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ENVIRONMENTAL CONTAMINANTS IN BALD EAGLESON THE COAST OF BRITISH COLUMBIA:EXPOSURE AND BIOLOGICAL EFFECTS
by
JOHN EDWARD ELLIOTT
B.Sc., Carleton University, Ottawa, 1979M.Sc., The University of Ottawa, 1989
A THESIS IN PARTIAL FULFILMENT OF THE REQUIREMENTS FORTHE DEGREE OF DOCTOR OF PHILOSOPHY
in
THE FACULTY OF GRADUATE STUDIES
THE FACULTY OF AGRICULTURE
Department of Animal Science
We accept this thesis as/conforming to he required standard
The University of British Columbia
October, 1995c John Edward Elliott , 1995
In presenting this thesis in partial fulfilment of the requirements for an advanced
degree at the University of British Columbia, I agree that the Library shall make it
freely available for reference and study. I further agree that permission for extensive
copying of this thesis for scholarly purposes may be granted by the head of my
department or by his or her representatives, It is understood that copying or
publication of this thesis for financial gain shall not be allowed without my written
permission.
(Signature)
Department of
________________
The University of British ColumbiaVancouver, Canada
Date
DE-6 (2188)
Abstract
Attracted by abundant food and nesting sites, a large (about 4,000 pairs) Bald Eagle
(Haliaeetus leucocephalus) population breeds and winters around the Strait of Georgia on the
Pacific coast of Canada. Eagle habitat has been extensively modified by logging and
waterfront development, while industrial effluents have contaminated food chains. Until
recently, most pulp mills on the British Columbia coast used elemental chlorine bleaching and
did not secondarily treat effluents, thus releasing chlorine containing chemicals, particularly
polychiorinated dibenzo-p-dioxins (PCDDs) and polychiorinated dibenzofurans into the local
environment. As top predators, Bald Eagles are exposed to elevated levels of PCDDs, PCDFs
and the chemically related polychiorinated biphenyls (PCB5) and organochlorine pesticides.
This thesis addressed spatial and temporal trends in chlorinated hydrocarbon exposure of Bald
Eagles and toxicological consequences at treatment populations near pulp mills in the Strait of
Georgia and in industrial areas of the Fraser River delta, and at reference areas on west coast
Vancouver Island, Johnstone Strait and the Queen Charlotte Islands.
Initial research during 1990-199 1 focused on eagles found dead or dying and determined
that the majority of birds tested had low liver organochlorine levels (< 5 mg/kg, N =59). A
small proportion (< 5 %) had levels of DDE, polychiorinated biphenyls (PCBs) and chiordane
related compounds potentially diagnostic of acute poisoning. A larger proportion had
PCDD/PCDF levels of possible concern; four of 19 eagles tested had TEQ5WHO > 1000 rig/kg,
all of which were adults in poor body condition found near pulp mills during the breeding
season.
In 1992, in ovo exposure to a gradient of environmental contaminants was studied by
collecting eggs (N = 25) for laboratory incubation. Hatching success was not significantly
different between eggs from pulp mill versus reference sites. A hepatic cytochrome P450 1A
(CYP1A) cross-reactive protein was induced sixfold in chicks from near a pulp mill at Powell
River compared to those from a reference site (p < 0.05); hepatic EROD and BROD activities
11
were also significantly higher in chicks from pulp mill nests compared to reference sites
(p <0.0005 and p < 0.02, respectively). Residual yolk sacs from near pulp mill sites had
greater concentrations of 2,3,7,8-substituted PCDDs and PCDFs than reference areas. The
hepatic CYP1A cross-reactive protein and EROD and BROD activities were positively
correlated with concentrations of 2,3,7, 8-TCDD, 2,3,7, 8-TCDF and toxic equivalents (TEQs)
in yolk sacs. No concentration-related effects on histological or morphological parameters were
found. Using hepatic CYP1A expression as a biomarker, a no-observed-effect-level (NOEL) of
100 ng/kg and a lowest-observed-effect-level (LOEL) of 210 ng/kg TEQ5WHO on a whole egg
(wet weight basis) were suggested for Bald Eagle chicks.
To investigate spatial patterns, trends and sources of contaminants to Bald Eagles, eggs
were also collected during incubation, 1990-92, at the treatment and reference areas and
analyzed for chlorinated hydrocarbons. Data on Bald Eagle avian and fish prey items from the
study area were compiled and used as input to a bioaccumulation model. The model accurately
predicted 2,3,7, 8-TCDD levels in eagle eggs based on dietary concentrations, but was less
accurate for other PCDDs and PCDFs. Using the LOEL levels in eagle eggs derived from the
above study, concentrations of 2,3,7,8-TCDD in prey fish of 0.5 ng/kg and in fish-eating birds
of 10 ng/kg are suggested as ecosystem guidelines to avoid TCDD-like toxicity in Bald Eagles.
At all of the treatment and reference areas, Bald Eagle breeding success was measured
for three years and blood samples of nestling eagles were collected for contaminant analysis.
Average 3-year eagle productivity was high at most Strait of Georgia study sites, but was
significantly lower at reference sites. Using nestling plasma lipid content as a marker of body
condition, food supply appeared to be the main factor limiting eagle productivity on the British
Columbia coast. However, at a sample of eagle nests adjacent to the dioxin fishery-closure
zone near the pulp mill at Crofton, low productivity was probably not caused by low food
availability. The cause of the low reproductive rate at Crofton has not been determined;
however, a toxicological explanation has not been ruled out.
111
Key Words: Bald Eagle, bioaccumulation, CYP 1A, mortality, reproductive rate, 2,3,7,8-
tetrachlorodibenzo—p-dioxin
iv
Table of Contents
Abstract
Table of Contents
List of Tables
List of Figures.
List of Appendices
Abbreviations
Acknowledgements
General Introduction
Hypotheses and Objectives
Overview of the Thesis
Chapter 1 Chlorinated hydrocarbon liver levels and autopsyEagles found dead or debilitated, 1989-93
Materials and MethodsResultsDiscussion
Chapter 2 Biological effects of chlorinated hydrocarbons in
Materials and MethodsResultsDiscussion
Chapter 3 Bioaccumulation of chlorinated hydrocarbons and mercuryin eggs and prey of Bald Eagles
Materials and MethodsResults
Page11
vii
• ix
xii
xlii
xiv
17
• 18
data
Bald
for Bald
Eagle chicks
• . 19
• . 1923
• . 32
• . 41
• . 41• . 48
60
70
7077
Discussion 89
V
Page
Chapter 4 Influence of contaminants and food supply on Bald Eagle productivity 102
Materials and Methods 103Results 110Discussion 120
General Summary and Conclusions 131
References 136
vi
List of Tables
Page
Table 1.1 Organochlorine residue levels, geometric mean ± 95% confidence intervals,in livers of Bald Eagles found dead on the coast of British Columbia,1988 - 1993 28
Table 1.2 Non-ortho and mono-ortho PCBs in Bald Eagle livers collected from BritishColumbia (ng/kg, wet weight) 29
Table 1.3 Concentrations of select PCDDs and PCDFs in Bald Eagle livers collected fromthe south coast of British Columbia (ng/kg, wet weight) 30
Table 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) 31
Table 2.1 PCDD and PCDF concentrations (nglkg, lipid weight basis) in yolk sacs ofBald Eagle chicks collected in 1992 from British Columbia 51
Table 2.2 Concentrations of non-ortho PCB congeners in yolk sacs of Bald Eagleembroys collected in 1992 from British Columbia 52
Table 2.3 Organochlorine pesticide concentrations (geometric means 95% confidenceintervals, range in brackets) in yolk sacs of Bald Eagle chicks collectedin 1992 from British Columbia 53
Table 2.4 Outcome of artificial incubation of Bald Eagle eggs collected from BritishColumbia, 1992 54
Table 2.5 Histological examination of immune system tissues in Bald Eagle chicks(mean ± SD) 55
Table 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) 56
Table 2.7 Concentration-effect relationships between biochemical and morphologicalmeasurements with chlorinated hydrocarbon yolk sac levels in Bald Eagle chicks 59
Table 2.8 Comparison of regression (r2) values of some hepatic biochemical parameterson TEQs derived from three sets of toxic equivalence factors (TEF5) 60
Table 3.1 Mean PCDD/PCDF lvels (ng/kg, wet weight) in fish collected near three pulpmills on the Strait of Georgia, British Columbia 75
vi’
Page
Table 3.2 PCDD and PCDF levels (ng/kg, wet weight) in waterbird and seabird speciesfrom the British Columbia coast 76
Table 3.3 (PCDD) and (PCDF) residue levels (wet weight basis) in Bald Eagle eggsfrom British Columbia, 1990 - 1992 78
Table 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) 80
Table 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) 81
Table 3.6 Non-ortho PCBs in Bald Eagle eggs (ng/kg, wet weight) collected from BritishColumbia, 1992 84
Table 3.7 Eggshell thickness data, mean ± SD, (range in brackets) for Bald Eagles collectedfrom British Columbia, 1990-1992 85
Table 3.8 A simulation of PCDD/PCDF levels in Bald Eagle eggs at Crofton, 1990,based on concentrations in the diet 86
Table 3.9 Characterization of British Columbia pulp mills discussed in this paper . . . . 101
Table 4.1 Correlation Matrix (r values) for percent plasma lipid and selected hydrocarbonin Bald Eagle nestlings from British Columbia, 1993-94 109
Table 4.2 Nest success and production of young for Bald Eagles at nine study areas onthe British Columbia coast (1992-94) 111
Table 4.3 PCDD/PCDF levels, geometric means and 95% confidence interval (ng/kg,wet weight) in blood plasma of Bald Eagle chicks from the coast of BritishColumbia, 1993-94 116
Table 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, 1993-94 118
Table 4.5 Non-ortho PCB levels, geometric mean and 95% confidence interval (ng/kg,wet weight) in blood plasma of Bald Eagle chicks from the coast of BritishColumbia, 1993-94 120
vi”
List of FiguresPage
Figure 1. Molecular structure and position numbering of polychiorinated dibenzo-pdioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs) 2
Figure 2. Molecular structure of the major organochiorine pesticides 3
Figure 3. Molecular mechanism proposed for TCDD and related chemicals 7
Figure 1.1 Locations of Bald Eagles collected from British Columbia, 1989-93, andanalyzed for chlorinated hydrocarbons (N = 59) 20
Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed compared to the completeset of birds received 24
Figure 1.3 Numbers of eagles showing different DDE and PCBs in livers (N =59) 24
Figure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month 25
Figure 1.5 DDE and PCB residue levels in relation to body condition 26
Figure 1.6 PCB congeners in Bald Eagle livers expressed as percent of total PCBscompared for birds in good and poor body condition (N =9, for each group) . . 36
Figure 2.1 Locations where Bald Eagle eggs were collected for artificial incubation 42
Figure 2.2 Residue levels of major PCDDs and PCDFs in yolk sacs of Bald 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) 49
Figure 2.3 PCB congeners in yolk sacs of Bald Eagle chicks from British Columbia, 1992,expressed as percent of total PCBs. Values represent means of two to eightanalyses per collection site. Isomers are identified according to their IUPACnumber 50
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 58
lx
Page
Figure 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) . . 66
Figure 3.1 Locations where Bald Eagle eggs were collected for analysis 71
Figure 3.2 PCB congeners in Bald Eagle from British Columbia, 1990-1992, expressed aspercent of total PCBs. Values represent means of three to eight analysesper collection site. Congeners are identified according to their IUPAC number 82
Figure 3.3 Plot of the first and second principle components (PCi and PC2). 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 83
Figure 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) 84
Figure 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 88
Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections.At Langara Island, the survey circumsribed the coastline of the island 104
Figure 4.2 Bald Eagle nest sites and dioxin fishery closure areas at Powell River,Nanaimo and Crofton 106
Figure 4.3 Bald Eagle productivity compared between samples of nest located adjacentto shorelines inside and outside of dioxin fishery closure areas on the BritishColumbia coast 112
x
Page
Figure 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 113
Figure 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 114
Figure 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) 117
Figure 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) 119
Figure 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 120
xl
List of Appendices
Page
Appendix 1-1 Organochiorine pesticide and PCB levels in Bald Eagle liverscollected from British Columbia (mg/kg wet wt.) 38
Appendix 2-1 Selected morphological measurements in Bald Eagle chickscollected in 1992 from British Columbia 69
Appendix 4-1 Productivity, % lipid and selected chlorinated hydrocarbonresidue levels in plasma of individual Bald Eagle chicks collectedfrom the coast of British Columbia, 1993-94 128
xii
Abbreviations
Ah aryl hydrocarbon NWRC National Wildlife ResearchCentre
AHH aryl hydrocarbonhydroxylase OC Organochiorine pesticide
ANCOVA analysis of covariance PCA principle component analysis
ANOVA analysis of variance PCB polychiorinated biphenyl
BMF biomagnification factor PCDD polychiorinated dibenzo-pdioxin
BROD benzyloxyresorufin 0-dealkylase PCDF polychiorinated dibenzofuran
CWS Canadian Wildlife Service PWRC Pacific Wildlife ResearchCentre
CYP1A cytochrome P450 1ASAS Trademark, SAS Institute
CYP2B 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 dibenzofuran
EROD ethoxyresorufin 0-deethylase TEF toxic equivalent factor
GLEMEDS Great Lakes embryo TEQ TCDD toxic equivalentmortality edema anddeformities syndrome WHO World Health Organization
HCB hexachlorocyclobenzene
HCH hexachiorocyclohexane
LOEL lowest-observed-effect-level
NOEL no-observed-effect-level
xl”
Acknowledgements
I would like to thank my supervisory committee, Kim Cheng, Gail Beliward, Ross
Norstrom and Tom Sullivan for overall guidance and support. I would like to acknowledge the
financial and personal support of the Canadian Wildlife Service, and to personally thank Steve
Wetmore at the Pacific Wildlife Research Centre and Keith Marshall at the National Wildlife
Research Centre for their advice and support over the years.
