Aquatic Pollution-Induced Immunotoxicity in Wildlife Species1

Toxicological Sciences, May 1997

The potential for chemicals to adversely affect human immunologic health has traditionally been evaluated in rodents, under laboratory conditions. These laboratory studies have generated valuable hazard identification and immunotoxicologic mechanism data; however, genetically diverse populations exposed in the wild may better reflect both human exposure conditions and may provide insight into potential immunotoxic effects in humans. In addition, comparative studies of species occupying reference and impacted sites provide important information on the effects of environmental pollution on the immunologic health of wildlife populations. In this symposium overview, Peter Hodson describes physiological changes in fish collected above or below the outflows of paper mills discharging effluent from the bleaching process (BKME). Effects attributable to BKME were identified, as were physiological changes attributable to other environmental factors. In this context, he discussed the problems of identifying true cause and effect relationships in field studies. Mohamed Faisal described changes in immune function of fish collected from areas with high levels of polyaromatic hydrocarbon contamination. His studies identified a contaminant-related decreases in the ability of anterior kidney leukocytes to bind to and kill tumor cell line targets, as well as changes in lymphocyte proliferation in response to mitogens. Altered proliferative responses of fish from the contaminated site were partially reversed by maintaining fish in water from the reference site. Peter Ross described studies in which harbor seals were fed herring obtained from relatively clean (Atlantic Ocean) and contaminated (Baltic Sea) waters. Decreased natural killer cell activity and lymphoproliferative responses to T and B cell mitogens, as well as depressed antibody and delayed hypersensitivity responses to injected antigens, were identified in seals fed contaminated herring. In laboratory studies, it was determined that rats fed freeze-dried Baltic Sea herring had higher virus titers after challenge with rat cytomegalovirus (RCMV) than rats fed Atlantic Ocean herring; perinatal exposure of rats to oil extracted from Baltic herring also reduced the response to challenge with RCMV. Keith Grassman reported an association between exposure to polyhalogenated aryl hydrocarbons and decreased T cell immunity in the offspring of fish-eating birds (herring gulls and Capsian terns) at highly contaminated sites in the Great Lakes. The greatest suppression of skin test responses to phytohemagglutinin injection (an indicator of T cell immunity) was consistently found at sites with the highest contaminant concentrations. Judith Zelikoff addressed the applicability of immunotoxicity studies developed in laboratory-reared fish for detecting altered immune function in wild populations. She presented data from studies done in her laboratory with environmentally relevant concentrations of metals as examples. Although the necessity of proceeding with caution when extrapolating across species was emphasized, she concluded that published data, and results presented by the other Symposium participants, demonstrate that assays similar to those developed for use in laboratory rodents may be useful for detecting immune system defects in wildlife species directly exposed to toxicants present in the environment.

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Aquatic Pollution-Induced Immunotoxicity in Wildlife Species1

FUNDAMENTAL AND APPLIED TOXICOLOGY 1 Aquatic Pollution-Induced Immunotoxicity in Wildlife Species Robert W. Luebke 0 1 Peter V. Hodson 0 ! Mohamed Faisal 0 2 Peter S. Ross 0 3 J Keith A. Grasman 0 Judith Zelikoff' 0 0 Aquatic Pollution-Induced Immunotoxicity in Wildlife Species. Luebke , R. W., Hodson, P. V., Faisal, M., Ross, P. S., Grasman, K. A., and Zelikoff, J. (1997). Fundam. Appl. Toxicol. 37, 1-15 1 lmmunotoxicology Branch, U.S. EPA , Research Triangle Park , North Carolina 27711; tSchool of Environmental Studies, Queen's University Kingston, Ontario, Canada K7L 3N6; %School of Marine Science, Virginia Institute of Marine Science, The College of William and Mary , Gloucester Point, Virginia 23062 , USA 2 NYU Medical Center, Institute of Environmental Medicine , Tuxedo, New York 10987 , USA 3 Seal Rehabilitation and Research Centre , Pieterburen , The Netherlands ; department of Biological Sciences, Wright State University , Dayton, Ohio 45435 The potential for chemicals to adversely affect human immunologic health has traditionally been evaluated in rodents, under laboratory conditions. These laboratory studies have generated valuable hazard identification and immunotoxicologic mechanism data; however, genetically diverse populations exposed in the wild may better reflect both human exposure conditions and may provide insight into potential immunotoxic effects in humans. In addition, comparative studies of species occupying reference and impacted sites provide important information on the effects of environmental pollution on the immunologic health of wildlife populations. In this symposium overview, Peter Hodson describes physiological changes in fish collected above or below the outflows of paper mills discharging effluent from the bleaching process (BKME). Effects attributable to BKME were identified, as were physiological changes attributable to other environmental factors. In this context, he discussed the problems of identifying true cause and effect relationships in field studies. Mohamed Faisal described changes in immune function of fish collected from areas with high levels of polyaromatic hydrocarbon contamination. His studies identified a contaminant-related decreases in the ability of anterior kidney leukocytes to bind to and kill tumor cell line targets, as well as changes in lymphocyte proliferation in response to mitogens. Altered proliferative responses offish from the contaminated site were partially reversed by maintaining fish in water from the reference site. Peter Ross described studies in which harbor seals were fed herring obtained from relatively clean (Atlantic Ocean) - ' Symposium held at the 35th Annual Meeting of the Society of Toxicology (SOT), Anaheim, CA. Sponsored by the lmmunotoxicology Specialty Section of the SOT. 2 To whom correspondence should be addressed at Mail Drop 92, U.S. EPA, Research Triangle Park, NC 27711. E-mail: luebke.robert® epamail.epa.gov. 3 Present address: Contaminant Sciences Section, Institute of Ocean Sciences, P.O. Box 6000, Sidney, BC, Canada V8L 4B2. and contaminated (Baltic Sea) waters. Decreased natural killer cell activity and lymphoproliferative responses to T and B cell mitogens, as well as depressed antibody and delayed hypersensitivity responses to injected antigens, were identified in seals fed contaminated herring. In laboratory studies, it was determined that rats fed freeze-dried Baltic Sea herring had higher virus titers after challenge with rat cytomegalovirus (RCMV) than rats fed Atlantic Ocean herring; perinatal exposure of rats to oil extracted from Baltic herring also reduced