Relative Concentrations of Cytochrome P450-Active Organochlorine Compounds in Liver and Muscle of Rainbow Trout From Lake Ontario

J. Great Lakes Res. 18(4):759-765 Internat. Assoc. Great Lakes Res., 1992

RELATIVE CONCENTRATIONS OF CYTOCHROME P450-ACTIVE ORGANOCHLORINE COMPOUNDS IN LIVER AND MUSCLE OF RAINBOW TROUT FROM LAKE ONTARIO

David M. Janz 1 , Tracy L. Metcalfe, and Chris D. Metcalfe2

Environmental and Resource Studies Program Trent University Peterborough, Ontario K9J 7B8 and G. Douglas Haffner

Department of Biology, University of Windsor Windsor, Ontario N9B 3P4 ABSTRACT. The teleost liver is the primary target organ for the induction of cytochrome P450dependent monooxygenases by chemicals in fish. However, monitoring information on the residues of monooxygenase-inducing organochlorine compounds in fish generally consists of data on the concentrations in muscle tissue or the whole body. We determined the relative concentrations of cytochrome P450-active PCB congeners and organochlorine insecticides in the muscle and liver of rainbow trout (Oncorhynchus mykiss) sampled from the Ganaraska River on Lake Ontario. There were no sigmficant differences in the mean concentrations of 24 PCB congeners in muscle and liver tissues when concentrations were expressed on a wet weight or a lipid weight basis. Of the 7 other potential cytochrome P450-active organochlorine compounds quantified, only dieldrin was present at significantly higher concentrations in muscle than liver samples, and endrin was present at significantly higher concentrations in liver than muscle. The demonstration ofequivalent concentrations of most ofthese organochlorine compounds in liver and muscle suggests that data on contaminant levels in muscle may be used to evaluate the potential for induction of hepatic monooxygenases in salmonids. INDEX WORDS: PCBs, coplanar, monooxygenase, P450, pesticide, Lake Ontario, rainbow trout.

INTRODUCTION

in the Great Lakes has been to protect the health of humans consuming sportfish from these lakes (Swain 1983). Therefore, most monitoring data consists of information on the concentrations of persistent contaminants in the muscle or whole body of sportfish. However, another element of contaminant monitoring in the Great Lakes is directed at protection of the health of fish species. PCBs and organochlorine insecticides are acutely toxic to fish in mg/kg doses or at mg/L concentrations (Rand and Petrocelli 1985). At lesser doses relative to those causing acute toxicity, subtle toxic responses have been observed in fish. For instance, PCBs and several organochlorine insecticides are active in fish as inducers or modifiers of the activity of hepatic cytochrome P450dependent monooxygenases (Law and Addison

The contamination of fish species from Lake Ontario with PCBs and organochlorine pesticides has been well documented (Suns et al. 1983, Clark et al. 1984, Jaffe and Hites 1986, Oliver and Niimi 1988, Hallett 1988, Niimi and Oliver 1989a). The greatest contamination occurred during the late 1960s and early 1970s, but although inputs into the lake ecosystem have declined since then, many Lake Ontario fish species still contain /l-g/g concentrations of these persistent contaminants (Baumann and Whittle 1988, Borgmann and Whittle 1991). The primary emphasis of monitoring programs lCurrent address: Faculty of Pharmaceutical Sciences, University of British Columbia, Vancouver, B.C., Canada V6T lZ3. 2To whom correspondence should be addressed.

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1981, Vodicnik et al. 1981, Kleinow et al. 1987, Gooch et ai. 1989, Janz and Metcalfe 1991). There is some evidence that these biochemical changes also occur in Great Lakes fish in response to in situ exposure to persistent contaminants, including PCBs and organochlorine insecticides. Induction of a cytochrome P450-dependent monooxygenase has been observed in the hepatic tissues of lake trout sampled from western Lake Ontario (Luxon et al. 1987). Changes in the activity of these hepatic enzymes may induce chronic toxic effects in fish. For instance, increased activity of hepatic monooxygenases in wild fish from polluted environments has been associated with reduced reproductive potential (Munkittrick et ai. 1991, Casillas et al. 1991). There is also evidence that the induction or inhibition of monooxygenase activity alters chemical biotransformation pathways in teleosts, and therefore may alter the response of fish to other toxic or carcinogenic compounds (Erickson et al. 1986, Loveland et al. 1984). Since the teleost liver is a major target organ for the induction of monooxygenases (Kleinow et al. 1987), data on the concentration of contaminants in liver tissue may indicate the potential for monooxygenase induction in wild fish. However, very few monitoring data are available on the concentrations of persistent contaminants in the hepatic tissues of fish. The objective of this study was to determine the relative concentrations of 24 PCB congeners, including non-ortho substituted congeners 77, 126 and 169, and 7 other chlorinated organic chemicals in liver and muscle tissues of Lake Ontario rainbow trout (Oncorhynchus mykiss), and to determine whether there are intertissue differences in the distribution of these compounds. These data will be useful in determining whether organspecific monitoring data are necessary to properly assess the potential for monooxygenase induction, and associated chronic toxic effects, in Great Lakes salmonids. METHODS

