Atlantic Tomcod Microgadus tomcod: A Model Species for the ... · exposures to chemical contaminants. First, the presence of a contaminant in an ecosys-tem is not synonymous with

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Atlantic Tomcod Microgadus tomcod: A ModelSpecies for the Responses of Hudson River Fish to

Toxicants

ISAAC WIRGIN

New York University, School of Medicine, Department of Environmental Medicine57 Old Forge Road, Tuxedo, New York 10987 USA

[email protected]

R. CHRISTOPHER CHAMBERS

Howard Marine Sciences Laboratory, Northeast Fisheries Science Center, NOAA Fisheries Service74 Magruder Road, Highlands, New Jersey 07732 USA

Abstract.—Despite recent successes in eliminating or reducing many pointsources of chemical contaminants, sediments in the Hudson River Estuary arestill highly contaminated with lipophilic and highly persistent polychlorinatedbiphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinateddibenzofurans (PCDFs), and polycyclic aromatic hydrocarbons (PAHs). Thesehave been shown to bioaccumulate to high levels in resource species and otherkey ecological components of the Hudson River food web. Resource managersand stewards must consider the possible toxic effects of these pollutants on theHudson River biota, including its fish community; however, few studies havedirectly investigated these effects. A series of toxicological studies on Atlantictomcod Microgadus tomcod from the Hudson River Estuary have demonstratedprofound and broad-based changes in response to local contaminants. Levels ofcontaminants in the tissues of different life stages of tomcod from the HudsonRiver Estuary far exceed those in tomcod from other Atlantic Coast estuaries.More importantly, a combination of field and laboratory studies has demonstratedmolecular to population level perturbations in tomcod from the Hudson River,all of which are consistent with chemical exposures and many of which appear tobe mechanistically linked. These effects include induction of hepatic expressionof cytochrome P4501A1 mRNA, high levels of hepatic DNA damage, somaticmutations at an oncogene locus critical to the initiation of chemicalcarcinogenesis, elevated prevalence of gross and histologically defined hepatictumors, truncated age structure, and dramatic resistance at the molecular andorganismal levels to halogenated aromatic hydrocarbons (HAHs). Resistance, apopulation level effect, was observed in the toxic responses of tomcod embryosand larvae to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and TCDD-like PCBs,but not PAHs. Because young of the year tomcod are a critical node in theHudson River food chain, their evolved resistance to HAHs and high body burdenof these and related contaminants has likely resulted in the trophic transfer ofthese contaminants to secondary and tertiary consumers of the Hudson River,

American Fisheries Society Symposium 51:331–365, 2006© 2006 by the American Fisheries Society

332 WIRGIN AND CHAMBERS

including important resource species, and an elevated tissue burden of thesecontaminants in those consumers. In total, these studies are consistent with thehypothesis that exposure to Hudson River-borne contaminants has significantlyaltered its tomcod population and perhaps evoked broad change in the HudsonRiver fish community.

Introduction

The assessment of risk from chemical con-tamination to fish and other populations inthe Hudson River Estuary is currently ofgreat interest to New York State and UnitedStates federal resource stewards such as theNational Oceanic and Atmospheric Admin-istration (NOAA) Fisheries and the U.S.Fish and Wildlife Service. One responsibil-ity of these agencies is to assess potentialecological damage to fish populations fromexposure to xenobiotics and, if damage hasoccurred, to assign a monetary value to thoseperturbations. According to the Departmentof Interior Natural Resource Damage As-sessment (NRDA) regulations, measures ofreduced viability to fish populations couldinclude mutations, liver and skin cancers,physiological abnormalities, skeletal defor-mities, histopathological lesions, fin erosion,impaired reproductive success, decreasedability to capture prey and avoid predators,and death [43 CFR 11.62(f)(1)(i)]. Thesebiological endpoints could be evaluated bydirect observations of affected fish in na-ture, by assays of in situ caged fish, or bytoxicity testing of fish in the laboratory [43CFR 11.62 (f)(2)(i-iv)]. In order to confi-dently link these changes in viability due toa xenobiotic of interest, a toxic response infree-ranging fish and in fish exposed undercontrolled laboratory conditions should bedemonstrated. These last criteria, whichcombine field and laboratory studies, forma far more convincing case for ecologicaldamage to natural populations from chemi-cal exposure than implied by the demon-stration of bioaccumulation of contaminantsalone. In this review, we marshal evidence

from a variety of levels of biological organi-zation that support our view of Atlantictomcod Microgadus tomcod from the HudsonRiver as a sensitive investigative model foridentifying and evaluating the complex re-sponses of fishes to environmental contami-nants.

Chemical Contaminants of Concernin the Hudson River Estuary

The extent of risk to fish or other taxa fromchemical pollutants depends on the iden-tity of the specific chemical; the route, dose,and duration of exposure; and its toxicityin the target species. Uncertainties in as-signing risks from exposure can be due tospecies, population, and individual differ-ences in susceptibilities to contaminants.This may result from variation in thechemical’s bioavailability, its uptake orefflux, or its toxicity at the target receptor.The fish community in the Hudson RiverEstuary has been chronically exposed formore than half a century to high and some-times extreme levels of chemical contami-nants (Wirgin and Waldman 1998). Thesechemical contaminants include organic andinorganic pollutants. Some of the organicpollutants of greatest concern for ecologicalhealth consequences include polychlorinatedbiphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinateddibenzofurans (PCDFs), polycyclic aromatichydrocarbons (PAHs), pesticides, and vari-ous estrogen-mimicking endocrine disruptors(Wirgin et al. 2006). Among the metal pol-lutants of greatest concern within the Estu-ary because of their concentrations or known

333RESPONSES OF HUDSON RIVER FISH TO TOXICANTS

toxicities are cadmium (Cd), mercury (Hg),chromium (Cr), nickel (Ni), and arsenic (As).The high sediment loads of these contami-nants led to the designation of five U.S.federal Superfund sites in the estuary. ThePCBs Superfund site, the longest such sitein the nation, is the entire 200 mi of theHudson River from Hudson Falls, New Yorkto the Battery at the southern tip of Man-

hattan (Figure 1). The Diamond Alkali2 ,3 ,7 ,8-tetrachlorodibenzo-p -d ioxin(TCDD)Superfund site on the Passaic Riverresulted from chemical releases from a singleherbicide manufacturing facility and theirsubsequent transport to much of downstreamNewark Bay. In this paper, we focus on thebiological consequences of exposure to PCBs,PCDD/Fs, and PAHs.

Figure 1. A map of the Hudson River Estuary showing the locations mentioned in the text and three U.S.federal Superfund sites.


PCBs, PCDD/Fs, and PAHs all are aromatic(ring structure) hydrocarbon (AH) compounds,however, because PCBs and PCDD/Fs havechlorine substitutions, they are often termedhalogenated aromatic hydrocarbons (HAHs).AHs are regarded as generally lipophilic andhydrophobic in character and tend to accu-mulate in the sediments of aquatic systemswhere they become bioavailable to benthicorganisms. Chlorine dramatically alters thechemical properties of HAHs compared toPAHs by making HAHs relatively more re-sistant to biological degradation and hencemore likely to biomagnify in the food chainand to persist in the environment. IndividualPCBs and PCDD/Fs that differ in the num-ber and position of their chlorine substituentsare termed congeners. In fishes and other ver-tebrates, individual congeners usually differin their toxicities due to their different abili-ties to bind the aryl hydrocarbon receptor(AHR) at the cellular level. Usually, the greaterthe affinity of a compound to the AHR, thegreater is its toxicity. TCDD, which has thegreatest affinity for AHR among all AHs, alsohas the greatest toxicity. Among PCBs, thosethat are coplanar are able to most efficientlybind AHR and are thus usually the mosttoxic. Knowledge of the HAH congener com-position in contaminated environmental ma-trices (water, sediment, biota) enables moreaccurate evaluation of the toxicities of thesematrices. Individual PAH compounds alsodiffer in structure from one another by con-taining variable numbers (three or more) andconfigurations of benzene rings and sometimesby the presence of non-halogenated additions.These differences also result in different tox-icities that are mediated by AHR binding.

Problems in Evaluating Toxicity ofChemical Contaminants to Natural

Populations

Several issues complicate assessments of thedamage to natural populations caused by

exposures to chemical contaminants. First,the presence of a contaminant in an ecosys-tem is not synonymous with itsbioavailability to any given species. Thus,bioavailability must be empirically verifiedfor the species or tissue of interest. Second,there are large variations among and oftenwithin taxa in their biological responses totoxic chemicals. This reality also compli-cates the selection of an appropriate modelspecies with which to evaluate these pro-cesses. Third, a wide range of responses—from the sublethal to lethal and expressedat levels from the molecular, cellular, histo-logical, physiological, behavioral, organismal,and population to community endpoints—could be evaluated. Endpoints at lower lev-els of biological organization may be moresensitive, dose-responsive, specific to indi-vidual or classes of contaminants, and nearlyimmediate in their response. Those at higherlevels may be less sensitive, less chemical-specific, with substantial lag times before afull response is exhibited. Yet these higher-order endpoints may be of more direct, ifnot greater, ecological relevance. Fourth, be-cause there is usually large interindividualvariation in toxic responses within naturalpopulations, large numbers of specimensmust be analyzed to ensure statistical con-fidence in the results. Fifth, chemical pol-lutants almost always co-occur in stressedurban or industrialized ecosystems. Formany of the contaminants of concern, tox-icities of individual contaminants have beendefined in model fish species under con-trolled laboratory conditions. However, be-cause of the typical co-occurrence of con-taminants in the environment and in bio-logical receptor species, predicting the tox-icities of these contaminants in nature willbe very difficult. These chemicals may ex-hibit additive toxicities, but often their com-bined toxicities are synergistic or antago-nistic. Thus, using laboratory testing ofsingle compounds as a benchmark to evalu-ate toxicity in fish exposed in nature may


over- or underestimate realized toxic effects.Sixth, individuals from chronically andhighly exposed environments may becomeresistant to the effects of chemical contami-nants. This outcome not only makes it dif-ficult to interpret the results from field sur-veys, but it also requires that some account-ing be made for the costs associated withevolutionary resistance. While resistance maybe beneficial in the short term to the af-fected population, it may hypersensitize in-dividuals in the population to other stres-sors, negatively impact its viability after siteremediation, and result in it serving as ameans of transferring contamination tohigher trophic levels. Because the endpointis ecological rather than immediate, andbecause the link to ecosystem functions—including human health—is complex andindirect, long-term, well-funded investiga-tions are necessary in order to fairly ap-praise these effects.