the response to challenge with RCMV. Keith Grassman reported an association between exposure to polyhalogenated aryl hydrocarbons and decreased T cell immunity in the offspring of fish-eating birds (herring gulls and Capsian terns) at highly contaminated sites in the Great Lakes. The greatest suppression of skin test responses to phytohemagglutinin injection (an indicator of T cell immunity) was consistently found at sites with the highest contaminant concentrations. Judith Zelikoff addressed the applicability of immunotoxicity studies developed in laboratory-reared fish for detecting altered immune function in wild populations. She presented data from studies done in her laboratory with environmentally relevant concentrations of metals as examples. Although the necessity of proceeding with caution when extrapolating across species was emphasized, she concluded that published data, and results presented by the other Symposium participants, demonstrate that assays similar to those developed for use in laboratory rodents may be useful for detecting immune system defects in wildlife species directly exposed to toxicants present in the environment e 1997 sod«y or With few exceptions, immunotoxicity studies have been conducted under laboratory conditions in rodents. As such, many toxicologists are only vaguely aware of the immunotoxicity studies done under the more natural condition of environmental exposure to pollutants. The purpose of this symposium was threefold: to disseminate information on the immunologic health of selected wildlife species in contaminated habitats; to demonstrate the plausibility of using altered immune responses of directly exposed species as a biomarker to predict the effects of toxic environmental pol0272-0590/97 $23.00 Copyright © 1997 by the Society of Toxicology. All rights of reproduction in any form reserved. lutants; and to foster collaboration between field and laboratory immunotoxicologists. Dr. Peter Hodson, past-president of SETAC and Director of The School of Environmental Studies at Queen's University, will address trends in aquatic pollution and provide specific examples based on pulp mill effluents. The next three speakers will discuss the effects of polluted environments on feral populations of fish (Dr. Mohamed Faisal, Virginia Institute of Marine Science), birds (Dr. Keith Grasman, Wright State University), and seals (Dr. Peter Ross, Institute of Ocean Sciences) living in these contaminated environments. The final speaker (Judith Zelikoff, New York University Medical Center) will discuss how well laboratory immunotoxicological studies in wildlife species compare with data generated in field studies, with an emphasis on fish models. All of the speakers in this symposium are actively involved in evaluating immune function in wildlife species under field or laboratory conditions and are uniquely suited to discuss the ecological, immunological, and logistical aspects of these studies. TRENDS IN RESEARCH ON WATER POLLUTION: EFFECTS ON FISH (P. V. Hodson and C M . Couillard) In North America, the adverse effects of water pollution have evolved from fish kills to less obvious sublethal effects. Chemicals and pathogens still threaten aquatic life, but more often through subtle impairment of biochemical, physiological, immune, and behavioral processes. Pollution control demands an ability to recognize and measure sublethal responses, to interpret their impacts on fish populations, and to establish cause-effect relationships to direct appropriate remediation. While laboratory studies can unequivocally relate chemical exposures to specific effects, field studies are the only way to verify proposed cause-effect relationships and to validate relevance to populations. Given the complexity of natural ecosystems, field studies are also the most susceptible to bias and misinterpretation. Epidemiology is the study of the causes and correlates of diseases and is emerging as a critical component of environmental toxicology. Fox (1991) has reviewed the principles of epidemiology as applied to field studies of pollution and the criteria for testing the strength of proposed cause-effect relationships (Susser, 1986) : time order, the cause must precede the effect and the effect must diminish with remediation; strength of association, statistical strength and clear exposure-response relationships; specificity, in effects of a cause and in causes of an effect; consistency on replication, over time, among cases; and coherence, with theory, biology, the facts, and the statistics; does the proposed relationship make sense? These criteria are often difficult to apply. Potential confounding factors include old sources of effluent which preclude time order studies, sources that are located at inconvenient sites, multiple sources of pollution, migration of test species among sites, and poor knowledge of the test species and of the ecosystem at test sites. This paper reviews field studies of the effects of bleached kraft mill effluent (BKME) on fish to illustrate the problems in relating cause and effect unequivocally and to identify factors to consider when interpreting biological responses. A landmark study of BKME in the Baltic Sea (Sodergren etai, 1992) established a suite of effects on fish that included accumulation of chemicals specific to BKME; induction of liver mixed-function oxygenases (MFO); physiological signs of stress, ionoregulatory failure, and immune dysfunction; histopathological abnormalities and deformities; decreased recruitment; and shifts in population and community structure. The intensity of effects decreased with secondary treatment, with effluent dilution, and after substitution of C1O2 for Cl2 as a bleaching agent. As a result, the Swedish EPA set new regulations for pulp mill effluents to limit the discharge of chlorinated organic compounds. The applicability of these results to North America was questioned (Mehrle et ai, 1989) because the Baltic is brackish, unlike most receiving waters in North America; the mill was undergoing modernization; Cl2 was the primary bleaching agent; there was no secondary effluent treatment; and the "reference" mill (not using chlorine bleaching) discharged effluent under different receiving water conditions. In response, related studies of white sucker (Catostomus commersoni) were launched at Canadian bleached kraft mills, including one on the St. Maurice River, Quebec. The river provided an ideal "upstream-downstream" study because dams prevented fish from migrating between reference and exposed sites, and the only other human presence in the watershed was the town of LaTuque; the river was bounded by a wilderness park and commercial forest. A preliminary study of 8-15 fish per site showed clear patterns of response parallel to a downstream dilution gradient (Hodson et ai, 1992a) . Most importantly, serum testosterone and estradiol concentrations decreased in exposed fish and recovered to reference values by 100 km distance, which mirrored hormone effects and impaired sexual maturation of Lake Superior white suckers exposed to BKME (Munkittrick et al., 1991). These results appeared to confirm the effects reported in Sweden. The case for BKME