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Environmental contaminants in bald eagles on the coast of British Columbia: exposure and biological effects Elliott, John E. 1995

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ENVIRONMENTAL CONTAMINANTS IN BALD EAGLES ON THE COAST OF BRITISH COLUMBIA: EXPOSURE AND BIOLOGICAL EFFECTS by  JOHN EDWARD ELLIOTT B.Sc., Carleton University, Ottawa, 1979 M.Sc., The University of Ottawa, 1989  A THESIS IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE FACULTY OF GRADUATE STUDIES THE FACULTY OF AGRICULTURE Department of Animal Science  We accept this thesis as/conforming to he required standard  The University of British Columbia October, 1995 c John Edward Elliott  ,  1995  In presenting this thesis in partial fulfilment of the requirements for an advanced degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department  or  by  his  or  her  representatives,  It  is  understood  that  copying  or  publication of this thesis for financial gain shall not be allowed without my written permission.  (Signature)  Department  of  The University of British Columbia Vancouver, Canada  Date  DE-6 (2188)  Abstract  Attracted by abundant food and nesting sites, a large (about 4,000 pairs) Bald Eagle (Haliaeetus leucocephalus) population breeds and winters around the Strait of Georgia on the Pacific coast of Canada.  Eagle habitat has been extensively modified by logging and  waterfront development, while industrial effluents have contaminated food chains. Until recently, most pulp mills on the British Columbia coast used elemental chlorine bleaching and did not secondarily treat effluents, thus releasing chlorine containing chemicals, particularly polychiorinated dibenzo-p-dioxins (PCDDs) and polychiorinated dibenzofurans into the local environment. As top predators, Bald Eagles are exposed to elevated levels of PCDDs, PCDFs and the chemically related polychiorinated biphenyls (PCB5) and organochlorine pesticides. This thesis addressed spatial and temporal trends in chlorinated hydrocarbon exposure of Bald Eagles and toxicological consequences at treatment populations near pulp mills in the Strait of Georgia and in industrial areas of the Fraser River delta, and at reference areas on west coast Vancouver Island, Johnstone Strait and the Queen Charlotte Islands. Initial research during 1990-199 1 focused on eagles found dead or dying and determined that the majority of birds tested had low liver organochlorine levels (< 5 mg/kg, N =59). A small proportion (< 5 %) had levels of DDE, polychiorinated biphenyls (PCBs) and chiordane related compounds potentially diagnostic of acute poisoning. A larger proportion had PCDD/PCDF levels of possible concern; four of 19 eagles tested had TEQ5WHO > 1000 rig/kg, all of which were adults in poor body condition found near pulp mills during the breeding season. In 1992, in ovo exposure to a gradient of environmental contaminants was studied by collecting eggs (N = 25) for laboratory incubation. Hatching success was not significantly different between eggs from pulp mill versus reference sites. A hepatic cytochrome P450 1A (CYP1A) cross-reactive protein was induced sixfold in chicks from near a pulp mill at Powell River compared to those from a reference site (p < 0.05); hepatic EROD and BROD activities  11  were also significantly higher in chicks from pulp mill nests compared to reference sites (p <0.0005 and p < 0.02, respectively). Residual yolk sacs from near pulp mill sites had greater concentrations of 2,3,7,8-substituted PCDDs and PCDFs than reference areas. The hepatic CYP1A cross-reactive protein and EROD and BROD activities were positively correlated with concentrations of 2,3,7, 8-TCDD, 2,3,7, 8-TCDF and toxic equivalents (TEQs) in yolk sacs. No concentration-related effects on histological or morphological parameters were found. Using hepatic CYP1A expression as a biomarker, a no-observed-effect-level (NOEL) of 100 ng/kg and a lowest-observed-effect-level (LOEL) of 210 ng/kg TEQ5WHO on a whole egg (wet weight basis) were suggested for Bald Eagle chicks. To investigate spatial patterns, trends and sources of contaminants to Bald Eagles, eggs were also collected during incubation, 1990-92, at the treatment and reference areas and analyzed for chlorinated hydrocarbons. Data on Bald Eagle avian and fish prey items from the study area were compiled and used as input to a bioaccumulation model. The model accurately predicted 2,3,7, 8-TCDD levels in eagle eggs based on dietary concentrations, but was less accurate for other PCDDs and PCDFs. Using the LOEL levels in eagle eggs derived from the above study, concentrations of 2,3,7,8-TCDD in prey fish of 0.5 ng/kg and in fish-eating birds of 10 ng/kg are suggested as ecosystem guidelines to avoid TCDD-like toxicity in Bald Eagles. At all of the treatment and reference areas, Bald Eagle breeding success was measured for three years and blood samples of nestling eagles were collected for contaminant analysis. Average 3-year eagle productivity was high at most Strait of Georgia study sites, but was significantly lower at reference sites. Using nestling plasma lipid content as a marker of body condition, food supply appeared to be the main factor limiting eagle productivity on the British Columbia coast. However, at a sample of eagle nests adjacent to the dioxin fishery-closure zone near the pulp mill at Crofton, low productivity was probably not caused by low food availability. The cause of the low reproductive rate at Crofton has not been determined; however, a toxicological explanation has not been ruled out.  111  Key Words: Bald Eagle, bioaccumulation, CYP 1A, mortality, reproductive rate, 2,3,7,8tetrachlorodibenzo—p-dioxin  iv  Table of Contents Page Abstract  11  Table of Contents List of Tables List of Figures  vii • ix  .  List of Appendices  xii  Abbreviations  xlii  Acknowledgements  xiv  General Introduction Hypotheses and Objectives  17  Overview of the Thesis  • 18  Chapter 1 Chlorinated hydrocarbon liver levels and autopsy data for Bald Eagles found dead or debilitated, 1989-93 Materials and Methods Results Discussion Chapter 2  Biological effects of chlorinated hydrocarbons in Bald Eagle chicks  Materials and Methods Results Discussion Chapter 3 Bioaccumulation of chlorinated hydrocarbons and mercury in eggs and prey of Bald Eagles  •  .  •  .  •  .  •  .  •  .  •  .  19 19 23 32 41 41 48 60  70 70 77 89  Materials and Methods Results Discussion  V  Page Chapter 4 Influence of contaminants and food supply on Bald Eagle productivity Materials and Methods Results Discussion  102 103 110 120  General Summary and Conclusions  131  References  136  vi  List of Tables Page Table 1.1  Organochlorine residue levels, geometric mean ± 95% confidence intervals, in livers of Bald Eagles found dead on the coast of British Columbia, 1988 1993  28  Non-ortho and mono-ortho PCBs in Bald Eagle livers collected from British Columbia (ng/kg, wet weight)  29  -  Table 1.2  Table 1.3  Concentrations of select PCDDs and PCDFs in Bald Eagle livers collected from the south coast of British Columbia (ng/kg, wet weight) 30  Table 1.4  Comparison of TEQs calculated from select PCDDs, PCDFs, non-ortho and mono-ortho PCBs levels in Bald Eagle livers collected from the south coast of British Columbia (ng/kg, wet weight)  31  PCDD and PCDF concentrations (nglkg, lipid weight basis) in yolk sacs of Bald Eagle chicks collected in 1992 from British Columbia  51  Concentrations of non-ortho PCB congeners in yolk sacs of Bald Eagle embroys collected in 1992 from British Columbia  52  Organochlorine pesticide concentrations (geometric means 95% confidence intervals, range in brackets) in yolk sacs of Bald Eagle chicks collected in 1992 from British Columbia  53  Outcome of artificial incubation of Bald Eagle eggs collected from British Columbia, 1992  54  Histological examination of immune system tissues in Bald Eagle chicks (mean ± SD)  55  Measurement of hepatic cytochrome P450 and porphyrin parameters and vitamin A in plasma and liver of Bald Eagle chicks collected in 1992 from British Columbia (mean ± SD)  56  Table 2.1  Table 2.2  Table 2.3  Table 2.4  Table 2.5  Table 2.6  Table 2.7  Concentration-effect relationships between biochemical and morphological measurements with chlorinated hydrocarbon yolk sac levels in Bald Eagle chicks 59  Table 2.8  Comparison of regression (r ) values of some hepatic biochemical parameters 2 on TEQs derived from three sets of toxic equivalence factors (TEF5)  60  Mean PCDD/PCDF lvels (ng/kg, wet weight) in fish collected near three pulp mills on the Strait of Georgia, British Columbia  75  Table 3.1  vi’  Page Table 3.2  PCDD and PCDF levels (ng/kg, wet weight) in waterbird and seabird species from the British Columbia coast  76  (PCDD) and (PCDF) residue levels (wet weight basis) in Bald Eagle eggs from British Columbia, 1990 1992  78  Table 3.4 Organochiorine and PCB residue levels (mg/kg, wet weight) in Bald Eagle eggs from the British Columbia coast, 1990-1992, expressed as geometric means and 95% confidence intervals (range in brackets)  80  Table 3.5 Mercury residue levels (mg/kg, wet weight) in Bald Eagle eggs from locations on the British Columbia coast, 1990-1992, expressed as geometric means and 95 % confidence intervals (range in brackets)  81  Table 3.6 Non-ortho PCBs in Bald Eagle eggs (ng/kg, wet weight) collected from British Columbia, 1992  84  Table 3.3  -  Table 3.7 Eggshell thickness data, mean ± SD, (range in brackets) for Bald Eagles collected from British Columbia, 1990-1992 85 Table 3.8  A simulation of PCDD/PCDF levels in Bald Eagle eggs at Crofton, 1990, based on concentrations in the diet  86  Table 3.9  Characterization of British Columbia pulp mills discussed in this paper  Table 4.1  Correlation Matrix (r values) for percent plasma lipid and selected hydrocarbon in Bald Eagle nestlings from British Columbia, 1993-94 109  .  .  Table 4.2 Nest success and production of young for Bald Eagles at nine study areas on the British Columbia coast (1992-94) Table 4.3  Table 4.4  PCDD/PCDF levels, geometric means and 95% confidence interval (ng/kg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94  .  .  101  111  116  Organochlorine pesticide and PCB levels, geometric means and 95% confidence interval (tg/kg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94 118  Table 4.5 Non-ortho PCB levels, geometric mean and 95% confidence interval (ng/kg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94  vi”  120  List of Figures Page Figure 1.  Molecular structure and position numbering of polychiorinated dibenzo-p dioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs)  2  Figure 2.  Molecular structure of the major organochiorine pesticides  3  Figure 3.  Molecular mechanism proposed for TCDD and related chemicals  7  Figure 1.1 Locations of Bald Eagles collected from British Columbia, 1989-93, and analyzed for chlorinated hydrocarbons (N = 59)  20  Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed compared to the complete set of birds received  24  Figure 1.3 Numbers of eagles showing different DDE and PCBs in livers (N =59)  24  Figure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month  25  Figure 1.5 DDE and PCB residue levels in relation to body condition  26  Figure 1.6 PCB congeners in Bald Eagle livers expressed as percent of total PCBs compared for birds in good and poor body condition (N =9, for each group)  .  .  36  Figure 2.1 Locations where Bald Eagle eggs were collected for artificial incubation  42  Figure 2.2 Residue levels of major PCDDs and PCDFs in yolk sacs of Bald Eagles collected from British Columbia in 1992. Vertical bars represent geometric means of two to five analyses per collection site along with the 95 % confidence interval. Means which do no share the same lower case letter were significantly different (p < 0.05)  49  Figure 2.3 PCB congeners in yolk sacs of Bald Eagle chicks from British Columbia, 1992, expressed as percent of total PCBs. Values represent means of two to eight analyses per collection site. Isomers are identified according to their IUPAC number  50  Figure 2.4 Exposure-response relationships between 2378-TCDD or log 2378-TCDF concentrations in yolk sacs of Bald Eagles and hepatic (A) EROD activity (B) CYP1A concentrations and (C) BROD activity  58  lx  Page Figure 2.5 The contribution of various chlorinated hydrocarbon groups to the sum of TCDD toxic equivalents (TEQ) in Bald Eagle yolksacs from coastal British Columbia, 1992 (N values and variances are in the tables), compared to values for common terns from the Netherlands. Toxic equivalents factors for PCDDs/PCDFs from Safe (1990) and for PCBs from Ahlborg et a!. (1994)  .  .  66  Figure 3.1 Locations where Bald Eagle eggs were collected for analysis  71  Figure 3.2 PCB congeners in Bald Eagle from British Columbia, 1990-1992, expressed as percent of total PCBs. Values represent means of three to eight analyses per collection site. Congeners are identified according to their IUPAC number  82  Figure 3.3 Plot of the first and second principle components (PCi and PC2). PCB congener concentrations for all individual egg analyses were expressed as percent total PCBs and arcsine transformed. Principle components analysis was then undertaken using a group of 6 congeners (66, 99, 118, 170, 180, 194) considered to be markers of Aroclor sources. 75 % of the matrix variance was explained by PCi and 15 % by PC2  83  Figure 3.4 The contribution of various chlorinated hydrocarbon groups to the sum of TCDD toxic equivalents (TEQs) in Bald Eagle eggs from coastal British Columbia, 1990-1992 (N values and variances are in the tables). Toxic equivalents for PCDDs/PCDFs from Safe (1990) and for PCBs from Ahlborg et a!. (1994)  84  Figure 3.5 Concentration of 2,3,7,8-TCDD predicted in Bald Eagle eggs based on the percent of fish-eating birds in the diet. Prediction is based on a bioaccumulation model described in the text and the simulation is based on data from Crofton, British Columbia, 1987-1992  88  Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections. At Langara Island, the survey circumsribed the coastline of the island  104  Figure 4.2 Bald Eagle nest sites and dioxin fishery closure areas at Powell River, Nanaimo and Crofton  106  Figure 4.3 Bald Eagle productivity compared between samples of nest located adjacent to shorelines inside and outside of dioxin fishery closure areas on the British Columbia coast  112  x  Page Figure 4.4 Productivity at Bald Eagle nest sites on the British Columbia coast as a function of contaminant concentrations in plasma samples from nestlings raised in that territory, for: A) the log of TEQsWHO, B) the log of DDE. The subpopulations included: East Vancouver Island, Powell River, Barkley Sound, Clayoquot Sound, Johnstone Strait, Fraser Delta, lower Fraser Valley and Langara Island 113 Figure 4.5 Comparison of mean productivity of Bald Eagles at sites on the coast of British Columbia with the mean percent lipid in plasma samples of nestling eagles at each site  114  Figure 4.6 Residue levels of selected PCDDs and PCDFs in plasma samples of bald eagle nestlings collected on the British Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.3. Means that do not share the same lower case letter are significantly different (p <0.05)  117  Figure 4.7 Residue levels of selected PCBs in plasma samples of Bald Eagle nestlings collected on the British Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.4. Means that do not share the same lower case letter are significantly different (p < 0.05)  119  Figure 4.8 Trends in 2,3,7,8-TCDD in eggs of eagles, herons and cormorants at Crofton, British Columbia. The likely trend in eagles is extrapolated back to 1987, based on the mean 2,3,7,8-TCDD ratio of eagles:herons, 1990-1992  120  xl  List of Appendices Page Appendix 1-1 Organochiorine pesticide and PCB levels in Bald Eagle livers collected from British Columbia (mg/kg wet wt.)  38  Appendix 2-1 Selected morphological measurements in Bald Eagle chicks collected in 1992 from British Columbia  69  Appendix 4-1 Productivity, % lipid and selected chlorinated hydrocarbon residue levels in plasma of individual Bald Eagle chicks collected from the coast of British Columbia, 1993-94  xii  128  Abbreviations Ah  aryl hydrocarbon  NWRC  National Wildlife Research Centre  AHH  aryl hydrocarbon hydroxylase  OC  Organochiorine pesticide  ANCOVA  analysis of covariance  PCA  principle component analysis  ANOVA  analysis of variance  PCB  polychiorinated biphenyl  BMF  biomagnification factor  PCDD  polychiorinated dibenzo-p dioxin  BROD  benzyloxyresorufin 0dealkylase  PCDF  polychiorinated dibenzofuran  CWS  Canadian Wildlife Service  PWRC  Pacific Wildlife Research Centre  CYP1A  cytochrome P450 1A SAS  CYP2B  cytochrome P450 2B  Trademark, SAS Institute Inc.  DDE  1, 1-dichioro ethylene bis (pchiorophenyl)  SYSTAT  Trademark, Systat Inc.  TCDD  tetrachioro dibenzo-p-dioxin  DDT  1,1, 1-trichloro-2,2-bis(pchlorophenyl)ethane  TCDF  tetrachioro dibenzofuran  EROD  ethoxyresorufin 0-deethylase  TEF  toxic equivalent factor  GLEMEDS  Great Lakes embryo mortality edema and deformities syndrome  TEQ  TCDD toxic equivalent  WHO  World Health Organization  HCB  hexachlorocyclobenzene  HCH  hexachiorocyclohexane  LOEL  lowest-observed-effect-level  NOEL  no-observed-effect-level  xl”  Acknowledgements I would like to thank my supervisory committee, Kim Cheng, Gail Beliward, Ross Norstrom and Tom Sullivan for overall guidance and support. I would like to acknowledge the financial and personal support of the Canadian Wildlife Service, and to personally thank Steve Wetmore at the Pacific Wildlife Research Centre and Keith Marshall at the National Wildlife Research Centre for their advice and support over the years. A project of this sort depends on the assistance of a great many people. Specific contributions are acknowledged at the end of each chapter, however, the support of a number of people deserves special consideration: Ian Moul was a valuable co-worker in virtually all phases of the field work; George Compton contributed his considerable tree climbing and bush-whacking skills. Mary Simon, Henry Won and Suzanne Trudeau are thanked for their work on the chemistry and biochemistry, Ken Langelier was a great help in the wildlife health aspects and suggested the initial work on Bald Eagles. I am very grateful to Laurie Wilson for the many technical and scientific roles she undertook for me. Shelagh Bucknell and Pam Whitehead are thanked for their assistance and patience in typing of tables and preparation of figures, respectively. I would also like to acknowledge my friends and coworkers both at UBC and CWS for making this PhD experience more rewarding and enjoyable. I also wish to thank my parents for imparting a sense of what is important in life. Most importantly, I am most grateful to the patience and support of my wife Christine and my children, Kyle, Siobhan, Frazer and Alicia.  xiv  Introduction  Pollution of the environment by toxic substances has become a global problem with ecological, economic and political consequences. Chlorinated hydrocarbons such as polychiorinated dibenzo-p-dioxins (PCDDs), polychiorinated dibenzofurans (PCDFs), polychiorinated biphenyls (PCBs) and DDT (1,1,1 -trichloro-2,2-bis[p-chlorophenyl]ethane) have attracted a great deal of attention from both the scientific community and the general public. Among the best known and most dramatic effects has been the impact of DDT and other organochiorine pesticides on reproduction and survival of birds of prey, such as eagles, Ospreys (Pandion haliaetus) and falcons. These birds, particularly the Bald Eagle (Haliaeetus leucocephalus) and the Peregrine Falcon (Falco peregrinus), have become symbols of environmental awareness and reminders of ecological consequences of short-sighted use of chemical technology. Although most Bald Eagle populations have recovered from the effects of DDT, reproduction and survival in some areas are impaired by chemicals, such as PCBs, which can function toxicologically like TCDD. A great deal of laboratory research has been conducted on PCDDs and related compounds; however, little is known of exposure and effects on wildlife. This thesis focused on the Bald Eagle population resident around British Columbia’s Strait of Georgia and on exposure to and the consequences of the widespread pollution of that area by PCDDs and PCDFs from forest industry sources.  Chlorinated hydrocarbons Structures Chlorinated hydrocarbons are organic compounds with chlorine substituents. This thesis is concerned primarily with the polychiorinated aromatics, those with chiorines substituted on aromatic ring structures, and to a lesser extent with some non-aromatic organochiorine pesticides, such as hexachiorocyclohexane (HCH). Ecotoxicologically, the most important  1  polychiorinated aromatics are the PCDDs, PCDFs, PCBs, and some of the organochiorine pesticides such as DDT. The structures of the PCDDs, PCDFs and PCBs are represented in Figure 1. The PCDDs and PCDFs obtain a mainly rigid, planar configuration, which determines their biological behaviour. For the PCBs, the molecular conformation depends on the chlorine substituents. Those congeners without ortho-chiorines energetically obtain a mainly planar conformation, those with di-ortho chlorine substituents are non-planar and those with mono  ortho substituents are intermediary. Thus the non-ortho PCBs are approximate stereoisomers of PCDDs and PCDFs and if chlorinated laterally, exhibit similar biological behaviour (Safe 1984).  1  9 8  4  6  2,3,7,8  dibenzo-p-dioxin  -  Tetrachlorodibenzo-p-dioxin  S  -Cl  7  c1  dibenzofuran 3  2  2,3,7,8 2’  -  Tetrachlorodibenzofuran  3’  4  PCB 126  biphenyl  33’44’5  -  Penta  Figure 1. Molecular structure and position numbering of polychiorinated dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs).  2  Organochiorine pesticides fall into three structural groups (Figure 2). DDT is similar in structure to the PCBs, in that it has two chlorine-substituted benzene rings, in this case joined on an ethane backbone. Dieldrin, mirex and the chiordane-related compounds, including heptachlor epoxide, all belong to the cyclodiene group. The third group are the chlorinated benzenes and cyclohexanes.  DIC I-I LO RO DI PHE N YLET HAN ES  cI_OH_O_  ci  DDT, DDD Dicofol Perthane Methoxychior Methiochior  ci CYCLODIENES  Aidrin, Dieldrin Heptachlor Chiorciane Endosulfan  CHLORINATED BENZENES CYCLOHEXANES  HCB, HCH Lindane (a-BHC) 6 (Cl) Cl  Figure 2. Molecular structure of the major organochiorine pesticides. Sources PCDDs g PCDFs. Neither PCDDs nor PCDFs are deliberately produced commercially, but are formed either as by-products during the synthesis of other chemicals, such as chiorophenolic biocides, or during combustion of chlorine containing wastes. Incineration of municipal and industrial wastes is the major global source of dioxins, which can be transported long distances and subsequently deposited in soils and lake sediments (Czuczwa et al. 1984). Although combustion produces a fairly uniform mixture of PCDD and PCDF isomers, physical and chemical atmospheric processes favour the deposition and accumulation of less toxic higher chlorinated compounds, which then dominate in sediments (Hites 1990). Elevated contamination by more toxic and persistent isomers such as 2,3,7,8-TCDD was previously associated with use, production or waste storage of chiorophenoxy acid herbicides, particularly 2,4,5-T (see:  3  Baughman and Meselson 1973; Fanelli et at. 1980; Powell 1984). However, relatively recent studies showed that effluents from kraft pulp mills using elemental chlorine bleaching contained 2,3,7,8-TCDD and 2,3,7,8-TCDF (Kuehl et at. 1987), which caused contamination of fish and wildlife in receiving waters (Rogers et at. 1989; Elliott et at. 1989a). Elevated HxCDDs (hexachioro dibenzo-p-dioxins) in effluents and foodchains can result from pulp mill digestion of tetrachiorophenol-contaminated woodchips (Elliott et al. 1989a; Luthe et at. 1990). Use and production of 2,4,5-T and most chlorophenols has been regulated in North America. Pulp mills in Canada, but not necessarily in the USA or elsewhere, now use alternative bleaching methods, which have substantially reduced formation of TCDD and TCDF. PCBs. PCBs were used for a variety of purposes which can be divided into ‘closed circuit’ uses such as in electrical transformers and capacitors and in heat transfer and hydraulic systems, and into ‘open circuit uses’ such as the formulation of lubricating and cutting oils, pesticides, plastics, paints, inks, adhesives, etc. More than one billion (l0) kg PCBs were produced worldwide (Tanabe 1988). Until 1977, over 90 % of the production was in the U.S.A., after which it switched to Europe and Japan. Some 40 million kg PCBs have been imported into Canada; the most recent inventory accounted for about 24 million and assumed that the remaining 16 million kg had been lost to the Canadian environment (Environment Canada 1985). Open circuit uses of PCBs were voluntarily restricted by industry in 1973 and all uses of PCBs have been regulated by governments in North America since 1977. Organochlorine pesticides. OC pesticides are synthetic compounds widely used to control agricultural and forest pests and the transmission of vector-borne diseases. The most abundant OC pesticide in the environment is DDE, the major persistent metabolite of DDT. Other compounds commonly detected in wildlife include DDD, DDT, dieldrin, heptachlor epoxide, mirex, photomirex, toxaphene, oxychiordane, cis- and trans-chlordane, cis- and trans-nonachlor, endrin, HCB, and HCH isomers. DDT, a broad-spectrum insecticide, was first used in North America in the 1940s in public health campaigns to control lice (Carson 1962). From the 1940s until the early 1970s, large quantities of DDT were sprayed to control forest insect pests in  4  British Columbia (Nigam 1975) and in the northwest USA (Henny 1977). Major restrictions on the use of most organochlorine pesticides (ie. DDT, dieldrin, endrin, heptachior, HCH and toxaphene) in Canada and the USA were first implemented in the early 1970s, with further controls imposed throughout the 1970s and 1980s (Noble and Elliott 1986). Heptachior continued to be used in Oregon until 1974 (Henny et at. 1983) and significant amounts of chiordane, lindane, dicofol and toxaphene were used until the early 1980s in California (Ohlendorf and Miller 1984). A few minor applications of chlordane, lindane, dieldrin and heptachlor (eg. seed treatment, termite control) are still permitted in Canada and the USA. In Mexico, some restrictions on the use of DDT, BHC, dieldrin and heptachior were imposed in 1980 (Burton and Phiogene 1986). Organochlorines can be transported over vast distances by atmospheric and oceanic vectors; as such, ongoing use in Asia may now be the main source of OCs to the Canadian environment, particularly the Pacific coast (Elliott et at. 1 989b). Information on OC use in Asian countries bordering the north Pacific is scarce. Since the 1950’s, DDT and HCH have been used extensively on rice, cotton and vegetable crops, but in the 1970s, many countries began to replace them with organophosphorus compounds. As in North America, agricultural uses of OCs are subject to regulation in most north Pacific countries, but the degree of compliance varies. The People’s Republic of China manufactures OC pesticides; however, the production and use of two, DDT and HCH, were banned there in 1983 (Wolfe et al. 1984). In Japan, production and use of DDT and HCH were prohibited in 1971, but the use of chlordane for termite control was permitted until the late 1 980s (Tanabe et al. 1989). Korea also prohibited the use of DDT in the early 1970s (Phillips and Tanabe 1989). However, in Hong Kong (where many pesticides are still formulated), there appears to be continued input of DDT into coastal waters, despite restrictions imposed in 1988 (Phillips and Tanabe 1989). Food chain bioaccumulation For a substance to bioaccumulate, the following physico-chemical traits are necessary: 1) lipid solubility evident by a high octanol/water partition coefficient; 2) resistence to metabolic attack.  5  PCDDs  PCDFs. Food chain bioaccumulation of PCDDs and PCDFs generally  requires a 2,3,7,8-substitution pattern, as congeners lacking that substitution pattern are metabolized in birds, mammals and fish (Van den Berg et al. 1993a). Accumulation of non2,3,7,8-substituted PCDDs has been reported in some invertebrate species, particularly crustaceans (Norstrom and Simon 1991). PCBs. Among homeotherms, tissue retention of PCB congeners varies with development of the cytochrome P450 system and capacity to metabolize different compounds. In general, mono and non-ortho PCBs are metabolized by CYP1A enzymes, while di-ortho congeners are degraded by CYP2B enzymes (Boon et al. 1987; Brown 1994). Organochiorines. The relative capacity of organochlorines to bioaccumulate has been extensively studied in the Herring Gull (Larus argentatus) by Norstrom and co-workers (Norstrom et at. 1986; Clark et at. 1987; Braune and Norstrom 1989). The more slowly degraded and therefore more accumulative OCs in birds are DDE, mirex and oxychlordane, with heptachlor expoxide, dieldrin and HCH compounds being more rapidly cleared. Effects of chlorinated hydrocarbons PCDDs, PCDFs  PCBs. This group of compounds causes similar toxic symptoms in  most species studied (Safe 1990). Dose-related responses include: irnmunotoxicity, liver enlargement and other signs of hepatotoxicity such as porphyria, induction of drug-metabolizing enzymes, reproductive toxicity and cancer promotion (Safe 1984). The toxicity of the individual compounds varies greatly with the molecular structure. The most toxic compound is 2,3,7,8TCDD, which is often used as a model for studying the effects of these chemicals. The more toxic furan and biphenyl congeners all exhibit a structural similarity to 2,3,7, 8-TCDD. Many of the toxic effects caused by this class of compounds are believed to be mediated by a cytosolic receptor found in many tissues, known as the aryl hydrocarbon (Ah) receptor (Landers and Bunce 1991). The Ah-receptor mediated mode of action is represented schematically in Figure 3. Traditional toxicology studies have focused on single chemicals in test organisms. However, environmental exposure to chlorinated hydrocarbons involves a multitude of  6  compounds. To provide a practical method of dealing with this, the study of Ah-receptor mediated structure-activity relationships has produced an additive scheme for estimating the toxicity of complex mixtures of these chemicals through use of “TCDD Toxic Equivalence Factors” (TEFs). Each individual compound is assigned a TEF relative to 2,3,7,8-TCDD, essentially a ratio of its relative toxicity based on one or more endpoints. Analytically determined concentrations are multiplied by the TEF, the results summed to produce the “TCDD Toxic Equivalents” or TEQs. TEFs published by Safe (1990) are widely used; however, those reported by Ahlborg et al (1994), which attribute lower relative toxicity to the mono-ortho PCBs, appear more relevant for most birds (Brunstrom and Andersson 1988; Bosveld et al. 1992; Kennedy et al. 1994).  Xenobiotic ligand (TCDD, etc)  INCREASED METABOLISM OF DRUGS AND ENVIRONMENTAL CHEMICALS  TOXICITY  Figure 3. Molecular mechanism proposed for TCDD and related chemicals.  The lipophiic  xenobiotic ligand, such as TCDD, enters the cell by passive diffusion through the lipo-protein cell membrane and binds with the Ah-receptor (AbR); the AhR releases a heat shock protein (HSP 90) as it binds with the ligand. The ligand-receptor complex then associates with the nuclear translocating protein (Arnt) and moves into the nucleus, where it interacts with dioxin responsive elements (AhREs) on the genome, which alters the transcription of specific niRNAs. The resulting proteins then mediate the biochemical and toxic responses observed with TCDD exposure (after Okey et al. 1994).  