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Dose-response analysis of the effects of Aroclor 1260 treatment on immune and endocrine endpoints in… Aloysius, Herve 2001

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DOSE-RESPONSE ANALYSIS OF THE EFFECTS OF AROCLOR 1260 TREATMENT ON IMMUNE AND ENDOCRINE ENDPOINTS IN ADULT M A L E RATS by HERVE ALOYSIUS B.Sc, University of British Columbia, Vancouver, British Columbia, Canada, 1997 A THESIS SUBMITTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF MASTER OF SCIENCE in THE FACULTY OF GRADUATE STUDIES FACULTY OF PHARMACEUTICAL SCIENCES Divisions of Pharmacology & Toxicology and Pharmaceutical Chemistry We accept-this4hesis as conforming to the requ^eiLstandard THE UNIVERSITY OF BRITISH COLUMBIA November 2001 © Herve Aloysius, 2001 In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department or by his or her representatives. It is understood that copying or publication of this thesis for financial gain shall not be allowed without my written permission. Department of The University of British Columbia Vancouver, Canada Date / W JLI, &O0 I DE-6 (2/88) ii ABSTRACT Polychlorinated biphenyls (PCBs) are prevalent environmental contaminants with a wide range of biological effects in wildlife, experimental animals, and humans. In order to quantify the toxicity associated with environmental PCB exposure, endocrine and immune endpoints were evaluated in a dose-response study using Aroclor 1260. Nine-week old male rats were treated with Aroclor 1260 in corn oil at dosages ranging from 0.025 to 156 mg/kg/day. Rats were treated by oral gavage for 10 consecutive days and killed two days after the last treatment. Eight days prior to sacrifice, rats received sheep red blood cells (2x108 SRBCs) by iv injection. Blood for determination of anti-SRBC IgM titer by enzyme-linked immunosorbent assay (ELISA) was collected 2 days prior to sacrifice, and at the time of death. Hepatic microsomes were prepared for subsequent cytochrome P450 (CYP) determinations. Only rats treated with the highest dosage of Aroclor 1260 exhibited decreased body weight gain. Aroclor 1260 had no effect on thymus, testis, ventral prostate or seminal vesicle weights, but liver weight was increased in rats treated at dosages greater than 0.13 mg/kg/day. Treatment with Aroclor 1260 decreased serum thyroxine (T4) levels in a dose-dependent manner, but serum luteinizing hormone (LH) and testosterone levels were unchanged. Aroclor 1260 also elicited dose-dependent suppression of humoral immunity as assessed by anti-SRBC IgM titer. In addition, total hepatic C Y P content was increased at dosages greater than 1.25mg/kg/day. Immunoblot analysis indicated a dose-dependent induction of hepatic CYP2B enzymes up to a dosage of 15.6 mg/kg/day. Hepatic CYP1A protein content was also induced by Aroclor 1260 at dosages greater than 3.13mg/kg/day, but to a lesser degree than CYP2B. Determination of the various testosterone hydroxylase activities in hepatic microsomes iii revealed that hepatic CYP2B-, CYP3A-, and CYP2A1 -mediated activities were induced whereas CYP2C11 was suppressed. E D 5 0 values of 0.66, 1.52, 3.99, 4.69, 5.38, 6.36 and 21.3 mg/kg/day were determined for total hepatic CYP induction, humoral immunity suppression, hepatic CYP2B2 induction, decrease in serum T4 levels, hepatic CYP2B1 induction, relative liver weight increase, and hepatic CYP1A2 induction, respectively. In summary, the results indicate that significant immunosuppressive, endocrine, and biochemical effects are produced in male rats by exposure to relatively low dosages (less than 6.25 mg/kg/day) of Aroclor 1260. iv TABLE OF CONTENTS Title Page page Abstract i i Table of Contents iv List of Tables vii List of Figures ix Appendices xi List of Abbreviations xii Acknowledgements xiv 1. INTRODUCTION 1 1.1. Physical chemical properties of PCBs 1 1.2. Biotransformation of PCBs 5 1.3. Biological effects of PCBs 9 1.3.1. Effects of PCBs on immune function 10 1.3.1.1. Effects of PCB treatment on immune function 10 1.3.1.2. Mechanism of action of PCBs as immunotoxicants 12 1.3.2. Effects of PCBs on endocrine function 15 1.3.2.1. Effects of PCB treatment on gonadotropin and sex hormone levels 15 1.3.2.2. Estrogenic activity of PCBs 19 1.3.2.3. Effects of PCBs on thyroid hormone homeostasis and function 21 1.3.2.4. Mechanisms of PCB-induced thyroid hormone modulation 23 1.3.3. Effects of PCBs on cytochrome P450 (CYP)enzymes 26 1.3.3.1. Effects of PCB treatment on C Y P induction 27 1.3.3.2. Mechanism of action of PCBs as C Y P inducers 30 1.3.4. Strain, species, and sex differences in effects elicited by PCBs 32 1.3.5. Summary of dose-response trends 33 1.4. Environmental PCB exposure 36 1.4.1. Wildlife exposure 36 1.4.2. Human exposure 41 1.5. Hypotheses 44 1.6. Objectives 44 2. E X P E R I M E N T A L 45 2.1. Reagents 45 2.2. Preparation of SRBC membrane antigens 47 2.3. Animal treatment 48 2.4. Sample handling and serum anti-SRBC IgM analysis by ELISA 49 2.5. Hepatic microsome preparation 50 2.6. Determination of serum T4 and testosterone concentrations by radioimmunoassay 51 V 2.7. Determination of total C Y P concentration 51 2.8. Determination of total protein concentration 52 2.9. Sodium dodecyl sulphate-polyacrylamide gel electrophoresis (SDS-PAGE) 52 2.10. Immunoblot analysis 5 3 2.11. Testosterone hydroxylase assay 55 2.12. Statistical analysis 57 2.13. Assay validation 58 2.13.1. Validation of the ELISA assay to measure serum anti-SRBC IgM titer 59 2.13.2. Validation of the radioimmunoassays 61 2.13.2.1. Validation of the T4 radioimmunoassay 61 2.13.2.2. Validation of the testosterone radioimmunoassay 61 2.13.3. Validation of the immunoblot analysis for CYP1A and CYP2B protein levels 62 2.13.4. Validation of the testosterone hydroxylase assay 68 3. RESULTS 85 3.1. Effects of the vehicle (corn oil) 85 3.2. Dose-response effects of Aroclor 1260 treatment 88 3.2.1. Effects of Aroclor 1260 treatment on body weight gain 88 3.2.2. Effects of Aroclor 1260 treatment on organ weights 88 3.2.3. Effects of Aroclor 1260 treatment on antibody response 91 3.2.4. Effects of Aroclor 1260 treatment on serum testosterone levels 94 3.2.5. Effects of Aroclor 1260 treatment on serum T4 levels 94 3.2.6. Effects of Aroclor 1260 treatment on total hepatic C Y P content 98 3.2.7. Effects of Aroclor 1260 treatment on specific hepatic C Y P isozymes 100 3.2.7.1. Effects of Aroclor 1260 on hepatic C Y P protein levels 100 3.2.7.1.1. Effects of Aroclor 1260 on CYP2B protein levels 100 3.2.7.1.2. Effects of Aroclor 1260 on CYP1A protein levels 105 3.2.7.2. Effects of Aroclor 1260 on CYP-dependent enzyme activities 109 4. DISCUSSION 119 4.1. Dose-response effects of Aroclor 1260 treatment on body weight gain 119 4.2. Dose-response effects of Aroclor 1260 treatment on organ weights 120 4.3. Dose-response effects of Aroclor 1260 treatment on antibody response 125 4.4. Dose-response effects of Aroclor 1260 treatment on serum gonadotropin and sex hormone levels 127 4.5. Dose-response effects of Aroclor 1260 treatment on serum thyroxine (T4) levels 128 4.6. Dose-response effects of Aroclor 1260 treatment on hepatic C Y P enzyme levels 131 4.7. Summary of results 136 4.8. Conclusions 137 4.9. Future studies 138 5. REFERENCES 140 6. APPENDICES 162 6.1. Appendix I 162 6.2. Appendix II 163 6.3. Appendix III 165 6.4. Appendix IV 166 vii LIST OF TABLES Table page 1. IUP A C designation of PCB congeners 3 2. Congener composition (weight percent of chlorobiphenyl) of selected Aroclor mixtures 5 3. Major cytochrome P450 inducers 27 4. Summary of dose-response effects of PCBs in rodents 35 5. Total PCB concentrations in milk from Inuit women and in fatty tissues of species from the arctic Quebec food web 37 6. PCB levels in the arctic polar bear 38 7. Geographical regions used to describe PCB levels among polar bear populations 39 8. Quantitative and qualitative analysis of PCBs in Aroclor 1260 and a human breast milk extract 43 9. Determinations of the anti-SRBC IgM titer of the control sample on the same day 61 10. Calibration curves for CYP1A and CYP2B protein determinations 63 11. Intra-assay coefficients of variation for the immunoblot analysis of C Y P 1A and CYP2B enzymes 67 12. Inter-assay coefficients of variation of the immunoblot analysis of C Y P 1A and CYP2B enzymes 67 13. Mean A U C ratios of the testosterone metabolite standards at various concentrations 71 14. Intra-assay coefficient of variation (%) of the testosterone hydroxylase assay for testosterone metabolite standards 77 15. Inter-assay coefficient of variation (%) of the testosterone hydroxylase assay for testosterone metabolite standards 78 16. Linearity of the testosterone hydroxylase assay with respect to protein (enzyme) concentration 80 vii i 17. Effects of treatment with corn oil on body weight gain, and liver weight in adult male rats 85 18. Effects of treatment with corn oil on absolute thymus, testes, seminal vesicle, and prostate weights in adult male rats 85 19. Effects of treatment with corn oil on serum anti-SRBC IgM levels, testosterone concentrations and T4 concentrations in adult male rat 86 20. Effects of treatment with corn oil on protein concentration and total C Y P content in hepatic microsomes from adult male rats 86 21. Effects of treatment with corn oil on hepatic CYP2B and CYP1A protein levels 86 22. Effects of treatment with corn oil on testosterone hydroxylase activities in hepatic microsomes from adult male rats 87 23. Effects of treatment with Aroclor 1260 on body weight gain, and liver weight in adult male rats 89 24. Effects of treatment with Aroclor 1260 on absolute thymus, testes, seminal vesicle, and prostate weights in male rats 90 25. Effects of treatment with Aroclor 1260 on serum anti-SRBC IgM units in adult male rats 92 26. Effects of treatment with Aroclor 1260 on serum testosterone concentration in adult male rats 95 27. Effects of treatment with Aroclor 1260 on serum T4 concentration in adult male rats 95 28. Effects of treatment with Aroclor 1260 on protein concentration and total C Y P content in hepatic microsomes from adult male rats 99 29. Effects of treatment with Aroclor 1260 on specific hepatic CYP2B and CYP1A content 102 30. Effects of treatment with Aroclor 1260 on hepatic CYP-mediated testosterone hydroxylase activities in adult male rats 113 31. Summary of dose-response effects on immune and endocrine parameters in adult male rats 136 LIST OF FIGURES Figure page 1. Chemical structure of PCBs 2 2. Metabolism of PCBs 8 3. Regulation of gonadrotropin secretion 17 4. Molecular pathway of estrogen action 18 5. Chemical structure of estradiol and hydroxylated PCBs 20 6. Chemical structures of thyroid hormones 25 7. Study design 49 8. Regiospecific hydroxylation of testosterone by C Y P enzymes 55 9. Determination of the anti-SRBC IgM titer to yield 1.0 OD of the control sample 60 10. Mean calibration curves for quantitation of CYP1A1, CYP1A2, and CYP2B enzymes 64 11. Representative chromatogram of a mixture of hydroxylated testosterone metabolites 70 12. Calibration curves of testosterone metabolites 72 13. Linearity of the testosterone hydroxylase assay: effects of increasing protein (total CYP) concentration 81 14. Effects of treatment with Aroclor 1260 on humoral immunity response in adult male rats 93 15. Effects of treatment with Aroclor 1260 on serum testosterone concentration in adult male rats 96 16. Effects of treatment with Aroclor 1260 on serum T4 concentration in adult male rats 97 17. Representative immunoblot of hepatic microsomes from adult male rats (probed for CYP2B) 101 18 Effects of Aroclor 1260 treatment on hepatic CYP2B1 protein levels in adult male rats 19. Effects of Aroclor 1260 treatment on hepatic CYP2B2 protein levels in adult male rats 20. Representative immunoblot of hepatic microsomes from adult male rats (probed for CYP1 A) 21. Effects of Aroclor 1260 treatment on hepatic CYP1 A l protein levels in adult male rats 22. Effects of Aroclor 1260 treatment on hepatic CYP1A2 protein levels in adult male rats 23. Representative chromatogram of testosterone hydroxylation reactions catalyzed by hepatic C Y P enzymes in a corn oil-treated adult male rat 24. Representative chromatogram of testosterone hydroxylation reactions catalyzed by hepatic C Y P enzymes in an adult male rat after treatment with Aroclor 1260 25. Effects of treatment with Aroclor 1260 on hepatic CYP-mediated testosterone hydroxylase activities in adult male rats xi APPENDICES Table page 1.1 .Effects of treatment with corn oil on relative thymus, testes, seminal vesicle, and prostate weights in adult male rats 162 I. 2.Effects of treatment with Aroclor 1260 on relative thymus, testes, seminal vesicle, and prostate weights in adult male rats 162 II. 1 .Effects of treatment with corn oil on serum L H and TSH levels in adult male rats 163 II. 2.Effects of Treatment with Aroclor 1260 on serum L H levels in adult male rats 164 III. 1 .Effects of corn oil treatment on hepatic CYP1A and CYP2B protein levels in adult male rats 165 III. 2.Effects of corn oil treatment on hepatic CYP1A and CYP2B protein levels in adult male rats 165 IV. 1.Effects of corn oil treatment on hepatic CYP-mediated testosterone hydroxylase activities in adult male rats 166 V.2.Effects of Aroclor 1260 treatment on hepatic CYP-mediated testosterone hydroxylase activities in adult male rats 167 xii LIST OF ABBREVIATIONS A h aryl hydrocarbon AP alkaline phosphatase A U C area under the curve C A R constitutively active receptor C Y P cytochrome P450 D A G 1,2-^«-diacylglycerol D T H delayed-type hypersensitivity ELISA enzyme-linked immunosorbant assay EROD ethoxyresorufin O-dealkylase FSH follicle-stimulating hormone GnRH gonadotropin-releasing hormone IPs inositol-1,4,5-triphosphate L H luteinizing hormone L O A E L lowest observed adverse effect level L O E L lowest observed effect level M A P K mitogen-activated protein kinase M C 3 -methylcholanthrene N H R nuclear hormone receptor N K natural killer NIOSH National Institute for Occupational Safety and Health N P L National Priority List N O A E L no observed adverse effect level ORD outer-ring deiodinase PB phenobarbital P B R E M phenobarbital-responsive enhancer module PCBs polychlorinated biphenyls P C N pregnenolone 16a-carbonitrile PFC plaque-forming cell PIP 2 phosphatidylinositol-4,5-biphosphate PL phospholipase PROD pentoxyresorufin O-dealkylase P X R pregnane x receptor Rb retinoblastoma RBF-1 retinoblastoma binding factor 1 RIA radioimmunoassay R X R retinoid x receptor SDS-PAGE sodium dodecyl sulfate polyacrylamide gel electrophc SRBC sheep red blood cell T3 triiodothyronine T4 thyroxine TCDD 2,3,7,8-tetrachlorodibenzo-p-dioxin TEF toxic equivalency factor T R H thyroid releasing hormone TSH thyroid-stimulating hormone xii i UDP-GT UDP-glucuronosyltransferase xiv ACKNOWLEDGMENTS I would like to thank my supervisor, Dr. Stelvio Bandiera, for his constant support and guidance during the most challenging and stimulating period of my academic experience. I greatly appreciate the constructive feedback of my research committee, Dr. Stelvio Bandiera, Dr. Gail Bellward, Dr. Thomas Chang, Dr. Keith McErlane and Dr. John Sinclair. I am grateful to Dr. Daniel Desaulniers (Health Canada, Ottawa, ON) for analyzing L H and TSH in the serum samples, and to The National Institute of Diabetes and Digestive and Kidney Diseases (NIDDK) for providing the assay reagents. I thank Dr. Stephen Safe (Texas A & M University) for providing us with Aroclor 1260, and Dr. Paul Thomas (Rutgers-The State University of New Jersey) for providing us with monoclonal antibody to rat CYP1A1/2. Many thanks to my lab partners, Ludger Ickenstein, Vincent Pillay, and Kelsey Miller, my classmates, Vincent Tong and Ted Lakowski, the staff from Dr. McNeil 's lab, especially Mary Battell and Violet Yuen, Matthew Fedoruk and Simon Cowell from the prostate center (VGH), for their support and helpful advice. I would like to express my gratitude to my family for giving me the strongest moral and material support over the last three years. I send special thanks to my father, Richard Aloysius and, my mother, Therese Essovi, for showing me the meaning of hard work and perseverance. Dear Vivian, I could not have maintained my personal equilibrium if your understanding and patient love had not carried away all the stress that comes with grad school. Thank you. Thank you, Lord Jesus Christ, for you endless love. 1 INTRODUCTION Polychlorinated biphenyls (PCBs) are ubiquitous environmental contaminants with biological activity in wildlife, experimental animals, and humans (Safe, 1989; Safe, 1994). PCBs cause immune suppression (Harper et al, 1995), endocrine disruption (Li and Hansen 1996; Desaulniers et al, 1999), and various biochemical changes including the induction of hepatic microsomal cytochrome P450 (CYP) enzymes (Parkinson et al, 1983) in experimental animals. There is increasing interest on the part of the public and regulatory agencies in recent years in determining whether the toxicological and biochemical effects of PCBs occur at dosages that reflect environmental levels. Obtaining complete dose-response data provides a reference point for assessing the biological risks associated with environmental PCB exposure. Using Aroclor 1260, a commercial PCB mixture consisting mainly of ortho-?CB congeners (Safe, 1994), the present study sets out to investigate the levels of PCB exposure at which endocrine, immune and biochemical effects are observed in a widely used rodent model. 1.1. PHYSICAL A N D C H E M I C A L PROPERTIES OF PCBs The accumulation of PCBs in the ecosystem originated from their wide use in manufacturing processes. The advantageous physical and chemical properties of PCBs such as high viscosity, insulating ability, and high heat resistance were exploited in various industrial applications. PCBs were used as dielectric fluids in transformers and large capacitors, plasticizers, heat-exchange fluids, and carbonless copy paper (Bennett and Albro, 1973). 2 PCBs are aromatic hydrocarbons manufactured by the chlorination of biphenyl in the presence of a suitable catalyst. The general chemical formula of PCBs is Ci2Hio- nCl n (Figure 1). There are ten possible chlorine substitution sites on the two benzene rings of PCBs, which generates 209 possible PCB compounds called congeners. The term "homolog" is used for PCB congeners with the same number of chlorine atoms. PCB congeners that have chlorine atoms at the 2, 6, 2', and 6' positions are called ortho-substituted or ortho-PCBs. Congeners that are chlorinated at the 3, 5, 3', and 5' positions are meta-substituted, and at the 4 and 4' positions are /?ara-substituted. The IUPAC designation of PCB congeners uses a numbering system from 1 to 209, based on the number and position of chlorine atoms in the congener. For example, mono-substituted PCBs congeners are assigned numbers 1 through 3, and tetra-substituted congeners are assigned numbers 40 through 81 and the deca-substituted congener is assigned number 209 (Table 1). Thus, 2-monochlorobiphenyl is PCB#1, and 2,3,2',3'-tetrachlorobiphenyl is PCB#40. 3 2 2 3 5 6 6 5 Figure 1. Chemical structure of PCBs. 3 Table 1. IUPAC designation of PCB congeners (adapted from Ballschmitter and Zell, 1980) Substitution IUPAC No. Mono- 1-3 . Di - 4-15 Tri- 16-39 Tetra- 40-81 Penta- 82-127 Hexa- 128-169 Hepta- 170-193 Octa- 194-205 Nona- 206-208 Deca- 209 The physical-chemical properties of PCB congeners depend on their degree and pattern of substitution. Rotation of the benzene rings results in conformational isomers that can assume various degree of planarity although absolute planarity is the least favored PCB conformation. PCB co-planarity is reduced by the degree of ortho-substitution because of hindrance to rotation about the pivot bond (McKinney and Singh, 1981). In general, ortho-FCBs are non-planar while non-ortho-PCBs are relatively co-planar and most acutely toxic because PCB co-planarity and lateral chlorine substitution influence specific binding behaviors and consequent biological activity (McKinney and Waller, 1994). For example, ortho-VCB congeners bind to the aryl hydrocarbon (Ah) receptor with low affinity because receptor-binding affinity is determined by the relative ability of PCB isomers to adopt a relatively planar conformation (McKinney and Singh, 1981). In contrast to di-or?/zo-PCBs, mono-ort/jo-PCBs and non-ortho-FCBs are able to 4 assume a more planar conformation due to the absence of the steric hindrance factor and bind to the Ah receptor with greater affinity. The Aroclors® are commercial PCB mixtures marketed by Monsanto Corporation from 1930 to 1977. The Aroclors are identified by a four-digit code that contains information on the parent molecule (first two digits, 12 for biphenyl) and the chlorine content by percent weight (last two digits). For example, the chlorine contents of Aroclors 1242 and 1260 are 42% and 60%, respectively. The congener compositions of the Aroclor mixtures vary because of their degree of chlorination (Table 2). Of the 209 possible PCB congeners, 132 congeners were identified in all Aroclors (Aroclors 1242, 1248, 1254, and 1260) by gas chromatography with electron capture detection (Safe, 1994). Non-planar congeners such as 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153) and 2,3,4,5,2',4',5'-heptachlorobiphenyl (PCB#180) are mainly found in highly chlorinated mixtures such as Aroclor 1260 (Safe, 1994). Batch-to-batch variations in the congener composition of the Aroclors are common. Due to the complexity of the mixtures and the difficulty associated with chemical analysis, manufacturers were unable to tightly regulate the compositions of the PCB mixtures that they produced. Nevertheless, the compositions of Aroclor mixtures produced for industry have been characterized. 