effects was supported by consistency on replication (Sweden, St. Maurice, Lake Superior), strong dilution-response relationships, and specificity of causes for a "syndrome" of effects. However, these results were not conclusive. The St. Maurice preliminary study was one observation, and upstreamdownstream studies violate statistical assumptions of independence of treatments; i.e., events upstream influence conditions downstream. Furthermore, there were obvious ecological gradients associated with river morphology: the reference site was a reservoir, sites 2 and 35 km downstream were riverine, and the site 100 km downstream was the reservoir of the next dam. Hormone effects were unreliable because fish were sampled in August, during the interspawning period when concentrations are lowest. Finally, the river was used for log driving, and the influence of floating and sunken logs was unknown. A more ambitious study was carried out between 1990 and 1993 to determine if physiological results from 1989 could be repeated August to August and over three seasons (prespawning in May, interspawning in August, and sexual maturation in November). Data on population demographics were collected to test the significance of physiological responses, and the prevalence of pathologies was measured to determine if BKME increased rates of disease. The Gatineau River was added as a reference for ecological gradients. It flows southwest from the same geographic area as the St. Maurice, and although it has roughly half the drainage area and half the annual flow of the St. Maurice, the water quality, fish communities, and watershed land use are similar. Two dams and reservoirs, separated by about 100 km, isolate a stretch of river, with the town of Maniwaki providing a source of nutrients. This second study sampled several thousand fish from each river. Those in the St. Maurice downstream of the mill were exposed to BKME, as suggested by elevated levels of tissue chlorophenols and induction of liver MFO activity (Gagnon et al., 1995) . In contrast, reference fish upstream of the mill and from all sites on the Gatineau showed no signs of chemical exposure. In November, sexually maturing fish from the St. Maurice had lower concentrations of serum sex steroids downstream of the mill, confirming the 1989 study (Gagnon et al, 1994) . Generally, physiological responses of fish were similar from August 1989 to August 1990 at all sites in the St. Maurice (Hodson et al, 1992b) , although seasonal changes in certain endpoints, including MFO induction, were consistent among sites. A role for effluent in other observed alterations cannot be ruled out, but it would not be a simple effect because site to site differences in various endpoints varied significantly with season, indicating that factors other than site and exposure to effluent were important. Similarly, for some effluent-associated responses, the pattern of change in the Gatineau River was the same as in the St. Maurice, indicating an effect of ecological gradients rather than BKME. In contrast, MFO induction and liver and muscle glycogen concentrations exhibited a pattern unique to the St. Maurice, consistent with an effect of BKME. The small number of responses strongly associated with BKME exposure contrasts sharply with the multiple physiological responses attributed to BKME by the Swedish study. In terms of population demographics, downstream fish in both rivers grew more quickly than reference (upstream) fish (Gagnon et al., 1995) . In the St. Maurice, fish at downstream sites were larger when they first reached sexual maturity (age to maturity), but age to maturity itself did not vary. Relative gonad weights of males (normalized to fish weight) were larger in fish exposed to BKME relative to reference fish, while gonads of females were smaller, and the relationship between fecundity and female body weight was disrupted near the pulp mill. These responses were confirmed by resampling in 1992 and 1993 and indicated slight impairment of reproduction, unlike dramatic effects observed in Lake Superior (Munkittrick et al., 1991). Gatineau white suckers sampled at downstream sites showed very different responses to increased growth rates. Age to maturity of both sexes was lower, relative gonad sizes increased, and fecundity of females increased. In brief, both sexes invested more energy in reproduction at downstream relative to upstream sites, a classic response to nutrient enrichment and increased growth rates (Trippel and Harvey, 1989) . Therefore, the effect of BKME on white sucker in the St. Maurice was the absence of such an investment: the failure to mature earlier and to increase gonad size and fecundity at downstream sites was a negative effect of BKME exposure. Measurements of pathology were examined in a similar way, comparing the prevalence and characteristics of each pathology among sites, among seasons, and between exposed and reference rivers. The covariance of different pathologies was also measured as was the influence of factors such as size and sex. For example, the severity of some lesions was size related and the largest fish were found at the site closest to the pulp mill. By applying analysis of covariance, the dependence of severity on size was factored out of the site to site comparison. In the St. Maurice River, changes in fin morphology, liver color, fat abundance, and infection with parasites showed site-related changes that might be interpreted as a response to BKME in the absence of further statistical analyses (Couillard et al., 1995) . However, logistic regressions of prevalence on size and Wilcoxon tests for differences in prevalence between sexes demonstrated that liver color and tapeworm infestation varied mainly with fish size. Similarly, fish from the Gatineau showed that fin morphology and fat abundance were more a function of ecological gradients present in both rivers than of exposure to BKME. Only the prevalence of encysted nematode larvae, least common at the site closest to the pulp mill, suggested a BKME effect. Because this parasite was absent from the Gatineau River, it is uncertain whether infestation rates in the St. Maurice are due only to BKME. The life cycle of this parasite involves oligochaetes and fish eating birds, both of which may vary in abundance with habitat. A variety of histological changes were also observed (CouiUard and Hodson, 1996) . In liver, spleen, and kidney, pigmented macrophage aggregates (PMA or melanomacrophage centers) were most prevalent at the most BKMEexposed site, 2 km downstream of the mill in the St. Maurice. PMA in all tissues increased with fish age and size, but in liver, the rate of formation of PMA with size (slope) increased most rapidly in BKME-exposed fish. In kidney and spleen, slopes did not change but intercepts were highest in exposed fish; i.e., for a given age, PMA were more prevalent in fish from downstream sites. Hence, despite an age bias, there was a real, residual effect of site, suggesting a response to BKME; there were no equivalent changes among sites in the Gatineau River. It has been proposed that an increased prevalence of PMA in exposed fish may be related to MFO induction and