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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 EAGLESON THE COAST OF BRITISH COLUMBIA:EXPOSURE AND BIOLOGICAL EFFECTSbyJOHN EDWARD ELLIOTTB.Sc., Carleton University, Ottawa, 1979M.Sc., The University of Ottawa, 1989A THESIS IN PARTIAL FULFILMENT OF THE REQUIREMENTS FORTHE DEGREE OF DOCTOR OF PHILOSOPHYinTHE FACULTY OF GRADUATE STUDIESTHE FACULTY OF AGRICULTUREDepartment of Animal ScienceWe accept this thesis as/conforming to he required standardThe University of British ColumbiaOctober, 1995c John Edward Elliott , 1995In presenting this thesis in partial fulfilment of the requirements for an advanceddegree at the University of British Columbia, I agree that the Library shall make itfreely available for reference and study. I further agree that permission for extensivecopying of this thesis for scholarly purposes may be granted by the head of mydepartment or by his or her representatives, It is understood that copying orpublication of this thesis for financial gain shall not be allowed without my writtenpermission.(Signature)Department of________________The University of British ColumbiaVancouver, CanadaDateDE-6 (2188)AbstractAttracted 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 thePacific coast of Canada. Eagle habitat has been extensively modified by logging andwaterfront development, while industrial effluents have contaminated food chains. Untilrecently, most pulp mills on the British Columbia coast used elemental chlorine bleaching anddid not secondarily treat effluents, thus releasing chlorine containing chemicals, particularlypolychiorinated dibenzo-p-dioxins (PCDDs) and polychiorinated dibenzofurans into the localenvironment. As top predators, Bald Eagles are exposed to elevated levels of PCDDs, PCDFsand the chemically related polychiorinated biphenyls (PCB5) and organochlorine pesticides.This thesis addressed spatial and temporal trends in chlorinated hydrocarbon exposure of BaldEagles and toxicological consequences at treatment populations near pulp mills in the Strait ofGeorgia and in industrial areas of the Fraser River delta, and at reference areas on west coastVancouver Island, Johnstone Strait and the Queen Charlotte Islands.Initial research during 1990-199 1 focused on eagles found dead or dying and determinedthat the majority of birds tested had low liver organochlorine levels (< 5 mg/kg, N =59). Asmall proportion (< 5 %) had levels of DDE, polychiorinated biphenyls (PCBs) and chiordanerelated compounds potentially diagnostic of acute poisoning. A larger proportion hadPCDD/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 breedingseason.In 1992, in ovo exposure to a gradient of environmental contaminants was studied bycollecting eggs (N = 25) for laboratory incubation. Hatching success was not significantlydifferent 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 PowellRiver compared to those from a reference site (p < 0.05); hepatic EROD and BROD activities11were 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 hadgreater concentrations of 2,3,7,8-substituted PCDDs and PCDFs than reference areas. Thehepatic CYP1A cross-reactive protein and EROD and BROD activities were positivelycorrelated 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 werefound. Using hepatic CYP1A expression as a biomarker, a no-observed-effect-level (NOEL) of100 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, eggswere also collected during incubation, 1990-92, at the treatment and reference areas andanalyzed for chlorinated hydrocarbons. Data on Bald Eagle avian and fish prey items from thestudy area were compiled and used as input to a bioaccumulation model. The model accuratelypredicted 2,3,7, 8-TCDD levels in eagle eggs based on dietary concentrations, but was lessaccurate for other PCDDs and PCDFs. Using the LOEL levels in eagle eggs derived from theabove study, concentrations of 2,3,7,8-TCDD in prey fish of 0.5 ng/kg and in fish-eating birdsof 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 measuredfor 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 wassignificantly lower at reference sites. Using nestling plasma lipid content as a marker of bodycondition, food supply appeared to be the main factor limiting eagle productivity on the BritishColumbia coast. However, at a sample of eagle nests adjacent to the dioxin fishery-closurezone near the pulp mill at Crofton, low productivity was probably not caused by low foodavailability. The cause of the low reproductive rate at Crofton has not been determined;however, a toxicological explanation has not been ruled out.111Key Words: Bald Eagle, bioaccumulation, CYP 1A, mortality, reproductive rate, 2,3,7,8-tetrachlorodibenzo—p-dioxinivTable of ContentsAbstractTable of ContentsList of TablesList of Figures.List of AppendicesAbbreviationsAcknowledgementsGeneral IntroductionHypotheses and ObjectivesOverview of the ThesisChapter 1 Chlorinated hydrocarbon liver levels and autopsyEagles found dead or debilitated, 1989-93Materials and MethodsResultsDiscussionChapter 2 Biological effects of chlorinated hydrocarbons inMaterials and MethodsResultsDiscussionChapter 3 Bioaccumulation of chlorinated hydrocarbons and mercuryin eggs and prey of Bald EaglesMaterials and MethodsResultsPage11vii• ixxiixliixiv17• 18dataBaldfor BaldEagle chicks• . 19• . 1923• . 32• . 41• . 41• . 4860707077Discussion 89VPageChapter 4 Influence of contaminants and food supply on Bald Eagle productivity 102Materials and Methods 103Results 110Discussion 120General Summary and Conclusions 131References 136viList of TablesPageTable 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 28Table 1.2 Non-ortho and mono-ortho PCBs in Bald Eagle livers collected from BritishColumbia (ng/kg, wet weight) 29Table 1.3 Concentrations of select PCDDs and PCDFs in Bald Eagle livers collected fromthe south coast of British Columbia (ng/kg, wet weight) 30Table 1.4 Comparison of TEQs calculated from select PCDDs, PCDFs, non-orthoand mono-ortho PCBs levels in Bald Eagle livers collected from the southcoast of British Columbia (ng/kg, wet weight) 31Table 2.1 PCDD and PCDF concentrations (nglkg, lipid weight basis) in yolk sacs ofBald Eagle chicks collected in 1992 from British Columbia 51Table 2.2 Concentrations of non-ortho PCB congeners in yolk sacs of Bald Eagleembroys collected in 1992 from British Columbia 52Table 2.3 Organochlorine pesticide concentrations (geometric means 95% confidenceintervals, range in brackets) in yolk sacs of Bald Eagle chicks collectedin 1992 from British Columbia 53Table 2.4 Outcome of artificial incubation of Bald Eagle eggs collected from BritishColumbia, 1992 54Table 2.5 Histological examination of immune system tissues in Bald Eagle chicks(mean ± SD) 55Table 2.6 Measurement of hepatic cytochrome P450 and porphyrin parameters andvitamin A in plasma and liver of Bald Eagle chicks collected in 1992 fromBritish Columbia (mean ± SD) 56Table 2.7 Concentration-effect relationships between biochemical and morphologicalmeasurements with chlorinated hydrocarbon yolk sac levels in Bald Eagle chicks 59Table 2.8 Comparison of regression (r2) values of some hepatic biochemical parameterson TEQs derived from three sets of toxic equivalence factors (TEF5) 60Table 3.1 Mean PCDD/PCDF lvels (ng/kg, wet weight) in fish collected near three pulpmills on the Strait of Georgia, British Columbia 75vi’PageTable 3.2 PCDD and PCDF levels (ng/kg, wet weight) in waterbird and seabird speciesfrom the British Columbia coast 76Table 3.3 (PCDD) and (PCDF) residue levels (wet weight basis) in Bald Eagle eggsfrom British Columbia, 1990- 1992 78Table 3.4 Organochiorine and PCB residue levels (mg/kg, wet weight) in Bald Eagleeggs from the British Columbia coast, 1990-1992, expressed as geometricmeans and 95% confidence intervals (range in brackets) 80Table 3.5 Mercury residue levels (mg/kg, wet weight) in Bald Eagle eggs from locationson the British Columbia coast, 1990-1992, expressed as geometric means and95 % confidence intervals (range in brackets) 81Table 3.6 Non-ortho PCBs in Bald Eagle eggs (ng/kg, wet weight) collected from BritishColumbia, 1992 84Table 3.7 Eggshell thickness data, mean ± SD, (range in brackets) for Bald Eagles collectedfrom British Columbia, 1990-1992 85Table 3.8 A simulation of PCDD/PCDF levels in Bald Eagle eggs at Crofton, 1990,based on concentrations in the diet 86Table 3.9 Characterization of British Columbia pulp mills discussed in this paper . . . . 101Table 4.1 Correlation Matrix (r values) for percent plasma lipid and selected hydrocarbonin Bald Eagle nestlings from British Columbia, 1993-94 109Table 4.2 Nest success and production of young for Bald Eagles at nine study areas onthe British Columbia coast (1992-94) 111Table 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 BritishColumbia, 1993-94 116Table 4.4 Organochlorine pesticide and PCB levels, geometric means and 95% confidenceinterval (tg/kg, wet weight) in blood plasma of Bald Eagle chicks fromthe coast of British Columbia, 1993-94 118Table 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 BritishColumbia, 1993-94 120vi”List of FiguresPageFigure 1. Molecular structure and position numbering of polychiorinated dibenzo-pdioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs) 2Figure 2. Molecular structure of the major organochiorine pesticides 3Figure 3. Molecular mechanism proposed for TCDD and related chemicals 7Figure 1.1 Locations of Bald Eagles collected from British Columbia, 1989-93, andanalyzed for chlorinated hydrocarbons (N = 59) 20Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed compared to the completeset of birds received 24Figure 1.3 Numbers of eagles showing different DDE and PCBs in livers (N =59) 24Figure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month 25Figure 1.5 DDE and PCB residue levels in relation to body condition 26Figure 1.6 PCB congeners in Bald Eagle livers expressed as percent of total PCBscompared for birds in good and poor body condition (N =9, for each group) . . 36Figure 2.1 Locations where Bald Eagle eggs were collected for artificial incubation 42Figure 2.2 Residue levels of major PCDDs and PCDFs in yolk sacs of Bald Eaglescollected from British Columbia in 1992. Vertical bars represent geometricmeans of two to five analyses per collection site along with the 95 %confidence interval. Means which do no share the same lower caseletter were significantly different (p < 0.05) 49Figure 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 eightanalyses per collection site. Isomers are identified according to their IUPACnumber 50Figure 2.4 Exposure-response relationships between 2378-TCDD or log 2378-TCDFconcentrations in yolk sacs of Bald Eagles and hepatic (A) EROD activity(B) CYP1A concentrations and (C) BROD activity 58lxPageFigure 2.5 The contribution of various chlorinated hydrocarbon groups to the sum ofTCDD toxic equivalents (TEQ) in Bald Eagle yolksacs from coastal BritishColumbia, 1992 (N values and variances are in the tables), compared tovalues for common terns from the Netherlands. Toxic equivalents factors forPCDDs/PCDFs from Safe (1990) and for PCBs from Ahlborg et a!. (1994) . . 66Figure 3.1 Locations where Bald Eagle eggs were collected for analysis 71Figure 3.2 PCB congeners in Bald Eagle from British Columbia, 1990-1992, expressed aspercent of total PCBs. Values represent means of three to eight analysesper collection site. Congeners are identified according to their IUPAC number 82Figure 3.3 Plot of the first and second principle components (PCi and PC2). PCBcongener concentrations for all individual egg analyses were expressed aspercent total PCBs and arcsine transformed. Principle components analysiswas 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 variancewas explained by PCi and 15 % by PC2 83Figure 3.4 The contribution of various chlorinated hydrocarbon groups to the sum ofTCDD toxic equivalents (TEQs) in Bald Eagle eggs from coastal BritishColumbia, 1990-1992 (N values and variances are in the tables). Toxicequivalents for PCDDs/PCDFs from Safe (1990) and for PCBs fromAhlborg et a!. (1994) 84Figure 3.5 Concentration of 2,3,7,8-TCDD predicted in Bald Eagle eggs based onthe percent of fish-eating birds in the diet. Prediction is based on abioaccumulation model described in the text and the simulation is based ondata from Crofton, British Columbia, 1987-1992 88Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections.At Langara Island, the survey circumsribed the coastline of the island 104Figure 4.2 Bald Eagle nest sites and dioxin fishery closure areas at Powell River,Nanaimo and Crofton 106Figure 4.3 Bald Eagle productivity compared between samples of nest located adjacentto shorelines inside and outside of dioxin fishery closure areas on the BritishColumbia coast 112xPageFigure 4.4 Productivity at Bald Eagle nest sites on the British Columbia coast as afunction of contaminant concentrations in plasma samples from nestlings raisedin that territory, for: A) the log of TEQsWHO, B) the log of DDE. Thesubpopulations included: East Vancouver Island, Powell River, BarkleySound, Clayoquot Sound, Johnstone Strait, Fraser Delta, lower Fraser Valleyand Langara Island 113Figure 4.5 Comparison of mean productivity of Bald Eagles at sites on the coastof British Columbia with the mean percent lipid in plasma samples of nestlingeagles at each site 114Figure 4.6 Residue levels of selected PCDDs and PCDFs in plasma samples of baldeagle nestlings collected on the British Columbia coast, 1993-1994.N sizes and error estimates are in Table 4.3. Means that do not share thesame lower case letter are significantly different (p <0.05) 117Figure 4.7 Residue levels of selected PCBs in plasma samples of Bald Eagle nestlingscollected on the British Columbia coast, 1993-1994. N sizes and errorestimates are in Table 4.4. Means that do not share the same lower caseletter are significantly different (p < 0.05) 119Figure 4.8 Trends in 2,3,7,8-TCDD in eggs of eagles, herons and cormorants atCrofton, British Columbia. The likely trend in eagles is extrapolated backto 1987, based on the mean 2,3,7,8-TCDD ratio of eagles:herons, 1990-1992 120xlList of AppendicesPageAppendix 1-1 Organochiorine pesticide and PCB levels in Bald Eagle liverscollected from British Columbia (mg/kg wet wt.) 38Appendix 2-1 Selected morphological measurements in Bald Eagle chickscollected in 1992 from British Columbia 69Appendix 4-1 Productivity, % lipid and selected chlorinated hydrocarbonresidue levels in plasma of individual Bald Eagle chicks collectedfrom the coast of British Columbia, 1993-94 128xiiAbbreviationsAh aryl hydrocarbon NWRC National Wildlife ResearchCentreAHH aryl hydrocarbonhydroxylase OC Organochiorine pesticideANCOVA analysis of covariance PCA principle component analysisANOVA analysis of variance PCB polychiorinated biphenylBMF biomagnification factor PCDD polychiorinated dibenzo-pdioxinBROD benzyloxyresorufin 0-dealkylase PCDF polychiorinated dibenzofuranCWS Canadian Wildlife Service PWRC Pacific Wildlife ResearchCentreCYP1A cytochrome P450 1ASAS Trademark, SAS InstituteCYP2B cytochrome P450 2B Inc.DDE 1, 1-dichioro ethylene bis (p- SYSTAT Trademark, Systat Inc.chiorophenyl)TCDD tetrachioro dibenzo-p-dioxinDDT 1,1, 1-trichloro-2,2-bis(p-chlorophenyl)ethane TCDF tetrachioro dibenzofuranEROD ethoxyresorufin 0-deethylase TEF toxic equivalent factorGLEMEDS Great Lakes embryo TEQ TCDD toxic equivalentmortality edema anddeformities syndrome WHO World Health OrganizationHCB hexachlorocyclobenzeneHCH hexachiorocyclohexaneLOEL lowest-observed-effect-levelNOEL no-observed-effect-levelxl”AcknowledgementsI would like to thank my supervisory committee, Kim Cheng, Gail Beliward, RossNorstrom and Tom Sullivan for overall guidance and support. I would like to acknowledge thefinancial and personal support of the Canadian Wildlife Service, and to personally thank SteveWetmore at the Pacific Wildlife Research Centre and Keith Marshall at the National WildlifeResearch Centre for their advice and support over the years.A project of this sort depends on the assistance of a great many people. Specificcontributions are acknowledged at the end of each chapter, however, the support of a numberof people deserves special consideration: Ian Moul was a valuable co-worker in virtually allphases of the field work; George Compton contributed his considerable tree climbing andbush-whacking skills. Mary Simon, Henry Won and Suzanne Trudeau are thanked for theirwork on the chemistry and biochemistry, Ken Langelier was a great help in the wildlife healthaspects and suggested the initial work on Bald Eagles. I am very grateful to Laurie Wilson forthe many technical and scientific roles she undertook for me. Shelagh Bucknell and PamWhitehead are thanked for their assistance and patience in typing of tables and preparation offigures, respectively.I would also like to acknowledge my friends and coworkers both at UBC and CWS formaking this PhD experience more rewarding and enjoyable.I also wish to thank my parents for imparting a sense of what is important in life. Mostimportantly, I am most grateful to the patience and support of my wife Christine and mychildren, Kyle, Siobhan, Frazer and Alicia.xivIntroductionPollution of the environment by toxic substances has become a global problem withecological, economic and political consequences. Chlorinated hydrocarbons such aspolychiorinated dibenzo-p-dioxins (PCDDs), polychiorinated dibenzofurans (PCDFs),polychiorinated biphenyls (PCBs) and DDT (1,1,1 -trichloro-2,2-bis[p-chlorophenyl]ethane) haveattracted 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 otherorganochiorine pesticides on reproduction and survival of birds of prey, such as eagles,Ospreys (Pandion haliaetus) and falcons. These birds, particularly the Bald Eagle (Haliaeetusleucocephalus) and the Peregrine Falcon (Falco peregrinus), have become symbols ofenvironmental awareness and reminders of ecological consequences of short-sighted use ofchemical 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 canfunction toxicologically like TCDD. A great deal of laboratory research has been conducted onPCDDs 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 ofGeorgia and on exposure to and the consequences of the widespread pollution of that area byPCDDs and PCDFs from forest industry sources.Chlorinated hydrocarbonsStructuresChlorinated hydrocarbons are organic compounds with chlorine substituents. This thesisis concerned primarily with the polychiorinated aromatics, those with chiorines substituted onaromatic ring structures, and to a lesser extent with some non-aromatic organochiorinepesticides, such as hexachiorocyclohexane (HCH). Ecotoxicologically, the most important1polychiorinated aromatics are the PCDDs, PCDFs, PCBs, and some of the organochiorinepesticides such as DDT.The structures of the PCDDs, PCDFs and PCBs are represented in Figure 1. ThePCDDs and PCDFs obtain a mainly rigid, planar configuration, which determines theirbiological behaviour. For the PCBs, the molecular conformation depends on the chlorinesubstituents. Those congeners without ortho-chiorines energetically obtain a mainly planarconformation, those with di-ortho chlorine substituents are non-planar and those with monoortho substituents are intermediary. Thus the non-ortho PCBs are approximate stereoisomers ofPCDDs and PCDFs and if chlorinated laterally, exhibit similar biological behaviour (Safe1984).Figure 1. Molecular structure and position numbering of polychiorinateddibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs).9 186 4dibenzo-p-dioxin 2,3,7,8 - Tetrachlorodibenzo-p-dioxin-ClSc173 2 2’ 3’4dibenzofuran 2,3,7,8 - Tetrachlorodibenzofuranbiphenyl PCB 12633’44’5 - Penta2Organochiorine pesticides fall into three structural groups (Figure 2). DDT is similar instructure to the PCBs, in that it has two chlorine-substituted benzene rings, in this case joinedon an ethane backbone. Dieldrin, mirex and the chiordane-related compounds, includingheptachlor epoxide, all belong to the cyclodiene group. The third group are the chlorinatedbenzenes and cyclohexanes.Figure 2. Molecular structure of the major organochiorine pesticides.SourcesPCDDs g PCDFs. Neither PCDDs nor PCDFs are deliberately produced commercially,but are formed either as by-products during the synthesis of other chemicals, such aschiorophenolic biocides, or during combustion of chlorine containing wastes. Incineration ofmunicipal and industrial wastes is the major global source of dioxins, which can be transportedlong 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, physicaland chemical atmospheric processes favour the deposition and accumulation of less toxic higherchlorinated compounds, which then dominate in sediments (Hites 1990). Elevated contaminationby 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:DIC I-I LO RO DI PHE N YLET HAN ES cI_OH_O_ ciCYCLODIENESDDT, DDDDicofolPerthaneMethoxychiorMethiochiorciCHLORINATED BENZENESCYCLOHEXANESAidrin, DieldrinHeptachlorChiorcianeEndosulfan(Cl)6HCB, HCHLindane (a-BHC)Cl3Baughman and Meselson 1973; Fanelli et at. 1980; Powell 1984). However, relatively recentstudies showed that effluents from kraft pulp mills using elemental chlorine bleaching contained2,3,7,8-TCDD and 2,3,7,8-TCDF (Kuehl et at. 1987), which caused contamination of fish andwildlife 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 oftetrachiorophenol-contaminated woodchips (Elliott et al. 1989a; Luthe et at. 1990). Use andproduction of 2,4,5-T and most chlorophenols has been regulated in North America. Pulp millsin 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 ‘closedcircuit’ uses such as in electrical transformers and capacitors and in heat transfer and hydraulicsystems, 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 wereproduced worldwide (Tanabe 1988). Until 1977, over 90 % of the production was in theU.S.A., after which it switched to Europe and Japan. Some 40 million kg PCBs have beenimported into Canada; the most recent inventory accounted for about 24 million and assumed thatthe remaining 16 million kg had been lost to the Canadian environment (Environment Canada1985). Open circuit uses of PCBs were voluntarily restricted by industry in 1973 and all uses ofPCBs have been regulated by governments in North America since 1977.Organochlorine pesticides. OC pesticides are synthetic compounds widely used to controlagricultural and forest pests and the transmission of vector-borne diseases. The most abundantOC pesticide in the environment is DDE, the major persistent metabolite of DDT. Othercompounds 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 NorthAmerica in the 1940s in public health campaigns to control lice (Carson 1962). From the 1940suntil the early 1970s, large quantities of DDT were sprayed to control forest insect pests in4British Columbia (Nigam 1975) and in the northwest USA (Henny 1977). Major restrictions onthe use of most organochlorine pesticides (ie. DDT, dieldrin, endrin, heptachior, HCH andtoxaphene) in Canada and the USA were first implemented in the early 1970s, with furthercontrols imposed throughout the 1970s and 1980s (Noble and Elliott 1986). Heptachior continuedto 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 andMiller 1984). A few minor applications of chlordane, lindane, dieldrin and heptachlor (eg. seedtreatment, termite control) are still permitted in Canada and the USA. In Mexico, somerestrictions on the use of DDT, BHC, dieldrin and heptachior were imposed in 1980 (Burton andPhiogene 1986).Organochlorines can be transported over vast distances by atmospheric and oceanicvectors; as such, ongoing use in Asia may now be the main source of OCs to the Canadianenvironment, particularly the Pacific coast (Elliott et at. 1 989b). Information on OC use inAsian countries bordering the north Pacific is scarce. Since the 1950’s, DDT and HCH havebeen used extensively on rice, cotton and vegetable crops, but in the 1970s, many countriesbegan to replace them with organophosphorus compounds. As in North America, agriculturaluses of OCs are subject to regulation in most north Pacific countries, but the degree ofcompliance varies. The People’s Republic of China manufactures OC pesticides; however, theproduction and use of two, DDT and HCH, were banned there in 1983 (Wolfe et al. 1984). InJapan, production and use of DDT and HCH were prohibited in 1971, but the use of chlordanefor termite control was permitted until the late 1 980s (Tanabe et al. 1989). Korea also prohibitedthe use of DDT in the early 1970s (Phillips and Tanabe 1989). However, in Hong Kong (wheremany pesticides are still formulated), there appears to be continued input of DDT into coastalwaters, despite restrictions imposed in 1988 (Phillips and Tanabe 1989).Food chain bioaccumulationFor 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 metabolicattack.5PCDDs PCDFs. Food chain bioaccumulation of PCDDs and PCDFs generallyrequires a 2,3,7,8-substitution pattern, as congeners lacking that substitution pattern aremetabolized in birds, mammals and fish (Van den Berg et al. 1993a). Accumulation of non-2,3,7,8-substituted PCDDs has been reported in some invertebrate species, particularlycrustaceans (Norstrom and Simon 1991).PCBs. Among homeotherms, tissue retention of PCB congeners varies with developmentof 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 aredegraded by CYP2B enzymes (Boon et al. 1987; Brown 1994).Organochiorines. The relative capacity of organochlorines to bioaccumulate has beenextensively 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 slowlydegraded and therefore more accumulative OCs in birds are DDE, mirex and oxychlordane, withheptachlor expoxide, dieldrin and HCH compounds being more rapidly cleared.Effects of chlorinated hydrocarbonsPCDDs, PCDFs PCBs. This group of compounds causes similar toxic symptoms inmost species studied (Safe 1990). Dose-related responses include: irnmunotoxicity, liverenlargement and other signs of hepatotoxicity such as porphyria, induction of drug-metabolizingenzymes, reproductive toxicity and cancer promotion (Safe 1984). The toxicity of the individualcompounds varies greatly with the molecular structure. The most toxic compound is 2,3,7,8-TCDD, which is often used as a model for studying the effects of these chemicals. The moretoxic furan and biphenyl congeners all exhibit a structural similarity to 2,3,7, 8-TCDD. Many ofthe toxic effects caused by this class of compounds are believed to be mediated by a cytosolicreceptor found in many tissues, known as the aryl hydrocarbon (Ah) receptor (Landers and Bunce1991). 