Sampling

Nine rainbow trout (four males and five females), each weighing approximately 3 kg, were collected during their spawning run in the spring of 1989 at a fish ladder on the Ganaraska River. The Ganaraska River enters the eastern basin of Lake

Ontario at Port Hope, Ontario, Canada. A section of muscle tissue was removed from the dorsal area of each fish between the dorsal fin and the midline. Livers were excised after removal of the gall bladder. Samples were stored on ice in solvent-washed aluminum foil and frozen at -4°C within 6 hours of collection. All samples remained frozen until analysis. Sample preparation

Samples were prepared for analysis as described by Macdonald and Metcalfe (1990). Briefly, tissue samples of approximately 6 g were ground with sodium sulphate and extracted into hexane with a Soxhlet apparatus. Lipids were removed from the extract by gel permeation chromatography. Extracts were then fractionated into three subfractions by column chromatography with activated silica gel. The column was eluted with 40 mL of hexane to yield fraction A, which contained PCBs, DDE, hexachlorobenzene, heptachlor, mirex, photomirex, and octachlorostyrene. The column was then eluted with 40 mL of 25070 methylene chloride in hexane (v/v) to yield fraction B, which contained DDT, DDE, DDD, lindane, chlordane isomers and metabolites, and heptachlor epoxide. Further elution with 40 mL of 40% methylene chloride in hexane yielded fraction C, which contained dieldrin, endrin, and DDD. Further cleanup of liver samples from fractions Band C was performed by adding 0.1 mL of concentrated H 2S04 to the subfractions and vortexing for 1 minute. Non-ortho substituted PCB congeners were isolated from fraction A by column chromatography on a carbon/silica column, as described by Lazar et ai. (1992). Briefly, a 5% carbon in silica gel mixture was prepared from AX-21 activated carbon (Anderson Development Company) and silica gel (100-200 I'm mesh, Supe1co), and activated for 24 h at 130°C. A 2 cm bed of carbon/silica was added to a 0.6 cm x 20 cm glass column. The column was rinsed with toluene, methylene chloride, and hexane. An aliquot of extract (fraction A) was added to the top of the column, and the column was eluted with 30 mL of hexane to yield fraction 1. This fraction primarily contained diortho substituted congeners. The column was then eluted with 40 mL of methylene chloride to yield fraction 2, primarily containing mono-ortho substituted congeners. The column was then inverted and eluted with 30 mL of toluene to yield fraction 3. This final fraction contained the non-ortho sub-

P450-ACTIVE COMPOUNDS IN LIVER AND MUSCLE OF TROUT

stituted PCB congeners analyzed separately from other congeners. Chemical Analysis

All compounds were analyzed by high resolution gas chromatography using a Varian model 3500 GC with a 30 m DB-5 column and an electron capture (63Ni) detector. The GC conditions were as described by Macdonald and Metcalfe (1990). Muscle and liver samples were analyzed for the 24 PCB congeners and 7 other organochlorine compounds listed in Table 1. Quantification of all PCB congeners, except the non-ortho substituted compounds, was done as described by Macdonald and Metcalfe (1990) using standards purchased from the National Research Council, Halifax. The congeners analyzed represent approximately 500,70 by weight of the total congeners in Aroclors 1242, 1254, and 1260. Quantification of the non-ortho substituted congeners (77, 126, 169) was by comparison to standards prepared gravimetrically in our laboratory from individual compounds purchased from Ultra Scientific, Rhode Island. Quantification of mirex, photomirex, p,p'-DDT, p,p'-DDD, p,p'-DDE, endrin, and dieldrin was made by comparison to standards obtained from the Wildlife Toxicology and Surveys Branch, Canadian Wildlife Service, Hull. Quality Assurance

