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In Ovo induction of hepatic ethoxyresorufin o-deethylase activity by 2,3,7,8-tetrachlorodibenzo-P-dioxin… Sanderson, John Thomas 1994

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INOVO INDUCTION OF HEPATICETHOXYRESORUFIN 0-DEETHYLASE ACTIVITYBY 2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN INFOUR AVIAN SPECIESbyJOHN THOMAS SANDERSONDrs., Free University of Amsterdam, 1989A THESIS IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREEOF DOCTOR OF PIHLOSOPHYinTHE FACULTY OF GRADUATE STUDIESTHE FACULTY OF PHARMACEUTICAL SCIENCESDivision of Pharmacology and ToxicologyWe accept this thesis as conforming to the required standardThe University of British ColumbiaMay 1994© John Thomas Sanderson, 1994In presenting this thesis in partial fulfilment of the requirements for an advanceddegree at the University of British Columbia, I agree that the Library shall make itfreely available for reference and study. I further agree that permission for extensivecopying of this thesis for scholarly purposes may be granted by the head of mydepartment or by his or her representatives. It is understood that copying orpublication of this thesis for financial gain shall not be allowed without my writtenpermission.(Signature___________________________Department CL&(The University of British ColumbiaVancouver, CanadaDate 2 1 i!LDE-6 (2188)ABSTRACTThe Canadian Wildlife Service monitors levels of polychiorinated aromatic hydrocarbonsin great blue heron (Ardea herodias) and double-crested cormorant (Phalacrocorax auritus)eggs in British Columbia (B.C.) as biological indicators of environmental contamination. Thisthesis assessed the spatial and temporal effects of environmental contamination withpolychiorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs) onhepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities and other biologicalparameters in heron hatchlings from colonies in B.C and cormorant hatchlings from coloniesacross Canada. Also, the induction of hepatic EROD activities by environmental levels ofpolychiorinated aromatic hydrocarbons were compared to in ovo dose-response studies, using2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), in the laboratory. Finally, the differences inEROD-inducing potency of TCDD among avian species were compared with the affinities ofTCDD for the avian hepatic Ah receptors.In the herons, levels of PCDDs and PCDFs had decreased considerably since 1988,particularly in the Crofton colony. Concomitantly, there was a decrease in hepatic ERODactivity and incidence of chick edema, an increase in body weight and organ weights, andimprovement of the reproductive success of the Crofton colony. The reduction in severity ofthe effects observed in the contaminated colonies in the recent collections, accompanied by thedecline in levels of PCDDs and PCDFs, was consistent with the dose-response relationshipsdetermined in 1988. Regression analysis demonstrated a significant positive relationshipbetween hepatic EROD activity (pmole/minlmg protein) and TCDD level (r2=O.49;p<O.00005; n=54). In the cormorants, hepatic EROD activity was positively related to thesum of TCDD-toxic equivalents (r2=O.69; p<O.00005; n25). Immunoblots showed cross-reactivity of mono specific antibodies raised against rat cytochrome P-45 0 1 Al (CYP 1 Al)with a hepatic microsomal protein in the heron and cormorant hatchlings. Staining of the bandincreased with increased EROD activity and increased exposure of the birds.These results support the use of avian hepatic microsomal EROD activity as an index ofCYP1A1 induction by environmental levels of polychiorinated aromatic hydrocarbons, and asa useful biomarker of the extent of exposure to such chemicals. Furthermore, the induction of— ii —AbstractCYP 1A1 observed in the heron and cormorant indicates that the Ah receptor-mediatedprocess, by which TCDD and related chemicals exert many of their toxicities, has beenactivated.The in ovo hepatic EROD-inducing potency of TCDD, as an indicator of sensitivity toAh receptor-mediated toxicities, was determined in the domestic chicken (Galius gallus),domestic pigeon (Columba livia), great blue heron and double-crested cormorant. Dose-response curves were produced by injecting various doses of3H-TCDD (in corn oil) into theair sac of developing eggs, during the latter third part of incubation. After hatching, yolk,blood and a portion of liver were analysed for3H-TCDD, using liquid scintillation counting.There were no major differences in organ distribution of TCDD among hatchlings of the fouravian species. The ED50 for EROD induction by TCDD was about 1 ‘/2 order of magnitudelower in the domestic chick (0.1 pg/kg), a species known to be sensitive to TCDD toxicity,than in the heron and cormorant (3-10 pg/kg). The pigeon was similar to the wild birdspecies. Consistent with this, receptor binding studies showed that the affinity of3H-TCDDfor the hepatic Ah receptor was about 10 to 15 times higher in the chick (Kd=O.7S and 1.6nM) than in the other avian species (pigeon: Kd=l 1 and 14 nM; heron: Kd=lO and 20 nM;cormorant: Kd=l2 and 16 nM). Receptor binding affinities in pigeon, heron and cormorantwere similar to that reported in human placenta. Subcutaneous edema was observed inTCDD-treated hatchlings of the chick, heron and cormorant, but not in the pigeon hatchling,at the dose-range examined.Comparison of the laboratory dose-response curves with environmental exposures toTCDD and related chemicals indicates that hepatic EROD induction is at the lower end of thelinear part of the dose-response curves in both the great blue heron and double-crestedcormorant. A further increase in levels of TCDD and similarly acting compounds would leadto a large increase in hepatic EROD activity and further increases in other Ah receptor-mediated toxicities, such as body weight loss and edema.Key Words: Avian, Ah receptor, biological monitoring, dose-response, double-crestedcormorant, ethoxyresorufin 0-deethylase, great blue heron, 2,3,7, 8-tetrachlorodibenzo-p-dioxin, toxic equivalents.— In —TABLE OF CONTENTSpageABSTRACT iiTable of Contents ivList of Tables ViiList of Figures ixList of Abbreviations xiiAcknowledgments xii1. INTRODUCTION 1LA. TOXICOLOGY OF 2,3,7, 8-TETRA CHLORODJBENZO-P-DJOXINAND REM TED 1POLYHALOGENA TED AROMA TIC HYDROCARBONSl.A. 1. Sources of Polyhalogenated Aromatic Hydrocarbons 1l.A. 1.1. Sources ofPolyhalogenatedAromatic Hydrocarbons in the Strait ofGeorgia 31.A.2. Toxicities of Polychlorinated Aromatic Hydrocarbons 41 .A.3. Mechanism of Action of TCDD and Related Chemicals 4l.A.3. 1. Genetic Regulation of the Induction ofCYPJAJ - Evidencefor the Existence ofan 5Aryl Hydrocarbon Receptorl.A. 3.2. Ah Receptor Characteristics and Function 6l.A.3.3. Presence ofAh Receptor in Avian Speciesl.A.3.4. Structure-Activity Relationships 101.A.4. Environmental Toxicities of Polychlorinated Aromatic Hydrocarbons in Birds 12l.B. BIOLOGICAL MONITORING OF ENVIRONMENTAL CONTAMINATION 131.B. 1. Terms Used in the Field of Biological Monitoring 131.B.2. Why use Sentinel Species? 141 .B.3. Historical Use of Sentinel Species 141.B.4. Criteria for a Good Sentinel Species 151.B.5. Great Blue Herons and their Eggs as Sentinels 151.B.6. Use of CYP1A1 Induction in Biological Monitoring 17l.B.6.l. Introduction 17l.B. 6.2. Use ofCYP1A 1 Induction in Biological Monitoring 18l.B. 6.3. Use ofCYP1A I Induction in Biological Monitoring ofBirds 201.B. 6.4. Toxicological Implications ofEnvironmental Induction ofCYP1AJ 211. C.. HYPOTHESESAND OBJECTIVES 23- iv -Table ofContents2. MATERIALS AND METHODS 252.1. Experimental Birds 252.2. Incubation Conditions 292.3. Injection Procedure 292.4. Dose-Response Experiments 312.5. Tissue Preparations 312.6. Ethoxyresorufin 0-Deethylase Assay 332.7. Total Cytochrome P-450 Determinations 362.8. Protein Determinations 362,9,3H-TCDD Analyses 362.10. Ah Receptor Binding Assay 372.11. Immunochemical Cross-Reactivity Study 382.12. Environmental Chemical Analyses 392.13. TCDD-Toxic Equivalents 402.14. Statistical Analyses 413. RESULTS 423.A. ENVIRONMENTAL MONITORING STUDY 423 A. 1. Levels of Chemical Contamination in Monitored Bird Eggs 423.A.2. Gross Abnormalities and Edema 473 .A.3. Morphological Measurements, Hepatic Microsomal Cytochrome P-450 Content and EROD 49Activity3.A.4. Regression of Morphological Measurements on Chemical Contamination Levels 523.A.5. Regression of Hepatic Microsomal Cytochrome P-450 Content and EROD Activity on 56Chemical Contamination Levels3.A.6. Immunochemical Cross-Reactivity Study 593.B. IN OVO DOSE-RESPONSE STUDY 623 .B. 1. Effects of3H-TCDD on Length of Incubation, Duration of Pip and Mortality 623 .B.2. Effects of3H-TCDD on Body and Organ Weights 633.B.3. Incidence of Edema and other Gross Abnormalities 663.B.4. Induction of Hepatic Microsomal Cytochrome P-450 Content 673.B.5. Induction of Hepatic Microsomal Ethoxyresorufin 0-Deethylase Activity 693.B.6.3H-TCDD Organ Distribution 753.B.7. Relationship between Hepatic Uptake of3H-TCDD and Duration ofIn Ovo Exposure 793.B.8. Comparison of Hepatic EROD Activities with Hepatic3H-TCDD Concentrations in Four 81Avian Species: Detailed Dose-Response CurvesTable ofContents3.C. AHRECEPTOR BINDING STUDY 863.C. 1. Specific Binding of3H-TCDD in Avian Hepatic Cytosolic Preparations 863. C.2. Woolf and Scatchard analyses of3H-TCDD Specific Binding 904. DISCUSSION 934.A. ENVIRONMENTAL MONiTORING STUDY 934.A. 1. Rapid Decline of Levels of PCDDs and PCDFs in Great Blue Heron Eggs 934.A.2. Negative effects of PCDDs, PCDFs and PCBs in Monitored Bird Hatchuings. 944.A.3. Use of TEFs and TEQs in Biological Monitoring of Avian Species 954.A.4. Use of Avian Hepatic EROD Activities in Biological Monitoring 974.A.5. Immunochemical Cross-Reactivity Study 984.B. INOVO DOSE-RESPONSE STUDY 1004.B. 1. Effects of3H-TCDD on Incubation Time and Morphological Measurements 1004.B.2. Incidence of Edema 1004.B3. Induction of Hepatic Microsomal EROD Activity 1024.C. AHRECEPTOR BINDING STUDY 1034.C. 1. Ah Receptor Binding Characteristics of TCDD in Four Avian Species 1034.C.2. Comparison of Au Receptor Affinities in Avian Livers with Human Placenta 1044. C.3. Ah Receptor Binding Affinities as Determinants of the Toxic Potency of TCDD 1054.D. TOXICOLOGICAL IMPLICATIONS OF ENVIRONMENTAL INDUCTION OF CYP1A 1 1064.E. FUTURE DIRECTIONS 1095. CONCLUSIONS 1105.A. ENVIRONMENTAL MONITORING STUDY 1105.B. DOSE-RESPONSE/AHRECEPTOR BINDING STUDY 1116. REFERENCES 112- vi -LIST OF TABLESTable page1.1. Criteria for the involvement of an Ah receptor in the mediation of the toxicities of TCDD. 61.2. Examples of enzyme systems affected by enviromnental contaminants and used as 17biomarkers in biological monitoring studies.2.1. Hatchability of domestic chicken eggs under different injection conditions. 292.2. Hatchability of domestic pigeon eggs under different injection conditions. 302.3. Injection conditions used in the in ovo dose-response study. 302.4. TCDD-toxic equivalency factors for PCDDs, PCDFs and PCBs. 403.1. Levels of PCDDs, PCDFs and PCBs and sum of TEQs in great blue heron eggs 443.2. Levels of PCDDs, PCDFs and PCBs and TEQ in double-crested cormorant eggs 463.3. Incidence of edema in great blue heron hatchlings. 483.4. Morphological measurements, hepatic microsomal cytochrome P-450 content and EROD 50activity in great blue heron hatchlings.3.5. Morphological measurements, hepatic microsomal cytochrome P-450 content and EROD 52activity in double-crested cormorant hatchlings.3.6. Linear regression of morphological measurements and hepatic microsomal EROD activity 53on TCDD and TEQ level in the great blue heron.3.7. Linear regression of morphological measurements, hepatic microsomal cytochrome P450 55content and EROD activity on TEQ level in the double-crested cormorant.3.8. In ovo duration of exposure to vehicle and3H-TCDD in great blue heron and double- 62crested connorants.3.9. Effect of3H-TCDD on egg mortality: comparison of four avian species. 633.10. Effects of3H-TCDD on body and organ weights in the domestic chicken hatchling. 643.11. Effects of3H-TCDD on body and organ weights in the domestic pigeon hatchling. 643.12. Effects of3H-TCDD on body and organ weights in the great blue heron hatchling. 653.13. Effects of3H-TCDD on body and organ weights in the double-crested cormorant hatchling. 653.14. Effect of3H-TCDD on incidence of edema: comparison of four avian species. 663.15. Increase in total hepatic microsomal cytochrome P-450 content by3H-TCDD: comparison 67of the in ovo dose-response curves in four avian species.3.16. Estimated no-observed-effect and lowest-observed-effect levels of3H-TCDD for hepatic 74EROD induction in the four avian species.3.17. Distribution of3H-TCDD among liver, yolk and blood in the domestic chicken hatchling. 76- vii -List of Tables3.18. Distribution of3H-TCDD among liver, yolk and blood in the domestic pigeon hatchling. 763.19. Distribution of3H-TCDD among liver, yolk and blood in the great blue heron hatchling. 773.20. Distribution of3H-TCDD among liver, yolk and blood in the double-crested cormorant 77hatchling.3.21. Linear correlations of blood with liver and yolk concentrations of3H-TCDD in the 78hatchlings of four avian species.3.22. Percentages of the injected doses of3H-TCDD found in livers and yolks of the hatchlings of 79four avian species.3.23. ED50 values for hepatic EROD induction by TCDD in the hatchlings of four avian species. 813.24. Sunimary of the Ah receptor binding characteristics of TCDD in four avian species 92- viii -LIST OF FIGURESFigure page1.1. The structures of several classes of 2,3,7,8-substituted polychiorinated aromatic 2hydrocarbons.1.2. Proposed mechanism of action of TCDD and structurally similar chemicals 71.3. Comparison of the ontogeny of Ah receptor and aryl hydrocarbon hydroxylase inducibility 10in the developing chick embryo.1.4. Effects of chlorine substitution at different positions on the dibenzo-p-dioxin and dibenzo 11furan rings on the relative binding affinities of PCDD and PCDF congeners for the Ahreceptor. Suggested dimensions of the Ah receptor binding site.2.1. Location of three great blue heron colonies on the Strait of Georgia, British Columbia. 262.2. Candling scheme for the development of the great blue heron embryo. 282.3. Hepatic EROD activity versus ethoxyresorufin concentration in the great blue heron chick. 342.4. Hepatic EROD activity versuspH in the great blue heron chick. 342.5. Hepatic EROD activity versus microsomal protein concentration in the great blue heron 35chick.2.6. Resorufin fonnation versus reaction time in the great blue heron chick. 353.1. Contribution of PCDDs, PCDFs and PCBs to the sum of TCDD-toxic equivalents in great 43blue heron eggs from three colonies in British Columbia, monitored from 1988 to 1992.3.2. Contribution of PCDDs, PCDFs and PCBs to the sum of TCDD-toxic equivalents in double- 47crested cormorant eggs from five colonies across Cana4a.3.3. Hepatic EROD activities in great blue heron hatchlings from three colonies in British 49Columbia, monitored from 1988 to 1992.3.4. Hepatic EROD activities in double-crested connorant hatchlings from five colonies across 51Canada.3.5. Regression of yolk-free body weight on TCDD level in great blue heron hatchlings. 543.6. Regression of hepatic EROD activity on TCDD level in great blue heron hatchlings. 573.7. Regression of hepatic EROD activity on TCDD level in great blue heron hatchlings from 58the Vancouver colony alone.3.8. Regression of hepatic EROD activity on TCDD-toxic equivalents in double-crested 60cormorant hatchlings.- ix -List ofFigures3.9. Immunochemical cross-reactivity of monoclonal rat CYP1A1 antibodies with a hepatic 61microsomal protein in great blue heron and double-crested cormorant hatchlingscontaminated with various levels of polychiorinated aromatic hydrocarbons.3.10. Increase in hepatic microsomal cytochrome P-450 content by3H-TCDD: comparison of the 68in ovo dose-response curves in the domestic chicken, domestic pigeon, great blue heron anddouble-crested connorant hatchling.3.11. Hepatic microsomal ethoxyresorufin 0-deethylase induction by3H-TCDD: in ovo dose- 69response curve in the domestic chicken.3.12. Hepatic microsomal ethoxyresorufin 0-deethylase induction by3H-TCDD: in ovo dose- 70response curve in the domestic pigeon.3.13. Hepatic microsomal ethoxyresorufin 0-deethylase induction by3H-TCDD: in ovo dose- 71response curve in the great blue heron.3.14. Hepatic microsomal ethoxyresorufin 0-deethylase induction by3H-TCDD: in ovo dose- 72response curve in the double-crested cormorant.3.15. The induction of avian hepatic microsomal EROD activity by3H-TCDD: comparison of the 73in ovo dose-response curves in the domestic chicken, domestic pigeon, great blue heron anddouble-crested cormorant hatchling.3.16. Immunochemical cross-reactivity of monoclonal rat CYP1A1 antibodies with a hepatic 74microsomal protein in domestic chicken and great blue heron hatchlings exposed in ovo tovarious doses of3H-TCDD.3.17. Hepatic3H-TCDD concentration versus injected dose: comparison of four avian species. 753.18. Percentage of the injected dose found in the liver of the great blue heron hatchling, after 80various durations of in ovo exposure to3H-TCDD.3.19. Percentage of the injected dose found in the liver of the double-crested cormorant hatchling, 80after various durations of in ovo exposure to3H-TCDD.3.20. Hepatic microsomal EROD activity versus hepatic3H-TCDD concentration: in ovo dose- 82response curve in the domestic chicken.3.21. Hepatic microsomal EROD activity versus hepatic3H-TCDD concentration: in ovo dose- 83response curve in the domestic pigeon.3.22. Hepatic microsomal EROD activity versus hepatic3H-TCDD concentration: in ovo dose- 84response curve in the great blue heron.3.23. Hepatic microsomal EROD activity versus hepatic3H-TCDD concentration: in ovo dose- 85response curve in the double-crested cormorant.3.24. Comparison of the specific binding of3H-TCDD in cytosolic fractions of adult human 86placenta and hatchling chicken liver, analysed by velocity sedimentation on sucrose densitygradients.-xList ofFigures3.25. The effect of dextran-coated charcoal on the specific binding of3H-TCDD in the liver 87cytosol of the domestic pigeon hatching.3.26. Specific binding of3H-TCDD in the 9-10 S region of domestic chick liver cytosol. 883.27. Specific binding of3H-TCDD in the 9-10 S region of domestic pigeon liver cytosol. 883.28. Specific binding of3H-TCDD in the 9-10 S region of great blue heron liver cytosol. 893.29. Specific binding of3H-TCDD in the 9-10 S region of double-crested cormorant liver 89cytosol.3.30. Specific binding of3H-TCDD in the 9-10 S region of adult great blue heron liver cytosol. 903.31. Saturation, Scatchard and Woolf plots for the specific binding of3H-TCDD in hepatic 91cytosol of the great blue heron hatching.4.1. Comparisons of the induction of CYP1A1 by environmental levels of TCDD in great blue 108heron hatchlings and TEQ in double-crested cormorant hatchlings with the respective inovo dose-response curves constructed in the laboratory.- xi -LIST OF ABBREVIATIONS3-MC 3-methylcholanthrene Kd apparent equilibrium dissociationconstantAh aryl hydrocarbon LD50 lethal dose for 50% ofanimalsAHH aryl hydrocarbon hydroxylase n value sample sizeANOVA analysis ofvariance NADPB /3-nicotinamide adenine dinucleotidephosphate (reducedform)Bmax apparent total number ofspecific binding p value level ofstatistical significance (type Isites error)BSA bovine serum albumin PCB polychiorinated biphenylCAT catalase PCDD polychiorinated dibenzo-p-dioxincpm counts per minute PCDF polychiorinated dibenzofuranCYPIA1 cytochrome P-450 lÀ] PROD pentoxyresorufin 0-deethylaseCYP2B1/2 cytochrome P-450 2B1 and 2B2 r value correlation coefficientDI day ofincubation r2 value coeficient ofdeterminationdpm disintegrations per minute SEM standard error ofthe meanED50 dose at which half-maximal effect is TCDD 2,3,7, 8-tetrachlorodibenzo-p-dioxinattainedEDTA ethylenediamine tefraacetic acid TCD1? 2,3, 7,8-tetrachlorodibenzofuranEOCL extractable organic chlorine TEF toxic equivalencyfactorEROD ethoxyresorujin 0-deethylase TEQ 2,3,7, 8-tetrachlorodibenzo-p-dioxin-toxicequivalentGC-MS gas chromatography-mass specfroscopy TRIS 2-amino-2-(hydroxymethyl)-1, 3-propanediolHEPES (N-[2-hydroxyethylJpiperazine-N’-[2- UDP-GT uridyldiphosphate-glucuronyltransferaseethanesulfonic acid])-xii -ACKNOWLEDGMENTSI am very grateful to Dr. Gail Beliward for her excellent supervision, expert advice andguidance throughout my graduate years. Thank you for making my graduate experience anenjoyable and successful one.I also thank the members of my graduate research committee, Drs. Stelvio Bandiera,Kimberly Cheng, Jack Diamond, Leslie Hart, Neil Hartman and Alan Mitchell for their input.Furthermore, I would like to thank Dr. Robert Thies for his time and effort in preparing slidesand scans. Finally, I would like to thank my friends in the Faculty for making my time inVancouver and in the graduate program a very enjoyable experience.- xiii -1. INTRODUCTIONEnvironmental contamination with persistent toxic chemicals is ofgreat concern. Inthe past, environmental contamination was generally a local problem. However, with theincrease of urbanization, industrialization and intensive agricultural practices, which havefollowed the dramatic increase in human population, persistent environmentalcontaminants have become a global concern. Increased levels of toxic metals, persistentpesticides and polyhalogenated aromatic hydrocarbons can now be found in all parts ofthe world.As concern for human and environmental health has increased, more attention hasbeen given to research in environmental toxicology. Increased knowledge of the sources,fate and toxicities ofchemicals in the environment is essential in order to identfy, examineand remediate environmental contamination problems, and to regulate future use ordisposal of the contaminant(s) in question.This thesis focuses on a spec/Ic contamination problem in the Strait of Georgia.Increased levels of2,3,7, 8-tetrachlorodibenzo-p-dioxin and other polychlorinated dibenzop-dioxins and dibenzofurans, mainlyfrom the pulp andpaper mills in the Strait, are beingassociated with the decline in populations of coastal fish-eating birds, such as the greatblue heron (Elliott et al., 1989) and double-crested cormorant. This thesis demonstratesthe use of these fish-eating birds as biological monitors of environmental contaminationwith polychiorinated aromatic hydrocarbons and examines in greater detail thecomparative sensitivity ofseveral bird species to 2,3,7, 8-tetrachlorodibenzo-p-dioxin.1 .A. TOXICOLOGY OF 2,3,7, 8-TETRACHLORODIBENZO-P-DIOXINAND RELATED POLYHALOGENA TED AROMATIC HYDROCARBONS1.A.1. Sources of Polyhalogenated Aromatic HydrocarbonsThe sources of a large number of polyhalogenated aromatic hydrocarbons have beenreviewed by Brinkman and De Kok (1980) and De Voogt and Brinkman (1989). Thesechemicals are almost entirely of anthropogenic origin.-1-IntroductionThe chemical 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is a member of thepolychlorinated dibenzo-p-dioxins, which consists of 75 congeners. Polychlorinated dibenzop-dioxins (PCDDs) and the structurally similar polychlorinated dibenzoftirans (PCDFs)(Figure 1.1), of which there exist 135 possible congeners, were never produced forcommercial purposes. They are produced as trace contaminants in the synthesis of certainchlorinated chemicals, such as chlorophenols, which are used as wood preservatives,insecticides and fungicides, and 2,4,5-trichlorophenoxyacetic acid, which is used widely as abroad-spectrum herbicide and defoliant. Trace amounts of PCDDs and PCDFs are producedby pulp and paper mills that use chlorine bleaching. PCDDs and PCDFs are also producedduring combustion processes involving chlorinated organic and inorganic chemicals; theseinclude automobile exhaust fumes, hospital and municipal waste incinerators and, to a muchlesser extent, natural sources such as forest fires.CCl2,3,7,8 -Tetrachlorodibenzo -p -dioxinc.Cl33(44’— TetrachloroozoxybenzeneCICl2,3,6,7 —Tetrochloronaphtholene 2,3 ,6,7—TetrachlorobiphenyleneFigure 1.1. The structures of several classes of laterally-substituted polychiorinated aromatic hydrocarbons(Poland and Knutson, 1982).9 0 IClC02,3,7,8 —Tetrochorodbenzofuron3,3’, 4,4’-Tetrachlorobiphenyl5 4-2-IntroductionPolychiorinated biphenyls (PCBs) are structurally related to the PCDDs and PCDFs;theoretically, there exist 209 congeners. PCBs have been manufactured commercially andused world-wide for a variety of purposes. The properties of PCBs, such as thermal stability,low flammability, high electrical resistance and chemical inertness made them ideal for use astransformer fluids and heat-transfer fluids, lubricants and plasticizers. Commercial PCBmixtures are known to contain trace amounts ofPCDFs.Other structurally related chemicals have been detected in the environment.Polybrominated biphenyls have been used commercially as flame retardants in syntheticmaterials. Polyhalogenated naphthalenes have been produced commercially, but productionhas decreased since the 1950s.l.A.].]