula) had seasonal and species differences in enzyme activities (SOD and CAT), and enzyme activity decreased after fish were held for 1 month in laboratory tanks while hepatic lipid peroxidation increased during some seasons but not others (Ferreira et al. 2005). The investigators concluded that differences in response between mullet and flounder were likely influenced by fasting (flounders did not feed during 1-month observation) and suggested that seasonal variation in compensatory mechanisms could explain some of the results. Numerous other studies of wild fish collected from contaminated environments have measured biomarkers of oxidative stress, but linkages between exposure to specific substances and oxidative injury are not consistent. For example, smallmouth bass, Micropterus dolomieu, collected from a PCB-contaminated site were found to have lower hepatic EROD and SOD activity compared with those from an uncontaminated site (Anderson et al. 2003). In Atlantic salmon, Salmo salar, from the Baltic Sea, PCB concentrations were correlated with hepatic lipid levels but not hepatic CYP1A gene transcription or EROD activity (Hansson et al. 2006). An example of the difficulties in making linkages among biomarkers of exposure and effect for oxidative stress include experiments with juvenile sea bream, Sparus aurata, exposed (4 d) to concentrations of a mixture of PAHs [known inducers of CYP1A and oxidative stress; (Kopecka-Pilarczyk and Correia 2009)]. Changes in hepatic biochemistry were assessed, but no PAH–concentration relations were found for EROD, enzyme activity (CAT, GST, and SOD), or lipid peroxidation; although evidence of PAH metabolism was documented through dose-related increase in PAH metabolites in the bile. The complexity of the physiological responses and differences among fish, life history stages, seasons, and sex combined with the complexity of contaminant exposure scenarios make interpretation of specific contaminant exposure and oxidative injury in wild fish difficult. While there are reports that PCBs can induce CYP1A and EROD in fish (e.g., Roy et al. 2001) there does not appear to be any evidence of toxic effects in tissues generated by oxidative stress consequent to PCB exposure in wild fish.

Effects of PCBs on reproduction in wild fish The process of producing viable offspring that can survive through early life stages (embryonic/larval development) and recruit into the population is an important aspect for the persistence of a healthy fish population. It is possible that toxicants can interfere with reproductive system physiology in adult fish and influence the production/quality of gametes, spawning behavior, fertilization success, or parental care processes (e.g.,


nest tending); effects on these processes could affect population survival. Toxicant-induced reproductive system dysfunction has been documented along with population declines of fathead minnow, Pimephales promelas, after whole-lake exposures to environmentally relevant concentrations of the synthetic estrogen 17 α-ethinylestradiol (Kidd et al. 2007); however, similar studies have not been conducted for PCBs. There does not appear to be any link between fish tissue PCB burdens and the presence of a healthy reproducing fish population, and numerous studies have reported elevated PCB levels in fish collected from otherwise healthy reproducing populations (e.g., Smith 1998, Barnthouse et al. 2003, Baldigo et al. 2006, Barnthouse et al. 2009, Rypel and Bayne 2010). Nonetheless, the presence of PCBs in wild fish and the potential for PCBs to cause reproductive system dysfunction is a question recognized in the literature and the target of numerous investigations in wild fish (e.g., de Lafontaine et al. 2002, Baldigo et al. 2006, Hinck et al. 2006a, b, 2007a, b, 2009a, b, c). Fish reproductive system physiology is regulated by complex interactions between hormones, tissues, and environmental factors; and is vulnerable to modulation by toxicants including PCBs. Toxicants act at low levels of biological organization (i.e., toxic chemicals interact directly with biological molecules); while the potential number of interactions between all 209 PCB congeners, their metabolites, and all possible biological molecules is enormous, there are relatively few biochemical pathways that have been investigated in the context of PCB (or other toxicant) reproductive system toxicity (Di Giulio and Hinton 2008). Among these, disruption of fish reproductive hormone levels (e.g., assessed in blood plasma) and the examination of estrogenic or androgenic effects have been examined in wild fish exposed to PCBs (e.g., Baldigo et al. 2006, Hinck et al. 2007a, 2009a). A toxicant may affect reproductive hormones or processes they control by interfering with the production or metabolism of hormones or by affecting hormone action (e.g., competing for receptor binding sites). However, based on data from the USGS BEST program that reported the ratios of 17β-estradiol (E2)/11-ketotestosterone (11-KT) in blood plasma in wild fish collected from large U.S. river basins, there was no evidence of a relation between total PCBs concentrations in fish and alteration in these sex hormone levels (Figure 6a). Evidence of environmental modulation of reproductive hormone levels (and other endpoints of reproductive system physiology) has been reported in fish from the Hudson River (USA); however, no relation with these endpoints and PCB concentrations was found in males or females of the four evaluated species (Baldigo et al. 2006). Numerous studies have investigated changes in biomarkers of reproductive endocrine system function in wild fish collected from environments with elevated levels of PCBs. One biomarker recognized to be consistent with exposure to environmental estrogens in wild fish is the abnormal induction of the lipoprotein vitellogenin (Vg) or transcripts of Vg genes (vtg) in male fish (Miracle et al. 2006). Female fish produce Vg as a part of their normal reproductive physiology in hepatocytes in response to endogenous estrogen (E2) binding to the estrogen receptor (ER), and subsequent ER–E2 binding to promoters of estrogen-responsive genes including vtg (Hiramatsu et al. 2006). Male fish generally transcribe vtg and have Vg only

PCBs in wild fish

in small amounts, but changes in endogenous estrogen levels or exposure to environmental estrogens can cause high induction (Henry et al. 2009). Metabolism of PCBs into OH-PCBs (as discussed above) can increase affinity to the ER; however, based on receptor assays, the estrogenicity of these OH-PCBs is orders of magnitude lower than endogenous estrogens (Layton et al. 2002), and OH-PCB exposure in wild fish is likely exceedingly low (e.g., Ueno et al. 2007). Male fish collected from environments contaminated by PCBs have been found with induced Vg (e.g., Baldigo et al. 2006, Hinck et al. 2009a), which could indicate a link between PCB exposure and Vg induction; however, male fish collected from other environments with similar PCB concentrations have been shown to not have Vg induction (Hinck et al. 2009a, Figure 6b). Investigations of PCB and other contaminant concentrations in tissues of white sturgeon, A. transmontanus, in various locations within the Columbia River Basin (USA) found no relations between PCBs and plasma Vg (Feist et al. 2005). No relation was observed in wild male blood plasma Vg levels with total PCBs concentrations in whole fish collected from sites within the Mississippi and Southeastern River Basins based on data from the USGS BEST program (Figure 6b). Abnormal induction of Vg in fish has been linked to exposure to anthropogenic

Ratio of blood plasma E2/11-KT

11

estrogenic substances [e.g., 17α-ethinylestradiol (Kidd et al. 2007)] and also to natural substances [e.g., toxic cyanobacteria Microcystis (Rogers et al. 2011)], and the presence of these or other substances is a more likely explanation of Vg induction than that of PCBs (Hinck et al. 2009a). The incidence of abnormalities in wild fish gonads and links to anthropogenic contaminants has generated considerable concern within the public and environmental science community. The presence of oocytes with focal or multifocal presentation upon histopathologic examination of testes (ovotestes) has been diagnostic of the intersex condition in wild fish (Jobling et al. 2002, Hinck et al. 2009a); however, the mechanisms responsible for intersex or the frequency of natural intersex occurrence in wild fish are unknown. Intersex in female fish has been documented (e.g., sperm or early stages of sperm development), and alterations in the normal progression or maturation of gametes has been reported in wild fish of both sexes (e.g., Hinck et al. 2009a). The link between anthropogenic chemical exposure and intersex in wild fish is perhaps strongest for fathead minnows, in which ovotestes were observed in males after several years of whole-lake exposure to 5 ng/L of 17α-ethinylestradiol, which was not observed in fish from unexposed control lakes (Kidd

20

(A)

15 10 5 0

Plasma vitellogenin (mg/ml)

3

(B)

2

1

0

Intersex index (score)


DOI 10.3109/10408444.2015.1038498

2

(C)

1

0 0

0.5

1.5 2.5 1 2 Total PCB concentration (µg/g) based on wet weight of whole fish composite

3

3.5

Figure 6. Reproductive endpoints evaluated in all 376 male common carp, C. carpio, and 275 male largemouth bass, M. salmoides, collected from sites in the Mississippi and Southeastern River Basins (USGS BEST). The USGS BEST program investigated fish health and contaminant levels in these river systems from 1995 to 2002. Specifics of the sampling sites and study details are available on the website indicated above. In this figure, the maximum concentration of total PCBs reported for a whole fish composite sample at each site was used to relate reproductive endpoint to total PCB concentration among all sites from both river basins. No relation is evident between total PCB concentration and (A) ratio of plasma E2/11-KT; (B) plasma Vg concentration; and (C) intersex score.

12

T. B. Henry

et al. 2007). However, a link between PCB exposure and gonad abnormalities in wild fish has not been demonstrated. Based on data from the USGS BEST program that reported the presence of intersex in wild male fish collected from two large U.S. river basins, there was no evidence of a relation of intersex with total PCBs concentrations in composite whole fish samples collected from the same sites (Figure 6c).