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 of6compounds. To provide a practical method of dealing with this, the study of Ah-receptormediated structure-activity relationships has produced an additive scheme for estimating thetoxicity of complex mixtures of these chemicals through use of “TCDD Toxic EquivalenceFactors” (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. Analyticallydetermined concentrations are multiplied by the TEF, the results summed to produce the “TCDDToxic Equivalents” or TEQs. TEFs published by Safe (1990) are widely used; however, thosereported 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)INCREASEDMETABOLISM OFDRUGS ANDENVIRONMENTALCHEMICALSTOXICITYFigure 3. Molecular mechanism proposed for TCDD and related chemicals. The lipophiicxenobiotic ligand, such as TCDD, enters the cell by passive diffusion through thelipo-protein cell membrane and binds with the Ah-receptor (AbR); the AhR releases aheat shock protein (HSP 90) as it binds with the ligand. The ligand-receptor complexthen associates with the nuclear translocating protein (Arnt) and moves into the nucleus,where it interacts with dioxin responsive elements (AhREs) on the genome, which altersthe transcription of specific niRNAs. The resulting proteins then mediate the biochemicaland toxic responses observed with TCDD exposure (after Okey et al. 1994).7Although the toxicology of dioxins and related compounds continues to be extensivelystudied in laboratory mammals, there are less data on avian species. Bird studies have focusedon embryos, as the most sensitive life stage (Peterson et at. 1993). Chicken embryos areparticularly sensitive: the LD50 for 2,3,7, 8-TCDD, administered into the air sac of the chickenembryo, was reported as 250 ng/kg (ppt) egg (Alired and Strange, 1977). An LD50 for 2,3,7,8-TCDD in chicken embryos of about 200 ng/kg was determined more recently by both Henschelet at. (in preparation) and Janz (1995) using air cell and yolk sac injection. They also reported avery steep dose response curve, with no mortality at 100 ng/kg and complete mortality at 300ng/lcg. Injection of 2,3,7,8-TCDD or similar compounds into developing chickens causes atoxicity syndrome which includes, in addition to mortality, beak and other deformities, thymicand 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 ofcardiovascular malformations (Cheung et at., 1981). In domestic turkey embryos, non-orthoPCB congeners that bind the Ah receptor and thus act by a similar toxic mechanism to 2,3,7,8-TCDD, also cause gross deformities and mortality, but not the other symptoms seen in chickenembryos (Brunstrom and Lund 1988). In embryos of other avian species, such as Ring-neckedPheasants (Phasianus coichicus) and Eastern Bluebirds (Siatia sialis) injected with 2,3,7,8-TCDD, sublethal effects observed in chickens were not observed, rather mortality was the mostsensitive endpoint (Nosek et at. 1992; Martin et at. 1989). The LD5O for 2,3,7,8-TCDD was1100 ng/kg egg in the pheasant embryo and between 1000 and 10,000 ng/kg egg in the EasternBluebird 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 interspecificsensitivity among birds to the TCDD-isostereomer, PCB congener 77 (34-34). Chickens werethe 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).8Adults 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 formulationof 2,4,5-T fed to bobwhites for 18 weeks produced no effects on egg production or survival ofembryos. 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 2,3,7,8-TCDD for 10 weeks exhibited mortality and signs of wasting syndrome; egg production was alsoreduced and hatchabiity of eggs was < 2 %. Pheasants dosed with 0.1 pg/kg/week for 10weeks exhibited no adverse effects. Daily feeding of PCB congeners 126 (34-345) and 105 (234-24) for up to eight weeks caused hepatic porphyria, thymic atrophy (PCB 126 only) and markedmicrosomal cytochrome P450 enzyme induction in Japanese Quail (Coturnix coturnix), but noporphyria, 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 abird of prey.Field studies of PCDDs. PCDFs and PCBs in birds. In the Great Lakes, a toxicsyndrome observed in a number of fish-eating bird species, is referred to as GLEMEDS (GreatLakes embryo mortality, edema and deformities syndrome), and has been attributed to exposureto PCBs, PCDDs and PCDFs (Gilbertson et al. 1991). The syndrome was first recognized inLake 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 high2,3,7,8-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 also9contained 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) tobiological effects in colonial waterbird populations. Recently, Bosveld et at. (1994) and Van denBerg et at. (1994) reported high PCB levels in eggs of fish-eating birds breeding in the Rhineestuary, which correlated with various endpoints of exposure and toxicity, including CYP1Ainduction 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 studytoxicant exposure and effects (Elliott et at. 1989; Whitehead et at. 1 992a). Failure of a GreatBlue Heron colony in 1987 at Crofton, British Columbia coincided with a three-fold increase inmean 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 reproductiveoutcome among individual birds was determined (Elliott et at. 1989a). Heron embryos, collectedin 1988 at colonies with high, intermediate and low levels of PCDD and PCDF contaminationand incubated in the laboratory, did not exhibit any significant differences in hatching successamong 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 ofheron colony failure at Crofton in the late 1980s on the British Columbia coast and would havemasked other potential factors (Elliott et at. 1 989a); however, intensive observation of heronnests showed that mean time spent incubating was lower and greater between-nest variability inincubation 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 incubation10behaviour. Chemically mediated aberrant parental behaviour has been reported for a number ofspecies 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 andColumbia rivers of the British Columbia interior, contained significantly higher levels of 2,3,7,8-TCDD than eggs from nests upstream of the mills (Whitehead et at. 1993). Studies of ospreyproductivity 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 (Aixsponsa) 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 thelimited chemical data provided, Wood Ducks appear to be more sensitive to the effects of TCDDthan other wild bird species.Organochiorine pesticides. The acute vertebrate toxicity of DDT is low, the LD50to the Japanese Quail was 595 mg/kg (ppm). The cyclodienes are much more acutely toxic tovertebrates; for example, the LD50 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 thathas been best characterized in wildbirds (Anderson et at. 1975; Blus et at. 1974; Newton andBogan 1974; Blus et at. 1980; Custer et at. 1983; Elliott et at. 1988). DDE affects calciummetabolism 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 someraptors (Fyfe et a!. 1988; Peakall et at. 1991; Wiemeyer et a!. 1993), The chronic toxicology ofother organochiorines to wild birds has not been established. Dieldrin has been implicated in11reproductive 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 inScottish Golden Eagles (Aquila chrysaetos), However, the association of this level may have hadmore 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 wasassociated with effects on reproduction of American Kestrels (Henny et al. 1983). Egg levels ofHCB >5.0 mg/kg in Herring Gull chicks were associated with embryo mortality (Boersma et al.1986).The Bald EagleNatural historyThe Bald Eagle is an endemic North American member of the genus haliaeetus, the seaeagles. Bald Eagles are sexually dimorphic, adult females average 5.3 kg and 221 cm, and males4.3 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 considerablyfrom 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). Thebreeding 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 alternatenest is not clear, but may be to reduce parasite loads (Stalmaster 1987). Nests are always locatedin proximity to water. Nest trees are usually the dominant or codominant tree in the area inorder to provide a clear view of the territory and clear flight paths to feeding areas. Femaleeagles lay from one to three eggs, two being most common. Eggs are incubated for about 35days, and the chicks are dependent on their parents at the nest for food and protection for another72 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 ofreaching adult age are variable and may be less than 10 in some populations, such as in12Alaska, and as high as 50 % in more southern locations. Bald Eagles do not attain adult plumageuntil their 5th year, when they normally begin breeding (McCollough 1989)Eagles have a number of physical adaptations as predators. They have excellent visionand can reportedly detect other eagles flying at 23 to 65 km distance (Shlaer 1972). They killusing their powerful feet and talons, while food is torn apart by a large beak. They are powerfulflyers, particularly adapted for soaring in open country. Bald Eagles are opportunistic foragersand 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 beimportant in some areas (Vermeer et al. 1989; Knight et at. 1990; Watson et al. 1991).Population trends and critical factorsLike many other large predatory animals, Bald Eagle populations declined during the pastcentury over much of their North American breeding range (Stalmaster, 1987). Habitat loss anddegradation combined with intentional and accidental killing contributed to poor productivity andloss of breeding stock. In the early 1950s, populations of eagles and other birds of prey began todisappear from many areas. Classic work by Charles Broley (1947, 1958) showed a precipitousdecline 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 belowsustainable levels in many areas of the U.S. and Canada (Stalmaster, 1987). The low breedingsuccess of Bald Eagles and other birds of prey, which began in North America in the early1950s, coincided with the introduction of DDT and other organochlorine pesticides. The widelyaccepted paradigm for decline of the Bald Eagle and other North American raptor populationsstates that DDE persists, bioaccumulates and impairs reproduction via the mechanism of reducedeggshell quality (Grier, 1982; Peakall et at., 1991). Wiemeyer et al. (1984) determined that inBald Eagles, reproductive failure approached 100 % when DDE egg levels were greater than 15mg/kg. DDE egg levels of 5 mg/kg were associated with 10 % eggshell thinning, whilepopulations with less than 3 mg/kg exhibited no significant shell thinning and normal productionof young. However, those values were based on analyses of adled eggs which may tend to havehigher than average residues and may bias the estimate of critical values. In other birds of prey,13particularly European populations, loss of breeding stock to acute dieldrin poisoning, has beensuggested to be more critical than DDE-induced shell thinning (Newton et a!. 1992). Bald Eagleshave also been acutely poisoned by other pesticides, including dieldrin, and heavy metals, such asmercury and lead (Reichel et al. 1984), athough these effects were probably less critical topopulation decline. At any rate, eagle productivity has improved and populations have increasedin most areas, following strict regulation by the early 1970s of organochiorine use in NorthAmerica (Grier, 1982; Wiemeyer et a!. 1993). As a result in July, 1995, the U.S. Fish andWildlife 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 atnearby inland locations (Bowerman 1993), although, at least for Lake Superior populations,reduced food delivery to nestlings was an important factor (Dykstra 1994). Eagle populations inMaine generally exhibit low productivity, which has been related to high contamination by PCBsand DDE (Welch 1994). Along the lower Columbia River, low Bald Eagle breeding successcorrelated with high egg and plasma levels of DDE and PCBs; moderately high levels of 2,3,7,8-TCDD (tetrachloro dibenzo-p-dioxin) were also present in those eggs (Anthony et al. 1993).British Columbia Eagle PopulationsBased 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 onconcern for long term conservation of some populations (British Columbia Conservation DataCentre 1995).Hodges et at. (1984) estimated the resident breeding population of Bald Eagles on theBritish 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). BaldEagles are lured by the rich food resources and high biological productivity, both terrestial andmarine, of the Strait of Georgia, which is essentially a large estuary with nutrient input fromnumerous rivers, particularly the Fraser (LeBlond 1989); those rivers are also major salmon14spawning sites, which attract thousands of eagles each winter. Millions of waterbirds andshorebirds migrate through and winter in the region, which provides the major food supply forBald Eagles and falcons. The basin is surrounded by temperate rain forests which have beenextensively 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). Withincreased population growth and commercial activity, especially of coastal and estuarine areas inthe Georgia basin, habitat for Bald Eagle roosting and nesting will be continually threatened. Inaddition to those stesses, there are major pollutant inputs, particularly from pulp mills and otherwood processing industries.Coastal Bald Eagle populations in British Columbia apparently did not experience themajor declines that occurred elsewhere during the organochlorine era. Anecdotal information(based on discussion with naturalists, farmers and fisherman) suggests that in the Fraser Riverdelta, 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 ofnests since 1974, which they attributed mainly to increasing food supply in the form of Glaucous-winged gulls (Larus glaucescens). However, data derived from such comparisons requirescautious interpretation, as it may be more indicative of increased survey intensity and ability tofind nests (Henny and Anthony 1989). In the Okanagan Lakes region of interior BritishColumbia, Bald Eagles have disappeared as a breeding species (Cannings, 1987); orchard areasof the Okanagan valley received heavy DDT applications and wildlife samples from that area arestill highly contaminated (Elliott et al. 1994).Problem StatementAs a predator feeding at the top of marine and estuarine food chains, Bald Eagles areexposed to an array of persistent environmental chemicals, particularly chlorinated hydrocarbonsand mercury. There is a considerable body of literature on levels of organochiorine pesticides andtotal PCBs in tissues of Bald Eagles. However, there is very little published data on levels ofindividual PCB congeners, particularly the toxic non-ortho PCBs, or on levels of other significant15environmental contaminants including polychiorinated dibenzo-p-dioxins (PCDDs) andpolychiorinated dibenzofurans (PCDFs) in Bald Eagles. In addition, whereas significant progresshas been made in detennining critical levels of DDT-related compounds and mercury for BaldEagle eggs (Wiemeyer et al. 1993), there is no such information for other chlorinatedcompounds.The Strait of Georgia provides an interesting location to investigate the effects of PCDDsand PCDFs on eagle populations. Previous studies in the area showed that fish-eating birds, suchas Great Blue Herons, Double-crested Cormorants, Western Grebes (Aechmophorus occidentalis)and Common Mergansers (Mergus merganser), all of which are potential Bald Eagle prey, werecontaminated with high levels of PCDDs and PCDFs, but relatively low levels of otherorganochlorines (Elliott et at., 1989; 1992; Whitehead et at., 1990; 1992). Concentrations ofPCDDs and PCDFs in Western Grebes and in Surf Scoters (Melanita perspicillata), another eagleprey item, collected near some British Columbia coastal mills in 1990 were high enough towarrant advisories against their consumption by people (Whitehead et al. 1990). In Great Blueherons, episodes of poor breeding success in the late 1980s at a colony near a kraft pulp millwere associated with sublethal effects on embryos, including edema, reduced body weight andEROD induction which correlated well with levels of 2,3,7,8-TCDD (Beliward et al. 1990; Hartet at. 1991; Sanderson et at. 1994a). Coastal Bald Eagle populations feed heavily on marinebirds 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 suchas 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, suchas lead shot and pesticides, acquired by waterfowl feeding in other distant areas, such as thewestern 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 theLower Fraser Valley (Elliott et at. submitted).16There is, therefore, potential for exposure of Strait of Georgia Bald Eagles to potentiallyharmful levels of chlorinated organics and other toxicants. Positioned at the top of the food weband with a high public profile, Bald Eagles are an excellent sentinel species and indicator ofecosystem health. Thus, further research is warranted.Hypotheses 4 Objectives:Mortality studyHypothesis: The accumulation of persistent chlorinated hydrocarbons will affect the survival ofBald 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 ofchlorinated hydrocarbons versus other causes of death; to determine spatial and possibly temporaltrends in contamination.Embiyotoxiciry studyHypothesis: 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 gradientof chlorinated hydrocarbon pollutants and to relate the degree of exposure to biomarkers such asCYP1A induction; to document exposure by chemical analysis of yolk sacs.Bioaccumulation studyHypothesis: 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 toppredators in marine and estuarine food chains.Objectives: To determine spatial and temporal patterns of chlorinated hydrocarbons in BaldEagle eggs and to relate those levels to the diet and to sources; to determine criticalconcentrations of contaminants, particularly PCDDs and PCDFs in the eagle diet.Productivity study17Hypothesis: The accumulation of persistent chlorinated hydrocarbons in Bald Eagles impairsoverall reproduction through toxicity to embryos, reduce survival of nestlings or impaireddevelopment of the reproductive system.Objectives: To determine breeding success of a representative sample of eagles in the Strait ofGeorgia and reference locations and to relate breeding success to chlorinated hydrocarbon levelsin nestling blood samples; to examine the role of other factors critical to breeding success,particularly food supply.Overview of the thesisThis thesis represents the results of a four year field and laboratory study of chlorinatedhydrocarbon exposure and effects in Bald Eagle populations on the coast of British Columbia. Inthe first chapter, mortality and the role of chlorinated hydrocarbons are examined throughautopsy and liver residue analysis of eagles found dead and dying from 1989 to 1993 in BritishColumbia. Chapter two presents the results of a laboratory incubation study of in ovo effects ofPCDDs, PCDFs and PCBs in an environmental exposure gradient. Contaminant levels in yolksacs 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-effect-level (LOEL) for TCDD-toxic equivalents in eagle eggs. Chapter three presents contaminantresidue levels for Bald Eagle eggs and prey items. Patterns, trends and sources are discussed anda simple bioaccumulation model used to relate levels in eagles to those in their food chain. InChapter four, the results of productivity studies and contaminant levels in nestling plasmasamples are presented. Relationships between breeding success, contaminant levels and othervariables, particularly food supply, are discussed.18CHAPTER 1CHLORINATED HYDROCARBON LIVER LEVELS AND AUTOPSYDATA FOR BALD EAGLES FOUND DEAD OR DEBILITATED, 1989-1993.The objective of this study was to determine the degree of chlorinated hydrocarbonexposure of adult and juvenile Bald Eagles and to assess spatial trends in contamination.Statistical examination of relationships among environmental contaminant levels and cause ofdeath 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 notincluded as part of the thesis. In this chapter, the results of autopsies and analyses of PCBs andorganochiorine 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 arepresented and evaluated.Materials and MethodsSample collectionSpecimens collected for this study were part of an overall investigation into the healthstatus of Bald Eagles in British Columbia. Carcasses were obtained by writing to potentialcooperators, including government and non-government agencies, veterinarians and wildliferehabilitators and by placing advertisements in periodicals. Sick, injured and deceased BaldEagles were thus obtained from all of the above sources. Specimens were received and initiallyexamined at the Pacific Wildlife Research Centre and then shipped on ice to the IslandVeterinary Hospital, Nanaimo, British Columbia, where they received a complete autopsy byDr. K.M. Langelier.19The 484 eagles received were grouped by geographical area as follows: lower Fraservalley, Strait of Georgia, Johnstone Strait, west coast Vancouver Island and north coast. Atotal of 59 individuals were analyzed for organochiorines and PCBs (Figure 1.1, see alsoAppendix 1.1). Specimens for analyses were selected in order to provide a reasonablyrepresentative sub-sample, based on age, sex, and collection location. Other criteria were alsoconsidered such as a preliminary diagnosis of non-specific poisoning or proximity of the carcassto an industrial pollutant source. Some eagles were also analyzed for organochiorines duringinvestigations of suspected poisonings by lead or anticholinesterase pesticides. Birds foundFigure 1.1 Locations where eagles were collected in British Columbia, 1989-93,and analyzed for chlorinated hydrocarbons (N = 59).20dead during the breeding season in the Strait of Georgia, and therefore likely to be residentbirds, 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 theStrait of Georgia or Johnstone Strait 2) collection date in late spring or summer, i.e. residentbirds 3) breeding age birds 4) high organochiorine levels. Criteria were set to maximizechances of analyzing eagles which had been exposed to pulp mill pollutants.Based on elevated levels of total PCBs, nine samples were selected for high resolutionGC/MS analysis of non-ortho PCB congeners. Linear regressions were determined betweenconcentrations of non-ortho PCBs and total PCBs for the nine livers analyzed, in order toestimate values for the other ten livers which had been analyzed for PCDDs and PCDFs andthus to estimate TCDD toxic equivalents. Regressions were not significant for PCBs 77, 81and 37, but were significant for PCBs 126 and 169:PCB 126 (ng/kg) = 92 [sum-PCB5 (mg/kg)] + 310, r2 = 0.660, p<O.OlPCB 169 (ng/kg) = 27 [sum-PCBs (mg/kg)] + 75, r2 = 0.584, p <0.05Chemical analysisCarcasses were stored at -20° C until postmortem examination. Tissue samples werefrozen at -20°C in chemically-cleaned (acetone/hexane) glass jars, frozen, and shipped to theNational Wildlife Research Centre (NWRC), Hull, Quebec, for analysis in the laboratory ofDr. 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 sodiumsulfate and colunm extracted with 50% methylene chloride in hexane. After extraction, theeluate was concentrated on a rotovapor, further mixed with hexane and a 0.5 ml sample takenfor lipid determination (removal of solvent and weighing of residue). The remaining extractwas then cleaned up and separated into three fractions by Florisil chromatography. Thefractions were analyzed by gas chromatography-electron capture detector using a 60m DB-521capillary 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 PCBcongeners was effected by using a calibrated internal PCB standard solutions. Detection limitswere 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 resolutionGC/MS using a Hewlett-Packard 5987B with a 30 m DB-5 capillary GC column according tomethods described in Norstrom et al. (1990) and Norstrom and Simon (1991). The methodemployed gel permeation-carbon chromatographic clean-up and the use of13C2-labelled internalstandards for quantification.Analysis of PCDD/PCDFs and non-ortho PCB in livers from other years were carriedout according to methods in Letcher et al. (in press). The method involves neutral extractionfollowed by removal of lipids and biogenic compounds by gel permeation chromatography andalumina column cleanup. Separation of PCDDs, PCDFs and non-ortho PCBs from othercontaminants was achieved using a carbon/fibre column; further separation of PCDDs/PCDFsfrom the non-ortho PCBs was effected by Florisil column chromatography. Quantitation wasperformed with a VG Autospec high resolution mass spectrometer linked to a HP 5890 Series IIdata system. Each sample was spiked with‘3C12-labelled 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 ofinternal standard recoveries. Two other‘3C12-labelled standards (1 ,2,3,4-TCDD and 123789-HxCDD) were added to the cleaned PCDD/PCDF extracts and PCB 112 to the non-ortho PCBfraction, just prior to analysis to serve as recovery standards, for quantification of internalstandard recoveries. Recoveries of132-PCDDs/PCDFs/non-ortho PCBs were calculated bycomparing the integrated areas of the labelled internal standards and the areas of the recoverystandards in the samples to the areas of those compounds measured in the external standard22mixture, analyzed along with the samples. Results were generally accepted when recoveries oflabelled standards were between 70% and 120%.Statistical analysisOrganochiorine pesticide data were transformed to common logarithms and geometricmeans 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 multiplecomparison procedure (MCP). To test for an association between residue levels and cause ofdeath, 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 (Ahlborgetal. 