A project of this sort depends on the assistance of a great many people. Specific
contributions are acknowledged at the end of each chapter, however, the support of a number
of people deserves special consideration: Ian Moul was a valuable co-worker in virtually all
phases of the field work; George Compton contributed his considerable tree climbing and
bush-whacking skills. Mary Simon, Henry Won and Suzanne Trudeau are thanked for their
work on the chemistry and biochemistry, Ken Langelier was a great help in the wildlife health
aspects and suggested the initial work on Bald Eagles. I am very grateful to Laurie Wilson for
the many technical and scientific roles she undertook for me. Shelagh Bucknell and Pam
Whitehead are thanked for their assistance and patience in typing of tables and preparation of
figures, respectively.
I would also like to acknowledge my friends and coworkers both at UBC and CWS for
making this PhD experience more rewarding and enjoyable.
I also wish to thank my parents for imparting a sense of what is important in life. Most
importantly, I am most grateful to the patience and support of my wife Christine and my
children, Kyle, Siobhan, Frazer and Alicia.
xiv
Introduction
Pollution of the environment by toxic substances has become a global problem with
ecological, economic and political consequences. Chlorinated hydrocarbons such as
polychiorinated dibenzo-p-dioxins (PCDDs), polychiorinated dibenzofurans (PCDFs),
polychiorinated biphenyls (PCBs) and DDT (1,1,1 -trichloro-2,2-bis[p-chlorophenyl]ethane) have
attracted a great deal of attention from both the scientific community and the general public.
Among the best known and most dramatic effects has been the impact of DDT and other
organochiorine pesticides on reproduction and survival of birds of prey, such as eagles,
Ospreys (Pandion haliaetus) and falcons. These birds, particularly the Bald Eagle (Haliaeetus
leucocephalus) and the Peregrine Falcon (Falco peregrinus), have become symbols of
environmental awareness and reminders of ecological consequences of short-sighted use of
chemical technology.
Although most Bald Eagle populations have recovered from the effects of DDT,
reproduction and survival in some areas are impaired by chemicals, such as PCBs, which can
function toxicologically like TCDD. A great deal of laboratory research has been conducted on
PCDDs and related compounds; however, little is known of exposure and effects on wildlife.
This thesis focused on the Bald Eagle population resident around British Columbia’s Strait of
Georgia and on exposure to and the consequences of the widespread pollution of that area by
PCDDs and PCDFs from forest industry sources.
Chlorinated hydrocarbons
Structures
Chlorinated hydrocarbons are organic compounds with chlorine substituents. This thesis
is concerned primarily with the polychiorinated aromatics, those with chiorines substituted on
aromatic ring structures, and to a lesser extent with some non-aromatic organochiorine
pesticides, such as hexachiorocyclohexane (HCH). Ecotoxicologically, the most important
1
polychiorinated aromatics are the PCDDs, PCDFs, PCBs, and some of the organochiorine
pesticides such as DDT.
The structures of the PCDDs, PCDFs and PCBs are represented in Figure 1. The
PCDDs and PCDFs obtain a mainly rigid, planar configuration, which determines their
biological behaviour. For the PCBs, the molecular conformation depends on the chlorine
substituents. Those congeners without ortho-chiorines energetically obtain a mainly planar
conformation, those with di-ortho chlorine substituents are non-planar and those with mono
ortho substituents are intermediary. Thus the non-ortho PCBs are approximate stereoisomers of
PCDDs and PCDFs and if chlorinated laterally, exhibit similar biological behaviour (Safe
1984).
Figure 1. Molecular structure and position numbering of polychiorinateddibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs) and biphenyls (PCBs).
9 1
8
6 4
dibenzo-p-dioxin 2,3,7,8 - Tetrachlorodibenzo-p-dioxin
-ClS
c17
3 2 2’ 3’
4
dibenzofuran 2,3,7,8 - Tetrachlorodibenzofuran
biphenylPCB 126
33’44’5 - Penta
2
Organochiorine pesticides fall into three structural groups (Figure 2). DDT is similar in
structure to the PCBs, in that it has two chlorine-substituted benzene rings, in this case joined
on an ethane backbone. Dieldrin, mirex and the chiordane-related compounds, including
heptachlor epoxide, all belong to the cyclodiene group. The third group are the chlorinated
benzenes and cyclohexanes.
Figure 2. Molecular structure of the major organochiorine pesticides.
Sources
PCDDs g PCDFs. Neither PCDDs nor PCDFs are deliberately produced commercially,
but are formed either as by-products during the synthesis of other chemicals, such as
chiorophenolic biocides, or during combustion of chlorine containing wastes. Incineration of
municipal and industrial wastes is the major global source of dioxins, which can be transported
long distances and subsequently deposited in soils and lake sediments (Czuczwa et al. 1984).
Although combustion produces a fairly uniform mixture of PCDD and PCDF isomers, physical
and chemical atmospheric processes favour the deposition and accumulation of less toxic higher
chlorinated compounds, which then dominate in sediments (Hites 1990). Elevated contamination
by more toxic and persistent isomers such as 2,3,7,8-TCDD was previously associated with use,
production or waste storage of chiorophenoxy acid herbicides, particularly 2,4,5-T (see:
DIC I-I LO RO DI PHE N YLET HAN ES cI_OH_O_ ci
CYCLODIENES
DDT, DDDDicofolPerthaneMethoxychiorMethiochior
ci
CHLORINATED BENZENES
CYCLOHEXANES
Aidrin, DieldrinHeptachlorChiorcianeEndosulfan
(Cl)6
HCB, HCHLindane (a-BHC)
Cl
3
Baughman and Meselson 1973; Fanelli et at. 1980; Powell 1984). However, relatively recent
studies showed that effluents from kraft pulp mills using elemental chlorine bleaching contained
2,3,7,8-TCDD and 2,3,7,8-TCDF (Kuehl et at. 1987), which caused contamination of fish and
wildlife in receiving waters (Rogers et at. 1989; Elliott et at. 1989a). Elevated HxCDDs
(hexachioro dibenzo-p-dioxins) in effluents and foodchains can result from pulp mill digestion of
tetrachiorophenol-contaminated woodchips (Elliott et al. 1989a; Luthe et at. 1990). Use and
production of 2,4,5-T and most chlorophenols has been regulated in North America. Pulp mills
in Canada, but not necessarily in the USA or elsewhere, now use alternative bleaching methods,
which have substantially reduced formation of TCDD and TCDF.
PCBs. PCBs were used for a variety of purposes which can be divided into ‘closed
circuit’ uses such as in electrical transformers and capacitors and in heat transfer and hydraulic
systems, and into ‘open circuit uses’ such as the formulation of lubricating and cutting oils,
pesticides, plastics, paints, inks, adhesives, etc. More than one billion (l0) kg PCBs were
produced worldwide (Tanabe 1988). Until 1977, over 90 % of the production was in the
U.S.A., after which it switched to Europe and Japan. Some 40 million kg PCBs have been
imported into Canada; the most recent inventory accounted for about 24 million and assumed that
the remaining 16 million kg had been lost to the Canadian environment (Environment Canada
1985). Open circuit uses of PCBs were voluntarily restricted by industry in 1973 and all uses of
PCBs have been regulated by governments in North America since 1977.
Organochlorine pesticides. OC pesticides are synthetic compounds widely used to control
agricultural and forest pests and the transmission of vector-borne diseases. The most abundant
OC pesticide in the environment is DDE, the major persistent metabolite of DDT. Other
compounds commonly detected in wildlife include DDD, DDT, dieldrin, heptachlor epoxide,
mirex, photomirex, toxaphene, oxychiordane, cis- and trans-chlordane, cis- and trans-nonachlor,
endrin, HCB, and HCH isomers. DDT, a broad-spectrum insecticide, was first used in North
America in the 1940s in public health campaigns to control lice (Carson 1962). From the 1940s
until the early 1970s, large quantities of DDT were sprayed to control forest insect pests in
4
British Columbia (Nigam 1975) and in the northwest USA (Henny 1977). Major restrictions on
the use of most organochlorine pesticides (ie. DDT, dieldrin, endrin, heptachior, HCH and
toxaphene) in Canada and the USA were first implemented in the early 1970s, with further
controls imposed throughout the 1970s and 1980s (Noble and Elliott 1986). Heptachior continued
to be used in Oregon until 1974 (Henny et at. 1983) and significant amounts of chiordane,
lindane, dicofol and toxaphene were used until the early 1980s in California (Ohlendorf and
Miller 1984). A few minor applications of chlordane, lindane, dieldrin and heptachlor (eg. seed
treatment, termite control) are still permitted in Canada and the USA. In Mexico, some
restrictions on the use of DDT, BHC, dieldrin and heptachior were imposed in 1980 (Burton and
Phiogene 1986).
Organochlorines can be transported over vast distances by atmospheric and oceanic
vectors; as such, ongoing use in Asia may now be the main source of OCs to the Canadian
environment, particularly the Pacific coast (Elliott et at. 1 989b). Information on OC use in
Asian countries bordering the north Pacific is scarce. Since the 1950’s, DDT and HCH have
been used extensively on rice, cotton and vegetable crops, but in the 1970s, many countries
began to replace them with organophosphorus compounds. As in North America, agricultural
uses of OCs are subject to regulation in most north Pacific countries, but the degree of
compliance varies. The People’s Republic of China manufactures OC pesticides; however, the
production and use of two, DDT and HCH, were banned there in 1983 (Wolfe et al. 1984). In
Japan, production and use of DDT and HCH were prohibited in 1971, but the use of chlordane
for termite control was permitted until the late 1 980s (Tanabe et al. 1989). Korea also prohibited
the use of DDT in the early 1970s (Phillips and Tanabe 1989). However, in Hong Kong (where
many pesticides are still formulated), there appears to be continued input of DDT into coastal
waters, despite restrictions imposed in 1988 (Phillips and Tanabe 1989).
Food chain bioaccumulation
For a substance to bioaccumulate, the following physico-chemical traits are necessary: 1)
lipid solubility evident by a high octanol/water partition coefficient; 2) resistence to metabolic
attack.
5
PCDDs PCDFs. Food chain bioaccumulation of PCDDs and PCDFs generally
requires a 2,3,7,8-substitution pattern, as congeners lacking that substitution pattern are
metabolized in birds, mammals and fish (Van den Berg et al. 1993a). Accumulation of non-
2,3,7,8-substituted PCDDs has been reported in some invertebrate species, particularly
crustaceans (Norstrom and Simon 1991).
PCBs. Among homeotherms, tissue retention of PCB congeners varies with development
of the cytochrome P450 system and capacity to metabolize different compounds. In general,
mono and non-ortho PCBs are metabolized by CYP1A enzymes, while di-ortho congeners are
degraded by CYP2B enzymes (Boon et al. 1987; Brown 1994).
Organochiorines. The relative capacity of organochlorines to bioaccumulate has been
extensively studied in the Herring Gull (Larus argentatus) by Norstrom and co-workers
(Norstrom et at. 1986; Clark et at. 1987; Braune and Norstrom 1989). The more slowly
degraded and therefore more accumulative OCs in birds are DDE, mirex and oxychlordane, with
heptachlor expoxide, dieldrin and HCH compounds being more rapidly cleared.
Effects of chlorinated hydrocarbons
PCDDs, PCDFs PCBs. This group of compounds causes similar toxic symptoms in
most species studied (Safe 1990). Dose-related responses include: irnmunotoxicity, liver
enlargement and other signs of hepatotoxicity such as porphyria, induction of drug-metabolizing
enzymes, reproductive toxicity and cancer promotion (Safe 1984). The toxicity of the individual
compounds varies greatly with the molecular structure. The most toxic compound is 2,3,7,8-
TCDD, which is often used as a model for studying the effects of these chemicals. The more
toxic furan and biphenyl congeners all exhibit a structural similarity to 2,3,7, 8-TCDD. Many of
the toxic effects caused by this class of compounds are believed to be mediated by a cytosolic
receptor found in many tissues, known as the aryl hydrocarbon (Ah) receptor (Landers and Bunce
1991). The Ah-receptor mediated mode of action is represented schematically in Figure 3.
Traditional toxicology studies have focused on single chemicals in test organisms.
However, environmental exposure to chlorinated hydrocarbons involves a multitude of
6
compounds. To provide a practical method of dealing with this, the study of Ah-receptor
mediated structure-activity relationships has produced an additive scheme for estimating the
toxicity of complex mixtures of these chemicals through use of “TCDD Toxic Equivalence
Factors” (TEFs). Each individual compound is assigned a TEF relative to 2,3,7,8-TCDD,
essentially a ratio of its relative toxicity based on one or more endpoints. Analytically
determined concentrations are multiplied by the TEF, the results summed to produce the “TCDD
Toxic Equivalents” or TEQs. TEFs published by Safe (1990) are widely used; however, those
reported by Ahlborg et al (1994), which attribute lower relative toxicity to the mono-ortho PCBs,
appear more relevant for most birds (Brunstrom and Andersson 1988; Bosveld et al. 1992;
Kennedy et al. 1994).
Xenobiotic ligand
(TCDD, etc)
INCREASEDMETABOLISM OFDRUGS ANDENVIRONMENTALCHEMICALS
TOXICITY
Figure 3. Molecular mechanism proposed for TCDD and related chemicals. The lipophiicxenobiotic ligand, such as TCDD, enters the cell by passive diffusion through the
lipo-protein cell membrane and binds with the Ah-receptor (AbR); the AhR releases aheat shock protein (HSP 90) as it binds with the ligand. The ligand-receptor complex
then associates with the nuclear translocating protein (Arnt) and moves into the nucleus,where it interacts with dioxin responsive elements (AhREs) on the genome, which alters
the transcription of specific niRNAs. The resulting proteins then mediate the biochemicaland toxic responses observed with TCDD exposure (after Okey et al. 1994).