Developmental toxicity Early life history stages of fish are often considered vulnerable to negative effects of anthropogenic toxicants, and these life history stages are commonly targeted in regulatory tests to determine substance toxicity (e.g., U.S. EPA and Organization for Economic Cooperation and Development standardized tests for aquatic toxicity). Declines of lake trout, Salvelinus namaycush, populations in the Great Lakes of North America have been attributed in part to exposure to DL chemicals and the susceptibility of early life history stages to these substances (Spitsbergen et al. 1991, Tillitt et al. 2008). DL chemicals can cause toxicity in early life stages of fish by binding to the AhR and induction of CYP1A enzyme systems (discussed above), a process to which young fish are particularly susceptible (Walker and Peterson 1991, Tillitt et al. 2008). Early life stage toxicity in fish resulting from TCDD exposure is frequently indicated by clinical signs that include pericardial edema (Hill et al. 2004), impairment of skeletal development and skeletal abnormalities (Henry et al. 1997), and neurotoxicity (Hill et al. 2004). The potential for PCB-induced early life stage toxicity in fish is strongly dependent on the configuration of the PCB congener, and congeners with planar configuration (PCB 77, 81, 126, and 169) are most potent while the mono-ortho PCBs did not cause early life stage toxicity in the tested fish species (Zabel et al. 1995a, b, c). Based on early life stage toxicity, the TEF for PCB-126 is 0.005, and the other planar PCBs have TEF values that are at least an order of magnitude lower [based on van den Berg et al. (1998) Table 1]. Toxicity of PCB-126 is indicated by a lethal dose predicted to kill 50% of the most sensitive species tested (lake trout) of 29,000 pg/g egg (Zabel et al. 1995b), and the other tested species appeared to be more than ten times less vulnerable than lake trout (Elonen et al. 1998). The TEF values are derived from toxicity tests in which the PCB was administered into the early embryo by injection into the yolk sac, and studies have shown that toxicity (mortality) after co-injection of PCBs with other DL substances can be predicted by summation of the TEF values of the respective substances (Walker and Peterson 1991). DL substances present in the yolk (i.e., after injection) are absorbed by the developing tissues of the fish as yolk is depleted, and toxicity appears during this period and declines rapidly as fish begin feeding exogenously (Foekema et al. 2012). For the non-DL PCBs (e.g., PCB-153) toxicity can occur in fish at early life history stages; however, this does not appear to be mediated by AhR and does not present the typical clinical signs of early life stage toxicity (Broyles and Noveck 1979), and toxic effects only appear at concentrations that are not environmentally relevant. Laboratory investigations of DL early life stage mortality have revealed some of the intricate mechanisms of toxicity, but there is no evidence of these toxic effects occurring in


wild fish that can be attributed to PCB exposure. Of particular concern is the nature of PCB exposure in the laboratory (e.g., aqueous exposure or via yolk injection) and how these compare with actual exposure scenarios in wild fish (Zint et al. 1995). Aqueous PCBs can be absorbed by fish embryos during laboratory exposures (e.g., Koponen et al. 2000, Chambers et al. 2012), but the extent to which PCBs are absorbed by wild fish embryos is unknown. An aqueous exposure of rainbow trout, Oncorhynchus mykiss, embryos to PCB-77 in glass beakers (Koponen et al. 2000) led to complete absorption of all added PCBs (1 μg/L), and highest exposure (100 μg/L) resulted in absorption of 14% of all added PCBs (majority of PCBs were indicated to be sorbed to glass). If similar absorption of PCBs (Aroclor 1254) by zebrafish embryos occurred in the study of Li et al. (2014), then the amount of PCBs to which embryos were exposed (not measured) was likely extremely high because nominal PCB concentrations in water were 0.125–1 mg/L; exposure solutions were replaced daily. While the results of Li et al. (2014) provide interesting information on cardiac development in zebrafish, the exposure scenario does not enable any conclusions to be drawn about effects of PCB exposure in wild fish embryos. Absorption of aqueous radiolabeled PCB-126 added to beakers containing shortnose sturgeon, A. brevirostum, or Atlantic sturgeon, A. oxyrinchus, occurred rapidly, and after 24-h exposure to 10 μg/L the mean reported embryo concentrations were 12 and 27 ng/g, respectively, in both species (Chambers et al. 2012). There is no information on PCB concentrations in wild fish embryos that are attributed solely to exposure of the embryo within the abiotic environment (i.e., absorption of PCBs from the environment by the embryos), and the concentrations of PCBs in wild fish embryos are more likely to be attributed to maternal transfer during oocyte development (Russell et al. 1995). The concentrations of total PCBs have been documented in eggs of wild fish and appear to vary among fish species, maternal tissues, and lipid levels (Russell et al. 1995), and PCB hydrophobicity (Fisk and Johnston 1998). When PCB concentrations are corrected for tissue lipid levels, similar concentrations of PCBs are expected in eggs as observed in maternal tissues (Russell et al. 1995, Foekema et al. 2014). The concentrations of PCBs in embryos of fish from the Great Lakes (U.S.) have been monitored over time, and declining concentrations have been documented (Mac et al. 1993) that are consistent with observations of declining PCBs in older life history stages of fish (discussed above). Elevated concentrations of PCBs in lake trout embryos have been reported for samples collected in the 1970 s (Mac et al. 1993) including concentrations up to 10 mg/g (total PCBs) in lake trout from Lake Michigan (Stauffer 1979). A total PCB concentration of 240 ng/g was reported in walleye, Stizostedion vitreum, eggs from Lake Superior (Fisk and Johnston 1998). These concentrations are consistent with recent measurements of total PCBs in adult fish (males and females) among large river systems in North America [e.g., ⬍ 3 μg/g; USGS BEST, (Figure 3)] and the expectation that similar concentrations of total PCBs are to be found in the eggs of these fish (Russell et al. 1995). If tissue PCB concentrations of adult fish are used to indicate likely concentrations of PCBs in embryos, of greatest interest (i.e., potential toxicity to embryos) is the


DOI 10.3109/10408444.2015.1038498

concentration of PCB-126 in which maximum concentrations in wild fish reported from PCB-contaminated environments were ⬍ 3 ng/g [based on measured concentrations of PCB 126 of 3 ng/g wet weight of liver in Atlantic tomcod (Yuan et al. 2001, also discussed above)]. The potential for PCBs to induce early life stage mortality in fish is dependent on the PCB concentration in the embryo. In shortnose sturgeon, A. transmontanus, ∼12 ng of PCB-126/g egg reduced hatching by ∼30% (to 63% survival), but tested concentrations of ⬎ 12 ng/g egg did not affect hatching of Atlantic sturgeon, A. oxyrinchus (Chambers et al. 2012). When PCB concentrations were considerably higher in eggs of lake trout (e.g., in 1979), relations between PCB concentrations and embryo mortality were significant, and this correlation was suggested as evidence of a potential link between PCB exposure and population decline in wild lake trout (Mac et al. 1993). However, a critical review of the evidence conducted to evaluate the potential that PCBs (and other contaminants) were responsible for early life stage mortality of lake trout during the late 1970 s indicated that PCB concentrations were considerably lower than what would have been required to cause mortality (Fitzsimons 1995). Present PCB levels in embryos of wild lake trout are orders of magnitude below those required for induction of early life stage mortality, and other fish species that are reported to be much less vulnerable to early life mortality are at even lower risk of early life stage mortality from PCB exposure.

Changes in population genetics linked to toxicant exposure in wild fish Potential alteration in population genetics of wild fish resulting from toxicant exposure is of scientific interest and perhaps ecological relevance in the context of the sustainability of wild fish populations. Perhaps the most well-documented polymorphism in genetic sequence in a population of wild fish that inhabit a location with elevated levels of toxicants known to induce the AhR (discussed above) are Atlantic tomcod from the Hudson River (Wirgin et al. 2011). Fish from this population are reported to be more tolerant to inducers (e.g., TCDD and some PCBs) of CYP1A enzymes and therefore may have reduced production of ROS and subsequent oxidative stress (Yuan et al. 2006a, b). Examination of Hudson River tomcod from this population revealed a polymorphism (six-base deletion) in the DNA sequence that codes for an AhR, and this genetic alteration was not found in populations of Atlantic tomcod from locations with lower levels of toxicants known to induce CYP1A (Wirgin et al. 2011). Changes in genetic sequence of AhR genes in other fish species (e.g., Atlantic killifish Fundulus heteroclitus) have been reported to be consistent with greater tolerance (i.e., ability to resist induction of CYP1A) to dioxin (Hahn et al. 2004) and perhaps also to PCB-126 (Nacci et al. 2010). Recent investigations of Atlantic killifish indicate that considerable differences exist among subpopulations from contaminated environments regarding the nature of their resistance (Clark et al. 2014), and ongoing investigations are certain to reveal greater insight into Ah-mediated toxicity, resistance, and this interesting aspect of wild fish toxicology (Proestou et al. 2014, Reitzel et al. 2014).

PCBs in wild fish

13

Important questions regarding observations of genetic differences in populations of wild fish include the forces responsible for the selective pressure that resulted in survival of these specific genetic polymorphisms, and the effects on the population as a result of the persistence of these polymorphisms. It is possible that a toxicant-induced induction of CYP1A could generate sufficient stress on wild populations such that those individuals, which had lower levels of CYP1A induction based on mutations in AhR genes, had a competitive advantage, higher survival, and contributed more offspring to subsequent generations. However, although the occurrence of wild fish populations that are tolerant of toxicants is being documented as further investigations are undertaken (Brammell et al. 2013), the agents responsible for the selective pressure within these populations are difficult to ascertain. The documentation of now numerous independent wild fish populations with resistance to AhR-mediated toxicants has led some researchers to propose that genotypes capable of conferring resistance were present in fish populations prior to anthropogenic pollutants rather than being caused by the pollutants themselves (Whitehead et al. 2012). A population of Atlantic killifish, F. heteroclitus, previously demonstrated to be more resistant to CYP1A induction, was exposed to pathogenic bacteria in the laboratory, but the results did not show higher incidence of disease compared with fish from a control site [i.e., no evidence of a compensatory trade-off (e.g., immunosuppression) for the enhanced resistance to toxicant exposure (Nacci et al. 2009)]. While reduced CYP1A induction could be advantageous in environments contaminated by CYP1A inducers, genetic polymorphisms should not be offered as an explanation for why CYP1A induction is not observed in fish from these environments unless a thorough analysis of genotypes has been conducted (i.e., as in Wirgin et al. 2011). Some fish species have CYP1A genes with naturally low levels of induction and EROD activity (Parente et al. 2011), and natural variation in CYP1A induction among populations of wild fish is to be expected and should not be inappropriately linked to the presence of toxicants.