Tissue Burdens of Contaminants inHudson River Fish Populations

Water and sediment concentrations of thesecontaminants are routinely monitored andtheir burdens in tissues have been deter-mined in some Hudson River fish popula-tions. Initial analysis focused on commer-cially or recreationally important resourcespecies, most often striped bass Moronesaxatilis. Because of the wide-ranging andseasonal movements of striped bass withinthe river and along the coast for spawningand feeding, especially in older age classes,the tissue burden of contaminants in thisspecies provides a picture of spatially inte-grated contaminant levels from much of thetidal estuary and beyond (Sloan et al. 1995;Farley and Thomann 1998). Additionally,interest has also centered on species fromlocales within the Hudson River ecosystemwith the greatest known concentrations ofindividual contaminants, such as brown

bullhead Ameiurus nebulosus, goldfishCarassius auratus, pumpkinseed Lepomisgibbosus, and yellow perch Perca flavescens(Armstrong and Sloan 1988) from the mainstem of the Hudson River at Stillwater, ap-proximately 25 mi south of Ft. Edward,below the Federal Dam at Troy-Albany, and85 mi south of Ft. Edward at Catskill, andin Atlantic tomcod (Yuan et al. 2001), stripedbass, and bluefish Pomatomus saltatrix(Skinner et al. 1996, 1997) from the NewarkBay complex, which comprises the westernestuary. Generally, tissue burdens of PCBsare high in fish from the upper HudsonRiver (above the Troy Dam) (Sloan et al.2002). Often, but not always, the level ofPCBs burdens in fish tissue correspondsdirectly with distance from the PCBs pointsources with some punctuation associatedwith secondary inputs of PCBs in the NewYork City area (Farley and Thomann 1998).Long-term trends consist of a very rapidinitial decline in PCBs burdens followingthe cessation of release of PCBs in 1977,followed by a much slower decrease overthe past two decades with several rapid,short-term increases associated with episodicinputs from point sources (Sloan et al. 2005).Unfortunately, these analyses of PCBs wereusually reported on an Aroclor (commer-cially available PCBs mixtures manufac-tured by Monsanto), not on a congener spe-cific basis, so it is difficult to evaluate thepotential toxicity of the specific tissue bur-dens to Hudson River populations.

Fish from Newark Bay and its associatedtributaries have exhibited very high levelsof total PCDD/Fs and PCBs. PCDD con-gener patterns were dominated by the mosttoxic congener TCDD. Levels of TCDD infish and shellfish from Newark Bay were attimes the highest ever reported in the lit-erature (Rappe et al. 1991; Yuan et al. 2001).Most, if not all, of the TCDD in NewarkBay and its tributaries originated from asingle herbicide manufacturing plant along


the shores of the Passaic River, which pro-duced chlorinated phenols from the late1940s until 1969. It has been estimated that4–8 kg of TCDD have been deposited inNewark Bay from releases from the PassaicRiver facility (Bopp et al. 1991).

Ecological Damage to HudsonRiver Fish Populations

Evidence of fish die-offs due to chemicals israre but death due to any cause in the earlylife stages of fishes is rarely observed innature. When such evidence does exist, it iscertainly difficult to confidently ascribe itto a source. However, Cook et al. (2003)combined results of laboratory testing ofyoung life stages of lake trout Salvelinusnamaycush with retrospective analysis ofTCDD-like chemicals in sediments andmodeling of sediment bioaccumulation fac-tors for TCDD-like chemicals in eggs, toconclude that environmental levels of TCDDfrom the 1930s through 1950s were suffi-ciently high in Lake Ontario to prohibitsuccessful natural reproduction of lake troutand likely resulted in their extirpation fromthat lake by 1960.

There is no evidence of direct, chemicallyinduced lethality to fish in the Hudson Riveryet experimental results demonstrating agenetic basis to resistance in some HudsonRiver ichthyofauna (discussed below) wouldimplicate chemically mediated lethality inthe past. Nevertheless, it is noteworthy thatspecies diversity in the Hudson River is un-usually high for a temperate estuary (Smithand Lake 1990), reflecting a combination ofresident and seasonally marine migrant ma-rine species from more southern climes aswell as a large number of invasive species(Daniels et al. 2005). Extirpation of nativeHudson River species is unknown with theprobable exception of rainbow smeltOsmerus mordax which may be due to glo-

bal warming (J. Waldman, Queens College,personal communication). Thus, evaluation oftoxicities in nature is likely to require atten-tion to more subtle sublethal effects.

Heightened interest in toxicity of chemicals inimpacted ecosystems often follows reports ofreduced abundance in economically or ecologi-cally important species. In the mid-1970s,populations of striped bass along the Atlanticcoast, including the Hudson River, declinedto near historic low levels. Contaminants inthe Hudson River were viewed as a factor thatmay have contributed to these radical declinesin population size (Mehrle et al. 1982; Westinet al. 1983). Extensive evaluation of the toxiceffects of a variety of chemicals, including PCBs,in young as well as adult striped bass, showedthat although PCBs may reduce survivorshipof larvae (Mehrle and Ludke 1983; Westin etal. 1985), they were not responsible for thepopulation’s decline. Instead, overfishing wasimplicated as the important mechanism forpopulation reductions (Barnthouse et al. 2003)because severe limitations of harvests resultedin strong rebounds of striped bass adults alongmuch of the Atlantic coast (Striped Bass Tech-nical Committee 2003). Similarly, abundanceof shortnose sturgeon Acipenser brevirostrumand Atlantic sturgeon A. oxyrinchus in theEstuary declined to historically low levels overthe past 40 years, but overfishing rather thanchemical pollution is believed to be the causefor these declines as well.

One manifestation of possible chemical-inducedtoxicity in Hudson River fish populations wasthe elevated prevalence of pathological aberra-tions in brown bullhead from the upper reachesof the Hudson River near Griffin Island andStillwater at river mile (rm) 175 compared tothose collected near Corinth (rm 200), justupstream of PCBs point sources (Kim et al.1989; Bowser et al. 1990). Brown bullhead fromcontaminated sites exhibited 60-fold higherlevels of total PCBs in skeletal muscle, sig-nificantly higher splenic and renal


hemosiderous (pigmentation from hemoglo-bin degradation), and bile-duct hyperplasiathan fish from the less impacted upriversite. The high incidence of these patholo-gies is consistent with the role of liver andkidney in detoxifying chemical contaminants.In view of the susceptibility of brown bull-head to liver and skin cancer at other con-taminated sites (Baumann 1998), it is sur-prising that there are no reports of tumorsin this species from the upper Hudson River.Furthermore, the proportion of female tomale brown bullhead at the contaminatedsites (0.75) far exceeded that at the less con-taminated ones (0.55) suggesting the pos-sible impact of endocrine disrupting chemi-cals. However, the impacts of these histo-logical and demographic changes on popu-lation processes were not investigated fur-ther.

Epizootics of Neoplasia in NorthAmerican Fish Populations

In the early 1980s, reports surfaced of epi-zootics of neoplasia in fish populations fromcontaminated urbanized and industrializedestuaries and rivers in different regions inthe United States (reviewed in Wirgin andWaldman 1998). The most publicized of theepizootics and sites included English soleParophrys vetulus from Puget Sound, Wash-ington, winter flounder Pseudo-pleuronectesamericanus from Boston Harbor, Massachu-setts, brown bullhead from industrialized

tributaries of Lake Ontario (Baumann etal. 1987) and the Anacostia River, Wash-ington, D.C. (Pinkney et al. 2004), andAtlantic tomcod from the Hudson River(Dey et al. 1993). All of these affected taxahave benthic lifestyles as juveniles andadults. In almost all cases, liver or skintumors were observed. Chemical etiologieswere proposed for these epizootics becauseof the direct contact of skin with contami-nated environmental matrices, the role ofthe liver in the detoxification of chemicalcontaminants, the lipophilicity of most can-cer-causing xenobiotics, and the depositionof hydrophobic pollutants in the bottom

Figure 2. A: An adult one-year-old Atlantictomcod from the Hudson River. B: Liver from anadult tomcod from the Hudson River with severallarge hepatocellular carcinomas and a tumor massinvolving most of one of three liver lobes.

A

B


stratum. This explanation is consistent withthe finding of a significant decrease in thefrequency of liver tumors in a brown bull-head population from the Black River, atributary of Lake Erie, after the closing of alocal coking facility resulted in reduced PAHpollution (Baumann and Harshbarger 1995).

Empirical determination of the etiology ofenvironmentally induced tumors in fishpopulations is difficult. While associationsbetween environmental or tissue contami-nant levels and prevalence or severity ofcancer or noncancer liver lesions have beenmade, most often for PAHs (Johnson et al.1993) and sometimes PCBs and insecti-cides and their derivatives (Myers et al.1994), controlled laboratory studies that ex-perimentally attempt to induce tumors infish from natural populations have beenunsuccessful with few exceptions (Vethaaket al. 1996). This failure to induce tumorsmay have resulted from insufficient dura-tion of exposure, exposure of inappropriatelife stages, exposure to single chemicals ormixtures different than those present innature, or combinations of these factors.

