The results of the present study demonstrate increased chemical contaminant exposure in outmigrant juvenile chinook salmon during their relatively brief residence in polluted urban estuaries when compared to juvenile chinook salmon that reside in minimally contaminated estuaries. Chemical analyses showed concentrations of anthropogenic contaminants in stomach contents of salmon from two urban estuaries, the Duwamish Waterway and the Puyallup River estuary, were much greater than the concentrations in stomach contents of fish from the hatcheries or the nonurban Nisqually estuary. Moreover, there were consistently higher concentrations of hepatic PCBs and biliary FACs (an indicator of exposure to AHs) in juvenile salmon collected from the Duwamish Waterway and the Puyallup estuary than in salmon from the hatcheries of the Nisqually estuary. The exposure of juvenile salmon to chemical contaminants was sufficient to elicit responses at the biochemical level. For example, significantly higher hepatic P450 monooxygenase activities and xenobiotic-DNA adduct levels were observed in salmon collected from the Duwamish Waterway and the Puyallup estuary than in fish from the hatcheries or the Nisqually estuary. However, evidence for chemical contaminant exposure was not found in fish from all three urban estuaries studied. For example, fish from the Snohomish estuary showed very little evidence of chemical contaminant exposure. This finding is consistent with the documented differences in sediment contamination in these aquatic urban environments (i.e., the urban sediments of the Duwamish Waterway and the Puyallup estuary, particularly in the waterways, are significantly more contaminated than those in the Snohomish estuary (Stein et al. 1992)). Moreover, the level of chemical exposure in fish persisted for months for some contaminants. Specifically, the elevated body burden of PCBs were found to remain stable for at least 3 months after exposure in contaminated urban estuaries. Concomitant with the increased exposure of juvenile chinook salmon to contaminants in some of these urban estuaries was evidence of immune dysfunction, reduced survival, and possibly reduced growth relative to animals taken directly from the hatcheries or from the Nisqually estuary. Thus, brief exposure to chemical contaminants in urban estuaries appears to be of sufficent magnitude to impair significant biological processes in juvenile chinook salmon. The long-term consequences of these effects warrant future studies.
In the present study, fish from the Duwamish Waterway and the Puyallup estuary were of comparable size to fish sampled from their respective hatcheries. Moreover, there was often close agreement between the percent of tagged salmon from the hatchery and the percent of tagged salmon recovered in the estuary, suggesting that hatchery fish constituted the major portion of our sample from the estuary. In the Nisqually system, tagged juvenile salmon taken from the Nisqually estuary were confirmed to be from the Kalama Creek Hatchery. However, juvenile salmon from the estuary were much larger than juveniles from the Kalama Creek Hatchery, suggesting that juvenile salmon other than hatchery fish were also being sampled in the estuary. This finding, when coupled with allowances for mixing with juvenile salmon from the adjacent McAllister Creek system, suggests that chinook smolts captured in the Nisqually estuary reflect a more uniform mixture of natural and hatchery-produced salmon in the estuary. In contrast, it appears that many of the fish captured in the Snohomish estuary were not of Skykomish Hatchery origin. Beauchamps et al. (1987) report additional mixing of juvenile chinook from the Stillaguamish River, 25 km to the north of the Snohomish River estuary, and also had difficulty identifying the source of juvenile salmon captured in the Snohomish estuary.
Results of chemical analyses of stomach contents indicated diet as a possible route of xenobiotic exposure of juvenile salmon residing in polluted urban estuaries. The diet of juvenile chinook salmon from the Duwamish Waterway consisted of copepods, amphipods, insects, annelids and small fish; while chinook from the Puyallup River estuary (Milwaukee, Blair, and City Waterways) consumed similar prey, except annelids, and small fish were not found. In the Snohomish River estuary, salmon fed primarily on crustaceans and small fish and salmon from the Nisqually estuary preferred insects and annelids (P. Plesha, NMFS. Pers. commun., Sept. 1990). The results of taxonomic examination of stomach contents of juvenile salmon from the estuaries in the present study are supported by results of earlier studies (Meyer et al. 1981a,b; Weitkamp and Campbell 1980; Shreffler 1989), which showed preferences for copepods, chironomid insects, and gammarid amphipods by juvenile salmon in a number of Puget Sound estuaries. Many of these food organisms have been shown to be a potential source of contaminants in the food chain. For example, Brown et al. (1985) reported that amphipods collected in the Duwamish Waterway contained elevated concentrations of AHs (1.3 ± 0.20 µg/g dry wt.) and PCBs (0.24 ± 0.01 µg/g dry wt.), whereas AH and PCB concentrations in amphipods from a reference site were only 0.12 ± 0.04 and 0.01 ± 0.0006 µg per gram dry wt., respectively, indicating that amphipods can bioaccumulate anthropogenic contaminants from their environment. Moreover, uptake of AHs and PCBs from sediment by amphipods was demonstrated when two amphipod species, Rhepoxynius abronius and Eohaustorius washingtonianus, from clean sites were exposed to Duwamish Waterway sediments for 7 days (Reichert et al. 1985, Varanasi and Stein 1991). For salmon feeding on these and similar organisms in urban estuaries, diet may represent a significant route of exposure to sediment-associated AHs and CHs.