. Sources ofFolyhalogenatedAromatic Hydrocarbons in the Strait of GeorgiaIn the Strait of Georgia, the polychiorinated aromatic hydrocarbons, such as PCDDs,PCDFs and PCBs, are the major polyhalogenated aromatic hydrocarbons contaminants,because of the predominant presence of forest-related industries. The Strait of Georgia is thelocation of nine large Kraft pulp and paper mills and numerous other lumber processingindustries. Pulp and paper mills that use high concentrations of molecular chlorine in thebleaching process produce trace amounts of PCDDs and PCDFs, which are discarded in theeffluent (Kuehl et a!., 1987). Furthermore, the wide use of chlorophenol-treated lumber in theforest industry has added to the environmental contamination with PCDDs and PCDFs (Elliottet a!., 1989). The presence of PCBs is also of concern. The sources of PCBs in the Strait ofGeorgia have not been studied in any detail. However, PCB contamination is typicallyassociated with the human activities in urban areas, such as Vancouver and Victoria. Likelyurban sources of PCBs, and also PCDDs and PCDFs in the Strait of Georgia, includeautomobile exhaust fumes, waste incinerators (Rappe et at., 1989), and leaking transformersand capacitors. With the forecasted rapid growth of these urban areas, levels ofcontamination with PCDDs, PCDFs and PCBs are expected to increase. The followingsections focus on the toxicities of the polychiorinated aromatic hydrocarbons, but it must bekept in mind that these toxicities are produced by a larger group of halogenated aromatichydrocarbons.-3-Introduction1.A.2. Toxicities of Polychiorinated Aromatic HydrocarbonsLaboratory studies have reported numerous effects of certain PCDDs, PCDFs, PCBsand other related halogenated aromatic hydrocarbons in mammalian and avian species(Kimbrough, 1974; Poland et al., 1979; Kociba and Schwetz, 1982; Poland and Knutson,1982; Safe, 1984; BrunstrOm and Reutergárdh, 1986; Safe, 1986; Nikolaidis et al., 1988;Nosek et cii., 1 992a). These include weight loss, edema, hepatotoxicity, immunotoxicity (e.g.thymic atrophy), decreased reproductive success, teratogenicity, promotion of cancer, hepaticporphyria and enzyme induction. This range of toxic responses is mediated by a commonmechanism of action, in which binding to the cytosolic aryl hydrocarbon (Ah) receptor is aprerequisite (Bandiera et a!., 1982; Poland and Knutson, 1982). However, the fill spectrumof toxicities is not usually elicited in any one given species. There are considerable sex, age,strain and species differences in the resultant toxic effects and in the toxic potency of thesechemicals. The overt toxicities common to most laboratory animals are the immunotoxiceffects and body weight loss caused by PCDDs, PCDFs and PCBs, On a biochemical level,the induction of cytochrome P-450 1A1 (CYP1A1) is a common and sensitive response tothese polychlorinated aromatic hydrocarbons. Chloracne and other dermal toxicities arecharacteristic of polychlorinated aromatic hydrocarbon toxicity in humans, Rhesus monkeys,rabbits and hairless mice (Poland and Knutson, 1982). On the other hand, the mostcharacteristic response in avian species, such as the domestic chicken, is the formation of asubcutaneous jelly-like’ edema (Flick et a!., 1972; Norback and Allen, 1973; Firestone, 1973).The entire suite of toxicities of polychlorinated aromatic hydrocarbons in avian species hasoften been termed ‘chick-edema disease’ in the past, because its manifestations were firstdocumented in chickens exposed to polychiorinated dibenzo-p-dioxins via contaminated feedin 1957 (Firestone, 1973).1.A.3. Mechanism of Action of TCDD and Related ChemicalsThe mechanism of toxicity of TCDD and related chemicals has received a great deal ofattention in recent years and has been reviewed extensively (Goldstein and Safe, 1989; Safe,1986; Landers and Bunce, 1991). TCDD does not bind covalently to DNA or other cellularmacromolecules (Poland and Glover, 1979), nor does it undergo metabolic activation to a-4-Introductionreactive intermediate (Neal et a!., 1979). Instead, the toxic potency of the polychlorinatedaromatic hydrocarbons, correlates closely with the binding affinity of these chemicals for thehepatic Ah receptor (Bandiera et at., 1984; Mason et a!., 1985; 1986; Safe, 1986). TCDDappears to be the most toxic of this group of chemicals, and it has the highest affinity for theAh receptor found so far. For these reasons, TCDD is generally used by investigators as arepresentative for the whole group of polyhalogenated aromatic hydrocarbons with affinity forthe Ah receptor.l.A. 3.1. Genetic Regulation of the Induction of Cytochrome P-450 lÀ] - Evidencefor theExistence ofan Aryl Hydrocarbon ReceptorA well documented effect of TCDD and related chemicals is the ability to induce hepaticCYP1A1 and CYP1A2 (Poland and Glover, 1973; Poland et at., 1974; Bradlaw andCasterline, 1979; Yoshihara et at., 1981; Nagayama et at., 1983; Rifkind eta!., 1984). Theseisoenzymes are relatively selectively responsible for ethoxyresorufin 0-deethylase (EROD)activity found in micro somal preparations of mammalian and avian livers (Burke and Mayer,1974; Kaminsky et a!,, 1983; Parkinson et a!., 1983; Brunström and Andersson, 1988; Lubetet a!., 1990). There exists a strong correlation between the toxicity of polychiorinatedaromatic hydrocarbons and their ability to induce hepatic CYP1A1 (Poland and Glover, 1973;Poland et a!., 1979). Nebert et at. (1972) and Thomas et a!. (1972) disovered that certaininbred strains of mice responded differently to CYP1A1 induction by 3-methylcholanthrene(3-MC). 3-MC, a less potent inducer of CYP1A1 and CYP1A2 than TCDD, was able toinduce CYP1A1 in C57BL16 mice (termed ‘responsive’), but not in DBAJ2 mice (termed ‘non-responsive’). This responsiveness appeared to be a simple autosomal dominant trait,genetically controlled by the aryl hydrocarbon (Ah) locus. The allele for responsiveness isdesignated Ahb and for non-responsiveness Ahd. TCDD was able to induce CYP1A1 in the‘non-responsive’ DBA/2 mice, but the ED50 (10 nmollkg) was approximately ten times higherthan that in the ‘responsive’ C57BL/6 mice (1 nmollkg). Furthermore, it was observed that theED50 for other TCDD-mediated toxicities, such as body weight loss, thymic atrophy, hepaticporphyria, production of cleft-palate, and the LD50 were also about ten times higher in theDBAJ2 mice (Poland and Knutson, 1982). Using3H-TCDD in ligand binding studies, Poland-5-Jnfroductionand coworkers (1976) detected a protein in hepatic cytosol of C57BL/6 mice with all theproperties of a pharmacological receptor (Table 1.1) and termed it the aryl hydrocarbon (Ah)receptor. Only recently, Okey eta!. (1989) have been able to detect an Ah receptor in DBAI2mice. In line with the 10-fold lower potency of TCDD in this strain, the affinity for the Ahreceptor was 10 times lower in the DBAI2 mouse. These two mouse strains, differing only inthe gene that codes for the Ah receptor (the Ah locus), have been of valuable use in manystudies aimed at investigating the role of the Ah receptor in TCDD-mediated toxicities.Table 1.1. Criteria for the involvement of an aryl hydrocarbon receptor in the mediation of the toxicities of2,3,7,8-tetrachlorodibenzo-p-dioxin and other related chemicals.1. Existence of a finite number of specific binding sites (Bm; i.e. saturable) in thefmole/mg range.2. High-affinity ligand binding (in the nM range) similar to that of sex-hormones. (TCDDshould have an equilibrium dissociation constant (Kd) which corresponds to the in vivoED50 of CYP1A1 induction.)3. Stereoselective binding of the ligand to the receptor.4. Specificity in tissue and organ response to the receptor ligand.5. Correlation between receptor affinity or receptor occupancy, and the magnitude of theresponse. (3-MC and other polycyclic aromatic hydrocarbons that induce CYP1A1should also bind to the receptor.)Adapted from Poland and Knutson (1982) and Safe (1986).l.A. 3.2. Ah Receptor Characteristics and FunctionThe unoccupied Ah receptor is a cytosolic protein, which is associated with the heat-shock protein, hsp 90, and possibly other proteins; it has a combined molecular mass of about270 kDa. The molecular mass of the ligand-binding subunit is about 95 kDa in mice (Polandand Glover, 1987; 1990; Poland et a!., 1991). Molecular masses vary considerably amongrodents, ranging from 95 kDa in the mouse to 124 kDa in the hamster (Poland and Glover,1987) and 130 kDa in the deer mouse (Peromyscus maniculatus) (Poland eta!., 1991).-6-IntroductionUnoccupied Ah receptor cannot interact with DNA. Once occupied by a ligand,however, the ligand-Ah receptor complex translocates to the nucleus where it specificallyrecognizes Mi-responsive elements (AhREs) on the DNA (Figure 1.2). These regions are alsoknown as dioxin-responsive elements/enhancers (DREs) or xenobiotic regulatory elements(XREs). Nuclear ligand-Ali receptor complex is associated with a protein called the Allreceptor nuclear translocator (Arnt) protein (Reyes et al., 1992) and has a molecular mass ofabout 175 kDa (Prokipcak and Okey, 1988). The approximate molecular mass of the Arntprotein is 87 kDa (Reyes et al., 1992) and of the ligand-binding subunit is 95 kDa. Combined,the masses correspond closely to the 175 lcDa mass of the nuclear receptor. This indicatesthat no other proteins appear to be associated with the nuclear form of receptor. Theassociation of Arnt appears to be required for the Ah receptor to translocate to the nucleus.TCDD,[MC, etc.]IN CR EASEDMETABOLISM ‘OF DRUGS ANDENVIRONMENTALCHEMICALS9,TOXICITY -4--——Figure 1.2. Proposed mechanism of action of TCDD and structurally similar chemicals. TCDD enters thecell by passive diffusion and binds to the Ah receptor (AhR); the ligand-receptor complex associates with theAhR nuclear translocator (Arnt) and translocates to the nucleus, where it interacts with Ah-responsiveelements (AhRE) on the DNA. This leads to increased or decreased transcription of specific mRNAs, resultingin altered protein synthesis and a number of biological and toxicological responses (Okey et at., 1994).-* TCDDINCREASEDCYP1A1— INCREASEDCYP1A2CELL GROWTH &DIFFERENTIATION-7-IntroductionOnce the nuclear Ah receptor-ligand complex binds to Ah-responsive elements on theDNA, it leads to the increased or decreased transcription of specific mRNAs, which aresubsequently translated into proteins. The best characterized response that has shown to bedirectly mediated by the Ah receptor is the induction of CYP1A1 (Nebert and Gonzalez,1987). Other gene products that are responsive to TCDD include CYP1A2 (Quattrochi andTukey, 1989), UDP-glucuronosyl transferase, epidermal growth factor, plasminogen activatorinhibitor-2 and interleukin-1f3 (reviewed by Okey et at., 1994). Recently, TCDD has beenshown to increase protein tyrosine phosphorylation (Clark et at., 1991; DeVito et a!., 1994)and protein tyrosine kinase activities (Bombick et at., 1988), which play an imporatant role incell growth and differentiation. It is not known how specific changes in gene expression,induced by TCDD and similarly acting chemicals, lead to toxicities. It is likely that theinduction of CYP1A1, for instance, is not directly responsible for most Ah receptor-mediatedtoxicities, but is part of a pleiotropic response (Poland and Knutson, 1982).So far, it has not been possible to isolate the intact receptor, because of its instability.However, Bradfield et at. (1991) have been able to determine the partial amino acid sequenceof the N-terminal region of the ligand-binding subunit of the Ah receptor in C57BL/6J mice.This has allowed for the cloning of mouse Ah receptor (Burbach et a!., 1992; Ema et at.,1992). These cloning experiments have indicated that the Ah receptor does not have ‘zinc-fingers’ as binding domains for DNA, which is a property of the steroid hormone receptors.Therefore, the Ah receptor does not belong to the steroid receptor superfamily as was thoughtpreviously (Evans, 1988). Instead, it is a member of the helix-loop-helix DNA-bindingproteins (Burbach et at., 1992; Ema et at., 1992).It is speculated that the Ah receptor plays an important physiological role in cell growthand differentiation, particularly during fetal development. Although an endogenous ligand forthe Ah receptor has not been identified at this time, it is regarded likely by a number ofinvestigators that such a ligand exists (Nebert, 1991). At present, clues are being sought insteroid structures (Beresford, 1993) and derivatives of 5 -hydroxytryptophan (Ranug et at.,1987; Helferich and Denison, 1991).-8-IntroductionSpecific ligand-binding characteristics of the Ah receptor have been investigated usingvarious methods. The high affinity of TCDD has made it the ligand of choice for measuringAh receptor levels and affinities in different tissues and organisms. The saturable binding of3H-TCDD (with a high specific activity) in cytosolic fractions is commonly measured byvelocity sedimentation analysis on sucrose density gradients (Okey et a!., 1979), which is themethod of choice in this thesis, or by hydroxylapatite adsorption (Gasiewicz and Neal, 1982;Poellinger eta!., 1985).l.A. 3.3. Presence ofAh Receptor in Avian SpeciesThe Ah receptor was first detected in the mouse (Poland and Knutson, 1982) andsubsequently in several other rodent species (Denison et a!., 1986a). Ah receptor has alsobeen detected in a rainbow trout hepatoma cell line (Lorenzen and Okey, 1990). Attempts todetect Ah receptor have been made in a limited number of avian species. The chicken embryowas the only bird species to show significant receptor levels (Denison et a!., 1986a;BrunstrOm and Lund, 1988). Denison et a!. (1986a) mentioned low levels of specific bindingin quail, but did not specifj the species used or show any data. Ah receptor was not detectedin the turkey embryo (Meleagris gallopavo) (Brunstrom and Lund, 1988), pigeon (Columbalivia) and Jungle fowl (species of Jungle fowl not indicated) (Denison eta!., 1986a).Hepatic Ah receptor is present in the chick embryo throughout development (Denison eta!., 1 986c). Highest levels (40-60 fmole/mg cytosolic protein) are present at the early stagesof morphological differentiation, from day 5 to 9 of incubation (Figure 1.3). These decreasedto levels of about 15 nmole/mg protein which are maintained throughout the the rest ofdevelopment and after hatch. Consistent with the presence of Ah receptor, aryl hydrocarbonhydroxylase (AHH, another CYP1A1-associated activity) inducibility was present in the chickembryonic liver from day 6 or 7 and was maintained throughout development and after hatch.-9-IntroductionFigure 1.3. Comparison of the ontogeny of Ah receptor and AHH inducibilty (by 3,4,3’,4’-tetrachlorobiphenyl)in the developing chick embryo (Denison eta!., 1986c).l.A. 3.4. Structure-Activity RelationshipsStructure-activity relationships have been developed for a number of polychiorinatedaromatic hydrocarbons. PCDDs and PCDFs have the highest affinities for the Ah receptor,when substituted with chiorines at all four lateral (2,3,7 and 8) positions (indicated by plussigns in Figure 1.4) (Poland et at., 1976; Bandiera eta!., 1984; Mason et al., 1985; Mason eta!., 1986; Safe, 1986). In these studies, the binding affinities of the chemicals were notdetermined directly; they were derived from the relative abilities of the chemicals to displacespecific binding of3H-TCDD from the Ah receptor in rat liver cytosol. Increased chlorinesubstitution at the other positions, or substitution at less than the four lateral positions,decreases the binding affinities of the PCDDs and PCDFs (Figure 1.4). Certain PCBs with theability to attain a coplanar conformation, such as the non-ortho-PCBs and mono-ortho-PCBs,also have affinities for the Ah receptor (Bandiera et a!., 1982; Safe, 1984). Mono-orthoPCBs, for which the coplanar conformation is energetically less favourable, have lowerC>I0II5 79 1113151719h 2468embryo neonateDAYS OF DEVELOPMENT-10-Introductionbinding affinities for the Ah receptor. Di-ortho-PCBs, which are unable to reach a coplanarconformation due to steric hindrance, have negligible affinities for the Ah receptor. Severalvariables influence the binding affinity of the polychlorinated aromatic hydrocarbons. Theseare steric conformation (i.e. molecular dimesion and planarity) and lipophilicity, and to a lesserextent electron density of the aromatic ring and hydrogen bonding capacity.1< IOAI x — — — — — - — xI I I 3AI _.... _! _. _i_,. _.i..Lr — — — — -x 2Figure 1.4. Top: Differential effects of chlorine substituents, at different positions in the dibenzo-p-dioxin anddibenzofuran molecules, on the relative receptor binding affinities of the PCDD and PCDF congeners (Safe,1986). Bottom: the putative dimensions (3x10 A) of the binding site on the cytosolic Ah receptor, derivedfrom the structure-affinity relationships of the PCDDs, PCDFs and PCBs (Poland and Knutson, 1982).Within each chemical class, there exists a correlation between the receptor affinity andAHH or EROD-inducing potency of the PCDDs (Poland and Glover, 1980; Mason et al.,1986), PCDFs (Bandiera et at., 1984; Mason et at., 1985) and PCBs (Bandiera et at., 1982;Leece et at., 1985). Furthermore, the in vitro AHH or EROD-inducing potencies of a numberof PCDDs and PCDFs in H4IIE rat hepatoma cells could be correlated with in vivo toxicitiessuch as body weight loss and thymic atrophy (summarized by Safe, 1990). In all structure-activity/toxicity studies, TCDD appeared to be the most toxic chemical and the most avidbinder to the Ah receptor (Safe, 1986; 1990). The large number of structure-activity studiesof the polychlorinated aromatic hydrocarbons has allowed Safe (1990) to derive toxic+ ÷- 11 -Jntroductionequivalency factors (TEFs) for the PCDDs, PCDFs and PCBs. These TEFs are based onmammalian studies, including human epidemiological studies, and take numerous end-pointsinto account (e.g. receptor binding affinity, CYP1A1 induction, body weight loss,immunotoxicity, teratogenicity). TEFs rank the polychiorinated aromatic hydrocarbons thatbind to the Ah receptor in order of toxic potency, relative to that of TCDD. The potency ofTCDD is set at unity; the TEFs for the other PCDDs, PCDFs and PCBs, which are less toxic,are a fraction of unity. These TEFs can be used to calculate TCDD-toxic equivalents (TEQs)by multiplying the concentration of the chemical in question by its TEF value. These TEQsare usefll for toxicological assessments of complex mixtures of polychiorinated aromatichydrocarbons in the environment (discussed in section 4.A.3).1.A.4. Environmental Toxicities of Polychiorinated Aromatic Hydrocarbons in BirdsEnvironmental levels of certain PCDDs, PCDFs and PCBs are strongly implicated as thecausative factors in the so-called Great Lakes embryo mortality, edema and deformitiessyndrome (GLEMEDS), observed in double-crested cormorants (Phalacrocorax auritus),herring gulls (Larus argentatus), Forste?s terns (Sternaforsteri) and other fish-eating birds inthe Great Lakes basin (reviewed by Gilbertson et at., 1991). Gross symptoms of GLEMEDSinclude increased embryo and chick mortality, growth retardation, the occurrence of jelly-like’edema, congenital deformities, feminization of embryos and abnormal parental behaviourregarding care of eggs and chicks. On a biochemical level, the syndrome includes theinduction of hepatic CYP1A1 and hepatic porphyria, and decreased thyroid function andhepatic retinoid levels. These toxicities are consistent with those observed in laboratorystudies. Taken together, these effects are detrimental to the reproductive success of avianspecies and have led to declines in populations of several species of fish-eating bird in theGreat Lakes. Similar effects have been associated with elevated levels of polychlorinatedaromatic hydrocarbons in San Francisco Bay (Hoffman et at., 1986; Ohlendorf and Fleming,1988) and the Rhine estuary of the Netherlands (Van den Berg et at., 1987; 1992). Of thepolychlorinated aromatic hydrocarbons, which are full or partial agonists of the Ah receptor,the 2,3,7,8-substituted PCDDs and PCDFs, the non-ortho-PCBs, PCB-77, 126 and 169, andthe mono-ortho-PCBs, PCB-105 and 118 can be found in relatively high concentrations in- 12 -Introductionfish-eating birds in Canada (Norstrom, 1988; Beliward et a!., 1990). In the Netherlands, thepattern of PCB contamination is slightly different: the major mono-ortho-PCB contaminantsare PCB-105, 118, 156, 157 and 167 (Van den Berg eta!., 1992; Bosveld eta!., 1993).Recently, several sublethal effects associated with environmental levels ofpolychiorinated aromatic hydrocarbons, particularly TCDD, have been reported in great blueherons exhibiting reproductive failure in British Columbia (Elliott et a!., 1989; Beliward et a!.,1990; Hart et a!., 1991). The most prominent effects observed in these herons were increasedhepatic EROD activities and incidence of edema, growth retardation and a decreasedreproductive success (measured by number of fledglings) in birds from the most contaminatedcolony. The results of the continued monitoring of these great blue herons are presented inthe environmental monitoring section (section 3 .A) of this thesis.1. B. BIOLOGICAL MONITORING OF ENVIRONMENTALCONTAMINATION1.B.1. Terms used in the Field of Biological MonitoringDefinitions of terms used in the field of biological monitoring have been formulated byseveral researchers (reviewed by O’Brien et a!., 1993). Distinctions have been made amongterms such as biological indicator, monitor and sentinel species.Biological Indicator: An organism whose characteristics are used to point out the presenceor absence of environmental conditions.Biological Monitor: An organism in which changes in known characteristics can be measuredto assess the extent of environmental contamination so that conclusions on the health impactfor other species or the environment as a whole can be made.Sentinel Species: An organism in which changes in known characteristics can be measured toassess the extent of environmental contamination and its implications for human health and toprovide an early warning of those implications.Biomarker: A known characteristic which responds to environmental contamination and canbe used to determine the extent of exposure of a species chosen for biological monitoring.- 13 -IntroductionIn the case of the great blue herons and double-crested cormorants examined in thisthesis, they can be categorised as biological monitors and possibly sentinels; hepatic ERODactivity can be referred to as a biomarker.1.B.2. Why use Sentinel Species?The main question that is asked of environmental contamination is whether it isadversely affecting the health of the environment as a whole, and humans, in particular. Untilrecently, the emphasis of assessing environmental contamination has been on performingchemical analyses of compartments such as soil, water and air. Knowledge of levels ofcontamination in the environment is essential, but has many limitations. Chemical levels per sedo not provide any information on the bioavailability of the contamination to the livingsystems present in the environment. Also, they are not able to assess the toxicological impactsuch levels have on these living systems, which may differ considerably in susceptibility. Howcan the impact of a certain level of contamination in the environment on wildlife and humansbe assessed? The problems with assessing toxicities of environmental contamination inhumans are obvious: humans cannot be exposed in controlled experiments for ethical reasons.Furthermore, knowledge of contamination levels and subtle biological effects in humans arecomplicated by the lack of availabilty of tissues. Epidemiological studies, although veryuseful, have significant limitations. Also, toxicological assessments of environmentalcontamination in humans do not provide information on the health of other species, withwhich we share the environment. The use of one or several sentinel species is a practicalmethod that avoids these problems to a certain extent.1.B.3. Historical Use of Sentinel SpeciesA classical application of an animal as a biological indicator is the use of canaries bymineworkers. Canaries are more sensitive to carbon monoxide poisoning than humans, whichmade them early-warning indicators of increasing CO concentrations in the poorly ventilatedmines (National Research Council, 1991). There are also examples of the use ofepidemiological observations made in domestic animals such as pets and cattle, which sharethe environment closely with humans. One of the first observations of the toxic effects of-14-Introductionsmog (smoke from industrial sources mixed with fog) were made in cattle that died during anepisode of heavy smog in London, United Kingdom. Such events were later associated withhuman deaths in London (National Research Council, 1991).1.B.4. Criteria for a good sentinel species:Criteria for the use of sentinel species to monitor environmental contamination havebeen given by Gilbertson et al. (1987) and O’Brien et al. (1993):1. Size: Sufficiently large to allow adequate tissue samples.2. Longevity: Sufficiently long-lived to provide information over a prolonged period of timeand to be able to accummulate environmental contaminants.3. Availability: The animal should be numerous and sampling should have a negligibleimpact on the total population and population dynamics.4. Sensitivity: First of all, the animal should be able to express the effects that one hasdecided to measure. The animal should be sensitive enough for environmentally relevantlevels to have measurable effects. It should, however, not be so sensitive that the specieswould disappear from the area of study at a relatively small increase in contamination.5. Residency: The animal should be a year-round resident in the area of contamination.6. Distribution: The animal should have a broad ecological and geographical distribution sothat comparisons can be made over a large number of locations.7. Type of Contamination: The choice of sentinel also depends on the type ofcontamination in question. Fish and shell fish would be appropriate for waterbournecontaminants; land mammals such as voles would be suitable for airbourne or groundcontaminants. Persistent lipophilic chemicals with the potential to bioaccumulate andbiomagnif’ are better monitored in longer-lived species at the top of the food chain, suchas larger fish and fish-eating birds.1.B.5. Great Blue Herons and their Eggs as Sentinels:Birds have been recognized as useful sentinels of environmental contamination(Gilbertson et al., 1987). In 1974, the herring gull was proposed by the Canadian WildlifeService as a sentinel of organochiorine contamination of the Great Lakes (Mineau et a!.,-15-Introduction1984). The National Contaminant Biomonitoring Program of the U.S. Fish and WildlifeService proposed the use of black-crowned night herons as biological indicators of thechemical contamination of estuaries and wetlands (Custer et al., 1991).Since the 1970s, the Canadian Wildlife Service has been using the great blue heron as asentinel of environmental contamination with polychiorinated aromatic hydrocarbons in theStrait of Georgia. Its size, availability and lifespan fit the criteria mentioned in section 1 .B .4.The great blue heron is a wading bird, which feeds on small inshore fish (Butler 1991), placingit at a high trophic level within the food chain. For this reason, the heron and other fish-eatingbirds biomagnify highly persistent, lipophilic chemicals, such as polychiorinated aromatichydrocarbons. The heron is also relatively long-lived and bioaccummulates these chemicals.After reaching breeding age, the heron is a year-round resident of the Strait of Georgia.Breeding colonies are located in large trees in estuarine or coastal areas with large eelgrassbeds. The female heron lays three to four eggs per clutch. In females, these chemicals arepassed on to the egg by deposition into the yolk. In 1977, the Canadian Wildlife Servicestarted monitoring polychlorinated aromatic hydrocarbon levels in great blue heron eggs(Whitehead, 1989; Elliott et al., 1992), and more recently, in double-crested cormorant eggs.The use of bird eggs in biological monitoring has certain advantages: they are easy to collectand relatively large numbers can be obtained without an impact on the population. Impact onthe population is offset by the fact that competition for food among hatchlings is reduced ifonly one or two eggs are taken from a clutch of four (fledging success is usually about 50% inthe wild). Another advantage of eggs is that they can be incubated artificially for examinationof teratogenic and other adverse effects in the embryos. Eggs can also be used for controlledlaboratory studies, such as the in ovo dose-response study in this thesis (Results, 3.B).However, one must be aware that not all persistent chemicals are transferred from adult to theegg or that levels in the egg do not represent the degree of exposure; this appears to be thecase for certain toxic metals such as cadmium, chromium and lead (Leonzio and Massi, 1989).