Investigations of PCB-induced organ system dysfunction in wild fish Effects of laboratory PCB exposures on physiology of various body systems have been investigated, which indicate that toxicity does not occur at PCB concentrations reported in wild fish. Dietary exposure to seven PCBs caused morphological alterations in thyroid tissue detectable upon histopathology in seabass, Dicentrarchus labrax; however, physiological adjustments in thyroid hormone enabled hormone levels to be maintained in tissues (Schnitzler et al. 2011). The potential effect of PCBs on the hypothalamic-pituitary-thyroid (HPT) axis has been investigated in various fish species; while changes in some indices of the HPT axis have been correlated with PCB concentrations in wild fish (Brar et al. 2010), these changes appear to be within the range for fish to compensate without any negative effects on physiology [hormone levels, growth, and mortality (e.g., Besselink et al. 1996, Brown et al. 2004, Buckman et al. 2007b)]. Immunosuppression and increased vulnerability to infectious disease have been considered possible consequences of


14

T. B. Henry

contaminant exposure in wild fish. Brown bullhead, Ameiurus nebulosus, exposed in the laboratory to PCBs (i.p. injection) were reported to have altered immune system function and increased susceptibility to challenge with Edwardsiella ictaluri, a bacterial pathogen of fish (Iwanowicz et al. 2009). Although the results of the Iwanowicz et al. (2009) study are interesting within the context of the experiment, the exposure scenario is not representative of wild fish and does not enable conclusions to be made regarding the effects of PCB exposure on immune system function in wild fish. The single PCB injection into the peritoneal cavity differs from the chronic PCB absorption and tissue distribution that occurs in wild fish, and the effects of the i.p. injection were not investigated [i.e., fish injected with vehicle control (vegetable oil) were not compared with uninjected control fish]. Further, the fish used in the study were survivors from a previous E. ictaluri infection episode that killed 40% of the fish in the facility; consequently, the status of the immune system in these fish compared with that in wild fish cannot be determined. Other laboratory studies that have investigated PCB exposure and disease resistance in fish have reported inconsistent results with some indicating no effect on disease resistance (e.g., Spitsbergen et al. 1988, Powell et al. 2003) or a minor indication of reduced disease resistance (48% survival in fish treated with PCB compared with 68% control fish survival, after exposure to pathogen Aeromonas salmonicida) in Arctic char during a fasting experiment (Maule et al. 2005). Collectively, these laboratory studies demonstrate the complexity of investigating effects of toxicants on immune system function in fish, but are unable to provide conclusions on the effects of PCB exposure on immune system and disease resistance in wild fish. There are few studies that have compared fish immune system function or pathogen resistance in wild fish between PCB-contaminated and uncontaminated sites. No significant alterations in immune system function were found in adult walleye collected from a PCB-contaminated site relative to an uncontaminated site in Wisconsin USA (Barron et al. 2000). Physiology of the fish immunity is complex and toxicants may influence components of the immune system, but there is no consistent evidence that PCB exposure has led to immune system dysfunction and reduced pathogen resistance in wild fish. Further investigation may reveal greater understanding of the potential interactions between PCB exposure and fish immune system response to infectious pathogens. Tissue histopathology of wild fish collected from PCBcontaminated locations has been conducted in numerous studies (e.g., USGS BEST), and examinations have included most fish tissue types. At some locations increased numbers of lesions (e.g., pre-neoplastic foci and neoplasia) have been reported in fish from contaminated environments (e.g., Dey et al. 1993, Barron et al. 2000). Lesions detected during histologic examination of wild fish, however, have been extremely difficult to link to toxicant exposure because of the natural variation in tissues (e.g., species, age, sex, condition, and season), the presence of normal parasitic infection in tissues, and the presence of numerous types of toxicants at the same location (Dey et al. 1993, Anderson et al. 2003).


Conclusions The presence of persistent organic chemicals such as PCBs in the environment has generated considerable concern regarding the potential for these substances to impact aquatic ecosystems and fish health. While production of PCBs has been banned in most countries, and concentrations of PCBs have declined in surface waters and sediments, PCBs continue to contaminate wild fish tissues and concerns for negative effects on fish health persist. Concentrations of PCBs in wild fish have decreased and continue to decline in the environment and there is little evidence that PCBs have had any widespread effects on health or survival of wild fish. Toxicity of PCBs has been investigated thoroughly in laboratory experiments, but there does not appear to be any consistent evidence of PCB-induced toxicity in wild fish populations from PCB-contaminated environments or in large-scale geographic assessments (e.g., USGS BEST). Perhaps the most vulnerable fish life history stage to PCBinduced toxicity are fish embryos; however, evidence indicates that concentrations of PCBs in wild fish embryos are considerably below those reported to cause toxicity. The investigation of the toxicity of PCBs in fish has been conducted for nearly 50 years and the evidence accumulated over this period indicates that the presence of PCBs in wild fish have been of minimal toxicity. As in any field of scientific investigation, and perhaps in particular within toxicology, some aspects (e.g., pathways of toxicity) might not have been investigated to the same extent as others and new research methods will be applied in the future to provide additional information on the toxicology of PCBs in wild fish. Nonetheless, it is important for environmental scientists, regulators, and the public to recognize that PCBs in wild fish are likely of minimal toxicity based on the accumulated evidence.

Acknowledgments The author thanks Professor A. Fisk (University of Windsor, ON, Canada) for independent review and comment on the manuscript prior to submission, and the thoughtful suggestions of five anonymous reviewers who evaluated the submitted manuscript.

Declaration of interest Funding for the preparation of this manuscript was provided by the General Electric Company; although representatives of the company reviewed the manuscript, the opinions are exclusively those of the author and do not necessarily reflect the views of the sponsor. The author’s affiliation is as shown on the cover page and the author has sole responsibility for the writing and content of the paper.

References Abramowicz DA. (1990). Aerobic and anaerobic biodegradation of PCBs–a review. Crit Rev Biotech, 10, 241–9. Adams BA, Cyr DG, Eales JG. (2000). Thyroid hormone deiodination in tissues of American plaice, Hippoglossoides platessoides: characterization and short-term responses to polychlorinated biphenyls (PCBs) 77 and 126. Comp Biochem Physiol C Pharmacol, 127, 367–78. Anderson MJ, Cacela D, Beltman D, Teh SJ, Okihiro MS, Hinton DE, et al. (2003). Biochemical and toxicopathic biomarkers assessed in smallmouth bass recovered from a polychlorinated biphenylcontaminated river. Biomarkers, 8, 371–93.


DOI 10.3109/10408444.2015.1038498

Andersson PL, Berg AH, Olsson PE, Tysklind M. (1998). Distribution of selected polychlorinated biphenyls (PCBs) in brain and liver of Arctic char (Salvelinus alpinus). Mar Environ Res, 46, 501–4. Ankley GT, Tillitt DE, Giesy JP, Jones PD, Verbrugge DA. (1991). Bioassay-derived 2,3,7,8-tetrachlorodibenzo-paradioxin equivalents in PCB-containing extracts from the flesh and eggs of Lake-Michigan Chinook salmon (Oncorhynchus tshawytscha) and possible implication for reproduction. Can J Fish Aquat Sci, 48, 1685–90. Antunes P, Amado J, Vale C, Gil O. (2007a). Influence of the chemical structure on mobility of PCB congeners in female and male sardine (Sardina pilchardus) from Portuguese coast. Chemosphere, 69, 395–402. Antunes P, Gil O, Reis-Henriques MA. (2007b). Evidence for higher biomagnification factors of lower chlorinated PCBs in cultivated seabass. Sci Total Environ, 377, 36–44. Ashley JTF, Horwitz R, Steinbacher JC, Ruppel B. (2003). A comparison of congeneric PCB patterns in American eels and striped bass from the Hudson and Delaware River estuaries. Mar Poll Bull, 46, 1294–308. Ashley JTF, Secor DH, Zlokovitz E, Wales SQ, Baker JE. (2000). Linking habitat use of Hudson River striped bass to accumulation of polychlorinated biphenyl congeners. Environ Sci Technol, 34, 1023–29. Babut M, Miege C, Villeneuve B, Abarnou A, Duchemin J, Marchand P, Narbonne JF. (2009). Correlations between dioxin-like and indicators PCBs: potential consequences for environmental studies involving fish or sediment. Environ Poll, 157, 3451–56. Baldigo BP, Sloan RJ, Smith SB, Denslow ND, Blazer VS, Gross TS. (2006). Polychlorinated biphenyls, mercury, and potential endocrine disruption in fish from the Hudson River. Aquat Sci, 68, 206–28. Barnthouse LW, Glaser D, Young J. (2003). Effects of historic PCB exposures on the reproductive success of the Hudson River striped bass population. Environ Sci Technol, 37, 223–28. Barnthouse LW, Glaser D, DeSantis L. (2009). Polychlorinated biphenyls and Hudson River white perch: implications for population-level ecological risk assessment and risk management. Integ Environ Assess Manage, 5, 435–44. Barron MG, Anderson MJ, Cacela D, Lipton J, Teh SJ, Hinton DE, et al. (2000). PCBs, liver lesions, and biomarker responses in adult walleye (Stizostedium vitreum vitreum) collected from Green Bay, Wisconsin. J Great Lakes Res, 26, 250–71. Berg K, Puntervoll P, Klungsoyr J, Goksoyr A. (2011). Brain proteome alterations of Atlantic cod (Gadus morhua) exposed to PCB 153. Aquat Toxicol, 105, 206–17. Besselink HT, van Beusekom S, Roex E, Vethaak AD, Koeman JH, Brouwer A. (1996). Low hepatic 7-ethoxyresorufin-O-deethylase (EROD) activity and minor alterations in retinoid and thyroid hormone levels in flounder (Platichthys flesus) exposed to the polychlorinated biphenyl (PCB) mixture, Clophen A50. Environ Poll, 92, 267–74. Bettinetti R, Quadroni S, Manca M, Piscia R, Volta P, Guzzella L, et al. (2012). Seasonal fluctuations of DDTs and PCBs in zooplankton and fish of Lake Maggiore (Northern Italy). Chemosphere, 88, 344–51. Bhavsar SP, Fletcher R, Hayton A, Reiner EJ, Jackson DA. (2007). Composition of dioxin-like PCBs in fish: an application for risk assessment. Environ Sci Technol, 41, 3096–102. Black DE, Gutjahr-Gobell R, Pruell RJ, Bergen B, McElroy AE. (1998). Effects of a mixture of non-ortho-and mono-ortho-polychlorinated biphenyls on reproduction in Fundulus heteroclitus (Linnaeus). Environ Toxicol Chem, 17, 1396–404. Bodiguel X, Maury O, Mellon-Duval C, Roupsard F, Le Guellec AM, Loizeau V. (2009). A dynamic and mechanistic model of PCB bioaccumulation in the European hake (Merluccius merluccius). J Sea Res, 62, 124–34. Boon JP, Hoekstein M, Van Ommen B. (1989). Monooxygenase and glutathione-s-transferase GST activities in flounder Platichthys flesus after a single injection with a technical PCB-mixture Clophen A-40 characterization of the GST pattern. Mar Environ Res, 28, 143–43. Borga K, Di Guardo A. (2005a). Comparing in measured and predicted PCB concentrations Arctic seawater and marine biota. Sci Total Environ, 342, 281–300. Borga K, Fisk AT, Hargrave B, Hoekstra PF, Swackhamer D, Muir DCG. (2005b). Bioaccumulation factors for PCBs revisited. Environ Sci Technol, 39, 4523–32. Borga K, Gabrielsen GW, Skaare JU, Kleivane L, Norstrom RJ, Fisk AT. (2005c). Why do organochlorine differences between arctic regions vary among trophic levels? Environ Sci Technol, 39, 4343–52. Brammell BF, Price DJ, Birge WJ, Elskus AA. (2004). Apparent lack of CYP1A response to high PCB body burdens in fish from a chronically contaminated PCB site. Mar Environ Res, 58, 251–5.