5 Table 2: Congener composition (weight percent of chlorobiphenyl) of selected Aroclor mixtures (adapted from Ballschmiter et al, 1989) Congener type Aroclor 1016 Aroclor 1221 Aroclor 1232 Aroclor 1242 Aroclor 1248 Aroclor 1254 Aroclor 1260 Monochlorobiphenyl 2 50 26 1 nd nd nd Dichlorobiphenyl 19 35 29 13 1 nd nd Trichlorobiphenyl 57 4 24 45 2 1 nd Tetrachlorobiphenyl 22 1 15 31 49 15 nd Pentachlorobiphenyl nd nd nd 10 27 53 12 Hexachlorobiphenyl nd nd nd nd 2 26 42 Heptachlorobiphenyl nd nd nd nd nd 4 38 Octachlorobiphenyl nd nd nd nd nd nd 7 Nonachlorobiphenyl nd nd nd nd nd nd 1 nd: not detected 1.2. BIOTRANSFORMATION OF PCBS PCBs persist in biological tissues due to their high lipophilicity and relatively slow rate of biotransformation. The factors that determine the rate of PCB biotransformation are the chlorine substitution pattern (number and position) and the animal species (Murk et al, 1994; Hansen, 1987). In animal studies, individual PCB congeners were rapidly absorbed (Albro and Fishbein, 1972; Tanabe et al, 1981; Allen et al., 1974) in the gastrointestinal tract after oral exposure, with peak serum concentrations reached in approximately 2-5 hours (Clevenger et al, 1989; Borchard et al, 1975). Data 6 from various pharmacokinetic studies suggest that the distribution of PCBs is biphasic, with rapid clearance from the blood and accumulation in the liver and various organs (WHO, 1993). In humans, PCBs are found in highest concentrations in the adipose tissue. Due to its high fat content, breast milk contains significant amounts of PCBs, which are transferred to infants through breast-feeding (McLachlan, 1993; Jacobson et al, 1984; Ando et al, 1985). Newborn animals are also exposed to PCBs via placental transfer (Jacobson et al, 1984). A l l PCB congeners do not have the same biological half-life because some congeners are more easily metabolized than others. The metabolism of PCBs in humans is not well characterized. Chromatographic analysis of PCB congeners in human tissues demonstrated that certain congeners such as 2,4,5,2',4',5*-hexachlorobiphenyl (PCB#153) and 2,3,4,2',4',5'-hexachlorobiphenyl (PCB#138) were persistent in human adipose tissues (Jensen and Sundstrom, 1974; Safe, 1994). Based on mechanistic studies in experimental animals, it has been recognized that the higher chlorinated PCBs and congeners with meta-, and para-substituted positions are more likely to bio-accumulate due to their slow metabolism by the microsomal C Y P system. Typically, para-substituted congeners containing five to ten chlorine atoms are resistant to biological degradation. Conversely, PCB congeners lacking meta- and para-substitution are susceptible to metabolism (Letcher et al, 2000). PCBs are mainly hydroxylated by CYP enzymes, via an arene oxide intermediate, (Schnellmann et al, 1983; Safe et al, 1975; Gardner et al, 1973) although alternate pathways may be involved in PCB metabolism (Wolff et al, 1992; Brown, 1994). In rodents, CYP1A and CYP2B enzymes are responsible for the formation of the arene oxide intermediate. Planar PCBs are substrates for CYP1A enzymes while non-planar 7 PCBs are the favored substrates for CYP2B enzymes (Poland et al, 1976; Poland and Knutson, 1982; Letcher et al, 2000). The reactive arene oxide intermediates can bind covalently to tissue macromolecules thereby causing cellular necrosis, mutagenicity and carcinogenicity (Safe, 1989). Alternatively, the arene oxide can be converted to phenols, dihhydrodiols, or glutathione conjugates (Figure 2). Further metabolism of the phenols results in the formation of catechols or phenol conjugates. Dihydrodiols are also converted to catechols via a pathway catalyzed by dehydrogenase. Glutathione conjugates can be further converted via the mercapturic acid pathway into methylsulfonyl metabolites, some of which are selectively retained in the lung (example: 4-methylsulfonyl derivatives) thereby causing respiratory toxicity (Haraguchi et al, 1997; Brandt and Bergman, 1987). Finally, the polar PCB metabolites are eliminated from the body as glucuronides or sulfates (Kurachi and Mio, 1983). There are species differences in the rate of metabolism of PCBs probably due to differences in the basal levels of particular C Y P isozymes (Murk et al, 1994; Hansen, 1987; WHO, 1993). For example, the in vitro rate of metabolism of 3,4,3',4'-tetracholorobiphenyl (PCB#77) by hepatic microsomes from the harbour porpoise and the Wistar rat were comparable. However, flounder and trout liver microsomes did not metabolize 3,4,3',4'-tetracholorobiphenyl (PCB#77). The metabolic rate correlated with ethoxyresorufin-O-deethylase (EROD), a CYPlA-mediated enzymatic activity, for all species except the trout suggesting that planar PCBs are substrates for CYP1A enzymes (Murke r / . , 1994). 8 Dechlorination products CI, CYP CI 0-5 oxidation COOH CI 0-5 Microbial degradation products multiple pathways Methyl sulfonyl metabolites! Thioethers GSH Macromolecular adducts <[o-5^p Arene Oxide Intermediate [protein, RNA, DNA] Rearomatization Phenols Epoxide {hydrolase H 2 0 GlucuronidationJ Sulfation Phenol-conjugates and derivatives Dehydrogenase catechols Dihydrodiols Figure 2. Metabolism of PCBs (adapted from Safe, 1984 and Safe, 1994) 9 1.3. BIOLOGICAL EFFECTS OF PCBs The biological activity of commercial PCB mixtures is the result of the effects of individual congeners, which in turn exhibit structure-dependent toxicity. Like polychlorinated dibenzo-p-dioxins (PCDDs), planar PCBs bind to the A h receptor and cause greater acute toxicity than non-planar congeners (Safe, 1994). Polychlorinated dibenzofurans (PCDFs), PCDDs and PCBs cause toxic responses ranging from decreased body weight to thyroid and liver cancers and immunosuppression. For compounds that elicit dioxin-like toxicity, there is a correlation between toxicity and Ah binding affinity. The toxic equivalency factor (TEF) model used for the risk assessment of PCDDs and PCDFs is based on the principle of Ah receptor-mediated toxicity. The TEF model is useful for assessing the relative toxicity of planar PCB congeners (Safe, 1994). In this model, all classes of compounds are compared to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which has the highest affinity for the Ah receptor and a TEF value of 1. The TEF model assumes that all PCB congeners share a common mechanism of toxic action and that toxic effects caused by individual congeners are additive. While non-planar PCBs are less effective at inducing Ah receptor-mediated effects, they elicit important non dioxin-like responses in biological systems. Non planar di-ort/20-PCBs accumulate in the brain, decrease neuronal dopamine content, and perturb intracellular second messenger systems central to the normal functioning and growth of neurons more effectively than planar PCBs (Olivero and Ganey, 2000; Kodavanti et al, 1996; Kodavanti et al, 1995; Kodavanti et al, 1994; Shain et al, 1991). Such PCB-induced neurotoxic effects might be mediated by changes in thyroid hormones (Chauhan et al, 2000; Goldey and Crofton, 1998; Goldey et al, 1995). 10 PCBs also elicit other biological effects including immune suppression (Harper et al, 1993; Rodriguez et al, 1997; Tryphonas et al, 1991b; Green et al, 1975), reproductive deficits (Linder et al, 1974; L i and Hansen, 1996), developmental effects (Linder et al, 1974; Sager, 1983) hematological changes (Bruckner et al, 1973) and hepatic effects (Carter, 1985; Sanders et al, 1974) in laboratory animals. The spectrum of PCB-induced responses is quite broad. Our study will focus mainly on the immune, endocrine, and biochemical effects of PCBs. 1.3.1. EFFECTS OF PCBs ON I M M U N E FUNCTION 1.3.1.2. EFFECTS OF PCB TREATMENT ON I M M U N E FUNCTION Thymus-derived lymphocytes, called T cells, are involved in immune defense by regulating cellular response and antibody production. There are two types of T cells with different functions. The first class of T cells consists of CD+4 cells involved in antibody production and the delayed-type hypersensitivity response (DTH). The second category of T cells comprises CD+8 cells, which mediate cytotoxic responses. CD+4 cells are subdivided into two subsets, differentiated by the types of cytokines produced. The TH1 subset produces interleukin 2 (IL-2) and interferon gamma (IFN-y). The TH2 subset produces IL-4 and IL-5. Stimulation of the TH1 subset is necessary for induction of protective immunity, whereas stimulation of the TH2 subset is associated with pathology (Scott etal, 1989). Changes in plasma cytokine levels may be indicative of PCB-induced alterations in immune function. Daniel et al (2001) reported plasma cytokine levels in 146 patients that were occupationally exposed to PCB levels. There was a weak negative association 11 between PCB levels and plasma IFN-y levels in individuals exposed to 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153), 2,3,4,2',4',5'-hexachlorobiphenyl (PCB#138) and 2,3,4,5,2',4',5'-heptachlorobiphenyl (PCB#180). Plasma levels of the congeners considered (PCB#138, PCB#153, and PCB#180) were above background levels in over 70% of subjects. PCB exposure was associated with mild immune dysfunction in the subjects studied. In another study by Exon et al. (1985), male Sprague Dawley rats of weanling age were exposed to 50 or 500 ppm Aroclor 1254 in the feed for 10 weeks, and IL-2 and natural killer (NK) cell activities were determined in rat splenocytes. There was a 25% and 50% suppression of IL-2 and N K cell activities, respectively, at a dosage of 500 ppm. Similar studies demonstrated that PCBs can suppress T cell function (Tryphonas et al, 1991; Fujimaki et al, 1997). Early studies indicated that PCBs affect humoral as well as cell-mediated immune responses. In rodents, antibody response to experimentally administered antigens was a reliable indicator of PCB-induced immune dysfunction. Aroclor 1260 exposure caused a dose-related decrease in plaque forming cells (PFCs) and a 35% decrease in antibody response to trinitrophenyl-lipopolysaccharide in mice at a dosage of 500 mg/kg. The ED50 value was calculated to be 354 mg/kg (Harper et al, 1995). A significant decrease in PFCs was also found in mice treated with Aroclor 1254, 1248, 1242 at dosages of 50-500 mg/kg, with ED50 values of 520, 379 and 507 mg/kg respectively (Harper et al, 1995). PCB treatment also affected cell-mediated immune responses. In rats exposed to Aroclor 1254 oral daily administration, for 5 to 15 weeks, there was a 20% and 40% decrease in thymus weight and N K cell activity, respectively. The lowest observed adverse effect level (LOAEL) values for changes thymus weight and N K cell activity 12 were both 10 mg/kg/day, and the dosage of maximal effects for N K cell activity was 25 mg/kg (Smialowicz et al, 1989). 1.3.1.2. M E C H A N I S M OF ACTION OF PCBs AS IMMUNOTOXICANTS The mechanism of action of PCBs as immunotoxic agents is still under investigation. PCBs are non-genotoxic tumor promoters in the liver and thyroid (AMAP, 1998; Whysner et al, 1998). Xenobiotics impair immune function by suppressing tumor gene expression thereby deregulating normal cell growth mechanisms such as the differentiation of B-lymphocytes (Rought et al, 1999). The biotransformation of xenobiotics leads to an increased formation of mutagenic and reactive intermediates, which can result in mutations of tumor suppressor genes. Examples of tumor suppressor genes that can be mutated by reactive intermediates are the retinoblastoma (Rb) and the p53 genes. A regulatory element in the Rb tumor suppressor gene has been well characterized. It is the so-called E2F site located directly downstream of the essential Rb binding factor 1 (RBF-1). Deletion of the E2F site increased Rb promoter activity in Rb-positive cell lines. The E2F protein was also found to bind to this site and downregulate Rb expression (Ohtani-Fujita et., al 1994). The absence of D N A adducts, in PCB-treated rats, does not support a genotoxic mechanism of immunosuppression (Whysner et al., 1998). However, dioxin-like PCBs (co-planar PCBs) possess significant tumor promoting capacity in rodents (Safe, 1989) as revealed by an increase in altered hepatic foci in animals treated with single planar PCB congeners such as 3,4,3',4'-tetrachlorobiphenyl (PCB#77) (Sargent et al, 1991; Bager et al, 1997). Likewise, non-13 dioxin-like di-ort/zo-substituted PCBs, such as 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153), were shown to be tumor promoters (Hemming et al, 1993). Planar PCBs cause humoral immunosuppression via a mechanism(s) that is mediated by the A h receptor (Okey, 1990; Safe, 1994; Silkworth and Grabstein, 1982; Silkworm and Antrim, 1985; Silkworth et al, 1986). Congener planarity and Ah gene expression correlated with PCB-induced immunotoxicity (Silkworth and Grabstein, 1982). Male mice possessing (C57BL/6-Ah + ) or lacking (DBA/2-Ah ") the A h phenotype were treated with planar 3,4,3',4'-tetrachlorobiphenyl (PCB#77) or non-planar 2,5,2',5'-tetrachlorobiphenyl (PCB#52) at the dosages of 0, 10, or 100 mg/kg ip, for two days before and two days after iv immunization with sheep red blood cells (SRBC). The planar PCB congener caused thymic atrophy and decreased anti-SRBC antibody-forming cell number (five days after immunization) per 106 spleen cells and per spleen by 79% and 90%>, respectively, in C57BL/6 but not in DBA/2 mice. The non-planar PCB congener did not produce effects in either strain. Because thymic atrophy was not always associated with severe immunosuppression, it was suspected to be a poor indicator of Ah-mediated humoral immunosuppression. The observed decrease in antibody response was thought to result from the impairment of B-lymphocyte differentiation (into antibody-producing cells) by Ah-mediated gene activation (Silkworth and Antrim, 1985). Antibody production appears to depend on the A h phenotype in rodents. Individual PCB congeners have been shown to interact to produce non-additive effects that correlate with Ah gene expression. The non-planar PCB, 2,5,2',5'-tetrachlorobiphenyl (PCB#52), and the planar congener 3,4,3',4'-tetrachlorobiphenyl (PCB#77), when administered together, caused a greater than additive decrease in the 14 total number of lymphocytes and antibody-producing B-cells in rats. There was a correlation between the Ah receptor-binding affinity of PCBs and their immunotoxic potency (Sargent et al, 1991). According to Sargent et al (1991), the synergistic immune effects observed with the combination of the two congeners mentioned earlier, were a result of increased gene expression. Immunosuppression by PCBs is not entirely dependent on dioxin-like characteristics. For instance, non-planar PCBs cause alterations in cellular immune response independently of their affinity for the Ah receptor. Neutrophils are leukocytes mainly involved in non-specific immune functions including inflammatory response, infection, and tissue injury. The non-planar PCB congener, 2,4,2',4'-tetrachlorobiphenyl (PCB#47), induced rat peritoneal neutrophil degranulation and superoxide production, as defensive responses to cytotoxicity, whereas the planar PCB congener, 3,4,3',4'-tetrachlorobiphenyl (PCB#77) did not (Brown and Ganey, 1995). These effects were abolished in the absence of calcium, suggesting that altered neutrophil function occurred through signal transduction pathways that are calcium-dependent, and not Ah-dependent (Brown and Ganey, 1995). Such pathways involve the activation of phospholipases C (PLC) and A 2 (PLA 2 ) (Tithof et al, 1994; Tithof et al, 1996; Olivero and Ganey, 1999). PLC hydrolyzes phosphatidylinositol-4,5-biphosphate (PIP2), resulting in the production of the second messengers inositol-1,4,5-triphosphate (IP3) and 1,2-src-diacylglycerol (DAG). IP3 mobilizes intracellular calcium in the endoplasmic reticulum, while D A G activates protein kinase C. Overall, PIP2 hydrolysis results in superoxide production and neutrophil degranulation via mechanisms that involve PLA2-dependent release of arachidonic acid. PCB-induced activation of PLA2 in turn depends on several pathways 15 involving tyrosine kinase, protein kinase C, mitogen-activated protein kinase (MAPK) , and M A P K kinase (Olivero and Ganey, 1999). 1.3.2. EFFECTS OF PCBs ON ENDOCRINE FUNCTION The literature contains extensive information on the endocrine effects of PCBs in experimental animals. PCB congeners and metabolites can bind to the estrogen receptor and transthyretin1, a thyroxine (T4) transport protein, because PCBs share structural similarities with these biological molecules as will be discussed below. By binding to the T4 transport protein and the estrogen receptor, PCBs can disrupt normal serum hormone levels. Currently, there is no clear indication that PCB compounds also bind to the androgen receptor or any other receptor involved in hormone homeostasis. Therefore, the following discussion will focus on the effects of PCBs on reproductive and thyroid hormones. 1.3.2.1. EFFECTS OF PCBs ON GONADOTROPIN A N D SEX H O R M O N E L E V E L S The secretion of the follicle-stimulating hormone (FSH) and the luteinizing hormone (LH) is mediated by steroid hormones (Figure 3). FSH and L H secretion are regulated by gonadal products that suppress the secretion of the gonadotropin-releasing hormone (GnRH) by negative feedback. The most important gonadal steroid hormones involved are testosterone in males and estradiol in females. Estradiol and testosterone decrease FSH and L H release by several mechanisms, including decrease in FSH and L H synthesis by transcription repression, alteration of the number of GnRH receptors on gonadotrophs, and decrease of GnRH secretion via interaction with endorphin neurons in 1 Tranthyretin is present in all vertebrates and can be used as model for other T4-binding proteins (Chauhan a/., 2000). 16 the hypothalamus (Figures 3 and 4). In experimental studies, PCBs induced an estrogen-mediated increase in uterine weight in the rat (Li et al, 1998; L i and Hansen, 1996). Neonatal exposure to Aroclor 1242 or 1254 (0.4-5.2 mg/day) from birth to day 25 by daily injection increased adult testis size and sperm production in the rat. These effects resulted from neonatal hypothyroidism and were effectively reversed with T4 replacement (Cooke et al, 1996). Estrogenic PCBs alter FSH pituitary isoforms as well as pituitary and serum levels of L H . Desaulniers et al (1999) exposed adult Sprague Dawley male rats to 3,4,5,3',4'-pentachlorobiphenyl (PCB#126) at dosages of 6.25 to 400 ug/kg/day, ip, for two days, and found a significant decrease in serum T4 levels, at a dosage of 25 ug/kg/day. Serum T4 levels were decreased by 70% from control levels in the highest dosage group. In addition, serum L H concentrations were decreased by 60% at a dosage of 400 (xg/kg/day, while pituitary L H concentrations were increased at this dosage of PCB#126. The"re were no significant changes in serum FSH levels suggesting that changes in serum L H may be a more sensitive indicator of PCB-induced reproductive toxicity in male rats. At a dosage of 25 mg/kg/day, 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153), a non-planar PCB congener, caused less or equal endocrine disrupting effects when compared to PCB#126 (Desaulniers et al, 1999). The effects of PCBs on serum gonadotropin and sex hormone levels can result in important reproductive deficits. PCBs inhibited ovulation and reduced serum progesterone, FSH, and L H levels by more than 80%), 60%, and 80%>, respectively, in gonadotropin-primed immature female rats treated with 3,4,5,3',4',-pentachlorobiphenyl (PCB#126) and 2,5,2',5'-tetrachlorobiphenyl (PCB#52) at dosages of 114-457 ug/kg (Gao et al, 2000). Moreover, PCB#126 caused a 17 3.5-fold increase in serum estradiol at a dosage of 457 ug/kg (Gao et al, 2000). The effects of PCBs on sex steroids can lead to sex-specific neurobehavioural toxicity (Hany etal, 1999). Dopamine Norepinephrine Testosterone/Estrad io 1 Testosterone Estradiol Stimulate Inhibit y^- ^ Figure 3. Regulation of gonadotropin secretion 18 inactive receptor dimerization estrogen-response element transcription protein mRNA altered cell function decreased gonadotrpins and sex hormones decreased T3 & T4 and increased TSH stores Figure 4. Molecular pathway of estrogen action The biological action of estradiol is mediated by the estrogen receptor (ER) , located within target cells. In the absence of estradiol, the E R resides in an inactive conformation (monomeric form) and is stabilized by heat-shock (hs) proteins bound to it. The interaction o f estradiol or xenoestrogens with the E R induces a conformational change in the E R . A cascade of events is thereby initiated that induces dissociation of hs proteins and dimerization of the estradiol/xenoestrogen-ER complex. The dimerized receptor complex binds to specific D N A responsive elements within regulatory regions of hormone-responsive genes (gene X ) and interacts with other regulatory proteins to alter gene transcription. Changes in gene transcription lead to altered m R N A and protein concentrations that ultimately result in a change in cellular function (adapted from Gaido et al, 1997). 19 1.3.2.2. ESTROGENIC ACTIVITY OF PCBs PCBs and their metabolites possess the ability to interfere with normal estrogenic activity. In vitro, the hydroxylated PCBs, 2,4,6-trichloro-4'-biphenylol and 2,3,4,5-tetrachloro-4'-biphenylol, were estrogenic and bound to the estrogen receptor competitively (Arcaro et al, 1998). The PCB metabolites were assessed for their ability to inhibit H-labelled estradiol ( H-E 2) binding to the estrogen receptor, and to induce estrogen-dependent cell proliferation (induction of foci). Both hydroxylated PCBs inhibited 3 H - E 2 binding with I C 5 0 values of 0.079 uM and 0.015 uM for 2,4,6-trichloro-4'-biphenylol and 2,3,4,5-tetrachloro-4'-biphenylol, respectively (Acaro et al, 1998). Foci were induced in a dose-dependent manner with EC50 values of 0.22 uM and 0.72 u M for 2,4,6-trichloro-4'-biphenylol and 2,3,4,5-tetrachloro-4'-biphenylol, respectively (Arcaro et al, 1998). Similar studies reported estrogenic activities of catechol metabolites of selected PCBs (Garner et al, 1999). A limited number of parent PCBs (eg. PCB#104, PCB#184, and PCB#185) were found to be estrogenic with slightly higher IC50 values (ranging from 0.4-1.3 uM) than hydroxylated congeners in the estrogen receptor competitive binding assay (Matthews and Zacharewski, 2000). Other hydroxy-PCBs identified in human serum, such as 2,2',3,4',5,5'-hexachloro-4-biphenylol and 2,3,3',4',5-pentachlorobiphenylol, are considered to be anti-estrogenic because they inhibit estrogen-mediated cell proliferation and/or reporter gene activity in vitro (Moore etal, 1997). The ability of PCBs and/or their metabolites to bind to the estrogen receptor is structure-dependent. The structures of estrogens consist of rigid polycyclic molecules with important asymmetry. Environmental estrogens often share structural similarities 20 with the phenolic A ring of estradiol (Figure 5). McKinney and Waller (1994) postulated that the affinity of PCBs for the estrogen receptor can be increased following metabolism. This may partly explain the fact that some hydroxylated PCBs are estrogenic, although their parent congeners may lack estrogenicity. Figure 5. Chemical structures of estradiol and hydroxylated PCBs 21 1.3.2.3. EFFECTS OF PCBs ON THYROID HORMONE HOMEOSTASIS A N D FUNCTION As mentioned earlier, disruptions in thyroid hormone homeostasis can result in appreciable reproductive and neuronal toxicity. Serum T4 levels declined to less than 50% of control values in rats fed more than 36 mg/kg/day of a PCB extract from National Priority List (NPL) sites for two days. A maximal decrease in serum T4 levels was observed at a dosage of 72 mg/kg/day. The TEF of the environmental mixture and that of Aroclor 1248 were comparable (Hansen et al, 1995). There was an increase in thyroid follicular cell size and a 40% decrease in the colloid area compared to control animals. Likewise, L i and Hansen (1996) reported a dose-dependent decrease in serum T4 levels in rats exposed to air, dust, and soil PCB extracts. The decrease in serum T4 concentration was linear in the 10-100 mg/kg/day range. There was also a 30% increase in uterine wet weight at dosages of 346 mg/kg and 175 mg/kg for the soil and air extracts, respectively. Although the composition of the soil extract was not identical to that of Aroclor 1260, the TEF of the soil extract was slightly higher than that of Aroclor 1260, suggesting that the commercial mixture would be less potent. A decrease in serum T4 levels was reported in female, but not male, Sprague Dawley rats exposed to 3,4,3',4'-tetrachlorobiphenyl (PCB#77) at the dosage of 0.75 mg/kg/day for ninety days (Desaulniers et al, 1997). 