Accordingly, stomach contents of salmon collected in the present study were analyzed for selected chemicals to determine the potential for uptake of contaminants through the diet. Concentrations of AHs in stomach contents showed a very marked increase (>16,000-fold greater) in salmon from the Duwamish Waterway and sites in the Puyallup estuary compared to salmon from the hatcheries or the Nisqually estuary. Interestingly, the broad spectrum of AHs analyzed in salmon from the Duwamish Waterway and the Puyallup estuary included, in addition to unsubstituted AHs, a substantial contribution of alkylated AHs in both sampling years (Table 6). A single analysis of stomach contents of juvenile salmon from contaminated urban estuaries for all the alkylated AHs listed in Table 2 can result in the quantitation of hundreds of individual analytes (not shown). For example, analysis of chromatographic results of stomach contents of juvenile salmon from the Duwamish Waterway or the Puyallup estuary identified about 30 alyklated C3-phenanthrenes (Appendix Fig. A-1-2). Moreover, a comparison of analyses of stomach contents, which consist primarily of benthic and epibenthic invertebrates, in salmon from the urban sites and a sediment sample from the Duwamish Waterway showed marked similarities in profiles of the aklkylated C3-phenanthrenes (Appendix Fig. A-3). These results support findings of earlier studies (Varanasi and Stein 1991, Stein et al. 1992) that showed benthic organisms inhabiting contaminated urban estuaries accumulate many of the same chemicals associated with the sediment. This similarity in profiles of contaminants in sediment and stomach contents again emphasizes the importance of diet as a route of exposure to sediment-associated contaminants in pelagic fish.
Salmon from the Duwamish Waterway were held to assess the depuration of PCBs in fish from a PCB-contaminated site. The results showed little evidence for a substantial decline in the level of PCBs in the whole body (i.e., body burden) during 3 months of holding in filtered seawater in the laboratory. These results are consistent with an earlier study with rainbow trout showing that the body burden of PCBs did not change with increasing size or after prolonged starvation (Lieb et al. 1974) and indicating that the majority of PCBs are not readily excreted. The results suggest that PCBs accumulated by juvenile salmon during their residency in a contaminated estuary are retained after the salmon leave for the open ocean. Moreover, because PCBs are associated with numerous significant biological effects, such as immunological dysfunction, reproductive impairment, and cancer (Safe 1984), the retention of PCBs and potential for such chronic effects warrant further investigation.
In the present study, a suite of biochemical (hepatic monooxygenase activities and xenobiotic-DNA adduct levels) and chemical (hepatic PCBs and biliary FACs) indicators of xenobiotic exposure and sublethal effects were used to evaluate exposure of juvenile chinook salmon as they migrate out into urban estuaries. The use of these bioindicators to assess xenobiotic exposure in marine species is well documented (Stein et al. 1992, Collier and Varanasi 1991, Krahn et al. 1986, Varanasi et al. in press). The results of the present study showed that these bioindicators were generally responsive to differences in the level of contamination in salmon from the various urban and nonurban estuaries within Puget Sound. The findings also show that these indices gave consistent information regarding the amount of exposure juvenile chinook salmon were undergoing. For example, the bioindicators of chemical contaminant exposure for fish sampled from the Duwamish Waterway and the Puyallup estuary were consistently higher than for fish taken from the Nisqually estuary, whereas the bioindicators of exposure for salmon from the Snohomish estuary, another urban estuary, were only occasionally higher (biliary FACs) than for salmon from the Nisqually estuary. This is consistent with the documented differences in sediment contamination in these aquatic urban environments, (i.e., the urban environments of the Duwamish Waterway and the Puyallup estuary, particularly in the waterways, are significantly more contaminated than the Snohomish estuary (Stein et al. 1992)). In addition, some of the bioindicators of exposure (bile FACs and contaminants in stomach contents) showed that juvenile salmon sampled in the Nisqually estuary were, in general, less exposed to chemical contaminants than juveniles from the hatchery. This is interesting in that hatchery practices may also contribute to the exposure of juvenile salmon to contaminants, albeit to much lower levels than seen in the urban estuaries. However, the significance of this background exposure to the future well being of juvenile chinook salmon is not clear at this time.