The complexity of the neoplastic processmakes the experimental induction of tumorsdifficult. First, chemical carcinogenesis isknown to be a multistep process includinginitiation (when DNA in a single cell isdamaged by environmental mutagens),promotion (when cells with damaged DNAselectively outcompete neighboring cells withnormal DNA), and progression (when cellswith damaged DNA become genetically un-stable and spontaneously acquire additionalgenetic alterations) (Lengauer et al. 1998).DNA damage to initiate this process prob-ably must occur at specific genes, bothoncogenes and tumor suppressor genes(Manam et al. 1992). Carcinogenesis oftenrequires exposure to distinctly differentclasses of chemical carcinogens, those thatinitiate DNA damage and those that pro-

mote cells with damaged DNA. In naturalenvironments such as the Hudson River,metabolites of PAHs probably serve as ini-tiating agents by adducting to DNA (Krieket al. 1998; Stein et al. 1994), and PCBs(Cogliano 1998) and PCDD/Fs (Huff et al.1994) likely act as promoting agents. Sec-ond, early life stages are most sensitive totoxicity and carcinogenicity requiring thathusbandry practices of vulnerable targetspecies are sufficiently developed to allowfor rearing of fish from the embryo/larvaethrough to the adult stage. This is unchartedterritory for many species of marine and es-tuarine fish. Third, the development of tu-mors often requires that the rate of growthand sexual maturation in fish held in thelaboratory closely reflects that in their naturalenvironments, a difficult requirement formany taxa of concern.

Atlantic Tomcod from the HudsonRiver: A Model for Carcinogenesis?

In the 1970s and early 1980s, Atlantic tom-cod (Figure 2A) from the Hudson River ex-hibited one of the highest prevalences oftumors ever seen in natural populations ofany organism. Tumors were initially ob-served incidentally in tomcod collected fromthe Hudson River in routine biomonitoringprograms conducted by the utility compa-nies sited along the river. Smith et al. (1979)originally reported that the livers of 25% ofadult tomcod collected in 1978 between rm25 and rm 71 during their spawning sea-son (late December to February) exhibitedgross neoplastic nodules and hepatocellularcarcinomas (Figure 2B). In later studies, thepresence of gross and histologically identi-fied liver lesions was related to the age andsize of the fish. The prevalences of grosshepatic lesions in adult tomcod from theHudson River ranged from 32% for age 1fish to 78% for age 2 individuals, while suchlesions were absent in adult tomcod from


the Pawcatuck River, a less contaminatedwaterway along the Connecticut–Rhode Is-land border (Dey et al. 1993). Histopatho-logical examination revealed hepatomas(neoplastic nodules and hepatocellular car-cinomas) in 44% of age-1 and 93% of age-2Hudson River tomcod. In comparison, theincidence of histologically defined liver tu-mors in age-1 and age-2 tomcod from thePawcatuck River was 3% and 10%, respec-tively. Furthermore, histopathological evalu-ation failed to reveal preneoplastic or neo-plastic lesions in tomcod from five estuar-ies of varying environmental quality in thesouthern Gulf of St. Lawrence (Couillard etal. 1999).The lesions observed in the Hudson Rivertomcod covered a continuum from baso-philic foci with little cellular alteration tohighly invasive hepatocellular carcinomasaffecting the entire liver. The lesion mor-phology in tomcod was similar to chemi-cally induced liver lesions observed in ro-dents used as toxicological models. A posi-tive association was detected between theprevalence of liver lesions and the size of

fish within an age-group, indicating thathigh growth rate may enhance tumorigen-esis. Gross hepatic lesions and hepatomaswere absent in juvenile young of the year(age-0) tomcod collected in summers fromthe lower Hudson River Estuary (Dey et al.1993) suggesting that duration of exposure,source of contaminants, or processes associ-ated with sexual maturation in the fallmonths might play a role in the neoplasticprocess in this population. Environmentalfactors operating on a broad scale in theHudson River ecosystem may contribute tothe propensity of tomcod to exhibit chemi-cally induced tumors.

The presence of preneoplastic and neoplas-tic lesions in Hudson River tomcod has alsobeen associated with other abnormalities andpathologies of the liver, including signifi-cantly elevated hepatic lipid levels, perhapsindicative of increased detoxification activi-ties in livers of Hudson River tomcod(Cormier et al. 1989). Lipophilic contami-nants such as PAHs and PCBs would beexpected to occur at higher levels in these

Figure 3. The distribution of hierarchical levels of biological organization and how the response time andecological relevancy of biomarkers covary across this hierarchy.


fish during growth spurts, one of which oc-curs in the first autumn of age-0 fish asthey approach sexual maturation.

A short-term laboratory study did not de-tect liver cancers in juvenile age-0 HudsonRiver tomcod (Cormier and Racine 1990).Juvenile tomcod were collected in June fromthe Hudson River and Sheepscot River,Maine, and reared until December in cleanlaboratory water. The expectation was thatthe Hudson River tomcod, four to six monthsold at time of collection, would develop livercancer because of early life stage environ-mental exposures to contaminants and thattomcod from the Sheepscot River would not.The lack of tumors in the Hudson Rivertomcod might have been due to insufficientexposures through feeding on contaminatedprey or, more likely, to reduced growth andan absence of sexual maturation in theselaboratory-reared fish.

Use of Biomarkers to Determine theEtiology of Disease and Health of

Ecosystems

Biomarkers and bioindicators have been usedas alternatives to controlled laboratory stud-ies to investigate the etiology of cancers orother perturbations in natural populations.Biomarkers in fishes have also been used toevaluate the overall health of impactedaquatic ecosystems. Because the presenceof contaminants in the environment doesnot necessarily mean that they arebioavailable to potentially impacted popu-lations, the use of biomarkers andbioindicators confirms bioavailability andusually does so in the milieu of the naturalenvironment. Biomarkers have been definedas functional measures of exposure to envi-ronmental stressors and are often measuredat the suborganismal level of biological or-

ganization (Adams 2002). In contrast,bioindicators can be viewed as measurableresponses at higher levels of biological or-ganization from the organismal through thecommunity levels. Thus, it has been sug-gested that biomarkers can be viewed asmeasures of exposure and bioindicators asmeasures of effects. It is our contention,however, that any response at or above themolecular level is indicative of a biologicaleffect. Thus, we will use the term ‘biomarker’in this chapter for all biological responsesto exposures.

Biomarker responses can be quantified ata hierarchy of levels of biological organiza-tion from the molecular to the communitylevels (Figure 3). Responses at lower levelsof complexity (molecular) are immediate,sensitive, and specific to toxicant exposurerelative to responses at higher levels of or-ganization (population, community). Con-versely, responses at higher levels have moreecological relevancy but are delayed, dif-fuse, and difficult to assign to specific stres-sors relative to responses at lower levels oforganization. Current research efforts arefocused on identifying mechanistic linkagesbetween responses at different levels of bio-logical organization (Johnson and Collier2002). Success in these efforts would in-clude the capability to predict populationand community effects from perturbationsat lower levels of organization. Biomarkerscan be evaluated in natural populations,in situ enclosures, or laboratory settingsusing single or multiple contaminants orcomplex contaminated environmental ma-trices. Over the past decade, biomarkerstudies in tomcod have employed a varietyof investigative methods and have beenimplemented at multiple levels of biologi-cal organization in order to determine theetiology of tumors in the Hudson Riverpopulation and to evaluate the effects of


pollutants on tomcod and the Hudson Riverfish community.

Life History Characteristicsof Atlantic Tomcod

Tomcod lend themselves well to biomarkerstudies for several reasons. Tomcod are usu-ally very common in estuaries of the north-eastern United States and Atlantic Canada.The Hudson River Estuary supports thesouthernmost spawning population of tom-cod (Klauda et al. 1988). The egg stage oftomcod is benthic and prolonged (at least 1month), larvae at hatching have a large yolkcomplement from their mothers, and juve-niles and adults are bottom-dwelling om-nivorous feeders on benthic invertebrates.Tomcod thus have the potential tobioaccumulate high levels of lipophilic con-taminants through direct contact with sedi-ments, maternal transfer, and consumptionof benthic prey. They have a lipid-rich liverthat further serves to accumulate these pol-lutants. Tomcod are resident in the HudsonRiver year round, although they do undergoseasonal movements upstream to the saltfront for spawning (Klauda et al. 1988).The tissue burdens and biological responsesof tomcod, therefore, reflect exposure histo-ries that are specific to, and perhaps inte-grated across, much of their natal estuaries.This is in contrast to species such as mum-michog (also known as Atlantic killifish)Fundulus heteroclitus which have a morelimited range during their life history andtherefore have tissue burdens and biomarkerresponses that reflect more localized signalsof contaminant levels. Additionally, recentadvances in tomcod husbandry now allowfor multi-generational, controlled laboratorystudies of the responses of various life stagesof tomcod to contaminants. Furthermore,the timing of events in the tomcod life his-tory (e.g., the winter spawning of tomcod isunique in the Hudson River) results in the

young life stages of tomcod being a keyprey for major resource and other ecologi-cally important species within the lower es-tuary in spring and early summer. The widedistribution and high abundance of juve-nile tomcod throughout the tidal estuarycombined with its role as prey for commonfreshwater, estuarine, and marine species,place tomcod at a critical node in the HudsonRiver food web (Chambers and Witting,2006, unpublished).

Tissue Burdens of HAHs in Tomcod

Chemical carcinogenesis is a multistage pro-cess and in natural populations may re-quire high levels of chronic exposure toDNA-damaging PAHs and to tumor pro-moting PCBs and PCDD/Fs. Tomcod fromthe Hudson River Estuary bioaccumulatemuch higher levels of PCBs and PCDD/Fsand are exposed to significantly higher lev-els of PAHs than tomcod from elsewhere.

Initial studies of tissue burdens in tomcodcompared hepatic levels of PCBs andPCDD/Fs on a congener-and-sex specificbasis among adult tomcod collected fromthe Hudson River at Garrison, New York(rm 51), the Hackensack River, New Jersey,(one of two major tributaries to NewarkBay in the western Hudson River Estuary;Figure 1), and the Miramichi and Margareerivers in the southern Gulf of St. Lawrence,Canada (Courtenay et al. 1999). HepaticPCDD burdens in tomcod from theHackensack River were dominated byTCDD, which provided greater than 80%of total TCDD toxic equivalency quotients(hereafter, “TEQ,” which is a summarymeasure of the total toxicity of PCDDs,PCDFs, and coplanar PCBs in environmen-tal samples scaled to the toxicity of TCDDand in fishes is usually assessed by earlylife stage toxicities, gene induction, or both[Van den Berg et al. 1998]). Levels of he-


patic TCDD in tomcod from the HackensackRiver were among the highest ever reportedin a natural population. TEQs derived fromPCDD/Fs were seven- to eightfold higherin tomcod from the Hackensack River thanin those from the Hudson River and 250-fold higher than in tomcod from theMargaree River. Surprisingly, total copla-nar PCB levels were very similar betweentomcod from the Hackensack River and theHudson River and 40 and 90-fold higherthan in tomcod from the Miramichi Riverand Margaree River, respectively. The con-tribution of PCB77, expressed as lipid-nor-malized/body weight (ng/kg wet weight),comprised about 82% of the total coplanarPCBs in Hudson River tomcod. Hepaticlevels of all of these HAH contaminantswere much higher in male than female tom-cod (2.4-to 4.6-fold), suggesting that a size-able amount of these contaminants isshunted into tomcod eggs and, importantly,potentially transferred to and affecting earlylife stages of offspring.

