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Effects on early in ovo 2,3,7,8-tetrachlorodibenzo-P-dioxin exposure on perinatal thyroid and sex steroid… Janz, David Michael 1994

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EFFECTS OF EARLY IN OVO 2,3,7,8TETRACHLORODIBENZO-P-DIOXIN EXPOSURE ON PERINATAL THYROID AND SEX STEROID HORMONE LEVELS IN THREE AVIAN SPECIES by  DAVID MICHAEL JANZ B.Sc., Simon Fraser University, 1987 M.Sc., Trent University, 1991 A THESIS IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE FACULTY OF GRADUATE STUDIES THE FACULTY OF PHARMACEUTICAL SCIENCES Division of Pharmacology and Toxicology  We accept this thesis as conforming to the required standard  The University of British Columbia September 1995 © David Michael Janz,  1995  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.  (Signature)  Dopartiicr of  vI  rlioy wimC€  The University of British Columbia Vancouver, Canada Date  DE-6 (2/88)  ç  l C  I  (  Ctfli ff  ABSTRACT Halogenated aromatic hydrocarbons (HAHs) are a class of highly toxic and persistent environmental pollutants that include the polychlorinated dibenzo-p-dioxins (PCDDs), polychiorinated  dibenzofiirans  and  (PCDFs),  polychlorinated  biphenyls  (PCBs).  Contamination of aquatic systems with these chemicals is believed to be responsible for impaired reproductive capacity in species at the top of the food web, such as fish-eating birds. A highly sensitive and specific biomarker of exposure to HAHs is the induction of cytochrome P4501A1 (CYP1A1) and associated hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activity.  This thesis examined a number of toxicologically relevant hormonal  endpoints, such as plasma thyroid and sex steroid hormone concentrations and hepatic estrogen receptor (ER) levels, as potential biomarkers of HAH exposure and toxic effect during the perinatal period in three avian species, in comparison to hepatic EROD induction. Since thyroid and sex steroid hormones are important in perinatal growth and development in avian species, perturbations in levels of these hormones during this period may be a factor in the decreased reproductive success observed in wild fish-eating bird species inhabiting aquatic systems contaminated with HAHs. Fertile eggs of the domestic chicken (Gallus gallus), pigeon (Columba livia), and great blue heron (Ardea herodias) were exposed early in incubation to the most toxic HAH, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Eggs were artificially incubated and sacrificed at relevant time points during the perinatal period.  In chicken embryos and hatchlings  exposed to 0.1 p.g/kg of TCDD via air cell injection, hepatic EROD activity was induced halfmaximally to maximally at all time points, indicating activation of the aryl hydrocarbon receptor (AhR) complex, an initial step in the mediation of TCDD toxicity. There were no , 17f3T ) , total 4 3 effects of TCDD exposure on hatchability, plasma thyroid hormone (total T estradiol, or testosterone concentrations, hepatic ER levels, or a number of body and skeletal growth-related parameters in chickens.  In pigeons (1 .ig/kg), hepatic EROD was induced 11  half-maximally at hatch and day 7 after hatch.  There were significant reductions in  hatchability and several growth-related parameters in TCDD-exposed pigeons  (p<O.O1),  but  no effect on plasma thyroid hormone, estradiol, or testosterone concentrations, or hepatic ER levels. In great blue herons (2 .tg/kg) sacrificed at hatch and day 7, hepatic EROD activity was elevated three- and two-fold above controls, respectively, similar to EROD induction observed in heron hatchlings exposed to environmental levels of HAHs in the Strait of Georgia, BC.  In TCDD-exposed herons, there were elevated hepatic ER concentrations  (Bmax) and decreased ER affinities (Kd) at hatch (p<O.05), but not on day 7 after hatch. H]-TCDD concentrations decreased 14-fold between hatch and day 7. There was no 3 Liver [ effect of TCDD exposure on plasma thyroid hormone, estradiol,  or testosterone  concentrations in herons. In a separate experiment, pigeons were exposed during the latter third part of incubation to a TCDD dose (3 FIg/kg) which would cause high embryo mortality if exposure occurred early in incubation. In this experiment, hepatic EROD was induced half-maximally 3 and estradiol levels were decreased at hatch, and estradiol at hatch and day 7. Plasma total T . Similar to herons, hepatic ER (p<O.O ) was elevated on day 7 in TCDD-exposed pigeons 5 concentrations were increased at hatch (p<O. 01). Overall, the results of this thesis consistently show that in chickens, pigeons, and great blue herons exposed to TCDD early in incubation, thyroid and sex steroid hormone concentrations are not as sensitive as hepatic EROD activity, body growth, or mortality as biomarkers of exposure. However, the results suggest that TCDD may influence hepatic ER levels during the perinatal period in certain avian species.  111  TABLE OF CONTENTS page Abstract  ii  Table of Contents  iv  List of Tables  vii  List of Figures  ix  List of Abbreviations  Xl  Acknowledgments  1.  xii  INTRODUCTION  1  l.A. TOXICOLOGY OF HALOGENATED AROMATIC HYDROCARBONS  1  1 .A. 1. Sources of Halogenated Aromatic Hydrocarbons  1  1 .A.2. Environmental Distribution and Fate of HAHs  4  1.A.3. Mechanism of Action of HAHs  5  1 .B. TOXICITY OF HAHs iN WILD FISH-EATING BIRDS  9  1 .B. 1. Great Lakes Ecosystem  10  1.B.2. Strait of Georgia, British Columbia  12  iC. BIOMARKERS OF HAH EXPOSURE AND EFFECT IN AVIAN SPECiES 1.C.1. CYP1A1 Induction as a Biomarker of TCDD Exposure  14 14  in Avian Species i.C.2. Thyroid Hormones as Biomarkers of TCDD Exposure  17  and Effect in Avian Species 1 .C.3. Sex Steroid Hormones and Receptors as Biomarkers of  18  TCDD Exposure and Effect in Avian Species iD. HYPOTHESES AN]) OBJECTIVES  20  1.D.i. Hypotheses  20  1.D.2. Objectives  20 iv  2. MATERIALS AND METHODS  21  2.1. Experimental Birds  21  2.2. Preparation of TCDD Stock Solutions  21  2.3. Injection procedure  22  2.4. Early In Ovo [ H]-TCDD Distribution in Chicken Eggs 3  25  2.5. Reproducibility 3 of[ H ]-TCDD Injection  26  2.6. Dose-Response Experiments in Chickens and Pigeons  27  2.7. In Ovo TCDD Exposure Experiments  29  2.8. Preparation of Hepatic Microsomal and Cytosolic Fractions  33  2.9. Ethoxyresorufin O-Deethylase (EROD) Assay  34  2.10. Protein Assay  34  2.11. EROD/Total Cytochrome P450 Comparison Experiment  35  2.12. Hepatic Estrogen Receptor Assay  36  2.13. Radioimmunoassays  39  2.14. [ H]-TCDD Concentrations in Liver and Yolk 3  44  2.15. Adult Great Blue Herons  45  2.16. Statistical Analyses  47  3. RESULTS  48  3.A. AVIAN EMBRYOS AND HATCHLINGS  48  3 .A. 1. Fertility and Hatchability  48  3.A.2. Body Growth  50  3.A.3. EROD Activity  53  3.A.4. Plasma Thyroid Hormone Concentrations  56  3.A.5. Plasma Sex Steroid Hormone Concentrations  65  3.A.6. Hepatic Estrogen Receptor (ER) Levels  66  V  3.A.7. [ H1-TCDD Concentrations in Liver and Yolk 3 3.B. ADULT GREAT BLUE HERONS  78 80  3.B.1. Acute TCDD Exposure Experiment  80  3.B.2. Seasonal Plasma Hormone Concentrations  82  4. DISCUSSION  86  4.A. AVIAN EMBRYOS AND HATCHLINGS  86  4.A. 1. Hepatic EROD Induction in Avian Embryos and Hatchlings  86  4.A. 1.1. Chicken Embryos and Hatchlings  86  4.A. 1.2. Pigeon Hatchlings  89  4.A. 1.3. Great Blue Heron Hatchlings  90  4.A.2. [ H]-TCDD Concentrations in Liver and Yolk of Avian Hatchlings 3  92  4.A.3. Thyroid Hormones and Growth in Avian Embryos and Hatchlings  93  4.A. 3.1. Chicken Embryos and Hatchlings  93  4.A.3.2. Pigeon Hatchlings  95  4.A. 3.3. Great Blue Heron Hatchlings  97  4.A.4. Sex Steroid Hormones and Receptors in Avian Embryos and Hatchlings 4.A. 4.1. Chicken Embryos and Hatchlings  98 99  4.A. 4.2. Pigeon Hatchlings  101  4.A. 4.3. Great Blue Heron Hatchlings  102  4.B. ADULT GREAT BLUE HERONS  105  4.B.1. Acute Effects of TCDD in Adult Great Blue Herons  106  4.B.2. Seasonal Hormone Levels in Great Blue Herons  107  4.C. FUTURE RESEARCH  108  5. CONCLUSIONS  111  6. REFERENCES  113 vi  LIST OF TABLES Table  page  2.1.  Hatchability of chicken eggs injected into the air cell with 0, 12.5, or 25 itL of corn oil during early incubation.  23  2.2.  Hatchability of chicken eggs injected into the yolk with 0, 12.5, or 25 j.iL of corn oil during early incubation. Hatchability of pigeon eggs injected into the yolk with 0, 10, or 20 iiL of  24  2.3.  25  corn oil during early incubation.  2.4.  2.5.  Concentrations of [ H}-TCDD in chicken embryo, yolk, and albumin, 48 3 and 96 hours following either air cell or yolk injection of 0.01 pg of H]-TCDD on embryonic day 4.5. 3 [ Hatchability of chickens exposed in ovo to 0.1 jig/kg or 0.3 .tg/kg of TCDD, or corn oil vehicle, via air cell injection on embryonic day 4.5.  27  28  2.6.  Hatchability of pigeons exposed in ovo to 1, 2, or 3 p.g/kg of TCDD or corn oil vehicle via air cell injection on embryonic day 3.5.  28  2.7.  Comparison of hepatic total cytochrome P450 and EROD levels in chicken hatchlings using microsomes prepared in Tris/KC1 vs.  36  2.8,  HEDGM buffer and/or CO 2 asphyxiation. Performance characteristics of plasma hormone RIAs.  43  2.9.  Parallelism between dilutions of avian plasma and standard hormone solutions for each hormone RJA.  44  3.1  Developmental periods during which mortality was observed to occur in pigeons exposed to 1 jig/kg of TCDD.  49  3.2.  Body weights of chicken embryos and hatchlings exposed to 0.1 .tg/kgofTCDD.  50  3.3.  Crown-rump length, tibia length, and culmen length in chicken hatchlings exposed to 0.1 jig/kg of TCDD. Body measurements of pigeons exposed to I jig/kg of TCDD. Body measurements of pigeons exposed to 3 jig/kg of TCDD. Body measurements of great blue heron hatchlings exposed to 2 jig/kg of TCDD.  51  3.4. 3.5. 3.6.  VII  52 53 55  Table  page  3.7.  Hepatic estrogen receptor affinities and concentrations in female chicken embryos and hatchlings exposed to 0.1 p.g/kg of TCDD.  76  3.8.  Hepatic estrogen receptor affinities and concentrations in female pigeon  77  hatchlings exposed to 1 p.g/kg of TCDD. 3.9.  Hepatic estrogen receptor affinities and concentrations in female pigeon hatchlings exposed to 3 .tg/kg of TCDD.  77  3.10.  Hepatic estrogen receptor affinities and concentrations in great blue heron hatchlings exposed to 2 .tg/kg of TCDD.  78  3.11.  Concentrations of [ H]-TCDD in the liver and yolk of chicken, pigeon, 3  79  and great blue heron hatchlings. 3.12.  Hepatic EROD activities, plasma total T 4 and T 3 concentrations, and  80  plasma 4 /T ratio in adult great blue herons after 14 day exposure to 3 T 20 gig/kg of TCDD or corn oil vehicle. 3.13.  Plasma 1 7f3-estradiol and testosterone concentrations, and estrogen receptor affinities and concentrations in adult great blue herons after  81  14 day exposure to 20 rig/kg of TCDD or corn oil vehicle. 3.14.  Concentrations of [ H1-TCDD in liver, kidney, muscle, and fat tissues 3 of adult great blue herons after 14 day exposure to a single i.p. dose of 20 p.g/kg of TCDD.  viii  82  LIST OF FIGURES Figure  1.1. 1.2. 2.1. 2.2.  page  Representative structures of halogenated aromatic hydrocarbons. Mechanism of action of TCDD and related chemicals. Effect of cytosolic protein concentration on specific binding of H]-estradiol to hepatic estrogen receptor. 3 [ Linearity of nonspecific binding between 0.16 nM and 20 nM H1-estradiol in hepatic cytosol. 3 [  2 7 38 38  3.1.  Body weights of great blue heron hatchlings fed for seven days following hatch after in ovo exposure to 2 jig/kg of TCDD or corn oil vehicle.  54  3.2.  Hepatic EROD induction in chicken embryos and hatchlings exposed in ovo to 0.1 jig/kg of TCDD.  57  3.3.  Hepatic EROD induction in pigeon hatchlings exposed in ovo to 1 jig/kg of TCDD.  58  3.4.  Hepatic EROD induction in pigeon hatchlings exposed in ovo to 3 jig/kg of TCDD.  58  3.5.  Hepatic EROD induction in great blue heron hatchlings exposed in ovo to 2 jig/kg of TCDD.  59  3.6.  Plasma total 4 T total 3 , T and 4 , /T ratio in chicken embryos and 3 T hatchlings exposed in ovo to 0.1 jig/kg of TCDD. Plasma total T , total T 4 , and 4 3 /T ratio in pigeon hatchlings exposed 3 T in ovo to 1 jig/kg of TCDD. Plasma total T , total T 4 , and 4 3 /T ratio in pigeon hatchlings exposed 3 T in ovo to 3 jig/kg of TCDD.  60  3.7. 3.8. 3.9. 3.10. 3.11.  Plasma total T , total 3 4 T and T , 4 / 3 T ratio in naturally-raised pigeon hatchlings sacrificed on day 1, day 3, and day 6 after hatch. Plasma total T , total T 4 , and 4 3 /T ratio in great blue heron hatchlings 3 T exposed in ovo to 2 jig/kg of TCDD. Plasma 1 73-estradiol concentrations in female chicken embryos and hatchlings exposed in ovo to 0.1 jig/kg of TCDD.  ix  61 62 63 64 67  Figure 3.12. 3 13. 3.14. 3.15. 3.16. 3.17. 3.18. 3.19.  page Plasma 1713-estradiol concentrations in female and male pigeon hatchlings exposed in ovo to 1 p.g/kg of TCDD. Plasma 1 713-estradiol concentrations in female and male pigeon  hatchlings exposed in ovo to 3 rig/kg of TCDD. Plasma 1 7f3-estradiol concentrations in naturally-raised pigeon hatchlings sacrificed on day 1, day 3, and day 6 after hatch. Plasma I 7f3-estradiol concentrations in great blue heron hatchlings exposed in ovo to 2 i.gfkg of TCDD. Plasma testosterone concentrations in male chicken embryos and hatchlings exposed in ovo to 0.1 .tg/kg of TCDD. Plasma testosterone concentrations in great blue heron hatchlings exposed in ovo to 2 fig/kg of TCDD. Saturabiity of [ H1-estradiol specific binding, Woolf analysis, and 3 Scatchard analysis in a hepatic cytosol sample.  68 69 70 71 72 73 74 83  3.20.  Total 4 T and T 3 concentrations, and 4 /T ratio in adult great blue 3 T heron plasma samples collected monthly between September and May. 173-Estradiol concentrations in female and male adult great blue heron plasma samples collected monthly between September and May.  84  3.21.  Testosterone concentrations in female and male adult great blue heron plasma samples collected monthly between September and May.  85  x  LIST OF ABBREVIATIONS Ah AHH AhR AhRE ANOVA ARNT Bm CNS cpm CYP DCC DES dpm E EDTA ER ERE EROD GC-MS GLEMEDS HAH HEPES hsp Kd 0 K 50 LD LSC NADPH NSB p PCB PCDD PCDF 2 r R1A SEM 3 T 4 T TCDD TRIS TSH UBC UDP  aryl hydrocarbon aryl hydrocarbon hydroxylase aryl hydrocarbon receptor aryl hydrocarbon responsive element analysis of variance Ah receptor nuclear translocator protein apparent concentration of specific receptor binding sites central nervous system counts per minute cytochrome P450 dextran-coated charcoal diethylstilbestrol disintegrations per minute embryonic day of development, or day of incubation ethylenediamine tetraacetic acid estrogen receptor estrogen responsive element ethoxyresorufin O-deethylase gas chromatography-mass spectrometry Great Lakes embryo mortality, edema, and deformities syndrome halogenated aromatic hydrocarbon (N-[2-hydroxyethyl]piperazine-N’-[2-ethanesulfonic acid]) heat shock protein apparent equilibrium dissociation constant octanol:water partition coefficient dose causing mortality in 50% of a population of test animals liquid scintillation counting 13-nicotinamide adenine dinucleotide phosphate (reduced) nonspecific binding probability of a Type I error, CL polychlorinated biphenyl polychiorinated dibenzo-p-dioxin polychiorinated dibenzofuran coefficient of determination radioimmunoassay standard error of the mean triiodothyronine thyroxine 2,3,7, 8-tetrachlorodibenzo-p-dioxin 2-amino-2-(hydroxymethyl)- 1,3 -propanediol thyroid stimulating hormone University of British Columbia, Vancouver, BC uridine diphosphate xi  ACKNOWLEDGMENTS I am grateflul to Dr. Gail Beliward for her excellent supervision during my doctoral program.  Thank you for always being available for discussion and advice, allowing me  independence in my research project, and in general taking a genuine interest in my development as a scientist. I also thank you for your valuable advice on matters outside of the thesis. I thank my supervisory committee, Dr. Frank Abbott (Chair), Dr. Stelvio Bandiera, Dr. Kim Cheng, Dr. Jack Diamond, and Dr. Rob Thies, for helpftul advice concerning the thesis. I am  gratefhl to John Elliott, Canadian Wildlife Service, for arranging the collection of great blue heron eggs. I would also like to thank Dr. Thomas Sanderson, who was a valuable source of technical and theoretical advice and discussion throughout the thesis. This thesis was funded by an NSERC Postgraduate Scholarship and Eco-Research Doctoral Fellowship, and I thank NSERC and the Tn-Council Secretariat for these scholarships. Finally, I would like to thank Susan for her love and patience during the last four years. Our daughter Maia, born during the first year of the doctorate, and son Aidan, born during the last year, have provided us with incredible joy. Aidan.  xl’  This thesis is dedicated to Susan, Maia, and  1. INTRODUCTION 1 .A. TOXICOLOGY OF HALOGENATED AROMATIC HYDROCARBONS 1.A.1. Sources of Halogenated Aromatic Hydrocarbons The halogenated aromatic hydrocarbons (HAHs) are a class of highly toxic and persistent environmental contaminants which include the polychlorinated dibenzo-p-dioxins (PCDDs), polychiorinated dibenzofurans (PCDFs), polychiorinated biphenyls (PCBs), and other structurally similar compounds (Fig. 1.1).  The HAHs exist in the environment as  complex mixtures of hundreds of different chemicals which vary considerably in toxicity and resistance to environmental and metabolic degradation, The most toxic HAH, and also the most toxic anthropogenic chemical known, is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD; Fig. 1.1). Several of the HAHs, most notably the PCBs, had widespread industrial use in transformers, capacitors, heat transfer and hydraulic fluids, and other applications. Between their initial discovery in the 1 920s and eventual ceasing of production in the 1 970s, about 1.5 million metric tons of PCBs were produced worldwide, and it is estimated that 20-30% of this amount has entered the environment (De Voogt and Brinkman 1989). In contrast, the PCDDs and PCDFs are unwanted by-products of the large scale industrial chlorine chemistry of this century. PCDDs were initially identified as by-products in the production of trichiorophenols and the herbicides 2,4,5-trichlorophenoxyacetic acid (2,4,5-T) and 2,4-dichiorophenoxyacetic acid (2,4-D). Highly chlorinated PCDDs and PCDFs  are also formed during the production of the wood preservatives pentachiorophenol and tetrachlorophenol (Fiedler et al. 1990). Although initially thought to be restricted to a few reactions involving structurally related chemicals, thrther evidence of the release of PCDDs and PCDFs came when trace quantities of these compounds were measured in fly ash and air emissions from trash-burning 1  CI CI  9  (0  I  6  5  4  2,3 ,7,8-tetrachlorodibenzo-p-dioxin  CI.  ,cI  C[  ,cI  2,3,7, 8-tetrachlorodibenzofuran  ‘ii  CI.  3,3 ‘,4,4’-tetrachlorobiphenyl  Figure 1.1. Representative structures of PCDDs, PCDFs, and PCBs. The chemicals shown  are tetrachiorinated congeners of each class, and are known to have high affinity for the aryl 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is the most toxic hydrocarbon receptor. halogenated aromatic hydrocarbon, and is used in this thesis as a test compound.  2  incinerators (Olie et a!. 1977). PCDDs and PCDFs are also released into the environment during incineration of other chlorinated waste products, such as hospital and chemical waste, and sewage sludge (Fiedler et at. 1990).  The exhaust from motor vehicles using leaded  gasoline contains PCDDs and PCDFs, as well as polybrominated dibenzo-p-dioxins and dibenzofl.irans. Unleaded gasoline contains much less chlorine, and considerably lower levels of HAHs are released upon combustion of this fuel (Marklund et at. 1990). Nevertheless, the large number of motor vehicles used in today’s society will cause this to remain a significant source of PCDDs and PCDFs. More recently, PCDDs and PCDFs were discovered to be formed during the bleaching of pulp and paper with molecular chlorine. Although first suggested in the early 1970s, the formation and release of PCDDs and PCDFs from bleached Kraft pulp and paper mills was not confirmed until the mid-1980s (Kuehi eta!. 1987; Amendola et at. 1989). Canada has one of the largest pulp and paper industries in the world, and many mills employ the Kraft bleaching process.  Levels of PCDDs and PCDFs in biota occurring near certain mills in  Canada have exceeded federal health guidelines for safe consumption, resulting in fisheries closures. The elevated level of environmental contamination with HAHs is a 20th century phenomenon resulting from widespread industrial chlorine chemistry and the combustion of chlorine-containing fhels. The range of sources identified to date indicates that most industrial processes utilizing chlorine will generate HAHs at some point during production, use, or disposal (Webster and Commoner 1994).  The large growth of these industrial processes  during the 20th century has resulted in greatly increased levels of HAHs in both abiotic and biotic environments. Advances in analytical chemistry over the last two decades have revealed the presence of HAHs in virtually every compartment of the global biosphere.  3  1.A.2. Environmental Distribution and Fate of BARs The environmental distribution and fate of HAHs is dictated by their physical, chemical, and biological properties. These compounds exhibit high octanol:water partition coefficients (Kow), resulting in extremely low water solubilities and high lipophilicities. Most HAHs are chemically stable and resistant to environmental degradation via photolysis and  hydrolysis.  In addition, many of the HAHs are resistant to metabolic processes in vivo,  resulting in limited excretion from organisms once exposed. As a result, HAHs are extremely persistent and tend to bioaccumulate in biota (Connell 1988). Combustion sources of HAHs release these compounds into the atmosphere, where they are adsorbed onto airborne particulates or, in the case of lower halogenated compounds, are present as vapor. The stability of many HAHs results in long range atmospheric transport, and subsequent wet and dry deposition in remote areas (Koester and Hites 1992).  The  atmospheric transport of HAHs has resulted in their ubiquitous presence in the global environment. Due to the low water solubilities and low vapor pressures of HAHs, these compounds preferentially partition into soil and sediments. The half-life of TCDD in soil is in the order of 1-2 decades, and possibly longer in sediments (Fiedler et at. 1990).  In aquatic systems,  extremely low HAH levels in water are bioconcentrated by phytoplankton and zooplankton, and then biomagnified at each successive trophic level in the food web (Oliver and Niimi 1988). For example, levels of PCBs in fish-eating birds inhabiting the lower Great Lakes, animals near the top of the food web, are several million times greater than the PCB concentration in water (Clark et at. 1988).  Furthermore, the accumulation of HAHs in  sediment acts as a reservoir, providing a continuous source of these compounds to aquatic systems. Humans also occupy a high trophic level and obviously are not excluded from the accumulation of HAHs.  Humans are exposed to HAHs primarily via the diet.  These  compounds distribute mainly to body lipids, particularly adipose tissue. Elimination of HAHs 4  in humans is primarily via the feces for many of the more persistent and toxic HAHs, such as TCDD, which are metabolically degraded to a very small extent (Webster and Connett 1991). A second, and potentially more insiduous exposure route in humans is via in utero and lactational transfer to the fetus and infant, respectively (Sheeter 1991). In a similar fashion, birds, fish, and other oviparous vertebrates accumulate HAHs and transfer them to eggs. The result is a continuous exposure to HAHs from the first stages of development throughout life. Unfortunately, the widespread release, environmental distribution, and exposure of living organisms to HAHs cannot be viewed without concern. In contrast to the physical and chemical stability of certain HAHs, they are far from biologically inert, and are in fact some of the most toxicologically potent chemicals known.  