-16-Introduction1.B.6. Use of CYP1A1 Induction in Biological Monitoring1.B. 6.1. IntroductionIn the search for suitable biological effects (biomarkers) to measure in response tocontaminant exposures in biological monitoring programs, enzyme induction has been studiedextensively (reviewed by Payne, 1984; Payne eta!., 1987; Rattner eta!., 1989; Peakall, 1992).Alterations in enzyme activities are often sensitive, specific and rapid responses to certainchemicals. Furthermore, many enzymes play critical roles in detoxication and bioactivationprocesses and in homeostasis. Therefore, effects on these enzymes would be expected to haveadverse effects (section 1 .B .6.4). These are important criteria in order for a biological effectto be of use as an early warning of an environmental health hazard. Examples of enzymeactivities that arealtered by exposures to certain environmental chemicals are listed in Table1.2.Table 1.2. Examples of enzyme systems affected by environmental contaminants and used as biomarkers inbiological monitoring studies.Environmental Enzyme system(s) Effect ReferenceContaminant(s)Lead aminolevulinic acid inhibition Scheuhammer,dehydratase 1987.DDT, Mirex, toxaphene CYP2B1/2 induction of aldrin Payne et at., 1987.di-ortho-PCBs epoxidaseCertain polyhalogenatedaromatic hydrocarbons CYP1A1/2 induction of EROD Payne eta!., 1987.(i.e. andAHHPCDDs, PCDFs, nonortho- PCBs, mono-orthoPCBs)Organophosphate and acetyicholinesterase inhibition Hardy et at., 1987.carbamate pesticidesAdapted from Peakall, 1992.-17-IntroductionThe induction of CYP1A1 is used as a biomarker of exposure to polychiorinatedaromatic hydrocarbons that activate the Ah receptor for several reasons (some from DeVito eta!., 1994):1. It is a relatively sensitive response to exposure to these polyhalogenated aromatichydrocarbons.2. It is a rapid response to polyhalogenated aromatic hydrocarbons.3. It is relatively selective response to these polyhalogenated aromatic hydrocarbons andcertain polycyclic aromatic hydrocarbons (i.e. benzo[a]pyrene).4. It is an easily quantifiable response, which can increase greatly over a given dose-range.5. It is a response common to most eukaryotic organisms examined to date.6. It is of toxicological relevance (section 1.B.6.4).1.B. 6.2. Use of CYPJA] Induction in BiologicalMonitoring StudiesRattner eta!. (1989) suggested that the induction of cytochrome P-450 enzymes such asCYP1A1 could be used as biomarkers of exposure to persistent polycyclic andpolyhalogenated aromatic hydrocarbon in birds and other wildlife. CYP1A1-mediatedactivities, such as EROD and AHH are already applied widely as biomarkers of exposure tothese chemicals in fish (Lee et a!., 1980; Payne, 1984; Payne et al., 1987). However, despitethe extensive use in fish, detailed examinations of the nature and/or levels of contaminationand their relationship to the biological responses measured in individual organisms are lacking.Also, these biomonitoring studies are rarely continued over time. In order to strengthen anyconclusions regarding cause-effect relationships, it should be established whether long-termchanges in levels of environmental contaminants (i.e. over a period of a few years) arereflected in changes in the biological responses measured in the monitored species. To myknowledge there are no systematic studies that correlate levels of specific contaminants withbiological effects in individual organisms and continue to monitor levels and effects over time.Andersson et a!. (1988) monitored perch (Percafluviatilis) for the impact of bleachedkraft mill effluent from a pulp and paper mill on the Baltic coast of Sweden. They determinedextractable organic chlorine (EOCL) content of perch, a relatively stationary fish, from foursites at increasing distances from the mill (2-10 km) and compared these levels with several-18-Introductionbiological responses in the fish. A gradient in EOCL content existed (300-50 mg/kg), whichdecreased with increased distance. Similarly, effects on certain biological responses followedthe pollution gradient. Most of these effects were not specific for a particular contaminant(e.g. blood glucose, muscle glycogen and blood lactate level) and could have been caused byvariations in environmental conditions, such as current, water temperature and the availabilityof nutrients. The only relatively specific effect that appeared to be distance-related (andEOCL-related) was the induction of EROD. A study such as this indicates that bleached kraftmill effluent may be affecting normal physiological functions in perch. However, the nature ofthe contamination is poorly characterized (what is in EOCL?). Furthermore, pooled chemicalanalyses do not account for the ofien- considerable inter-individual variabilty in exposure andconsequently, in variability in the biological effects with which one is trying to associate theexposure.A study by Hodson et al. (1992) is slightly more detailed in approach. They determinedbiological effects in white suckers (Catostomus commersoni) at one site upstream and foursites at increasing distances downstream of a bleached kraft pulp mill on the St. Maurice riverin Quebec. A small number of fish were chosen for specific chemical analyses, in order tocorrelate levels of selected PCDDs and PCDFs with the observed effects in individual fish.Similar to the Swedish study, levels of these chemicals decreased with distance downstreamfrom the site of the mill, Effects on several measured parameters followed the profile of thepollution gradient. However, just like in the Swedish study, many of the measured responseswere nonspecific (e.g. hematocrit, serum glucose, liver-somatic index). Hodson andcoworkers observed a correlation between AHH activity and the levels of measured PCDDsand PCDFs (r=0.71-0.72; n=10), and also the TEQ derived from these chemicals (r=0.73;n=10), using TEFs based on rainbow trout data (Walker and Peterson, 1992). Thisrelationship is based on a relatively small sample size, and the significance level, which isprobably p<O.05, is not explicitely stated. The authors admit that this relationship is notstrong and is mainly determined by two samples of low TEQ content. Furthermore themethod by which AHH activities were determined is questionable. Activities were determinedin S-9 fractions of liver homogenates, which contains both cytosolic and microsomal enzymes,as opposed to microsomes, which contain mainly cytochrome P-450 embedded in endoplasmic-19-Introductionreticulum. Also activities were expressed in fluorescence units/mg proteinl2o mm, indicatingthat the reaction was carried out for 20 mm. In our experience, induced CYP1A1 activitiesare constant over only a couple of minutes of reaction time and then decrease rapidly.Measuring induced enzyme activities over a longer time period would result in anunderestimation of the actual activity. The use of fluorescence units, instead of determiningamounts of formed product using standard curves, does not account for differences insensitivity of the spectrofluorometer. If the method of the enzymatic assay is scientificallyincorrect this raises concerns regarding the rest of the study.The two studies described above are examples of the more recent and elaborateapplications of CYP1A1 induction and other responses as biomarkers of exposure toenvironmental contamination in fish. They should be considered preliminary studies, whichare inconclusive by themselves. Clearly, more detailed investigations and better experimentaldesigns are required to establish these types of cause-effect relationships with confidence.1.B.6.3. Use of CYPJA] Induction in Biological Monitoring ofBirdsThere are a few examples of the use of avian CYP1A1 as a biomarker of exposure toenvironmental contamination with polychiorinated aromatic hydrocarbons. In the search forsuitable biomarkers of exposure in black-crowned night herons (Nycticorax nycticorax),HoffiTlan et al. (1993) examined a number of morphological and biochemical measurements inherons from several colonies in the Great Lakes, with various levels of PCB contamination.AHH activities were elevated and femur length-to-body weight ratios were reduced in pippingbirds from the most contaminated colonies compared with a captive colony kept at thePatuxent Wildlife Research Center, MD. In a ftirther study, Rattner et al. (1993) examinedCYP1A1 induction, as a sensitive biomarker of PCB contamination, in a large number ofpipping black-crowned night herons collected from several sites across the United States.They demonstrated a significant (p<O.O5) correlation between EROD activity and total PCBconcentration, measured in individual birds (r=0.72; n=57). One criticism of this study is thattotal PCB concentrations in biological systems do not provide much information on theconcentration of individual congeners, which differ considerably in persistence and toxicity.- 20-IntroductionEarlier, Beliward et al. (1990) reported a highly significant positive correlation betweenenvironmental levels of TCDD or TEQ and hepatic EROD activity in 31 great blue heronhatchlings, in the Strait of Georgia. Detailed analyses of levels of a wide variety oforganochiorine contaminants, including PCDDs, PCDFs, PCBs and pesticides, indicated thatTCDD (r0,57) and the sum of TEQs (r=0.55-0.57) were the only measurements thatcorrelated significantly (p<O.OO1) with this response. Furthermore, TCDD and TEQ levelswere negatively correlated with yolk-free body weight and several other morphologicalmeasurements (Hart et al., 1991). These findings suggest that hepatic EROD activity in greatblue herons can be used as a specific biomarker of contamination of their habitats with TCDDand similarly acting chemicals. In combination with the data presented in this thesis(Sanderson et at., 1994a, 1994b), these studies are examples of a systematic and detailedapproach to the use of a biological monitoring species, which has not been reported before.1.B. 6.4. Toxicological Implications ofEnvironmental Induction of CYP1A 1The cytochrome P-450 enzyme system plays a central role in the detoxication andbioactivation of a wide variety of xenobiotic chemicals. The cytochrome P-450 familyconsists of many different isoenzymes, which have distinct but overlapping substratespecificities. Generally, cytochrome P-450 enzymes oxidate lipophilic compounds to morewater-soluble metabolites, which are readily excreted or conjugated by phase TI-type enzymes,such as glutathione S-transferases, UDP-glucuronosyl transferases and sulfonyl transferases,However, cytochrome P-450 enzymes can also convert certain chemicals into metabolites(usually reactive intermediates) with increased toxicity compared to the parent chemical.Therefore, enzyme induction can work two ways: it can offer increased protection to theorganism by increasing the rate of elimination of a toxic chemical, or it can increase thesusceptibilty of the organism to a potentially toxic chemical by increasing the rate ofproduction of a toxic metabolite or reactive intermediate. Small amounts of reactiveintermediates are continually produced by organisms and are rapidly detoxified by conjugatingenzymes such as glutathione S-transferases. These enzymes can be saturated, however, whenthere is a large increase in the production of reactive intermediates due to enzyme induction.-21-IntroductionIt is the balance between bioactivation and detoxication pathways which determine thetoxicological outcome.CYP1A1, for instance, is partly responsible for the transformation of polycyclic aromatichydrocarbons, such as benz[a]pyrene to at least one mutagenic and carcinogenic metabolite(Gelboin, 1980). Sustained induction of CYP1A1 by persistent inducers such as PCDDs,PCDFs and PCBs, would increase the susceptibility of the induced organism to the mutagenicand carcinogenic effects of benz{ajpyrene, and similar polycyclic aromatic hydrocarbons(loannides and Parke, 1990).An example of enhanced toxicity caused by enzyme induction in birds is brieflysummarized by Walker et at. (1991): the pesticide prochloraz, dosed orally to red-leggedpartridges (Alectoris rufa), induced hepatic EROD activity, aldrin epoxidase activity andcytochrome P-450 content. Induced partridges were more susceptible to the toxicity of thepesticide malathione, resulting in increased inhibition of serum carboxyesterase activity. Thiswas attributed to the increased oxidation of malathione to malaoxon, the active metabolite,instead of the normal detoxication route via ester formation. (It should be noted that theinvolvement of CYP1A1 was not explicitely proved in this study.)CYP1A enzymes also appear to be involved in the metabolism of hormones and otherendogenous chemicals required for homeostasis. Therefore, enzyme induction could lead toaltered metabolism of these endogenous compounds, which could interfere with normalphysiological ftinctions, such as reproductive capacity and growth (Rattner et at., 1984).Another important toxicological implication of CYP1A1 induction is related to thegenetic mechanism by which it is induced (section 1.A.3). CYP1A1 induction is under directregulation of the Ah receptor. Numerous studies have shown that binding to the Ah receptoris not only a prerequisite for CYP1A1 induction but also for many of the toxicities caused byTCDD and other structurally related chemicals. These toxicities do not necessarily have dose-response curves identical to that of CYP1A1 induction (DeVito et a!., 1994); nor do theynecessarily all occur at the same time. These considerations, however, do not take away fromthe fact that the induction of CYP1A1 is an indication that the Ah receptor-mediated process,by which these chemicals exert their toxicities, has been activated. This last point forms thebasis for the hypotheses in this thesis.- 22 -Introduction1. C. HYPOTHESES AND OBJECTIVESHypotheses1. Hepatic EROD activity, as an index of CYP1A1 activity, is a useful biomarker of theextent of exposure of fish-eating birds to polychiorinated aromatic hydrocarbons.Furthermore, environmental induction of hepatic EROD activity indicates that the Ahreceptor-mediated process by which TCDD and related chemicals exert their toxicities hasbeen activated.2. Species differences exist among birds in the hepatic EROD-inducing potency of TCDD,which are partly explained by differences among avian species in the affinity of TCDD forthe hepatic Ah receptor. Consequently, the relative sensitivity to EROD induction can beused as an indicator of potential sensitivity to other Ah receptor-mediated toxicities.ObjectivesA. EnvironmentalMonitoring Study1. To demonstrate that hepatic ethoxyresorufin 0-deethylase (EROD) activity, as an index ofCYP1A1 activity, can be used as a sensitive tool in the biological monitoring of exposureof great blue heron and double-crested cormorant embryos to polychiorinated aromatichydrocarbons.2. To confirm the observations made in the great blue heron hatchlings in 1988 (Bellward etal., 1990; Hart et al., 1991) and to determine whether the observations presented in thisthesis are consistent with the dose-effect relationships determined in 1988; particularlysince the remedial process changes implemented by the Crofton pulp and paper millbetween 1988 and 1990, which have led to a decreased discharge of PCDDs and PCDFsinto the Strait of Georgia (Whitehead et a!., 1992).B. In Ovo Dose-Response Study3. To characterize the induction profile of hepatic EROD activity in the great blue heron anddouble-crested cormorant upon in ovo exposure to3H-TCDD. The precocial chicken, a- 23 -Introductionspecies known to be highly sensitive to the toxicities of TCDD, is included as a positivecontrol. The altricial pigeon was included as a ftirther control, because it resembles thesemi-altricial heron and thily altricial cormorant, developmentally. These dose-responsestudies are compared with the induction due to environmental exposure to TCDD andrelated chemicals.C. Ah Receptor Binding Study4. To determine whether differences among avian species in the hepatic EROD-inducingpotency of TCDD can be explained by differences in affinity of TCDD for the avianhepatic Ah receptors.- 24 -2. MATERIALS AND METHODS2.1. Experimental Birds2. l.A. ENVIRONMENTAL MONITORING STUDYGreat Blue HeronsGreat blue heron eggs (Ardea herodias) were collected from two colonies in the Straitof Georgia, British Columbia (B.C.) (Figure 2.1). The Vancouver colony is located on theUniversity Endowment Lands of the University of British Columbia in an urban setting, andwas sampled in the spring of 1990 and 1992. The Crofton colony is situated near a large pulpand paper mill, and was sampled in 1991. Both colonies had elevated levels of PCDDs andPCDFs in 1988, when the environmental monitoring study began. In 1988, the relativelyuncontaminated Nicomekl colony was also included as a ‘control’ site (Figure 2.1). Thesources of the polychiorinated aromatic hydrocarbon contamination of these sites are knownto a certain extent (Elliott et al., 1989). The herons from the Vancouver colony forage in theFraser estuary, an urban and industrial area which receives effluents of 13 pulp and paper millsand 20 lumber treatment plants. The herons from the Crofton colony feed in an inter-tidalarea near an underwater pipe that discharges the waste water of the Crofton pulp mill.Chlorine bleaching and use of chlorophenols in wood treatment are the most likely sources ofPCDDs and PCDFs in these herons. Two or three eggs were taken per clutch, one of whichwas sent to the National Wildlife Research Centre in Hull, Quebec, for chemical analysis. Theremaining eggs, which were at various stages of natural incubation, were incubated undercontrolled conditions in the Department of Animal Science at the University of BritishColumbia (see section 2.2). After pipping, the chicks were transferred to a hatcher andsacrificed within 24 hours after hatching. It has been shown that levels of chemicalcontamination are similar among eggs from the same clutch (Bellward et at., 1990).Therefore, it is reasonable to compare contamination levels in one egg and hepatic microsomalEROD activity in the matched egg from the same clutch.- 25 -Materials and MethodsFigure 2.1. Locations of three great blue heron colonies in the Strait of Georgia, British Columbia: Nicomeki,Vancouver (U B.C.) and Crofton.Double-Crested CormorantsDouble-crested cormorant eggs (Phalacrocorax auritus) were collected from fivecolonies across Canada. The Saskatchewan colony is from Last Mountain Lake, located in aprovincial park in Saskatchewan, remote from any direct industrial discharges. The threecolonies in British Columbia, Chain Islands, Christy Islet and Crofton, are locatedin the Strait- 26-Materials and Methodsof Georgia. The Chain Islands are located near the city of Victoria. Christy Islet is located inHowe Sound, which receives effluent from two large kraft pulp mills. The cormorants atCrofton nest on the superstructure of the effluent discharge pipe of the Crofton pulp andpaper mill. The Lake Ontario colony is located on Pigeon Island, an area known to have highlevels of PCDDs, PCDFs and PCBs. The cormorant eggs were collected and handled thesame way as the herons above. The incubation conditions of the herons and cormorants aredescribed in section 2.2.2. LB. IN 0 VO DOSE-RESPONSE STUDYWhite Leghorn chicken eggs (Gallus gallus) were obtained from a local distributor andplaced in the incubator on the same day. Pigeon eggs (Rock dove; Columba livia) wereobtained from a breeding colony at the Department of Animal Science and kept at 10°C untilsufficient eggs were collected for the experiments, not storing eggs for longer than a week.Great blue heron eggs were gathered from heron colonies near Chilliwack and Little River,B.C., two areas of relatively low chemical contamination, and transported to our laboratorywithin four hours. Double-crested cormorant eggs were collected from a large colony inBuffalo Narrows, an area in northern Saskatchewan, remote from industrial activity. Theseeggs were placed in the incubator within twelve hours after collection. All eggs used in thedose-response study were artificially incubated at the Faculty of Pharmaceutical Sciences.The age of the heron and cormorant eggs were estimated by candling at regular intervals(Figure 2.2).2.1. C. AH RECEPTOR BINDING STUDYChicken, pigeon and double-crested cormorant eggs were obtained from the samesources as for the dose-response study. Great blue heron eggs were collected from arelatively clean colony near Holden Lake, B.C., and transported to our laboratory within fourhours.-27-Materials and MethodsClear Egg25% Blood Vsl’s.2 50% BV’s75-100% BV’sEmbryo visible50% Embryo‘ Egg darkishDark, still By’Dark, no BV’s0PippingHatching2426 28Figure 2.2. A development scheme of the great blue heron embryo, based on candling observations. Thissheme could be also be applied to double-crested connorant embryos. Abbreviations: Blood vsl’s and BVsblood vessels; e.g. 50% embryo = half of egg taken up by embryo.2.1.D. JUVENILEAND ADULTGREATBLUE HERONSGreat blue herons were raised from the hatchling stage by Darrin Bennett at theDepartment of Animal Science. These herons were fed a controlled diet of uncontaminatedherring (Clupea harengus), trout (Oncorhynchus mykiss) and mackerel (Scomber scombrus)(Bennett, 1993; Bennett and Hart, 1993). Six 4 month-old juvenile herons were sacrificed inSeptember, 1988. These herons were used for several immunoblotting experiments and foranalysis of hepatic microsomal cytochrome P-450 content and EROD activities. Furthermore,they were used to validate the ethoxyresorufin 0-deethylase assay for great blue herons (seesection 2.6). In May 1993, a 2 year-old cross-billed adult heron, raised from hatchling byDarrin Bennet, was sacrificed. Hepatic cytosolic preparations from this heron were includedin the Ah receptor binding study.