2,4,4'-Trichlorobiphenyl (PCB#28) treatment did not affect serum T4 levels at a dosage of 3.7 mg/kg/day (Desaulniers et al, 1997). Goldey et al. (1995) reported a dose-dependent decrease in circulating T4 levels (up to 10%) of control) in offspring of Long-Evans rats that received Aroclor 1254 orally at dosages of 1, 4, or 10 mg/kg. A moderate decline in serum triiodothyronine (T3) 22 levels was noted at dosages of 4 and 10 mg/kg. The decrease in serum T4 levels was associated with hearing deficits, suggesting that endocrine disruption affected the clochea (Goldey et al, 1995). A study by Byrne et al. (1987) demonstrated that Aroclor 1254 not only suppressed serum T4 and T3 levels, but also affected the kinetics of thyroid hormone metabolism. Adult female Sprague Dawley rats were exposed to Aroclor 1254 in the feed for five to seven months at the dosages of 0, 1, 5, 10, and 50 ppm. Aroclor 1254 treatment caused a decrease in serum T4 and T3 concentrations in a dose-dependent manner. Serum T4 concentration was suppressed by approximately 50% on day 14, and by more than a 75% on day 140, at a dosage of 50 ppm. There was a moderate decline in serum T3 levels (-29% decrease) on day 25, and a large decrease (86% decrease) on day 158. The kinetics of T4 metabolism were examined by measuring the thyroid response to exogenous TSH injections. Abnormalities in the disappearance of T4 and T3 from the serum were observed. There was a faster disappearance of T4 and T3 from the serum in PCB-treated animals compared to control animals. Although, T3 was not an early indicator of Aroclor 1254 exposure, long-term exposure resulted in a significant suppression of serum T3 levels. These results account for the fact that subchronic and acute studies previously mentioned (Desaulniers et al, 1997; Hansen et al, 1995; L i and Hansen, 1996) did not report a decrease in serum T3 levels, because the duration of exposure in such studies was relatively short. Placental PCB transfer from PCB-treated animals to neonates was also shown to result in significant thyroid hormone effects in neonates (Seo and Meserve, 1995). Female Sprague Dawley rats received Aroclor 1254 in the diet at the dosages of 125 ppm 23 and 250 ppm, from the first day of pregnancy to birth. The neonates exhibited a 30% decrease in body weight fifteen days after birth (day 15) at a dosage of 250 ppm, and a moderate decrease (-5%) at a dosage of 125 ppm. Pups were killed on day 15, and trunk blood was analyzed for serum T4 and T3 concentrations. There was a 1.5-fold increase in thyroid weight, and greater than 90% suppression of serum T4 levels, but no change in serum T3 levels at a dosage of 125 ppm (Seo and Meserve, 1995). 1.3.2.4. MECHANISMS OF PCB-INDUCED THYROID H O R M O N E M O D U L A T I O N Different mechanisms have been proposed to account for the effects of PCBs on thyroid hormone regulation. PCBs can mimic the action of estradiol (Figure 4), and estradiol has been shown to regulate thyroid releasing hormone (TRH) gene expression and 5'-iodothyronine deiodinase activity, and to up-regulate the T R H receptors (Desaulniers et al, 1999). A more widely accepted view is that PCBs, as well as other enzyme inducers and antithyroid drugs accelerate thyroid hormone turnover via a common mechanism (McClain et al, 1989; Liu et al, 1995; Barter and Klaassen, 1994; Hood et al, 1999b) namely, UDP-glucuronosyltransferase (UDP-GT) induction. UDP-GT catalyzes the formation of T4-glucuronide and is primarily responsible for the glucuronidation of T4. There are four classes of UDP-GT inducers known to date. Representative UDP-GT inducers of each class are clofibrate, phenobarbital (PB), pregnenolone-16a-carbonitrile (PCN) and 3-methylcholanthrene (3-MC). Upon UDP-GT induction, T4 degradation is accelerated (converted to T3) by glucuronidation followed by biliary excretion. T4 and T3 feedback-inhibit the synthesis and secretion of the thyroid-stimulating hormone (TSH). The accelerated T4 turnover leads to an increase 24 in plasma T3, which causes a decrease in the number of receptors for the T R H 2 . Because TSH is stimulated by TRH, there is increased inhibition of TSH synthesis and secretion and a probable rise in TSH stores. Only PB has been shown to cause an increase in serum TSH and to promote tumor growth. P C N also increases serum TSH whereas 3MC and PCBs do not (Liu et al, 1995; Hood et al, 1999a). It is unclear why serum TSH concentration is not increased with 3MC and PCB-treatment. Furthermore, serum T3 levels are not significantly affected by PB, PCN, 3MC, and PCBs (Liu et al, 1995). The conversion of T4 to T3, catalyzed by outer-ring deiodinase (ORD) type-I and -II enzymes, was reduced rather than increased in the thyroid, kidney, liver, brown adipose tissue, pituitary, and brain. Consequently, increased T4 turnover does not account for the fact that serum T3 concentration is maintained in 3MC- or PCB-treated rats (Hood and Klaassen, 2000a). Determination of T3-UDP-GT activities in rats fed a diet containing PB (300-2400 ppm), P C N (200-1600 ppm), 3MC (50-400 ppm), or PCB (25-200 ppm) for seven days, indicated that enzyme inducers affected T3-UDP-GT activity differently. PB and P C N increased T3-UDP-GT activity by 90% and 120-200%, respectively, whereas Aroclor 1254 and 3MC did not have any appreciable effect (Hood and Klaassen, 2000b). PCBs decrease serum T4 levels by other mechanisms as well. Hydroxylated PCBs bind transthyretin with relatively high affinity, thereby displacing T4 (Brouwer and Van Den Berg, 1986). Displaced T4 is eliminated via T4 glucuronidation leading to a decrease in serum T4 levels. Non-hydroxylated PCBs exhibit lower binding activity compared to hydroxylated PCBs. In addition, di- and tri-ortho congeners bind transthyretin with high affinity, while mono- and fully or/Ao-substituted congeners do not 2 T3 binds to its nuclear receptors that in turn regulate the synthesis of an inhibitory protein 25 (Chauhan et al, 1999). Fully weta-substituted congeners also promote increased transthyretin binding activity (Chauhan et al, 1999). In general, the substitution patterns of PCBs (di-meto-substitution in one or both rings) associated with high binding affinity to transthyretin were similar to the diiodophenolic ring of T4 (Figure 6). On the other hand, PCB congener patterns (single meta-substitution on one or both rings) similar to the monoiodophenolic ring of T3 (Figure 6), exhibited low binding affinities to transthyretin, consistent with the low binding activity of T3 (Chauhan et al, 1999). N H 2 H O 3 C H 2 — C H — C O O H Triiodothyronine (T3) I H O 0 C H 2 C H — C O O H Thyroxine (T4) N H 2 Figure 6. Chemical structures of thyroid hormones 26 1.3.3. EFFECTS OF PCBs ON C Y T O C H R O M E P450 (CYP) E N Z Y M E S C Y P enzymes are the major steroid and xenobiotic-transforming enzymes in the body. Although present in various body compartments, C Y P enzymes are found at their highest concentration in the liver. C Y P enzymes are involved in numerous phase I biotransformation pathways. Examples of reactions catalyzed by C Y P enzymes include aliphatic and aromatic hydroxylation, double bond epoxidation, heteroatom (S-, O- and N-) dealkylation, oxidative group transfer, and cleavage of esters (Klaassen, 1996). Certain environmental chemicals induce C Y P activity levels, thereby accelerating their own biotransformation. C Y P inducers are classified based on their mechanism of action and the types of enzymes induced. The five representative C Y P inducers are (1) 3-methylcholanthrene (MC), (2) phenobarbital (PB), (3) pregnenolone 16a-carbonitrile (PCN) and glucocorticoids, (4) ethanol, and (5) peroxisome proliferators (Table 3). Results from studies involving single PCB congeners demonstrated structure-dependent induction of C Y P enzymes by PCBs. Planar PCBs are MC-type inducers (see future paragraphs), and mainly induce hepatic CYP1A enzymes. Conversely, some non-planar PCBs are PB-type inducers and primarily induce C Y P 2B enzymes, whereas others are mixed-type inducers. Mixed-type inducers induce both C Y P 1A and C Y P 2B enzymes (Parkinson et al, 1983; Parkinson et al, 1981). Some PCBs do not induce any C Y P enzymes (Safe, 1990; Safe, 1994). 27 Table 3. Major C Y P inducers (derived from Okey, 1990; Gerbal-Chaloin et al, 2001; Waxman, 1999; Omiecinski et al, 1999) CYP isoforms induced Inducer category Rat Human Polycyclic aromatic 1A1 1A1 Compounds 1A2 1A2 Phenobarbital 1B1 1B1 2B1/2B2 2B6 2C8 2C9 PCN/Glucocorticoids 3A23 (3A1) 3A4 Ethanol 2E1 2E1 Peroxisome proliferators 4A1 1.3.3.1. EFFECTS OF PCB TREATMENT ON CYP INDUCTION Single PCB congeners have been shown to induce C Y P enzymes in a structure-dependent manner. Early Studies by Parkinson et al (1981, 1983) reported induction of several hepatic C Y P enzymes in one-month old male Wistar rats exposed to individual PCB congeners by i.p. injection, at dosages of 150 umol/kg/day for two days, or a single injection at a dosage of 125-500 umol/kg. Treatment with 3,4,3',4'-tetrachlorobiphenyl (PCB#77), 3,4,5,4'-tetrachlorobiphenyl (PCB#81), 3,4,5,3',4'-pentachlorobiphenyl (PCB#126), or 3,4,5,3',4',5'-hexachlorobiphenyl (PCB#169) induced CYP1A1 and CYP1A2 enzymes by approximately 30-50-fold, but had minimal effect on CYP2A and CYP2B enzymes. In comparison, treatment with 2,4,2',4'-tetrachlorobiphenyl (PCB#47) or 2,5,2',5'-tetrachlorobiphenyl (PCB#52) induced CYP2B enzymes but did not effect CYP1A enzymes. Other congeners such as 2,3,4,6,3',4'-hexachlorobiphenyl (PCB#158) 28 and 2,3,4,5,2',3',4'-heptachlorobiphenyl (PCB#170) induced CYP1A and CYP2B enzymes. These studies demonstrated that planar PCB congeners (no ortho substituents) preferentially induce CYP1A enzymes, whereas di-ort/zo-substituted non-planar PCB congeners (with one ortho substituent on each ring) preferentially induce CYP2B enzymes. Mono-ortho and di-ortho PCB congeners that are able to assume a planar conformation induce both CYP1A and CYP2B enzymes (Parkinson et al, 1981; Parkinson et al., 1983). A dose-dependent increase in CYP2B-mediated pentoxyresorufin O-dealkylase (PROD) activity was observed in Sprague Dawley rats treated with 2,3,3',4',6-pentachlorobiphenyl (PCB#110), ip (Li et al, 1998). PROD activity was increased by 4-fold and 7.6-fold, at dosages of 32 mg/kg and 96 mg/kg, respectively. These results suggested that PCB#110, a non-planar PCB congener, induced CYP2B1 and CYP2B2 enzymes (Li et al, 1998). Ikegwuonu et al. (1996) reported that 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153), a non-planar congener, induced CYP2A1, CYP2B1, CYP2B2, CYP2C6, and CYP3A1 in Fischer F344 and Wistar Furth rats. By comparing the effects of PCB#153 to that of PB, it was concluded that similar C Y P isozymes were induced by both chemical treatments, and that non-planar PCBs can also induce CYP3A enzymes. PCB mixtures elicit more complex biochemical effects when compared to single PCB congeners. There was a four-fold increase in total hepatic C Y P content in three-week old male Long Evans rats treated with Aroclor 1254 by ip injection at a dosage of 1500 umol/kg. Hepatic CYP2A1, CYP2B, CYP1A1, and CYP1A2 enzymes were increased by 5-, 60-, 33-, and 14-fold, respectively (Parkinson et al, 1982). Oral 29 administration of Aroclor 1260 to adult male rats, at dosages of 0 mg/kg/day-50 mg/kg/day, for seven days resulted in increased liver weight, and a dose-dependent increase in total hepatic C Y P enzymes. Aroclor 1260 induced hepatic CYP1A, CYP2A, CYP2B, and CYP3A enzymes, but suppressed CYP2C11 enzymes (Ngui and Bandiera, 1999). Maximal induction was observed at a dosage of 5 mg/kg/day for hepatic CYP2B enzymes and 20 mg/kg/day for hepatic CYP3A enzymes. Results from Ngui and Bandiera's study (1999) suggested that hepatic CYP2B induction was a sensitive marker of Aroclor 1260 exposure. A previous study by Harris et al. (1993) with various Aroclor mixtures reported similar results. In that study, immature Wistar male rats received ip injections of Aroclor 1232, Aroclor 1242, Aroclor 1248, Aroclor 1254, or Aroclor 1260 at the dosages 10, 40, 160, 480, and 2000 mg/kg. EROD activity was induced in the following order, Aroclor 1254>Aroclor 1248>Aroclor 1242>Aroclor 1260>Aroclor 1232. However, Aroclor 1260 was the most potent inducer of PROD activity. PROD activity increased in a dose-dependent manner (maximal increase at 160 mg/kg), and was a more sensitive indicator of Aroclor 1260 exposure (ED50 value of 37 mg/kg) than EROD activity ( E D 5 0 value of 48 mg/kg). Environmental PCB mixtures exhibit effects on hepatic C Y P enzyme that are comparable to those of commercial PCB mixtures. For example, a 17% increase in liver weight was observed in twenty-day old female rats that received ip injections of PCBs extracted from NPL sites, at dosages greater than 36 mg/kg/day for two days (Hansen et al, 1995). There was a four-fold induction in hepatic CYP2B enzymes, with maximal effect at a dosage of 96 mg/kg/day. Hepatic CYP1A1 increased by approximately twenty-fold at a dosage of 12 mg/kg/day, indicating that the extracts studied were 30 comparable to Aroclor 1242. Additional studies by L i and Hansen (1996) demonstrated that PCB mixtures extracted from NPL sites (air, dust and soil) caused dose-dependent increases in EROD and PROD activities in Sprague Dawley rats that received dosages between 0 and 346 mg/kg/day for two days, ip. The dose-response curves obtained were quasi-linear in the 10-100 mg/kg range. 1.3.3.2. M E C H A N I S M OF ACTION OF PCBs AS C Y P INDUCERS The mechanism of action of MC-type inducers has well been characterized. M C -type inducers bind to the Ah receptor, a cytosolic receptor known to interact with a variety of xenobiotics. The inducer-receptor complex translocates into the nucleus through binding to the aryl receptor nuclear translocase (Arnt). The resulting complex in turn binds to D N A responsive elements, which leads to transcription of CYP1A1 and CYP1A2 genes and other genes involved in cell growth and differentiation (Whitlock, 1999; Rowlands and Gustafsson, 1997; Hankinson, 1995). There is a broad structural diversity among 'PB-type' inducers. This feature appears to rule out the existence of a single receptor that 'PB-type' inducers might bind in order to induce hepatic enzymes. Several models ranging from mechanisms involving the metabolism of an endogenous substrate by CYP2B enzymes, to the maintenance of a 'repressor precursor' were proposed but have not been tested extensively (Waxman and Azaroff, 1992; Ortiz de Montellano, 1995). Because previous studies demonstrated that PB-responsiveness was inhibited by the glucocorticoid/progesterone antagonist RU486, PB-type inducers were thought to cause the accumulation of an endogenous steroid that in turn induced the expression of CYP2B1 and CYP2B2 genes through binding to a 31 steroid receptor. PB-dependent proteins bind to D N A near the promoter region of CYP genes. Early experiments in the bacterium Bacillus megaterium led to the discovery of PB-responsive elements (Ortiz de Montellano, 1995). The 5' promoter region of CYP2B genes was thought to contain a 15-base pair sequence, known as the Barbie box, that is similar to 5' enhancer regions in mammalian genes. Transcription of the Barbie box is inactivated by a repressor protein, which is displaced by PB-dependent protein(s) upon PB-induction. Recent studies have reported that members of the CYP2A, CYP2B, and CYP3A families, which are inducible by a variety of compounds including steroids, are induced via the pregnane x receptor (PXR), and the constitutively active receptor (CAR) (Blumberg and Evans, 1998; Savas et al, 1999; Sueyoshi and Negishi, 2000). P X R and C A R are members of the nuclear hormone receptor (NHR) family. P X R and C A R can translocate into the nucleus, form heterodimers with the NHR retinoid x receptor (RXR) and bind D N A responsive elements upstream of CYP genes. For instance, the C A R : R X R complex mediates the induction of CYP2B enzymes by binding to a 51-bp PB-responsive gene in the CYP2B6 gene called the PB-responsive enhancer module (PBREM), which is also activated by the PXR. Likewise, the P X R : R X R dimer activates CYP3A genes by binding to nuclear receptor sites contained in the xenobiotic-responsive module (XREM) found in human CYP3A4 genes. Regulation through these newly discovered receptors accounts for the array of compounds that act as PB-type inducers (Goodwin et al, 2001; Drocourt et al, 2001; Sueyoshi and Negishi, 2000; Blumberg and Evans, 1998; Blumberg etal, 1998). 32 1.3.4. STRAIN, SPECIES, A N D SEX DIFFERENCES IN EFFECTS ELICITED B Y PCBs The inducibility of C Y P enzymes by PCBs depends on species and strain differences in basal C Y P expression. CYP1A is inducible by TCDD and planar PCBs in all vertebrate species examined to date. PB-type inducers, and hence non-planar PCBs, induce CYP2B1 and CYP2B2 enzymes in the rat. CYP2B1 and CYP2B2 enzymes are primarily expressed in rodents, while CYP2B6 is present in humans at low levels (Omiecinski et al, 1999). The induction of CYP2B6 enzymes by PB, but not PCBs, has been demonstrated in humans (Drocourt et al, 2001; Goodwin et al, 2001). Furthermore, C Y P inducibility exhibits important strain differences in the rat. For example, 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153) induced hepatic C Y P 2 A L CYP2B1, CYP2B2, CYP2C6, and CYP3A1 in the rat with significant strain differences between Fischer F344 and Wistar Furth rats. Strain differences in C Y P induction by 2,4,5,2',4',5'-hexachlorobiphenyl were caused by genetic differences between Fischer and Wistar rats (Ikegwuonu et al, 1996). Sex differences in C Y P induction are endocrine-dependent. Results by Ikegwuonu et al. (1996) suggested that 2,4,5,2',4',5'-hexachlorobiphenyl induced hepatic C Y P enzymes through a strain-dependent endocrine sensitive PB-stimulated pathway. This pathway was sensitive to growth hormone, T3, and testosterone, and strain differences were more apparent in females than males. PB-type induction is sexually dimorphic because it is influenced by sex hormones. For example, pre-pubertal estradiol exposure attenuates CYP2B1 induction whereas pre-pubertal testosterone exposure stimulates CYP2B1 induction (Chang et al, 1997). Some C Y P enzymes are more 33 endocrine-sensitive than others. For example, CYP3A18 is more sensitive to G H hormone suppression than CYP3A9 in adult rats (Kawai et al., 2000). 1.3.5. S U M M A R Y OF DOSE-RESPONSE TRENDS Numerous studies have shown that various PCB congeners and mixtures cause biochemical, endocrine, and immune effects in experimental animals (Table 4) yet, few studies have established complete dose-response curves for these effects. The dose-response curves for the effects of Aroclor 1260 on hepatic C Y P induction can be predicted, based on the data available from a number of studies, quoted earlier, that used various dosages of PCBs (see Table 4). It is more difficult to predict dose-response effects for immune and endocrine endpoints because there are no complete dose-response curves for Aroclor 1260. Harper et al. (1995) reported that antibody response decreased at dosages greater than 50 mg/kg, in PCB-treated mice. However, the dosages used were high and are not comparable to environmental PCB exposure. Desaulniers et al.'s (1999) findings indicated that serum L H levels declined rapidly to a plateau at relatively low dosages using a single PCB congener, but not a PCB mixture. L i and Hansen (1996) reported the full dose-response curves for serum T4 levels in rats exposed acutely to PCBs extracted from the environment. Although the results of such studies are informative, very little is known about the dosages at which these effects are observed with prolonged exposure to environmental PCB mixtures. Goldey et al. (1995) studied the long-term effects of Aroclor 1254 on serum T4 concentrations, but did not report complete dose-response curves for this endpoint. 34 A previous study in our laboratory reported dose-response curves of the effects of Aroclor 1260 on hepatic C Y P enzyme levels (Ngui and Bandiera, 1999). The results of this study indicated that Aroclor 1260 caused 55-and 16-fold increases in hepatic CYP2B1 and CYP2B2 enzymes, respectively, between cumulative doses3 of 2.5 mg/kg and 35 mg/kg. Based on these considerations, we set out to obtain full dose-response curves for the immune and endocrine effects of PCBs in rats after a ten-day exposure. The duration of PCB treatment was chosen to match previous work performed in our laboratory. It was necessary to choose a broad range of dosages to obtain the entire dose-response curves for the endpoints investigated. The LD50 of Aroclor 1260 ranges between 1.26 and 2.0 g/kg (Safe, 1994). To prevent Aroclor 1260 toxicity the highest dosage used was 156 mg/kg/day, equal to a cumulative dosage of 1.56 g/kg. Typically, dosages were selected at five-fold increments. However, additional dosages were included to observe mid-dose range effects for specific endpoints. For example, a dosage of 1.25 mg/kg was chosen because hepatic CYP2B induction leveled off at a dosage 5 mg/kg/day (Ngui and Bandiera, 1999). A n additional dosage group between the dosages of 0.625 mg/kg/day and 3.13 mg/kg/day was required to fully characterize the CYP2B dose-response curve. 3 The cumulative dose is defined as the total amount of PCB administered over a given period, and does not represent tissue PCB levels. 