7  Although the toxicology of dioxins and related compounds continues to be extensively studied in laboratory mammals, there are less data on avian species. Bird studies have focused on embryos, as the most sensitive life stage (Peterson et at. 1993). Chicken embryos are particularly sensitive: the LD 50 for 2,3,7, 8-TCDD, administered into the air sac of the chicken embryo, was reported as 250 ng/kg (ppt) egg (Alired and Strange, 1977). An LD 50 for 2,3,7,8TCDD in chicken embryos of about 200 ng/kg was determined more recently by both Henschel  et at. (in preparation) and Janz (1995) using air cell and yolk sac injection. They also reported a very steep dose response curve, with no mortality at 100 ng/kg and complete mortality at 300 ng/lcg. Injection of 2,3,7,8-TCDD or similar compounds into developing chickens causes a toxicity syndrome which includes, in addition to mortality, beak and other deformities, thymic and bursa inhibition, edema and liver lesions (Brunstrom and Andersson 1988; Brunstrom 1990). The heart is a sensitive target organ as only 9 ng/kg caused an increase in the incidence of cardiovascular malformations (Cheung et at., 1981). In domestic turkey embryos, non-ortho PCB congeners that bind the Ah receptor and thus act by a similar toxic mechanism to 2,3,7,8TCDD, also cause gross deformities and mortality, but not the other symptoms seen in chicken embryos (Brunstrom and Lund 1988). In embryos of other avian species, such as Ring-necked Pheasants (Phasianus coichicus) and Eastern Bluebirds (Siatia sialis) injected with 2,3,7,8TCDD, sublethal effects observed in chickens were not observed, rather mortality was the most sensitive endpoint (Nosek et at. 1992; Martin et at. 1989). The LD5O for 2,3,7,8-TCDD was 1100 ng/kg egg in the pheasant embryo and between 1000 and 10,000 ng/kg egg in the Eastern Bluebird embryo, in both cases via albumin injection (Nosek et at. 1992; Martin et at. 1989). Brunstrom & Reutergardh (1986), using mortality as an endpoint, reported marked interspecific sensitivity among birds to the TCDD-isostereomer, PCB congener 77 (34-34). Chickens were the most sensitive, followed by turkeys (30 X less sensitive) and pheasants (100 X less sensitive) and then by Mallards, Goldeneyes, domestic ducks, geese, Herring Gulls and Black-headed Gulls (>1000 X less sensitive).  8  Adults of avian species were much less sensitive to TCDD than embryos; 25 to 50 gIkg (ppb) body weight caused mortality in chickens (Greig et al. 1973), while 25 pg/kg caused 75% mortality in ring-necked pheasant hens (Nosek et al. 1993). In other studies with adult birds, acute oral toxicity of 2,3,7,8-TCDD ranged from 15 pg/kg body weight in Northern Bobwhite (Colinus virginianus) to greater than 810 pg/kg body weight in the Ringed Turtle Dove (Streptopelia risoria), (Hudson et al. 1984). There are few published studies of the chronic effects of dioxin-like compounds in birds. Kenega and Norris (1983) reported that a diet containing 0.3 or 3 ng/kg TCDD in a formulation of 2,4,5-T fed to bobwhites for 18 weeks produced no effects on egg production or survival of embryos. However, 50 % mortality did occur within 5 days at a dietary level of 167 ng/kg. Nosek et al. (1992) showed that Ring-necked Pheasants dosed with 1.0 ug/kg/week of 2,3,7,8TCDD for 10 weeks exhibited mortality and signs of wasting syndrome; egg production was also reduced and hatchabiity of eggs was < 2 %. Pheasants dosed with 0.1 pg/kg/week for 10 weeks exhibited no adverse effects. Daily feeding of PCB congeners 126 (34-345) and 105 (23424) for up to eight weeks caused hepatic porphyria, thymic atrophy (PCB 126 only) and marked microsomal cytochrome P450 enzyme induction in Japanese Quail (Coturnix coturnix), but no porphyria, and only minor P450 induction in American Kestrels (Falco sparvarius) (Elliott et al. 1990; 1991). This is the only available laboratory study involving TCDD-like compounds in a bird of prey. Field studies of PCDDs. PCDFs and PCBs in birds. In the Great Lakes, a toxic syndrome observed in a number of fish-eating bird species, is referred to as GLEMEDS (Great Lakes embryo mortality, edema and deformities syndrome), and has been attributed to exposure to PCBs, PCDDs and PCDFs (Gilbertson et al. 1991). The syndrome was first recognized in Lake Ontario gull and tern populations in the early 1970s (Gilbertson and Fox, 1977). Subsequent retrospective analysis of archived Herring Gull eggs revealed the presence of high 2,3,7,8-TCDD concentrations in eggs of Lake Ontario gulls collected in the early and mid 1970s, which likely contributed to poor reproduction (Gilbertson et al. 1991). However those eggs also  9  contained high levels of other known embryotoxins, including PCBs and HCB (Mineau et at. 1984; Bishop et at. 1992). A number of recent studies in the Great Lakes: (Kubiak et at. 1989; Tilett et at. 1992; Yamashita et at. 1993; Rattner et at. 1994) related exposure to PCBs, particularly the non-ortho 126 (345-34) and the mono-orthos 105 (234-34) and 118 (245-34) to biological effects in colonial waterbird populations. Recently, Bosveld et at. (1994) and Van den Berg et at. (1994) reported high PCB levels in eggs of fish-eating birds breeding in the Rhine estuary, which correlated with various endpoints of exposure and toxicity, including CYP1A induction and embryonic growth. In British Columbia, Great Blue Herons (Ardea herodias) and Double-crested Cormorants (Phalacrocorax auritus) breeding near pulp mills have been used as sentinel species to study toxicant exposure and effects (Elliott et at. 1989; Whitehead et at. 1 992a). Failure of a Great Blue Heron colony in 1987 at Crofton, British Columbia coincided with a three-fold increase in mean egg levels of 2,3,7,8-TCDD over the previous year when reproduction was normal; however, no statistically significant relationship between contaminant levels and reproductive outcome among individual birds was determined (Elliott et at. 1989a). Heron embryos, collected in 1988 at colonies with high, intermediate and low levels of PCDD and PCDF contamination and incubated in the laboratory, did not exhibit any significant differences in hatching success among the three sites. There were, however, a number of sublethal effects in heron chicks, which correlated with their 2,3,7,8-TCDD levels, including induction of hepatic EROD (ethoxyresorufin-O-deethylase) activity, edema and lower embryonic weight (Bellward et at. 1990; Hart et at. 1991; Sanderson et at. 1994) and brain abnormalities (Henshel et at. 1995). Disturbance by people and/or Bald Eagles (Norman et al. 1989) was probably the main cause of heron colony failure at Crofton in the late 1980s on the British Columbia coast and would have masked other potential factors (Elliott et at. 1 989a); however, intensive observation of heron nests showed that mean time spent incubating was lower and greater between-nest variability in incubation time occured at a contaminated versus a control heron colony in 1988 (Moul 1990). The strong possibility exists, therefore, of a contamiiiant-related effect on adult incubation  10  behaviour. Chemically mediated aberrant parental behaviour has been reported for a number of species in both laboratory (Peakall and Peakall 1973; McArthur et at. 1983) and field studies (Cooke et at. 1976; Mineau et at. 1984; Kubiak et at. 1989). Eggs of ospreys nesting downstream of bleached-kraft pulp mills on the Thompson and Columbia rivers of the British Columbia interior, contained significantly higher levels of 2,3,7,8TCDD than eggs from nests upstream of the mills (Whitehead et at. 1993). Studies of osprey productivity showed a trend of lower productivity at downstream compared to upstream sites; however, there were a number of confounding factors, particularly relating to food supply. White & Hoffman (1991) recently reported poor reproductive success in Wood Ducks (Aix sponsa) contaminated with TCDD and TCDF from a 2,4,5-T waste disposal site in Arkansas. Mean levels in Wood Duck eggs were 70 to 75 ng/kg for both TCDD and TCDF. Based on the limited chemical data provided, Wood Ducks appear to be more sensitive to the effects of TCDD than other wild bird species. Organochiorine pesticides. The acute vertebrate toxicity of DDT is low, the LD 50 to the Japanese Quail was 595 mg/kg (ppm). The cyclodienes are much more acutely toxic to vertebrates; for example, the LD 50 of endrin to California Quail is 1.1 mg/kg (Hudson et at. 1984). Cyclodiene insecticides have been implicated in many avian mortality incidences, particularly of birds of prey (reviewed in Noble et at. 1993). Liver residues of dieldrin, chlordane and heptachior epoxide associated with mortality are in the order of 3-10 mg/kg (Cooke et at. 1982). The effects of DDE on eggshell thickness and quality is the toxicological endpoint that has been best characterized in wildbirds (Anderson et at. 1975; Blus et at. 1974; Newton and Bogan 1974; Blus et at. 1980; Custer et at. 1983; Elliott et at. 1988). DDE affects calcium metabolism by interfering with carbonic anhydrase metabolism at the shell gland (Cooke 1983). Critical egg levels of DDE vary widely among species and have been established for some raptors (Fyfe et a!. 1988; Peakall et at. 1991; Wiemeyer et a!. 1993), The chronic toxicology of other organochiorines to wild birds has not been established. Dieldrin has been implicated in  11  reproductive effects, not via eggshell thinning, but rather embryotoxicity. Lockie et al. (1969) suggested that dieldrin levels in eggs greater than 1.0 mg/kg were associated with egg failure in Scottish Golden Eagles (Aquila chrysaetos), However, the association of this level may have had more to do with its indication of lethal dieldrin residues in adult birds as suggested by Newton (1986) for European Sparrowhawks. Heptachlor epoxide, at egg levels > 1.5 mg/kg was associated with effects on reproduction of American Kestrels (Henny et al. 1983). Egg levels of HCB >5.0 mg/kg in Herring Gull chicks were associated with embryo mortality (Boersma et al. 1986).  The Bald Eagle Natural history The Bald Eagle is an endemic North American member of the genus haliaeetus, the sea eagles. Bald Eagles are sexually dimorphic, adult females average 5.3 kg and 221 cm, and males 4.3 kg and 207 cm (Stalmaster 1987). Breeding adults are thought to form life-long pair bonds; the average breeding life span is about 20-25 years. Breeding success may vary considerably from year to year depending on factors such as disturbance and food supply (Stalmaster 1987). In the Pacific northwest, Bald Eagles are year-round residents (Hancock 1964). The breeding season can last from February until August, although nests are maintained year round (Herrick 1932). Eagles often have more than one nest in a territory; the function of the alternate  nest is not clear, but may be to reduce parasite loads (Stalmaster 1987). Nests are always located in proximity to water. Nest trees are usually the dominant or codominant tree in the area in order to provide a clear view of the territory and clear flight paths to feeding areas. Female eagles lay from one to three eggs, two being most common. Eggs are incubated for about 35 days, and the chicks are dependent on their parents at the nest for food and protection for another 72 to 96 days (Herrick 1932). Adults appear to intially remain with the chicks on fledging; subsequent juvenile dispersal patterns can be complex (McClelland et al. 1994). Chances of reaching adult age are variable and may be less than 10  12  in some populations, such as in  Alaska, and as high as 50 % in more southern locations. Bald Eagles do not attain adult plumage until their 5th year, when they normally begin breeding (McCollough 1989) Eagles have a number of physical adaptations as predators. They have excellent vision and can reportedly detect other eagles flying at 23 to 65 km distance (Shlaer 1972). They kill using their powerful feet and talons, while food is torn apart by a large beak. They are powerful flyers, particularly adapted for soaring in open country.  Bald Eagles are opportunistic foragers  and predators. In the northwest, birds, particularly gulls and waterfowl, marine and aquatic fish, and invertebrates make up the bulk of the diet for most birds, although mammals can be important in some areas (Vermeer et al. 1989; Knight et at. 1990; Watson et al. 1991). Population trends and critical factors Like many other large predatory animals, Bald Eagle populations declined during the past century over much of their North American breeding range (Stalmaster, 1987). Habitat loss and degradation combined with intentional and accidental killing contributed to poor productivity and loss of breeding stock. In the early 1950s, populations of eagles and other birds of prey began to disappear from many areas. Classic work by Charles Broley (1947, 1958) showed a precipitous decline in productivity of a Florida population from a high of 89 % nest success in 1942 to 14 % in 1952. During the 1960s and 1970s, eagle productivity was subsequently found to be below sustainable levels in many areas of the U.S. and Canada (Stalmaster, 1987). The low breeding success of Bald Eagles and other birds of prey, which began in North America in the early 1950s, coincided with the introduction of DDT and other organochlorine pesticides. The widely accepted paradigm for decline of the Bald Eagle and other North American raptor populations states that DDE persists, bioaccumulates and impairs reproduction via the mechanism of reduced eggshell quality (Grier, 1982; Peakall et at., 1991). Wiemeyer et al. (1984) determined that in Bald Eagles, reproductive failure approached 100 % when DDE egg levels were greater than 15 mg/kg. DDE egg levels of 5 mg/kg were associated with 10 % eggshell thinning, while populations with less than 3 mg/kg exhibited no significant shell thinning and normal production of young. However, those values were based on analyses of adled eggs which may tend to have higher than average residues and may bias the estimate of critical values. In other birds of prey,  13  particularly European populations, loss of breeding stock to acute dieldrin poisoning, has been suggested to be more critical than DDE-induced shell thinning (Newton et a!. 1992). Bald Eagles have also been acutely poisoned by other pesticides, including dieldrin, and heavy metals, such as mercury and lead (Reichel et al. 1984), athough these effects were probably less critical to population decline. At any rate, eagle productivity has improved and populations have increased in most areas, following strict regulation by the early 1970s of organochiorine use in North America (Grier, 1982; Wiemeyer et a!. 1993). As a result in July, 1995, the U.S. Fish and Wildlife Service changed the status of most Bald Eagle populations in the continental U.S.A. from endangered to threatened. However, breeding success remains below maintenance levels at some regional ‘hotspots’. Along the Great Lakes shoreline, productivity is lower and contaminant levels higher than at nearby inland locations (Bowerman 1993), although, at least for Lake Superior populations, reduced food delivery to nestlings was an important factor (Dykstra 1994). Eagle populations in Maine generally exhibit low productivity, which has been related to high contamination by PCBs and DDE (Welch 1994). Along the lower Columbia River, low Bald Eagle breeding success correlated with high egg and plasma levels of DDE and PCBs; moderately high levels of 2,3,7,8TCDD (tetrachloro dibenzo-p-dioxin) were also present in those eggs (Anthony et al. 1993). British Columbia Eagle Populations Based on a 1984 report, no Canadian Bald Eagle populations are listed as threatened (COSEWIC 1995). In British Columbia, most Bald Eagle populations are “blue listed”, based on concern for long term conservation of some populations (British Columbia Conservation Data Centre 1995). Hodges et at. (1984) estimated the resident breeding population of Bald Eagles on the British Columbia coast to be about 9,000 birds. An estimated 30,000 eagles winter on the coast, mainly in the river estuaries surrounding the Strait of Georgia (Farr and Dunbar, 1988). Bald Eagles are lured by the rich food resources and high biological productivity, both terrestial and marine, of the Strait of Georgia, which is essentially a large estuary with nutrient input from numerous rivers, particularly the Fraser (LeBlond 1989); those rivers are also major salmon  14  spawning sites, which attract thousands of eagles each winter. Millions of waterbirds and shorebirds migrate through and winter in the region, which provides the major food supply for Bald Eagles and falcons. The basin is surrounded by temperate rain forests which have been extensively exploited for wood fibre. The impact on Bald Eagles of habitat modification, especially the clearing of nest trees, has received some attention (Bunnel et al. 1994). With increased population growth and commercial activity, especially of coastal and estuarine areas in the Georgia basin, habitat for Bald Eagle roosting and nesting will be continually threatened. In addition to those stesses, there are major pollutant inputs, particularly from pulp mills and other wood processing industries. Coastal Bald Eagle populations in British Columbia apparently did not experience the major declines that occurred elsewhere during the organochlorine era. Anecdotal information (based on discussion with naturalists, farmers and fisherman) suggests that in the Fraser River delta, eagles were less common in the 1960s and 1970s than at present. In 1987, Vermeer et al. (1989) resurveyed areas of the southern Gulf Islands where nests had been counted previously (Hancock 1964; Trenholm and Campbell 1975) and reported a 30 % increase in the number of nests since 1974, which they attributed mainly to increasing food supply in the form of Glaucouswinged gulls (Larus glaucescens). However, data derived from such comparisons requires cautious interpretation, as it may be more indicative of increased survey intensity and ability to find nests (Henny and Anthony 1989). In the Okanagan Lakes region of interior British Columbia, Bald Eagles have disappeared as a breeding species (Cannings, 1987); orchard areas of the Okanagan valley received heavy DDT applications and wildlife samples from that area are still highly contaminated (Elliott et al. 1994). Problem Statement As a predator feeding at the top of marine and estuarine food chains, Bald Eagles are exposed to an array of persistent environmental chemicals, particularly chlorinated hydrocarbons and mercury. There is a considerable body of literature on levels of organochiorine pesticides and total PCBs in tissues of Bald Eagles. However, there is very little published data on levels of individual PCB congeners, particularly the toxic non-ortho PCBs, or on levels of other significant  15  environmental contaminants including polychiorinated dibenzo-p-dioxins (PCDDs) and polychiorinated dibenzofurans (PCDFs) in Bald Eagles. In addition, whereas significant progress has been made in detennining critical levels of DDT-related compounds and mercury for Bald Eagle eggs (Wiemeyer et al. 1993), there is no such information for other chlorinated compounds. The Strait of Georgia provides an interesting location to investigate the effects of PCDDs and PCDFs on eagle populations. Previous studies in the area showed that fish-eating birds, such as Great Blue Herons, Double-crested Cormorants, Western Grebes (Aechmophorus occidentalis) and Common Mergansers (Mergus merganser), all of which are potential Bald Eagle prey, were contaminated with high levels of PCDDs and PCDFs, but relatively low levels of other organochlorines (Elliott et at., 1989; 1992; Whitehead et at., 1990; 1992). Concentrations of PCDDs and PCDFs in Western Grebes and in Surf Scoters (Melanita perspicillata), another eagle prey item, collected near some British Columbia coastal mills in 1990 were high enough to warrant advisories against their consumption by people (Whitehead et al. 1990). In Great Blue herons, episodes of poor breeding success in the late 1980s at a colony near a kraft pulp mill were associated with sublethal effects on embryos, including edema, reduced body weight and EROD induction which correlated well with levels of 2,3,7,8-TCDD (Beliward et al. 1990; Hart et at. 1991; Sanderson et at. 1994a). Coastal Bald Eagle populations feed heavily on marine birds such as Western Grebes and Glaucous-winged Gulls and on larger fish (Knight et at., 1990). Eagles are therefore exposed to even higher dietary contaminant levels than species such as herons and cormorants which eat mainly smaller fish. In winter, after salmon runs are over, Bald Eagles eat mainly waterfowl (Watson et at., 1991) and thus are exposed to toxicants, such as lead shot and pesticides, acquired by waterfowl feeding in other distant areas, such as the western USA. Lead poisoning is a major cause of death for British Columbia Bald Eagles (Elliott et at. 1 992a), while pesticides are an important mortality factor in local areas such as the Lower Fraser Valley (Elliott et at. submitted).  16  There is, therefore, potential for exposure of Strait of Georgia Bald Eagles to potentially harmful levels of chlorinated organics and other toxicants. Positioned at the top of the food web and with a high public profile, Bald Eagles are an excellent sentinel species and indicator of ecosystem health. Thus, further research is warranted.  Hypotheses 4 Objectives: Mortality study Hypothesis: The accumulation of persistent chlorinated hydrocarbons will affect the survival of Bald Eagles, particularly if fat stores are depleted during periods of environmental stress. Objective: To investigate bald eagle mortality in British Columbia and specifically the role of chlorinated hydrocarbons versus other causes of death; to determine spatial and possibly temporal trends in contamination. Embiyotoxiciry study Hypothesis: Accumulated chlorinated hydrocarbons are transferred from females into eggs, where they negatively affect growth, development and survival of embryos. Objectives: To examine the health of Bald Eagle embryos exposed to an environmental gradient of chlorinated hydrocarbon pollutants and to relate the degree of exposure to biomarkers such as CYP1A induction; to document exposure by chemical analysis of yolk sacs. Bioaccumulation study Hypothesis: Chlorinated hydrocarbons, particularly PCDDs and PCDFs from pulp mill sources, are accumulating at high concentrations in bald eagle eggs as a result of their position as top predators in marine and estuarine food chains. Objectives: To determine spatial and temporal patterns of chlorinated hydrocarbons in Bald Eagle eggs and to relate those levels to the diet and to sources; to determine critical concentrations of contaminants, particularly PCDDs and PCDFs in the eagle diet. Productivity study  17  Hypothesis: The accumulation of persistent chlorinated hydrocarbons in Bald Eagles impairs overall reproduction through toxicity to embryos, reduce survival of nestlings or impaired development of the reproductive system. Objectives: To determine breeding success of a representative sample of eagles in the Strait of Georgia and reference locations and to relate breeding success to chlorinated hydrocarbon levels in nestling blood samples; to examine the role of other factors critical to breeding success, particularly food supply. Overview of the thesis This thesis represents the results of a four year field and laboratory study of chlorinated hydrocarbon exposure and effects in Bald Eagle populations on the coast of British Columbia. In the first chapter, mortality and the role of chlorinated hydrocarbons are examined through autopsy and liver residue analysis of eagles found dead and dying from 1989 to 1993 in British Columbia. Chapter two presents the results of a laboratory incubation study of in ovo effects of PCDDs, PCDFs and PCBs in an environmental exposure gradient. Contaminant levels in yolk sacs are presented with the results of biomarker assays, such as CYP1A, in embryonic tissues. The data are used to estimate a no-observed-effect-level (NOEL) and a lowest-observed-effectlevel (LOEL) for TCDD-toxic equivalents in eagle eggs. Chapter three presents contaminant residue levels for Bald Eagle eggs and prey items. Patterns, trends and sources are discussed and a simple bioaccumulation model used to relate levels in eagles to those in their food chain. In Chapter four, the results of productivity studies and contaminant levels in nestling plasma samples are presented. Relationships between breeding success, contaminant levels and other variables, particularly food supply, are discussed.  18  CHAPTER 1 CHLORINATED HYDROCARBON LIVER LEVELS AND AUTOPSY DATA FOR BALD EAGLES FOUND DEAD OR DEBILITATED, 1989-1993.  The objective of this study was to determine the degree of chlorinated hydrocarbon exposure of adult and juvenile Bald Eagles and to assess spatial trends in contamination. Statistical examination of relationships among environmental contaminant levels and cause of death was a secondary objective. Preliminary reports on toxicants such as lead (Elliott et al. 1992a) and anticholinesterase pesticides (Elliott et al. in press[b]) have been made, but are not included as part of the thesis. In this chapter, the results of autopsies and analyses of PCBs and organochiorine pesticides in livers of 59 eagles found dead in British Columbia over the period, 1988 to 1993, and results from a subset of 19 birds analyzed for PCDDs and PCDFs are presented and evaluated.  Materials and Methods Sample collection Specimens collected for this study were part of an overall investigation into the health status of Bald Eagles in British Columbia. Carcasses were obtained by writing to potential cooperators, including government and non-government agencies, veterinarians and wildlife rehabilitators and by placing advertisements in periodicals. Sick, injured and deceased Bald Eagles were thus obtained from all of the above sources. Specimens were received and initially examined at the Pacific Wildlife Research Centre and then shipped on ice to the Island Veterinary Hospital, Nanaimo, British Columbia, where they received a complete autopsy by Dr. K.M. Langelier.  19  The 484 eagles received were grouped by geographical area as follows: lower Fraser valley, Strait of Georgia, Johnstone Strait, west coast Vancouver Island and north coast. A total of 59 individuals were analyzed for organochiorines and PCBs (Figure 1.1, see also Appendix 1.1). Specimens for analyses were selected in order to provide a reasonably representative sub-sample, based on age, sex, and collection location. Other criteria were also considered such as a preliminary diagnosis of non-specific poisoning or proximity of the carcass to an industrial pollutant source. Some eagles were also analyzed for organochiorines during investigations of suspected poisonings by lead or anticholinesterase pesticides. Birds found  Figure 1.1 Locations where eagles were collected in British Columbia, 1989-93, and analyzed for chlorinated hydrocarbons (N = 59).  20  dead during the breeding season in the Strait of Georgia, and therefore likely to be resident birds, were considered to have priority for analysis. Concentrations of PCDDs and PCDFs were determined in nineteen liver samples. Criteria for selection of samples for PCDD/PCDF analysis were as follows: 1) collected in the Strait of Georgia or Johnstone Strait 2) collection date in late spring or summer, i.e. resident birds 3) breeding age birds 4) high organochiorine levels. Criteria were set to maximize chances of analyzing eagles which had been exposed to pulp mill pollutants. Based on elevated levels of total PCBs, nine samples were selected for high resolution GC/MS analysis of non-ortho PCB congeners. Linear regressions were determined between concentrations of non-ortho PCBs and total PCBs for the nine livers analyzed, in order to estimate values for the other ten livers which had been analyzed for PCDDs and PCDFs and thus to estimate TCDD toxic equivalents. Regressions were not significant for PCBs 77, 81 and 37, but were significant for PCBs 126 and 169: PCB 126 (ng/kg)  =  92 [sum-PCB5 (mg/kg)] + 310, r 2  PCB 169 (ng/kg)  =  27 [sum-PCBs (mg/kg)] + 75, r 2  =  =  0.660, p<O.Ol 0.584, p <0.05  Chemical analysis Carcasses were stored at -20° C until postmortem examination. Tissue samples were frozen at -20°C in chemically-cleaned (acetone/hexane) glass jars, frozen, and shipped to the National Wildlife Research Centre (NWRC), Hull, Quebec, for analysis in the laboratory of Dr. Ross Norstrom. Organochlorines in liver were analyzed according to methods described previously (Norstrom et al. 1988), except that PCBs were reported as the sum of 28 congener peaks. Briefly, 2-4 gram sections of liver were dehydrated by grinding with excess anhydrous sodium sulfate and colunm extracted with 50% methylene chloride in hexane. After extraction, the eluate was concentrated on a rotovapor, further mixed with hexane and a 0.5 ml sample taken for lipid determination (removal of solvent and weighing of residue). The remaining extract was then cleaned up and separated into three fractions by Florisil chromatography. The fractions were analyzed by gas chromatography-electron capture detector using a 60m DB-5 21  capillary column (Superco Inc.). Fraction 1 contained PCBs, p,p’-DDE, hexachlorobenzene, pentachioroberizene, tetrachlorobenzenes and mirex. Fraction 2 contained cis-chiordane, oxychiordane, trans-nonachior, and beta-hexachiorocyclohexane. Fraction 3 contained dieldrin. Recoveries of these compounds by this method ranged from 82-94%. Quantification of PCB congeners was effected by using a calibrated internal PCB standard solutions. Detection limits were 0.005 mg/kg for organochiorine pesticides and 0.0025 mg/kg for PCB congeners. Livers from 1990 collections were analyzed for PCDDs/PCDFs by low resolution GC/MS using a Hewlett-Packard 5987B with a 30 m DB-5 capillary GC column according to methods described in Norstrom et al. (1990) and Norstrom and Simon (1991). The method employed gel permeation-carbon chromatographic clean-up and the use of 13 -labelled 1 C 2 internal standards for quantification. Analysis of PCDD/PCDFs and non-ortho PCB in livers from other years were carried out according to methods in Letcher et al. (in press). The method involves neutral extraction followed by removal of lipids and biogenic compounds by gel permeation chromatography and alumina column cleanup. Separation of PCDDs, PCDFs and non-ortho PCBs from other contaminants was achieved using a carbon/fibre column; further separation of PCDDs/PCDFs from the non-ortho PCBs was effected by Florisil column chromatography. Quantitation was performed with a VG Autospec high resolution mass spectrometer linked to a HP 5890 Series II data system. Each sample was spiked with ‘ -labelled 1 C 3 2 PCDD and PCDF congeners (TCDD/TCDF to HpCDD/HpCDF and OCDD) and non-ortho PCBs (PCBs 77, 126 and 169) internal standards, prior to lipid extraction, for internal standard quantitation and calculation of internal standard recoveries. Two other ‘ -labelled 1 C 3 2 standards (1 ,2,3,4-TCDD and 123789HxCDD) were added to the cleaned PCDD/PCDF extracts and PCB 112 to the non-ortho PCB fraction, just prior to analysis to serve as recovery standards, for quantification of internal standard recoveries. Recoveries of -PCDDs/PCDFs/non-ortho 12 PCBs were calculated by C 13 comparing the integrated areas of the labelled internal standards and the areas of the recovery standards in the samples to the areas of those compounds measured in the external standard  22  mixture, analyzed along with the samples. Results were generally accepted when recoveries of labelled standards were between 70% and 120%. Statistical analysis Organochiorine pesticide data were transformed to common logarithms and geometric means and 95 % confidence intervals were calculated with the data grouped by collection site. Differences among sites were tested by a two-way ANOVA followed by Tukey’s multiple comparison procedure (MCP). To test for an association between residue levels and cause of death, birds were grouped into 12 categories (Figure 1.2) and analyzed by a one-way ANOVA. All statistical tests were done using SYSTAT. A value of p < 0.05 was used throughout. TCDD-toxic equivalents (TEQ5) were calculated using three different sets of TEFs, Safe’s (1990), chick embryo hepatocyte (CEH) (Kennedy et al. in press) and WHO (Ahlborg etal. 1994).  Results  Autopsy results The diagnosed cause of death for each individual bird analyzed for organochiorines is included in Appendix 1-1. Autopsy results for the 59 Bald Eagles analyzed for organochiorines were compared to the total of 484 examined in the broader study with the causes grouped into twelve categories (Figure 1.2). The graphs indicate that the subset for analysis was reasonably representative of the range of mortality factors. Only two minor categories, falling from the nest and infectious disease, were not represented. There were no statistically significant associations between any of the chlorinated hydrocarbon levels and cause of death. However, given the relatively small sample size, even within the Strait of Georgia, and the variance in the residue levels, the probability of detecting a significant association was low.  23  Clinical diagnosis  liii  D  trauma  186  electrocution undetermined  46 veh. collision  14..  30  ‘  j25  I24  eagle attack fell from nest  j2  drowning  *:.•(•n•O  ófbátdéáglej):.::..:.....  I  0  117  I•  I  15 20 Percent Bald eagles analyzed for OCsIPCBs, N59 5  18  •••,•••••.•••••..  •.•..  