7
Although the toxicology of dioxins and related compounds continues to be extensively
studied in laboratory mammals, there are less data on avian species. Bird studies have focused
on embryos, as the most sensitive life stage (Peterson et at. 1993). Chicken embryos are
particularly sensitive: the LD50 for 2,3,7, 8-TCDD, administered into the air sac of the chicken
embryo, was reported as 250 ng/kg (ppt) egg (Alired and Strange, 1977). An LD50 for 2,3,7,8-
TCDD in chicken embryos of about 200 ng/kg was determined more recently by both Henschel
et at. (in preparation) and Janz (1995) using air cell and yolk sac injection. They also reported a
very steep dose response curve, with no mortality at 100 ng/kg and complete mortality at 300
ng/lcg. Injection of 2,3,7,8-TCDD or similar compounds into developing chickens causes a
toxicity syndrome which includes, in addition to mortality, beak and other deformities, thymic
and bursa inhibition, edema and liver lesions (Brunstrom and Andersson 1988; Brunstrom 1990).
The heart is a sensitive target organ as only 9 ng/kg caused an increase in the incidence of
cardiovascular malformations (Cheung et at., 1981). In domestic turkey embryos, non-ortho
PCB congeners that bind the Ah receptor and thus act by a similar toxic mechanism to 2,3,7,8-
TCDD, also cause gross deformities and mortality, but not the other symptoms seen in chicken
embryos (Brunstrom and Lund 1988). In embryos of other avian species, such as Ring-necked
Pheasants (Phasianus coichicus) and Eastern Bluebirds (Siatia sialis) injected with 2,3,7,8-
TCDD, sublethal effects observed in chickens were not observed, rather mortality was the most
sensitive endpoint (Nosek et at. 1992; Martin et at. 1989). The LD5O for 2,3,7,8-TCDD was
1100 ng/kg egg in the pheasant embryo and between 1000 and 10,000 ng/kg egg in the Eastern
Bluebird embryo, in both cases via albumin injection (Nosek et at. 1992; Martin et at. 1989).
Brunstrom & Reutergardh (1986), using mortality as an endpoint, reported marked interspecific
sensitivity among birds to the TCDD-isostereomer, PCB congener 77 (34-34). Chickens were
the most sensitive, followed by turkeys (30 X less sensitive) and pheasants (100 X less sensitive)
and then by Mallards, Goldeneyes, domestic ducks, geese, Herring Gulls and Black-headed Gulls
(>1000 X less sensitive).
8
Adults of avian species were much less sensitive to TCDD than embryos; 25 to 50 gIkg
(ppb) body weight caused mortality in chickens (Greig et al. 1973), while 25 pg/kg caused 75%
mortality in ring-necked pheasant hens (Nosek et al. 1993). In other studies with adult birds,
acute oral toxicity of 2,3,7,8-TCDD ranged from 15 pg/kg body weight in Northern Bobwhite
(Colinus virginianus) to greater than 810 pg/kg body weight in the Ringed Turtle Dove
(Streptopelia risoria), (Hudson et al. 1984).
There are few published studies of the chronic effects of dioxin-like compounds in birds.
Kenega and Norris (1983) reported that a diet containing 0.3 or 3 ng/kg TCDD in a formulation
of 2,4,5-T fed to bobwhites for 18 weeks produced no effects on egg production or survival of
embryos. However, 50 % mortality did occur within 5 days at a dietary level of 167 ng/kg.
Nosek et al. (1992) showed that Ring-necked Pheasants dosed with 1.0 ug/kg/week of 2,3,7,8-
TCDD for 10 weeks exhibited mortality and signs of wasting syndrome; egg production was also
reduced and hatchabiity of eggs was < 2 %. Pheasants dosed with 0.1 pg/kg/week for 10
weeks exhibited no adverse effects. Daily feeding of PCB congeners 126 (34-345) and 105 (234-
24) for up to eight weeks caused hepatic porphyria, thymic atrophy (PCB 126 only) and marked
microsomal cytochrome P450 enzyme induction in Japanese Quail (Coturnix coturnix), but no
porphyria, and only minor P450 induction in American Kestrels (Falco sparvarius) (Elliott et al.
1990; 1991). This is the only available laboratory study involving TCDD-like compounds in a
bird of prey.
Field studies of PCDDs. PCDFs and PCBs in birds. In the Great Lakes, a toxic
syndrome observed in a number of fish-eating bird species, is referred to as GLEMEDS (Great
Lakes embryo mortality, edema and deformities syndrome), and has been attributed to exposure
to PCBs, PCDDs and PCDFs (Gilbertson et al. 1991). The syndrome was first recognized in
Lake Ontario gull and tern populations in the early 1970s (Gilbertson and Fox, 1977).
Subsequent retrospective analysis of archived Herring Gull eggs revealed the presence of high
2,3,7,8-TCDD concentrations in eggs of Lake Ontario gulls collected in the early and mid 1970s,
which likely contributed to poor reproduction (Gilbertson et al. 1991). However those eggs also
9
contained high levels of other known embryotoxins, including PCBs and HCB (Mineau et at.
1984; Bishop et at. 1992). A number of recent studies in the Great Lakes: (Kubiak et at. 1989;
Tilett et at. 1992; Yamashita et at. 1993; Rattner et at. 1994) related exposure to PCBs,
particularly the non-ortho 126 (345-34) and the mono-orthos 105 (234-34) and 118 (245-34) to
biological effects in colonial waterbird populations. Recently, Bosveld et at. (1994) and Van den
Berg et at. (1994) reported high PCB levels in eggs of fish-eating birds breeding in the Rhine
estuary, which correlated with various endpoints of exposure and toxicity, including CYP1A
induction and embryonic growth.
In British Columbia, Great Blue Herons (Ardea herodias) and Double-crested Cormorants
(Phalacrocorax auritus) breeding near pulp mills have been used as sentinel species to study
toxicant exposure and effects (Elliott et at. 1989; Whitehead et at. 1 992a). Failure of a Great
Blue Heron colony in 1987 at Crofton, British Columbia coincided with a three-fold increase in
mean egg levels of 2,3,7,8-TCDD over the previous year when reproduction was normal;
however, no statistically significant relationship between contaminant levels and reproductive
outcome among individual birds was determined (Elliott et at. 1989a). Heron embryos, collected
in 1988 at colonies with high, intermediate and low levels of PCDD and PCDF contamination
and incubated in the laboratory, did not exhibit any significant differences in hatching success
among the three sites. There were, however, a number of sublethal effects in heron chicks,
which correlated with their 2,3,7,8-TCDD levels, including induction of hepatic EROD
(ethoxyresorufin-O-deethylase) activity, edema and lower embryonic weight (Bellward et at.
1990; Hart et at. 1991; Sanderson et at. 1994) and brain abnormalities (Henshel et at. 1995).
Disturbance by people and/or Bald Eagles (Norman et al. 1989) was probably the main cause of
heron colony failure at Crofton in the late 1980s on the British Columbia coast and would have
masked other potential factors (Elliott et at. 1 989a); however, intensive observation of heron
nests showed that mean time spent incubating was lower and greater between-nest variability in
incubation time occured at a contaminated versus a control heron colony in 1988 (Moul 1990).
The strong possibility exists, therefore, of a contamiiiant-related effect on adult incubation
10
behaviour. Chemically mediated aberrant parental behaviour has been reported for a number of
species in both laboratory (Peakall and Peakall 1973; McArthur et at. 1983) and field studies
(Cooke et at. 1976; Mineau et at. 1984; Kubiak et at. 1989).
Eggs of ospreys nesting downstream of bleached-kraft pulp mills on the Thompson and
Columbia rivers of the British Columbia interior, contained significantly higher levels of 2,3,7,8-
TCDD than eggs from nests upstream of the mills (Whitehead et at. 1993). Studies of osprey
productivity showed a trend of lower productivity at downstream compared to upstream sites;
however, there were a number of confounding factors, particularly relating to food supply.
White & Hoffman (1991) recently reported poor reproductive success in Wood Ducks (Aix
sponsa) contaminated with TCDD and TCDF from a 2,4,5-T waste disposal site in Arkansas.
Mean levels in Wood Duck eggs were 70 to 75 ng/kg for both TCDD and TCDF. Based on the
limited chemical data provided, Wood Ducks appear to be more sensitive to the effects of TCDD
than other wild bird species.
Organochiorine pesticides. The acute vertebrate toxicity of DDT is low, the LD50
to the Japanese Quail was 595 mg/kg (ppm). The cyclodienes are much more acutely toxic to
vertebrates; 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 that
has been best characterized in wildbirds (Anderson et at. 1975; Blus et at. 1974; Newton and
Bogan 1974; Blus et at. 1980; Custer et at. 1983; Elliott et at. 1988). DDE affects calcium
metabolism by interfering with carbonic anhydrase metabolism at the shell gland (Cooke 1983).
Critical egg levels of DDE vary widely among species and have been established for some
raptors (Fyfe et a!. 1988; Peakall et at. 1991; Wiemeyer et a!. 1993), The chronic toxicology of
other organochiorines to wild birds has not been established. Dieldrin has been implicated in
11
reproductive effects, not via eggshell thinning, but rather embryotoxicity. Lockie et al. (1969)
suggested that dieldrin levels in eggs greater than 1.0 mg/kg were associated with egg failure in
Scottish Golden Eagles (Aquila chrysaetos), However, the association of this level may have had
more to do with its indication of lethal dieldrin residues in adult birds as suggested by Newton
(1986) for European Sparrowhawks. Heptachlor epoxide, at egg levels > 1.5 mg/kg was
associated with effects on reproduction of American Kestrels (Henny et al. 1983). Egg levels of
HCB >5.0 mg/kg in Herring Gull chicks were associated with embryo mortality (Boersma et al.
1986).
The Bald Eagle
Natural history
The Bald Eagle is an endemic North American member of the genus haliaeetus, the sea
eagles. Bald Eagles are sexually dimorphic, adult females average 5.3 kg and 221 cm, and males
4.3 kg and 207 cm (Stalmaster 1987). Breeding adults are thought to form life-long pair bonds;
the average breeding life span is about 20-25 years. Breeding success may vary considerably
from year to year depending on factors such as disturbance and food supply (Stalmaster 1987).
In the Pacific northwest, Bald Eagles are year-round residents (Hancock 1964). The
breeding season can last from February until August, although nests are maintained year round
(Herrick 1932). Eagles often have more than one nest in a territory; the function of the alternate
nest is not clear, but may be to reduce parasite loads (Stalmaster 1987). Nests are always located
in proximity to water. Nest trees are usually the dominant or codominant tree in the area in
order to provide a clear view of the territory and clear flight paths to feeding areas. Female
eagles lay from one to three eggs, two being most common. Eggs are incubated for about 35
days, and the chicks are dependent on their parents at the nest for food and protection for another
72 to 96 days (Herrick 1932). Adults appear to intially remain with the chicks on fledging;
subsequent juvenile dispersal patterns can be complex (McClelland et al. 1994). Chances of
reaching adult age are variable and may be less than 10 in some populations, such as in
12
Alaska, and as high as 50 % in more southern locations. Bald Eagles do not attain adult plumage
until their 5th year, when they normally begin breeding (McCollough 1989)
Eagles have a number of physical adaptations as predators. They have excellent vision
and can reportedly detect other eagles flying at 23 to 65 km distance (Shlaer 1972). They kill
using their powerful feet and talons, while food is torn apart by a large beak. They are powerful
flyers, particularly adapted for soaring in open country. Bald Eagles are opportunistic foragers
and predators. In the northwest, birds, particularly gulls and waterfowl, marine and aquatic fish,
and invertebrates make up the bulk of the diet for most birds, although mammals can be
important in some areas (Vermeer et al. 1989; Knight et at. 1990; Watson et al. 1991).
Population trends and critical factors
Like many other large predatory animals, Bald Eagle populations declined during the past
century over much of their North American breeding range (Stalmaster, 1987). Habitat loss and
degradation combined with intentional and accidental killing contributed to poor productivity and
loss of breeding stock. In the early 1950s, populations of eagles and other birds of prey began to
disappear from many areas. Classic work by Charles Broley (1947, 1958) showed a precipitous
decline in productivity of a Florida population from a high of 89 % nest success in 1942 to 14 %
in 1952. During the 1960s and 1970s, eagle productivity was subsequently found to be below
sustainable levels in many areas of the U.S. and Canada (Stalmaster, 1987). The low breeding
success of Bald Eagles and other birds of prey, which began in North America in the early
1950s, coincided with the introduction of DDT and other organochlorine pesticides. The widely
accepted paradigm for decline of the Bald Eagle and other North American raptor populations
states that DDE persists, bioaccumulates and impairs reproduction via the mechanism of reduced
eggshell quality (Grier, 1982; Peakall et at., 1991). Wiemeyer et al. (1984) determined that in
Bald Eagles, reproductive failure approached 100 % when DDE egg levels were greater than 15
mg/kg. DDE egg levels of 5 mg/kg were associated with 10 % eggshell thinning, while
populations with less than 3 mg/kg exhibited no significant shell thinning and normal production
of young. However, those values were based on analyses of adled eggs which may tend to have
higher than average residues and may bias the estimate of critical values. In other birds of prey,
13
particularly European populations, loss of breeding stock to acute dieldrin poisoning, has been
suggested to be more critical than DDE-induced shell thinning (Newton et a!. 1992). Bald Eagles
have also been acutely poisoned by other pesticides, including dieldrin, and heavy metals, such as
mercury and lead (Reichel et al. 1984), athough these effects were probably less critical to
population decline. At any rate, eagle productivity has improved and populations have increased
in most areas, following strict regulation by the early 1970s of organochiorine use in North
America (Grier, 1982; Wiemeyer et a!. 1993). As a result in July, 1995, the U.S. Fish and
Wildlife Service changed the status of most Bald Eagle populations in the continental U.S.A.
from endangered to threatened.
However, breeding success remains below maintenance levels at some regional ‘hotspots’.