35 Table 4. Summary of dose-response effects of PCBs in rodents PCB congener/ mixture used (dosage range) Effect (change at maximal dosage tested) Lowest dosage at which effect was observed Dosage of maximal effect Species Reference Aroclor 1260 (50-500 mg/kg) Antibody response (35% decrease) 50 mg/kg 500 mg/kg (ED50=354 mg/kg) Mice Harper etal, 1995 Aroclor 1254 (7 days/week for 5-15 weeks N K cell activity (<40% decrease) 10 mg/kg/day 25 mg/kg/day Rat Smialowicz et al, 1989 Aroclor 1254 (7 days/week for 5-15 weeks Thymus weight (<20% decrease) 10 mg/kg 10 mg/kg/day Rat Smialowicz et al, 1989 PCB 126 (6.25 ug/kg/day-400 u/kg/d for 2 days), Serum T4 levels (30% decrease) 25 fxg/kg/day 400 ug/kg/day Rat Desaulniers et al, 1999 PCB 126 (6.25 ug/kg/day-400 u/kg/day for 2 days), Decrease in serum L H levels (60%) decrease) 400 ug/kg/day N / A Rat Desaulniers et al, 1999 Aroclor 1254(1-10 mg/kg) Serum T4 levels (10% decrease) 1 mg/kg N / A Rat Goldey et al., 1995 Environmental PCB extracts Serum T4 levels (50% decrease) >36 mg/kg 72 mg/kg Rat Hansen et al., 1995 PCB 110(0-96 mg/kg) PROD activity (7.6-fold increase), increased liver weight 32 mg/kg 96 mg/kg Rat Ikegwuonu et al, 1996 Aroclor 1260 (0-50 mg/kg/day for 7 day) Hepatic CYP2B enzymes (55-fold and 16-fold induction in CYP2B1 and CYP2B2, respectively) 2.5 mg/kg/day 5 mg/kg/day Rat Ngui and Bandiera, 1999 Environmental extracts (0-346 mg/kg) Hepatic CYP2B enzymes (4-fold increase) 36 mg/kg 96 mg/kg Rat L i and Hansen, 1996 N / A : not available 36 1.4. E N V I R O N M E N T A L PCB EXPOSURE 1.4.1. WILDLIFE EXPOSURE The presence of detectable PCB concentrations in environmental and biological samples is a result of poor PCB disposal practices in the past. PCBs can be detected in air, water, soil samples, and a wide spectrum of wildlife species. For example, total PCB levels in soil, dust, and air from NPL sites were extracted in hexane and quantified. The amounts of sample used were 612 m 3 for air and lOOg for soil and dust. PCB concentrations in extracts of soil, dust, and air samples were reported to be 47, 21 and 5 mg/ml of hexane extract, respectively (the quantity of hexane used was not specified in the original paper) indicating that the soil sample initially contained the highest PCB concentration (Li and Hansen, 1996). PCB levels in environmental extracts depend on the location. Highly contaminated areas contain relatively high levels of PCBs. For example, atmospheric PCB levels in an electro-industrial plant in Belakrajina (Yugoslavia) were 2000 ug/m , but only 2 to 5 ug/m in nearby residential areas (Safe, 1994). PCBs are among the most prevalent organochlorine contaminants in wildlife species. Senthilkumar et al. (1999) determined oganochlorine levels in the blubber of Ganges River dolphins and found that PCBs were the second most prevalent organochlorine contaminant, after l,l,l-trichloro-2,2-bis[p-chlorophenyl]ethane (DDT). PCB levels varied from 1.5 to 25 ug/kg lipid weight and the most abundant PCB congeners were PCB#138, PCB#153 and PCB#118. In Hudson Bay, Canada, the muscle 37 and skin of anadromous fish4 contained up to 57 ug/kg PCBs by lipid weight (Arctic Monitoring Assessment Programme (AMAP, 1998). Tissue levels of PCBs vary among species according to their trophic levels. PCBs bioaccumulate at the top of the food chain in the adipose tissues of large organisms that eat a variety of smaller organisms. A study by Dewailly et al. (1993) reported total PCB levels in the adipose tissue of various species from the arctic aquatic food web (Table 5). Polar bear fat contained the highest concentration of PCBs while Arctic char muscle contained 70 times less PCBs than the polar bear. The polar bear is at the top of the arctic marine food web and has relatively high PCB concentrations. Hepatic PCB content was as high as 83 mg/kg lipid weight in the arctic polar bear (Table 6). Table 5. Total PCB concentration (ug/g, lipid basis) in milk from Inuit women and in fatty tissues of species from the arctic Quebec food web (adapted from Dewailly et al. (1993)) Species Polar bear Human Beluga Seal Arctic Human and fat milk blubber blubber char milk tissues (Arctic) muscle (south) N 35 107 16 16 9 16* 7.00±1.28 1.05±0.15 1.00+0.47 0.53±0.69 0.15+0.04 0.16+0.04 (*) Sixteen pools of milk, six samples each. 4 Anadromous fish live in both salt and fresh water, and include whitefish, Arctic char, trout, and salmon. Table 6. PCB levels in the Arctic polar bear (adapted from A M A P , 1998) Location and date n a tissue PCB level (jag/kg lipid wt)b Canada, 1969 6-7 fat (1890-4500) Canada, 1982-84 6-20 fat (3240-8250) Svalbard, 1978-89 16 liver 13000 (100-780000) 5 liver 12000 (200-35000) 7 fat 31000 (2900-90000) 3 fat 15000 (4100-21000) R l c , 1993 17 fat 7948 . R2, 1988-90 9 fat 2763 R3, 1989-90 25 fat 20256 R4, 1989-90 12 fat 5191 R5, 1989-90 21 fat 8632 R6, 1989-90 13 fat 4566 R7, 1990 10 fat 4280 R8, 1989-90 10 fat 3062 R9, 1989-90 18 fat 5985 RIO, 1989-90 5 fat 6819 R l l , 1989-90 29 fat 5565 R12, 1989-90 33 fat 5942 R13, 1990-91 12 fat 10873 R14, 1989-91 9 fat 7049 R15, 1990 18 fat 24316 R16, 1990 14 fat 22735 Svalbard, 1990-94 8 fat 11400 (4810-18300) 25 fat 16400 (5250-36700) 23 fat 15700 (4790-41500) 20 fat 28100 (6960-80300) 9 fat 16600 (5540-27700) (a) Pools or individuals (b) Total PCB levels (c) Locations are described in Table 7 39 Table 7. Geographical regions used to measure PCB levels among polar bear populations (adapted from A M A P , 1998) Sampling Sampling period Geographical location region R l April 1993 Wrangel Island(Russia)-Chukchi Sea R2 March 1988- Bering Sea & Bering Strait south of March 1990 Arctic Circle, Chukchi Sea & Goodhope Bay, Alaskan coast to 155°W R3 April 1989- McLure Strait & the adjacent Arctic May 1993 Ocean R4 December 1989- Amundsen Gulf & Beaufort Sea to May 1990 135°W R5 April 1989- Viscount Mellville Sound west of May 1990 100°W R6 December 1989- Queen Maud Gulf & Larsen Sound May 1990 R7 January 1990- Barrow Strait & Cornwallis Island May 1990 R8 December 1989- Gulf of Boothia May 1990 Baffin Bay north of 72°N-Lancaster R9 April 1989-June 1990 Sound-Jones Sound-Kane Basin-Thule-Ellesmere Island RIO December 1989- Southern Baffin Bay and Northern January 1990 Davis Strait R l l October 1989- Foxe Basin & Hudson Strait west of April 1990 72.5°W R12 August 1989- Western Hudson Bay (Cape September 1990 Churchill area) R13 January 1990- Eastern Hudson Bay (Belcher April 1991 Islands) R14 December 1989- Davis Strait below the Arctic Circle March 1991 & Hudson Strait east of 72.5°W R15 January 1990- East coast of Greenland near July 1990 Scoresbysund R16 March 1990- Svalbard (Norway) April 1990 40 Mossner and Ballschmiter (1997) reported PCB concentrations in marine mammals from three different locations, the North Atlantic, the North Pacific, and the Bering Sea/Arctic Ocean. PCB levels were measured in the blubber of seals (harbor seal, northern fur seal), toothed whales (beluga whale, common dolphin, pilot whale), and baleen whales (bowhead whale). PCB#28, PCB#52, PCB#101, PCB#118, PCB#138, PCB#153 and PCB#180 were used as indicators of PCB exposure and represented approximately 25% of total PCB levels in all species. Comparison of PCB levels in the various species examined suggested that marine mammals from the western North Atlantic were 15 times more contaminated than the animals from the eastern North Pacific and the Bering Sea/Arctic Ocean. There was a positive association between trophic level and PCB bioaccumulation. Total PCB levels were increased at higher trophic levels in the North Atlantic (harbor seal > common dolphin > pilot whale) and the North Pacific/Arctic Ocean (northern fur seal > beluga whale > bowhead whale). The harbor seal and northern fur seal are predators and are at a high trophic level. The beluga whale feeds on a variety of prey, especially herring, squid and cod, and are at a higher trophic level than the bowhead and pilot whales, which are close to the bottom of the food web. There was also an increase in PCB metabolism at higher trophic levels, except for persistent congeners such as 4,4'-chloro-substituted PCBs. 41 1.4.2. H U M A N EXPOSURE Humans are exposed to PCBs through the ingestion of environmental samples, namely air, drinking water, and food contaminated with PCBs. However, the primary source of human PCB exposure is via food, especially, meat, poultry, and fish (WHO, 1993). In the general population, detectable levels of PCBs were present in blood adipose tissue, ovarian follicular fluid, and breast milk (Jensen, 1983; Jensen, 1983; Safe, 1994; Jarrell et al, 1993). Chemical analysis of human breast milk extracts revealed that the composition of the human breast milk extract was different from that of Aroclor 1260. However compared to other Aroclor mixtures, Aroclor 1260 was more similar to breast milk in terms of congener composition. Certain PCB congeners such as 2,4,5,4',-tetrachlorobiphenyl (PCB#74) were present in human breast milk at high concentrations (3.7-11% of total PCBs), but were minor components of Aroclor 1260 (0.03% of total PCBs). Nevertheless, the three ortho-VCBs that are prevalent in human breast milk (27.3% of total PCBs), namely, 2,4,5,2',3',4'-hexachlorobiphenyl (PCB#138), 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153) and 2,3,4,5,2',4',5'-heptachlorobiphenyl (PCB#180) are also major components (25.2% of total PCBs) of Aroclor 1260 (Table 8). Previous studies determined that the average total PCB concentration in human breast milk fat ranges between 0.5 and 4.0 ppm per lipid weight (Jensen, 1983; Jensen, 1987; Rogan et al, 1987; Wickizer et al, 1981). Given the human breast milk fat concentration of 2.5-4.5%), PCB concentration in whole breast milk ranges between 10 and 180 ppb. Kutz et al. (1991) reported that data published by the National Human Adipose Tissue Survey (NHATS), conducted in 1982, revealed that approximately 94% of the general US population had detectable PCB levels in the fat. Adipose PCB concentrations were less 42 than 1 ppm in 66% of the population, greater than 1 ppm in 29% of the population, and greater than 3 ppm in 5% of the population. The percentage of individuals with detectable PCB levels was lowest in the 0-14 year-old-age group. It was also found that the percentage of the US population with PCB concentrations between 1 and 3 ppm decreased from over 62% in 1972 to less than 2% in 1984 (Kutz et al., 1991). Certain populations are exposed to significantly higher PCB levels compared to the general population. In general, PCB levels are over ten times higher in occupationally exposed individuals than the general population (Wollf, 1985). According to the National Institute for Occupational Safety and Health (NIOSH) 1977 report, approximately 12,000 individuals per year were exposed to PCBs in the workplace. PCB levels in human plasma were higher than background levels in over 70% of subjects that were occupationally exposed to PCBs.. Background levels were defined from the analysis of plasma samples of individuals who were not exposed to PCBs. The congeners measured were PCB#138, PCB#153, and PCB#180 (Daniel et al., 2001). The body burden of PCBs is also high in Arctic populations that consume larger amounts of fish than the general population. Total PCB levels in human milk from women from Southern Quebec represented a tenth of the values reported for Arctic Quebec (Table 5). According to the 1998 A M A P report, the average daily intake (ADI) of PCBs from traditional food exceeded the tolerable intake (TDI) of 50 uk/kg in 16% of Inuit women from Baffin Island. 43 Table 8. Quantitative and qualitative analysis of PCBs in Aroclor 1260 and a human breast milk extract (adapted from Safe, 1994). Congener Percentage Percentage Congener Percentage Percentage name in Aroclor in human name in Aroclor in human 1260 milk 1260 milk PCB-018 0.12 - PCB-118 0.49 6.5 PCB-017 0.05 - PCB-134 0.35 PCB-024 0.01 - PCB-114 - 0.33 PCB-016 0.04 - PCB-131 0.07 -PCB-029 0.02 - PCB 122 0.12 0.53 PCB-026 0.02 - PCB-146 1.3 1.9 PCB-028 0.04 8.8 PCB-153 9.6 12.0 PCB-021 0.01 - PCB-141 2.5 0.29 PCB-033 0.09 2.2 PCB-176 0.33 -PCB-053 0.04 - PCB-137 0.22 0.87 PCB-022 0.01 0.65 PCB-130 - 0.59 PCB-045 0.07 - PCB-138 6.5 10.0 PCB-046 0.02 0.25 PCB-158 0.70 0.55 PCB-052 0.25 1.9 PCB-129 0.20 -PCB-043 0.02 - PCB-178 1.2 -PCB-049 0.06 0.66 PCB-175 0.49 -PCB-048 0.29 0.37 PCB-187 4.5 1.5 PCB-044 0.11 0.78 PCB-183 2.3 1.4 PCB-037 0.04 2.9 PCB-128 0.47 0.33 PCB-042 0.04 - PCB-167 0.16 0.85 PCB-041 0.25 1.3 PCB-185 4.1 0.11 PCB-040 0.0 - PCB-174 5.5 0.39 PCB-100 0.02 - PCB-177 1.9 0.61 PCB-074 0.03 11.0 PCB-171+202 1-2 0.37 PCB-070+076 0.15 0.61 PCB-156 0.45 4.87 PCB-095 2.7 - PCB 173 0.06 -PCB-091 0.07 - PCB-200 0.78 -PCB-056+060 0.14 0.71 PCB-157 - 0.47 PCB-084 0.65 - PCB-172 0.78 0.31 PCB-101 • 2.5 0.97 PCB-180 9.1 5.3 PCB-099 0.13 4.8 PCB-193 0.47 0.19 PCB-119 - 0.08 PCB-191 0.10 0.90 PCB-083 0.04 - PCB-199 0.33 -PCB-097 0.45 - PCB-170 6.8 5.3 PCB-087 0.45 0.82 PCB-201 2.9 0.85 PCB-085 0.13 - PCB-203 3.1 0.79 PCB-136 1.4 - PCB-196 2.5 0.18 PCB-110 1.7 1.0 PCB-189 0.15 2.4 PCB-154 0.02 - PCB-195 3.1 0.31 PCB-082 0.11 - PCB-207 0.08 -PCB-151 2.5 0.59 PCB-194 1.7 0.48 PCB-144+135 1.5 0.51 PCB-205 0.11 0.06 PCB-107 0.03 0.31 PCB-206 0.85 0.24 PCB-149 7.4 - PCB-209 0.06 0.09 44 1.5. HYPOTHESES Previous studies demonstrated that PCBs elicit a spectrum of biological effects, including suppression of humoral immune response (Harper et al, 1995), changes in serum levels of reproductive hormones (Desaulniers et al, 1999), alteration of thyroid hormone homeostasis and function (Byrne et al, 1987), and induction of C Y P enzymes (Ngui and Bandiera, 1999) in experimental animals. However, it is not known whether the effects mentioned above occur at low dosages (less than 6.25 mg/kg/day) comparable to environmental PCB levels. The specific hypotheses to be tested in my research study are the following: (1) Aroclor 1260 suppresses humoral immune response in rats at low dosages. (2) Aroclor 1260 disrupts endocrine function in rats at low dosages. (3) Hepatic CYP2B induction is a bio-marker of Aroclor 1260 exposure in rats. 1.6. OBJECTIVES In order to determine PCB concentrations at which immune and endocrine effects are observed in the rat, a common rodent model, it is necessary to obtain dose-response data for each endpoint studied. The specific objectives of this study are: (1) To investigate dose-response effects of treatment with Aroclor 1260 on antibody response in rats (anti-SRBC). (2) To investigate dose-response effects of treatment with Aroclor 1260 on serum T4, TSH, L H , and testosterone levels in rats. (3) To investigate dose-response effects of treatment with Aroclor 1260 on hepatic CYP1A and CYP2B induction in rats. 45 2. EXPERIMENTAL 2.1. REAGENTS Reagents were obtained from the following sources: Anachemia (Montreal, Quebec, Canada): Hydrochloric acid, hydrogen peroxide. BDH Chemicals (Toronto, Ontario, Canada): Ethylenediaminetetraacetic acid (EDTA), disodium salt; Folin and Ciocalteu phenol reagent; magnesium chloride; potassium chloride; sodium azide; sodium carbonate, anhydrous; sodium hydroxide; sucrose; sulfuric acid. Bio-Rad Laboratories (Mississauga, Ontario, Canada): Acrylamide 99.9%; N,N'-methylene-bis-acrylamide (BIS); P-mercaptoethanol; N,N,N',N'-tetramethylethylenediamine (TEMED). Biosource International (Camarillo, California, U.S.A.): Horseradish peroxidase-conjugated mouse (monoclonal) anti-rat IgM, affinity purified; alkaline phosphatase conjugated, goat F(ab')2 anti-mouse IgG, gamma and light chain specific, affinity purified; alkaline phosphatase conjugated, goat F(ab')2 anti-rabbit IgG, gamma and light chain specific, affinity purified. Boehringer Mannheim Canada Ltd. (Laval, Quebec, Canada): Nicotinamide adenine dinucleotide phosphate, tetrasodium salt (NADPH). Carnation Inc. (Toronto, Ontario, Canada): Skim milk powder Colorado Serum Co. (Denver, Colorado, U.S.A.): Sheep blood diluted 1:1 in Alsever's solution. 46 Fischer Scientific Ltd. (Vancouver, British Columbia, Canada): Acetonitrile, HPLC-grade; methylene chloride HPLC-grade; glycerin; glycine; methanol; methanol, HPLC-grade; potassium phosphate monobasic; sodium chloride; sodium citrate; sodium dodecyl sulphate (SDS); sodium phosphate. ICN Biomedicals Canada Ltd. (St-Laurent, Quebec, Canada): Bovine serum albumin (BSA), globulin and fatty acid free, fraction V ; testosterone radioimmunoassay (RIA) kit, antibody coated tube; total thyroxine (T4) RIA kit, antibody coated tube; tris(hydroxymethyl)aminomethane (Trizma base). J.T. Baker Chemical Co. (Phillipsburg, New Jersey, U.S.A.): Sodium dithionite Mandel Scientific Company Ltd. (Edmonton, Alberta, Canada) Blotting paper; nitrocellulose membrane. Pierce (Rockford, Illinois, U.S.A.): 4-Nitro-blue tetrazolium chloride; bicinchoninic acid (BCA) protein assay kit. Praxair (Vancouver, British Columbia, Canada): Carbon monoxide gas, 99.5% purity; nitrogen gas. Schwarz/Mann Biotech (Cleveland, Ohio, U.S.A.): Ammonium persulphate Sigma Chemical Co. (St. Louis, Missouri, U.S.A.): Bromphenol blue; corn oil; cupric sulfate pentahydrate; dimethyl sulfoxide; o-phenylenediamine; sodium potassium tartrate tetrahydrate; tris(hydroxymethyl)aminomethane hydrochloride (Trizma HC1); poloxyethylene sorbitan monolaurate (Tween 20). 47 Steraloids Inc. (Wilton, New Hamsphire, U.S.A.): Testosterone; 2a-hydoxytetosterone; 2 P-hydoxytetosterone; 6 P -hydoxytetosterone; 7 a-hydoxytetosterone; 11 P-hydoxytetosterone; 16 a-hydoxytetosterone; 16 P-hydoxytetosterone. Xymotech Biosystems (Mt. Royal, Quebec, Canada): 5-bromo-4-chloro-3-indolyl phosphate disodium salt. Dr. S. M. Bandiera (Faculty of Pharmaceutical Sciences, University of British Columbia, Vancouver, British Columbia, Canada): Purified rat CYP1A1; Purified rat CYP2B1; polyclonal rabbit anti-rat CYP2B1 IgG. Dr. P. Thomas (Rutgers University, Piscataway, New Jersey, U.S.A.) Monoclonal mouse anti-rat CYP1A1 IgG, CD 5.2 Mab. Dr. S. Safe (Texas A&M University, College Station, Texas, U.S.A.) Aroclor 1260 (same batch as that used and analyzed by Jason Ngui). The congener composition of the Aroclor 1260 sample used in this study was similar to the composition of the Aroclor 1260 reported by Frame et al. (1996) and Newman et al. (1998). 2.2. PREPARATION OF SRBC M E M B R A N E ANTIGENS The procedure for SRBC membrane antigen preparation is similar to that described elsewhere (Temple et al, 1993). A l l procedures were conducted at 4 °C. Sheep blood was spun at 1,000 x g for 15 min. The plasma was removed carefully. Cells were washed in 0.9% saline, spun at 1000 x g for 10 min, and re-suspended in 0.05 M 48 Tris-HCl buffer, containing 0.1 m M EDTA, pH 7.6. Cells from 25 ml of sheep blood were suspended in 40 ml of Tris-HCl buffer. This procedure was repeated four times to remove all visible traces of hemoglobin. The pellet was washed with saline solution, spun at 1000 x g for 10 min and re-suspended in 2.5 ml 0.1% SDS. The SRBC membrane preparation was stored at -20 °C as 0.5 ml aliquots. 2.3. A N I M A L T R E A T M E N T Adult male Long-Evans rats (7-8 weeks old, 250-300 g) were obtained from Charles River Canada, Inc. (Montreal, Quebec). Animals were housed two per cage, and given Certified Rodent Diet (5002, PMI® Feeds, Inc., Richmond Indiana) and water ad libitun. Animals were divided into ten treatment groups (n=6) and two control groups. After one week of acclimatization, treatment groups received Aroclor 1260 in corn oil (2.5 ml/kg/day) by oral gavage at dosages of 0.025, 0.13, 0.625, 1.25, 3.13, 6.25, 15.6, 31.3, 78.1, and 156 mg/kg/day. One control group (n=6) was treated with the corn oil vehicle at a dosage of 2.5 ml/kg/day and a second group was untreated (n=9). Animals were treated daily for ten days and killed two days after the last oral dose of PCB. On the fourth day of treatment, all animals were injected with sheep red blood cells (2><108 SRBCs). Body weights were measured before each treatment. Blood samples were, collected on day 10 for anti-SRBC IgM determination and at the time of death for hormone assays (Figure 7). Liver, thymus, testes, seminal vesicle and ventral prostate were carefully dissected, blotted dry, and weighed. Liver samples were used for the preparation of hepatic microsomes. 49 1 st dose of i.v. injection of last dose of kil l and collect blood Aroclor 1260 SRBC (2x108 cells) Aroclor 1260 liver, thymus, testes blood sample prostate, and seminal taken vesicle Figure 7. Study design 2.4. S A M P L E HANDLING A N D SERUM ANTI-SRBC IgM ANALYSIS B Y ELISA Animals were bled by tail nicking six days after SRBC immunization. Blood samples were allowed to clot at room temperature for 1-1.5 hour. Serum was separated by centrifugation at 850 x g for 20 min. Serum IgM was titrated by direct non-competitive ELISA because SRBC-specific IgM was not available to generate a calibration curve. Typically, ten serial (two-fold) dilutions of each sample were assayed. The SRBC membrane antigen preparation was diluted in 0.1 M sodium carbonate-bicarbonate buffer, pH 9.5, to a protein concentration of 0.67 (ig/ml. The diluted antigen preparation was added (150 ul/per well) to a Nunc-Immuno microtiter plate (A/S Nunc, Roskilde, Denmark) and incubated overnight at 4 °C. The next day, each well was washed three times with 150 ul of 0.01% Tween, for 1-2 min each time. Each well was blocked with 150 ul of 0.05% Tween in phosphate-buffered saline (PBS) containing 0.137 M NaCl, 2.6 mM KC1, 8.1 mM sodium phosphate, and 0.15 M potassium phosphate, pH 7.4, at 37°C in a shaking water bath for 2 hours. After blocking, ten (two-50 fold) serial dilutions of a serum sample were carried out and a 150 ul aliquot of each serum dilution was loaded into a well. A l l dilutions were loaded in duplicate. This procedure was repeated for each serum sample, and the wells were incubated for 2 hours at 37°C with shaking. Each well was washed as before and incubated with 150 ul of horseradish peroxidase-conjugated mouse anti-rat antibody (500 ng/ml) for 1 hour at 37°C with shaking. Finally, each well was washed thoroughly and incubated with 150 ul of peroxidase substrate consisting of 92.2 mM o-phenylenediamine, and 0.01% hydrogen peroxide in 0.1 M sodium citrate buffer, pH 5.0, for 10 min. The sodium citrate buffer contained 0.035 M citric acid monohydrate and 0.065 M trisodium citrate dihydrate. The reaction was stopped by the addition of 40 (al of 4 M sulfuric acid to each well and optical density (OD) ie absorbance values were measured at 490 nm using a BioTek plate reader (Bio-Tek Instruments, Winooski, Vermont, USA). 