infectious  120  ...::  10  25  10  5  0  15  20  25  Percent Bald eagles submitted to Island Veterinary Hospital, N484  Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed compared to the complete set of birds received.  Organochiorines and total-PCBs Organochiorine pesticides and total PCBs were generally low; most eagles had DDE and PCB levels < 5.0 mg/kg (Figure 1.3). However, a few birds had elevated levels of DDE and total PCBs (>50 mg/kg) and chiordane-related chemicals (>1.0 mg/kg) (Appendix 1.1)  44.%l 25-  I.  •.:  DDE :::totaI PCBS  —  —•  ••.•• .••.,  ‘.-.••••  -  -•  -.  8% -  1-Li  -  ..  -  ,-10  Concentration (mg/kg, wet wt.)  Figure 1.3  Numbers of Bald Eagles showing different DDE and PCB levels in livers (N=59) 24  Quantifiable levels of total PCBs, trans-nonachlor and oxychiordane were present in all 59 samples analyzed while DDE was present in 98 % of the samples. There were quantifiable levels of DDD, heptachlor epoxide and dieldrin in 96 %; DDT, hexachlorobenzene (HCB) mirex, beta-hexachlorocyclohexane (b-HCH) trans-chlordane and cis-nonachior in 92 %; octachiorostyrene (OCS) in 80 %; and photomirex in 50 % of the samples. Significantly elevated geometric mean residue levels were measured in Johnstone Strait samples, followed by the Strait of Georgia, with the other four sites all being lower. Mean DDE levels were significantly higher in samples from Johnstone Strait compared to the lower Fraser valley.  1000  :  :  :  :  :  -  100— E  ci)  1  :  :  :  :  :  :  :  :  :  :  :  0  Ainddatapoints .  0.0001  Inn  I  I  I  -.  maverages  .  I  I  I  I  I  I  —  —-:  -  •—  —  ci,  -  :  -  :  :  :  :  : : : :  nd. data points  C.) -  —  I  Jan  .  Feb.  Mar.  Apr.  May  I  I  Jun.  Jul.  Aug  Sep.  Oct.  I  I  Nov.  Dec.  Month  Figure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month  25  Individuals with elevated organochiorines were found mainly in late spring or early summer (Appendix 1.1). Concentrations of DDE and total PCBs in eagle livers tended to increase throughout the winter, peak in April, level off and even decline slightly in summer (Figure 1.4).  1000-  =  =  =  -  -  =  100  =  =  -  .:  =  =  -  C  E  -:  0) 0)  _E  10—  C  1-  =  C  C  :::,::  =  =  :  ::  =  =  ::  :  C  :C  o =  =  =  C  =  :  =  =  ----.4  ::  =  =  =  =  E E E 4E E E E :::::.:::.:::.a:::.:  =  C  =  =  =  ,  E E ..  0  C.)  0.1  0.01—  100-  I  I  I  I  I  —  -  -  -  10  :::::::::::::::::::::::::. 0) C  o =  -  -  — ..  1  —  —  q  —  -  —-.  .....  I  C C.)  c.  -  —  a)  C 0  I  _  01—  C.)  —..  0.01—  —— ——.  1  0  I  I  2  3  I  I  4  5  *  Bald eagle body conthtion *  Scale: 0  poor, 5  -  excellent  Figure 1.5 Liver DDE and PCB residue levels in relation to body condition  26  Although eagles with higher residue levels of DDE and PCBs tended to weigh less than those with lower residues, for neither DDE (ANOVA, f=3.28,  p=O.077) nor PCBs  (f=3.5,  p =0.068) was the relationship significant. However, comparison of DDE and PCBs with a numeric scoring of body condition did produce statistically significant negative relationships for DDE (f=7.4 p=O.009) and PCBs (f=8.5 p =0.005 (Figure 1.5). Non-ortho PCBs  Levels of three non-ortho PCB congeners and two mono-ortho PCB congeners, PCBs 105 (234-(234-34) and 118 (245-34), which are present at relatively high concentrations and  also considered to be partial Ah-receptor agonists (Safe 199), are presented in Table 1.2. For the non-ortho PCBs, in most samples, the pattern was of PCB 126 > 77 > 169 > 81 > 37. There were some exceptions; in three cases (Dent Island, Nanaimo and Port Hardy, 1990), PCB 77 > 126. In one case, Campbell River, PCB 169 >77.  PCDDs and PCDFs  The most contaminated individuals were from near pulp mill sites, Powell River and Campbell River or nearby areas, such as Bowser and Sechelt (Table 1.3). In most samples, highest levels were of HxCDD followed by PnCDD; after that, the relative levels of TCDD, TCDF and PnCDF were very variable.  Toxic equivalents  TEQ results varied widely among the three sets of factors. Highest values were consistently produced using the chick embryo hepatocyte derived numbers, followed by Safe’s and then the WHO TEFs (Table 1.4). The TEQ5WHO ranged from 53 to 2740 ng/kg. Two birds had liver TEQs 110 > 2000 ng/kg, while an additional two birds had liver 1000 ng/kg.  27  TEQSWHO  >  NJ  1.31  ab 146  73  3.2  Strait of  (0.819-29.7) 1.01  (0.685-16.5) ab 0775  (73-77) 70  1  Northern  *  a.b  -  -  (0.29-1.64) 0.429  (62-73) 73  (2-11)  2.1  0.689’  67  4.6  0.42  (0.255-5.07) 0.067  (0.042-0.149)  0.079  0.073  (0.058-1.78)  0.321  (0.057-0.156)  0.094  (0.022-0.096)  0.046  nonachior  trans-  0.011  (0.006-0.033)  0.014  0.013  (0.008-0.337)  0.052  (0.01-0.031)  0.018  (0.005-0.020)  0.01  Oxychlordane  0.005  (0.003-0.015)  0.006  0.007  (0.006-0.153)  0.030  (0.004-0.014)  0.008  (0.002-0.009)  0.004  Mirex  0.002  (0.005-0.016)  0.009  0.013  (0.007-0.270)  0.042  (0.006-0.021)  0.011  (0.001-0.016)  0.003  B-HCH  insufficient sample size to calculate confidence interval  means that do not share the same letter are significantly different (P<0.05). NOTE: significant differences among sites were found only for DDE  Interior*  4  North Coast  Vancouver Is.*  3.6  West Coast  2  (2.0-5.1)  Strait  1.14  4.93  36 • 3 b  75  3.2  Johnstone  9  (0.057-1.56)  (0.811-2.62)  (72-74)  (2.8-3.7)  Georgia  33  (0.186-1.58)  (0.253-1.47)  0.542  0.609a  71 (70-72)  4.1  DDE  Total PCBs  0 2 % H  (3.3-5.1)  10  Lower Fraser  % fat  Valley  N  -  0.009  (0.007-0.023)  0.013  0.013  (0.005-0.188)  0.032  (0.008-0.027)  0.014  (0.002-0.03)  0.007  Dieldrin  0.011  (0.003-0.057)  0.014  0.011  (0.006-0.097)  0.023  (0.009-0.02)  0.014  (0.006-0.016)  0.01  HCB  Organochiorine residue levels (mg/kg, wet weight), geometric mean ± 95% confidence intervals, in livers from Bald Eagles found dead in British Columbia, 1988 1993.  Location  Table 1.1  Table 1.2  Location  Selected non-ortho and total PCBs in Bald Eagle livers collected from the south coast of British Columbia (wet weight). Date  Sex! Age  BC(a)  Initial Etiology  Non-ortho PCBs #771)  #126  Total PCBs  #169  (ng/kg) Port Hardy  27 Jun/89  F/A  1  Undet.  Port Hardy  6 Mar/90  M/3y  1  Undet.  Port Hardy  2 Apr/90  M!ly  0  Inanit.  Port Hardy  May/93  F/A  1  Port Hardy  May/93  F/A  Campbell R.  9 Jun/90  Campbell R.  1270  Comments  (mg/kg)  1170  221  6.42  349  86  0.425  2300  4800  2180  43.8  Inanit.  2000  2550  533  65  4  Trauma  1070  1490  368  12  F/2y  5  Tauma  357  88  0.515  31 Jul/90  F/3y  0  Trauma  248  688  160  7.52  Campbell R.  16 Apr/93  F/A  4  Electro.  738  9960  2640  71.7  Powell R.  26 Apr/90  M/A  1  Electro.  5820  1690  60  *  Hg tox.  Powell R.  18 Jun/90  F/A  3  Tox.Pb  499  130  2.06  *  Pb exp.  Comox  13 Jun/90  M/3y  4  Trauma  561  148  2.72  *  Pb exp.  Denman Isl.  19 Jul/90  M/A  3  Asphyx.  426  108  1.26  *  Pb exp.  Bowser  7 Jul/90  F/4y  1  Tox.Pb  2640  759  25.4  *  Pb tox.  Coombs  3 Mar/90  M/A  1  Tox.Pb  361  90  0.558  *  Pb tox.  Nanoose  26 Apr/90  F/A  2  Trauma  496  129  2.02  *  Pb exp.  Nanaimo  8 Feb/90  F/A  5  Electro.  370  56  4.56  Sechelt  7 May/90  F/A  3  Electro.  783  213  5.15  Dent Isl.  5 Apr/93  F/ly  1  Drown.  870  472  138  9.27  Victoria  14 Nov/92  M/A  1  Trauma  1620  2240  523  7.94  -  -  -  -  -  -  -  -  -  395 -  *(c)  *  Pb exp.  Hg exp/Pb-exp. *  Body Condition: 0-emaciated, 1-thin, 2-fair, 3-good, 4-very good, 5-excellent = values not calculated since regression not significant (c) * non-ortho PCBs calculated from regression equations Non-ortho PCB Minimum Detection Limit (MDL) = 3 ng/kg wet wt; mono-o,iho PCB MDL = approx. 0.5 pg/kg wet. wgt. A = adult, ly, 2y, 3y = age of subadults Undet. = Undetermined, Inanit. = Inanition, Electro. = Electrocution, Tox. Pb = Toxicosis, Asphyx. = Asphyxiation, Drown. = Drowning (a)  -  -  -  29  Date  Sex/Age  BC(a)  Initial Etiology 2378TCDD 12378PnCDD  123678HxCDD 2378TCDF  23478/ 13489PnCDF  Comments  Concentrations of selected PCDDs and PCDFs in Bald Eagle livers collected from the south coast of British Columbia (ng/kg, wet wt.)  -  -  15* F/A Undet. 27 June/89 1 Port Hardy 33 71 81 101 1 M/3y Undet. Port Hardy 6 Mar/90 5 10 20 9 trace 49* 2 Apr/90 M/3y Inanit. Port Hardy 0 77 241 280 41 20* F/A Inanit. 1 Port Hardy May/93 54 232 350 3 11* F/A Trauma 4 Port Hardy May/93 17 141 49 45 F/2y Trauma 5 18 Campbell R. 9 June/90 25 37 33 8 10* F/3y Trauma Campbell R. 31 July/90 0 30 76 56 1 Pb-exp. 105* F/A Electro. 4 212 Campbell R. 16 Apr/93 2120 793 8 26 Apr/90 M/A Electro. 1 392 Powell R. 1420 4360 3 375 Hg-exp. F/A Tox:Pb 3 Powell R. 18 June/90 41 83 184 63 27 Hg-tox/Pb-tox. M/3y Trauma 4 Comox 13 June/90 21 51 169 60 17 Pb-exp. M/A Asphyx. 3 Denman Isi. 19 July/90 49 295 92 78 30 Hg-exp/Pb-tox. F/4y Tox:Pb 1 Bowser 7 July/90 263 2050 603 15 152 Hg-exp/Pb-tox. M/A Tox:Pb 3 Mar/90 1 Coombs 6 9 10 42 6 Pb-tox. F/A Trauma 2 Nanoose 26 Apr/90 23 57 28 90 11 Pb-exp. 5* F/A Electro. 5 8 Feb/90 25 Nanaimo 40 51 15 F/A Electro. 7 May/90 Sechelt 3 29 199 24 936 138 Hg-exp./Pb-tox. 13* 1 F/ly Drown. Dent Is!. 5 Apr/93 4 208 18 35 18* 1 14 Nov/92 M/A Trauma Victoria 30 108 77 15 (a) BC Body Condition: 0-emaciated, 1-thin, 2-fair, 3-good, 4-very good, 5-excellent (b) * 13489-PnCDF not included (c) trace = <2 ng/kg wet wt. MDL Minimum Detection Limit (signal/noise) = 3 ng/kg wet wt. Undet. = Undetermined, Inanit. = Inanition, Electro. = Electrocution, Tox. Pb = Lead Toxicosis, Asphyx. = Asphyxiation, Drown. = Drowning A = adult, ly, 2y, 3y = age of subadults  Location  Table 1.3  F/A  27 June/89 6 Mar/90 2 Apr/90 May/93 May/93 9 June/90 31 July/90 16 Apr/93 26 Apr/90  Port Hardy  Port Hardy  Port Hardy  Port Hardy  Port Hardy  Campbell R.  Campbell R.  Campbell R.  Powell R.  14 Nov/92  Victoria  A = adult ly 2y, 3y = age of subadults Undet. = ‘Undetermined, Inanit. = Inanition, Electro.  (I)  M/A  F/ly  F/A  F/A  F/4y  M/A  1  1  3  5  2  1  1  3  4  3  1  4  0  5  4  1  0  1  1  0 BC  =  (k)  (d)  (b)  =  Toxicosis, Asphyx.  -  1110  282  815  472  274  99  4220  335  334  384  6550  7460  711  130  1080  4120  5490  95  852  Safe  2100  453  1180  832  491  221  5960  552  621  677  10100  13100  1140  268  1920  7430  11700  204  1560  CEHOI)  394  110  417  129  135  60  1430  205  155  193  2740  2440  197  83  302  832  1220  53  276  WHO’  Pb-exp  Hg-exp/Pb-exp  Hg-exp/Pb-exp  Pb-exp  Hg-tox/Pb-exp  **  **  +  **  +  Hg-exp/Pb-exp  Pb-exp  + Pb-tox  +  +  +  +  + Hg-exp  **  **  +  **  **  **  +0)  **)  Comments  =  Asphyxiation, Drown.  =  Drowning  2378-TCDF, 23478/13489-PnCDF PCB con eners #118 and 105 BC Bocy Condition: 0-emaciated, 1-thin, 2-fair, 3-good, 4-very good, 5-excellent TEFs from Kennedy et al. in press ** 13489-PnCDF not inchided in TEQ calculations PCB con eners #126 & 169 calculated from regression equations; #77 not mcluded in ‘EQ calculations  Trauma  Drown.  Electro.  Electro.  Trauma  Tox:Pb  Tox:Pb  Asphyx.  Trauma  Tox:Pb  Electro.  Electro.  Trauma  Trauma  Trauma  Inanit.  Inanit.  Undet.  Undet.  Initial Etiology  Electrocution, Tox. Pb  2378-TCDD, 12378-PnCDD 123678-HxCDD PCB coneners #77, 126 and 129 * PCDIj, PCDFs, non-ortho PCBs Minimum Detection Limit çsnal/noise = 3( s from safe 1 TEFs from Ahltorg et al., 1994  5 Apr/93  Dent Isl.  (a) (e) (e)  8 Feb/90  26 Apr/90  Nanoose 7 May/90  3 Mar/90  Coombs  Nanaimo  F/A  7 July/90  Bowser  Sechelt  M/A  19 July/90  Denman Isl.  M/3y  Comox  F/A  18 June/90 13 June/90  Powell R.  M/A  F/A  F/3y  F/2y  F/A  F/A  M/3y  M/3y  Sex/Age  TEQs  (d) levels in 5 Comparison of TEQs calculated from selected pCDDS(a), PCDFs(b), non-ortho and mono-ortho PCB wt.(e)). Bald Eagle livers collected from the south coast of British Columbia (nglkg, wet  Date  Location  Table 1.4  Discussion Chlorinated hydrocarbon levels in livers of Bald Eagles tested for this study were generally low; however, a small number of birds found dead or debilitated from the Strait of Georgia or northern Johnstone Strait had elevated PCDDs, PCDFs, PCBs and organochlorines. Higher liver levels of lipid soluble contaminants in sick or dead birds do not necessarily mean that their death was a direct result of toxicity due to those chemicals. Most of the eagles with higher chlorinated hydrocarbon levels were in poor body condition, indicating lipid and contaminant mobilization. Body weight was negatively correlated with liver organochiorine levels in other studies (Cooke et al. 1982; Reichel et al. 1980). A variety of factors can contribute to weight loss, including: seasonal utilization of fat stores, poor foraging abilities of juvenile birds, a debilitating injury, disease or toxicosis, and the anorexic effects of chemicals such as lead, dieldrin and TCDD. Discriminating among these factors in a sample of wild birds is difficult. Body weight loss per se can, however, be symptomatic of toxicity. A number of the birds with elevated chlorinated hydrocarbon levels were also lead exposed or poisoned (Appendix 1.1). Chronic lead-poisoned birds exhibit wasting and extreme loss of body weight and appear clinically to have starved (U.S. Fish and Wildlife Service 1986). Dieldrin exposure can induce fasting (Heinz and Johnson, 1982). Weight loss due to fasting, referred to as wasting syndrome, is the cause of death in acutely TCDD-exposed mammals (Peterson et al, 1984) and birds (Nosek et al. 1992). Bald Eagles with the highest PCDDs/PCDFs and TEQ5WHO (2440, 2730 and 1430 ng/kg) were found in the vicinity of pulp mills on south east Vancouver Island. Proximal causes of death were electrocution in two cases and lead poisoning in the third. However, in one particular case, an adult male eagle from Powell River, total TCDD-toxic equivalents were calculated to be from 2740 TEQ5WHO to 6550 ng/kg TEQSSafe. A sample of the solvent extract was tested in a chick embryo hepatocyte bioassay (Kennedy et al. 1993) and TEQs were estimated at 13,100 ng/kg. This bird was also in very thin body 32  condition, perhaps indicating that it suffered from wasting syndrome. There are no data on tissue levels of TCDD-like compounds which could be diagnostic of acute toxicity. LD s 50 reported for 2,3,7,8-TCDD are 240 ng/kg in chicken and 1350-2 180 ng/kg in pheasant embryos (Peterson et al. 1993). Lethal doses in adult birds are estimated to be one to two orders of magnitude higher (ibid. 1993). A pattern of increasing mean contaminant levels in spring partly reflects normal seasonal lipid dynamics. Late summer and fall deposition and winter mobilization of fat is typical of temperate climate species, adapted for winter survival (Stalmaster and Gessaman 1984). Seasonal deposition and mobilization of lipids and lipid soluble contaminants such DDE, PCBs and dieldrin was shown in three species of predatory birds monitored for many years in Great Britain (Cooke et al. 1982). Starvation and associated lipid and contaminant mobilization can result from reduced foraging ability caused by debilitating injury or disease. Starvation without injury or disease should be more common among juvenile birds which, particularly during their first winter, are less efficient at finding food (Todd et al. 1982). However, only one of nine eagles with liver DDE levels > 10 mg/kg was a juvenile, a firstyear male found in 1990 at Port Hardy, in very poor condition and believed to have starved. The age ratio of birds selected for analysis is somewhat skewed towards adults, because of greater conservation interest in birds which have reached breeding age. Juvenile eagles, particularly first-year birds, may also have lower chlorinated hydrocarbon levels than adults, as they have had less time to reach pharmacokinetic equilibrium with dietary residues, which took up to two years in Great Lakes herring gulls (Anderson and Hickey, 1976). Juvenile eagles eat more fish (Stalmaster 1987), which would also tend to have lower contaminant levels than fish eating birds, which are eaten more often by adults (Chapter 3). Only one bird had > 100 mg/kg DDE in liver, the level suggested by Cooke et a!. (1982) as indicative of acute poisoning, although two other birds had liver DDE levels of 91 and 96 mg/kg. None of the birds had PCB levels in livers > 100 mg/kg, considered indicative of toxicity (Cooke et a!. 1982). One bird had levels of oxychiordane in liver > 2 mg/kg and trans-nonchior levels > 7 mg/kg. Diagnostic liver levels of oxychiordane are not available;  33  brain levels of 1.1  -  5.0 mg/kg indicate acute toxicity (Stickel et al. 1979). In an earlier  sample of nine eagles found dead, 1969 to 1973, from British Columbia, one bird had 179 mg/kg DDE and 23.7 mg/kg dieldrin (Friis, 1974), well above the level of 5 to 10 mg/kg dieldrin in liver, indicative of acute poisoning (Cooke et al. 1982). None of the eagles in the present sample had elevated dieldrin levels, indicating an improvement in dieldrin contamination of the eagle foodchain. However, the presence of potentially toxic levels of DDE in livers of British Columbia Bald Eagles more than 20 years after DDT was heavily restricted in North America raises questions regarding sources. A number of hypotheses have been suggested in the literature to account for sources of continuing high levels of DDT in the environment. Recent data show that DDT can persist at high levels in soils and foodchains in areas of former intensive use or manufacturing (Blus et al. 1987; Elliott et a!. 1994). Eagles may also acquire some DDT from feeding on migrant waterbirds, which are exposed to ongoing use in Latin American wintering areas (Fyfe et a!. 1990). Finally, on the Pacific coast, elevated DDE levels in seabirds, such as storm-petrels, important seasonal prey items of eagles nesting on their colonies, indicates long-range transport from recent use in Asian countries (Elliott et a!. 1989). Elevated PCBs in some eagle livers likely originate from industrial sources in the Georgia basin, as PCBs were significantly elevated in samples from the Strait of Georgia, compared to other sites in both egg (Chapter 3) and nestling plasma samples (Chapter 4). Although most toxic effects of TCDD are thought to be mediated via the Ah receptor, it is possible that the anorexic effects of TCDD are not Ah-receptor mediated (Tuomisto and Pohjanvirta 1991). Therefore, it would be interesting to know if any biomarkers of Ah-like toxicities were activated in eagles with high liver TEQs. Indirect indications, at least of CYP1A induction, may be inferred from examination of TCDD/TCDF ratios, which varied greatly between eagles with high versus low TCDD exposure. For example, TCDF levels are much lower in the three birds with the highest TCDD levels (212, 392, 263 ng/kg in liver); the mean TCDD/TCDF ratio for those three birds was 74. In contrast, the mean TCDD:TCDF ratio is 0.17 for the three birds with the lowest TCDD levels (5,6,4 ng/kg in liver). The  34  TCDD/TCDF ratio in the high TCDD birds is also markedly different from ratios observed in eggs. Mean ratios in eagle eggs were 0.58 at Powell River and 0.32 in Jolmstone Strait (Table 2.1). This shifting ratio may indicate that hepatic cytochrome P450 enzymes have been induced in birds exposed to elevated TCDD levels; consequently, TCDF has been metabolized (Van den Berg et at. 1993). A hepatic CYP 1 A cross-reactive protein was shown to be present and inducible in Bald Eagle chicks (Chapter 3) and should, therefore, also be inducible in adult eagles. CYP1A1 was recently shown to be the protein responsible for TCDF metabolism in rats and humans (Tai et at. 1993). Alternatively, higher liver TCDD concentrations in more highly exposed birds may be evidence of the dose-related increase in liver retention of TCDD, reported for rats (Abraham et a!. 1988). Inducibility of a hepatic binding protein, possibly CYP1A2, has been suggested as a mechanism for increased TCDD retention at higher doses (Van den Berg et a!. 1993). CYP1A enzymes can also metabolize certain PCB congeners and thus alter the PCB pattern (Brown 1994). The PCB congener pattern between birds classified as good versus poor body condition is compared in Figure 1.6. As discussed above, birds in poor condition have higher chlorinated hydrocarbon levels in liver, because of lipid and contaminant mobilization, and thus, hepatic P450 enzymes may have been induced. Differences in mean percent total PCBs were not significantly different for any of the congeners measured (t-test, p <0.05); however, a consistent trend is apparent, whereby the percent contribution of the lower chlorinated compounds was consistently lower and the higher chlorinated compounds consistently higher in the poor condition group. CYP1A induction should increase the metabolism of non-ortho and mono-ortho PCBs but not those with two or more ortho chiorines (Brown 1994). In particular, compounds such as PCBs 118 (245-34) and 99 (245-24) and 70 (245-4) which have been suggested as indicators of CYP1A metabolism (Brown 1994), as well as 60 (234-4) and 101 (245-25), appear lower in the poor condition group. From this indirect evidence, it appears that at least hepatic CYP1A enzymes were induced in eagles, suggesting the possible activation of other Ah-mediated processes.  35  20  15 0  C.) 0  F° 0  4-.  0 0  5  0  rjZ  qç’  of PCB congeners  Figure 1.6 PCB congeners in Bald Eagle livers expressed as percent of total PCBs compared for birds in good and poor body condition (N=9, for each group).  In conclusion, the majority of eagles found dead in this study had relatively low (< 5 mg/kg) levels of DDE and PCBs, and even lower levels of other organochiorines. However, a few birds had DDE levels diagnostic of acute poisoning, more than 20 years after regulatory restrictions on DDT usage in North America. At least one eagle found near a bleached kraft pulp mill had liver TEQWHO levels potentially indicative of acute toxicity. Differences in TCDD/TCDF ratios in birds with high 2,3,7,8-TCDD levels may indicate hepatic cytochrome P450 induction and TCDF metabolism in those birds. Because of the selection criteria, samples analyzed for PCDDs and PCDFs were biased towards birds with a higher probability of such exposure. Nevertheless, 4/19 (21 %) of eagles tested had > 1,000 ng/kg TEQSWHO in their livers. All of those birds were of reproductive age  36  found during the breeding season. This may indicate that acute exposure to TCDD-like compounds has removed a component of the breeding eagle population in the Strait of Georgia.  Acknowledgements  Dr. K.M. Langelier performed the final autopsies. Working in the laboratory of Dr. R. Norstrom, M. Simon and H. Won did the chemical analysis. L. Wilson, P. Sinclair and I. Moul assisted in procurring of carcasses. I thank all those people who submitted birds for the study. Funding was provided by the Canadian Wildlife Service.  37  Date  5  4 y. Ad. Ad. 4 y.  F F F  M  11 Feb/93  Mar/93  Campbell R.  Campbell R.  18 Jun/90  28 May/90  4 Feb/92  13 Jun/90  19 Jul/90  7 Jul/90  3 Mar/90  Powell R.  Comox  Comox  Comox  Denman Isl.  Bowser  Coombs  26 Apr/90  3 y. Ad.  F  9 Feb/93  Campbell R.  Powell R.  Ad.  F  31 Jul/90  Campbell R.  11 Feb/93  Ad.  M  9 Jun/90  Campbell R.  9 Apr/90  2 y. 3 y.  F  9 Jun/90  Campbell R.  Merville  Ad.  M  18 Apr/90  Campbell R.  4 y. Ad.  M F M  M  I y. 3 y. Ad.  n/a  Ad. Ad.  F  3 y. 3 y. Ad.  F  M  M  F  F  Ad.  18 Apr/90  M  16 Apr/93  F  F  Ad.  5  2 y. Ad.  M  M  1  1  3  4  0  5  3  1  1  3  3  2  4  0  5  5  5  5  4  5  5  3  5  5  Ad.  4  n/a  Campbell R.  Powell R.  3 4  M  3 y. Ad.  Age  M  F  F  Sex  Campbell R.  STRAIT OF GEORGIA  FRASER VALLEY Coal Harbour 12 Mar/90 Coal Harbour 26 Feb/91 New West. 28 Mar/92 Chilliwack 26 Apr/93 DeRoche 27 Dec/92 Delta 11 Dec/90 Ladner 2 Feb/92 27 Jan/92 Westham Isl. Tsawwassen 13 Dec/91 Pt. Roberts 1 Feb/92 Mean S.D. N  Location  n/a 2.50  3.98  3.86  3.15  5.80  Toxic:Pb  Toxic:Pb  Asphyx.  Power.Coll.  Inanit.  Veh.Coll.  Toxic:Pb  Electro.  Undeter. trauma  Toxic:Pb.  Electro.  Gunshot  Undeter. trauma Electro.  n/a 4.40 3.70 3.95 3.70 3.75 3.18 3.41  5.45  Undeter. trauma Undeter. trauma  5.00  Intra.Agg.  Intra.Agg.  nia 4.55  Electro.  10  3.0  2.9  1.8  3.6  2.9  3.5  2.5  1.9  1.6  5.8  4.8  3.0  5.4 5.4  3.9 4.0  3.2  3.1  72.8  72.5  75.8  71.9  76.1  71.2  75.6 74.8  79.0  69.5 71.0  75.6  68.0  73.8  72.0  66.3  69.0 69.4  0.529  1.002 15.743  2.283  5.746  0.950  1.649  91.405  3  1.191  0.151  0.420  9.838 0.188  1.186  0.352  0.624  0.513  56.997  0.701 10  1.27  10  76.9  0.914  71.1  4.3 1.2  4.1  2.351  0.774  0.256  0.016  1.164  0.391  1.660  1.510  0.438  0.576  DDE  68.8  69.7  72.0  72.4  71.6  72.4  70.5  69.9  72.7  70.6  0 2 % H  5.6  4.1  Veh.CoIl. Undeter.  5.2  3.3  2.7  5.7  5.6  4.1  3.8  2.5  % Fat  Undeter.  Undeter.  Undeter.  Gunshot  Electro.  Undet.  Trauma  Electro.  Initial Etiology  3.75  5.00 4.90 5.30 n/a 3.75 n/a 6.14 n/a n/a  4.09  Wt(kg)  0.004 0.014 0.049 0.156 0.013 0.009 0.100 0.033 0.012 0.783 0.007  0.003  0.005  0.090  0.012  0.009  0.014  0.011  0.439  10  0.008  0.013  0.033  0.011  0.015  0.001  0.013  0.005  0.019  0.017  0.007  0.010  Oxychlor.  0.043  2.207  0.081  0.161  0.327  0.075  0.142  2.656  0.281  0.079  0.016  0.031  0.018  0.457  0.075  0.050  0.043  0.079  0.122  10  0.039  0.062  0.120  0.043  0.045  0.005  0.058  0.022  0.137  0.080  0.045  0.066  t-Nonachl.  Appendix 1-1. OCs and PCB levels in Bald Eagle livers collected from British Columbia (mg/kg wet wt.)  60.047 2.059 0.654 3.025 2.729 1.261 25.423 0.558  3.241  0.591  0.143  0.354  0.142  7.516  1.031  0.515  1.156  0.761  71.789  10  0.617  0.909  2.104  1.206  0.614  0.042  1.089  0.249  1.331  1.521  0.318  0.617  Total PCBs  Pb-tox.  Hg-exp/Pb-tox.  Hg-explPb-exp.  Pb-exp.  Hg-toxlPb-tox.  Hg-exp.  Pb-exp.  Pb-tox.  Pb-exp.  Pb-exp.  Pb-exp.  Phorate  Pb exp.  Parathion  Comments  3  n/a 1 y. Ad.  n/a F M  15 Jan/93  5 Apr/93  14 Nov/92  Dent 1st.  Victoria  Saanich  2  Tahsis  7 Jul/90  F  Ad.  WE COAST VANCOUVER ISLAND 5.45  Undeter. trauma  2.5  71.7  9  76.7  73.4  9  N  5  3.4 2.3  2.58  Inamt.  Trauma  2.3  3.70  4.50  S.D.  1  4  75.9  78.4  71.8  79.2  72.2  73.6  73.0  74.9  Ad. Ad.  2.6  1.7  9.3  1.4  4.4  3.7  4.8  3.7  n/a  Port Hardy  Infection  Veh.Coll.  4.55 4.50  4 4  Inanit.  2.50  Undeter.  Undeter.  No diagnosis  Suspect toxic.  0  n/a  3.52  n/a  4.70  Mean  n/a  Port Hardy F  Ad.  F  1 May/90  20 Oct/92  Port Hardy  Port Hardy F  1 y.  M  2 Apr/90  Port Hardy Ad.  3 y.  M  6 Mar/90  Port Hardy  F  Ad.  F  27 Jun/89 1  1  Ad.  Port Hardy  n/a  Ad.  14 Feb/93  Sointula  F  12 May/93  Port McNeil n/a  4  1.389  9  60.859  34.648  186.183  16.355  1.462  1.893  95.540  0.662  8.505  0.623  0.611  33  33  JOHNSTONE STRAIT  17.977  3.74  7.543  15.301  5.675  0.103  1.164  2.713  0.663  0  0.422  1.741  33  73.1  77.2  75.2  70.3  74.8  72.4  78.7  72.3  68.8  72.3  11.461  1.2  5.4  4.5  3.6  2.9  3.1  2.1  1.9  2.5  3.8  75.6  1.162  0.423  10.588  3.734  DDE  S.D.  Trauma/Inanit.  Drown.  Trauma/Electro.  Gunshot  Toxic:Pb  Gunshot  Drown.  Electro.  Gunshot  1.4  Undeter. trauma  73.0  83.7  66.6  70.4  0 2 % H  3.5  3.20  n/a  n/a  3.64  3.64  2.40  3.86  3.41  4.50  3.8  3.3  Electro.  Inanit  5.8  3.7  % Fat  Mean  1  1  2  3 y. Ad.  M M  12 Feb/90  Duncan  6 Mar/90  Ladysmith  3 0  n/a  14 May/93  Gabriola Isi.  Ad.  31 Jan/91  Gabriola Isl. M  M  n/a  Gabriola Isl.  2  n/a  M  n/a  Yellowpoint  4  3.64  2  Ad.  26 Apr/90  Nanoose  6.15  3.10  Electro.  n/a 5.45  Initial Etiology Elertro.  Wt(kg)  5  M  11 Mar.93  Union Bay  1  Ad.  1 y.  M  16 Mar/93  Nanaimo  5  F  Ad.  F  8 Feb/90  Nanaimo  3  BC  Ad.  Ad.  F  Age  F  Date  7 May/90  Sechelt  Location  Sex  Appendix 1-1, cont.  0.018  9  0.779  0.436  1.280  0.159  0.012  0.011  2.345  0.005  0.089  0.013  0.008  33  0.153  0.068  0.126  0.227  0.003  0.013  0.039  0.004  0.003  0.004  0.005  0.010  0.015  0.008  0.037  0.004  Oxychior.  0.099  9  2.736  1.761  7.416  1.304  0.110  0.097  6.181  0.033  0.603  0.066  0.043  33  0.566  0.271  0.414  0.417  0.006  0.102  0.152  0.029  0.007  0.054  0.032  0.115  0.080  0.026  0.409  0.166  t-Nonachl.  1.451  9  22.117  14.638  65.008  12.010  1.256  1.719  43.746  0.425  6.421  0.829  0.328  33  15.827  6.607  7.948  9.269  0.123  1.356  1.663  0.367  0.172  0.712  0.679  2.022  0.727  0.282  4.558  5.152  Total PCBs  Comnients  Hg-exp.  Pb-exp.  Pb-tox.  Pb-exp.  PB-exp.  Pb-exp.  Hg-exp/PB-exp.  BC n/a  -  -  11 May/90 -  4.77 Intra.Agg.  Body Condition: 0-emaciated, 1-thin, 2-fair, 3-good, 4-very good, 5-excellent ND not detected; detection limit = 0.0005 mg.kg wet wt not available  Smithers  NORTHERN INTERIOR 3 2.1  73.1  4  67.2  4  N Ad.  65.5 71.5 2.55  F  3.9 2.7  65.1  66.7  5.0  Veh.Coll.  Undeter. trauma  5.7  7.7  1.89  3.86  n/a  Electro.  Electro.  S.D.  4 y.  M  n/a  4.55  Mean  3 3  Ad.  M  29 Jul/90  Sandspit  4 Mar/91  Qu.Char.Isl.  5  n/a  F  n/a  Qu.Char.Isl.  5  2 y.  M  1 Feb/90  Ocean Falls  NORTH COAST  N  2  0 2 % H  2  % Fat  2.01  Electro.  Initial Etiology  1.30  4.00  Wt(kg)  S.D.  3  69.7  Ad.  M  3.8  Age  Sex  Mean  Date  21 Jan/91 67.7  Jokervile  5.1  Location  Appendix 1-1, cont.  0.42  4  0.877  1.428  0.452  2.639  1.871  0.750  2  0.324  1.065  0.740  DDE  0.011  4  0.006  0.015  0.008  0.012  0.025  0.015  2  0.005  0.014  0.009  Oxychlor.  0.067  4  0.022  0.083  0.049  0.079  0.109  0.093  2  0.022  0.076  0.054  t-Nonachl.  0.429  4  0.259  0.743  0.359  0.868  1.06 1  0.683  2  0.518  0.932  0.414  Total PCBs  Pb exposed  Pb-exp.  Pb-exp.  Comments  CHAPTER 2 BIOLOGICAL EFFECTS OF CHLORINATED HYDROCARBONS IN BALD EAGLE CHICKS  This study of embryotoxicily was designed to investigate whether in ovo exposure to PCDDs, PCDFs and PCBs was impacting hatching success and affecting a variety of biochemical and morphological parameters in Bald Eagle chicks. The aim of the study was also to estimate concentrations of PCDDs and PCDFs in Bald Eagle eggs which would be indicative of no-observed-efffects (NOEL) and lowest-observed-effect (LOEL) levels The results presented in this Chapter represent an extensive collaborative study with other laboratories. Contributions of those laboratories and of the principle investigators are identified in the Materials and Methods, while technical contributions are included in the Acknowledgements. The concept, study design, field work, statistical analyses, calculations, graphic representations and other manipulations of data were performed by me. A version of this chapter has been accepted for publication (Elliott et al. in press).  Materials and Methods Sample collection Bald Eagle eggs were collected from 20 nests (Figure 2.1). At three sites, Crofton (designated as location 3), Nanaimo (4,5), Powell River (6-9), sample nests were all within a 25 km radius of a kraft pulp and paper mill, and generally within the effluent impact zone of the mills, as defined by fisheries closures due to dioxin contamination (Harding and Pomeroy 1990). Eggs were collected from two nests in the Fraser River estuary (Map Nos. 1-2); at least 500 km downstream from where effluent is discharged into the Fraser River from four kraft pulp and paper mills. An area of the west coast of Vancouver Island, Clayoquot Sound (Map Nos. 10-14), was used as a reference site; there are no major industrial discharges to the  41  sound, although there is some fish processing and lumber yarding around Ucluelet Inlet. Further details on pollutant sources to Bald Eagles are discussed in Chapter 3.  Figure 2.1  -  Locations where Bald Eagle eggs were collected for artificial incubation. 42  Usually one egg was taken from each nest; the smallest egg in the clutch, presumably the second egg, was selected. At five nests in the Powell River area both eggs were taken. Because of the wide variability in nesting dates of Bald Eagles within and among areas, collecting at each site was scheduled for the estimated midpoint of incubation. The nests were accessed by a professional tree climber. Eggs were placed initially into a portable thermos. The temperature was maintained between 25 and  300  C using hotwater bottles, replenished as  required from thermos bottles. Within eight hours of collection, the eggs were transferred into a battery powered CurfewTM incubator kept at a temperature of 34°C. The eggs were rotated about hourly and turned on their long axis twice daily. Within 72 hours (normally within 24  -  48 hours) the eggs were brought to the  laboratory at the Department of Animal Science, University of British Columbia, where they were candled to determine fertility and placed into a Humidaire incubator maintained at 37.2°C with a relative humidity of 82-84 %. The eggs were rotated once per hour and turned twice a day in opposite directions on their long axis. At pipping the eggs were placed into a hatcher.  