Along the Great Lakes shoreline, productivity is lower and contaminant levels higher than at
nearby inland locations (Bowerman 1993), although, at least for Lake Superior populations,
reduced food delivery to nestlings was an important factor (Dykstra 1994). Eagle populations in
Maine generally exhibit low productivity, which has been related to high contamination by PCBs
and DDE (Welch 1994). Along the lower Columbia River, low Bald Eagle breeding success
correlated with high egg and plasma levels of DDE and PCBs; moderately high levels of 2,3,7,8-
TCDD (tetrachloro dibenzo-p-dioxin) were also present in those eggs (Anthony et al. 1993).
British Columbia Eagle Populations
Based on a 1984 report, no Canadian Bald Eagle populations are listed as threatened
(COSEWIC 1995). In British Columbia, most Bald Eagle populations are “blue listed”, based on
concern for long term conservation of some populations (British Columbia Conservation Data
Centre 1995).
Hodges et at. (1984) estimated the resident breeding population of Bald Eagles on the
British Columbia coast to be about 9,000 birds. An estimated 30,000 eagles winter on the coast,
mainly in the river estuaries surrounding the Strait of Georgia (Farr and Dunbar, 1988). Bald
Eagles are lured by the rich food resources and high biological productivity, both terrestial and
marine, of the Strait of Georgia, which is essentially a large estuary with nutrient input from
numerous rivers, particularly the Fraser (LeBlond 1989); those rivers are also major salmon
14
spawning sites, which attract thousands of eagles each winter. Millions of waterbirds and
shorebirds migrate through and winter in the region, which provides the major food supply for
Bald Eagles and falcons. The basin is surrounded by temperate rain forests which have been
extensively exploited for wood fibre. The impact on Bald Eagles of habitat modification,
especially the clearing of nest trees, has received some attention (Bunnel et al. 1994). With
increased population growth and commercial activity, especially of coastal and estuarine areas in
the Georgia basin, habitat for Bald Eagle roosting and nesting will be continually threatened. In
addition to those stesses, there are major pollutant inputs, particularly from pulp mills and other
wood processing industries.
Coastal Bald Eagle populations in British Columbia apparently did not experience the
major declines that occurred elsewhere during the organochlorine era. Anecdotal information
(based on discussion with naturalists, farmers and fisherman) suggests that in the Fraser River
delta, eagles were less common in the 1960s and 1970s than at present. In 1987, Vermeer et al.
(1989) resurveyed areas of the southern Gulf Islands where nests had been counted previously
(Hancock 1964; Trenholm and Campbell 1975) and reported a 30 % increase in the number of
nests since 1974, which they attributed mainly to increasing food supply in the form of Glaucous-
winged gulls (Larus glaucescens). However, data derived from such comparisons requires
cautious interpretation, as it may be more indicative of increased survey intensity and ability to
find nests (Henny and Anthony 1989). In the Okanagan Lakes region of interior British
Columbia, Bald Eagles have disappeared as a breeding species (Cannings, 1987); orchard areas
of the Okanagan valley received heavy DDT applications and wildlife samples from that area are
still highly contaminated (Elliott et al. 1994).
Problem Statement
As a predator feeding at the top of marine and estuarine food chains, Bald Eagles are
exposed to an array of persistent environmental chemicals, particularly chlorinated hydrocarbons
and mercury. There is a considerable body of literature on levels of organochiorine pesticides and
total PCBs in tissues of Bald Eagles. However, there is very little published data on levels of
individual PCB congeners, particularly the toxic non-ortho PCBs, or on levels of other significant
15
environmental contaminants including polychiorinated dibenzo-p-dioxins (PCDDs) and
polychiorinated dibenzofurans (PCDFs) in Bald Eagles. In addition, whereas significant progress
has been made in detennining critical levels of DDT-related compounds and mercury for Bald
Eagle eggs (Wiemeyer et al. 1993), there is no such information for other chlorinated
compounds.
The Strait of Georgia provides an interesting location to investigate the effects of PCDDs
and PCDFs on eagle populations. Previous studies in the area showed that fish-eating birds, such
as Great Blue Herons, Double-crested Cormorants, Western Grebes (Aechmophorus occidentalis)
and Common Mergansers (Mergus merganser), all of which are potential Bald Eagle prey, were
contaminated with high levels of PCDDs and PCDFs, but relatively low levels of other
organochlorines (Elliott et at., 1989; 1992; Whitehead et at., 1990; 1992). Concentrations of
PCDDs and PCDFs in Western Grebes and in Surf Scoters (Melanita perspicillata), another eagle
prey item, collected near some British Columbia coastal mills in 1990 were high enough to
warrant advisories against their consumption by people (Whitehead et al. 1990). In Great Blue
herons, episodes of poor breeding success in the late 1980s at a colony near a kraft pulp mill
were associated with sublethal effects on embryos, including edema, reduced body weight and
EROD induction which correlated well with levels of 2,3,7,8-TCDD (Beliward et al. 1990; Hart
et at. 1991; Sanderson et at. 1994a). Coastal Bald Eagle populations feed heavily on marine
birds such as Western Grebes and Glaucous-winged Gulls and on larger fish (Knight et at.,
1990). Eagles are therefore exposed to even higher dietary contaminant levels than species such
as herons and cormorants which eat mainly smaller fish. In winter, after salmon runs are over,
Bald Eagles eat mainly waterfowl (Watson et at., 1991) and thus are exposed to toxicants, such
as lead shot and pesticides, acquired by waterfowl feeding in other distant areas, such as the
western USA. Lead poisoning is a major cause of death for British Columbia Bald Eagles
(Elliott et at. 1 992a), while pesticides are an important mortality factor in local areas such as the
Lower Fraser Valley (Elliott et at. submitted).
16
There is, therefore, potential for exposure of Strait of Georgia Bald Eagles to potentially
harmful levels of chlorinated organics and other toxicants. Positioned at the top of the food web
and with a high public profile, Bald Eagles are an excellent sentinel species and indicator of
ecosystem health. Thus, further research is warranted.
Hypotheses 4 Objectives:
Mortality study
Hypothesis: The accumulation of persistent chlorinated hydrocarbons will affect the survival of
Bald Eagles, particularly if fat stores are depleted during periods of environmental stress.
Objective: To investigate bald eagle mortality in British Columbia and specifically the role of
chlorinated hydrocarbons versus other causes of death; to determine spatial and possibly temporal
trends in contamination.
Embiyotoxiciry study
Hypothesis: Accumulated chlorinated hydrocarbons are transferred from females into eggs,
where they negatively affect growth, development and survival of embryos.
Objectives: To examine the health of Bald Eagle embryos exposed to an environmental gradient
of chlorinated hydrocarbon pollutants and to relate the degree of exposure to biomarkers such as
CYP1A induction; to document exposure by chemical analysis of yolk sacs.
Bioaccumulation study
Hypothesis: Chlorinated hydrocarbons, particularly PCDDs and PCDFs from pulp mill sources,
are accumulating at high concentrations in bald eagle eggs as a result of their position as top
predators in marine and estuarine food chains.
Objectives: To determine spatial and temporal patterns of chlorinated hydrocarbons in Bald
Eagle eggs and to relate those levels to the diet and to sources; to determine critical
concentrations of contaminants, particularly PCDDs and PCDFs in the eagle diet.
Productivity study
17
Hypothesis: The accumulation of persistent chlorinated hydrocarbons in Bald Eagles impairs
overall reproduction through toxicity to embryos, reduce survival of nestlings or impaired
development of the reproductive system.
Objectives: To determine breeding success of a representative sample of eagles in the Strait of
Georgia and reference locations and to relate breeding success to chlorinated hydrocarbon levels
in nestling blood samples; to examine the role of other factors critical to breeding success,
particularly food supply.
Overview of the thesis
This thesis represents the results of a four year field and laboratory study of chlorinated
hydrocarbon exposure and effects in Bald Eagle populations on the coast of British Columbia. In
the first chapter, mortality and the role of chlorinated hydrocarbons are examined through
autopsy and liver residue analysis of eagles found dead and dying from 1989 to 1993 in British
Columbia. Chapter two presents the results of a laboratory incubation study of in ovo effects of
PCDDs, PCDFs and PCBs in an environmental exposure gradient. Contaminant levels in yolk
sacs are presented with the results of biomarker assays, such as CYP1A, in embryonic tissues.
The data are used to estimate a no-observed-effect-level (NOEL) and a lowest-observed-effect-
level (LOEL) for TCDD-toxic equivalents in eagle eggs. Chapter three presents contaminant
residue levels for Bald Eagle eggs and prey items. Patterns, trends and sources are discussed and
a simple bioaccumulation model used to relate levels in eagles to those in their food chain. In
Chapter four, the results of productivity studies and contaminant levels in nestling plasma
samples are presented. Relationships between breeding success, contaminant levels and other
variables, particularly food supply, are discussed.
18
CHAPTER 1
CHLORINATED HYDROCARBON LIVER LEVELS AND AUTOPSYDATA FOR BALD EAGLES FOUND DEAD OR DEBILITATED, 1989-1993.
The objective of this study was to determine the degree of chlorinated hydrocarbon
exposure of adult and juvenile Bald Eagles and to assess spatial trends in contamination.
Statistical examination of relationships among environmental contaminant levels and cause of
death was a secondary objective. Preliminary reports on toxicants such as lead (Elliott et al.
1992a) and anticholinesterase pesticides (Elliott et al. in press[b]) have been made, but are not
included as part of the thesis. In this chapter, the results of autopsies and analyses of PCBs and
organochiorine pesticides in livers of 59 eagles found dead in British Columbia over the period,
1988 to 1993, and results from a subset of 19 birds analyzed for PCDDs and PCDFs are
presented and evaluated.
Materials and Methods
Sample collection
Specimens collected for this study were part of an overall investigation into the health
status of Bald Eagles in British Columbia. Carcasses were obtained by writing to potential
cooperators, including government and non-government agencies, veterinarians and wildlife
rehabilitators and by placing advertisements in periodicals. Sick, injured and deceased Bald
Eagles were thus obtained from all of the above sources. Specimens were received and initially
examined at the Pacific Wildlife Research Centre and then shipped on ice to the Island
Veterinary Hospital, Nanaimo, British Columbia, where they received a complete autopsy by
Dr. K.M. Langelier.
19
The 484 eagles received were grouped by geographical area as follows: lower Fraser
valley, Strait of Georgia, Johnstone Strait, west coast Vancouver Island and north coast. A
total of 59 individuals were analyzed for organochiorines and PCBs (Figure 1.1, see also
Appendix 1.1). Specimens for analyses were selected in order to provide a reasonably
representative sub-sample, based on age, sex, and collection location. Other criteria were also
considered such as a preliminary diagnosis of non-specific poisoning or proximity of the carcass
to an industrial pollutant source. Some eagles were also analyzed for organochiorines during
investigations of suspected poisonings by lead or anticholinesterase pesticides. Birds found
Figure 1.1 Locations where eagles were collected in British Columbia, 1989-93,and analyzed for chlorinated hydrocarbons (N = 59).
20
dead during the breeding season in the Strait of Georgia, and therefore likely to be resident
birds, were considered to have priority for analysis.
Concentrations of PCDDs and PCDFs were determined in nineteen liver samples.
Criteria for selection of samples for PCDD/PCDF analysis were as follows: 1) collected in the
Strait of Georgia or Johnstone Strait 2) collection date in late spring or summer, i.e. resident
birds 3) breeding age birds 4) high organochiorine levels. Criteria were set to maximize
chances of analyzing eagles which had been exposed to pulp mill pollutants.
Based on elevated levels of total PCBs, nine samples were selected for high resolution
GC/MS analysis of non-ortho PCB congeners. Linear regressions were determined between
concentrations of non-ortho PCBs and total PCBs for the nine livers analyzed, in order to
estimate values for the other ten livers which had been analyzed for PCDDs and PCDFs and
thus to estimate TCDD toxic equivalents. Regressions were not significant for PCBs 77, 81
and 37, but were significant for PCBs 126 and 169:
PCB 126 (ng/kg) = 92 [sum-PCB5 (mg/kg)] + 310, r2 = 0.660, p<O.Ol
PCB 169 (ng/kg) = 27 [sum-PCBs (mg/kg)] + 75, r2 = 0.584, p <0.05
Chemical analysis
Carcasses were stored at -20° C until postmortem examination. Tissue samples were
frozen at -20°C in chemically-cleaned (acetone/hexane) glass jars, frozen, and shipped to the
National Wildlife Research Centre (NWRC), Hull, Quebec, for analysis in the laboratory of
Dr. Ross Norstrom.
Organochlorines in liver were analyzed according to methods described previously
(Norstrom et al. 1988), except that PCBs were reported as the sum of 28 congener peaks.
Briefly, 2-4 gram sections of liver were dehydrated by grinding with excess anhydrous sodium
sulfate and colunm extracted with 50% methylene chloride in hexane. After extraction, the
eluate was concentrated on a rotovapor, further mixed with hexane and a 0.5 ml sample taken
for lipid determination (removal of solvent and weighing of residue). The remaining extract
was then cleaned up and separated into three fractions by Florisil chromatography. The
fractions were analyzed by gas chromatography-electron capture detector using a 60m DB-5
21
capillary column (Superco Inc.). Fraction 1 contained PCBs, p,p’-DDE, hexachlorobenzene,
pentachioroberizene, tetrachlorobenzenes and mirex. Fraction 2 contained cis-chiordane,
oxychiordane, trans-nonachior, and beta-hexachiorocyclohexane. Fraction 3 contained dieldrin.
Recoveries of these compounds by this method ranged from 82-94%. Quantification of PCB
congeners was effected by using a calibrated internal PCB standard solutions. Detection limits
were 0.005 mg/kg for organochiorine pesticides and 0.0025 mg/kg for PCB congeners.