2.5. HEPATIC MICROSOME PREPARATION Immediately after decapitation, each liver was removed, perfused, and homogenized in 20 ml of ice-cold 0.05 M Tris/KCl buffer, pH 7.5, containing 1.15% KC1, (4 ml/g wet liver weight) in a Potter Elvehjem tissue grinder. Liver microsomes were prepared according to the method of Thomas et al. (1983). The homogenates were centrifuged at 9,000 x g at 5°C for 20 min in a Beckman Model J2-21 centrifuge (Beckman Instruments, Palo alto, CA, USA). The supernatants, the S9 fractions, were filtered. The S9 fractions were spun at 105,000 x g at 5°C for 60 min using either a Beckman L5-50 or Beckman L8-60 ultracentrifuge. The pellets were freed from glycogen and lipid, and then re-suspended in ice-cold 10 m M EDTA, 1.15 % KC1, pH 7.4. The 51 suspensions were spun at 105,000 * g at 5°C for 60 min and the pellets were saved. The microsomal pellets were suspended in small volumes (2-3 times pellet volume) of ice-cold 0.25 M sucrose and stored as 1.5 ml aliquots at -75°C. 2.6. DETERMINATION OF S E R U M T4 A N D TESTOSTERONE CONCENTRATIONS B Y RADIOIMMUNOASSAY Trunk blood samples drawn at the time of death were allowed to clot at room temperature for 1-1.5 hour. Serum was separated by centrifugation at 850 x g for 20 min. Serum T4 and testosterone concentrations were measured with antibody-coated tube RIA kits supplied by ICN Biomedicals Canada Ltd. (ICN Biomedicals Canada Ltd., St-Laurent, Quebec, Canada). 2.7. DETERMINATION OF TOTAL C Y P CONCENTRATION The total C Y P concentrations of the hepatic microsome samples were measured from the reduced carbon monoxide difference spectrum as described by Omura and Sato (1964a). Each liver microsome was diluted either 1:20 for control samples or 1:50 for highly induced samples (samples from rats treated with Aroclor 1260 at dosages greater than 3.13 mg/kg/day) in 0.1 M sodium phosphate buffer, pH 7.4, containing 20% glycerol, 0.1 m M EDTA. The diluted sample was reduced with sodium dithionite in reference and sample cuvettes. The aliquot in the sample cuvette was saturated with carbon monoxide. The difference absorption spectrum was obtained using a S L M -Aminco DW-2 spectrophotometer (SLM Instruments Inc. Urbana, IL, U.S.A.). Total 52 C Y P concentration was determined using the difference between A450 and A 4 9 0 , and a molar absorptivity of 91 cm /mmol as described by Omura and Sato (1964b). 2.8. DETERMINATION OF TOTAL PROTEIN CONCENTRATION The protein concentration of the SRBC membrane antigen preparation was determined by the bicinchoninic acid (BCA) assay using a B C A assay kit supplied by Pierce. Protein concentration was measured in each hepatic microsome sample as described by Lowry et al. (1951). Spectrophotometric analysis was performed at a wavelength of 650 nm using a Shimadzu UV-160 UV-Visible recording spectrophotometer (Shimadzu Corporation, Kyoto, Japan). BSA standards were used to generate calibration curves. 2.9. SODIUM D O D E C Y L SULPHATE P O L Y A C R Y L A M I D E GEL ELECTROPHORESIS (SDS-PAGE) Hepatic microsomal proteins were resolved by sodium dodecyl sulfate polyacrylamide gel electrophoresis (SDS-PAGE) using a Hoefer SE 600 vertical slab gel unit (Hoefer Scientific Instruments, San Francisco, CA, USA) according to the procedure of Laemmli (1970). The gels used were 0.75 mm thick, and consisted of a stacking gel and a separating gel. The stacking gel was 1 cm high, and contained 3% (w/v) BIS (22.2%-0.6% w/w), 0.125 M Tris-HCl (pH 6.8), 0.1% (w/v) of SDS, 0.08% of AP, and 0.05%) (v/v) of TEMED. The separating gel was 12.5 cm high and consisted of 7.5% (w/v) BIS (22.2%-0.6% w/w), 0.375 M Tris-HCl (pH 8.8), 0.1% (w/v) of SDS, 0.042% 53 (w/v) of AP , and 0.03% (w/v) of TEMED. Hepatic microsome samples were diluted in sample dilution buffer containing 0.062 M Tris-HCl (pH 6.8), 1.0% SDS, 0.001% bromphenol blue, 10% glycerol, and 5.0%> (3-mercaptoethanol. The samples were heat-denatured by boiling for 2 min. A volume of 20 ul of each sample was injected into a well in the stacking gel. The samples were subjected to electrophoresis at a constant current of 11 mA per gel for approximately 1 hour to effect migration through the stacking gel, and at 22 mA for 2.5 hours, with constant cooling, until the dye front reached the bottom of the separating gel. 2.10. I M M U N O B L O T ANALYSIS Proteins separated by SDS-PAGE were transferred electrophoretically (0.4 A for 2 hours at 4°C) onto a 13 x 14 cm nitrocellulose membrane according to the procedure of Towbin et al. (1979), using a Hoefer transfer apparatus, model TE52. After completion of the transfer, the membrane was soaked in blocking buffer consisting of 1% (w/v) BSA, 3% (w/v) skim milk powder in modified PBS, pH 7.4, and stored overnight. Modified PBS contained 0.137 M NaCl, 2.6 m M KC1, 8.1 m M sodium phosphate, 0.15 M potassium phosphate, and 0.2 m M EDTA. The blocking buffer was discarded. The membrane was incubated with primary antibody (monoclonal mouse anti-rat CYP1A1 IgG at 1.0 ug/ml or polyclonal rabbit anti-rat CYP2B1 IgG at 2.0 ug/ml) in antibody dilution buffer consisting of 1%> (w/v) BSA, 3% (w/v) skim milk powder, and 0.05% (v/v) Tween 20 in modified PBS, at 37°C in a shaking water bath for 2 hours. The membrane was washed three times with washing buffer (0.05%> (v/v) Tween 20 in modified PBS) and incubated with secondary antibody (either alkaline phosphatase 54 conjugated, goat F(ab')2 anti-mouse IgG or alkaline phosphatase conjugated, goat F(ab')2 anti-rabbit IgG), at 1:3,000 dilution in antibody dilution buffer for 2 hours. The membrane was washed and rinsed with distilled water. The membrane was incubated with substrate solution (0.03% (w/v) NBT and 0.015% (w/v) BCIP) in 0.1M Tris-HCl buffer containing 0.5 m M M g C l 2 , pH 9.5, at room temperature for 2.5 min, under dim light. The substrate solution was discarded to stop the reaction. The membrane was rinsed with distilled water and blotted dry. To ensure that the intensities of the immuno-reactive bands were within the linear range of the calibration curves, microsome samples were typically loaded onto the gel at concentrations ranging from 0.05 to 1.0 nmol total CYP per lane. The staining intensity of the protein bands on the nitrocellulose membrane was measured by computer image analysis with a pdi 420oe densitometer (PDI, Inc., New York, N Y , U.S.A.) equipped with an A G F A Arcus II scanner using the pdi Quantity One® 3.0 software. The contour quantity (OD x mm2), that is the optical density times the contour area, was determined for each band and was divided by the contour quantity of a known standard used as the internal standard. A calibration curve, a plot of the ratio of OD x m m 2 values of the purified standards and the internal standard versus CYP protein amount per lane was used to determine CYP1A and CYP2B protein concentrations in all samples. 55 2.11. TESTOSTERONE H Y D R O X Y L A S E A S S A Y O H O H 16a H H 19 14 IS H Cytochrome P450 10 O' NADPH, 0 2 Figure 8. Regiospecific hydroxylation of testosterone by C Y P enzymes. Arrows indicate selected positions for hydroxylation (taken from Ickeinstein, 1999). Testosterone hydroxylase activities were measured in hepatic microsomes as described by Sonderfan et al. (1987) with modifications as reported by Ngui and Bandiera (1999). The oxidation of testosterone is catalyzed by C Y P enzymes in the presence of N A D P H and oxygen, and results in the formation of hydroxylated products that are readily resolved by HPLC (Figure 8). The reaction mixture consisted of 50 |iil of hepatic C Y P enzymes at a concentration of 6 nmol total CYP/ml, 0.92 ml of 50 m M potassium phosphate buffer, pH 7.4, containing 3 m M M g C l 2 , and 10 ul of 100 m M N A D P H . The mixture was preincubated at room temperature for 10 min. The reaction was initiated at 37°C by the addition of 20 ul of 12.5 m M testosterone, and was terminated after 5 min with the addition of 6 ml of dichloromethane and vigorous mixing. The mixture was spiked with 50 ul of 50 m M 11 p-hydroxytestosterone, which served as an internal standard, and mixed thoroughly. Each sample was spun at 2000 rpm for 1 min. The aqueous layer was removed by aspiration, and the remaining organic phase was dried under a gentle stream 56 of nitrogen gas. The reaction products were reconstituted with 200 ul of HPLC-grade methanol, and filtered through 13 mm, 0.45 um syringe filters. The filtrates were analyzed as 10 jul aliquots by reverse phase HPLC at 40 °C at a total flow rate of 2 ml/min. A Supelcosil (Supelco, Bellefonte, PA, U.S.A.) LC-18 octyldecylsilane column (5 urn particle size, 15 cm x 4.6 mm), was preceded by a Supelcosil LC-18 guard column (40 um particle size, 2 cm x 4.6 cm), and connected to a Shimadzu LC-6A binary gradient HPLC system equipped with SPD-6A U V detector, a CTO-6A oven, and an SIL-6B autosampler. Each sample was eluted with the following time program: 100% solvent A (methanol:water:acetonitrile, 35:64:1) from 0 to 10 min, followed by a linear gradient of solvent B (methanol:water:acetonitrile, 80:18:2) from 0 to 100% from 10 to 29 min, 100% solvent B from 29 to 31 min, followed by a linear gradient to 100% solvent A from 31 to 32 min, and equlibration with 100%) solvent A till 34 min. The area under the curve (AUC) was integrated for each metabolite using a Shimadzu CR501 data processor. The metabolites were identified in each sample by comparing the retention times of the metabolite peaks to those of a known standard metabolite mixture. A calibration curve was generated for each testosterone standard by plotting the A U C ratio (standard metabolite internal standard) versus the amount of metabolite per tube. For a given metabolite, the amount of metabolite formed in each microsomal sample was calculated using the slope of the appropriate calibration curve. Finally, the rate of metabolite formation was computed by dividing the amount of metabolite formed per mg protein microsomal protein or per nmol total C Y P by the reaction time. 57 2.12. STATISTICAL ANALYSIS Data were analyzed by one way analysis of Variance (ANOVA), and differences between means were tested by the Student Newman-Keuls test (SNK). A p value less than 0.05 was considered statistically significant. Linear regression analysis was performed using Microsoft Excel version 97 (Microsoft Corporation, Redmond, W A , U.S.A.). A l l dose response curves were fitted using Sigmaplot version 1.01 (Jandell Scientific, San Rafael, CA) using equation (1) below. E = Em a x/(1 + exp(-(D- E D 5 0 )/A)) (1) Where: Em ax is the maximal response, ED50 is the dose at half-maximal response, and A is a constant. The use of equation (1) for dose-response curve fitting does not rely on any assumption about the mechanism of action of the drug. The induction of hepatic enzymes by PCBs is A h receptor mediated; however, the immune and endocrine effects of PCBs occur through multiple mechanisms, which have not been well characterized. Therefore, knowledge of such mechanisms was not required and the same equation was used to fit all dose-response data. 58 2.13. A S S A Y VALIDATION A l l assays were validated to ensure that the responses measured were reproducible and accurate. The intra- and inter-assay coefficients of variation, the limit of detection, and the limit of quantitation of each assay were determined as part of the assay validation process. The coefficient of variation is a measure of the precision and reproducibility of the assay. The intra-assay coefficient of variation is obtained by measuring the amounts of analyte or standard, five to six times, on the same day. For example, if the assay involves the determination of amount of analyte from a calibration curve, then five to six calibration curves, generated on the same day, are required for a determination of the intra-assay coefficient of variation. The calibration curve is a plot of response (example: the A U C ratio) versus amount of analyte. The calibration curve is obtained with varying concentrations of authentic standards. For a given standard, the intra-assay coefficient of variation is calculated by dividing the standard deviation by the mean (of five to six measurements) multiplied by a 100%. Alternatively, the mean slopes of five to six calibrations curves generated on the same day can be used to calculate the intra-assay coefficient of variation. The inter-assay coefficient of variation is calculated by a similar method, except that calibration curves or measurements obtained on five to six different days are used. The limits of detection (LOD) and the limits of quantitation (LOQ) were determined using the intra-assay coefficients of variation at different concentrations of analytes. The LOQ is the smallest amount of analyte that can be quantified reliably and is equal to the smallest amount of analyte measured in the assay, with a coefficient of 59 variation of less than 15%. The LOD is defined as the smallest amount of analyte that can be detected reproducibly with an assay response that is at least three times above the background noise of the assay. The coefficient at the LOD may be greater than 15%. Assay recovery is the percentage of analyte present in a sample following sample handling. It is calculated by dividing the initial amount of analyte present by the amount of analyte present in the sample after processing, multiplied by a 100%). 2.13.1. VALIDATION OF THE ELISA A S S A Y TO M E A S U R E ANTI-SRBC IgM TITER The ELISA for measuring anti-SRBC IgM titer was validated using an experimental control sample because anti-SRBC IgM standards were unavailable. The control sample was pooled sera from two SRBC-treated eight-week old male rats. The Q animals received 10 SRBCs by iv tail injection and were killed on the seventh day. It was not important to draw a blood sample on the sixth day as previously described in our study design (Figure 7), because any sample containing the anti-SRBC IgM can serve as an adequate control. The control sample was included in all ELISA assays. The optical density at 490 nm (OD490) of ten serial (two-fold) dilutions of the control sample were measured and plotted against the anti-SRBC IgM titer (the dilution factor) on a log-logit scale. A n example of plot is shown in Figure 9. The IgM titer or dilution factor that yields 1.0 OD was extrapolated from the linear portion of the curve. The intra-assay coefficient of variation of the ELISA for measuring anti-SRBC IgM was calculated from six anti-SRBC IgM titer determinations of the control sample conducted on the same day (see Table 9). The mean anti-SRBC IgM titer (to yield 1.0 OD) of the control sample was 60 75 ± 2, and the intra-assay coefficient of variation, calculated as the percent ratio of the standard deviation to the mean, was 6.5%. The inter-assay coefficient of variation of the ELISA was determined by measuring the anti-SRBC IgM titer of the control sample on six different days. The mean anti-SRBC IgM titer of the control sample was 108 ± 4, and the inter-assay coefficient of variation was 8.8%). These data indicate that the day-to-day variation was greater than the same-day variation and may have been due to the placement of the control sample in the outside wells of the plate. -0.6 H 1 • , , , , , 0 0.5 1 1.5 2 2.5 3 3.5 Log (anti-SRBC IgM titer) Figure 9. Determination of the anti-SRBC IgM titer to yield 1.0 OD of the control sample The logarithm of the OD490 values of ten serial (two-fold) dilutions of the control sample were plotted against the logarithm of their respective dilution factors (log (anti-SRBC IgM titer)). The linear portion of the curve was used to extrapolate a dilution factor or anti-SRBC IgM titer to yield 1.0 OD of 73. Table 9. Determinations of the anti-SRBC IgM titer of the control sample on the same day 61 Trial # 1 2 3 4 5 6 IgM titer to yiel 1.0 OD 73 66 77 74 81 79 2.13.2. VALIDATION OF THE RADIOIMMUNOASSAYS The T4 and testosterone radioimmnunoassays (RIAs) were validated by the manufacturer as reported on the information inserts included in the kits. 3.1.2.1. VALIDATION OF THE T4 RADIOIMMUNOASSAY The recovery of the T4 radioimmunoassay (RIA) was reported to range from 91% to 104%). The intra-assay coefficient of variation of the T4 RIA was reported to be between 3.3 and 8.1%, and its inter-assay coefficient of variation was between 5.3 and 11.4%). The sensitivity (LOQ value) of the T4 assay was determined to be 7.6 ng/ml. 3.1.2.2. VALIDATION OF THE TESTOSTERONE RADIOIMMUNOASSAY The recovery of the testosterone RIA was reported to be 100.4%>. The intra-assay of variation of the testosterone RIA was reported to be between 9.6%> and 13.0%>, and the inter-assay coefficient of variation was between 6.8 and 15.2%. The sensitivity (LOQ value) of the testosterone assay was 0.1 ng/ml. 62 2.13.3. V A L I D A T I O N OF THE IMMUNOBLOT ANALYSIS FOR CYP1A A N D CYP2B PROTEIN L E V E L S The intra-assay coefficient of variation of the immunoblot analysis of hepatic C Y P 1 A and CYP2B protein amounts was calculated from five calibration curves generated on the same day. The calibration curves were obtained by plotting the ratio of the integrated intensities (standard:internal standard) versus the amount of C Y P protein per lane (Table 10 and Figure 10). The integrated intensity of a given C Y P protein band was calculated as the product of the optical density of the band and of its area (mm2). The intra-assay coefficients of variation for CYP1A1, CYP1A2, and CYP2B1 were calculated at six protein concentrations. The amounts of C Y P standards used were 0.0125, 0.025, 0.05 pmol per lane for CYP1A2 protein, and 0.0313, 0.0625, and 0.125 pmol per lane for CYP1A1 and CYP2B1 proteins. Because CYP2B1 and CYP2B2 react equally well with anti-CYP2Bl IgG, the validation for CYP2B1 also applies to CYP2B2. The intra-assay coefficients of variation of the immunoblot analysis for CYP1A1, CYP1A2, and CYP2B1 proteins are summarized in Table 11. The intra-assay coefficients of variation of the immunoblot analysis were used to determine the limit of quantitation (LOQ) and limit of detection (LOD) values. The LOD values for CYP1A1 and CYP1A2 proteins were 0.0313 and 0.0125 pmol per lane, respectively (see Table 11). The LOD value for CYP2B1 and CYP2B2 protein was less than 0.0313 pmol per lane, because the intra-assay coefficient of variation was less than 15% at all CYP2B1 protein concentrations. The LOQ values for CYP1A1, CYP1A2, and CYP2B1 proteins were 0.0625, 0.025, and 0.0313 pmol per lane, respectively. bo c <L) CD .3 t5 I o a \< CN • OH u 13 Si) CD 00 CD '"s 1 o CD CN J5 < "d u B, S-l C/3 M d CD <L) o a O OH " I ^ ^ 2 \< U a m CN o as oo o o CN o © i n o o O o i—i CN o © o © © © -H -fl -H -fl -H -fl C-- o O N o oo CN VO m CN o oo CN © CN © © © 4^ i n CN i n i n m CN i n © o CN m © © © © © i n CN m r—1 CN m o o o o © © CN © m i n CN i n i n m CN m r-» © o <—1 CN CI © o o © O i n o CN m i n o CN oo o © o CN '—i © © o © o © -H -fl -H -H -fl -H r - CN CN oo m o t-> O < m o 1—1 CN oo © o O o >—< T-H © CN o o CN CN i n o O O o o o ,—1 CN o © © © © © -H -H -H -H -H -H CN r - ,—i © O N i n O N i n o O CN i n © i n © © © i n © 64 Figure 10. Mean calibration curves for quantitation of CYP1A1 (graph A), CYP1A2 (graph B), and CYP2B (graph C) enzymes. The integrated intensities (optical density x contour area) of the CYP1A and CYP2B protein bands were quantified by densitometry. The ratio of integrated intensity of the C Y P standard to the integrated intensity of the internal standard was calculated for each standard (see Table 10). .2 a 1 « •§ u > I o I o • i—i ^—» o ^ a cn -a . a d 1 " .2 .g '53 a g CN <: < U o a O H 'o I ? i 3 > .g '53 +-» o <3 < u o a O h • 6 s-oo m i n CN >n m CN o o <—i o © o 0s- 0s-i n I—H CN *—H i n CN o © i n CN o i n o 0s- 0s- 0s-i n O N i n rn o CN i—i CO m CN i n m CN o o '—i o © © CN t4—I o a <H-H <u S-, o ts O > PQ g CN S O H «3 ^ a a l o o 11—I • I—I S+H H—; o U <+H g § .a o u a o < & 0s- 0s- 0s-<—1 oo ^r' 00 i n '—i CN i n CN i n CN i n m i n © ' 0s- o x i n © CN CN CN m m i n o 68 In addition, the inter-assay coefficient of variation of the immunoblot analysis for CYP1A1 and CYP2B1 proteins was determined using calibration curves obtained on five different days. The inter-assay coefficient of variation was calculated at each standard concentration as described previously. The protein concentrations used were 0.125, 0.375, and 0.5 pmol/lane for both CYP1A1 and CYP2B1 (Table 12). The inter-assay coefficients of variation were greater than 15% for all C Y P standards, except for the 0.125 pmol/lane CYP2B1 standard. The reasons for such results were unknown. 2.13.4. VALIDATION OF THE TESTOSTERONE H Y D R O X Y L A S E A S S A Y The intra-assay coefficient of variation of the testosterone hydroxylase assay was determined using six concentrations of a mixture of authentic testosterone metabolites containing 60-, 7a-, 16a-, 160-, 2a-, 2|3-hydroxytestosterone, and androstenedione. The concentrations used were 6.25, 12.5, 25, 50, 100 and 250 uM of each testosterone metabolite. Testosterone metabolites were quantified at all concentrations of metabolite standard five times in the same assay on the same day. A representative chromatogram of testosterone metabolites is shown in Figure 11. The testosterone metabolite to internal standard (11 (3-hydroxytestosterone) A U C ratios were calculated for each metabolite. The mean A U C ratios and standard deviation values were computed, and the calibration curve, a plot of A U C ratio versus amount of testosterone metabolite per tube, was generated for each metabolite (Figure 12). The intra-assay coefficient of variation was calculated at each concentration for each testosterone metabolite by dividing the standard deviation (of the mean A U C ratio) by the mean A U C ratio (Table 14). Likewise, the inter-assay coefficient of variation was determined using six concentrations of the testosterone metabolites on five different days (Table 15). 70 Figure 11. Representative chromatogram of a mixture of hydroxylated testosterone metabolites. A testosterone metabolite standard mix was added to a reaction mixture containing 0.25M sucrose instead of hepatic microsomes. The concentration of each metabolite was 1 nmol/ml. The metabolites were extracted by dichloromethane and resolved by HPLC. And: Androstenedione; 15(3, 60,7a, 16a, 2a, and 2(3 correspond to 150-, 60-, 7a- , 16a-, 2a- , and 20-hydroxy-testosterone respectively. on fl O S-l S3 CD o c o o CO o ' S > 00 CD "o •§ a CD fl O >-i CD CO O +j 00 CD -t-» CD -fl -t—» «4-l O oo O S-l o o S-l o <: CD 3 A . i n CD fl. CN CD X) fl CD fl fl. i n © CD i n CN CD fl o a a . i n CN i n CN CN cn 0 0 CN i n cn o O o © o o © © o © o © -H -fl -H - d -H -fl -fl CN oo <o fl i n r--CN 0 0 oo >n cn i n oo i n CN CN CN CN CN CN O N O N r- r- oo o 1 o O o o o © o o p p o o © © o O © o © -H -H -H -H -H -H -fl O N ,—i CN O N 1 T—1 CN oo cn i n O N o O N O N O N © © ^ O © o © cn r - CN CN oo o o o O o .cn o o o © o o o © © © © © o o -H -fl -H -H -H -H -fl cn 0 0 cn CN CN o o CN O N cn O N cn cn cn •<fr i n cn i n i n o o o © © © o cn ^J- i n oo VO © o o o o o o o o © p o o o © o © © © o -fl -H -H -H -H -H -H cn cn o o i n i n m »—I i n CN oo 'O <—1 CN CN CN '—1 CN CN © © © © © © O i n oo 0 0 O O CN © o 1 o o o O o o © o o o o o © o © © o o -fl -H -H -H -H -fl -H ,—i oo ,—i oo oo ,—1 re o i—i O N o CN o *—I i—i o i—i 1—1 © © © © © © o cn oo cn cn o o o o © o o o o o O © © © © © o -H -fl -H -H +1 -H fl m cn i> o cn i n i n o i n r-- t~-o o o o o o © © o © © o J2 H o oa CN CN CD fl O "3 CD fl CD -t-» oo O T3 72 Figure 12. Calibration curves of testosterone metabolites Graphs A , B, C, D, E, F, and G are the calibration curves for 6p-, 7a-, 16a-, 16(3-, 2a-, 2p-hydroxy-testosterone, and androstenedione, respectively. 