1.A.3. Mechanism of Action of HAlls  Initial studies examining the toxicity of HAHs in laboratory animals revealed a similar pattern of biological and toxicological responses. These effects include weight loss, edema, endocrine disorders, teratogenicity, decreased reproductive success, hepatic porphyria, immunotoxicity, promotion of cancer, and altered levels of certain metabolic enzymes (reviewed in Goldstein and Safe 1989). mechanism  of action  of HAHs  and  Extensive research in the early 1 970s into the polycyclic  aromatic  hydrocarbons  (e.g.  3-  methylcholanthrene) culminated in the isolation of a stereospecific, high affinity receptor protein in hepatic cytosol of mice, named the aryl hydrocarbon (Ah) receptor (Poland et aL 1976). An excellent correlation emerged between the toxic potencies of individual HAHs and their binding affinities for the Ah receptor (Poland et aL 1979; Poland and Knutson 1982). Extensive structure-activity relationships were developed for PCDDs, PCDFs, and PCBs which revealed the basis for the stereochemical selectivity of the Ah receptor (Poland et at. 1979; Bandiera et aL 1982, 1984). Within each chemical class, the most toxic congeners have halogen atoms occupying at least three of the four lateral positions, with zero or a minimal number of halogens at ortho positions. This pattern of halogen substitution results in 5  a planar or nearly planar conformation of the molecule with respect to the two aromatic rings, and is required for high affinity binding to the Ah receptor (AhR). TCDD binds to the AhR with the highest affinity of any chemical (Kd< 1.5 nM), and has the greatest toxic potency of the HAHs. For these reasons, TCDD is used as a prototype” compound for toxicological studies of HAHs, and is used in this regard in this thesis.  Thus, HAHs are classified  biologically based on their ability to bind to the AhR, and are more accurately referred to as “TCDD-like I{AHs” or “TCDD and related chemicals”. It is now a general consensus that most, if not all, of the major effects of exposure to TCDD and related chemicals are initially mediated by binding to the AhR (Roberts 1991), although controversy still exists regarding this aspect of TCDD toxicity (Weber and Stahl 1995; Okey et al. 1995). Indeed, many important steps occurring between ligand binding to the AhR and expression of overt toxicities remain to be elucidated.  In addition, there are  significant species and tissue specificities to the effects of TCDD exposure which remain largely unexplained. However, the binding of TCDD-like HAHs to the AhR as a prerequisite for toxicity is supported by the abundant scientific evidence available and by the majority of the scientific community (Poland and Knutson 1982; Safe 1990; Whitlock 1990; Landers and Bunce 1991; Okey eta!. 1994, 1995; Hankinson 1995). The AhR in its unbound form is found in the cytoplasm (Fig. 1.2), associated with the 90 k.Da heat shock protein, hsp9o (Perdew 1988), and possibly other proteins (Perdew 1992). TCDD and related chemicals passively diffuse into the cell and bind to AhR in the cytosol, causing release of hsp9o. It is unclear whether dissociation of hsp9O occurs in the cytoplasm or nucleus, suggesting a possible active role of hsp90 in transporting the AhR into the nucleus (Hankinson 1995).  In the cytosol, the receptor-ligand complex is associated with an Ah  receptor nuclear translocator protein (ARNT) which transforms the receptor-ligand complex to an activated form able to bind to DNA (Hofthian et at. 1991). Phosphorylation may play a role in activation of AhR complex, although the exact nature of this role is not clear at present (Hankinson 1995).  The activated receptor-ligand complex binds to specific recognition 6  Ligand (e.g. TCDD)  .  Figure 1.2. Proposed mechanism of action of TCDD and related halogenated aromatic hydrocarbons. TCDD passively diffuses into the cell and binds to the inactivated aryl hydrocarbon receptor (AhR). In the inactive state, the AhR is associated with the heat shock protein, hsp9O. After binding, hsp9O is released, and the receptor-ligand complex is transformed to an activated AhR complex via association with the AhR nuclear translocator protein (ARNT). The activated AhR complex translocates to the nucleus, where it interacts with aryl hydrocarbon responsive elements (AhREs) located on DNA, resulting in increased or decreased transcription of specific mRNAs and subsequent protein synthesis. The induction of cytocbrome P4501A1 (CYP1A1) is the best characterized of these responses, but most, if not all of the many other biological and toxicological responses to TCDD exposure are also mediated via the AhR. Thus, the AhR acts as a ligand-activated transcription factor or “genomic switch”, affecting the synthesis of specific proteins.  7  sequences on the DNA, known as Ah-responsive elements (AhRE), located upstream of regulatory genes. Binding of the activated AhR complex to AhRE results in increased or decreased transcription of specific mRNAs, and subsequent translation of proteins (Fig. 1.2). Thus, TCDD-Eike chemicals act as genomic switches, affecting the synthesis of specific proteins. The best characterized response to TCDD exposure directly mediated by the AhR is the induction of cytochrome P4501A1 (CYP1A1; Nebert and Gonzalez 1987), and associated monooxygenase  activities,  mainly  ethoxyresorufin  0-deethylase  (EROD)  and  aryl  hydrocarbon hydroxylase (AHH). Other gene products under direct transcriptional regulation by the AhR include the closely related CYP1A2, UDP-glucuronosyltransferase, the Ya subunit , and aldehyde (NQO ) of glutathione S-transferase, NAD(P)H:quinone oxidoreductase 1 dehydrogenase-3 (reviewed in Okey et a!. 1994). Other genes whose expression is altered by TCDD-like HAHs include epidermal growth factor receptor, interleukin-1f3, and plasminogen activator inhibitor-2. It is not known whether altered expression of these genes, including induction of CYP1A1, is directly responsible for any toxic effects of TCDD and related chemicals (Okey eta!. 1994). The mechanism of action of sex steroid hormones, thyroid hormones, and retinoic acid (vitamin A) is functionally similar to the above mechanism of action of TCDD, and it was thought for several years that the AhR was a member of the steroid/thyroid/retinoic acid receptor superfamily (Evans 1988). However, recent cloning of mouse AhR revealed that the ligand-binding subunit of the AhR does not possess the “zinc finger” DNA-binding motif characteristic of steroid hormones, but rather has a basic helix-loop-helix DNA-binding motif (Burbach et a!. 1992; Ema eta!. 1992). Nevertheless, TCDD has been shown to affect a number of receptor systems, such as estrogen receptor (ER). Although TCDD does not bind to the ER, and estrogens do not bind to the AhR, TCDD has been shown to decrease the binding of 17f3-estradiol to ER in rats and mice (Romkes et a!. 1987; Umbreit and Gallo 1988), and decrease hepatic and uterine ER 8  concentrations in CD-i mice (DeVito et at. 1992). It is possible that the effects of TCDD-like HAHs on ER may involve allosteric mechanisms and cooperativity of proteins at the gene level (DeVito et aL 1991). The AhR appears to be widely distributed in mammalian and nonmammalian species (Okey et al. 1994). Phylogenetic studies show that the AhR apparently evolved at least 450 million years ago, during early vertebrate evolution (Hahn et at. 1994). The AhR has been measured in several mammalian species and cell lines, including human tissues and cell cultures (reviewed in Okey et at. 1994), and in rainbow trout hepatoma cells (Lorenzen and Okey 1990).  In birds, AhR has been detected in chicken (Gallus gallus) embryos and  hatchlings, pigeon (Columba livia), great blue heron (Ardea herodias), double-crested cormorant (Phalacrocorax auritus), and quail, but not in turkey embryo (Meleagris gallopavo), or jungle fowl (Denison et at. i986b; Brunström and Lund 1988; Sanderson and Bellward 1995).  The lack of detection in certain species may be a function of the low  receptor concentrations and/or instability of the AhR.  1 .B. TOXICITY OF HAHs IN WILD FISH-EATING BIRDS Although the majority of toxicological research of TCDD and related chemicals has focused on laboratory animals, there is evidence that environmental levels of these compounds may be adversely affecting certain wild animal populations. A particularly susceptible group of animals, mentioned previously, are fish-eating birds which inhabit areas contaminated with HAHs. This thesis is concerned with examining potential adverse effects and mechanisms of toxicity of TCDD in avian species.  9  1.B.1. Great Lakes Ecosystem The lower Great Lakes are one of the most HAH contaminated aquatic systems in the world.  High environmental levels of PCDDs, PCDFs, and PCBs are believed to be  responsible for the impaired reproductive capacity observed since the mid-i 960s (Keith 1966) in several colonial fish-eating bird species, such as herring gulls (Larus argentatus), doublecrested cormorants, Forster’s terns (Sterna forsteri), and common terns (Sterna hirundo). Surveys conducted in the early I 970s by the Canadian Wildlife Service found hatchabilities of less than 20%, and production of less than one fledged young per 10 nests in some herring gull colonies on Lake Ontario (Gilbertson 1983). Symptoms of the high embryo and hatchling mortality observed in wild birds resembled those of “chick edema disease”, a series of outbreaks in poultry farms in the U.S. and elsewhere caused by HAH-contaminated feed (Schmittle et a!. 1958; Firestone 1973). Gross toxicities observed in both chickens and wild piscivorous birds included elevated embryo and hatchling mortality, edema (mainly subcutaneous, abdominal, and pericardial), teratogenicity, and weight loss. Hence the term “Great Lakes embryo mortality, edema and deformities syndrome” (GLEMEDS) was used to describe it (reviewed in Gilbertson et aL 1991). Characteristic biochemical effects of GLEMEDS included induction of CYP1A1mediated AHH activity, hepatic porphyria, altered thyroid function, and decreased retinoic acid levels. Although concentrations of the most toxic planar HAHs could not be quantified at that time, archived samples reanalysed in the late 1 980s revealed high levels of TCDD and related chemicals, confirming the parallel between chick edema disease and GLEMEDS (Gilbertson et a!. 1991). These field studies were supported by similar findings in subsequent laboratory studies which compared hatchlings from artificially incubated herring gull eggs from Lake Ontario to eggs from relatively uncontaminated reference colonies in New Brunswick and northern Alberta (Gilbertson and Fox 1977).  10  Although a causal relationship emerged between concentrations of TCDD-like HAHs and direct toxic effects on embryos and hatchlings of fish-eating birds, it became apparent that anomalous parental behavior during incubation may also be playing a role. Independent field experiments were conducted in which eggs from contaminated and clean colonies were exchanged, and toxicity was compared to unexchanged eggs (Fox et at. 1978; Kubiak et a!. 1989). In both experiments, poor nest attentiveness and reduced heat applied to the eggs by parents from contaminated sites resulted in delayed hatching and decreased hatchability. Furthermore, in contaminated eggs incubated by parents from reference sites, hatchabiity was greatly increased, while uncontaminated eggs incubated by parents from the contaminated sites displayed reduced hatchability. Thus, GLEMEDS appeared to be a result of factors both intrinsic and extrinsic to the egg (Fox et aL 1978; Kubiak et at. 1989; Gilbertson et at. 1991). The ultimate rationale for studying the effects of environmental contamination on biological systems is to determine whether exposure of humans and other species to this contamination has an adverse impact on health.  Humans cannot be used directly for this  purpose for obvious ethical reasons. If adverse effects are observed in wild species at the top of the food web, such as piscivorous birds, then humans may also be at risk to deleterious health outcomes. Therefore, wildlife biomonitoring programs are used to assess the quality of environmental health. In particular, chemical analysis of eggs from colonial fish-eating birds in the Great Lakes has been advocated as an indicator of chemical contamination of the aquatic environment (Clark et aL 1988; Fox et al. 1991). In this context, these birds are referred to as sentinel species, biomonitoring species, or bioindicator species.  In the Great Lakes, fish-  eating birds and other wild species have been used as indicators of environmental quality since the 1970s (Fox et a!. 1991). Similar effects to those observed in the Great Lakes have been documented in fisheating birds inhabiting other aquatic systems contaminated with TCDD and related chemicals, such as the Rhine estuary in the Netherlands (Van den Berg et at. 1994) and San Francisco Bay (Hoffman et a!. 1986). Locally, impaired reproductive success has been observed in fish 11  eating birds inhabiting the Strait of Georgia, British Columbia, and will be the focus of the next section.  1.B.2. Strait of Georgia, British Columbia The Strait of Georgia is a highly productive marine ecosystem separating the mainland of southwestern British Columbia from Vancouver Island. Several species of non-migratory colonial fish-eating birds inhabit this area, including great blue herons and double-crested cormorants. This region also has a large number of bleached Kraft pulp mills which discharge their effluent into the Strait.  As mentioned previously, effluent from the bleached Kraft  process is known to contain PCDDs and PCDFs. The Canadian Wildlife Service has measured contaminant levels in great blue heron eggs as an indicator of environmental quality in the Strait of Georgia since 1977.  The  biological characteristics of great blue herons in this area make them especially suitable as biomonitoring species. This species is a year round resident in the Strait of Georgia, so it is expected that the entire body burden of HAHs is accumulated from local sources. Herons are locally abundant, and collection of one or two eggs per nest does not have adverse effects on the population. This species nests colonially, making it relatively easy to collect reasonable numbers of eggs. Since TCDD and related chemicals are highly persistent and lipophilic, and tend to biomagnify in each successive level in the food web, the use of a relatively longer-lived species occupying a high trophic level, such as the heron, is desirable. In addition, great blue herons have a broad geographical distribution, allowing comparisons to be made between different regions. In 1987, a great blue heron colony near Crofton on the east coast of Vancouver Island was observed not to fledge a single offspring from 57 active nests. The heron colony nests and feeds less than one km from the outfall of a bleached Kraft pulp mill. Chemical analysis of embryos from the colony revealed a threefold increase in concentrations of TCDD compared to the previous year (Elliott et at. 1989). Elevated levels of PCDDs and PCDFs were also 12  reported in fish, crabs, bivalve molluscs, and sediment collected near the effluent discharge in 1987 (Norstrom et a!. 1988). A study was initiated in 1988 to determine whether TCDD contamination was adversely affecting the reproductive capacity of herons from the Crofton colony (Bellward et a!. 1990; Hart eta!. 1991). Paired eggs from Crofton and two other heron colonies nesting in the Strait of Georgia, known to have intermediate and low TCDD contamination, were collected. One egg from each nest was analysed for PCDDs and PCDFs (Norstrom et at. 1990), and the other was artificially incubated in the laboratory until hatching. Beliward eta!. (1990) found a high correlation between hepatic microsomal EROD activities in day old hatchlings and TCDD concentrations in the paired eggs. As mentioned previously, elevated EROD activity is a measure of CYP1A1 induction, a process known to be under direct regulation by the Ah receptor. Although several other PCDDs and PCDFs were detected in heron eggs, multiple regression analysis indicated that TCDD alone was the only predictor variable for EROD activity (Bellward et a!. 1990). Although there was no significant mortality in the artificially incubated Crofton eggs, there were significant negative relationships observed between concentrations of TCDD and the following growth-related parameters: yolk-free body weight, tibia length, tibia weight, plasma calcium concentration, beak length, and kidney and stomach weight (Hart et aL 1991). In addition, subcutaneous edema was observed in four of twelve Crofton hatchlings, and there were fewer down follicles on the heads of hatchlings. In all, these findings are consistent with the known effects of TCDD and related chemicals on wild fish-eating birds. Between 1990 and 1992, fl.irther studies using artificially incubated great blue heron and double-crested cormorant eggs from various locations in the Strait of Georgia were conducted as part of the environmental monitoring program (Sanderson et at. 1994a,b). As part of Canadian federal legislation aimed at limiting discharges of persistent organochlorine compounds into the environment, the Crofton mill began implementing process changes in 1988.  These changes initially involved monitoring wood chips for polychlorophenol 13  2 contamination, and removal of contaminated chips as feed stock. Soon after this, 50% C10 substitution in place of molecular chlorine in the bleaching process was introduced. By 1991, levels of TCDD and related PCDDs had fallen to approximately 10% of the 1987 levels (Whitehead et at. 1992). Concomitant with the fall in PCDD and PCDF levels in heron eggs, hepatic EROD activities and incidence of edema were decreased, body weights were increased, and the reproductive success of the Crofton colony was improved (Sanderson et at. 1 994a). This remarkably rapid change was believed to be due mainly to prey selection preferences by the adult herons, which feed primarily on one and two year old fish (Butler 1990). These juvenile fish would have been exposed to lower levels of PCDDs and PCDFs, assuming bioaccumulation primarily via effluent, resulting in a rapidly decreased level of exposure to herons.  BIOMARKERS OF HAH EXPOSURE ANT) EFFECT IN AVIAN  1 .C.  SPECiES 1.C.1. CYP1A1 Induction as a Biomarker of TCDD Exposure in Avian Species  Biological responses that change in a predictable, reproducible, and dose-related manner to xenobiotic exposure are referred to as biomarkers. Certain enzyme activities are commonly used in toxicological studies as biochemical markers because they are often sensitive, specific, and rapidly elicited responses.  Since chemical analysis of TCDD and  related compounds is extremely expensive and only conducted in a few laboratories worldwide, the use of appropriate biomarkers of exposure for these contaminants is especially relevant.  The induction of CYPIA1 and associated enzymes (EROD, AHH) have been  advocated as biomarkers in fish for several years (Payne 1976; Payne et a!. 1987; Luxon et at.  14  1987). More recently, induction of EROD and ANH have been proposed as biomarkers in birds and other wildlife (Mineau et al. 1984; Rattner et a!. 1989; Kubiak et a!. 1989). The induction of CYP1A1 is a suitable biomarker of exposure to TCDD and related Ah receptor-activating chemicals for several reasons. It is a sensitive response, that is, the dose-response curve for CYP1A1 induction is found to the left of most other responses to TCDD exposure. CYP1A1 induction is also a very specific response, since binding of TCDD like ligands to the Ah receptor is a prerequisite for induction, and constitutive levels of CYP1A1 are low (Beresford 1993).  This response also occurs very rapidly, reaching  maximum induction 24-48 hours after exposure. Determination ofCYP1A1-mediated enzyme activities is relatively inexpensive and easily measured in the laboratory. Finally, CYPIA1 induction has significant toxicological relevance. In knowing that CYP 1 Al activities have increased, then activation of the Ah receptor-mediated process has occurred. Since TCDD and related chemicals exert most, if not all, of their toxic effects by initially binding to the Ah receptor, then it is possible that other toxicities may also occur. The excellent relationships observed between levels of TCDD and related chemicals and hepatic microsomal EROD activities in great blue heron hatchlings (Beliward et a!. 1990; Sanderson et a!.  1994a) and double-crested cormorants (Sanderson et aL  1994b)  demonstrated the use of hepatic EROD induction in wild piscivorous birds as a biomarker of the extent of exposure to TCDD and related chemicals. The measured EROD activities in wild hatchlings accurately followed temporal changes in environmental concentrations of TCDD and related chemicals in accordance with a typical dose-response relationship. Immunoblots using monoclonal mouse antibodies against rat CYP1A1 recognized a protein in hepatic microsomal preparations from herons and cormorants (Sanderson and Bellward 1995). Intensity of the stained band on the immunoblot increased with increasing EROD activity, indicating that ethoxyresorufin is a suitable substrate for avian CYP I Al mediated enzyme activities.  -  In addition, hepatic Ah receptor was detected in herons and  cormorants (Sanderson and Beliward 1995). In all, these studies show the validity of using 15  hepatic EROD activity in fish-eating birds as a sensitive and specific biomarker of exposure to TCDD and related chemicals (Beliward et aL 1990; Sanderson et aL 1994a,b; Sanderson and Beliward 1995). Although CYP1A1 induction shows that the Ah receptor-mediated process has been activated, elevated enzyme activities associated with CYP 1 Al are not necessarily responses with deleterious consequences for the organism (Beresford 1993).  In general, enzymes  associated with the cytochromes P450 are so-called “phase I” enzymes which metabolize lipophilic endogenous and exogenous chemicals to more water soluble compounds able to be excreted or conjugated by phase II enzymes prior to excretion. However, certain cytochrome P450-mediated reactions can lead to metabolites which are more toxic than the parent compound.  For example, CYP1A1 is responsible for metabolizing benzo(a)pyrene to its  ultimate carcinogenic form (Gelboin 1980). Therefore, sustained induction of this isozyme by exposure to TCDD and related chemicals may increase the risk of developing cancer in individuals exposed to the procarcinogen. Although TCDD and related chemicals are potent inducers of CYP1A1, they are not readily oxidized by CYP1A1-associated enzymes (Beresford 1993). Since CYP1A1 induction in fish-eating birds and other wild species is not a directly toxic effect per Se, it is referred to as a biomarker of exposure to TCDD and related chemicals. it is of importance to develop sensitive and specific biomarkers that can indicate exposure and toxic effect in response to TCDD. Such biomarkers have the potential to add greater toxicological relevance to biomonitoring studies. In addition, determination of hepatic EROD and AHH activities requires sacrifice of experimental animals.  