1994).ResultsAutopsy resultsThe diagnosed cause of death for each individual bird analyzed for organochiorines isincluded in Appendix 1-1. Autopsy results for the 59 Bald Eagles analyzed for organochiorineswere compared to the total of 484 examined in the broader study with the causes grouped intotwelve categories (Figure 1.2). The graphs indicate that the subset for analysis was reasonablyrepresentative of the range of mortality factors. Only two minor categories, falling from thenest and infectious disease, were not represented.There were no statistically significant associations between any of the chlorinatedhydrocarbon levels and cause of death. However, given the relatively small sample size, evenwithin the Strait of Georgia, and the variance in the residue levels, the probability of detectinga significant association was low.23Clinical diagnosistraumaelectrocutionundeterminedeagle attackfell from nestdrowning_____________________liii186________ J49_ __ _ _ __46_30j25_I24122120_18117PercentBald eagles submitted to Island Veterinary Hospital, N484Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed comparedOrganochiorines and total-PCBsto the complete set of birds received.Organochiorine pesticides and total PCBs were generally low; most eagles had DDE andPCB levels < 5.0 mg/kg (Figure 1.3). However, a few birds had elevated levels of DDE andtotal PCBs (>50 mg/kg) and chiordane-related chemicals (>1.0 mg/kg) (Appendix 1.1)..•.: DDE:::totaI PCBS.-••.••••••-••.•• .••., ‘.-.••••- -• -.8%- -... -Figure 1.3 Numbers of Bald Eagles showing different DDE and PCB levels in livers (N=59)Dveh. collision14.. ‘j2 *:.•(•n•O ófbátdéáglej):.::..:..... ...::•.•..•••,•••••.•••••..I I I•infectiousBald eagles analyzed for OCsIPCBs, N590 5 10 15 20 25 0 5 10 15 20 25Percent25- I. — —•44.%l1-LiConcentration (mg/kg, wet wt.),-1024Quantifiable levels of total PCBs, trans-nonachlor and oxychiordane were present in all59 samples analyzed while DDE was present in 98 % of the samples. There were quantifiablelevels 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 Straitsamples, followed by the Strait of Georgia, with the other four sites all being lower. MeanDDE levels were significantly higher in samples from Johnstone Strait compared to the lowerFraser valley.1000 : : : : : -100—ci) E : : : : : : : : : : :1 0______Ainddatapoints.-. .maverages0.0001 I I I I I I I I IInn ——-: - •——ci,--: : : : : : : : : nd. data pointsC.)averages-— I . I I I IJan Feb. Mar. Apr. May Jun. Jul. Aug Sep. Oct. Nov. Dec.MonthFigure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month25Individuals with elevated organochiorines were found mainly in late spring or earlysummer (Appendix 1.1). Concentrations of DDE and total PCBs in eagle livers tended toincrease throughout the winter, peak in April, level off and even decline slightly in summer(Figure 1.4).1000- = = = - - = .: - = = - = =100 E C-:0)0)_E 10— C C C :::,:: : :: :: C : :Co----.41- = = = C = = = : = = = = = :: = = = = = C = = , =E E E 4E E E E E E:::::.:::.:::.a:::.: ..0C.)0.10.01—— I I I I I100- ---10:::::::::::::::::::::::::.0) - -——-.C .. ..... — I -o 1 —= q-— ICa) —C.)C _0c. 01—C.) —..0.01— —— ——. I I I I0 1 2 3 4 5*Bald eagle body conthtion* Scale: 0 poor, 5 - excellentFigure 1.5 Liver DDE and PCB residue levels in relation to body condition26Although eagles with higher residue levels of DDE and PCBs tended to weigh less thanthose 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 anumeric scoring of body condition did produce statistically significant negative relationships forDDE (f=7.4 p=O.009) and PCBs (f=8.5 p =0.005 (Figure 1.5).Non-ortho PCBsLevels of three non-ortho PCB congeners and two mono-ortho PCB congeners, PCBs105 (234-(234-34) and 118 (245-34), which are present at relatively high concentrations andalso 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 PCDFsThe most contaminated individuals were from near pulp mill sites, Powell River andCampbell 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 equivalentsTEQ results varied widely among the three sets of factors. Highest values wereconsistently produced using the chick embryo hepatocyte derived numbers, followed by Safe’sand then the WHO TEFs (Table 1.4). The TEQ5WHO ranged from 53 to 2740 ng/kg. Twobirds had liver TEQs110 > 2000 ng/kg, while an additional two birds had liver TEQSWHO >1000 ng/kg.27Table1.1Organochiorineresiduelevels(mg/kg,wetweight),geometricmean±95%confidenceintervals,inliversfromBaldEaglesfounddeadinBritishColumbia,1988-1993.LocationN%fat%H20Total PCBsDDEtrans-OxychlordaneMirexB-HCHDieldrinHCBnonachiorLowerFraser104.1710.609a0.5420.0460.010.0040.0030.0070.01Valley(3.3-5.1)(70-72)(0.253-1.47)(0.186-1.58)(0.022-0.096)(0.005-0.020)(0.002-0.009)(0.001-0.016)(0.002-0.03)(0.006-0.016)Straitof333.273146ab1.310.0940.0180.0080.0110.0140.014Georgia(2.8-3.7)(72-74)(0.811-2.62)(0.057-1.56)(0.057-0.156)(0.01-0.031)(0.004-0.014)(0.006-0.021)(0.008-0.027)(0.009-0.02)Johnstone93.2753•36b4.930.3210.0520.0300.0420.0320.023Strait(2.0-5.1)(73-77)(0.685-16.5)(0.819-29.7)(0.058-1.78)(0.008-0.337)(0.006-0.153)(0.007-0.270)(0.005-0.188)(0.006-0.097)WestCoast23.6700775ab1.010.0730.0130.0070.0130.0130.011VancouverIs.*NorthCoast44.6670.689’1.140.0790.0140.0060.0090.0130.014(2-11)(62-73)(0.29-1.64)(0.255-5.07)(0.042-0.149)(0.006-0.033)(0.003-0.015)(0.005-0.016)(0.007-0.023)(0.003-0.057)Northern12.1730.4290.420.0670.0110.0050.0020.0090.011Interior*a.b-meansthatdonotsharethesameletter aresignificantlydifferent(P<0.05).NOTE:significant differencesamongsiteswerefoundonlyforDDE*-insufficientsamplesizetocalculateconfidenceintervalNJTable 1.2 Selected non-ortho and total PCBs in Bald Eagle livers collected from the southcoast of British Columbia (wet weight).(a)- 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 equationsNon-ortho PCB Minimum Detection Limit (MDL) = 3 ng/kg wet wt; mono-o,iho PCB MDL = approx. 0.5 pg/kgwet. wgt.A = adult, ly, 2y, 3y = age of subadultsUndet. = Undetermined, Inanit. = Inanition, Electro. = Electrocution, Tox. Pb = Toxicosis, Asphyx. =Asphyxiation, Drown. = DrowningLocation Date Sex! BC(a) InitialAge EtiologyTotal CommentsPCBsNon-ortho PCBs#771) #126 #169(ng/kg) (mg/kg)Port Hardy 27 Jun/89 F/A 1 Undet. 1270 1170 221 6.42Port Hardy 6 Mar/90 M/3y 1 Undet. - 349 86 0.425Port Hardy 2 Apr/90 M!ly 0 Inanit. 2300 4800 2180 43.8Port Hardy May/93 F/A 1 Inanit. 2000 2550 533 65Port Hardy May/93 F/A 4 Trauma 1070 1490 368 12Campbell R. 9 Jun/90 F/2y 5 Tauma - 357 88 0.515Campbell R. 31 Jul/90 F/3y 0 Trauma 248 688 160 7.52Campbell R. 16 Apr/93 F/A 4 Electro. 738 9960 2640 71.7Powell R. 26 Apr/90 M/A 1 Electro. - 5820 1690 60Powell R. 18 Jun/90 F/A 3 Tox.Pb - 499 130 2.06Comox 13 Jun/90 M/3y 4 Trauma - 561 148 2.72Denman Isl. 19 Jul/90 M/A 3 Asphyx. - 426 108 1.26Bowser 7 Jul/90 F/4y 1 Tox.Pb - 2640 759 25.4Coombs 3 Mar/90 M/A 1 Tox.Pb - 361 90 0.558Nanoose 26 Apr/90 F/A 2 Trauma - 496 129 2.02Nanaimo 8 Feb/90 F/A 5 Electro. 395 370 56 4.56Sechelt 7 May/90 F/A 3 Electro. - 783 213 5.15Dent Isl. 5 Apr/93 F/ly 1 Drown. 870 472 138 9.27Victoria 14 Nov/92 M/A 1 Trauma 1620 2240 523 7.94*(c)*Pb exp.* Hg tox.* Pb exp.* Pb exp.* Pb exp.* Pb tox.* Pb tox.* Pb exp.* Hgexp/Pb-exp.29Table1.3ConcentrationsofselectedPCDDsandPCDFsinBaldEagleliverscollectedfromthesouthcoastofBritishColumbia(ng/kg, wetwt.)23478/Initial2378-12378-123678-2378-13489-LocationDateSex/AgeBC(a)EtiologyTCDDPnCDDHxCDDTCDFPnCDFCommentsPortHardy27June/89F/A1Undet.33718110115*PortHardy6Mar/90M/3y1Undet.510920tracePortHardy2Apr/90M/3y0Inanit.772412804149*PortHardyMay/93F/A1Inanit.54232350320*PortHardyMay/93F/A4Trauma17491414511*CampbellR.9June/90F/2y5Trauma182537338CampbellR.31July/90F/3y0Trauma305676110*Pb-exp.CampbellR.16Apr/93F/A4Electro.21279321208105*Powell R.26Apr/90M/A1Electro.392142043603375Hg-exp.PowellR.18June/90F/A3Tox:Pb41831846327Hg-tox/Pb-tox.Comox13June/90M/3y4Trauma21511696017Pb-exp.DenmanIsi.19July/90M/A3Asphyx.49922957830Hg-exp/Pb-tox.Bowser7July/90F/4y1Tox:Pb263603205015152Hg-exp/Pb-tox.Coombs3Mar/90M/A1Tox:Pb6910426Pb-tox.Nanoose26Apr/90F/A2Trauma2357902811Pb-exp.Nanaimo8Feb/90F/A5Electro.254051155*Sechelt7May/90F/A3Electro.2919993624138Hg-exp./Pb-tox.Dent Is!.5Apr/93F/ly1Drown.4182083513*Victoria14Nov/92M/A1Trauma30771081518*(a)BC-BodyCondition:0-emaciated,1-thin,2-fair,3-good,4-verygood,5-excellent(b)*13489-PnCDFnotincluded(c)trace=<2ng/kgwetwt.MDL-MinimumDetectionLimit(signal/noise)=3ng/kgwetwt.Undet.=Undetermined,Inanit.=Inanition,Electro.=Electrocution,Tox.Pb=LeadToxicosis,Asphyx.=Asphyxiation,Drown.=DrowningA=adult,ly,2y,3y=ageofsubadultsTable1.4Comparisonof TEQscalculatedfromselectedpCDDS(a),PCDFs(b),non-orthoandmono-orthoPCB5(d)levelsinBaldEagleliverscollectedfromthesouthcoastof BritishColumbia(nglkg, wetwt.(e)).TEQsLocationDateSex/AgeBC0InitialEtiologySafeCEHOI)WHO’CommentsPortHardy27June/89F/A1Undet.8521560276**)PortHardy6Mar/90M/3y1Undet.9520453+0)PortHardy2Apr/90M/3y0Inanit.5490117001220**PortHardyMay/93F/A1Inanit.41207430832**PortHardyMay/93F/A4Trauma10801920302**Campbell R.9June/90F/2y5Trauma13026883+Campbell R.31July/90F/3y0Trauma7111140197**Pb-expCampbell R.16Apr/93F/A4Electro.7460131002440**PowellR.26Apr/90M/A1Electro.6550101002740+Hg-expPowell R.18June/90F/A3Tox:Pb384677193+Hg-tox/Pb-expComox13June/90M/3y4Trauma334621155+Pb-expDenmanIsl.19July/90M/A3Asphyx.335552205+Hg-exp/Pb-expBowser7July/90F/4y1Tox:Pb422059601430+Hg-exp/Pb-expCoombs3Mar/90M/A1Tox:Pb9922160+Pb-toxNanoose26Apr/90F/A2Trauma274491135+Pb-expNanaimo8Feb/90F/A5Electro.472832129**Sechelt7May/90F/A3Electro.8151180417+Hg-exp/Pb-expDentIsl.5Apr/93F/ly1Drown.282453110**Victoria14Nov/92M/A1Trauma11102100394**(a)2378-TCDD,12378-PnCDD123678-HxCDD(b)2378-TCDF,23478/13489-PnCDF(e)PCBconeners#77,126and129(d)PCBconeners#118and105(e)*PCDIj,PCDFs,non-orthoPCBsMinimumDetectionLimitBC-BocyCondition:0-emaciated,1-thin,2-fair,3-good,4-verygood, 5-excellentçsnal/noise=3(TEFsfromKennedyetal.inpresssfromsafe1**13489-PnCDFnotinchidedinTEQcalculations(I)TEFsfromAhltorgetal.,1994(k)PCBconeners#126&169calculatedfromregressionequations;#77notmcludedin‘EQcalculationsA=adultly2y,3y=ageof subadultsUndet.=‘Undetermined,Inanit.=Inanition,Electro.=Electrocution,Tox.Pb=Toxicosis,Asphyx.=Asphyxiation,Drown.=DrowningDiscussionChlorinated hydrocarbon levels in livers of Bald Eagles tested for this study weregenerally low; however, a small number of birds found dead or debilitated from the Strait ofGeorgia 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 meanthat their death was a direct result of toxicity due to those chemicals. Most of the eagles withhigher chlorinated hydrocarbon levels were in poor body condition, indicating lipid andcontaminant mobilization. Body weight was negatively correlated with liver organochiorinelevels in other studies (Cooke et al. 1982; Reichel et al. 1980). A variety of factors cancontribute to weight loss, including: seasonal utilization of fat stores, poor foraging abilities ofjuvenile birds, a debilitating injury, disease or toxicosis, and the anorexic effects of chemicalssuch as lead, dieldrin and TCDD. Discriminating among these factors in a sample of wildbirds is difficult.Body weight loss per se can, however, be symptomatic of toxicity. A number of thebirds 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 weightand appear clinically to have starved (U.S. Fish and Wildlife Service 1986). Dieldrin exposurecan induce fasting (Heinz and Johnson, 1982). Weight loss due to fasting, referred to aswasting 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 andTEQ5WHO (2440, 2730 and 1430 ng/kg) were found in the vicinity of pulp mills on south eastVancouver Island. Proximal causes of death were electrocution in two cases and lead poisoningin the third. However, in one particular case, an adult male eagle from Powell River, totalTCDD-toxic equivalents were calculated to be from 2740 TEQ5WHO to 6550 ng/kg TEQSSafe. Asample 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 body32condition, perhaps indicating that it suffered from wasting syndrome. There are no data ontissue levels of TCDD-like compounds which could be diagnostic of acute toxicity. LD50sreported for 2,3,7,8-TCDD are 240 ng/kg in chicken and 1350-2 180 ng/kg in pheasantembryos (Peterson et al. 1993). Lethal doses in adult birds are estimated to be one to twoorders of magnitude higher (ibid. 1993).A pattern of increasing mean contaminant levels in spring partly reflects normalseasonal lipid dynamics. Late summer and fall deposition and winter mobilization of fat istypical of temperate climate species, adapted for winter survival (Stalmaster and Gessaman1984). Seasonal deposition and mobilization of lipids and lipid soluble contaminants suchDDE, PCBs and dieldrin was shown in three species of predatory birds monitored for manyyears in Great Britain (Cooke et al. 1982). Starvation and associated lipid and contaminantmobilization 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 first-year 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 ofgreater conservation interest in birds which have reached breeding age. Juvenile eagles,particularly first-year birds, may also have lower chlorinated hydrocarbon levels than adults, asthey have had less time to reach pharmacokinetic equilibrium with dietary residues, which tookup to two years in Great Lakes herring gulls (Anderson and Hickey, 1976). Juvenile eagles eatmore fish (Stalmaster 1987), which would also tend to have lower contaminant levels than fisheating 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 91and 96 mg/kg. None of the birds had PCB levels in livers > 100 mg/kg, considered indicativeof toxicity (Cooke et a!. 1982). One bird had levels of oxychiordane in liver > 2 mg/kg andtrans-nonchior levels > 7 mg/kg. Diagnostic liver levels of oxychiordane are not available;33brain levels of 1.1 - 5.0 mg/kg indicate acute toxicity (Stickel et al. 1979). In an earliersample of nine eagles found dead, 1969 to 1973, from British Columbia, one bird had 179mg/kg DDE and 23.7 mg/kg dieldrin (Friis, 1974), well above the level of 5 to 10 mg/kgdieldrin in liver, indicative of acute poisoning (Cooke et al. 1982). None of the eagles in thepresent sample had elevated dieldrin levels, indicating an improvement in dieldrincontamination of the eagle foodchain. However, the presence of potentially toxic levels ofDDE in livers of British Columbia Bald Eagles more than 20 years after DDT was heavilyrestricted in North America raises questions regarding sources. A number of hypotheses havebeen suggested in the literature to account for sources of continuing high levels of DDT in theenvironment. Recent data show that DDT can persist at high levels in soils and foodchains inareas of former intensive use or manufacturing (Blus et al. 1987; Elliott et a!. 1994). Eaglesmay also acquire some DDT from feeding on migrant waterbirds, which are exposed toongoing use in Latin American wintering areas (Fyfe et a!. 1990). Finally, on the Pacificcoast, elevated DDE levels in seabirds, such as storm-petrels, important seasonal prey items ofeagles nesting on their colonies, indicates long-range transport from recent use in Asiancountries (Elliott et a!. 1989). Elevated PCBs in some eagle livers likely originate fromindustrial sources in the Georgia basin, as PCBs were significantly elevated in samples from theStrait 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, itis possible that the anorexic effects of TCDD are not Ah-receptor mediated (Tuomisto andPohjanvirta 1991). Therefore, it would be interesting to know if any biomarkers of Ah-liketoxicities were activated in eagles with high liver TEQs. Indirect indications, at least ofCYP1A induction, may be inferred from examination of TCDD/TCDF ratios, which variedgreatly between eagles with high versus low TCDD exposure. For example, TCDF levels aremuch lower in the three birds with the highest TCDD levels (212, 392, 263 ng/kg in liver); themean TCDD/TCDF ratio for those three birds was 74. In contrast, the mean TCDD:TCDFratio is 0.17 for the three birds with the lowest TCDD levels (5,6,4 ng/kg in liver). The34TCDD/TCDF ratio in the high TCDD birds is also markedly different from ratios observed ineggs. Mean ratios in eagle eggs were 0.58 at Powell River and 0.32 in Jolmstone Strait (Table2.1). This shifting ratio may indicate that hepatic cytochrome P450 enzymes have beeninduced 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 presentand inducible in Bald Eagle chicks (Chapter 3) and should, therefore, also be inducible in adulteagles. CYP1A1 was recently shown to be the protein responsible for TCDF metabolism inrats and humans (Tai et at. 1993). Alternatively, higher liver TCDD concentrations in morehighly 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, possiblyCYP1A2, 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 PCBpattern (Brown 1994). The PCB congener pattern between birds classified as good versus poorbody condition is compared in Figure 1.6. As discussed above, birds in poor condition havehigher chlorinated hydrocarbon levels in liver, because of lipid and contaminant mobilization,and thus, hepatic P450 enzymes may have been induced. Differences in mean percent totalPCBs 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 lowerchlorinated compounds was consistently lower and the higher chlorinated compoundsconsistently higher in the poor condition group. CYP1A induction should increase themetabolism 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 wellas 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 wereinduced in eagles, suggesting the possible activation of other Ah-mediated processes.3520150C.)0F°04-.005PCB congenersFigure 1.6 PCB congeners in Bald Eagle livers expressed as percent of totalPCBs 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 (< 5mg/kg) levels of DDE and PCBs, and even lower levels of other organochiorines. However, afew birds had DDE levels diagnostic of acute poisoning, more than 20 years after regulatoryrestrictions on DDT usage in North America. At least one eagle found near a bleached kraftpulp mill had liver TEQWHO levels potentially indicative of acute toxicity. Differences inTCDD/TCDF ratios in birds with high 2,3,7,8-TCDD levels may indicate hepatic cytochromeP450 induction and TCDF metabolism in those birds.Because of the selection criteria, samples analyzed for PCDDs and PCDFs were biasedtowards birds with a higher probability of such exposure. Nevertheless, 4/19 (21 %) of eaglestested had > 1,000 ng/kg TEQSWHO in their livers. All of those birds were of reproductive age0rjZ qç’ of36found during the breeding season. This may indicate that acute exposure to TCDD-likecompounds has removed a component of the breeding eagle population in the Strait of Georgia.AcknowledgementsDr. 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 thestudy. Funding was provided by the Canadian Wildlife Service.37MeanS.D.N1010101010STRAITOFGEORGIA43.75Electro.4.176.956.9975niaIntra.Agg. wetwt.)LocationDateSexAgeWt(kg)InitialEtiology%Fat%H20DDEOxychlor.t-Nonachl.TotalPCBsCommentsFRASERVALLEYCoalHarbour12Mar/90F3y.34.09Electro.2.570.60.5760.0100.0660.617CoalHarbour26Feb/91FAd.45.00Trauma3.872.70.4380.0070.0450.318NewWest.28Mar/92Mn/a44.90Undet. 10Campbell R.16Apr/93FAd.Campbell R.18Apr/90MAd.Campbell R.18Apr/90MAd.Campbell R.9Jun/90F2y.Campbell R.9Jun/90F3y.Campbell R.31Jul/90F3y.Campbell R.9Feb/93M3y.Campbell R.11Feb/93MAd.Campbell R.Mar/93MAd.Merville11Feb/93FAd.PowellR.9Apr/90F3y.PowellR.26Apr/90MAd.PowellR.18Jun/90FAd.Comox28May/90FAd.Comox4Feb/92n/aIy.Comox13Jun/90M3y.DenmanIsl.19Jul/90MAd.Bowser7Jul/90F4y.Coombs3Mar/90MAd.0.4390.12271.7890.0110.0790.761Pb-exp.0.0140.0431.156Pb-exp.0.0090.0500.5150.0120.0751.0310.0900.4577.516Pb-exp.0.0050.0180.1420.0030.0310.3540.0040.0160.1430.0140.0790.591Pb-tox.0.0490.2813.241Pb-exp.0.1562.65660.047Hg-exp.0.0130.1422.059Hg-toxlPb-tox.0.0090.0750.6540.1000.3273.0250.0330.1612.729Pb-exp.0.0120.0811.261Hg-explPb-exp.0.7832.20725.423Hg-exp/Pb-tox.0.0070.0430.558Pb-tox.JOHNSTONESTRAITAppendix1-1,cont.LocationDateSexAgeBCWt(kg)InitialEtiology%Fat%H20DDEOxychior.t-Nonachl.Total PCBsComnientsMeanS.D.Sechelt7May/90FAd.3n/aElertro.3.770.43.7340.0040.1665.152Hg-exp/PB-exp.Nanaimo8Feb/90FAd.55.45Electro.5.866.610.5880.0370.4094.558Nanaimo16Mar/93M1y.13.10Inanit3.383.70.4230.0080.0260.282UnionBay11Mar.93FAd.56.15Electro.3.873.01.1620.0150.0800.727Nanoose26Apr/90FAd.23.64Undeter.trauma1.475.611.4610.0100.1152.022Pb-exp.Yellowpointn/aMAd.44.50Gunshot3.872.31.7410.0050.0320.679GabriolaIsl.n/aMn/a23.41Electro.2.568.80.4220.0040.0540.712GabriolaIsl.31Jan/91MAd.33.86Drown.1.972.300.0030.0070.172PB-exp.GabriolaIsi.14May/93Mn/a02.40Gunshot2.178.70.6630.0040.0290.367Ladysmith6Mar/90M3y.23.64Toxic:Pb3.172.42.7130.0390.1521.663Pb-exp.Duncan12Feb/90MAd.23.64Gunshot2.974.81.1640.0130.1021.356Pb-tox.Saanich15Jan/93n/an/a3n/aTrauma/Electro.3.670.30.1030.0030.0060.123Dent1st.5Apr/93F1y.1n/aDrown.4.575.25.6750.2270.4179.269Victoria14Nov/92MAd.13.20Trauma/Inanit.5.477.215.3010.1260.4147.9483.573.17.5430.0680.2716.6071.23.7417.9770.1530.56615.827333333333333PortMcNeil12May/93FAd.44.70Suspecttoxic.4.873.00.6110.0080.0430.328Sointula14Feb/93n/aAd.n/an/aNodiagnosis3.773.60.6230.0130.0660.829PortHardy27Jun/89FAd.13.52Undeter.4.472.28.5050.0890.6036.421PortHardy6Mar/90M3y.1n/aUndeter.1.479.20.6620.0050.0330.425PortHardy2Apr/90M1y.02.50Inanit.9.371.895.5402.3456.18143.746PortHardy1May/90FAd.44.55Veh.Coll.1.778.41.8930.0110.0971.719Pb-exp.PortHardy20Oct/92FAd.44.50Infection2.675.91.4620.0120.1101.256PortHardyn/aFAd.44.50Trauma3.473.416.3550.1591.30412.010PortHardyn/aFAd.13.70Inamt.2.376.7186.1831.2807.41665.0083.774.934.6480.4361.76114.6382.32.5860.8590.7792.73622.117999999MeanS.D.N WECOASTVANCOUVERISLANDTahsis7Jul/90FAd.55.45Undeter.trauma2.571.71.3890.0180.0991.451Hg-exp.NORTHCOASTOceanFalls1Feb/90M2y.54.55Qu.Char.Isl.n/aFn/a5n/aQu.Char.Isl.4Mar/91MAd.3n/aSandspit29Jul/90M4y.33.86MeanS.D.N3.869.71.0651.302.010.324222Electro.7.766.70.750Electro.5.765.11.871Undeter.trauma3.965.52.639Veh.Coll.2.771.50.4525.067.21.4281.892.550.877444NORTHERNINTERIORSmithers11May/90FAd.34.77Intra.Agg.BC-BodyCondition:0-emaciated,1-thin,2-fair,3-good,4-verygood,5-excellentn/a-notavailableND-notdetected;detectionlimit=0.0005mg.kgwetwtAppendix1-1,cont.MeanS.D.NJokervile21Jan/91MAd.34.00Electro. 40.0110.0540.0760.0220.0930.1090.0790.0490.0830.022 40.0670.4140.9320.5180.683Pb-exp.1.0610.868Pb-exp.0.3590.7430.259 40.429Pbexposed2.173.10.42CHAPTER 2BIOLOGICAL EFFECTS OF CHLORINATED HYDROCARBONSIN BALD EAGLE CHICKSThis study of embryotoxicily was designed to investigate whether in ovo exposure toPCDDs, PCDFs and PCBs was impacting hatching success and affecting a variety ofbiochemical and morphological parameters in Bald Eagle chicks. The aim of the study wasalso to estimate concentrations of PCDDs and PCDFs in Bald Eagle eggs which would beindicative of no-observed-efffects (NOEL) and lowest-observed-effect (LOEL) levelsThe results presented in this Chapter represent an extensive collaborative study withother laboratories. Contributions of those laboratories and of the principle investigators areidentified in the Materials and Methods, while technical contributions are included in theAcknowledgements. The concept, study design, field work, statistical analyses, calculations,graphic representations and other manipulations of data were performed by me. A version ofthis chapter has been acceptedfor publication (Elliott et al. in press).Materials and MethodsSample collectionBald 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 a25 km radius of a kraft pulp and paper mill, and generally within the effluent impact zone ofthe mills, as defined by fisheries closures due to dioxin contamination (Harding and Pomeroy1990). Eggs were collected from two nests in the Fraser River estuary (Map Nos. 1-2); at least500 km downstream from where effluent is discharged into the Fraser River from four kraftpulp and paper mills. An area of the west coast of Vancouver Island, Clayoquot Sound (MapNos. 