73 77 c o > o t3 .Si CD o O S i s Q O X> fl fl. i n X> fl fl. CN CD X> fl o s fl. XI fl o a A . i n © CD •§ s fl. m CN CD x> fl o s i n CN CD "3 X) 03 0 -^_, X> 1 fl i n i n i n CN i n i n i n © © © © o © © i n i n i n m i n m i n CN CN CN >—< CN CN CN © © © o © © © CN cn cn CN O N CN cn CN 0 0 CN CN CN cn -fl fl O r-; cn cn CN CN CN m cn CN CN CN 0 0 N O o O N i n N O CN in -fl fl fl ea a N O ca N O S CN O N CN O cn CN CN 0 0 oa CN i n cn N O N O N O cn © N O O N CN N O cn cn cn CN N O ,—i cn cn o -a c < -fl CD fl cr 0 ) S-l a> CD a _ - 3 CD 13 - A CD fl -4—" O CD fl O CD fl A ' G CD S-l » CD oo ' O S-l • f l JJ "3 T 3 fl fl < 78 OO u CO o a a o CO o oo s? CO CO <a CO 03 ">> O i-H X? o CO (D H—> <D X -4-» O 0s-G O s C > U-t o G <U o o 5? CO CO c3 i -<u -*-» G H o o G 'o IS <u o U •3 o a C N <U x es o a X G O a G . © X G .+-» "o a G . IT) C N 0) G C N ' I—I N O O i n d o C N o o C N C N C N C N — i o C N C N C N O C N N O C N O N © O N i n i n oo N O i n N O N O i n C N C N O N O N C N O N oo o i n C N i n N O r-->—i -a d C N N O O O ex N O a N O C O . N O C N C N el o • i—i T3 el -t—» CO O T 3 s C N C N T3 <U Ifl ' G G c3 et cr <u S-l > <D CO el * O H 13 -*-» o -4—* o el "—' d 79 The linearity of the testosterone hydroxylase assay with respect to protein (enzyme) concentration was evaluated using an Aroclor 1260-treated rat liver microsomal sample, A4, from the 156 mg/kg/day dosage group. The total C Y P content of this sample was 1.90 nmol CYP/mg protein, and its testosterone hydroxylase activity values (data shown in a later section) were close to the mean values for the 156 mg/kg/day dosage group. The testosterone hydroxylase activities of sample A4 were measured at four protein concentrations, namely 6, 3, 1.5, and 0.75 nmol total CYP/ml. The A U C ratios of each testosterone metabolite over a five-minute incubation period were calculated (Table 16), and a plot of A U C ratio versus total C Y P concentration was generated for each metabolite. Regression analysis indicated that all enzymatic activities were linear with respect to enzyme concentration, except for 2a-hydroxylase and 20-hydroxylase (Figure 13). The R z value was 1.00 for androstenedione formation, 60-, 7a-, 16a-, 160-hydroxylase activities, and was 0.92 for 20-hydroxylase activity. The R 2 value for 2a-hydroxylase was not determined because only one data point was obtained. 80 CO > Pi o o o o o o o o id d CN O N O O d o o d o o U NO c-> o r- ON NO • 3 - ON 0 0 ON < 1—1 NO CN ON oo o CN CN o O O © © © N d <u _d °5 o O . o O N O N CI O C N O N C N CI •—i © © O o T3 d oo O N O O N O O O o O H CO CO CO KS <D CO o T3 d o (-1 I D H ^ CO o in in •<d-m CN o >—1 o CN o © © © T3 d d o ON O CN OO o T3 d CN in o T3 d T3 d in ON CI © d o CO O +-» CO <D o I CO. <  — * CD -G +-> o <D d H •8 CO. NO a NO CO. NO a CN CO. CN <D d o <D d u -t—» CO O O o3 -*-» <D U .2 <u H - | + 3 >-> d • d 81 Figure 13. Linearity of the testosterone hydroxylase assay: effects of increasing protein (total CYP) concentration. Various testosterone hydroxylase activities were measured in sample A4 of the 156 mg/kg/day dosage group, at the concentrations of 0.75, 1.5, 3.0, and 6.0 nmol total CYP/ml. Graphs A , B, C, D, E and F represent 60-, 7a-, 16a-, 160-, 20-hydroxylase activities and androstenedione formation respectively. 84 (E) o S-l u < 1.600 -| 1.400 -1.200 -1.000 -0.800 -0.600 -0.400 - • 0.200 -0.000 -R z - 0.92 —r-2 —r-6 Total CYP concentration (nmol/ml) 0 A ! 1 1 , 0 2 4 6 8 Total C Y P concentration (nmol/ml) 85 3. RESULTS 3.1. EFFECTS OF THE V E H I C L E (CORN OIL) Rats were treated with various dosages of Aroclor 1260 in corn oil. To determine whether the corn oil vehicle had an effect on any of the endpoints measured, an untreated group of rats was included in the study. Corn oil treatment did not affect any of the endpoints investigated (Tables 17-22). Table 17. Effects of treatment with corn oil on body weight gain, and liver weight in adult male rats Treatment Final body weight (g) Body weight gain (g) Absolute liver weight (g) Relative liver weight (% of final body weight) Untreated 329 ± 15 85 + 4 14.0 ±0.5 4.3 ±0.1 Corn oil (2.5 ml/kg/day) 324 ± 23 78 ± 7 14.0 ±0.6 4.3 ±0.1 Values are expressed as mean ± SEM. The corn oil group consisted of six rats and the untreated group consisted of nine rats. Animals were treated with the corn oil vehicle by oral gavage once daily for ten days Table 18. Effects of treatment with corn oil on absolute thymus, testes, seminal vesicle, and prostate weights in adult male rats Dosage of corn oil (ml/kg/day) Thymus weight (g) Testes weight (g) Seminal vesicle weight (g) Ventral prostate weight (g) Untreated 0.52 ±0.02 3.20 + 0.12 0.94 ± 0.06 0.27 + 0.01 Corn oil (2.5 ml/kg/day) 0.47 ± 0.02 3.16 ± 0.12 0.87 ±0.05 0.27 ±0.01 Values are expressed as mean ± SEM. The corn oil group consisted of six rats and the untreated group consisted of nine rats. Animals were treated with the corn oil vehicle by oral gavage once daily for ten days. 86 Table 19. Effects of treatment with corn oil on serum anti-SRBC IgM levels, testosterone concentrations and T4 concentrations in adult male rats Treatment IgM titer to yield Testosterone T4 (ng/ml) 1.0 OD (ng/ml) Untreated 532 ± 35 5.58 ± 0.97 41.4 ±7 .3 Corn oil 483 ± 80 5.79 ± 1.56 45.9 ±4 .1 Values are expressed as mean ± SEM. The corn oil group consisted of six rats and the untreated group consisted of nine rats. Animals were treated with the corn oil vehicle by oral gavage once daily for ten days Table 20. Effects of treatment with corn oil on protein concentration and total C Y P content in hepatic microsomes from adult male rats Treatment Total C Y P Protein Total C Y P Relative liver concentration concentration Content weight (% of (nmol/ml) (mg/ml) (nmol/mg final body protein) weight) Untreated 24.9 ± 3.6 33.8 ±3 .4 0.72 ± 0.08 4.3 ±0 .1 Corn oil 25.5 ±2.2 36.7 ±3.0 0.70 ±0.02 4.3 ±0 .1 Values are expressed as mean ± SEM. The corn oil group consisted of six rats and the untreated group consisted of nine rats. Animals were treated with the corn oil vehicle by oral gavage once daily for ten days. Table 21. Effects of treatment with corn oil on hepatic CYP2B and CYP1A protein levels Treatment CYP2B1 (pmol/mg CYP2B2 (pmol/mg CYP1A2 (pmol/mg protein) protein) protein) Untreated 5 ± 0 6 ± 1 3.5 ±0 .9 Corn oil 6 ± 2 4 ± 1 2.0 ± 0.4 Values are expressed as mean ± SEM. The corn oil group consisted of six rats and the untreated group consisted of nine rats. Animals were treated with the corn oil vehicle by oral gavage once daily for ten days 87 CO -s i-t 13 a -+-• 13 o3 a o CO &> a o CO O I-I o a o O H a> .g CO O 03 ! D CO jo o >-< 13 > ^ - G d O CO <D O o o d a 13 is o o to CN d H o O H 00 a o 3 >, '-+-> o o3 o '3 a >^  O <u O H 00 d o T3 <D d <u -*-» CO O s-i 13 C O . a CN co. N O a N O d 4> 03 O NO CN 0 0 2 4 ! oo c-~ ^ -H o >o CN O o CN -H OO m m -H o N O C O CN -H O C N O N C O -H i 3 <D 13 i) j -d o O 0 0 —< 4 1 o C O C N 2 -H NO O NO O N C O i n CN 4 1 oo N O N O 4 1 o O N O i—i a 0 0 O N N O O r--1—I CN -H CN -H >n N O CN C O CN 4 1 O o >n m CN N O 4 1 CO 5? 13 O H a d £ 00 eg 13 a) id o d o <u 00 03 > d d <u . d -*-» 13 03 03 co M » •—i 03 03 t- >-< X ° rr, CO <H-H o I 13 D e o O CO <u •55 > d rd 8 d O o 00 u , r G rd +-• o x I'S O T 3 - d t3 L S-i CO l> c3 d S '5 co < 03 i3 ^ <L> 08 CO K CO CD <U O H x d <D <+H <u o G '55 ^ o 88 3.2. DOSE-RESPONSE EFFECTS OF AROCLOR 1260 T R E A T M E N T 3.2.1. EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON B O D Y WEIGHT G A I N Aroclor 1260 decreased body weight gain in the 156 mg/kg/day dosage group only (Table 23). One animal in this dosage group died on day twelve. The mean body weight gain of the group was decreased by 65% compared to corn oil-treated animals. 3.2.2. EFFECTS OF A R O C L O R 1260 TREATMENT ON O R G A N WEIGHT The effects of Aroclor 1260 treatment on organ weight are summarized in Tables 23, and 24. Absolute liver weight was increased in a dose-dependent manner by Aroclor 1260 at dosages of 6.25 mg/kg/day and greater. Compared to the corn oil-treated group, there was a small but significant 1.2-fold increase in relative liver weight following treatment with Aroclor 1260 at a dosage of 3.13 mg/kg/day. Aroclor 1260 treatment had no effect on absolute thymus, testes, seminal vesicle, or ventral prostate weights (Table 24). 89 03 •~ CD 13 3 13 03 43 .SP '53 > 03 00 .2? 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EFFECTS OF A R O C L O R 1260 TREATMENT ON ANTIBODY RESPONSE Aroclor 1260 treatment suppressed the primary antibody response to SRBC (Table 25, and Figure 14). Serum anti-SRBC IgM concentrations were assessed as the serum IgM titer or dilution factor that yields 1.0 OD. The suppression of antibody response was a sensitive endpoint of Aroclor 1260 exposure as shown by a dose-dependent decrease in anti-SRBC IgM titer at dosages greater than 1.25 mg/kg/day. A 72%-80% maximal decrease in antibody production was achieved at dosages of 31.3 mg/kg/day and greater. 5? ~So OO o N O 1-1 o o <+-< o CD bO a CO o Q NO in oo CO cn N O in i n <N NO cn cn m C M in C N N O cn o i n (N o o O * cn NO ^ 0 0 -fl CN £ -H o C N -H 3 5 C N -H oo ^ C N -H o 1 -H in oo m -H ON ^ 5 oo ^ -H CD ^ u Q a o o J> b o J D bO ;fl cd ^ 3 s ^  ^ A 5? S ^ i d bO 'S ^ bO "bp bo N O i n 1 — 1 C O CD > . o « X >> CD rfl CO 2^ ™ <^ * B co O <4H CD O b o cd > cd bO CO —I fl 2? O H ^ bO ° C 2 © - f l N O -fl £^ CD 1 — 1 03 >-i W o CD co 3 w C O r ; + 1 - A G cd CD a CO 03 T3 CU CO CO CD s-O H X CD CD o3 CO CD fl CD in o © V & & o CD £ bo i2 T3 CD td CD cd +7* a rfl o fl cd T3 > < 93 700 600 500 400 S o © ffl 5 I «3 fc 300 200 100 o © o o o o o Dosage of Aroclor 1260 (mg/kg/day) Figure 14. Effects of treatment with Aroclor 1260 on humoral immunity response in adult male rats. Adult male rats were sensitized with SRBC after four days of treatment with Aroclor and blood for determination of anti-SRBC IgM was collected six days later. Each treatment group consisted of six rats, except for the 156 mg/kg/day dosage (n=5) groups. Blood samples were collected by tail nicking and analyzed for anti-SRBC IgM by ELISA (refer to figure 7). The mean anti-SRBC IgM titer of the corn oil-treated group was 483 ± 80. Data is expressed as mean anti-SRBC IgM titer to yield 1.0 OD ± SEM. 94 3.2.4. EFFECTS OF A R O C L O R 1260 TREATMENT ON S E R U M TESTOSTERONE L E V E L S Aroclor 1260 treatment had no appreciable effect on serum testosterone concentrations as shown in Table 26, and Figure 15. There was a relatively large degree of variability in mean serum testosterone concentrations among individual rats (2.0-13.5 ng/ml). Our results suggested that serum testosterone concentration was insensitive to Aroclor 1260 treatment. Serum L H concentrations were also unaffected by Aroclor 1260 treatment as indicated by the results obtained from Dr. Desaulniers' laboratory (see data in Appendix II). 3.2.5. EFFECTS OF A R O C L O R 1260 TREATMENT ON S E R U M T H Y R O X I N E (T4) L E V E L S There was a decline in serum T4 levels following Aroclor 1260 exposure. Serum T4 levels decreased to 60% of control values at a dosage of 6.25 mg/kg/day (Table 27 and Figure 16). Aroclor 1260 suppressed serum T4 levels to approximately 25% of control values at a dosage of 31.3 mg/kg/day. Although serum TSH concentrations were only affected at a dosage of 15.6 mg/kg/day, a moderate increase in serum TSH levels at higher dosages was apparent (Appendix II). The suppression of serum T4 levels was a sensitive endocrine endpoint of Aroclor 1260 exposure. 13 0 0 o NO CN o CD 0 0 03 GO NO IT) OO co co N O in in CN NO co co in CN in CN NO CO © in CN o o U oo OO CN -H O N o « J o in -H © ^ CN ^ NO -H o © co -H in in in f- o ^ -H CO CO oo NO © -H r--NO in O N O CO -H in <—i NO +1 NO o in +! 0 0 CN K -H N O O N ^ m' -H a CD O O g C O £ S3 r=! CD O a o o 0 0 o S H 0 0 CD 0 0 03 CO O 13 o? ~ob 44 "ob NO m 1 00 CD >^ ^ 43 CD rd « ,2 -ti 13 S H CD .a s co O < 4 H CD O 0 0 13 5 » 03 23 M 03 s-< S 0 0 CO +1 3 CD o NO CN o 43 T 3 CD S3 "03 GO |D 03 j3 T 3 CD CD s-i 00 CD co > <D On + J CD 1 - 1 CD i n a C O »-l CD - M * 3 03 13 > < -3 13 03 S3 O CD O S3 o o CD CO S3 O O NO CN CD O S H < 43 O o to w i> CN CU 3 « 0? T3 ~oo 44 "ob a, o NO CN o S H < < + H o CD 0 0 03 CO O Q NO in 0 0 t--co CO NO in in CN NO CO CO in CN in CN NO CO © in CN © S3 S H o o * CO ^ © • ^ —' +1 * in ° . ^ < ^ — -H CO ^ in 1—1 - H -H * co ^ •< M — -H * —1 -H © <=> •n ^ co -H NO O N 3 -H co ^ m -H 0 0 in © ^ m -H -H CD ^ X\ l B S3 O O CD 0 0 R C D CD t+H o ? ^ s ^ o3 o3 O 13 i n 0 0 "3 ^ ob ~5b'co NO w in — 1 CO CD >-> •=1' -*-» T3 " O H x >? CD .TH + J T3 S H CD co O T H - H CD O 0 0 13 ^ -4— ' 03 .eg 0 0 S3 g 0 o o 0 S^43 0 0 0 ^ S3 o3 2 <=» +3 NO CD ^ 03 S H W O . o s J w <; CO M +1 - | ^ S3 > 1 0 a -n P CD a CD © 2 3 03 G 13 CD CD S H 91 U O S ^ 5b U ^ 03 <D 2 H 03 -if in « CD +-> O •2 3 e 03 13 > < 8 96 12 10 4 H 2 H T 3 "S i-l fl-fl o o i n CN o o i n CN NO i n m CN —< i n CN NO i n Dosage of Aroclor 1260 (mg/kg/day) NO m Figure 15. Effects of Aroclor 1260 treatment on serum testosterone concentration in adult male rats. Data is expressed as mean serum testosterone concentration ±SEM. Each group consisted of six rats, except for the 156 mg/kg/day dosage (n=5) group. 97 70 i 60 H 50 H I fl o c3 — J 40 fl co g oo c 30 i 20 10 o o o o o o o Dosage of Aroclor 1260 (mg/kg/day) Figure 16. Effects of treatment with Aroclor 1260 on serum T4 concentration in adult male rats. The mean serum T4 concentration of the corn oil-treated group was 45.9 ±4 .1 ng/ml. Data is expressed as mean ± SEM. Each treatment group consisted of six rats, except for the highest dosage (n=5) groups. 98 3.2.6. EFFECTS OF A R O C L O R 1260 TREATMENT ON TOTAL C Y P CONTENT Total hepatic C Y P content was calculated by dividing the total C Y P concentration by the protein concentration. As shown in Table 28, total hepatic C Y P content increased with increasing dosages of Aroclor 1260. Total hepatic C Y P content increased at a dosage of 1.25 mg/kg/day or greater in a dose-dependent manner. There was approximately a 1.6-fold increase in total hepatic C Y P content at a dosage of 1.25 mg/kg/day and 3-fold increase at a dosage of 156 mg/kg/day. 99 Table 28. Effects of treatment with Aroclor 1260 on protein concentration and total C Y P content in hepatic microsomes from adult male rats Dosage of Aroclor Total C Y P Protein 1260 concentration concentration (mg/kg/day) (nmol/ml) (mg/ml) Total C Y P content Relative liver (nmol/mg protein) Weight (% of final body weight) Corn oil 0.025 0.13 0.625 1.25 3.13 6.25 15.6 31.3 78.1 156 25.5 ±2 .2 29.2 ±3 .2 30.3 ±3 .2 36.2 ±4 .0 45.7 ±6 .2 70.1 ±5 .2 85.9 ±7.1 84.7 ±8 .0 82.1 ±8 .8 97.0 ±4 .9 106 ± 8 36.7 ±3 .0 35.8 ±2 .0 39.6 ±2 .6 37.3 ±2 .4 38.1 ±3.2 43.9 ±3 .4 46.5 ±2 .7 52.7 ±5 .3 43.5 ±2.1 52.5 ±3.9 51.5 ± 4.1 0.70 ± 0.02 0.81 ±0.06 0.74 ± 0.07 0.90 ± 0.06 1.13 ±0.08* 1.53 ±0.06* 1.85 ±0.09* 1.66 ±0.10* 1.93 ±0.14* 1.76 ± 0.11* 2.07 ±0 .21* 4.3 ±0.1 4.7 ±0.1 4.5 + 0.1 4.6 ±0.1 4.7 ±0.1 5.1 ± 0 . 1 * 5.7 + 0.1* 6.1 ±0 .2* 6.4 ±0 .2* 7.6 ±0 .2* 8.8 ±0 .5* Values are expressed as mean ± SEM. Each treatment group consisted of six rats except for the 156 mg/kg/day dosage group (n=5). Adult male rats were treated with Aroclor 1260 in corn oil by oral gavage once daily for ten days. (*) Significantly different from the corn oil-treated group (p<0.05). 100 3.2.7. EFFECTS OF A R O C L O R 1260 TREATMENT ON SPECIFIC C Y P ISOZYMES 3.2.7.1. EFFECTS OF A R O C L O R 1260 ON CYP PROTEIN L E V E L S 3.2.7.1.1. EFFECTS OF A R O C L O R 1260 ON CYP2B PROTEIN L E V E L S Figure 17 shows a representative immunoblot of hepatic microsomes from Aroclor 1260-treated rats probed with polyclonal anti-CYP2Bl IgG. The CYP2B1 IgG reacts with rat CYP2B1 and CYP2B2 equally. Various concentrations of CYP2B1 standards were included in the analysis to generate a calibration curve, which was used to measure hepatic CYP2B1 and CYP2B2 content of the microsomal samples. In lanes containing microsomal samples, the upper and lower bands correspond to CYP2B2 and CYP2B1, respectively. In some lanes, two faint bands appear below the CYP2B1 band and may correspond to proteins with shared antigenic sites as CYP2B1 or CYP2B2 (example: CYP2B3). The CYP2B1 and CYP2B2 bands are faint for the corn-oil treated rat, and are intensified at dosages of Aroclor 1260 greater than 0.13 mg/kg/day. Immunoquantitation of CYP2B protein bands revealed that Aroclor 1260 caused a large and dose-dependent induction of hepatic CYP2B protein levels. Hepatic CYP2B content was expressed in pmol/mg protein (Table 29, Figures 18 and 19) or in pmol/nmol total C Y P (Appendix III). At dosages of 0.625 mg/kg/day and 31.1 mg/kg/day, hepatic CYP2B1 enzyme was induced by 13-fold and 230-fold, respectively. There was a 17-fold induction of hepatic CYP2B2 at a dosage of 0.625 mg/kg/day and up to 200-fold induction at a dosage of 31.3 mg/kg/day. CYP2B induction was highly sensitive to Aroclor 1260 exposure. Maximal CYP2B induction was attained at a dosage of 15.6 mg/kg/day. 101 Lane 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 Figure 17. Representative immunoblot of hepatic microsomes from adult male rats (probed for CYP2B) The primary antibody used was polyclonal rabbit anti-rat CYP2B1 IgG at 2.0 ug/ml. The secondary antibody used was alkaline phosphatase conjugated, goat F(ab')2 anti-rabbit IgG at 1:3,000 dilution. Lanes 1 and 2 contain hepatic microsomes from rats treated with Aroclor 1260 at the dosages of 156 and 78.1 mg/kg/day, respectively. Lane 3 contains hepatic microsomes from a rat treated with corn oil. Lanes 4 through 13 contain hepatic microsomes from rats treated with Aroclor 1260 at the dosages of, 0.025, 0.13, 0.625, 1.25, 3.13, 6.25, 15.6, 31.3, 78.1, and 156 mg/kg/day, respectively. Lanes 14 through 18 contain 0.0313, 0.0625, 0.125, 0.25 (internal standard), and 0.375 pmol of CYP2B1 standards, respectively. Hepatic microsomes were loaded at total C Y P concentrations of 0.05 nmol/ml for lanes 1 and 2, 1 nmol/ml for lanes 3 through 5, 0.25 nmol/ml for lane 6, 0.1 nmol/ml for lane 7, 0.05 nmol/ml for lanes 8 and 9, and 0.025 nmol/ml for lanes 10 through 13. 102 Table 29. Effects of treatment with Aroclor 1260 on specific hepatic CYP2B and CYP1A content Dosage of Aroclor 1260 (mg/kg/day) CYP2B1 (pmol/mg protein) CYP2B2 (pmol/mg protein) CYP1A2 (pmol/mg protein) CYP1A1 (pmol/mg protein) Corn oil 6 ± 2 4 ± 1 2.0 ±0 .4 n.d. 0.025 11 ± 6 10± 1 3.0 ±0 .7 n.d. 0.13 18 ± 9 8 ± 2 2.8 ±0 .7 n.d. 0.625 75 ± 11* 67 ± 10* 3.4 ±0 .5 n.d. 1.25 189 ± 4 8 * 119 ± 2 7 * 3.0 ±0 .7 n.d. 3.13 637±102* 360 ± 2 2 * 3.8 ±0 .7 11.3 ± 3.3* 6.25 702 ± 78* 576 ±97* 13.5 ±3.0* 21.8 ± 1.9* 15.6 • 952±153* 602 ± 79* 12.4 ± 1.6* 30.8 ±3 .4* 31.3 1390±163* 808 ± 151* 23.4 ±4.4* 60.3 ±5 .6* 78.1 1330±199* 796±115* 31.9 ± 7.5* 64.4 ±8 .9* 156 1430 ±297* 929± 136* 35.7 ±3 .7* 93.3 ± 15.3* Values are expressed as mean ± SEM. Each treatment group consisted of six rats except for the 156 mg/kg/day dosage group (n=5). Adult male rats were treated with Aroclor 1260 in corn oil by oral gavage once daily for ten days. (*) Significantly different from the corn oil-treated group (p<0.05). n.d.: not determined (<LOQ=0.0625 pmol/lane) 103 2000 i 1800 -1600 -0.01 0.1 1 10 100 1000 Dosage(mg/kg/ day) Figure 18. Effects of Aroclorl260 treatment on hepatic CYP2B1 protein levels in adult male rat. Mean CYP2B1 protein levels ± S E M in liver microsomes from adult male rats after treatment with Aroclor1260. The hepatic CYP2B1 content of the corn oil-treated group was 6 ± 2 pmol/mg protein. Each group consisted of six rats, except for the highest dosage (n=5) groups. 104 1200 i 1000 H 0.01 0.1 1 10 100 1000 Dosage(mg/kg/day) Figure 19. Effects of Aroclor1260 treatment on hepatic CYP2B2 protein levels in adult male rats. Mean CYP2B2 protein levels ± S E M in liver microsomes from adult male rats after treatment with Aroclor1260. The hepatic CYP2B2 content of the corn oil-treated group was 4 ± 1 pmol/mg/protein. Each group consisted of 6 rats, except for the highest dosage (n=5) groups. 105 3.2.7.1.2. EFFECTS OF A R O C L O R 1260 ON CYP1A PROTEIN L E V E L S Figure 20 is a representative immunoblot of hepatic microsomes from Aroclor 1260-treated rats probed with monoclonal an t i -CYPlAl IgG. Various concentrations of CYP1A1 standards were included on the gel to generate a calibration curve, which was used to measure hepatic CYP1A1 content of the microsomal samples. A separate calibration curve was obtained with CYP1A2 standard to calculate hepatic CYP1A2 content (immunoblot not shown). The upper band represents CYP1A1 and the lower band corresponds to CYP1A2. The CYP1A1 band is less intense than the CYP1A2 band in lanes containing hepatic microsomes from rats treated with Aroclor 1260 at dosages less than 3.13 mg/kg/day. The staining intensity of the 1A1 band increased with increasing dosages of Aroclor 1260. Hepatic CYP1A was expressed in pmol/mg protein (Table 29, Figures 21 and 22) or in pmol/nmol total C Y P (Appendix III). CYP1A1 protein levels, which could not be measured in control samples, increased linearly with increasing dosages of Aroclor 1260 (Table 29, Figures 21 and 22). There was a 5.7-fold increase in hepatic CYP1A1 content following treatment at a dosage of 3.13 mg/kg/day and a 47-fold increase following treatment at a dosage of 156 mg/kg/day. The increase in hepatic CYP1A1 levels did not level off. Thus, CYP1A1 protein levels did not reach a maximum. Hepatic CYP1A2 content remained unchanged up to a dosage of 6.25 mg/kg/day (6.8-fold increase) and only increased moderately to a plateau at a dosage of 31.3 mg/kg/day (12-fold increase). 106 Lane 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 Figure 20. Representative immunoblot of hepatic microsomes from adult male rats (probed for CYP1 A) The primary antibody used was monoclonal mouse anti-rat CYP1A1 IgG at 1.0 ug/ml. The secondary antibody used was alkaline phosphatase conjugated, goat F(ab')2 anti-mouse IgG at 1:3,000 dilution. Lanes 1 and 2 contain hepatic microsomes from untreated rats. Lane 3 contains hepatic microsomes from a corn oil treated rat. Lanes 4 through 13 contain hepatic microsomes from rats treated with Aroclor 1260 at the dosages of, 0.025, 0.13, 0.625, 1.25, 3.13, 6.25, 15.6, 31.3, 78.1 and 156 mg/kg/day, respectively. Lanes 14 through 18 contain 0.12, 0.250, 0.375, 0.500 and 0.375 pmol of CYP1A1 standards, respectively. Hepatic microsomes were loaded at total C Y P concentrations of 1 nmol/ml for lanes 1 through 9, and 0.5 nmol/ml for lanes 10 through 13. 