Procedural blanks were extracted and analyzed to provide quality assurance for analysis of individual PCB congeners. The "Limits of Detection" (LODs), determined by the method of Keith et al. (1983), were 0.3 ng/mL-l.O ng/mL for the diortho and mono-ortho substituted PCB congeners and individual organochlorine insecticides, and 0.01 ng/mL-O.03 ng/mL ng/mL for the nonortho substituted PCB congeners. Congeners 77, 126, and 169 were recovered from spiked samples with 1000,70 efficiency after enrichment on the carbon/silica column (Lazar et al. 1992). Data Treatment

Concentrations in ng per g wet weight and p-g per g lipid weight were calculated for each compound analyzed in the rainbow trout tissues. For each compound, the mean and standard deviation were determined from a sample size of nine fish, with the exception of the non-ortho PCB congeners 77, 126, and 169, where the sample size was three fish. Paired

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t-tests were used to compare differences between muscle and corresponding liver concentrations for each chemical. Statistical significance was measured at the 95% level of confidence (p < 0.05). RESULTS AND DISCUSSION

The rainbow trout collected for analysis averaged 2,989 ± 1,155 g (mean ± SD) in size. The age of all fish was estimated to be at least 5 y old, as determined from the age-length relationships for Ganaraska River rainbow trout (Ontario Ministry of Natural Resources, unpublished data). Lipid contents of the two tissues were similar, averaging 3.9% in muscle (2.0% - 6.3%) and 3.8% in liver (2.7% - 5.1%). All 24 PCB congeners and 7 organochlorine compounds analyzed were present at concentrations above limits of detection in both the muscle and liver tissues. The mean concentrations of total PCB congeners (sum of 24 congeners) were 751 ng/g and 556 ng/g wet weight in muscle and liver, respectively. These congener-specific data are consistent with mean PCB concentrations of approximately 1,500 ng/g previously reported in muscle tissue from Ganaraska River rainbow trout (Johnson et al. 1989), and indicate that the 24 congeners analyzed represent approximately half of total PCBs measured by packed column (low resolution) gas chromatography. The concentrations of total PCB congeners in the Ganaraska trout are also consistent with the concentrations of total PCB congeners (sum of 90 congeners) in the muscle tissue of rainbow trout from western Lake Ontario (Niimi and Oliver 1989a). The majority of PCBs analyzed in the trout tissues were di-ortho and mono-ortho substituted congeners. There were no significant differences (p < 0.05) between concentrations of individual di-ortho and mono-ortho substituted PCBs in muscle tissue and liver tissue, on either a wet weight or a lipid weight basis (Table 1). Congener 153 was present at the greatest concentration relative to other congeners, as it comprised approximately 17% of the total congeners in muscle and liver. Congeners 101, 110, 138, 151, and 180 were also major congeners analyzed in both muscle and liver tissue. The mono-ortho congeners 105, 118, and 156 together accounted for approximately 12% of the total congeners in the tissues. The relative proportions of PCB congeners reported here are very similar to the congener patterns reported for rainbow trout from western Lake Ontario

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TABLE 1. Mean concentrations (standard deviations in brackets), expressed on both a wet weight and lipid weight basis, for PCB congeners and other organochlorine compounds analyzed in muscle and liver tissue of Lake Ontario rainbow trout. Each value represents the mean of n = 9 fish, except for the non-ortho PCB congeners, where n = 3 fish. Significant differences between concentrations in muscle and liver are indicated by # for wet weight values, and <1> for lipid weight values. Compound

Concentration in Muscle wet weight (ng/g) lipid weight (p,g/g)

a) Mono- and di-ortho substituted PCB Congeners: 2.1 (1.0) 0.06 18 31(28)' (4.0) 10.5 0.30 52 19.6 (6.5) 0.57 11.1 (3.8) 49 0.33 (2.8) 44 8.2 0.23 62.2 (20.9) 1.88 101 87 29.3 (9.5) 0.88 22.5 (8.5) 0.71 105b (27.2) 110 87.8 2.64 64.6 (22.0) 1.96 118b (7.2) 19.9 0.61 151 (41.0) 153 135.0 4.10 96.5 (25.1) 2.91 138 (2.7) 8.4 0.24 156b (22.0) 180 66.2 2.03 (12.9) 1.06 170 34.3 24.1 (9.5) 0.75 201 26.9 (10.7) 0.84 196 (2.7) 0.21 6.6 195 (4.8) 11.6 0.36 194 (1.7) 0.12 3.8 209