Sample preparation Within 24 hours of hatching, the birds were weighed, blood drawn by cardiac puncture using a heparinized syringe and the bird sacrificed by decapitation. The yolk sac was removed and frozen. The liver was removed, weighed and separated as follows: 0.25 g from tip of left lobe for Vitamin A analysis, 0.10 g from tip of the right lobe for porphyrin analysis; these samples were then frozen. The remaining liver was used to prepare microsomes. Various organs were removed and morphological measurements performed (Hart et al. 1991): body, yolk-free body, liver, heart, kidney (sum of both), yolk, stomach, intestine, bursa, adrenal (sum of both), spleen and tibia (wet, dry, ash) weights and tibia length The following tissues were fixed in 10 % buffered formalin for histological examination: right kidney, bursa, thymus, spleen, gonads, lung, heart, intestines, thyroid and adrenal glands. Tissues were processed routinely and embedded in paraffin blocks. Sections were cut at 6 urn and stained with 43  hematoxylin and eosin and examined by light microscopy. The amount of lymphoid tissue was estimated based on follicular size and cell density of cortex and medulla in the bursa, on the density of white pulp in the spleen and on the thickness of the cortex and cell density in the thymus. The number of mitoses in all lymphoid organs and the number of necrotic cells in the bursa and thymus were counted in five fields at 600 X magnification. The level of extramedullary haematopoiesis was assessed in the spleen.  Chemical analysis Bald Eagle yolk sacs were analyzed for PCDDs, PCDFs and non-ortho PCBs at the National Wildlife Research Centre, Hull, Quebec, in the laboratory of Dr. R.J. Norstrom. The analyses were carried out on a VG Autospec high resolution mass spectrometer linked to a HP 5890 Series II data system using ‘ C-labeled internal standards after gel permeation/carbon 3 chromatographic cleanup, essentially as described for livers in Chapter 1. Organochiorines and other PCBs were determined using GC/MSD (high resolution GC/low resolution MS) (Letcher et al. in press).  Biochemical assays Microsome preparation: Microsomes were prepared as described in Bellward et al. 1990. Briefly, livers were homogenized in 25 ml TRIS-KCL buffer using a teflon pestle; the homogenate was centrifuged at 10,000 g for 20 minutes, the precipitate discarded and the supernatant further centrifuged at 100,000 g for 60 minutes. The microsomal pellet was suspended in 20 ml of 10 mM EDTA (ethylenediamine tetraacetic acid), 1. 15% KCL, pH = 7.4, buffer at 4°C and homogenized; the homogenate was spun in an ultra-centrifuge as described above and the resulting microsomal pellet resuspended in 0.5 ml of 0.25 M sucrose. Aliquots of 100 ul were stored in cryovials in liquid nitrogen until assayed. Cytochrome P450-related activity: Ethoxyresorufin 0-deethylase and benzyloxyresorufin 0-deethylase activity in liver microsomes were determined using the method of Klotz et al.  44  (1984), adapted to a fluorescence multi-well plate reader. The standard reaction mixture for Bald Eagle microsomes contained 0.1 M TRIS-HC1, pH 8.0, containing 0.1 M NaC1, 10 mM of MgC1 , 2 uM 7-ethoxyresorufin or 1.5 uM 7-benzyloxyresorufin and approximately 200 g 2 of microsomal protein in a final volume of 500 uL. After a pre-incubation period of 5 minutes at 37°C, the reaction was initiated by the addition of NADPH (final concentration 0.6 mM) to the sample well (the blank did not receive NADPH). The reaction was stopped after 20 minutes by the addition of 1.0 ml of cold methanol. The amount of resorufin formed was measured in a fluorescence plate reader, using an excitation wavelength of 530 nm and an emission wavelength of 590 nm. Hepatic microsomal total protein was measured using a modification of Lowry’s method (Peterson 1977). Immunoblotting: Based on the original western blot method developed by Towbin et al. (1979), hepatic microsomal proteins were separated on sodium dodecyl sulfate polyacrylamide gels (SDS-PAGE, 9% acrylamide) and electrophoretically transferred to Rad-free membranes (Schleicher & Schuell, Keene, NH). Aroclor 1254-induced rat liver microsomes (prepared from commercially available postmitochondrial supernatant, Molecular Toxicology Inc., Annapolis, MD) were used as standards. Immunodetection of CYP1A was performed using monoclonal antibody 1-12-3 prepared against scup cytochrome P45O1A1 which recognizes CYP1A in all taxonomic groups of vertebrates examined so far (Park et al. 1986, Stegeman 1989). The secondary antibody was a goat anti-mouse IgG linked to alkaline phosphatase. Immnunoreactive proteins were detected by chemiluminescence (Rad-Free, Schleicher & Schuell, Keene, NH) and the light intensities of the inimunoreactive protein bands were quantified by video imaging densitometry (UVP Gel Documentation System 7500, San Gabriel, CA). This work was carried out in the laboratory of Dr. S.W. Kennedy. Cytoebrome P4502B (CYP2B) levels were determined by protein immunoblotting using rabbit polyclonal antibody 7-94 against scup P450B (a CYP2B like protein), which recognizes CYP2B proteins (Stegeman 1989). Methods were as described above, but with Bio-Rad goat anti-rabbit alkaline phosphatase-linked secondary antibody and using NBT (Nitro blue tetrazolium) and BCIP (5-bromo 4-chioro 3-indoyl phosphate) for colour development. 30 g 45  of samples were loaded in each well. Scup microsomes containing known amounts of P450B were included for quantitation in each gel. Since equivalence of cross-reactivity for the antibody between scup and eagle is unknown, numbers are relative and not absolute. Scup standards insure the linearity of response of the system and are necessary for normalizing between blots and runs. Analysis of developed blots was performed using a Kodak DCS 200 digital camera system and the NIH Image 1.55 densitometry software. This assay was performed in the laboratory of Dr. J.J. Stegeman. Liver vitamin A analysis: Samples of liver (300 to 500 mg) were dehydrated to a pink powder by grinding with anhydrous sodium sulphate. The internal standard, retinyl acetate (40 ng/20 uL methanol) was added to the equivalent of 0.20 g of liver and the vitamin A compounds were extracted with 10 mls of a 1:9 dichloromethane:methanol solvent mixture in an amber vial. After centrifugation (10 mins at 600 rpm at 10°C) the supernatant was filtered through a 0.2 urn Acrodisc LC13 PVDF filter (Gelman) and a 20 ul aliquot was analyzed in duplicate by non-aqueous reverse phase HPLC. Separation of retinol, retinyl acetate and retinyl palmitate was achieved with a 15 cm long, 5 urn ODS Zorbax column with 100 % methanol at 1 ml/min for 5.5 minutes followed by a linear gradient which brought the mobile phase to 30 % dichioromethane and 70 % methanol within 0.5 mm. This composition was held until the end of the run at a flow rate of 2.0 ml/min. With these conditions, retinol, retinyl acetate and retinyl palmitate had retention times of 3.1, 4.2, and 9.7 minutes, respectively. Plasma vitamin A analysis: The internal standard, retinyl acetate was added to 100 ul of serum. The retinol-protein complex was dissociated by the addition of 200 ul of acetonitrile. The retinol was extracted twice using 4 mIs and 1 ml of hexane. The organic and aqueous phases were separated by centrifugation, and the combined organic phases were evaporated to dryness under a stream of pure nitrogen. The residues were reconstituted in 1 ml of methanol, filtered through a 0.2 urn Acrodisc LC13 PVDF filter (Gelman) and a 50 uL aliquot was analyzed in duplicate by HPLC using the colunrn described above for liver. With 100 % methanol as the mobile phase and a flow rate of 1 ml/min, retinol and retinyl acetate had retention times of 3.3 and 4.5 mm, respectively. 46  Hepatic porphyrins: Porphyrin levels in liver were determined using the method of Kennedy and James (Kennedy and James 1993). This method involves extraction in duplicate using a mixture (1:1) iN hydrochloric acid/acetonitrile. The porphyrins were then concentrated on Sep-Pak Plus t C18 cartridges followed by separation and quantification by HPLC.  Statistical analysis The SYSTAT software package was used for statistical analyses of all data. Data are presented on a lipid weight basis as suggested by Hebert and Keenleyside (1995), when there are significant relationships between wet weight contaminant concentrations and percent lipid. For example, using only data from pulp mill sites (to minimize the influence of location), 2,3,7,8-TCDD concentrations (wet weight) in yolk sacs were highly significantly correlated with percent lipid (linear regression, r 2 =0.772, p < 0.0001, N  =  11). Chemical residue data  were transformed to common logarithms and geometric means and 95 % confidence intervals were calculated with the data grouped by collection site. Contaminant levels were compared among location with a one-way analysis of variance (ANOVA); significant differences were determined using Tukey’s multiple comparison procedure (MCP).  Data were also compared  on the basis of a pulp mill versus non-pulp mill grouping and significant differences identified using Student’s t-test. In order to avoid a bias, for comparison among sites and between pulp mill and non-pulp mill sites, only the results from the second or smallest egg were used from the Powell River nests, thus giving a total sample size of 14. Concentration-effect relationships were determined using coefficients of determination (r ) using least-squares linear regression. 2 Unless stated otherwise, a value of p < 0.05 was considered statistically significant in all analyses. TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalency factors proposed by Ahlborg et al. (1994), and referred to here as the WHO (World Health Organization) TEFs. For comparison, TEFs proposed by Safe (1990) and Kennedy et al. (in press) were also used.  47  Results  Chemical contaminant levels PCDDs and PCDFs. Data are presented on an individual nest basis in Table 2.1. The eight PCDDIPCDF congeners which exhibited significant differences among sites are grouped and compared in Figure 2.2. Congeners with a 2,3,7,8-substitution pattern were dominant; however, there were traces of l,2,3,4,6,7,9-HpCDD (5  -  10 ng/kg) in some yolk sacs from  Powell River, in both yolk sacs from the Fraser estuary and in the yolk sac from Nanaimo. Likewise, trace amounts of 1,2,4,6,7,8-HxCDF (5 ng/kg) and 1,2,3,4,6,8,9-HpCDF (ND  -  -  10 ng/kg), 1,2,4,6,8,9-HxCDF (10  -  100  150 ng/kg) had a similar geographical distribution.  Concentrations of 2,3,7,8-TCDD, 1,2,3,7,8-PnCDD, and 2,3,7,8-TCDF were highest in the yolk sacs from Powell River and were significantly higher than in yolk sacs from the Fraser Delta or west Vancouver Island. Those same major congeners were not statistically separable between Powell River and east Vancouver Island. Comparison between pulp mill (Powell River + east Vancouver Island) and non-pulp mill (Fraser Delta + West Vancouver Island) showed that concentrations were significantly higher (p < 0.005) at pulp mill sites for all the congeners in Figure 2.2, except 1,2,3,4,6,7, 8-HpCDD. Although not statistically different from other sites, highest concentrations of 1,2,3,4,6,7,8-HxCDD and OCDD (331 ng/kg) were in yolk sacs from the Fraser Delta.  48  TCDD  2500  2,000  1,500  1,000  500  1120 101.,  a  160  :::  140  12378-PCDF  CDD  I— S  120 100 80 ab  4-  .:  0) U)  4OabI  b  4-  U)  16700 200  117400 123678-HCDD  C) 8,00C  150  ab  0)  z  6OOC 100 4,00C 50 2,00C  cT 0 5800  1 234678HCDD  1996  01)1.,  400  a  300  200  10:  I, —‘s,  \C3•  4.  .*1  \•  Figure 2.2 - Residue levels of major PCDDs and PCDFs in yolk sacs of Bald Eagles collected from British Columbia in 1992. Vertical bars represent geometric means of two to five analyses per collection site along with the 95 % confidence interval. Means which do not share the same lower case letter were significantly different (p < 0.05).  49  ç C  4 t__  o —  a  CD  —.  e —  C’  o  .-  C•)  —  Cl)O  c_)  o  7  -  .9 OL?  -  — 0 0-  )  7 77  —  -  4)  7  0  7  ,  7  77  0  76?  ...  .  “‘-  v  0  -•.  .  :  .  .  .  .  .  .  : :  .  (11  -  .  .  .  :  .  .  . .  .  :  .  C)  %totalPCBs N) 0  .  .  .  .  :  .  .  .  : :  .  W  .  .  .  .  .  .  .  .  :  .  :  .  .  : :  :Q 00 D o  :u.  -  (71  g  —.  cj  (•) CD  o  I  0  •  0  C)  C)  .  0  o  CD  —  CD  )  CD  CD  CD  CD  0  -‘  C  0  — C)  .  CD  •  \ .1  0  •  0  ..  •  c  •  0’  CD  —.  -•  ..  CD  C  CD  CD CD  E.  CD  tj  Z  —  0  CD  CD  c-  t CD  C)  0 CD  5  C.’ ..  CD  Z  S  •  C)  CD  -e  ..  —• -  ..  .  z  .  i  CD  —•  CD  •  ‘.‘  .  4  .  V  CD  CD  —.  CD  CD  >  vr  j’-  ..  8020 9630 12800 504 394 567  3630 2400 3670 409 450 675 1070  2450 1470 2250 218 306 353 629 323  22 23 12 23 13 16 20 16  Scuttle Bay  Limekiln Bay  Limekiln Bay  Bawden Bay  White Pine 1  White Pine 4  Thornton Cr.  Mercantile Creek  8  9  9  10  11  12  13  14  *  8670  3840  2700  18  Scuttle Bay  8  ND  18.1  ND  272  150  189  170  261  35.2  49.3  ND  15.9  16.2  68.6  35.4  60.9  137  66.5  46.9  37.7  83.7  49.2  76.7  79.8  65.7 64.6  79.6  60.4  69.2  83.2  83.4  187  575  1076  102  OCDD  127  91.4  62.9  1-2 Fraser Delta; 3-5 East Vancouver Island; 6-9 Powell River; 10-14 West Vancouver Island  11.9  10900  3850  3560  18  Evenden Point  7  235  139  67.3  140  534  8390  2630  2670  14  Evenden Point  7  395  4230  1130  1120  13  Ball Pt.  6  93.2  149  19.1  6020  1170  1130  8.8  Ball Pt.  6  256  430  1023  144  1234678 HpCDD  98.5  152  1150  9380  3950  2460  10  Winchelsea Is.  5  4050  2030  1450  11  Jack Point  4  8400  2510  1800  11  Crofton  3  88.4  3200  1010  646  14  River Rd.  2  39.4  2400  1340  868  123789 HxCDD  123678 HxCDD  12378 PnCDD  11  2378 TCDD  % fat  Brunswick Pt.  Location  360  1070  465  305  672  7970  6900  11590  16700  24100  18800  1880  1670  2190  2930  923  238  661  2378TCDF  13.8  52.5  50.3  17.6  13.5  49.4  82.0  32.1  15.2  80.8  324  101  60.4  62.8  99.2  30.0  37.4  43.8  12478PnCDF  19.5  52.5  34.2  23.4  22.9  178  114  212  217  424  305  53.6  39.9  93.2  64.2  47.6  26.6  48.3  12378PnCDF  91.0  207  130  86.2  88.9  953  651  933  1040  1220  823  296  339  820  267  375  170  348  23478 PnCDF  169  264  47.7  141  39.8  21.8  17.5  145  135  153  153  280  306  100  124  163  316  4.39  69.9  ND  5.86  4.04  42.5  82.0  26.3  16.9  31.5  278  103  62.7  52.0  342  59.1  24.2  100  130  Total HpCDF  Total HxCDF  PCDD and PCDF concentrations (nglkg, lipid weight basis) in yolk sacs of Bald Eagle chicks collected in 1992 from British Columbia.  1  Map No*.  Table 2.1  Table 2.2  Concentrations of non-ortho PCB congeners, geometric mean and 95% confidence interval (range in brackets), in yolk sacs of Bald Eagle chicks collected in 1992 from British Columbia. PCB congener, (Lg/kg, lipid weight basis)  Location  N  Fraser Delta  2  12.6 ±1.4  66.9 ±0.42  3.23 0.77-13.5 (2.31-4.50)  4.79 1.6-14.3 (3.71-6.17)  26.9 23-31.5 (26-27.9)  40.0 9.17-175 (28.4-56.4)  5.60 3.85-8.15 (5.14-6.12)  East Vancouver Island  3  10.7 ±0.48  60.1 ±2.24  0.63 0.18-2.17 (0.32-1.23)  3.00 2.55-3.47 (2.74-3.25)  19.4 17.8-21.2 (18.5-20.3)  40.9 26.1-64.3 (33.6-53.8)  7.63 3.61-16.1 (5.14-11.6)  Powell River  8  16.0 ±4.2  64.6 ±4.9  1.23 0.61-2.6 (0.51-5.54)  3.56 2.45-5.18 (1.84-6.77)  33.4 23.1-48.2 (21.3-73.8)  50.0 37.3-56.6 (34.5-72.9)  8.90 7.14-11 (6.15-13.4)  West Vancouver Island  5  17.6 ±3.9  63.2 ±2.5  1.04 0.74-1.44 (0.32-5.54)  3.27 1.97-5.43 (1.64-5.10)  38.1 24.5-59.2 (23.3-58.5)  36.2 20.6-63.8 (18.7-71.6)  5.84 3.22-10.6 (3.24-11.8)  Lipid Moisture % % (mean ± SD)  #37 (34-4)  #77 (34-34)  #81 (345-4)  #126 (345-34)  #169 (345-345)  Organochiorines. Total PCBs, DDT-related and other major organochiorines detected in Bald Eagle yolk sacs are presented in Table 2.3. As with the PCB congeners, no significant differences in mean concentrations occurred among sites for any of the organochiorine compounds. The pattern was relatively consistent among yolk sacs with total PCBs related  >  chiordane-related  >  dieldrin  >  B-HCH  >  HCB  >  >  DDT  mirex. The exception to this  pattern was the yolk sac from White Pine Cove No. 1, where DDE levels were greater than total PCBs. The PCB/DDE ratio was generally much lower in yolk sacs from the west coast of Vancouver Island than from other sites.  Artificial hatching success and condition of embryos A total of 25 Bald Eagle eggs were collected for incubation, of which one was cool to the touch at the time of collection (there was a recently hatched chick in the nest) while a  52  3  8  5  East Vancouver Island  Powell River  West Vancouver Island 192 70.5-524 (63.9-488)  400 211-760 (117-1052)  125 63.1-247 (46.6-206)  93.9 65.1-136 (51.2-194)  127 83.3-194 (97.5-147)  73.5 20.3-267 (54.5-99.2)  364 113-1160 (278-477)  2  Fraser Delta  559 450-694 (490-615)  DDE  Total PCBs  1.13 0.79-1.62 (0.62-2.37)  0.89 0.58-1.36 (0.48-1.24)  9.77 5.4-17.7 (4.7-15)  1.12 0.72-1.74 (0.85-1.29)  0.91 0.12-7.01 (0.56-1.46)  oxychlordane  13.5 9.44-19.2 (6.47-22)  9.46 5.86-15.3 (7.56-12.6)  3.82 1.54-9.45 (3.09-4.71)  iransnonachior  0.78 0.64-0.96 (0.75-0.82)  0.99 0.64-1.52 (0.61-1.54)  1.15 0.93-1.42 (1.07-1.31)  1.69 1.25-2.30 (1.58-1.82)  heptachiorepoxide  2.1 1.44-3.08 (1.92-2.30)  1.59 0.60-4.22 (0.35-2.68)  2.01 1.48-2.73 (1.74-2.41)  2.30 0.53-9.96 (1.64-3.23)  Dieldrin Mirex  0.66 0.42-1.02 (0.44-1.08)  0.68 0.49-0.95 (0.47-1.20)  0.63 0.33-1.19 (0.42-0.71)  0.45 0.10-2.0 (0.32-0.64)  Organoclorine centration (mg/kg, lipid weight basis)  1.19 0.67-2.12 (0.60-1.89)  1.31 0.89-1.92 (0.80-2.02)  1.91 0.96-3.83 (1.53-2.96)  0.68 0.22-2.09 (0.52-0.88)  B-HCH  0.75 0.53-1.07 (0.57-1.24)  0.70 0.43-1.16 (0.25-1.15)  1.13 0.46-2.78 (0.64-1.62)  0.62 0.35-1.10 (0.55-0.71)  HCB  Organochiorine concentrations, geometric means and 95% confidence intervals, (range in brackets) in yolk sacs of Bald Eagle chicks collected in 1992 from British Columbia.  N  Location  Table 2.3  second egg was possibly shaken as it was lowered from the nest. Therefore, 23 of the eggs were possibly viable when placed into the incubator. A total of 18 eggs hatched for an overall success rate for artificial incubation of 78.3 %. Eliminating the possibly shaken egg from Northwest Bay, 16 eggs were collected from pulp mill sites of which 11 hatched for a hatching rate of 69 % (Table 2.4). Of eight eggs collected from non-pulp mill sites, seven hatched for a hatching rate of 88 %. This difference in hatching success between pulp mill and non-pulp mill sites was not, however, significant (Chi 2 test). One chick (Ball Point A) was edematous at hatching. Of the eggs which failed to hatch, one was infertile (Powell River area), two were addled (both from the same nest in the Powell River area), three were early (first quarter of development) embryos (one each from east Vancouver Island, Powell River and west Vancouver Island) and one was a late (last quarter of development) embryo (Alberni Inlet).  Table 2.4  Outcome of artificial incubation of Bald Eagle eggs collected from British Columbia, 1992.  Location  Treatment  Fraser Delta West Vancouver Island  No. collected  No. hatched  Non-pulp mill  2  2  100  Non-pulp mill  6  5  83  8  7  88  (Mean, non-pulp mill) East Vancouver Island  Pulp mill  3b  3  100  Powell River  Pulp mill  12  8  67  Alberni Inlet  Pulp mill  1  0  0  16  11  69  (Mean, pulp mill) a b  % success  2 pulp mill versus non-pulp mill difference not significant, chi 4 eggs were collected, 1 was eliminated as possibly shaken  54  =  1.402  Morphological and histological measurements  No significant differences occurred among sites for mean values of any of the measured morphological parameters, whether expressed as actual values or as percent yolk-free body weight. For the 18 chicks measured morphological measurements (mean ± SD) were as follows: body weight (88 ± 9.4 g), yolk-free body weight (83 ± 8.3 g), liver (1.9 ± 0.29 g), right kidney (0.70±0.12 g), intestine (2.1±0.27 g), heart (0.56±0.09 g), adrenal glands (0.04±0.02 g), spleen (0.077±0.025 g), bursa (0. 152±0.039 g), yolk (5.7±2.3 g), thyroid glands (0.075±0.024 g), dry tibia weight (0.057±0.005 g), tibia length (26.8 ±0.85 mm), tarsus length (20.5±1.48 mm), wing chord (29.4 ± 2.0 mm). Selected parameters are compared among sites in Appendix 2.1. For the tissues examined histologically, variations among individual birds were seen only for lymphoid organs (Table 2.5). Variations were observed within and between sites in amount of lymphoid tissue, the number of cells in mitosis, the number of necrotic cells and the degree of extramedullary hematopoiesis. However, no significant differences among sites occurred for mean values of any of the measured parameters. The amount of lymphoid tissue in the spleen was constant among individual birds. Table 2.5  Bursa  Spleen  Histological examination of immune system tissues in Bald Eagle chicks (Mean ± SD). Fraser Delta (N=2)  East Van. Isl. (N=3)  Powell River (N=8)  West Van. Isi. (N=5)  3.0 ± 0.0  3.0 ± 0.0  3.0 ± 1.1  1.8 ± 0.84  No. necrotic cells”  90 ± 28  109 ± 28  142 ± 85  105 ± 30  No. cells in mitosisb  29 ± 11  43 ± 5.5  50 ± 3.1  40 ± 14  No. cells in mitosisb  15 ± 9.2  16 ± 7.5  19 ± 6.7  6 ± 3.9  1.5 ± 0.71  1.3 ± 0.58  2.2 ± 0.64  1.4 ± 0.55  Amount of lymphoid tissue’  3.0 ± 0  3.7 ± 0.58  2.9 ± 3.8  2.0 ± 0.71  No. necrotic cells’  28 ± 3.5  54 ± 21  66 ± 19  64 ± 18  No. cells in mitosisb  10 ± 8.5  24 ± 22  11 ± 5.4  21 ± 8  Amount of lymphoid tissuea  Degree of E.M.C Thymus  a -  b -  -  d -  based on follicular size and cell density of cortex and medulla. The amount varied from small (1) to large (4). per 5 fields at 600x. e.m. extramedullary hematopoiesis, based on the amount of hematopoietic tissue. Amount varied from small (1) to large (3). based on the thickness of the cortex and cell density. The amount varied from small (1) to large (4). -  55  Biochemical measurements  Mean concentrations of CYP1A were sixfold greater (p <0.05) in chicks from Powell River compared to west Vancouver Island (Table 2.6). Mean concentrations of a CYP2B-like protein were two to three-fold higher in livers from Strait of Georgia sites compared to west Vancouver Island; however, the differences were not significant. Mean EROD activity was eight-fold higher in east Vancouver Island compared to Fraser delta and mean BROD activity was nearly nine-fold higher in Powell River than Fraser delta chicks; however, the differences were not significant, likely in part due to small sample sizes and large variabilities. However, both hepatic EROD and BROD were significantly induced, if datafor all chicks collected near pulp mills were pooled compared to non-pulp mills sites (p <0.0005 and p < 0.02, respectively). Mean uroporphyrin and Vitamin A levels did not differ significantly among sites, although liver retinyl palmitate levels were about one-half in chicks from the Fraser delta compared to west Vancouver Island.  Table 2.6  Measurement of hepatic cytochrome P450 and porphyrin parameters and vitamin A in plasma and liver of Bald Eagle chicks collected in 1992 from British Columbia (Mean ± SD). Fraser Delta (N = 2)  East Vancouver Is. (N = 3)  Powell River (N = 8)  West Vancouver Is. (N = 5)  (± 35)  25 (± 12)  NA  48 (± 30)  36 (± 34)  18 (± 13)  EROD (pmol/min/mg protein)  1.2 (± 0.92)  9.3 (± 4.6)  9.0 (± 5.4)  1.8 (± 1.8)  BROD (pmol/min/mg protein)  6.6 (± 0)  35 (± 14)  56 (± 27)  25 (± 24)  Uroporphyrins (pmol/g)  10 (± 1.4)  8.0 (± 0)  12 (± 3.8)  8.2 (± 1.5)  Retinol-plasma (g/1)  320 (± 2)  315 (± 76)  350 (± 76)  380 (± 93)  0.65 (± 0.07)  0.60 (± 0.15)  0.65 (0.13)  0.67 (0.12)  19 (± 6.9)  28 (± 7.3)  29 (± 8.4)  37 (± 13)  CYP1A (std. vol. equiv. [id])  NA  CYP2B equivalents (pmol/mg)  Retinol-liver (gIg) Retinyl palmitate-liver (gIg) a,b -  NA  a,b 15  means that do not share the same superscript are significantly different among sites. not assayed.  -  56  b 44  (± 2.3)  Concentration-effect relationships Data from the complete set of 18 Bald Eagle chicks were used to examine relationships between measured biological parameters and contaminant exposure. The gradient of exposure from lowest to highest was 16-fold for 2,3,7,8-TCDD and 80-fold for 2,3,7,8-TCDF. Regression analysis was performed using both normal and log-transformed chemical residue data; results are presented in Table 2.7 for each parameter based on which form of the residue data gave the best fit (greatest r 2 value) to the regression curve. Highly significant positive regressions were found between hepatic CYP1A and most of the individual PCDD, PCDF and PCB compounds in yolk sacs; however, the best fits were with log 2,3,7,8-TCDF and 2,3,7,8-TCDD (Table 2.7, Figure 2.4). No significant regressions were found between a CYP2B-like protein and yolk sac concentrations of any of the chemical parameters measured. For EROD, the best r 2 value was with 2,3,7,8-TCDD, while the strongest regression for BROD was found with log 2,3,7,8-TCDF. Hepatic urophorphyrin also showed a significant positive regression on 2,3,7,8-TCDD, log-2,3,7,8-TCDF and log-TEQs. Hepatic retinyl-palmitate levels showed a weakly significant positive regression with log-PCB 126, but not with any other chemical parameters. The hepatic cytochrome P450 and porphyrin parameters all regressed more strongly with either 2,3,7, 8-TCDD or log 2,3,7, 8-TCDF than with TCDD-TEQ5 estimated using three different TEFs (Table 2.8). Among the morphological parameters, a weakly significant positive regression was found between yolk-free body weight and log PCB 126. Yolk sac weight negatively regressed with both total PCBs and log TEQs. A weakly significant positive regression was also 2 determined for density of thymic lymphoid tissue with log 2,3,7,8-TCDD (r 0.02) and log TEQ5WHO (Table 2.7).  57  =  0.320, p <  (A)  r2  =  0.748  *  E E 0  2  0  0 uJ  500  1,000  1 500  2,000  2,500  3.000  3,500  3,000  3,500  2378-TCDD (nglkg, lipid basis) 40  (B)  0  500  1,000  1 .500  2,000  2,500  2378-TCDD (nglkg, lipid basis) ()1  r2  =  0.721 *  0)  E C  2 0  2 0  1,000 I  10,000 I  I  I  2,3,7,8-TCDF (nglkg, lipid basis)  Figure 2.4 Exposure-response relationships between 2378-TCDD or log 2378-TCDF concentrations in yolk sacs of Bald Eagles and hepatic (A) EROD activity (B) CYP1A concentrations and (C) BROD activity. -  58  ci,  (+)  (+) (+)  18  13  17  18  EROD  BROD  Uroporphyrins  Retinylpalmitate liver  18  Thymic lymphoid tissue NS  NS  0.034 0.191  NS  NS  0.032  0.059  <0.02  <0.002  0.601 0.316  0.708  <0.0005  0.748  0.091  0.045  0.042  0.028  0.298  0.721  0.114  NS  0.082  TEQs, according to Ahlborg (1994) NS - not significant  1  (-)  18  Yolk sac (+)  (+)  18  Yolk-free body weight  (+)  (+)  14  0.887  <0.0001  0.850  (+)  CYP2B  2 r  p  2 r  NS  NS  NS  NS  <0.02  <0.0005  <0.0005  NS  <0.0001  p  Log 2,3,7,8-TCDF  Slope  14  N  2,3,7,8-TCDD  0.058  0.098  0.247  0.26  0.194  0.396  0.297  0.057  0.371  2 r  NS  NS  <0.04  <0.03  NS  <0.03  <0.02  NS  <0.03  p  Log PCB126  0.09  0.269  0.023  0.097  0.122  0.346  0.588  0.255  0.576  2 r  NS  <0.03  NS  NS  NS  <0.006  <0.0001  NS  <0.002  p  Total PCBs  0.250  0.128  0.072  0.202  0.232  0.549  0.633  0.136  0.728  2 r  <0.04  NS  NS  NS  <0.05  <0.004  <0.0005  NS  <0.0005  p  Log TEQswHO 1  Concentration-effect relationships between biochemical and morphological measurement with chlorinated hydrocarbon levels in yolk sacs of Bald Eagle chicks.  CYP1A  Parameter  Table 2.7  Table 2.8  Comparison of regression (r ) values of some hepatic biochemical parameters on 2 TEQs derived from three sets of toxic equivalence factors (TEFs). Toxic Equivalent Factors  Parameter  TCDD/F’  2 Safe  3 CEH  4 WHO  P450 1A  0.887  0.687  0.759  0.805  EROD  0.748  0.529  0.607  0.633  BROD  0.601  0.427  0.515  0.549  Uroporphyrin  0.316  0.107  0.162  0.232  ‘Best r 2 value (either 2,3,7,8-TCDD or 2,3,7,8-TCDF) 2 S afe (1990) Chick embryo hepatocyte (S. Kennedy, person. comm.) 3 Ahlborg et at. 1994 4  Discussion Bald Eagle chicks collected from nests near pulp mills were exposed to elevated concentrations of potent embryotoxic PCDD and PCDF congeners, compared to chicks from reference nests. Symptoms of TCDD-like exposure, such as have been observed in field studies of fish-eating birds (Hoffman et at. 1986; 1987; Kubiak et at. 1989; Bosveld et at. 1994; Van den Berg et at. 1994b; Elliott et at. 1989a; Bellward et at. 1990; Hart et at. 1991; Sanderson et at. 1994a; Whitehead et at. 1992b), were not found in Bald Eagle chicks. Laboratory hatching success did not differ between eggs from pulp mill versus reference sites. However, hepatic CYP1A levels were significantly higher in eagle chicks from pulp mill sites and regressed positively on yolk sac concentrations of 2,3,7,8-TCDD and 2,3,7,8-TCDF. Induction of CYP1A can be linked primarily to PCDDs and PCDFs acquired by the female parent from local sources, as breeding Bald Eagles on the Pacific coast are year round residents (Hancock 1964). Yolk sacs contained high concentrations of the toxic non-ortho PCBs, 126 and 77, although regressions with biochemical and morphological parameters were weak and inconsistent compared to TCDD and TCDF. Concentrations of total PCBs and other organochiorines in eagle yolk sacs also varied little among sites.  60  Laboratory hatching success Except for one edematous chick, no signs were apparent in either the hatched eaglets or in failed eggs of GLEMEDS (Great Lakes embryo mortality, edema, and deformities syndrome) (Gilbertson et at. 1991), such as reported for fish-eating birds in the Great Lakes and elsewhere (Hoffman et at. 1986; 1987; Kubiak et at. 1989; Bosveld et a!. 1994; Van den Berg et at. 1994b; Elliott et at. 1989a; Bellward et at. 1990; Hart et a!. 1991; Sanderson et a!. 1994a; 1994b; Whitehead et a!. 1992b; White and Seginak 1994), which is similar to the toxic syndrome caused by TCDD in chicken embryos. In embryos of other avian species, such as ring-necked pheasants (Phasianus cotchicus), mortality is the most sensitive response to TCDD exposure and the symptoms seen in chickens at lower doses are not observed (Nosek et at. 1993). However, there were no significant differences in laboratory hatching success of eagle eggs among sites or between pulp mill and non-pulp mill areas. The overall artificial hatching success of 78.3 % was comparable to the average of 75 % (range 62  -  87 %) reported for wild  and captive Bald Eagles from a number of studies (Stalmaster 1987). The absence of deformities and other GLEMEDS symptoms in Bald Eagle chicks from this study is likely doserelated; some eagle chicks with deformed bills have been found in the Great Lakes basin (Bowennan et at. 1994), where at least some addled Bald Eagle eggs had much higher total PCB levels than any of the fresh eggs from the Strait of Georgia.  Patterns and trends of PCDD, PCDF and PCB contaminants in yotk sacs Local pulp mill and chiorophenol inputs account for the particular pattern and elevated levels of 2,3,7,8-substituted PCDDs and PCDFs in Bald Eagles and other wildlife from the Strait of Georgia (Elliott et at. 1989a; Whitehead et at. 1990; 1992b), compared to similar samples from other North American and European sites (Van den Berg et at. 1994b; Yamashita et at. 1993; Hebert et at. 1994). In particular, Bald Eagle yolk sacs contained high concentrations of 2,3,7, 8-TCDF, which is reported elsewhere at only nominal levels in wildlife samples. High TCDF levels such as in the eagle yolk sacs from Powell River reflect exposure to prey items contaminated by local pulp mill discharges (Harding and Pomeroy 1990). 61  Elevated TCDF levels have also been reported in tissues of common mergansers (Mergus merganser) and herring gulls breeding near a bleach kraft pulp mill in Quebec (Champoux 1993). Assuming that 2,3,7,8-TCDF should be cleared quickly from the body (Braune et al. 1989; Norstrom et al. 1976), the presence of this chemical in eggs likely results, therefore, from recent exposure and direct yolk deposition of contaminated lipids as suggested previously for herons (Elliott et al. 1 989a). Accumulation of TCDF in eagle tissues is probably not linked to the low absolute EROD activity found in Bald Eagle chicks (Table 6); a recent study compared EROD induction with in vitro capability to metabolize PCB 77, and concluded that low EROD activity does not reflect reduced capability to metabolize typical CYP1A substrates, such as PCB 77 or 2,3,7,8-TCDF (Murk et al. 1994). Recent exposure and direct shunting of dietary lipids to the yolk may also explain the presence of non-2,3,7,8 substituted PCDDs and PCDFs in eagle yolk sacs. Fish are able to metabolize most compounds of this type (Sjim et al. 1989), leading to low levels in the diet of fish-eating species; birds are also likely capable of further metabolizing them. The presence of elevated levels of 1,2,3,4,6,7,8-HpCDD and OCDD in the yolk sac from River Road in the Fraser River delta is consistent with reports of high concentrations of those contaminants in sediments from near the nest site (Tuominen and Sekela 1992). Elevated levels of higher chlorinated dioxins in Fraser estuarine sediments are indicative of the intensive past use of chlorophenol wood preservatives at industrial sites in the Fraser delta (Drinnen et al. 1991). In contrast to the well-defined local point sources of PCDDs and PCDFs, the uniformity among sites in concentrations of PCBs and other organochiorines in eagle yolk sacs reflects the importance of diffuse atmospheric inputs for those compounds (Elliott et al. 1989b). The geographically uniform PCB congener pattern contrasts with the finding of significant differences in the percent contribution of certain congeners in great blue herons between Crofton and Vancouver in 1987 (Elliott et al. 1989a). Because of their restricted seasonal movement and diet, herons appear to be better indicators of local PCB contamination than eagles.  