10-14), was used as a reference site; there are no major industrial discharges to the41sound, 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.42Usually one egg was taken from each nest; the smallest egg in the clutch, presumablythe 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 wereaccessed 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 asrequired from thermos bottles. Within eight hours of collection, the eggs were transferred intoa battery powered CurfewTM incubator kept at a temperature of 34°C. The eggs were rotatedabout hourly and turned on their long axis twice daily.Within 72 hours (normally within 24- 48 hours) the eggs were brought to thelaboratory at the Department of Animal Science, University of British Columbia, where theywere candled to determine fertility and placed into a Humidaire incubator maintained at37.2°C with a relative humidity of 82-84 %. The eggs were rotated once per hour and turnedtwice a day in opposite directions on their long axis. At pipping the eggs were placed into ahatcher.Sample preparationWithin 24 hours of hatching, the birds were weighed, blood drawn by cardiac punctureusing a heparinized syringe and the bird sacrificed by decapitation. The yolk sac was removedand frozen. The liver was removed, weighed and separated as follows: 0.25 g from tip of leftlobe for Vitamin A analysis, 0.10 g from tip of the right lobe for porphyrin analysis; thesesamples were then frozen. The remaining liver was used to prepare microsomes. Variousorgans 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 tissueswere fixed in 10 % buffered formalin for histological examination: right kidney, bursa, thymus,spleen, gonads, lung, heart, intestines, thyroid and adrenal glands. Tissues were processedroutinely and embedded in paraffin blocks. Sections were cut at 6 urn and stained with43hematoxylin and eosin and examined by light microscopy. The amount of lymphoid tissue wasestimated based on follicular size and cell density of cortex and medulla in the bursa, on thedensity of white pulp in the spleen and on the thickness of the cortex and cell density in thethymus. The number of mitoses in all lymphoid organs and the number of necrotic cells in thebursa and thymus were counted in five fields at 600 X magnification. The level ofextramedullary haematopoiesis was assessed in the spleen.Chemical analysisBald Eagle yolk sacs were analyzed for PCDDs, PCDFs and non-ortho PCBs at theNational Wildlife Research Centre, Hull, Quebec, in the laboratory of Dr. R.J. Norstrom. Theanalyses were carried out on a VG Autospec high resolution mass spectrometer linked to a HP5890 Series II data system using‘3C-labeled internal standards after gel permeation/carbonchromatographic cleanup, essentially as described for livers in Chapter 1. Organochiorines andother PCBs were determined using GC/MSD (high resolution GC/low resolution MS) (Letcheret al. in press).Biochemical assaysMicrosome 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; thehomogenate was centrifuged at 10,000 g for 20 minutes, the precipitate discarded and thesupernatant further centrifuged at 100,000 g for 60 minutes. The microsomal pellet wassuspended 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 describedabove and the resulting microsomal pellet resuspended in 0.5 ml of 0.25 M sucrose. Aliquotsof 100 ul were stored in cryovials in liquid nitrogen until assayed.Cytochrome P450-related activity: Ethoxyresorufin 0-deethylase and benzyloxyresorufin0-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 forBald Eagle microsomes contained 0.1 M TRIS-HC1, pH 8.0, containing 0.1 M NaC1, 10 mMof MgC12, 2 uM 7-ethoxyresorufin or 1.5 uM 7-benzyloxyresorufin and approximately 200 gof microsomal protein in a final volume of 500 uL. After a pre-incubation period of 5 minutesat 37°C, the reaction was initiated by the addition of NADPH (final concentration 0.6 mM) tothe sample well (the blank did not receive NADPH). The reaction was stopped after 20minutes by the addition of 1.0 ml of cold methanol. The amount of resorufin formed wasmeasured in a fluorescence plate reader, using an excitation wavelength of 530 nm and anemission wavelength of 590 nm. Hepatic microsomal total protein was measured using amodification 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 polyacrylamidegels (SDS-PAGE, 9% acrylamide) and electrophoretically transferred to Rad-free membranes(Schleicher & Schuell, Keene, NH). Aroclor 1254-induced rat liver microsomes (preparedfrom commercially available postmitochondrial supernatant, Molecular Toxicology Inc.,Annapolis, MD) were used as standards. Immunodetection of CYP1A was performed usingmonoclonal antibody 1-12-3 prepared against scup cytochrome P45O1A1 which recognizesCYP1A in all taxonomic groups of vertebrates examined so far (Park et al. 1986, Stegeman1989). 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 werequantified 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 usingrabbit polyclonal antibody 7-94 against scup P450B (a CYP2B like protein), which recognizesCYP2B proteins (Stegeman 1989). Methods were as described above, but with Bio-Rad goatanti-rabbit alkaline phosphatase-linked secondary antibody and using NBT (Nitro bluetetrazolium) and BCIP (5-bromo 4-chioro 3-indoyl phosphate) for colour development. 30 g45of samples were loaded in each well. Scup microsomes containing known amounts of P450Bwere included for quantitation in each gel. Since equivalence of cross-reactivity for the antibodybetween scup and eagle is unknown, numbers are relative and not absolute. Scup standardsinsure the linearity of response of the system and are necessary for normalizing between blotsand runs. Analysis of developed blots was performed using a Kodak DCS 200 digital camerasystem and the NIH Image 1.55 densitometry software. This assay was performed in thelaboratory of Dr. J.J. Stegeman.Liver vitamin A analysis: Samples of liver (300 to 500 mg) were dehydrated to a pinkpowder by grinding with anhydrous sodium sulphate. The internal standard, retinyl acetate (40ng/20 uL methanol) was added to the equivalent of 0.20 g of liver and the vitamin Acompounds were extracted with 10 mls of a 1:9 dichloromethane:methanol solvent mixture inan amber vial. After centrifugation (10 mins at 600 rpm at 10°C) the supernatant was filteredthrough a 0.2 urn Acrodisc LC13 PVDF filter (Gelman) and a 20 ul aliquot was analyzed induplicate by non-aqueous reverse phase HPLC. Separation of retinol, retinyl acetate andretinyl 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 mobilephase to 30 % dichioromethane and 70 % methanol within 0.5 mm. This composition was helduntil the end of the run at a flow rate of 2.0 ml/min. With these conditions, retinol, retinylacetate 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 ofserum. 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 aqueousphases were separated by centrifugation, and the combined organic phases were evaporated todryness 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 wasanalyzed 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 hadretention times of 3.3 and 4.5 mm, respectively.46Hepatic porphyrins: Porphyrin levels in liver were determined using the method ofKennedy and James (Kennedy and James 1993). This method involves extraction in duplicateusing a mixture (1:1) iN hydrochloric acid/acetonitrile. The porphyrins were then concentratedon Sep-Pak Plus t C18 cartridges followed by separation and quantification by HPLC.Statistical analysisThe SYSTAT software package was used for statistical analyses of all data. Data arepresented on a lipid weight basis as suggested by Hebert and Keenleyside (1995), when thereare 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 correlatedwith percent lipid (linear regression, r2 =0.772, p < 0.0001, N = 11). Chemical residue datawere transformed to common logarithms and geometric means and 95 % confidence intervalswere calculated with the data grouped by collection site. Contaminant levels were comparedamong location with a one-way analysis of variance (ANOVA); significant differences weredetermined using Tukey’s multiple comparison procedure (MCP). Data were also comparedon the basis of a pulp mill versus non-pulp mill grouping and significant differences identifiedusing Student’s t-test. In order to avoid a bias, for comparison among sites and between pulpmill and non-pulp mill sites, only the results from the second or smallest egg were used fromthe Powell River nests, thus giving a total sample size of 14. Concentration-effect relationshipswere determined using coefficients of determination (r2) using least-squares linear regression.Unless stated otherwise, a value of p < 0.05 was considered statistically significant in allanalyses.TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalency factorsproposed by Ahlborg et al. (1994), and referred to here as the WHO (World HealthOrganization) TEFs. For comparison, TEFs proposed by Safe (1990) and Kennedy et al. (inpress) were also used.47ResultsChemical contaminant levelsPCDDs and PCDFs. Data are presented on an individual nest basis in Table 2.1. Theeight PCDDIPCDF congeners which exhibited significant differences among sites are groupedand 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 fromPowell 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 - 10 ng/kg), 1,2,4,6,8,9-HxCDF (10 - 100ng/kg) and 1,2,3,4,6,8,9-HpCDF (ND - 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 theyolk sacs from Powell River and were significantly higher than in yolk sacs from the FraserDelta or west Vancouver Island. Those same major congeners were not statistically separablebetween Powell River and east Vancouver Island. Comparison between pulp mill (PowellRiver + 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 thecongeners in Figure 2.2, except 1,2,3,4,6,7, 8-HpCDD. Although not statistically differentfrom other sites, highest concentrations of 1,2,3,4,6,7,8-HxCDD and OCDD (331 ng/kg) werein yolk sacs from the Fraser Delta.48Figure 2.2 - Residue levels of major PCDDs and PCDFs in yolk sacs of Bald Eagles collectedfrom British Columbia in 1992. Vertical bars represent geometric means of two to fiveanalyses per collection site along with the 95 % confidence interval. Means which do notshare the same lower case letter were significantly different (p < 0.05).TCDD25002,0001,5001,0005001120 12378-PCDF101., I—160 a140 : : : S12010080ab4OabI bCDD4-.:0)U)4-U)C)0)z20015010050016700 117400 123678-HCDD8,00C6OOC4,00C2,00Cab5800 1996cT1 234678HCDD01)1.,400 a300a20010:I,\C3•—‘s, 4. .*1 \•49)C•..g0’%totalPCBs-N)CCD0j’-CD0(11C)(710vr:u.76?:QD00o__CD-‘CD>____W___00..V.0___...o-eo::::c—77:::CD.••47....C)C)_c_),v7....—CD..ZçCl)O..:(•)0CDCCD_0...CDCDCD•7...cj..tjtzCD—.CDCD________________0c-a4):-•oCDS_-::.0o____________.._—.Z..CD——C•)—“‘--•....C)\.1‘.‘_____E.CD•:CD7...CDCD77.:C)—.CDCD..t__4e7......CDC.’—•——.)..CD.CD..—•.-—0-.•--.0o0-.—•_C)CDCDC’.9-05OL?.0IiTable2.1PCDDandPCDFconcentrations(nglkg,lipidweightbasis)inyolksacsofBaldEaglechickscollectedin1992fromBritishColumbia.MapLocation%fat2378123781236781237891234678OCDD2378-12478-12378-23478TotalTotalNo*.TCDDPnCDDHxCDDHxCDDHpCDDTCDFPnCDFPnCDFPnCDFHxCDFHpCDF1BrunswickPt.118681340240039.414410266143.848.334810024.22RiverRd.146461010320088.41023107623837.426.61702641693Crofton1118002510840015243057592330.047.637513059.14JackPoint1114502030405098.5256187293099.264.22673163425WinchelseaIs.1024603950938014993.283.4219062.893.282016352.06BallPt.8.811301170602014067.383.2167060.439.933912462.76BallPt.1311201130423013962.969.2188010153.62961001037EvendenPoint1426702630839023591.460.4188003243058233062787EvendenPoint18356038501090026112779.62410080.8424122028031.58ScuttleBay1827003840867017065.779.81670015.2217104015316.98ScuttleBay2224503630802018964.676.71159032.121293315326.39LimekilnBay2314702400963015035.449.2690082.011465113582.09LimekilnBay12225036701280027268.683.7797049.417895314542.510BawdenBay23218409504ND16.237.767213.522.988.917.54.0411WhitePine11330645039418.115.946.930517.623.486.221.85.8612WhitePine416353675567NDND66.546550.334.213039.8ND13ThorntonCr.206291070115019.149.3137107052.552.520714169.914MercantileCreek1632339553411.935.260.936013.819.591.047.74.39*1-2FraserDelta;3-5EastVancouverIsland;6-9PowellRiver;10-14WestVancouverIslandTable 2.2 Concentrations of non-ortho PCB congeners, geometric mean and 95%confidence interval (range in brackets), in yolk sacs of Bald Eagle chickscollected in 1992 from British Columbia.PCB congener, (Lg/kg, lipid weight basis)Location N Lipid Moisture #37 #81 #77 #126 #169% % (34-4) (345-4) (34-34) (345-34) (345-345)(mean ± SD)Fraser 2 12.6 66.9 3.23 4.79 26.9 40.0 5.60Delta ±1.4 ±0.42 0.77-13.5 1.6-14.3 23-31.5 9.17-175 3.85-8.15(2.31-4.50) (3.71-6.17) (26-27.9) (28.4-56.4) (5.14-6.12)East 3 10.7 60.1 0.63 3.00 19.4 40.9 7.63Vancouver ±0.48 ±2.24 0.18-2.17 2.55-3.47 17.8-21.2 26.1-64.3 3.61-16.1Island (0.32-1.23) (2.74-3.25) (18.5-20.3) (33.6-53.8) (5.14-11.6)Powell 8 16.0 64.6 1.23 3.56 33.4 50.0 8.90River ±4.2 ±4.9 0.61-2.6 2.45-5.18 23.1-48.2 37.3-56.6 7.14-11(0.51-5.54) (1.84-6.77) (21.3-73.8) (34.5-72.9) (6.15-13.4)West 5 17.6 63.2 1.04 3.27 38.1 36.2 5.84Vancouver ±3.9 ±2.5 0.74-1.44 1.97-5.43 24.5-59.2 20.6-63.8 3.22-10.6Island (0.32-5.54) (1.64-5.10) (23.3-58.5) (18.7-71.6) (3.24-11.8)Organochiorines. Total PCBs, DDT-related and other major organochiorines detected inBald Eagle yolk sacs are presented in Table 2.3. As with the PCB congeners, no significantdifferences in mean concentrations occurred among sites for any of the organochiorinecompounds. The pattern was relatively consistent among yolk sacs with total PCBs > DDTrelated > chiordane-related > dieldrin > B-HCH > HCB > mirex. The exception to thispattern was the yolk sac from White Pine Cove No. 1, where DDE levels were greater thantotal PCBs. The PCB/DDE ratio was generally much lower in yolk sacs from the west coast ofVancouver Island than from other sites.Artificial hatching success and condition of embryosA total of 25 Bald Eagle eggs were collected for incubation, of which one was cool tothe touch at the time of collection (there was a recently hatched chick in the nest) while a52Table2.3Organochiorineconcentrations,geometricmeansand95%confidenceintervals,(range inbrackets)inyolksacsof BaldEaglechickscollectedin1992fromBritishColumbia.LocationOrganoclorinecentration(mg/kg,lipidweightbasis)NTotalDDEirans-oxychlordaneheptachior-DieldrinMirexB-HCHHCBPCBsnonachiorepoxideFraserDelta236473.53.820.911.692.300.450.680.62113-116020.3-2671.54-9.450.12-7.011.25-2.300.53-9.960.10-2.00.22-2.090.35-1.10(278-477)(54.5-99.2)(3.09-4.71)(0.56-1.46)(1.58-1.82)(1.64-3.23)(0.32-0.64)(0.52-0.88)(0.55-0.71)EastVancouver35591279.461. egg was possibly shaken as it was lowered from the nest. Therefore, 23 of the eggswere possibly viable when placed into the incubator. A total of 18 eggs hatched for an overallsuccess rate for artificial incubation of 78.3 %. Eliminating the possibly shaken egg fromNorthwest Bay, 16 eggs were collected from pulp mill sites of which 11 hatched for a hatchingrate of 69 % (Table 2.4). Of eight eggs collected from non-pulp mill sites, seven hatched for ahatching rate of 88 %. This difference in hatching success between pulp mill and non-pulp millsites was not, however, significant (Chi2 test). One chick (Ball Point A) was edematous athatching. Of the eggs which failed to hatch, one was infertile (Powell River area), two wereaddled (both from the same nest in the Powell River area), three were early (first quarter ofdevelopment) embryos (one each from east Vancouver Island, Powell River and westVancouver 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 BritishColumbia, 1992.Location Treatment No. No. %collected hatched successFraser Delta Non-pulp mill 2 2 100West Vancouver Island Non-pulp mill 6 5 83(Mean, non-pulp mill) 8 7 88East Vancouver Island Pulp mill 3b 3 100Powell River Pulp mill 12 8 67Alberni Inlet Pulp mill 1 0 0(Mean, pulp mill) 16 11 69a pulp mill versus non-pulp mill difference not significant, chi2 = 1.402b 4 eggs were collected, 1 was eliminated as possibly shaken54Morphological and histological measurementsNo significant differences occurred among sites for mean values of any of the measuredmorphological parameters, whether expressed as actual values or as percent yolk-free bodyweight. For the 18 chicks measured morphological measurements (mean ± SD) were asfollows: body weight (88 ± 9.4 g), yolk-free body weight (83 ± 8.3 g), liver (1.9 ± 0.29 g), rightkidney (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), thyroidglands (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 arecompared among sites in Appendix 2.1.For the tissues examined histologically, variations among individual birds were seenonly for lymphoid organs (Table 2.5). Variations were observed within and between sites inamount of lymphoid tissue, the number of cells in mitosis, the number of necrotic cells and thedegree of extramedullary hematopoiesis. However, no significant differences among sitesoccurred for mean values of any of the measured parameters. The amount of lymphoid tissuein the spleen was constant among individual birds.Table 2.5 Histological examination of immune system tissues in Bald Eagle chicks (Mean± SD).Fraser Delta East Van. Isl. Powell River West Van. Isi.(N=2) (N=3) (N=8) (N=5)Bursa Amount of lymphoid tissuea 3.0 ± 0.0 3.0 ± 0.0 3.0 ± 1.1 1.8 ± 0.84No. necrotic cells” 90 ± 28 109 ± 28 142 ± 85 105 ± 30No. cells in mitosisb 29 ± 11 43 ± 5.5 50 ± 3.1 40 ± 14Spleen No. cells in mitosisb 15 ± 9.2 16 ± 7.5 19 ± 6.7 6 ± 3.9Degree of E.M.C 1.5 ± 0.71 1.3 ± 0.58 2.2 ± 0.64 1.4 ± 0.55Thymus Amount of lymphoid tissue’ 3.0 ± 0 3.7 ± 0.58 2.9 ± 3.8 2.0 ± 0.71No. necrotic cells’ 28 ± 3.5 54 ± 21 66 ± 19 64 ± 18No. cells in mitosisb 10 ± 8.5 24 ± 22 11 ± 5.4 21 ± 8a- based on follicular size and cell density of cortex and medulla. The amount varied from small (1) to large (4).b- per 5 fields at 600x.- e.m. - extramedullary hematopoiesis, based on the amount of hematopoietic tissue. Amount varied from small (1) to large (3).d- based on the thickness of the cortex and cell density. The amount varied from small (1) to large (4).55Biochemical measurementsMean concentrations of CYP1A were sixfold greater (p <0.05) in chicks from PowellRiver compared to west Vancouver Island (Table 2.6). Mean concentrations of a CYP2B-likeprotein were two to three-fold higher in livers from Strait of Georgia sites compared to westVancouver Island; however, the differences were not significant. Mean EROD activity waseight-fold higher in east Vancouver Island compared to Fraser delta and mean BROD activitywas nearly nine-fold higher in Powell River than Fraser delta chicks; however, the differenceswere 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 nearpulp 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 deltacompared to west Vancouver Island.Table 2.6 Measurement of hepatic cytochrome P450 and porphyrin parameters and vitaminA in plasma and liver of Bald Eagle chicks collected in 1992 from BritishColumbia (Mean ± SD).Fraser Delta East Vancouver Powell River West Vancouver(N = 2) Is. (N = 3) (N = 8) Is. (N = 5)CYP1A (std. vol. equiv. [id]) NA 15a,b (± 35) 25 (± 12) 44b (± 2.3)CYP2B equivalents (pmol/mg) 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)Retinol-liver (gIg) 0.65 (± 0.07) 0.60 (± 0.15) 0.65 (0.13) 0.67 (0.12)Retinyl palmitate-liver (gIg) 19 (± 6.9) 28 (± 7.3) 29 (± 8.4) 37 (± 13)a,b- means that do not share the same superscript are significantly different among sites.NA - not assayed.56Concentration-effect relationshipsData from the complete set of 18 Bald Eagle chicks were used to examine relationshipsbetween measured biological parameters and contaminant exposure. The gradient of exposurefrom 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 residuedata; results are presented in Table 2.7 for each parameter based on which form of the residuedata gave the best fit (greatest r2 value) to the regression curve.Highly significant positive regressions were found between hepatic CYP1A and most ofthe individual PCDD, PCDF and PCB compounds in yolk sacs; however, the best fits werewith log 2,3,7,8-TCDF and 2,3,7,8-TCDD (Table 2.7, Figure 2.4). No significant regressionswere found between a CYP2B-like protein and yolk sac concentrations of any of the chemicalparameters measured. For EROD, the best r2 value was with 2,3,7,8-TCDD, while thestrongest regression for BROD was found with log 2,3,7,8-TCDF. Hepatic urophorphyrin alsoshowed 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-PCB126, but not with any other chemical parameters. The hepatic cytochrome P450 and porphyrinparameters all regressed more strongly with either 2,3,7, 8-TCDD or log 2,3,7, 8-TCDF thanwith TCDD-TEQ5 estimated using three different TEFs (Table 2.8).Among the morphological parameters, a weakly significant positive regression wasfound between yolk-free body weight and log PCB 126. Yolk sac weight negatively regressedwith both total PCBs and log TEQs. A weakly significant positive regression was alsodetermined for density of thymic lymphoid tissue with log 2,3,7,8-TCDD (r2 = 0.320, p <0.02) and log TEQ5WHO (Table 2.7).57(A)EE0200uJ(B)0 500 1,000 1 .500 2,000 2,500 3,000 3,5002378-TCDD (nglkg, lipid basis)()10)EC2020I I I I2,3,7,8-TCDF (nglkg, lipid basis)Figure 2.4 - Exposure-response relationships between 2378-TCDD or log 2378-TCDFconcentrations in yolk sacs of Bald Eagles and hepatic (A) EROD activity (B)CYP1A concentrations and (C) BROD activity.r2 = 0.748 *500401,000 1 500 2,000 2,5002378-TCDD (nglkg, lipid basis)3.000 3,500r2 = 0.721*1,000 10,00058Table2.7Concentration-effectrelationshipsbetweenbiochemicalandmorphological measurementwithchlorinatedhydrocarbonlevelsinyolksacsof BaldEaglechicks.1TEQs,accordingtoAhlborg(1994)NS-notsignificantci,Parameter2,3,7,8-TCDDLog2,3,7,8-TCDFLogPCB126TotalPCBsLogTEQswHO1NSloper2pr2pr2pr2pr2pCYP1A14(+)0.850<0.00010.887<0.00010.371<0.030.576<0.0020.728<0.0005CYP2B14(+)0.082NS0.114NS0.057NS0.255NS0.136NSEROD18(+)0.748<0.00050.708<0.00050.297<0.020.588<0.00010.633<0.0005BROD13(+)0.601<0.0020.721<0.00050.396<0.030.346<0.0060.549<0.004Uroporphyrins17(+)0.316<0.020.298<0.020.194NS0.122NS0.232<0.05Retinyl-18(+)0.059NS0.028NS0.26<0.030.097NS0.202NSpalmitateliverYolk-freebody18(+)0.032NS0.042NS0.247<0.040.023NS0.072NSweightYolksac18(-)0.034NS0.045NS0.098NS0.269<0.030.128NSThymic18(+)0.191NS0.091NS0.058NS0.09NS0.250<0.04lymphoidtissueTable 2.8 Comparison of regression (r2) values of some hepatic biochemical parameters onTEQs derived from three sets of toxic equivalence factors (TEFs).Toxic Equivalent FactorsParameter TCDD/F’ Safe2 CEH3 WHO4P450 1A 0.887 0.687 0.759 0.805EROD 0.748 0.529 0.607 0.633BROD 0.601 0.427 0.515 0.549Uroporphyrin 0.316 0.107 0.162 0.232‘Best r2 value (either 2,3,7,8-TCDD or 2,3,7,8-TCDF)2Safe (1990)3Chick embryo hepatocyte (S. Kennedy, person. comm.)4Ahlborg et at. 1994DiscussionBald Eagle chicks collected from nests near pulp mills were exposed to elevatedconcentrations of potent embryotoxic PCDD and PCDF congeners, compared to chicks fromreference nests. Symptoms of TCDD-like exposure, such as have been observed in fieldstudies 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 sitesand 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 femaleparent 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, 126and 77, although regressions with biochemical and morphological parameters were weak andinconsistent compared to TCDD and TCDF. Concentrations of total PCBs and otherorganochiorines in eagle yolk sacs also varied little among sites.60Laboratory hatching successExcept for one edematous chick, no signs were apparent in either the hatched eaglets orin failed eggs of GLEMEDS (Great Lakes embryo mortality, edema, and deformitiessyndrome) (Gilbertson et at. 1991), such as reported for fish-eating birds in the Great Lakesand elsewhere (Hoffman et at. 1986; 1987; Kubiak et at. 1989; Bosveld et a!. 1994; Van denBerg 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 toxicsyndrome caused by TCDD in chicken embryos. In embryos of other avian species, such asring-necked pheasants (Phasianus cotchicus), mortality is the most sensitive response to TCDDexposure 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 eagleeggs among sites or between pulp mill and non-pulp mill areas. The overall artificial hatchingsuccess of 78.3 % was comparable to the average of 75 % (range 62 - 87 %) reported for wildand captive Bald Eagles from a number of studies (Stalmaster 1987). The absence ofdeformities and other GLEMEDS symptoms in Bald Eagle chicks from this study is likely dose-related; 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 totalPCB levels than any of the fresh eggs from the Strait of Georgia.Patterns and trends of PCDD, PCDF and PCB contaminants in yotk sacsLocal pulp mill and chiorophenol inputs account for the particular pattern and elevatedlevels of 2,3,7,8-substituted PCDDs and PCDFs in Bald Eagles and other wildlife from theStrait of Georgia (Elliott et at. 1989a; Whitehead et at. 1990; 1992b), compared to similarsamples from other North American and European sites (Van den Berg et at. 1994b; Yamashitaet at. 1993; Hebert et at. 1994). In particular, Bald Eagle yolk sacs contained highconcentrations of 2,3,7, 8-TCDF, which is reported elsewhere at only nominal levels in wildlifesamples. High TCDF levels such as in the eagle yolk sacs from Powell River reflect exposureto prey items contaminated by local pulp mill discharges (Harding and Pomeroy 1990).61Elevated TCDF levels have also been reported in tissues of common mergansers (Mergusmerganser) and herring gulls breeding near a bleach kraft pulp mill in Quebec (Champoux1993). 