In environmental  monitoring studies it would be beneficial to have less invasive biomarkers which would not necessarily cause death of animals, but still provide adequate information regarding the extent of exposure and toxicity.  A major goal of this thesis is to investigate a number of  toxicologically relevant endpoints as potential biomarkers of TCDD exposure and effect in avian species. 16  1.C.2.  Thyroid Hormones as Biomarkers of TCDD Exposure and Effect in Avian  Species  TCDD and related HAHs have been reported to affect thyroid function in laboratory animals and avian species. TCDD has been shown to decrease serum thyroxine (T ) levels, 4 increase biliary excretion of thyroxine glucuronide, increase serum TSH levels, increase 13 lj uptake by the thyroid gland, and increase thyroid gland weight in rats (Bastomsky 1977a,b). Herring gull populations in the lower Great Lakes have displayed evidence of thyroid dysfunction related to HAH exposure. Histopathological examination of thyroid glands from adult birds revealed epithelial hyperplasia, increased mass, and decreases in follicular diameter, epithelial area, and colloidal vacuolization when compared to relatively uncontaminated colonies (Moccia et a!. 1986). Spear and Moon (1985) observed decreases in serum T 4 and triiodothyronine (T ) in adult ring doves (Streptopetia risoria) injected intraperitoneally with 3 3,3 ‘,4,4’-tetrachlorobiphenyl, a TCDD-like PCB congener. A significant negative correlation between plasma free T 4 levels and PCB concentrations in yolk was reported in great cormorant (Phatacrocorax carbo) hatchlings from the Netherlands (Van den Berg et at. 1994).  A positive correlation was observed between plasma total T 4 levels and EROD  activity in common terns, also in the Netherlands (Murk et at. 1994). Plasma thyroid hormone (T , T 4 ) concentrations have been shown to increase 3 significantly in birds at hatch or soon after hatching.  In precocial species, such as the  domestic chicken, plasma T 3 levels increase five-fold on the day of hatch (Scanes et 4 and T at. 1987). In altricial ring doves, plasma T 4 and T 3 levels gradually increase during the first week after hatch (McNabb 1987), reflecting the relatively delayed developmental profile of altricial avian species. Thyroid hormones are known to be involved in perinatal growth and development in birds (King and May 1984; Scanes et aL 1987), as well as initiation and maintenance of egg production in adults (Lien and Siopes 1989).  If TCDD and related  chemicals decrease circulating T 4 and/or T 3 levels during the critical perinatal period, this 17  may be a causative factor in the reduced growth, delayed hatching, and embryo and hatchling mortality observed in wild fish-eating birds inhabiting aquatic systems contaminated with HAHs.  1.C.3. Sex Steroid Hormones and Receptors as Biomarkers of TCDD Exposure and Effect The perinatal period in higher vertebrates is critical with respect to sexual differentiation of the central nervous system (CNS), and gonadal hormone secretion has the greatest influence on sex differences (MacLusky and Naftolin 1981). The intrinsic pattern of CNS sexual development in mammals and birds exists in the homogametic sex. In mammals, the intrinsic pattern is female, and CNS differentiation of male primary and secondary sexual characteristics occurs largely by secretion of testicular androgens during embryonal development. The masculinizing action of androgens on the brain appears to be exclusively a result of their aromatization to 1 7f3-estradiol (McCarthy 1994). These early “organizational” effects of gonadal secretion result in permanent and irreversible differences in CNS function between sexes. Therefore, alterations in gonadal secretions during the critical perinatal period may have consequences that persist throughout adult life. Perinatal TCDD exposure was observed to cause a chronic alteration in the ancirogenic status of male rats that lasted through to sexual maturation (Mably et a!. 1 992a).  These  effects included decreased serum testosterone and 5ct-dihydrotestosterone levels, decreased seminal vesicle and ventral prostate weights,  delayed testicular descent,  reduced  spermatogenesis, and demasculinization and feminization of adult sexual behavior (Mably et aL 1992a,b,c).  Moreover, the observed alterations in sexually mature male reproductive  behavior occurred when essentially all TCDD had been eliminated (Mably et at. 1992b). The authors concluded that the male reproductive system in rats was the most sensitive organ system studied to date (Mably et at. 1992a,b,c).  18  As mentioned previously, TCDD and related chemicals also appear to antagonize the effects of endogenous estrogens in laboratory animals, primarily by affecting estrogen receptor levels (Romkes et at. 1987; Umbreit and Gallo 1988; DeVito et at. 1992). Since in birds the heterogametic sex is female, feminization and demasculinization of the female embryo by ovarian estrogen secretions is imperative for sexual differentiation of the avian CNS (MacLusky and Naftolin 1981; Jost 1983).  If TCDD and related chemicals have  antiestrogenic effects in birds, exposure to these compounds during the perinatal period could be responsible for certain behavioral characteristics of the reproductive dysfunction observed in wild fish-eating bird species, in a manner analogous to the studies in male rats (Mably et aL 1992a,b,c).  In addition, the sex steroid hormones are important in perinatal growth and  development. Furthermore, piscivorous birds inhabiting areas contaminated with TCDD and related chemicals are chronically exposed during embryonic development via the yolk, and during adulthood via consumption of contaminated prey items. Thus, TCDD may be altering sex steroid hormone levels in avian populations during both the early developmental or “organizational” period, and during the breeding or “activational” period. These alterations in endocrine homeostasis may be contributing to the decreased reproductive success observed in certain fish-eating bird populations.  19  1 .D. HYPOTHESES AND OBJECTIVES  1.D.1. Hypotheses  General Hypothesis: TCDD can decrease the reproductive capacity of wild fish-eating birds by causing alterations in endocrine function during the perinatal period.  Sub-hypotheses: 1. TCDD causes alterations in plasma thyroid hormone concentrations during the perinatal  period in wild fish-eating birds, and these alterations are negatively affecting perinatal growth and development. Furthermore, the changes in thyroid hormone status during the perinatal period can be used as sensitive biomarkers of exposure and toxic effect in avian biomonitoring studies.  2.  TCDD causes alterations in plasma 1 713-estradiol and testosterone concentrations, and  hepatic estrogen receptor levels during the perinatal and adult periods in wild fish-eating birds. These perturbations can be used as biomarkers of exposure and toxic effect in biomonitoring studies ofjuvenile and adult fish-eating birds.  1.D.2. Objectives 1.  To develop an early in ovo TCDD exposure technique that does not cause excessive  mortality in domestic and wild avian embryos and hatchlings, and can be used reliably to study the effects of TCDD during the perinatal period. 2. To determine effects of early in ovo TCDD exposure on plasma T , 4 3 /T ratio, 17133 T , T 4 estradiol, and testosterone concentrations, hepatic estrogen receptor levels, induction of hepatic microsomal EROD activity, overall body and skeletal growth, and edema in domestic  chicken embryos and hatchlings, domestic pigeon hatchlings, and great blue heron hatchlings. 3. To determine the effects of acute (14 day) TCDD exposure on plasma T ,T 4 ,T 3 4 / 3 T ratio, 1 7f3-estradiol, and testosterone concentrations, hepatic estrogen receptor levels, and induction of hepatic microsomal EROD activity in adult great blue herons raised in captivity.  20  2. MATERIALS AND METHODS 2.1. Experimental Birds Fertile White Leghorn chicken eggs were purchased from a local supplier and incubated on the day of delivery. Pigeon eggs were obtained from a breeding colony at the Department of Animal Science, UBC. Pigeon eggs were collected daily from Monday to Friday and stored upright at 10°C until pickup on Friday afternoon. Eggs were kept at room temperature for two to three hours prior to incubation. Cracked or otherwise damaged eggs were discarded. Great blue heron eggs were collected on April 5, 1994 from a breeding colony located near Chilliwack, BC, an area of relatively low HAH contamination. A permit was obtained from the Canadian Wildlife Service which allowed collection of wild avian eggs for scientific purposes.  A total of 33 heron eggs were collected from 8 nests between 1100-1300 hrs.  Eggs were transported to UBC in a portable incubator held at 3 0°C, and placed in the laboratory incubator by 1630 hrs.  Heron eggs were allowed to acclimate to incubation  conditions for one day before candling observations were conducted. Chicken, pigeon, and heron eggs were artificially incubated at 37.5°C and 55% relative humidity in a Humidaire (New Madison, OH) Model 21 incubator. All eggs were incubated at the Faculty of Pharmaceutical Sciences, UBC.  Eggs were placed horizontally and  automatically tipped 90° every hour. The average incubation periods for these species were 21, 18, and 28 days for chicken, pigeon, and heron, respectively. Eggs were candled every 23 days to observe any mortality.  2.2. Preparation of TCDD Stock Solutions Tritiated 2,3,7,8-tetrachlorodibenzo-p-dioxin 3 ([ H ]-TCDD; 8 tg/mL in toluene) and nonradioactive TCDD (1 mg, crystalline) were purchased from Cambridge Isotope Laboratories (Andover, MA). The specific activity of [ H]-TCDD was 40 CiJmmol, with a 3 21  radiochemical purity of 97%. A stock solution of I mg/mL of unlabelled TCDD was prepared in toluene.  Stock solutions of 3 [ H ]-TCDD used for injecting bird eggs were prepared by  adding a constant amount of radiolabelled TCDD solution to varying amounts of unlabelled TCDD solution, in 2 mL of corn oil. The corn oil solutions were vortex-mixed for 1 minute, and the toluene was evaporated under a gentle stream of nitrogen gas at 80°C (Sanderson and Bellward 1995).  Stock solutions had a [ H]-TCDD activity of 0.67 Cilmmol, and a 3  concentration of 16.7 j.tCi/mL. All other chemicals were purchased from Sigma Chemical Co. (St. Louis, MO), unless otherwise specified.  2.3. Injection procedure Preliminary experiments were conducted in chicken and pigeon eggs to establish an early in ovo injection procedure that would not result in excessive mortality in control (corn oil vehicle injected) eggs. An early in ovo exposure protocol was desirable in order to mimic environmental exposure conditions as closely as possible, and to expose avian embryos to TCDD during ontogeny of endocrine fhnction. In addition, the limited number of wild (heron) eggs collected did not allow excessive control mortality to occur.  Previous work by  Sanderson (1994) showed that injection of 100 iiL of corn oil into the air cell of chicken eggs, or 50 tL into pigeon eggs at approximately the midpoint of incubation resulted in 75-100% mortality. Therefore, the approach used here was to minimize the volume of corn oil injected into eggs at relatively early stages of development, and observe any adverse effects on hatchability.  A comparison was made between injection into either the air cell or yolk of  chicken eggs as potential routes of administration. Eggs were injected into the air cell with corn oil using the following procedure. Eggs were candled, and the border of the air cell was marked with a pencil. The egg was held upright and the blunt end was pierced into the air cell with a sterile tack.  Corn oil was  injected into the air cell with a 25 i.tL Hamilton (Reno, NV) glass syringe, using a stereotaxic frame to ensure that the inner membrane of the air cell was not punctured. Prior to piercing 22  the egg, the surface of the egg shell and all injection equipment coming into contact with the egg were sterilized with 70% ethanol, which was allowed to evaporate. The corn oil was warmed slightly on a hot plate prior to drawing into the syringe to reduce viscosity. The solution was injected slowly, and the hole was sealed with sterile bone wax (Lukens Medical Corp., Rio Rancho, NM).  The egg was rotated for approximately 30 seconds to evenly  distribute the injected corn oil. Eggs were then returned to the incubator and allowed to sit upright for 1.5 hours. Finally, eggs were placed horizontally on incubator trays. Chicken eggs were injected into the air cell with 12.5 or 25 j.iL of corn oil on embryonic day 0 (E0) or on E4.5. Control eggs were either untreated or sham-injected. Embryonic day 4.5 was chosen as a potential time of injection for future experiments in chickens since blood vessels and the embryo become clearly visible by candling at this time of development in the chicken, therefore ensuring fertility at the time of injection in experimental eggs. Effects on hatchability in this experiment are shown in Table 2.1.  Table 2.1. Hatchabilities of chicken eggs injected on embryonic day 0 or embryonic day 4.5 into the air cell with 12.5 or 25 p.L of corn oil, sham-injected, or untreated.  Embryonic Day of Injection 0 4.5  Treatment Untreated  Sham-injected  12.5 iiL  25 .tL  6/8  7/8  6/8  5/8  5/5  7/8  8/8  A separate experiment was conducted in chicken eggs injected into the yolk with 12.5 or 25 jiL of corn oil on E4.5. The yolk injection method of McLaughlin et aL (1963) was  23  used. This technique is similar to the above procedure for air cell injection, except that the egg is held horizontally and the syringe needle is inserted through the inner air cell membrane and directly into the yolk sac. All other steps in this procedure are as described previously for air cell injection. Controls were left untreated in this experiment. Effects on hatchability are shown in Table 2.2.  Table 2.2. Hatchabilities of chicken eggs injected on embryonic day 4.5 into the yolk with 12.5 or 25 1 i.L of corn oil.  Treatment  Hatchability  Untreated  6/8  12.5 L  4/8  25 iL  7/8  Pigeon eggs were injected into the air cell with 10 or 20 pL of corn oil on E3.5. Embryonic day 3.5 was chosen for injections in pigeons for similar reasons as described in the chicken. However, the pigeon has a shorter incubation period, and blood vessels are easily observed by E3 .5. Controls were either sham-injected on E3 .5, or left untreated. Effects on hatchability of pigeons in this experiment are shown in Table 2.3.  In conclusion, injection of 12.5 and 25 j.tL of corn oil into the air cell or yolk sac of chicken eggs on E4.5, or 10 and 20 tL of corn oil into the air cell of pigeon eggs on E3.5, had no adverse effect on hatchability when compared to sham-injected or untreated controls.  24  Table 2.3. Hatchabilities of pigeon eggs injected on embryonic day 3.5 into the air cell with 10 or 20 tL of corn oil, sham-injected, or untreated.  Treatment  Hatchability  Untreated  9/10  Sham-injected  10/10  lOpL  7/10  201J.L  9/10  2.4. Early In Ovo [ HJ-TCDD Distribution in Chicken Eggs 3 An experiment was conducted to compare the early in ovo distribution of [ H]-TCDD 3 to the embryo, yolk, and albumin following injection into either the air cell or yolk of fertile chicken eggs. Eggs were injected on E4.5 with 20 p.L of a 0.5 ng/jiL solution of [ H]-TCDD 3 in corn oil. Eggs were injected into either the air cell or yolk using the methods described previously. Controls were injected with the corn oil vehicle. At 48 and 96 hours following injection (i.e. E6.5 and E8.5), the egg was opened at the blunt end and the contents were emptied into a plastic weighboat. The embryo was removed, blotted dry, and placed in a small glass vial. The yolk sac was collected and placed in a glass vial. Finally, the albumin was collected using a 5 mL Gilson pipette and stored in a glass vial. Embryos, yolk sacs, and albumin samples were stored at -20°C until analysis by liquid scintillation counting (LSC). Embryo weight (mean ± SEM.) was 0.43 ± 0.03 g on E6.5, and 1.01 ± 0.05 g on E8.5. There was no embryo mortality in any of the treatment groups. Embryos were thawed, 250 iL of distilled water was added and the embryo was homogenized for approximately 20 seconds with an Omni 2000 homogenizer (DiaMed, Mississauga, ON). Approximately 200 mg of the homogenate, in duplicate, was transferred to scintillation vials. Individual yolks  25  were homogenized for 20 seconds, and approximately 250 mg aliquots, in duplicate, were added to scintillation vials. Albumin from individual eggs was homogenized for 20 seconds, and approximately 750 mg aliquots, in duplicate, were added to scintillation vials. A 10 mL volume of BioFluor scintillation cocktail (Du Pont, Mississauga, ON), a high efficiency emulsifier for aqueous samples, was added to each vial and vortex-mixed for 5-10 seconds. Vials were allowed to sit in the dark for one hour to reduce chemiluminescence. Radioactivity was measured by LSC on a Packard Tricarb Model 4530.  Background  radioactivity for each injection technique was determined in embryo, yolk, and albumin collected from corn oil injected eggs using the methods described above. Mean background cpm from each compartment (n=2) was subtracted from the total cpm. Quenching ranged from 11-15% in embryos, 19-21% in yolk, and 17-22% in albumin samples. Quenching was corrected for in each sample using the internal standard method. Concentrations of [ H]3 TCDD in each compartment are shown in Table 2.4.  2.5. Reproducibility of 1 H1-TCDD Injection 3 An experiment was conducted to determine the reproducibility of injecting small volumes of [ H]-TCDD in corn oil using a 25 pL Hamilton glass syringe. Either 10 or 20 iiL 3 of a 20 pg/iL [ H]-TCDD solution in corn oil was injected directly into scintillation vials. 3 After injection of the corn oil, the syringe needle tip was either touched to the side of the vial or removed without contact with the vial surface. A 10 mL volume of BioFluor scintillation cocktail was added, and radioactivity was measured on a Packard Tricarb 6530. Coefficients of variation for each treatment were less than 1% (n=6), indicating excellent reproducibility. There was significantly greater cpm for each volume when the needle was touched to the side of the vial (p<O.Ol); however, this difference represented only 2% of total cpm for the 10 .tL volume, and 1% of total cpm for the 20 iiL volume.  26  Table 2.4. Concentrations of [ H]-TCDD in chicken embryo, yolk, and albumin 48 and 96 3 hours following either air cell or yolk injection of 10 ng 3 of[ H ]-TCDD on embryonic day 4.5.  1 1 3 {1-TCDD Concentration (pg/g) Technique  Compartment  48 hr  96 hr  Air Cell  Embryo  40.0 ± 4.7  67.3 ± 9.8  Injection  Yolk  41.5±4.1  66.1±6.7  Albumin  29.4±3.5  48.6±3.8  Yolk  Embryo  93.5 ±6.4  100.0±6.0  Injection  Yolk  229.3 ± 77.1  226.4 ± 29.6  Albumin  97.2 ± 10.6  124.0 ± 9.4  All data are means ± SEM of 6 eggs.  From the above preliminary experiments in chickens and pigeons, an early in ovo exposure procedure was chosen. Air cell injection was chosen over yolk injection due to the less invasive nature of this exposure route. Eggs would be injected into the air cell on E4.5 (chickens) or E3.5 (pigeons) with between 10-20 .tL of corn oil or jiii TCDD in corn oil.  2.6. Dose-Response Experiments in Chickens and Pigeons Experiments were conducted using readily available chicken and pigeon eggs in order  27  to determine appropriate early in ovo doses of TCDD that would not cause excessive embryo mortality or reduced hatchability. Chicken eggs were injected into the air cell with 0.1 jig/kg or 0.3 jig/kg of TCDD, or corn oil vehicle, on E4.5.  Hatchability of chickens in this  experiment is shown in Table 2.5. Pigeon eggs were injected into the air cell with 1, 2, or 3 jig/kg of TCDD, or corn oil, on E3.5. The results of this experiment are shown in Table 2.6.  Table 2.5. Hatchability of chickens exposed in ovo to 0.1 jig/kg or 0.3 jig/kg of TCDD, or corn oil vehicle, via air cell injection on embryonic day 4.5. Mortality is the percent mortality above controls.  Treatment  Elatchability  Control  8/8  0.1 jig/kg  7/8  12.5  0.3 jig/kg  1/8  87.5  Percent Mortality  Table 2.6. Hatchability of pigeons exposed in ovo to 1, 2, or 3 jig/kg of TCDD or corn oil vehicle via air cell injection on embryonic day 3.5. Mortality is the percent mortality above controls.  Treatment  Hatchability  Control  12/16  1 jig/kg  10/16  12.5  2 jig/kg  3/8  37.5  3 jig/kg  1/16  68.8  28  Percent Mortality  From these experiments, appropriate early in ovo doses were chosen that would not cause excessive mortality in TCDD-exposed eggs (i.e. below the LD ), but that 50 would have the greatest potential to cause toxicological effects in embryos and hatchlings. The doses chosen were 0.1 tgfkg in chickens, and 1 tg/kg in pigeons.  2.7. In Ovo TCDD Exposure Experiments  Chicken Fertile chicken eggs were injected into the air cell with 0.1 p.g/kg egg of [ H]-TCDD 3 or corn oil vehicle on E4.5. Injection volumes of TCDD ranged from 10-13 iL; control eggs received 11 pL of corn oil. Eggs were artificially incubated until sacrifice on El 7 (4 days before hatch), El 9 (2 days before hatch), day of hatch (H), day 2 after hatch (D2), and day 4 after hatch (D4). These time points were chosen in order to characterize the perinatal time course of plasma thyroid and sex steroid hormone concentrations, which are known to peak at approximately the time of hatch in the precocial chicken (Scanes et al. 1987). Blood and tissues from individual chicken embryos sacrificed on E17 and E19 were collected by the following procedure. Eggs were candled and large blood vessels near the surface of the egg shell were marked with a pencil. A 25 mm 2 hole was carefully cut in the egg shell with a Dremel tool (Emerson Electric Co., Racine, WI). The shell was removed with forceps, exposing the shell membrane and underlying blood vessel. The blood vessel was pierced with a 27 gauge needle, and blood was drawn into a heparinized Natelson blood collecting tube (Fisher Scientific, Vancouver, BC). Blood was emptied into heparinized 12x 75mm glass test tubes, on ice. Tubes were centrifuged at 3000xg and 4°C for 20 minutes in an IEC Centra MP4R refrigerated centrifuge (International Equipment Co., Needham Hts., MA).  Several aliquots (50-200 pL) of plasma were collected in 1 mL Nunc CryoTubes  (Kamstrup, DK) and stored at -20°C.  In all experiments with avian hatchlings (including  pigeon and heron), blood and tissue collection was performed between 0900 and 1300 hours  29  to control for possible circadian variations in plasma hormone levels (Newcomer 1974). In addition, the photoperiod was controlled at 14 h light: 10 h dark for all experiments. After blood collection, the egg was opened at the blunt end and the embryo was emptied into a plastic weigh boat. The embryo was separated from the yolk sac, blotted dry, and weighed. The gall bladder was carefully removed using a hemostat after opening the body cavity.  The liver was removed, blotted dry, weighed, and placed in a beaker containing  approximately 20 mL of ice cold HEDGM buffer (25 mM HEPES, 1.5 mM EDTA, 1.0 mM dithiothreitol, 20 mM sodium molybdate, 10% v/v glycerol, pH 7.6) until preparation of microsomal and cytosolie fractions.  The sex of embryos was determined by visual  examination of the gonads (Greenwood 1924). Chicken hatchlings were sacrificed on H, D2, and D4. The D2 and D4 hatchlings were housed in hatching trays on the floor of incubators, using the same temperature and relative humidity as described for eggs; these birds were not fed in this experiment.  