Several studies (reviewed in Buhler and Williams 1989) have shown that exposure of fish to certain chemical contaminants, such as 4-5 ring AHs and certain PCB congeners, substantially induces (increases) hepatic cytochrome P-450-dependent monooxygenases. The monooxygenases are enzymes that catalyze biotransformation of anthropogenic chemicals to more polar compounds to enhance their excretion. However, in addition to the detoxication of toxic chemicals, the monooxygenases also activate some chemicals to more toxic metabolites that readily and covalently bind to cellular macromolecules (Lutz 1979, Conney 1982). Our results showing enhanced hepatic AHH activity in juvenile outmigrant chinook salmon from the Duwamish Waterway and the Puyallup estuary demonstrate that these fish were exposed to inducers, presumably the AHs and PCBs, in their environment. The biological consequences of induction of hepatic monooxygenases are not fully delineated; however, the induction of certain monooxygenases, such as the one measured by the AHH assay, are linked to the currently accepted mechanism for the toxic effects induced by environmental contaminants including coplanar PCBs, dibenz-p-dioxins, and dibenzofurans (Safe 1990).
The measurement of hepatic xenobiotic-DNA adducts to evaluate contaminant exposure was useful, but the levels were not consistent for each year of the study. The covalent binding of a chemical carcinogen to DNA is believed to be a critical step in the multistep process of chemical carcinogenesis (Swenberg et al. 1985, Poirier et al. 1991). Previous results with fish have shown that some of the adducts detected by the 32P-postlabeling assay, used here to measure DNA adducts, may be due primarily to AHs, some of which are known to be carcinogenic in fish (Hendricks et al. 1985, Schiewe et al. 1991). Additionally, our previous (Varanasi et al. 1989b, Stein et al. 1992) results demonstrate that hepatic xenobiotic-DNA adducts also provide a measure of cumulative exposure to compounds (e.g., AHs) which are not accumulated in tissues of fish because of their extensive metabolism to more easily excreted metabolites. Levels of hepatic xenobiotic-DNA adducts are unique because they not only are a measure of the exposure and binding of genotoxic agents to DNA, a critical cellular macromolecule, but they also persist in the tissues and thus can serve as a long-term and cumulative indicator of exposure to contaminants. Thus, while biliary FACs provide an estimate of recent exposure to AHs and their biotransformation, DNA adducts provide a long-term cumulative estimate of exposure to certain major classes of environmental contaminants.
Salmon from the Duwamish Waterway in 1989 had adduct levels that were 1.8 times greater than levels in salmon from the Duwamish Waterway in 1990. This is consistent with the lower hepatic PCB levels in fish sampled from the Duwamish in 1990, compared to 1989, indicating a lower magnitude or duration of exposure to this class of contaminants in fish in 1990 compared to fish in 1989. In contrast, hepatic DNA adduct concentrations in salmon from the Puyallup estuary were significantly higher than adduct concentrations in salmon from the hatcheries in 1990, but not in 1989 (Fig. 5). The higher levels of hepatic xenobiotic-DNA adducts in fish from the Puyallup estuary in 1990, compared to 1989, appears to be at least in part due to greater exposure to AHs, as reflected in contaminants in stomach contents and bile FACs. Differences in chemical contaminant exposure may be a result of where the fish were sampled. Salmon in 1989 were predominantly from the Milwaukee Waterway whereas in 1990 a majority of the fish were from the City Waterway.