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 previouslyfor herons (Elliott et al. 1 989a). Accumulation of TCDF in eagle tissues is probably not linkedto the low absolute EROD activity found in Bald Eagle chicks (Table 6); a recent studycompared EROD induction with in vitro capability to metabolize PCB 77, and concluded thatlow 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 thepresence of non-2,3,7,8 substituted PCDDs and PCDFs in eagle yolk sacs. Fish are able tometabolize most compounds of this type (Sjim et al. 1989), leading to low levels in the diet offish-eating species; birds are also likely capable of further metabolizing them. The presence ofelevated levels of 1,2,3,4,6,7,8-HpCDD and OCDD in the yolk sac from River Road in theFraser River delta is consistent with reports of high concentrations of those contaminants insediments from near the nest site (Tuominen and Sekela 1992). Elevated levels of higherchlorinated dioxins in Fraser estuarine sediments are indicative of the intensive past use ofchlorophenol 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 uniformityamong sites in concentrations of PCBs and other organochiorines in eagle yolk sacs reflects theimportance of diffuse atmospheric inputs for those compounds (Elliott et al. 1989b). Thegeographically uniform PCB congener pattern contrasts with the finding of significantdifferences in the percent contribution of certain congeners in great blue herons betweenCrofton and Vancouver in 1987 (Elliott et al. 1989a). Because of their restricted seasonalmovement and diet, herons appear to be better indicators of local PCB contamination thaneagles.62Biochemical responsesThe results of this study confirm for another avian wildlife species the value of CYP1Ainduction, particularly as measured by western blotting, as a sensitive marker of exposure toTCDD-like compounds. Absolute EROD activities in these embryonic Bald Eagle microsomeswere 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 andherons (Sanderson et at. l994a; Whitehead et al. 1992b). Interspecific variation of this type isnot surprising as there is increasing evidence that cytochrome P450 isoforms vary substantiallyeven among closely related species (Yamashita et at. 1992).Absolute BROD activity was about five-fold higher than EROD in livers of Bald Eaglechicks, while differences in rates from least to most contaminated individuals was similar forthe two activities. As with EROD, the best r2 values were found between BROD and 2,3,7,8-TCDF or 2,3,7,8-TCDD. BROD is considered a relatively specific marker of CYP2B1 activityin phenobarbital-induced rats (Burke et at. 1994). However, Rattner et al. (1993) recentlyreported that, while phenobarbital treatment of black-crowned night-heron embryos caused a2,000-fold increase in a CYP2B-like protein, there was only a threefold increase in BRODactivity. 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-reactivewith putative fish CYP2B and rat CYP2B are present in at least some groups of birds, but thesubstrate specificities may be quite different. The results suggest the presence of a CYP2Bisoform in Bald Eagles. Although Bosveld and Van den Berg (1994) in a recent reviewconcluded that there is no evidence of chlorinated hydrocarbon-inducible CYP2B isoforms inbirds which cross-react with mammalian CYP2B antibodies, further experiments using purifiedCYP enzymes and antibodies are required for a better understanding of substrate specificities.Uroporphyrin levels in chicks from the various sites were similar. Although there wasa significant concentration-effect relationship between uroporphyrin levels and both 2,3,7,8-TCDD and -TCDF, this finding must be treated cautiously as normal uroporphyrin levels inavian livers range from 5-25 pmol/g (Fox et al. 1988). PCBs have been reported to cause63accumulation of porphyrins in chick embryo hepatic cell cultures (Kennedy et al. 1995) and inliver and other tissues of adult birds of common laboratory species (Elliott et al. 1990), but notapparently in captive predatory birds (Elliott et al. 1991). In previous field studies, hepaticporphyrins 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 retinylpalmitate did not differ among sites, although a weakly significant positive relationship betweenhepatic retinyl-palmitate and PCB 126 (34-345) was found. In contrast, laboratory data for ratsreport 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 variedamong colonies and the molar ratio of retinol to retinyl palmitate correlated positively withTEQs in eggs (Spear et al. 1990). Van den Berg et al also reported a non-significant reductionin hepatic retinyl palmitate in cormorants from a contaminated relative to a reference site in theNetherlands.Morphological and histological parametersMorphological and histological measurements did not differ among sites. However, aswith retinyl-palmitate and PCB 126, a number of weakly significant exposure-responserelationships occurred, at the p < 0.05 level (Table 2.8), which are likely not meaningfulbiologically. For example, yolk-free body weight appeared to increase with PCB 126 levels inyolk sacs; this contrasts to data from a number of field studies which report statisticallysignificant negative relationships between PCDDs/PCDFs or PCBs and embryonic weight andother 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 yolkweight is consistent with similar findings for cormorants from British Columbia (Sanderson etal. 1994b), but contrasts with reports of a positive relationship for cormorants from theNetherlands (Van den Berg et al. l994b). The positive relationship between density of thymic64lymphoid tissue and log 2,3,7,8-TCDD is in contrast to reports from a number of laboratorystudies that TCDD and related compounds cause atrophy of the thymus with depletion oflymphocytes (Elliott et al. 1990; Nikolaides et al. 1988). Therefore, it is likely that thesefindings in Bald Eagles are spurious in nature due in part to the relatively small sample size andlarge number of variables analyzed.Comparison of toxic equivalentsAs reported previously for great blue heron chicks in the Strait of Georgia (Bellward etal. 1990; Sanderson et al. 1994a), regression of the biochemical endpoints against 2,3,7,8-TCDD or 2,3,7, 8-TCDF produced the best coefficients of determination (r2). This contrasts todata 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, exposureto PCDDs and PCDFs relative to PCBs was low in all of those studies, whereas the reversewas 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. Therelative contribution of the major Ah-receptor active congeners in yolk sacs of Bald Eagles iscompared among sites and to common terms from the Netherlands in Figure 2.5. TotalTEQ5WHO and the pattern of contributors was similar between Bald Eagles from west VancouverIsland and the common terms; however, PCDDs and PCDFs made a much greater contributionin 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 andPCDFs suggests that TEFs derived from mammals such as Safe’s (1990) TEF’s tend tooverestimate the toxicity of the both the mono-ortho and non-ortho PCBs, in all avian specieswith the possible exception of the chicken ((Brunstrom 1990; Brunstrom and Anderson 1988;6514000xCl)Cl)(U0Cl)CwI-Figure 2.5 - The contribution of various chlorinated hydrocarbon groups to the sumof TCDD toxic equivalents (TEQ) in Bald Eagle yolksacs from coastal BritishColumbia, 1992 (N values and variances are in the tables), compared to values forcommon terns from the Netherlands. Toxic equivalents factors for PCDDs/PCDFsfrom 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 ofeither TCDD or TCDF and TEQs using the different TEFs. The WHO-TEFs, which givelower weighting to the mono-ortho PCBs, produced r2 values which were closest to thosedetermined 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 asensitive biomarker for other deleterious Ah-receptor mediated responses (Hoffman et al. 1987;121086420Li mono-o-PCBs1--.LJ non-o-PCBsPCDFsEl other-PCDDs1•TCDD.Bald Eagle\ \I— Common Tern66Bosveld 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 levelsof 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 resultto a whole egg, wet weight basis, (dividing by a mean factor of 60, based on comparison forBald Eagles of a yolk sac and whole egg analyzed from the same nest), mean TEQ5PCDD,PCDFSwere 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 asuggested 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 PowellRiver, where mean TEQ5WHO in yolk sacs, on a lipid weight basis, were about 12,600 ng/kg, orabout 210 ng/kg, on a wet weight basis in the whole egg. This is suggested as a lowestobserved-effect-level (LOEL).In conclusion, Bald Eagle chicks collected near pulp mills were exposed to elevatedconcentrations of PCDDs and PCDFs which correlated with induction of a hepatic CYP1Across-reactive protein. Levels of PCBs and other organochiorines did not vary among sites andwere less important in the CYP1A induction.67AcknowledgementsMany people contributed their time to the success of this project. I would especiallythank I. Moul and G. Compton for assistance in the field. C. Kuehier suggested the incubationconditions. 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 ofVeterinary 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 arethanked for their technical assistance with the biochemistry, which was performed in thelaboratory of S. Trudeau (NWRC). A. Lorenzen performed the CYP1A assay. B. Woodinperformed the CYP2B assay. J. Smith provided advice on the statistics. S. Bucknell typed thetables and P. Whitehead assisted with drafting figures. The research was supported by theCanadian Wildlife Service and th Wildlife Toxicology Fund of Environment Canada and by theNational Science and Engineering Research Council of Canada.68Appendix 2.1 Selected morphological measurements in Bald Eagle chicks collected in1992 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 + 10.6Relative liver weight 2.3 + 0.19 2.3 + 0.34 2.4 + 0.34 2.3 + 0.32(as % body weight)Tarsus length (mm) 19.0 + 0.53 20.6 + 1.41 20.6 + 1.16 20.7 + 2.26Tibia length (mm) 26.5 + 0.53 27.5 + 1.16 26.5 + 0.69 26.8 + 1.09NOTE: 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 orwing lengths.69CHAPTER 3BIOACCUMULATION OF CHLORINATED HYDROCARBONS ANDMERCURY IN EGGS AND PREY OF BALD EAGLESThe purpose of the bioaccumulation study was to measure chlorinated hydrocarbonlevels, particularly for PCDDs and PCDFs, in eagle eggs in order to determine spatial andtemporal patterns and trends, and to relate the levels to critical concentrations in their foodusing a simple model. At issue was the determination of site specific concentrations of PCDDsand PCDFs in representative sentinel food items, such as forage fish andfish-eating birds, thatwould not adversely affect Bald Eagles. The development of guidelines for chlorinatedhydrocarbon levels in dietary items of eagles should have broader applicability in other NorthAmerican jurisdictions.Materials and MethodsSample collectionFrom 1990 to 1992, a total of 32 Bald Eagle eggs were collected at six sites on thesouth coast of British Columbia (Figure 3.1). Four treatment areas were selected based onproximity of eagle breeding sites to industrial pollutant sources. The lower Fraser valley nearVancouver is a heavily urbanized and industrialized area that receives wastes from numerouslocal and upstream pulp, paper and lumber mills and wood treatment operations. The Croftonand Powell River areas each receive effluent inputs from local kraft pulp mills. Nanaimo is anurbanized area with a large kraft mill and other wood milling and yarding operations. Themain reference site was an area of northern Johnstone Strait, with little industrial activity otherthan lumber yarding. Three single eggs were also obtained from 1) Clayoquot Sound on thewest 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 inthe Queen Charlotte’s archipelago, remote from any industrial activity.70•1 C C.) I I ISuitable nests were located by ground, boat and aerial surveys, during which nests werescored numerically to estimate access, suitability for climbing and land tenure. In 1990 and1991, in an effort to obtain fresh eggs, collections were made during the first two weeks ofApril in the lower Fraser valley and the Strait of Georgia, during the first week of May on thewest coast of Vancouver Island and during the third week in May in Johnstone Strait. Normallya single fresh egg was collected and only from nests with at least two eggs, except at StillwaterBay in 1992, when both eggs were taken. The two eggs collected in 1994 were addled, theywere retrieved from nests in June or July during blood sampling of nestlings. To encouragecontinued incubation of the remaining egg and thus to minimize the impact of collection, timenear the nest and in the nest tree was minimized. Eggs collected in 1990 and 1991 wererefrigerated 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 in1992 were initially incubated as part of another study (Chapter 2); the failed eggs from thisstudy 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 analysesWhole eggs were homogenized and prepared for analysis at NWRC. Aliquots fororganochlorine pesticides and PCBs were analyzed according to methods described in Norstromet al. (1988) and outlined in Chapter 1, except that total PCB levels are reported as the sum of28 congener peaks (24 listed in Figure 3.2, plus trace amounts of PCBs 137, 195, 200 and206). Eggs collected in 1990 and 1991 were analyzed for PCDDs/PCDFs by low resolutionGC/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 nonortho PCB analyses of eggs collected in 1992 were carried out on a VG Autospec highresolution mass spectrometer linked to a HP 5890 Series II data system according to methodsdescribed by Letcher et al. (in press), also as outlined in Chapter 1. Mercury was analyzed atthe NWRC by cold vapour atomic absorption according to methods described by Scheuhammer72& Bond (1991), and methyl mercury was extracted as described in Callum and Ferguson(1981).Eggshell thickness measurementEggshells were air-dried in the laboratory for two weeks or more. Using a ballmicrometer, shell thickness was measured at the equator of the shell, including the membrane;five readings were made and averaged.Statistical treatmentFor each location, data were combined for all years in order to give a larger samplesize. Chemical residue data were transformed to common logarithms and geometric means and95 % confidence intervals determined. Data were also converted to common logarithms andSAS routines used to perform a one-way analysis of variance followed by Tukey’s multiplecomparison procedure (MCP) to determine significant differences in mean residue levels amongsites. For determination of statistical differences among sites for percent PCB congeners, anarcsine transformation was used, followed by ANOVA and Tukey’s MCP. Unless otherwiseindicated, a significance level of p <0.05 was applied to all statistical tests. Patterns of allchlorinated hydrocarbons and other the major PCB congeners as percent total PCBs wereanalyzed using principle components analysis (PCA) in SAS. As for the other statisticalanalyses, residue concentrations were transformed to common logarithms, while for the percentPCB congener contributions, an arcsine transformation was used.Toxic equivalents (TEQs) were estimated using standard toxic equivalent factors forPCDDs and PCDFs as suggested in Safe (1990), except that for the mono-ortho and non-orthoPCBs, the World Health Organization toxic equivalents (WHO-TEFs, Ahlborg et al. 1994).Bioaccumulation modelIn order to relate PCDD and PCDF levels in Bald Eagle eggs to their diet, a simplebioaccumulation model was used (modified after U.S. Environmental Protection Agency 1993and US Fish and Wildlife Service, 1994). The model assumes: 1) breeding Bald Eagles areyear-round residents and, therefore, acquire most of their contaminant burden from local73sources 2) levels in eagle eggs are in equilibrium with those in the female’s diet. The modelhas the form:BEE = BMF [F1(X) + F2(X) ... + FN(XN)]BEE = Contaminant concentration in Bald Eagle eggBMF = Biomagnification factor for a given contaminantF1 = Fraction of item one in dietX1 = Contaminant concentration in item oneFN = Fraction of the Nth itemXN = Contaminant concentration in the Nth itemAs input, we used data on PCDD and PCDF levels in avian and fish prey from nearpulp mills and at reference sites on the British Columbia coast, summarized in Tables 3.1 and3.2. Estimates of Bald Eagle diet composition were taken from Knight et al. (1990), Vermeeret al. (1989) and Watson et al. (1991). The eagle diet was divided into components, whichvaried 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-eatingwaterfowl) 3) omnivorous gulls 4) non-salmonid fish 5) salmonid fish. Biomagnification factorsdetermined 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), 2,3,7,8-TCDF (1.4, estimated), 2,3,4,7 ,8-PnCDF (4.5). The biomagnification factors for herring gullswere similar to those estimated for great blue herons to forage fish at Crofton, 25 and 10respectively for 2,3,7,8-TCDD and 1 ,2,3,6,7,8-HxCDD (Elliott et al. 1989a). Where only eggor liver data was available for a species, the inter-tissue ratios in Braune and Norstrom (1989)were used to convert to whole body concentrations.74Table3.1MeanPCDD/PCDFlevels(ng/kg, wetweight)infishcollectednearthreepulpmillsontheStraitofGeorgia, BritishColumbia.AreaSpeciesN*TissueCollection%2,3,7,812378TotalTotal2,3,7,823467DataperiodlipidTCDDPnCDDHxCDDHpCDDTCDFPnCDFSourceNanaiinoEnglishSole4/4FilletJan-Feb.19902.<1.013<0.051NanaimoEnglishSole4/4LiverJan-Feb.19908.52.66.6441.6581.71NanaimoChinookSalmonFilletJan-Feb.19906.5<1.0<1.04.2<1.047<1.02CroftonEnglishSole2/2FilletJan-Feb.19902.<0.051CroftonEnglishSole2/2LiverJan-Feb.1990105.08.064.<3.0892.01CroftonArrowtoothFlounder1/1FilletJan-Feb.19903.4<0.5<1.5<1.0<1.57<0.51CroftonArrowtoothFlounderLiverJan-Feb.19903.63.5<2.59.0<1.563<1.01CroftonRockFishFilletJan-Feb.19901.81.7<1.09.1<1.017<1.02Powell RiverEnglishSole3/3FilletJan-Feb.19901.4<0.5<1.01.5<1.015<0.51PowellRiverEnglishSole3/3LiverJan-Feb.19905.73.08.0987.41379.01Powell RiverChinookSalmonFilletJan-Feb.19904.,19902-EnvironmentCanada,*-NumberanalyzedINumbercollectedLiiTable 3.2 PCDD and PCDF levels (ng/kg, wet weight) in waterbird and seabird speciesfrom the British Columbia coast.Species Location Year Tissue N % Fat 2378- 12378- 123678- OCDD 2378- 23478- RefTCDD PnCDD HxDDD TCDF PnCDFFish eating birdsWestern Grebe Port Alberni 1989 liver1992 bm*Nanaimo 1992 bmPowell River 1992 bmAlert Bay 1992 bmDouble-crested Fraser Delta 1990 eggsConuorantFraser Delta 1992 eggsCrofton 1990 eggsCrofton 1992 eggsNanaimo 1992 eggsGreat Blue Fraser Delta 1990 eggsHeronFraser Delta 1992 eggsCrofton 1990 eggsCrofton 1992 eggsRhinoceros Johnstone 1990 eggsAuldet Strait10 4.7 87 4.2 268 4.8 810 5.6 108 6.4 4210 5.6 105 6.3 1025 6.1 1912 18 1<15 217 38 1<22 69 <2.4 2<8.3 41 <1.4 2<9.5 230 9.9 2<18 <0.9 <2.4 2ND ND 11 22332Invertebrate feedersBufflehead Fraser Delta 1990 bmCrofton 1990 bmAlert Bay 1992 bmSurf Scoter Fraser Delta 1989 bmPort Alberni 1989 liverCrofton 1990 bmNanaimo 1992 bmPowell River 1992 bmAlert Bay 1992 bmGlaucous- Nanaimo 1992 eggswinged Gull1 3.4 182 2.1 9.411 3.7 <110 4.1 <1.15 3.1 2410 2.8 5.16 2.5 0.58 3.1 <1.67 1.6 <210 8.8 <1.0<1 17<2 29<3 <3<1.9 <3.122 30<0.5 3.30.7 2<3.1 <4.2<6.5 <6<1.0 22<7 10 <1 1<10 28 3.5 1<10 1.4 <1 2<11 13 <1.5 111 123 13 13 40 1 17.3 7.4 1.3 2<8.2 12 <1.8 2<22 3.2 <2.8 2<1.0 <1.0 1.0 25 7.3 117 385 2498 3.8 25 66 643 4.6 2.9 <2.9 8.65 7.1 4.4 <2.6 212 3.1 <1.6 <3.2 <3.56 4.5 25 42 6414 11 ND ND ND 242 65 ND 2 11 225 47 ND 1 7 222 38 ND ND ND 245 57 5 15 16 210 11 ND 4 1223 229 3 10 3240 47 1 5 73 3 1 27 2*bm- breast muscleND - not detected (detection limit = 0.5 ng/kg)References: 1 - Whiteheact et al. 1990; 2 - Elliott et al., 1995b; 3 - Whitehead et al. 1992.76ResultsPCDDs and PCDFsMajor 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 othertwo compounds (Table 3.3). All eggs contained detectable levels of the three major PCDDcongeners. Lesser concentrations of 1,2,3,4,6,7,8-HpCDD and OCDD were found in mosteggs. The only PCDFs consistently detected in Bald Eagle eggs were 2,3,7,8-TCDF and2,3,4,7,8-PnCDF. Eggs from Johnstone Strait contained significantly less 2,3,7,8-TCDD thandid eggs from other sites. Concentrations of 1,2,3,7,8-PnCDD were significantly higher ineggs from Crofton than either the lower Fraser valley or Johnstone Strait. Concentrations of1,2,3,6,7,8-HxCDD and 2,3,4,7,8-PnCDF were significantly lower in Fraser valley eggs thanfrom the pulp mill sites, but did not differ significantly from Johnstone Strait.OrganochiorinesQuantifiable residues of DDE, DDD, trans-nonachlor, cis-nonachlor oxychlordane, cischiordane, heptachlor epoxide, dieldrin, mirex, fi-HCH and HCB were found in all eggsanalyzed (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 wereabout 50 % of mirex concentrations. Organochiorine levels were generally highest in eggs fromPowell River, although differences were significant in only one case: trans-nonachlor wassignificantly higher at Powell River and Nanaimo than the Fraser valley. Mean concentrationsof cis-chlordane were significantly higher in eggs from Johnstone Strait than other sites exceptPowell River and were also significantly lower at Crofton than all other sites.MercuryHighest mean concentrations of total mercury were in eggs from Johnston Strait and theFraser valley and were significantly higher than those from Nanaimo and Crofton, but notPowell River (Table 3.5). Methyl-mercury was also determined in the eleven eggs from 1990and constituted an average of 88 % (SD = 11, range 73 - 100%) of the total mercury present inthose Bald Eagle eggs.77Table3.3Polycliloiinateddibenzodioxin(PCDD)andpolychiorinateddibenzofuran(PCDF)residuelevels(wetweight basis)inBaldEagleeggsfromBritishColumbia,1990-1992.PCDDandPCDFLevels(ng/kg)(geometricmeanand95%confidenceinterval)Nest(MapNo.*)Year%liyid%moisture237812378123678123789237823478(mean±SD)TCDDPnCDDHxCDDHxCDDTCDFPnCDFLowerFraserValleyBrunswickPt.(1)19905.182.542374242313AnnacisIs.(2)19905.982.75855112511212ChahalisFlats(3)19906.183.158525528914Island20(4)19915.382.351717ND16NDCheamIsland(5)19914.982.023615ND732AgassizBridge(6)199141151811310Mean5.682.44422023321.33925a±0.6±0.5(30-63)(7-57)(14-76)(0.5-3.8)(14-105)(1.3-22)NanaimoCanoxy(7)19903.085.75910919852916Leask(8)19904.783.0639925093618Canso(9)19904.284.182116346122931JackPt.(10)19904.584.27913322764920NorthwestBay(11)19924.981.21422371185MaudeIsland(12)19914.583.270104173211935SoutheyIsland(13)19914.283.328297956512JackPoint(10)19924.722Mean4.483.445266134”341432±0.6±1.4(26-78)(35-122)(68-264)(1.5-7.6)(26-70)(8-27)CroftonR.Pringle(14)19904.482.7104211374101627Southey(15)19905.980.011014931072634Crofton(16)19916.082.85110817316022Mean5.481.8842150”270L3229227”±0.9±1.6(29-243)(65-346)(99-742)(0.1-6)(6-154)(16-47)Table3.3cont...PCDDandPCDFLevels(ng/kg)(geometricmeanand95%confidenceinterval)Nest(MapNo.*)Year%liyid%moisture237812378123678123789237823478(mean±SD)TCDDPnCDDHxCDDHxCDDTCDFPnCDFPowellRiverKellyPt.(17)19906.180.09812924475927Convent(18)19905.782.488128372159737Lund(19)19915.082.14159186711024Powell River(20)19923.781.932471162185Stiliwater(A)(21)19925.879.678104143316648Stiliwater(B)(21)19926.779.381106146316850GrisePoint (22)19926.179.61018801Mean5.680.749ä71b,c170b48527b±1.0±1.4(23-105)(37-138)(103-261)(2-9)(49-147)(15-49)JohnstoneStraitPlumper5(23)19914.882.822397255810Plumper8(24)19914.383.312281673397Pearce3(23)19916.080.632641115803Pearce5(26)19914.483.31143732375HarbiedonIsland(27)19912.885.4153379ND6810SwansonIsland(28)19914.185.41026521295OwlIsland(29)19913.983.31125431Mean4.384.1iSh3578’1.3947±1.0±4.1(10-22)(26-48)(51-118)(0.5-3.4)(33-66)(6-11)PocahontasPt(30)19922.488.1175438174BerrymanPt(31)19925.981.410167ND74LangaraIs.**19946.580.1253ND52MapNos.30-Albemi Inlet,31-ClayoquotSound.