Livers from 1990 collections were analyzed for PCDDs/PCDFs by low resolution
GC/MS using a Hewlett-Packard 5987B with a 30 m DB-5 capillary GC column according to
methods described in Norstrom et al. (1990) and Norstrom and Simon (1991). The method
employed gel permeation-carbon chromatographic clean-up and the use of13C12-labelled internal
standards for quantification.
Analysis of PCDD/PCDFs and non-ortho PCB in livers from other years were carried
out according to methods in Letcher et al. (in press). The method involves neutral extraction
followed by removal of lipids and biogenic compounds by gel permeation chromatography and
alumina column cleanup. Separation of PCDDs, PCDFs and non-ortho PCBs from other
contaminants was achieved using a carbon/fibre column; further separation of PCDDs/PCDFs
from the non-ortho PCBs was effected by Florisil column chromatography. Quantitation was
performed with a VG Autospec high resolution mass spectrometer linked to a HP 5890 Series II
data system. Each sample was spiked with‘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 of
internal 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 PCB
fraction, just prior to analysis to serve as recovery standards, for quantification of internal
standard recoveries. Recoveries of13C12-PCDDs/PCDFs/non-ortho PCBs were calculated by
comparing the integrated areas of the labelled internal standards and the areas of the recovery
standards in the samples to the areas of those compounds measured in the external standard
22
mixture, analyzed along with the samples. Results were generally accepted when recoveries of
labelled standards were between 70% and 120%.
Statistical analysis
Organochiorine pesticide data were transformed to common logarithms and geometric
means and 95 % confidence intervals were calculated with the data grouped by collection site.
Differences among sites were tested by a two-way ANOVA followed by Tukey’s multiple
comparison procedure (MCP). To test for an association between residue levels and cause of
death, birds were grouped into 12 categories (Figure 1.2) and analyzed by a one-way ANOVA.
All statistical tests were done using SYSTAT. A value of p < 0.05 was used throughout.
TCDD-toxic equivalents (TEQ5) were calculated using three different sets of TEFs,
Safe’s (1990), chick embryo hepatocyte (CEH) (Kennedy et al. in press) and WHO (Ahlborg
etal. 1994).
Results
Autopsy results
The diagnosed cause of death for each individual bird analyzed for organochiorines is
included in Appendix 1-1. Autopsy results for the 59 Bald Eagles analyzed for organochiorines
were compared to the total of 484 examined in the broader study with the causes grouped into
twelve categories (Figure 1.2). The graphs indicate that the subset for analysis was reasonably
representative of the range of mortality factors. Only two minor categories, falling from the
nest and infectious disease, were not represented.
There were no statistically significant associations between any of the chlorinated
hydrocarbon levels and cause of death. However, given the relatively small sample size, even
within the Strait of Georgia, and the variance in the residue levels, the probability of detecting
a significant association was low.
23
Clinical diagnosis
trauma
electrocution
undetermined
eagle attack
fell from nest
drowning
_____________________
liii186
_________J49_____________
46
__________
30
_________
j25
________
I24
________122_______
120
______
18
117
Percent
Bald eagles submitted to Island Veterinary Hospital, N484
Figure 1.2 Diagnosed cause of death for Bald Eagles analyzed compared
Organochiorines and total-PCBs
to the complete set of birds received.
Organochiorine pesticides and total PCBs were generally low; most eagles had DDE and
PCB levels < 5.0 mg/kg (Figure 1.3). However, a few birds had elevated levels of DDE and
total PCBs (>50 mg/kg) and chiordane-related chemicals (>1.0 mg/kg) (Appendix 1.1)
..
•.: DDE
:::totaI PCBS
.-••.••••••-
••.•• .••., ‘.-.•••• - -• -.
8%- -
____.
.. -
Figure 1.3 Numbers of Bald Eagles showing different DDE and PCB levels in livers (N=59)
D
veh. collision
14.. ‘
j2 *:.•(•n•O ófbátdéáglej):.::..:..... ...::
•.•.. •••,•••••.•••••..
I I I•infectious
Bald eagles analyzed for OCsIPCBs, N59
0 5 10 15 20 25 0 5 10 15 20 25Percent
25- I. — —•
44.%l
1-Li
Concentration (mg/kg, wet wt.)
,-10
24
Quantifiable levels of total PCBs, trans-nonachlor and oxychiordane were present in all
59 samples analyzed while DDE was present in 98 % of the samples. There were quantifiable
levels of DDD, heptachlor epoxide and dieldrin in 96 %; DDT, hexachlorobenzene (HCB)
mirex, beta-hexachlorocyclohexane (b-HCH) trans-chlordane and cis-nonachior in 92 %;
octachiorostyrene (OCS) in 80 %; and photomirex in 50 % of the samples.
Significantly elevated geometric mean residue levels were measured in Johnstone Strait
samples, followed by the Strait of Georgia, with the other four sites all being lower. Mean
DDE levels were significantly higher in samples from Johnstone Strait compared to the lower
Fraser valley.
1000 : : : : : -
100—ci) E : : : : : : : : : : :
1 0
______
Ainddatapoints. -. . maverages
0.0001 I I I I I I I I I
Inn —
—-: - •—
—
ci,
-
__
- : : : : : : : : : nd. data pointsC.)
averages
-
— I . I I I I
Jan Feb. Mar. Apr. May Jun. Jul. Aug Sep. Oct. Nov. Dec.
Month
Figure 1.4 DDE and PCB residue levels in Bald Eagle livers by collection month
25
Individuals with elevated organochiorines were found mainly in late spring or early
summer (Appendix 1.1). Concentrations of DDE and total PCBs in eagle livers tended to
increase throughout the winter, peak in April, level off and even decline slightly in summer
(Figure 1.4).
1000- = = = - - = .: - = = - = =
100 E C
-:
0)
0)_E 10— C C C :::,:: : :: :: C : :C
o
----.41- = = = C = = = : = = = = = :: = = = = = C = = , =
E E E 4E E E E E E:::::.:::.:::.a:::.: ..
0C.)
0.1
0.01— — I I I I I
100- -
- -
10
:::::::::::::::::::::::::.0) - -
— —-.
C .. ..... — I -
o 1 —=
q- — ICa) —C.)C _0c. 01—
C.) —..
0.01— ————. I I I I
0 1 2 3 4 5*
Bald eagle body conthtion
* Scale: 0 poor, 5 - excellent
Figure 1.5 Liver DDE and PCB residue levels in relation to body condition
26
Although eagles with higher residue levels of DDE and PCBs tended to weigh less than
those with lower residues, for neither DDE (ANOVA, f=3.28, p=O.077) nor PCBs (f=3.5,
p =0.068) was the relationship significant. However, comparison of DDE and PCBs with a
numeric scoring of body condition did produce statistically significant negative relationships for
DDE (f=7.4 p=O.009) and PCBs (f=8.5 p =0.005 (Figure 1.5).
Non-ortho PCBs
Levels of three non-ortho PCB congeners and two mono-ortho PCB congeners, PCBs
105 (234-(234-34) and 118 (245-34), which are present at relatively high concentrations and
also considered to be partial Ah-receptor agonists (Safe 199), are presented in Table 1.2.
For the non-ortho PCBs, in most samples, the pattern was of PCB 126 > 77 > 169 >
81 > 37. There were some exceptions; in three cases (Dent Island, Nanaimo and Port Hardy,
1990), PCB 77 > 126. In one case, Campbell River, PCB 169 >77.
PCDDs and PCDFs
The most contaminated individuals were from near pulp mill sites, Powell River and
Campbell River or nearby areas, such as Bowser and Sechelt (Table 1.3). In most samples,
highest levels were of HxCDD followed by PnCDD; after that, the relative levels of TCDD,
TCDF and PnCDF were very variable.
Toxic equivalents
TEQ results varied widely among the three sets of factors. Highest values were
consistently produced using the chick embryo hepatocyte derived numbers, followed by Safe’s
and then the WHO TEFs (Table 1.4). The TEQ5WHO ranged from 53 to 2740 ng/kg. Two
birds had liver TEQs110 > 2000 ng/kg, while an additional two birds had liver TEQSWHO >
1000 ng/kg.
27
Tab
le1.
1O
rgan
ochi
orin
ere
sidu
ele
vels
(mg/
kg,
wet
wei
ght)
,ge
omet
ric
mea
n±
95%
conf
iden
cein
terv
als,
inliv
ers
from
Bal
dE
agle
sfo
und
dead
inB
ritis
hC
olum
bia,
1988
-19
93.
Loc
atio
nN
%fa
t%H20
Tot
alPC
Bs
DD
Etr
ans-
Oxy
chlo
rdan
eM
irex
B-H
CH
Die
ldri
nH
CB
nona
chio
r
Low
erF
rase
r10
4.1
710.
609a
0.54
20.
046
0.01
0.00
40.
003
0.00
70.
01
Val
ley
(3.3
-5.1
)(7
0-72
)(0
.253
-1.4
7)(0
.186
-1.5
8)(0
.022
-0.0
96)
(0.0
05-0
.020
)(0
.002
-0.0
09)
(0.0
01-0
.016
)(0
.002
-0.0
3)(0
.006
-0.0
16)
Stra
itof
333.
273
146
ab
1.31
0.09
40.
018
0.00
80.
011
0.01
40.
014
Geo
rgia
(2.8
-3.7
)(7
2-74
)(0
.811
-2.6
2)(0
.057
-1.5
6)(0
.057
-0.1
56)
(0.0
1-0.
031)
(0.0
04-0
.014
)(0
.006
-0.0
21)
(0.0
08-0
.027
)(0
.009
-0.0
2)
John
ston
e9
3.2
753•36b
4.93
0.32
10.
052
0.03
00.
042
0.03
20.
023
Stra
it(2
.0-5
.1)
(73-
77)
(0.6
85-1
6.5)
(0.8
19-2
9.7)
(0.0
58-1
.78)
(0.0
08-0
.337
)(0
.006
-0.1
53)
(0.0
07-0
.270
)(0
.005
-0.1
88)
(0.0
06-0
.097
)
Wes
tC
oast
23.
670
0775
ab
1.01
0.07
30.
013
0.00
70.
013
0.01
30.
011
Van
couv
erIs
.*
Nor
thC
oast
44.
667
0.6
89
’1.
140.
079
0.01
40.
006
0.00
90.
013
0.01
4
(2-1
1)(6
2-73
)(0
.29-
1.64
)(0
.255
-5.0
7)(0
.042
-0.1
49)
(0.0
06-0
.033
)(0
.003
-0.0
15)
(0.0
05-0
.016
)(0
.007
-0.0
23)
(0.0
03-0
.057
)
Nor
ther
n1
2.1
730.
429
0.42
0.06
70.
011
0.00
50.
002
0.00
90.
011
Inte
rior
*
a.b
-m
eans
that
dono
tsh
are
the
sam
ele
tter
are
sign
ific
antly
diff
eren
t(P
<0
.05
).N
OT
E:
sign
ific
ant
diff
eren
ces
amon
gsi
tes
wer
efo
und
only
for
DD
E
*-
insu
ffic
ient
sam
ple
size
toca
lcul
ate
conf
iden
cein
terv
al
NJ
Table 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. = Drowning
Location Date Sex! BC(a) InitialAge Etiology
Total CommentsPCBs
Non-ortho PCBs
#771) #126 #169
(ng/kg) (mg/kg)
Port Hardy 27 Jun/89 F/A 1 Undet. 1270 1170 221 6.42
Port Hardy 6 Mar/90 M/3y 1 Undet. - 349 86 0.425
Port Hardy 2 Apr/90 M!ly 0 Inanit. 2300 4800 2180 43.8
Port Hardy May/93 F/A 1 Inanit. 2000 2550 533 65
Port Hardy May/93 F/A 4 Trauma 1070 1490 368 12
Campbell R. 9 Jun/90 F/2y 5 Tauma - 357 88 0.515
Campbell R. 31 Jul/90 F/3y 0 Trauma 248 688 160 7.52
Campbell R. 16 Apr/93 F/A 4 Electro. 738 9960 2640 71.7
Powell R. 26 Apr/90 M/A 1 Electro. - 5820 1690 60
Powell R. 18 Jun/90 F/A 3 Tox.Pb - 499 130 2.06
Comox 13 Jun/90 M/3y 4 Trauma - 561 148 2.72
Denman Isl. 19 Jul/90 M/A 3 Asphyx. - 426 108 1.26
Bowser 7 Jul/90 F/4y 1 Tox.Pb - 2640 759 25.4
Coombs 3 Mar/90 M/A 1 Tox.Pb - 361 90 0.558
Nanoose 26 Apr/90 F/A 2 Trauma - 496 129 2.02
Nanaimo 8 Feb/90 F/A 5 Electro. 395 370 56 4.56
Sechelt 7 May/90 F/A 3 Electro. - 783 213 5.15
Dent Isl. 5 Apr/93 F/ly 1 Drown. 870 472 138 9.27
Victoria 14 Nov/92 M/A 1 Trauma 1620 2240 523 7.94
*(c)
*
Pb exp.
* Hg tox.
* Pb exp.
* Pb exp.
* Pb exp.
* Pb tox.
* Pb tox.
* Pb exp.
* Hgexp/Pb-exp.
29
Tab
le1.
3C
once
ntra
tions
ofse
lect
edPC
DD
san
dPC
DFs
inB
ald
Eag
leliv
ers
coll
ecte
dfr
omth
eso
uth
coas
tof
Bri
tish
Col
umbi
a(n
g/kg
,w
etw
t.)
2347
8/In
itial
2378
-12
378-
1236
78-
2378
-13
489-
Loc
atio
nD
ate
Sex/
Age
BC(a
)E
tiolo
gyT
CD
DPn
CD
DH
xCD
DT
CD
FPn
CD
FC
omm
ents
Por
tH
ardy
27Ju
ne/8
9F
/A1
Und
et.