62  Biochemical responses The results of this study confirm for another avian wildlife species the value of CYP1A induction, particularly as measured by western blotting, as a sensitive marker of exposure to TCDD-like compounds. Absolute EROD activities in these embryonic Bald Eagle microsomes were low, although the overall degree of induction from lowest to highest exposure groups, from six to eight fold, was the same as that observed for other species such as cormorants and herons (Sanderson et at. l994a; Whitehead et al. 1992b). Interspecific variation of this type is not surprising as there is increasing evidence that cytochrome P450 isoforms vary substantially even among closely related species (Yamashita et at. 1992). Absolute BROD activity was about five-fold higher than EROD in livers of Bald Eagle chicks, while differences in rates from least to most contaminated individuals was similar for the two activities. As with EROD, the best r 2 values were found between BROD and 2,3,7,8TCDF or 2,3,7,8-TCDD. BROD is considered a relatively specific marker of CYP2B1 activity in phenobarbital-induced rats (Burke et at. 1994). However, Rattner et al. (1993) recently reported that, while phenobarbital treatment of black-crowned night-heron embryos caused a 2,000-fold increase in a CYP2B-like protein, there was only a threefold increase in BROD activity. In contrast, 3-methylcholanthrene treatment increased BROD six to fourteen-fold. Based on that work and other recent reports (Yamashita et at. 1992), isoforms cross-reactive with putative fish CYP2B and rat CYP2B are present in at least some groups of birds, but the substrate specificities may be quite different. The results suggest the presence of a CYP2B isoform in Bald Eagles. Although Bosveld and Van den Berg (1994) in a recent review concluded that there is no evidence of chlorinated hydrocarbon-inducible CYP2B isoforms in birds which cross-react with mammalian CYP2B antibodies, further experiments using purified CYP enzymes and antibodies are required for a better understanding of substrate specificities. Uroporphyrin levels in chicks from the various sites were similar. Although there was a significant concentration-effect relationship between uroporphyrin levels and both 2,3,7,8TCDD and -TCDF, this finding must be treated cautiously as normal uroporphyrin levels in avian livers range from 5-25 pmol/g (Fox et al. 1988). PCBs have been reported to cause 63  accumulation of porphyrins in chick embryo hepatic cell cultures (Kennedy et al. 1995) and in liver and other tissues of adult birds of common laboratory species (Elliott et al. 1990), but not apparently in captive predatory birds (Elliott et al. 1991). In previous field studies, hepatic porphyrins were elevated in adult herring gulls from more polluted areas of the Great Lakes (Fox et al. 1988), but not in great blue heron embryos exposed to elevated PCDDs and PCDFs (Beliward et al. 1990). Plasma and liver retinoid concentrations and the molar ratios of retinol to retinyl palmitate did not differ among sites, although a weakly significant positive relationship between hepatic retinyl-palmitate and PCB 126 (34-345) was found. In contrast, laboratory data for rats report that PCDDs, PCDFs and PCBs caused depletion of liver retinoid stores (Chen et al. 1992). In field studies, such as with herring gulls in the Great Lakes, yolk retinoids varied among colonies and the molar ratio of retinol to retinyl palmitate correlated positively with TEQs in eggs (Spear et al. 1990). Van den Berg et al also reported a non-significant reduction in hepatic retinyl palmitate in cormorants from a contaminated relative to a reference site in the Netherlands.  Morphological and histological parameters Morphological and histological measurements did not differ among sites. However, as with retinyl-palmitate and PCB 126, a number of weakly significant exposure-response relationships occurred, at the p < 0.05 level (Table 2.8), which are likely not meaningful biologically. For example, yolk-free body weight appeared to increase with PCB 126 levels in yolk sacs; this contrasts to data from a number of field studies which report statistically significant negative relationships between PCDDs/PCDFs or PCBs and embryonic weight and other morphological characteristics (Hoffman et al. 1986; 1987; Van den Berg et al. 1994b; Hart et al. 1991; Sanderson et al. 1994a). The negative relationship between PCBs and yolk weight is consistent with similar findings for cormorants from British Columbia (Sanderson et al. 1994b), but contrasts with reports of a positive relationship for cormorants from the Netherlands (Van den Berg et al. l994b). The positive relationship between density of thymic 64  lymphoid tissue and log 2,3,7,8-TCDD is in contrast to reports from a number of laboratory studies that TCDD and related compounds cause atrophy of the thymus with depletion of lymphocytes (Elliott et al. 1990; Nikolaides et al. 1988). Therefore, it is likely that these findings in Bald Eagles are spurious in nature due in part to the relatively small sample size and large number of variables analyzed.  Comparison of toxic equivalents As reported previously for great blue heron chicks in the Strait of Georgia (Bellward et al. 1990; Sanderson et al. 1994a), regression of the biochemical endpoints against 2,3,7,8TCDD or 2,3,7, 8-TCDF produced the best coefficients of determination (r ). This contrasts to 2 data for other avian species and locations, where non-ortho and mono-ortho PCBs or TEQs, commonly using Safe’s (1990) TEFs, provided the best statistical fit to CYP1A parameters (Van den Berg et al. 1994b; Sanderson et al. 1994a; Rattner et al. 1993). However, exposure to PCDDs and PCDFs relative to PCBs was low in all of those studies, whereas the reverse was true for Bald Eagles. Data on fish-eating birds in the Great Lakes region (Kubiak et al. 1989; Yamashita et al. 1993) and in the Rhine estuary of The Netherlands (Van de Berg et al. 1994b, Bosveld and Van den Berg 1994) indicated that PCB congeners, in particular PCB 126 (34-345) and PCB 118 (245-34), were the major contributors to TCDD-like toxicity. The relative contribution of the major Ah-receptor active congeners in yolk sacs of Bald Eagles is compared among sites and to common terms from the Netherlands in Figure 2.5. Total TEQ5WHO and the pattern of contributors was similar between Bald Eagles from west Vancouver Island and the common terms; however, PCDDs and PCDFs made a much greater contribution in the Bald Eagles from the Strait of Georgia and the Fraser Delta. Further comparison of avian laboratory data on relative toxicity of PCBs to PCDDs and PCDFs suggests that TEFs derived from mammals such as Safe’s (1990) TEF’s tend to overestimate the toxicity of the both the mono-ortho and non-ortho PCBs, in all avian species with the possible exception of the chicken ((Brunstrom 1990; Brunstrom and Anderson 1988;  65  14 0 0 0  Li mono-o-PCBs 1  12  -.  x Cl) Cl)  LJ non-o-PCBs PCDFs  10  (U  El other-PCDDs  1  •TCDD 8 6 .  0 Cl)  C  w  4 2  I0 \  \  I —  Common Tern  Bald Eagle  Figure 2.5 The contribution of various chlorinated hydrocarbon groups to the sum of TCDD toxic equivalents (TEQ) in Bald Eagle yolksacs from coastal British Columbia, 1992 (N values and variances are in the tables), compared to values for common terns from the Netherlands. Toxic equivalents factors for PCDDs/PCDFs from Safe (1990) and for PCBs from Ahlborg et al. [1994]. -  Kennedy et al. 1994; Bosveld et al. 1992). In Table 2.8, three sets of TEFs were compared; biochemical parameters in Bald Eagle livers were regressed against yolk sac concentrations of either TCDD or TCDF and TEQs using the different TEFs. The WHO-TEFs, which give lower weighting to the mono-ortho PCBs, produced r 2 values which were closest to those determined using the individual contaminants. These results suggest that in Bald Eagle chicks, PCBs are relatively less toxic than TCDD for the endpoints measured. A number of fish-eating bird studies concluded that embryonic CYP1A induction is a sensitive biomarker for other deleterious Ah-receptor mediated responses (Hoffman et al. 1987; 66  Bosveld and Van den Berg 1994; Bosveld et al. 1994; Sanderson et al. 1994a; Rattner et a!. 1994). In Bald Eagle chicks from west Vancouver Island, low EROD activity and low levels of the CYP1A cross-reactive protein indicate background exposure to TCDD-like compounds. On a lipid weight basis, TEQ5WHO in yolk sacs were about 6,000 ng/kg. Converting this result to a whole egg, wet weight basis, (dividing by a mean factor of 60, based on comparison for Bald Eagles of a yolk sac and whole egg analyzed from the same nest), mean  TEQ5PCDD,PCDFS  were about 15 ng/kg in west Vancouver Island eggs. If we include the PCB contribution, TEQsWHO in the west Vancouver Island reference area were about 100 ng/kg. This is a suggested no-observed-effect-level (NOEL) in Bald Eagle eggs, using CYP1A as a marker. Likewise, levels of the CYP1A cross-reactive protein were significantly higher at Powell River, where mean TEQ5WHO in yolk sacs, on a lipid weight basis, were about 12,600 ng/kg, or about 210 ng/kg, on a wet weight basis in the whole egg. This is suggested as a lowest observed-effect-level (LOEL). In conclusion, Bald Eagle chicks collected near pulp mills were exposed to elevated concentrations of PCDDs and PCDFs which correlated with induction of a hepatic CYP1A cross-reactive protein. Levels of PCBs and other organochiorines did not vary among sites and were less important in the CYP1A induction.  67  Acknowledgements Many people contributed their time to the success of this project. I would especially thank I. Moul and G. Compton for assistance in the field. C. Kuehier suggested the incubation conditions. M.S. Bhatti and A. Roble assisted with dissecting and initial processing of embryos. Dr. H. Philibert undertook the histology at the University of Saskatchewan, Western College of Veterinary Medicine. M. Simon, M. Mulvihill and A. Idrissi performed the chemical analysis. W. Ko prepared the microsomes. F. Maisonneurve, G. Sans-Cartier and K. Williams are thanked for their technical assistance with the biochemistry, which was performed in the laboratory of S. Trudeau (NWRC). A. Lorenzen performed the CYP1A assay. B. Woodin performed the CYP2B assay. J. Smith provided advice on the statistics. S. Bucknell typed the tables and P. Whitehead assisted with drafting figures. The research was supported by the Canadian Wildlife Service and th Wildlife Toxicology Fund of Environment Canada and by the National Science and Engineering Research Council of Canada.  68  Appendix 2.1  Selected morphological measurements in Bald Eagle chicks collected in 1992 from British Columbia.  Parameter  Fraser Delta  East Van. Island  Powell River  West Van. Island  (N=2)  (N=3)  (N=8)  (N=5)  Yolk-free body weight  78.8 + 10.3  87.3 + 4.1  84.3 + 7.8  78.1 + 10.6  Relative liver weight (as % body weight)  2.3 + 0.19  2.3 + 0.34  2.4 + 0.34  2.3 + 0.32  Tarsus length (mm)  19.0 + 0.53  20.6 + 1.41  20.6 + 1.16  20.7 + 2.26  Tibia length (mm)  26.5 + 0.53  27.5 + 1.16  26.5 + 0.69  26.8 + 1.09  NOTE: No significant differences were detected among locations for body or yolk-free body, liver, kidney, intestine, heart, adrenal, yolk, tibia and thyroid weights; tibia, tarsus, culman or wing lengths.  69  CHAPTER 3 BIOACCUMULATION OF CHLORINATED HYDROCARBONS AND MERCURY IN EGGS AND PREY OF BALD EAGLES  The purpose of the bioaccumulation study was to measure chlorinated hydrocarbon levels, particularly for PCDDs and PCDFs, in eagle eggs in order to determine spatial and temporal patterns and trends, and to relate the levels to critical concentrations in their food using a simple model. At issue was the determination of site specific concentrations of PCDDs and PCDFs in representative sentinel food items, such as forage fish and fish-eating birds, that would not adversely affect Bald Eagles. The development of guidelines for chlorinated hydrocarbon levels in dietary items of eagles should have broader applicability in other North American jurisdictions.  Materials and Methods Sample collection From 1990 to 1992, a total of 32 Bald Eagle eggs were collected at six sites on the south coast of British Columbia (Figure 3.1). Four treatment areas were selected based on proximity of eagle breeding sites to industrial pollutant sources. The lower Fraser valley near Vancouver is a heavily urbanized and industrialized area that receives wastes from numerous local and upstream pulp, paper and lumber mills and wood treatment operations. The Crofton and Powell River areas each receive effluent inputs from local kraft pulp mills. Nanaimo is an urbanized area with a large kraft mill and other wood milling and yarding operations. The main reference site was an area of northern Johnstone Strait, with little industrial activity other than lumber yarding. Three single eggs were also obtained from 1) Clayoquot Sound on the west coast of Vancouver Island in an area where lumber cutting is the only industrial activity 2) lower Alberni Inlet, a bleached-kraft pulp mill is at the head of the inlet 3) Langara Island in the Queen Charlotte’s archipelago, remote from any industrial activity. 70  II  I  C C.)  •1  Suitable nests were located by ground, boat and aerial surveys, during which nests were scored numerically to estimate access, suitability for climbing and land tenure. In 1990 and 1991, in an effort to obtain fresh eggs, collections were made during the first two weeks of April in the lower Fraser valley and the Strait of Georgia, during the first week of May on the west coast of Vancouver Island and during the third week in May in Johnstone Strait. Normally a single fresh egg was collected and only from nests with at least two eggs, except at Stillwater Bay in 1992, when both eggs were taken. The two eggs collected in 1994 were addled, they were retrieved from nests in June or July during blood sampling of nestlings. To encourage continued incubation of the remaining egg and thus to minimize the impact of collection, time near the nest and in the nest tree was minimized. Eggs collected in 1990 and 1991 were refrigerated until the contents were removed and placed into chemically-cleaned (acetone/hexane) glass jars with aluminum foil lid-liners and then frozen. The eggs collected in 1992 were initially incubated as part of another study (Chapter 2); the failed eggs from this study were removed from the incubator and then treated the same as eggs from other years. Frozen eggs were shipped to the CWS National Wildlife Research Centre (NWRC) in Ottawa. Chemical analyses Whole eggs were homogenized and prepared for analysis at NWRC. Aliquots for organochlorine pesticides and PCBs were analyzed according to methods described in Norstrom et al. (1988) and outlined in Chapter 1, except that total PCB levels are reported as the sum of 28 congener peaks (24 listed in Figure 3.2, plus trace amounts of PCBs 137, 195, 200 and 206). Eggs collected in 1990 and 1991 were analyzed for PCDDs/PCDFs by low resolution GC/MS using a Hewlett-Packard 5987B machine with a 30 m DB-5 capillary GC column (Norstrom and Simon 1991); the method is described in Chapter 1. PCDD/PCDF and non  ortho PCB analyses of eggs collected in 1992 were carried out on a VG Autospec high resolution mass spectrometer linked to a HP 5890 Series II data system according to methods described by Letcher et al. (in press), also as outlined in Chapter 1. Mercury was analyzed at the NWRC by cold vapour atomic absorption according to methods described by Scheuhammer  72  & Bond (1991), and methyl mercury was extracted as described in Callum and Ferguson (1981). Eggshell thickness measurement Eggshells were air-dried in the laboratory for two weeks or more. Using a ball micrometer, shell thickness was measured at the equator of the shell, including the membrane; five readings were made and averaged. Statistical treatment For each location, data were combined for all years in order to give a larger sample size. Chemical residue data were transformed to common logarithms and geometric means and 95 % confidence intervals determined. Data were also converted to common logarithms and SAS routines used to perform a one-way analysis of variance followed by Tukey’s multiple comparison procedure (MCP) to determine significant differences in mean residue levels among sites. For determination of statistical differences among sites for percent PCB congeners, an arcsine transformation was used, followed by ANOVA and Tukey’s MCP. Unless otherwise indicated, a significance level of p <0.05 was applied to all statistical tests. Patterns of all chlorinated hydrocarbons and other the major PCB congeners as percent total PCBs were analyzed using principle components analysis (PCA) in SAS. As for the other statistical analyses, residue concentrations were transformed to common logarithms, while for the percent PCB congener contributions, an arcsine transformation was used. Toxic equivalents (TEQs) were estimated using standard toxic equivalent factors for PCDDs and PCDFs as suggested in Safe (1990), except that for the mono-ortho and non-ortho PCBs, the World Health Organization toxic equivalents (WHO-TEFs, Ahlborg et al. 1994). Bioaccumulation model In order to relate PCDD and PCDF levels in Bald Eagle eggs to their diet, a simple bioaccumulation model was used (modified after U.S. Environmental Protection Agency 1993 and US Fish and Wildlife Service, 1994). The model assumes: 1) breeding Bald Eagles are year-round residents and, therefore, acquire most of their contaminant burden from local  73  sources 2) levels in eagle eggs are in equilibrium with those in the female’s diet. The model has the form:  BEE =  BMF 1 (X + 2 [F ) (X F ) BEE  + FN(XN)]  Contaminant concentration in Bald Eagle egg  =  BMF  ...  =  Biomagnification factor for a given contaminant  1 F  =  Fraction of item one in diet  1 X  =  Contaminant concentration in item one  FN  =  Fraction of the Nth item  XN  =  Contaminant concentration in the Nth item  As input, we used data on PCDD and PCDF levels in avian and fish prey from near pulp mills and at reference sites on the British Columbia coast, summarized in Tables 3.1 and 3.2. Estimates of Bald Eagle diet composition were taken from Knight et al. (1990), Vermeer et al. (1989) and Watson et al. (1991). The eagle diet was divided into components, which varied among sites based on availability of contaminants data: 1) fish-eating waterfowl (grebes, cormorants, herons and mergansers) 2) non-fish-eating waterfowl (invertebrate and plant-eating waterfowl) 3) omnivorous gulls 4) non-s almonid fish 5) salmonid fish. Biomagnification factors determined in Lake Ontario Herring gulls relative to forage fish (Braune and Norstrom, 1989) were used: 2,3,7,8-TCDD (21), 1,2,3,7,8-PnCDD (10), 1,2,3,6,7,8-HxCDD (16), 2,3,7,8TCDF (1.4, estimated), 2,3,4,7 ,8-PnCDF (4.5). The biomagnification factors for herring gulls were similar to those estimated for great blue herons to forage fish at Crofton, 25 and 10 respectively for 2,3,7,8-TCDD and 1 ,2,3,6,7,8-HxCDD (Elliott et al. 1989a). Where only egg or liver data was available for a species, the inter-tissue ratios in Braune and Norstrom (1989) were used to convert to whole body concentrations.  74  Lii  Rock Fish  English Sole  English Sole  Chinook Salmon  Crofton  Powell River  Powell River  Powell River  *  2  1  -  -  -  Number analyzed I Number collected  Environment Canada, unpubl. data  Harding & Pomeroy, 1990  Data Source:  Arrowtooth Flounder  1/1  Arrowtooth Flounder  Crofton  Crofton  Jan-Feb. 1990  Liver  2/2  English Sole  Crofton  4.3  Jan-Feb. 1990 Jan-Feb. 1990 Jan-Feb. 1990  Fillet Liver  3/3 3/3 Fillet  5.7  Jan-Feb. 1990  Fillet 1.4  1.8  Jan-Feb. 1990  3.6  3.4  10  2.1  6.5  8.5  Liver  Jan-Feb. 1990  Jan-Feb. 1990  Fillet  2/2  English Sole  Crofton  Fillet  Jan-Feb. 1990  Fillet  Chinook Salmon  Nanaimo  Jan-Feb. 1990  Liver  4/4  English Sole  Nanaimo  period 2.9  % lipid  Collection  Jan-Feb. 1990  English Sole  Nanaiino  Tissue  Fillet  4/4  Species  N*  2.2  3.0  < 0.5  1.7  3.5  < 0.5  5.0  1.0  < 1.0  2.6  1.0  TCDD  2,3,7,8  2.4  8.0  <1.0  <1.0  < 2.5  <1.5  8.0  2.0  < 1.0  6.6  1.5  PnCDD  12378  2.5  98  1.5  9.1  9.0  <1.0  64.  4.0  4.2  44  4.5  HxCDD  Total  9.4  7.4  <1.0  < 1.0  < 1.5  <1.5  < 3.0  1.0  < 1.0  1.6  <1.0  HpCDD  Total  62  137  15  17  63  7  89  9  47  58  13  TCDF  2,3,7,8  1.3  9.0  <0.5  <1.0  <1.0  < 0.5  2.0  <0.05  < 1.0  1.7  <0.05  PnCDF  23467  2  1  1  2  1  1  1  1  2  1  1  Source  Data  Mean PCDD/PCDF levels (ng/kg, wet weight) in fish collected near three pulp mills on the Strait of Georgia, British Columbia.  Area  Table 3.1  Table 3.2 PCDD and PCDF levels (ng/kg, wet weight) in waterbird and seabird species from the British Columbia coast. Species  Location  Year  Tissue  N  % Fat  2378TCDD  12378PnCDD  123678HxDDD  OCDD  2378TCDF  23478PnCDF  Ref  Port Alberni  1989  liver  5  7.3  117  385  249  <15  217  38  1  1992  bm*  8  3.8  25  66  64  <22  69  <2.4  2  Nanaimo  1992  bm  3  4.6  2.9  <2.9  8.6  <8.3  41  <1.4  2  Powell River  1992  bm  5  7.1  4.4  <2.6  21  <9.5  230  9.9  2  Alert Bay  1992  bm  2  3.1  <1.6  <3.2  <3.5  <18  <0.9  <2.4  2  Fraser Delta  1990  eggs  6  4.5  25  42  64  ND  ND  11  2  Fraser Delta  1992  eggs  10  4.7  8  14  11  ND  ND  ND  2  Crofton  1990  eggs  7  4.2  26  42  65  ND  2  11  2  Crofton  1992  eggs  8  4.8  8  25  47  ND  1  7  2  Nanaimo  1992  eggs  10  5.6  10  22  38  ND  ND  ND  2  Fraser Delta  1990  eggs  8  6.4  42  45  57  5  15  16  2  Fraser Delta  1992  eggs  10  5.6  10  10  11  ND  4  1  2  Crofton  1990  eggs  5  6.3  102  223  229  3  10  32  3  Crofton  1992  eggs  5  6.1  19  40  47  1  5  7  3  Johnstone Strait  1990  eggs  12  18  1  3  3  1  27  2  2  Fraser Delta  1990  bm  1  3.4  18  <1  17  <7  10  <1  1  Crofton  1990  bm  2  2.1  9.4  <2  29  <10  28  3.5  1  Alert Bay  1992  bm  11  3.7  <1  <3  <3  <10  1.4  <1  2  Fraser Delta  1989  bm  10  4.1  <1.1  <1.9  <3.1  <11  13  <1.5  1  Port Alberni  1989  liver  5  3.1  24  22  30  11  123  13  1  Crofton  1990  bm  10  2.8  5.1  <0.5  3.3  3  40  1  1  Nanaimo  1992  bm  6  2.5  0.5  0.7  2  7.3  7.4  1.3  2  Powell River  1992  bm  8  3.1  <1.6  <3.1  <4.2  <8.2  12  <1.8  2  Alert Bay  1992  bm  7  1.6  <2  <6.5  <6  <22  3.2  <2.8  2  Nanaimo  1992  eggs  10  8.8  <1.0  <1.0  22  <1.0  <1.0  1.0  2  Fish eating birds Western Grebe  Double-crested Conuorant  Great Blue Heron  Rhinoceros Auldet  Invertebrate feeders Bufflehead  Surf Scoter  Glaucouswinged Gull  *bm breast muscle ND not detected (detection limit = 0.5 ng/kg) References: 1 Whiteheact et al. 1990; 2 Elliott et al., 1995b; 3 -  -  -  -  76  -  Whitehead et al. 1992.  Results PCDDs and PCDFs Major PCDD contaminants were 1,2,3,6,7,8-HxCDD > 1,2,7,8-PnCDD > 2,3,7,8-  TCDD, except in the lower Fraser valley, where 2,3,7,8-TCDD was the greater than the other two compounds (Table 3.3). All eggs contained detectable levels of the three major PCDD congeners. Lesser concentrations of 1,2,3,4,6,7,8-HpCDD and OCDD were found in most eggs. The only PCDFs consistently detected in Bald Eagle eggs were 2,3,7,8-TCDF and 2,3,4,7,8-PnCDF. Eggs from Johnstone Strait contained significantly less 2,3,7,8-TCDD than did eggs from other sites. Concentrations of 1,2,3,7,8-PnCDD were significantly higher in eggs from Crofton than either the lower Fraser valley or Johnstone Strait. Concentrations of 1,2,3,6,7,8-HxCDD and 2,3,4,7,8-PnCDF were significantly lower in Fraser valley eggs than from the pulp mill sites, but did not differ significantly from Johnstone Strait. Organochiorines Quantifiable residues of DDE, DDD, trans-nonachlor, cis-nonachlor oxychlordane, cis chiordane, heptachlor epoxide, dieldrin, mirex, fi-HCH and HCB were found in all eggs analyzed (Table 3.4). DDT was found in the majority of the eggs at low levels, generally < 0.01 mg/kg. Photomirex was detected in 65 % of the eggs; where present concentrations were about 50 % of mirex concentrations. Organochiorine levels were generally highest in eggs from Powell River, although differences were significant in only one case:  trans-nonachlor was  significantly higher at Powell River and Nanaimo than the Fraser valley. Mean concentrations of cis-chlordane were significantly higher in eggs from Johnstone Strait than other sites except Powell River and were also significantly lower at Crofton than all other sites. Mercury  Highest mean concentrations of total mercury were in eggs from Johnston Strait and the Fraser valley and were significantly higher than those from Nanaimo and Crofton, but not Powell River (Table 3.5). Methyl-mercury was also determined in the eleven eggs from 1990 and constituted an average of 88 % (SD = 11, range 73 those Bald Eagle eggs. 77  -  100%) of the total mercury present in  Crofton R. Pringle (14) Southey (15) Crofton(16) 1990 1990 1991 Mean  1990 1990 1990 1990 1992 1991 1991 1992 Mean  4.4 5.9 6.0 5.4 ±0.9  4.5 4.9 4.5 4.2 4.7 4.4 ±0.6  3.0 4.7 4.2  5.6 ±0.6  Mean  Nanaimo Canoxy (7) Leask (8) Canso (9) Jack Pt. (10) Northwest Bay (11) Maude Island (12) Southey Island (13) Jack Point (10)  5.1 5.9 6.1 5.3 4.9  1990 1990 1990 1991 1991 1991  Lower Fraser Valley Brunswick Pt. (1) Annacis Is. (2) Chahalis Flats (3) Island 20 (4) Cheam Island (5) Agassiz Bridge (6)  82.7 80.0 82.8 81.8 ±1.6  83.4 ± 1.4  22  85.7 83.0 84.1 84.2 81.2 83.2 83.3  82.4 ±0.5  82.5 82.7 83.1 82.3 82.0  %liyid %moisture (mean ± SD)  Year  -  104 110 51 842 (29-243)  (26-78)  452  59 63 82 79 14 70 28  (30-63)  211 149 108 150” (65-346)  (35-122)  66  109 99 116 133 22 104 29  55 52 7 6 15 202 (7-57)  442  37  58 58 51 23 41  12378 PnCDD  42  2378 TCDD  374 310 173 270L (99-742)  (68-264)  134”  198 250 346 227 37 173 79  (14-76)  332  55 17 15 18  42 112  123678 HxCDD  (0.1-6)  32  10 7 1  (1.5-7.6)  341  5 9 12 6 1 2 5  5 2 ND ND 1 1.3 (0.5-3.8)  4  123789 HxCDD  23 112 89 16 73 13 392  2378 TCDF  16 26 60 292 (6-154)  (26-70)  432  29 36 29 49 18 119 65  (14-105)  PCDD and PCDF Levels (ng/kg) (geometric mean and 95% confidence interval)  27 34 22 27” (16-47)  (8-27)  35 12  5  16 18 31 20  (1.3-22)  5a  13 12 14 ND 2 10  23478 PnCDF  Polycliloiinated dibenzodioxin (PCDD) and polychiorinated dibenzofuran (PCDF) residue levels (wet weight basis) in Bald Eagle eggs from British Columbia, 1990 1992.  Nest (Map No.*)  Table 3.3  -  88.1 81.4 80.1  2.4 5.9 6.5  1992 1992 1994 -  83.3 84.1 ±4.1  3.9 4.3 ±1.0  Mean  1991  82.8 83.3 80.6 83.3 85.4 85.4  -  -  11  17 10 2  (10-22)  iSh  54 16 5  25 35 (26-48)  39 28 64 43 33 26  38 7 3  (51-118)  78’  43  72 167 111 73 79 52  (103-261)  (37-138)  (23-105) 22 12 32 11 15 10  170b 71b,c  59  49ä  80.7 ±1.4  123678 HxCDD  47 104 106 18  129 128  12378 PnCDD  1 ND ND  1.39 (0.5-3.4)  1  5 3 5 2 ND 1  (2-9)  4  1  15 7 2 3 3  7  123789 HxCDD  7 7 5  (33-66)  47  58 39 80 37 68 29  85 (49-147)  59 97 110 18 166 168  2378 TCDF  PCDD and PCDF Levels (ng/kg) (geometric mean and 95% confidence interval)  244 372 186 116 143 146 80  98 88 41 32 78 81 10  2378 TCDD  80.0 82.4 82.1 81.9 79.6 79.3 79.6  4.8 4.3 6.0 4.4 2.8 4.1  5.7 5.0 3.7 5.8 6.7 6.1 5.6 ±1.0  6.1  %liyid %moisture (mean ± SD)  1991 1991 1991 1991 1991 1991  1990 1990 1991 1992 1992 1992 1992 Mean  Year  Map Nos. 30 Albemi Inlet, 31 Clayoquot Sound. ** Queen Charlotte Islands ab,c means that do not share the same letter are significiantly different (p 0.05)  Pocahontas Pt (30) Berryman Pt (31) Langara Is.**  Johnstone Strait Plumper 5 (23) Plumper 8 (24) Pearce 3 (23) Pearce 5 (26) Harbiedon Island (27) Swanson Island (28) Owl Island (29)  Powell River Kelly Pt. (17) Convent (18) Lund (19) Powell River (20) Stiliwater (A) (21) Stiliwater (B) (21) Grise Point (22)  Nest (Map No.*)  Table 3.3 cont...  4 4 2  (6-11)  10 7 3 5 10 5  27b (15-49)  27 37 24 5 48 50  23478 PnCDF  6  8  Lower Fraser Valley  Nanaimo  4 a b  a,b  -  means that do not share the same letter are significiantly different (p  0.034  2.97  1.91  1  Langara Island  0.044  5.12  3.86  1  Clayoquot Sound  0.028  5.14  4.47  1  a 229  -  0.024-0.058 (0.026-0.112)  Alberni Inlet  33a  1.64 6.66 (1.31-8.70)  0.071 0.035-0.143 (0.029-0.215)  0.039a 0.014-0.110 (0.024-0.50)  2.77a 1.48-5.17 (2.07-3.26)  0.o52 0.035-0.079 (0.023-0.101)  1.46-3.58 (1.22-5.95)  2.5&’ 1.78-3.64 (1.70-5.34)  7  Johnstone Strait  5.08 3.88-6.65 (3.32-6.96)  7  2.23-10.2 (3.47-6.38)  3.28-6.79 (1.80-7.14)  0.058 0.041-0.083 (0.030-0.075)  DDD  3.13a 1.65-5.92 (.672-8.52)  2j7a 1.07-4.41 (.90-4.14)  ab 268  1.49-4.8 (1.08-6.21)  DDE  Total PCBs  0.05)  0.215  0.293  0.151  0.144-0.301 (0.142-0.453)  0.32k’ 0.234-0.438 (0.192-0.478)  0.039  0.037  0.028  0.047” 0.031-0.079 (0.033-0.119)  0.046-0.094 (0.041-0.109)  b 0066  0.020-0.062 (0.028-0.044)  ab 0036  0.121-0.217 (0.143-0.179)  162 • b  004 a 7 . b  0.036-0.060 (0.027-0.076)  ab 0245  0.186-0.323 (0.148-0.432)  0.03 0.020-0.044 (0.018-0.049)  cisnonachior  0.142k 0.098-.204 (0.082-0.234)  transnonachior  0.095  0.047  0.024  o.0z4 0.005-0.117 (0.001-0.081)  0.034-0.062 (0.030-0.075)  a 0046  0.020-0.550 (0.027-0.041)  0.033a  0.042a (0.029-0.062) (0.014-0.062)  0.037a 0.023-0.059 (0.022-0.082)  oxychiordane  0.024  0.06  0.038  0.020-0.048 (0.017-0.071)  a 0031  0.044a 0.031-0.064 (0.021-0.065)  0.03k 0.014-0.065 (0.021-0.038)  0.042a 0.029-0.061 (0.018-0.064)  0.037a 0.019-0.073 (0.020-0.091)  dieldrin  0.062  0.038  0.019  0.o2 0.014-0.028 (0.014-0.044)  0.016-0.048 (0.015-0.063)  a 0028  0.016k 0.006-0.042 (0.010-0.023)  o.olsa 0.005-0.047 (0.001-0.030)  0.009 0.002-0.045 (0.001-0.038)  mirex  0.061  0.026  0.024  0.013 0.003-0.052 (0.001-0.046)  0.013-0.047 (0.007-0.052)  a 0025  0.018k 0.008-0.040 (0.013-0.024)  0.022a 0.011-0.042 (0.004-0.053)  0.001-0.024 (0.001-0.032)  o.oosa  beta11CR  0.051  0.019  0.001  0.024a 0.016-0.036 (0.015-0.055)  o.olza 0.003-0.046 (0.001-0.031)  o.oiz 0.008-0.020 (0.010-0.015)  0.012 0.004-0.038 (0.001-0.033)  0.025a 0.017-0.039 (0.016-0.032)  HCB  Organochiorine and PCB residue levels (mg/kg, wet weight) in Bald Eagle eggs from the British Columbia coast, 1990-1992, expressed as geometric means and 95% confidence intervals (range in brackets).  Powell River  Crofton  N  Location  Table 3.4  Table 3.5 Mercury residue levels (mg/kg, wet weight) in Bald Eagle eggs from locations on the British Columbia coast, 1990-1992, expressed as geometric means and 95% confidence intervals (range in brackets). Lower Fraser Valley (N=6)  Nanaimo (N=8)  Crofton (N=3)  0.25&’ 0.186 0.358 (0.170 0.400)  0.147a 0.110 0.198 (0.070 0.240)  0.191a 0.096 0.384 (0.150 0.260)  -  -  a,b -  NA  -  -  -  -  Powell River (N=7)  Johnstone Strait (N=7)  Alberni Inlet (N=1)  Clayoquot Sound (N=1)  Langara Island (N=1)  0.08  0.17  NA  b 0294  0.174 (0.150  -  -  0.296 0.380)  means that do not share the same letter are significiantly different (p not analyzed  0.236 (0.220  -  -  0.367 0.440)  0.05)  -  Polychiorinated biphenyls Mean sum-PCB concentrations were significantly different only between Powell River, where they were higher, and Johnstone Strait (Table 3.2). There were a number of significant differences in mean concentrations of individual PCB congeners: PCBs 170 (2345-234), 171 (2346-234), 182 (2345-246), 201 (2356-2345) and 203 (23456-245) were significantly higher in  eggs from Crofton, Nanaimo and Powell River than Johnstone Strait; PCBs 180 (2345-245), 183 (2346-245) and 194 (2345-2345) were significantly higher at Crofton, Nanaimo and Powell River than both Johnstone Strait and the Fraser valley; PCBs 153 (245-245) and 128 (234-234) were significantly higher only at Powell River compared to Johnstone Strait and the Fraser valley and PCB 138 (234-245) was significantly higher at Powell River than Johnstone Strait. The percent contribution of individual congeners was determined and compared among sites (Figure 3.2). The major peaks were 153, 138, 180, 182, 118 (245-34) and 99 (245-24), which together contributed 64 % of the total PCBs present in all eggs. There were a number of statistically significant differences among sites in percent contribution of individuals PCBs. Percent contribution of a number of the lower chlorinated congeners, including PCBs 66 (2434), 101 (245-25), 99, 87 (234-25), 118 and 105 (234-34), was significantly higher at both Johnstone Strait and the lower Fraser Valley than the other three sites. In addition, among those compounds, percent contribution of PCBs 99 and 118 were significantly higher at Powell River than at Crofton. In contrast, the percent contribution of a number of the higher  81  chlorinated congeners, PCBs 183, 180, 170, 203 and 194, was significantly lower in eggs from the lower Fraser Valley and Johnstone Strait than Nanaimo, Crofton and Powell River. This geographic trend of differences in the PCB pattern was supported by results of principle components analysis. Principle components analysis of the PCB pattern was carried out using only congeners, 170, 180 and 194, which are indicative of Aroclor 1260 (Mullin et al. 1984), PCBs 99 and 118, indicative of Aroclor 1254, and PCB 66, considered indicative of Aroclor 1242. Two significant principle components were apparent which explained 90 % of the total variance among individual egg analyses. The first component (PC 1) explained 75 % and the second component (PC2) explained 15 %. As shown in Figure 3.3, the Johnstone Strait and Fraser Valley eggs clump separately from the other locations, although there is some overlap, particularly of some samples from Powell River.  