**-Queen CharlotteIslandsab,c-meansthatdonotsharethesameletteraresignificiantlydifferent(p0.05)Table3.4OrganochiorineandPCBresiduelevels(mg/kg,wetweight)inBaldEagleeggsfromtheBritishColumbiacoast,1990-1992,expressedasgeometricmeansand95%confidenceintervals(range inbrackets).LocationNTotalPCBsDDEDDDtrans-cis-oxy-dieldrinmirexbeta-HCBnonachiornonachiorchiordane11CRLower6268ab2j7a0.0580.142k0.030.037a0.037a0.009o.oosa0.025aFraser1.49-4.81.07-4.410.041-0.0830.098-.2040.020-0.0440.023-0.0590.019-0.0730.002-0.0450.001-0.0240.017-0.039Valley(1.08-6.21)(.90-4.14)(0.030-0.075)(0.082-0.234)(0.018-0.049)(0.022-0.082)(0.020-0.091)(0.001-0.038)(0.001-0.032)(0.016-0.032)Nanaimo84ab3.13a0.o520245ab004.7ab0.042a0.042ao.olsa0.022a0.0123.28-6.791.65-5.920.035-0.0790.186-0.3230.036-0.060(0.029-0.062)0.029-0.0610.005-0.0470.011-0.0420.004-0.038(1.80-7.14)(.672-8.52)(0.023-0.101)(0.148-0.432)(0.027-0.076)(0.014-0.062)(0.018-0.064)(0.001-0.030)(0.004-0.053)(0.001-0.033)Crofton2.77a0.039a•162b0036ab0.033a0.03k0.016k0.018ko.oiz2.23-10.21.48-5.170.014-0.1100.121-0.2170.020-0.0620.020-0.5500.014-0.0650.006-0.0420.008-0.0400.008-0.020(3.47-6.38)(2.07-3.26)(0.024-0.50)(0.143-0.179)(0.028-0.044)(0.027-0.041)(0.021-0.038)(0.010-0.023)(0.013-0.024)(0.010-0.015)Powell75.0833a0.0710.32k’0066b0046a0.044a0028a0025ao.olzaRiver3.88-6.651.64-6.660.035-0.1430.234-0.4380.046-0.0940.034-0.0620.031-0.0640.016-0.0480.013-0.0470.003-0.046(3.32-6.96)(1.31-8.70)(0.029-0.215)(0.192-0.478)(0.041-0.109)(0.030-0.075)(0.021-0.065)(0.015-0.063)(0.007-0.052)(0.001-0.031)Johnstone72.5&’229a0.047”o.0z40031a0.o20.0130.024aStrait1.78-3.641.46-3.580.024-0.0580.144-0.3010.031-0.0790.005-0.1170.020-0.0480.014-0.0280.003-0.0520.016-0.036(1.70-5.34)(1.22-5.95)(0.026-0.112)(0.142-0.453)(0.033-0.119)(0.001-0.081)(0.017-0.071)(0.014-0.044)(0.001-0.046)(0.015-0.055)Alberni14.475.140.0280.1510.0280.0240.0380.0190.0240.001InletClayoquot13.865.120.0440.2930.0370.0470.060.0380.0260.019SoundLangara11.912.970.0340.2150.0390.0950.0240.0620.0610.051Islanda,b-meansthatdonotsharethesameletteraresignificiantlydifferent(p0.05)Table 3.5 Mercury residue levels (mg/kg, wet weight) in Bald Eagle eggs from locations onthe British Columbia coast, 1990-1992, expressed as geometric means and 95%confidence intervals (range in brackets).Lower Fraser Nanaimo Crofton Powell River Johnstone Strait Alberni Clayoquot LangaraValley (N=8) (N=3) (N=7) (N=7) Inlet Sound Island(N=6) (N=1) (N=1) (N=1)0.25&’ 0.147a 0.191a 0294b0.186 - 0.358 0.110 - 0.198 0.096 - 0.384 0.174 - 0.296 0.236 - 0.367 0.08 0.17 NA(0.170 - 0.400) (0.070 - 0.240) (0.150 - 0.260) (0.150- 0.380) (0.220- 0.440)a,b- means that do not share the same letter are significiantly different (p 0.05)NA - not analyzedPolychiorinated biphenylsMean 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 significantdifferences 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 ineggs 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 PowellRiver 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 Fraservalley and PCB 138 (234-245) was significantly higher at Powell River than Johnstone Strait.The percent contribution of individual congeners was determined and compared amongsites (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 ofstatistically significant differences among sites in percent contribution of individuals PCBs.Percent contribution of a number of the lower chlorinated congeners, including PCBs 66 (24-34), 101 (245-25), 99, 87 (234-25), 118 and 105 (234-34), was significantly higher at bothJohnstone Strait and the lower Fraser Valley than the other three sites. In addition, amongthose compounds, percent contribution of PCBs 99 and 118 were significantly higher at PowellRiver than at Crofton. In contrast, the percent contribution of a number of the higher81chlorinated congeners, PCBs 183, 180, 170, 203 and 194, was significantly lower in eggs fromthe lower Fraser Valley and Johnstone Strait than Nanaimo, Crofton and Powell River. Thisgeographic trend of differences in the PCB pattern was supported by results of principlecomponents analysis. Principle components analysis of the PCB pattern was carried out usingonly 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 Aroclor1242. Two significant principle components were apparent which explained 90 % of the totalvariance among individual egg analyses. The first component (PC 1) explained 75 % and thesecond component (PC2) explained 15 %. As shown in Figure 3.3, the Johnstone Strait andFraser Valley eggs clump separately from the other locations, although there is some overlap,particularly of some samples from Powell River.Figure 3.2 PCB congeners in Bald Eagle from British Columbia, 1990-1992, expressed as percent of totalPCBs. Values represent means of three to eight analyses per collection site. Congeners are identifiedaccording to their IUPAC number.252050p.. p. .. p..PCB Congener Number82Concentrations of six non-ortho PCB congeners were determined in eight eggs collectedin 1992 (Table 3.6) and in two eggs collected in 1994. The pattern was consistent in the 1992samples 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 ofPCBs 126 and 77 and sum-PCBs for the ten eggs in Table 3.3, in order to estimate values inthe whole data set for estimation of TCDD toxic equivalents (TEQs):PCB 126 (ng/kg) = 156 [sum-PCBs (mg/kg)] + 78, r2=0.634, p<O.OlPCB 77 (ng/kg) 69 [sum-PCBs (mg/kg)] + 85, r2 = 0.505, p <0.04PCBs66j99V 118L Lower Fraser Valley C Crofton J Johnstone StraitN Nanaimo P Powell RiverFigure 3.3 Plot of the first and second principle components (PC 1 and PC2). Selected PCB congenersonly, 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 arcsinetransformed. A total of 75% of the matrix variance was explained by PC 1 and 15 % by PC2.T PCBs170180194z-3 -2.5 -2 -1.5 -1 -0.5 0PRIN2PCB 990.5 1 1.5PCB 6683Table 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-189Northwest Bay (13) 6.2 24 146 323 65 22Jack Point (10) 31 52 349 709 131 41Powell River (20) 13 42 207 544 131 35Stiliwater A (21) 26 105 720 1354 285 84Stillwater B (21) 24 107 684 1326 277 97Grise Point (22) 52 46 387 547 121 41Pocohantas Point (30) 15 51 459 685 114 35Berrryman Point (31) 23 74 691 754 135 39Herrling Is. 27 96 576 314 47 5Langara Is. 5 32 310 585 203 1* 10-13- Nanaimo,60020-22 - Powell River, 30 - Albemi Inlet, 31- Clayoquot Sound.500—400c,)0)U)CwI— 200çcclIE non-O-PCBsmono-O-PCBsLI PCDFsother-PCDDsmTCDDii10000’Cl’$Figure 3.4 The contribution of various chlorinated hydrocarbon groups to the sum of TCDD toxicequivalents (TEQs) in Bald Eagle eggs from coastal British Columbia, 1990-1992 (N values and variancesare in the tables). Toxic equivalents for PCDDs/PCDFs from Safe 1990 and for PCBs from Ahlborg etal. 1994.84Table 3.7 Eggshell thickness data, mean ± SD, (range in brackets) for Bald Eagle eggs collectedfrom British Columbia, 1990-1992.Area Collection N Shell thicknessPeriod (mm)Lower Fraser Valley 1990-91 6 0.558 ± 0.024Nanaimo 1990-92 5 0.587 ± 0.035Crofton 1990-91 3 0.583 ± 0.024Powell River 1990-92 5 0.590 ± 0.038Johnstone Strait 1991 6 0.569 ± 0.036Percent Difference fromprel947*-8.3 ± 4.3(-14.6 to +1.5)-3.6 ± 6.4(-8,6 to +5.2)-4.2 ± 4.7(-9.7 to -1.5)-3.1 ± 7.0(-11.3 to +6.7)-6.6 ± 6.3(-12.9 to + 3.5)* pre-1947 value - 0.6088 (Anderson & Hickey, 1972)Bioaccumulation of PCDDs and PCDFs from preyAn example output from the model is shown in Table 3.8, based onCrofton, the location with the best data base of contaminants in prey items.local data, results on gulls and salmonids from nearby Nanaimo were used.1990 data fromIn the absence ofThe putative dietTCDD toxic equivalents (TEQs)Highest mean TEQs0were in eggs from Crofton, followed by Powell River, both ofwhich were significantly greater than Johnstone Strait. The relative contribution of PCDDs andPCDFs to total TEQs0,64 %, was also highest at Crofton and was lowest, 47 %, in thelower Fraser Valley eggs, as shown in Figure 3.4.Eggshell thickness resultsNeither mean eggshell thickness nor the percentage difference from the pre-1946average 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 orother organochlorines.85Table3.8AsimulationofPCDDandPCDFlevelsinBaldEagleeggsatCrofton,1990, basedonconcentrationsinthediet.Contaminantconcentrationindietaryitems(ng!kg, wetweight)(Fractionof thatiteminsimulateddiet)Birds(0.475)Fish(0.525)ContaminantconcentrationinbaldeagleeggsChemicalBMF’Non-fishGullsHeronsCormorantsNon-SalmonidsCalculatedMeasuredMeaneatingbirds2(0.25)(0.05)(0.025)salmonids3(0.125)ValueValue(0.15)(0.4)2378-TCDD2173100302111710712378-PnCDD10152204620.582180123678-HxCDD1611723070172.52843422378-TCDF1.4350.5100.53715312123478-PnCDF4.52135120.50.58311BMF-biomagnificationfactor2BuffleheadandSurfScoterSole,flounder,rockfishconsisted of 52.5 % fish, and 47.5 % birds, mainly gulls and non-fish-eating species; fish-eating birds comprised only 6 % (herons and comorants). The model accurately predicted2,3,7,8-TCDD levels in eagle eggs, but concentrations of other compounds, such as 1,2,3,7,8-PnCDD, 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 2,3,7,8-TCDD level in the forage fish prey was relatively high, while levels for the other PCDDs andPCDFs were near the detection limit; thus, a small difference in forage fish concentrationswould translate to a large error in the estimated BMF. The other major source of error is theputative 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. Forexample, Knight et at. (1990) reported that western grebes, which can accumulate extremelyhigh PCDD/PCDF levels (Table 3.2), were the main prey item of Bald Eagles in the PugetSound 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 ineagle eggs with an increasing fish-eating bird diet. Feeding on fish-eating birds may accountfor extremely high liver levels of PCDDs, PCDFs and other chlorinated hydrocarbons of adulteagles found dead or dying near Powell River and other areas of the Strait of Georgia(Chapter 1).875*Crofton, 1993C) ±Crofton 1990)K0 10 20 30 40 50 60 70 80 90Percent fish-eating birds in dietFigure 3.5 Concentration of 2,3 ,7,8-TCDD predicted in Bald Eagle eggs based on the percent offish-eating birds in the diet. Prediction is based on a bioaccumulation model described in the text andthe simulation is based on data from Crofton, British Columbia.88DiscussionThe data presented in this chapter show that Bald Eagle eggs collected near bleachedkraft pulp mills in the Strait of Georgia contained higher levels of 2,3,7,8-TCDD and -TCDFwhen compared to other locations on the British Columbia coast. Total PCB levels were alsohighest in eggs from the Strait of Georgia, reflecting greater industrialization. Concentrationsof organochiorine pesticides, including DDE, in eagle eggs were relatively consistent amongsites. Total-mercury levels were significantly higher in eggs from the Fraser Valley andJohnstone Strait than the Strait of Georgia.Patterns and sources of PCDDs/PCDFsThe formation of 2,3,7,8-TCDD and 2,3,7,8-TCDF during molecular chlorine bleachingof 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 minimizeTCDD/TCDF formation (Table 3.9), which has resulted in declining PCDD levels, particularlyof 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 werestill elevated in 1992, suggesting that efforts to reduce TCDF contamination have been lesssuccessful. In birds, 2,3,7,8-TCDF is quite quickly cleared from the body (Braune andNorstrom 1989; Van den Berg et at. 1994). Other studies of wild birds have reported low2,3,7 ,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 reportedin 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 fromnest locations downstream of pulp mills in the British Columbia interior contained 2,3,7,8-TCDF levels up to 68 ng/kg (Whitehead et at. 1993). The high TCDF levels in eggs of eaglesand ospreys likely reflect a combination of recent exposure and direct yolk deposition ofcontaminated dietary lipids, as suggested previously for great blue herons (Elliott et at. 1989a).89Until 1989, up to several million kg of chiorophenolic compounds were used annuallyby the British Columbia forest industry, particularly on the coast, to prevent sap staining ofundried lumber. Although HxCDDs and HpCDDs predominate as dioxin contaminants inchlorophenol mixtures, HxCDDs are further produced in large amounts when chlorophenolcontaminated woodchips are pulped (Luthe et al. 1990). Monitoring chip supplies forchiorophenols, followed by a regulatory ban on their use as anti-sapstains, produced significantHxCDD 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 atJack Point near Nanaimo.In Fraser valley eagle eggs low HxCDD : TCDD ratios are consistent with lowerHxCDD concentrations in sediments and biota downstream of Fraser river pulp mills, theputative 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, lesseramounts of chlorophenol antisapstain agents were use by lumber operations on the upper Fraserand Thompson Rivers. Osprey eggs collected in 1991 from nests located downstream of thepulp 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 containedvery high levels of 1 ,2,3,4,6,7,8-HpCDD and OCDD, indicative of direct chlorophenolicinputs, rather than via pulp milling of contaminated wood chips.Although there are no pulp or large saw mills on northern Johnstone Strait (only logsorting facilities), PCDD/PCDF levels in eagles were relatively high. A non-kraft pulp milllocated 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 hepatapancreasfrom near the mill site (Harding and Pomeroy 1990). The PCCD/PCDF pattern in JohnstoneStrait 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 southern90movements is unlikely as resident Bald Eagles on the Pacific coast remain on territory for mostof the year (Frenzel et al. 1989). Residents may leave breeding territories periodically duringthe fall and winter to feed at salmon spawning sites; however, Pacific salmon, even from nearpulp 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 (Table3.3, 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 localizedeven 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 (Thomson1981), which may conceivably transport some suspended sediment-bound PCDDs and PCDFs.A sediment sample from Louchborough Inlet, a fjord off of central Johnston Strait, wasreported to have levels of higher chlorinated PCDDs comparable to those near industrial sites inthe Fraser delta (Harding 1990). However, eagle prey species, such as western grebes and surfscoters collected from Johnstone Strait in mid-March 1992, timed to obtain birds which hadwintered on site, had very low PCDD/PCDF levels, while samples of the same speciescollected near pulp mills showed the typical pulp mill PCDD/PCDF signature. Johnstone StraitBald Eagles may still be exposed to contaminants from the Strait of Georgia by feeding onwaterfowl during spring migration along the coast towards their northern breeding grounds.Rhinoceros auklets, large numbers of which breed in northern Johnston Strait, contained lowPCDD levels, although the mean 2,3,7,8-TCDF concentration was quite high and couldpartially account for this compound in Johnston Strait eagle eggs.The pattern of HxCDD > PnCDD > TCDD in Strait of Georgia wildlife differs from thatreported at other locations such as the Great Lakes (Ankley et al. 1993), interior rivers ofBritish 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 Georgiablue heron eggs clustered separately from Great Lakes herring gulls and other biota, based onhigher PnCDD and HxCDD concentrations, attributed to chlorophenol sources. However, a91sample 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, perhapsindicating 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, recalculated based on 17 % lipid in common murre eggs (Noble and Elliott 1986; Cederberg etal. 1991), similar to the Strait of Georgia pattern. Grey Heron (Ardea cineria) livers from theNetherlands also had a pattern somewhat similar to the Strait of Georgia, which was attributedmainly to chlorophenols (Van den Berg et al. 1987).European wildlife samples, at least from The Netherlands, appear to have higher2,3,4,7,8-PnCDF concentrations (Bosveld et at. 1994; Van den Berg et a!. 1994b) compared tothose from North America (Elliott et at. 1989a; Hebert et at. 1994). This compound is aknown contaminant in PCB mixtures (Van den Berg et al. 1985), which would explain itsassociation with areas of PCB contamination (Hebert et a!. 1994) and its tendency to correlateclosely with PCB congeners in eggs (Elliott et at. 1989). Bosveld et at. (1994) suggested thathigher 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 birdeggs from industrialized areas of the Great Lakes. However, direct comparison of lipid-normalized whole egg to yolksac concentrations may overestimate concentrations in yollcsacs.For example, in Bald Eagles, concentrations of chlorinated hydrocarbons were three-fold higheron a lipid weight basis in a single yolksac compared to the sibling whole egg. On a wet weightbasis, total PCB levels in Great Cormorant (Phatacrocorax carbo) eggs from the contaminatedBiesbosch colony in the Netherlands (Van Hattum et at. 1993 cited in Bosveld and Van denBerg, 1994) were similar, about 23 mg/kg, to those in double-crested cormorants from highlycontaminated Hamilton Harbour in the Great Lakes (Bishop et al. 1992). Therefore,differences in PCB formulations or other sources may account for higher PnCDF levels inEuropean wildlife samples, rather than higher PCB levels.92Patterns and sources of organochiorines and mercuryThe uniformity in OC residues indicates similar dietary exposure among mostindividuals. The few eggs with distinctly lower organochiorines are probably individual eaglesfeeding on larger amounts of fish, non-fish-eating birds or mammals. Based on OC patterns inseabird eggs, Elliott et al. (1989) concluded that atmospheric sources were dominant over awide area of the British Columbia coast. However, local sources can still pre-empt theinfluence of atmospheric input: DDE levels in heron eggs were significantly higher in coloniesfrom the Fraser delta (0.49 mg/kg), an area of intensive farming, than non-agriculturallocations (0.11 mg/kg) (Elliott et al. 1989; Whitehead, 1989). In fact, the mean DDE level intwo eagle eggs collected within the Fraser delta, 3.86 mg/kg, is significantly higher than thefour eggs from upstream of the main agricultural areas, 1.63 mg/kg DDE. High DDE levelscontinue to be reported in wildlife from areas of former high DDT use, such as orchards (Bluset al. 1987; Elliott et al. 1994).After the DDT-related compounds, chiordanes were present at the highest concentrationsin 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, consideredto 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 tooxychlordane levels in tissues shows a lower specific capacity to metabolize chlorinatedhydrocarbon compounds (Kawano et al. 1986; Yamashita et al. 1993).The concentrations of chlordane-related and heptachlor epoxide compounds found hereare similar to those reported in addled Bald Eagle eggs collected in the early 1980s from avariety of United States locations (Wiemeyer et al. 1993). Concentrations of mirex anddieldrin were somewhat higher in those U.S. Bald Eagle eggs collected a decade earlier than inthe fresh eggs from the British Columbia coast in 1992. Mean DDE and PCB levels wereabout three-fold higher in eagle eggs from the lower Columbia River than the lower FraserRiver (Anthony et al. 1993). Dietary differences may partly account for this; eagles in the93lower Columbia reportedly ate more Western Grebes (Watson et al. 1991), which tend to havehigh levels of chlorinated hydrocarbons (Table 3.7), while Fraser estuary eagles ate a largeproportion 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). Differencesin organochiorine levels in estuarine biota also reflect differences in agricultural and industrialdevelopment of the respective watersheds. Areas of intensive agriculture, particularly fruitorchards 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 accountsfor higher PCB concentrations, evident in Osprey eggs collected in the upper reaches of eachwatershed (Whitehead et al. 1993).Higher mercury levels in Bald Eagle eggs from the Fraser estuary are consistent withdata in herons from that site (Elliott et al. 1989a), and with Fimreite et al. ‘s (1971) findings ofhigher mercury in aquatic versus coastal marine fish. Elevated mercury levels in fish-eatingbirds were associated with industrial, including pulp mill, sources by Fimreite et al. Based onthe levels in eagle eggs, any mercury discharges from Crofton and Nanaimo pulp mills havenot had a lasting impact in local food chains. Highest mercury levels were in the JohnstoneStrait eagle eggs. A great proportion of fish in the diet may explain higher mercury levels inJohnstone Strait and the lower Fraser Valley, as suggested below to account for the PCBpattern at those sites.Polychiorinated bihenylsMean total PCBs in Bald Eagle eggs were highest near the three pulp mill sites, whichcontrasts with data on great blue herons, in which highest PCBs were from colonies in theFraser delta near Vancouver (Elliott et al. 1989a; Whitehead 1989). However, the PCBconcentration in the single Bald Eagle egg from an industrial site in the Fraser delta, 6.21, wasin the same range as the eggs from near the pulp mill sites; other Fraser valley Bald Eagle eggswere from agricultural or woodland locations and PCB levels were 50 % lower. The PCBpattern in great blue herons varied significantly among sites which was attributed to local94differences in Aroclor inputs (Elliott et al. 1989a). Variability in PCB congener patterns inwildlife 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) considereddietary differences as the best explanation for individual variation in PCB patterns. This mayalso apply in Bald Eagles given the consistent differences in the PCB pattern between JohnstoneStrait and lower Fraser valley eggs compared to those from the Strait of Georgia sites. TotalPCB levels were also lower in the Johnstone Strait and lower Fraser valley eggs. The mostlikely explanation is of more fish in the diet of Fraser and Johnstone Strait eagles and thusgreater exposure to the lower chlorinated PCBs. Higher total mercury levels at those two sitesare also consistent with more fish in their diet. The PCA results can be used to support thisexplanation; however, alternatively the differences among sites may also be explained bydiffering local Aroclor inputs. Fraser delta eagle eggs, like Great Blue Heron eggs, containmore of PCB 66, indicative of Aroclor 1242 input, while the pulp mill areas, includingCrofton, 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 areamay 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 higherconsumption of fish which have low capacity to metabolize PCBs (Brown 1994). The ratio atmost sites of non-ortho PCBs 126:77 was 2:1, except in the eggs from Alberni Inlet andClayoqot Sound, where the ratio is closer to 1:1, and the egg from Herrling Island in theFraser valley, where the ratio was 1:2. Although the ratios vary somewhat, the other nonortho 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 tissuescompared to egg were caused by reduced metabolic capability in embryos. Levels of more95rapidly metabolized compounds such as PCB 77 may also be higher in eagle eggs as a result ofdirect 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 inmethodology. Total PCB numbers in Wiemeyer et al. (1993) were probably determined asAroclor estimates based on the analytical references. Determination of total PCBs based forexample on Aroclor 1254:1260 overestimate total PCBs, based on the sum of congeners, byabout two-fold (Turle et al. 1991).Toxicological significance of PCDD and PCDF levelsThe bioaccumulation model was developed in order to estimate critical concentrations of2,3,7,8-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 aremore easily monitored than eagles. Levels in the monitoring species should indicate a degreeof foodchain contaminant which should result in accumulation in bald eagle eggs less than thesuggested NOEL from Chapter 2.Using the same BMIF of 21, the average 2,3 ,7,8-TCDD concentration in forage fishconsumed by great blue herons in 1990 at Crofton would have been about 5 ng/kg. With thepostulated eagle diet in Table 3.8, TEQ5PCDD,PCDF. in eagle eggs were calculated as 193 ng/kgversus the measured value of 248 ng/kg. If an average value of 115 ng/kg TEQs0 for nonortho and mono-ortho PCB contribution at Crofton, 1990 is included, the total TEQs0were308 and 355 ng/kg, calculated and measured respectively, both of which exceed the LOEL (210ng/kg), determined for Bald Eagle embryos (Chapter 2). If the data from Crofton, 1992, areused the estimated mean value of 1 ng/kg in forage fish gives a calculated TEQWIIO value ineagle eggs of 194 ng/kg (79 TEQsPCDD,PCDFS + 115 TEQsPCBS), less than the LOEL, but stillgreater than the NOEL of 100 ng/kg. Therefore, assuming that both the ratios of otherPCDDs/PCDFs and PCB levels remain constant, a maximum value of 0.5 ng/kg 2378-TCDDin forage fish is suggested as site-specific dietary concentration in the Strait of Georgia, toavoid adverse toxic effects of TCDD-like chemicals in Bald Eagle populations. Thecorresponding concentration of 2,3,7,8-TCDD in Great Blue Heron eggs, to avoid TCDD96toxicity in both herons and top predators, such as the Bald Eagle, in the Crofton area is 10ng/kg. At other areas in the Strait of Georgia, given that ratios of PCDDs, PCDFs and PCBsare similar, a value of 10 ng/kg in double-crested or pelagic cormorants, would also indicatethat levels in local foodchains should not cause toxicity in Bald Eagles, given a typical diet, asshown in Table 3.8. The utility of colonial waterbirds as sentinel species for monitoring oftoxic contaminants has been demonstrated in many studies (Gilbertson et al. 1987). Given thatthe 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, CYP1Ainduction, then these critical values suggested for prey items, should provide a reasonablemargin of safety.The above values would be effective in areas with contaminant profiles which aresimilar to the Strait of Georgia. However, as shown in Figure 3.3, in Common Tern eggs,PCDDs made only a minor contribution to the TEQs0,relative to the non-ortho PCBs(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 denBerg 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 lowerColumbia river, were 32 ng/kg, less than those found in eagle eggs near pulp mills on the Straitof Georgia. However, total PCB levels were 12.7 mg/kg, more than two-fold higher than thehighest mean concentrations in Table 3.4. Thus, TEQs110 in Bald Eagle eggs from the LowerColumbia 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 andproductivity 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 nonortho 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 et97at. 1994b). However, the data on Bald Eagle chicks reported in Chapter 4 suggests that PCBcongeners are less potent relative to PCDDs and PCDFs in Ah-receptor mediated biomarkerresponses, such as CYP1A induction. Nevertheless, total PCB levels up to 119 mg/kg havebeen reported in recent years in adled Bald Eagle eggs from the Great Lakes region (Bowermanet al, 1994); that egg would have contained about 18,500 ng/kg of PCB 126 using theregression from the Result section above. PCB concentrations of that degree may partlyaccount 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 todetermine total PCB concentrations in foraging fish and a sentinel fish-eating bird, which wouldresult in a PCB contribution to TEQ5 in eagle eggs less than the NOEL of 100 ng/kg. Usingthe BMF for PCBs of 30 from Braune and Norstrom (1980), assuming constant ratios of nonortho 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 aresuggested, would be necessary to achieve TEQs0 less than the NOEL of 100 ng/kg in BaldEagle eggs. This value for forage fish is much lower than 0.2-0.4 mg/kg suggested by Harriset 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 fishwould be required to avoid accumulation of toxic levels in eagles. However, at most sites thecontribution of PCDDs and PCDFs to TEQs is considerably less than at Crofton and isprobably 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 PCDFsand a comprehensive dataset on PCBs would enable determination of better guidelines forPCBs.Toxicological signjficance of organochtorine and mercury levelsWiemeyer et at. (1993) determined that DDE was the chemical contaminant mostassociated with reduced breeding success of Bald Eagles in the United States during the period1969 - 1984. Production of young began to decrease at DDE levels > 3.6 mg/kg, and furtherdecreased at > 6.3 mg/kg. DDE levels of 16 mg/kg were associated with fifteen percent98eggshell 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 UnitedStates. 