—I. I I • I I • I • I I • I • I • • • I0 2 4 6 8 10121416182022Incubation Time (days)- 28-Materials and Methods2.2. Incubation ConditionsChicken, pigeon and great blue heron eggs were artificially incubated at 37.5°C and55% relative humidity. Double-crested cormorant eggs were incubated at 35.6°C and arelative humidity of 37%. Chicken eggs were incubated in the upright position in a Humidaire(model 21) forced-air tipping incubator. The other species were incubated on their sides androlled every 75 minutes in a Curfew (RX 200) incubator. The average lengths of incubationfor chicken, pigeon, great blue heron and double-crested cormorant eggs are 21, 19, 28 and28 days, respectively. Hatchabilities under untreated circumstances (or treatment with vehiclealone) were about 90%, 70%, 85% and 77%, respectively.2.3. Injection ProcedurePreliminary control experiments were performed in order to establish a vehicle, time andvolume of egg injection that would not affect length of incubation, duration of pip andhatchability. Conditions were tested in readily available chicken and pigeon eggs. Vehiclesand volumes tested were corn oil (12.5, 25, 50, 100 p1) and 1,4-dioxane (25, 50 p1). Thesetwo vehicles are potential solvents for TCDD. Times of injection were chosen at day ofincubation (DI) 0, 9, 13 and 16 for chicken and DI 0, 9, 14 for pigeon eggs. These three time-points represent early, half-way and two-thirds way through the incubation of each species.Effects on hatchabilty are shown in Table 2.1 for chickens and Table 2.2 for pigeons.Table 2.1. Hatchability of chicken eggs injected with corn oil or 1,4-dioxane at various times duringincubation.Day of Incubation Corn Oil Corn Oil 1,4-Dioxane(5Opl) (loOpl) (5Opl)0- 0/16 2/89- 0/8 2/413 6/8 0/8 3/616 8/8 6/14 2/9Control (uninjected) hatchability: 17/20 (85%).- 29 -Materials and MethodsTable 2.2. Hatchability of pigeon eggs injected with corn oil or 1,4-dioxane at various times duringincubation.Day of Incubation Corn Oil Corn Oil Corn Oil 1,4-Dioxane(12.5 p1) (25 pl) (50 pl) (25 p1)0 4/6 ---9 6/6 5/6 1/4 -14 6/6 4/6 2/9 2/6Control (uninjected) hatchability: 7/11(63%).For the dose-response study, it was decided to inject the bird eggs in the latter thirdperiod of incubation, using corn oil as a vehicle. Vehicle volumes and injection times for thefour bird species are listed in Table 2.3. The injection time for herons was accurate withinabout six days; for cormorants within about three days (as determined by the candling schemein Figure 2.2). The egg shells were pierced through the air sac with a tack. A 50 p1 glasssyringe was inserted into the air sac with the aid of a stereotaxic frame to ensure that theneedle would not pierce the air sac membrane. The tack, syringe needle and surface of theegg were sterilized with 70% ethanol before each injection. The volume was injected slowly;then the hole was sealed with sterile bonewax, Fine Science Tools. The eggs were returned tothe incubator and kept in the upright position for 1 Y2 hour to increase contact of the corn oilwith the air sac membrane, allowing absorption of the TCDD through the blood vessels.Table 2.3. Injection conditions used in the in ova dose-response study (volumes are medians ± range).Bird Species Injection Time Injection Volume(days of incubation) (p1)Domestic Chicken 16 27± 3Domestic Pigeon 14 13 ± 3Great Blue Herona 22 34±4Double-Crested Cormoranta 22 22± 2aThe age of great blue herons and double-crested cormorants were estimated from the candling scheme shownin figure 2.2.-30-Materials and MethodsThe injection times chosen allowed for the birds to be exposed to TCDD during thelatter third period of their incubation. This would be sufficient to maximally induce hepaticEROD activity (Rifkind et a!., 1985), without causing excessive mortality due to exposure tovehicle and the teratogenic effects of TCDD, which occur after injection earlier duringincubation. We were permitted to use only a small number of wild bird eggs per season,which did not allow us to use experimental conditions with high mortality.2.4. Dose-Response ExperimentsTritiated 2,3,7, 8-tetrachlorodibenzo-p-dioxin (3H-TCDD) was purchased fromCambridge Isotope Laboratories. The stock solution dissolved in toluene had a concentrationof 8.0 pg/mI, an activity of 1.0 mCi/mi and radiochemical purity of 97%. Unlabeled TCDD(1mg) in crystalline form was purchased from Cambridge Isotope Laboratories and wasdissolved in toluene to make up a 1 mg/mi solution. Dilutions were prepared by adding aconstant amount of radiolabeled TCDD solution and increasing amounts of unlabeled TCDDsolution to 2 ml of corn oil. The corn oil solutions were mixed for 1 mm using a Vortex(Fisher Scientific), and the toluene was evaporated at 80°C under a gentle flow of nitrogengas (Union Carbide).Chicken eggs were injected with3H-TCDD doses of 0, 0.01, 0.03, 0.125, 0.25, 0.5, 1and 3 pg/kg, with n=8-16 per group. Pigeon eggs were injected with 0, 0.25, 1, 3, 10 and 100pg/kg egg, (n=8-14). Great blue heron eggs were treated with doses of 0, 0.5, 1, 3, 10 and100 pg/kg egg, (n=8-16). Double-crested cormorant eggs were injected with 0, 1, 3, 10, and100 pg/kg egg (n=6-10).2.5. Tissue Preparations2. 5.A. ENVIRONMENTAL MONITORING STUDYEach hatchling was weighed and blood was drawn by cardiac puncture. The bird wasthen decapitated, and the abdominal cavity was immediately opened to remove the liver andattached gallbladder. The gallbladder was carefully separated from the liver without puncture.The liver was immediately immersed in 25 ml TRIS-KC1 buffer (0.05 M TRIS, 1.15% KC1,pH=7.5) at 4°C for microsome preparation. Morphological measurements on the hatchlings- 31 -Materials and Methodswere performed in the laboratory of Dr. L.E. Hart and Dr. K.M. Cheng at the Department ofAnimal Science.Microsomal Preparations: All steps of the microsomal preparation were performed at4°C. Avian livers were individually homogenized in 25 ml TRIS-KC1 buffer in a 50 ml glassPotter-Elvehjem homogenizing tube, using a Teflon pestle (Fisher Scientific). The pestle waspassed through the homogenate five times at a speed sufficiently high to homogenize the liverwithout causing excessive foaming. After a 1 mm break (homogenizing creates heat) this wasrepeated. The homogenate was centrifuged at 10 000 g for 20 mm in a model J2-21 Beckmancentrifuge, using a JA- 17 rotor. The precipitate was discarded and the supernatantcentrifuged at 100 000 g for 60 mm in a L5-50 Beckman ultra-centrifuge, using a 50.2 Tirotor. The microsomal pellet was suspended in 20 ml of 10 mM ethylenediamine tetraaceticacid (EDTA), 1.15% KC1, pH=7.4, buffer at 4°C and homogenized. The Teflon pestle waspassed through the homogenate four times, avoiding foaming. The homogenate was thenrecentrifuged in the ultra-centrifuge, as described above. The resulting microsomal pellet wasresuspended in about 0.5 ml of 0.25 M sucrose using a 1 ml syringe with 25 gauge needle.Aliquots of 100 p1 of the final microsomal preparation were frozen in 1 ml cryotubes (Nunc)and stored at -80°C for subsequent assays. The final protein content was about 10 mg/mi.2. 5.B. IN OVO DOSE-RESPONSE STUDYHatchlings were sacrificed and treated the same way as in the environmental monitoringstudy, except that several organs were collected for3H-TCDD analysis: immediately afterremoval of the gallbladder (the pigeon does not have a gallbladder), the liver was weighed andabout 50 mg of tissue was separated for 3H-TCDD analysis. The remaining liver wasimmersed in TRIS-KCL for microsomal preparation as described in the previous paragraph.The yolk sac and both kidneys were removed, weighed and stored in glass vials at -20°C foranalysis. Whole blood was stored at -20°C in small glass test tubes rinsed with heparin.- 32 -Materials and Methods2.5. C. AH RECEPTOR BINDING STUDYHatchlings were weighed and decapitated and the abdominal cavity was opened. Aftercareflul separation of the gallbladder, the liver was immediately removed and immersed in icecold HEDGM buffer (25 mM HEPES, 1.5 mM EDTA, 1 mM dithiothreitoi, 10% glycerol and20 mM sodium molybdate, pH=7. 8) at a 1:1 weight-to-volume ratio.Cytosolic Preparations: All further steps were performed at 40C. Cytosolic fractionswere prepared by homogenizing the livers, using a Teflon pestle and centrifuging thehomogenate at 10 000 g for 20 mm, using a JA-17 rotor with foam adaptors. The supernatantunderwent an ultra-centrifugation step at 100 000 g for 1 h, using a Ty 65 rotor. Theresulting cytosolic supernatant was drawn from the pellet using a 20 gauge needle, as not todisturb pellet and lipid layer, and was frozen in 100 p1 aliquots in cryovials at -80°C. Thepellet was discarded. The final cytosolic protein concentration ranged from 15-30 mg/mi.2.6. Ethoxyresorufin 0-Deethylase AssayHepatic EROD activities, as measures of CYP 1A1/2 activities, were determined by amodification of the direct fluorometric assay described by Burke and Mayer (1974). AShimadzu RF-540 spectrofluorometer was used with an excitation wavelength of 550 nm (slitwidth: 2 nm) and an emission wavelength of 582 nm (slit width: 2 nm). The assay wasperformed at 37°C instead of at the body temperature of birds (42°C) because activities(pmol/min!mg protein; mean ± SEM) in great blue heron microsomes were 1 ‘/2 times higher at37°C (86 ± 6) than at 42°C (59 ± 3). Also, a temperature of 37°C was chosen in order tocompare with other published studies. The fluorescence cuvette contained 1.93 ml HEPESMgCl2 buffer (0.1 M HEPES, 5 mM MgC12, pH=7.8) and 10 p1 of 1 mM ethoxyresorufindissolved in dimethyl suifoxide. The cuvette was warmed up to 37°C for 5 mm before addingthe microsomal protein and NADPH. Microsomal protein (50 p1 in 0.25 M sucrose) wasadded at a concentration of 6 mg/mi for non-induced and 1 or 2 mg/mi for induced samples.The reaction was started with 10 pl of 50 mM NADPH (prepared freshly in cold HEPESMgC12 buffer and kept on ice). The total reaction volume was 2 ml. The increase influorescence due to product formation was measured after 5 mm for non-induced and 2 mmfor induced samples. The amount of resorufin produced in the cuvette during the reaction was- 33 -Materials and Methodscalculated from a standard resorufin curve measured in the presence of microsomal protein.All samples were tested in duplicate or triplicate and standard curves were prepared daily.Substrate concentration and pH were optimal (Figure 2.3 and 2.4). Reaction rates were linearwith protein concentration and time using the conditions described above (Figure 2.5 and 2.6).EROD activities were preferably expressed on the basis of microsomal protein content forinterpretation purposes (see section 3 .A.5).0.ECE0E0.0wa0.100806040200 1 2 3 4 5 6 10 15 20[Ethoxyresorufin] (uM)Figure 2.3. Dependency of hepatic EROD activity of 4 pooled great blue heron microsomes (triplicatedeterminations) on the concentration of ethoxyresorufin (final concentration in cuvette); other assay conditionsare the final assay conditions described in section 2.6.110 -r: 100I ::f 70I07.2 7.4 7.6 7.8 8.0pHFigure 2.4. Dependency of hepatic EROD activity on pH in 4 pooled great blue heron microsomes (triplicate).- 34 -Materials andMethodsCE0E0.C0C)•000.C0100080060040020000 3 6 9 12[Protein] (mglml)Figure 2.5. Dependency of hepatic EROD activity on protein concentration in 4 pooied great blue heronmicrosomes (triplicate determinations).1000C000.E0E0.C04-C)•000.C20000000.0 2.5 5.0 7.5 10.0 12.5 15.0 17.5 20.0Time (mm)Figure 2.6. Amount of resorufin formed over time in 4 pooled great blue heron microsomes (in tripliclate).- 35 -Materials and Methods2.7. Total Cytochrome P-450 DeterminationsHepatic cytochrome P-450 concentrations were determined by difference spectroscopy(Omura and Sato, 1964). Measurements were performed at room temperature using an SLMAminco DW-2C dual-beam spectrophotometer. Microsomal protein (3-15 mg/mi) wasdiluted between 10 and 15 times in sodium phosphate buffer (0.1 M sodium phosphate, 20 %glycerol, 0.1 mM EDTA, pH=7.4) to obtain a volume of 1 ml. Equal volumes were dividedover a 0.5 ml quartz sample and reference cuvette. After recording the baseline spectrum, afew mg of sodium dithionite were added to both sample and reference cuvette and allowed todissolve for 2 mm. Then carbon monoxide was bubbled through the sample cuvette for 60 s,at a rate of 1 bubble per second. Finally, a spectrum was recorded from 325-625 nm, using ascan speed of 2 nm/s. Cytochrome P-450 concentrations were calculated from the differencein absorbance between 450 nm and 490 nm, using a molar extinction coefficient of 91 cm1xmM4. Insufficient sample was available to perform duplicate determinations.2.8. Protein DeterminationsMicrosomal protein concentrations were determined according to the method ofBradford (1976). Bovine serum albumin was used as a standard and the Bio Rad kit asreagent. Microsomal protein was diluted 100-fold in distilled water. Standard curves wereprepared over a concentration range corresponding to 0-20 mg/mi of final microsomal protein.Absorbance was measured at 595 nm on a Hewlett-Packard 8452A diode-arrayspectrophotometer. All determinations were performed in duplicate.2.9.3H-TCDD AnalysesPortions of 50 mg of avian liver were thawed and 250 pl of distilled water and 250 p1 ofacetone were added (Curtis et at., 1990). The liver portions were then homogenized, using an0mm 2000 homogenizer, DiaMed. The resulting homogenate was extracted twice with 1 mlof hexane by mixing vigorously for 60 s on a Vortex. Liquid phases were separated bycentrifugation at 3000 rpm for 5 mm. The 1 ml hexane phases were pooled and 800 plaliquots were taken in duplicate for liquid scintillation counting on a Beckman LS 6000TA.The liquid scintillation cocktail used was Biofluor, a high efficiency emulsifier for aqueous-36-Materials andMethodssamples (DuPont). Tissue recoveries were determined by adding a comparable amount of3H-TCDD, dissolved in hexane, to 50 mg amounts of untreated avian livers. The3H-TCDDsolution was applied to the livers dropwise and allowed to evaporate for an hour. The liverswere then frozen overnight and thawed the next day, before extraction by the methoddescribed above. Recoveries were 81 ± 1%. Quenching (20-30%) was corrected for usingthe internal standard method.Whole yolk sacs were thawed and homogenised for 15-20 s; then 50-100 mg aliquotswere taken, in duplicate, and added directly to the liquid scintillation cocktail. Quenching,which varied from sample to sample and ranged from 10-30%, was corrected for using theinternal standard method.Whole blood samples were thawed and duplicate 25 p1 aliquots were added directly toliquid scintillation cocktail. The liquid scintillation vials were kept in the dark for 48 h beforecounting to reduce chemiluminescence to less than 5%. Quenching in these samples wasabout 25%.2.10. Ah Receptor Binding AssayThe Ah receptor binding assays were performed in the laboratory of A.B. Okey,University of Toronto, ON. Aliquots of 500 p1 of cytosolic protein, diluted to 5 mg/mi inHEDGM buffer (25 mM HEPES, 1.5 mM EDTA, 1 mM dithiothreitol, 10% glycerol (v/v)and 20 mM sodium molybdate, pH=7.8 at 40C), were incubated with various concentrationsof3H-TCDD (0.25-40 nM; specific activity: 50 Cilmmole) in the presence and absence of a100-fold molar excess of unlabeled 2,3,7,8-tetrachlorodibenzofuran (TCDF). All steps wereperformed at 4°C. Ligand and competitor were dissolved in dimethyl sulfoxide and added tothe cytosol in a volume of 10 p1. After a 2 h incubation, 300 p1 aliquots of cytosol wereadded directly to continuous (10-30%) sucrose density gradients, prepared in HEDGM buffer.The incubation mixture was not treated with dextran-coated charcoal, which is often used toremove non-specific binding, because of the sensitive nature of the specific binding in theavian cytosolic fractions (see Results and Discussion sections). 14C-formaldehyde-labeledbovine serum albumin (BSA; sedimentation coefficient: 4.4 S) and14C-formaldehyde-labeledcatalase (CAT; 11.3 S) were included as internal sedimentation markers. The sucrose density-37-Materials andMethodsgradients were centrifuged in a vertical VTi 65 rotor at 372 000 g for 2 hours (Tsui and Okey,1981). Then 25 fractions of 200 4u1 were collected from each gradient. The apparentconcentration of Ah receptors (Bm) and the apparent equilibrium dissociation constant (Kd)were determined by Woolf plot analysis (Cressie and Keightley, 1981). Scatchard plots(Scatchard, 1949) are also presented. Preference was given to Woolf plot analysis becausethe only experimentally derived parameter required is the amount of specifically bound ligand.Free ligand is simply determined by subtracting the amount specifically bound from the totalamount added. Also, Woolf plots are less sensitive to outliers and to data that are notperfectly linear (Okey et a!., 1989).2.11. Immunochemical Cross-Reactivity StudyImmunoblots were performed in the laboratory of S.M. Bandiera. A small number ofavian hepatic microsomes were probed on immunoblots with antibodies raised against ratcytochrome P-450 1A1 and 2B1/2 (section 2.1 1A and B). The CYP1A1 antibodies were amixture of four monoclonal mouse antibodies that specifically recognised rat CYP1A1 (a giftfrom P.E. Thomas, Rutgers University, N.J.). The CYP2B1/2 antibodies were polyclonalrabbit antibodies that recognized both rat CYP2B1 and CYP2B2 (prepared by S.M. Bandiera,University of British Columbia). Microsomal proteins (20 ,ug/well) were separated usingsodium dodecylsulfate-polyacrylamide gel electrophoresis (SDS-PAGE) according to themethod described by Laemmli, (1970). Constant currents of 10-12 mA/gel for the stackinggel and 20-25 mA/gel for the separating gel were applied. After separation, the blot waseither stained with 1% Coomassie blue R-250 or prepared for immunoblotting (Towbin eta!.,1979). Briefly, the proteins on the gel were electrophoretically transferred to a nitrocellulosemembrane, at 4°C. The nitrocellulose membrane was then incubated overnight in blockingbuffer, at 4°C. The next day the membrane was incubated with the primary antibody (2 pgIgG/ml for anti-CYP1A1 and 50 pg IgG/ml for anti-CYP2B1/2) for 2 h, washed 3 times for10 mm, and incubated with secondary antibody for 2 h, all at 3 7°C. The secondary antibodywas a horseradish peroxidase-conjugated goat anti-rabbit (used at a dilution of 1:1000) oranti-mouse F(ab)’2 (4 year-old stock, used at a dilution of 1:100). As substrate for thehorseradish peroxidase, 0.0 18% 1-chloro-4-naphthol was used.- 38 -Materials and Methods2.1 l.A. ENVIRONMENTAL MONITORING STUDYSelected hepatic microsomes from great blue heron and double-crested cormoranthatchlings were probed with monoclonal antibodies raised against rat CYP1A1 and polyclonalantibodies raised against rat CYP2B 1 and CYP2B2. Three micro somal samples were selectedper species: one sample per site was chosen from a site of low, intermediate, and highcontamination and respective hepatic EROD activity.IVlicrosomal samples from one male andone female juvenile great blue heron, sacrificed in 1988 (see section 2.1 .D), were probed withthe same antibodies as the hatchlings mentioned above.2. 11.B. IN 0 VO DOSE-RESPONSE STUDYChicken microsomal samples were chosen from eggs treated with 0, 0.l25 and 1.0pg/kg3H-TCDD. Great blue heron samples were chosen from eggs treated with 0, 1.0 and10 pg/kg3H-TCDD. These two species were probed with the monoclonal rat CYP1A1antibodies only. Double-crested cormorants and pigeons from the dose-response study werenot used in the immunochemical cross-reactivity study, because the dose-responseexperiments in these two species had not been performed at the time that the antibodies wereavailable.2.12. Environmental Chemical AnalysesEnvironmental levels of chemicals in the eggs were analysed by Ross J. Norstrom andcoworkers (Mary Simon and Henry T. Won) at the National Wildlife Research Centre(Canadian Wildlife Service) in Hull, Quebec. The laboratory of Ross Norstrom specializes inthe detection of trace levels of polyhalogenated aromatic hydrocarbons in environmentalsamples. The method used is described briefly below: PCDD and PCDF determinations werecarried out in matched eggs of great blue herons and double-crested cormorants by gaschromatography-mass spectroscopy (GC-MS) using‘3C-labelled internal standards after gelpermeation-carbon chromatographic clean-up, as described in Norstrom et al. (1990) andNorstrom and Simon (1991). Non-ortho-PCBs were determined by a modification of thePCDD method: the sample obtained from the gel permeation-carbon chromatographic cleanup was evaporated on a Brinkman Rotavapor to a small volume (0.5-1 ml) and transferred to- 39 -Materials and Methodsthe top of a 10 g alumina column (1 cm i.d.x 12 cm long, basic alumina, activated, FisherScientific). The column was eluted with 70 ml dichioromethane/hexane (1:1). The eluate wasevaporated nearly to dryness, made up to about 2 ml with hexane and chromatographed on acarbon column similar to that described in Norstrom and Simon (1991), except that the lengthwas 6 cm instead of 3 cm. The carbon column was eluted sequentially with 80 ml 5%dichloromethane/95% hexane (Fl), 200 ml dichioromethane (F2) and back-eluted with 110 mltoluene (F3), at a flow rate of 5 mllmin. Fl contained the multi-ortho-PCBs. F2 containedthe mono-ortho-PCBs 105 and 118. F3 contained the non-ortho-PCBs 77, 126 and 169.Only results of F2 and F3 analyses are reported here, because these fractions contained thePCBs with Ah receptor affinity. F2 was analysed by GC and electron capture detection asdescribed by Elliott eta!. (1989). F3 was analysed by GC-MS using 13C-labeled standards.2.13. TCDD-Toxic EquivalentsSums of TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalencyfactors (TEFs) proposed by Safe (1990) (Table 2.4). The TEQs in great blue herons in 1988,Table 2.4. 2,3,7,8-Tetrachlorodibenzo-p-dioxin-toxic equivalency factors (TEFs) for the majorpolychiorinated aromatic hydrocarbons found in great blue herons and double-crested cormorants in Canada.Polychlorinated TEF’ Polychlorinated TEP Polychlorinated TEP’dibenzo-p-clioxins dibenzoflirans biphenylsbPCDDs PCDFs PCBs2,3,7,8-TCDD 1 2,3,7,8-TCDF 0.1 PCB-77 0.01l,2,3,7,8-PnCDD 0.5 2,3,4,7,8-PnCDF 0.5 PCB-126 0.11,2,3,4,7,8-HxCDD 0.1 1 ,2,3,4,7,8-HxCDD 0.1 PCB-169 0.051,2,3,6,7,8-HxCDD 0.1 1,2,3,6,7,8-HxCDD 0.1 PCB-105 0.0011,2,3,7,8,9-HxCDD 0.1 2,3,4,6,7,8-HxCDD 0.1 PCB-118 0.0011,2,3,4,6,7,8-HpCDD 0.01OCDD 0.001aThFs were obtained from Safe, 1990.bPCB nomenclature from Ballschmiter and Zell, 1980.-40-Materials andMethodsreported by Beliward et al. in 1990, were recalculated using these TEFs. Analyses of themajor non-ortho-PCBs, PCB-126 and PCB-169 were not available for the great blue heroneggs, but were estimated from the sum ofPCB congeners (PCB) as follows:PCB 126 (ng/kg) = 29.6 + 231 x PCB (mg/kg); r=0.77; n=9.PCB 169 (ng/kg) = 12.2 + 26.9 x EPCB (mg/kg); r=0.84; n9.These equations were obtained from analytical data on great blue heron eggs from B.C.collected by the Canadian Wildlife Service between 1988 and 1990.2.14. Statistical AnalysesThe Number Cruncher Statistical System (NCSS) software package (Hintze, 1990) wasused for statistical analyses of the data. Arithmetic means were calculated and presented withstandard errors of the mean (SEM). Means of two groups were compared using Student’s ttest. Means of more than two groups were compared using one-way analysis of variance(ANOVA). Statistically significant differences were detected using the Newman-Keulsmultiple range test. Coefficients of determination (r2) were calculated using least squareslinear regression analysis. Correlation coefficients (r) were determined using least squareslinear correlation analysis. Concentrations of chemicals were transformed logarithmically inregression analyses to conform to the way in which pharmacological dose-response curves arepresented, and not necessarily to obtain the best linear relationships. A value of p<O.0l wasconsidered statistically significant for the analyses in the environmental monitoring study,because of the large number of variables compared. A value of p<O.O5 was consideredstatistically significant in the in ovo dose-response study. Differences in incidences ofoccurrence (mortality or edema) were not analysed statistically. A suitable test was notavailable for these data, because of the small sample sizes (greater than 20% less than 5) andlow incidences (often less than 1) (Zar, 1984).