Subsequent studies of tissue burdens haveexamined hepatic, congener-specific levelsof HAHs in juvenile age-0 tomcod (5–9months of age) from multiple sites withinthe Hudson River Estuary complex and inAtlantic Canada. Age-0 tomcod were col-lected from 19 locales in the Hudson RiverEstuary (between RM1 and RM107), New-ark Bay, the Hackensack River, theMiramichi River, and age-matched, labora-tory-spawned F1 juvenile tomcod of HudsonRiver descent (Yuan et al. 2001; Yuan 2003;Fernandez et al. 2004) (Table 1). Juvenilefish from the Hackensack River and New-ark Bay exhibited 930 times higher totalhepatic TEQs than similarly aged tomcodfrom the Miramichi River. Mean total TEQsin juvenile tomcod collected from the HudsonRiver were 62-fold higher than in tomcodfrom the Miramichi River and sevenfoldhigher than in the laboratory-reared F1Hudson River reared under clean conditions.

Coplanar PCB burdens expressed as totalTEQs using fish toxic equivalency factors(TEFs) (a relative measure of toxicity ofindividual PCDDs, PCDFs, and coplanarPCBs congeners compared to that of TCDDwhich is assigned a value of 1) were 58-and 19-fold higher in tomcod from theHackensack River/Newark Bay than in ju-veniles from the Miramichi River and labo-ratory-reared F1 Hudson River controls, re-spectively. Similarly, mean TEQs from co-planar PCBs were 30- and nine-fold higherin age-0 tomcod from the Hudson Rivercompared to age-0 juveniles from theMiramichi River or laboratory-reared F1Hudson River fish, respectively. However,coplanar PCBs only contributed 2% of thetotal TEQs in tomcod from the HackensackRiver/Newark Bay compared to a mean of25.6% in tomcod from the Hudson River.The contribution of coplanar PCBs to totalTEQs among locales in the Hudson Rivervaried greatly; the lowest contribution (1.5%)was at Yonkers (rm 17), where TCDD con-tributions were high, and the highest con-tribution (57%) was in fish from HaverstrawBay (rm 37). Surprisingly, there were notrends in total hepatic (coplanar andnoncoplanar) PCBs levels with river milesin the Hudson River. Variation in PCBslevels in tomcod among collection sites wasgreat, but the levels did not decrease withdistance downstream and away from thehistoric point sources of PCBs release intothe upper Hudson River.

Furthermore, a direct mixing model (Satheret al. 2001) was used to estimate the rela-tive contributions of three commerciallyavailable PCBs mixtures (Aroclor 1242[A1242], A1254, A1260) to the hepatic PCBsprofiles in age-0 tomcod from multipleHudson River sites (Fernandez et al. 2004).The last two digits in the numerical desig-nation of Aroclors denote the percentage ofchlorine that they contain, such that A1242contains fewer chlorine substituents than

343RESPONSES OF HUDSON RIVER FISH TO TOXICANTSTa

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A1254 or A1260. A1242 was the predomi-nant mixture released at the upstream GEmanufacturing facilities, whereas A1254 andA1260 predominate in NYC municipal andindustrial effluents. As expected, a linearincrease in the A1242 contribution with rivermile was observed in age-0 tomcod collectedbetween rm 0 and rm 80, a result whichwas consistent with its upriver sources. Un-expectedly, age-0 tomcod collected upriverof rm 80 exhibited a PCBs congener pat-tern with decreasing A1242 characteristics.Because this shift in pattern occurred nearthe zone of seasonal transition from brack-ish to freshwater, it was hypothesized thatthis was due to increased depuration or de-creased uptake by tomcod of lower chlori-nated PCBs congeners in freshwater.

Similar concentrations of PCDD/PCDFswere observed in age-0 juveniles as reportedfor adults from the same locales (Yuan etal. 2001, Yuan 2003; Fernandez et al. 2004).Juveniles from the Hackensack River/New-ark Bay exhibited slightly higher levels ofPCDD/Fs than adults, indicating that ex-posure occurs early in their life cycles. Ap-proximately 90% of the total TEQs in tom-cod from these two sites was contributedby one congener, TCDD, most likely origi-nating from the single industrial point sourceon the Passaic River (Figure 1; Bopp et al.1991). Total TEQs were highest in age-0juveniles from the Hackensack River (1052ng/kg) compared to a mean of 62 ng/kgTEQs in the 19 samples from the HudsonRiver. There were some TEQ “hotspots” inthe Hudson River with fish from Yonkers(rm 17) exhibiting 420 ng/kg total TEQs.The TCDD contribution to total TEQs wasmuch smaller in the Hudson River than inthe Hackensack River/Newark Bay, rangingfrom 42% at rm 17 in Yonkers to 8% at rm107.

Interestingly, PCBs congener patterns dif-fered between age-0 juvenile and adult tom-

cod collected from the same sites within theHudson River Estuary (Fernandez et al.2004). The di- to tetra substituted PCBcongeners dominated the PCBs composi-tion in age-0 juvenile tomcod, whereas thepenta- to nonachlorinated substituted con-geners, particularly those with di-ortho sub-stitutions, predominated in adults. This dif-ference in patterns between life stages mayhave been due to dietary or metabolic varia-tion between age-0 juvenile and adult tom-cod and may have implications for differ-ences in PCBs toxicity in young versus olderlife stages and for those consumers of thisresource.

Because of their long-term persistence withintissues, ratios of total PCDDs to PCDFscan provide information on sources of con-taminants and the movement of fish withinand between estuaries. The ratios of totalPCDDs:PCDFs expressed on a wet weightbasis were very different in juvenile tomcodfrom the Hackensack River/Newark Baycompared to those from the Hudson River.In tomcod from the Hackensack River/New-ark Bay the ratio of PCDDs:PCDFs variedbetween 1.5 and 1.9, while PCDFs concen-trations exceeded those of PCDDs in eightof nine collections from the Hudson River(Yuan et al. 2001; Yuan 2003; Fernandez etal. 2004). These results suggest that tom-cod from Newark Bay and the main-stemHudson River are exposed to different lev-els and sources of PCDDs/Fs and, giventhe persistence of these contaminants, thatexchange of tomcod between Newark Bayand the Hudson River is rare. This sug-gests that spawning populations of tomcodin the tributaries of Newark Bay are prob-ably distinct from those in the Hudson River.

Hepatic burdens of PCDDs and PCDFs,expressed as TEQs, were very similar intomcod from most Hudson River sites. Ona wet weight basis, however, the contribu-tion of TCDD to total PCDDs among


Hudson River collections decreased withdistance upstream from lower Manhattan.This pattern supports the assessment of anincreased bioavailability of upstream sourcesof PCDDs other than TCDD, most likelyoctachlorodibenzo-p-dioxins (Bopp et al.1998). In summary, these studies of tissueburdens in juvenile tomcod demonstrate thatafter only 5 to 9 months of environmentalexposure, juvenile tomcod from the HudsonRiver Estuary exhibit contaminant levelsthat are comparable to those found in adults.

Levels of PCDD/Fs and PCBs in unfertil-ized eggs of tomcod collected from theHudson River (RM50) and from the lowerMiramichi River (Loggieville, NewBrunswick) were analyzed and compared tothe hepatic levels of their mothers (Roy etal. 2001). Total TEQs from PCDD/Fs were8- to 15-fold higher in Hudson River tom-cod eggs compared to those from the

Miramichi River, while coplanar PCBs were73- to 79-fold higher in eggs from HudsonRiver tomcod than those from the MiramichiRiver. Lipid-adjusted TEQs from totalPCDD/Fs and total coplanar PCBs werevery similar in the unfertilized eggs and inthe livers of their mothers. This similaritysuggests that the maternal transfer of thesepersistent lipophilic contaminants to devel-oping embryos is substantial and could con-tribute to various early life stage toxicitiesin tomcod of Hudson River origin.

Bile Metabolites of PAHs in Tomcod

PAHs are very difficult to measure in tis-sues of fish because they are metabolizedvery rapidly (Varanasi and Stein 1991). Asurrogate measure of exposure has been de-veloped, termed “fluoresent aromatic com-pounds” (FACs), that quantifies bile me-

Figure 4. Relative levels of hepatic cytochrome P4501A1 (CYP1A1) mRNA expression in age-0 juvenileAtlantic tomcod collected from 1994 to 1998 and in 2000 from 40 sites in the Hudson River Estuarycomplex including two locales in the Hackensack River and one in Newark Bay (see Figure 1). Thenumerals along the x-axis refer to location (in river miles) in the Hudson River upstream of the southern tipof Manhattan Island. The mean (+ 95% CI) relative CYP1A1 mRNA concentrations are displayed.


tabolites of PAHs that are comprised offour to six aromatic rings (includingbenzo[a]pyrene (B[a]P)) by using high per-formance liquid chromatography with de-tection at an excitation/emission wavelengthpair of 380/430 nm (Krahn et al. 1986).Levels of FACs were significantly higher inadult tomcod from the Hudson River com-pared to those from three other cleaner riv-ers (Wirgin et al. 1994). For example, FACslevels were approximately eightfold higherin tomcod from the Hudson River than inthe Miramichi River.