PCBs in wild fish

15

Brammell BF, Price DJ, Birge WJ, Elskus AA. (2013). Lack of CYP1A responsiveness in species inhabiting chronically contaminated habitats: two varieties of resistance? Comp Biochem Physiol C Toxicol Pharmacol, 157, 212–9. Brar NK, Waggoner C, Reyes JA, Fairey R, Kelley KM. (2010). Evidence for thyroid endocrine disruption in wild fish in San Francisco Bay, California, USA. Relationships to contaminant exposures. Aquat Toxicol, 96, 203–15. Breivik K, Alcock R. (2002). Emission impossible? The challenge of quantifying sources and releases of POPs into the environment. Environ Int, 28, 137–8. Breivik K, Sweetman A, Pacyna JM, Jones KC. (2002). Towards a global historical emission inventory for selected PCB congeners–a mass balance approach 1. Global production and consumption. Sci Total Environ, 290, 181–98. Brown JF. (1992). Metabolic afterations of PCB residues in aquatic fauna–distributions of cytochrome P4501A-like and P4502B-like activities. Mar Environ Res, 34, 261–66. Brown JF, Wagner RE. (1990). PCB movement, dechlorination, and detoxication in the Acushnet Estuary. Environ Toxicol Chem, 9, 1215–33. Brown SB, Evans RE, Vandenbyllardt L, Finnson KW, Palace VP, Kane AS, et al. (2004). Altered thyroid status in lake trout (Salvelinus namaycush) exposed to co-planar 3,3′,4,4′,5-pentachlorobiphenyl. Aquat Toxicol, 67, 75–85. Broyles RH, Noveck MI. (1979). Uptake and distribution of 2,5,2′, 5′-tetrachlorobiphenyl in developing lake trout. Toxicol Appl Pharmacol, 50, 291–98. Buckman AH, Brown SB, Small J, Muir DCG, Parrott J, Solomon KR, Fisk AT. (2007a). Role of temperature and enzyme induction in the biotransformation of polychlorinated biphenyls and bioformation of hydroxylated polychlorinated biphenyls by rainbow trout (Oncorhynchus mykiss). Environ Sci Technol, 41, 3856–63. Buckman AH, Fisk AT, Parrott JL, Solomon KR, Brown SB. (2007b). PCBs can diminish the influence of temperature on thyroid indices in rainbow trout (Oncorhynchus mykiss). Aquat Toxicol, 84, 366–78. Buckman AH, Wong CS, Chow EA, Brown SB, Solomon KR, Fisk AT. (2006). Biotransformation of polychlorinated biphenyls (PCBs) and bioformation of hydroxylated PCBs in fish. Aquat Toxicol, 78, 176–85. Campbell LM, Muir DCG, Whittle DM, Backus S, Norstrom RJ, Fisk AT. (2003). Hydroxylated PCBs and other chlorinated phenolic compounds in lake trout (Salvelinus namaycush) blood plasma from the Great Lakes Region. Environ Sci Technol, 37, 1720–25. Campfens J, MacKay D. (1997). Fugacity-based model of PCB bioaccumulation in complex aquatic food webs. Environ Sci Technol, 31, 577–83. Carlson DB, Williams DE. (2001). 4-hydroxy-2 ′,4′,6′-trichlorobiphenyl and 4-hydroxy-2 ′,3 ′,4 ′,5 ‘-tetrachlorobiphenyl are estrogenic in rainbow trout. Environ Toxicol Chem, 20, 351–58. Chambers RC, Davis DD, Habeck EA, Roy NK, Wirgin I. (2012). Toxic effects of PCB126 and TCDD on shortnose sturgeon and Atlantic sturgeon. Environ Toxicol Chem, 31, 2324–37. Chang BV, Liu WG, Yuan SY. (2001). Microbial dechlorination of three PCB congeners in river sediment. Chemosphere, 45, 849–56. Chevreuil M, Carru AM, Chesterikoff A, Boet P, Tales E, Allardi J. (1995). Contamination of fish from different areas of the River Seine (France) by organic (PCB and pesticides) and metallic (Cd, Cr, Cu, Fe, Mn, Pb, and Zn) micropollutants. Sci Total Environ, 162, 31–42. Clark BW, Matson CW, Jung D, Di Giulio RT. (2010). AHR2 mediates cardiac teratogenesis of polycyclic aromatic hydrocarbons and PCB126 in Atlantic killifish. Aquat Toxicol, 99, 232–40. Clark BW, Cooper EM, Stapleton M, Di Giulio RT. (2014). Compound- and mixture-specific differences in resistance to polycyclic aromatic hydrocarbons and PCB-126 among Funddulus heteroclitus subpopulations throughout the Elizabeth River Estuary (Virginia, USA). Environ Sci Technol, 47, 10556–66. Connolly JP, Zahakos HA, Benaman J, Ziegler CK, Rhea JR, Russell K. (2000). A model of PCB fate in the Upper Hudson River. Environ Sci Technol, 34, 4076–87. Cook PM, Robbins JA, Endicott DD, Lodge KB, Guiney PD, Walker MK, et al. (2003). Effects of aryl hydrocarbon receptor-mediated early life stage toxicity on lake trout populations in Lake Ontario during the 20th century. Environ Sci Technol, 37, 3864–77. Couillard CM, Wirgin II, Lebeuf M, Legare B. (2004). Reduction of cytochrome P4501A with age in Atlantic tomcod from the St. Lawrence Estuary, Canada: relationship with emaciation and possible effect of contamination. Aquat Toxicol, 68, 233–47.