Another factor that may contribute to differences in hepatic DNA-adduct levels in fish in 1989 and 1990 may be related to the sensitivity of the technique to document chemical contaminant exposure after a relative short exposure period. Based on our studies, benthic fish that reside in these contaminated environments for longer periods of time than salmon have hepatic xenobiotic-DNA adduct and PCB levels that are approximately 10 times higher than those observed in salmon from these same contaminated environments (Stein et al. 1992). Differences in adduct levels after long-term exposure periods in benthic resident fish are more easily discernable than adduct levels in migratory fish that reside in these urban environments for a shorter time span. Although benthic bottom fish and salmon represent different species, species differences among adduct levels for animals from the same environment or exposed to the same contaminants are very small (Stein et al. 1992). The fact that we can observe significant differences in hepatic xenobiotic-DNA adduct levels in fish from the contaminated urban estuaries after the relatively short exposure period reflects the strength of the method to serve as bioindicator of exposure to contaminants, as well as to confirm that significant exposure has occurred to salmon in these urban environments.
The biological effects that were monitored in this study demonstrated that salmon exposed to contaminants in an urban estuary, specifically the Duwamish Waterway, suffered greater mortality and exhibited greater immune dysfunction than salmon from the hatcheries or a minimally contaminated estuary. We also demonstrated that salmon from the Duwamish Waterway grew less than fish from the hatchery. A similar effect was observed for fish from the Nisqually estuary compared to fish from the hatchery; thus we could not separate an estuary from a contaminant-related effect on growth at this time. Although it is premature to extrapolate the significance of these findings to the long-term consequences, the implications are that juvenile salmon that must outmigrate through an urban estuary may be less equipped to face the multiple challenges of survival as they enter the ocean environment.
The immune system is responsible for protecting an organism against infectious diseases and neoplastic cells. The three main components of the immune system are cell-mediated immunity, humoral immunity, and macrophage function. All of these components have been shown to be perturbed by interaction with environmental contaminants (Dean et al. 1986). In the present study, we examined the effect of contaminant exposure on humoral immunity of juvenile chinook salmon. Alterations in the humoral immunity may result in a suppressed host ability to fight diseases. This is the first field study to examine the primary and secondary humoral immune response of juvenile chinook salmon from an urban estuary. Although there were no differences in the concentrations of naturally occurring immunoglobulins between juvenile chinook salmon from an urban estuary and juveniles from its respective hatchery, there was significant evidence of altered in vivo production of primary antibodies to the specific antigen TNP-KLH and in the secondary in vitro response of plaque-forming cells (B cells) to TNP-KLH and TNP-LPS. The altered primary in vivo response and suppressed secondary in vitro PFC response may be linked to differences in exposure to contaminants, as described previously.
The differences observed in the immune response of chinook salmon from the Duwamish Waterway and the Nisqually estuary may be due to factors other than contaminants. For example, genetic differences between salmon from these two river systems could account for the differences in their immune responses. However, salmon from the Green River Hatchery are cultured at the Kalama Creek Hatchery for release; therefore, strain differences are unlikely to be a factor. Another factor which could account for the differences in immune response is that the composition of salmon caught from the estuary is not representative of the salmon released from the hatchery and this results in population differences between the hatchery and estuary. However, we have shown that a large portion of hatchery salmon are captured in the estuary. It appears, therefore, that contaminant exposure plays an important role in bringing about a differential reponse in the immune system of juvenile salmon.
To more causally link the relationship between contaminants and altered immune function, salmon from the Green River Hatchery acclimated to saltwater in the laboratory were injected with an organic solvent extract of a contaminated sediment from the Duwamish Waterway (DWSE). As we observed in the field-exposed juvenile salmon, the ability of plaque-forming cells (B-cells) to produce specific antibodies to TNP-LP was suppressed. These findings (M. Arkoosh E. Clemons, and E. Casillas, NMFS, Unpubl. manuscr.) suggest that the immunosuppression observed with salmon from the Duwamish Waterway was most likely due to chemical contaminants and not due to other environmental variables. The concentrations of biliary FACs from DWSE-injected salmon were comparable to the values of biliary FACs in salmon sampled from urban estuaries, suggesting that the dose of DWSE administered to salmon in the laboratory study was environmentally relevant. Interestingly, the lymphoid organs affected in the laboratory and field studies were different. This difference may reflect differences in length of exposure of contaminants or differences in route of exposure (Ward et al. 1985). In general it appears that both field and laboratory exposure of juvenile salmon to contaminants have produced similar results, in that immunomodulation (suppression) of the in vitro secondary response was observed. Suppression of immunological memory after exposure to chemical contaminants is not without precedent. For example, a previous laboratory study demonstrated that rainbow trout exposed to aflatoxin B1 were also found to have a suppressed humoral secondary response (Arkoosh and Kaattari 1987). The consequences of a suppressed immune system are not clear at this time. Results of an earlier laboratory study where investigators injected a commercial PCB mixture into channel catfish (Ictalurus punctatus) showed that disease resistance to Aeromonas hydrophilia (Jones et al. 1979) was much reduced. We have yet to undertake this type of disease challenge with salmon passing through an urban estuary or exposed to selected environmentally relevant contaminants in the laboratory. But, because of the stresses encountered by these salmon entering a marine environment, perturbation of this important physiological system may have significant implications on their long-term health and survival.