Although Hudson River tomcod have beenexposed to higher levels of PAHs comparedto tomcod from elsewhere, this does neces-sarily mean that these compounds are bio-logically active. Parent PAH compounds arenot inherently mutagenic and carcinogenic.They become mutagenic once they are meta-bolically converted to forms that are highlyreactive with cellular biomolecules, at whichtime they can initiate cancer and cause otherperturbations by damaging DNA (Stein etal. 1994). Vertebrates, including fishes, havea system of enzymes (Phase I enzymes) thatoxidize PAHs and some other lipophilic com-pounds to form more polar metabolites(Stegeman and Hahn 1994), which are thenconjugated by other enzymes (Phase II en-zymes) to carrier molecules that facilitatethe excretion of these conjugated metabo-lites from the body (George 1994). Some-times, however, the metabolites of Phase Ienzymes escape the conjugating activity ofPhase II enzymes, which can result in dam-age to DNA.

Cytochrome P4501A1 Expressionin Environmentally Exposed Tomcod

Cytochrome P4501A1 (CYP1A1) is a ma-jor Phase I enzyme gene whose productsmetabolize PAHs and other compounds toreactants that can form adducts and dam-

age DNA. Levels of CYP1A1 expression infishes can be dramatically increased (in-duced) at the mRNA, protein, and enzymeactivity levels by exposure to many formsof HAHs and PAHs (Stegeman and Hahn1994). Levels of CYP1A1 expression infishes, therefore, are widely used to quan-tify exposure to these compounds and theirearly biological effect. Induced CYP1A1activity is needed to metabolize PAHs;HAHs are largely refractory to CYP1A1 me-tabolism. Persistent HAH-induced CYP1A1activity, however, can result in increasedDNA damage. This process is known togenerate reactive oxygen species that cancause mutations in DNA by modifying itsbases (Schlezinger et al. 1999). Oxidizedbases have also been implicated as a causeof cancer in fish populations from contami-nated locales (Malins et al. 1990).

AHs differ in their efficacies in inducingCYP1A1 due to their differing abilities tobind AHR. PCDDs, the most active AHs,often elicit induction at low parts per tril-lion levels, whereas PAHs are the least ac-tive AHs (induction occurs at ppm levels).It is thought that the AHR pathway medi-ates most, if not all, biological responses toAH contaminants including transcriptionof CYP1A1 and a battery of other xenobioticmetabolism genes. As a result, expressionlevels of CYP1A1 are believed to be predic-tive of most toxic responses to these con-taminants. Since TCDD best binds AHR,other molecules that are structurally simi-lar to TCDD share an ability to bind AHRand share potency in eliciting AHR-medi-ated toxicities. A numeric system based onempirically derived toxicity and CYP1A1expression data in mammals, birds, andfishes has been developed to compare thetoxicities of individual PCDDs, PCDFs, andcoplanar PCBs congeners (Van den Berg etal. 1998). Because it is most toxic, TCDDis given a Toxic Equivalency Factor (TEF)of 1. Toxicities of all other PCDD/Fs and


coplanar PCBs congeners are scaled to thatexhibited by TCDD and have values lessthan 1. Noncoplanar PCBs also have beenshown to exert toxicities, most likely througha mechanism independent of the AHR path-way and therefore are not assigned TEFvalues.

Levels of hepatic CYP1A1 mRNA expres-sion have been found to vary among tom-cod from different estuaries (Kreamer et al.1991). Wirgin et al. (1994) compared adult

tomcod collected from five Atlantic coastrivers and found, not unexpectedly, thatCYP1A1 mRNA expression was highest bya considerable degree in tomcod from theHudson River. Levels were 16- and 28-foldhigher in Hudson River tomcod than in tom-cod from the cleaner Saco River (Maine)and the Margaree River (Nova Scotia), re-spectively. These results confirmed that AHsare bioavailable to tomcod from the HudsonRiver and that AHs elicit a molecular-leveleffect that can be mechanistically linked to

Figure 5. Autoradiograms of thin-layer chromatograms of hepatic DNA adducts detected by 32Ppostlabeling analysis in Atlantic tomcod collected from 1987 to 1993 in five Atlantic coast estuaries.Locations of tomcod collections are (north to south) St. Lawrence River, Quebec; Miramichi River, NewBrunswick; Margaree River, Nova Scotia; Saco River, Maine; and Hudson River, New York. DNAadducts were visualized by storage phosphor imaging (Reproduced from Wirgin et al. [1994] withpermission from Environmental Health Perspectives).


perturbations at higher levels of biologicalorganization, in this case at the level ofDNA damage.

Levels of hepatic CYP1A1 mRNA in tom-cod also varied among locations within anecosystem although no significant trend be-tween CYP1A1 mRNA levels and locationin the Hudson River was evident. CYPIA1expression was quantified for age-0 juveniletomcod collected from 40 sites in the HudsonRiver (RM1 to RM107) and from theHackensack River and Newark Bay (Yuanet al. 2001; Yuan 2003). Levels of CYP1A1mRNA gene expression were found to varyby as much as 23- to 34-fold among siteswithin the estuary (Figure 4). As expected,levels of CYP1A1 and HAH contaminantswere highest in tomcod collected from theHackensack River and Newark Bay. In tom-cod from the Hudson River, CYP1A1 washighest at the most downriver sites and dis-played secondary peaks just downstream ofmunicipalities such as Newburgh, New York(rm 61) and Kingston, New York (rm 80).The significant and consistent site-to-sitedifferences in the levels of the CYP1A1biomarker support the utility of this andother biomarkers as indicators of themicrogeographic distribution of bioavailablecontaminants in this large system. Althoughlevels of hepatic HAHs (expressed as TEQs)and hepatic CYP1A1 were highest in juve-nile age-0 tomcod from Newark Bay/Hackensack, there was not, however, a sig-nificant among-site relationship between lev-els of CYP1A1 mRNA and hepatic bur-dens of HAHs in tomcod from the HudsonRiver (Yuan et al. 2001). This may havebeen due to exposure to environmental in-ducers other than HAHs (such as PAHs),resistance to CYP1A1 mRNA inducibilityor a combination of these and other factors.

The age at which induced CYP1A1 mRNAexpression occurs in environmentally exposedHudson River tomcod is yet to be fully de-

termined. CYP1A1 mRNA expression wasnot induced in environmentally exposedHudson River tomcod larvae (1–2 monthsold) relative to matched, similarly aged, anduntreated laboratory-reared F1 larvae ofHudson River descent (Roy et al. 2002).CYP1A1 mRNA expression in these labo-ratory-reared tomcod was highly inducible,however, by waterborne exposure to B[a]P.In view of the high levels of CYP1A1 mRNAand HAH burdens in wild 5- to 9-monthold juvenile tomcod from the Hudson River(Yuan et al. 2001; Yuan 2003), these re-sults suggest that 1) the HAHs in younglarvae that might have been transferred tothe egg and onto the yolk sac larvae fromthe mother have been diluted by growthand depuration to levels that do not induceCYP1A1 expression, and 2) bioaccumulationof contaminants has yet to occur in larvaealthough at 2 months of age these fish wouldhave been consuming prey, albeit pelagiczooplankton, for most of this period. Anincrease in tissue burdens of HAHs andCYP1A1 mRNA may be associated withthe habitat and diet shift of tomcod associ-ated with their transition from pelagic lar-vae to the benthic associated juvenile lifestage.

DNA Damage in Tomcod from theHudson River

Although these studies indicate that AHsaccumulate and are biologically active inHudson River tomcod, they do not demon-strate injury to individual fish nor evidenceof historical effects on the population. Useof the 32P- postlabeling technique has indi-cated that tomcod from the Hudson Riversuffered higher levels of hepatic DNA dam-age than tomcod from cleaner rivers. The32P-postlabeling technique is a very sensi-tive assay for quantifying levels of hepaticDNA adducts in environmentally exposedfish (~ 1 adduct in 10 9 to 10 10 nucleotides)


(Stein et al. 1994). DNA adducts are DNAbases to which the reactive metabolites ofPAHs or other compounds have covalentlybound. Presumably the presence of thesebulky DNA adducts results in errors dur-ing DNA replication and therefore the in-troduction of mutations (Wirgin andTheodorakis 2002). Among Hudson Rivercontaminants of concern, 32P postlabelingis sensitive to PAHs but not to PCDD/Fsor PCBs exposure because the latter com-pounds are not usually biotransformed toreactive metabolites (Wirgin et al. 1994).Nevertheless, it is difficult to determine towhich individual PAHs the organism hasbeen exposed because the 32P postlabelingassay detects exposure to many differentDNA-bound PAH metabolites.

DNA adducts in fishes provide a cumula-tive, lifetime measure of exposure to thesecompounds because fishes are inefficient atrepairing damaged DNA compared to, forexample, rodents (Stein et al. 1993). More-over, DNA adducts have proven to be aninformative tool for biomonitoring becausethey are unambiguous indicators ofgenotoxicity and have been clearly associ-ated with health effects such as neoplasia,tissue damage, and early life stage toxici-ties (Stein et al. 1994; Kriek et al. 1998,Wirgin and Theodorakis 2002). Importantly,the formation of DNA adducts, if incarcinogenically important genes such asoncogenes or tumor suppressor genes, isprobably a necessary first step in the initia-tion of chemical-caused neoplasia. Addition-ally, the levels of hepatic CYP1A1 and he-patic DNA adducts should be correlatedbecause the generation of DNA adducts ismechanistically linked to prior activity ofCYP1A1 in order to metabolically trans-form PAHs.

The levels of hepatic DNA adducts weresignificantly higher in adult tomcod fromthe Hudson River than in adult tomcod

from four cleaner rivers (Wirgin et al. 1994)(Figure 5). Hudson River tomcod exhibited40-fold higher levels of hepatic DNA ad-ducts than tomcod from the least impactedlocale, the Margaree River. Similarly, levelsof DNA adducts were significantly higherin tomcod from the AH-contaminated St.Lawrence River than in all of the other riv-ers except the Hudson River. This confirmsthat adult tomcod from the Hudson Riverhave been exposed to much higher levels ofPAHs than tomcod from other locales andthat these almost certainly have had detri-mental biological effects such as the initia-tion of neoplasia or early life stage toxici-ties. In later studies (Yuan 2003), levels ofhepatic DNA adducts were compared injuvenile tomcod collected in the year 2000from 10 sites within the Hudson River (rm1 to rm 107). Although hot spots were iden-tified and there were fourfold differences inlevels of DNA adducts among sites, no geo-graphic pattern among river locales wasobserved. Interestingly, overall levels of he-patic DNA adducts in juvenile fish wereapproximately an order of magnitude lessin 2000 than that previously reported foradult Hudson River tomcod collected be-tween 1989 and 1993 (Wirgin et al. 1994).This result is consistent with cumulativelifetime exposure and inefficient DNA re-pair in fish, but it is also consistent with apossible decrease in PAH levels during the7- to 11-years period between collections.