16

T. B. Henry

Courtenay SC, Grunwald CM, Kreamer GL, Fairchild WL, Arsenault JT, Ikonomou M, Wirgin, II. (1999). A comparison of the dose and time response of CYP1A1 mRNA induction in chemically treated Atlantic tomcod from two populations. Aquat Toxicol, 47, 43–69. Dabrowska H, Bernard E, Barska I, Radtke K. (2009). Inter-tissue distribution and evaluation of potential toxicity of PCBs in Baltic cod (Gadus morhua L.). Ecotoxicol Environ Saf, 72, 1975–84. Daouk T, Larcher T, Roupsard F, Lyphout L, Rigaud C, Ledevin M, et al. (2011). Long-term food-exposure of zebrafish to PCB mixtures mimicking some environmental situations induces ovary pathology and impairs reproduction ability. Aquat Toxicol, 105, 270–78. de Boer J, Dao QT, van Leeuwen SPJ, Kotterman MJJ, Schobben JHM. (2010). Thirty year monitoring of PCBs, organochlorine pesticides and tetrabromodiphenylether in eel from The Netherlands. Environ Poll, 158, 1228–36. De Laender F, Van Oevelen D, Frantzen S, Middelburg JJ, Soetaert K. (2010). Seasonal PCB bioaccumulation in an arctic marine ecosystem: a model analysis incorporating lipid dynamics, food-web productivity and migration. Environ Sci Technol, 44, 356–61. de Lafontaine Y, Gilbert NL, Dumouchel F, Brochu C, Moore S, Pelletier T, et al. (2002). Is chemical contamination responsible for the decline of the copper redhorse (Moxostoma hubbsi), an endangered fish species, in Canada? Sci Total Environ, 298, 25–44. Deshpande AD, Dockum BW, Cleary T, Farrington C, Wieczorek D. (2013). Bioaccumulation of polychlorinated biphenyls and organochlorine pesticides in young-of-the-year bluefish (Pomatomus saltatrix) in the vicinity of a superfund site in New Bedford Harbor, Massachusetts, and in the adjacent waters. Mar Pollut Bull, 72, 146–64. Dey WP, Peck TH, Smith CE, Kreamer GL. (1993). Epizoology of hepatic neoplasia in Atlantic tomcod (Microgadus tomcod) from the Hudson River Esturary. Can J Fish Aquat Sci, 50, 1897–907. Di Giulio RT, Hinton DE. (2008). The Toxicology of Fishes. Boca Raton, FL: CRC Press, Taylor & Francis Group. Di Giulio RT, Washburn PC, Wenning RJ, Winston GW, Jewell CS. (1989). Biochemical responses in aquatic animals: a review of determinants of oxidative stress. Environ Toxicol Chem, 8, 1103–24. Doering JA, Wiseman S, Beitel SC, Tendler BJ, Giesy JP, Hecker M. (2012). Tissue specificity of aryl hydrocarbon receptor (AhR) mediated responses and relative sensitivity of white sturgeon (Acipenser transmontanus) to an AhR agonist. Aquat Toxicol, 114, 125–33. Doering JA, Giesy JP, Wiseman S, Hecker M. (2013). Predicting the sensitivity of fishes to dioxin-like compounds: possible role of the aryl hydrocarbon receptor (AhR) ligand binding domain. Environ Sci Pollution Res, 20, 1219–24. Duffy JE, Carlson E, Li Y, Prophete C, Zelikoff JT. (2002). Impact of polychlorinated biphenyls (PCBs) on the immune function of fish: age as a variable in determining adverse outcome. Mar Environ Res, 54, 559–63. Elonen GE, Spehar RL, Holcombe GW, Johnson RD, Fernandez JD, Erickson RJ, Tietge JE, Cook PM. (1998). Comparative toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin to seven freshwater fish species during early life-stage development. Environ Toxicol Chem, 17, 472–83. El-Shaarawi AH, Backus S, Zhu R, Chen Y. (2011). Modelling temporal and spatial changes of PCBs in fish tissue from Lake Huron. Environ Mon Assess, 173, 611–23. Fairey R, Taberski K, Lamerdin S, Johnson E, Clark RP, Downing JW, Newman J, Petreas M. (1997). Organochlorines and other environmental contaminants in muscle tissues of sportfish collected from San Francisco Bay. Mar Poll Bull, 34, 1058–71. Feist GW, Webb MAH, Gundersen DT, Foster EP, Schreck CB, Maule AG, Fitzpatrick MS. (2005). Evidence of detrimental effects of environmental contaminants on growth and reproductive physiology of white sturgeon in impounded areas of the Columbia River. Environ Health Perspect, 113, 1675–82. Fernandez MP, Ikonomou MG, Courtenay SC, Wirgin II. (2004). Spatial variation in hepatic levels and patterns of PCBs and PCDD/Fs among young-of-the-year and adult Atlantic tomcod (Microgadus tomcod) in the Hudson River estuary. Environ Sci Technol, 38, 976–83. Ferreira M, Moradas-Ferreira P, Reis-Henriques MA. (2005). Oxidative stress biomarkers in two resident species, mullet (Mugil cephalus) and flounder (Platichtkys flesus), from a polluted site in River Douro Estuary, Portugal. Aquat Toxicol, 71, 39–48. Fisher MA, Eversole R, Mehne C, Means JC, Delong C, Mihalko D, Ide CF. (2008). Liver cyp1A protein expression and pigmented


macrophage aggregates as indicators of polychlorinated biphenyl exposure in carp Cyprinus carpio L. from the Kalamazoo River Superfund site, Michigan. J Fish Biol, 72, 1960–75. Fisher MA, Mehne C, Means JC, Ide CF. (2006). Induction of CYP1A mRNA in carp (Cyprinus carpio) from the Kalamazoo River polychlorinated biphenyl-contaminated superfund site and in a laboratory study. Arch Environ Contam Toxicol, 50, 14–22. Fisk AT, Johnston TA. (1998). Maternal transfer of organochlorines to eggs of walleye (Stizostedion vitreum) in Lake Manitoba and western Lake Superior. J Great Lakes Res, 24, 917–28. Fitzsimons JD. (1995). A critical review of the effects of contaminants on early life stage (ELS) mortality of lake trout in the Great Lakes. J Great Lakes Res, 21, 267–76. Flammarion P, Fortier V, Migeon B, Morfin P, Garric J. (1999). Cytochrome P450 1A induction in gudgeon (Gobio gobio): laboratory and field studies. Biomarkers, 4, 59–71. Foekema EM, Fischer A, Parron ML, Kwadijk C, de Vries P, Murk AJ. (2012). Toxic concentrations in fish early life stages peak at a critical moment. Environ Toxicol Chem, 31, 1381–90. Foster GD, Miller CV, Huff TB, Roberts E. (2003). Pesticides, polycyclic aromatic hydrocarbons, and polychlorinated biphenyls in transport in two atlantic coastal plain tributaries and loadings to Chesapeake bay. J Environ Sci Health A Tox Hazard Subst Environ Eng, 38, 1177–200. French TD, Petro S, Reiner EJ, Bhavsar SP, Jackson DA. (2011). Thirtyyear time series of pcb concentrations in a small invertivorous fish (Notropis Hudsonius): an examination of post-1990 trajectory shifts in the lower Great Lakes. Ecosystems, 14, 415–29. Foekema EM, Parron ML, Mergia MT, Carolus ERM, vd Berg JHJ, Kwadijk C, et al. (2014). Internal effect concentrations of organic substances for early life development of egg-exposed fish. Ecotoxicol Environ Saf, 101, 14–22. Fu CT, Wu SC. (2005). Bioaccumulation of polychlorinated biphenyls in mullet fish in a former ship dismantling harbour, a contaminated estuary, and nearby coastal fish farms. Mar Poll Bull, 51, 932–9. Fu CT, Wu SC. (2006). Seasonal variation of the distribution of PCBs in sediments and biota in a PCB-contaminated estuary. Chemosphere, 62, 1786–94. Furukawa K, Fujihara H. (2008). Microbial degradation of polychlorinated biphenyls: biochemical and molecular features. J Biosci Bioengin, 105, 433–49. Gerpe M, Kling P, Berg AH, Olsson PE. (2000). Arctic char (Salvelinus alpinus) metallothionein: cDNA sequence, expression, and tissue-specific inhibition of cadmium-mediated metallothionein induction by 17 beta-estradiol, 4-OH-PCB 30, and PCB 104. Environ Toxicol Chem, 19, 638–45. Greenfield BK, Teh SJ, Ross JRM, Hunt J, Zhang GH, Davis JA, et al. (2008). Contaminant concentrations and histopathological effects in Sacramento splittail (Pogonichthys macrolepidotus). Arch Environ Contam Toxicol, 55, 270–81. Grimes AC, Erwin KN, Stadt HA, Hunter GL, Gefroh HA, Tsai HJ, Kirby ML. (2008). PCB126 exposure disrupts zebrafish ventricular and branchial but not early neural crest development. Toxicol Sci, 106, 193–205. Grinwis GCM, van den Brandhof EJ, Engelsma MY, Kuiper RV, Vaal MA, Vethaak AD, et al. (2001). Toxicity of PCB-126 in European flounder (Platichthys flesus) with emphasis on histopathology and cytochrome P4501A induction in several organ systems. Arch Toxicol, 75, 80–7. Grinwis GCM, Vethaak AD, Wester PW, Vos JG. (2000). Toxicology of environmental chemicals in the flounder (Platichthys flesus) with emphasis on the immune system: field, semi-field (mesocosm) and laboratory studies. Toxicol Lett (Shannon),112–113, 289–301. Guengerich FP. (2006). Cytochrome P450s and other enzymes in drug metabolism and toxicity. AAPS J, 8, E101–11. Guengerich FP. (2008). Cytochrome P450 and chemical toxicology. Chem Res Toxicol, 21, 70–83. Gunawickrama SHNP, Aarsaether N, Orbea A, Cajaraville MP, Goksoyr A. (2008). PCB77 (3,3′, 4,4′-tetrachlorobiphenyl) co-exposure prolongs CYP1A induction, and sustains oxidative stress in B(a)P-exposed turbot, Scophthalmus maximus, in a long-term study. Aquat Toxicol, 89, 65–74. Hahn ME, Karchner SI, Franks DG, Merson RR. (2004). Aryl hydrocarbon receptor polymorphisms and dioxin resistance in Atlantic killifish (Fundulus heteroclitus). Pharmacogenetics, 14, 131–43. Hahn ME, Karchner SI, Shapiro MA, Perera SA. (1997). Molecular evolution of two vertebrate aryl hydrocarbon (dioxin) receptors