Individual  hatchlings were removed from the incubator, weighed, and euthanized with CO 2 for approximately 45 seconds. The chest cavity was opened and blood was drawn by cardiac puncture using an ice cold 1 mL heparinized glass syringe (Becton-Dickinson, Rutherford, NJ). Blood was collected in heparinized glass test tubes on ice, and plasma was prepared and stored as described previously. After blood collection, the gall bladder was removed and the liver was excised, weighed, and placed in ice cold HEDGM buffer. The yolk sac was removed and weighed. The sex of birds was determined by visual examination of the gonads. For day of hatch birds, crown-rump length, tibia length, and length of exposed culmen were measured with Vernier calipers, following the standard bird measurement methods of Baldwin eta!. (1931).  Pigeon Two experiments were conducted in pigeon hatchlings comparing effects of early (E3.5) vs. late (E14) in ovo injection of TCDD. In the early in ovo experiment, pigeons were 30  injected into the air cell with 1 p.g/kg of [ H]-TCDD or corn oil on E3.5. Injection volumes 3 of TCDD ranged from 9-15 pL; control eggs received 10 .iL of corn oil. The pigeon is an altricial species, and it was expected that increases in plasma hormone levels associated with perinatal growth and development, such as thyroid hormones, would be delayed in comparison to precocial species such as the chicken (McNabb 1987). Therefore, in these experiments pigeons were sacrificed on day of hatch (N), and fed for 7 days following hatch (D7). Pigeon hatchlings (squabs) were fed a starter diet that simulated crop milk (Yang and Vohra 1987). The diet (glkg dry ingredients) consisted of 215 g glucose, 88.7 g soybean oil, 610.6 g isolated soybean protein (92%), 40.0 g 2 •2H 4 CaHPO 0 , 17.1 g calcium-free mineral mixture, 13.0 g CaCO , and 15.6 g vitamin diet fortification mixture. All ingredients were 3 purchased from ICN Biomedicals (Aurora, OH). Squabs were fed a slurry consisting of 14% diet and 86% distilled water blended with a hand-held homogenizer. Fresh diet was prepared every 48 hours or sooner as required. The slurry was warmed to 37.5°C, and squabs were force fed into the crop using a stainless steel gavage tube attached to a 10 mL plastic syringe. In this experiment, squabs were not fed for the first 12-18 hours following hatch, then were fed 3 times daily at approximately 1000, 1600, and 2200 hours. The last feeding was on the morning of D6, 24 hours prior to sacrifice. Volumes of the slurry fed to squabs ranged from 2-8 mL per feeding, and each bird received a total volume of 84 mL. Squabs were housed in hatching trays on the floor of incubators, at the same temperature and relative humidity used for egg incubation. Pigeons were sacrificed, and blood and tissues were collected on H and D7 using the same procedures described for chicken hatchlings. Like chickens, the sex of pigeon squabs is easily determined by visual examination of the gonads. Body weight, yolk-free body weight, crown-rump length, wing length, tibia length, and length of exposed culmen (beak) were measured in H and D7 squabs (Baldwin et at. 1931). In the second experiment in pigeons, eggs were injected during the latter third part of incubation (E14) with 3 pg/kg of [ H]-TCDD. Squabs raised to D7 received an additional 3 31  feeding on the afternoon of D6, for a total volume of 92 mL of diet. All other procedures were the same as those described for the first pigeon experiment. Plasma samples were also collected from pigeon hatchlings raised naturally at the pigeon colony where experimental eggs were collected (Dept. of Animal Science, UBC). Blood was collected by cardiac puncture from one, three, and six day old squabs (n=4-5). Plasma was prepared and stored as described previously.  The sex of the naturally-raised  pigeons was not determined. These samples were used to determine plasma hormone levels in naturally-raised pigeon squabs, in order to compare with results obtained from the squabs fed an artificial diet in the two experiments.  Great Blue Heron The experimental approach used for the heron hatchlings was similar to the pigeon experiments. Great blue herons are semi-altricial, and it was expected that hormone levels would increase within the first week after hatch.  Therefore, one group of herons was  sacrificed at hatch, and the other fed for one week prior to sacrifice. Of the 33 heron eggs collected, 29 were fertile and at various developmental stages. Eggs were candled and categorized into a range of ages, based on a candling scheme developed for the great blue heron in our laboratory (Sanderson and Bellward 1995). The youngest embryos were used for TCDD treatment, since an early in ovo TCDD exposure was desired. A total of 12 of the more developed eggs were used as controls, and received 10 iL of corn oil via air cell injection. A total of 17 eggs were injected into the air cell with 2 .tgfkg of [ H]-TCDD. 3  This dose was chosen based on relative EROD induction potencies of  chickens, pigeons, and herons exposed in ovo to TCDD in a previous study performed in our laboratory (Sanderson and Bellward 1995). The volume of injected TCDD ranged from 1317 pL. Eggs were artificially incubated at 37.5°C and 55% relative humidity as described previously, and candling observations were made each day.  32  One group of control and TCDD-treated hatchlings were fed for one week prior to sacrifice. Herons were fed a diet of whole rainbow trout (Oncorhynchus mykiss) purchased from Albion Fisheries (Vancouver, BC). The trout was thawed and ground using an electric meat grinder. Fresh diet was prepared each day. Herons were fed 3-4 times daily between 1000 and 2400 hours. Body weights were measured each morning before feeding. The last feeding was on the evening of D6. Herons received a total of 240 g of food prior to sacrifice. The hatchlings were housed for the first day after hatch on the floor of incubators using egg incubation conditions.  For Dl to D7, they were housed in Curfew (Essex, UK) Model  RX200 incubators maintained at 35°C and 46% relative humidity. The procedures used for collecting blood and tissues from heron hatchlings were the same as those described previously for chickens and pigeons. Determining the sex of herons was not possible by visual examination since the gonads were in a variety of stages of development.  2.8. Preparation of Hepatic Microsomal and Cytosolic Fractions All steps of the microsomal and cytosolic preparations were performed at 4°C. Individual livers were combined with fresh HEDGM buffer at a 1:2 ratio (i.e. 33% homogenate), minced with scissors, and homogenized using a 15 mL Potter-Elvehjem homogenizing tube with Teflon pestle. The homogenization procedure included 5 passes of the pestle at a moderate speed, followed by a one minute rest on ice, followed by a further 5 passes. The homogenate was centrifuged at 10,000xg for 20 minutes in a Beckman J2-21 centrifuge, using a JA17 fixed-angle rotor.  The resulting supernatant was centrifuged at  100,000xg for 60 minutes in a Beckman L5-50 ultracentrifuage, using a Ty65 fixed-angle rotor. The supernatant from this centrifhgation is the cytosolic fraction, while the pellet is the microsomal fraction. The cytosol was collected in an ice-cold 1 mL glass syringe with a 22g needle, with care being taken to avoid collecting any of the lipid layer. Cytosol was aliquoted (0.5-1 mL) into 1 mL cryovials, immediately frozen in liquid nitrogen, and stored at -80°C. 33  The microsomal pellet was suspended in a 10 mM EDTA, 1.15% KCI, pH 7.4 buffer using a glass rod.  The pellet was homogenized with 5 passes of the pestle, and centrifuged for  another hour at 100,000xg. The resulting “washed” microsomal pellet was resuspended in 0.5 mL of 0.25 M sucrose using an ice-cold I mL syringe with a 25g needle. Microsomes were aliquoted (100 .tL) into cryovials and stored at -80°C.  2.9. Ethoxyresorufin O-Deethylase (EROD) Assay Hepatic EROD activity was determined using the direct fluorometric procedure of Burke and Mayer (1974), optimized in our laboratory by Sanderson (1994). A 2 mL reaction mixture was prepared in Heilma (Concord, ON) quartz cuvettes, consisting of 1.93 mL 2l buffer (0.1 M HEPES, 5 mM MgC1 HEPES-MgC , pH 7.8) and 10 j.tL of 1 mM 72 ethoxyresorufin dissolved in DMSO. The cuvette was warmed to 37°C for 5 minutes prior to adding microsomal preparation and NADPH. Microsomal samples were diluted in 0.25 M sucrose to give protein concentrations of 2 mg/mL for induced (TCDD-treated) and 6 mg/mL for uninduced (control) samples. A 50 jtL volume of microsomal protein was added to the cuvette. The reaction was started by adding 10 iiL of 50 mM NADPH in 2 HEPES-MgCI buffer.  The temperature was maintained at 3 7°C, and the reaction mixture was mixed  continuously with a magnetic stirbar. The increase in fluorescence due to resorufin formation was measured in a Shimadzu RF-540 spectrofluorometer with an excitation wavelength of 550 nm, an emission wavelength of 582 nm, and slit widths of 2 nm.  Reaction times were 5  minutes for umnduced samples and 2 minutes for induced samples. measured in duplicate. Standard curves were prepared daily, using 0.05  All samples were -  1.0 iiM resorufin in  DMSO in the presence of uninduced chicken microsomal protein, and in the absence of NADPH.  2.10. Protein Assay  Protein concentrations were determined in microsomal and cytosolic samples using the 34  dye-binding method of Bradford (1976). The protein assay kit was purchased from BioRad (Hercules, CA). Standard curves were prepared over a range of 10-40 g of bovine serum albumin. Absorbances from duplicate samples were measured at 595 nm using a Beckman DU-64 spectrophotometer.  2.11. EROD/Total Cytochrome P450 Comparison Experiment A preliminary experiment was conducted to compare and validate procedures used in this thesis with standard procedures used in our laboratory. It was of interest to determine whether the use of CO 2 asphyxiation or HEDGM (molybdate) homogenizing buffer in these experiments had any effect on hepatic total cytochrome P450 content or catalytic (EROD) activity, in comparison to the standard laboratory procedure (i.e. decapitation and Tris/KC1 homogenizing buffer).  The molybdate-containing buffer was used in order to stabilize  estrogen receptors in the cytosolic preparation (Dahmer et aL 1984). This experiment was conducted using untreated chicken hatchling liver. To compare the effects of buffer (TrisfKCI vs. HEDGM) on hepatic total P450 content and hepatic EROD activity, birds were decapitated and microsomes were prepared in each buffer as described previously.  The third group of hatchlings were asphyxiated with C0 , livers were 2  homogenized with HEDGM buffer, and microsomes were prepared. Protein concentrations in the microsomal samples ranged from 9.6  -  16.2 mg/mL. EROD activities were determined as  described previously. Total cytochrome P450 concentrations in the microsomal samples were determined using difference spectroscopy (Omura and Sato 1964).  Microsomes were diluted 1/10 in  sodium phosphate buffer (100 mM sodium phosphate, 0.1 mM EDTA, 20% v/v glycerol, pH 7.4) to give a 1 mL volume which was divided between the sample and reference quartz cuvettes. Initially, a baseline spectrum was recorded on a SLM Aminco DW-2C dual beam spectrophotometer. Sodium dithionite (5-10 mg) was added to each cuvette and mixed with a pasteur pipette until dissolved.  Carbon monoxide was then bubbled through the sample 35  cuvette for one minute at approximately I bubble per second. The absorbance spectrum from 325-625 nm was recorded, and cytochrome P450 concentrations were calculated from the  difference in absorbance between 450-490 nm, using a molar extinction coefficient of 91 cm 1 m M. The results of this experiment are shown in Table 2.7.  There were no significant  differences in cytochrome P450 content or EROD activity among any of the treatments.  Table 2.7. Comparison of hepatic total cytochrome P450 concentrations and hepatic EROD activities in chicken hatchlings using (i) microsomes prepared in TristKCl buffer after decapitation (TRIS), (ii) microsomes prepared in HEDGM buffer after decapitation (HEDGM), or (iii) microsomes prepared in HEDGM buffer after CO 2 asphyxiation 2 (C0 ) .  Total cytochrome P450  TRIS  HEDGM  2 CO  0.30 ± 0.02  0.30 ± 0.01  0.35 ± 0.02  213 ± 18  232 ± 12  260 ± 14  (nmollmg protein) EROD activity (pmollmin/mg  protein)  All data are means ± SEM of 5-6 birds.  2.12. Hepatic Estrogen Receptor (ER) Assay Estrogen receptor affinities (Kd) and concentrations (Bmax) were measured in hepatic cytosols using saturation analysis as described by Leake and Habib (1987), with modifications. 3 [2,4,6,7H ]-Estradiol (87 Cilmmol; 99% radiochemical purity) was purchased from Du Pont.  Stock solutions of 3 [ H ]-estradiol were prepared in absolute ethanol at  36  concentrations of 0.05 .tM, 0.5 jiM, and 2 jiM. A stock solution of 2 p.M [ H]-estradiol plus 3 200 p.M diethylstilbestrol (DES) was also prepared in absolute ethanol. Stock solutions were stored in glass vials at -20°C. A series of 10 working solutions were prepared from the stock solutions in HED buffer (25 mM HEPES, 1.5 mM EDTA, 0.25 mM dithiothreitol, pH 7.6). The [ H]-estradiol concentration range used for the assay was 0.05 3  -  12 nM. The working  solutions were stored in glass vials at 4°C for a maximum of one week.  Dextran-coated  charcoal (DCC) was prepared in HEG buffer (25 mM HEPES, 1.5 mM EDTA, 10% v/v glycerol, pH 7.6), with a final concentration of 0.5% w/v activated charcoal and 0.05% w/v Dextran T-70 (Pharmacia, Uppsala, Sweden). The DCC slurry was mixed thoroughly on a stir plate at 4°C prior to use. The assay was performed in 12x75 mm borosilicate glass test tubes, on ice. For each sample, duplicate 150 jiL aliquots of cytosol were added, diluted in HED buffer to give a protein concentration of 1 mg/mL for pigeon samples and 2 mg/mL for chicken and heron samples. A preliminary experiment was conducted to examine the effect of varying cytosolic protein concentration on specific ER binding at low, intermediate, and high [ H]-estradiol 3 concentrations. Specific binding was linear for each radioligand concentration between 1, 2, and 4 mg/mL cytosolic protein (Fig. 2.1). A 50 jiL volume of each working solution of 3 [ H ]-estradiol was added to cytosol preparations, as well as to duplicate tubes containing 150 jiL of HED buffer for total and blank counts. Non-specific binding (NSB) was determined in the presence of 100-fold molar excess of DES at the three highest [ HJ-estradiol concentrations (3, 6, and 12 nM). The 3 mean NSB of these three tubes was subtracted from the total binding for each of the 10 3 [ H ]estradiol concentrations to determine specific binding (Leake and Habib 1987). The validity of this method was tested by examining the linearity of NSB over the concentration range of radioligand used in the assay (Fig. 2.2). Tubes were vortex-mixed and incubated for 18 hours at 4°C. After incubation, 200 jiL of DCC was added to tubes to remove unbound [ H]-estradiol. After DCC addition, 3 37  60  50 0. 0) C C  40•  •  O.l6nM  •  1.25nM lOnM  30  0  .4-  0  a)0  20  C’)  10  0  0  1  2  3  4  5  Protein Concentration (mglmL)  Figure 2.1. Effect of cytosolic protein concentration on specific binding of 0.16 nM, 1.25 nM, and 10 nM [ H]-estradiol to estrogen receptor in pooled chicken hatchling hepatic 3 cyto so!.  150  0  -50  C 0  z  0  5  10  15  20  25  Free [3H]-Estradiol (nM)  Figure 2.2. Linearity of nonspecific binding between 0.16 nM and 20 nM [ H]-estradiol 3 in pooled hepatic cytosol from pigeon hatch!ings.  38  tubes were vortex-mixed and allowed to sit for 15 minutes at 4°C. The DCC was pelleted by centrifl.igation at l000xg and 4°C for 10 minutes. The DCC separation method was observed to remove between 97-99% of free radioligand at each concentration. A 200 tL aliquot was removed from each tube and added to 6 mL plastic scintillation vials (Simport Plastics, Beloeil, PQ). For total counts, 200 jiL was removed from the “buffer only” tubes prior to DCC addition. centrifligation.  For blank counts, 200 L was removed after DCC incubation and A 4 mL volume of BioFluor scintillation cocktail was added to each vial.  Radioactivity was measured on a Beckman LS 6000TA scintillation counter with a 3 [ H ] counting efficiency of approximately 60%. Specific binding at each [ H]-estradiol concentration was calculated as described 3 previously, after subtracting blank counts. estradiol concentration (free cpm  =  Specific binding was plotted against free 3 [ H ]-  total cpm  -  bound cpm) to determine saturability of  receptor binding. Specifically-bound and free cpm were converted to molar quantities, and apparent estrogen receptor affinities (Kd) and concentrations (Bmax) were determined by Scatchard (Scatchard 1949) and Woolf (Haldane 1957) analyses.  Least-squares linear  regression was performed on the Scatchard and Woolf plots to determine Kd and Bmax values for individual samples.  2.13. Radioimmunoassays  Thyroxine (T ) 4 Total T 4 concentrations were determined in unextracted plasma from avian embryos and hatchlings using an antibody-coated tube radioimmunoassay (RIA), with 1]-T 125 as [ 4 antigen. The RIA kits for total T 4 were purchased from ICN Biomedicals (Costa Mesa, CA). Literature values for total T 4 in precocial chicken and altricial ring dove (Streptopelia risoria) embryos and hatchlings (McNabb 1987), as well as preliminary results in chicken hatchlings, indicated that plasma concentrations of T 4 were at the lower end of the T 4 standard curve (5  39  -  15 ng/mL). Therefore, modifications were made to the standard RIA protocol in order to improve the sensitivity of the assay. The T 4 standards (10 .tL, in triplicate) and avian plasma samples (50 iiL, in duplicate) were added to polypropylene antibody-coated tubes, and I mL of phosphate gelatin buffer, pH 7.0 (assay buffer; ICN Biomedicals) was added.  An additional 40 j.tL of the  hormone-free (0 ng/mL) standard was added to standards to make equivalent assay volumes. Tubes were vortex-mixed and preincubated in the absence of 1]-T 125 at 4°C for 24 hours. [ 4 After preincubation, 0.5 mL of the 4 1]-T 1 [ 25 tracer was added, and tubes were vortex-mixed and incubated for a further 24 hours at 4°C.  After incubation, tubes were aspirated and  antibody-bound tracer was counted for one minute in a Packard Cobra gamma counter. The standard curve was curve fitted using a weighted logit/log algorithm.  Trilodothyronine (T ) 3 Total T 3 concentrations were determined in unextracted plasma using a solid phase 1]-T 1 [ 3 25 RIA kit (ICN Biomedicals), modified similarly to the total T 4 RIA.  The T 3  standards (100 .tL, in triplicate) and avian plasma samples (200 pL) were added to assay tubes, and 1 mL of assay buffer was added. An additional 100 tL of buffer was added to standard tubes to make equivalent assay volumes. Due to limited plasma collection, samples from chickens and pigeons were not assayed in duplicate. Heron hatchling samples were run in duplicate. All remaining steps were identical to those described for the total T 4 RIA.  1 7/3-Estradiol 1713-Estradiol concentrations were determined in unextracted plasma using a double antibody [ 1]-estradiol 1 25 RIA kit (ICN Biomedicals), with modifications.  The estradiol  standards (50 .tL, in triplicate) and avian plasma samples (100 p.L for chickens, 200 .tL for pigeons and herons) were added to 12x75 mm glass test tubes. Standards were brought up to 40  an equivalent volume as samples with assay buffer. For chicken samples, only female plasma was assayed, and samples were run in duplicate whenever possible. Both female and male pigeon samples were assayed, although plasma availability did not allow duplicates to be run. The heron samples collected at H were not run in duplicate; the D7 heron samples were run in duplicate. The estradiol antibody solution was diluted 1/2 with assay buffer, and 0.5 mL was added to each tube, except the nonspecific binding tubes which contained buffer only. Tubes were vortex-mixed and preincubated at 4°C for 24 hours. After preincubation, 0.5 mL of H]-estradiol tracer was added, tubes were vortex-mixed, and incubated for a further 24 3 [ hours at 4°C.  25 1]-estradiol 1 After the second incubation, 0.5 mL of antibody-bound [  precipitating solution was added, tubes were vortex-mixed, and centrifuged at l000xg and 4° C for 30 minutes in a Beckman J-6B centrifiuge. After centrifugation, the supernatant was 25 was counted. The standard curve 1]-estradiol 1 aspirated, and remaining antibody-bound [ was curve fitted using a weighted logit/log algorithm.  Testosterone  Plasma levels of testosterone were below the quantitation limit for RIA using unextracted plasma in chickens, pigeons, and herons.  In order to measure testosterone,  plasma samples were extracted into diethyl ether and concentrated 3-5 fold using the following procedure (McMaster et a!. 1992). Plasma aliquots were added to 16x 150 mm borosilicate glass test tubes (200-250 p.L for chicken males, 250-400 tL for pigeon males, 400-500 pL for herons). The volume was brought up to 1 mL with RIA assay buffer. A 5 mL volume of diethyl ether was added, and tubes were vortex-mixed vigorously for 60 seconds. Tubes were allowed to stand for 10 minutes to achieve separation of the upper organic layer and lower aqueous layer. The aqueous layer was “snap-frozen” by immersing the tube into liquid nitrogen for 20 seconds. Tubes were removed and warmed in the hand for 10 seconds, gently mixed, and refrozen for 20 seconds. The organic phase was then poured 41  into 13x 100 mm test tubes, and evaporated at 50°C under a gentle stream of nitrogen. A further 5 mL of ether was added to the aqueous phase and extracted again. After combining extracts and evaporating to dryness, the tube walls were rinsed with 700 tL of ether and evaporated.  Depending on the original volume of plasma added, extracted steroids were  reconstituted in 50-100 tL of RIA assay buffer in order to achieve a concentration factor of 3-5. Extracts were frozen at -20°C until assayed. Extraction efficiency of the procedure was determined using a C]-testosterone 14 (Du Pont) spiking solution, and ranged from 95-97%. [ Plasma testosterone concentrations were measured in the extracted plasma using a double antibody 1]-testosterone 125 RIA kit (ICN Biomedicals). For this assay, the standard [ kit procedure was used. A 50 p.L volume of standards and samples was added to 12x75 mm test tubes. Sex steroid binding globulin inhibitor solution (100 pL) was added to each tube and mixed gently for 10 seconds. The [1 25 1]-testosterone tracer was added (0.5 mL) to each tube. With the exception of the NSB tubes, 0.5 mL of anti-testosterone antibody was added to each tube and vortex-mixed. Tubes were incubated in a water bath at 37°C for 2 hours. After incubation, 100 pL of the precipitating antiserum (second antibody) was added, vortexmixed, and incubated for one hour at 3 7°C. Tubes were centrifuged at l000xg and 4°C for 30 minutes. The supernatant was aspirated and remaining antibody-bound 1]-testosterone 125 [ was counted.  Quality Control Several procedures were performed to determine the performance characteristics of each RIA. Intra- and interassay variability was measured using avian plasma when possible, or plasma controls purchased from ICN Biomedicals (Table 2.8). The commercially available controls were routinely assayed alongside samples in each RIA. The lower detection limit of each assay was determined as two standard deviations of the binding observed with the hormone-free standard. The midrange value (ED ) is defined as the hormone concentration 50 at which B/B 0 is 50% (Table 2.8). 42  Table 2.8. Performance characteristics of plasma hormone RIAs.  