While high levels of overall DNA damageare almost certainly deleterious to a popu-lation, it is the genomic location of the dam-age that is most predictive of its carcino-genic potential. Oncogenes and tumor sup-pressor genes are known targets of PAH-DNA adducts in mammals and fish (re-viewed in Wirgin and Theodorakis 2002)and their mutagenic activation and inacti-vation, respectively, is necessary for initia-tion of chemical carcinogenesis. In fact, thespecific sensitive parts of these genes and


patterns of mutation are often identical be-tween mammals and fish (Bailey et al.1996). For example, codons 12–13 and 62–63 of the K-ras oncogene in rodents, hu-mans, and rainbow trout Oncorhynchusmykiss are vulnerable to PAH-induced mu-tations. DNA from liver tumors of tomcodfrom the Hudson River was found to havean activated K-ras oncogene, whereas therewas no evidence of K-ras activation in DNAfrom normal livers of tomcod from theHudson River or Saco River (Wirgin et al.1989). In total, results from these studiesare consistent with a chemical basis to theinitiation of the high prevalence of tumorsin tomcod and the sensitivity of the HudsonRiver population to toxicity from these con-taminants.

CYP1A1 Expression in ChemicallyTreated Tomcod

Experimental studies were conducted in or-der to calibrate the responses from field-exposed fish and to determine the role ofdifferent classes of AHs in eliciting re-sponses, particularly CYP1A1 induction. Inthese studies, adult tomcod from the HudsonRiver and Miramichi River were returned tothe laboratory, depurated in clean water forextended periods (usually 20, but up to 300d), and treated under controlled experimen-tal conditions with AHs of concern. Althoughthis period was undoubtedly insufficient todepurate HAHs to background levels,CYP1A1 mRNA were decreased to basallevels (Kreamer et al. 1991). As expected,hepatic CYP1A1 mRNA expression in adulttomcod from both populations was dose-responsive and highly inducible (up to 500-fold) by exposure to PAHs (B[a]P and beta-naphthoflavone [B-NF]). Tomcod from thetwo populations, however, differed dramati-cally in their sensitivities to CYP1A1 mRNAinduction by coplanar PCBs or TCDD; tom-cod from the Miramichi River were highly

sensitive, whereas tomcod from the HudsonRiver were insensitive. In these initial stud-ies with limited doses of PCB77 and TCDD,there was no evidence of CYP1A1 induc-tion by these chemicals in tomcod from theHudson River (Wirgin et al. 1992; Courtenayet al. 1999).

In more extensive dose–response studies,significant differences in the sensitivity ofhepatic CYP1A1 mRNA inducibility wasobserved between adult tomcod from thetwo aforementioned populations for threedifferent coplanar PCBs congeners (PCB77,PCB81, and PCB126) and TCDD. When

Vehicle

B[a]P

PCB77

Miramichi Hudson

Figure 6a. A slot blot hybridization of cyto-chrome P4501A1 (CYP1A1) mRNA expression inlaboratory-reared, F1

early yolk-sac tomcod larvaefrom Miramichi River and Hudson River parents.Larvae were waterborne-exposed to 1 ppm B[a]Pfor 48 hr, 1 ppm PCB77 for 144 hr, or acetonevehicle for 144 hr and immediately sacrificed.CYP1A1 mRNA expression is depicted for fourlarvae for each treatment group in each population


higher doses of these chemicals were used,significant hepatic CYP1A1 induction wasobserved in Hudson River tomcod, albeitrequiring 100-fold higher doses than thatneeded to significantly induce CYP1A1 inMiramichi River tomcod (Yuan 2003). Ad-ditionally, the extra-hepatic inducibility ofCYP1A1 mRNA was compared in brain,heart, gills, gonads, intestine, kidney, andspleen in adult and juvenile tomcod fromthe two populations. For all tissues tested,tomcod from the Hudson River were ap-proximately 100-fold less sensitive toCYPIA1 mRNA induction than tomcodfrom the Miramichi River (Yuan 2003).Thus, tomcod from the Hudson River ex-hibited significantly less sensitivity for allHAHs tested in all tissues tested. Further-more, this difference between populationsin sensitivity to coplanar PCBs, but notB[a]P-induced CYP1A1 induction, was also

observed for PCB77 in waterborne exposedembryos, early yolk sac larvae, and late yolksac larvae (Wirgin unpublished data). Thus,tomcod from the Hudson River exhibitedsignificantly less sensitivity for all HAHs,in all tissues, and all life stages tested.

Yuan (2003) addressed the issue of whetherresistance to CYP1A1 inductibility byHAHs, but not PAHs occurs throughoutthe Hudson River tomcod population bycollecting juvenile age-0 tomcod from sevensites (rm 1 to rm 107) in the Hudson Riverand one site in the Miramichi River, depu-rating them in clean laboratory water, andtreating them with PCB77 or B[a]P. Fishfrom all Hudson River sites showed stronginduction of CYP1A1 mRNA after expo-sure to B[a]P, but not after exposure toPCB77. Juvenile age-0 tomcod from theMiramichi River showed strong induction

0

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Unexposed B[a]P PCB77 Unexposed B[a]P PCB77

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Figure 6b. A comparison of relative cytochrome P4501A1 (CYP1A1) mRNA expression in 1 ppm B[a]P,1 ppm PCB77, and acetone vehicle waterborne-exposed laboratory-reared, F1 early yolk-sac tomcodlarvae from Miramichi River and Hudson River parents (n=8 larvae/treatment group in each population).The mean (+ 95% CI) CYP1A1 mRNA concentrations are displayed.


for both chemicals. This demonstrates thatjuvenile tomcod from throughout the HudsonRiver population exhibit impaired CYP1A1inducibility and further strengthens the ar-gument that this is due to a population-wide genetic adaptation.

Researchers working on mummichog (alsoknown as Atlantic killifish) from highly pol-luted Atlantic Coast estuaries have madesimilar, but not identical, observations. Kil-lifish from the Hudson River (Elskus et al.1999), Newark Bay (Prince and Cooper1995b), New Bedford Harbor, Massachu-setts (Nacci et al. 1999; Bello et al. 2001),and the Elizabeth River, Virginia (Van Veldand Westbrook 1995) exhibited reducedCYP1A protein or EROD inducibility com-pared to conspecifics from reference sites forAHs. Killifish from these highly pollutedlocales also were resistant to overt adultlethality or early life stage toxicities inducedby these and other AH compounds (Princeand Cooper 1995a; Nacci et al. 1999; Meyeret al. 2002). In most cases, in contrast totomcod, resistance to gene induction andtoxicities was observed for both PAHs andHAHs in these killifish populations.

Resistance as a Genetic Adaptationor a Physiological Acclimation?

Resistance in tomcod and other fish speciesprobably resulted from chronic exposuresto highly toxic AHs, other contaminants,or combinations of these. Resistance to toxi-cants can result from genetic adaptationsor physiological acclimations. If genetic, re-sistance is a population-level effect that istransmitted across generations and that per-sists long after site remediation. Develop-ment of genetic adaptations requires thatvariant resistant individuals exist, albeit atlow frequencies, in the population prior tochronic exposure to chemical pollutants(Nacci et al. 1999). Chemical exposure re-sults in the selection of individuals withresistant phenotypes and an increase in theirfrequency over time in the population. Studiesshowed that inter-individual variation inCYP1A1 inducibility was high within tom-cod populations (Courtenay et al. 1994) aswas variation in the structure of theCYP1A1 gene (Roy et al. 1995). If resis-tance has a physiological basis, it willquickly be lost as exposures to the opera-

Figure 7. Effects of dose of a mixture of coplanar PCBs congeners and two controls (water (0 x groups),and acetone (contaminant vehicle)) on development of replicate populations of F2 embryos of Hudson Riverand Miramichi River ancestry. The x represents concentrations of four coplanar PCB congeners (PCB77,PCB81, PCB126, and PCB169) measured in the livers of adult tomcod from the Hudson River Estuary.


tive toxicant diminish in frequency and in-tensity.

In either case, it is thought that resistanceis associated with a cost, such as an in-creased sensitivity to other stressors or re-duced performance (e.g., fecundity, growth,etc.) relative to individuals lacking resis-tance when both groups are in environmentsfree of the toxicant. For example, killifishfrom the AH-resistant Elizabeth River popu-lation are significantly more susceptible thanthose from a nearby clean reference site toacute phototoxicity from a PAH, ambientUV light, and low oxygen conditions (Meyerand Di Giulio 2003). Additionally, a cad-mium (Cd)-resistant population of the oli-gochaete worm Limnodrilus hoffmeisteri in-habited a highly contaminated locale (Cd,Ni, cobalt) in the Hudson River (FoundryCove [Figure 1]). Following remediation ef-forts, the genetically adapted resistant phe-notype in the population was rapidly dis-placed by Cd-sensitive worms that were ei-ther at low background levels prior to theclean-up or rapidly immigrated to the sitesuggesting that selective pressure againstthe resistant phenotype was intense in theremediated environment (Levinton et al.2003).