(0.02) (0.15) (0.30) (0.19) (0.09) (1.18) (0.52) (0.22) (1.53) (1.25) (0.40) (2.43) (1.55) (0.12) (1.31) (0.75) (0.55) (0.62) (0.15) (0.27) (0.10)

b) Non-ortho substituted PCB congeners: (0.58) 77 3.32 (0.36) 126 0.86 (0.23) 169 0.35

0.085 0.022 0.009

(0.015) (0.009) (0.006)

c) Organochlorines: p,p'-DDE 115.0 p,p'-DDD 11.0 p,p'-DDT 15.7 mirex 66.0 photomirex 43.3 dieldrin 10.8 endrin 0.5

3.52 0.33 0.46 2.04 1.33 0.31 0.02

(1.59) (0.13) (0.14) (1.23) (0.89) (0.09)<1> (0.01)<1>

(11.6) (1.0) (2.4) (18.7) (15.7) (1.6)# (0.2)#

Concentration in Liver wet weight (ng/g) lipid weight (p,g/g) 2.0 10.1 13.9 8.0 5.3 44.9 22.9 19.3 65.4 49.2 15.8 93.4 72.6 6.1 45.4 27.8 19.4 19.5 5.2 8.3 1.8 2.96 0.68 0.30 105.0 7.9 10.3 43.9 31.5 0.3 1.5

(0.9) (5.8) (10.7) (7.0) (3.4) (39.4) (19.0) (7.2) (50.0) (40.8) (14.7) (59.1) (39.3) (3.4) (31. 7) (24.4) (17.0) (18.6) (4.4) (7.7) (1.6)

0.05 0.27 0.35 0.20 0.14 1.10 0.57 0.48 1.63 1.21 0.38 2.35 1.85 0.18 1.13 0.68 0.48 0.47 0.13 0.20 0.05

(0.02) (0.13) (0.18) (0.12) (0.06) (0.70) (0.34) (0.19) (0.87) (0.73) (0.27) (1.03) (0.71) (0.08) (0.57) (0.44) (0.31) (0.34) (0.08) (0.09) (0.03)

(0.42) (0.27) (0.08)

0.078 0.018 0.008

(0.011) (0.007) (0.002)

(39.8) (3.3) (3.7) (27.7) (22.9) (0.3)# (0.4)#

2.86 0.21 0.27 1.10 0.78 0.01 0.04

(1.08) (0.07) (0.06) (0.51) (0.41) (0.01)<1> (0.01)<1>

a) Two congeners coelute on DB-5 column b) Mono-ortho congeners

(Niimi and Oliver 1989a), except that in the latter study, concentrations of congener 101 exceeded congener 110. The non-ortho substituted PCBs analyzed included congeners 77, 126, and 169. These compounds were present in both liver and muscle tissues at concentrations greater than detection limits

(Table 1). Because of the additional cleanup step, concentrations could be reported to three significant figures. The concentration of congener 77 was greatest relative to the other non-ortho substituted congeners, followed by congener 126. No differences were seen between the concentrations of these compounds in liver and muscle, on a wet