During each year of the study, survival was lower in juvenile salmon from the urban estuaries compared to the survival of fish from the hatcheries or from the reference Nisqually estuary. Although extensive mortalities were found in all groups during the first year (1989) of the study (data not shown), including fish sampled from the hatcheries, mortalities of fish from two of the urban estuaries (the Duwamish Waterway and the Puyallup estuary) were always higher than for fish from the reference estuary (the Nisqually estuary) or the hatcheries. However, confidence in the interpretation of the data for 1989 was reduced because of the high mortality in all groups, including fish from the hatcheries (controls). During this first year, salmon were held for less than 30 days before mortalities reduced the survivorship to less than 30% in all groups. Mortalities during 1989 were, in part, attributed to mechanical problems with the seawater system leading to reduced water flows to the tanks and to an air leak that resulted in extensive gas bubble disease in all groups of salmon held for the survival and growth portion of the study. In 1990 and 1991, improvements in the holding facilities and care increased survival for all groups. For example, survival rates of juvenile salmon from the hatcheries and the Nisqually estuary were greater than 80% over an approximate 40- and 80-day period, respectively. Moreover, the mortality rates of fish from the Duwamish Waterway and Puyallup estuary in 1990 and Duwamish Waterway in 1991 were consistently higher than for fish from the Nisqually estuary or the hatcheries. The cause of the higher mortalities of salmon from the urban estuaries is unclear at this time, but it may be linked to increased or greater physiological stress (e.g.,immune dysfunction) in fish exposed to chemical contaminants in the urban environments.
The rate of growth of juvenile salmon from urban environments also appeared to be lower than that of fish from the hatcheries. However, the significance of this finding is uncertain because the growth rate of fish from the reference Nisqually estuary in 1990 was also lower than that of fish from the hatchery. Overall salmon from the Kalama Creek Hatchery and the Nisqually estuary grew much less than salmon from the Green River Hatchery and the Duwamish Waterway. One possible explanation for the reduced growth of fish from the Nisqually River was the presence of the parasite, Nanophyetus salmincola (Lee Harrell, NMFS. Pers. commun., July 1990). Nanophyetus salmincola is a digenic trematode that infects salmonids via the freshwater snails Juga plicifera (Bennington and Pratt 1960). This parasite infects the kidney of salmonids and may affect their survival in seawater. Nanophyetus salmincola was found to infect juvenile coho salmon (Oncorhynchus kisutch) in the Chehalis River system and cause extensive mortalites as they entered the estuarine environment (Schroder and Fresh 1992). The prevalence of this parasite in fish from the Nisqually system, particularly for juveniles from the estuary, was extremely high (M. Myers, NMFS, Pers. commun., July 1990). Thus, the reduced growth in fish from the Nisqually system may have been in part attributable to the high parasitic infection, although this has yet to be confirmed.
In addition, the growth of juvenile chinook salmon appeared to be affected by the number of fish in each tank. In the beginning of the studies, distributions of lengths and weights of fish were more uniform. However, toward the end of the studies, variability in length and weight between individuals was dramatically increased. A few of the individual salmon were very large at the conclusion of the studies while growth of many other fish was almost negligible. Salmon in tanks with high mortality of juveniles may actually experience higher rates of growth because of decreased fish densities. Fish densities would necessarily need to be continually adjusted to correct this problem. In this study, maintaining a constant number of fish was not possible, because the number of fish in the growth study was also used to determine survival. Reimers (1973) observed little change in size of chinook salmon in the Sixes River estuary from June through August, which he attributed to high densities. This problem was also recognized by Schroder and Fresh (1992) in a study of salmonid growth in the Chehalis estuary. Future studies should consider the impact of both stocking densities and parasitic infections on the measurement of growth of juvenile salmonids in order to more accurately assess the impact of xenobiotic exposure.