DOI 10.3109/10408444.2015.1038498

(AHR1 and AHR2) and the PAS family. Proc Natl Acad Sci U S A, 94, 13743–8. Hansson MC, Persson ME, Larsson P, Kjellman C, von Schantz T. (2006). Polychlorinated biphenyl load, aryl hydrocarbon receptor, and cytochrome P4501a1 induction in a wild population of Atlantic salmon (Salmo salar) from the Baltic Sea. Environ Toxicol Chem, 25, 2197–207. Haranczyk M, Puzyn T, Ng EG. (2010). On enumeration of congeners of common persistent organic pollutants. Environ Poll, 158, 2786–89. Hawker DW, Connell DW. (1988). Octanol water partition-coefficients of polychlorinated biphenyl congeners. Environ Sci Technol, 22, 382–7. Henry TB, McPherson JT, Rogers ED, Heah TP, Hawkins SA, Layton AC, Sayler GS. (2009). Changes in the relative expression pattern of multiple vitellogenin genes in adult male and larval zebrafish exposed to exogenous estrogens. Comp Biochem Physiol A Mol Integr Physiol, 154, 119–26. Henry TR, Spitsbergen JM, Hornung MW, Abnet CC, Peterson RE. (1997). Early life stage toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in zebrafish (Danio rerio). Toxicol Appl Pharmacol, 142, 56–68. Henshel DS, Sparks DW, Simon TP, Tosick MJ. (2006). Age structure and growth of Semotilus atromaculatus (Mitchill) in PCB-contaminated streams. J Fish Biol, 68, 44–62. Hestermann EV, Stegeman JJ, Hahn ME. (2000). Relative contributions of affinity and intrinsic efficacy to aryl hydrocarbon receptor ligand potency. Toxicol Appl Pharmacol, 168, 160–72. Hill AJ, Bello SM, Prasch AL, Peterson RE, Heideman W. (2004). Water permeability and TCDD-induced edema in zebrafish early-life stages. Toxicol Sci, 78, 78–87. Hinck JE, Blazer VS, Denslow ND, Echols KR, Gale RW, Wieser C, et al. (2008). Chemical contaminants, health indicators, and reproductive biomarker responses in fish from rivers in the Southeastern United States. Sci Total Environ, 390, 538–57. Hinck JE, Blazer VS, Denslow ND, Echols KR, Gross TS, May TW, et al. (2007a). Chemical contaminants, health indicators, and reproductive biomarker responses in fish from the Colorado River and its tributaries. Sci Total Environ, 378, 376–402. Hinck JE, Blazer VS, Denslow ND, Myers MS, Gross TS, Tillitt DE. (2007b). Biomarkers of contaminant exposure in northern pike (Esox lucius) from the Yukon River Basin, Alaska. Arch Environ Contam Toxicol, 52, 549–62. Hinck JE, Blazer VS, Schmitt CJ, Papoulias DM, Tillitt DE. (2009a). Widespread occurrence of intersex in black basses (Micropterus spp.) from US rivers, 1995–2004. Aquat Toxicol, 95, 60–70. Hinck JE, Norstrom RJ, Orazio CE, Schmitt CJ, Tillitt DE. (2009b). Persistence of organochlorine chemical residues in fish from the Tombigbee River (Alabama, USA): continuing risk to wildlife from a former DDT manufacturing facility. Environ Poll, 157, 582–91. Hinck JE, Schmitt CJ, Blazer VS, Denslow ND, Bartish TM, Anderson PJ, et al. (2006a). Environmental contaminants and biomarker responses in fish from the Columbia River and its tributaries: spatial and temporal trends. Sci Total Environ, 366, 549–78. Hinck JE, Schmitt CJ, Chojnacki KA, Tillitt DE. (2009c). Environmental contaminants in freshwater fish and their risk to piscivorous wildlife based on a national monitoring program. Environ Mon Assess, 152, 469–94. Hinck JE, Schmitt CJ, Echols KR, May TW, Orazio CE, Tillitt DE. (2006b). Environmental contaminants in fish and their associated risk to piscivorous wildlife in the Yukon River Basin, Alaska. Arch Environ Contam Toxicol, 51, 661–72. Hinck JE, Schmitt CJ, Ellersieck MR, Tillitt DE. (2008). Relations between and among contaminant concentrations and biomarkers in black bass (Micropterus spp.) and common carp (Cyprinus carpio) from large US rivers, 1995–2004. J Environ Mon, 10, 1499–518. Hiramatsu N, Matsubara T, Fujita T, Sullivan CV, Hara A. (2006). Multiple piscine vitellogenins: biomarkers of fish exposure to estrogenic endocrine disruptors in aquatic environments. Mar Biol, 149, 35–47. Ho CH, Liu SM. (2011). Effect of coplanar PCB concentration on dechlorinating microbial communities and dechlorination in estuarine sediments. Chemosphere, 82, 48–55. Howell NL, Suarez MP, Rifai HS, Koenig L. (2008). Concentrations of polychlorinated biphenyls (PCBs) in water, sediment, and aquatic biota in the Houston Ship Channel, Texas. Chemosphere, 70, 593–606. Hughes AS, Vanbriesen JM, Small MJ. (2010). Identification of structural properties associated with polychlorinated biphenyl dechlorination processes. Environ Sci Technol, 44, 2842–48.

PCBs in wild fish

17

Imamoglu I, Li K, Christensen ER. (2002). Modeling polychlorinated biphenyl congener patterns and dechlorination in dated sediments from the Ashtabula River, Ohio, USA. Environ Toxicol Chem, 21, 2283–91. Imm P, Knobeloch L, Anderson HA, Great Lakes Sport Fish Consortium. (2005). Fish consumption and advisory awareness in the Great Lakes Basin. Environ Health Perspect, 113, 1325–9. Isin E, Guengerich FP. (2007). Complex reactions catalysed by cytochrome P450 enzymes. Biochim Biophys Acta 1770, 314–29. Iwanowicz LR, Blazer VS, McCormick SD, VanVeld PA, Ottinger CA. (2009). Aroclor 1248 exposure leads to immunomodulation, decreased disease resistance and endocrine disruption in the brown bullhead, Ameiurus nebulosus. Aquat Toxicol, 93, 70–82. Iwanowicz LR, Lerner DT, Blazer VS, McCormick SD. (2005). Aqueous exposure to Aroclor 1254 modulates the mitogenic response of Atlantic salmon anterior kidney T-cells: indications of short- and long-term immunomodulation. Aquat Toxicol, 72, 305–14. Jobling S, Beresford N, Nolan M, Rodgers-Gray T, Brighty GC, Sumpter JP, Tyler CR. (2002). Altered sexual maturation and gamete production in wild roach (Rutilus rutilus) living in rivers that receive treated sewage effluents. Biol Reprod, 66, 272–81. Jorgensen EH, Foshaug H, Andersson P, Burkow IC, Jobling M. (2002). Polychlorinated biphenyl toxicokinetics and P4501A responses in anadromous Arctic charr during winter emaciation. Environ Toxicol Chem, 21, 1745–52. Jorgensen EH, Vijayan MM, Killie JEA, Aluru N, Aas-Hansen O, Maule A. (2006). Toxicokinetics and effects of PCBs in Arctic fish: a review of studies on Arctic charr. J Toxicol Environ Health-Part A-Curr Iss, 69, 37–52. Kidd KA, Blanchfield PJ, Mills KH, Palace VP, Evans RE, Lazorchak JM, Flick RW. (2007). Collapse of a fish population after exposure to a synthetic estrogen. Proc Nat Acad Sci U S A, 104, 8897–901. Kim J, Rhee GY. (1997). Population dynamics of polychlorinated biphenyl-dechlorinating microorganisms in contaminated sediments. Appl Environ Microbiol, 63, 1771–76. Kopecka-Pilarczyk J, Correia AD. (2009). Biochemical response in gilthead seabream (Sparus aurata) to in vivo exposure to a mix of selected PAHs. Ecotoxicol Environ Saf, 72, 1296–302. Koponen K, Lindstrom-Seppa P, Kukkonen JVK. (2000). Accumulation pattern and biotransformation enzyme induction in rainbow trout embryos exposed to sublethal aqueous concentrations of 3,3’,4,4’tetrachlorobiphenyl. Chemosphere, 40, 245–53. Kutz FW, Wood PH, Bottimore DP. (1991). Organochlorine pesticides and polychlorinated-biphenyls in human adipose tissue. Rev Environl Contam Toxicol, 120, 1–82. Lake JL, Pruell RJ, Osterman FA. (1992). An examination of dechlorination processes and pathways in New Bedford Harbor sediments. Mar Environ Res, 33, 31–47. Larssen T, Duan L, Mulder J. (2011). Deposition and leaching of sulfur, nitrogen and calcium in four forested catchments in China: implications for acidification. Environ Sci Technol, 45, 1192–8. Larsson M, van den Berg M, Brenerova P, van Duursen MB, van Ede KI, Lohr C, et al. (2015). Consensus toxicity factors for PCDDs, PCDFs and PCBs combining in silico models and extensive in vitro screening of AhR-mediated effects in human and rodent cells. Chem Res Toxicol. doi: 10.1021/tx500434j Layton AC, Sanseverino J, Gregory BW, Easter JP, Sayler GS, Schultz TW. (2002). In vitro estrogen receptor binding of PCBs: measured activity and detection of hydroxylated metabolites in a recombinant yeast assay. Toxicol Appl Pharmacol, 180, 157–63. Letcher RJ, Bustnes JO, Dietz R, Jenssen BM, Jorgensen EH, Sonne C, et al. (2010). Exposure and effects assessment of persistent organohalogen contaminants in arctic wildlife and fish. Sci Total Environ, 408, 2995–3043. Li HX, Drouillard KG, Bennett E, Haffner GD, Letcher RJ. (2003). Plasmaassociated halogenated phenolic contaminants in benthic and pelagic fish species from the Detroit River. Environ Sci Technol, 37, 832–9. Li M, Wang X, Zhu J, Zhu S, Hu X, Zhu C, et al. (2014). Toxic effects of polychlorinated biphenyls on cardiac development in zebrafish. Mol Biol Rep, doi: 10.1007/s11033–014-3692–6 Lopes C, Perga ME, Peretti A, Roger MC, Persat H, Babut M. (2011). Is PCBs concentration variability between and within freshwater fish species explained by their contamination pathways? Chemosphere, 85, 502–8. Ma Q. (2001). Induction of CYP1A1. The AhR/DRE paradigm: transcription, receptor regulation, and expanding biological roles. Curr Drug Metabol, 2, 149–64.