P450-ACTIVE COMPOUNDS IN LIVER AND MUSCLE OF TROUT weight basis or on a lipid weight basis. The concentrations and relative proportions of these compounds are consistent with data reported by Niimi and Oliver (1989b) for rainbow trout from western Lake Ontario. In other basins of the Great Lakes, congener 126 may be present in fish at a greater concentration relative to congener 77 (Huckins et al. 1988). Seven chlorinated pesticides were also analyzed in muscle and liver tissue. DDE was present at greater concentrations than DDT and DDD (Table 1), which is consistent with other monitoring data for Lake Ontario salmonids (Clark et al. 1984, Oliver and Niimi 1988, Niimi and Oliver 1989a). With the exception of endrin and dieldrin, concentrations of all organochlorine compounds in the liver and muscle were not significantly different on a wet weight or a lipid weight basis. The mean dieldrin concentration was significantly greater in muscle than in liver (p < 0.05). Conversely, the mean concentration of endrin was significantly greater in liver than in muscle (p < 0.05). The reasons for these differences in the distribution of structurally similar dieldrin and endrin are unclear, but may be due to differences in the pharmacokinetics of each compound. Alternatively, the data may reflect epoxidative metabolism of dieldrin to endrin in the liver. The analysis of chlorinated hydrocarbons in liver tissue may have implications regarding the induction of hepatic cytochrome P450-dependent monooxygenases. Non-ortho substituted PCBs are strong inducers of hepatic aryl hydrocarbon hydroxylase (AHH) and 7-ethoxyresorufin-Odeethylase (EROD) activity in fish (Kleinow et al. 1987, Gooch et al. 1989, Janz and Metcalfe 1991). In addition, the mono-ortho substituted PCB, congener 118, is a moderately potent inducer of hepatic EROD and AHH activity in rainbow trout (Skaare et al. 1991). Therefore, hepatic concentrations of the non-ortho congeners 77, 126, and 169 and the mono-ortho congeners 105, 118, and 156 are particularily relevant to the analysis of monooxygenase inducing potential in Lake Ontario salmonids. Induction of AHH and EROD catalysts in fish have been related to enhanced activity of a P450 isozyme similar to the mammalian P450IAI enzyme (Stegeman 1989). The concentrations of di-ortho substituted congeners in the tissues of Ganaraska rainbow trout are also of interest, even though exposure to diortho PCBs appears to cause no change in activity, or a decline in activity of teleost monooxygenases

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(Vodicnik et al. 1981, Melancon et al. 1981). The induction of hepatic monooxygenases in fish seems to only occur in response to exposure to 3methylcholanthrene (MC) type inducers of the P450IA gene subfamily. The insensitivity of teleosts to phenobarbital (PB) type inducing compounds, which includes the di-ortho PCBs, may be related to the absence of isozymes in fish which are homologous to the mammalian P450IIB gene family (Stegeman 1989). In any case, contamination of fish by di-ortho PCB congeners is significant because these compounds may alter biotransformation pathways by causing a decline in the activity of some monooxygenases. In addition, mammalian studies have shown that di-ortho PCBs may act as antagonists to the induction of P450IAI catalysts by planar aromatic hydrocarbons (Biegel et al. 1989).lhus, contamination of Lake Ontario fish by a mixture of di-ortho, mono-ortho, and non-ortho substituted congeners may involve complex interactions influencing the expression of P450 genes. Data in the literature on the induction of monooxygenases in fish by mirex, dieldrin, and DDT and metabolites (DDD, DDE) are ambiguous. Kleinow et al. (1987) pointed out in their review that responses by teleosts to these PB-type inducing compounds range from slight induction of hepatic monooxygenases, to no effect, to inhibition of activity. Despite these ambiguities, the concentrations of these compounds in trout tissues were reported here because of their potential for altering the activity of monooxygenases in Lake Ontario salmonids. To our knowledge, the monooxygenase inducing potentials of photomirex and endrin have not been evaluated experimentally, but these compounds were analyzed in this study because of their close structural similarity to mirex and dieldrin, respectively. Melancon et al. (1989) injected rainbow trout with a radiolabelled PCB mixture and reported that hepatic EROD activity was only elevated when total PCB concentrations in muscle or liver exceeded 300 ng/g wet weight. Based upon these data, we would assume that hepatic monooxygenases would be induced above normal activity in the rainbow trout from the Ganaraska River. It was not practical to determine this directly, since the fish were in spawning condition. In conclusion, concentrations of PCB congeners and other organochlorine chemicals in Lake Ontario rainbow trout do not differ significantly between muscle and liver tissues. In the absence of

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data on hepatic concentrations of these compounds, it may be possible to substitute data on concentrations in muscle to evaluate the potential for alterations to monooxygenase activity in salmonids from contaminated environments; particularily, if the data are lipid normalized.