Total T 4  Total T 3  Estradiol  Testosterone  (ng/mL)  (nglmL)  (pg/mL)  (pglmL)e  Intraassay  78.9 (6)  2.81 (6)  56.4 (5)  720 (6)  variabilitya  CV. 3.7%  C.V. 4.6%  C.V. 8.1%  C.V. 4.4%  Interassay  79.1 (12)  2.83 (12)  55.3 (10)  695 (12)  variabiityb  CV. 5.4%  C.V. 5.7%  C.V. 10.5%  C.V. 7.1%  Detection limitc  0.43 (3)  0.0023 (3)  1.12 (3)  1.89 (3)  Midrange valued  40.2 (3)  1.36 (3)  19.94 (3)  594 (3)  All data are means (n in parentheses). a Coefficient of variation (C.V.; SD/mean) of a pooled chicken plasma sample assayed n times in one assay. b C.V. of a plasma sample assayed in two separate assays. C  The sensitivity is defined as 2 SD of the binding observed with the hormone-free (zero)  standard. d Defined as the hormone concentration at which B/B 0 is 50% e Determined using commercially available control sera of human origin.  Serial dilutions of avian plasma were assayed and compared to the commercial (human origin) hormone standards supplied with each RIA kit. Plasma samples were diluted with the hormone-free standard.  Parallelism between dilutions of adult heron plasma and  hormone standards was demonstrated for each MA (Table 2.9).  43  Table 2.9. Parallelism between dilutions of avian plasma and standard hormone solutions for , 3 , total T 4 each hormone RIA. Adult great blue heron plasma dilutions were used for total T and estradiol assays. Eastern meadowlark plasma was used for the testosterone assay.  Total T 3 (nglmL)  4 (nglmL) Total T Dilution  Observed  Expected  Dilution  Observed  Expected  6.61  188 1:1.25  151  150  1:1.33  4.85  4.96  1:1.67  110  113  1:2  3.26  3.31  1:2.5  66  75  1:4  1.67  1.65  Testosterone (pglmL)  Estradiol (pg/mL) Dilution  Observed  Expected  Dilution  Observed  Expected  5.24  203 1:1.33  146  152  1:2  2.82  2.62  1:2  96  102  1:4  1.54  1.31  1:4  49  51  1:8  0.82  0.66  2.14. [ 111-TCDD Concentrations in Liver and Yolk 3 H]-TCDD were measured in the liver and yolk of chickens, 3 Concentrations of [ H]-TCDD-treated chicken hatchlings (H), 3 pigeons and herons by LSC. Liver and yolk from [ day 3 after hatch (D3) pigeons, and H and D7 herons were collected, stored in glass vials, and frozen at -20°C until analysis.  44  Approximately 50 mg of liver was combined with 250 p.L of acetone and 250 iiL of H]3 distilled water and homogenized with an Omni 2000 homogenizer for 20-30 seconds. [ TCDD was extracted into I mL of hexane by vortexing for 60 seconds.  Tubes were  centrifhged at l000xg for 5 minutes, and the hexane phase was collected. The extraction procedure was repeated again, and the hexane phases were pooled. A 800 p.L volume, in duplicate, was added to scintillation vials, 10 mL of BioFluor was added, and radioactivity was counted on a Beckman LS 6000TA scintillation counter. H]3 Efficiency of the liver extraction procedure was determined by adding 10 pL of [ TCDD dissolved in hexane directly to untreated 50 mg liver samples in glass test tubes, and allowing to evaporate for one hour. Samples were frozen overnight at -20°C, and extracted the next day by the same procedure. Efficiencies ranged from 8 1-84%. Quenching in the extracted liver samples ranged from 12-22% in all species, and was corrected in each sample using the internal standard method. Yolk samples were homogenized for 20-3 0 seconds, and approximately 75 mg of yolk, in duplicate, was added directly to scintillation vials. A 15 mL volume of BioFluor was added, and radioactivity was counted. Quenching ranged from 14-33% in chickens, 6-25% in pigeons, and 5-24% in herons, and was corrected for in each sample using the internal standard method.  2.15. Adult Great Blue Herons Herons were raised from hatchlings since 1990 by the Department of Animal Science, UBC, with the intention of creating a successful breeding colony. Breeding was unsuccessful in the 13 herons after up to 3-4 years, and unfortunately it was not economically feasible to continue maintaining the heron colony. The herons were not permitted to be released back to their natural setting because survival was unlikely in these captive birds. In order to collect as much information as possible in the adult herons within the existing monetary constraints, an acute (14 day) TCDD exposure experiment was conducted in February 1994. 45  Seven herons (2 females, 5 males) were located at the UBC San Rafael Research Aviary near White Rock, BC.  These birds served as controls, and were injected  intraperitoneally (i.p.) with 1 mL of corn oil. Six herons (2 females, 4 males) were located in an outdoor aviary at UBC. These birds were injected i.p. with 20 pg/kg of 3 [ H ]-TCDD in corn oil. Herons were fed a diet of whole herring (Clupea harengus) and rainbow trout for the 2 weeks prior to sacrifice. On the day of sacrifice, herons were weighed, and blood was drawn from wing veins using a 10 mL syringe and 21 gauge needle and collected in heparinized glass test tubes, on ice. Birds were decapitated and trunk blood was collected in glass test tubes, on ice. The body cavity was opened, and the gall bladder was tied off with surgical thread. The liver was perfused through the portal vein with ice-cold Tris/KCI buffer (0.05 M Tris, 1.15% KCL, pH 7.5) in a 10 mL syringe with a 21 gauge needle.  The liver was removed, weighed, and  approximately 5 g was placed in a beaker containing 20 mL of ice-cold Tris/KC1 buffer for microsomal preparation as described previously.  Other liver samples were collected,  immediately frozen in liquid nitrogen, and stored at -80°C for future analyses, including estrogen receptor levels and 3 [ H ]-TCDD concentrations.  Other organs and tissues were  collected and stored at -80°C (kidney, gonads, thyroid glands, spleen, heart, intestine, breast muscle, fat, and skin). Hepatic EROD activities, and plasma total T , 17f3-estradiol, and 3 , total T 4 testosterone concentrations were determined as described previously.  Hepatic estrogen  receptor levels were determined as described previously, except that frozen liver samples were used.  Liver was thawed, cytosols were prepared, and the estrogen receptor assay was  performed on the same day to avoid repeat freeze-thawing of livers. Concentrations of [ H]-TCDD in liver, kidney, fat, and breast muscle were measured 3 using the same extraction procedure described for hatchlings. Quenching ranged from 2-7% in liver and kidney, 15-27% in fat, and 13-22% in muscle, and was corrected in each sample using the internal standard method. Extraction efficiency for breast muscle tissue was 84%. 46  Plasma samples were collected from wing veins of the adult herons on a monthly basis between September 1991 and April 1992 by the Department of Animal Science, UBC, and stored at -20°C. These samples were made available in order to measure seasonal trends , T 3 , estradiol, testosterone) in the adult herons. In addition, 4 in plasma hormone levels (T plasma samples were collected in May 1993 from the herons. Concentrations of plasma total , total T 4 T , estradiol, and testosterone were determined as described previously. 3  2.16. Statistical Analyses Statistical comparisons between control and TCDD-treated groups were made using one way analysis of variance (ANOVA), with a significance level (ce) of 0.05.  Multiple  comparisons were made using the Newman-Keuls test. Hatchability of control and TCDD treated birds was compared using chi-square analysis. Homogeneity of the hatchabilities of different batches of pigeon eggs was tested using heterogeneity chi-square contingency tests (Zar 1984). All statistical tests were performed using the Minitab statistical program (Minitab Inc., State College, PA).  47  3. RESULTS 3.A. AVJAN EMBRYOS AND HATCHLTNGS 3.A.1. Fertility and Hatchability Fertility of chicken eggs, determined on E4.5, was 93.7%. Hatchability of control eggs was 93.6% (44/47), and hatchability of eggs injected with 0.1 p.g/kg of TCDD was 90.8% (69/76). There was no significant difference in the hatchability of control and TCDD treated chicken eggs. Overall fertility of pigeon eggs in the two experiments, determined on E3 .5, was 79.9%. Hatchabilities of control and TCDD-treated pigeon eggs in the early in ovo TCDD  exposure experiment (1 pg/kg of TCDD injected into the air cell on E3.5) were 73.7% (70/95) and 57.9% (66/114), respectively.  There was significantly reduced hatchability in  TCDD-treated pigeons in this experiment (p<O.O5). The time period during incubation in which mortality occurred, based on candling observations, was compared between control and TCDD-treated groups (Table 3.1). Overall, there was greater mortality in the TCDD-exposed pigeons during all time periods (Table 3.1). There was significantly greater embryo mortality in TCDD-exposed pigeons between E14 and pipping (p<O.O5). A similar trend was seen for the period between injection (E3.5) and E9, however this result was not statistically significant. Mortality between E9-E14 was not tested statistically due to the lack of mortalities in control embryos (Table 3.1). In the late in ovo TCDD exposure experiment in pigeons (3 rig/kg of TCDD injected into the air cell on E 14), hatchabilities of control and TCDD-treated eggs were 79.4% (5 0/63) and 70.5% (5 5/78), respectively. Hatchabilities of control and TCDD-exposed pigeons were not significantly different in this experiment. The hatchabilities of different batches of pigeon eggs received each week for these experiments were tested for homogeneity. Seven batches of eggs were used for the early 48  Table 3.1. Developmental periods during which mortality was observed to occur in pigeons exposed to 1 .tg/kg of TCDD via air cell injection on embryonic day 3.5 (E3.5). Mortality was observed by candling during four developmental periods: (i) between day of injection (E3 .5) and approximately E9 (range E7-E1O), (ii) between E9 and E14, (iii) between E14 and time of pipping the shell (-‘E18), and (iv) between pipping and hatching from the shell.  Mortality Developmental Period  Control  TCDD  p valuea  E3.5-E9  4/77  16/104  O.OS<p<O.lO  E9 E14  0/72  4/87  Not tested  E14-Pipping  5/71  16/82  O.Ol<p<O.O5  Pipping Hatch  7/84  11/66  0. 1O<p<O.25  -  -  a Differences between control and TCDD-exposed pigeons were compared using chi-square analysis.  exposure experiment, and four batches were used for the late exposure experiment. Heterogeneity chi-square contingency tests indicated that the hatchabilities of eggs from each experiment were not significantly different. Fertility of the heron eggs was 29/33 (87.9%).  After the herons hatched, it was  possible to determine the approximate embryonic day of injection for control and TCDD exposed birds.  Control eggs were injected between E13-E19 (mean E16), and TCDD  exposed eggs were injected between E9-E18 (mean E13). This represents approximately the midpoint of the 28 day incubation period.  Hatchabilities of control and TCDD-exposed  herons were 100% (12/12) and 82.4% (14/17), respectively.  49  3.A.2. Body Growth Chicken Body weights of chicken embryos and hatchlings are shown in Table 3.2. There was no significant effect of TCDD exposure on body weight at any of the time points. In addition, there was no effect of TCDD exposure on crown-rump length, tibia length, or culmen length measured in chickens sacrificed on the day of hatch (Table 3.3).  Subcutaneous edema, a  common response to TCDD exposure in avian species, was not observed at any time point in this experiment.  One of the TCDD-treated embryos sacrificed on embryonic day 19 was  anencephalic.  Table 3.2. Body weights of chicken embryos (E17, E19) and hatchlings (H, D2, D4) exposed to 0.1 .tg/kg of TCDD or corn oil vehicle via air cell injection on embryonic day 4.5.  Body Weight (g) n  Control  n  0.1 rig/kg TCDD  E17  26  18.5 ± 0.3  27  19.3 ± 0.3  E19  17  25.8±0.9  15  25.7±0.5  Hatch  23  41.5±0.5  23  41.9±0.5  Day2  10  36.0±0.3  12  35.6±0.7  Day4  18  31.3±0.6  18  32.1±0.7  Day  All data are means ± SEM.  50  Table 3.3.  Crown-rump length, tibia length, and length of exposed culmen in chicken  hatchlings exposed to 0.1 jig/kg of TCDD via air cell injection on embryonic day 4.5.  Body Measurement  Control  0.1 jig/kg TCDD  Crown-rump length (mm)  95.2 ± 0.5  94.8 ± 0.4  Tibia length (mm)  25.6 ± 0.3  25.2 ± 0.2  Culmenlength(mm)  16.5±0.1  16.6±0.1  All data are means ± SEM of 23 birds.  Pigeon Body measurements in pigeons from the early in ovo TCDD exposure experiment are shown in Table 3.4. There was one mortality among control hatchlings and 2 mortalities in TCDD-exposed pigeons fed to day 7, At hatch, there were significant decreases in crownrump, tibia, wing, and culmen lengths in pigeons exposed to TCDD (p<O.Ol). On day 7, yolkfree body weight was decreased, liver to body weight ratio was elevated, and crown-rump, tibia, wing, and culmen lengths were significantly decreased in TCDD-exposed birds (p<0.Ol). There was no subcutaneous edema observed in pigeons at hatch or day 7. In the late in ovo TCDD exposure experiment in pigeons, there were 5 mortalities in control birds, and 11 mortalities in TCDD-treated birds fed to day 7. Yolk-free body weight was significantly lower in TCDD-exposed birds on day 7 (Table 3.5). Liver to body weight ratio was elevated in comparison to controls at hatch and day 7 (p<O.Ol). In TCDD-exposed pigeons, crown-rump length and culmen length were decreased at hatch and day 7, and tibia and wing lengths were decreased on day 7 (p<O.Ol). Subcutaneous edema was not observed in pigeons at either time point.  51  Table 3.4. Body measurements of pigeons exposed to 1 j.ig/kg of TCDD via air cell injection on embryonic day 3.5.  Hatch Parameter  Day 7  Control  TCDD  Control  TCDD  (n=16)  (n=22)  (n13)  (n17)  14.6 ± 0.3  14.2 ± 0.3  27.4 ± 0.5  25.7 ± 0.3’  3.3 ±0.8  3.5 ±0.7  2.4 ±0.1  3.0 ±0.03a  Crown-rump length (mm)  78.5 ± 0.6  75.2 ± 0.7a  93.8 ± 0.7  89.5 ± o.8a  Tibialength(mm)  16.1 ±0.2  15.2±0.2a  23.4±0.2  .O±O.2a 22  Wing length (mm)  31.7 ± 0.3  29.8 ± 0.4a  43.8 ± 0.7  41.5 ±  Culmen length (mm)  10.7 ± 0.1  9.7 ± 0.2k  15.4 ± 0.1  14.3 ± 0.2a  Yolk-free body weight (g) Liver weight (%)  o.5’  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.Ol).  Body weights (including yolk) measured in naturally-raised captive pigeon squabs were greater than body weights of pigeons fed the artificial diet in the two experiments. Body weights (mean ± SEM; n=4-5) were 25.0 ± 1.8 g, 46.7 ± 4.9 g, and 127.8 ± 14.2 g in naturally-raised squabs sacrificed on day 1, day 3, and day 6, respectively.  Great Blue Heron In herons, there was one mortality in control hatchlings and three mortalities in TCDD-exposed hatchlings fed to day 7.  There were no differences between control and  TCDD-treated herons in body weights recorded daily from hatch to day 7 (Fig. 3.1). Crown rump length, tibia length, and wing length were decreased in TCDD-exposed herons at hatch  52  Table 3.5. Body measurements of pigeons exposed to 3 p.g/kg of TCDD via air cell injection on embryonic day 14.  Day 7  Hatch Parameter  Yolk-freebodyweight(g)  Control  TCDD  Control  TCDD  (n=17)  (n21)  (n25)  (n=22)  13.8 ±0.5  14.8±0.3  34.1 ±0.5  a 9 31.1 ±O.  3.1±0.1  3.5±0.1’  2.1±0.1  .5±O.ia 2  Crown-rump length (mm)  77.0 ± 0.8  74.9 ± 0.5a  99.7 ± 0.7  a 95.8 ± 10  Tibia length (mm)  15.3 ± 0.3  14.8 ± 0.2  23.9 ± 0.3  22.4 ± 0.4’  Wing length (mm)  30.7 ± 0.4  30.1 ± 0.4  46.8 ± 0.4  44.5 ± 0.6a  Culmen length (mm)  10.3 ± 0.1  15.8 ± 0.1  14.7 ± 0.2’  Liverweight(%)  9.8 ± 0.2  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.0l).  (p<O.0S), and culmen length was reduced on day 7 (p<O.Ol; Table 3.6). Subcutaneous edema was observed on the thighs and neck of one of the TCDD-exposed herons on day of hatch.  3.A.3. EROD Activity Hepatic EROD activity in chicken embryos and hatchlings exposed to TCDD was induced 13, 15, 34, and 43-fold above control activities on E19, day of hatch, day 2, and day 4, respectively (p<O.OO1; Fig. 3.2).  Hepatic EROD activities in TCDD-exposed chickens  were 2.2, 6.5, 13.2, and 18.2 nmol/min/mg microsomal protein on E19, day of hatch, day 2, and day 4, respectively. Control (basal) EROD activities were significantly lower  53  (p<O.O5) on  200  150  -  CI) 100 >  0 50  0-  I  I  I  I  H  Dl  D2  D3  D4  D5  Day After Hatch  I  I  D6  D7  Figure 3.1. Body weights of great blue heron hatchlings exposed in ovo to 2 p.g/kg of TCDD (•) or corn oil vehicle (.) via air cell injection at approximately the midpoint of incubation. Herons were fed a diet of ground whole rainbow trout. Body weights were measured before feeding each morning from day of hatch to day 7. Values are means ± SEM of 5 birds.  54  Table 3.6. Body measurements of great blue heron hatchlings exposed to 2 rig/kg of TCDD via air cell injection at approximately the midpoint of incubation.  Hatch Parameter  Yolk-free body weight (g) Liverweight(%)  Day 7  Control  TCDD  Control  TCDD  (n=6)  (n=6)  (n=5)  (n=5)  48.2 ± 1.3  50.1 ± 1.8  146.9 ± 5.3  146.9 ± 3.4  2.5±0.1  2.6±0.1  4.7±0.2  5.6±0.2  Crown-rump length (mm)  135.1 ± 0.9  131.1 ±0.9a  210.8 ± 5.2  219.6 ± 0.3  Tibia length (mm)  27.5 ± 0.4  26.4 ± 0.3a  44.8 ± 1.1  44.1 ± 0.8  Wing length (mm)  53.6 ± 0.4  51.9 ± O.6  89.3 ± 2.2  86.0 ± 1.4  Culmenlength(mm)  14.1 ±0.3  13.6±0.2  25.3 ±0.5  b 3 23.1 ±o.  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.O5). b Significantly different from control by ANOVA (p<O.OI).  E19 (0.17 nmol/min/mg microsomal protein; p<O.O5) in comparison to hatch, day 2, and day 4 (0.43, 0.39, and 0.43 nmol/min/mg microsomal protein, respectively). In the early in ovo TCDD exposure experiment in pigeons, hepatic EROD was induced 15- and 6-fold above control activities at hatch and day 7, respectively (p<O.OO ; Fig. 1 3.3). Hepatic EROD activities in TCDD-exposed pigeons were 2.2 nmol/min/mg microsomal protein at hatch and 1.5 nmol/min/mg microsomal protein on day 7. Control EROD activities approximately doubled between hatch (0.15 nmol/min/mg microsomal protein) and day 7 (0.27 nmol/min/mg microsomal protein; p<O.O5).  55  In the late in ovo TCDD exposure experiment in pigeons, hepatic EROD was induced 14- and 10-fold above control activities at hatch and day 7, respectively (p<O.OO1; Fig. 3.4). Hepatic EROD activities in TCDD-exposed pigeons were 2.5 nmol/min/mg microsomal protein at hatch and 4.1 nmollminlmg microsomal protein on day 7. Similarly, control EROD activities in pigeons roughly doubled between hatch (0.17 nmollmin/mg microsomal protein) and day 7 (0.42 nmol/min/mg microsomal protein; p<O.OOl). In herons, hepatic EROD was induced 3-fold above controls at hatch (p<O.05) and 2fold on day 7 (p<O.O1; Fig. 3.5). Hepatic EROD activities in TCDD-exposed herons were 0.73 nmol/min/mg microsomal protein at hatch and 0.17 nmollmin/mg microsomal protein on day 7. In contrast to control EROD activities in pigeons, control EROD activity on day 7 in herons (0.09 nmol/min/mg microsomal protein) was approximately half of the control EROD activity at hatch (0.22 nmol/min/mg microsomal protein; p=O.O7).  3.A.4. Plasma Thyroid Hormone Concentrations Plasma total T 3 concentration and 4 /T ratio increased approximately five-fold 3 T between E19 and hatch in chickens (Fig. 3.6). Plasma total T /T ratio 3 T , total T 4 , and 4 3 were not affected by TCDD exposure in chickens at any time point (Fig. 3.6). In the early TCDD exposure experiment in pigeons, there was no effect of TCDD treatment on plasma total T 4 and T 3 concentrations, or on plasma 4 /T ratio (Fig. 3.7). In 3 T the late TCDD exposure experiment in pigeons, plasma total T 3 concentration was significantly lower in hatchlings (p<0.O5), but not at day 7 (Fig. 3.8). There was no effect of TCDD exposure on plasma total T 4 or plasma 4 /T ratio in this experiment (Fig. 3.8). 3 T Plasma thyroid hormone concentrations in naturally-raised pigeon hatchlings are shown in Figure 3.9. Plasma total T 4 and T 3 concentrations in the naturally-raised day 1 pigeons were similar to day of hatch pigeons fed the artificial diet in the two experiments (Figs. 3.7-3.9). Total T 4 and T 3 levels were approximately 2-3 fold higher in the naturally  56  25 I  0 C)  015  E  C  U Q10  w  C-) -  Cci  0  F  E19  1  I  H  I  D2  D4  Figure 3.2. Hepatic EROD induction in chicken em ryos’19) and hatchlings (H, D2, D4) exposed in ovo to 0.1 pg/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 4.5. Values are means ± SEM of 5-8 birds.  57  0.  E .E 3 E 0  E  C  U  0 w  C.)  2  1  4-  0.  =0  HATCH  DAY7  Figure 3.3. Hepatic EROD induction in pigeon hatchlings exposed in ovo to 1 .tg/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 3.5. Values are means ± SEM of 8-9 birds.  0. 0)  HATCH  DAY7  Figure 3.4. Hepatic EROD induction in pigeon hatchlings exposed in ovo to 3 rig/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 14. Values are means ± SEM of 8-11 birds.  58  1.2 0)10  0.8 0  E U  0  0.4  w  C)  c0.2 0  x  0.0  HATCH  DAY7  Figure 3.5. Hepatic EROD induction in great blue heron hatchlings exposed in ovo to 2.ig/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection at approximately the midpoint of incubation. Values are means ± SEM of 5-6 birds. Significantly different from control using one way ANOVA: a, p<O.O5; b, p<O.Ol.  59  30  E  0)  20 I cUl5 0  110 Cu  E  a.  -J  0 6.0  E  5.0  4.0  I  3.o 0  CU  2.0  E ol.0 CU  0_0.0  -  0.5 0 4-  CU  0.4  1—• 0.3 Co  I— CU 0.2  E 0  CS0.1 0 0.0-  I  I  I  I  I  E17  E19  H DAY  D2  04  Figure 3.6. Plasma total T , and 4 3 /T ratio in chicken embryos (E17, E19) and 3 T , total T 4 hatchlings (H, D2, D4) exposed in ovo to 0.1 rig/kg of TCDD () or corn oil vehicle (.) via air cell injection on embryonic day 4.5. Values are means ± SEM of 7-13 birds.  60  20 -J  E  15  I. 10 0  F  E  0  5  0 0  HATCH  DAY7  2.5  -J  E  2.0  -  0) t  C) 1.5 I—  m  0  i— 1.0  E  0  a  0.5  0.0  -  HATCH  DAY7  0.25  0.20  0.15  0.10  0.05  0.00  HATCH  DAY 7  Figure 3.7. Plasma total T , total T 4 , and 4 3 /T ratio in pigeon hatchlings exposed in ovo to 3 T 1 tgfkg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 3.5. Values are means ± SEM of 10 birds. 61  20 -J  E  -  15  I 10 40 I  E05 a0  HATCH  DAY7  HATC H  DAY7  HATCH  DAY 7  3.0 -J  E  0)  £2.0 C)  I  4-  0 I  c 1.0  E 0  c  0  0.0 0.4  0.3  rr  ‘1• I— C’) 0.2  I  E 0  0.1  0.0  Figure 3.8. Plasma total T , total T 4 , and 4 3 /T ratio in pigeon hatchlings exposed in ovo to 3 T  3 pg/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 14. Values are means ± SEM of 10 birds. a, significantly different from control by ANOVA (p<O.05). 62  50  E  40  0)  30 0  I 20  E 0  10  0  DAY1  DAY3  DAY6  DAY1  DAY3  DAY6  DAY1  DAY3  DAY6  5.0 -J  E  4.0  0) C  3.0 4-  0 I— 2.0  E  (I)  a-  1.0  0.0 0.3  0 0.2 I—  c)  I 0.1  0.0  Figure 3.9. Plasma total T , total T 4 , and 4 3 1T ratio in naturally-raised pigeon hatchlings 3 T sacrificed on day I (n=1-3), day 3 (n=4-5), and day 6 (n=4) after hatch. Values are means ± SEM.  63  60  250 E  .S40  -  I. 0  I— c 20  E 0  -  io 0 6.0  HATCH  DAY7  HATCH  DAY7  25.0  E  c) 3.0 4-.  0  I— c 2.0  E (0  i.o 0.0 0.3  —  0  (U  a  0.2  I— c) I— (U  E CU  0.1  0  0.0  —  HATCH  DAY 7  Figure 3.10. Plasma total T , total T 4 , and 4 3 /T ratio in great blue heron hatchlings 3 T exposed in ovo to 2 p.g/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection at approximately the midpoint of incubation. Values are means ± SEM of 5-6 birds. a, significantly different from control by ANOVA (p<O.O5).  64  raised day 6 pigeons in comparison to day 7 pigeons fed the artificial diet. Plasma 4 /T 3 T ratios were similar in naturally-fed and artificially-fed pigeons (Figs. 3.7-3.9). In herons, plasma T 4 concentration increased 3-fold between hatch and day 7 (Fig. 3.10). Plasma total T 4 and T 3 concentrations were not affected by TCDD exposure in herons at hatch or day 7.  Plasma 4 /T ratio was elevated in TCDD-treated herons on day 7 3 T  (p<O.05), but not at hatch (Fig. 3.10).  