3371
8110
11
5*
Por
tH
ardy
6M
ar/9
0M
/3y
1U
ndet
.5
109
20tr
ace
Por
tH
ardy
2A
pr/9
0M
/3y
0In
anit
.77
241
280
4149
*
Por
tH
ardy
May
/93
F/A
1In
anit
.54
232
350
320
*
Por
tH
ardy
May
/93
F/A
4T
raum
a17
4914
145
11*
Cam
pbel
lR
.9
June
/90
F/2
y5
Tra
uma
1825
3733
8C
ampb
ell
R.
31Ju
ly/9
0F
/3y
0T
raum
a30
5676
110
*P
b-ex
p.C
ampb
ell
R.
16A
pr/9
3F
/A4
Ele
ctro
.21
279
321
208
105*
Pow
ell
R.
26A
pr/9
0M
/A1
Ele
ctro
.39
214
2043
603
375
Hg-
exp.
Pow
ell
R.
18Ju
ne/9
0F
/A3
Tox
:Pb
4183
184
6327
Hg-
tox/
Pb-
tox.
Com
ox13
June
/90
M/3
y4
Tra
uma
2151
169
6017
Pb-
exp.
Den
man
Isi.
19Ju
ly/9
0M
/A3
Asp
hyx.
4992
295
7830
Hg-
exp/
Pb-
tox.
Bow
ser
7Ju
ly/9
0F
/4y
1T
ox:P
b26
360
320
5015
152
Hg-
exp/
Pb-t
ox.
Coo
mbs
3M
ar/9
0M
/A1
Tox
:Pb
69
1042
6P
b-to
x.N
anoo
se26
Apr
/90
F/A
2T
raum
a23
5790
2811
Pb-
exp.
Nan
aim
o8
Feb
/90
F/A
5E
lect
ro.
2540
5115
5*
Sec
helt
7M
ay/9
0F
/A3
Ele
ctro
.29
199
936
2413
8H
g-ex
p./P
b-to
x.D
ent
Is!.
5A
pr/9
3F
/ly
1D
row
n.4
1820
835
13*
Vic
tori
a14
Nov
/92
M/A
1T
raum
a30
7710
815
18*
(a)B
C-
Bod
yC
ondi
tion:
0-em
acia
ted,
1-th
in,
2-fa
ir,
3-go
od,
4-ve
rygo
od,
5-ex
celle
nt(b
)*
1348
9-Pn
CD
Fno
tin
clud
ed(c)
trac
e=
<2
ng/k
gw
etw
t.M
DL
-M
inim
umD
etec
tion
Lim
it(s
igna
l/noi
se)
=3
ng/k
gw
etw
t.U
ndet
.=
Und
eter
min
ed,
Inan
it.=
Inan
ition
,E
lect
ro.
=E
lect
rocu
tion,
Tox
.Pb
=L
ead
Tox
icos
is,
Asp
hyx.
=A
sphy
xiat
ion,
Dro
wn.
=
Dro
wni
ngA
=ad
ult,
ly,
2y,
3y=
age
ofsu
badu
lts
Tab
le1.
4C
ompa
riso
nof
TE
Qs
calc
ulat
edfr
omse
lect
edpC
DD
S(a)
,PC
DFs
(b),
no
n-o
rth
oan
dm
ono-
orth
oP
CB
5(d)
leve
lsin
Bal
dE
agle
liver
sco
llec
ted
from
the
sout
hco
ast
ofB
ritis
hC
olum
bia
(ngl
kg,
wet
wt.(
e)).
TE
Qs
Loc
atio
nD
ate
Sex/
Age
BC0
Initi
alE
tiolo
gyS
afe
CEH
OI)
WH
O’
Com
men
ts
Port
Har
dy27
June
/89
F/A
1U
ndet
.85
215
6027
6**)
Port
Har
dy6
Mar
/90
M/3
y1
Und
et.
9520
453
+0)
Por
tH
ardy
2A
pr/9
0M
/3y
0In
anit.
5490
1170
012
20*
*
Port
Har
dyM
ay/9
3F
/A1
Inan
it.41
2074
3083
2*
*
Port
Har
dyM
ay/9
3F
/A4
Tra
uma
1080
1920
302
**
Cam
pbel
lR
.9
June
/90
F/2y
5T
raum
a13
026
883
+
Cam
pbel
lR
.31
July
/90
F/3y
0T
raum
a71
111
4019
7*
*Pb
-exp
Cam
pbel
lR
.16
Apr
/93
F/A
4E
lect
ro.
7460
1310
024
40**
Pow
ell
R.
26A
pr/9
0M
/A1
Ele
ctro
.65
5010
100
2740
+H
g-ex
p
Pow
ell
R.
18Ju
ne/9
0F
/A3
Tox
:Pb
384
677
193
+H
g-to
x/Pb
-exp
Com
ox13
June
/90
M/3
y4
Tra
uma
334
621
155
+Pb
-exp
Den
man
Isl.
19Ju
ly/9
0M
/A3
Asp
hyx.
335
552
205
+H
g-ex
p/Pb
-exp
Bow
ser
7Ju
ly/9
0F
/4y
1T
ox:P
b42
2059
6014
30+
Hg-
exp/
Pb-e
xp
Coo
mbs
3M
ar/9
0M
/A1
Tox
:Pb
9922
160
+Pb
-tox
Nan
oose
26A
pr/9
0F
/A2
Tra
uma
274
491
135
+Pb
-exp
Nan
aim
o8
Feb/
90F
/A5
Ele
ctro
.47
283
212
9*
*
Sech
elt
7M
ay/9
0F/
A3
Ele
ctro
.81
511
8041
7+
Hg-
exp/
Pb-e
xp
Den
tIs
l.5
Apr
/93
F/l
y1
Dro
wn.
282
453
110
**
Vic
tori
a14
Nov
/92
M/A
1T
raum
a11
1021
0039
4**
(a)
2378
-TC
DD
,12
378-
PnC
DD
1236
78-H
xCD
D(b
)23
78-T
CD
F,
2347
8/13
489-
PnC
DF
(e)
PCB
con
ener
s#7
7,12
6an
d12
9(d
)PC
Bco
nen
ers
#118
and
105
(e)
*P
CD
Ij,
PC
DF
s,no
n-or
tho
PCB
sM
inim
umD
etec
tion
Lim
itB
C-
Boc
yC
ondi
tion:
0-em
acia
ted,
1-th
in,
2-fa
ir,
3-go
od,
4-ve
rygo
od,
5-ex
celle
ntçs
nal
/nois
e=
3(T
EFs
from
Ken
nedy
etal
.in
pres
ss
from
safe
1*
*13
489-
PnC
DF
not
inch
ided
inT
EQ
calc
ulat
ions
(I)T
EFs
from
Ahl
torg
etal
.,19
94(k
)PC
Bco
nen
ers
#126
&16
9ca
lcul
ated
from
regr
essi
oneq
uatio
ns;
#77
not
mcl
uded
in‘E
Qca
lcul
atio
nsA
=ad
ult
ly2y
,3y
=ag
eof
suba
dults
Und
et.
=‘U
ndet
erm
ined
,In
anit.
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aniti
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ctro
.=
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ctro
cutio
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ox.
Pb=
Tox
icos
is,
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ion,
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row
ning
Discussion
Chlorinated hydrocarbon levels in livers of Bald Eagles tested for this study were
generally low; however, a small number of birds found dead or debilitated from the Strait of
Georgia or northern Johnstone Strait had elevated PCDDs, PCDFs, PCBs and organochlorines.
Higher liver levels of lipid soluble contaminants in sick or dead birds do not necessarily mean
that their death was a direct result of toxicity due to those chemicals. Most of the eagles with
higher chlorinated hydrocarbon levels were in poor body condition, indicating lipid and
contaminant mobilization. Body weight was negatively correlated with liver organochiorine
levels in other studies (Cooke et al. 1982; Reichel et al. 1980). A variety of factors can
contribute to weight loss, including: seasonal utilization of fat stores, poor foraging abilities of
juvenile birds, a debilitating injury, disease or toxicosis, and the anorexic effects of chemicals
such as lead, dieldrin and TCDD. Discriminating among these factors in a sample of wild
birds is difficult.
Body weight loss per se can, however, be symptomatic of toxicity. A number of the
birds with elevated chlorinated hydrocarbon levels were also lead exposed or poisoned
(Appendix 1.1). Chronic lead-poisoned birds exhibit wasting and extreme loss of body weight
and appear clinically to have starved (U.S. Fish and Wildlife Service 1986). Dieldrin exposure
can induce fasting (Heinz and Johnson, 1982). Weight loss due to fasting, referred to as
wasting syndrome, is the cause of death in acutely TCDD-exposed mammals (Peterson et al,
1984) and birds (Nosek et al. 1992). Bald Eagles with the highest PCDDs/PCDFs and
TEQ5WHO (2440, 2730 and 1430 ng/kg) were found in the vicinity of pulp mills on south east
Vancouver Island. Proximal causes of death were electrocution in two cases and lead poisoning
in the third. However, in one particular case, an adult male eagle from Powell River, total
TCDD-toxic equivalents were calculated to be from 2740 TEQ5WHO to 6550 ng/kg TEQSSafe. A
sample of the solvent extract was tested in a chick embryo hepatocyte bioassay (Kennedy et al.
1993) and TEQs were estimated at 13,100 ng/kg. This bird was also in very thin body
32
condition, perhaps indicating that it suffered from wasting syndrome. There are no data on
tissue levels of TCDD-like compounds which could be diagnostic of acute toxicity. LD50s
reported for 2,3,7,8-TCDD are 240 ng/kg in chicken and 1350-2 180 ng/kg in pheasant
embryos (Peterson et al. 1993). Lethal doses in adult birds are estimated to be one to two
orders of magnitude higher (ibid. 1993).
A pattern of increasing mean contaminant levels in spring partly reflects normal
seasonal lipid dynamics. Late summer and fall deposition and winter mobilization of fat is
typical of temperate climate species, adapted for winter survival (Stalmaster and Gessaman
1984). Seasonal deposition and mobilization of lipids and lipid soluble contaminants such
DDE, PCBs and dieldrin was shown in three species of predatory birds monitored for many
years in Great Britain (Cooke et al. 1982). Starvation and associated lipid and contaminant
mobilization can result from reduced foraging ability caused by debilitating injury or disease.
Starvation without injury or disease should be more common among juvenile birds which,
particularly during their first winter, are less efficient at finding food (Todd et al. 1982).
However, only one of nine eagles with liver DDE levels > 10 mg/kg was a juvenile, a 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 of
greater conservation interest in birds which have reached breeding age. Juvenile eagles,
particularly first-year birds, may also have lower chlorinated hydrocarbon levels than adults, as
they have had less time to reach pharmacokinetic equilibrium with dietary residues, which took
up to two years in Great Lakes herring gulls (Anderson and Hickey, 1976). Juvenile eagles eat
more fish (Stalmaster 1987), which would also tend to have lower contaminant levels than fish
eating birds, which are eaten more often by adults (Chapter 3).
Only one bird had > 100 mg/kg DDE in liver, the level suggested by Cooke et a!.
(1982) as indicative of acute poisoning, although two other birds had liver DDE levels of 91
and 96 mg/kg. None of the birds had PCB levels in livers > 100 mg/kg, considered indicative
of toxicity (Cooke et a!. 1982). One bird had levels of oxychiordane in liver > 2 mg/kg and
trans-nonchior levels > 7 mg/kg. Diagnostic liver levels of oxychiordane are not available;
33
brain levels of 1.1 - 5.0 mg/kg indicate acute toxicity (Stickel et al. 1979). In an earlier
sample of nine eagles found dead, 1969 to 1973, from British Columbia, one bird had 179
mg/kg DDE and 23.7 mg/kg dieldrin (Friis, 1974), well above the level of 5 to 10 mg/kg
dieldrin in liver, indicative of acute poisoning (Cooke et al. 1982). None of the eagles in the
present sample had elevated dieldrin levels, indicating an improvement in dieldrin
contamination of the eagle foodchain. However, the presence of potentially toxic levels of
DDE in livers of British Columbia Bald Eagles more than 20 years after DDT was heavily
restricted in North America raises questions regarding sources. A number of hypotheses have
been suggested in the literature to account for sources of continuing high levels of DDT in the
environment. Recent data show that DDT can persist at high levels in soils and foodchains in
areas of former intensive use or manufacturing (Blus et al. 1987; Elliott et a!. 1994). Eagles
may also acquire some DDT from feeding on migrant waterbirds, which are exposed to
ongoing use in Latin American wintering areas (Fyfe et a!. 1990). Finally, on the Pacific
coast, elevated DDE levels in seabirds, such as storm-petrels, important seasonal prey items of
eagles nesting on their colonies, indicates long-range transport from recent use in Asian
countries (Elliott et a!. 1989). Elevated PCBs in some eagle livers likely originate from
industrial sources in the Georgia basin, as PCBs were significantly elevated in samples from the
Strait of Georgia, compared to other sites in both egg (Chapter 3) and nestling plasma samples
(Chapter 4).
Although most toxic effects of TCDD are thought to be mediated via the Ah receptor, it
is possible that the anorexic effects of TCDD are not Ah-receptor mediated (Tuomisto and
Pohjanvirta 1991). Therefore, it would be interesting to know if any biomarkers of Ah-like
toxicities were activated in eagles with high liver TEQs. Indirect indications, at least of
CYP1A induction, may be inferred from examination of TCDD/TCDF ratios, which varied
greatly between eagles with high versus low TCDD exposure. For example, TCDF levels are
much lower in the three birds with the highest TCDD levels (212, 392, 263 ng/kg in liver); the
mean TCDD/TCDF ratio for those three birds was 74. In contrast, the mean TCDD:TCDF
ratio is 0.17 for the three birds with the lowest TCDD levels (5,6,4 ng/kg in liver). The
34
TCDD/TCDF ratio in the high TCDD birds is also markedly different from ratios observed in
eggs. Mean ratios in eagle eggs were 0.58 at Powell River and 0.32 in Jolmstone Strait (Table
2.1). This shifting ratio may indicate that hepatic cytochrome P450 enzymes have been
induced in birds exposed to elevated TCDD levels; consequently, TCDF has been metabolized
(Van den Berg et at. 1993). A hepatic CYP 1 A cross-reactive protein was shown to be present
and inducible in Bald Eagle chicks (Chapter 3) and should, therefore, also be inducible in adult
eagles. CYP1A1 was recently shown to be the protein responsible for TCDF metabolism in
rats and humans (Tai et at. 1993). Alternatively, higher liver TCDD concentrations in more
highly exposed birds may be evidence of the dose-related increase in liver retention of TCDD,
reported for rats (Abraham et a!. 1988). Inducibility of a hepatic binding protein, possibly
CYP1A2, has been suggested as a mechanism for increased TCDD retention at higher doses
(Van den Berg et a!. 1993).