25  20  5  0 p..  p.  ..  p..  PCB Congener Number Figure 3.2 PCB congeners in Bald Eagle from British Columbia, 1990-1992, expressed as percent of total PCBs. Values represent means of three to eight analyses per collection site. Congeners are identified according to their IUPAC number.  82  Concentrations of six non-ortho PCB congeners were determined in eight eggs collected in 1992 (Table 3.6) and in two eggs collected in 1994. The pattern was consistent in the 1992 samples and the 1994 sample from Langara Island, with 126 (345-34)> 77 (34-34) > 169 (345-345) > 81 (345-4) > 37 (34-4). However, in the 1994 sample from Herrling Island, 77 > 126 > 81 > 169 > 37. Linear regressions were determined between concentration of PCBs 126 and 77 and sum-PCBs for the ten eggs in Table 3.3, in order to estimate values in the whole data set for estimation of TCDD toxic equivalents (TEQs): PCB 126 (ng/kg) PCB 77 (ng/kg)  T  =  156 [sum-PCBs (mg/kg)] + 78, r =0.634, p<O.Ol 2 69 [sum-PCBs (mg/kg)] + 85, r 2 = 0.505, p <0.04  PCBs  170 180  194  z  PCBs 66  j99 V  118  -3 -2.5 -2 -1.5 -1 -0.5 0 PRIN2 PCB 99  0.5  1  1.5  PCB 66  L  Lower Fraser Valley  C  Crofton  N  Nanaimo  P  Powell River  J  Johnstone Strait  Figure 3.3 Plot of the first and second principle components (PC 1 and PC2). Selected PCB congeners only, considered indicative of various Aroclor inputs (Mullin et al. 1984) were included in the analysis. Concentrations for all individual egg analyses were expressed as percent total PCBs and arcsine transformed. A total of 75% of the matrix variance was explained by PC 1 and 15 % by PC2. 83  Table 3.6Non-ortho PCBs in Bald Eagle eggs (ng/kg, wet weight) collected from British Columbia, 1992. Nest (Map No.*)  *  PCB-37  PCB-81  PCB-77  PCB-126  PCB-169  PCB-189  Northwest Bay (13)  6.2  24  146  323  65  22  Jack Point (10)  31  52  349  709  131  41  Powell River (20)  13  42  207  544  131  35  Stiliwater A (21)  26  105  720  1354  285  84  Stillwater B (21)  24  107  684  1326  277  97  Grise Point (22)  52  46  387  547  121  41  Pocohantas Point (30)  15  51  459  685  114  35  Berrryman Point (31)  23  74  691  754  135  39  Herrling Is.  27  96  576  314  47  5  Langara Is.  5  32  310  585  203  1  10-13  -  Nanaimo, 20-22  -  Powell River, 30  -  Albemi Inlet, 31  -  Clayoquot Sound.  600 lIE non-O-PCBs 500 —400 c,)  mono-O-PCBs  LI PCDFs other-PCDDs mTCDD  0) U)  C  w I—  200 100  ii  0 çcc  0’  $  Cl’  Figure 3.4 The contribution of various chlorinated hydrocarbon groups to the sum of TCDD toxic equivalents (TEQs) in Bald Eagle eggs from coastal British Columbia, 1990-1992 (N values and variances are in the tables). Toxic equivalents for PCDDs/PCDFs from Safe 1990 and for PCBs from Ahlborg et al. 1994.  84  TCDD toxic equivalents (TEQs) Highest mean TEQs 0 were in eggs from Crofton, followed by Powell River, both of which were significantly greater than Johnstone Strait. The relative contribution of PCDDs and PCDFs to total TEQs , 64 %, was also highest at Crofton and was lowest, 47 %, in the 0 lower Fraser Valley eggs, as shown in Figure 3.4. Eggshell thickness results Neither mean eggshell thickness nor the percentage difference from the pre-1946 average for the Pacific North West of 0.6088 mm differed significantly among sites (Table 3.7). There were no significant regressions between eggshell thickness and DDE or other organochlorines. Table 3.7  Eggshell thickness data, mean ± SD, (range in brackets) for Bald Eagle eggs collected from British Columbia, 1990-1992.  Area  Collection Period  N  Shell thickness (mm)  Percent Difference from prel947*  Lower Fraser Valley  1990-91  6  0.558 ± 0.024  -8.3 ± 4.3 (-14.6 to +1.5)  Nanaimo  1990-92  5  0.587 ± 0.035  -3.6 ± 6.4 (-8,6 to +5.2)  Crofton  1990-91  3  0.583 ± 0.024  -4.2 ± 4.7 (-9.7 to -1.5)  Powell River  1990-92  5  0.590 ± 0.038  -3.1 ± 7.0 (-11.3 to +6.7)  1991  6  0.569 ± 0.036  -6.6 ± 6.3 (-12.9 to + 3.5)  Johnstone Strait  *  pre-1947 value 0.6088 (Anderson & Hickey, 1972) -  Bioaccumulation of PCDDs and PCDFs from prey An example output from the model is shown in Table 3.8, based on 1990 data from Crofton, the location with the best data base of contaminants in prey items. In the absence of local data, results on gulls and salmonids from nearby Nanaimo were used. The putative diet  85  Sole, flounder, rock fish  Bufflehead and Surf Scoter  2  2  4.5  23478-PnCDF  biomagnification factor  35  1.4  2378-TCDF  -  11  16  123678-HxCDD  BMF  1  10  12378-PnCDD  1  7  1  0.5  7  5  3  (0.25)  2 eating birds (0.15)  Gulls  Non-fish  21  BMF’  2378-TCDD  Chemical  35  12  0.5  10  0.5  37  17  2  46 70  2  (0.4)  3 salmonids  0.5  15  2.5  0.5  1  (0.125)  Salmonids  Fish (0.525)  Non-  30  (0.025)  Cormorants  230  220  100  (0.05)  Herons  Birds (0.475)  (Fraction of that item in simulated diet)  Contaminant concentration in dietary items (ng!kg, wet weight)  8  31  284  82  117  Value  Calculated  31  21  342  180  107  Value  Measured Mean  bald eagle eggs  Contaminant concentration in  Table 3.8 A simulation of PCDD and PCDF levels in Bald Eagle eggs at Crofton, 1990, based on concentrations in the diet.  consisted of 52.5 % fish, and 47.5 % birds, mainly gulls and non-fish-eating species; fisheating birds comprised only 6 % (herons and comorants). The model accurately predicted 2,3,7,8-TCDD levels in eagle eggs, but concentrations of other compounds, such as 1,2,3,7,8PnCDD, were less accurately predicted. BMFs for the compounds, other than 2,3,7,8-TCDD, are a possible source of error. Being derived from a Lake Ontario food chain, the 2,3,7,8TCDD level in the forage fish prey was relatively high, while levels for the other PCDDs and PCDFs were near the detection limit; thus, a small difference in forage fish concentrations would translate to a large error in the estimated BMF. The other major source of error is the putative eagle diet, particularly the relative importance of fish and non-fish-eating birds. The example in Table 3.8 approximates an average coastal Bald Eagle diet; however, individual eagles or sub-populations can prey on greater amounts of fish-eating birds. For example, Knight et at. (1990) reported that western grebes, which can accumulate extremely high PCDD/PCDF levels (Table 3.2), were the main prey item of Bald Eagles in the Puget Sound area. Eagles nesting near great blue heron colonies may also prey on chicks and adults (Norman et al. 1989). Figure 3.5 shows how 2,3,7,8-TCDD concentrations would increase in eagle eggs with an increasing fish-eating bird diet. Feeding on fish-eating birds may account for extremely high liver levels of PCDDs, PCDFs and other chlorinated hydrocarbons of adult eagles found dead or dying near Powell River and other areas of the Strait of Georgia (Chapter 1).  87  5 *Crofton, 1993 ±Crofton 1990  C)  )K  0  10  20  30  40  50  60  70  80  90  Percent fish-eating birds in diet Figure 3.5 Concentration of 2,3 ,7,8-TCDD predicted in Bald Eagle eggs based on the percent of fish-eating birds in the diet. Prediction is based on a bioaccumulation model described in the text and the simulation is based on data from Crofton, British Columbia.  88  Discussion The data presented in this chapter show that Bald Eagle eggs collected near bleached kraft pulp mills in the Strait of Georgia contained higher levels of 2,3,7,8-TCDD and -TCDF when compared to other locations on the British Columbia coast. Total PCB levels were also highest in eggs from the Strait of Georgia, reflecting greater industrialization. Concentrations of organochiorine pesticides, including DDE, in eagle eggs were relatively consistent among sites. Total-mercury levels were significantly higher in eggs from the Fraser Valley and Johnstone Strait than the Strait of Georgia. Patterns and sources of PCDDs/PCDFs The formation of 2,3,7,8-TCDD and 2,3,7,8-TCDF during molecular chlorine bleaching of wood pulp is a well known phenomenon (Kuehi et at. 1987; Luthe et al. 1990). By 1991, all pulp mills studied here had implemented bleaching technology changes designed to minimize TCDD/TCDF formation (Table 3.9), which has resulted in declining PCDD levels, particularly of 2,3,7 ,8-TCDD, in sediments and biota near the mills (Whitehead et at. 1992; Elliott et a!. 1995). Concentrations of 2,3,7,8-TCDF in eagle eggs from Nanaimo and Powell River were still elevated in 1992, suggesting that efforts to reduce TCDF contamination have been less successful. In birds, 2,3,7,8-TCDF is quite quickly cleared from the body (Braune and Norstrom 1989; Van den Berg et at. 1994). Other studies of wild birds have reported low 2,3,7 ,8-TCDF concentrations from the Great Lakes (Hebert et at. 1994; Ankley et at. 1993) and Europe (Van Den Berg et at. 1987). However, elevated TCDF levels have been reported in fish, invertebrates, and waterfowl near both riverine and marine pulp mills (Mah et a!. 1989; Harding and Pomeroy, 1990; Table 3.1; Champoux 1993). Osprey eggs collected from nest locations downstream of pulp mills in the British Columbia interior contained 2,3,7,8TCDF levels up to 68 ng/kg (Whitehead et at. 1993). The high TCDF levels in eggs of eagles and ospreys likely reflect a combination of recent exposure and direct yolk deposition of contaminated dietary lipids, as suggested previously for great blue herons (Elliott et at. 1989a).  89  Until 1989, up to several million kg of chiorophenolic compounds were used annually by the British Columbia forest industry, particularly on the coast, to prevent sap staining of undried lumber. Although HxCDDs and HpCDDs predominate as dioxin contaminants in chlorophenol mixtures, HxCDDs are further produced in large amounts when chlorophenol contaminated woodchips are pulped (Luthe et al. 1990). Monitoring chip supplies for chiorophenols, followed by a regulatory ban on their use as anti-sapstains, produced significant HxCDD reductions in effluents and foodchains at the Crofton mill site (Whitehead et al. 1992). A reduction in PCDD levels in eagle eggs is apparent, particularly between 1990 and 1992 at Jack Point near Nanaimo. In Fraser valley eagle eggs low HxCDD : TCDD ratios are consistent with lower HxCDD concentrations in sediments and biota downstream of Fraser river pulp mills, the putative sources of PCDDs and PCDFs at that site (Mah et al. 1989; Whitehead et al. 1993; Harfenist et al. 1995). Due to the cooler, dryer climate of the British Columbia interior, lesser amounts of chlorophenol antisapstain agents were use by lumber operations on the upper Fraser and Thompson Rivers. Osprey eggs collected in 1991 from nests located downstream of the pulp mill on the Thompson River at Kamloops had mean values of 47:3:22 ng/kg, TCDD:PnCDD:HxCDD (Whitehead et al. 1993). In contrast, some osprey eggs contained very high levels of 1 ,2,3,4,6,7,8-HpCDD and OCDD, indicative of direct chlorophenolic inputs, rather than via pulp milling of contaminated wood chips. Although there are no pulp or large saw mills on northern Johnstone Strait (only log sorting facilities), PCDD/PCDF levels in eagles were relatively high. A non-kraft pulp mill located to the west at Port Alice reported non-detectable PCDD/PCDF levels in effluents (Anonymous 1994), and only trace amounts, 4 ng/kg of 2,3,7,8-TCDF, in crab hepatapancreas from near the mill site (Harding and Pomeroy 1990). The PCCD/PCDF pattern in Johnstone Strait eagle eggs is similar to the Strait of Georgia, which is the most likely source; however, the exposure route is not clear. Acquisition of contaminants during seasonal southern  90  movements is unlikely as resident Bald Eagles on the Pacific coast remain on territory for most of the year (Frenzel et al. 1989). Residents may leave breeding territories periodically during the fall and winter to feed at salmon spawning sites; however, Pacific salmon, even from near pulp mill sites, contained low PCDD/PCDF levels, with the exception of some 2,3,7,8-TCDF. Eagle eggs from the west coast of Vancouver Island also had low PCDD/PCDF levels (Table 3.3, Chapter 2), probably indicating that they had not dispersed to more contaminated sites. Long range transport is unlikely as a major vector, as pulp mill pollution is relatively localized even within the Strait of Georgia (Elliott et al. 1989a; Harding and Pomeroy, 1990). There is, however, an estuarine surface flow out of the Georgia Strait through Johnstone Strait (Thomson 1981), which may conceivably transport some suspended sediment-bound PCDDs and PCDFs. A sediment sample from Louchborough Inlet, a fjord off of central Johnston Strait, was reported to have levels of higher chlorinated PCDDs comparable to those near industrial sites in the Fraser delta (Harding 1990). However, eagle prey species, such as western grebes and surf scoters collected from Johnstone Strait in mid-March 1992, timed to obtain birds which had wintered on site, had very low PCDD/PCDF levels, while samples of the same species collected near pulp mills showed the typical pulp mill PCDD/PCDF signature. Johnstone Strait Bald Eagles may still be exposed to contaminants from the Strait of Georgia by feeding on waterfowl during spring migration along the coast towards their northern breeding grounds. Rhinoceros auklets, large numbers of which breed in northern Johnston Strait, contained low PCDD levels, although the mean 2,3,7,8-TCDF concentration was quite high and could partially account for this compound in Johnston Strait eagle eggs. The pattern of HxCDD > PnCDD > TCDD in Strait of Georgia wildlife differs from that reported at other locations such as the Great Lakes (Ankley et al. 1993), interior rivers of British Columbia (Whitehead et al. 1993) and elsewhere in North America (Elliott et al. 1995a). Hebert et al. (1994) used principal components analysis to show that Strait of Georgia blue heron eggs clustered separately from Great Lakes herring gulls and other biota, based on higher PnCDD and HxCDD concentrations, attributed to chlorophenol sources. However, a  91  sample of common merganser eggs from downstream of a pulp mill in Quebec had a pattern, 24:28:40 ng/kg TCDD:PnCDD:HxCDD, similar to that observed in British Columbia, perhaps indicating a chlorophenol and a pulp mill source (Champoux 1993). Baltic Sea Common Murre (Uria aalge) eggs contained 27:45:59 mg/kg TCDD: PnCDD: HxCDD (wet weight, re calculated based on 17 % lipid in common murre eggs (Noble and Elliott 1986; Cederberg et al. 1991), similar to the Strait of Georgia pattern. Grey Heron (Ardea cineria) livers from the Netherlands also had a pattern somewhat similar to the Strait of Georgia, which was attributed mainly to chlorophenols (Van den Berg et al. 1987). European wildlife samples, at least from The Netherlands, appear to have higher 2,3,4,7,8-PnCDF concentrations (Bosveld et at. 1994; Van den Berg et a!. 1994b) compared to those from North America (Elliott et at. 1989a; Hebert et at. 1994). This compound is a known contaminant in PCB mixtures (Van den Berg et al. 1985), which would explain its association with areas of PCB contamination (Hebert et a!. 1994) and its tendency to correlate closely with PCB congeners in eggs (Elliott et at. 1989). Bosveld et at. (1994) suggested that higher PCB levels in European wildlife samples explained the elevated 2,3,7,8-PnCDF levels; they determined that lipid-normalized PCB concentrations in Common Tern (Sterna hirundo) yolksacs from the Rhine-Meuse estuary were two to three-fold higher than in fish-eating bird eggs from industrialized areas of the Great Lakes. However, direct comparison of lipidnormalized whole egg to yolksac concentrations may overestimate concentrations in yollcsacs. For example, in Bald Eagles, concentrations of chlorinated hydrocarbons were three-fold higher on a lipid weight basis in a single yolksac compared to the sibling whole egg. On a wet weight basis, total PCB levels in Great Cormorant (Phatacrocorax carbo) eggs from the contaminated Biesbosch colony in the Netherlands (Van Hattum et at. 1993 cited in Bosveld and Van den Berg, 1994) were similar, about 23 mg/kg, to those in double-crested cormorants from highly contaminated Hamilton Harbour in the Great Lakes (Bishop et al. 1992). Therefore, differences in PCB formulations or other sources may account for higher PnCDF levels in European wildlife samples, rather than higher PCB levels.  92  Patterns and sources of organochiorines and mercury The uniformity in OC residues indicates similar dietary exposure among most individuals. The few eggs with distinctly lower organochiorines are probably individual eagles feeding on larger amounts of fish, non-fish-eating birds or mammals. Based on OC patterns in seabird eggs, Elliott et al. (1989) concluded that atmospheric sources were dominant over a wide area of the British Columbia coast. However, local sources can still pre-empt the influence of atmospheric input: DDE levels in heron eggs were significantly higher in colonies from the Fraser delta (0.49 mg/kg), an area of intensive farming, than non-agricultural locations (0.11 mg/kg) (Elliott et al. 1989; Whitehead, 1989). In fact, the mean DDE level in two eagle eggs collected within the Fraser delta, 3.86 mg/kg, is significantly higher than the four eggs from upstream of the main agricultural areas, 1.63 mg/kg DDE. High DDE levels continue to be reported in wildlife from areas of former high DDT use, such as orchards (Blus et al. 1987; Elliott et al. 1994). After the DDT-related compounds, chiordanes were present at the highest concentrations in eagle eggs. Among chiordanes, trans-nonachlor was consistently the dominant component, constituting a mean of 67 % (SD =5.3, range 51-77  %) of the total. Oxychiordane, considered  to be the most stable metabolite (Nomeir & Hajjar 1987), made a mean contribution of 13 % (SD =5, range 0.2-27  %). Some authors have suggested that a high ratio of trans-nonachlor to  oxychlordane levels in tissues shows a lower specific capacity to metabolize chlorinated hydrocarbon compounds (Kawano et al. 1986; Yamashita et al. 1993). The concentrations of chlordane-related and heptachlor epoxide compounds found here are similar to those reported in addled Bald Eagle eggs collected in the early 1980s from a variety of United States locations (Wiemeyer et al. 1993). Concentrations of mirex and dieldrin were somewhat higher in those U.S. Bald Eagle eggs collected a decade earlier than in the fresh eggs from the British Columbia coast in 1992. Mean DDE and PCB levels were about three-fold higher in eagle eggs from the lower Columbia River than the lower Fraser River (Anthony et al. 1993). Dietary differences may partly account for this; eagles in the  93  lower Columbia reportedly ate more Western Grebes (Watson et al. 1991), which tend to have high levels of chlorinated hydrocarbons (Table 3.7), while Fraser estuary eagles ate a large proportion of Glaucous-winged Gulls which tend to have low organochiorine levels (Table 3.7), probably because in that area they consume mainly garbage (Vermeer et al. 1989). Differences in organochiorine levels in estuarine biota also reflect differences in agricultural and industrial development of the respective watersheds. Areas of intensive agriculture, particularly fruit orchards are more prevalent in the Columbia basis and account for high DDT (Rinella et al. 1993). Hydroelectric development is much greater on the Columbia river and likely accounts for higher PCB concentrations, evident in Osprey eggs collected in the upper reaches of each watershed (Whitehead et al. 1993). Higher mercury levels in Bald Eagle eggs from the Fraser estuary are consistent with data in herons from that site (Elliott et al. 1989a), and with Fimreite et al. ‘s (1971) findings of higher mercury in aquatic versus coastal marine fish. Elevated mercury levels in fish-eating birds were associated with industrial, including pulp mill, sources by Fimreite et al. Based on the levels in eagle eggs, any mercury discharges from Crofton and Nanaimo pulp mills have not had a lasting impact in local food chains. Highest mercury levels were in the Johnstone Strait eagle eggs. A great proportion of fish in the diet may explain higher mercury levels in Johnstone Strait and the lower Fraser Valley, as suggested below to account for the PCB pattern at those sites. Polychiorinated bihenyls Mean total PCBs in Bald Eagle eggs were highest near the three pulp mill sites, which contrasts with data on great blue herons, in which highest PCBs were from colonies in the Fraser delta near Vancouver (Elliott et al. 1989a; Whitehead 1989). However, the PCB concentration in the single Bald Eagle egg from an industrial site in the Fraser delta, 6.21, was in the same range as the eggs from near the pulp mill sites; other Fraser valley Bald Eagle eggs were from agricultural or woodland locations and PCB levels were 50 % lower. The PCB pattern in great blue herons varied significantly among sites which was attributed to local  94  differences in Aroclor inputs (Elliott et al. 1989a). Variability in PCB congener patterns in wildlife in the the Green Bay area were also attributed to different industrial Aroclor sources (Ankley et a!. 1993). However, in British grey herons, Boumphrey et al. (1993) considered dietary differences as the best explanation for individual variation in PCB patterns. This may also apply in Bald Eagles given the consistent differences in the PCB pattern between Johnstone Strait and lower Fraser valley eggs compared to those from the Strait of Georgia sites. Total PCB levels were also lower in the Johnstone Strait and lower Fraser valley eggs. The most likely explanation is of more fish in the diet of Fraser and Johnstone Strait eagles and thus greater exposure to the lower chlorinated PCBs. Higher total mercury levels at those two sites are also consistent with more fish in their diet. The PCA results can be used to support this explanation; however, alternatively the differences among sites may also be explained by differing local Aroclor inputs. Fraser delta eagle eggs, like Great Blue Heron eggs, contain more of PCB 66, indicative of Aroclor 1242 input, while the pulp mill areas, including Crofton, generally contain more Aroclor 1260 peaks, again similar to Great Blue Herons (Elliott et al. 1989a). A preponderance of lower chlorinated PCBs in the Johnstone Strait area may be indicative of greater atmospheric sources over local industrial inputs (Eisenreich et al. 1981). The single egg from the lower Fraser analyzed for non-ortho PCBs  ,  Herrling Island,  also had a lower ratio of PCBs 126:77 than eagle eggs from other areas, also suggesting higher consumption of fish which have low capacity to metabolize PCBs (Brown 1994). The ratio at most sites of non-ortho PCBs 126:77 was 2:1, except in the eggs from Alberni Inlet and Clayoqot Sound, where the ratio is closer to 1:1, and the egg from Herrling Island in the Fraser valley, where the ratio was 1:2. Although the ratios vary somewhat, the other non ortho PCB levels such as 169, are consistently less than either 77 or 126 in Bald Eagle eggs. Bosveld and Van Den Berg (1994) suggested that lower levels of PCB 77 in adult tissues compared to egg were caused by reduced metabolic capability in embryos. Levels of more  95  rapidly metabolized compounds such as PCB 77 may also be higher in eagle eggs as a result of direct deposition of dietary lipids to egg yolk, as suggested above for 2,3,7,8-TCDF. Comparison of total PCB levels to those in the literature is confounded by changes in methodology. Total PCB numbers in Wiemeyer et al. (1993) were probably determined as Aroclor estimates based on the analytical references. Determination of total PCBs based for example on Aroclor 1254:1260 overestimate total PCBs, based on the sum of congeners, by about two-fold (Turle et al. 1991). Toxicological significance of PCDD and PCDF levels The bioaccumulation model was developed in order to estimate critical concentrations of 2,3,7,8-TCDD and other contaminants in forage fish (eg. sculpin, perch and flounder species) or fish-eating birds (herons, cormorants, waterfowl), components of the foodchain which are more easily monitored than eagles. Levels in the monitoring species should indicate a degree of foodchain contaminant which should result in accumulation in bald eagle eggs less than the suggested NOEL from Chapter 2. Using the same BMIF of 21, the average 2,3 ,7,8-TCDD concentration in forage fish consumed by great blue herons in 1990 at Crofton would have been about 5 ng/kg. With the postulated eagle diet in Table 3.8,  TEQ5PCDD,PCDF. in  eagle eggs were calculated as 193 ng/kg  versus the measured value of 248 ng/kg. If an average value of 115 ng/kg TEQs 0 for non 0 were ortho and mono-ortho PCB contribution at Crofton, 1990 is included, the total TEQs 308 and 355 ng/kg, calculated and measured respectively, both of which exceed the LOEL (210 ng/kg), determined for Bald Eagle embryos (Chapter 2). If the data from Crofton, 1992, are used the estimated mean value of 1 ng/kg in forage fish gives a calculated  TEQWIIO  value in  eagle eggs of 194 ng/kg (79 TEQsPCDD,PCDFS + 115 TEQsPCBS), less than the LOEL, but still greater than the NOEL of 100 ng/kg. Therefore, assuming that both the ratios of other  PCDDs/PCDFs and PCB levels remain constant, a maximum value of 0.5 ng/kg 2378-TCDD in forage fish is suggested as site-specific dietary concentration in the Strait of Georgia, to avoid adverse toxic effects of TCDD-like chemicals in Bald Eagle populations. The corresponding concentration of 2,3,7,8-TCDD in Great Blue Heron eggs, to avoid TCDD  96  toxicity in both herons and top predators, such as the Bald Eagle, in the Crofton area is 10 ng/kg. At other areas in the Strait of Georgia, given that ratios of PCDDs, PCDFs and PCBs are similar, a value of 10 ng/kg in double-crested or pelagic cormorants, would also indicate that levels in local foodchains should not cause toxicity in Bald Eagles, given a typical diet, as shown in Table 3.8. The utility of colonial waterbirds as sentinel species for monitoring of toxic contaminants has been demonstrated in many studies (Gilbertson et al. 1987). Given that the embryonic life stage appears to be the most sensitive to TCDD-like effects (Peterson et al. 1993) and that the NOEL from Chapter 2 was derived using a very sensitive endpoint, CYP1A induction, then these critical values suggested for prey items, should provide a reasonable margin of safety. The above values would be effective in areas with contaminant profiles which are similar to the Strait of Georgia. However, as shown in Figure 3.3, in Common Tern eggs, PCDDs made only a minor contribution to the TEQs , relative to the non-ortho PCBs 0 (Kubiak et al. 1989; Harris et al. 1993). In yolksacs of fish-eating birds from the Netherlands, TEQs were also dominated by PCBs compared to PCDDs and PCDFs (Bosveld 1994; Van den Berg 1994). There are few published data on PCDD and PCDF levels in Bald Eagle eggs. Mean 2,3 ,7,8-TCDD levels in live fresh Bald Eagle eggs collected in 1985-87 from the lower Columbia river, were 32 ng/kg, less than those found in eagle eggs near pulp mills on the Strait of Georgia. However, total PCB levels were 12.7 mg/kg, more than two-fold higher than the highest mean concentrations in Table 3.4. Thus, TEQs 110 in Bald Eagle eggs from the Lower Columbia River would be dominated by the PCB contribution. Other studies have reported high PCB levels in Bald Eagle egg and plasma samples; however, because of correlations with DDE, no clear statistical relationships between PCBs and productivity were determined (Wiemeyer et al. 1984; 1993; Bowerman 1993, Dystra 1994; Welch 1994). Recent studies of PCB toxicity in other avian species have focused on the non ortho PCBs, particularly 126 and 77, and certain mono-oilho PCBs, such as 118 and 105, which are partial Ah-receptor agonists and thus cause TCDD-like toxicity in laboratory animals (Safe, 1990) and apparently in wildlife (Kubiak et al. 1989; Bosveld et al. 1994; Sanderson et  97  at. 1994b). However, the data on Bald Eagle chicks reported in Chapter 4 suggests that PCB congeners are less potent relative to PCDDs and PCDFs in Ah-receptor mediated biomarker responses, such as CYP1A induction. Nevertheless, total PCB levels up to 119 mg/kg have been reported in recent years in adled Bald Eagle eggs from the Great Lakes region (Bowerman et al, 1994); that egg would have contained about 18,500 ng/kg of PCB 126 using the regression from the Result section above. PCB concentrations of that degree may partly account for the poor productivity and reports of deformed young in the Great Lakes region. Although the data are not shown here, the same modelling approach can be used to determine total PCB concentrations in foraging fish and a sentinel fish-eating bird, which would result in a PCB contribution to TEQ5 in eagle eggs less than the NOEL of 100 ng/kg. Using the BMF for PCBs of 30 from Braune and Norstrom (1980), assuming constant ratios of non ortho and mono-ortho PCBs to total PCBs, for Crofton (assuming  TEQSDD/PCDFS  =  79 ng/kg)  site-specific values of 0.01 mg/kg in forage fish and 0.3 mg/kg in fish-eating bird eggs are suggested, would be necessary to achieve TEQs 0 less than the NOEL of 100 ng/kg in Bald Eagle eggs. This value for forage fish is much lower than 0.2-0.4 mg/kg suggested by Harris et at. (1993) to produce a NOEL in Forster’s Tern eggs in Green Bay, Michigan. However, eagles feed at a higher trophic level than terns; therefore, a lower target level in forage fish would be required to avoid accumulation of toxic levels in eagles. However, at most sites the contribution of PCDDs and PCDFs to TEQs is considerably less than at Crofton and is probably in the order of 25 ng/kg, in which cases a higher PCB contribution could be tolerated. Application of this or more sophisticated models to sites with both lower PCDDs and PCDFs and a comprehensive dataset on PCBs would enable determination of better guidelines for PCBs. Toxicological signjficance of organochtorine and mercury levels Wiemeyer et at. (1993) determined that DDE was the chemical contaminant most associated with reduced breeding success of Bald Eagles in the United States during the period 1969  -  1984. Production of young began to decrease at DDE levels > 3.6 mg/kg, and further  decreased at > 6.3 mg/kg. DDE levels of 16 mg/kg were associated with fifteen percent  98  eggshell thinning, a threshold related to population declines in other raptors (Noble et at. 1993). Wiemeyer et at. (1993) also found a highly significant relationship (r  =  0.912, p <  0.0001) between DDE and shell thickness in a large sample of Bald Eagle eggs from the United States. Mean DDE levels in the eggs in Table 2.2 were all less than 16 mg/kg, although 31 % (11/35) contained > 3.6 mg/kg and nine percent (3/35) had > 6.3 mg/kg. Although mean eggshell thickness was less than the pre-1946 mean at all sites, there was no significant relationship between DDE and eggshell thickness, likely because of the narrow range of DDE concentrations. Although quantitative data are limited, there were no reports of widescale declines of coastal eagle populations in British Columbia, as occured in other areas of North America during the organochlorine era. However, Vermeer et al. (1989) reported an increase in Bald Eagles nests in the southern Strait of Georgia between the mid-1970s and the late 1980s. They attributed eagle population growth to increased prey populations, particularly glaucous-winged gulls, populations of which had increased due to greater availability of human refuse. However, in the 1970s, DDE and other organochlorines were also likely much higher in Strait of Georgia eagle eggs. In Great Blue Heron eggs from a Fraser delta colony, DDE declined from a mean of 2.0 mg/kg in 1977 to 0.42 mg/kg in 1990 (Whitehead 1989; Canadian Wildlife Service, unpublished data). In Pelagic and Double-crested Cormorant eggs from Mandarte Island in the southern Strait of Georgia, DDE decreased by factors of five and ten respectively between the early 1970s and the late 1980s  (Elliott et at. 1989a). Organochlorine levels in  Bald Eagle eggs are currently about ten-fold higher than in those of marine and fish-eating birds from the Pacific coast (Elliott et at 1989a; 1989b). If the ten-fold difference was constant over time, then during the late 1970s mean DDE levels in Bald Eagle eggs from the Fraser delta would have been about 25 mg/kg, high enough to cause nest failures and reduced productivity. It is probable, therefore, that the population increase reported by Vermeer et at. (1989) was partly due to declining DDE levels. In the Okanagan valley of interior British Columbia, Bald Eagles declined as a breeding species between the 1930s and 1970s (Cannings et at. (1987). Although habitat loss was likely a factor, the extremely high DDE levels in  99  Okanagan valley foodchains (Elliott et al. 1994) probably continue to impact Bald Eagle reproduction in that area. None of the Bald Eagle eggs analyzed in this study had mercury levels > 0.5 mg/kg (wet weight), determined by Wiemeyer et al. (1993) to be associated with effects on productivity. In conclusion, Bald Eagle eggs collected in the Strait of Georgia contained elevated levels of PCDDs and PCDFs; the pattern was similar to that measured in other components of food chain and indicative of both bleached kraft pulp mill and chiorophenol sources. Relatively high PCDDs and PCDFs in a supposed reference area in northern Johnstone Strait probably resulted from feeding on waterbirds migrating north from the Strait of Georgia. Recommended site specific concentrations of 2,3,7 ,8-TCDD are 0.5 ng/kg in forage fish and 10 ng/kg in sentinel fish-eating bird eggs in the Strait of Georgia are suggested to avoid accumulation of potentially harmful levels in Bald Eagle eggs. Likewise, total PCB concentrations of 0.01 mg/kg in forage fish and 0.3 mg/kg in fish-eating bird eggs are suggested as maximum concentrations to prevent accumulation of potentially harmful PCB levels in Bald Eagle populations.  