- 4] -3. RESULTS3.A. ENVIRONMENTAL MONITORING STUDYThe aim of the environmental monitoring study, which began in 1988, was todetermine correlations between biological effects in great blue heron and double-crestedcormorant hatchlings and levels of environmental contamination in their matched eggs.The use of hepatic EROD activity as biomarker of exposure to polychlorinated aromatichydrocarbons was also examined. My contribution to this study started in 1989.Monitoring data in great blue herons in 1988 were published by Bellward et al. (1990) andHart et at. (1991). Some of these data were included in this thesis in recalculated form.Some of the morphological measurements of great blue herons and double-crestedcormorants were performed by L.E.Hart and coworkers at the Department of AnimalScience. Finally, the chemical residue analyses, an essentialpart of the monitoringproject,were performed by R.J. Norstrom (Canadian Wildlfe Service). I have clearly indicated inthe main text and in footnotes in tables, raw data that were not obtained by me. All otherdata, statistical analyses, calculations, graphic representations and other manipulations ofthese data were performed by me. I have included these data in the main body of the thesis,instead of in an appendix, for continuity and clarity. The purpose of the environmentalmonitoring project as presented in this thesis, is to provide an overall picture of the effectsthat changing levels of polychlorinated aromatic hydrocarbons have had on great blueherons in the last few years, with an extension to include the double-crested cormorant.This work has led to two publications (Sanderson et at., 1994a; 1994b).3.A.1. Levels of Chemical Contamination in Monitored Bird EggsGreat Blue HeronIn continuation of the monitoring study that began in 1988, chemical contaminant levelswere measured in great blue heron eggs from a Vancouver colony in 1990 and 1992, and aCrofton colony in 1991. Levels in these colonies were elevated in 1988 (Bellward et at.,1990). PCDD and PCDF levels in the Vancouver (1990) and Crofton (1991) colonies haddecreased significantly since 1988 (Figure 3.1). The levels in Vancouver decreased further- 42-Results700600500• 4003000Figure 3.1. The contribution of environmental levels of several classes of chemicals with Ah receptor affinityto the sum of TCDD-toxic equivalents (TEQ; nglkg egg wet weight) in great blue heron eggs from colonies inthe Strait of Georgia in British Columbia between 1988 and 1992 (n values are listed in table 3.1). The TEQin Crofton was significantly higher in 1988 than in 1991 (p<0.0 1, Student t-test). The TEQ in Vancouver wasnot different in 1988, 1990 and 1992. The TEQ in Crofton in 1988 was higher than in all other colonies andyears (p<O.01, one-way ANOVA and Newman-Keuls multiple range test). Abbreviations: non-ortho-PCBs(non-o-PCB5), mono-ortho-PCBs (mono-o-PCB5).from 1990 to 1992 (Figure 3.1 and Table 3.1). TCDD levels (mean ± SEM) in Vancouverand Crofton were 135 ± 50 and 211 ± 44 ng/kg, respectively, in 1988 (Beliward et al, 1990),whereas the levels in Vancouver in 1990 and 1992, and Crofton in 1991 were 42 ± 7, 10 ± 1and 16 ± 6 nglkg, respectively (Table 3.1). In fact, TCDD levels in Crofton (1991) andVancouver (1992) were not different from those in the Nicomeki colony in 1988 (10 ± 1ng/kg), which served as a relatively uncontaminated reference site in the studies by BeliwardNicomekl’88 Vancouver’SR Vancouver’90 Vancouver’92 Crotton88 Crofton9l- 43 -ResultsTable 3.1. Polychiorinated dibenzo-p-dioxin (PCDD), dibenzofuran (PCDF) and biphenyl (PCB) levels andsum of TCDD-toxic equivalents (TEQ) (mean ± SEM; ng/kg egg wet weight) in great blue heron eggs.Great Blue Heron ColonyConcentration of Crofton (1991) Vancouver (1990) Vancouver (1992)Chemical (ngfkg) (n=7) (n=7) (n=7)PCDDs2378-TCDD 16± 6 42± 7 10± 1b12378-PnCDD 37± 17 45 ± 7 10± 3123478-HxCDD (1 ± 1)C (3 ± l)C n.d.123678-HxCDD 46±26 57±13 ll±3123 789-ITxCDD (2 ± 2)C (3 ± 1)C n.d.1234678-HpCDD (2± 1)C (4± 2)C (2± 1)COCDD 2±1 5±1 (O±O)bCPCDFs2378-TCDF 4± 1 15 ± 11 4± 2C23478-PnCDF 9±3 16±2 (i±i)C123478-HxCDF (0±0)C (1±l)C n.d.123678-HxCDF (0±0)C (1±0)C n.d.234678-HxCDF n.d. (1 ± l)C n.d.PCBsdPCB126d 10113d 29354d 32788dPCB169d 20 ± 2d 43 ± 10dPCB-105(ug/kg) 5±1 22±6 30± 10PCB-118(uglkg) 40±8 144±37 170±55TEQe(ng/kg) 100± 16 278±52 253 ±75Note 1: chemical levels were analysed by R.J.Norstrom at the National Wildlife Research Centre, Hull, Que.Note 2: above data were published by Sanderson et al. (1994a).n.d.: not detectable. Values were below the limit of detection, defined by a signal-to-noise ratio of 3:1.aSijfjctly different from Crofton (1991) and Vancouver (1992), using one-way ANOVA (p<O.Ol) andNewman-Keuls multiple range test.bsigpjficanuy different from Vancouver (1990), using Student t-test (p<O.O1).CLevels between parentheses are averages of values of which more than half were below the limit of detection.dLevels of PCB-126 and 169 were estimated from the sum of PCB congeners for each individual egg (seeMaterials and Methods, section 2.13). PCB-77 levels could not be estimated.eTEQs were calculated using the TEF values proposed by Safe, (1990).- 44 -Resultset al. (1990) and Hart et a!. (1991). The TEQ was significantly lower in Crofton (1991) thanin Crofton (1988) (Figure 3.1). The TEQ was not significantly different among Vancouver(1988), (1990) and (1992). The contribution of TCDD and other PCDDs and PCDFs to theTEQ decreased in the colonies between 1988 and 1992 (Figure 3.1). Despite the drop inPCDD and PCDF levels in the Vancouver colony since 1988, the TEQ remained statisticallythe same in 1990 and 1992, due to close to doubling of the contribution of the non-ortho andmono-ortho-PCBs. In the other colonies, the contribution of the non-ortho and mono-orthoPCBs to the TEQ remained similar throughout the years.Double-Crested CormorantMono- and non-ortho-PCBs accounted for the most striking differences in chemicalcontamination among the five cormorant colonies (Table 3.2). PCB levels were significantlyhigher in the Lake Ontario colony than the other colonies. PCB levels in Chain Islandssamples were intermediate. The Crofton colony was not included in the comparison, becauseonly one egg was analysed for PCBs. Levels of PCDDs and PCDFs were generally lower inthe Saskatchewan colony than all the other colonies (Table 3.2). The TEQ was six timeshigher in the Lake Ontario than in Saskatchewan and Christy Islet eggs, and 2.4-fold higherthan in those from the Chain Islands. PCBs were by far the largest contributor to the TEQfound in the cormorant eggs from the five colonies (Figure 3.2). Despite the relatively lowtoxic equivalency factor of the mono-ortho-PCBs, PCB-105 and PCB-1 18 (Table 2.4), theycontributed considerably to the TEQ because of their much higher concentration (present inpg/kg) than the non-ortho-PCBs. However, levels of individual PCB congeners tend to behighly correlated, and TEFs vary depending on the end-point chosen. Furthermore, thecongener-specific TEFs, which are based on mammalian studies (Safe, 1990), may beconsiderably different for avian species. Therefore, care must be taken in assigning relativeimportance of contribution to the total TEQ among congeners.- 45 -ResultsTable 3.2. Polychiorinated dibenzo-p-dioxin, dibenzofuran and biphenyl levels and sum of TCDD-toxicequivalents (TEQ) (mean ± SEM; ng/kg egg wet weight) in double-crested cormorant eggs.Double-Crested Cormorant ColonyConcentration of Saskatchewan Chain Christy Islet Crofton LakeChemical (nglkg) (n=12) Islands (n=7) (n=7) Ontario(n=7) (n=5)PCDDs2378-TCDD 6±la 20±5b 39±6c 26±2bc 36±5c12378-PeCDD 10± 3a 42 ± 12b 42± 4b 42 ± 3b 32 ± 2b123478-HxCDD (2± 1)d (2± 1)d n.d. n.d. 6± 1123678-HxCDD 12±3a 61±14b 69± lob 65±8b 37±7a123789-HxCDD (3 ± l)d 4± 1 (3 ± 0)d 6 ± 1 12±31234678-HpCDD (6 ± 2)d 7±2 (2 ± 0)d (1 ± 0)d 27± 8OCDD (7± 4)d n.d. n.d. n.d. 25 ± 6PCDFs2378-TCDF n.d. (0± 0)d 3 ± 1 2± 1 (0± 0)d23478-PeCDF (2± 1)ad 16± 4b 12± lb 11 ± lb 29± 4c123478-HxCDF n.d. 3 ± 2 (1 ± 0)d (1 ± 0)d 10±2123678-HxCDF n.d. 3 ± 1 (0 ± 0)d (1 ± 0)d 4 ± 1234678-HxCDF n.d. (1 ± 1)d (0± 0)d (1 ± 0)d (3 ± 0)dPCBs (n=12) (n=4) (n=3) (n=l)e (n=5)PCB-77 88±13a 71±12a 40±16a 23e 318±55bPCB-l26 547±105a 1123±239b 287±28a J89e 3577±354cPCB-169 91±46a 128±25a 52±5a 24e 412±70bPCB-105(ug/kg) 20±4a 79±15b 26±6a Joe 168±18cPCB-118( 142±33a 400±74b 131±28a 56e 983±107cTEQ(ng/kg) 250±50a 672±73b 276±14a J3]e 1606±118cNote 1: chemical levels were analysed by R.J.Norstrom at the National Wildlife Research Centre, Hull, Que.Note 2: above data were published by Sanderson eta!., (1994b).n.d.: not detectable. Values were below the limit of detection, defined by a signal to noise ratio of 3:1(approximately 0.5-2 ng/kg). Non-detectable levels were considered to be zero in the TEQ calculations.a,b,cColomes without overlapping superscripts are significantly different from each other, using one-wayANOVA (p<0.01) and Newinan-Keuls multiple range test.djvels in parentheses are averages of values of which more than half were at or below the limit of detection,defined by a signal to noise ratio of 3:1 (approximately 0.5-2 ng/kg).eNot included in the one-way ANOVA analysis of the PCBs because n=1.- 46-Results1750-S0).S 12501000750:0Figure 3.2. The contribution of environmental levels of several classes of chemicals with Ah receptor affinityto the sum of TCDD-toxic equivalents (ngfkg egg wet weight) in double-crested cormorant eggs from coloniesacross Canada (n values and statistics are shown in table 3.2). Abbreviations: non-ortho-PCBs (non-o-PCBs),mono-ortho-PCBs, (mono-o-PCBs).3.A.2. Gross Abnormalities and EdemaGreat Blue HeronThe incidence of jelly-like’ subcutaneous edema in the great blue heron chicks,monitored between 1988 and 1992, is summarized in Table 3.3. The 1988 observations werepublished by Hart et a!. (1991). One of the 7 Vancouver (1990) chicks had subcutaneousedema on the breast, indicative of chick-edema disease, and fluid in the brain cavity. Anotherhad a blocked cloaca, causing uric acid to accumulate in the intestine. None of the 7 Crofton(199!) chicks showed signs of edema or other abnormalities. This is in contrast to 1988,Saskatchewan Chain islands Christy Isiet Crofton Lake Ontario-47-Resultswhen the incidence of edema (4/12) in the Crofton colony was significantly higher (Hart eta!.,1991) (Table 3.3). None of the 7 Vancouver (1992) chicks showed signs of edema, but onechick had an unresorbed yolk-sac.Table 3.3. Incidence of edema in great blue heron hatchlings from three colonies in British Columbiamonitored between 1988 and 1992.Great Blue HeronColony Incidence of EdemaNicomekl (1988) 0/11Vancouver (1988) 2/13Vancouver (1990) 1/7Vancouver (1992) 0/7Crofton (1988) 4/12Crofton (1991) 0/7Data from 1988 were published by Hart eta!. (1991).Data from 1990-1992 were published by Sanderson eta!. (1994a).Double-Crested CormorantOne of the 7 Crofton cormorant chicks had a brain malformation (asymmetric brain).One of the 11 Lake Ontario cormorant chicks had a deformed liver (outwardly curled liverlobes). None of the chicks showed signs of jelly-lik& edema as in chick-edema disease.However, under our incubation conditions, freshly hatched cormorant chicks often had‘watery edema, which was mainly found in the abdominal cavity. We observed similar ‘watery’edema in pigeon and heron chicks, when examined directly after hatching. This edema couldbe easily distinguished from the jelly-1ike’ edema which was found near the neck, chest andlegs. In the case of the double-crested cormorant chicks this ‘watery’ edema appeared to be anartefact of excessive humidity in the incubator.- 48 -Results0.ECECEa>C)0wC)4-0.I3.A3. Morphological Measurements, Hepatic Microsomal Cytochrome P-450 Contentand EROD ActivityGreat Blue HeronTibia length and tibia wet and dry weight were significantly greater in Crofton (1991)(Table 3.4) than in the Crofton colony in 1988 (Hart et aL, 1991). In the Vancouver colony,tibia dry weight was greater in 1990 and 1992 than in 1988. Intestine wight, tibia wet weightand tibia length were significantly greater in Crofton (1991) than in Vancouver (1990) and(1992) (Table 3.4). The mean hepatic microsomal EROD activity (pmole/minlmg protein) ofthe Crofton (1991) chicks had decreased significantly compared with the same colony in 1988,and was not different from Vancouver (1990) and (1992) (Figure 3.3). The mean hepaticEROD activities of the Vancouver colony in 1990 and 1992 were not significantly differentfrom 1988 (Figure 3.3), as was the case for the TEQ (Figure 3.1).4003002001000Figure 3.3. Comparison of hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities in great blueheron hatchlings from three colonies in British Columbia, monitored between 1988 and 1992. (*): ERODactivities were significantly higher in Crofton (1988) than in the other colonies and other years (p<O.Ol, one-way ANOVA and Newman-Keuls multiple range test).Nlcomekl88 Vancouver’BS Crofton88 Vancouver90 Croftongi Vancouver92- 49 -ResultsTable 3.4. Morphological measurements, hepatic microsomal cytochrome P-450 content and ethoxyresorufin0-deethylase (EROD) activity (mean ± SEM) in great blue heron hatchlings.Note 1: the following measurements were performed by L.E.Hart at the department of Animal Science,University of British Columbia: adrenal, kidney, stomach, intestine and tibia weights, and tibia length.Note 2: above data were published by Sanderson et a!. (1994a).aThere were no significant differences between groups for (not all shown in table): body, yolk-free body, yolk,liver, adrenal, kidney, stomach, tibia ash and thy, heart, spleen and bursa weights, and total hepaticmicrosomal cytochrome P450 content and EROD activity.bsigpjficanijy different from Vancouver (1990) and Vancouver (1992), using one-way ANOVA (1<0.01) andNewman-Keuls multiple range test.Great Blue Heron ColonyCrofton (1991) Vancouver (1990) Vancouver (1992)Parametera (n=7) (n=7) (n=7)Bodyweight(g) 53.9±2.5 49.1± 1.7 52.7± 1.5Yolk-free body weight(g) 49.1 ± 2.1 43.9 ± 1.5 47.6 ± 1.5Yolkweight(g) 4.8±0.6 5.2±0.8 5.1±0.3Liver weight (g) 1.1 ± 0.10 0.99 ± 0.03 1.3 ± 0.07Adrenal weight (mg) 21 ±2 20± 2 12± 3Kidney weight (g) 0.80 ± 0.03 0.63 ± 0.01 0.73 ± 0.05Stomach weight (g) 2.8 ± 0.2 2.3 ± 0.1 2.8 ± 0.3Intestine weight (g) 1.8 ± 008b 1.5 ± 0.04 1.5 ± 0.06Tibia length (mm) 28.9 ± 02b 26.9 ± 0.2 27.5 ± 0.5Tibia wet weight (g) 0.28 ±0•2b 0.21 ± 0.01 0.24 ± 0.02Tibia dry weight (mg) 53 ± 3 51 ± 1 52 ± 2Tibiaashweight(mg) 11±0.5 10±0.2 10±1Total hepatic microsomal 0.11 ± 0.01 0.10 ± 0.02 0.10 ± 0.02cytochrome P-450 content(nmollmg protein)Hepatic EROD activity 1285 ± 363 1565 ± 223 774 ± 144(pmollminlnmol P-450)- 50-ResultsIECE00.>..>0wC)4-.0.Double-Crested CormorantTotal hepatic microsomal cytochrome P-450 content was 2-3 times higher in the highlycontaminated Lake Ontario colony than in the other colonies (Table 3.5). Hepatic microsomalEROD activities (pmole/minlmg protein) were increased 4-7 fold in the Lake Ontario colonycompared with the other colonies (Figure 3.4). Hepatic EROD activities (pmole/minlnmole P450) were statistically higher in the Lake Ontario and statistically lower in the relativelyuncontaminated Saskatchewan colony than in the other colonies (Table 3.5). Yolk weightwas statistically higher in Saskatchewan and Crofton than in the other cormorant colonies.There were no differences in body, yolk-free body, any of the other organ weights, and winglength among the colonies.25002000150010005000Figure 3.4. Comparison of hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities in double-crested cormorant hatchlings from five colonies across Canada. (*): EROD activities were significantly higherin Lake Ontario than in the other colonies (p<O.O1, one-way ANOVA and Newman-Keuls multiple range test).Saskatchewan Chain Islands Christy islet Crofton Lake Ontario- 51 -ResultsTable 3.5. Morphological measurements, hepatic microsomal cytochrome P-450 content and ethoxyresorufin0-deethylase (EROD) activity (mean ± SEM) in double-crested cormorant hatchlings.Double-Crested Cormorant ColonySaskatchewan Chain Christy Islet Crofton Lake OntarioParametera (n=12) Islands (n=7) (n=7) (n=5)(n=7)Bodyweight(g) 35.6±0.6 34.0± 1.1 32.8±0.5 36.0± 1.5 32.5±0.6Yolk-free body 31.7 ± 0.7 30.9 ± 1.6 30.0 ± 0.5 32.3 ± 1.2 29.9 ± 0.6weight (g)Yolkweight(g) 3.9±0,2 2.5±0.4 2.8±0.3 3,704b 2.6±0.3Wing length (mm) 19.1 ± 0.2 18,6 ± 0.2 18.8 ± 0.2 19.0 ± 0.2 18.5 ± 0.2Liver weight (g) 0.83 ± 0.04 0.89 ± 0.04 0.77 ± 0.04 0.80 ± 0.07 0.89 ± 0.03Total hepatic P450 0.16 ± 0.01 0.14± 0.01 0.11 ± 0.02 0.12 ± 0.01 0.30 ± 0.02ccontent (nmole/mgprotein)HepaticEROD 1785247d 3530± 603 4815±588 3395 ±434 7538±348cactivity (pmole/min/nmole P-450)Note 1: the following measurements were performed by L.E.Hart at the department of Animal Science, UBC,B.C.: adrenal, kidney, stomach, intestine and tibia weight; tibia and wing length.Note 2: above data were published by Sanderson et a!. (1994b).aThere were no significant differences among colonies for (not all shown in table): body, yolk-free body, liver,stomach, heart, kidney, intestine, adrenal, spleen and bursa weights; gape width; tarsus, culmen, toe and winglength.b5igpjfjcantly different from colonies without superscripts, using one-way ANOVA (p<0.01) and NewmanKeuls multiple range test.Csigpjfictly higher than all other colonies (p<O.O1).dSigrjfictly lower than all other colonies (p<O.Ol).3.A.4. Regression of Morphological Measurements on Chemical Contamination LevelsGreat Blue HeronThe morphological measurements of the present study were combined with the data setacquired in 1988 in order to determine whether the recent observations were consistent withthe dose-effect relationships determined in 1988 (Hart et al., 1991). Linear regression- 52 -Resultsanalyses of the morphological characteristics measured in the heron chicks of 1988-1992 onboth levels of TCDD and TCDD-toxic equivalents (TEQ) in the matched eggs showed severalhighly significant relationships (p<O.Ol; n=54) with negative regression coefficients (slopes)(Table 3.6). The following morphological measurements were negatively dependent onTCDD and TEQ level: body, yolk-free body (Figure 3.5) and intestine weights, tibia wet, dryand ash weights, and tibia length. Regression of stomach weight on TCDD level wassignificant (p<O.Ol), but on TEQ level was not. All the morphological measurements thatTable 3.6. Linear regression of morphological measurements and hepatic EROD activity great blue heronchicks on log-transformed TCDD and TEQ levels in matched eggs, monitored between 1988 and 1992 (n=54).Great Blue Heron TCDD (n=54) TEQ (n=54)(nglkg) (ng/kg)Parametera r2 value slope p value r2 value slope p valueBodyweight(g) 0.21 -4.1 00006b 0.11 -5.0 O.O1l3Yolk-free body wt, (g) 0.29 -4.4 0•0005b 0.18 -5.7 0002bLiver weight (g) 0.10 -0.097 0.02 0.02 -0.069 0.35Adrenal weight (mg) 0.08 4.6 0.04 0.02 3.9 0.32Kidneyweight(g) 0.10 -0.068 0.02 0.07 -0.094 0.06Stomachweight(g) 0.14 -0.31 0.006k 0.08 -0.38 0.04Intestine weight (g)C 0.15 -0.19 00004b 0.13 -0.30 0.008’Tibialength(mm)C 0.37 -1.3 000005b 0.32 -2.0 •00005bTibia wet weight (g)c 0.20 -0.034 00009b 0.23 -0.061 00003bTibia dry weight (mg)C 0.27 -7.3 0•001b 0.18 -10 0.0011)Tibia ash weight (mg) 0.33 -1.2 0.000051 0.19 -1.6 0001bHepatic EROD activity 0.49 168 0.000051 0.25 202(pmollminlmg protein)Hepatic EROD activity 0.36 744 0.000051) 0.20 997 0001b(pmollminlnmol P-450)Note: above data were published by Sanderson et al. (1994a).aLine regressions on log-transformed TCDD and TEQ levels were not significant (p<O.O1) for the followingparameters (not all shown in table): liver, adrenal, kidney, yolk, heart, spleen and bursa weights. Regressionof stomach weight on TEQ was not significant.bA statistically significant regression (p<O.O1).csigpfficant positive correlation of parameter with body weight and yolk-free body weight (p<O.O1).- 53 -Resultswere negatively dependent on TCDD level at the p<0.Oi significance level in 1988 (n=3 1)(Hart et al., 1991), were so on both TCDD and TEQ levels with lower p values, when thedata from the present study were included (n=54). Furthermore, a negative relationshipbetween intestine weight and TCDD or TEQ level was found when the entire 1988-1992 dataset was analysed. This regression was not significant in 1988 (Hart et at., 1991). Finally,TCDD was the only chemical that negatively influenced the above measurements, againconsistent with the observations in 1988 (Beliward et at., 1990, Hart et at., 1991).600),50a) 4L>40-m 350 Nicomekl(1988)c, 30 L Vancouver(1988)D Crofton(1988)25 A Vancouver(1990)- I Crofton(1991)o 20 • Vancouver(1992)>-r20.290 I I1 10 100 1000TODD Concentration(ng/kg egg)Figure 3.5. Comparison of yolk-free body weight in great blue heron hatchlings with environmental levels of2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the matched eggs (n=54). Yolk-free body weight wasnegatively dependent on log-transformed TCDD level (r2=O.29; slope=-4.4; p<O.00005).-54-ResultsTable 3.7. Linear regression of morphological and biochemical measurements in double-crested cormorantchicks on log-transformed TCDD-toxic equivalents (TEQ) in matched eggs (n=25).Double-Crested Cormorant TEQ (nglkg)Parametera r2 value slope p valueBody weight (g) 0.25 -2.5 0.011Yolk-free body weight(g) 0.09 -1.3 0.15Yolk weight (g)C 0.44 -1.2 •00005bYolk weight (% body weight)C 0.36 -3.0 0001bYolk weight (%yolk-free body 0.37 -3.6weight)CWing length (mm)d 0.21 -0.62 0.024Total hepatic microsomal cytoebrome 0.45 +1.0 00003bP-450 content (nmole/mg protein)eHepatic microsomal EROD activity 0.69 +1506 0•0005b(pmole/minlmg protein)efgHepatic microsomal EROD activity 0.56 +4609(pmole/minlnmole P450)cNote: above data were published by Sanderson et al. (1994b).aThere was no significant regression on TEQ of (not all shown in table): body, yolk-free body, liver, stomach,heart, intestine, kidney, adrenal, spleen and bursa weights; gape width; tarsus, culmen, toe and wing length.bA statistically significant regression (p<O.Ol).CSjmilar regression of this parameter on PCB-126, 169, 105 and 118, but with lower r2 values than on TEQ.Regression of the parameter on PCB-77 or TCDD was not significant.dsignjfjcant regression of wing length on PCB-169 (r2= 0.28; slope=-O.88; p=0.007).esignificant regression, with positive slope, of parameter on PCB-77, 126, 169, 105 and 118, but with lower r2values than on TEQ.significant regression of hepatic microsomal EROD activity (pmole/minlmg protein) on untransformedhepatic microsomal P-450 content(r20.74; slope’+8993; pO.0000S).A significant regression of hepatic microsomal EROD activity on TEQ still existed when Lake Ontario datapoints were omitted(r2=O.50; slope=+748; p’0.0OO5; n=20).- 55 -ResultsDouble-Crested CormorantIn the double-crested cormorants there was a significant regression, with negative slope,of yolk weight (expressed in grams, % body weight and % yolk-free body weight) on TEQconcentration (p<O.Ol; n25) (Table 3.7). There was no significant regression of bodyweight, yolk-free body weight and wing length on TEQ at the p<O.Ol level. Regressionanalysis did show a statistically significant negative relationship between wing length andPCB-169 concentration (r2=0.28; p=O.O07). Three of the twenty five chicks included in theregression analyses had ‘watery’ edema. Omitting these three chicks from the analyses did notalter any of the coefficients of determination to a significant extent. Therefore, the presenceof watery edema did not appear to mask any potential effects of TEQ on these parameters.3.A.5. Regression of Hepatic Microsomal Cytochrome P-450 Content and ERODActivity on Chemical Contamination LevelsGreat Blue HeronSimilar to the morphological measurements, the hepatic microsomal EROD datadetermined in the present study were combined with the data of 1988 (Bellward et al., 1990).This resulted in a highly significant regression, with positive slope, of hepatic EROD activitiesof the 54 great blue heron chicks on the levels of TCDD contamination in their matched eggs(r2=0.49; p=O.00005) (Figure 3.6). This regression was also significant on TEQ (r2=0.25;pO.0001) (Table 3.6). In figure 3.6, the data points for Vancouver (1990) and (1992) andCrofton (1991) are presented using filled symbols for comparison with the data of 1988 (opensymbols). The values fitted the regression well, resulting in a typical dose-responserelationship, which confirmed the findings of Beliward et al. (1990).Vancouver Great Blue Heron ColonyIn the 1988 study (Bellward et al., 1990; Hart et al., 1991) three different great blueheron colonies were examined in the same year. Inclusion of the data from the present studyhas provided us with a sufficiently large sample size to examine one colony, Vancouver, overseveral years (1988, 1990 and 1992). Comparison of chemical contamination levels withbiological measurements in the Vancouver colony over time showed a significant regression,-56-Results600>41 C—I—>-I-4000).E30oE:fl0 o20010001 10 100 1000TCDD Concentration(ng/kg egg)Figure 3.6. Comparison of hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities (pmole/min/mg protein) in great blue heron hatchlings with environmental levels (ng/kg egg wet weight) of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in matched eggs of three colonies in British Columbia, monitoredbetween 1988 and 1992 (n=54). EROD activity was positively dependent on log-transformed TCDD level(r2=O.49; p<O.00005).o Ntcomekl(1988)t Vancouver(1988)- D Crofton(1988)A Vancouver(1990)- • Crofton(1991)• Vancouver(1992) Dr20.49AA D-57-Resultswith positive slope, of hepatic EROD activity on TCDD level (r2=O.33; p=O.OO2; n=26)(Figure 3.7) but not on TEQ(r2=O.20; p=O.O2). Regression of tibia length (r2=O.24; pO.O1),tibia dry weight (r2=O.36; p=O.OO1) and tibia ash weight (r2=O.40; p=O.0007) on TCDD levelresulted in statistically significant negative slopes. Regressions of body, yolk-free body,stomach, intestine and tibia wet weights on TCDD level were no longer significant, likely dueto the lower number of observations and overall lower levels of TCDD contamination in theVancouver colony. For the same reasons, regression of the morphological measurements onTEQ levels no longer resulted in significant relationships.6000 Vancouver(1988)-I— t Vancouver(1990)500-Vancouver(1992) 0o L. r2—O.33<- 4000E300 0w.E 0E 200-‘-a)0— 01001 10 100 1000TCDD Concentration(ng/kg egg)Figure 3.7. Comparison of hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities in great blueheron hatchlings with environmental levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in matched eggs of theVancouver colony between 1988 and 1992 (n=26). EROD activity was positively dependent on log-transformed TCDD level (r2=O.33; p<O.OO2).