CYP1A enzymes can also metabolize certain PCB congeners and thus alter the PCB
pattern (Brown 1994). The PCB congener pattern between birds classified as good versus poor
body condition is compared in Figure 1.6. As discussed above, birds in poor condition have
higher chlorinated hydrocarbon levels in liver, because of lipid and contaminant mobilization,
and thus, hepatic P450 enzymes may have been induced. Differences in mean percent total
PCBs were not significantly different for any of the congeners measured (t-test, p <0.05);
however, a consistent trend is apparent, whereby the percent contribution of the lower
chlorinated compounds was consistently lower and the higher chlorinated compounds
consistently higher in the poor condition group. CYP1A induction should increase the
metabolism of non-ortho and mono-ortho PCBs but not those with two or more ortho chiorines
(Brown 1994). In particular, compounds such as PCBs 118 (245-34) and 99 (245-24) and 70
(245-4) which have been suggested as indicators of CYP1A metabolism (Brown 1994), as well
as 60 (234-4) and 101 (245-25), appear lower in the poor condition group.
From this indirect evidence, it appears that at least hepatic CYP1A enzymes were
induced in eagles, suggesting the possible activation of other Ah-mediated processes.
35
20
15
0
C.)0
F°0
4-.
00
5
PCB congeners
Figure 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 (< 5
mg/kg) levels of DDE and PCBs, and even lower levels of other organochiorines. However, a
few birds had DDE levels diagnostic of acute poisoning, more than 20 years after regulatory
restrictions on DDT usage in North America. At least one eagle found near a bleached kraft
pulp mill had liver TEQWHO levels potentially indicative of acute toxicity. Differences in
TCDD/TCDF ratios in birds with high 2,3,7,8-TCDD levels may indicate hepatic cytochrome
P450 induction and TCDF metabolism in those birds.
Because of the selection criteria, samples analyzed for PCDDs and PCDFs were biased
towards birds with a higher probability of such exposure. Nevertheless, 4/19 (21 %) of eagles
tested had > 1,000 ng/kg TEQSWHO in their livers. All of those birds were of reproductive age
0rjZ qç’ of
36
found during the breeding season. This may indicate that acute exposure to TCDD-like
compounds has removed a component of the breeding eagle population in the Strait of Georgia.
Acknowledgements
Dr. K.M. Langelier performed the final autopsies. Working in the laboratory of Dr. R.
Norstrom, M. Simon and H. Won did the chemical analysis. L. Wilson, P. Sinclair and I.
Moul assisted in procurring of carcasses. I thank all those people who submitted birds for the
study. Funding was provided by the Canadian Wildlife Service.
37
Mea
n
S.D
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1010
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69.7
1.06
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22
2
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ctro
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766
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ctro
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452
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8
1.89
2.55
0.87
7
44
4
NO
RT
HE
RN
INT
ER
IOR
Smith
ers
11M
ay/9
0F
Ad.
34.
77In
tra.
Agg
.
BC
-B
ody
Con
ditio
n:0-
emac
iate
d,1-
thin
,2-
fair
,3-
good
,4-
very
good
,5-
exce
llen
tn/
a-
not
avai
labl
eN
D-
not
dete
cted
;de
tect
ion
limit
=0.
0005
mg.
kgw
etw
t
App
endi
x1-
1,co
nt.
Mea
n
S.D
.
N
Joke
rvil
e21
Jan/
91M
Ad.
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00E
lect
ro.
5.1
67.7
0.74
0
Loc
atio
nD
ate
Sex
Age
Wt(
kg)
Initi
alE
tiol
ogy
%F
at%H20
DD
EO
xych
lor.
t-N
onac
hl.
Tot
alPC
Bs
Com
men
ts
22
2
0.00
9
0.01
4
0.00
5
0.01
5
0.02
5
0.01
2
0.00
8
0.01
5
0.00
6 4
0.01
1
0.05
4
0.07
6
0.02
2
0.09
3
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9
0.07
9
0.04
9
0.08
3
0.02
2 4
0.06
7
0.41
4
0.93
2
0.51
8
0.68
3Pb
-exp
.
1.06
1
0.86
8Pb
-exp
.
0.35
9
0.74
3
0.25
9 4
0.42
9Pb
expo
sed
2.1
73.1
0.42
CHAPTER 2
BIOLOGICAL EFFECTS OF CHLORINATED HYDROCARBONS
IN BALD EAGLE CHICKS
This study of embryotoxicily was designed to investigate whether in ovo exposure to
PCDDs, PCDFs and PCBs was impacting hatching success and affecting a variety of
biochemical and morphological parameters in Bald Eagle chicks. The aim of the study was
also to estimate concentrations of PCDDs and PCDFs in Bald Eagle eggs which would be
indicative of no-observed-efffects (NOEL) and lowest-observed-effect (LOEL) levels
The results presented in this Chapter represent an extensive collaborative study with
other laboratories. Contributions of those laboratories and of the principle investigators are
identified in the Materials and Methods, while technical contributions are included in the
Acknowledgements. The concept, study design, field work, statistical analyses, calculations,
graphic representations and other manipulations of data were performed by me. A version of
this chapter has been acceptedfor publication (Elliott et al. in press).
Materials and Methods
Sample collection
Bald Eagle eggs were collected from 20 nests (Figure 2.1). At three sites, Crofton
(designated as location 3), Nanaimo (4,5), Powell River (6-9), sample nests were all within a
25 km radius of a kraft pulp and paper mill, and generally within the effluent impact zone of
the mills, as defined by fisheries closures due to dioxin contamination (Harding and Pomeroy
1990). Eggs were collected from two nests in the Fraser River estuary (Map Nos. 1-2); at least
500 km downstream from where effluent is discharged into the Fraser River from four kraft
pulp and paper mills. An area of the west coast of Vancouver Island, Clayoquot Sound (Map
Nos. 10-14), was used as a reference site; there are no major industrial discharges to the
41
sound, although there is some fish processing and lumber yarding around Ucluelet Inlet.
Further details on pollutant sources to Bald Eagles are discussed in Chapter 3.
Figure 2.1 - Locations where Bald Eagle eggs were collected for artificial incubation.
42
Usually one egg was taken from each nest; the smallest egg in the clutch, presumably
the second egg, was selected. At five nests in the Powell River area both eggs were taken.
Because of the wide variability in nesting dates of Bald Eagles within and among areas,
collecting at each site was scheduled for the estimated midpoint of incubation. The nests were
accessed by a professional tree climber. Eggs were placed initially into a portable thermos.
The temperature was maintained between 25 and 300 C using hotwater bottles, replenished as
required from thermos bottles. Within eight hours of collection, the eggs were transferred into
a battery powered CurfewTM incubator kept at a temperature of 34°C. The eggs were rotated
about hourly and turned on their long axis twice daily.
Within 72 hours (normally within 24 - 48 hours) the eggs were brought to the
laboratory at the Department of Animal Science, University of British Columbia, where they
were candled to determine fertility and placed into a Humidaire incubator maintained at
37.2°C with a relative humidity of 82-84 %. The eggs were rotated once per hour and turned
twice a day in opposite directions on their long axis. At pipping the eggs were placed into a
hatcher.
Sample preparation
Within 24 hours of hatching, the birds were weighed, blood drawn by cardiac puncture
using a heparinized syringe and the bird sacrificed by decapitation. The yolk sac was removed
and frozen. The liver was removed, weighed and separated as follows: 0.25 g from tip of left
lobe for Vitamin A analysis, 0.10 g from tip of the right lobe for porphyrin analysis; these
samples were then frozen. The remaining liver was used to prepare microsomes. Various
organs were removed and morphological measurements performed (Hart et al. 1991): body,
yolk-free body, liver, heart, kidney (sum of both), yolk, stomach, intestine, bursa, adrenal
(sum of both), spleen and tibia (wet, dry, ash) weights and tibia length The following tissues
were fixed in 10 % buffered formalin for histological examination: right kidney, bursa, thymus,
spleen, gonads, lung, heart, intestines, thyroid and adrenal glands. Tissues were processed
routinely and embedded in paraffin blocks. Sections were cut at 6 urn and stained with
43
hematoxylin and eosin and examined by light microscopy. The amount of lymphoid tissue was
estimated based on follicular size and cell density of cortex and medulla in the bursa, on the
density of white pulp in the spleen and on the thickness of the cortex and cell density in the
thymus. The number of mitoses in all lymphoid organs and the number of necrotic cells in the
bursa and thymus were counted in five fields at 600 X magnification. The level of
extramedullary haematopoiesis was assessed in the spleen.
Chemical analysis
Bald Eagle yolk sacs were analyzed for PCDDs, PCDFs and non-ortho PCBs at the
National Wildlife Research Centre, Hull, Quebec, in the laboratory of Dr. R.J. Norstrom. The
analyses were carried out on a VG Autospec high resolution mass spectrometer linked to a HP
5890 Series II data system using‘3C-labeled internal standards after gel permeation/carbon
chromatographic cleanup, essentially as described for livers in Chapter 1. Organochiorines and
other PCBs were determined using GC/MSD (high resolution GC/low resolution MS) (Letcher
et al. in press).
Biochemical assays
Microsome preparation: Microsomes were prepared as described in Bellward et al.
1990. Briefly, livers were homogenized in 25 ml TRIS-KCL buffer using a teflon pestle; the
homogenate was centrifuged at 10,000 g for 20 minutes, the precipitate discarded and the
supernatant further centrifuged at 100,000 g for 60 minutes. The microsomal pellet was
suspended in 20 ml of 10 mM EDTA (ethylenediamine tetraacetic acid), 1. 15% KCL, pH = 7.4,
buffer at 4°C and homogenized; the homogenate was spun in an ultra-centrifuge as described
above and the resulting microsomal pellet resuspended in 0.5 ml of 0.25 M sucrose. Aliquots
of 100 ul were stored in cryovials in liquid nitrogen until assayed.
Cytochrome P450-related activity: Ethoxyresorufin 0-deethylase and benzyloxyresorufin
0-deethylase activity in liver microsomes were determined using the method of Klotz et al.
44
(1984), adapted to a fluorescence multi-well plate reader. The standard reaction mixture for
Bald Eagle microsomes contained 0.1 M TRIS-HC1, pH 8.0, containing 0.1 M NaC1, 10 mM
of MgC12, 2 uM 7-ethoxyresorufin or 1.5 uM 7-benzyloxyresorufin and approximately 200 g
of microsomal protein in a final volume of 500 uL. After a pre-incubation period of 5 minutes
at 37°C, the reaction was initiated by the addition of NADPH (final concentration 0.6 mM) to
the sample well (the blank did not receive NADPH). The reaction was stopped after 20
minutes by the addition of 1.0 ml of cold methanol. The amount of resorufin formed was
measured in a fluorescence plate reader, using an excitation wavelength of 530 nm and an
emission wavelength of 590 nm. Hepatic microsomal total protein was measured using a
modification of Lowry’s method (Peterson 1977).
Immunoblotting: Based on the original western blot method developed by Towbin et al.
(1979), hepatic microsomal proteins were separated on sodium dodecyl sulfate polyacrylamide
gels (SDS-PAGE, 9% acrylamide) and electrophoretically transferred to Rad-free membranes
(Schleicher & Schuell, Keene, NH). Aroclor 1254-induced rat liver microsomes (prepared
from commercially available postmitochondrial supernatant, Molecular Toxicology Inc.,
Annapolis, MD) were used as standards. Immunodetection of CYP1A was performed using
monoclonal antibody 1-12-3 prepared against scup cytochrome P45O1A1 which recognizes
CYP1A in all taxonomic groups of vertebrates examined so far (Park et al. 1986, Stegeman
1989). The secondary antibody was a goat anti-mouse IgG linked to alkaline phosphatase.
Immnunoreactive proteins were detected by chemiluminescence (Rad-Free, Schleicher &
Schuell, Keene, NH) and the light intensities of the inimunoreactive protein bands were
quantified by video imaging densitometry (UVP Gel Documentation System 7500, San Gabriel,
CA). This work was carried out in the laboratory of Dr. S.W. Kennedy.
Cytoebrome P4502B (CYP2B) levels were determined by protein immunoblotting using
rabbit polyclonal antibody 7-94 against scup P450B (a CYP2B like protein), which recognizes
CYP2B proteins (Stegeman 1989). Methods were as described above, but with Bio-Rad goat
anti-rabbit alkaline phosphatase-linked secondary antibody and using NBT (Nitro blue
tetrazolium) and BCIP (5-bromo 4-chioro 3-indoyl phosphate) for colour development. 30 g
45
of samples were loaded in each well. Scup microsomes containing known amounts of P450B
were included for quantitation in each gel. Since equivalence of cross-reactivity for the antibody
between scup and eagle is unknown, numbers are relative and not absolute. Scup standards
insure the linearity of response of the system and are necessary for normalizing between blots
and runs. Analysis of developed blots was performed using a Kodak DCS 200 digital camera
system and the NIH Image 1.55 densitometry software. This assay was performed in the
laboratory of Dr. J.J. Stegeman.