ACKNOWLEDGMENTS Jill Duncan assisted on the preparation of samples and Mark Nanni assisted on the GC analysis. Arnie O'Donnell (Ontario Ministry of Natural Resources), Chris Williams, and Nakos Kotsanis assisted in collecting fish. We thank Art Niimi (Fisheries and Oceans, Burlington) for his suggestions on the manuscript and his advice. This work was supported by a Subvention Grant from the Department of Fisheries and Oceans, and a Great Lakes University Research Fund grant from the Department of Environment and the Natural Sciences and Engineering Research Council.

REFERENCES Baumann, P. C., and Whittle, D. M. 1988. The status of selected organics in the Laurentian Great Lakes; an overview of DDT, PCBs, dioxins, furans, and aromatic hydrocarbons. Aquatic Toxicology 11:241257. Beigel, L., Harris, M., Davis, D., Rosengren, R., Safe, L., and Safe, S. 1989. 2,2'4,4',5,5'-hexachlorobiphenyl as a 2,3,7,8-tetrachlorodibenzo-p-dioxin antagonist in C57BL/6J mice. Toxicol. Appl. Pharmacol. 97:561-571. Borgmann, V., and Whittle, D. M. 1991. Contaminant concentration trends in Lake Ontario lake trout (Salvelinus namaycush): 1977 to 1988. J. Great Lakes Res. 17:368-381. Casillas, E., Misitano, D., Johnson, L. L., Rhodes, L. D., Collier, T. K., Stein, J. E., McCain, B. R, and Varanasi, V. 1991. Inducibility of spawning and reproductive success of female English sole (Paraphorys vetulus) from urban and nonurban areas of Puget Sound, Washington. Marine Environ. Res. 31:99-122. Clark, J. R., DeValut, D., Bowden, R. J., and Weishaar, J. A. 1984. Contaminant analysis of fillets from Great Lakes coho salmon, 1980. J. Great Lakes Res. 10:38-47. Erickson, D. A., Goodrich, M. S., and Lech, J. J. 1986. The effect of ,8-napthoflavone and piperonyl butoxide on hepatic monooxygenase activity and the toxicity of rotenone to rainbow trout, Salmo gairdneri. Toxicologist 6: 160-161. Gooch, J. W., Elskus, A. A., Kloepper-Sams, P. J., Kahn, M. E., and Stegeman, J. J. 1989. Effects of ortho and non-ortho substituted polychlorinated bi-

phenyl congeners on the hepatic mono-oxygenase system in scup (Stenotomus chrysops). Toxicol. Appl. Pharmacol. 98:422-433. Hallet, D. J. 1988. Ecosystem management of persistent toxic chemicals - back to basics. In Toxic Contaminants and Ecosystem Health: A Great Lakes Focus, ed. M.S. Evans, pp. 491-509. John Wiley and Sons. Huckins, J. N., Schwartz, T. R., Petty, J. D., and Smith, L. M. 1988. Determination, fate, and potential significance of PCBs in fish and sediment samples with emphasis on selected AHH-inducing congeners. Chemosphere 17:1995-2016. Jaffe, R. and Hites, R. A. 1986. Anthropogenic, polyhalogenated organic compounds in non-migratory fish from the Niagara area and tributaries to Lake Ontario. J. Great Lakes Res. 12:63-71. Janz, D. M., and Metcalfe, C. D. 1991. Relative induction of aryl hydrocarbon hydroxylase by 2,3,7,8TCDD and two coplanar PCBs in rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 10:231-239. Johnson, A. E, Cox, C. M., and Vaillancourt, A. L. 1989. Contaminants in Ontario sport fish: Long term trends and future prospects. In Proc. Technology Transfer Conference, pp. 285-291. Ontario Ministry of the Environment. Keith, L. H., Crummett, W., Deegan, J., Libby, R. A., Taylor, J. K., and Wentler, G. 1983. Principles of environmental analysis. Anal. Chem. 55:2210-2218. Kleinow, K. M., Melancon, M. J., and Lech, J. J. 1987. Biotransformation and induction: Implications for toxicity, bioaccumulation and monitoring of environmental xenobiotics in fish. Env. Health Persp. 71 :105-119. Law, F. C. P. and Addison, R. F. 1981. Response of trout hepatic mixed-function oxidases to experimental feeding of ten known or possible chlorinated environmental contaminants. Bull. Environ. Contam. Toxicol. 27:605-609. Lazar, R., Edwards, R. C., Metcalfe, C. D., Metcalfe, T. L., Gobas, F. A. P. C., and Haffner, G. D. 1992. A simple, novel method for the quantitative analysis of coplanar (non-ortho substituted) polychlorinated biphenyls in environmental samples. Chemosphere 25:493-504. Loveland, P. M., Nixon, J. E., and Bailey, G. S. 1984. Glucuronides in bile of rainbow trout (Salmo gairdnen) injected with PH] aflatoxin B1 and the effects of dietary ,8-napthoflavone. Compo Biochem. Physiol. 78C:13-19. Luxon, P. L., Hodson, P. V., and Borgmann, V. 1987. Hepatic aryl hydrocarbon hydroxylase activity of lake trout (Salvelinus namaycush) as an indicator of organic pollution. Environ. Toxicol. Chem. 6:649-657.