- 58 -ResultsDouble-Crested CormorantRegression analysis demonstrated a highly significant positive relationship betweenhepatic microsomal EROD activity (pmole/minlmg protein) and TEQ (r2=0.69; p=O 00005;n=25) (Table 3.7), resulting in a typical dose-response curve (Figure 3.8). This regressionwas still highly significant when the data for the Lake Ontario colony were omitted (r2=0.50;p=O.0005; n=20) (Table 3.7), which was not the case for any other parameter measured in thisstudy. Significant regressions of hepatic EROD activity (pmole/minlmg protein) also existedon levels of PCB-105, 118, 77, 126 and 169, individually, although these resulted in lowercoefficients of determination. A significant positive relationship between total hepaticmicrosomal cytochrome P-450 content and TEQ was also found (r2=0.45; pO.0003). The 2-3 fold increase in hepatic cytochrome P-450 content found in the cormorant chicks withincreased EROD activity (Lake Ontario) partially masked the actual increase in activity whenexpressing hepatic EROD activity on the basis of cytochrome P-450 content. Expressinghepatic microsomal EROD activity on the basis of microsomal protein content, whichremained relatively constant in our study, gave a better reflection of changes in activity. Forthis reason, regression of hepatic EROD activity expressed per milligram microsomal proteinon TEQ resulted in higher coefficients of determination. In this thesis, EROD activities havebeen consistently expressed in pmollminlmg microsomal protein for interpretation purposes.3.A.6. Immunochemical Cross-Reactivity StudyImmunoblot analysis of great blue heron and double-crested cormorant microsomesdemonstrated the presence of a hepatic protein, that was cross-reactive with monoclonalantibodies raised against rat CYP1A1. Heron microsomes from Nicomekl (1988), Vancouver(1988) and Crofton (1988) (Bellward et a!., 1990) and cormorant microsomes fromSaskatchewan, Christy Islet and Lake Ontario, both sets of three colonies in order of increasedcontamination and EROD activity, showed increased band staining intensity in the immunoblotshown in figure 3.9. (Lanes 3,4,5 and lanes 6,7,8, respectively).Although pentoxyresorufin 0-depentylase (PROD) activity was also present in the 1988heron microsomes (Bellward et al., 1990), no cross-reactivity was observed with polyclonalantibodies raised against rat CYP2B 1 and CYP2B2.-59 -Results3.5 •0 Chain Islands__3.0- 1 Christy Islet VLJ Crofton- 0 2 5 - ‘ Saskatchewan VC) V Lake Ontario0E 2.0 r20.69Lii.—E 1.5 V-I- 01.0 00.50.0I I I30 100 300 1000 3000TCDD—Toxic Equivalents(ng/kg egg)Figure 3.8. Comparison of hepatic microsomal ethoxyresorufin 0-deethylase (EROD) activities (nmole/min/mg protein) in double-crested cormorant hatchlings with environmental levels (nglkg egg wet weight) ofTCDD-toxic equivalents (TEQ) in matched eggs (n=25). EROD activity was positively dependent on log-transformed TEQ level(r2=O.69; p<O.00005). This regression was still significant when the Lake Ontariodata points were omitted (r2=O.50; p=O.0005; n=20).- 60-Results234Figure 3.9. Immunoblot showing the cross-reactivity of monoclonal antibodies against rat cytochrome P-4501A1 with a protein in individual avian hepatic microsomes. The lanes contained 20 pg of hepatic microsomalprotein (avian hepatic EROD activity in pmole/minlmg protein are in parentheses):1. Uninduced male rat. 6. Saskatchewan cormorant (211).2. 3-MC-induced male rat. 7. Christy Islet cormorant (787).3. Nicomekl (1988) heron (56). 8. Lake Ontario cormorant (2538).4. Vancouver (1988) heron (248). 9. Buffer control.5. Crofton (1988) heron (434).56- 61 -3.B. IN OVO DOSE-RESPONSE STUDYThe main purpose of the in ovo dose-response study was to characterize the hepaticEROD induction response of TCDD in several avian species; this in order to delineatespecies dfferences in sensitivity of a response, which is under direct regulation of the Ahreceptor, and to compare the controlled induction ofEROD by TCDD with that caused byenvironmental exposures to polychiorinated aromatic hydrocarbons.3.B.1. Effects of3H-TCDD on Length of Incubation, Duration of Pip and MortalityDomestic chickens, domestic pigeons, great blue herons and double-crested cormorantswere exposed in ovo to various doses of3H-TCDD during the latter third part of incubation.Several incubation characteristics were measured: length of incubation, duration of pip andegg mortality. Length of incubation was not affected by treatment with any of the TCDDdoses in the chicken (21.4 days) and pigeon (19.1 days). Length of incubation of the wild birdspecies could not be determined reliably due to the uncertain age of the eggs, which wereestimated from the candling scheme in figure 2.2. Consequently, heron and cormorant eggswere exposed to TCDD for variable durations (Table 3.8).Table 3.8. The duration of exposure (days; mean ± SEM) to vehicle alone or3H-TCDD from time of injectionuntil day of hatch (day of sacrifice) in great blue heron and double-crested cormorant embryos.TCDD dose Great Blue Heron Double-Crested(ugIkg egg) n n Cormoranto 16 6.8 ± 1.2 8 3.3 ± 0.70.5 7 5.6±1.3 - -1.0 5 3.8±1.0 6 4.2±1.03.0 7 7.2 ± 1.2 5 4.2 ± 0.910 12 7.9± 1.1 8 3.8±0.7100 4 4.5±0.9 8 3.9±0.6Note: there were no significant differences between dose groups in the duration of exposure, using one-wayANOVA (p<O.05).- 62 -ResultsTable 3.9. Mortality in day-old chicken, pigeon, great blue heron and double-crested cormorant hatchlingsexposed in ovo to various doses of3H-TCDD during the latter third part of incubation.3H-TCDD Chicken Pigeon Great Blue Heron Double-Crested)ose Cormorant(pg/kg egg)incidence (%) incidence (%) incidence (%) incidence (%)o 3/16 (19) 2/14 (14) 0/16 (0) 0/8 (0)0.01 0/8 (0) - - -0.03 2/8 (25) - - -0.125 0/8 (0) - - -0.25 4/8 (50) 1/8 (13) - -0.5 3/8 (38) - 3/10 (30) -1.0 3/8 (38) 1/8 (13) 5/10 (50) 0/6 (0)3.0 3/16 (19) 2/8 (25) 1/8 (13) 3/8 (38)10 - 3/8 (38) 4/16 (25) 1/9 (11)100- 7/12 (58) 3/7 (43) 2/10 (20)Note: statistical analyses were not performed, due to the low incidences and small sample sizes.Duration of pip was not different in any of the treated birds compared with the controlbirds, for each of the four species. In the precocial chicken, the duration of pip (days; mean ±SEM) was 0.6 ± 0.1; in the semi-altricial heron it was 1.0 ± 0.1; in the fully altricial pigeonand cormorant they were 2.1 ± 0.2 and 2.0 ± 0.1, respectively. Effects of TCDD on eggmortality in the four bird species are shown in Table 3.9.3.B.2. Effects of311-TCDD on Body and Organ WeightsIn the chicken, body, yolk-free body, kidney and yolk weights were not significantlydifferent among the dose groups (Table 3.10). The only exception was an increase in liverweight in hatchlings from the highest dose group (3 pg/kg TCDD). Similarly, in the pigeon,heron and cormorant, there were no differences in bird weight or any of the measured organweights among the dose groups (Tables 3.11-3.13).- 63 -ResultsTable 3.10. Bird and organ weights (mean ± SEM) in day-old domestic chicken hatchlings exposed in ovo tovarious doses of3H-TCDD (uglkg egg), during the latter third part of incubation.Domestic Chicken3H-TCDD Egg weight Bird weight Liver weight Kidney weight Yolk weightDose (g) (g) (g) (g) (g)(jig/kg egg) no 13 51.2 ± 1.0 39.2 ± 0.8 0.72 ± 0.01 0.25 ± 0.01 5.01 ± 0.410.01 8 54.8± 1.6 41.7± 1.1 0.92±0.07 0.27±0.02 4.72±0.250.03 6 51.3 ± 1.5 37.3 ± 0.9 0.83 ± 0.07 0.22 ± 0.01 3.90 ± 0.630.125 8 53.5 ± 1.0 41.7 ± 1.1 0.87 ± 0.05 0.32 ± 0.02 5.44 ± 0.290.25 4 53.3 ± 0.6 42,4 ± 0.7 0.74 ± 0.06 0.26 ± 0.03 7.06 ± 1.070.50 8 52.8 ± 1.1 40.5 ± 1.1 0.77 ± 0.03 0.30 ± 0.01 6.67 ± 0.411.0 5 52.3 ± 1.0 41.6 ± 0.8 0.80 ± 0.03 0.27 ± 0.01 5.92 ± 0.313.0 13 56.5 ± 1.9 42.4 ± 1.2 1.05 ± o.o6a 0.30 ± 0.02 6.01 ± 0.57aSjjjjcant1y higher than control, using one-way ANOVA (p<O.05) and Newman-Keuls multiple range test.Table 3.11. Bird and organ weights (mean ± SEM) in day-old domestic pigeon hatchlings exposed in ovo tovarious doses of3H-TCDD (pg/kg egg), during the latter third part of incubation.Domestic Pigeon3H-TCDD Egg weight Bird weight Liver weight Kidney weight Yolk weightDose (g) (g) (g) (g) (g)(pg/kg egg) n0 12 20.3 ± 0.4 14.9 ± 0.5 0.47 ± 0.04 0.19 ± 0.01 1.80 ± 0.230.25 7 22.9±0.5 16.0± 1.4 0.58±0.04 0.19±0.02 1.77±0.331.0 7 21.2±0.5 14.8±1.1 0.49± 0.04 0.19± 0.01 1.82± 0.403.0 6 20.9 ± 0.8 15.6 ± 0.7 0.44 ± 0.02 0.16 ± 0.01 1.93 ± 0.2010 5 20.1 ± 1.5 14.6 ± 1.3 0.48 ± 0.09 0.19 ± 0.01 1.67 ± 0.43100 5 23.0 ± 2.5 14.2 ± 0.6 0.51 ± 0.01 0.17 ± 0.02 0.96 ± 0.23Note: there were no significant differences among dose groups (p<O.05), using one-way ANOVA.- 64 -ResultsTable 3.12. Bird and organ weights (mean ± SEM) in day-old great blue heron hatchlings exposed in ovo tovarious doses of3H-TCDD (ug/kg egg), during the latter third part of incubation.Great Blue Heron3H-TCDD Egg weight Bird weight Liver weight Kidney weight Yolk weightDose (g) (g) (g) (g) (g)(fig/kg egg) n0 16 69.7± 1.3 53.3± 1.1 1.21 ± 0.02 0.77± 0.04 5.42 ±0.440.5 7 68.0± 1.4 51.6± 1.4 1.12±0.08 0.71±0.04 6.10±0.821.0 5 67.5±2.6 52.2±2.2 1.17±0.10 0.74±0.04 5.74±0.423.0 7 70.1 ± 1.8 53.0 ± 1.4 1.27 ± 0.03 0.76 ± 0.04 5.46 ± 0.4810 12 69.8 ± 1.1 52.0 ± 1.0 1.18 ± 0.04 0.73 ± 0.03 5.58 ± 0.39100 4 72.7 ± 1.3 57.3 ± 0.8 1.36 ± 0.04 0.81 ± 0.07 5.66 ± 0.43Note: there were no significant differences among dose groups (p<O.O5), using one-way ANOVA.Table 3.13. Bird and organ weights (mean ± SEM) in day-old double-crested cormorant hatchlings exposedin ovo to various doses of3H-TCDD (ugfkg egg), during the latter third part of incubation.Double-Crested Cormorant3H-TCDD Egg weight Bird weight Liver weight Kidney weight Yolk weightDose (g) (g) (g) (g) (g)(pg/kg negg)0 8 43.6±0.8 32.1±0.8 0.79±0.06 0.56±0.3 2.71±0.201.0 6 45.9± 1.1 33.5± 1.1 0.92±0.05 0.55±0.04 3.10±0.423.0 5 42.5 ± 1.6 31.1 ± 1.9 0.88 ± 0.07 0.54 ± 0.05 2.46 ± 0.2310 8 43.4± 1.3 31.5± 1.1 0.94± 0.05 0.57± 0.02 2.82± 0.34100 8 42.6±0.9 30.8±0.8 0.92±0.04 0.52±0.02 2.90±0.46Note: there were no significant differences among dose groups (p<O.05), using one-way ANOVA.- 65-ResultsThe same statistical outcomes were obtained when organ weights were expressed as %body weight or % egg weight; note that there were no differences among treatment groups inegg weight at the time of injection and in bird weight after hatch.3.B.3. Incidence of Edema and other Gross AbnormalitiesThe incidence of edema increased dose-dependently in hatchlings of the domesticchicken, the great blue heron and the double-crested cormorant (Table 3.14). This wasparticularly clear in the double-crested cormorant. In each species, the incidence of edema inthe control group was zero. In the chick, the first cases of edema (2/8) were observed in thelowest dose group of 0.01 pg/kg TCDD. Similarly, in the heron, the first cases (4/7) occurredat the lowest injected dose of 0.5 pg/kg. In the cormorant, the first case of edema (1/5) wasnot observed until the 3 pg/kg dose group. Edema was not observed in the pigeon at any ofthe dose groups (0.25-100 pg/kg), indicating a remarkable species difference in sensitivity ofthis response to TCDD.Table 3.14. Incidence of subcutaneous edema in hatchlings of four bird species exposed in ovo to variousdoses of3H-TCDD (jig/kg egg), during the latter third part of incubation.3H-TCDD Chicken Pigeon Great Blue Heron Double-CrestedDose Cormorant(pg/kgegg)incidence (%) incidence (%) incidence (%) incidence (%)0 0/13 (0) 0/12 (0) 0/16 (0) 0/8 (0)0.01 2/8 (25)---0.03 2/6 (33)---0.125 3/8 (38)---0.25 2/4 (50) 0/7 (0)--0.5 5/8 (63)- 4/7 (57) -1.0 3/5 (60) 0/7 (0) 3/5 (60) 0/6 (0)3.0 6/13 (46) 0/6 (0) 3/7 (43) 1/5 (20)10- 0/5 (0) 9/12 (75) 3/8 (38)100- 0/5 (0) 1/4 (25) 6/8 (75)Note: statistical analyses were not performed, because of the low incidences and small sample sizes.- 66-ResultsOne of the 12 great blue herons treated with 10 pg/kg TCDD, hatched with anunresorbed yolk sac. Similarly, one of the 5 pigeons that hatched in the 100 pg/kg dose grouphad an unresorbed yolk sac.3.B.4. Induction of Ilepatic Microsomal Cytochrome P-450 Content3H-TCDD caused a dose-dependent increase in hepatic microsomal cytochrome P-450content in the hatchlings of all four avian species (Table 3.15). Basal (control) hepaticcytochrome P-450 contents were considerable different among bird species (Table 3.15) andwere 2-10 times lower than those found in untreated rat livers (1.0 nmollmg microsomalprotein). Cytochrome P-450 content of the liver was increased at considerably lower doses inthe chicken than in the other birds (Table 3.15 and Figure 3.10).Table 3.15. Increase in total hepatic cytochrome P-450 content (nmol/mg microsomal protein; mean ± SEM)in hatchlings by in ovo exposure to3H-TCDD during the latter third period of incubation: comparison of thedomestic chicken, domestic pigeon, great blue heron and double-crested cormorantTotal Hepatic Cytochrome P-450 Content(nmol/mg microsomal protein)3H-TCDD Chicken Pigeon Great Blue Double-CrestedDose (ug/kg Heron Cormorantegg)0 0.41±0.03 0.13±0.01 0.10±0.01 0.21±0.020.01 0.30 ± 0.03---0.03 0.44 ± 0.03---0.125 O.70±0.04a - --0.25 1.1±o.osa 0.16±0.03--0.5 1.1 ± o.o5a- 0.1 ± 0.01-1.0 1.3±o.o8a 0.75±0.13a 0.09±0.01 0.22±0.023.0 1.3 ± o.oSa 0.97 ± 0.12k’ 0.13 ± 0.01 0.20 ± 0.0410- 1.1 ± 0.20 ± o.o2a 0.36 ± o.o5a100- 1.7 ± 0.2’ 0.21 ± 0.07’ 0.86 ± O.06’asigcant1y higher than control, using one-way ANOVA (p<O.O1) and Newman-Keuls multiple range test.- 67-ResultsTCDD Dose (ug/kg egg)Figure 3.10. The increase in total hepatic microsomal cytochrome P-450 content in avian hatchlings after inovo exposure to various doses of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin, during the latter third part ofincubation: comparison of the domestic chick, domestic pigeon, great blue heron and double-crestedcormorant (n values and standard errors are shown in table 3.15).In the chicken hatchling, the maximum increase in hepatic cytochrome P-450 contentoccurred at 0.25-3 pg/kg and was 3.3-fold; in the heron it occurred at 10-100 pg/kg and was2.4-fold. In the pigeon and cormorant, no maximum level was attained at the highest doseexamined (100 pg/kg); hepatic cytochrome P-450 content was increased 4.2-fold and 13.7-fold, respectively.-7//• • chick- 0 pigeonA heronD cormorant• /••A0__I • I • I •14-I- -o12C) 10a0>LOb;E40.1 0.3 1 3 1030100- 68-Results3.B.5. Induction of Hepatic Microsomal Ethoxyresorufin 0-Deethylase ActivityA dose-dependent increase in hepatic microsomal EROD activity was observed in allfour species of bird (Figures 3.11-3.14). Absolute activities expressed in nmolJmin/mgmicrosomal protein were highest in chick (0.50 ± 0.07), intermediate in the cormorant (0.26 ±0.04) and lower in the pigeon (0.12 ± 0.02) and heron (0.062 ± 0.005).3H—TCDD Dose (twig/kg egg)Figure 3.11. Induction of hepatic microsomal ethoxyresorufin 0-deethylase in domestic chicken hatchlingsinjected in ovo with H-2,3,7,8-tetrachlorodibenzo-p-dioxin on day of incubation 16 (n values in parentheses).Control activities were 0.50 ± 0.07 nmol/min/mg protein. Highest observed induction was 29-fold abovecontrol, at 3 dug/kg.18>%_% 16:c>.-z-2 14o< ° 12OOCEc%%LU CII!(1 3)(5)(8)(4)(8)20(13) (6)(8)-VA I . I • I0.01 0.03 0.1 0.3 1 3 10- 69-Results>%-l—._>Q,.— .4—-*-0w.EE•%4%%-I--- q)a,r-7,/II I I I0.1 0.3 1 3 10 30 1003H—TCDD Dose (ug/kg egg)(5)(5)(6)(7) 3.12. Induction of hepatic microsomal ethoxyresorufin 0-deethylase in domestic pigeon hatchlingsinjected in ovo with various doses of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin on day of incubation 14 (n valuesin parentheses). Control activities were 0.12 ± 0.02 nmollminlmg protein. Highest observed induction was3 0-fold above control, at 100 pg/kg.(12) (7)I • I • I- 70-Results3H—TCDD Dose (ug/kg egg)Figure 3.13. Induction of hepatic microsomal ethoxyresorufin 0-deethylase in great blue heron hatchlingsinjected in ovo with H-2,3,7,8-tetrachlorodibenzo-p-dioxin during the latter third part of incubation (n valuesin parentheses). Control activities were 0.062 ± 0.005 nmol/min/mg protein. Highest observed induction was35-fold above control, at 100 pg/kg.(4)>%-I.—>-4-00lii0.—-Ic,ci=Ca,0L00)EEa,0E2. 0.3(12)(7)(5)(7)1 3 10 30 100- 7] -Results3H—TCDD Dose (ug/kg egg)Figure 3.14. Induction of hepatic microsomal ethoxyresorufin 0-deethylase in double-crested cormoranthatchlings injected in ovo with3H-2,3,7,8-tetrachlorodibenzo-p-dioxin during the latter third part ofincubation (n values in parentheses). Control activities were 0.26 ± 0.04 nmol/minlmg protein. Highestobserved induction was 22-fold above control, at 100 pg/kg.I I I(8)>%-Ia—>a—-4-C)<0liiC)a—-Ic,a)=a)-4-00c3)E.cp0EC6543210//r6)//I0.1 0.3(8)(5)I1 3 10 30 100- 72 -Results0-IC0C)w>00C,-oCI—-D0‘I1 3 10 30 100(ug/kg egg)Figure 3.15. Induction of hepatic ethoxyresorufin 0-deethylase activities by3H-2,3 ,7,8-tetrachlorodibenzo-p-dioxin: comparison of the dose-response curves in domestic chicken, domestic pigeon, great blue heron anddouble-crested cormorant hatchlings (n values and standard errors are shown in figures 3.11-3.14).A striking species difference existed in sensitivity to hepatic EROD induction by TCDD.The dose-response curve in the chicken was positioned between 1 and 2 orders of magnitudeto the left of those in the other avian species. In figure 3.15, it is not possible to determineED50 values for lack of an observed maximal response in the pigeon, heron and cormorant atthe highest dose of TCDD injected (100 pg/kg). However, no-observed-effect (NOEL) andlowest-observed-effect levels (LOEL), as approximate descriptors of the lower end of thedose-response curves, could be determined (Table 3.16).—7//I I.ADchickpigeonheroncormorant>%-II35302520151050lI I0.010.03 0.1 0.3TCDD Dose- 73 -ResultsTable 3.16. Estimated no- and lowest-observed effect levels (NOEL and LOEL) of 2,3,7,8-tetrachlorodibenzo-p-dioxin for hepatic ethoxyresorufin 0-deethylase induction in the domestic chicken and pigeon, great blueheron and double-crested cormorant.The immunoblot shown in figure 3.16, is an independent means of demonstrating(semi-quantitatively) the dose-dependent induction of CYP1A1 by TCDD in 2 of the 4 birdspecies treated with TCDD: the great blue heron and the domestic chicken. Cross-reactivitywas not seen in a male and female juvenile great blue heron that had been raised fromFigure 3.16. Immunoblot showing cross-reactivity of monoclonal antibodies against rat CYP 1A 1 with a hepaticmicrosomal protein in hatchlings of the domestic chicken and the great blue heron exposed in ovo to 2,3,7,8-tetrachlorodibenzo-p-dioxin.1. Rat (untreated).2. Rat (3-MC-treated).3. Chick (untreated).4. Chick (0.125 pg/kg TCDD).5. Chick (1 pg/kg TCDD).6. Heron (untreated).7. Heron (3 pg/kg TCDD).8. Heron (10 pg/kg TCDD).Avian Species NOEL (jig/kg egg) LOEL (jig/kg egg)Domestic chicken 0.01 0.03Domestic pigeon 0.25 1.0Great blue heron 1 3Double-crested cormorant 1 3hatchling stage and fed a controlled diet (see section 2.1.D).9. Male juvenile heron (untreated).10. Female juvenile heron (untreated).- 74 -Results0>C)CC)I-—•— chick—— pigeon3.B.6.3H-TCDD Organ DistributionIn order to determine whether the observed species difference in sensitivity to hepaticEROD induction by TCDD could be due to differences in the uptake of3H-TCDD by theliver, the concentrations of3H-TCDD in the liver, and also in yolk and blood, were measuredin the four species of bird. H-TCDD-derived radioactivity could be set equal to TCDDconcentration because of the slow metabolism of TCDD in avian embryos (Nosek et a!.,1992b). Tables 3.17-3.20 list the organ concentrations of3H-TCDD in chicken, pigeon,heron and cormorant, respectively. Concentrations of3H-TCDD in the livers of the birdswere similar among species and increased proportionally with increased dose (Figure 3.17).The same was tme for blood and yolk.10001001010.10.01TCDD Dose (ug/kg egg)Figure 3.17. 3H-TCDD concentrations found in livers were similar among bird species and increasedproportionally with dose (n values and standard errors are listed in tables 3.17-3.20).—A— heron—E— cormorantI I0.01 0.1 1I I10 100- 75 -ResultsTable 3.17. The distribution of3H-TCDD (mean ± SEM) among liver, yolk and blood in day-old domesticchicken hatchlings, exposed in ovo by injection into the air sac on day of incubation 16.Domestic Chicken 3H-TCDD Concentration3H-TCDD n Liver Yolk BloodDose (jig/kg (ng/g liver) (ng/g yolk) (ng/ml blood)egg)o 13 0 0 00.01 8 0.091 ± 0.005 0.018 ± 0.001 n.d.0.03 6 0.26 ± 0.02 0.069 ± 0.01 n.d.0.125 8 0.58±0.03 0.19±0.01 0.02±0.0020.25 4 0.92 ± 0.2 0.32 ± 0.05 0.03 ± 0.0040.5 5 2.3±0.2 0.73±0.05 0.07±0.011.0 5 4.2±0.2 1.4±0.1 0.10±0.013.0 13 12±0.4 5.3±0.3 0.35±0.03n.d.: not detectable.3H-TCDD concentrations were below limit of detection, defined as 10-times backgroundor about 200 cpm.Table 3.18. The distribution of3H-TCDD (mean ± SEM) among liver, yolk and blood in day-old domesticpigeon hatchlings, exposed in ovo by injection into the air sac on day of incubation 14.Domestic Pigeon 3H-TCDD Concentration3H-TCDD n Liver Yolk BloodDose (pg/kg (ng/g liver) (ng/g yolk) (ng/ml blood)egg)0 12 0 0 00.25 7 0.88 ± 0.1 1.1 ± 0.3 0.07 ± 0.0011.0 7 4.6±0.2 2.8±0.4 0.17±0.023.0 6 12 ± 4 6.1 ± 0.2 0.23 ± 0.0310 5 65±8 40± 16 1.4±0.2100 5 503 ±66 390±28 15±4- 76-ResultsTable 3.19. The distribution of3H-TCDD (mean ± SEM) among liver, yolk and blood in day-old great blueheron hatchlings, exposed in ovo by injection into the air sac during the latter third part of incubation.Great Blue 3H-TCDD ConcentrationHeron3H-TCDD n Liver Yolk BloodDose (pg/kg (nglg liver) (nglg yolk) (ng/ml blood)egg)0 16 0 0 00.5 7 2.1 ± 0.1 1.8 ± 0.2 0.15 ± 0.031.0 5 3.9±0.6 3.0±0.5 0.26±0.043.0 7 14 ± 0.9 13 ± 0.9 0.84 ± 0.0410 12 48±4 47±5 2.3 ±0.3100 4 463 ±66 399 ±42 20±3Table 3.20. The distribution of3H-TCDD (mean ± SEM) among liver, yolk and blood in day-old double-crested cormorant hatchlings, exposed in ovo by injection into the air sac during the latter third part ofincubation.Double-Crested 3H-TCDD ConcentrationCormorant3H-CDD n Liver Yolk BloodDose (pg/kg (ng/g liver) (nglg yolk) (ng/ml blood)egg)0 8 0 0 01.0 6 2.8±0.3 2.5±0.5 0.19±0.023.0 5 11±2 7.9±1.5 0.71±0.0810 8 31±5 19±3 2.0±0.3100 8 364± 53 263 ±50 31±6- 77-ResultsConcentrations of3H-TCDD correlated highly significantly among organs, within eachspecies (Table 3.21). These correlations indicated that3H-TCDD in corn oil, injected into theair sac, distributes itself in a predictable, dose-dependent manner over the compartmentsmeasured. Furthermore, the equations in table 3.21 may be used to estimate organconcentrations of TCDD in future air sac injection studies, by determining TCDDconcentrations in a small (50-100 p1) blood sample.Table 3.21. Linear correlation analysis of log-transformed blood concentrations with log-transformed liverand yolk concentrations in four bird species.Correlations between Blood and Organ Concentrations:x=log-blood concentration (ng TCDD/ml blood)y=log-organ concentration (ng TCDD/g tissue)Species Organ r value Correlation EquationChicken Liver 0.97 y=1.56+1.03xYolk 0.96 y1.24+1.14xPigeon Liver 0.96 y1.59+1.35xYolk 0.86 y1.32+1.08xGreat Blue Liver 0.96 y=1 .24+ 1 .02xHeron Yolk 0.98 y=1. 19+1 .06xDouble-Crested Liver 0.98 y=1. 17+0.95 lxCormorant Yolk 0.98 y=1.02+0.913xNote: all correlations were significant at the p<O.00005 level.Expressing the amount of3H-TCDD distributed to the organs as a percentage of thetotal dose injected (% Dose) showed no differences among dose groups for each species ofbird. It did indicate some differences among species: the percentage of the injected dosefound in the liver was 1.2-1.5 times higher in the pigeon than in the other species (Table 3.22).The % dose found in the yolk was significantly higher in the pigeon and great blue heron thanin the other species (Table 3.22).- 78 -ResultsTable 3.22. Percentages of the injected doses of3H-TCDD found in the livers and yolks in day old hatchlingsof four species of birds, exposed in ovo during the latter third period of incubation.Note: There were no differences among doses for each species in the percentage of the injected dose found inthe liver and yolk.asignificantly higher than in the other species, using one-way ANOVA (p<O.O1) and Newman-Keuls multiplerange test.3.B.7.Relationship betiveen Hepatic Uptake of3H-TCDD and Duration of In OvoExposureIn the great blue heron and double-crested cormorant, the duration of exposure to 3H-TCDD was variable, because of the uncertain age of the embryos at the time of injection. Theexposure times could be determined after the birds had hatched, within one day of accuracy.To see whether the duration of exposure had an effect on the concentration of3H-TCDDfound in the liver, hepatic TCDD concentrations in the heron and cormorant hatchling wereplotted against duration of in ovo exposure. At any one given dose, the duration of exposurehad no effect on the liver concentration of TCDD in each species (data not shown). However,these comparisons have a small number of observations. To obtain a sufficiently large samplesize for comparison, all birds of a given species needed to be included. In order to do so,differences in doses had to be accounted for; this was done by transforming ‘concentration ofTCDD in the liver’ to ‘percentage of the dose found in the liver’. The percentage of injecteddose found in the liver of the heron and cormorant hatchling was plotted against the durationof in ovo exposure to3H-TCDD in figures 3.18 and 3.19, respectively. No change wasobserved the heron liver (Figure 3.18). In the cormorant, there was a statistically significant(p=O.O3) increase in % dose found in liver over duration of exposure (Figure 3.19). However,judged from the slope (0.66 % per day), this increase was very small.Tissue Chicken Pigeon Great Blue Double-CrestedHeron CormorantLiver 8.9±0.5 11±la 7.8±0.4 7.1±0.6Yolk 17±2 30±4k 32±2a 15±2- 79 -Results181614121086- 0 0.5 pg/kg- t1g/kgi: 3 ag/kgD- V-I-a)>-JC.—a)Cl)0V 10 pg/kg0 100 ag/kgVVV_nVVI • I • I •Li yV0I • I • I4200 2 4 6 810121416Duration of Exposure (d)Figure 3.18. Percentage of the injected dose found in the liver of the great blue heron hatchling, after varioustimes of in ovo exposure to3H-TCDD (n=30). There was no relationship between the two variables (p<O.05).ts 1 gig/kgi: 3g/kg VV 10 ag/kg0 100/.Lg/kg 00L.a,>-JCa)(I)0181614121086420VDVVDx0I I I I Io 1 2345678Duration of Exposure (d)Figure 3.19. Percentage of the injected dose found in the liver of the double-crested cormorant hatchling, aftervarious times of in ovo exposure to3H-TCDD (n=27). The % dose in liver correlated positively with time ofexposure to hatch (r=O.43; y=4.3+O.66x; p=O.O3).