activation of compounds to reactive intermediates that promote lipid peroxidation and free radical formation. An alternative explanation might be erythrocyte damage by resin and fatty acids (Bushnell et ai, 1985) . Hence, a PMA response is coherent with a BKME exposure. Overall, these experiments demonstrated the importance of considering as many natural variables as possible in assessing the effects of effluents or contaminants on fish health. In Susser's (1986) terms, specificity in causes of an effect is a major criterion in judging the validity of proposed causeeffect relationships. Strength of association, consistency on replication, and even coherence with theory, biology, facts, and statistics may point to an effluent as a causative agent. However, this conclusion could be false if the effects of ecological gradients, fish biology (size, sex), and other disease processes (parasites) are not considered. In epidemiological terms, the simple upstream-downstream study of the St. Maurice did not clearly demonstrate the effects of BKME. Only by including a reference river and by considering the effects of season, fish size, and sex were the effects of BKME obvious. Our ability to relate cause and effect will be enhanced by research to • improve methods and sampling designs in "ecoepidemiology," • understand mechanisms of toxicity and effects, • develop biomarkers relevant to both mechanism of toxicity and to population level impacts, • identify natural confounding factors, • define threshold exposures causing toxicity and threshold responses leading to population level effects, and • understand the biology of the test species. Finally, the best methods available will be ineffective without sufficient effort to collect an adequate number of fish. IMMUN0T0XIC0LOGICAL STUDIES IN FERAL FISH POPULATIONS: EXAMPLES FROM THE CHESAPEAKE BAY (Mohamed Faisal) The influx of xenobiotics into the Chesapeake Bay has contributed to major damage of this aquatic environment. As a result, fin- and shellfish suffered from a wide variety of environmentally induced diseases. The Elizabeth River, in Virginia, is an example of how high concentrations of chemical pollutants can adversely influence the integrity of an ecosystem. For over two centuries, the Elizabeth River has been the site of heavy industry, military naval activities, shipping, and shipbuilding. As a result, astronomically high concentrations of polycyclic aromatic hydrocarbons (PAH) have been found in several sites on the river. Studies have shown that fish collected from the Elizabeth River suffer from cancers, cataracts, and skin ulcerations (Huggett et ai, 1992) . In 1990, Vogelbein and co-workers described a high incidence of malignant hepatocellular carcinoma in a population of Fundulus heteroclitus (Pisces: Cyprinodontidae) from a site in the Elizabeth River (Atlantic Wood) whose sediment PAH concentration was as high as 2200 mg PAH/kg dry wt sediment. Several reports have also described cancer epizootics in feral fish population collected from other sites heavily contaminated with PAH in the United States. The mechanism involved in this PAH-associated liver cancer in fish has received little attention. Therefore, a study was designed (Faisal etal, 1991a) to investigate the possible immunotoxic effects of PAH on fish natural cytotoxic cells (believed to be the fish equivalent to mammalian NK cells). Tumor-bearing F. heteroclitus were collected from the Atlantic Wood site and from a relatively nonpolluted reference site on the York River in Virginia. The cytotoxic activity of anterior kidney leukocytes (believed to be the fish bone marrow equivalent) was tested against the tumor cell line K562. The leukocytes from the tumor-bearing fish displayed a significant depression of tumorolytic activity as compared with leukocytes from healthy fish of the York River (Fig. 1A). Analysis of leukocyte-tumor cell conjugates indicated that leukocytes from Atlantic Wood fish were unable to recognize and subsequently bind to the tumor target cells (Fig. IB). Depuration of the Atlantic Wood fish in cleaner York River water for up to 28 weeks failed to reverse suppressed NCC activity. In another study (Faisal et ai, 1991b) , spot (Leiostomus xanthurus) were collected from five sites in the lower Chesapeake Bay system that represented a gradient of sediment PAH concentrations. The proliferative responses to mitogens by anterior kidney lymphocytes were assessed using [3H]thymidine uptake by replicating DNA (Fig. 2). The data show two different mitogen-dependent lymphocytic responses as the sediment PAH levels increase at the sampling 40 - I 32 centrations (r2 > 0.8). A similar correlation was also observed with 15 selected individual PAH compounds regardless of their molecular weights. By maintaining the fish in clean York River water for up to 24 weeks, it was possible in some cases to reverse the effects on proliferative responses. Further studies (Faisal and Huggett, 1993) have shown that incubation of isolated lymphocytes with the high-molecular-weight PAH, benzo[a]pyrene (BP), suppressed the lymphoproliferative response to Con A. The pathway for this suppression appears to involve metabolic processes mediated by the cytochrome P-450 system, since blocking of the reaction by a-naphthaflavone (a known P-450 inhibitor) reversed the suppression. HPLC analysis provided evidence that the microsomal fraction of isolated anterior kidney cells of spot were able to convert BP into its reactive metabolites (Faisal, 1994) . Among BP breakdown products, the proximate carcinogen, (—)-fran5-7,8-dihydroxy-7,8 dihydro-BP was identified. Investigations have also shown that the immunotoxic effects of contaminated sediments extended to shellfish populations. It was shown that hemocytes (believed to be analogous to the vertebrate macrophage) of oyster (Crassostrea virginica) collected form PAH-contaminated sites of the Elizabeth River have altered cytometric (Sami et al., 1992) and phenotypic (Sami et al., 1993) characteristics and suppressed macromolecular synthesis (Faisal and DemmerleSami, 1994) (Contribution No. 2067 from Virginia Institute of Marine Science). CONTAMINANT-RELATED IMMUNOSUPPRESSION IN HARBOR SEALS FED HERRING FROM THE BALTIC SEA (P. S. Ross, R. L. de Swart, H. H. Timmerman, P. J. H. Reijnders, H. van Loveren, J. G. Vos, and A. D. M. E. Osterhaus) In recent years, a spate of mass mortalities among different marine mammal species has been attributed to newly identified members of the genus Morbillivirus (Osterhaus et al., 1990; de Swart et al., 1995) . In 1988, approximately 20,000 harbor seals {Phoca vitulina) died following an outbreak of a previously unknown agent, later isolated and named phocid distemper virus (PDV) (Osterhaus et al., 1990). Despite the isolation of the virus from victims, it was not possible to preclude a contributory role of immunotoxic environmental contaminants to the disease event. Harbor seals, like other piscivorous marine mammals, are at the top of the aquatic food chain, and can accumulate high levels of lipophilic contaminants, including polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins, and polychlorinated dibenzofurans. Elevated