18

T. B. Henry

Mac MJ, Schwartz TR, Edsall CC, Frank AM. (1993). Polychlorinatedbiphenyls in Great Lakes lake trout and their eggs–relations to survival and congener composition 1979–1988. J Great Lakes Res, 19, 752– 65. Maervoet J, Covaci A, Schepens P, Sandau CD, Letcher RJ. (2004). A reassessment of the nomenclature of polychlorinated biphenyl (PCB) metabolites. Environ Health Perspect, 112, 291–4. Martin M, Richardson BJ, Lam PKS. (2003). Harmonisation of polychlorinated biphenyl (PCB) analyses for ecotoxicological interpretations of southeast Asian environmental media: what’s the problem? Mar Poll Bull, 46, 159–70. Marvin CH, Painter S, Charlton MN, Fox ME, Thiessen PAL. (2004). Trends in spatial and temporal levels of persistent organic pollutants in Lake Erie sediments. Chemosphere, 54, 33–40. Maule AG, Jorgensen EH, Vijayan MM, Killie JE, 2005. Aroclor 1254 exposure reduces disease resistance and innate immune responses in fasted Arctic charr. Environ Toxicol Chem, 24, 117–24. McElroy JA, Kanarek MS, Trentham-Dietz A, Robert SA, Hampton JM, Newcomb PA, et al. (2004). Potential exposure to PCBs, DDT and PBDEs from sport-caught fish consumption in relation to breast cancer risk in Wisconsin. Environ Health Perspect, 112, 156–62. Meador JP, Ylitalo GM, Sommers FC, Boyd DT. (2010). Bioaccumulation of polychlorinated biphenyls in juvenile chinook salmon (Oncorhynchus tshawytscha) outmigrating through a contaminated urban estuary: dynamics and application. Ecotoxicology, 19, 141–52. Miracle A, Ankley G, Lattier D. (2006). Expression of two vitellogenin genes (vg1 and vg3) in fathead minnow (Pimephales promelas) liver in response to exposure to steroidal estrogens and androgens. Ecotoxicol Environ Saf, 63, 337–42. Monosson E, Fleming WJ, Sullivan CV. (1994). Effects of the planar PCB 3,3’4,4’-tetrachlorobiphenyl (TCB) on ovarian development, plasmalevels of sex steroid hormones and vitellogenin, and progeny survival in the white perch (Morone americana). Aquat Toxicol, 29, 1–19. Muir D, Sverko E. (2006). Analytical methods for PCBs and organochlorine pesticides in environmental monitoring and surveillance: a critical appraisal. Anal Bioanal Chem, 386, 769–89. Na YR, Seok SH, Baek MW, Lee HY, Kim DJ, Park SH, et al. (2009). Protective effects of vitamin E against 3,3 ‘,4,4 ‘,5-pentachlorobiphenyl (PCB126) induced toxicity in zebrafish embryos. Ecotoxicol Environ Saf, 72, 714–19. Nacci D, Huber M, Champlin D, Jayaraman S, Cohen S, Gauger E, et al. (2009). Evolution of tolerance to PCBs and susceptibility to a bacterial pathogen (Vibrio harveyi) in Atlantic killifish (Fundulus heteroclitus) from New Bedford (MA, USA) harbor. Environ Poll, 157, 857–64. Nacci DE, Champlin D, Jayaraman S. (2010). Adaptation of the Estuarine Fish Fundulus heteroclitus (Atlantic Killifish) to Polychlorinated Biphenyls (PCBs). Estuar Coast, 33, 853–64. Nakayama K, Oshima Y, Hiramatsu K, Honjo T. (2004). Alteration of general behavior of male medaka, Oryzias latipes, exposed to tributyltin and/or polychlorinated biphenyls. J Fac Agr Kyushu Univ, 49, 85–92. Nebert DW, Puga A, Vasiliou V. (1993). Role of the Ah receptor and the dioxin-inducbile Ah gene battery in toxcity, cancer, and signaltransduction. Ann N Y Acad Sci, 685, 624–40. Olsson PE, Westerlund L, Teh SJ, Billsson K, Berg AH, Tysklind M, et al. (1999). Effects of maternal exposure to estrogen and PCB on different life stages of zebrafish (Danio rerio). Ambio, 28, 100–06. Parente TEM, Rebelo MF, da-Silva ML, Woodin BR, Goldstone JV, Bisch PM, et al. (2011). Structural features of cytochrome P450 1A associated with the absence of EROD activity in liver of the loricariid catfish Pterygoplichthys sp. Gene, 489, 111–8. Powell DB, Palm RC Jr, Skillman A, Godtfredsen K. (2003). Immunocompetence of juvenile Chinook salmon against Listonella anguillarum following dietary exposure to Aroclor 1254. Environ Toxicol Chem, 22, 285–95. Proestou DA, Flight P, Champlin D, Nacci D. (2014). Targeted approach to identify genetic loci associated with evolved dioxin tolerance in Atalntic killifish (Fundulus heteroclitus). BMC Evol Biol, 14, 7. http:www.biomedcentral.com/1471–2148/14/7 Quensen JF, Mueller SA, Jain MK, Tiedje JM. (1998). Reductive chlorination of DDE to DDMU in marine sediment microcosms. Science, 280, 722–4. Reiner EJ. (2010). The analysis of dioxins and related compounds. Mass Spectrom Rev, 29, 526–59. Reiner EJ, Clement RE, Okey AB, Marvin CH. (2006). Advances in analytical techniques for polychlorinated dibenzo-p-dioxins,


polychlorinated dibenzofurans and dioxin-like PCBs. Anal Bioanal Chem, 386, 791–806. Reiser DW, Greenberg ES, Helser TE, Branton M, Jenkins KD. (2004). In situ reproduction, abundance, and growth of young-of-year and adult largemouth bass in a population exposed to polychlorinated biphenyls. Environ Toxicol Chem, 23, 1762–73. Reitzel AM, Karchner SI, Franks DG, Evans BR, Nacci D, Champlin D, et al. (2014). (AHR) loci in populations of Atlantic killifish (Funduduls heteroclitus) inhabiting polluted and reference habitats. BMC Evolut Biol, 14. http://www.biomedcentral.com/1471–2148/14/6 Rogers ED, Henry TB, Twiner MJ, Gouffon JS, McPherson JT, Boyer GL, et al. (2011). Global gene expression profiling in larval zebrafish exposed to microcystin-LR and microcystis reveals endocrine disrupting effects of Cyanobacteria. Environ Sci Technol, 45, 1962–9. Roy NK, Courtenay S, Yuan ZP, Ikonomou M, Wirgin I. (2001). An evaluation of the etiology of reduced CYP1A1 messenger RNA expression in the Atlantic tomcod from the Hudson River, New York, USA, using reverse transcriptase polymerase chain reaction analysis. Environ Toxicol Chem, 20, 1022–30. Roy NK, Walker N, Chambers RC, Wirgin I. (2011). Characterization and expression of cytochrome P4501A in Atlantic sturgeon and shortnose sturgeon experimentally exposed to coplanar PCB 126 and TCDD. Aquat Toxicol, 104, 23–31. Russell RW, Lazar R, Haefner GD. (1995). Biomagnification of organochlorines in Lake Erie white bass. Environ Toxicol Chem, 14, 719–24. Rypel AL, Bayne DR. (2010). Do fish growth rates correlate with PCB body burdens? Environ Poll, 158, 2533–36. Sadraddini S, Azim ME, Shimoda Y, Bhavsar SP, Drouillard KG, Backus SM, Arhonditsis GB. (2011a). A Bayesian assessment of the PCB temporal trends in Lake Erie fish communities. J Great Lakes Res, 37, 507–20. Sadraddini S, Azim ME, Shimoda Y, Mahmood M, Bhavsar SP, Backus SM, Arhonditsis GB. (2011b). Temporal PCB and mercury trends in Lake Erie fish communities: a dynamic linear modeling analysis. Ecotoxicol Environ Saf, 74, 2203–14. Sanchez BC, Ralston-Hooper KJ, Kowalski KA, Inerowicz HD, Adamec J, Sepulveda MS. (2009). Liver proteome response of largemouth bass (Micropterus salmoides) exposed to several environmental contaminants: potential insights into biomarker development. Aquat Toxicol, 95, 52–9. Sarasquete C, Segner H. (2000). Cytochrome P4501A (CYP1A) in teleostean fishes. A review of immunohistochemical studies. Sci Total Environ, 247, 313–32. Schlezinger JJ, Stegeman JJ. (2001). Induction and suppression of cytochrome P450 1A by 3,3 ‘,4,4’,5-pentachlorobiphenyl and its relationship to oxidative stress in the marine fish scup (Stenotomus chrysops). Aquat Toxicol, 52, 101–15. Schlezinger JJ, Struntz WDJ, Goldstone JV, Stegeman JJ. (2006). Uncoupling of cytochrome P450 1A and stimulation of reactive oxygen species production by co-planar polychlorinated biphenyl congeners. Aquat Toxicol, 77, 422–32. Schnitzler JG, Thome JP, Lepage M, Das K. (2011). Organochlorine pesticides, polychlorinated biphenyls and trace elements in wild European sea bass (Dicentrarchus labrax) off European estuaries. Sci Total Environ, 409, 3680–86. Schulzbull DE, Petrick G, Kannan N, Duinker JC. (1995). Distribution of individual chlorobiphenyls (PCB) in solution and suspension in the Baltic Sea. Mar Chem, 48, 245–70. Schwarzenbach RP, Escher BI, Fenner K, Hofstetter TB, Johnson CA, von Gunten U, Wehrli B. (2006). The challenge of micropollutants in aquatic systems. Science, 313, 1072–77. Sinkkonen S, Paasivirta J. (2000). Degradation half-life times of PCDDs, PCDFs and PCBs for environmental fate modeling. Chemosphere, 40, 943–49. Smith DW. (1998). Evidence of strong lake trout recruitment despite high PCB (total Aroclors) concentrations. J Great Lakes Res, 24, 974–79. Sobek A, McLachlan MS, Borga K, Asplund L, Lundstedt-Enkel K, Polder A, Gustafsson O. (2010). A comparison of PCB bioaccumulation factors between an arctic and a temperate marine food web. Sci Total Environ, 408, 2753–60. Spitsbergen JM, Walker MK, Olson JR, Peterson RE. (1991). Pathological alterations in early life stages of lake trout, Salvelinus namaycush, exposed to 2,3,7,8-tetrachlorodibenzo-para-dioxin as fertilized eggs. Aquat Toxicol, 19, 41–71. Spitsbergen JM, Schat KA, Kleeman JM, Peterson RE. (1988). Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin or Aroclor 1254 on the