- 80-Results3.B.8. Comparison of Hepatic EROD Activities with Hepatic3H-TCDD Concentrationsin Four Avian Species: Detailed Dose-Response Curves.The concentration of TCDD in the liver was plotted against hepatic EROD activity foreach bird species, resulting in the detailed dose-response relationships shown in figures 3.20-3.23. Typical sigmoidal relationships were observed in the heron and cormorant, with (closeto) maximal responses at 100 pg/kg TCDD (Figure 3.22 and 3.23). The shape of the dose-response curve in the domestic chicken was less sigmoidal (Figure 3.20). Particularly at thehighest dose of 3 pg/kg, the response in the chick was highly variable. In the pigeon, therelationship between liver concentration of TCDD and hepatic EROD activity was linear,without an indication of a maximal response within the dose range examined (Figure 3.21).From figures 3.20-3.23, ED50 values for hepatic EROD induction could be determined withgreater reliability in the chicken, heron and cormorant, but not in the pigeon (Table 3.23).In the heron and cormorant, exposure times to3H-TCDD were variable (section 3.B.6).Duration of exposure should be recognized as a third variable in figures 3.20-3.23. However,in each species, there were no significant differences in duration of exposure, among doses(Table 3.8). Also, there appeared to be no systematic effect of duration of exposure onhepatic EROD activity within a given dose group (example in figure 3.22). In addition,maximal EROD induction by TCDD is known to take place within 24 hours in chickenembryos and be sustained for at least 7 days (Rifkind et a!., 1985). Therefore, variability inhepatic EROD induction due to variable exposure times is expected to be negligible.Table 3.23. ED50 values for hepatic microsomal EROD induction in the hatchlings of four avian speciesexposed in ovo to 2,3,7,8-tetrachlorodibenzo-p-dioxin during the latter third period of incubation.Species ED50 ED50(ng TCDD/g liver) (pg TCDD/kg egg)Domestic chicken 1-2 0.1Domestic PigeonGreat blue heron 30-50 3-10Double-crested cormorant 20-30 3-10Note: ED50 in pigeon could not be determined (see figure 3.21).- 8] -Results30—//D0 control>% 25 - 0.01 jig/kgA 0.03 pg/kg.— -I—V 0.125g/kgO 20-0.25 ag/kg_0 0.5 pg/kgE 1pg/kgD 3kg/kgBE r20.77-I— in0—I”CC V70_7// I • I • I0.1 0.3 1 3 10 30[3H—TCDD] (ng/g liver)Figure 3.20. Hepatic microsomal ethoxyresorulin 0-deethylase activities plotted against liver3H-TCDDconcentrations in the domestic chicken hatchling (controls: n13; treated: n49).- 82 -Results-v/I I3.5C control__0 0.25 pg/kg- 1kg/kg0 3g/kg 00 V 10kg/kg2.5- 0 100 pg/kg0 r2=O.88L .Dlii.El.5-I-ol.O 0 EJa,:tc 00.50.0I I3 10 30 100 300 1000[3H—TCDD] (ng/g liver)Figure 3.21. Hepatic inicrosomal ethoxyresorufin 0-deethylase activities plotted against liver3H-TCDDconcentrations in the domestic pigeon hatchling (controls: n=12; treated: n26).- 83 -Results3.0—//0 control2.5 0 0.51ug/kg 63ts 1/Lg/kgz 3g/kg0- 2.0 V 10 pg/kg100kg/kg 5E r2O.83 gVyaEIn-I— - VDV9V0.01 0I I0 1 3 10 30 100 300 1000[3H—TCDD] (ng/g liver)Figure 3.22. Hepatic microsomal ethoxyresorufin 0-deethylase activities plotted against liver3H-TCDDconcentrations in the great blue heron hatchling (controls: n16; treated: n=35). Numbers in parentheses,accompanying symbols, indicate duration of exposure of those data point; there appeared to be no systematiceffect of duration of exposure on hepatic EROD activity or3H-TCDD concentration (section 3.B.7).-84-Results8—//07 - 0 controlz 1g/kg>%•-I- Q).— —I-—>0 6-3g/kgV1Opg/kgU0) 5 (> 1OOg/kgO E r2=O.79rx0VQ%l%% 3.—_; DW°-E2a) D1z2D0I I1 3 10 30 100 300 1000[3H—TCDD] (ng/g liver)Figure 3.23. Hepatic microsomal ethoxyresorufin 0-deethylase activities plotted against liver3H-TCDDconcentrations in the double-crested cormorant hatchling (controls: n=8; treated: n27).- 85 -3. C. AH RECEPTOR BINDING STUDY3.C.1. Specific Binding of3H-TCDD in Avian Hepatic Cytosolic PreparationsAh receptor binding affinities of TCDD in hepatic cytosol were determined to explainthe observed differences in hepatic EROD-inducing potency of TCDD among the four birdspecies. The Ab receptor binding experiments were performed in the laboratory of A.B.Okeyat the University of Toronto. The experience of this laboratory with the detection of Ahreceptor in human tissues, such as placenta, allowed for use of this tissue as a positive controlin comparison studies with our avian liver samples. Considerably less specific binding wasdetected in chick liver cytosol compared with human placental cytosol (Figure 3.24).8000—0-- TCDD+TCDF6400 / —•— TCDD:::1600BSA CAT , ..00 5 10 15 20 25Fraction #12108D 6:. 4200 5 10 15 20 25Fraction #Figure 3.24. Comparison of specific binding of 2,3,7,8-tetrachlorodibenzo-p-dioxin (10 nM) in adult humanplacental cytosol (top) with domestic chick hatchling hepatic cytosol (bottom), analysed by velocitysedimentation on sucrose density gradients. TCDD*3HTCDD; TCDF=2,3,7,8-tetrachlorodibenzofuran._0_ TCDD+TCDFTCDD*-86-Results7005604202801400-0 25Fraction #Figure 3.25. Treatment with dextran-coated charcoal (1 mg charcoal/mg cytosolic protein) completelyremoves specific binding of 2,3,7,8-tetrachlorodibenzo-p-dioxin (1 nM) in pigeon hatchling hepatic cytosol.The effect of treatment with dextran-coated charcoal (1 mg charcoal/mg cytosolicprotein), generally used to remove non-specific binding, on the specific binding in pigeoncytosol is shown in figure 3.25. Specific binding of TCDD was completely lost after treatmentwith dextran-coated charcoal at 1 and 0.1 mg charcoal/mg cytosolic protein. For this reason,all radioligand binding studies were performed in the absence of charcoal.Examples of specific binding of TCDD in the 9-10 S region of avian cytosolic fractionsanalysed on sucrose density gradients are shown for day-old hatchlings of chick, pigeon, greatblue heron and double-crested cormorant in figures 3.26-3.29, respectively. Specific bindingin the same region was also observed in the hepatic cytosol of an adult heron (Figure 3.30).The specific binding occurred at sedimentation coefficient of 9-10 5, corresponding to that ofhuman cytosolic Ah receptor protein, and did not occur in the presence of 100-fold molarexcess of TCDF. Although the specific binding of TCDD in the avian cytosolic fractions(both adult and hatchling) was extremely small, the quality of the gradients (judged by thesedimentation profile of the BSA and catalase (CAT) markers) and the presence of TCDF ineach analysis made reliable quantitation of the specific binding possible.TCDD* + TCDFTCDD*TCDD*IF+DCCTCDD*+DCC5 10 15 20-87-Results0__ TCDD_u_ TCDD+TCDFFigure 3.26. Specific binding of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin (1 nM) in 9-10 S region of 1 day-olddomestic chick hepatic cytosol, analysed by velocity sedimentation on sucrose density gradients.0Figure 3.27. Specific binding of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin (1 nM) in 9-10 S region of 1 day-olddomestic pigeon hepatic cytosol, analysed by velocity sedimentation on sucrose density gradients.70060050040030020010000 5 10 15 20 25Fraction #1400120010008006004002000_._ TCDD*—ci— TCDD*+TCDF0 5 10 15 20 25Fraction #- 88 -Results0).Ca-coI__I 0.CI4C—0-u,IIIFigure 3.29. Specific binding of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin (1 nM) in 9-10 S region of 1 day-olddouble-crested cormorant hepatic cytosol, analysed by velocity sedimentation on sucrose density gradients._._ TCDD—LI— TCDD+TCDF‘•.I5040__30201000 5 10 15 20 25Fraction #Figure 3.28. Specific binding of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin (40 nM) in 9-10 S region of 1 day-old great blue heron hepatic cytosol, analysed by velocity sedimentation on sucrose density gradients.605040TCDD*30 TCDD* + TCDF201000 5 10 15 20 25Fraction #- 89 -Results45361 27 TCDD—El— TCDD*+TCDF18I900 5 10 15 20 25Fraction #Figure 3.30. Specific binding of3H-2,3,7,8-tetrachlorodibenzo-p-dioxin (1 nM) in 9-10 S region of adultgreat blue heron hepatic cytosol, analysed by velocity sedimentation on sucrose density gradients.3.C.2. Woolf and Scatchard Analyses of3H-TCDD Specific BindingAn example of a Scatchard, Woolf and saturation plot for the binding of3H-TCDD inthe hepatic cytosol of the great blue heron hatchling is illustrated in figure 3.31. The apparentequilibrium dissociation constants (Kcj) and apparent total number of specific binding sites(Bmax) are summarized for each bird species in table 3.24 . The results of the Woolf plotshad higher correlation coefficients (r) and were used to calculate 1d and Bm values (section2.10). The apparent specific binding affinity (Kcj) in the livers of the domestic chickenhatchlings were about an order of magnitude lower than that in the other species (Table 3.24).The apparent total number of binding sites (apparent Bm) were somewhat lower in thechicken hatchling than in the other species.- 90-Results0.0100.00800IN-D0.00400.0020.0000)E00I20.0_____5 155 10 15 20 25 30 35 40 45E3H-TCDD] nMKd-9.3 nMBmax—18.1 fmolelmgr—-0.970 10 20Bound (fmolelmg)600500400030000 20010000 1000 2000 3000 4000 5000 6000 7000 8000 9000Free (fmolelmg)Figure 3.31. Saturation, Scatchard and Woolf plot of TCDD specific binding in day-old heron liver cytosol.- 91 -ResultsTable 3.24. Specffic binding characteristics of3H-TCDD in four avian species: comparison of the apparentreceptor affinity (Td) and apparent number of receptors (Bmax), among the domestic chick, domestic pigeon,double-crested cormorant and great blue heron (chick and adult).Species Sample Woolf Plota Scatchard PlotKd Bm r Kd Bm r(nM) (fiuole/mg) (nM) (fmolelmg)Domestic chickb pool #1 1.6 10.5 0.99 1.1 9.5 -0.93pool #2 0.75 4.3 1.0 0.76 4.3 -0.96Domesticpigeon pool#l 13.9 24.4 0.96 10.9 22.7 -0.75chickb pool #2 10.8 21.2 0.96 18.0 28.4 -0.79Double-crested B8 15.8 41.8 1.0 14.8 40.5 -0.99cormorant chick C B29 12.0 43.9 0.99 15.1 49.9 -0.96Greatblueheron HL1A 10,1 18.7 1.0 9.3 18.1 -0.97chickc HUB 19.6 26,7 0.95 32.5 38.0 -0.69Great blue heron cross-bill 15.4 42.5 0.88 15.9 46.2 -0.76adultd duplicate 15.4 48.8 0.95 19.4 58.2 -0.72awooff plot analyses were used to derive apparent Kd and Bm values.bCiiick and pigeon specific binding characteristics were determined in cytosolic fractions prepared from twoseparate pools of four livers per pool.C}{eron and cormorant specific binding characteristics were determined in cytosolic preparations of twoindividual chick livers per species.dAdult heron specific binding characteristics were determined in duplicate in the cytosolic preparation of onecross-billed heron.- 92 -4. DISCUSSION4.A. ENVIRONMENTAL MONITORING STUDYThe environmental monitoring study was designed to determine the association ofconcentrations ofpolychiorinated aromatic hydrocarbons with biological effects in greatblue heron and double-crested cormorant embryos. This association was examined overseveral locations with dfferent levels of contamination and, in the case of the heron, overseveral years with changing levels of contamination. The use of hepatic EROD activitiesin these birds as sensitive biomarkers of exposure to polychlorinated aromatichydrocarbons is discussed, together with the use of TEFs and TEQs.4.A.1. Rapid Decline of Levels of PCDDs and PCDFs in Great Blue Heron EggsIn continuation of the monitoring of great blue herons in British Columbia, eggs werecollected from two colonies which had elevated levels of PCDDs and PCDFs in 1988. Levelsof these chemicals have declined rapidly in the Strait of Georgia since 1988 (Whitehead et al.,1992) and this was reflected in the levels found in the heron eggs in 1990-1992 (Table 3.1).The decrease in contamination with PCDDs and PCDFs followed remedial process changesimplemented by the Crofton pulp and paper mill between 1988 and 1990. The elimination ofthe use of pentachlorophenol-treated wood chips and several changes in the bleaching process,including dO2 substitution, have led to a decrease in the amounts of many PCDDs andPCDFs discharged into the environment (Whitehead et al., 1992). Fish-eating birds areexposed to PCDDs and PCDFs mainly through their diet. The rapid decline in contaminantlevels in the heron eggs, therefore, reflects a rapid decrease in contaminant levels in their prey.Great blue heron diet on the B.C. coast consists mainly of starry flounder (Platichthysstellatus), threespine stickleback (Gasterosteus aculeatus), Pacific staghorn sculpin(Leptocottus armatus) and shiner perch (Cymatogaster aggregata), of about one to two yearsof age (Butler, 1991). By 1991, these young fish would have hatched and grown in anenvironment less contaminated with PCDDs and PCDFs than previously. In addition, theseyoung fish would have had only a short time to accumulate persistent contaminants.- 93 -DiscussionTherefore, any changes in the input of PCDDs and PCDFs into the environment will be rapidlyreflected in young fish and subsequently in their predator, the great blue heron.4.A.2. Negative Effects of PCDDs, PCDFs and PCBs in Monitored Bird HatchlingsGreat Blue HeronBody, yolk-free body and several organ weights of great blue heron hatchlings werenegatively associated with levels of TCDD and TEQ in the matched eggs (Table 3.6). This isin accordance with the well documented growth retardation and body weight loss caused byTCDD and similarly acting compounds in many laboratory species. In the wild, Hoffman etat. (1986) found a similar correlation (r=-O.61; p=O.O4; n=12) between yolk-free embryoweight and PCB residue levels in black-crowned night herons (Nycticorax nycticorax) fromthe San Francisco Bay area. Other toxicities associated with PCB exposure in black-crownednight heron hatchlings from contaminated colonies in the Great Lakes included reduced femurlength-to-body weight ratios, decreased hepatic DNA concentrations, increased incidences ofedema and elevated hepatic aryl hydrocarbon hydroxylase activities (Hoffinan et at., 1993).Our observations of reduced body and organ weights, reduced tibia length and increasedincidence of edema, associated with TCDD (and TEQ) exposure in great blue heronhatchlings, are consistent with the findings of Hoffman and coworkers. These observationscontinue to support the notion that increased levels of polychlorinated aromatic hydrocarbons,with Ah receptor affinity, are negatively affecting the health of avian species in contaminatedareas.Double-Crested CormorantThe double-crested cormorant hatchlings showed no signs of overt toxicity, other thanthe induction of hepatic EROD activity. Yolk weight decreased with increased TEQ, but thetoxicological significance of this is not clear. Linear regression of body weight and winglength on TEQ resulted in negative relationships with significance levels between 0.01 and0.05. These may be indications that some effects on growth were occurring, but were notlarge enough to be unequivocal.- 94 -DiscussionVan den Berg et al. (1992) studied cormorant chicks (Phalacrocorax carbo) from twocolonies in the Netherlands with differing levels of contamination with PCDDs, PCDFs andPCBs and differing reproductive successes. They found a significant positive correlationbetween EROD and the sum of mono-ortho-PCBs (r=O.58; p<O.O2; n16), consistent withour results in double-crested cormorants in Canada. However, in contrast to the findings inthis thesis, they detected a positive correlation between levels of these PCBs and yolk weight(r=O.62; p<O.O2). In addition to a species difference, there were inevitably differences in localenvironmental conditions, such as climate and food supply, and differences in methodology,such as incubation conditions. These factors can all influence yolk weight and may be partiallyresponsible for our differing findings, although they do not necessarily explain opposingfindings. Furthermore, the Dutch study lacked a relatively uncontaminated reference colony;levels of mono-ortho-PCBs in the lowest contaminated cormorants from the Netherlands werecomparable to those found in the Lake Ontario colony. PCB levels in the most contaminatedcormorants from the Netherlands were ten-fold higher than in the Lake Ontario colony. Thisresulted in considerably higher TEQs found by Van den Berg et at. compared to those in thepresent study. Therefore, assuming a similar sensitivity of the two species of cormorant,increased toxicities would be expected in the Dutch study.Yamashita et al. (1993) found deformities, such as crossed-bills and clubbed feet, in 10-12 and 21-24 day old live double-crested cormorant embryos from colonies in the upper GreatLakes, which had total TEQ levels similar to those in our Lake Ontario colony. We did notdetect such deformities in the 11 Lake Ontario chicks that hatched in our laboratory. Thesample size used in our study was too small to detect deformities reliably at incidence rates of6-10%, as were found in the most contaminated Great Lakes colonies by Yamashita andcoworkers. Another possibility is that under the artificial incubation conditions used in ourstudy, deformed embryos may have died before hatching.4.A.3. The Use TEFs and TEQs in Biological Monitoring of Avian SpeciesGreat Blue HeronThe highly significant positive relationship between hepatic microsomal EROD activityin great blue heron hatchlings and TCDD and TEQ levels in the matched eggs confirms- 95 -Discussionprevious findings by Beliward et al. (1990). The coefficient of determination was increasedowing to the larger sample size. Total TEQ was not as good a predictor of EROD activity(r2=0.25) as TCDD alone (r2=0.49) (Table 3.6), again in accordance with Beliward et at.(1990). The same was true for any of the morphological characteristics. The reason for this islikely the presence of relatively high levels of TCDD in the great blue heron eggs in 1988,making TCDD the largest single contributor to the TEQ (Figure 3.1). The use of TEQsprobably resulted in lower r2 values because of the extra variability introduced by includingestimates of the non-ortho-PCBs (section 2.13) or because some TEF values are notappropriate for this particular species. The TEFs derived by Safe (1990) are based onmammalian studies and do not necessarily apply to avian species. Also, the use of TEFs in anadditive model does not take into consideration synergistic and antagonistic effects in acomplex mixture of chemicals, which act on the same receptor.Double-Crested CormorantUnlike in the great blue heron, the best coefficients of determination (r2) in the double-crested cormorant were obtained regression of the parameters on the sum of TCDD-toxicequivalents. As pointed out above, the nature of contamination with polychiorinated aromatichydrocarbons was considerably different in the herons in British Columbia. In the herons,TCDD was the major contributor to the total TEQ. In the cormorant, however, TCDD levelswere lower, and the relative contribution of other similarly acting chemicals, such as the non-and mono-ortho-PCBs, at relatively high concentrations increased in importance. This wasparticularly the case in the Lake Ontario colony. Furthermore, the detailed analysesperformed by R.J.Norstrom did not identify any polyhalogenated aromatic hydrocarbons withAh receptor affinity other than those listed in tables 3.1 and 3.2 (personal communication).Although in the cormorants, regressions of hepatic microsomal EROD activity and yolkweight on individual levels of PCB-105, 118, 126 and 169, were also significant, theseregressions resulted in lower r2 values than the regressions on TEQ. In other words, the TEQwas able to account for more of the variability in these parameters than any one particularchemical. For these reasons, comparisons with TEQ rather than one individual chemical arepreferred in order to determine the effects of a complex environmental mixture of chemicals- 96-Discussionon the double-crested cormorant. Furthermore, the regression of hepatic EROD activity onTEQ was still significant when the highly contaminated Lake Ontario birds were omitted(Figure 3.8, legend). In other words, TEQs can be used as predictors of avian hepatic ERODactivities over a relatively small concentration range and small degree of environmentalinduction. This further supports the applicabilty of TEQs as tools to predict an Ah receptor-mediated effect in double-crested cormorant embryos, exposed to the complex mixture ofpolychiorinated aromatic hydrocarbons found in the Canadian environment. In conclusion,despite the fact that the TEFs used to calculate the TEQs are based on mammalian studies(Safe, 1990), they were successfully applied to regress hepatic microsomal EROD activity onTEQ level in the double-crested cormorant.4.A.4. Use of Avian Hepatic EROD Activities in Biological MonitoringGreat Blue HeronA highly significant positive relationship between hepatic microsomal EROD activity ingreat blue heron hatchlings and TCDD and TEQ levels in their matched eggs, was found.This relationship was consistent over several locations and over several years. Figure 3.6demonstrates that the hepatic EROD activities in the contaminated birds of 1988 (Vancouverand Crofton) have decreased with the decline in contamination over the subsequent years,following a typical dose-response relationship. A similar pattern was observed with themorphological parameters which regressed negatively on TCDD and TEQ levels (e.g. yolk-free body weight in figure 3.5). Furthermore, concomitant with the decrease in chemicalcontamination and hepatic EROD activity in the Crofton colony, there was a decrease in theincidence of edematous hatchlings in both colonies (Table 3.3) and an increase in thereproductive success of the Crofton colony (observation by the Canadian Wildlife Service). Inaddition, the regression equations obtained through analysis of the Vancouver colony alonewere similar to those of the entire data set (section 3 .A.5). The above results are consistentwith our conclusion that the negative effects observed in the contaminated heron chicks in1988 were a result of the elevated levels of PCDDs and PCDFs and not, for instance, due tocolony-dependent factors, such as genetic differences or differences in food supply. Takentogether, our results provide further evidence that the health of the great blue herons was-97-Discussionaffected by the increased levels of TCDD and other PCDDs and PCDFs in 1988, but hasimproved over the last few years, since the remedial changes implemented by the Crofton pulpand paper mill in 1989, Finally, the above findings support the use of hepatic EROD activityas a biological indicator of exposure to chemicals whose toxicities are mediated by the Ahreceptor.Double-Crested CormorantA highly significant positive relationship between hepatic microsomal EROD activity indouble-crested cormorant hatchlings and TEQ level in the matched eggs was found in birdsfrom several locations across Canada (Figure 3.8). Hepatic EROD induction was observedbefore any obvious signs of toxicity, although trends were visible for decreased body and winglength (section 4.A.2 and Table 3.7). Hepatic EROD induction appears to be a sensitiveresponse to polychlorinated aromatic hydrocarbon exposure (expressed as TEQ), which canserve as a relatively ‘early-warning’ of the toxicities of these chemicals.4.A.5. Immunochemical Cross-Reactivity StudyThe results of the immunoblotting experiments demonstrated cross-reactivity of double-crested cormorant and great blue heron hepatic microsomes with rat CYP1A1 antibodies.The positive relationship between intensity of the stained band and in vitro hepatic microsomalEROD activity indicates that ethoxyresorufin is a suitable substrate for CYP 1A1 in these birdspecies and that induction of the enzyme has occurred. Ronis et a!. (1989) demonstrated thepresence of CYP1A1, using rat CYP1A1 antibodies, in six fish-eating bird species, includingthe great cormorant (Phalacrocorax carbo) and shag (Phalacrocorax aristotelis) collectedfrom the Irish Sea. The fact that cross-reactivity could be observed in these birds indicatesthat significant induction of the enzyme had taken place, considering that in the present studya band was not visible in the relatively uncontaminated double-crested cormorants fromSaskatchewan and great blue herons from Nicomekl (1988). Indeed, Ronis et a!. (1989)found significant levels of total PCBs in the brain and fat of the seabirds. In the greatcormorant (n=4), 1 and 2 ppm total PCBs were detected in the brain and 50 and 675 ppm infat (Ronis eta!., 1989). Walker and Ronis (1989) found the mean relative hepatic microsomal- 98 -DiscussionEROD activities (expressed in pmole/minlmg protein) of seven adult fish-eating bird species tobe 2.3-fold higher than that of the uninduced male rat, using a formula that corrects forspecies differences in liver weight as a percentage of body weight (below):Hepatic EROD activity, bird Liver/body weight, birdRelative activity =_______________________________x_________________________Hepatic EROD activity, male rat Liver/body weight, male ratThe results in this thesis suggest that the activities detected do not represent the normalphysiological or ‘uninduced’ state of these birds, but are the result of a certain degree ofinduction due to environmental exposure to persistent polyhalogenated aromatichydrocarbons. For example, hepatic EROD activities (pmole/min/mg protein; mean ± SEM)in our six unexposed juvenile great blue herons averaged 96 ± 9; their body and liver weights(kg; mean ± SEM) were 2.2 ± 0.1 and 0.03 ± 0.01, respectively. Using the formula above, theaverage relative EROD activity of these herons was about 0.3 compared to that of uninducedmale rat (rat data were obtained from Ronis and Walker, 1989). Most other cytochrome P45 0-associated activities, such as aldrin