concentrations of these and other contaminants have been associated with abortions and premature pupping in California sea lions (Zalophus californianus) (Delong et al., 1973) , tumors and decreased fecundity in Beluga whales {Delphinapterus leucas) in the St. Lawrence River (Martineau et al., 1987) , decreased fecundity in harbor seals in the Dutch Wadden Sea (Reijnders, 1980 and 1986) , impaired reproduction in ringed seals (Phoca hispida) in the Baltic Sea (Helle et al., 1976) , and skeletal lesions in harbor and grey (Halichoerus grypus) seals in the Baltic Sea (Bergman et al., 1992; Mortensen et al., 1992) . As such, contaminant levels in industrialized coastal areas have been high enough to elicit adverse biological effects in marine mammals. Until recently, however, evidence of polyhalogenated aromatic hydrocarbon (e.g., dioxins and PCB's)-induced immunotoxicity in mammals has been largely limited to laboratory studies under carefully controlled conditions. Numerous laboratory-based studies have demonstrated that dioxin-like compounds are particularly injurious to the mammalian immune system (Vos and Luster, 1989) . Exposure to low doses of such compounds can lead to an increased susceptibility to infectious disease (Vos et al., 1991; Thigpen et ai, 1975; House et al., 1990) . In an attempt to assess the effect of environmental contaminants on immune function in harbor seals, we carried out a 30-month captive feeding experiment (de Swart et al., 1994, 1995b,c,d; Ross et al., 1995, 1996a, 1996b) . During this time, 11 seals were fed herring from the relatively uncontaminated Atlantic Ocean and 11 seals were fed herring from the contaminated Baltic Sea. All seals had been captured as recently weaned pups from the relatively uncontaminated coast of northeastern Scotland. Blood was sampled on a regular basis and a series of immune function assays was carried out. In vitro tests consisted of mitogen-induced lymphocyte proliferation and natural killer cell assays. Antigenspecific responses following immunization with tetanus toxoid and rabies antigen were assessed by determining specific antibody levels and in vitro cellular proliferation. Delayedtype hypersensitivity (DTH) and antibody responses following immunization with ovalbumin provided an additional means of assessing the ability of the immune system as a whole to mount a specific immune response to a foreign antigen. Natural killer cell function in seals of the Baltic group began to decline in comparison with their Atlantic counterparts in the first six months of the study and remained consistently lower for the duration of the feeding study (Ross et al., 1996b) . T-lymphocyte function, as assessed by in vitro responses to the T-cell mitogens Con A and PHA, and the T- and B-cell pokeweed mitogen, began to decline within 1 year of the start of the study, and remained consistently lower during the second year (de Swart et al., 1994; de Swart et al., 1995b) . Specific responses to foreign antigen were also affected, as evidenced by reduced cellular proliferation to tetanus toxoid and rabies antigen in vitro (de Swart et al., AQUATIC POLLUTION-INDUCED IMMUNOTOX1CITY IN WILDLIFE SPECIES 1995b), and by reduced DTH and antibody responses to ovalbumin (Ross et ai, 1995) . While the in vitro tests provided evidence that certain facets of immune function were impaired by exposure to contaminants, the reduced specific responses to antigen suggested that the immune system as a whole was less able to mount a response to a foreign substance in the seals fed herring from the Baltic Sea. Elevated neutrophil numbers in the blood of seals in the Baltic group may have indicated an increased incidence of bacterial infections as a consequence of immunosuppression or alterations in stem cell production in the bone marrow (de Swart et ai, 1995c) . Other hematological and serum chemistry parameters were largely comparable between the two groups of seals (de Swart et ai, 1995c) . In an effort to corroborate and extend these findings, two parallel studies were carried out using PVG rats. Laboratory rats offer several advantages over seals, including the availability of specific immunological reagents, the capacity to carry out more intrusive testing, including virus challenge experiments, and their comparative ease of handling. In the first of these studies, adult rats were fed a diet consisting of freeze-dried herring from the same two batches used in the seal study (Ross et ai, 1996c) . While no contaminant-related effect on immune function could be detected, higher rat cytomegalovirus (RCMV) titers following challenge suggested that host resistance may have been affected. In the second study, pregnant rats were given a daily dose of oil extracted from the two batches of herring, and immune function was assessed in their offspring (Ross et ai, 1996d) . The pregnant rats were exposed to a similar dose of 2,3,7,8tetrachlorodibenzo-p-dioxin toxic equivalents (TEQ) as seals on a body weight basis per day (rats received 0.25 and 2.12 ng TEQ/kg bw/day in the Atlantic and Baltic groups, respectively). Perinatal exposure to the lipophilic contaminants present in the herring oil resulted in a transient thymus atrophy, reduced virus-associated natural killer cell responses following infection with RCMV, and reduced RCMV-specific immunoglobulin G titers. An additional, positive control, group was perinatally exposed to a daily dose of 134ng/kgbwTEQ. Offspring of this group displayed a more profound impairment in T-cell immunity than rats of the Baltic Sea group. Results of these experiments suggest that lipophilic contaminants in the Baltic Sea herring are immunotoxic to PVG rats, although perinatal exposure appears to lead to more severe effects. While the experimental design precludes an in-depth comparison, harbor seals appear to be more sensitive to the immunotoxic actions of the contaminants in the Baltic Sea herring than PVG rats. In addition, the sensitivity of the rat thymus to the actions of the contaminants suggests that this organ may also have been targeted in the seals. Contaminants in the Baltic Sea herring were immunotoxic to both laboratory rats and harbor seals, primarily affecting T-cell immunity and natural killer cell activity. While it is difficult to relate this to an increased susceptibility to virus infection, the observed reduction in immune function in our captive harbor seals may well be of clinical significance. Natural killer cells are a vital first line of nonspecific defense against virus infections and tumors, while T-cells are essential in the specific clearance of virus and coordination of an immune response. In the case of the 1988 PDV epizootic, a contaminant-related immunosuppression may have predisposed harbor seals in European waters to a more severe infection by a new virus than might otherwise have been the case. Since our captive seals accumulated concentrations of PCBs that were lower than those found in many populations of free-ranging harbor seals in Europe and North America, we conclude that contaminant-induced immunotoxicity represents a real threat to these and other marine mammal species (Ross et ai, 1996a) . In addition, free-ranging seals