We conducted experiments to determinewhether resistance in tomcod has a geneticor physiological basis. Adult tomcod werecollected from the Hudson and Miramichirivers and F1 and F2 pure Hudson River,pure Miramichi River, and F1 hybrid off-spring were produced and evaluated forCYP1A1 mRNA inducibility as embryosor larvae (Roy et al. 2001; Chambers andWirgin unpublished data). In initial stud-ies with a single dose of each chemical,CYP1A1 mRNA was significantly induc-ible in pure F1 larvae from both popula-tions with B[a]P, however, gene expressionwas significantly higher in larvae ofMiramichi River descent. For PCB 77, sig-

0.01 0.10 1.00 10.00 100.00

20 pg per g)

-5

-3

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0.01 0.10 1.00 10.00 100.00

-5

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Figure 8. Effects of population source (HudsonRiver vs. Miramichi River), contaminant (PCBs andTCDD), and dose (0.01 to 100 times the concentra-tions of PCBs or TCDD measured in the liver of adulttomcod from the Hudson River Estuary) on morpho-metric variables of 1-d old F1 tomcod larvae assummarized by the scores on the first principalcomponent axis (PCA 1). The controls (water andacetone (the contaminant vehicle)) did not differ intheir effects on PCA 1 values and are represented bythe solid horizontal lines. The sensitivity of tomcodfrom the Miramichi River, and the resistance ofHudson River tomcod to dose were best reflected inPCA 1, a composite of whose values increases withincreasing fish length, head and eye size, and yolk-sac depth. PCA 2 values (not shown) best separatedthe two source populations from one another.


nificant CYP1A1 induction was observedin larvae of both populations, althoughlevels of gene expression were significantlyless in larvae of Hudson River comparedto Miramichi River descent. Hybrids ex-hibited intermediate levels of gene expres-sion. In later, more extensive experiments,F1 and F2 embryos, early yolk sac larvae,and late yolk sac larvae from both popu-lations were exposed to more doses ofB[a]P and PCB77 and evaluated forCYP1A1 mRNA inducibility. Responsesin the F1 and F2 offspring were nearly iden-tical in all three early life stages. CYP1A1mRNA was highly inducible by treatmentwith PCB77 or B[a]P in both generationsof Miramichi River offspring. In both gen-erations of Hudson River offspring,CYP1A1 mRNA was highly inducible byB[a]P, but only slightly by PCB77 (Fig-ures 6a, 6b). Statistical analysis indicatedthat CYP1A1 mRNA inducibility was sig-

nificantly more sensitive to PCB77, butnot for B[a]P treatment in tomcod fromthe Miramichi River than in those fromthe Hudson River. These results indicatethat genetic adaptation has occurred inthe Hudson River tomcod population toHAHs, but not PAHs. This also suggeststhat more than one molecular pathway –likely an AHR-dependent and an AHR-independent pathway – may mediateCYP1A1 induction and perhaps higher-level biological responses in this popula-tion. Although levels of both PAHs andHAHs are unusually high in the HudsonRiver environment, it is possible thatCYP1A1 activity in detoxifying PAHs iscritical to organismic survival and popu-lation viability, but because HAHs inducebut are poorly metabolized and generatereactive oxygen species, the second alter-native pathway is shut down in the HudsonRiver population.

water

1 X

10 X

5 X

MiramichiHudsonwater

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10 20 30 400 10 20 30 400

Pro

po

rtio

n a

ctive

0.0

0.2

0.4

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Figure 9. Effect of population source (Hudson River vs. Miramichi River) and dose of PCBs mixture offour coplanar congeners (PCB77, PCB81, PCB126, and PCB169) on activity level of unfed F2 larvaeduring their first 40 d after hatching. Dose represents 0 to 100 times ‘x’, the concentration of thesecongeners measured in the livers of adult tomcod from the Hudson River Estuary.


Early Life stage Toxicities in PCBsand TCDD Exposed Tomcod

While CYP1A1 induction is predictive ofmany toxic responses to PCDD/Fs, PCBs,and PAHs because it is AHR mediated, noempirical evidence of resistance at higherlevels of biological organization had beenobserved in tomcod from the Hudson Riverpopulation. It is known that early life stagetoxicities, including craniofacial malforma-tions, multifocal hemorrhages, reducedgrowth, yolk sac edema, pericardial edema,and reduced blood flow (Hornung et al. 1996,1999) are very sensitive biological responses

to TCDD and dioxin-like PCBs in some,but not all species of fish (Elonen et al.1998). It is thought that the early life stagetoxic effects of individual PCB congenerscovary with CYP1A1 inducibility and aremediated by the AHR pathway. Thus, it islikely that early life stage toxic effects ofindividual PCB congeners can be predictedbased on their structural similarity toTCDD. These effects are also expected tobe additive and species-specific. In controlledlaboratory experiments, early life stages oflake trout are exquisitely sensitive to TCDD(Walker et al. 1991; Walker et al. 1994) anddioxin-like PCBs (Walker et al. 1996), andmuch more sensitive (16–180 fold) than other

Figure 10. A schematic of a cell depicting molecules that are part of the aryl hydrocarbon receptor(AHR) pathway and whose expression was quantified and compared between Atlantic tomcod from theMiramichi River and Hudson River. Abbreviations: AIP,-aryl hydrocarbon receptor interacting protein;AHRR,-aryl hydrocarbon receptor repressor; ARNT, -aryl hydrocarbon receptor nuclear translocator;CYP1A1, -cytochrome P4501A1; DRE, -dioxin response element; HSP90, -heat shock protein 90.


fish species tested (Elonen et al. 1998). Asa result, despite continued introductions ofcaptive-reared fish to restore wild popula-tions, it is believed that lake trout popula-tions in the Great Lakes have failed to re-bound to historical levels of abundance be-cause environmental exposures of young lifestages to these chemicals has severely com-promised recruitment. Thus, it is possiblethat chronic exposure to these chemicalsmay have affected early life stage success intomcod from the Hudson River and thismay consistently restrict recruitment in thispopulation as well. In fact, analyses of theabundance of spawning tomcod in theHudson River show an overall decline overthe past twenty years (M. Mattson, per-sonal communication, Normandeau Asso-ciates).

We undertook experimental studies of tom-cod as an investigative bridge betweenorganismal and population-level effects. Theeffects of exposure to environmentally rel-evant concentrations and identities of co-planar PCBs and TCDD were experimen-tally investigated in young life stages of labo-ratory-reared tomcod from the Hudson Riverand the Miramichi River. In two consecu-tive years, F1 and F2 embryos of HudsonRiver and Miramichi River ancestry wereexposed as embryos to graded doses of awaterborne mixture of coplanar PCBs(PCB77, PCB81, PCB126, and PCB169)or TCDD. Doses were selected based ontissue burdens and congener patterns pre-viously measured in wild tomcod from theHudson River (Courtenay et al. 1999; Yuanet al. 2001, Roy et al. 2001). Measures ofviability of embryos, larvae, and yolk saclarvae, and morphometric and behavioralvariables of yolk sac larvae were scored. Themorphometric variables, including measuresof curvature of the larvae, yolk sac volume,and the sizes of eyes, head, and jaws, werereduced by principle component analysis(PCA) and the resulting PCA scores were

evaluated via analysis of variance (ANOVA)in order to detect the impacts of toxicant,dose, population, and their interactive ef-fects on tomcod larvae. Behavior was sum-marized by quantifying the relative activity(swimming versus resting) of variously ex-posed yolk sac larvae.

Tomcod from the two source populationsdiffered dramatically in their sensitivities toHAH contaminants. F1 and F2 embryos ofMiramichi River tomcod showed a gradedresponse to PCBs dose with lower viabilityto hatching and greater levels of partial orfailed hatch-outs at higher doses (Figure7). Miramichi River embryos, exposed toPCBs doses at and above baseline (‘1 x’)that survived to hatching, did not survivethrough their first six weeks of larval life. Incontrast, Hudson River tomcod embryosdisplayed a consistently high level of vi-ability to hatching with minimal abnormali-ties at all ages regardless of PCBs dose,which included doses that were 10 to 100times greater those measured in environ-mentally exposed tomcod from the HudsonRiver.

These differences between populationsources in their responses to dose were alsoexpressed in the morphology of survivors.In this case, a parallel response occurred toPCBs and TCDD (Figure 8). Over 80% ofthe experiment-wide variation in morpho-metric features was captured in the firsttwo principal component (PC) axes. Totallength, head dimensions (including eye andjaw size), and yolk amount contributed mostto the first PC axis. Miramichi River yolksac larvae showed dramatic change in theirmorphology at doses at and above the nomi-nal baseline dose. In contrast, Hudson Riverindividuals showed no significant effect ofdoses of either PCBs or TCDD.

These experimental studies were also de-signed to test for more subtle and sublethal


behavioral responses. Vertical position inthe water column is critical to the successof the nearly passive earliest larval stages ofestuarine fishes. Because the Hudson Riveris a vertically stratified estuary of overlyingfreshwater flowing downstream and deeperseawater flowing upstream, vertical positionwithin the water column will affect the spa-tial distribution of larvae and the locationin the benthic habitat of recently transformedjuvenile fish. In the laboratory, untreatedlarval tomcod from both populations ex-hibit characteristic regular vertical move-ments in which they ascend from the bot-tom to near the surface of holding vessels.Presumably, this behavior is reflective ofsimilar movements in the natural environ-ment. Vertical movements during the first2–40 d of larval life for PCBs exposed F2Miramichi and Hudson River tomcod weremonitored while they resided in 400 mLbeakers. Activity of Miramichi River larvaewas sensitive to all doses of PCBs exposure(doses at 1–10 times those measured in tom-cod from the Hudson River), whereas tom-cod larvae of Hudson River descent wereunaffected by all doses of PCBs mix (up to100 x) (Figure 9). In total, these studiesindicate that tomcod from the Hudson Riverpopulation have undergone significant evo-lutionary change at gene(s) that mediateresponses at organismal and population lev-els to HAHs, perhaps due to chronic expo-sure to these contaminants.