DOI 10.3109/10408444.2015.1038498

resistance of rainbow trout, Salmo gairdneri Richardson, to infectious haematopoietic necrosis virus. J Fish Dis, 11, 73–83. Stapleton HM, Letcher RJ, Baker JE. (2001). Metabolism of PCBs by the deepwater sculpin (Myoxocephalus thompsoni). Environ Sci Technol, 35, 4747–52. Stauffer TM. (1979). Effects of DDT and PCBs on survival of lake trout eggs and fry in a hatchery and in Lake Michigan, 1973–1976. Trans Am Fish Soc, 108, 178–86. Suh JH, Kang JS, Yang KH, Kaminski NE. (2003). Antagonism of aryl hydrocarbon receptor-dependent induction of CYP1A1 and inhibition of IgM expression by di-ortho-substituted polychlorinated biphenyls. Toxicol Appl Pharmacol, 187, 11–21. Tillitt DE, Cook PM, Giesy JP, Heideman W, Peterson RE. (2008). Reproductive impairment of Great Lakes lake trout by dioxin-like chemicals. In: Di Giulio RT, Hinton DE, Eds. The Toxicology of Fishes. Boca Raton, FL: CRC Press, Taylor & Francis Group. Ueno D, Darling C, Alaee M, Campbell L, Pacepavicius G, Teixeira C, Muir D. (2007). Detection of hydroxylated polychlorinated biphenyls (OH-PCBs) in the abiotic environment: surface water and precipitation from Ontario, Canada. Environ Sci Technol, 41, 1841–48. UNEP. (2012). United Nations Environment Program Annual Report 2011. Published February 2012. ISBN: 978-92-807-3244-3. US EPA. (2007). Hudson River PCB Site Baseline Monitoring Report, Data Summary Report (2004–2006 data), data provided to US EPA 30 March 2007. US EPA. (2008). Hudson River PCB Site Baseline Monitoring Report, Data Summary Report (2007 data), data provided to US EPA 31 March 2008. Valdehita A, Fernandez-Cruz ML, Torrent F, Sericano JL, Navas JM. (2012). Differences in the induction of cyp1A and related genes in cultured rainbow trout Oncorhynchus mykiss. Additional considerations for the use of EROD activity as a biomarker. J Fish Biol, 81, 270–87. van den Berg M, Birnbaum L, Bosveld ATC, Brunstrom B, Cook P, Feeley M, et al. (1998). Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and wildlife. Environ Health Perspect, 106, 775–92. van den Berg M, Denison MS, Birnbaum LS, DeVito MJ, Fiedler H, Falandysz J, et al. (2013). Polybrominated dibenzop-dioxins, dibenozfurans, and biphenyls: inclusion in the toxicity equivalency factor concept for dioxin-like compounds. Toxicol Sci, 133, 197–208. van der Heijden SA, Jonker MTO. (2011). Intra- and Interspecies Variation in Bioconcentration Potential of Polychlorinated Biphenyls: are all lipids equal? Environ Sci Technol, 45, 10408–14. Van der Oost R, Beyer J, Vermeulen NPE. (2003). Fish bioaccumulation an biomarkers in environmental risk assessment: a review. Environ Toxicol Pharmacol, 13, 57–149. Vasilyeva GK, Strijakova ER. (2007). Bioremediation of soils and sediments contaminated by polychlorinated biphenyls. Microbiol, 76, 639–53. Walker MK, Peterson RE. (1991). Potencies of polychlorinated dibenzop-dioxin, dibenzofuran, and biphenyl congeners, relative to 2,3,7,8tetrachlorodibenzo-p-dioxin, for producing early life stage mortality in rainbow trout (Oncorhynchus mykiss). Aquat Toxicol, 21, 219–38. Walters DM, Mills MA, Cade BS, Burkard LP. (2011). Trophic Magnification of PCBs and Its Relationship to the Octanol-Water Partition Coefficient. Environ Sci Technol, 45, 3917–24.

PCBs in wild fish

19

White RD, Hahn ME, Lockhart WL, Stegeman JJ. (1994). Catalytic and immunochemical characterization of hepatic-microsomal cytochromes P450 in beluga whale (Delphinapterus leucas). Toxicol Appl Pharmacol, 126, 45–57. White RD, Shea D, Stegeman JJ. (1997). Metabolism of the aryl hydrocarbon receptor agonist 3,3′ 4,4′ tetrachlorobiphenyl by the marine fish scup (Stenotomus chrysops) in vivo and in vitro. Drug Metab Dispos, 25, 564–72. Whitehead A, Pilcher W, Champlin D, Nacci D. (2012). Common mechanism underlies repeated evolution of extreme pollution tolerance. Proc Royal Soc B 279, 427–33. Whyte JJ, Jung RE, Schmitt CJ, Tillitt DE. (2000). EthoxyresorufinO-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit Rev Toxicol, 30, 347–570. Wiegel J, Wu QZ. (2000). Microbial reductive dehalogenation of polychlorinated biphenyls. Fems Microbiol Ecol, 32, 1–15. Williams TD, Diab A, Ortega F, Sabine VS, Godfrey RE, Falciani F, et al. (2008). Transcriptomic responses of European flounder (Platichthys flesus) to model toxicants. Aquat Toxicol, 90, 83–91. Wirgin I, Roy NK, Loftus M, Chambers RC, Franks DG, Hahn ME. (2011). Mechanistic basis of resistance to PCBs in atlantic tomcod from the Hudson River. Science, 331, 1322–5. Wu QZ, Bedard DL, Wiegel J. (1996). Influence of incubation temperature on the microbial reductive dechlorination of 2,3,4,6-tetrachlorobiphenyl in two freshwater sediments. Appl Environ Microbiol, 62, 4174–9. Yuan ZP, Courtenay S, Chambers RC, Wirgin I. (2006a). Evidence of spatially extensive resistance to PCBs in an anadromous fish of the Hudson River. Environ Health Perspect, 114, 77–84. Yuan ZP, Courtenay S, Wirgin I. (2006b). Comparison of hepatic and extra hepatic induction of cytochrome P4501A by graded doses of aryl hydrocarbon receptor agonists in Atlantic tomcod from two populations. Aquat Toxicol, 76, 306–20. Yuan ZP, Wirgin M, Courtenay S, Ikonomou M, Wirgin I. (2001). Is hepatic cytochrome P4501A1 expression predictive of hepatic burdens of dioxins, furans, and PCBs in Atlantic tomcod from the Hudson River estuary? Aquat Toxicol, 54, 217–30. Yunker MB, Ikonomou MG, Sather PJ, Friesen EN, Higgs DA, Dubetz C. (2011). Development and validation of protocols to differentiate PCB patterns between farmed and wild salmon. Environ Sci Technol, 45, 2107–15. Zabel EW, Cook PM, Peterson RE. (1995a). Potency of 3,3′ 4,4′ 5-pentachlorobiphenyl (PCB-126), alone and in combination with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), to produce late trout early life stage mortality. Environ Toxicol Chem, 14, 2175–9. Zabel EW, Cook PM, Peterson RE. (1995b). Toxic equivalency factors of polychlorinated dibenzo-p-dioxin, dibenzofuran and biphenyl congeners based on early-life stage mortality in rainbow trout (Oncorhynchus mykiss). Aquat Toxicol, 31, 315–28. Zabel EW, Walker MK, Hornung MW, Clayton MK, Peterson RE. (1995c). Interactions of polychlorinated dibenzo-p-dioxin, dibenzofuran, and biphenyl congeners for producing rainbow trout early-life stage mortality. Toxicol App Pharmacol, 134, 204–13. Zint MT, Taylor WW, Carl L, Edsall CC, Heinrich J, Sippel A, et al. (1995). Do toxic substances pose a threat to rehabilitation of lake trout in the Great Lakes? A review of the literature. J Great Lakes Res, 21, 530–46. Zlokovitz ER, Secor DH. (1999). Effect of habitat use on PCB body burden in Hudson River striped bass (Morone saxatilis). Can J Fish Aquat Sci, 56, 86–93.

Toxicokinetics of chiral polychlorinated biphenyls across different species--a review.

A Critical Review about Human Exposure to Polychlorinated Dibenzo-p-Dioxins (PCDDs), Polychlorinated Dibenzofurans (PCDFs) and Polychlorinated Biphenyls (PCBs) through Foods.

Chiral polychlorinated biphenyls: absorption, metabolism and excretion--a review.

Goitres in rats fed polychlorinated biphenyls.

Polychlorinated biphenyls in Baltimore Harbor sediments.

Effects of polychlorinated biphenyls on poultry reproduction.

Mortality among workers exposed to polychlorinated biphenyls.

Polychlorinated biphenyls and links to cardiovascular disease.

Polychlorinated biphenyls: estimated serum half lives.

Toxic effects of polychlorinated biphenyls on cardiac development in zebrafish.

Immunoassay monitoring of polychlorinated biphenyls (PCBs) in the Great Lakes.

Alterations in drug metabolism in workers exposed to polychlorinated biphenyls.

Polychlorinated biphenyls in plankton from the Turku Archipelago.

Current progress in the determination of the polychlorinated biphenyls.

Polychlorinated dibenzo-p-dioxins and dibenzofurans and polychlorinated biphenyls in surface soil from the Tibetan Plateau.

Photochemical confirmation of mirex in the presence of polychlorinated biphenyls.

Polychlorinated biphenyls in residential dust: sources of variability.

Abundance and distribution of polychlorinated biphenyls (PCBs) in breast tissue.

Investigation of isomer specific polychlorinated biphenyls in printing inks.

In vitro effects of polychlorinated biphenyls on human platelets.

Polychlorinated biphenyls (PCBs) in bottom sediments: identification of sources.

Influence of polychlorinated dibenzofurans (PCDFs) and polychlorinated biphenyls (PCBs) to serum protein components in rats.

Emissions of polychlorinated biphenyls, polychlorinated dibenzo-p-dioxins, and polychlorinated dibenzofurans during 2010 and 2011 in Zurich, Switzerland.

Polychlorinated biphenyls (PCB's) in the rivers Avon and Frome.