are exposed to contaminants perinatally, predisposing them to more profound effects than those observed in our juvenile seals that were caught in uncontaminated waters prior to the study. Analysis of the profiles of contaminants in the Baltic Sea herring lipid and in the harbor seal blubber following 2.5 years on a diet of Baltic Sea herring suggests that PCBs represent the greatest aryl hydrocarbon (Ah) receptor-mediated threat to seals. PCBs represented over 50% of the total TEQs in the herring and over 90% in the seal blubber (Ross et ai, 1996a) . With the stabilizing environmental levels of PCBs in the last decade and the long life span of harbor seals (up to 40 years), a contaminant-induced immunotoxicity may represent a problem for marine mammals for many years yet. IMMUNOLOGICAL BIOMARKERS AND ENVIRONMENTAL CONTAMINANTS IN FISH-EATING BIRDS OF THE GREAT LAKES (K. A. Grasman, P. F. Scanlon, and G. A. Fox) Fish-eating birds are effective "sentinel species" to help assess the effects of toxic contaminants on the health of the Great Lakes ecosystem. Over the last 30 years, many studies have documented associations between organochlorines and physiological, reproductive, developmental, behavioral, and population-level problems in fisheating birds of the Great Lakes (Fox and Weseloh, 1987, Gilbertson et ai, 1991, Fox, 1993, Bowerman et ai, 1995) . Affected species have included the herring gull (Larus argentatus), the Caspian tern (Sterna caspia), the double-crested cormorant (Phalacrocorax auritus), and the bald eagle (Haliaeetus leucocephalus). Effects associated with PCBs continue at some sites today, including embryonic mortality, deformities, and low reproductive success. Banding studies of Caspian terns have shown low recruitment into the breeding population of terns raised at highly contaminated colonies (Ludwig, 1979, Mora et al, 1993) . In bald eagles inhabiting the shorelines of the Great Lakes and rivers open to Great Lakes anadromous fish, PCBs reduce reproduction below the level necessary to maintain a stable population (Bowerman et al, 1995). Similar effects could be occurring in other fish-eating species, but few long-term population studies are being conducted. Biomarkers are biochemical, physiological, or histological changes that measure effects of, or exposure to, toxic chemicals. Ecotoxicologists use biomarkers to clarify cause-effect relationships between toxic effects on the physiological and population levels and to provide tools for biomonitoring. At highly contaminated Great Lakes sites, organochlorines have been associated with altered liver MFO enzymes (i.e., cytochrome P-450's), liver porphyrins, thyroid mass and histology, and vitamin A homeostasis (Fox, 1993) . Many industrial chemicals, heavy metals, and pesticides found in the Great Lakes can suppress immune function in laboratory birds and mammals at concentrations comparable to those measured in wildlife at some Great Lakes sites (Thomas and Faith, 1985, Luster et al, 1987) . Exposing the developing immune system of mammals and birds to low levels of PHAHs (e.g., PCBs and 2,3,7,8-tetrachlorodibenzop-dioxin) has been shown to affect T-cell-mediated immunity. Toxic effects at several stages in T cell development contribute to thymic atrophy and to suppression of numerous T-cell functions (Clark et al, 1981, 1983; Vos and Luster, 1989; Tomar et al, 1991) . Exposure to PHAHs also suppresses antibody production in laboratory animals (Clark et al, 1981, 1983; Kerkvliet et al, 1990) and increases vulnerability to bacteria, viruses, and protozoan parasites (Friend and Trainer, 1970, Vos and Luster, 1989) . Similar effects have been suggested in wild marine mammals (Martineau et al, 1988; McGourty, 1988, Lahvis et al, 1995) . At highly contaminated sites in the Great Lakes, double-crested cormorants had an elevated incidence of eye infections associated with Pasteurella multocida (Ecological Research Services, 1991) . Children exposed perinatally to PCBs and dioxins in arctic Quebec experienced an increased incidence of middle ear infections (Dewailly et al, 1993) . Fish-eating birds are excellent wild species in which to study the immunotoxicity of PHAHs. Their high trophic level exposes them to high concentrations of contaminants that biomagnify. At some Great Lakes colonies, these birds are experiencing other effects associated with developmental exposure to PHAHs. Their colonial nesting habits provide large sample sizes and may facilitate the spread of disease. Reduced resistance to infections could reduce survival and contribute to the population-level impacts associated with current levels of PHAHs in Caspian terns and bald eagles. In a recent multiyear study (1991-1994), we found associations between PHAHs and suppression of T-cell-mediated immunity in herring gull and Caspian tem chicks at highly contaminated sites in the Great Lakes (Grasman, 1995; Grasman et al, 1996) . We also demonstrated that immunological and hematological variables are useful biomarkers for assessing contaminant-associated health effects in wild birds. Our investigations assessed immunocompetence using biomarkers of immunological structure (white blood cell counts; mass and histology of the thymus, bursa of Fabricius, and spleen) and biomarkers of immunological function [the PHA skin test for T-cell-mediated immunity and the sheep red blood cell (SRBC) hemagglutination test for antibody-mediated immunity]. A variety of contaminant concentrations, including coplanar PCBs, dioxin congeners, and DDE (1,1dichloro-2,2-bis-(/7-chlorophenyl)ethylene were measured in pooled liver or egg homogenates from each site. Alterations of white blood cell numbers were associated with organochlorine contaminants. In adult herring gulls, total white blood cell numbers and heterophil numbers increased as liver concentrations of DDE increased and as liver activity of ethoxyresorufin-Odeethylase (EROD; an index of Ah-receptor-mediated toxicity) decreased. Total lymphocytes increased as liver PCB concentrations increased. In adults, the heterophil to lymphocyte ratio decreased as liver EROD increased. In herring gull chicks, the heterophil to lymphocyte ratio increased as HG-TEQs (dioxin toxicity equivalents calculated using herring gull-specific induction equivalency factors) increased. This study showed a strong association between the burdens of organochlorine contaminants and suppressed T-cellmediated immunity. Immune function tests were conducted in herring gull and Caspian tern chicks at five sites for each species, with several sites replicated for three years. The PHA skin test has been shown to be a sensitive indicator of T-cell-mediated immunity in birds (Grasman et al, 1995) . Suppression of the PHA test was greatest at highly contaminated colonies in Lake Ontario and Saginaw Bay for both species and in western Lake Erie for herring gulls (Fig. 3). At high and low contamination sites where the PHA test was carried out for 2 or 3 years, the response was consistent among years. Suppression of this response was associated with several organochlorines including PCBs and DDE, but PCBs were found in the highest concentrations and were most closely associated with