3.A.5. Plasma Sex Steroid Hormone Concentrations 1 7/3-Estradiol Plasma 1 7f3-estradiol concentrations in female chicken embryos and hatchlings increased gradually between El 9 and day 4 (Fig. 3.11). Plasma estradiol was significantly lower in TCDD-exposed chicken hatchlings on day 4 (p<O.05), but at no other time point (Fig. 3.11). Plasma estradiol concentrations of 10-20 pg/mL were observed in female and male pigeons sacrificed at hatch and day 7. In the early TCDD exposure experiment in pigeons, plasma estradiol concentrations were not affected by TCDD treatment at hatch or day 7 in female or male hatchlings (Fig. 3.12).  In the late TCDD exposure experiment, plasma  estradiol was significantly decreased at hatch and elevated on day 7 (p<O.05) in TCDD-treated female pigeon hatchlings (Fig. 3.13).  There was no effect of TCDD exposure on plasma  estradiol concentrations in male pigeons in this experiment (Fig. 3.13). Plasma estradiol concentrations in the naturally-raised pigeon hatchlings (Fig. 3.14; sex unknown) were similar to plasma estradiol levels observed in pigeons raised on the artificial diet (Figs. 3.12 and 3.13). However, plasma estradiol concentrations approximately doubled between day 1 and day 6 in the naturally-raised pigeons, while in artificially-fed pigeons estradiol levels were similar or decreased slightly between hatch and day 7 (Figs. 3.12-3.14).  65  Plasma estradiol levels increased 2.5-fold between hatch and day 7 in great blue heron hatchlings (Fig. 3.15; female and male combined). There was no effect of TCDD exposure on plasma estradiol concentrations measured at hatch or day 7 in herons (Fig. 3.15).  Testosterone Plasma testosterone concentrations in male chicken embryos and hatchlings increased sharply between day 2 and day 4 (Fig. 3.16). TCDD exposure had no effect on testosterone levels in this experiment. Several of the chicken samples collected on El 7, El 9, hatch, and day 2 were below the quantitation limit for testosterone.  The plasma extracts were  concentrated 3-4 times in order to increase the concentration of testosterone in the samples. As a consequence, there were large amounts of lipid in the plasma extracts which may have interfered with reconstitution of the testosterone into buffer following ether extraction. The majority of pigeon samples were below the quantitation limit for testosterone. Again, there was a large amount of lipid in plasma extracts which may have interfered with reconstitution of the testosterone into buffer. Plasma testosterone concentrations of approximately 5-10 pg/mL were measured in male and female herons (Fig. 3.17).  There was no effect of TCDD exposure on plasma  testosterone concentrations at hatch or day 7 in herons.  3.A.6. Hepatic Estrogen Receptor (ER) Levels  H1-estradiol was observed in 3 Saturable, high affinity binding between 0.05-3 nM [ hepatic cytosols prepared from all three species of avian hatchling. A representative plot of HJ-estradiol specific binding in heron hatchlings, and corresponding 3 the saturability of [ Scatchard and Woolf plots, are shown in Figure 3.18.  The binding achieved at the two  H]-estradiol concentrations (6 and 12 nM) were not used for data analysis in 3 highest [ chickens, pigeons, and herons since receptor occupancy was already saturated at 1-3 nM, and ][ H because the data at these higher concentrations were highly variable. The number of 3 66  50  E  cci  a  •1-’  0  20-  cci  E  0  10 cc  0  I  I  I  I  I  E17  E19  H  D2  D4  DAY  Figure 3.11. Plasma 1713-estradiol concentrations in female chicken embryos (E17, E19) and hatchlings (H, D2, D4) exposed in ovo to 0.1 p.g/kg of TCDD () or corn oil vehicle (.) via air cell injection on embryonic day 4.5. Values are means ± SEM of 6-8 birds, a, significantly different from control by ANOVA (p<O.05).  67  25 -J  220  0) 0  315 -D Cu (0  w 2  (0  Cu5 0 0  HATCH  DAY7  HATCH  DAY7  25  MALE -J  E20  -D Cu 4-  (0  w  Cu  2 0  Cu  a-  5-  0—  Figure 3.12. Plasma 1713-estradiol concentrations in female (n=8) and male (n=6) pigeon hatchlings exposed in ovo to 1 pjg/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 3.5. Values are means ± SEM.  68  30  FEMALE -J  E  0) 0  20  0  a  a  I 4.’  Cl)  w  10 as  E C’)  as  0—  HATCH  DAY7  HATCH  DAY7  20 -J  E 0) 0  •-..  15  0  asio 4-  Cl)  w  as  E  (‘5 0 0  Figure 3.13. Plasma 1713-estradiol concentrations in female (n7-8) and male (n=6) pigeon hatchlings exposed in ovo to 3 .ig/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection on embryonic day 14. Values are means ± SEM. a, significantly different from control by ANOVA (p<O.O5).  69  35  -J  30  E  25  .220 I-’  015  w as  Eio 0  as  I  O_5  0  DAY1  DAY3  DAY6  Figure 3.14. Plasma 1713-estradiol concentrations in naturally-raised pigeon hatchlings (sex unknown) sacrificed on day 1 (n=3), day 3 (n=4), and day 6 (n=3) after hatch. Values are means ± SEM.  70  60  _b%5Q  -J  E  -40 0 -t  c30 Co  w  C20  E  Co  010  0  HATCH  DAY7  Figure 3.15. Plasma 17f3-estradiol concentrations in great blue heron hatchlings exposed in ovo to 2 gig/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection at approximately the midpoint of incubation. Values are means ± SEM of 5-6 birds.  71  -J  25  E  2O  ci 015  ci) ()  0 .I-’10 C,)  ci)  F  E  0 C  E17  E19  H  D2  D4  DAY Figure 3.16. Plasma testosterone concentrations in male chicken embryos (E17, E19) and hatchlings (H, D2, D4) exposed in ovo to 0.1 i.g/kg of TCDD () or corn oil vehicle (.) via air cell injection on embryonic day 4.5. Values are means ± SEM of 6-8 birds.  72  20 -J  E  C  0  0)  4-’  0 0 4-’ 0  ci)  F  0  0  0  HATCH  DAY7  Figure 3.17. Plasma testosterone concentrations in great blue heron hatchlings exposed in  ovo to 2 tg/kg of TCDD (solid bars) or corn oil vehicle (open bars) via air cell injection at approximately the midpoint of incubation. Values are means ± SEM of 5-6 birds.  73  C  0.08 0  L  U)  c)  -g  0.03  0  Free [3HJ-Estradiol (nM) 30  b 0 20 C D 0  a)  a,  Kd—0.14 nM  LL 10  U  Bmax—133 fmollmg  0.0  -—  0.5  1.0  1.5  2.0  2.5  3.0  3.5  Free [3HJ-Estradiol (nM)  a) a) U C  0  0.05  0.10  0.15  Bound [3H1-Estradiol (nM)  Figure 3.18. Representative plots of a) saturability of [ H]-estradiol specific binding, b) 3 Woolf analysis, and c) Scatchard analysis in hepatic cytosol prepared from a great blue heron hatchling.  74  estradiol concentrations used for data analysis was 8 for chickens and herons (0.05-3 nM), and 6-8 for pigeons. Overall, apparent ER affinities (Kd) ranged from 0.03  -  1.25 nM, and  apparent ER concentrations (Bmax) ranged from 15-250 finollmg cytosolic protein in the three avian species. In female chicken embryos and hatchlings, Kd ranged from 0.16-0.46 nM and Bm ranged from 16-50 fmol/mg protein (Table 3.7). Kd was significantly elevated in TCDD exposed E19 chickens (p<O.Ol) using Scatchard plots. Although the Woolf plot data showed a similar trend for El 9 chickens, the difference was not statistically significant (Table 3.7). Kd was decreased in TCDD-exposed day 2 chickens, as indicated by Woolf plots only (p<O.O5). There was no effect of TCDD treatment on Bmax values at any time point (Table 3.7). In the two experiments in female pigeons, Kd values differed between hatch and day 7, ranging from 0.09-0.26 nM at hatch, and from 0.69-1.9 nM on day 7. Bm ranged from 27-65 fmollmg protein at hatch, and from 25-249 finollmg protein on day 7. There was no effect of TCDD treatment on hepatic ER affinities or concentrations in the early in ovo experiment in pigeons (Table 3.8). In the late in ovo TCDD exposure experiment in pigeons, Bm was elevated in TCDD-treated birds at hatch (p<O.OO1), but no effect was seen at day 7 (Table 3.9). Kd values were not influenced by TCDD exposure in this experiment. In TCDD-exposed herons, Kd was significantly elevated at hatch using both Scatchard and Woolf analyses (p<O.O5; Table 3.10). Bmax was elevated in TCDD-treated hatchlings at hatch using Woolf analysis (p<O.O5). A similar but not statistically significant trend was observed in heron hatchlings using Scatchard analysis (Table 3.10). There was no effect of TCDD exposure on Kd or Bmax in day 7 herons.  75  Table 3.7. Hepatic estrogen receptor affinities (Kd) and concentrations (Bmax) in female chicken embryos and hatchlings exposed to 0.1 .i.g/kg of TCDD via air cell injection on embryonic day 4.5.  Woolf  —  Kd  Bm  Scatchard Kd  Bmax  Day  Dose  n  (nM)  (fmollmg)  n  (nM)  (fmollmg)  E19  Control  5  0.16±0.02  50±5  5  0.17±0.01  50±4  0.1 p.g/kg  5  0.22±0.03  45±7  5  ±o. 02 s 2 o. b  47±6  Control  5  0.18±0.04  17±3  5  0.16±0.01  16±3  0.1 rig/kg  5  0.22±0.03  18±2  5  0.19±0.02  18±2  Control  6  0.46±0.08  19±3  6  0.23±0.04  15±2  0,1 pg/kg  6  a 6 ±O.O 23 O.  19±2  6  0.19±0.02  19±2  Control  8  0.22±0.05  20±3  8  0.16±0.01  19±4  0.1 3.Lg/kg  7  0.20±0.04  17±2  7  0.16±0.03  17±2  Hatch  Day 2  Day4  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.OS). b Significantly different from control by ANOVA (p<O.Ol).  76  Table 3.8. Hepatic estrogen receptor affinities (Kd) and concentrations (Bmax) in female  pigeon hatchlings exposed to 1 jig/kg of TCDD via air cell injection on embryonic day 3.5.  Woolf  —  Kd  Bmax  Scatchard Kd  Bmax  Day  Dose  n  (nM)  (tholJmg)  n  (nM)  (finollmg)  Hatch  Control  8  0.09±0.02  29±4  8  0.09±0.02  29±4  1 jig/kg  8  0.26±0.09  30±7  8  0.22±0.10  27±8  Control  8  1.25±0.42  65±22  8  0.83±0.32  53±22  1 jig/kg  5  0.79±0.20  25±4  5  0.75±0.20  25±4  Day 7  All data are means ± SEM.  Table 3.9. Hepatic estrogen receptor affinities (Kd) and concentrations (Bmax) in female pigeon hatchlings exposed to 3 jig/kg of TCDD via air cell injection on embryonic day 14.  Woolf  —  Kd  Bmax  Scatchard Kd  Bmax  Day  Dose  n  (nM)  (fmol/mg)  n  (nM)  (fmollmg)  Hatch  Control  8  0.10±0.02  33±4  8  0.14±0.02  36±4  3 jig/kg  8  0.17±0.03  a 4 ± 63  8  0.16±0.01  65±5a  Control  9  0.69±0.09  78±20  9  0.97±0.17  94±25  3 jig/kg  1  0.79±0.14  98±10  9  1.90±0.66  249±110  Day 7  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.OOl).  77  Table 3.10. Hepatic estrogen receptor affinities (Kd) and concentrations (Bm) in great blue  heron hatchlings exposed to 2 p.g/kg of TCDD via air cell injection at approximately the midpoint of incubation.  Woolf  Scatchard  Kd  Bmax  Kd  Bmax  Day  Dose  n  (nM)  (fmollmg)  n  (nM)  (fmollmg)  Hatch  Control  6  0.06±0.02  38±2  6  0.03±0.001  37±3  2.tg/kg  6  0.13±0.02a  53±6a  6  ±O.Oia 4 O.0  50±6  Control  5  0.08±0.01  33±5  5  0.03±0.001  31±4  2 jig/kg  5  0.09±0.05  28±3  5  0.04±0.007  27±3  Day7  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.O5).  3.A.7. 1 H1-TCDD Concentrations in Liver and Yolk 3 Concentrations of [ H]-TCDD were determined using liquid scintillation counting in 3 the liver and yolk of day of hatch chickens, day 3 pigeons, and hatch and day 7 herons. Approximately 5-10% of the total injected dose was found in the liver of each species at the different time points (Table 3.11). Day of hatch chickens and herons accumulated 20.3% and 24.2% of the injected dose into the yolk, respectively. In day 3 pigeons, 10.6% of the dose was found in yolk.  The [ H]-TCDD concentration in heron liver decreased over 10-fold 3  between hatch and day 7. The percentage of the total [ H]-TCDD dose decreased 2-fold 3 between hatch and day 7 in herons (Table 3.11).  78  Table 3.11. Concentrations of [ H]-TCDD in the liver and yolk of chicken, pigeon, and great 3 blue heron hatchlings. Birds were exposed to [ H1-TCDD via air cell injection on embryonic 3 day 4.5 in chickens, embryonic day 3.5 in pigeons, and between embryonic days 9-18 in herons.  H1-TCDD 3 1  11]-TCDD 3 [  Percent of  Dose  Concentration  Injected  Species  (ng/g egg)  Tissue  ii  (ng/g wet wt.)  Dose  Chicken  0.1  Liver  8  0.50 ± 0.02  8.2 ± 0.4  (Hatch)  0.1  Yolk  8  0.21±0.01  20.3±0.2  Pigeon  1  Liver  6  3.01±0.35  6.0±0.7  (Day3)  1  Yolk  6  5.14±0.71  10.6± 1.8  Heron  2  Liver  6  11.32 ±0.84  9.9± 1.2  (Hatch)  2  Yolk  5  6.62±0.49  24.2±4.3  Herona  2  Liver  5  0.82±0.11  4.9 ±0.7  (Day_7) All data are means ± SEM. a Yolk was completely resorbed by day 7 in herons.  79  3.B. ADULT GREAT BLUE HERONS 3.B.1. Acute TCDD Exposure Experiment H]-TCDD via i.p. 3 Adult herons were exposed to a single dose of 20 pg/kg of [ injection and sacrificed after 14 days.  Hepatic EROD activity was induced 6-fold above  controls in TCDD-exposed herons (Table 3.12).  4 concentration was Plasma total T  (p<O.05).  significantly elevated in TCDD-treated herons  There was no effect of TCDD  exposure on plasma total T 3 concentration or 4 /T ratio (Table 3.12). 3 T  Table 3.12. Hepatic EROD activities, plasma total T 4 and T 3 concentrations, and plasma /T ratios in adult great blue herons after 14 day exposure to 20 pg/kg of TCDD or corn 3 T 4 oil vehicle via i.p. injection.  Treatment Control  n  20 pg/kg  7  96 ± 10  6  536 ± 45a  4 (ng/mL) Plasma total T  7  39 ± 4  6  b 55 ± 5  Plasma total T 3 (ng/mL)  7  3.7 ± 0.3  6  3.9 ± 0.4  ratio 4 P1asmaT/T  7  0.104±0.016  6  0.070±0.004  Parameter Hepatic EROD activity (pmol/minlmg protein)  All data are means ± SEM. a Significantly different from control by ANOVA (p<O.OOl). b Significantly different from control by ANOVA (p<O.O5).  80  Effects of TCDD exposure on plasma sex steroid hormone concentrations and estrogen receptor levels are shown in Table 3.13. Plasma estradiol concentration was similar among female and male herons. There was no effect of TCDD treatment on plasma estradiol levels in male herons, or in males and females combined. Plasma testosterone concentrations were highly variable in the herons. Hepatic ER affinities and concentrations were similar between female and male herons. There was no effect of TCDD exposure on ER levels in this experiment (Table 3.13).  Table 3.13. Plasma 1 7f3-estradiol and testosterone concentrations, and estrogen receptor affinities (Kd) and concentrations (Bmax) in female and male adult great blue herons after 14 day exposure to 20 p.g/kg of TCDD via i.p. injection.  Treatment Parameter Plasma testosterone (pg/mL)  Plasma estradiol (pglmL)  Hepatic ER affinity (nM)  Hepatic ER concentration (fmollmg protein)  Sex  n  Control  n  20 p.g/kg  Female  2  9.8 ± 0.8  2  45.7 ± 5.2  Male  5  13.8 ± 4.3  4  141 ±75  Female  2  37.8 ± 12.5  2  59.9 ± 1.6  Male  5  43.8±3.1  4  38.9±3.0  Female  2  0.15 ± 0.04  2  0.13 ± 0.01  Male  4  0.15 ±0.02  4  0.24 ±0.03  Female  2  38 ± 5  2  33 ± 2  Male  4  38 ±5  4  36 ± 8  All data are means ± SEM.  81  H]-TCDD were determined in the liver, kidney, breast muscle, 3 Concentrations of [ and fat of adult herons (Table 3.14).  The concentration of [ H]-TCDD in fat was 3  approximately 10-fold higher than that found in liver, kidney, and breast muscle.  Table 3.14. Concentrations of [ H]-TCDD in liver, kidney, muscle, and fat tissues of adult 3 great blue herons after 14 day exposure to a single i.p. dose of 20 fig/kg of TCDD.  Hj-TCDD 3 [ Concentration  Percent of  n  (ng/g wet wt.)  Injected Dose  Liver  6  3.55±0.38  1.9±0.2  Kidney  5  5.69±2.31  1.1±0.4  Breast muscle  6  7.74 ± 2.93  Fat  6  62.64 ±6.66  Tissue  —  —  All data are means ± SEM.  3.B.2. Seasonal Plasma Hormone Concentrations Thyroid and sex hormone concentrations were measured in adult heron plasma samples collected monthly between September 1991 and April 1992, and in May 1993. Plasma total T 3 concentrations and 4 /T ratios appeared to peak during the winter months 3 T (November-March; Fig. 3.19). Plasma 1 7f3-estradiol concentrations in female and male adult herons were similar (Fig. 3.20). Plasma testosterone concentrations in male herons appeared to peak in March (Fig. 3.21), corresponding to the time of breeding in this species.  82  100 -J  E  80  C  60 0  j_-  40  E  0  c 0  20  0 6.0  C  CV) I 3.0 0 I c 2.  E  U)  01. 0.  Sept  Oct  Nov  Dec  Jan  Feb  Mar  Apr” May  0.15  0  0  0.10  I— CV)  I  j  0.05  0.00  Figure 3.19. Total T 4 and total T 3 concentrations, and 4 /T ratio in adult great blue heron 3 T  plasma samples collected monthly between September 1991 and April 1992, and in May 1993. Values are means ± SEM of 10-13 birds. 83  250 -J  E 200 0) Q.  o  150  -D Co  w  E  0  50  0  Sept  Oct  Nov  Dec  Jan  Feb  Mar  Apr  “  May  150 -J  E  0) 0  100  0  -D  Cu  4-  CO  w 50 Cu  E Cd)  a0  Figure 3.20. 17f3-Estradiol concentrations in female (n=3-4) and male (n=7-9) adult great  blue heron plasma samples collected monthly between September 1991 and April 1992, and in May 1993. Values are means ± SEM.  84  600  FEMALE  -J  E  —..  0) 0.  500  400 0 a, Co 0 Co  I— 200 CU  E 0 CU  100  0  0  Apr  May  800 -J  E  MALE  700  0) ..S600 a, C 0 a,  -  -  0 0 Co  a,  FCU  200  Co CU  100  E  a-  0  -  -  Oct  Nov  Dec  Jan  Feb  Mar  Apr  May  Testosterone concentrations in female (n=2-4) and male (n=6-9) adult great blue heron plasma samples collected monthly between October 1991 and April 1992, and in May 1993. Values are means ± SEM.  Figure 3.21.  85  4. DISCUSSION 4.A. AVIAN EMBRYOS AND HATCHLINGS The induction of monooxygenase activities associated with CYP1AI, particularly hepatic EROD activity, is one of the most sensitive and specific biomarkers of exposure to TCDD and related HAHs. Although CYP1A1 induction itself cannot be considered a toxic effect, there is a clear molecular basis for the in vivo and in vitro correlations between CYP1A1 induction and a variety of toxicities associated with PCDDs, PCDFs, and PCBs in laboratory animals. Since the major toxic effects of these compounds are mediated initially by binding to the AhR, induction of hepatic EROD activity indicates that the AhR has been activated, resulting in the potential for toxic effects to occur. Although strong correlations between TCDD exposure, CYPIA1 induction, and toxic effects have been shown in laboratory animals (Mason et al. 1985, 1986), it largely remains to be seen whether these relationships exist in wild animals exposed under environmental conditions.  This thesis is  concerned with examining the relationships between CYP1A1 induction and potentially adverse effects on endocrine homeostasis and growth during the perinatal period in avian species.  4.A.1. Hepatic EROD Induction in Avian Embryos and Hatchlings 4.A. 1.1. Chicken Embryos and Hatchlings Hepatic EROD activities were induced 13- to 43-fold above control activities in chicken embryos and hatchlings (Fig. 3.2).  Detailed dose-response curves for EROD  induction in chicken hatchlings exposed in ovo to TCDD suggest that 30-fold induction of EROD activity is an approximately maximal response in this species (Sanderson and Bellward 1995). Therefore, the level of induction observed in this experiment represents between halfmaximal (E19, hatch) and maximal (day 2, day 4) EROD induction in the chicken. 86  This  significant degree of EROD induction indicates that the AhR has been activated, and that other AhR-mediated biological and toxicological responses may occur, depending on their respective dose-response curves. The hepatic EROD activities (expressed in nmollmin/mg microsomal protein) observed in chickens exposed to TCDD in this experiment (Fig. 3.2) are in agreement with those reported by Sanderson and Bellward (1995). Hepatic EROD induction measured in day of hatch chickens by Sanderson and Beliward using the same in ovo TCDD dose (0.1 p.g/kg via air cell injection) was similar to the present study (both approximately 7 nmol/minlmg protein), even though eggs were injected with TCDD at different times during incubation (E4.5 in this study vs. E16 in Sanderson and Bellward (1995)). This similarity reflects the rapid distribution of TCDD and subsequent biological persistence in target organs such as the liver, causing rapid and prolonged induction of CYP 1 Al. However, EROD induction in day 2 and day 4 chickens in this study was greater than the EROD induction observed at the highest dose of TCDD (3 gig/kg) used by Sanderson and Bellward (1995), a dose that would cause acute lethality if exposure occurred early in incubation.  The increasing hepatic EROD induction observed between E19 and day 4 in  chickens (Fig. 3.2) may be a result of mobilization of yolk and elevated exposure to TCDD as the yolk was resorbed. Since EROD induction is a rapidly elicited response, changes in liver TCDD concentrations due to yolk resorption would be reflected within the time course of this experiment. Control (constitutive) EROD activities in posthatch chickens (hatch, day 2, day 4) were 2-3 fold higher (p<O.O5) than in E19 embryos (Fig. 3.2). The ontogeny of the AhR in chicken embryo liver reveals a peak in AhR levels on E20 (day before hatch), and corresponding increase in control AHH activity between E20 and day 2 after hatch (Denison  et aL 1986a). Therefore, the increase in hepatic EROD induction observed between E19 and day 4 may also be due to an increased ability of chickens to respond to TCDD exposure postnatally. 87  E5 to E9, In addition, AhR levels are highest in chicken embryo liver between . Although no suggesting possible endogenous function(s) for the AhR (Denison et al. 1986a) ered, endogenous ligand or physiological fhnction for the AhR has been discov  there is  n 1982; speculation of AhR involvement in cell growth and differentiation (Poland and Knutso has been Whitlock 1990; Okey et al. 1994). Recently, an AhR-deficient mouse line reduction in constructed and characterized. The AhR-deficient mice displayed a 50 percent , suggesting liver size and decreased accumulation of lymphocytes in lymph nodes and spleen alguero involvement of the AhR in development of the liver and immune system (Fernandez-S etal. 1995). Although there were no significant differences in embryo mortality or hatchability that a between control and TCDD-exposed chickens, a preliminary experiment demonstrated ls. This is in threefold higher dose (0.3 .tg/kg) caused 87.5 percent mortality above contro 50 for TCDD in chicken embryos of 0.25 rig/kg (Mired and agreement with the published LD o mortality in Strange 1977), and illustrates the extremely steep dose-response curve for embry observed in this species when exposed to TCDD early in incubation. Similar effects have been cked pheasant other avian species exposed early in ovo to TCDD, such as the ring-ne EROD was (Phasianus coichicus) (Nosek et a!. 1993). These authors concluded that hepatic o mortality was the most sensitive biochemical response to TCDD exposure, and that embry the most sensitive sign of toxicity (Nosek et a!. 1993). in ovo Together with the EROD induction data, these results indicate that the early al toxic effects of exposure protocol used in this experiment is appropriate for studying subleth avoids the most TCDD in chicken embryos and hatchlings. Exposing eggs to TCDD on E4.5 established by E4 sensitive period for teratogenicity, since most of the organ primordia are Furthermore, ontogeny of endocrine function occurs between (Romanoff 1960). approximately E5 to ElO in the chicken (Thommes 1987).  88  4.A. 1.2. Pigeon Hatchlings In the early in ovo TCDD exposure experiment in pigeons, hepatic EROD was induced 15- and 6-fold above controls at hatch and day 7, respectively (Fig. 3.3). This corresponds to approximately half-maximal EROD induction at hatch, and 20 percent of maximal EROD induction on day 7, based on published TCDD dose-response curves for EROD induction in day of hatch pigeons which were approximately 30-fold above controls (Sanderson and Bellward 1995). However, this difference in fold induction above control is due in part to an approximate doubling of control EROD activity between these two time points (p<O.O5). Actual EROD activities in TCDD-exposed pigeons were similar between hatch (2.2 nmol/minlmg protein) and day 7 (1.5 nmol/min/mg protein). Furthermore, comparing these data (Fig. 3.3) to absolute EROD activities from the dose-response study in pigeons (Sanderson and Bellward 1995) suggests that EROD was at least half-maximally induced at each time point in the present study. There was significantly reduced hatchability in TCDD-exposed pigeons in the early exposure experiment (p<O.O5). Similar to chickens, a three-fold higher dose of TCDD (3  t  g/kg) resulted in greatly increased mortality in preliminary experiments in pigeons (Table 2.6). The data from Table 2.6 suggest that the LD 50 in this species is between 2-3 p.g/kg, approximately an order of magnitude greater than in the chicken. The order of magnitude difference in sensitivity to mortality parallels the approximate ten-fold difference in EROD induction potency by TCDD in this species (Sanderson and Bellward 1995).  