Acknowledgements Ian Moul, George Compton, Andre Breault, Dave Dunbar and Ray Caton assisted with collection of eggs. Mary Simon did the PCDD/PCDF and non-ortho PCB analysis, while Henry Won did the organochlorine pesticide analysis. John Smith provided statistical advice.  Funding was provided by the Canadian Wildlife Service and the British Columbia Ministry of Environment.  100  1  2  Bleached Production  470  550  2,125  1120  2,400  -  470  550  2,125  1120  1,100  (air dried tonnes)  ADt  Total Production  C,C,E,C,H, D,A  0 2 H  ,H 1 H 2  /C 8 D , 0 ,E 20 0  W,DC,R,EO, D,E,D  ,W, 0 C/D,E D,E,D  Bleaching Sequence  -  elememental chlorine, D chlorine dioxide, E 0 hot caustic extraction, W OEAS oxygen enhanced activated sludge CTMP chemi-thermomechanical pulp AAS aerobic activated sludge AS/ASB activated sludge/aerated stabilization basin  -  1918  Port Alice  -  CTMP Groundwood  1947  Port Alberni  -  Kraft CTMP 2 Groundwood  1912  Powell River  -  Kraft  1950  Harmac  -  Kraft Groundwood  1957  Crofton  c  Process  Initial Start-up  Mill  -  wash, H  -  Clarifier  Clarifier  Clarifier  AAS  5 AS/ASB  4 AAS  Aeration  1 OEAS  Secondary  -  1995  1990  1991  1991  5.5 (1989)  9.1 (1990)  6.0 (1989)  5.4 (1989)  pre CL-sub  <1.0 (1995)  2.2 (1993)  1.3 (1993)  1.1 (1993)  1.2 (1992)  post CL-sub  AOX (kg/Adt)  acid treat  1988-90  Chlorine Subs. linpiem.  sodium hypochlorate, A  Clarifiers  Clariflers Equaliser  Primary  Treatment  Neroutsos Inlet  Alberni Inlet  Malaspina Strait  Northumberla nd Channel  Stuart Channel  Receiving Environment  Table 3.9 Characterization of British Columbia pulp mills discussed in this paper. Data were obtained from Environmental Effects Monitoring (EEM) reports submitted by industry to Environment Canada.  CHAPTER 4 INFLUENCE OF CONTAMINANTS AND FOOD SUPPLY ON BALD EAGLE PRODUCTIVITY The results of the previous chapters showed that Bald Eagle populations in the Strait of Georgia were exposed to elevated levels of PCDDs and PCDFs relative to reference populations. Eggs collected in 1990 and 1991, particularly near Crofton, had higher PCDD/PCDF levels and modelling showed that theoretically eagles which preyed on a larger component offish-eating birds in the diet would acquire a substantial TCDD body burden. Among dead eagles examined between 1988 and 1993, about 20 % of a sample of 19 adults, found during the breeding season in the Strait of Georgia, contained TEQ5WHO > 1,000 ng/kg in livers. Thus, some component of the breeding population may be affected each year by chlorinated hydrocarbon toxicity. Eggs collected near pulp mills in 1992 and incubated in the laboratoiy did not exhibit significant effects on hatchability and most morphological and physiological endpoints, although a hepatic CYP1A cross-reactive protein was induced. For the work described in this chapter, I measured breeding success of Bald Eagles near three pulp mills in the Strait of Georgia, at two areas of the Fraser delta, and at reference sites on the west coast of Vancouver Island, in northern Johnstone Strait and in the Queen Charlotte Islands. The objective of the study was to determine occupancy of breeding territories, measure nest success and compare the results to chlorinated hydrocarbon levels in nestling plasma samples. Most previous studies of contaminants in Bald Eagles (for example Wiemeyer et al. 1993) used addled eggs, because of concern that collection offresh eggs would impact already poor reproduction. My initial studies on the coastal BC eagle breeding population (Chapters 2 and 3) used eggs collected during incubation; however, this resulted in an unacceptable level of nest abandonments, even when only two egg clutches were sampled and a single egg removed. Some researchers had previously used blood samples of nestling eaglets to obtain a more  102  randomized sample for contaminant analysis (Henny et a!. 1981; Frenzel 1985), an approach that has been used more frequently in recent years (Anthony et a!. 1993; Bowerman 1993; Dykstra 1994, Welch 1994). Blood sampling has the further advantage of not eliminating nests from productivity estimates from an area, and also permits determination of a direct relationship between contaminant levels in chicks and 5-year average productivity for the territory in which they were produced. Because of development of advanced high resolution gas chromatography/mass spectrometry (GC/MS) analytical techniques, beginning in 1993, the NWRC lab was able to quantify PCDDs, PCDFs and non-ortho PCBs in nestling Bald Eagle plasma samples.  Materials and Methods  Productivity Survey routes were flown in exposed or ‘treatment’ study areas, selected on the basis of eagle nest density near industrial pollutant sources: Crofton, Nanaimo and Powell River (pulp mills) and the lower Fraser valley (mixed industrial sources) (Figure 4.1). Reference or control sites were based on concentrations of nesting eagles remote from industrial point sources: Barkley and Clayoquot Sounds, northern Johnstone Strait and the Queen Charlotte Islands. Bald Eagle breeding success was estimated in each area by a standard two-flight approach (Fraser et al., 1983) using rotary aircraft (Bell jet/long ranger or Aerospatial Astar). A minimum of two observers were used. The first survey took place during incubation to determine the number of eagle pairs attempting to breed. Timing of this flight varied from late March in the Fraser delta to mid-May in the Queen Charlotte Islands. The second flight was timed to count nestlings at 5-8 weeks of age and took place between late May and early July. Mean productivity at each study area was determined by dividing the total number of young produced by the number of occupied breeding territories, as described in Postupalsky (1974).  103  Ce  Ce) CC  = ci  e C  Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections. At Langara Island, the survey circumscribed the coastline of the island.  104  Prey deliveries Prey deliveries were observed during 1995 at five nests in the Fraser delta and at four nests in Barkley Sound, using methods described in Dykstra (1994). From a blind, dawn to dusk observations were made using a 20-60X spotting scope. Prey deliveries and other nestling and adult behaviours were recorded. Observers were switched every four to eight hours. In the Fraser delta five nests were observed for five days each. In Barldey Sound, three nests were observed for five complete days each, while one nest was watched for part of each day and therefore was not included in statistical analyses. Sample collection Nests suitable for collecting were located by ground, boat and aerial surveys, when they were scored to estimate ground access and suitability for climbing; land tenure was also considered. Samples were collected when nestlings were 5-9 weeks old. Collections were made during the first week of June in the Fraser Valley and the Strait of Georgia, during late June or early July on the west coast of Vancouver Island, Powell River and Langara Island (Figure 4.2). Nests were accessed by a professional tree climber. Nestlings were lowered to the ground in a soft bag, weighed and aged by measuring the length of the eighth primary feather (Bortolotti 1984). Up to 24 ml of blood was drawn from the brachial vein (12 ml per wing) using a 12 ml sterile disposable syringe and a 21 gauge needle. Blood was transferred immediately to heperinized vacutainers and stored on ice. Samples were centrifuged within six hours of collection and plasma transferred to chemically cleaned (acetone/hexane) glass vials with teflon liners and then frozen. Chemical analysis Frozen plasma samples were shipped to the CWS National Wildlife Research Centre (NWRC) for analysis in the laboratory of Dr. R.J. Norstrom. For organochlorines, the samples (1 ml of each) were first deproteinized with 0.5 ml of methanol containing aldrin as an internal standard (Smrek et al. 1981). The plasma was then extracted with hexane and  105  •  LXZJ  Bald Eagle neat sites surveyed  Dioxin Fishery Closure Area  Powell River Pulp Mill  I  0  Crofton Pulp Mill  \rviaude I.  I  5  I  10km  VANCOUVER  .  ‘<—Winchelsea I.  Southey I.  centrifuged. The hexane extracts were passed through sodium sulphate, evaporated to 1 ml or less and separated into three fractions with hexane and methylene chloride on a florisil column. Analyses were performed by GC-electron capture detector with capillary-column separation on a Hewlett Packard 7673A. PCBs were quantitated as the sum of 33 major congener peaks. Quality assurance procedures included the simultaneous analysis of 6 diluted Herring Gull egg pool reference material samples (Tune et al 1991). Plasma samples (1.98  -  12.94 gram samples) were simultaneously analyzed for PCDDs,  PCDFs and non-ortho PCBs as follows: isotopically labelled internal standards (‘ 1 C 3 2 PCDDsIPCDFsInon-ortho PCBs) were added to the plasma, and allowed to equilibrate for 30 minutes. Saturated aqueous animonium sulphate and absolute ethanol were added to the spiked plasma, and the samples were then extracted four times with hexane. The hexane layers were combined, filtered through anhydrous sodium sulphate and the volume reduced for clean-up with by gel penneation chromatography (GPC) (Norstrom et a!. 1986). Lipids and biogenic materials were removed by GPC and alumina column clean-up. Separation of PCDDs, PCDFs and non-ortho PCBs from other contaminants was achieved using a carbon/fibre column (Norstrom and Simon 1991); further separation of PCDDs and PCDFs from the non-ortho PCBs was done with florisil column chromatography. Quantitation was performed with a VG Autospec double-focusing high resolution mass spectrometer linked to a HP 5890 Series II high resolution gas chromatograph. Recoveries of -PCDDsIPCDFs/non-ortho 12 PCBs were C 13 calculated by comparing the integrated areas of the labelled internal standards and the areas of the recovery standards in the samples to the areas of these compounds measured in an external standard mixture, analyzed along with the samples. Results were accepted when recoveries of -PCDDs/PCDFsInon-ortho 1 C 13 2 PCBs were between 70% and 120%. For a few Bald Eagle plasma samples, the internal standard recoveries were <70%, due to losses during lipid extraction. Lipid was determined by combining 1-2 ml of sample with 4 ml of hexane in a centrifuge tube, which was then extracted with an Ultra-Turrax homogenizer for 2 minutes. The contents of the tube were then centrifuged to separate the hexane and plasma layers,  107  similar to the method of Mes (1987). The hexane was then passed through sodium sulphate to remove any moisture. This process was repeated twice more and the sodium sulphate washed with hexane after the final extract. The three hexane extracts were combined on a pre-weighed aluminum dish, the hexane was then evaporated and the dish re-weighed to determine the amount of lipid. Lipid was then calculated on the basis of grams per ml plasma. Statistical analyses The SYSTAT software package was used for statistical analyses of all data. Wet weight chemical residue data were transformed to common logarithms and geometric means and 95 % confidence intervals were calculated with the data grouped by collection site. The majority of chlorinated hydrocarbons tested were significantly correlated with percent plasma lipid (Table 4.1). DDE was only weakly correlated with plasma lipid, while the higher chlorinated PCDDs and PCDFs were not significantly correlated. There was also a significant interaction between plasma lipid and sampling location. Therefore, for testing of differences among locations, all of the contaminants which correlated significantly with plasma lipids, were further transformed using an analysis of covariance (ANCOVA) to account for the effect of variation in plasma lipid content among individuals and locations (Hebert and Keenlyside, 1995). Differences among locations were then determined using Bonferroni’s test. In a few cases, percent plasma lipids were three to ten-fold greater than the mean of the other samples at that site; those samples were fatty in appearence and the nest contained fresh, partly eaten prey remains, indicating that the chick was sampled during or immediately after feeding. Those ‘outliers’ were not removed from the data, rather, it was assumed that they were corrected by the ANCOVA. Productivity measures were compared among locations with a one-way analysis of variance (ANOVA); significant differences were determined using Tukey’s multiple comparison procedure (MCP). Data were also compared on the basis of a pulp mill versus non-pulp mill grouping and significant differences identified using Student’s t-test. At each pulp mill site,  108  C  TEQa  PCB— 169  PCB— 126  PCB—77  PCB—37  PCB— 180  PCB— 153  PCB—118  PCB—99  SUM— PCBs  1— nonachlor  Mirex  HLB  DDE  TCDF  PnCDF  HxC  OCDF  TCDD  PnCDD  HXCDD  HpCDD  OCDD  Lipid  Lipid  —0.025  0.771  0.041  0.835 —0.035  —0.008  0.962 0.081  0.179  —0.005  0.964 —0.007  0.896  —0.047 —0.028  0.134  0.961 —0.004  TCDD OCDF  —0.058 —0.035  0.872  OCDD HpCDD HXCDD PnCDD  0.261  0.040  0.056  0.066  0.587  0.791  0.032  HxCLE  0.057  —0.021  0.927  0.884  0.841  —0.029  —0.031  0.942  PnCDF 0.569  DDE 0,868  HCB 0,898  0.983  0.923  0.941 0.532  0.528  0.681  0.799  0.854  0.897  0.624  0.872  0.866  0.934  0.849  0.654  0.956  0.938  —0.018  0.904  0.862  0.851  —0.062  0.007 —0.101 —0.042 —0.045  0.010  0.744  0.784  —0.034  0.580  0.534  0.634  —0.060  0.969  —0,008 —0,065 —0.055 —0.074  0.967  0.943  0.476  —0.085 —0.059 —0.101 —0.087 0.886  trans Mirex nonaclor  —0.077 —0.046 —0.077 —0.074  0.931  TCDF  0.995  0.942  0.844  0.629  0.955  0.931  0.005  —0.025  0.986  0.938  0.870  —0.040  —0.034  0.961  —PCBs  Sum  0.994  0.993  0.942  0.864  0.625  0.953  0.943  —0.006  —0.035  0.977  0.909  0.830  —0.043  —0.040  0.978  99  0.998  0.990  0.995  0.942  0.855  0.618  0.955  0.942  0.009  —0.028  0.983  0.925  0.851  —0.027  —0.027  0.973  118  0.991  0.996  0.998  0.994  0.948  0.845  0.618  0.949  0.936  0.004  —0.030  0.980  0.926  0.857  —0.037  —0.033  0.964  153 37 0.238  0.299  0.989  0.984  0.996  0.987  0.950  0.830  0.618  0.949  0.917  0.004  —0.023  0.982  0.949  0.888  0.394  0.289  0.284  0.278  0.300  0.316  0.200  0.135  0.091  0.444  0.364  0.054  0.183  0.328  0.393  0.442  —0.051 —0.170  —0.046 —0.082  0.942  150  Table 4.1 Correlation Matrix (r value) for percent plasma lipid and selected chlorinated hydrocarbon in bald eagle nestlings from British Columbia, 1993—94  77  0.394  0.961  0.969  0.965  0.959  0.963  0.959  0.909  0.792  0.565  0.925  0.943  0.048  —0.031  0.943  0.908  0.852  —0.008  —0.001  0.919  126  0.983  0.324  0.970  0.978  0.975  0.973  0.972  0.973  0.940  0.853  0.608  0.925  0.958  0.032  —0.051  0.954  0.908  0.840  —0.050  —0.046  0.943  169  0.958 —0.071  TEQs Productivity  0.964  0.933  0.330  0.900  0.906  0.901  0.899  0.895  0.894  0.925  0.865  0,544  0.860  0.929  0.043  0.038  0.014  0.135  0.927  0.987  0.977  —0.080  0.004  0.074  0.094  0.343 —0.094  0.986 —0.137  0.990 —0.114  0.990 —0.103  0.985 —0.104  0.990 —0.119  0.987 —0.122  0.932 —0.142  0.840 —0.128  0.600 —0.349  0.962  0.959  0.050  0.182  0.986 —0.074  0.952 —0.070  0.893 —0.086  —0.065 —0.019  0.870  0.842  0.779  —0.095 —0.005 —0.073  —0.081 —0.015 —0.083  0.867  productivity at nests adjacent to dioxin fishery closure areas was compared to nests adjacent to areas outside the closure area, using a one-way ANOVA. We treated the closure areas as an indication of the area impacted directly by PCDD and PCDF contaminants in the respective pulp mill effluents. Mean 3-year productivity at individual nests was also compared to contaminant levels in nestling blood samples from each nest using regression analysis. Unless stated otherwise, a value of p < 0.05 was considered statistically significant in all analyses. TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalency factors (TEF5) proposed by Ahlborg et al. (1994) and referred to here as the WHO (World Health Organization) TEFs.  Results Productivity Mean three-year productivity was highest at nests in the Fraser valley and delta and comparable along south-east Vancouver Island (Table 4.2). The number of young/occupied territory was lower (significantly compared to the lower Fraser valley) at nests around Powell River and at Langara Island. Lowest productivity was in Clayoquot Sound, Johnstone Strait and South Moresby. Productivity of eagle nests located along the shoreline adjacent to the dioxin fishery closures in the Crofton area was significantly lower than at nests located outside the closure area (Figures 4.2 and 4.3). There were no significant differences in productivity at nests adjacent to the closure areas compared to those outside the closure areas at both Nanaimo and Powell River. However, the four eagle nests closest to the mill on the north side (Powell River nest and three Gibson’s Beach Park nests) produced only two chicks between 1992 and 1994 in nine nesting attempts. In contrast the next five nests to the north (three Scuttle Bay nests, Kees Bay and Lund) during the same time frame produced 21 chicks in 15 nesting attempts. This difference, was not statistically significant, however, likely due to small sample sizes.  110  Table 4.2 Nest success and production of young for Bald Eagles at nine study areas on the British Columbia coast (1992-94). Study Area  Lower Fraser Valley  Year  No. occupied territories  Successful Nests  % Nest Success  No. young produced  1992 1993 1994  19 22 21  19 18 18  100 82 86  27 27 29  1.4 1.2 1.4 1.3a  12 14  1.3 1.3  Mean Fraser Delta  1993 1994  89 9 11  7 9  Mean South-east Vancouver Island  1991 1992 1993 1994  86 19 30 34 42  11 19 22 27  Mean Powell River  1992 1993 1994  1992 1993 1994  24 37 36  14 25 21  1992 1993 1994  36 35 30  16 20 8  1991 1992 1993 1994  23 43 35  12 10 2  2 10 3 13  6 26 34 31  0.75 0.97 0.92 0.88&  44 57 27  21 26 12  0.58 0.74 0.4 0.57c  52 23 57  12 14 3  33 39 8.8 42  1994  19  Langara Island  1994  22  5 13  0.52 0.33 0.09 0.31d  2 12 4 14  0.33 0.46 0.12 0.45 0.34’  31  South Moresby  -  18 36 33  27  Mean  a,b,c,d  58 68 58  0.90 1.00 1.00 1.00 0. 97ab  43  Mean Johnstone Strait  17 30 35 43  61  Mean Clayoquot Sound  58 63 65 64  l.3  63  Mean Baridey Sound  78 82  Young! occupied nest  26  6  0.32’  59  16  0.73  means in the column that do not share the same letter are significantly different (p<O.O5)  111  1.2 4-,  Cl)  ci)  z -c,  a)0.8 a  D C.) C)  o  0.6  Outside D 0  :  Inside  >-  o  Z  0.2 0  0  Figure 4.3 Bald Eagle productivity (mean and SD) compared between samples of nests located adjacent to shorelines inside and outside of dioxin fishery closure areas on the British Columbia coast. Sample sizes were: Powell River, N=20 inside and N=26 outside; Nanaimo, N =15 inside and N =13 outside; Crofton, N=9 inside and N=8 outside.  No significant correlations occurred between productivity and any of the PCDD, PCDF or PCB compounds measured or with TEQs (Figure 4.4a). For the organochiorine pesticides, log-DDE in nestling plasma regressed weakly with 3-year average productivity for each corresponding territory (r 2 = 0.128, p < 0.011, Figure 4.4b).  112  A 2.5  -  ••  2-  .  ci  C  •  .1  A D 4 0 AALI.  )  LZJs.c>  1-  A A I  0.1 B  I  1111111  11111111  1 10 TEQs WHO (ng/kg wet weight)  100  -  2.5L  ci)  ci) >  U 0  U)  E’i-  .  0  +1  2  A  A  0  0—  111111  1111111  ‘‘I  1111111  10 100 DDE (ug/g wet weight)  1  •  E. Van. I.  A  Barkley Sd.  ü  Johnstone Str.  •  Powell R.  •  Clayoquot Sd.  0  Fraser Delta  1000  A  Low. Fraser Vafley  Langara I.  Figure 4.4 Productivity at Bald Eagle nest sites on the British Columbia coast as a function of contaminant concentrations in plasma samples in nestlings raised in that territory, for: A) the log of , B) the log of DDE. The subpopulations included: East Vancouver Island, Powell River, 0 TEQs Barkley Sound, Clayoquot Sound, Johnstone Strait, Fraser Delta, Lower Fraser Valley, and Langara Island. 113  Mean percent lipid in plasma regressed positively on mean productivity among sites (Figure 4.5).  1.4  r2=O.423 1.2 0  a)  I  a) 0.8 D C.) 0  U  oO.6 c3) C 0  >0.2  0 0.01  log (% plasma lipid)  0.1  Figure 4.5 Comparison of mean productivity of Bald Eagles at sites on the coast of British Columbia with the mean percent lipid in plasma samples of nestling eagles at each site. Sampling sites were: Fraser Delta (N=5), Lower Fraser Valley (N=5), East Vancouver Island (N= 12), Powell River (N= 10), Barkley Sound (N=9), Clayoquot Sound (N=3), Johnstone Strait (N=4), Langara Island (N=5).  Prey deliveries At surveyed nests in the Fraser Delta, mean daily prey deliveries were greater at, 3.5 than at Baridey Sound nests, 2.4 deliveries per day; however the difference was not statistically significant, likely due in part to small sample sizes in this pilot study.  114  PCDD and PCDF levels in plasma Plasma samples from 52 Bald Eagle chicks were analyzed for PCDD and PCDF levels (Table 4.3). The pattern in plasma near pulp mill sites was generally 2,3,7, 8-TCDF > l,2,3,6,7,8-HxCDD > 1,2,3,7,8-PnCDD > OCDD > 2,3,7,8-TCDD. At other sites, OCDD was often comparable or greater than 1 ,2,3,6,7,8-HxCDD, while in the Fraser delta, OCDD and 1,2,3,4,6,7,8-HpCDD were the dominant congeners. Most samples also contained detectable amounts of 2,3 ,4,7,8-PnCDF. Because of the significant interaction with plasma lipid content, selected PCDDs and PCDFs are further presented as lipid-adjusted log-normalized mean values (Figure 4.6). Mean plasma TCDD concentrations were significantly higher at Powell River and East Vancouver Island than other sites. Mean concentrations of PnCDD, HxCDD and TCDF were also highest near the pulp mill sites at Powell River and along east Vancouver Island; however, the differences were not consistently significant from the Fraser Delta and Johnstone Strait. Highest mean levels of HpCDD and OCDD occurred in samples from the Fraser Delta, although the mean was not significantly different from east Vancouver Island.  PCBs in plasma Highest concentrations of total PCBs were in samples from Powell River and east Vancouver Island (Table 4.4), which on a lipid-adjusted basis were significantly greater than Clayoquot Sound and the Fraser valley (data not shown). Mean concentrations of individual PCB congeners generally followed the geographical pattern of the total PCBs; for example, highest concentrations of PCBs 153 (245-245) and 105 (234-34) were also at Powell River and east Vancouver Island and were significantly different from Clayoquot Sound and the lower Fraser Valley.  115  8  3  4  5  Barkley Sound  Clayoquot Sound  Johnstone Strait  Langara Island  *  -  ND  0.37 0.17-0.80  10  Powell River  0.08 .02-.28  *  0.02  0.01 .01-.03  0.02 .01-.05  0.90 .51-1.6  0.623 .37-1.1  -  0.05 0.02-0.13  0.05 0.01-0.19  0.37 0.12-1.1 0.03 ND-0.25  0.11 0.02-0.52  0.08 0.03-0.26  0.04 0.03-0.06  0.13 0.07-0.23  2.2 1.2-3.9 0.07 0.03-0.17  0.31 0.20-0.48  0.09 0.04-0.23  0.07 0.02-0.32 1.2 0.59-2.4  1.7 0.12-25  1234678 HpCDD  0.45 0.08-2.5  123678 HxCDD  Not detected, minimum detection limit 0.01-0.05 ng/kg, wet weight. values all the same  0.04 0.01-0.01  0.03 0.01-0.12  0.01 0-0.02  0.01 0.01-0.03  0.33 0.21-0.53  11  EastVancouverlsland  0.14 .1-.19  0.05 0.04-0.06  5  Lower Fraser Valley  0.23 .09-.59  0.07 0.04-0.31  5  Fraser Delta  12378 PnCDD  N  Location  2378 TCDD  0.18 0.11-0.28  0.42 0.23-0.76  0.64 0.06-6.4  0.57 0.36-0.92  0.56 0.40-0.77  1.1 0.7-1.6  0.30 0.19-0.47  2.4 0.19-31  OCDD  0.18 0.05-0.66  0.97 0.23-4.1  0.19 0.04-1.0  0.24 0. 17-0.34  4.54 2.69-7.66  2.8 2.0-3.8  0.19 0.15-0.25  0.11 0.02-0.69  2,3,7,8 TCDF  0.04 0.01-0.14  0.06 0.03-0.12  0.14 0.09-0.21  0.16 0.08-0.32  0.12 0.07-0.21  0.15 0.10-0.23  0.02 0.01-0.04  0.07 0.01-0.41  12378 PnCDF  0.08 0.01-0.47  0.09 0.02-0.34  0.14 0.08-0.23  0.16 0.08-0.32  0.27 0.13-0.55  0.21 0.13-0.35  0.02 0.01-0.04  0.11 0.04-0.28  23478 PnCDF  0.11 0.09-0.30  0.06 0.05-0.08  0.09 0.04-0.18  0.06 0.04-0.09  0.09 0.06-0.14  0.16 0.12-0.35  0.12 0.04-0.35  0.16 0.04-0.61  234678 HxCDF  0.10 0.04-0.28  0.04 0.01-0.18  0.07 0.04-0.12  0.05 0.03-0.09  0.05 0.03-0.09  0.16 0.07-0.38  0.08 0.04-0.17  0.13 0.07-0.25  OCDF  Table 4.3 PCDD/PCDF levels, geometric means and 95% confidence interval (ng/kg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94.  2378-TCDD  2378-TCDF  1 2378- PnCDD  23478- P, CD F  -U  a)  ci, -U -U  U) U)  1 23678-HCDD  OCDD  2.5 a-  -  2  1.5  a  nnm b  \\ii  \‘•  —  -.  ,.  bb  \  Figure 4.6 Residue levels of selected PCDDs and PCDFs in plasma samples of Bald Eagle nestlings collected on the British Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.3. Means that do not share the same lower case letter are significantly different (p <0.05).  117  Table 4.4  Organochiorine levels, geometric means and 95% confidence interval (tg/kg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94.  Location  N  Total PCBs  DDE  transnonachlor  Oxychiordane  Dieldrin  Mirex  HCB  Fraser Delta  5  17.8 4.6-69.6  14.4 8-26  0.5 0.1-2.6  0.3 ND-1.8  0.1 ND-0.1  0.1 ND-0.3  0.2 0.1-0.8  Lower Fraser Valley  5  11.2 6.4-19.7  9 4-20.3  0.5 0.3-0.7  0.1 ND-0.1  0.1 ND-0.1  0.1 ND-0.1  0.3 0.2-0.5  East Vancouver Island  10  30 18.8-47.5  11 0.6-17.3  2 1.2-3.1  0.2 0.1-0.3  0.2 0.1-0.3  0.1 0.1-0.2  0.3 0.2-0.5  Powell River  10  56 27-114  20.2 8.3-50  3 1.4-6.4  0.4 0.1-1.5  0.2 0.1-0.8  0.3 0.2-0.7  0.6 0.3-1.0  Barkley Sound  8  20 14-28.5  21.1 6.9-64.5  1.3 0.8-2  0.1 ND-0.4  0.1 ND-0.1  0.1 0.1-0.3  0.3 0.2-0.6  Clayoquot Sound  3  6.8 1.9-24.2  6.6 1.8-24  0.3 0.1-0.7  0.1  0.1  0.1  *  *  *  0.3 0.1-0.7  Johnstone Strait  4  14.3 6.2-33  7.3 2.7-19.4  1.2 0.8-1.9  0.1 ND-0.1  0.1 ND-0.1  0.1 ND-0.2  0.4 0.2-0.5  Langara Island  5  16.4 6.3-43  22.3 5.8-86  1.1 0.6-2.0  0.9 0.5-1.7  0.1 ND-0.4  0.3 0.1-1.4  0.8 0.3-2.2  ND * -  Not detected, minimum detection limit 0.01-0.05 nglkg, wet weight. values all the same -  In Bald Eagle plasma samples the general pattern of non-ortho PCB congeners was: 77 (34-34)  37 (34-4)  >  126 (345-34)  >  169 (345-345)  >  81(345-4) (Table 4.5). Highest  lipid-adjusted mean concentrations of individual congeners were generally in samples from Powell River or east Vancouver Island, although the highest mean concentrations of PCB 169 were from Langara Island (Figure 4.7). Organochiorines in plasma. Highest mean organochlorine pesticide levels were in samples from the Strait of Georgia region, including the Fraser Delta and from Langara Island (Table 4.4). Most lipid-adjusted plasma OC levels did not differ significantly among sites. Mean oxychiordane levels were significantly greater at Langara Island than either Johnstone Strait or the lower Fraser Valley.  118  PCB 126  PCB 169  PCB 37  PCB 77  .  —.  ‘,.  .  c,&  q•  c  \G3  Figure 4.7 Residue levels of selected PCBs in plasma samples of Bald Eagle nestlings collected on the British Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.4. Means that do not share the same lower case letter are significantly different (p <0.05).  119  Table 4.5 Non-ortho PCB levels, geometric mean and 95% confidence interval (nglkg, wet weight) in blood plasma of Bald Eagle chicks from the coast of British Columbia, 1993-94.  Location  N  PCB 37  PCB 81  PCB 77  PCB 126  PCB 169  PCB 189  Fraser Delta  5  1.80 0.94-3.45  0.71 0.22-2.35  9.45 2.66-33.5  4.01 1.29-12.5  0.68 0.21-2.18  0.13 0.05-0.37  Lower Fraser Valley  5  1.01 0.53-19.95  0.59 0.32-1.07  5.63 3.22-9.84  2.5 1.66-3.75  0.42 0.24-0.72  0.09 0.04-0.22  East Vancouver Island  11  29.4 24.5-35.2  0.98 0.64-1.51  16.6 12.5-22.1  6.06 2.98-13.3  1.42 0.24-6.72  0.53 0.38-0.74  Powell River  10  18.3 13-25.7  1.49 0.82-2.71  26.1 13-52.3  14.8 7.01-31.3  3.39 1.88-6.12  0.58 0.23-1.49  Barkley Sound  8  10.1 7.0-25.7  0.76 0.45-1.29  7.27 5.74-9.2  4.75 3.24-6.96  0.64 0.30-1.34  0.23 0.11-0.46  Clayoquot Sound  3  9.84 6.31-15.4  0.26 0.01-6.02  5.77 0.89-37.5  0.98 ND-212  0.52 0.05-5.17  0.16 0.07-0.38  Johnstone Strait  4  7.70 2.68-22,1  0.45 0.14-1.40  4.77 1.52-14.9  1.46 0.31-6.72  0.40 0.07-2.43  0.20 0.05-0.75  Langara Island  5  0.81 0.31-2.13  0.51 0.14-1.84  5.04 1.33-19  6.29 1.69-23.5  2.52 0.71-8.97  0.10 0.03-0.27  ND  -  Not detected, minimum detection limit 0.01-0.05 ng/kg, wet weight.  Mean mirex concentrations were significantly greater at Langara Island and Powell River than the lower Fraser Valley.  Discussion Higher concentrations of chlorinated hydrocarbons in Bald Eagle nestlings from the Strait of Georgia were not associated with significant effects on breeding success at most sites. With the exception of a sample of nests near Crofton, mean 3-year productivity at study sites around the strait, particularly the estuary of the Fraser river, was substantially higher than the 0.7 young/occupied territory, necessary to sustain an eagle population (Sprunt et al. 1973). In contrast, eagle productivity at the more remote reference sites was generally less than 0.7. Only at Langara Island at the north end of the Queen Charlotte archipelago, an area of high biological productivity, was eagle breeding success comparable to the Fraser delta and the Strait 120  of Georgia. Using nestling plasma lipid content as a marker of body condition, food supply is likely the main factor limiting eagle productivity on the British Columbia coast. However, low productivity at a sample of eagle nests adjacent to the dioxin fishery-closure zone at Crofton is probably not caused by differences in food availability. The geographic pattern of PCDDs and PCDFs in plasma is similar to that found in eagle eggs and is discussed in detail in Chapter 3. Essentially, elevated levels of TCDD, PnCDD, HxCDD and TCDF are associated with pulp mill sources. Elevated HpCDD and OCDD in the Fraser delta samples likely reflect heavy past use of chlorophenolic wood preservatives in that area, and some contribution from combustion sources. There are few published data on PCDD and PCDF levels in avian plasma. Blood samples of osprey nestlings taken in 1992 downstream of a bleached-kraft pulp mill on the Thompson River, in the interior of British Columbia, did not contain any lower chlorinated dioxins and furans (minimum detection limit  =  0.5 ng/kg, wet weight); sample sizes were  small, however, averaging about 3.6 ml of plasma. OCDD and HpCDD (0.1  -  1.0 ng/kg)  were detected in most osprey samples (Norstrom and Simon 1994). Osprey eggs from the same sites in 1991 contained relatively high concentrations of TCDD, TCDF, HpCDD and OCDD (Whitehead et al. 1993). In bald eagles, five of the six non-ortho compounds displayed a good correlation with plasma lipid content, while PCB 37 was only weakly correlated with plasma lipid. Ratios of PCB 37 relative to other congeners were high in eagle plasma compared to eggs or liver. High ratios of PCB 37 to other non-ortho PCB congeners were also found in osprey samples (Norstrom and Simon 1994) This suggests that PCB 37 may bind with plasma proteins. Corraborative data on PCDDs, PCDFs or non-ortho PCBs in avian blood samples from other studies is unavailable. However, studies of human subjects have shown that, although absolute levels on a lipid weight basis were much lower than those found in the eagle samples, OCDD was the major congener present (Papke et a!. 1990). In humans, blood:adipose tissue ratios are highest for OCDD compared to other PCDDs and PCDFs (Schechter et al. 1990). As we found  121  with eagles, OCDD did not partition with lipid in human blood; it is believed to bind primarily to serum protein components (Patterson et al. 1989). Published data on total PCBs and DDE in avian plasma samples is available from a number of studies. Mean concentrations of PCBs and DDE in plasma of nestling Bald Eagles from the lower Columbia River, 1984-86 were 0.04 and 0.05 mg/kg, wet weight, respectively, (Anthony et al. 1993); those levels were comparable to eagle plasma samples from Powell River and east Vancouver Island nests. Meanwhile, PCB and DDE levels in eggs were about three-fold higher in eagle eggs from the lower Columbia compared to the Strait of Georgia (Anthony et al. 1993; Chapter 2). However, plasma lipid levels were not reported for the lower Columbia; therefore, the low levels of PCBs and DDE in those samples may reflect low plasma lipid levels. Geometric mean levels of DDE and PCBs (wet weight) in eagle plasma samples collected between 1987 and 1993 from less contaminated areas of the Great Lakes were comparable to samples from our reference sites: DDE, 3-12 ng/kg and total PCBs 5-34 ng/kg (Bowerman 1993; Dykstra 1994).  