suppressed T-cell function. In herring gull chicks, the mass of the thymus gland, where T lymphocytes mature, decreased as liver EROD activity increased. Although thymus mass did not show any associations with individual organochlorines, the association of thymic atrophy with high EROD activity suggests that the complex mixtures of PHAHs affect the thymus by an Ahreceptor-mediated mechanism. There was no discernible as NCh • ^ S X v * ^ N % ^ ^ (27) WInn < ' (20) ^ " " - " w ^ Jonckheara T««t: P=0.0002 E sociation between contaminants and antibody-mediated immunity. However, in both herring gull and Caspian tern chicks, there were biologically significant differences among sites in anti-SRBC antibody titers. Contaminant effects on bursal mass were confounded by infestations with the fluke Cotylurus communis. The suppression of T-cell-mediated, but not antibody-mediated, immunity that we observed is consistent with the laboratory database on the immunotoxicity of PHAHs following perinatal exposure (Holliday and Luster, 1994) . Few other studies have examined immunotoxicological associations in wild birds, although several recent studies with wild \ < 1 15 HamH (13) WErte T (32) . . . A SagB (102) and captive marine mammals exposed to PHAHs have also found greater effects of PHAHs on T-cell-mediated immunity as compared to antibody-mediated immunity (de Swart etal, 1994, Lahvis et al., 1995, Ross et al., 1995) . Suppression of T-cell-mediated immunity also has been found in captive mink (Mustela visori) fed fish and water contaminated with bleached kraft mill effluent, which contains TCDD and chlorophenols (Smits et al., 1996) . Antibodymediated immunity was not measured in these mink. In conclusion, several studies with wild and captive wildlife species have shown associations between PHAHs and suppression of T-cell-mediated immunity. HOW CLOSE DO LABORATORY IMMUN0T0XIC0LOGY STUDIES COME TO PREDICTING POLLUTANTINDUCED EFFECTS IN FERAL POPULATIONS? (J. T. Zelikoff) In mammalian systems, variations in the integrity of disease resistance and the immune response are very sensitive indicators of toxic insult (Luster et al., 1988; Luster and Rosenthal, 1993) due, at least in part, to the complex nature of the immune system. While not all compounds within a given chemical class alter immune function, certain metals, such as cadmium (Cd), lead (Pb), and nickel (Ni), as well as some pesticides, PHAHs, and PAHs can act to modulate the immune response of a variety of animal species (including humans) at exposure levels below which other more commonly utilized endpoints of toxicity are not observed (Dean et al, 1986; Luster et al, 1988) . Chemically induced immune system alterations may be manifested as immunosuppression (possibly leading to increased susceptibility to challenge with infectious agents or tumor cells), autoimmunity, or hypersensitivity (Luster et al, 1988). To identify immune system changes brought about by chemical exposure, a panel of assays is typically used (Vos, 1980; Luster et al., 1988) . These assays have been organized into tiers to test for functional, as well as histopathological, parameters with the number and content of tiers dependent upon the particular reference examined. In general, tier I is composed of assays which consist of functional and pathological endpoints designed to identify potential immunotoxicants at concentrations which do not induce overt toxicity. Tier II further defines the immunotoxic effect and its potential mechanism(s) of action by examining such things as the specific cell type(s) affected, nonspecific immunity, and host resistance against challenge with bacteria, viruses, parasites, or tumors. The aforementioned assays, originally developed in rodents, have been adapted for use in a variety of animal species (Weeks et al., 1992) , including fish. In fact, some of the same criteria used to determine the immunotoxic potential of xenobiotics in mammalian systems can also be used to assess immunotoxicity in these alternate models. These include immunopathology (i.e., thymic atrophy), alterations in nonspecific (i.e., macrophage activity) and specific (i.e., antibody forming cell numbers) immune functions, and changes in immune system-regulated functions (i.e., host resistance challenge models). For example, our laboratory has employed immune assays that measure macrophage function (i.e., phagocytosis, oxyradical production, and activation), lymphoproliferative responses, host resistance against infectious agents, and antibody-forming cell (AFC) numbers, as well as circulating leukocyte counts and lymphoid organ cellularity and weights to assess metal-induced immunotoxicity in laboratory-reared Japanese medaka (Oryzias latipes); assays measuring anti-oxidant activity (i.e., catalase and superoxide dismutase) have also been used in this capacity (Zelikoff et al., 1996a) . Using the medaka model, our studies have demonstrated that waterborne exposure to Ni, at concentrations at or below those found in Ni-contaminated aquatic sites (U.S. Army, 1994) , significantly alters mitogen-stimulated lymphoproliferation in the kidneys of exposed 10-month old medaka (compared to controls). Con A-stimulated lymphoproliferative responses of medaka exposed in the laboratory to 50 and 125 ppb Ni for 7 days were reduced by ~ 2 5 % (compared to control values); a shorter exposure time (i.e., 1 day) also reduced Con A-stimulated blastogenesis, but only following exposure to the highest Ni concentration (Fig. 4A). Exposure for 1 day to 50 and 125 ppb Ni also reduced lymphoproliferation in response to allogeneic cell stimulation (Fig. 4B). Given that LPS-induced B-cell proliferation and AFC numbers were unaffected by Ni exposure (data not shown), Tcells appear to be more sensitive to the immunotoxic effects of Ni than are B-cell-mediated humoral responses in medaka. Interestingly, studies by Smialowicz et al. (1984) , in which mice were exposed to nickel chloride by intramuscular injection, suggest that, under the conditions used in their study, it is T-cell-dependent immune functions rather than B-cell responses which are primarily affected by Ni. Macrophage function has also proved to be a successful marker of metal-induced immunotoxicity in laboratory fish (Zelikoff et al., 1993, 1995, 1996b; Zelikoff, 1994; Anderson and Zeeman, 1995) . For example, in studies performed in our laboratory in which medaka kidney macrophage function was examined following waterborne exposure to Cd at the nominal concentrations of 6, 60, or 600 ppb (Zelikoff et al., 1996b) , results demonstrated that metal exposure for 5 days significantly enhanced (above control levels) the intracellular and extracellular production of superoxide (O2~), as well as that of hydrogen peroxide (H2O2); at the two lowest Cd concentrations, changes in macrophage function occurred in the absence 1.0 0.8 Nickel (ppb) of any changes in kidney cell number or viability, blood cell parameters (e.g., hematocrit, leukocrit), plasma immunoglobulin levels, or total plasma protein. Results with medaka, as well as laboratory investigations with "other fish species (Secombes et al., 1988; Warinner et al, 1988; Weeks et al., 1992) , support the