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 meaneggshell thickness was less than the pre-1946 mean at all sites, there was no significantrelationship between DDE and eggshell thickness, likely because of the narrow range of DDEconcentrations.Although quantitative data are limited, there were no reports of widescale declines ofcoastal eagle populations in British Columbia, as occured in other areas of North Americaduring the organochlorine era. However, Vermeer et al. (1989) reported an increase in BaldEagles nests in the southern Strait of Georgia between the mid-1970s and the late 1980s. Theyattributed eagle population growth to increased prey populations, particularly glaucous-wingedgulls, 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 Straitof Georgia eagle eggs. In Great Blue Heron eggs from a Fraser delta colony, DDE declinedfrom a mean of 2.0 mg/kg in 1977 to 0.42 mg/kg in 1990 (Whitehead 1989; Canadian WildlifeService, unpublished data). In Pelagic and Double-crested Cormorant eggs from MandarteIsland in the southern Strait of Georgia, DDE decreased by factors of five and ten respectivelybetween the early 1970s and the late 1980s (Elliott et at. 1989a). Organochlorine levels inBald Eagle eggs are currently about ten-fold higher than in those of marine and fish-eatingbirds from the Pacific coast (Elliott et at 1989a; 1989b). If the ten-fold difference was constantover time, then during the late 1970s mean DDE levels in Bald Eagle eggs from the Fraserdelta would have been about 25 mg/kg, high enough to cause nest failures and reducedproductivity. 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 BritishColumbia, Bald Eagles declined as a breeding species between the 1930s and 1970s (Canningset at. (1987). Although habitat loss was likely a factor, the extremely high DDE levels in99Okanagan valley foodchains (Elliott et al. 1994) probably continue to impact Bald Eaglereproduction 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 onproductivity.In conclusion, Bald Eagle eggs collected in the Strait of Georgia contained elevatedlevels of PCDDs and PCDFs; the pattern was similar to that measured in other components offood chain and indicative of both bleached kraft pulp mill and chiorophenol sources. Relativelyhigh PCDDs and PCDFs in a supposed reference area in northern Johnstone Strait probablyresulted from feeding on waterbirds migrating north from the Strait of Georgia. Recommendedsite specific concentrations of 2,3,7 ,8-TCDD are 0.5 ng/kg in forage fish and 10 ng/kg insentinel fish-eating bird eggs in the Strait of Georgia are suggested to avoid accumulation ofpotentially harmful levels in Bald Eagle eggs. Likewise, total PCB concentrations of 0.01mg/kg in forage fish and 0.3 mg/kg in fish-eating bird eggs are suggested as maximumconcentrations to prevent accumulation of potentially harmful PCB levels in Bald Eaglepopulations.AcknowledgementsIan Moul, George Compton, Andre Breault, Dave Dunbar and Ray Caton assisted withcollection of eggs. Mary Simon did the PCDD/PCDF and non-ortho PCB analysis, whileHenry Won did the organochlorine pesticide analysis. John Smith provided statistical advice.Funding was provided by the Canadian Wildlife Service and the British Columbia Ministry ofEnvironment.100Table3.9CharacterizationofBritishColumbiapulpmillsdiscussedinthispaper.DatawereobtainedfromEnvironmental EffectsMonitoring(EEM)reportssubmittedbyindustrytoEnvironment Canada.MillInitialProcessTotalBleachedBleachingTreatmentChlorineAOX(kg/Adt)ReceivingStart-upProductionProductionSequenceSubs.Environmentlinpiem.ADtPrimarySecondaryprepost(airdriedtonnes)CL-subCL-subCrofton1957Kraft2,4001,100C/D,E0,W,ClariflersOEAS11988-905.41.2StuartGroundwoodD,E,DEqualiser(1989)(1992)ChannelHarmac1950Kraft11201120W,DC,R,EO,ClarifiersAeration19916.01.1NorthumberlaD,E,D(1989)(1993)ndChannelPowellRiver1912KraftCTMP22,1252,125D80/C2,E0ClarifierAAS419919.11.3MalaspinaGroundwoodH1,H2(1990)(1993)StraitPortAlberni1947CTMP550550H20ClarifierAS/ASB519905.52.2AlberniInletGroundwood(1989)(1993)PortAlice1918470470C,C,E,C,H,ClarifierAAS1995<1.0NeroutsosD,A(1995)Inlet1c-elememental chlorine,D-chlorinedioxide,E0-hot causticextraction,W-wash,H-sodiumhypochlorate,A-acidtreat2OEAS-oxygenenhancedactivatedsludgeCTMP-chemi-thermomechanical pulpAAS-aerobicactivatedsludgeAS/ASB-activatedsludge/aeratedstabilizationbasinCHAPTER 4INFLUENCE OF CONTAMINANTS AND FOOD SUPPLYON BALD EAGLE PRODUCTIVITYThe results of the previous chapters showed that Bald Eagle populations in the Strait ofGeorgia were exposed to elevated levels of PCDDs and PCDFs relative to referencepopulations. Eggs collected in 1990 and 1991, particularly near Crofton, had higherPCDD/PCDF levels and modelling showed that theoretically eagles which preyed on a largercomponent 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 inlivers. Thus, some component of the breeding population may be affected each year bychlorinated hydrocarbon toxicity. Eggs collected near pulp mills in 1992 and incubated in thelaboratoiy did not exhibit significant effects on hatchability and most morphological andphysiological endpoints, although a hepatic CYP1A cross-reactive protein was induced. For thework described in this chapter, I measured breeding success of Bald Eagles near three pulpmills in the Strait of Georgia, at two areas of the Fraser delta, and at reference sites on thewest coast of Vancouver Island, in northern Johnstone Strait and in the Queen CharlotteIslands. The objective of the study was to determine occupancy of breeding territories, measurenest success and compare the results to chlorinated hydrocarbon levels in nestling plasmasamples.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 alreadypoor reproduction. My initial studies on the coastal BC eagle breeding population (Chapters 2and 3) used eggs collected during incubation; however, this resulted in an unacceptable level ofnest 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 more102randomized sample for contaminant analysis (Henny et a!. 1981; Frenzel 1985), an approachthat 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 nestsfrom productivity estimates from an area, and also permits determination of a directrelationship between contaminant levels in chicks and 5-year average productivity for theterritory in which they were produced. Because of development of advanced high resolutiongas chromatography/mass spectrometry (GC/MS) analytical techniques, beginning in 1993, theNWRC lab was able to quantify PCDDs, PCDFs and non-ortho PCBs in nestling Bald Eagleplasma samples.Materials and MethodsProductivitySurvey routes were flown in exposed or ‘treatment’ study areas, selected on the basis ofeagle nest density near industrial pollutant sources: Crofton, Nanaimo and Powell River (pulpmills) and the lower Fraser valley (mixed industrial sources) (Figure 4.1). Reference or controlsites 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-flightapproach (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 todetermine the number of eagle pairs attempting to breed. Timing of this flight varied from lateMarch in the Fraser delta to mid-May in the Queen Charlotte Islands. The second flight wastimed 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 youngproduced by the number of occupied breeding territories, as described in Postupalsky (1974).103Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections. At Langara Island,the survey circumscribed the coastline of the island.CeCe)CC=cieC104Prey deliveriesPrey deliveries were observed during 1995 at five nests in the Fraser delta and at fournests in Barkley Sound, using methods described in Dykstra (1994). From a blind, dawn todusk observations were made using a 20-60X spotting scope. Prey deliveries and other nestlingand adult behaviours were recorded. Observers were switched every four to eight hours. Inthe Fraser delta five nests were observed for five days each. In Barldey Sound, three nestswere observed for five complete days each, while one nest was watched for part of each dayand therefore was not included in statistical analyses.Sample collectionNests suitable for collecting were located by ground, boat and aerial surveys, when theywere scored to estimate ground access and suitability for climbing; land tenure was alsoconsidered. Samples were collected when nestlings were 5-9 weeks old. Collections weremade during the first week of June in the Fraser Valley and the Strait of Georgia, during lateJune 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 theground 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 transferredimmediately to heperinized vacutainers and stored on ice. Samples were centrifuged within sixhours of collection and plasma transferred to chemically cleaned (acetone/hexane) glass vialswith teflon liners and then frozen.Chemical analysisFrozen plasma samples were shipped to the CWS National Wildlife Research Centre(NWRC) for analysis in the laboratory of Dr. R.J. Norstrom. For organochlorines, thesamples (1 ml of each) were first deproteinized with 0.5 ml of methanol containing aldrin as aninternal standard (Smrek et al. 1981). The plasma was then extracted with hexane and105DioxinFisheryLXZJClosureArea•BaldEagleneatsitessurveyedSoutheyI.‘<—WinchelseaI..\rviaudeI.PowellRiverPulpMillCroftonPulpMillVANCOUVER0510kmIIIcentrifuged. The hexane extracts were passed through sodium sulphate, evaporated to 1 ml orless 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 ona 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 eggpool 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 (‘3C12-PCDDsIPCDFsInon-ortho PCBs) were added to the plasma, and allowed to equilibrate for 30minutes. Saturated aqueous animonium sulphate and absolute ethanol were added to the spikedplasma, and the samples were then extracted four times with hexane. The hexane layers werecombined, filtered through anhydrous sodium sulphate and the volume reduced for clean-upwith by gel penneation chromatography (GPC) (Norstrom et a!. 1986). Lipids and biogenicmaterials were removed by GPC and alumina column clean-up. Separation of PCDDs, PCDFsand 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-orthoPCBs was done with florisil column chromatography. Quantitation was performed with a VGAutospec double-focusing high resolution mass spectrometer linked to a HP 5890 Series II highresolution gas chromatograph. Recoveries of13C2-PCDDsIPCDFs/non-ortho PCBs werecalculated by comparing the integrated areas of the labelled internal standards and the areas ofthe recovery standards in the samples to the areas of these compounds measured in an externalstandard mixture, analyzed along with the samples. Results were accepted when recoveries of13C2-PCDDs/PCDFsInon-ortho PCBs were between 70% and 120%. For a few Bald Eagleplasma samples, the internal standard recoveries were <70%, due to losses during lipidextraction.Lipid was determined by combining 1-2 ml of sample with 4 ml of hexane in acentrifuge 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,107similar to the method of Mes (1987). The hexane was then passed through sodium sulphate toremove any moisture. This process was repeated twice more and the sodium sulphate washedwith hexane after the final extract. The three hexane extracts were combined on a pre-weighedaluminum dish, the hexane was then evaporated and the dish re-weighed to determine theamount of lipid. Lipid was then calculated on the basis of grams per ml plasma.Statistical analysesThe SYSTAT software package was used for statistical analyses of all data. Wet weightchemical 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 ofchlorinated hydrocarbons tested were significantly correlated with percent plasma lipid (Table4.1). DDE was only weakly correlated with plasma lipid, while the higher chlorinated PCDDsand PCDFs were not significantly correlated. There was also a significant interaction betweenplasma lipid and sampling location. Therefore, for testing of differences among locations, allof the contaminants which correlated significantly with plasma lipids, were further transformedusing an analysis of covariance (ANCOVA) to account for the effect of variation in plasmalipid content among individuals and locations (Hebert and Keenlyside, 1995). Differencesamong locations were then determined using Bonferroni’s test. In a few cases, percent plasmalipids were three to ten-fold greater than the mean of the other samples at that site; thosesamples 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 theANCOVA.Productivity measures were compared among locations with a one-way analysis ofvariance (ANOVA); significant differences were determined using Tukey’s multiple comparisonprocedure (MCP). Data were also compared on the basis of a pulp mill versus non-pulp millgrouping and significant differences identified using Student’s t-test. At each pulp mill site,108Table4.1CorrelationMatrix(rvalue) forpercent plasmalipidandselectedchlorinatedhydrocarboninbaldeaglenestlingsfromBritishColumbia,1993—94LipidOCDDHpCDDHXCDDPnCDDTCDDOCDFHxCPnCDFTCDFDDEHLBCMirex1—nonachlorSUM—PCBsPCB—99PCB—118PCB—153PCB—180PCB—37PCB—77PCB—126PCB—169TEQa99118153150370.9340.9420.9420.9420.9950.9930.9950.9420.238—0.046—0.082—0.051—0.1700.8880.4420.9490.3930.9820.32877126169TEQsProductivity0.9430.8670.958—0.071—0.046—0.081—0.015—0.083—0.050—0.095—0.005—0.0730.8400.7790.893—0.0860.9080.8420.952—0.0700.9540.8700.986—0.0740.9060.990—0.1140.9000.986—0.1370.3300.343—0.0940.9330.9770.0940.9640.9870.0740.9270.004—0.080transSumLipidOCDDHpCDDHXCDDPnCDDTCDDOCDFHxCLEPnCDFTCDFDDEHCBMirexnonaclor—PCBs—0.025—0.0080.7710.8720.961—0.0040.0320.9420.9310.5690,8680,8980.9690.9610.9780.9730.9640.835—0.035—0.058—0.0350.1340.791—0.031—0.077—0.046—0.077—0.074—0.060—0.034—0.040—0.027—0.0330.041—0.047—0.0280.1790.587—0.029—0.085—0.059—0.101—0.087—0.062—0.040—0.043—0.027—0.0370.9620.8960.0810.0660.8410.8860.4760.6340.7840.8510.8700.8300.8510.8570.964—0.0070.0560.8840.9430.5340.7440.8620.9230.9380.9090.9250.926—0.0050.0400.9270.9670.5800.0100.9040.9830.9860.9770.9830.9800.919—0.001—0.0080.8520.9080.9430.261—0.021—0,008—0,065—0.055—0.074—0.034—0.025—0.035—0.028—0.030—0.0230.183—0.031—0.051—0.065—0.0190.1820.0570.007—0.101—0.042—0.045—0.0180.005—0.0060.0090.0040.0040.0540.9410.5280.8540.8660.9380.9310.9430.9420.9360.5320.7990.8720.9560.9550.9530.9550.9490.6810.6240.6540.6290.6250.6180.6180.8970.8490.8440.8640.8550.8450.0480.0320.0430.0500.1350.9170.3640.9490.4440.6180.0910.8300.1350.9500.2000.9870.3160.9430.9250.5650.7920.9090.9590.9480.9940.9580.9250.6080.8530.9400.9730.9290.9590.0140.8600.9620.0380,5440.600—0.3490.8650.840—0.1280.9250.932—0.1420.8940.987—0.1220.9940.9900.9980.9960.3000.9630.9720.8950.990—0.1190.9980.9960.9840.2780.9590.9730.8990.985—0.1040.9910.9890.2840.9650.9750.9010.990—0.1030.2990.2890.9690.3940.9610.3940.9780.9700.3240.983productivity at nests adjacent to dioxin fishery closure areas was compared to nests adjacent toareas outside the closure area, using a one-way ANOVA. We treated the closure areas as anindication of the area impacted directly by PCDD and PCDF contaminants in the respectivepulp mill effluents. Mean 3-year productivity at individual nests was also compared tocontaminant levels in nestling blood samples from each nest using regression analysis. Unlessstated 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 HealthOrganization) TEFs.ResultsProductivityMean three-year productivity was highest at nests in the Fraser valley and delta andcomparable along south-east Vancouver Island (Table 4.2). The number of young/occupiedterritory was lower (significantly compared to the lower Fraser valley) at nests around PowellRiver and at Langara Island. Lowest productivity was in Clayoquot Sound, Johnstone Straitand South Moresby.Productivity of eagle nests located along the shoreline adjacent to the dioxin fisheryclosures in the Crofton area was significantly lower than at nests located outside the closurearea (Figures 4.2 and 4.3). There were no significant differences in productivity at nestsadjacent to the closure areas compared to those outside the closure areas at both Nanaimo andPowell River. However, the four eagle nests closest to the mill on the north side (Powell Rivernest and three Gibson’s Beach Park nests) produced only two chicks between 1992 and 1994 innine nesting attempts. In contrast the next five nests to the north (three Scuttle Bay nests, KeesBay and Lund) during the same time frame produced 21 chicks in 15 nesting attempts. Thisdifference, was not statistically significant, however, likely due to small sample sizes.110Table 4.2 Nest success and production of young for Bald EaglesBritish Columbia coast (1992-94).at nine study areas on thea,b,c,d- means in the column that do not share the same letter are significantly different (p<O.O5)Study Area Year No. Successful % Nest No. Young!occupied Nests Success young occupied nestterritories produced1992 19 19 100 27 1.41993 22 18 82 27 1.21994 21 18 86 29 1.4Lower Fraser ValleyFraser DeltaSouth-east VancouverIslandPowell RiverBaridey SoundClayoquot SoundJohnstone StraitSouth MoresbyLangara IslandMean 89 1.3a1993 9 7 78 12 1.31994 11 9 82 14 1.3Mean 86 l.31991 19 11 58 17 0.901992 30 19 63 30 1.001993 34 22 65 35 1.001994 42 27 64 43 1.00Mean 63 0. 97ab1992 24 14 58 18 0.751993 37 25 68 36 0.971994 36 21 58 33 0.92Mean 61 0.88&1992 36 16 44 21 0.581993 35 20 57 26 0.741994 30 8 27 12 0.4Mean 43 0.57c1992 23 12 52 12 0.521993 43 10 23 14 0.331994 35 2 57 3 0.09Mean 27 0.31d1991 6 2 33 2 0.331992 26 10 39 12 0.461993 34 3 8.8 4 0.121994 31 13 42 14 0.45Mean1994199419225133126596160.34’0.32’0.731111.24-,Cl)ci)z-c,a)0.8aDC.)C)o 0.6D0>-oZ 0.20Outside: Inside0Figure 4.3 Bald Eagle productivity (mean and SD) compared between samples of nests located adjacentto shorelines inside and outside of dioxin fishery closure areas on the British Columbia coast. Samplesizes 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, PCDFor 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 eachcorresponding territory (r2 = 0.128, p < 0.011, Figure 4.4b).112A2.5 -. 2- ••ci• .1A4D 0) AALI.C1- LZJs.c>AAI I 1111111 111111110.1 1 10 100B TEQs - WHO (ng/kg wet weight)2.5-Lci)ci)> U0U)E’i- .0+1 A2 A00— 111111 ‘‘I 1111111 11111111 10 100 1000DDE (ug/g wet weight)• E. Van. I. A Barkley Sd. ü Johnstone Str. A Low. Fraser Vafley• Powell R. • Clayoquot Sd. 0 Fraser Delta Langara I.Figure 4.4 Productivity at Bald Eagle nest sites on the British Columbia coast as a function ofcontaminant concentrations in plasma samples in nestlings raised in that territory, for: A) the log ofTEQs0,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 LangaraIsland.113Mean percent lipid in plasma regressed positively on mean productivity among sites(Figure 4.5).1.41.20a) Ia)0.8DC.)0oO.6c3)C0>-0.20Figure 4.5 Comparison of mean productivity of Bald Eagles at sites on the coast of British Columbia withthe 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 deliveriesAt surveyed nests in the Fraser Delta, mean daily prey deliveries were greater at, 3.5 thanat Baridey Sound nests, 2.4 deliveries per day; however the difference was not statisticallysignificant, likely due in part to small sample sizes in this pilot study.r2=O.423U0.01 log (% plasma lipid) 0.1114PCDD and PCDF levels in plasmaPlasma 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, OCDDwas often comparable or greater than 1 ,2,3,6,7,8-HxCDD, while in the Fraser delta, OCDDand 1,2,3,4,6,7,8-HpCDD were the dominant congeners. Most samples also containeddetectable amounts of 2,3 ,4,7,8-PnCDF.Because of the significant interaction with plasma lipid content, selected PCDDs andPCDFs are further presented as lipid-adjusted log-normalized mean values (Figure 4.6). Meanplasma TCDD concentrations were significantly higher at Powell River and East VancouverIsland than other sites. Mean concentrations of PnCDD, HxCDD and TCDF were also highestnear the pulp mill sites at Powell River and along east Vancouver Island; however, thedifferences 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 plasmaHighest concentrations of total PCBs were in samples from Powell River and eastVancouver Island (Table 4.4), which on a lipid-adjusted basis were significantly greater thanClayoquot Sound and the Fraser valley (data not shown). Mean concentrations of individualPCB 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 andeast Vancouver Island and were significantly different from Clayoquot Sound and the lowerFraser Valley.115Table4.3PCDD/PCDFlevels,geometricmeansand95%confidenceinterval(ng/kg, wetweight)inbloodplasmaof BaldEaglechicksfromthecoastofBritishColumbia,1993-94.LocationN2378123781236781234678OCDD2,3,7,81237823478234678OCDFTCDDPnCDDHxCDDHpCDDTCDFPnCDFPnCDFHxCDFFraserDelta50.070.230.451. River100.370.902.20.130.564.540.*0.12-1.10.01-0.190.23-0.760.23-4.10.03-0.120.02-0.340.05-0.080.01-0.18LangaraIsland50.,minimumdetectionlimit0.01-0.05ng/kg, wetweight.*-valuesallthesame2378-TCDD 2378-TCDF-Ua)ci,-U-UU)U)21.5ab bbnnm\\ii \‘• — -. ,. \Figure 4.6 Residue levels of selected PCDDs and PCDFs in plasma samples of Bald Eagle nestlingscollected on the British Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.3. Meansthat do not share the same lower case letter are significantly different (p <0.05).1 2378- PnCDD 23478- P, CD F1 23678-HCDD OCDD2.5 a- -117Table 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 BritishColumbia, 1993-94.Location N Total DDE trans- Oxychiordane Dieldrin Mirex HCBPCBs nonachlorFraser Delta 5 17.8 14.4 0.5 0.3 0.1 0.1 0.24.6-69.6 8-26 0.1-2.6 ND-1.8 ND-0.1 ND-0.3 0.1-0.8Lower Fraser 5 11.2 9 0.5 0.1 0.1 0.1 0.3Valley 6.4-19.7 4-20.3 0.3-0.7 ND-0.1 ND-0.1 ND-0.1 0.2-0.5East Vancouver 10 30 11 2 0.2 0.2 0.1 0.3Island 18.8-47.5 0.6-17.3 1.2-3.1 0.1-0.3 0.1-0.3 0.1-0.2 0.2-0.5Powell River 10 56 20.2 3 0.4 0.2 0.3 0.627-114 8.3-50 1.4-6.4 0.1-1.5 0.1-0.8 0.2-0.7 0.3-1.0Barkley Sound 8 20 21.1 1.3 0.1 0.1 0.1 0.314-28.5 6.9-64.5 0.8-2 ND-0.4 ND-0.1 0.1-0.3 0.2-0.6Clayoquot Sound 3 6.8 6.6 0.3 0.1 0.1 0.1 0.31.9-24.2 1.8-24 0.1-0.7 * * * 0.1-0.7Johnstone Strait 4 14.3 7.3 1.2 0.1 0.1 0.1 0.46.2-33 2.7-19.4 0.8-1.9 ND-0.1 ND-0.1 ND-0.2 0.2-0.5Langara Island 5 16.4 22.3 1.1 0.9 0.1 0.3 0.86.3-43 5.8-86 0.6-2.0 0.5-1.7 ND-0.4 0.1-1.4 0.3-2.2ND - Not detected, minimum detection limit 0.01-0.05 nglkg, wet weight.