epoxidase, aminopyrine N-demethylase (Walker andRonis, 1989) and erythromycin N-demethylase (Beliward et al., 1990) and total hepaticmicrosomal cytochrome P.45 0 content are also very low in fish-eating birds.The lack of cross-reactivity of double-crested cormorants and great blue herons withpolyclonal rat CYP2B 1/2 antibodies is consistent with other reports in fish-eating birds(Walker and Ronis, 1989). This lack of cross-reactivity, despite the presence of PRODactivity, suggests that either the avian CYP2B isoenzymes are quite dissimilar in structurefrom mammalian, or that PROD is metabolized by a different avian cytochrome P-450 thanCYP2B 1/2. In line with the latter point, Rattner et at. (1993) found that phenobarbitalinduced black-crowned night herons did not have elevated PROD activities, while 3-MCinduction led to elevated EROD, AHH and PROD activities. Bellward et at., (1990) found ahighly siginificant correlation between EROD and PROD activity in great blue heronmicrosomes (r=0.625; p<O.0001; n31). These finding suggest that pentoxyresorufin is, at- 99 -Discussionleast partly, metabolized by the same enzyme as ethoxyresorufin. As for the presence of aCYP2B-type enzyme in avian species, Rattner et at. (1993) detected a large increase in cross-reactivity of polyclonal antibodies raised against scup (Stenomatus chrysops) CYP2B with amicrosomal protein in the black-crowned night heron. Although they report an over 2000-fold increase in band density in the phenobarbital-treated herons (determined by densitometry),this number means little, because they did not detect any cross-reactivity in the untreatedmicrosomes (they compared the treated values to one-half the detection limit of theimmunoassay). Furthermore, an immunoblot was never shown in the publication, neither dothey mention which secondary antibody and detection system is used. Therefore, their resultsshould be treated with caution as preliminary findings.4.B. DOSE-RESPONSE STUDY4.B.1. Effects of3H-TCDD on Incubation Time and Morphological MeasurementsThe present study was designed to describe the dose-response relationship betweeninduction of hepatic microsomal EROD activity, an index of CYP1A1 induction, and in ovoTCDD exposure in several avian species. It was not designed to examine the teratogeniceffects or chronic effects of TCDD. Chicken, pigeon, great blue heron and double-crestedcormorant eggs were exposed to varying doses of3H-TCDD in the latter third of incubation.Exposure to TCDD at such a late time in incubation was not expected to cause any effects onmorphological parameters such as hatchling bird and organ weights, and on incubationparameters such as length of incubation, duration of pip and egg mortality. Indeed, no sucheffects were observed in the four bird species examined in the present study. The onlyexception was a slight but statistically significant (p<O.O5) increase in liver weight in thedomestic chick hatchling at the highest tested dose of 3 pg/kg (Table 3.10).4.B.2. Incidence of EdemaSurprisingly, jelly-like’ subcutaneous edema was observed in many of the treated birds,after a relatively brief time of in ovo exposure (Table 3.14). In the chick, heron andcormorant, the first cases were observed at doses less than or equal to those required for the-100-Discussioninduction of hepatic EROD activity. This indicates that the subcutaneous edema, which is arelatively selective response of avian exposure to TCDD and other chemicals that activate theAh receptor, is also a sensitive and rapid response. In line with the documented highsensitivity of this response in domestic chickens (Gallus gallus) (Flick et al., 1972; 1973;Firestone, 1973), we found the first cases of edema (3/8) in the lowest treatment group of0.01 pg/kg. Rifkind et al. (1985) found an incidence of subcutaneous edema of 28% in 19day old chicken embryos exposed to 0.04 pg/kg TCDD on DI 10. Although the treatmentconditions differed, these results are similar to those of our study, in which a 25% incidence ofedema was observed in chicken hatchlings exposed to 0.03 pg/kg TCDD on DI 16. The dose-dependency of the incidence of edema observed in the heron and chicken was not as clear as inthe cormorant hatchling. The sample sizes were likely too small to detect incidences of edemareliably. It is also possible that in the higher dose groups edematous or otherwise deformedembryos may have died. The occurrence of subcutaneous edema appears to be a responsewith considerable inter-individual variation and also inter-species variabilty. The pigeonembryo (Columba livia) was resistant to TCDD-induced edema within the dose rangeexamined (0.25-100 pg/kg); this despite the induction of hepatic EROD activity at doseshigher than 1 pg/kg. Resistance to the formation of subcutaneous edema has also beenobserved in the turkey embryo (Meleagris gallopavo) upon in ovo exposure to a range ofdoses from 0-1000 pg/kg 3,Y,4,4’-tetrachlorobiphenyl (PCB-77), also an Ah receptor agonist(Brunstrom and Lund, 1988) and a TEF of 0.05. Nosek et al. (1993) did not observe edemain ring-necked pheasant embryos (Phasianus coichicus) treated with doses ranging from 0-100 pg/kg TCDD. In the study by Nosek et al., pheasant embryos were exposed to TCDD onDI 0, resulting in a longer exposure earlier on in incubation compared with our study.Therefore, increased sensitivity to the teratogenic effects of TCDD would be expected.Indeed, they found that mortality was the most sensitive adverse effect of TCDD in thepheasant eggs, indicating that chicks died before expressing sublethal toxicities. Whether thesame is true for the pigeon embryos in our study is uncertain. Considering that the Ahreceptor binding affinity (section 3.C) and hepatic EROD inducing potency of TCDD (section3 .B) in the pigeon are similar to that in the heron and cormorant, which do exhibit edemaunder our experimental conditions, it is not likely. The lack of response in the pigeon embryo- 101 -Discussionmay be due to genetic differences in the pathway that leads to the formation of subcutaneousedema in this species or at this stage of development.4.B.3. Induction of Hepatic Microsomal EROD ActivityDetailed dose-response curves were produced for the induction of avian hepaticmicrosomal EROD activity by TCDD, in the late-stages of embryonic development. Thisallows for the comparison, among the four bird species, of the sensitivity of a response whichis under direct regulation by the Ah receptor, namely the induction of CYP1A1. The dose-response curve for hepatic micro somal EROD induction by TCDD in the domestic chickembryo was similar to that determined by Rifkind et a!. in 1985. Rifkind and coworkersexposed chicken eggs via the albumin and determined EROD activities in 9000 g supernatantfractions of pooled livers. Considering these methodological differences, the ED50 of 0.1pg/kg egg found in our study was comparable to their ED50 of 0.06 pg/kg. In this thesis, theED50 values reported for the dose-response curves in the pigeon, heron and cormorant wereestimated to be 30-50 times higher than that of chick. The more detailed dose-responsecurves showing hepatic TCDD concentration versus EROD activity (section 3.B.7; figures3.22 and 3.23) allow for a more accurate determination of ED50 values in the heron andcormorant. These ED50 values were 3-10 pg/kg in both wild bird species (Table 3.23).Recently, Nosek et a!. (1993) reported an ED50 of 0.312 pg/kg for the induction of hepaticmicrosomal EROD activity by TCDD in ring-necked pheasant hatchlings (Phasianuscoichicus). These chicks were exposed to TCDD in ovo on DI 0. Nosek et a!. found theLD50 in pheasant embryos to be 1.3-2.2 ng/kg, approximately 5-10 times higher than theLD50 of 0.25 pg/kg reported in the domestic chicken embryo (Mired and Strange, 1977).Clearly, the ring-necked pheasant embryo is less sensitive to both hepatic EROD induction andacute toxicity (death) than the domestic chick (Nosek et a!., 1993). However, using the ED50for hepatic EROD induction as an indicator of sensitivity to Ah receptor-mediated toxicities,the heron and cormorant appear to be considerably less sensitive (>10 times) than thepheasant. It should be noted that there was a difference in route of exposure between thepresent study and that ofNosek and coworkers. They exposed eggs by injection into the yolkor albumin, in contrast to our less invasive air sac injection. These different injection routes-102-Discussionare expected to result in differences in distribution (Janz and Beliward, unpublished data).Therefore, reliable comparisons of sensitivity can only be made if the actual concentrations ofTCDD in the target tissues are determined, as has been done in the present study. Otherstudies showing species differences among birds in sensitivity to chemicals with Ah receptoraffinity have been also reported (Brunström and Reutergárdh, 1986; Brunström and Lund,1988). The ED50 of 3,Y,4,4’-tetrachlorobiphenyl for hepatic AHH induction was about tentimes less in the turkey embryo compared with the chicken embryo (BrunstrOm and Lund,1988). The turkey embryo was also considerably less sensitive to 3,3,4,4’-tetrachlorobiphenyl-induced thymic atrophy. These species differences may be due torespective differences in Ah receptor affinity, although differences in metabolism cannot beruled out. Brunström and Lund (1988) attempted to detect Ah receptor in the turkey, but nospecific binding was observed.4. C. AHRECEPTOR BINDING STUDY4.C.1. Ah Receptor Binding Characteristics of TCDD in Four Avian SpeciesSpecific binding of3H-TCDD in the 9-10 S region of sucrose density gradients wasdetected in the hepatic cytosolic preparations of chick, pigeon, heron and cormorant. Thesedimentation characteristics of specific binding were similar to those of the Ah receptorfound in human placental cytosol (Manchester et al., 1987) and in rat and mouse liver cytosol(Denison et at., 1986b).The average values for the equilibrium dissociation constant and apparent number of Ahreceptors found in 1 day old chick hepatic cytosol in the present study (Kd=O.75 and 1.6 nM;Bmax4.3 and 10.4 fhole/mg protein) were comparable to those found by Brunström andLund in 1988 (Kdl .0 nM; Bm,l 1 fmole/mg protein). However, Brunström and Lund usedthe hydroxylapatite method (Poellinger et at., 1985) instead of sucrose density gradientanalysis to determine these binding parameters. The findings of BrunstrOm and Lund appearto be abberrant, because the extensive experience of Dr. Okey’s laboratory is that thehydroxylapatite assay gives Kd values about an order of magnitude lower than sucrose densitygradient analysis without affecting Bm values (personal communication). Consistent with- 103 -Discussionthis, Denison et at. (1986c) found a KdO. 16 nM and a Bma,S8 fhiole/mg protein in 7 dayold chick embryos, using the hydroxylapatite method. They also found that the number ofreceptors was relatively high at this early time in development and decreased to post-hatchvalues of about 10 to 15 fInole/mg protein (Figure 1.3), closer to the levels found in our 1 dayold chicks. Although it is difficult to compare binding data produced by very differentmethods, we can draw conclusions from comparison of the binding data within our study,because of consistent use of the sucrose density gradient analysis method. Furthermore, byderiving binding kinetics from sucrose density gradient analyses (in the presence and absenceof competitor), instead of by hydroxylapatite affinity analysis, we have greater confidence thatwe are actually measuring specific binding in the 9-10 S region, which is the expectedsedimentation coefficient of cytosolic Ah receptor.The pigeon, heron and cormorant hatchling had similar average Kd values, which werean order of magnitude higher than the chick (Table 3.24). Limited availability of cytosolicmaterial allowed us to do two determinations per species only. In the case of the pigeon andcormorant the Kd values for the two individuals were similar. In the case of the chick andheron, the Kd values were about two-fold apart.Hepatic Ah receptor concentrations appeared to be lower in the chicken hatchling thanin the other species (Table 3.24). Bm values, as a measure of receptor concentration, arestrongly dependent on the method of tissue preparation and assay conditions used (Denison ela!., 1986b; Manchester et at., 1987; Okey et at., 1989). It is also not clear what the biologicalsignificance of relatively small differences in receptor levels are among species. They do notappear to account for most tissue and species differences in susceptibility to TCDD (Whitlock,1990). Therefore, it would be unwise to assign too much importance to the differences inBm values in the present study.4.C.2. Comparison of Ah Receptor Affinities in Avian Livers with Human PlacentaAh receptor has been detected in a number of human tissues, such as lung (Roberts eta!., 1986), tonsil (Lorenzen and Okey, 1991) and placenta (Manchester et a!., 1987). Ahreceptor in human placenta had Kd values ranging from 5.5 to 7.9 nM and Bm values from122 to 239 fmole/mg protein (Manchester et al., 1987). In the same study, human placental-104-Discussiontissues were compared to rat liver Ah receptor which had a Kd=2.4 nM and a Bml87fmole/mg protein. These results were obtained using sucrose density gradient analyses,allowing for comparison with our findings. The binding affinity of cytosolic Ah receptor in thelivers of our chicken hatchlings was slightly higher than that of adult rat and about 5 timeshigher than that of human placenta. However, the binding affinity of the Ah receptor in thepigeon, heron and cormorant in our study, was generally 5 to 10 times lower than that of ratand about 1 to 3 times lower than human placenta. The relative binding affinities can be rank-ordered as follows: chick embryonic liver>adult rat liver>human placenta>pigeon embryonic1iverheron embryonic livercormorant embryonic liver. In other words, the wild birdsexamined in this thesis appear to have Ah receptor binding affinities for TCDD, at the lowerend of the affinity-spectrum, but in the same order of magnitude as those in human placenta.4.C.3. Ah Receptor Binding Affinities as Determinants of the Toxic Potency of TCDDIt is tempting to suggest that the binding affinity of TCDD and other related ligands forthe Ah receptor determines their toxic potency in a particular species. This appears to be truefor certain inbred strains of mice (section 1 .A. 3.1) and forms the basis of the TEF concept(section 1 .A. 3.4). However, most species differences in sensitivity to TCDD and similarlyacting chemicals cannot be explained solely on the basis of different Ah receptorconcentrations or affinities. There are several other known processes that take place afterinitial binding of TCDD to the Ah receptor, which may be different among species. The rateof translocation of ligand-receptor complex to the nucleus (translocation efficiency) andaffinity for the Ah-responsive elements on the DNA are also determinants in theresponsiveness of a particular cell or species to TCDD and other ligands for the Ah receptor(Okey et al., 1994). Furthermore, in cases other than TCDD, which is usually poorlymetabolized and excreted, pharmacokinetic differences should also be taken intoconsideration.Despite the fact that Ah receptor binding affinities are not able to fully predict the toxicpotency of a ligand and species differences in susceptibility, the binding of the ligand to thereceptor is a prerequisite for many of the toxicities caused by persistent polyhalogenated-105-Discussionaromatic hydrocarbons (Okey et a!., 1994). Therefore, Ah receptor binding affinities remainimportant determinants of the potential susceptibility of a species to these chemicals.4.D. TOXICOLOGICAL IMPLICATIONS OF ENVIRONMENTALINDUCTION OF CYPJA]The three studies in this thesis describe certain effects of environmental exposure ofthe great blue heron and double-crested cormorant to TCDD and related chemicals, therelative sensitivities of the heron, cormorant, chicken and pigeon to TCDD and the Ahreceptor binding affinities of TCDD in the four avian species. Comparison of the resultsof these three studies, has allowed me to put the relative susceptibilities of the great blueheron and double-crested cormorant embryo to TCDD into the context of exposures topolychiorinated aromatic hydrocarbons in the environmentThe relationship between hepatic micro somal EROD activity and logarithmic TEQ orTCDD level is of a curvilinear nature at the observed environmental concentration range inboth the heron and the cormorant (Figures 3.6 and 3.8). Such a curvilinear pattern suggeststhat we are measuring these avian hepatic EROD activities at the lower end of the steeplyrising part of their dose-response curves for induction by TCDD and related compounds. Thisis confirmed by the findings of the in ovo dose-response study. The environmental andlaboratory exposures are compared in figure 4.1. (It should be noted that the environmentalexposure is throughout incubation of the developing embryo as opposed to the latter third ofincubation in the dose-response study.) Under these circumstances, a further increase inenvironmental contamination with TCDD and related compounds would be expected to resultin a large increase in EROD activity and further activation of Ah receptor-mediated toxicities,such as edema and body weight loss; these were observed in the great blue heron chicks andshown to decrease with declines in TEQ. In the double-crested cormorant, these toxicitieswere not observed. The number of chicks examined was too small to detect incidences ofdeformities and other toxicities reliably. It could also be that adversely affected embryos diedbefore expressing any toxic symptoms post-hatch. Although overt toxicities were not-106-Discussionobserved in double-crested cormorant chicks at the p<O.Ol significance level, there appearedto be some trends: reduced body weights (p=O.Ol 1) and wing lengths (p=O.O24) wereassociated with increased TEQ levels. Therefore, increased TEQ levels would be expected tolead to increased Ah receptor-mediated toxicities, such as decreased body and organ weights,in the cormorant as well. The fact that the double-crested cormorant embryo appears to beslightly less sensitive to the toxicities of TCDD is consistent with the observed difference inEROD-inducing potency of TCDD (Figure 3.15). However, this difference is small; thelimited Ah receptor binding data do not indicate a clear difference in Ah receptor bindingaffinity between the two species (Table 3.24).The results of the present study have led us to suggest that elevated hepatic CYP1A1levels imply that other Ah receptor-mediated toxicities, with equal or greater sensitivity tochemicals such as TCDD than hepatic EROD induction, are also occurring. For instance, theED50 for immunosuppression by TCDD has been found to be one to two orders of magnitudelower than that for hepatic EROD induction in laboratory-raised deer mice (Peromyscusmaniculatus) (Gard et al., 1993). Evidently, certain Ah receptor-mediated toxicities occur atlower doses of TCDD than required for EROD induction. Which of these toxicities would beexhibited depends on the genetic make-up of species in question. In the double-crestedcormorant none were observed. In the case of the great blue heron, some of these toxicitiesappear to be growth retardation, the occurrence of jelly-lik& subcutaneous edema and,speculatively, decreased reproductive success.-107-Discussion>%%-4-c>a,I— :--0lii.E-I>%‘-I.-i-.00)HTCDD Dose (ug/kg egg)0.1 0.3 1Figure 4.1. Comparison of environmental exposures to TCDD in great blue heron hatchlings (top) and TCDDtoxic equivalents in double-crested cormorant hatchlings (bottom) with the respective in ovo dose-responsecurves constructed in the laboratory.-v/I I I IenvTron mentallevels2. I I0.1 0.3 1 3 10 30 100-v/I I Ien vi ron m e n t a Ilevels-7//II • I I3 10 30 100-108-Discussion4. E. FUTURE DIRECTIONSThis thesis has examined a limited number of end-points of toxicity of TCDD andrelated chemicals in avian embryos. The focus of this thesis was on the use of CYP1A1induction as a biomarker of exposure and toxicity of TCDD and related contaminants in wildbird embryos. The relatively short durations of exposure, during the latter third part ofincubation, were suitable in order to study CYP1A1 induction without causing excessmortality due to the teratogenic effects of TCDD. However, we are ultimately interested inthe environmental toxicities of TCDD and related chemicals in the wild birds. Therefore,exposures earlier during incubation are required in order to examine relevant toxicities such asteratogenesis, immunotoxicity, decreased reproductive capacity and effects on varioushormones. Dose-response relationships for these toxicities need to be determined andcompared with that for CYP1A1 induction, in order to draw stronger conclusions on thetoxicological implications ofCYP1A1 induction in the wild.In continuation of this project, Dave Janz is currently studying the effects of TCDD,injected early on during incubation, on several end-points of toxicity in domestic chicken andpigeon embryos. These include thyroid and sex hormones and estrogen receptor levels. In thefuture, these studies will be expanded to include wild bird species. In addition, otherchemicals with Ah receptor affinity, and also bleached kraft mill effluent, will be examined forthese embryotoxicities.- 109 -5. CONCLUSIONS5.A. ENVIRONMENTAL MONITORING STUDYAt the levels of TCDD and similarly acting chemicals recently found in the Strait ofGeorgia, effects in great blue heron and double-crested cormorant hatchlings were observed.Heron and cormorant chicks showed increased hepatic microsomal EROD activity in areas ofhighest contamination with polychlorinated aromatic hydrocarbons. This indicates that theinduction of CYP1A1 has taken place. In support of this, monoclonal antibodies against ratCYP1A1 recognized a microsomal protein in the livers of the avian chicks. The intensity ofthe band correlated positively with the EROD activity of the hepatic microsomes. Theinduction of CYP1A1 indicates that the Ah receptor-mediated process, by which TCDD andrelated polychlorinated aromatic hydrocarbons exert many of their known toxicities, has beenactivated. Consistent with this, great blue heron hatchlings demonstrated an increasedincidence of edema and decreased body weights, yolk-free body weights and organ weights inareas with increased contamination, such as Crofton in 1988. Obvious toxicities were notobserved in the limited number of cormorant hatchlings examined. The results of theenvironmental monitoring study were in good agreement with the findings of 1988. Ofparticular interest is the observation that hepatic EROD activities and other toxicities in theherons decreased as PCDD and PCDF contamination declined. We are reasonably confidentthat the negative effects observed in the great blue heron over the last few years have beenreversed and that the health of the chicks in the Crofton colony has improved, concomitantwith the decrease in levels of contamination observed since 1988.A significant positive relationship between hepatic EROD activity and TCDD or TEQlevel was found in great blue heron hatchlings. This relationship was consistent acrosslocation and time. A significant positive relationship between hepatic EROD activity and TEQwas found in double-crested cormorant hatchlings from different locations across Canada.This relationship was still significant when the highest contaminated cormorants from LakeOntario were omitted from the analysis. Furthermore, in the cormorant, the induction ofhepatic EROD activity occurred before any signs of overt toxicity. These observationssupport the use of EROD activities as relatively sensitive biomarkers of environmental-110-Conclusionsexposure of these wild, fish-eating birds to TCDD and related chemicals for spatial andtemporal comparisons.The successful application of TEFs and TEQs to predict the EROD induction responsein the double-crested cormorant indicates the potential usefulness of the TEQ method in thescreening of environmental exposure of these birds to TCDD and other chemicals which exerttheir toxicity by the same mechanism.5.B. DOSE-RESPONSE STUDY/AHRECEPTOR BINDiNG STUDYDetailed dose-response curves were constructed for avian hepatic microsomal ERODinduction by in ovo exposure to 3H-TCDD during the latter third part of incubation indomestic chickens, domestic pigeons, great blue herons and double-crested cormorants. Thisallows for the comparison of the sensitivity in four bird species of a response under directregulation by the Ah receptor. This dose-response study demonstrated a between 1 and 2orders of magnitude lower sensitivity of the pigeon, great blue heron and double-crestedcormorant in comparison with the domestic chick. Furthermore, the difference in sensitivityto hepatic EROD induction could be explained, at least partly, by about an order of magnitudelower affinity of TCDD for the Ah receptor in the pigeon, heron and cormorant than in thechick. The dose-response findings and the Ah receptor binding results were consistent withthe observed lower sensitivity of great blue heron and double-crested cormorant embryoscompared with domestic chicken embryos.Comparison of the laboratory dose-response curves with environmental exposures toTCDD and related chemicals indicates that environmental induction of hepatic EROD activityis at the lower end of the linear part of the dose-response curves in both the great blue heronand double-crested cormorant. 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