Levels of DDE in eaglets from most shoreline areas of the  Great Lakes, 20-25 ng/kg, were comparable to data for the Strait of Georgia and Langara Island. Eaglets from Lake Michigan had somewhat higher levels, 35 ng/kg, DDE, than other sites. Mean levels of total PCBs in nestling eagle blood samples from the Great Lakes shoreline were two-fold (Lake Superior) to four-fold (Lake Erie) higher than Strait of Georgia samples. Maine eagle blood samples, 1991-1992, particularly from estuarine sites, had up to 150 ng/kg DDE and 1,250 ng/kg total PCBs (Welch 1994). However, plasma lipid data were also not reported for either the Great Lakes or Maine samples. The potential influence of geographic variation in plasma lipids on contaminant levels is particularly relevant for some Great Lakes samples, as Dykstra (1994) determined that low food availability was the main cause of poor breeding success at the Lake Superior nests, compared to those inland. This was reflected in lower rates of prey delivery to nests, greater time spent away from the nests by adults and increased time spent by nestlings sleeping and resting. Concentrations of DDE, but not PCBs, in nestling plasma samples from Lake Superior also regressed negatively on mean 5-  122  year productivity at the respective territories, indicating that DDE may still have been a factor contributing to low productivity. Low eagle productivity at certain areas of the British Columbia coast, such as Barkley and Clayoquot Sounds, Johnstone Strait and South Moresby may also be caused by low food availability. Mean plasma lipids were significantly lower in nestlings from those sites, indicating chicks in poorer body condition. The significant association among sites between productivity and mean percent plasma lipids also suggests that in productive areas, chicks are fed more regularly, are in better body condition and are more likely to survive to fledging. Breast muscle of eagle chicks found dead at inland nests near Lake Superior had higher mean fat content than those found at shoreline nests (Kozie and Anderson 1991). The pilot study on prey deliveries failed to show a significant difference between samples of nests in the Fraser Delta and Barkley Sound, although there were significant differences between those sites in both mean 3-productivity and percent plasma lipids in nestlings. However, because of logistical difficulties in observing nests at more remote areas of the coast, where productivity is particularly low, observations in Barkley Sound were made at nests which tended to be more accessible and to have higher productivity. Food supply during breeding is a major factor affecting avian productivity, including raptors (Newton, 1980; Gardarsson and Einarsson 1994). In addition to Dykstra’s (1994) study of eagles, Shutt (1994) related breeding failure and poor body condition of both herring gull chicks and adults to lack of food at Lake Superior breeding colonies. Prey availability was critical to productivity of white-tailed sea eagles (Helander 1985), European sparrowhawks (Accipiter nisus) (Newton et al. 1986) and ospreys (Van Daele and Van Daele 1982). A minimum food supply was required for successful breeding of wedge-tailed eagles (Aquila audax) in Australia, while Hansen’s (1987) experiment showed that Bald Eagle nesting and fledging success could be increased by providing additional food. Bald Eagle breeding densities in Saskatchewan were related to availability of key prey species, which correlated with primary productivity (Dzus and Gerrard 1993). Fish eating birds, particularly gulls, are important prey species to north west eagles (Knight et al. 1990).  123  On the west coast of Vancouver Island, colony sizes and breeding success were lower for gulls and cormorants (Vermeer et al. 1992) than the Strait of Georgia with its more stable food regime (Vermeer et a!. 1989). The steep fjord-like topography of the shoreline and the islands of the west coast of Vancouver Island, Johnstone Strait and Moresby Island also limits prey availability and foraging opportunities, compared to the beaches and tidal mudflats of the Strait of Georgia, which harbour abundant bird populations (Vermeer 1983). Food concentrated along the highly productive La Perouse Bank, to the west of Barkley Sound is beyond the reach of Bald Eagles. Langara Island is the only site outside the Georgia basin with relatively high eagle productivity. This island lies at the bottom of the Alaska gyre, an area of summer upwelling (Thomson 1981), which creates high marine productivity, evident by a rich fauna of salmonids, seabirds and cetaceans. Low eagle productivity in Barldey and Clayoquot Sounds and Johnstone Strait is characterized by a high incidence of failed nesting attempts. Many nests had incubating adults during the activity flight, but were empty during the productivity flight. Without nest observations throughout the breeding cycle, we cannot determine at what stage those attempts failed, although some nests certainly failed during incubation, as we often observed nests with abandoned eggs during the later flight. A high incidence of nest failures, indicated by the ‘fledging ratio’ (young per successful nest/young per occupied nest) has been suggested as a criteria for contaminant impact on an eagle population (Colborn 1991). The fledging ratio was as high as 11 in bad years in Clayoquot Sound, where, at least PCDD/PCDF levels are lower. High rates of nest failure in those areas is probably caused by the presence during nest initiation in March and April of abundant food resources, such as Pacific Herring spawn (Clupea harengus) (Hay et a!. 1992) and wintering waterbird prey (Vermeer and Morgan 1992), which are not available in May and June and are not adequately replaced by other food items. With the present data, it is difficult to determine why eagle productivity is low in the Crofton area. In contrast to Clayoquot Sound and other areas, eagle nesting near Crofton should not be food stressed. A number of the Crofton area nests are situated on small islands  124  (Shoal and Willy Islands), virtually in the estuary of the Chemainus River. Numerous waterbirds, including flocks of several hundred White-winged Scoters (Melanitta fusca), feeding on the abundant shellfish, are present during the breeding season. Eagle productivity is also high in the area immediately to the north, where major habitat differences are not apparent. It is conceivable that in the immediate past, PCDD and PCDF exposure at Crofton and also possibly Powell River, Nanaimo and other pulp mill sites affected bald eagle reproduction. Health effects in Great Blue Herons at Crofton were attributed to PCDD and PCDF exposure in the late 1 980s (Elliott et al. 1989a; Sanderson et al. 1994a). Based on the extrapolation in Figure 4.8, levels of 2,3,7,8-TCDD and other chemicals would likely have been even higher in eagles than herons. PCDDs and PCDFs in eagle eggs collected in 1990 and 1991 and on a lipid-adjusted basis in the one eagle plasma sample from Crofton were comparable to those from Powell River, yet a reduction in mean productivity in the dioxin fishery closure area was not found. However, for pragmatic reasons, fishery closures from persistent pollutants such as dioxins must be defined over broad areas, even though there are wide gradients in contamination within the zones (Harding and Pomeroy 1990). For example, higher PCDD and PCDF concentrations were consistently found in invertebrates collected to the north than to the south of the Powell River mill (Dwernychuck et al. 1994). This corresponds, perhaps coincidentally, with poor productivity at the four eagle nests immediately north of that mill. At Crofton, eagle productivity was also particularly poor at Shoal and Willy Islands, the nests closest to the Crofton mill; those nests have often been active, but have rarely produced chicks. Adult eagles, presumed to be from nests near the pulp mills, have been observed to forage in the heron colonies at Crofton and Powell River (Norman et a!.; C. Burton, person. comm.), which would cause very high PCDD exposure (Chapter 3). However, by 1991 when the first eagle productivity surveys were done, PCDD and PCDF concentrations in fish eating birds at Crofton had decreased by an order of magnitude from the high levels of the late 1980s (Whitehead et a!. 1 992a; Figure 4.8). The rapid decline of PCDDs/PCDFs in fish-eating birds was ascribed to their feeding primarily on small fish, including many young-of-the-year age classes, in which reductions in local contaminant inputs  125  would be more quickly apparent. Sample sizes are small, nevertheless, mean PCDD/PCDF levels in eagle eggs decreased between 1990 and 1992 at Crofton and Nanaimo, although possibly at a slower rate than in herons and cormorants. As larger animals feeding at a higher trophic level, clearence of TCDD and other compounds may occur more slowly in eagles.  300  extrapolated  Ii  Heron  -*-  Eagle  4-.  -c  •Eagle  ci)  Cormorant  4-.  ci  -  150 U U I— 100 co I-. C) C”  50  0  1987  1988  1989  1990  1991  1992  1993  Figure 4.8 Trends in 2,3,7, 8-TCDD in eggs of eagles, herons and cormorants at Crofton, British Columbia. The likely trend in eagles is extrapolated back to 1987, based on the mean 2,3,7,8-TCDD ratio of eagles:herons, 1990-1992. Assuming that poor productivity at Crofton is contaminant-related, it is also conceivable that some adult eagles suffer chronic reproductive impairment due to past high PCDD/PCDF exposure in ovo or during early growth and development. Rats and monkeys, of both sexes, dosed with < 1 ug/kg of TCDD display abnormal reproductive function in laboratory studies (Peterson et a!. 1993). For example, rhesus monkeys fed 25 ppt of TCDD, showed significant  126  reproductive impairment, but no apparent health problems (Bowman et al. 1989). Male rats exposed both in utero and lactationally to as little as 0.064 ug/kg TCDD via maternal dosing had damaged reproductive systems (Mably et al. 1992); however, fertility was not affected. Mably et al. speculated that the high critical sperm volume of the rat would mitigate against reduced fertility; other animals, for example man, which have a lower critical sperm volume could be more affected. Although similar studies have not been done in birds, extrapolation from the mammalian models implies that Bald Eagles hatched and raised in the Crofton area, particularly during the period of highest PCDD/PCDF contamination, may also appear externally normal, but have reduced capability to reproduce. The potential for wildlife exposure to other chlorinated compounds of pulp mill origin has received little attention. Although no samples were analyzed from the Crofton area, waterfowl breast muscle tissues collected from 1990 to 1992 near various pulp mills on the British Columbia coast, including Nanaimo and Powell River, contained from 0.5 to 5 ag/kg pentachiorophenol and traces (<1.0-3.3 tg/kg) of di- and tetrachloroquaiacols (Canadian Wildlife Service 1994). Those compounds are considered indicative of bleached-kraft pulp mill contamination of receiving water, sediments and biota (Dwernychuck et al. 1994). Release of organochlorines (AOX) in pulp mill effluents has decreased significantly since the installation of secondary treatment systems at all British Columbia coastal pulp mills (see Table 3.9). Studies of fish collected from both bleached-kraft and non-kraft pulp mills in eastern Canada have also reported the presence of an unidentified factor(s) present in effluents of both mill types that induce CYP 1A and affect reproductive hormone levels (Carey et al. 1992). Presence of that factor was independent of either chlorine bleaching or secondary treatment. However, both chiorophenols and chloroquaiacols and the unidentified factor appear to be cleared fairly rapidly in fish, ie. within two weeks; therefore, it seems unlikely that Bald Eagles would accumulate significant amounts of this class of chemicals. Alternatively, the low productivity measured in nests adjacent to the dioxin closure area at Crofton may be explained as either a sampling artifact or the result of ecological factors that have not been identified. Because of the cost of helicopter surveys and the difficulty in locating  127  nests, the sample may not be representative of the area, implying that some productive nests were not surveyed each year. However, the probability of overlooking a significant number of productive versus unsuccessful nests in the Crofton area should be no different than in other areas. Although the Crofton area is surveyed at the end of the flight, after only 1.5 hours, observer fatigue should not be a factor. Because of the history of contamination, the Crofton area likely receives greater attention. Quality of nesting habitat near Crofton appears comparable or better than most areas of the survey route; there are large numbers of suitable nest trees in relatively undisturbed areas and only limited activity. Currently, I am unable to determine the cause for poor eagle productivity at nests adjacent to the dioxin fishery closure area at Crofton. It is probably not caused by low food supply. It may be caused by other ecological factors which we have failed to identify; however, the effect of contaminants whether from past or ongoing exposure cannot be ruled out. Further intensive work in this area is necessary to confirm the results and investigate causes. My conclusions agree with those of Dykstra (1994) that the role of food supply needs to be factored into any studies of the effects of contaminants or other habitat quality variable in studies of Bald Eagles. Measurement of plasma lipids may provide a useful surrogate for energetic status of eagle nestlings. Further work is required to determine the causes of the apparent low productivity in the Crofton area. Acknowledgments A special thanks to Ian Moul and George Compton for all of their support and assistance in the field. Chris Coker and Brenda Li-Pak-Tong are thanked for their field work on the prey deliveries. Ron McLaughlin (MacMillan-Bloedel) and Ken Stenerson (Scott Paper) are also thanked for personal and corporate financial support with helicopter surveys. Working in the laboratory of Dr. Ross Norstrom, Mary Simon did the PCDD/PCDF and non-ortho PCB analysis; Henry Won did the organochiorine and plasma lipid analyses.  128  0.5  1 1 2 2 1.67 1.67 2  Powell River Ball Pt. Ball Pt. Hardy IsI. Hardy IsI. Thunder Bay Thunder Bay Lund  .  1.33 1.67 1.67 1.67 1.67 1.67 1 1 1.75 1.75  1.5 1.57 1.33 1 1.33  Lower Fraser Valley Cheam Is. Herrling I Vedder Carey Is. Island 20  East Vancouver Island Southey IsI. Maude Isl. Maude Isl. \Valljs Pt. Wallis Pt. Wallis Pt. New Castle IsI. New Castle IsI. Harmac Harmac Lantzvile Maple Mt.  1 1.5 1 1 1.5  Product. 1 (chicks/ active terr.)  Fraser River Delta Alaksen Steveston I Deas Is. Windrow Peat Farm  Location, Nest  3.60 0.182 0.076 0.075 0.054 0.074 0.042  0.112 0.1 0.253  .  0.1 0.121 0.224 0.124 0.096 0.115 0.13 0.105  0.107 0.1 0.114 0.111 0.104  0.106 0.116 0.139 0.103 0.126  Lipid %  39.74 9.07 3.60 3.28 2.15 1.83 2.47  4.95 1.82 8.79  .  1.74 3.73 2.26 1.48 1.31 1.02 4.39 2.02  0.45 0.63 0.75 0.48 0.60  2.39 2.51 2.70 0.73 0.75  TEQ5WHO  6.09 0.73 0.37 0.27 0.25 0.13 0.29  0.25 0.41 0.22 0.16 0.18 0.18 0.37 0.13 0.50 0.45 0.76 1.71  ND 0.05 ND 0.05 0.06  ND 0.21 ND ND ND  2378TCDD  7.16 1.34 0.76 0.56 0.49 0.58 0.58  0.41 0.63 0.33 0.48 0.19 0.40 1.50 0.50 1.27 1.20 0.27 3.62  0.14 0.14 0.21 0.11 0.11  ND 0.66 0.37 0.19 0.11  12378PnCDD OCDD  12.91 4.33 1.52 1.22 1.14 0.80 2.09  1.13 1.87 1.03 0.82 1.88 0.28 4.23 1.70 0.19 2.34 0.32 8.16  ND 0.1 0.27 ND 0.16  ND 0.57 2.83 0.29 0.63  0.63 0.18 0.43 0.52 0.68 0.95 0.57  0.98 1.23 0.85 0.78 3.24 2.62 1.36 0.88 0.46 1.21 1.69 0.36  0.47 0.24 0.41 0.24 0.21  26.82 0.12 9.46 1.72 1.6  33.43 3.19 4.12 4.51 3.04 3.33 3.78  2.73 3.75 3.04 2.54 1.76 2.13 6.22 2.13 2.39 2.43 1.21 6.79  ND 0.25 0.2 0.17 0.15  ND 0.88 0.27 ND 0.06  2378TCDF  (ng/kg) (wet weight)  123678HxCDD  2.97 0.52 0.32 0.15 0.17 0.07 0.26  0.05 0.24 0.22 0.24 0.11 0.07 0.66 0.20 0.29 0.56 0.47 0.13  ND ND ND ND 0.03  0.15 0.2 ND ND 0.21  23478PnCDF  219 83.3 25.6 23.5 11.12 33.8 16.2  14.8 16.8 12.7 11.6 10.7 10.0 27.2 15.3 15.9 39.7 13.0 33.0  2.68 6.87 6.15 5.61 8.91  20.86 25.8 13.78 3.57 2.85  PCB -77  164. 51. 15.5 14.7 6.79 6.38 8.32  5.56 16.7 9.43 3.56 2.12 1.05 11.14 7.34 6.67 28.2 0.97 23.9  2 2.44 4.3 1.86 2.48  7.54 10.8 3.95 3.25 1  PCB -126  Residue levels (wet weight basis)  56.4 11.3 3.8 4 3.3 2.1 2.1  1 1 9.7  2 4.1 2.4 1.3 1.2 1.4 3.9 2.5  0.6 1.4 0.6 1.1 1.3  2.9 4.8 1.9 1.2 0.3  PCB -118  PCB -153  DDE  11.4 2.1 0.8 0.8 1 0.4 0.4  0.2 0.3 2.3  0.5 1.1 0.6 0.4 0.3 0.4 0.9 0.6  0.05 0.4 0.05 0.3 0.4  0.7 1.1 0.5 0.2 0.05  118.1 29.1 7.8 8 5.3 4.4 4.2  2.3 1.8 22.3  5.8 9 5.4 2.7 2.6 2.7 7.9 5.3  1.1 2.2 1.4 2.2 2  5.9 9.5 2.8 4 0.8  368.4 82.5 21.4 21.7 10.8 7.9 7.1  3.7 3.9 110.6  11.9 24.9 16.4 5.5 4.8 6.5 17.3 11  4.1 7.6 7.9 24.6 10  19.1 24.4 12.7 14.6 7.1  (pg/kg) (wet weight)  PCB -105  670 149 47 49.8 37.9 28.2 25.6  10.1 14.3 153  31.9 57.2 33.8 21.3 16.3 20.5 50.3 35.2  7.05 16.1 6.70 14.0 16.6  34.5 57.3 15.9 18.2 3.14  Total PCBs  Appendix 4-1. Productivity, % lipid and selected chlorinated hydrocarbon residue levels in plasma of individual Bald Eagle chicks collected from the coast of British Columbia, 1993-94  2  0.5 1 1.33 1.33  Johnstone Strait Plumper Isi. Fire 1st. Cracroft Pt. Cracroft Pt. 0.1 0.135 0.365 0.103 0.185  0.025 0.098 0.061 0.099  0.012 0.006 0.061  1.08 0.1 0.014 0.018 0.02 0.036 0.052 0.008  0.93 0.34 4.20 0.98 1.05  0.46 0.45 0.91 1.41  0.43 0.80 0.72  2 ND 0.88 1.11 0.43 1.0 1.31 0.62 1.82  2.66 2.56 2.97  TEQsWHO  ND ND ND ND ND  ND ND ND ND  ND ND ND  ND ND ND ND ND ND ND ND  0.28 0.22 0.20  2378TCDD  ND 0.035 0.36 0.1 ND  ND 0.15 0.26 0.28  ND ND ND  ND ND ND ND ND ND ND ND  0.89 0.71 0.78  12378PnCDD  -  -  OCDD  ND ND 0.21 0.16 ND  0.26 0.16 0.56 0.76  ND ND ND  ND ND 0.48 ND ND ND ND ND  2.1 1.94 3.01  0.18 0.33 0.13 0.14 0.17  0.35 0.72 0.30 0.39  0.37 1.86 0.37  0.46 1.18 1.00 0.57 0.55 0.40 0.20 0.83  0.59 0.62 0.84  0.1 0.07 1.03 0.19 0.14  0.71 0.31 2.02 2.02  ND 0.41 ND  ND ND ND ND 0.61 ND ND ND  5.50 2.86 3.15  0.025 0.015 0.63 0.11 0.1  ND ND 0.21 0.12  ND ND ND  ND ND ND ND ND ND ND ND  0.29 0.24 0.12  23478PnCDF  2.85 1.37 24.5 6.18 5.47  3.02 3.10 4.03 13.7  2.52 11 6.90  6.53 6.15 12 6.21 8.64 5.53 5.69 9.61  10.3 11.4 17.1  PCB -77  5.73 1.42 28.8 6.28 6.8  1.11 0.57 1.27 5.58  0.08 3.28 3.54  5.49 2.69 6.00 2.89 5.39 5.68 3.18 10.4  6.57 9.73 11.6  PCB -126  Residue levels (wet weight basis) 2378TCDF  (ng/kg) (wet weight)  123678HxCDD  Productivity means of 3 years data (1992, 1993, 1994), except Langara Island, 1994 only. ND not detected (minimum detection limit, PCDDs!PCDFs, 0.01 ng/kg)  1 2 1 2 1  0.5 1 1  Clayoquot Sound White Pine C. Gibson Cove Gibson Cove  Langara Island Lucy 1 Lucy 2 Guillemot Marguerite Cabin Bay  1.33 0.5 1 1 0.67 1 1 0.33  0.042 0.056 0.068  2 1 1  Lund Gibson’s B Gibson’s B.P.  Barkley Sound Santa Maria IsI. Numukamis Assits IsI. Assits Isi. Hissin Pt. Mercantile Cr. Mercantile Cr. Salmon Beach  Lipid %  Product.’ (chicks! active terr.)  Location, Nest  Appendix 4, cont...  2.1 0.5 3.9 1.2 1.8  0.9 0.9 1.6 1.8  0.4 0.8 0.9  1.5 0.8 1.1 1 1.6 1.9 1.2 2.9  2.4 2.6 3.4  PCB -118 PCB -153  DDE  0.3 0.05 0.7 0.2 0.3  0.2 0.3 0.4 0.4  0.05 0.2 0.2  0.3 0.2 0.2 0.2 0.4 0.3 0.3 0.5  0.5 0.5 0.7  5.3 1.1 7.7 3 4.3  1.6 1.4 3.3 3.8  0.8 1.3 1.5  3.2 1.9 2.6 2.5 4.1 4.8 2.9 6  4.9 5.6 7  62.2 4.8 57.1 12.2 26.7  4.4 4.3 10.3 14.6  3.7 7.7 10  21.1 7.8 9.8 9.7 29.1 15.4 11.7 48.0  8.1 10.9 15.2  Q.g/kg) (wet weight)  PCB -105  21.6 5.04 39.6 12.8 21.4  9.63 8.59 22 23.7  3.77 8.21 9.95  20.6 13.4 14.7 14.1 26 24.4 16.2 45.6  29.7 34.2 43.4  Total PCBs  GENERAL SUMMARY AN]) CONCLUSIONS The overall purpose of this research was to investigate the toxic hazard posed by chlorinated hydrocarbon contaminants to Bald Eagle populations breeding and wintering in the Strait of Georgia area of British Columbia.  The research tested a general hypothesis that as  top predators in marine and estuarine systems, Bald Eagles would bioaccumulate high levels of chlorinated hydrocarbons. Consequent to high exposure and as ensuing hypotheses, both survival and reproduction would be adversely affected. These hypotheses were tested by a number of field and laboratory studies. Adult exposure and mortality study The investigation began by collecting samples from the large number of Bald Eagles found dead or dying each year in British Columbia. Many sick birds and carcasses are turned in by concerned members of the public or individuals seeking taxidermy permits. Of 484 eagles examined in this study, 59 found between 1988 and 1993 were selected for organochlorine analysis. Of those birds 5% had liver residue levels of DDE and chlordane related compounds diagnostic of acute toxicity. Even this percentage is surprising and the long term persistence of OC pesticides and continued input from atmospheric sources and migratory birds is indicated. These findings reinforce the need for vigilance in both the enforcement of current regulations and scrutiny of new commercial chemicals. Of 19 Bald Eagles further analyzed for PCDDs, PCDFs and non-ortho PCBs, livers of four birds (21 %) contained TCDD-toxic equivalents (TEQsWHO) > 1 ,000 ng/kg. Birds with high PCDD and PCDF levels were found in the vicinity of bleached-kraft pulp mills. Most bird with elevated chlorinated hydrocarbon levels were in poor body condition indicating lipid and contaminant mobilization. Based on high TCDD/TCDF ratios in at least three eagles, hepatic CYP1A enzymes were likely induced 131  Study of biological effects in eagle chicks In order to assess embryotoxic effects of chlorinated hydrocarbons in Bald Eagles, eggs were collected within an exposure gradient and incubated in the laboratory. Yolk sacs of chicks collected near bleached-kraft pulp mills contained higher concentrations of PCDDs and PCDFs, although there were no significant effects on hatching success or morphological endpoints. Hepatic CYP1A levels were induced in chicks from pulp mill sites and correlated significantly with 2,3,7,8-TCDD, 2,3,7,8-TCDF and TEQ5WHO in yolk sacs. TEQsWHO associated with CYP1A induction and converted to a whole egg wet weigh basis, 210 ng/kg, were suggested as a LOEL for the Bald Eagle; TEQsWHO associated with background CYP1A levels were suggested as a NOEL for the Bald Eagle, 100 ng/kg. These findings suggest that the Bald Eagle embryo is perhaps an order of magnitude less sensitive to TCDD-like toxicity than the chicken embryo. The LD 50 for the chicken embryo is about 250 ng/kg (Alired and Strange 1977; Janz 1995), similar to the 210 ng/kg TEQ5WHO measured in eagle eggs without apparent effects on hatching success or histological, morphological and some biochemical endpoints. At 100 ng/kg TEQ5WHO in eagles, no significant CYP1A induction occurred, while two-fold AHH induction was measured at 10 ng/kg injected into chicken eggs. With regard to CYP1A induction, Bald Eagles appear somewhat more sensitive than Great Blue Herons and Double-crested Corinorants. In heron chicks, EROD activity was significantly induced (six-fold) at about 440 ng/kg, but not at 250 ng/kg TEQsWHO (Sanderson et at. 1992a). In cormorant chicks, significant eight-fold EROD induction occurred at 550 ng/kg but not at 217 ng/kg TEQsWHO (Sanderson et at. 1992b). Bioaccumulation study For this study, fresh Bald Eagle eggs were collected at a variety of locations on the British Columbia coast, representing different chlorinated hydrocarbon exposure scenarios. A data base of contaminant levels in Bald Eagle prey items, principally from pulp mill sites in the Strait of Georgia, was compiled using existing data. A simple model was used to examine the relationships between contaminant levels in Bald Eagles and their foodchain. The model accurately predicted 2,3,7, 8-TCDD levels in Bald Eagle eggs and was reasonably accurate for  132  other compounds. The model was used to estimate 2,3,7,8-TCDD and TEQWHO levels in forage fish and sentinel fish-eating bird species (herons, cormorants, grebes, mergansers), which would be protective of Bald Eagles consuming an average diet. The NOELs and LOELs generated in the above embryotoxicity study were used as critical values in eagle eggs. Concentrations of 0.5 ng/kg in forage fish and 10 ng/kg in fish-eating birds were suggested as site specific guidelines for the Strait of Georgia. The same approach was used to derive similar values for total PCBs, suggested to be 0.01 ng/kg in forage fish and 0.3 ng/kg in fish-eating birds. Productivity study The research described in the previous studies addressed acute toxicity of adult birds and determination of critical levels in eggs, associated with embryotoxicity. During the fourth part of this work, Bald Eagle breeding success was measured for up to three years at eight sites on the British Columbia coast. Because of annual variability, assessment of breeding success in Bald Eagles requires a minimum of three years data. Studies elsewhere showed that reproduction in birds of prey is a critical endpoint affected by chlorinated hydrocarbons in birds of prey (Newton 1979). In order, to relate productivity of individual nests to contaminant exposure, blood samples were taken from nestlings, to minimize the impact of sample collection. Bald Eagle productivity was highest overall at nests in the lower Fraser River valley and delta, while at four of five reference areas, selected for their remoteness from direct industrial input of pollutants, productivity was less than the level of 0.75 young/occupied nest considered necessary to sustain an eagle population. Only at Langara Island, an area of very high biological productivity, was eagle breeding success comparable to the Fraser valley and most Strait of Georgia sites. At the reference locations, low breeding success is likely due to low food availability, particularly during chick rearing. This was supported by finding of significantly lower nestling plasma lipid content at those sites and a significant positive regression between mean nestling plasma lipid levels and mean productivity among sites. Despite higher plasma levels of PCDDs and PCDFs, Bald Eagle productivity was relatively  133  high at nests near two pulp mill areas on the Strait of Georgia (Nanaimo, Powell River); at those sites, no significant differences in mean productivity occurred at nests adjacent to PCDD/PCDF fishery closure areas compared to nests outside of the closure area. However, productivity was significantly lower at nests inside the fishery closure area at one site, Crofton, than outside the dioxin closure. Low breeding success around Crofton likely is not due to low food availability; the area is rich in marine life. Data from biomonitoring studies of fish-eating birds showed that PCDD and PCDF levels in local food chains fell dramatically between 1989 and 1992, subsequent to modifications to the bleaching process employed by the mill and a ban on chlorophenolic anti sapstain usage. Alternative hypotheses to explain the low eagle productivity in the area include: first, the presence of a substance released in the mill effluents, that has contaminated local food chains and is either embryotoxic or capable of affecting parental breeding behaviour. Second, some eagle pairs may be reproductively impaired as a result of past exposure in ovo or during early development of the reproductive system, to elevated levels of 2,3,7, 8-TCDD and related chemicals. This last hypothesis requires further study and testing. In conclusion, during the recent past reproduction of Bald Eagles in the Strait of Georgia was probably reduced by exposure to significant chlorinated hydrocarbon levels, particularly DDE. Increases in nest occupancy reported for the southern Gulf Island between the early 1 970s and late 1 980s is typical of the population recoveries documented in many areas of North America and attributed to declining environmental DDE contamination. During the 1 980s and at least until the early 1 990s, eagles breeding and wintering near bleached-kraft pulp mills on the British Columbia coast were exposed to relatively high levels of PCDDs and PCDFs. At Crofton, the effects of this pollution may be continuing, although the mechanism is obscure. At other areas of the British Columbia coast, Bald Eagle breeding success appears to be influenced mainly by food supply. The effects of chlorinated hydrocarbons on Bald Eagle populations have to be considered in the context of multiple stresses, both chemical and otherwise, on survival and reproduction. Lead poisoning from ingestion of spent shot is a major cause of death for British  134  Columbia Bald Eagles; many eagles have also been sublethally poisoned, with probable consequences for longterm health and survival. In some areas, such as the Lower Fraser Valley, pesticide poisoning is a major cause of mortality. Bald eagles are also vulnerable to loss and disturbance of nest sites. Given these factors, and the growing human population of the Georgia Basin, maintenance of a healthy eagle population will require ongoing vigilance. Finally, although the Bald Eagle has some merits as a sentinel species of pollutant exposure and effects, it may be more cost-effective to monitor colonial fish-eating birds.  Future Directions Ecotoxicological work on Bald Eagles should further investigate the low reproductive rate measured at Crofton. All nests in the area from Cowichan Bay to Thetis Island should be located. A sample of nests including those nearest the mill, should be intensively observed to determine breeding behaviour and the timing of nest failures. Toxicological hypotheses can be tested by trapping adult eagles on their breeding territories to obtain blood samples for contaminant analyses and measurement of reproductive and thyroid hormones. Similar studies are required at a reference site, such as Barkley Sound, and also possibly at another pulp mill site, either Nanaimo or Powell River, depending on available funding. Laboratory research using in vitro cell cultures of primary hepatocytes from eagles or other raptors would provide data on sensitivity of raptors compared to more commonly studied laboratory species and sentinel species such as Herring Gulls. Alternatively, in vivo comparative toxicology with American Kestrels, particularly of TCDD effects on reproductive endpoints would be valuable. Further ecological work on the role of food supply in breeding success of British Columbia eagle populations should also be undertaken. A long term monitoring study of Bald Eagle reproduction at one or more of the remote sites could provide valuable information on fluctuations in coastal productivity, and the influence of large scale processes such as global warming.  135  References Abraham, K. R. Krowke and D. 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