*- values all the sameIn 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). Highestlipid-adjusted mean concentrations of individual congeners were generally in samples fromPowell River or east Vancouver Island, although the highest mean concentrations of PCB 169were from Langara Island (Figure 4.7).Organochiorines in plasma.Highest mean organochlorine pesticide levels were in samples from the Strait of Georgiaregion, including the Fraser Delta and from Langara Island (Table 4.4). Most lipid-adjustedplasma OC levels did not differ significantly among sites. Mean oxychiordane levels weresignificantly greater at Langara Island than either Johnstone Strait or the lower Fraser Valley.118PCB 126 PCB 169Figure 4.7 Residue levels of selected PCBs in plasma samples of Bald Eagle nestlings collected on theBritish Columbia coast, 1993-1994. N sizes and error estimates are in Table 4.4. Means that do notshare the same lower case letter are significantly different (p <0.05).PCB 37 PCB 77. ‘,. . c,& q• c \G3—.119Table 4.5 Non-ortho PCB levels, geometric mean and 95% confidence interval (nglkg, wetweight) 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 189Fraser Delta 5 1.80 0.71 9.45 4.01 0.68 0.130.94-3.45 0.22-2.35 2.66-33.5 1.29-12.5 0.21-2.18 0.05-0.37Lower Fraser 5 1.01 0.59 5.63 2.5 0.42 0.09Valley 0.53-19.95 0.32-1.07 3.22-9.84 1.66-3.75 0.24-0.72 0.04-0.22East Vancouver 11 29.4 0.98 16.6 6.06 1.42 0.53Island 24.5-35.2 0.64-1.51 12.5-22.1 2.98-13.3 0.24-6.72 0.38-0.74Powell River 10 18.3 1.49 26.1 14.8 3.39 0.5813-25.7 0.82-2.71 13-52.3 7.01-31.3 1.88-6.12 0.23-1.49Barkley Sound 8 10.1 0.76 7.27 4.75 0.64 0.237.0-25.7 0.45-1.29 5.74-9.2 3.24-6.96 0.30-1.34 0.11-0.46Clayoquot Sound 3 9.84 0.26 5.77 0.98 0.52 0.166.31-15.4 0.01-6.02 0.89-37.5 ND-212 0.05-5.17 0.07-0.38Johnstone Strait 4 7.70 0.45 4.77 1.46 0.40 0.202.68-22,1 0.14-1.40 1.52-14.9 0.31-6.72 0.07-2.43 0.05-0.75Langara Island 5 0.81 0.51 5.04 6.29 2.52 0.100.31-2.13 0.14-1.84 1.33-19 1.69-23.5 0.71-8.97 0.03-0.27ND - Not detected, minimum detection limit 0.01-0.05 ng/kg, wet weight.Mean mirex concentrations were significantly greater at Langara Island and PowellRiver than the lower Fraser Valley.DiscussionHigher concentrations of chlorinated hydrocarbons in Bald Eagle nestlings from theStrait 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 sitesaround the strait, particularly the estuary of the Fraser river, was substantially higher than the0.7 young/occupied territory, necessary to sustain an eagle population (Sprunt et al. 1973). Incontrast, eagle productivity at the more remote reference sites was generally less than 0.7. Onlyat Langara Island at the north end of the Queen Charlotte archipelago, an area of highbiological productivity, was eagle breeding success comparable to the Fraser delta and the Strait120of Georgia. Using nestling plasma lipid content as a marker of body condition, food supply islikely the main factor limiting eagle productivity on the British Columbia coast. However, lowproductivity at a sample of eagle nests adjacent to the dioxin fishery-closure zone at Crofton isprobably not caused by differences in food availability.The geographic pattern of PCDDs and PCDFs in plasma is similar to that found in eagleeggs 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 theFraser delta samples likely reflect heavy past use of chlorophenolic wood preservatives in thatarea, and some contribution from combustion sources.There are few published data on PCDD and PCDF levels in avian plasma. Bloodsamples of osprey nestlings taken in 1992 downstream of a bleached-kraft pulp mill on theThompson River, in the interior of British Columbia, did not contain any lower chlorinateddioxins and furans (minimum detection limit = 0.5 ng/kg, wet weight); sample sizes weresmall, 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 thesame sites in 1991 contained relatively high concentrations of TCDD, TCDF, HpCDD andOCDD (Whitehead et al. 1993).In bald eagles, five of the six non-ortho compounds displayed a good correlation withplasma lipid content, while PCB 37 was only weakly correlated with plasma lipid. Ratios ofPCB 37 relative to other congeners were high in eagle plasma compared to eggs or liver. Highratios 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 otherstudies is unavailable. However, studies of human subjects have shown that, although absolutelevels on a lipid weight basis were much lower than those found in the eagle samples, OCDDwas the major congener present (Papke et a!. 1990). In humans, blood:adipose tissue ratios arehighest for OCDD compared to other PCDDs and PCDFs (Schechter et al. 1990). As we found121with eagles, OCDD did not partition with lipid in human blood; it is believed to bind primarilyto serum protein components (Patterson et al. 1989).Published data on total PCBs and DDE in avian plasma samples is available from anumber of studies. Mean concentrations of PCBs and DDE in plasma of nestling Bald Eaglesfrom 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 PowellRiver and east Vancouver Island nests. Meanwhile, PCB and DDE levels in eggs were aboutthree-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 thelower Columbia; therefore, the low levels of PCBs and DDE in those samples may reflect lowplasma lipid levels.Geometric mean levels of DDE and PCBs (wet weight) in eagle plasma samplescollected between 1987 and 1993 from less contaminated areas of the Great Lakes werecomparable 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 theGreat Lakes, 20-25 ng/kg, were comparable to data for the Strait of Georgia and LangaraIsland. Eaglets from Lake Michigan had somewhat higher levels, 35 ng/kg, DDE, than othersites. Mean levels of total PCBs in nestling eagle blood samples from the Great Lakesshoreline were two-fold (Lake Superior) to four-fold (Lake Erie) higher than Strait of Georgiasamples. Maine eagle blood samples, 1991-1992, particularly from estuarine sites, had up to150 ng/kg DDE and 1,250 ng/kg total PCBs (Welch 1994). However, plasma lipid data werealso not reported for either the Great Lakes or Maine samples. The potential influence ofgeographic variation in plasma lipids on contaminant levels is particularly relevant for someGreat Lakes samples, as Dykstra (1994) determined that low food availability was the maincause of poor breeding success at the Lake Superior nests, compared to those inland. This wasreflected in lower rates of prey delivery to nests, greater time spent away from the nests byadults and increased time spent by nestlings sleeping and resting. Concentrations of DDE, butnot PCBs, in nestling plasma samples from Lake Superior also regressed negatively on mean 5-122year productivity at the respective territories, indicating that DDE may still have been a factorcontributing to low productivity.Low eagle productivity at certain areas of the British Columbia coast, such as Barkleyand Clayoquot Sounds, Johnstone Strait and South Moresby may also be caused by low foodavailability. Mean plasma lipids were significantly lower in nestlings from those sites,indicating chicks in poorer body condition. The significant association among sites betweenproductivity and mean percent plasma lipids also suggests that in productive areas, chicks arefed 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 meanfat content than those found at shoreline nests (Kozie and Anderson 1991). The pilot study onprey deliveries failed to show a significant difference between samples of nests in the FraserDelta and Barkley Sound, although there were significant differences between those sites inboth mean 3-productivity and percent plasma lipids in nestlings. However, because of logisticaldifficulties in observing nests at more remote areas of the coast, where productivity isparticularly low, observations in Barkley Sound were made at nests which tended to be moreaccessible and to have higher productivity.Food supply during breeding is a major factor affecting avian productivity, includingraptors (Newton, 1980; Gardarsson and Einarsson 1994). In addition to Dykstra’s (1994) studyof eagles, Shutt (1994) related breeding failure and poor body condition of both herring gullchicks and adults to lack of food at Lake Superior breeding colonies. Prey availability wascritical 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). Aminimum food supply was required for successful breeding of wedge-tailed eagles (Aquilaaudax) in Australia, while Hansen’s (1987) experiment showed that Bald Eagle nesting andfledging success could be increased by providing additional food.Bald Eagle breeding densities in Saskatchewan were related to availability of key preyspecies, which correlated with primary productivity (Dzus and Gerrard 1993). Fish eatingbirds, particularly gulls, are important prey species to north west eagles (Knight et al. 1990).123On the west coast of Vancouver Island, colony sizes and breeding success were lower for gullsand cormorants (Vermeer et al. 1992) than the Strait of Georgia with its more stable foodregime (Vermeer et a!. 1989). The steep fjord-like topography of the shoreline and the islandsof the west coast of Vancouver Island, Johnstone Strait and Moresby Island also limits preyavailability and foraging opportunities, compared to the beaches and tidal mudflats of the Straitof Georgia, which harbour abundant bird populations (Vermeer 1983). Food concentratedalong the highly productive La Perouse Bank, to the west of Barkley Sound is beyond the reachof Bald Eagles. Langara Island is the only site outside the Georgia basin with relatively higheagle productivity. This island lies at the bottom of the Alaska gyre, an area of summerupwelling (Thomson 1981), which creates high marine productivity, evident by a rich fauna ofsalmonids, seabirds and cetaceans.Low eagle productivity in Barldey and Clayoquot Sounds and Johnstone Strait ischaracterized by a high incidence of failed nesting attempts. Many nests had incubating adultsduring the activity flight, but were empty during the productivity flight. Without nestobservations throughout the breeding cycle, we cannot determine at what stage those attemptsfailed, although some nests certainly failed during incubation, as we often observed nests withabandoned 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 acriteria for contaminant impact on an eagle population (Colborn 1991). The fledging ratio wasas 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 nestinitiation 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 Morgan1992), which are not available in May and June and are not adequately replaced by other fooditems.With the present data, it is difficult to determine why eagle productivity is low in theCrofton area. In contrast to Clayoquot Sound and other areas, eagle nesting near Croftonshould not be food stressed. A number of the Crofton area nests are situated on small islands124(Shoal and Willy Islands), virtually in the estuary of the Chemainus River. Numerouswaterbirds, including flocks of several hundred White-winged Scoters (Melanitta fusca), feedingon the abundant shellfish, are present during the breeding season. Eagle productivity is alsohigh 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 andalso 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 inthe late 1 980s (Elliott et al. 1989a; Sanderson et al. 1994a). Based on the extrapolation inFigure 4.8, levels of 2,3,7,8-TCDD and other chemicals would likely have been even higher ineagles than herons. PCDDs and PCDFs in eagle eggs collected in 1990 and 1991 and on alipid-adjusted basis in the one eagle plasma sample from Crofton were comparable to thosefrom Powell River, yet a reduction in mean productivity in the dioxin fishery closure area wasnot found. However, for pragmatic reasons, fishery closures from persistent pollutants such asdioxins must be defined over broad areas, even though there are wide gradients incontamination within the zones (Harding and Pomeroy 1990). For example, higher PCDD andPCDF concentrations were consistently found in invertebrates collected to the north than to thesouth of the Powell River mill (Dwernychuck et al. 1994). This corresponds, perhapscoincidentally, with poor productivity at the four eagle nests immediately north of that mill. AtCrofton, eagle productivity was also particularly poor at Shoal and Willy Islands, the nestsclosest 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 inthe 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 andPCDF concentrations in fish eating birds at Crofton had decreased by an order of magnitudefrom the high levels of the late 1980s (Whitehead et a!. 1 992a; Figure 4.8). The rapid declineof 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 inputs125would be more quickly apparent. Sample sizes are small, nevertheless, mean PCDD/PCDFlevels in eagle eggs decreased between 1990 and 1992 at Crofton and Nanaimo, althoughpossibly at a slower rate than in herons and cormorants. As larger animals feeding at a highertrophic level, clearence of TCDD and other compounds may occur more slowly in eagles.3004-.-cci)— 100coI-.C)C” 500Figure 4.8 Trends in 2,3,7, 8-TCDD in eggs of eagles, herons and cormorants at Crofton, BritishColumbia. The likely trend in eagles is extrapolated back to 1987, based on the mean 2,3,7,8-TCDDratio of eagles:herons, 1990-1992.Assuming that poor productivity at Crofton is contaminant-related, it is also conceivablethat some adult eagles suffer chronic reproductive impairment due to past high PCDD/PCDFexposure 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 significantextrapolatedIi Heron-*- Eagle•EagleCormorant1987 1988 1989 1990 1991 1992 1993126reproductive impairment, but no apparent health problems (Bowman et al. 1989). Male ratsexposed both in utero and lactationally to as little as 0.064 ug/kg TCDD via maternal dosinghad 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 againstreduced fertility; other animals, for example man, which have a lower critical sperm volumecould be more affected. Although similar studies have not been done in birds, extrapolationfrom 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 appearexternally normal, but have reduced capability to reproduce.The potential for wildlife exposure to other chlorinated compounds of pulp mill originhas 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 theBritish Columbia coast, including Nanaimo and Powell River, contained from 0.5 to 5 ag/kgpentachiorophenol and traces (<1.0-3.3 tg/kg) of di- and tetrachloroquaiacols (CanadianWildlife Service 1994). Those compounds are considered indicative of bleached-kraft pulp millcontamination of receiving water, sediments and biota (Dwernychuck et al. 1994). Release oforganochlorines (AOX) in pulp mill effluents has decreased significantly since the installation ofsecondary treatment systems at all British Columbia coastal pulp mills (see Table 3.9). Studiesof fish collected from both bleached-kraft and non-kraft pulp mills in eastern Canada have alsoreported the presence of an unidentified factor(s) present in effluents of both mill types thatinduce CYP 1A and affect reproductive hormone levels (Carey et al. 1992). Presence of thatfactor was independent of either chlorine bleaching or secondary treatment. However, bothchiorophenols and chloroquaiacols and the unidentified factor appear to be cleared fairly rapidlyin fish, ie. within two weeks; therefore, it seems unlikely that Bald Eagles would accumulatesignificant amounts of this class of chemicals.Alternatively, the low productivity measured in nests adjacent to the dioxin closure areaat Crofton may be explained as either a sampling artifact or the result of ecological factors thathave not been identified. Because of the cost of helicopter surveys and the difficulty in locating127nests, the sample may not be representative of the area, implying that some productive nestswere not surveyed each year. However, the probability of overlooking a significant number ofproductive versus unsuccessful nests in the Crofton area should be no different than in otherareas. 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 Croftonarea likely receives greater attention. Quality of nesting habitat near Crofton appearscomparable or better than most areas of the survey route; there are large numbers of suitablenest trees in relatively undisturbed areas and only limited activity.Currently, I am unable to determine the cause for poor eagle productivity at nestsadjacent to the dioxin fishery closure area at Crofton. It is probably not caused by low foodsupply. 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 ruledout. Further intensive work in this area is necessary to confirm the results and investigatecauses.My conclusions agree with those of Dykstra (1994) that the role of food supply needs tobe factored into any studies of the effects of contaminants or other habitat quality variable instudies of Bald Eagles. Measurement of plasma lipids may provide a useful surrogate forenergetic status of eagle nestlings. Further work is required to determine the causes of theapparent low productivity in the Crofton area.AcknowledgmentsA special thanks to Ian Moul and George Compton for all of their support andassistance in the field. Chris Coker and Brenda Li-Pak-Tong are thanked for their field work onthe prey deliveries. Ron McLaughlin (MacMillan-Bloedel) and Ken Stenerson (Scott Paper) arealso thanked for personal and corporate financial support with helicopter surveys. Working inthe laboratory of Dr. Ross Norstrom, Mary Simon did the PCDD/PCDF and non-ortho PCBanalysis; Henry Won did the organochiorine and plasma lipid analyses.128Appendix4-1.Productivity,%lipidandselectedchlorinatedhydrocarbonresiduelevelsinplasmaofindividual BaldEaglechickscollectedfromthecoastof BritishColumbia,1993-94Residuelevels(wetweightbasis)Location,NestProduct.1LipidTEQ5-2378-12378-123678-OCDD2378-23478-PCBPCBPCBPCBPCBDDETotal(chicks/%WHOTCDDPnCDDHxCDDTCDFPnCDF-77-126-118-105-153PCBsactiveterr.)(ng/kg) (wetweight)(pg/kg)(wetweight)FraserRiverDeltaAlaksen10.1062.39NDNDND26.82ND0.1520.867.542.90.75.919.134.5StevestonI1.50.1162.510.210.660.570.120.880.225.810.\ValljsPt.1.670.1241.480.160.480.820.782.540.2411.63.561.,cont...LundGibson’sBGibson’sB.P.BarkleySoundSantaMariaIsI.NumukamisAssitsIsI.AssitsIsi.HissinPt.MercantileCr.MercantileCr.SalmonBeachClayoquotSoundWhitePineC.GibsonCoveGibsonCoveJohnstoneStraitPlumperIsi.Fire1st.Cracroft Pt.CracroftPt.LangaraIslandLucy1Lucy2GuillemotMargueriteCabinBay20.0422.660.2810.0562.560.2210.0682.970.201.331.08ND2ND0.50.10.88ND10.0141.11ND10.0180.43ND0.670.021.0ND10.0361.31ND10.0520.62ND0.330.0081.82ND0.50.0120.43ND10.0060.80ND10.0610.72ND0.50.0250.46ND10.0980.45ND1.330.0610.91ND1.330.0991.41ND10.10.93ND20.1350.34ND10.3654.20ND20.1030.98ND10.1851.05ND0.892.10.595.500.711.940.622.860.783.010.843.15NDND0.46NDNDND1.18NDND0.481.00NDNDND0.57NDNDND0.550.61NDND0.40NDNDND0.20NDNDND0.83NDNDND0.37NDNDND1.860.41NDND0.37ND0.260.350.710.160.720.310.560.302.020.760.392.020.2910.36.570.2411.49.730.1217.111.6ND6.535.49ND6.152.69ND126.00ND6.212.89ND8.645.39ND5.535.68ND5.693.18ND9.6110.4ND2.520.08ND113.28ND6.903.54ND3.021.11ND3.100.570.,NestProduct.’LipidTEQs-2378-12378-123678-OCDD2378-23478-PCBPCBPCBPCBPCBDDETotal(chicks!%WHOTCDDPnCDDHxCDDTCDFPnCDF-77-126-118-105-153PCBsactiveterr.)(ng/kg)(wetweight)Q.g/kg)(wetweight)ND0.150.260.28 ND0.0350.36 0.1NDND0.180.1ND0.330.,1993,1994),exceptLangaraIsland,1994only.2ND-notdetected(minimumdetectionlimit,PCDDs!PCDFs,0.01ng/kg)GENERAL SUMMARY AN]) CONCLUSIONSThe overall purpose of this research was to investigate the toxic hazard posed bychlorinated hydrocarbon contaminants to Bald Eagle populations breeding and wintering in theStrait of Georgia area of British Columbia. The research tested a general hypothesis that astop predators in marine and estuarine systems, Bald Eagles would bioaccumulate high levels ofchlorinated hydrocarbons. Consequent to high exposure and as ensuing hypotheses, bothsurvival and reproduction would be adversely affected. These hypotheses were tested by anumber of field and laboratory studies.Adult exposure and mortality studyThe investigation began by collecting samples from the large number of Bald Eaglesfound dead or dying each year in British Columbia. Many sick birds and carcasses are turnedin by concerned members of the public or individuals seeking taxidermy permits. Of 484eagles examined in this study, 59 found between 1988 and 1993 were selected fororganochlorine analysis. Of those birds 5% had liver residue levels of DDE and chlordanerelated compounds diagnostic of acute toxicity. Even this percentage is surprising and the longterm persistence of OC pesticides and continued input from atmospheric sources and migratorybirds is indicated. These findings reinforce the need for vigilance in both the enforcement ofcurrent regulations and scrutiny of new commercial chemicals.Of 19 Bald Eagles further analyzed for PCDDs, PCDFs and non-ortho PCBs, livers offour birds (21 %) contained TCDD-toxic equivalents (TEQsWHO) > 1 ,000 ng/kg. Birds withhigh PCDD and PCDF levels were found in the vicinity of bleached-kraft pulp mills. Mostbird with elevated chlorinated hydrocarbon levels were in poor body condition indicating lipidand contaminant mobilization. Based on high TCDD/TCDF ratios in at least three eagles,hepatic CYP1A enzymes were likely induced131Study of biological effects in eagle chicksIn order to assess embryotoxic effects of chlorinated hydrocarbons in Bald Eagles, eggswere collected within an exposure gradient and incubated in the laboratory. Yolk sacs ofchicks collected near bleached-kraft pulp mills contained higher concentrations of PCDDs andPCDFs, although there were no significant effects on hatching success or morphologicalendpoints. Hepatic CYP1A levels were induced in chicks from pulp mill sites and correlatedsignificantly with 2,3,7,8-TCDD, 2,3,7,8-TCDF and TEQ5WHO in yolk sacs. TEQsWHOassociated 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 CYP1Alevels 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 lesssensitive to TCDD-like toxicity than the chicken embryo. The LD50 for the chicken embryo isabout 250 ng/kg (Alired and Strange 1977; Janz 1995), similar to the 210 ng/kg TEQ5WHOmeasured in eagle eggs without apparent effects on hatching success or histological,morphological and some biochemical endpoints. At 100 ng/kg TEQ5WHO in eagles, nosignificant CYP1A induction occurred, while two-fold AHH induction was measured at 10ng/kg injected into chicken eggs. With regard to CYP1A induction, Bald Eagles appearsomewhat more sensitive than Great Blue Herons and Double-crested Corinorants. In heronchicks, EROD activity was significantly induced (six-fold) at about 440 ng/kg, but not at 250ng/kg TEQsWHO (Sanderson et at. 1992a). In cormorant chicks, significant eight-fold ERODinduction occurred at 550 ng/kg but not at 217 ng/kg TEQsWHO (Sanderson et at. 1992b).Bioaccumulation studyFor this study, fresh Bald Eagle eggs were collected at a variety of locations on theBritish Columbia coast, representing different chlorinated hydrocarbon exposure scenarios. Adata base of contaminant levels in Bald Eagle prey items, principally from pulp mill sites in theStrait of Georgia, was compiled using existing data. A simple model was used to examine therelationships between contaminant levels in Bald Eagles and their foodchain. The modelaccurately predicted 2,3,7, 8-TCDD levels in Bald Eagle eggs and was reasonably accurate for132other compounds. The model was used to estimate 2,3,7,8-TCDD and TEQWHO levels inforage 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 LOELsgenerated 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 assite specific guidelines for the Strait of Georgia. The same approach was used to derive similarvalues for total PCBs, suggested to be 0.01 ng/kg in forage fish and 0.3 ng/kg in fish-eatingbirds.Productivity studyThe research described in the previous studies addressed acute toxicity of adult birdsand determination of critical levels in eggs, associated with embryotoxicity. During the fourthpart of this work, Bald Eagle breeding success was measured for up to three years at eight siteson the British Columbia coast. Because of annual variability, assessment of breeding success inBald Eagles requires a minimum of three years data. Studies elsewhere showed thatreproduction in birds of prey is a critical endpoint affected by chlorinated hydrocarbons in birdsof prey (Newton 1979). In order, to relate productivity of individual nests to contaminantexposure, blood samples were taken from nestlings, to minimize the impact of samplecollection.Bald Eagle productivity was highest overall at nests in the lower Fraser River valley anddelta, while at four of five reference areas, selected for their remoteness from direct industrialinput of pollutants, productivity was less than the level of 0.75 young/occupied nest considerednecessary to sustain an eagle population. Only at Langara Island, an area of very highbiological productivity, was eagle breeding success comparable to the Fraser valley and mostStrait of Georgia sites. At the reference locations, low breeding success is likely due to lowfood availability, particularly during chick rearing. This was supported by finding ofsignificantly lower nestling plasma lipid content at those sites and a significant positiveregression between mean nestling plasma lipid levels and mean productivity among sites.Despite higher plasma levels of PCDDs and PCDFs, Bald Eagle productivity was relatively133high at nests near two pulp mill areas on the Strait of Georgia (Nanaimo, Powell River); atthose sites, no significant differences in mean productivity occurred at nests adjacent toPCDD/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 areais rich in marine life. Data from biomonitoring studies of fish-eating birds showed that PCDDand PCDF levels in local food chains fell dramatically between 1989 and 1992, subsequent tomodifications to the bleaching process employed by the mill and a ban on chlorophenolic antisapstain usage. Alternative hypotheses to explain the low eagle productivity in the areainclude: first, the presence of a substance released in the mill effluents, that has contaminatedlocal 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 orduring early development of the reproductive system, to elevated levels of 2,3,7, 8-TCDD andrelated chemicals. This last hypothesis requires further study and testing.In conclusion, during the recent past reproduction of Bald Eagles in the Strait ofGeorgia was probably reduced by exposure to significant chlorinated hydrocarbon levels,particularly DDE. Increases in nest occupancy reported for the southern Gulf Island betweenthe early 1 970s and late 1 980s is typical of the population recoveries documented in many areasof North America and attributed to declining environmental DDE contamination. During the1 980s and at least until the early 1 990s, eagles breeding and wintering near bleached-kraft pulpmills on the British Columbia coast were exposed to relatively high levels of PCDDs andPCDFs. At Crofton, the effects of this pollution may be continuing, although the mechanism isobscure. At other areas of the British Columbia coast, Bald Eagle breeding success appears tobe influenced mainly by food supply.The effects of chlorinated hydrocarbons on Bald Eagle populations have to beconsidered in the context of multiple stresses, both chemical and otherwise, on survival andreproduction. Lead poisoning from ingestion of spent shot is a major cause of death for British134Columbia Bald Eagles; many eagles have also been sublethally poisoned, with probableconsequences for longterm health and survival. In some areas, such as the Lower FraserValley, pesticide poisoning is a major cause of mortality. Bald eagles are also vulnerable toloss and disturbance of nest sites. Given these factors, and the growing human population ofthe 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 andeffects, it may be more cost-effective to monitor colonial fish-eating birds.Future DirectionsEcotoxicological work on Bald Eagles should further investigate the low reproductiverate measured at Crofton. All nests in the area from Cowichan Bay to Thetis Island should belocated. A sample of nests including those nearest the mill, should be intensively observed todetermine breeding behaviour and the timing of nest failures. Toxicological hypotheses can betested by trapping adult eagles on their breeding territories to obtain blood samples forcontaminant analyses and measurement of reproductive and thyroid hormones. Similar studiesare required at a reference site, such as Barkley Sound, and also possibly at another pulp millsite, either Nanaimo or Powell River, depending on available funding.Laboratory research using in vitro cell cultures of primary hepatocytes from eagles orother raptors would provide data on sensitivity of raptors compared to more commonly studiedlaboratory species and sentinel species such as Herring Gulls. 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