P450-ACTIVE COMPOUNDS IN LIVER AND MUSCLE OF TROUT Macdonald, C. R. and Metcalfe, C. D. 1990. Concentration and distribution of PCB congeners in isolated Ontario lakes contaminated by atmospheric deposition. Can. J. Fish. Aquat. Sci. 48:371-381. Melancon, M. J., Elcombe, C. R., Vodicnik, M. J., and Lech, J. J. 1981. Induction of cytochromes P450 and mixed-function oxidase activity by polychlorinated biphenyls and l3-napthoflavone in carp (Cyprinis carpio). Compo Biochem. Physiol 69C:219-226. _ _ _ _ , Turnquist, K. A., and Lech, J. J. 1989. Relation of hepatic microsomal monooxygenase activity to tissue PCBs in rainbow trout (Salmo gairdnen) injected with [l4C]PCBs. Environ. Toxicol. Chem. 8:777-782. Munkittrick, K. R., Portt, C., Van der Kraak, G. J., Smith, I. R., and Rokosh, D. 1991. Impact of bleached kraft mill effluent on liver MFO activity, serum steroids, and population characteristics of a Lake Superior white sucker population. Can. J. Fish. Aquat. Sci. 48:1371-1380. Niimi, A. J. and Oliver, B. G. 1989a. Distribution of poly-chlorinated biphenyl congeners and other halocarbons in whole fish and muscle among Lake Ontario salmonids. Environ. Sci. Technol. 23:83-88. ____ , and Oliver, B. G. 1989b. Assessment of relative toxicity of chlorinated dibenzo-p-dioxins, dibenzofurans, and biphenyls in Lake Ontario salmonids to mammalian systems using toxic equivalent factors (TEF). Chemosphere 18:1413-1423. Oliver, B. G., and Niimi, A. J. 1988. Trophodynamic analysis of polychlorinated biphenyl congeners and

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other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ. Sci. Tech. 22:388-397. Rand, G. M. and Petrocelli, S. R. 1985. Fundamentals of Aquatic Toxicology. Washington: Hemisphere Publ. Corp. Skaare, J. V., Jensen, E. G., Goks16r, A., and Egaas, E. 1991. Response of xenobiotic metabolizing enzymes of rainbow trout (Oncorhynchus mykiss) to the mono-ortho substituted polychlorinated PCB congener 2,3',4,4',5-pentachlorobiphenyl, PCB-118, detected by enzyme activities and immunochemical methods. Arch. Environ. Contam. Toxico/. 20:349352. Stegeman, J. J. 1989. Cytochrome P450 forms in fish: Catalytic, immunological and sequence similarities. Xenobiotica 19: 1093 Suns, K., Craig, G. R., Crawford, G., Rees, G. A., Tosine, H., and Osborne,' J. 1983. Organochlorine contaminant residues in spottail shiners (Notropis hudsonicus) from the Niagara River. J. Great Lakes Res. 9:335-340. Swain, W. R. 1983. An overview of the scientific basis for concern with polychlorinated biphenyls in the Great Lakes. In PCBs: Human and Environmental Hazards, ed. F.M. D'Itri and M.A. Kamrin, pp. 11-49. New York: Butterworth. Vodicnik, M. J., Elcombe, C. R., and Lech, J. J. 1981. The effect of various types of inducing agents on hepatic microsomal monooxygenase activity in rainbow trout. Toxicol. Appl. Pharmacol. 59:364-374.

Submitted: 30 March 1992 Accepted: 10 August 1992