Studies of the Mechanistic Basis ofResistance in Hudson River Tomcod

Because of its importance in mediating manytoxic responses to HAHs and PAHs, it ispossible that the mechanistic bases of re-sistance resides in the structure (amino acidsequence) or expression of molecules in theAHR pathway (Figure 10). Briefly, in unex-posed cells, AHR resides in the cytoplasmwhere it is bound to two molecules of heat

shock protein 90 (HSP90) and a single mol-ecule of AHR interacting protein (AIP, alsoknown as ZAP2 and Ara9). When lipo-philic AH contaminants (ligands) diffuseinto the cell, they bind AHR in the cyto-plasm, HSP90s and AIP are released, andthe ligand-AHR complex is translocated intothe nucleus by an unknown mechanism.Within the nucleus, ligand-AHR binds thearyl hydrocarbon nuclear translocator(ARNT) and this three-part complex bindsenhancer elements (dioxin response elements[DREs] also known as XREs and AHREs)upstream of AH responsive genes such asCYP1A1. Binding of the ligand-AHR-ARNT complex to these enhancer elementsactivates induced transcription of thesegenes. The aryl hydrocarbon receptor re-pressor (AHRR) serves to down-regulateactivation of the AHR pathway by bindingwith ARNT and the AHRR-ARNT com-plex competing with the ligand-AHR-ARNTcomplex for DRE binding sites. AHRR-ARNT binding to DREs does not activatetranscription of CYP1A1.

To investigate the possible role of alteredAHR pathway molecules in conferring re-sistance, AHR, ARNT and AHRR werecloned from tomcod and their DNA se-quences determined (Roy and Wirgin 1997;Roy and Wirgin 2002). For AHR, sequenceswere compared between tomcod from theHudson River and three populations in At-lantic Canada. All tomcod from the HudsonRiver shared AHR genotypes not seen intomcod from elsewhere. All tomcod fromthe Hudson River showed a two amino aciddeletion in the AHR that was absent intomcod from the three Canadian rivers(Yuan 2003). The significance of this ge-netic polymorphism has yet to be assessedin functional assays.

Cloning and characterization of these genesallowed for development of polymerase chainreaction (PCR)-based assays to quantify and


compare the expression of AHR, ARNT,and AHRR at the RNA level among mul-tiple life stages (embryo, larval, juvenile,and adult) and tissues of tomcod from theHudson River and Miramichi River popu-lations (Roy et al. 2002). The effects of PCBsor TCDD treatment on the expression ofthese genes were then evaluated. It was hy-pothesized that resistance of the HudsonRiver population results from decreasedbasal or induced expression of AHR andARNT or increased expression of AHRR.Levels of AHR and ARNT expression weresimilar between tomcod from the two riv-ers, except that AHR expression was sig-nificantly higher in environmentally exposedtomcod from the Hudson River comparedto those from the St. Lawrence or MiramichiRivers. Exposure to coplanar PCB77 orTCDD under controlled laboratory did notaffect levels of AHR in any life stage ortissue of tomcod from either population. Asexpected, AHRR expression was equally in-ducible in tomcod from the two popula-tions by B[a]P, but more so in tomcod fromthe Miramichi River than the Hudson Riverby PCB77. Levels of hepatic AHRR ex-pression also differed among environmen-tally exposed adult tomcod from theHudson, St. Lawrence, and Miramichi Riv-ers, with highest levels of gene expressionin tomcod from the Hudson River. Thus,levels of AHRR expression paralleled thatof CYP1A1 in both controlled laboratoryexperiments and environmental exposures.

Several important conclusions can bereached from these studies of the mechanis-tic basis of resistance. First, differential lev-els of expression of molecules in the AHRpathway at the RNA level did not seem tocontribute to the resistant phenotype in tom-cod from the Hudson River. If anything,AHR was higher and AHRR expression lowerin tomcod from the Hudson River comparedto those from more sensitive populations.Second, levels of AHRR expression paral-

leled those of CYP1A1 in these popula-tions in that it was higher in environmen-tally exposed tomcod from the HudsonRiver than elsewhere and it was less induc-ible by PCB77 in tomcod from the HudsonRiver than in those from the MiramichiRiver. This similarity in patterns of expres-sion between AHRR and CYP1A1 in thesepopulations may be due to both genes be-ing similarly regulated at the transcriptionallevel. Both genes are induced by exposureto AH compounds and binding of AHR-ARNT to enhancer elements in their regu-latory regions. Third, AHR expression wasnot inducible or down-regulated by AHtreatment in tomcod. These results contrastwith those in resistant killifish from theElizabeth River (Meyer et al. 2003), but arein agreement with those in resistant killi-fish from New Bedford Harbor (Powell etal. 2000). Thus, it does not seem that chronicexposure of tomcod in the Hudson River toAHs has down-regulated AHR expressionand thereby impaired inducibility ofCYP1A1. One caveat is that these analyseswere at the RNA level, and it is expressionof the proteins encoded by these genes thatare functionally important. Studies usingrodents have shown that exposure to HAHsdestabilizes AHR and results in down-regu-lation of its expression (Pollenz et al. 1998).Studies of AHR pathway proteins of tom-cod must be conducted to truly understandthe role of this pathway in resistant pheno-types.

Summary and Conclusions

Surprisingly few examples of toxic effects ofchemical contaminants on the Hudson Riverfish community have been reported. Com-pared to neighboring large estuaries, theHudson River hosts at least as rich an ich-thyofaunal community. In recent decades,abundance of glamour species in the HudsonRiver such as shortnose sturgeon and stripedbass are near historically high levels. How-


ever, one species, Atlantic tomcod exhibitsmanifestations of toxic damage consistentwith exposure to AHs at a hierarchy ofmechanistically linked biological endpoints,including altered gene expression, DNAdamage, liver cancer, a truncated age struc-ture, and reduced adult abundance. Link-ing the higher-level perturbations in a natu-ral population to exposure to single or evengroups of contaminants, however, is tenu-ous in the absence of empirical verification.Controlled laboratory exposure experimentsare needed to evaluate the sensitivity of naïvetomcod to environmentally relevant mixturesof PAHs, PCDD/Fs, and PCBs in induc-ing hepatic neoplasms or preneoplastic le-sions and to determine the effects of tumorson longevity of individuals and the overallpopulation age structure.

Despite extensive evidence of toxic effects,the Hudson River tomcod population is re-markably resistant to biological effects ofcoplanar PCBs and TCDD at a variety ofmolecular and organismal endpoints. Thesealtered biological responses are mechanisti-cally linked and suggest a role for the AHRpathway in conferring resistance. Resistanceseems to have largely a genetic basis, al-though shorter-term acclimation also con-tributes to modulated gene expression inthis population. Most of the known compo-nents of the AHR pathway have been char-acterized in tomcod. However, their expres-sion at the mRNA level does not seem toexplain the resistance phenomenon. Stud-ies are needed to characterize the expres-sion of these genes at the protein level be-cause it is here where the phenotype is de-termined.

Given their high levels of chronic exposureto and bioaccumulation of AHs, it is notsurprising that resistance has developed inthe Hudson River tomcod population. Infact, it has been suggested that resistancemay be common in fish populations that

are chronically exposed to high levels of AHs(Wirgin and Waldman 2004). Tissue bur-dens of HAHs in environmentally exposedtomcod from the Hudson River are at orexceed those that induce acute toxicity intomcod from other populations that we haveexamined. On the surface, the developmentof resistance seems inconsistent with in-creased DNA damage, elevated prevalenceof tumors, and a truncated age structure inthis population. It is possible that resis-tance has only very recently developed andthat the prevalence of tumors today is lessthan that in the early 1980s. Targeted fieldstudies are needed to document the preva-lence of tumors and preneoplastic lesions inthe current Hudson River tomcod popula-tion. Alternatively, as has been suggestedin the mammalian literature, the develop-ment of liver tumors may serve as a strat-egy of resistance to toxicants. In tomcodfrom the Hudson River, it is likely that atleast one spawning occurs in age-1 fish be-fore any cancer-associated mortality, thusnegating any selective load of cancer on thepopulation. Liver tumors may allow indi-viduals to survive in the face of chronicexposure to hepatotoxicants (Farber 1990;Farber and Rubin 1991; Yusuf et al. 1999).In fish and in mammals, hepatic tumorsand preneoplastic lesions often exhibit de-creased expression of Phase I and increasedexpression of Phase II enzymes comparedto normal tissues. These strategies may en-hance detoxification of environmentalhepatotoxicants in tumors compared to nor-mal tissues and provide them a selectiveadvantage.

While resistance does provide short-termbenefits to populations in the face of chronicexposure to toxicants, it does not come with-out costs at the population and communitylevels. Fish that are resistant to one chemi-cal may be more sensitive to the effects of asecond stressor than fish that are not resis-tant. Investigations into interactive effects


of multiple chemical stressors (AH and met-als) in adult and young life stages of tom-cod from the Hudson River are underwayand suggest that tomcod from the HudsonRiver may be more sensitive to metals thanthose from the Miramichi River (Sorrentinoet al. 2004; Sorrentino 2004). Furthermore,tomcod in the Hudson River are at the verysouthern extreme of the species’ geographicdistribution and the additional effects ofthermal stress during the warmer monthsare of special concern and warrant investi-gation.

After the removal of a strong selective agent(e.g., through site remediation), the perfor-mance of resistant populations may be di-minished compared to that of sensitivepopulations. This reduction in performancemay be due to the allocation of a dispro-portionate share of a limited energy budgettowards resistance. Thus, fecundity, growthrate, and survival may all be reduced in apopulation that has been challenged by andadapted to contaminants. These and otherdemographic parameters can be empiricallytested in controlled laboratory experimentswith tomcod as has been done for killifishfrom the Elizabeth River.

The evolution of resistance permits the sur-vival of populations in habitats that arechronically challenged by high levels of con-taminants, but given the demonstrated im-portance of young-of-the-year tomcod as preyto resource and other key species through-out the lower Hudson River Estuary, resis-tance by tomcod almost certainly results inelevated transfer of contaminants up thefood chain. Several aspects of the rate ofconsumption of tomcod by striped bass andother fishes are currently being investigatedby us, but the full impact of resistance intomcod on its immediate and secondaryconsumers warrants full evaluation. Ulti-mately, however, resistant populations arean aberration. Their persistence, as well as

repercussions throughout the communityand ecosystem in which they are embed-ded, are best considered as toxic responsesto chemical contaminants in their environ-ment.

Acknowledgments

We acknowledge support of the HudsonRiver Foundation, SBRP Grant ES10344,NIEHS Center Grant ES00260, and theNortheast Fisheries Center (NOAA Fisher-ies Service) for the studies described in thisreview.

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