Because of the difficulties of holding juvenile chinook salmon for extended periods of time, much of our effort was spent in trying to improve our husbandry techniques and to identify other reference sites for accurate evaluation of the growth impairment observed in chinook salmon from the Duwamish Waterway. Improved survival was accomplished by using metomidate as an anaesthetic, placing the juvenile salmon in darkened tanks, and minimizing their handling. Using these modifications, the survival over a 90-day period was improved from less than 10% in 1989 to greater than 85% in 1991. Despite these difficulties, growth of fish from the Green River Hatchery was found to be always significantly greater than that of fish from the Duwamish Waterway. But because of a lack of growth in fish from the reference estuary, we are not able to unequivocally state that the rate of growth was less in fish as a result of exposure to contaminants in an urban estuary. Further studies are ongoing to measure growth rates in salmon from other reference estuaries in Puget Sound.
The data presented in this 3-year study represent new information evaluating chemical contaminant exposure and associated effects in juvenile chinook salmon in urban estuaries of Puget Sound, Washington. On the basis of this study, juvenile chinook salmon that outmigrate through several urban estuaries show extensive evidence of exposure to chemical contaminants. The level of exposure not only persists for months for some contaminants, but also the magnitude of the exposure is concordant with the severity of the contamination in the estuaries studied. This was supported by the concentration of chemical contaminants measured in the stomach contents and tissues as well as in the increased levels of the bioindicators of exposure (bile FACs, xenobiotic-DNA adducts, and AHH activity) in urban-exposed juvenile salmon. The bioindicators provided information about the exposure of juvenile chinook salmon to chemicals and provided a measure of sublethal biological effects, some of which may be early signs of more serious effects. Concomitant with the increased chemical exposure, juvenile chinook salmon inhabiting these urban estuaries exhibited evidence of immune alterations, reduced survival, and possibly impaired growth relative to juveniles taken directly from the hatcheries or from the nonurban estuary. In particular, the observed immunosuppression could weaken the fishes resistence to pathogens and increase vulnerability to a wide variety of diseases. Suppression of immune function could, in part, account for the decreased survival of juvenile salmon from the urban estuaries. Because it is difficult to accurately assess the impact of chemical contaminant exposure on the proportion of salmon returning from contaminated and minimally contaminated environments, the long-term consequences of xenobiotic exposure and the subsequent effects to the health and well-being of the salmon populations can not yet be precisely stated. What is needed now are laboratory studies using controlled exposures of fish to sediment contaminants either by using model compounds or mixtures of contaminants extracted from sediments. This controlled exposure will enable us to more closely link the contaminants in the environment to effects in juvenile salmon and to evaluate the longer term potential for survival, either by using disease challenges or behavioral studies. The effects of chemical contaminant exposure clearly occur after a brief residency in polluted estuaries and should be evaluated as a contributing factor affecting future salmon returns.
The work reported here was a result of efforts by an interdisciplinary team of NMFS scientists. Participants in the study included Nicolaus Adams, Bernie Anulacion, Luara Berggren, Ethel Blood, Lisa Bogatski, Richard Bogar, Daryle Boyd, Jennie Bolton, Cristin Bryant, Katherine Dana, Tho Dang, Tara Felix-Slinn, Barbara French, Bill Gronlund, Mark Gustafson, Rebecca Hastings, Deborah Holstad, Anna Kagley, Lyndal Johnson, Thomas Merculief, Hannah Morris, Mark Myers, Paul Olson, Ronald Pierce, Herb Sanborn, Eunice Schnell, Robert Snider, Sean Sol, Dana Whitney, and Gladys Yanagida. Thanks also are extended to Drs. Walt Dickoff and Orlay Johnson of the Coastal Zone and Estuarine Studies Division of the NMFS for providing PIT tags and the loan of the tag detector. Jim Peacock provided the map illustration. We would also like to thank Dr. Mike Watson and Jerry Larrance of the U.S. EPA for providing thoughtful and critical reviews of this document.
We are indebted to the following: Dr. S. L. Kaattari, Oregon State University for providing standard anti-TNP rainbow trout sera;.Dr. G. Warr, University of South Carolina for giving us the monoclonal cell line (1-14 monoclonal cell line); and Washington State Department of Fisheries for providing unpublished data and granting us authorization to obtain juvenile fall chinook salmon from the state-operated hatcheries. We also thank the following hatchery managers for their cooperation: Dutch Henderson and Don Peterson of the Green River Hatchery, Bob Jateff and Ron Warren of the Puyallup Hatchery, Don Rerdnick of the Skykomish Hatchery, and Bill Thomas and the Nisqually Tribe of the Kalama Creek Hatchery.
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