In all, the  approximately half-maximal EROD induction and increased mortality in TCDD-exposed pigeons (16 percent above controls) demonstrates that the AhR has been activated, and signs of overt toxicity are occurring in hatchlings. In the late in ovo experiment in pigeons, hepatic EROD was induced 14- and 10-fold above controls at hatch and day 7, respectively (Fig. 3.4). The EROD activities in control and TCDD-exposed pigeons at hatch were similar between the early and late exposure experiments in pigeons, even though the dose was three-fold higher in the late exposure 89  experiment (Figs. 3.3 and 3.4). In contrast, TCDD-induced EROD activities were three-fold greater on day 7 in the late exposure experiment, indicating higher TCDD exposure in this experiment.  Similar to the early exposure experiment, control EROD activities doubled  between hatch and day 7  (p<O.Ol).  The late exposure experimental protocol used in this study was identical to that used by Sanderson and Bellward (1995). Therefore, direct comparisons can be made between the two studies.  The EROD activities at hatch (2.5 nmollmin/mg protein) and day 7 (4.1  nmollmin/mg protein) correspond to near-maximal induction, based on EROD activities reported by Sanderson and Beliward (1995).  However, these authors reported 30-fold  induction of EROD activities in pigeons at the highest dose used (100 .tgIkg), in comparison to the 15-fold induction observed in day of hatch pigeons in this study (3 .tgIkg).  Thus,  hepatic EROD activity was induced between half-maximally and maximally in this experiment. Hatchability was similar between control and TCDD-exposed pigeons in the late exposure experiment. The TCDD dose used in this experiment (3 igIkg) would cause high embryo mortality if exposure occurred early in incubation (Table 2.6).  The rationale for  conducting a late in ovo TCDD exposure experiment was to examine whether a higher dose would result in a greater occurrence of toxicities related to endocrine function, since it would avoid the teratogenicity and embryo mortality associated with early in ovo TCDD exposure. Based on the EROD activity in TCDD-exposed pigeons, it appears that an appropriate dose was used for examining sublethal effects in this experiment.  4.A. 1.3. Great Blue Heron Hatchlings Hepatic EROD activity was induced three- and two-fold above control herons at hatch and day 7, respectively (Fig. 3.5). In contrast to chickens and pigeons, control activities were somewhat decreased between hatch and day 7, although this result was not statistically significant (pO.O7).  Hepatic EROD activities in TCDD-exposed herons (mnol/min/mg  protein) decreased four-fold between hatch and day 7 (Fig. 3.5). 90  The reduced EROD  induction reflects the I 4-fold decrease in liver TCDD concentration occurring between hatch and day 7 (Table 3.11). This low level of EROD induction is indicative of the conservative approach used in selecting an appropriate in ovo dose (2 ig/kg) for the herons. The limited number of heron eggs collected did not permit excessive mortality to occur in embryos and hatchlings exposed to TCDD. The observed EROD induction corresponds to the beginning of the steep portion of the TCDD dose-response curve for EROD induction reported by Sanderson and Beliward (1995). Importantly, EROD activities in TCDD-exposed herons in this thesis were similar to the EROD induction reported in heron hatchlings exposed environmentally to TCDD and related chemicals in the Strait of Georgia (Sanderson et a!. 1 994a). As mentioned previously, the environmentally-exposed heron hatchlings displayed TCDD-related toxicities such as edema and decreased growth (Bellward et a!. 1990; Hart et a!. 1991; Sanderson et aL 1 994a). Levels of TCDD and related chemicals (expressed as the sum of TCDD toxic equivalents) in great blue heron and double-crested cormorant eggs collected from colonies in the Strait of Georgia between 1986 and 1992 ranged from approximately 0.1 to 0.7 ng/g (Elliott et a!. 1989; Bellward et a!. 1990; Sanderson et at. 1994a,b). In comparison, doublecrested cormorant eggs collected from Lake Ontario in 1991 had 1.6 nglg of total TCDD equivalents (Sanderson et a!. 1 994b). The nominal dose used in the present study in herons (2 ng TCDD/g egg) is slightly higher than these environmental levels; however, the actual bioavailable dose may be lower. In the preliminary experiment in chicken embryos comparing air cell and yolk injection methods (Table 2.4), only 16 and 26 percent of the total TCDD dose was accounted for in embryo, yolk, and albumin, 48 and 96 hours following injection. It was assumed that a considerable portion of the TCDD injected into the air cell was absorbed to the shell membrane and thus not bioavailable. If so, then the actual bioavailable dose of TCDD in herons was less than 2 nglg, resulting in a level of exposure similar to the environmental levels found in wild avian eggs from the Strait of Georgia. Thus, the present  91  study incorporates an exposure protocol which is relevant to examining the effects of enviromnental levels of TCDD and related chemicals in great blue heron hatchlings.  4.A.2. 3 [ H 1-TCDD Concentrations in Liver and Yolk of Avian Hatchlings Approximately 5-10 percent of the total injected dose of [ H]-TCDD was found in the 3 liver of day of hatch chickens, day 3 pigeons, and hatch and day 7 herons (Table 3.11). These results are similar to those reported by Sanderson and Bellward (1995), who exposed avian hatchlings to TCDD via air cell injection during the latter third part of incubation.  This  similarity in percentage of injected dose found in the liver between early and late in ovo exposure is a reflection of the rapid distribution and equilibration of TCDD in target organs such as the liver. The percentage of 3 [ H ]-TCDD in the yolk of chickens (20%) and herons (24%) at hatch was similar, reflecting the large accumulation of TCDD in the lipid-rich yolk of avian hatchlings (Table 3.11). The percentage of injected 3 [ H ]-TCDD dose was lower in the yolk of day 3 pigeons (11 percent), since a considerable portion of the yolk would have been resorbed in the three days following hatch. Although the yolk was completely resorbed by day 7 in herons, the liver wet weight concentration of 3 [ H ]-TCDD was 14-fold lower at this time point in comparison to hatch (Table 3.11).  The percentage of injected 3 [ H 1-TCDD dose remaining in the liver also  decreased by half between hatch and day 7. This was likely due to an approximate 6-fold increase in liver weight between these time points, and also because the herons were fed an uncontaminated diet during this period.  Thus, the TCDD was “diluted” by the growth  occurring during the course of this experiment. It is also possible that a portion of the liver TCDD burden was excreted during this period. Overall, these data indicate that significant levels of TCDD were accumulated in the liver of avian hatchlings after early in ovo exposure to TCDD via air cell injection. It should  92  also be noted that these concentrations are expressed in terms of wet tissue weight, and that lipid-corrected concentrations were not determined.  4.A.3. Thyroid Hormones, Growth, and Edema in Avian Embryos and Hatchlings Several effects of TCDD and related chemicals involve reduced growth and altered development, and early life stages of vertebrates, including birds, are particularly sensitive to these effects (Peterson et a!. 1993).  Thyroid hormones have an important influence on  perinatal growth and development in avian species (King and May 1984; Scanes et a!. 1987). Plasma levels of thyroid hormones are known to increase significantly at or soon after hatch in precocial and altricial birds, respectively (McNabb 1987; Scanes et at. 1987), reflecting the 3 and T 4 in perinatal growth and development. importance of T  In addition to plasma  , plasma 4 4 /T ratio was also determined in these 3 T 3 and total T concentrations of total T experiments.  Since T 3 is the physiologically active thyroid hormone (Brent 1994), and is  /T ratio can be used as a measure of the relative levels of 3 T , plasma 4 4 metabolized from T circulating thyroid hormones. In this section, effects of TCDD exposure on plasma thyroid hormone levels, edema, and body and skeletal growth in avian hatchlings will be discussed.  4.A. 3.1. Chicken Embryos and Hatchlings TCDD exposure had no effect on body weights of chicken embryos and hatchlings at any time point (Table 3.2), or on crown-rump length, tibia length, and culmen length measured in day of hatch chickens (Table 3.3). Chicken hatchlings were not fed, and had decreased body weights on day 2 and day 4 in comparison to day of hatch (Table 3.2). However, there was no hatchling mortality, and day 2 and day 4 hatchlings still had yolk remaining to be resorbed when sacrificed. In addition, subcutaneous edema was not observed at any time point. The absence of these characteristic signs of TCDD toxicity in chicken embryos and hatchlings is surprising given the significant EROD induction observed in this 50 for this species. experiment, and early in ovo TCDD dose which was just below the LD 93  /T 3 T , or plasma 4 3 , total T 4 There was no effect of TCDD exposure on plasma total T ratio in chicken embryos and hatchlings at any time point (Fig. 3.6). The peak in plasma total 3 concentration and 4 T 1T ratio on day of hatch is in agreement with studies in the domestic 3 T chicken (Scanes et at. 1987) and other precocial avian species (McNabb 1987). The lack of /T ratio 3 T , total T 4 , and plasma 4 3 effect of TCDD treatment in this study on plasma total T indicates that plasma thyroid hormone concentrations are not as sensitive a biomarker of TCDD exposure in comparison to hepatic EROD induction in chicken embryos and hatchlings. There is limited information regarding effects of TCDD and related chemicals on thyroid hormone levels in avian embryos and hatchlings. Van den Berg et aL (1994) found a negative correlation between free plasma T 4 levels and PCB concentrations in the yolk of great cormorant hatchlings. It was hypothesized that hydroxylated PCB metabolites may be interfering with binding of T 4 to the plasma transport protein transthyretin, resulting in increased excretion of T 4 (Brouwer and Van den Berg 1986; Van den Berg et aL 1994). Alternatively, induction of UDP-glucuronosyltransferase by TCDD, known to be under direct 4 as the regulation by the AhR (Okey et at. 1994), could cause increased excretion of T glucuronide conjugate. Murk et aL (1994) reported positive correlations between hepatic 4 levels and UDP-glucuronosyltransferase activities in EROD activity and both plasma total T the common tern. Alterations of thyroid hormone levels by TCDD and related chemicals may be due to a number of possible mechanisms.  4 TCDD is known to increase the biliary excretion of T  3 glucuronide (Bastomsky 1977a,b). This glucuronide, but has no effect on excretion of T 4 levels in rodents, but not have may be reflected in the ability of TCDD to decrease serum T 3 levels. TCDD has also been reported to increase TSH any consistent effects on serum T 4 levels (Bastomsky 1 977a), probably as a homeostatic response to lowered plasma T concentrations. As mentioned previously, TCDD and related chemicals, and their metabolites, may also compete with T 4 for binding to transthyretin, resulting in the increased clearance and 94  excretion ofT 4 (Brouwer and Van den Berg 1986). Although these seem the most plausible mechanisms, it is also possible that TCDD influences thyroid hormone levels by altering the activity of Type I and/or Type II T 4 5-deiodinases responsible for conversion of T 4 to T 3 in liver and brain, respectively (Brent 1994).  4.A. 3.2. Pigeon Hatchlings The chicken is one of the most sensitive animals with respect to TCDD toxicity (Peterson et at. 1993), is well characterized biologically, and eggs are readily available and inexpensive. For these reasons, fertile chicken eggs provide an excellent model for studying embryotoxic effects of TCDD. However, the chicken is a precocial species, while fish-eating birds such as herons, cormorants, and raptors are altricial or semi-altricial. Furthermore, most of the piscivorous birds studied to date appear to be on the resistant side of the spectrum with respect to TCDD toxicity (Gilbertson et a!. 1991; Bosveld and Van den Berg 1994; Giesy et a!. 1994). Therefore, an altricial species such as the pigeon provides a more physiologically relevant model for studying the effects of TCDD exposure perinatally in wild avian species. In this regard, the pigeon was also utilized to develop techniques to be used in studying wild fisheating birds such as the great blue heron. In the early in ovo exposure experiment in pigeons, there were decreases in body and skeletal growth in TCDD-treated hatchlings (Table 3.4).  The skeletal growth parameters  (crown-rump, tibia, wing, and culmen lengths) were significantly decreased in TCDD-exposed pigeons at hatch and day 7  (p<O.Ol).  Yolk-free body weight was decreased and liver to body  weight ratio was increased on day 7 (p<O.Ol), but not at hatch (Table 3.4). These effects are consistent with the known effects of TCDD on perinatal growth in avian species. However, TCDD exposure had no effect on plasma total T , total T 4 , or plasma 4 3 /T ratio in this 3 T experiment (Fig. 3.7). In addition, there was no incidence of subcutaneous edema at either time point. The pigeon appears to be resistant to the edematous response, since Sanderson  95  and Beliward (1995) reported no incidences of edema in pigeons exposed in ovo to TCDD at doses as high as 100 p.g/kg. Overall, the decreased hatchability, half-maximal EROD induction, hepatomegaly, and reductions in body and skeletal growth indicate that AhR-mediated biological and toxicological responses are occurring in TCDD-exposed pigeons which parallel effects observed in wild avian species. The lack of effect of TCDD on thyroid hormone levels in this experiment suggests that this response is not as sensitive a biomarker as EROD induction, body growth, and mortality in pigeon hatchlings exposed early in ovo to TCDD. In the late exposure experiment in pigeons, significant decreases in body and skeletal growth, and elevated liver to body weight ratio were observed in TCDD-exposed hatchlings (Table 3.5). Similar to the early exposure experiment, edema was not observed at either time point. Plasma total T 3 concentration was significantly lower in TCDD-exposed pigeons at hatch (p<O.O5), but not at day 7 (Fig. 3.8). Since T 3 is the physiologically active thyroid hormone, this decrease in T 3 at hatch may have toxicological relevance. The results from this experiment suggest that there may be adverse effects on thyroid hormone levels in pigeons exposed to TCDD during the latter third part of incubation. Although the body weights of pigeons increased approximately 2-3 fold between hatch and day 7 in the two experiments (Tables 3.4 and 3.5), this growth is significantly less than the ten-fold increase in body weight observed over this time period in naturally-raised pigeons (this study) or in pigeons raised on artificial crop milk (Yang and Vohra 1987; Aggrey and Cheng 1993). The reduced growth in these experiments may be due to a variety of factors. Pigeons were not fed for the first 12-18 hours following hatch, and for the 12-24 hour period prior to sacrifice. There were three feedings per day, compared to five feedings per day given by Yang and Vohra (1987). Finally, pigeons were housed using egg incubation conditions (37.5°C and 55% relative humidity), which may have resulted in dehydration and heat stress. This was particularly evident in the late exposure experiment, in which there were five mortalities in control pigeons and eleven mortalities in TCDD-treated pigeons fed to day 7. 96  Nevertheless, control and TCDD-exposed pigeons were raised using the same conditions, so that valid comparisons can be made between groups despite less than optimal growth. Comparison of plasma total T 3 and T 4 levels in experimental pigeons at hatch and day 7 (Figs. 3.7 and 3.8) with naturally-raised pigeons on day 1 and day 6 (Fig. 3.9) may reflect differences in body growth during this time period. Although total T 3 and T 4 concentrations were similar between experimental and naturally-raised pigeons at hatch, the thyroid hormone levels were 2-3 fold higher in naturally-raised day 6 pigeons in comparison to experimental day 7 pigeons (Figs. 3.7-3.9). However, thyroid hormone levels in pigeon squabs raised to an optimal body weight on the artificial diet are not known.  4.A. 3.3. Great Blue Heron Hatchlings Body weights in control and TCDD-exposed herons fed to day 7 were not different when measured daily (Fig. 3.1). The growth observed in herons from hatch to day 7 was similar to that reported in heron hatchlings fed a similar diet (Bennett 1993). Yolk-free body weights and liver to body weight ratios were similar between control and TCDD-exposed herons at hatch and day 7 (Table 3.6). Crown-rump length, tibia length, and wing length were significantly reduced at hatch in TCDD-treated herons, but not on day 7 (Table 3.6). These results suggest slight growth-related effects of TCDD exposure in this experiment. There was one incidence of subcutaneous edema out of six TCDD-treated herons at hatch. Edema was not observed in herons fed to day 7.  Sanderson and Beliward (1995)  observed considerably higher incidences of edema in great blue heron hatchlings exposed to TCDD.  These authors reported a 43 to 75 percent incidence of subcutaneous edema in  herons exposed to in ovo doses of between 0.5 to 10 p.g/kg. Furthermore, edema was shown to be a sensitive response in herons, occurring in approximately the same dose range as induction of EROD activity (Sanderson and Beliward 1995). Although plasma T 4 / 3 T ratio was significantly elevated in day 7 herons exposed to TCDD (p<O.05), there were no consistent changes in plasma thyroid hormone levels in this 97  experiment (Fig. 3.10). Plasma total T 3 and T 4 levels increased 2-3 fold between hatch and day 7. This increase was expected, based on the limited information available on thyroid hormone levels in altricial avian hatchlings (McNabb 1987). This increase during the first week after hatch in herons corresponds to the peak observed in the precocial chicken at hatch, and reflects the delayed development of altricial species. Overall, the results from this experiment show that at environmentally relevant TCDD exposure causing slight elevations in EROD and decreases in growth, there is no effect on plasma thyroid hormone concentrations. Thus, thyroid hormone levels are not as sensitive a biomarker as EROD induction in great blue heron hatchlings.  4.A.4. Sex Steroid Hormones and Receptors in Avian Embryos and Hatchlings  Although the sex steroid hormones are also involved in perinatal growth and development, a more important role involves sexual differentiation of the avian CNS. In birds the heterogametic sex is female, and feminization and demasculinization of the female embryo is dependent on ovarian estrogen secretions (MacLusky and Naftolin 1981; Jost 1983). These gonadal secretions determine, in an irreversible manner, the sexual phenotype of the brain and secondary sexual characteristics including behavior.  Alterations in sex steroid hormone  function during the perinatal period in birds may have consequences that persist throughout adult life, and may be a factor in the decreased reproductive success observed in certain wild avian populations. The main focus of this section is to examine hepatic estrogen receptor affinities and concentrations, and plasma 1 713-estradiol levels in female avian embryos and hatchlings exposed in ovo to TCDD. In addition, plasma testosterone levels were determined in male embryos and hatchlings as a measure of androgenic status during the perinatal period.  98  4.A. 4.1. Chicken Embryos and Halchlings ed steadily Plasma estradiol levels in female chicken embryos and hatchlings increas -exposed between E19 and day 4 (Fig. 3.11). Although significantly decreased in TCDD ol at the earlier female hatchlings on day 4 (p<O.05), there were no alterations in plasma estradi 4 (Fig. 3.2), time points. Based on the increasing EROD induction between E19 and day experiment, which may be indicative of elevated exposure to TCDD during the course of this It is possible effects on plasma estradiol could be expected to occur at the later time point(s). at further time that more significant effects of TCDD on estradiol levels may have occurred in comparison points in juvenile chickens (e.g. one to four weeks postnatally). Nevertheless, ent, plasma with the high level of EROD induction observed in the chicken in this experim the perinatal estradiol levels do not appear to be a sensitive biomarker of exposure during period in TCDD-exposed female birds. ) were The plasma estradiol levels measured in this experiment (15-35 pg/mL values are in compared to literature values for female chicken embryos and hatchlings. These red between E17 agreement to those reported by Tanabe et aL (1979) of 15-25 pg/mL measu (1986), plasma and day 7 after hatch. However, in a subsequent study by Tanabe et a!. al period in female estradiol levels of 54-181 pg/mL were reported over the same perinat ol concentrations chickens. A study in female chicken embryos reported plasma estradi 1400 pg/mL on increasing throughout incubation, from approximately 900 pg/mL on E7.5 to and Brazzill E17.5 (Woods and Brazzill 1981). The estradiol levels reported by Woods in sexually mature (1981) seem unusually high; in contrast, plasma estradiol concentrations human females rarely exceed 1000 pg/mL unless measured during pregnancy. red by MA) The large differences in reported estradiol levels in chickens (all measu preparation, and/or may be due to a variety of factors. Differences in sample collection, were determined in storage could influence values. Estradiol concentrations in this study tradiol as antigen, whereas in the studies by Tanabe et al 25 1]-es 1 unextracted plasma using [ ed into diethyl ether (1979, 1986) and Woods and Brazzill (1981) plasma samples were extract 99  and measured using [ H]-estradiol as antigen. It is possible that the unextracted avian plasma 3 used in this study contained compounds that interfere with binding to the antiserum, resulting in lower measured values. Recovery experiments were not performed in the present study to examine this possibility.  The antigen used to generate antiserum in this study (6-keto-  estradiol-1713-6-oxime-BSA) was different to that employed by Tanabe et at (1979, 1986) and Woods and Bra.zzill (1981). However, the specificity of both antisera for estradiol was high, and cross reactivities with related steroids were reported to be very low. It is also possible that the time of year/photoperiod used in the respective studies differed, which could influence plasma estradiol levels. Hepatic estrogen receptor (ER) levels were determined to examine the sensitivity of this response in female chicken embryos and hatchlings (Table 3.7).  