107 120 i 100 H 0 H 1 1 1 o o o <-> O O —i o Dosage(mg/kg/day) Figure 21. Effects of treatment with Aroclor 1260 on CYP1 A l protein levels in adult male rats. Mean CYP1 A l protein levels ± S E M in liver microsomes from adult male rats after treatment with Aroclor1260. Each group consisted of six rats, except for the highest dosage (n=5) groups. 108 45 i Figure 22. Effects of treatment with Aroclor 1260 on CYP1A2 protein levels in adult male rats. Mean hepatic CYP1A2 protein levels ± S E M in liver microsomes from adult male rats after treatment with Aroclorl260. The hepatic CYP1A2 content of the corn oil-treated group was 2.0 ± 0.4 pmol/mg protein. Each group consisted of six rats, except for the highest dosage (n=5) groups. 109 3.2.7.2. EFFECTS OF A R O C L O R 1260 ON CYP-DEPENDENT E N Z Y M E ACTIVITIES The representative chromatograms of testosterone hydroxylation by hepatic microsomes from corn oil-treated and PCB-treated animals are shown in Figures 23 and 24, respectively. Enzyme activities are expressed in nmol metabolite/min/mg protein (Table 30) or in nmol metabolite/min/nmol total CYP (Appendix IV). As shown in Table 30, Aroclor 1260 exposure caused a dose-dependent elevation in CYP2B-catalyzed enzyme activities, namely the rates of formation of 16a-, 16p-hydroxytestosterone, and androstenedione. There were 2- and 51-fold induction of testosterone 16a-and 16P-hydroxylase activities, respectively, and a 3-fold increase in androstenedione formation following treatment with Aroclor 1260 at a dosage of 6.25 mg/kg/day. The testosterone 16p-hydroxylase activity increased at dosages greater than 0.625 mg/kg/day and leveled off at a dosage of 6.25 mg/kg/day. The increase in testosterone 16a-hydroxylase activity and androstenedione formation was less pronounced and was observed at dosages of 1.25 and 0.625 mg/kg/day, respectively. CYP3A enzymes catalyze testosterone 6p-and 2p-hydroxylase activities whereas CYP2A1 catalyzes testosterone 7a-hydroxylase activity (Waterman et al, 1991). Aroclor 1260 exposure induced testosterone 6P-, 2p- and 7a-hydroxylase activities at dosages of 1.25, 0.625 and 6.25 mg/kg/day, respectively. Maximal induction of these activities was attained at a dosage of 6.25mg/kg/day for testosterone 6p- and 2P-hydroxylase, and 31.3 mg/kg/day for testosterone 7a-hydroxylase. There was a decrease in testosterone 2a-hydroxylase activity (up to 93% decrease), which indicates a suppression of CYP2C11 activity because CYP2C11 catalyzes the formation of 2a-hydroxy-testosterone (Waterman et al, 1991). 110 Unexpectedly, a moderate decline in enzyme activities (example: testosterone 16p-hydroxylase activity) was noted at increasing dosages of Aroclor 1260, instead of a plateau. The testosterone hydroxylase assay is linear with respect to total C Y P concentration (see 'validation of the testosterone hydroxylase assay'). It is possible that the decrease in enzyme activity is caused by Aroclor 1260 treatment I l l Figure 23. Representative chromatogram of testosterone hydroxylation reactions catalyzed by hepatic C Y P enzymes in a corn oil-treated adult male rat (sample K l ) . A 20 ul volume of 12.5 m M testosterone was added to a reaction mixture containing 0.3 nmol of hepatic microsomes from an animal treated with corn oil at a dosage of 2.5 ml/kg/day for ten days. The metabolites formed over a period of five minutes were spiked 2.5 umol of the internal standard, 11 P-hydoxytestosterone, and were extracted by dichloromethane and resolved by HPLC. 112 OJ C O Figure 24. Representative chromatogram of testosterone hydroxylation reactions catalyzed by hepatic C Y P enzymes in an adult male rat after treatment with Aroclor 1260 sample A4). A 20 ul volume of 12.5 m M testosterone was added to a reaction mixture containing 0.3 nmol of hepatic microsomes from an animal treated with Aroclor 1260 at a dosage of 156.25 mg/kg/day for ten days. The metabolites formed over a period of five minutes were spiked 2.5 umol of the internal standard, 1 ip-hydoxytestosterone, and were extracted by dichloromethane and resolved by HPLC. CD G O CD a CD +J C O O C O . 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PI <D PI cd CD a CO <D <« co C M s ° } H T 3 O H CD S .22 CD kH cd CO cd 13 C CD CD , 0 ' i d 13 CD CD PI O CD 00 cd > cd 00 co CD J 3 "cd > a ^ O s-i CD O O H PS 43 O _ H S-H - - H 00 O 114 Figure 25. Effects of treatment with Aroclor 1260 on hepatic CYP-catalyzed testosterone hydroxylase activities in adult male rats. Graphs A , B, C, D, E, F, and G represent 160-, 60-, 7a-, 16a-, 20-, 2a-hydroxylase activities, and androstenedione formation respectively. The mean enzyme activity values of the corn oil-treated group were 1650 ±250 , 236 ± 2 5 , 2770 ±411 ,66 ± 8, 2590 ±366, 143 ± 2 0 , and 1740 ± 180'pmol metabolite/min/mg protein for 60-, 7a-, 16a-, 160-, 2a-, 20-hydroxylase activities, and androstenedione formation, respectively. (A) Dosage (mg/kg/day) 1400 n < S U _ 111 g o P H 6 0 1200 A IOOO H 800 H 600 400 200 J i o o o o o Dosage (mg/kg/day) o o o (D) -4—» 6000 n 5000 H 4000 -0 <u •s s 1 g 3000 „ a < a a "s w 3 a 2000 iooo H o o o o o o o Dosage (mg/kg/day) (E) Dosage (mg/kg/day) (G) 6000 i 1000 H o -i • 1 1 1 1 , - i ^ o o o Dosage (mg/kg/day) 119 4. DISCUSSION It has been known for several years that PCBs have the potential to cause immune and endocrine alterations (Silkworth et al, 1986; Harper et al, 1995; Desaulniers et al, 1999; Byrne et al, 1985). However, few studies examined a range of dosages at which PCB-induced biological effects are observed. In order to investigate PCB exposure levels that result in significant biological effects, we have undertaken a dose-response study for the immune and endocrine effects of PCBs. PCB exposure levels vary with several factors such as animal species and geographical location. Earlier studies by Ngui (1997) indicated that hepatic PCB congener levels in polar bears were comparable to levels in rats treated with Aroclor 1260 at cumulative dosages of 3.5 mg/kg-70 mg/kg. This dose range compares to dosages between 0.025 mg/kg/day and 6.25 mg/kg/day (cumulative doses of 0.25 and 62.5 mg/kg, respectively) in our study. 4.1. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON B O D Y WEIGHT G A I N Body weight gain was unchanged at all dosages of Aroclor 1260, except for the 65% decrease observed at a dosage of 156 mg/kg/day. The toxicity observed with the 156 mg/kg/day dosage group was due to the fact that the dosage of Aroclor 1260 administered was within the LD50 value range reported for Aroclor 1260 (LD50 = 1.26-2.0 g/kg). Compared to Aroclor 1242 and Aroclor 1248, which both have L D 5 0 values between 0.79 g/kg and 1.27 g/kg, Aroclor 1260 is relatively less toxic. 120 Several xenobiotics, including PCBs, PCDFs and PCDDs, produce A h receptor-mediated toxic responses including a waste syndrome that is characterized by a progressive loss in body weight. The cause of this weight loss is not known. The most toxic compounds are relatively planar and share significant structural similarity with TCDD. Aroclor 1260 contains few dioxin-like PCBs (eg. PCB#126) and consists mainly of non-planar di-ortho-PCBs, which do not elicit significant dioxin-like effects (Safe, 1994). In agreement with present study, Mayes et al. (1998) reported that body weight was unchanged in Sprague Dawley treated with Aroclor 1260 (TEQ contribution of 7.2 ppm) at a dietary concentration of 100 ppm for 24 months. In contrast, body weight was decreased in rats treated with Aroclor 1254 (TEQ contribution of 47.6 ppm) at 50 and 100 ppm using a similar dietary regimen (Mayes et al, 1998) 4.2. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON O R G A N WEIGHTS Aroclor 1260 treatment caused an increase in liver weight, as shown in Table 23. There was a dose-dependent increase in absolute liver weight at larger dosages while relative liver weight increased at dosages of 3.13 mg/kg/day and greater. The dose-response data for relative liver weight increase was fitted to a sigmoidal curve with equation (1). The ED50 and E m a x values for the relative liver weight increase were found to be 6.36 ± 3.46 mg/kg/day and 8.63 ± 0.39 %, respectively. 121 E = Em a x/(1 + exp(-(D- E D 5 0 )/A)) (1) Where: E m a x is the maximal response, ED50 is the dose at half-maximal response, and A is a constant. Aroclor 1260-induced liver enlargement resulted from a number of morphological and biochemical changes. Chu et al. (1996) observed an increase in relative liver weight in seven-week old female and male rat, following a thirteen-week dietary exposure to the non-planar PCB, 2,4,5,2',4',5'-hexachloribiphenyl (PCB#153), at a dosage of 50 ppm. There was a 1.4-fold increase in relative liver weight in males and a 1.2-fold increase in females. Gross pathology analysis indicated that some animals treated with PCB#153 had fatty and enlarged livers. Electron microscopy examinations revealed that there was proliferation of the endoplasmic reticulum and increased hepatocyte size in PCB-treated animals, which both contributed to liver enlargement. Other adverse effects of PCB#153 include a reduction in lung and liver vitamin A , changes in biogenic amine concentrations, increased hepatic uroporphyrin and urinary ascorbic acid levels (Chu et al, 1996). Likewise, the toxicity of the planar PCB congener, 2,3,4,3',4'-pentachlorobiphenyl (PCB#105), was investigated in male and female rats following dietary exposure at levels between 0 and 50 ppm for 13 weeks (Chu et al, 1998). Biochemical and morphological effects similar to those observed with PCB#153 treatment (Chu et al, 1996) were detected. However, PCB#105 resulted in a greater increase in liver weight in females (1.4-fold increase) compared to males (1.2-fold increase), at a dosage of 50 ppm (Chu et al, 1998). 122 Studies involving commercial and environmental PCB mixtures indicated that relative liver weight was also increased at similar dosages. Relative liver weight was increased by 1.5-fold in female offspring (at day 21 after birth) of female Long-Evans rats exposed to Aroclor 1254 in the diet at a dosage of 40 ppm for 50 days before mating until birth (Hany et al, 1999). A seven-day exposure to Aroclor 1260 by oral gavage resulted in a 1.2-fold increase in relative liver weight at a dosage of lOmg/kg/day (Ngui and Bandiera, 1999). Relative liver weight was maximally increased by 2.2-fold in female, but not male, rats exposed to Aroclor 1254 or Aroclor 1260 in the diet for 24 months at dosages between 25 and 100 ppm. Aroclor 1016 and Aroclor 1242 were less potent and increased relative liver weight (up to 1.4-fold increase) at dosages between 100 and 200 ppm (Mayes et al, 1998). Hansen et al.(1995) reported a 1.2-fold increase in relative liver weight at a dosage of 32 mg/kg in female pre-pubertal rats acutely exposed to a landfill soil extract with a TEF value similar to that of Aroclor 1248. Aroclor 1260 treatment did not affect thymus weight in our study. Thymus atrophy may not be a reliable indicator of immunotoxicity. Studies by L i et al. (1998) demonstrated that pre-pubertal female rats that were exposed to purified PCB#110 for two days exhibited an increase in liver weight at a dosage of 8 mg/kg/day, but no change in thymus weight. On the other hand, Silkworm et al. (1984) reported a suppression of humoral immune response with severe thymic atrophy in rats exposed to 3,4,3',4'-tetrachlorobiphenyl (PCB#77). Similar effects were observed with 2,3,4,5,3',4'-hexachlorobiphenyl (PCB#156), but with no decrease in thymus weight. These studies suggested that the two PCB congeners suppressed humoral immune response through a mechanism that was independent of the thymus. The thymus atrophy seen with PCB#77 123 could result from cytotoxicity (Silkworm et al, 1984). The planar PCB congener, 2,3,4,3',4'-pentachlorobiphenyl (PCB#105) caused a 30% decrease in thymus weight in female and male seven-week old rats following dietary exposure for thirteen weeks, at a dosage of 50 ppm (Chu et al, 1998). The decrease in thymus weight was explained by a reduction in cortical and medullary size and was considered a poor indicator of immunosuppression. Ross et al. (1997) reported a decrease in thymus weight in rats exposed prenatally to Atlantic Sea herring oil spiked with TCDD or Baltic Sea herring oil environmentally contaminated with PCBs, PCDFs and PCDDs. Pregnant rats were administered uncontaminated Atlantic Sea herring oil or Baltic Sea herring oil or Atlantic Sea herring oil spiked with TCDD, by oral gavage daily for forty one days starting from day 6 of gestation. The cumulative doses of TCDD toxic equivalents (TEQs) on day 11 were 2.03, 19.9 and 1440 ng TEQ/kg body weight for the Atlantic Sea herring oil, the Baltic Sea herring oil and the TCDD-spiked oil, respectively. Only the TCDD-spiked oil caused thymus atrophy (a 40%> decrease in thymus weight on day 25). Thymus cell number was reduced in pups from TCDD-treated mothers at all time points, and only on day 59 in pups from mothers treated with Sea Baltic herring oil (20% decrease in thymus cell number compared to Atlantic Sea). The results of Ross et al. (1997) suggested an indirect relationship between thymus weight and thymus cellularity. Our study did not detect any sign of reproductive toxicity following Aroclor 1260 exposure, as testis, ventral prostate and seminal vesicle weights were unchanged. Acute i.p. exposure of pre-pubertal female rats to environmental PCB mixtures extracted from the soil caused a 1.3-fold increase in uterine weight at a dosage of 346 mg/kg. The increase in uterine weight was attributed to the estrogenic effects of PCBs (Li and 124 Hansen, 1996). Aroclor 1254 treatment did not result in inhibitory reproductive effects, but caused a significant increase in testis weight (Cooke et al, 1996). Adult testis weight was increased by 23% in male rats that were exposed to Aroclor 1254 daily from birth to day 25 at a dosage of 1.6 mg/day. This effect, which was mediated by hypothyroidism (decreased serum T4 levels) and subsequent Sertoli cell proliferation, was effectively reversed by T4 administration. Daily sperm production also increased by 40% with Aroclor 1254 treatment at a dosage of 1.6 mg/kg. Aroclor 1254 was more effective at inducing hypothyroidism and increasing testis weight compared to Aroclor 1242 (Cooke et al, 1996). Conversely, Hany et al. (1999) reported a decrease in testis weight in the offspring of female Long-Evans rats exposed to Aroclor 1254 (treated 50 days before mating until birth), at a dosage of 40 ppm. There was a 40%> decrease in adult testis weight, which was associated with a decrease in serum testosterone levels. Reduced testes weight was a result of PCB-induced testosterone depletion during gestation (Hany et al, 1999). According to Cooke et al. (1996), PCB treatment prolongs Sertoli cell mitogenesis during a narrow developmental window. The different results of Cooke et al. (1996) study and Hany et al. (1999) can be explained by differences in study design. Neonates received Aroclor 1254 directly by subcutaneous injections in the Cooke et al. (1996) study, whereas they were exposed pre- and postnatally in Hany et al. (1999) study. Furthermore, the effects of PCB treatment on testis weight may be the combination of stimulatory hypothyroidism-mediated and dioxin-like estrogenic effects of PCBs. In our study, no changes in testis weight were observed probably because adult reproductive organs, which have undergone sexual maturation, may be insensitive to PCB exposure. 125 Our results complement previous reports and further indicate that exposure time is relevant in evaluating the effects of PCBs on reproductive organs. The effects of PCBs on organ morphology in young animals have received great attention from researchers. Humans and wildlife species may be exposed to PCBs during adulthood, rather than early developmental stages. Examples are individuals who are exposed to PCBs in the workplace. Although the effects of PCBs on organ weights were investigated in previous studies, our data indicated that of all the organs examined, liver was the most sensitive organ to short-term Aroclor 1260 exposure in adult rats (ED50 = 6.36 ± 3.46 mg/kg/day). Relative liver weight was increased significantly at relatively low dosages (>1.25 mg/kg/day) that are comparable to environmental PCB levels. 4.3. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON A N T I B O D Y RESPONSE There was a dose-dependent suppression of the antibody response to SRBC with Aroclor 1260 treatment at dosages greater than 1.25 mg/kg/day. Anti-SRBC IgM titers were decreased by 80% from control values at a dosage of 31.3 mg/kg/day. These results were somewhat similar to those reported by Harper et al. (1995) for B6C3F1 mice treated with single doses of Aroclor 1260. Antibody response decreased by 50% from control values at a dosage of 100 mg/kg and leveled off at a dosage of 500 mg/kg. Tryphonas et al. (1991) reported a suppression of the primary antibody response to SRBC in rhesus monkeys exposed to Aroclor 1254 at dosages of 5, 20, 40 and 80 ug/kg/day for twenty-three months. In that study, serum IgM levels were decreased by 80% at the highest dosage. The trend in our study was comparable to that in other studies, although it is 126 important to point out that the lowest dosages at which the effects were observed varied with length of exposure and the animal species. In our study, antibody response was decreased following exposure at a cumulative dose of 12.5 mg/kg (subchronic exposure) whereas suppression of antibody response was observed at cumulative doses of 100 mg/kg (acute dose) and 3.45 mg/kg (chronic exposure) by Harper et al. (1995) and Tryphonas et al. (1991), respectively. The mechanism by which PCBs suppress antibody response is still unclear. A number of studies have shown that the antibody response to SRBC in mice depends on PCB planarity and the Ah receptor (see section 1.3.1.2). However, studies by Harper et al. (1995) reported that Aroclor 1260 was a more potent humoral immunosuppressant than Aroclor 1254, Aroclor 1248 and Aroclor 1242. Aroclor 1260, which contains fewer co-planar PCBs than the other Aroclor mixtures examined, would be expected to be the least potent immunnosuppressant if antibody response were primarily dependent on PCB planarity. This was not the case. Sargent et al. (1991) showed that although a planar PCB congener was more effective at reducing the total number of antibody-producing B-cells, the combination of planar and non-planar PCB congeners produced a synergistic rather than additive decrease in B-cell number (see section 1.3.1.2). In our study, antibody response was sensitive to Aroclor 1260 exposure. The ED50 and E m a x values, calculated by fitting the anti-SRBC IgM data to a sigmoidal curve (equation 1), were 1.52 ± 0.37 mg/kg/day and 512 ± 32, respectively. The low ED50 value for the anti-SRBC IgM titer dose-response curve indicates that humoral immune response is suppressed at relatively low dosages of Aroclor 1260. Our data indicate that Aroclor 1260 compromised immunocompetence despite the absence of thymic atrophy, 127 and more importantly, that antibody response was a highly sensitive and early indicator of immunotoxicity following Aroclor 1260 exposure. 4.4. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON GONADOTROPIN A N D SEX HORMONE L E V E L S Various studies have shown that PCBs can act as estrogens in rodents and decrease gonadotropin release from the pituitary. Desaulniers et al. (1999) demonstrated the planar PCB, PCB#126, caused a 60% decrease in serum L H concentrations, and a two-fold increase in pituitary L H levels in male rats following treatment at a dosage of 400 |ag/kg/day for two days. There was no change in serum testosterone levels. Because L H and FSH both stimulate testosterone and estradiol release, a decrease in serum L H levels is expected to result in decreased serum testosterone concentrations. In our study, Aroclor 1260 did not affect serum testosterone or L H levels (Appendix II). Serum testosterone concentration may be insensitive to PCB-treatment in adult animals. Ovulation was inhibited and serum progesterone, FSH, and L H concentrations were reduced by 80%>, 60%>, and 80%, respectively in gonadotropin-primed immature rats treated with PCB#126 and PCB#52 at dosages of 114-457 ug/kg, ip (Gao et al, 2000). Hany et al. (1999) reported a 60%> decrease in serum testosterone concentration in offspring of Aroclor 1254-treated rats on postnatal day 170, which further suggests that PCB exposure during early developmental stages is more effective at suppressing serum testosterone levels. Because serum testosterone and L H levels were insensitive to Aroclor 1260 exposure in adult male rats in our study, we suggest that Aroclor 1260 may cause reproductive effects in young animals, but not in adults. 128 4.5. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON S E R U M T H Y R O X I N E (T4) L E V E L S Serum T4 concentration was sensitive to treatment with Aroclor 1260. Serum T4 levels were decreased by 60% from control values following treatment at a dosage of 6.25 mg/kg/day. At a dosage of 31.3 mg/kg/day, serum T4 concentrations were decreased by 75% from control values (Table 27, Figure 16). There was no further decline in serum T4 at dosages greater than 31.3 mg/kg/day. As with previous endpoints, the serum T4 dose-response curve was fitted to a sigmoidal curve (equation 1). The ED50 and E m a x values for the decrease in serum T4 levels were calculated to be 4.69 ± 0.79 mg/kg/day and 49.3 ± 1.82 ng/ml, respectively, suggesting that serum T4 was a sensitive indicator of endocrine toxicity. By comparison, L i and Hansen (1996) reported a 38% and 85% decrease in serum T4 levels in pre-pubertal female rats following acute exposure to environmental PCB mixtures by ip injection at dosages of 32 mg/kg and 346 mg/kg, respectively. The PCB mixture used was extracted from soil samples taken from NPL sites. The dose-response curves (for serum T4) for reconstituted PCB mixtures form air and dust samples from N P L sites were similar in terms of shape and slope to that of the soil sample although the TEQ values of the samples were different. L i and Hansen's (1996) results indicate that the observed decrease in serum total T4 levels is not a primarily dioxin-like response because the TEQ model would predict that the NPL soil sample (with the highest TEQ value) to be the most potent. In our study, serum T4 levels were decreased by 35% and 75%> at dosages of 3.13 mg/kg/day and 31.3 mg/kg/day, respectively (cumulative doses of 31.3 mg/kg and 313 mg/kg, respectively). In other words, the slope of our T4 dose-129 response curve is similar to those of the dose-response curves for the air, dust, and soil samples mentioned above. Therefore, our results indicate that Aroclor 1260 can be used as model PCB mixture when assessing the effects of environmental PCB exposure on thyroid hormones. Furthermore, inclusion of higher doses of PCBs, namely cumulative doses of 781 mg/kg and 1560. mg/kg (the highest dose in the L i and Hansen's study was 346 mg/kg), was necessary for obtaining the full dose-response curve for T4 decrease. L i and Hansen observed a quasi-linear decrease in serum T4 levels at all doses examined, whereas our results suggest a plateau effect at dosages greater than 31.3 mg/kg/day. PCB-induced hypothyroidism can indirectly result in reproductive and neuro-behavioral effects. Cooke et al. (1996) observed a dose-dependent decrease in serum T4 levels in male rats exposed to Aroclor 1254 or Aroclor 1242 by daily subcutaneous injections, from birth to day 25. Serum T4 levels were decreased by 50% and 75% at dosages of 0.4 mg/day and 3.2 mg/day, respectively, following Aroclor 1254 treatment. Compared to Aroclor 1254, Aroclor 1242 was less potent. Serum T4 levels were still suppressed by Aroclor 1254 on day 45, but were normal in Aroclor 1242-treated animals on the same day. Serum testosterone levels were unchanged in PCB-treated animals (Cooke et al, 1996). The decrease in serum T4 levels in PCB-treated animals was thought to result in Sertoli cell proliferation, leading to an increase in testis weight. Aroclor 1242 and Aroclor 1254 caused a 13%) and 23%) increase in testis weight, respectively, at a dosage of 1.6 mg/day. As discussed with other endocrine effects, the time of PCB exposure was critical in the study by Cook et al. (1996). A n increase in testis weight is observed when PCB treatment begins at birth, but not if treatment begins on day 12 after birth. This explains the fact that we did not see an increase in testis 130 weight, because rats were exposed during adulthood in our study. Changes in thyroid hormone homeostasis can also lead to neurobehavioral effects. PCB-induced hypothyroidism caused neurotoxic and hearing deficits in young rats (Goldey et al, 1995; Olivero and Ganey, 2000). However, it is not known whether Aroclor 1260 caused any neuro-behavioral effects in our animals, because these effects were not examined. As discussed earlier, the mechanisms by which PCBs decrease serum T4 levels include induction of UDP-GT enzymes and interference with thyroxine transport (Hood and Klaassen, 2000; Barter and Klassen, 1994; Chauhan et al, 2000). The decrease in serum T4 concentration results in an increase in serum TSH levels because of feedback-inhibition of TSH release by T4. PCBs are less effective at increasing serum TSH levels however. In sub-chronic studies, rats receiving PCB in their diets showed only a moderate increase in serum TSH levels compared to other UDP-GT inducers such as PB (Hood and Klaassen, 2000; Barter and Klassen, 1994). A twenty one-day dietary exposure to Aroclor 1254 at 250 ppm resulted in a 40% increase in serum TSH levels and a 30%> increase in thyroid weight (Barter and Klassen, 1994). Similar studies by Hood and Klaassen (2000) reported no change in serum TSH levels in rats treated with Aroclor 1260 in the diet at 25, 50, 100 and 200 ppm daily for seven days. The increase in serum TSH levels may depend on the duration of exposure. In our study, rats were exposed for a slightly longer period of time (ten days) and a detectable, but statistically insignificant, increase in serum TSH levels was observed (Appendix II). This suggests that lengthening our exposure time may result in an increase in serum TSH levels and a possible increase in thyroid weight. The studies described above tested the effects of Aroclor 1254 on thyroid gland weight, but did not examine the effects of Aroclor 1260 131 extensively. We predict that exposure to Aroclor 1260 for a longer duration (longer than ten days) can cause significant increases in serum TSH and thyroid weight. 4.6. DOSE-RESPONSE EFFECTS OF A R O C L O R 1260 T R E A T M E N T ON HEPATIC C Y P E N Z Y M E L E V E L S Aroclor 1260 exposure caused an induction in hepatic C Y P enzyme levels. Total C Y P content was increased approximately two-fold at dosages greater than 1.25 mg/kg/day. Ngui and Bandiera (1999) reported similar findings. In their study, total C Y P was content increased by approximately two-fold at dosages greater than 5 mg/kg/day (for seven days). Parkinson et al. (1981) tested 13 PCB congeners for their ability to induce hepatic C Y P enzymes in one-month old male Wistar rats. There was a 2.6-fold increase in total hepatic CYP content following a two-day exposure to 2,4,5,2',4',5'-hexachlorobiphenyl (PCB#153), one of the most potent C Y P inducers of all the congeners tested, at a dosage of 150 umol/kg/day (Parkinson et al, 1981). A four-fold induction in total hepatic CYP content was observed in three-week old male Long Evans exposed to Aroclor 1254 by i.p. injection, at a dosage of 1500 umol/kg (Parkinson etal, 1983). Our results indicated an induction of hepatic CYP2B protein levels in rats treated with Aroclor 1260 (Figures 17-19). There was a 230- and 200-fold induction (expressed on a protein basis) or 91- and 77-fold induction (expressed per total C Y P concentration, see Appendix III) of hepatic CYP2B1 and CYP2B2 enzyme levels, respectively, at dosages greater than 0.625 mg/kg/day. The CYP2B1 and CYP2B2 curves were fitted to sigmoidal curves with equation (1). The ED50 values for hepatic CYP2B1 and CYP2B2 132 induction were determined to be 5.38 ± 1.28 mg/kg/day and 3.99 ± 0.72 mg/kg/day, respectively. The E m a x values for CYP2B1 and CYP2B2 induction were 1290 ± 96 and 778 ± 49 pmol/mg protein, respectively, which indicates that there was greater induction of hepatic CYP2B1 compared to CYP2B2. Ngui and Bandiera (1999) observed a lesser degree of hepatic CYP2B1 and CYP2B2 induction (55- and 16-fold, respectively) with hepatic C Y P content expressed in pmol/nmol total CYP, but similar patterns, that is CYP2B1>CYP2B2. Furthermore, similar to our findings, Ngui and Bandiera reported the induction of hepatic CYP2B enzymes at dosages greater than 0.5 mg/kg/day, with maximal induction reached at 5 mg/kg/day. In our study, however, the maximal induction of hepatic CYP2B protein levels was achieved at a dosage of 15.6 mg/kg/day. The increase in hepatic CYP2B protein levels was further supported by our testosterone hydroxylase data. There was a 51-fold increase in testosterone 16(3-hydroxylase activity. Androstenedione formation and testosterone 16oc-hydroxylase activity increased by 3.2- and 2-fold, respectively. Similar to the induction in hepatic CYP2B protein levels, the rates of formation of androstenedione, 16a-hydroxytestosterone and 16p-hydroxytestosterone, which are all catalyzed by CYP2B enzymes, increased at dosages greater than 0.625 mg/kg/day. The rate of formation of 7a-hydroxytestosterone, catalyzed by CYP2A1 enzymes, was increased by 5-fold at dosages greater than 6.25 mg/kg/day. The increases in the rate of formation androstenedione and 16a-hydrox-testosterone were more moderate than the increase in testosterone 16p-hydroxylase activity because of CYP2C11 suppression, as suggested by a large decrease (up to 93% decrease at a dosage of 156.25 mg/kg/day) in testosterone 2a-hydroxylase activity. The testosterone hydroxylation data indicates a moderate 2-fold 133 and 5-fold induction of testosterone 6P- and 2 p-hydroxylase activities, respectively, at dosages greater than 6.25 mg/kg/day, which is indicative of hepatic CYP3A induction. The ED50 and E D m a x values for testosterone hydroxylase activities measured in this study are summarized in Table 31. There was a significant increase in hepatic CYP1A protein levels, but to a lesser degree than CYP2B protein levels. Hepatic CYP1A1 protein levels increased linearly by 47-fold at a dosage of 156 mg/kg/day. There was also a moderate increase in hepatic CYP1A2 protein levels (up to 12-fold) at dosages greater than 6.25 mg/kg/day. Likewise, Ngui and Bandiera (1999) reported a dose-dependent increase in hepatic CYP1A1 protein levels at dosages greater than 5 mg/kg/day and 3-fold induction of hepatic CYP1A2 protein levels at a dosage of 50 mg/kg/day. The ED50 and E m a x values could not be determined for CYP1 A l because the induction did not level off. Several studies examined the effects of single PCB congeners on hepatic C Y P induction (Li et al, 1998; Chu et al, 1998; Chu et al, 1996). Exposure to PCB mixtures is relevant when evaluating the environmental effects of PCBs. Acute intraperitoneal exposure of twenty-day old female rats to a reconstituted PCB mixture (from N P L site soil) for two days caused 18-fold and 3-fold induction in hepatic CYP1A (measured as EROD) and CYP2B (measured as PROD) levels, respectively, at a dosage 12mg/kg/day. Maximum CYP1A and CYP2B induction were achieved at dosages of 36 mg/kg/day and 48 mg/kg/day, respectively. The TEF values of the Aroclor 1248 and the soil extracts (TEF values of 0.01101 and 0.00863, respectively) were similar (Hansen et al, 1995). The C Y P induction patterns observed in the Hansen et al.'s (1995) study did not compare to our results because we found a greater induction of hepatic CYP2B than CYP1A. The 134 greater CYP1A induction in the Hansen et al's (1995) study was due to the presence of PCDDs and PCDFs in the NPL soil samples. Parkinson et al. (1983) reported a 60-, 33-, 14-, and 5-fold induction of hepatic CYP2B, CYP1A1, CYP1A2, and CYP2A1, respectively, following i.p. exposure to Aroclor 1254, at a dosage of 1500 umol/kg. In addition, PROD activity was found to be a more sensitive indicator of Aroclor 1260 exposure than EROD activity (Harris et al, 1993). In a study by Harris et al. (1993), Aroclor 1260 induced PROD activity to a greater degree compared to Aroclor 1232, Aroclor 1242, Aroclor 1248 and Aroclor 1254. Maximal induction was observed at an Aroclor 1260 dosage of 160 mg/kg. PCBs induce a variety of hepatic enzymes including C Y P enzymes, epoxide hydrolase, cytochrome bs, and NADPH-cytochrome C Y P reductase (Madhukar and Matsumura, 1979; Parkinson et al, 1983). Hepatic CYP enzymes are the most sensitive mixed-function oxidase system to PCB treatment (Madhukar and Matsumura, 1979). Planar dioxin-like PCBs, such as PCB#126, induce CYP1A enzymes whereas non-planar PCBs, such as PCB#110, induce CYP2B enzymes (Li et al, 1998). Some PCBs, such as PCB#153, are mixed-type inducers and induce both CYP1A and CYP2B enzymes (Chu et al, 1996) whereas other PCBs do not induce any CYP enzymes (Safe, 1994). Planar PCBs induce C Y P enzymes by binding to the Ah receptor. CYP2A, CYP2B and CYP3A enzymes are induced via the SXR, P X R and CAR, which non-planar PCBs may bind to. PCB mixtures elicit complex induction patterns that involve both dioxin-like and non-dioxin-like mechanisms. Our results compare to previous findings and confirm that hepatic CYP2B induction is a sensitive bio-marker of Aroclor 1260 exposure. We found that hepatic 135 CYP2B enzymes were induced by Aroclor 1260, over a wider dose range (cumulative dose range of 0-313 mg/kg) than previously reported (cumulative dose range of 0-35 mg/kg) by Ngui and Bandiera (1999). The CYP induction pattern observed with Aroclor 1260 treatment is not similar to that of N P L sites samples, which induce CYP1A to a greater extent than CYP2B due to significant concentrations of PCDDs and PCDFs. However, results from studies involving environmental PCB mixtures indicated that CYP2B induction was linear over a wide dose range. 4.7. S U M M A R Y OF RESULTS Table 31. Summary of dose-response effects on immune and endocrine parameters in adult male rats Effect L O E L Emax, maximal effect observed. E D 5 0 (mg/kg/day) Body weight change 156 mg/kg/day Relative liver weight 3.13 mg/kg/day 8.63 ± 0.39% 6.36 ±3.46 Antibody response (anti-SRBC IgM titer) 1.25 mg/kg/day 512 ± 32 1.52 ±0.37 Serum testosterone levels No change Serum T4 levels 6.25 mg/kg/day 49.3 ± 1.8 ng/ml 4.69 ±0.79 Total C Y P content 1.25 mg/kg/day 1.86 ±0.05 nmol total CYP/mg protein 0.66 ± 0.20 CYP1A1 protein levels 3.13 mg/kg/day n.d. n.d. CYP1A2 protein levels 6.25 mg/kg/day 33.8 ± 2.1 pmol/mg protein 21.3 ±3 .3 CYP2B1 protein levels 0.625 mg/kg/day 1290 ± 96 pmol/mg protein 5.38 ±1.28 CYP2B2 protein levels 0.625 mg/kg/day 778 ± 49 pmol/mg protein 3.99 ±0.72 Testosterone 60-hydroxylase 1.25 mg/kg/day 3870 ± 158 nmol metabolite/min/mg protein 0.45 ± 0.43 Testosterone 7a-hydroxylase 6.25 mg/kg/day 1080 ± 103 nmol metabolite/min/mg protein 7.74 ±3.57 Testosterone 16a-hydroxylase 1.25 mg/kg/day 5680 ± 280 nmol metabolite/min/mg protein 0.23 ±0.19 Testosterone 160-hydroxylase 0.625 mg/kg/day 2700 ± 168 nmol metabolite/min/mg protein 1.24 ±0.27 Testosterone 2a-hydroxylase 3.13 mg/kg/day 2350 ± 115 nmol metabolite/min/mg protein 7.69 ± 1.69 Testosterone 20-hydroxylase 0.625 mg/kg/day 573 ± 25 nmol metabolite/min/mg protein 2.16 ±0.46 Androstenedione formation 0.625 mg/kg/day 4690 ± 236 nmol metabolite/min/mg protein 0.53 ±0.25 n.d.: not determined c ue to linear increase in CYP1A1. 137 The endpoints measured in this study are summarized in Table 31. Treatment with Aroclor 1260 resulted in an increase in total hepatic C Y P content and liver weight at dosages greater than 1.25 mg/kg/day. Aroclor 1260 decreased body weight at a dosage of 156 mg/kg/day. Based on ED50 values the antibody response to SRBC was the most sensitive endpoint. There was a decrease in antibody response in a dose-dependent manner at dosages greater than 1.25 mg/kg/day, but no effect on thymus weight following treatment with Aroclor 1260. Aroclor 1260 suppressed serum T4 levels at dosages greater than 6.25 mg/kg/day, but serum L H and testosterone levels were unaffected. Testes, ventral prostate, and seminal vesicle weights were also unchanged. Based on L O E L values the most sensitive endpoint was hepatic CYP2B induction. There was a large and dose-dependent induction of hepatic CYP2B protein levels and 16(3-hydroxylase activity up to dosages of 15.6 mg/kg/day and 6.25mg/kg/day, respectively. Hepatic CYP1A1 protein levels were also increased at dosages greater than 3.13 mg/kg/day, whereas CYP1A2 protein levels remained unchanged up to a dosage of 6.25 mg/kg/day. Hepatic CYP2A1 and CYP3A enzymes were induced, but CYP2C11 enzyme levels were suppressed as indicated by the testosterone hydroxylation data. 4.8. CONCLUSIONS (1) Of the endocrine endpoints examined in the present study, only serum T4 levels were affected by exposure to Aroclor 1260 in adult male rats. Serum T4 was a sensitive and reliable indicator of endocrine toxicity, following Aroclor 1260 treatment. (2) Reduction of humoral immunity, as assessed by the anti-SRBC IgM titer, was a highly sensitive and early response to exposure to Aroclor 1260 in adult male rats. A 138 significant decrease in humoral immunity was observed following treatment at a dosage of 1.25 mg/kg/day (1.25 ppm), which approximates reported PCB levels in biological and environmental samples. (3) Induction of hepatic CYP2B and, to a lesser degree, CYP1A was a sensitive hepatic marker of exposure to Aroclor 1260 in adult male rats. Aroclor 1260 treatment induced hepatic CYP2B enzyme levels over a wider dosage range (0-15.6 mg/kg/day) than previously reported. 4.9. FUTURE STUDIES This study demonstrates that Aroclor 1260, a commercial PCB mixture consisting of non-planar PCBs, causes immune suppression and endocrine toxicity in adult male rats at relatively low dosages (less than 5mg/kg/day). Antibody response and serum total T4 concentration were sensitive indicators of endocrine and immune toxicity, respectively. Serum L H and testosterone levels were unaffected however, and we did not see any evidence of reproductive toxicity. Therefore, more sensitive endpoints are needed for direct assessment of the effects of PCBs and other xenobiotics on reproductive function. The duration of PCB exposure can be increased, so that long-term effects of PCB treatment can be measured. For example, changes in serum TSH levels and thyroid gland weight may be observed in a chronic study where dosages smaller than the ones administered in this study are used. In addition, effects on testis weight are not apparent in sub-chronic studies although we suspect that testes may be insensitive to PCB treatment in adult animals, regardless of the duration of exposure. In utero, neonatal or 139 life-long exposure is needed to examine PCB-induced reproductive and neuro-behavioral effects. The main objective of this study was to determine the dosages at which PCBs cause endocrine and immune effects. Dose-response analysis is not only the best approach to this question, but also provides a reference point for assessment of the biological alterations following environmental PCB exposure. In fact, environmental samples contain significant amounts of other xenobiotics in addition to PCBs. 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Archives of Biochemistry and Biophysics 369: 11-23, 1999. Waxman, D.J. and Azaroff, L. Phenobarbital induction of cytochrome P-450 gene expression. Journal of Biochemistry 281: 577-592, 1992. Waterman, M.R. and Johnson, E.F. (eds). Methods in Enzymology, vol. 206. Cytochrome P450. Academic Press, Inc. California. 1991. Welch, R . M . , Levin, W., Kuntzman, R., Jacobson, M . and Conney, A . H . Effect of hallogenated hydrocarbon insecticide on the metabolism and uterotropic action of estrogens in rats and mice. Toxicology and Applied Pharmacology 19: 234-246, 1971. Wickizer, T .M. , Brilliant, L .B. , Copeland, R., and Tilde, R. 1981. Polychlorinated biphenyl contamination of nursing mothers' milk in Michigan. American Journal of Public Health 71: 132-137, 1981. Whitlock, Jr., J.P. Induction of cytochrome P4501A1. Annual review of Pharmacology and Toxicology 39: 103-125, 1999. 161 Wolff, M.S. Occupational exposure to polychlorinated biphenyls (PCBs). 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Effect of treatment with corn oil on relative thymus, testes and ventral prostate weights in adult male rats Relative thymus weight (% final body weight) Relative testes weight (% final body weight) Relative ventral prostate weight (% final body weight) Untreated 0.16 ±0.01 0.97 ±0.09 0.082 ± 0.004 Corn oil (2.5 ml/kg/day) 0.14 ±0.01 0.97 ±0.07 0.082± 0.003 Values are expressed as mean ± SEM. The corn-oil group consisted of six rats and the untreated group consisted of nine rats. Adult male rats were treated with corn oil by oral gavage once daily for ten days. Table 1.2. Effect of treatment with Aroclor 1260 on relative thymus, testes and ventral prostate weights in adult male rats Dosage of Relative Relative Relative ventral Aroclor 1260 thymus weight testes weight prostate weight (% (mg/kg/day) (% final body (% final final body weight) weight) body weight) Corn oil 0.14 ±0.01 0.97 ± 0.07 0.082± 0.003 0.025 0.13 ±0.01 1.06 ±0.03 0.099± 0.010 0.13 0.14 ±0.01 1.04 ±0.01 0.088± 0.006 0.625 0.16 ±0.01 1.00 ±0.05 0.077± 0.003 1.25 0.18 ±0.01 1.00 ±0.01 0.086± 0.006 3.13 0.17 ±0.01 0.94 ± 0.02 0.088± 0.007 6.25 0.17 ±0.01 1.04 ±0.03 0.081± 0.009 15.6 0.14 ±0.02 0.99 ±0.03 0.092± 0.008 31.3 0.15 ±0.01 1.02 ±0.05 0.092± 0.005 78.1 0.16 ±0.01 1.06 ±0.04 0.074± 0.006 156 0.13 ±0.02 1.23 ±0.03 0.077± 0.003 Values are expressed as mean ± SEM. Each treatment group consisted of six rats except for the 156 mg/kg dosage (n=5) groups. Adult male rats were treated with Aroclor 1260 in corn oil by oral gavage once daily for ten days. 163 6.2. APPENDIX II The TSH and L H assays were validated elsewhere (Desaulniers et al, 1997). The intra-assay coefficients of variation were determined from six measurements to be 8.5% and 11.8%) for TSH and L H respectively. The sensitivity of the TSH and L H assays were 0.04 and 0.07 ng/tube respectively. Table 11.1. Effects of treatment with corn oil on serum L H and TSH levels in adult male rats Untreated Corn oil (2.5ml/kg/day) L H (ng/ml) 0.84 ± 0.23 0.65 ± 0.06 TSH (ng/ml) 2.51 ± 0.56 2.71 ± 0.83 Values are expressed as mean ± SEM. The corn oil-treated group consisted of six rats and the untreated group contained nine rats. Adult male rats were treated with corn oil by oral gavage once daily for ten days 5? "5b o N O C N T - H O o o CD 00 cd G O o Q N O O O ro NO in in C N N O m CN m C N N O © in C N o o U oo oo N O o o -H oo en m O N O in in O N m oo © © 4 ! O N o o +i O N ™ in o © +1 N O © © -H C N O O O N © © © +1 N O N O O O © o +1 O O © © +1 C N O N « r-~ © © +1 in N O N O © © +1 X .-J ^ ll 2 £p CD S3 C O W C--C N ° ^ C N C N K +1 C N C N K 4 in ON ° ° . CO © N O -H # N O m °° O N -H in -H N O P in 41 oo © cn 41 © r— O to 41 in N O © N O © C N 41 © © m 41 r- © C N 41 53 g» C O -S CD 43 00 <D !§ N to 3- ^ CD J^' Ob -s cd S3 oo cd o o -a tin oo '3 ob "ob'co N O in cd " O H « c2 ° x >> CD r=! 43 H_i T3 S-i CD X g o CD 00 co 13 g CD cd .co cso 00 — 13 g ° o | s I . a e © +3 N O 2 O 1 — 1 cd . o C O ^ 41 , S3 cd CD cd T3 CD CO C O (D S-I O H X CD CD co CD 3 in © © V o !> Ob <2 1 3 2 £ CD <D kH Cd +7* a cd 13 > < 2 o o 165 APPENDIX III Table III.l. Effect of corn oil treatment on hepatic CYP1A and CYP2B protein levels in adult male rats CYP1A1 (pmol/nmol total CYP) CYP1A2 (pmol/nmol total CYP) CYP2B1 (pmol/nmol total CYP) CYP2B2 (pmol/nmol total CYP) Untreated n.d. 4.3 ±0.8 6.6 ± 0.2 7.6 ± 1.5 Corn oil (2.5 ml/kg/day) n.d. 3.5 ±0.9 8.3 ±3 .0 5.8 ±0 .9 Values are expressed as mean ± SEM. The corn oil-treated group consisted of six rats and the untreated group contained nine rats. Adult male rats were treated with corn oil by oral gavage once daily for ten days, n.d.: not determined (< LOQ=0.0625 pmol/lane) Table III.2. Effects of Aroclor 1260 treatment on hepatic CYP1A and CYP2B protein levels in adult male rats Dosage CYP1A1 CYP1A2 CYP2B1 CYP2B2 (ml/kg/day) (pmol/nmol total (pmol/nmol total (pmol/nmol total (pmol/nmol total CYP) CYP) CYP) CYP) Corn oil n.d. 3.5 ±0 .9 8.3 ±3 .0 5.8 ±0 .9 0.025 n.d. 3.7 ± 0.6 14.6 ±9 .4 12.2 ± 1.9 0.13 n.d. 4.0 ±0 .7 25.3 ± 13 11.7±2.5 0.625 n.d. 3.8 ±0.7 84.0 ± 13* 73.6 ±8 .4* 1.25 n.d. 2.6 ±0.5 168 ± 3 8 * 102±19* 3.13 7.3 ± 2.1* 2.5 ±0 .4 414 ± 6 0 * 232± 16* 6.25 12.0 ± 1.2* 7.2 ± 1.5* 381 ± 3 8 * 314 ± 4 7 * 15.6 18.3 ± 1.3* 7.5 ±0.9* 585 ± 9 8 * 372 ± 60* 31.3 32.1 ±3 .6* 12.1 ±2.2* 723 ± 75* 405 ± 49* 78.1 36.4 ± 4 . 1 * 18.2 ±3 .8* 756 ± 9 2 * 447 ± 4 3 * 156 44.7 ±5 .7* 17.3 ± 1.2* 668 ± 70* 464 ± 8 5 * Values are expressed as mean ± SEM. Each treatment group consisted of six rats except for the 156 mg/kg dosage (n=5) groups. Adult male rats were treated with Aroclor 1260 in corn oil by oral gavage once daily for ten days. 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