usefulness of macrophage functions to predict immunotoxicity in a teleost system. A host resistance challenge assay was also performed in the aforementioned Cd studies. In these investigations, medaka were challenged intraperitoneally with an LD30 dose of the fish pathogen Yersinia ruckeri immediately following exposure to 6, 60, or 600 ppb Cd; in these studies, mortality due to Cd exposure alone was 0, ~ 1 5 % , and 45% at 6, 60, and 600 ppb Cd, respectively. While exposure to Cd at 6 and 60 ppb had no effect on susceptibility of the fish to bacterial challenge, exposure to 600 ppb Cd shortened time to death and increased the incidence of bacterially induced mortality compared to fish exposed to clean water or to water containing the two lower Cd concentrations (unpublished observations). Because some evidence has been presented which suggests that fish with increased antibody titers to bacterial pathogens may be more resistant to the toxic effects of environmental pollutants, a small pilot study was performed in our laboratory wherein medaka were challenged with an LD0—LDM concentration of Y. ruckeri and then exposed to Cd at 600 ppb. Interestingly, we noted that medaka challenged with a dose of bacteria which had no effect on fish survival (LD0) proved more resistant to the lethal effects of 600 ppb Cd (26% vs 40% mortality in unchallenged Cdexposed fish), while the incidence of mortality more than doubled in those fish challenged with an LD30 dose of Y. ruckeri (91% vs 40% mortality, respectively). These findings support the notion that a low, sublethal infection with a bacterial pathogen may provide some degree of "protection" to the host against the toxicity of aquatic pollutants. In addition, they also suggest that fish which survive a potentially lethal bacterial infection may be more susceptible to the lethality associated with exposure to a high concentration of aquatic metal pollutants. Further laboratory studies examining antibody titers in similarly challenged medaka are currently being performed to better understand the mechanisms by which the observed effects may have occurred. While the above studies demonstrate that certain immune functions are altered by metal exposure under laboratory conditions, the question remains as to the utility of the aforementioned endpoints for predicting toxicological hazards in feral populations exposed to environmental contaminants under "natural conditions." Studies performed in our laboratory with kidney cells recovered from feral fish populations have demonstrated that some of the same immune structural/ functional assays used with medaka to demonstrate the immunotoxic effects of metals, e.g., blood cellularity, hematocrit/leukocrit, phagocytosis, and oxyradical production, can be successfully employed to demonstrate immunotoxic effects of organochlorine-contaminated aquatic environments on feral populations of smallmouth bass (Anderson et al., 1997) . In these studies, Percoll-isolated kidney macrophages demonstrated reduced phagocytic uptake offish serum-opsonized latex particles, diminished extracellular O2" production, and increased numbers of blood granulocytes compared to fish taken from a clean water reference site; significant reductions in spleen- and liver-derived superoxide dismutase activity was also observed. Results from other field studies in which some of the same macrophage functional parameters were used to predict immunotoxicity in fish living in a PAH-contaminated site support our findings (Weeks and Warinner, 1986; Weeks et al., 1992) . In addition, studies by Dr. Faisal (presented previously in this Symposium) clearly demonstrate the applicability of immune endpoints other than macrophage function (e.g., NCC activity and lymphoproliferation) for successfully predicting PAH-associated immunotoxicity in wild populations of fundulus (Fundulus heteroclitus) and spot (Leiostomus xanthurus). Thus, while it is true that stresses such as capture, handling, and captivity can modulate immune responses either directly or indirectly via the neuroendocrine system (Tomasso et al., 1983; Ellsaesser and Clem, 1986; Flory, 1990) , immunological parameters such as those described above appear useful for predicting the immunotoxic effects of contaminated aquatic environments in feral fish populations. Some of these same immune parameters have also been used successfully to predict the immunotoxicity of polluted aquatic environments in feral populations of fish-eating birds and harbor seals. As discussed previously in this Symposium, Dr. Grasman, who examined biomarkers of immunological structure (i.e., white blood cell counts; mass and histology of the thymus, bursa of Fabricius, and spleen) and function (i.e., PHA skin test for T-cell-mediated immunity and SRBC hemagglutination test for antibody-mediated immunity) in fish-eating birds, demonstrated that birds eating PCB-contaminated fish had increased white blood cell numbers, decreased thymic mass, and suppressed T-cell-mediated immunity. Furthermore, Dr. Ross has shown that feral populations of seals fed herring from the PCB-contaminated Baltic Sea demonstrated significant reductions (compared to those fed herring from the relatively clean Atlantic Ocean) in natural killer cell activity, T-lymphocyte function, and delayed type hypersensitivity responses, as well as in host immunocompetence; results from these field studies were in general agreement with those laboratory studies performed by Dr. Ross with herring-fed PVG rats. As summarized in Table 1, PCB-associated immunotoxic effects observed in both of the aforementioned field studies appear to correlate well with rodent/wildlife studies performed in the laboratory under' controlled conditions. The purpose of this paper was to provide perspective regarding the utility of laboratory-developed immunotoxicological assays for predicting the toxicological hazards of environmental pollutants on directly exposed feral populations. While we recognize that comparing toxicant-induced responses in different species is difficult and should be approached with caution due to differences in such things as host metabolism and pharmacokinetics, results from wildlife studies presented in this symposium, coupled with those from other investigations, clearly demonstrate the usefulness of laboratory-developed immune assays to predict immunomodulating effects of xenobiotics in "real-world" situations. (Supported by U.S. Army Medical Research and Materiel Command, Contract No. DAMD-17-93-C-3059.) DISCLAIMER This report has been reviewed by the Environmental Protection Agency's Office of Research and Development and approved for publication. Approval does not signify that the contents necessarily reflect the views and policies of the Agency, nor does mention of trade names or commercial products constitute endorsement or recommendation for use. and Risk Assessment (F. J. Dwyer, T. R. Doane, and M. L. 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Robert W. Luebke, Peter V. Hodson, Mohamed Faisal, Peter S. Ross, Keith A. Grasman, Judith Zelikoff. Aquatic Pollution-Induced Immunotoxicity in Wildlife Species1, Toxicological Sciences, 1997, 1-15, DOI: 10.1093/toxsci/37.1.1