Although analysis of  receptor binding data is traditionally performed using Scatchard plots (Scatchard 1949), the Woolf plot approach (Haldane 1957) has been shown to be more reliable and less sensitive to outliers than Scatchard analysis (Cressie and Keightley 1981).  Thus, both Scatchard and  Woolf plots were used to analyse the ER binding data. In general, the results obtained from each method of analysis were similar for the three bird species (Tables 3.7-3.10). Apparent hepatic ER concentrations (Bmax) were generally low in chickens, but measurable at all time points. Bmax values were 2-3 fold higher in E19 embryos than in day of hatch, day 2, and day 4 chickens (Table 3.7).  TCDD exposure had no effect on ER  concentrations at any time point, although receptor affinities (Kd) were elevated on E19 (Scatchard analysis, p<0.Ol) and decreased on day 2 (Woolf analysis, p<O.O5) in TCDD treated birds.  Regardless of these alterations in receptor affinity, there was no consistent  pattern of TCDD affecting ER levels in this experiment in chickens. Hepatic estrogen receptors have previously been studied in chicken embryos as a convenient endocrinological model. The Kd values determined in this experiment (0.16-0.46 nM) are in agreement with the reported literature values of 0.1-0.7 nM in chicks (Mester and  100  Baulieu 1972; Lazier 1987). Immature chickens are also reported to have low hepatic ER concentrations, consistent with the Bmax values shown in Table 3.7 (Lazier 1987). Testosterone levels increased sharply between day 2 and day 4 in male chickens, reaching 10-15 pg/mL on day 4. Plasma testosterone concentrations in male embryos and hatchlings were below the quantitation limit in several samples collected between El 7 and day 2, resulting in large variability in these data (Fig. 3.16). As mentioned previously, extracted plasma samples in this study contained large amounts of yolk lipids which may have interfered with reconstitution of testosterone into the assay buffer.  However, radioimmunoassay of  testosterone in unextracted plasma from these birds also resulted in levels below the quantitation limit.  Nevertheless, there was no effect of TCDD exposure on plasma  testosterone levels in this experiment. A study in male chickens reported a significant peak in testosterone levels occurring on day 1 after hatch, with plasma concentrations reaching 300 pg/mL (Tanabe et aL 1979). However, in a subsequent study by the same author, plasma testosterone did not peak at hatch but remained relatively constant throughout the perinatal period in chickens at between 53109 pg/mL (Tanabe et a!. 1986). In a separate study in chickens, plasma testosterone was shown not to peak at hatching, with levels of approximately 5-11 pg/mL (determined by GC MS) occurring between E19 and day 1 after hatch (Corbier et a!. 1992). The reason for these large discrepancies in reported testosterone levels in male chicken embryos and hatchlings is not known, but may involve methodological differences. Overall, plasma estradiol concentrations and ER levels in females, and plasma testosterone concentrations in males, indicate that these endpoints are not as sensitive as EROD induction as a biomarker of exposure to TCDD in the chicken.  4.A. 4.2. Pigeon Hatchlings Plasma estradiol concentrations of 10-20 pg/mL were similar in female and male pigeon hatchlings when measured at hatch and day 7 (Figs. 3.12 and 3.13). Plasma estradiol 101  levels in naturally-raised pigeon hatchlings were similar to the hand-fed birds (Fig. 3.14; sex unknown).  TCDD exposure did not affect plasma estradiol levels in the early in ovo  experiment in pigeons (Fig. 3.12). There was no significant effect of TCDD treatment on hepatic ER affinities or concentrations in female hatchlings at either time point (Table 3.8). Hepatic ER concentrations in TCDD-exposed female pigeons on day 7 were approximately half of those measured in controls (Table 3.8); however, this difference was not statistically significant. It is interesting to note that control ER affinities were approximately an order of magnitude higher in pigeons at hatch in comparison to day 7. The physiological importance of this difference is not known, especially since Kd values were not measured in pigeon embryos. It is possible that greater ER affinities prenatally may be involved in growth and development, andJor sexual differentiation of the CNS. A similar difference in control hepatic ER affinities between hatch and day 7 was observed in the late in ovo TCDD exposure experiment in pigeons (Table 3.9).  In this  experiment, plasma estradiol levels were significantly decreased in TCDD-exposed female pigeons at hatch, and elevated on day 7 (p<O.05; Fig. 3.13). A similar trend was observed in male pigeons but was not statistically significant (Fig. 3.13).  The differences in TCDD  exposed female estradiol levels correspond to significantly elevated ER concentrations in female pigeons at hatch (p<O.05; Table 3.9). Bmax values in TCDD-exposed female pigeons remained elevated on day 7, although not significant statistically.  These results suggest  perturbations in estrogen homeostasis caused by TCDD during the perinatal period in the late exposure experiment in pigeons.  4.A. 4.3. Great Blue Heron Hatchlings Plasma estradiol concentrations in heron hatchlings (male and female combined) increased approximately 2.5-fold between hatch and day 7 (Fig. 3.15). There was no effect of TCDD exposure on estradiol levels in this experiment. However, there were effects of TCDD treatment on hepatic ER levels in herons at hatch (Table 3.10). Kd values were elevated in 102  TCDD-exposed herons at hatch (p<O.O5), and ER concentrations were elevated in heron hatchlings as indicated by Woolf analysis (p<O.05). A similar but not statistically significant trend for Bmax values was observed in TCDD-exposed herons at hatch using Scatchard analysis (Table 3.10). Hepatic ER concentrations and affinities were similar between control and TCDD-treated herons on day 7. The lack of effects observed on day 7 in herons may be a reflection of the large decrease in liver TCDD concentration between hatch and day 7 (Table 3.11). The toxicological significance of the alterations in hepatic ER levels in TCDD-exposed herons at hatch is not known, but these results parallel those observed in the late in ovo TCDD exposure experiment in pigeons, and suggest effects on estrogen homeostasis during the perinatal period in these species. TCDD has been shown to modulate a number of receptor systems in mammals without acting as an agonist, including receptors for estrogen, progesterone, epidermal growth factor, glucocorticoids, and prolactin (Goldstein and Safe 1989; Landers and Bunce 1991). With respect to ER, studies in mammals indicate that TCDD acts as an antiestrogen by downregulating uterine and hepatic ER concentrations without affecting ER affinity for estradiol (DeVito el a!.  1991).  Furthermore, dose-response relationships for ER  downregulation by a variety of TCDD-like HAHs are consistent with their affinity for the AhR, suggesting AhR involvement (Romkes et a!. 1987). In contrast, great blue herons and pigeons exposed in ovo to TCDD in this study exhibited significantly elevated hepatic ER concentrations at hatch (Tables 3.9 and 3.10). In addition, day of hatch herons exposed to TCDD had decreased ER affinities (Table 3.10). The exact mechanism(s) by which TCDD affects ER levels are not clear, but several hypotheses involving heterologous regulation of ER can be proposed (reviewed in Safe et aL 1991). The core sequence of the AhRE (GCGTG) is similar to that of the estrogen responsive element (ERE) in humans (White and Gasiewicz 1993), indicating the high conservancy of DNA binding domains, and suggesting that the activated AhR complex may bind to the ERE, thus directly affecting structural genes induced by ER. 103  This potential mechanism has  implications with respect to the estrogen responsiveness of organisms or tissues being studied. In sexually mature females, elevated estradiol levels would result in high ER occupancy in target tissues such as the uterus and liver, and the effects of low TCDD exposure on ER levels would be minimal. In contrast, in immature females, and certainly during the avian perinatal period, the relatively low estradiol levels and ER occupancy may result in a greater ability of the activated AhR complex to interact with the ER gene, causing perturbations in ERmediated protein synthesis.  This illustrates the considerable dose-, species-, and tissue-  specific nature of TCDD toxicities. The dose dependency and species differences associated with TCDD toxicity, as well as the complexity of TCDD modulation of ER levels, may help explain the upregulation and reduced affinity of ER observed in the present study. The doses of TCDD used in this study are considerably lower than those utilized in mammalian  studies examining ER  downregulation. As mentioned previously, there may be a relationship between TCDD dose and the levels of estradiol in the animal being studied, due to relative abilities of the AhR complex and ER to “compete” for estrogen responsive elements on the DNA.  Thus, the  discrepancy between ER concentrations in this study (upregulation) and studies in rodents (downregulation) may be a dose-related phenomenon. Since the liver is not the primary target organ for the effects of ER, tissue-specific differences may also occur in avian liver which differ from mammalian liver or uterus. Finally, there may be species differences in the ER responses to TCDD exposure during the perinatal period that are specific to avian species. Alternatively, the AhR complex could induce the synthesis of proteins which may indirectly modulate ER by a variety of mechanisms, referred to generally as “cross talk”. These modulatory proteins may decrease the stability of the nuclear ER, or influence estrogen induced gene transcription (Safe et aL 1991). An example of this would be altered activities of protein kinases as modulators (Matsumura 1994), such as protein kinase C or protein tyrosine kinase, which can regulate receptor function through phosphorylation.  Another  possible short term regulatory mechanism may involve allosteric effects of TCDD or the AhR 104  complex on ER.  Posttranscriptionally, TCDD exposure may affect protein levels by  decreasing the stability of mRNA, or have effects fi.irther downstream from mRNA translation. In addition, TCDD has been shown to increase 2- and 16c&estradiol hydroxylation in vitro (Gierthy et a!. 1988). The 2- and 1 6c-hydroxylations of estradiol are believed to be the main metabolic pathways for this compound, and the main cytochrome P450 isozymes involved are CYP1A2, inducible by TCDD and related chemicals, and CYP2C11 (Martucci and Fishman 1993). Thus, increased metabolism of estradiol by TCDD could also decrease the responsiveness of ER due to reduced ligand concentrations. In this study, there were no consistent effects of TCDD exposure on plasma estradiol levels.  It is possible that any  TCDD-induced decreases in plasma estradiol that may have occurred could have been compensated for rapidly via feedback control, by increasing the release of luteinizing hormone by the pituitary gland. Thus, there could have been important perinatal time periods during which plasma estradiol levels were decreased by TCDD exposure and not detected due to the limited number of time points examined in this study. Plasma testosterone concentrations of 5-10 pg/mL were observed in heron hatchlings (male and female combined). There were no differences between control and TCDD-exposed herons at hatch or day 7 in this experiment (Fig. 3.17). Thus, plasma testosterone levels in great blue heron hatchlings are not as sensitive a biomarker of TCDD exposure as EROD induction.  4.B. ADULT GREAT BLUE HERONS  The herons used in this study were artificially incubated and raised from the hatchling stage at the Department of Animal Science, UBC (Bennett 1993). Eggs were collected from heron colonies in British Columbia, located at Little River, Chilliwack, Crofton, and on the UBC Endowment Lands (now Pacific Spirit Regional Park). Eggs were collected in 1990 and 1991, and by the time of the acute TCDD exposure study the birds were 3-4 years old and 105  presumably adults. It should be noted that eggs were collected from colonies varying in HAH contamination: the Crofton and UBC eggs from relatively contaminated colonies, and the Chilliwack and Little River eggs from relatively uncontaminated colonies. However, when separating the herons into two groups for the acute TCDD exposure experiment, an approximately equal number of birds from each original location were included in each group. It is not known what effects, if any, this HAH contamination had on the herons, especially with respect to failed attempts at breeding, although no obvious differences were observed between birds from the various locations.  4.B.1. Acute Effects of TCDD in Adult Great Blue Herons  Adult herons were exposed to a single intraperitoneal dose of 20 p.g/kg of TCDD and sacrificed after 14 days.  Hepatic EROD was induced approximately six-fold in TCDD  exposed herons. This level of EROD induction suggests a moderate TCDD dose, although there is no information on maximal EROD activity in adults of this species. However, EROD induction of 20-40 fold or greater above controls is common in the large number of vertebrate species studied to date. The liver concentration of [ H1-TCDD in adult herons (3.55 nglg, wet weight) was 3 intermediate between liver concentrations measured in day of hatch (11.32 ng/g) and day 7 (0.82 nglg) herons (Tables 3.11 and 3.14).  H]3 Only 1.9 percent of the intraperitoneal [  TCDD dose in adult herons was found in the liver, 14 days after exposure. The concentration of [ H]-TCDD in adipose tissue collected from the breast of adult herons was almost 20-fold 3 greater than the liver [ H]-TCDD concentration (Table 3.14), due to the preferential 3 distribution of TCDD to fat. Plasma total T 4 concentration was significantly elevated in TCDD-exposed herons (p<O.O5). There was no significant effect of TCDD treatment on plasma total T 3 or plasma /T ratio, although 4 3 T 4 /T ratio was somewhat lower in TCDD-exposed herons (Table 3 T 3.12). 106  Plasma estradiol levels were similar between control and TCDD-exposed herons, and were also similar between female and male herons (Table 3.13). Hepatic ER concentrations and affinities in adult herons were similar in magnitude to those measured in heron hatchlings (Tables 3.10 and 3.13). Acute TCDD exposure in this experiment had no effect on hepatic ER levels in males, or males and females combined. Hepatic ER levels in female herons could not be tested statistically due to the small sample size. Plasma testosterone concentrations were highly variable in the herons (Table 3.13), possibly reflecting different stages of maturation in these birds.  Prior to and during this  experiment a few of the herons were making attempts at nest-building, and presumably breeding, and these factors may have influenced testosterone levels in certain birds.  4.B.2. Seasonal Hormone Levels in Great Blue Herons  The plasma samples used for these analyses were collected by the Department of Animal Science, UBC, monthly between September 1991 and April 1992. The herons were between one and two years old when these samples were collected, and therefore were not fully mature. The May samples were collected in 1993, when the herons were 2-3 years old. The possibility of the herons being sexually immature is important, especially with respect to the sex steroid hormone analyses. Plasma total T 3 and 4 /T ratio approximately doubled during the winter months 3 T (December-February) in comparison to spring and autumn months (Fig. 3.19), reflecting increased metabolic demands on the herons during the colder time of the year. In comparison, the lower 4 /T ratio observed in TCDD-exposed herons in the acute study (0.07), which 3 T was conducted in February, may have had physiological relevance to the herons. Plasma estradiol levels were similar between female and male herons, and appeared to be elevated during the spring months, although not to a great extent (Fig. 3.20). The estradiol concentrations measured in May (Fig. 3.20) are not necessarily comparable to the remaining data for September to April, since these samples were collected in the subsequent year. 107  Plasma testosterone levels were highly variable in female herons, partly due to the small sample sizes available for analysis (Fig. 3.21). In contrast, plasma testosterone levels in male herons displayed a distinct increase between January to March, corresponding to the time when breeding occurs in this species (Butler 1991).  4.C. FUTURE RESEARCH This thesis compared the induction of CYP1AI-associated EROD activity, a known sensitive and specific biomarker of exposure to TCDD and related chemicals, to a number of hormonal endpoints with potential toxicological relevance to the reproductive success of wild avian species, particularly piscivorous birds. The early in ovo exposure protocol used has relevance to environmental exposure of wild avian species. These methods could be used to examine potential effects of TCDD and related chemicals in other wild avian species susceptible to adverse reproductive outcomes due to their position in the food web. Other susceptible avian species include raptors such as bald eagles (Haliaeetus leucocephalus) and ospreys (Pandion haliaetus). A number of other diving piscivorous birds, such as cormorants, mergansers, and grebes, may also be sensitive to adverse effects of perinatal TCDD exposure. These birds generally consume larger fish than herons, which may result in higher exposure to TCDD via the diet, and thus greater quantities of TCDD transferred to eggs. Although TCDD is prototypical of the HAHs, environmental media contains complex mixtures of these compounds, many of which may act as partial agonists or even antagonists of the Ah receptor. An important goal of this research is to attempt to simulate environmental exposure conditions as closely as possible. In ovo exposure of avian species to mixtures of HAHs, such as extracts from contaminated prey items (fish), may represent a more environmentally realistic exposure scenario.  108  In addition, effects on hatchlings fed a  contaminated diet postnatally for longer periods (e.g. 4-6 weeks) would also increase our understanding of adverse effects of HAH exposure in wild avian species. Although environmental levels of the highly toxic PCDDs and PCDFs have decreased in recent years due to process changes implemented in pulp and paper mills in Canada, there is evidence that other products of the bleaching process may still be affecting animals inhabiting local areas.  Since there are over 30,000 different chemicals released in pulp mill effluent,  identification of compounds acting via the AhR-mediated mechanism, or possibly other mechanisms, is important but problematic.  A potential approach would be to fractionate  whole effluent based on molecular weight and/or polarity, and to screen these fractions for toxicity using an in vitro test. A variety of cell lines are routinely used for this purpose, and in recent years several avian cell culture systems have been employed.  Alternatively,  identification and isolation of individual components of pulp mill effluent would allow structure-activity relationships to be conducted on selected compounds with the highest potential for toxicity. Once the number of possible toxic constituents has been reduced, these compounds could be tested more rigorously using the in ovo exposure route. Examination of the persistence and toxicity of these compounds in avian embryos and hatchlings is another potential research direction. Attempting to elucidate possible endocrinological mechanisms of toxicity in avian species is difficult due to the complexity in hormonal systems and interactions between different systems, especially when modulated by TCDD and related chemicals. We believe the endpoints used in this thesis were relevant to this goal, but limited with respect to the large variety of effects of TCDD on endocrine homeostasis.  TCDD and related chemicals are  believed to exert many of their toxicities via alterations in cell growth and differentiation. The genetic regulation of these processes is currently a topic of great interest in many scientific disciplines, including toxicology. An important area of investigation would be to examine the effects of TCDD and related chemicals on growth factors and receptors during the perinatal period in avian species. 109  Another hypothesis for the high embryo and hatchling mortality observed in piscivorous birds exposed in ovo to HAHs involves hormonal control of calcium homeostasis. A common effect observed in these wild birds is an inability to successfUlly break open and exit from the egg shell. A balance between circulating calcium levels and calcium deposited in bone is essential in avian species at the time of hatch. Sufficient calcium needs to be available for the “pipping tooth” which enables the bird to pierce the egg shell during hatching. In addition, calcium is essential for skeletal muscle contraction. The hatching process requires considerable energy, and sufficient availability of circulating calcium would be necessary for this to occur.  Thus, calcium homeostasis plays an important role in the avian hatching  process, and TCDD-induced alterations in parathyroid hormone, which mobilizes calcium, and/or calcitonin, which controls deposition into bone, could have toxicological relevance in fish-eating bird species. The results in great blue heron and pigeon hatchlings in this study suggest that TCDD exposure causes perturbations in estrogen receptor levels during the perinatal period. Since sexual differentiation of the female avian brain occurs during this period, there may be irreversible effects on the development of sexual function caused by early in ovo TCDD exposure. Further experiments are needed to examine the effects of TCDD exposure on ER levels during earlier embryonic development, for example between embryonic day 5 and hatch. Further experiments could be conducted in which adult female sexual function is examined These experiments would be important in  following early in ovo TCDD exposure.  determining whether TCDD has adverse effects on the organizational and activational periods of sexual development in female avian species.  110  5. CONCLUSIONS The early in ovo exposure protocol used in this thesis provides a usefi.il and reliable tool for studying the effects of persistent environmental contaminants during the perinatal period in avian species.  T ratio) in chicken embryos and 4 / 3 , T 3 , T 4 Plasma thyroid hormone concentrations (T hatchlings, pigeon hatchlings, and great blue heron hatchlings are not as sensitive a biomarker as EROD induction when these birds are exposed early in incubation to TCDD.  Plasma I 7f3-estradiol and testosterone concentrations in chicken embryos and hatchlings, pigeon hatchlings, and great blue heron hatchlings are not as sensitive a biomarker as EROD induction when these birds are exposed early in incubation to TCDD.  In pigeons exposed during the latter third part of incubation to a TCDD dose that would cause excessive mortality if exposure occurred earlier in incubation, there were effects of TCDD on plasma thyroid hormones, plasma I 73-estradiol, and hepatic estrogen receptor concentrations. The difference in responses between early and late in ovo exposure to TCDD reflects the high sensitivity to embryo mortality which occurs when eggs are exposed early in incubation.  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