Liver vitamin A analysis: Samples of liver (300 to 500 mg) were dehydrated to a pink
powder by grinding with anhydrous sodium sulphate. The internal standard, retinyl acetate (40
ng/20 uL methanol) was added to the equivalent of 0.20 g of liver and the vitamin A
compounds were extracted with 10 mls of a 1:9 dichloromethane:methanol solvent mixture in
an amber vial. After centrifugation (10 mins at 600 rpm at 10°C) the supernatant was filtered
through a 0.2 urn Acrodisc LC13 PVDF filter (Gelman) and a 20 ul aliquot was analyzed in
duplicate by non-aqueous reverse phase HPLC. Separation of retinol, retinyl acetate and
retinyl palmitate was achieved with a 15 cm long, 5 urn ODS Zorbax column with 100 %
methanol at 1 ml/min for 5.5 minutes followed by a linear gradient which brought the mobile
phase to 30 % dichioromethane and 70 % methanol within 0.5 mm. This composition was held
until the end of the run at a flow rate of 2.0 ml/min. With these conditions, retinol, retinyl
acetate and retinyl palmitate had retention times of 3.1, 4.2, and 9.7 minutes, respectively.
Plasma vitamin A analysis: The internal standard, retinyl acetate was added to 100 ul of
serum. The retinol-protein complex was dissociated by the addition of 200 ul of acetonitrile.
The retinol was extracted twice using 4 mIs and 1 ml of hexane. The organic and aqueous
phases were separated by centrifugation, and the combined organic phases were evaporated to
dryness under a stream of pure nitrogen. The residues were reconstituted in 1 ml of methanol,
filtered through a 0.2 urn Acrodisc LC13 PVDF filter (Gelman) and a 50 uL aliquot was
analyzed in duplicate by HPLC using the colunrn described above for liver. With 100 %
methanol as the mobile phase and a flow rate of 1 ml/min, retinol and retinyl acetate had
retention times of 3.3 and 4.5 mm, respectively.
46
Hepatic porphyrins: Porphyrin levels in liver were determined using the method of
Kennedy and James (Kennedy and James 1993). This method involves extraction in duplicate
using a mixture (1:1) iN hydrochloric acid/acetonitrile. The porphyrins were then concentrated
on Sep-Pak Plus t C18 cartridges followed by separation and quantification by HPLC.
Statistical analysis
The SYSTAT software package was used for statistical analyses of all data. Data are
presented on a lipid weight basis as suggested by Hebert and Keenleyside (1995), when there
are significant relationships between wet weight contaminant concentrations and percent lipid.
For example, using only data from pulp mill sites (to minimize the influence of location),
2,3,7,8-TCDD concentrations (wet weight) in yolk sacs were highly significantly correlated
with percent lipid (linear regression, r2 =0.772, p < 0.0001, N = 11). Chemical residue data
were transformed to common logarithms and geometric means and 95 % confidence intervals
were calculated with the data grouped by collection site. Contaminant levels were compared
among location with a one-way analysis of variance (ANOVA); significant differences were
determined using Tukey’s multiple comparison procedure (MCP). Data were also compared
on the basis of a pulp mill versus non-pulp mill grouping and significant differences identified
using Student’s t-test. In order to avoid a bias, for comparison among sites and between pulp
mill and non-pulp mill sites, only the results from the second or smallest egg were used from
the Powell River nests, thus giving a total sample size of 14. Concentration-effect relationships
were determined using coefficients of determination (r2) using least-squares linear regression.
Unless stated otherwise, a value of p < 0.05 was considered statistically significant in all
analyses.
TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalency factors
proposed by Ahlborg et al. (1994), and referred to here as the WHO (World Health
Organization) TEFs. For comparison, TEFs proposed by Safe (1990) and Kennedy et al. (in
press) were also used.
47
Results
Chemical contaminant levels
PCDDs and PCDFs. Data are presented on an individual nest basis in Table 2.1. The
eight PCDDIPCDF congeners which exhibited significant differences among sites are grouped
and compared in Figure 2.2. Congeners with a 2,3,7,8-substitution pattern were dominant;
however, there were traces of l,2,3,4,6,7,9-HpCDD (5 - 10 ng/kg) in some yolk sacs from
Powell River, in both yolk sacs from the Fraser estuary and in the yolk sac from Nanaimo.
Likewise, trace amounts of 1,2,4,6,7,8-HxCDF (5 - 10 ng/kg), 1,2,4,6,8,9-HxCDF (10 - 100
ng/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 the
yolk sacs from Powell River and were significantly higher than in yolk sacs from the Fraser
Delta or west Vancouver Island. Those same major congeners were not statistically separable
between Powell River and east Vancouver Island. Comparison between pulp mill (Powell
River + east Vancouver Island) and non-pulp mill (Fraser Delta + West Vancouver Island)
showed that concentrations were significantly higher (p < 0.005) at pulp mill sites for all the
congeners in Figure 2.2, except 1,2,3,4,6,7, 8-HpCDD. Although not statistically different
from other sites, highest concentrations of 1,2,3,4,6,7,8-HxCDD and OCDD (331 ng/kg) were
in yolk sacs from the Fraser Delta.
48
Figure 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 five
analyses per collection site along with the 95 % confidence interval. Means which do notshare the same lower case letter were significantly different (p < 0.05).
TCDD2500
2,000
1,500
1,000
500
1120 12378-PCDF101., I—160 a140 : : : S
120
100
80ab
4OabI b
CDD
4-
.:0)U)
4-U)
C)
0)z
200
150
100
50
0
16700 117400 123678-HCDD
8,00C
6OOC
4,00C
2,00C
ab
5800 1996
cT
1 234678HCDD01)1.,
400 a
300
a
200
10:
I,\C3• —‘s, 4. .*1 \•
49
)C
•..
g0
’
%to
talP
CB
s-
N)
CC
D0
j’-
CD0
(11
C)
(71
0vr
:u.
76
?:Q
D0
0o
__
CD-‘
CD>
____
W
___
00
..
V.
0
___
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-eo
::
::
c—
77
::
:CD
.
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47
..
.
.
C)
C)
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c_)
,v
7.
..
.—
CD..
Zç
Cl)O
..
:(•
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CDC
CD
_
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CDCD
CD•
7.
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cj
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tz
CD—
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CD
__
__
__
__
__
______
0c-
a4)
:-•
oCD
S
_
-:
:.
0o
__
__
__
__
__
__
..
_
—.
Z.
.CD
——
C•)
—“‘-
-•.
..
.C
)\
.1‘.‘
__
__
_
E.CD
•
:
CD7
..
.
CDCD
77
.:
C)—
.
CDCD
..
t__
4e
7...
..
.CD
C.’
—•
——
.)
..
CD.
CD..
—•
.-
—0
-.
•-
-
.0
o0-
.—
•
_
C)
CDCD
C’.9
-0
5
OL?
.
0
Ii
Tab
le2.
1PC
DD
and
PCD
Fco
ncen
trat
ions
(ngl
kg,
lipid
wei
ght
basi
s)in
yolk
sacs
ofB
ald
Eag
lech
icks
coll
ecte
din
1992
from
Bri
tish
Col
umbi
a.
Map
Loc
atio
n%
fat
2378
1237
812
3678
1237
8912
3467
8O
CD
D23
78-
1247
8-12
378-
2347
8T
otal
Tot
alN
o*.
TC
DD
PnC
DD
HxC
DD
HxC
DD
HpC
DD
TC
DF
PnC
DF
PnC
DF
PnC
DF
HxC
DF
HpC
DF
1B
runs
wic
kPt
.11
868
1340
2400
39.4
144
102
661
43.8
48.3
348
100
24.2
2R
iver
Rd.
1464
610
1032
0088
.410
2310
7623
837
.426
.617
026
416
9
3C
roft
on11
1800
2510
8400
152
430
575
923
30.0
47.6
375
130
59.1
4Ja
ckP
oint
1114
5020
3040
5098
.525
618
729
3099
.264
.226
731
634
2
5W
inch
else
aIs
.10
2460
3950
9380
149
93.2
83.4
2190
62.8
93.2
820
163
52.0
6B
all
Pt.
8.8
1130
1170
6020
140
67.3
83.2
1670
60.4
39.9
339
124
62.7
6B
all
Pt.
1311
2011
3042
3013
962
.969
.218
8010
153
.629
610
010
3
7E
vend
enPo
int
1426
7026
3083
9023
591
.460
.418
800
324
305
823
306
278
7E
vend
enP
oint
1835
6038
5010
900
261
127
79.6
2410
080
.842
412
2028
031
.5
8Sc
uttle
Bay
1827
0038
4086
7017
065
.779
.816
700
15.2
217
1040
153
16.9
8Sc
uttle
Bay
2224
5036
3080
2018
964
.676
.711
590
32.1
212
933
153
26.3
9L
imek
ilnB
ay23
1470
2400
9630
150
35.4
49.2
6900
82.0
114
651
135
82.0
9L
imek
iln
Bay
1222
5036
7012
800
272
68.6
83.7
7970
49.4
178
953
145
42.5
10B
awde
nB
ay23
218
409
504
ND
16.2
37.7
672
13.5
22.9
88.9
17.5
4.04
11W
hite
Pine
113
306
450
394
18.1
15.9
46.9
305
17.6
23.4
86.2
21.8
5.86
12W
hite
Pine
416
353
675
567
ND
ND
66.5
465
50.3
34.2
130
39.8
ND
13T
horn
ton
Cr.
2062
910
7011
5019
.149
.313
710
7052
.552
.520
714
169
.9
14M
erca
ntil
eC
reek
1632
339
553
411
.935
.260
.936
013
.819
.591
.047
.74.
39
*1-
2F
rase
rD
elta
;3-
5E
ast
Van
couv
erIs
land
;6-
9Po
wel
lR
iver
;10
-14
Wes
tV
anco
uver
Isla
nd
Table 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 in
Bald Eagle yolk sacs are presented in Table 2.3. As with the PCB congeners, no significant
differences in mean concentrations occurred among sites for any of the organochiorine
compounds. The pattern was relatively consistent among yolk sacs with total PCBs > DDT
related > chiordane-related > dieldrin > B-HCH > HCB > mirex. The exception to this
pattern was the yolk sac from White Pine Cove No. 1, where DDE levels were greater than
total PCBs. The PCB/DDE ratio was generally much lower in yolk sacs from the west coast of
Vancouver Island than from other sites.
Artificial hatching success and condition of embryos
A total of 25 Bald Eagle eggs were collected for incubation, of which one was cool to
the touch at the time of collection (there was a recently hatched chick in the nest) while a
52
Tab
le2.
3O
rgan
ochi
orin
eco
ncen
trat
ions
,ge
omet
ric
mea
nsan
d95
%co
nfid
ence
inte
rval
s,(r
ange
inbr
acke
ts)
inyo
lksa
csof
Bal
dE
agle
chic
ksco
llect
edin
1992
from
Bri
tish
Col
umbi
a.
Loc
atio
nO
rgan
oclo
rine
cent
ratio
n(m
g/kg
,lip
idw
eigh
tba
sis)
NT
otal
DD
Eir
ans-
oxyc
hlor
dane
hept
achi
or-
Die
ldri
nM
irex
B-H
CH
HC
BPC
Bs
nona
chio
rep
oxid
e
Fra
ser
Del
ta2
364
73.5
3.82
0.91
1.69
2.30
0.45
0.68
0.62
113-
1160
20.3
-267
1.54
-9.4
50.
12-7
.01
1.25
-2.3
00.
53-9
.96
0.10
-2.0
0.22
-2.0
90.
35-1
.10
(278
-477
)(5
4.5-
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
)
Eas
tV
anco
uver
355
912
79.
461.
121.
152.
010.
631.
911.
13Is
land
450-
694
83.3
-194
5.86
-15.
30.
72-1
.74
0.93
-1.4
21.
48-2
.73
0.33
-1.1
90.
96-3
.83
0.46
-2.7
8(4
90-6
15)
(97.
5-14
7)(7
.56-
12.6
)(0
.85-
1.29
)(1
.07-
1.31
)(1
.74-
2.41
)(0
.42-
0.71
)(1
.53-
2.96
)(0
.64-
1.62
)
Pow
ell
Riv
er8
400
93.9
13.5
1.13
0.99
1.59
0.68
1.31
0.70
211-
760
65.1
-136
9.44
-19.
20.
79-1
.62
0.64
-1.5
20.
60-4
.22
0.49
-0.9
50.
89-1
.92
0.43
-1.1
6(1
17-1
052)
(51.
2-19
4)(6
.47-
22)
(0.6
2-2.
37)
(0.6
1-1.
54)
(0.3
5-2.
68)
(0.4
7-1.
20)
(0.8
0-2.
02)
(0.2
5-1.
15)
Wes
tV
anco
uver
519
212
59.
770.
890.
782.
10.
661.
190.
75Is
land
70.5
-524
63.1
-247
5.4-
17.7
0.58
-1.3
60.
64-0
.96
1.44
-3.0
80.
42-1
.02
0.67
-2.1
20.
53-1
.07
(63.
9-48
8)(4
6.6-
206)
(4.7
-15)
(0.4
8-1.
24)
(0.7
5-0.
82)
(1.9
2-2.
30)
(0.4
4-1.
08)
(0.6
0-1.
89)
(0.5
7-1.
24)
second egg was possibly shaken as it was lowered from the nest. Therefore, 23 of the eggs
were possibly viable when placed into the incubator. A total of 18 eggs hatched for an overall
success rate for artificial incubation of 78.3 %. Eliminating the possibly shaken egg from
Northwest Bay, 16 eggs were collected from pulp mill sites of which 11 hatched for a hatching
rate of 69 % (Table 2.4). Of eight eggs collected from non-pulp mill sites, seven hatched for a
hatching rate of 88 %. This difference in hatching success between pulp mill and non-pulp mill
sites was not, however, significant (Chi2 test). One chick (Ball Point A) was edematous at
hatching. Of the eggs which failed to hatch, one was infertile (Powell River area), two were
addled (both from the same nest in the Powell River area), three were early (first quarter of
development) embryos (one each from east Vancouver Island, Powell River and west
Vancouver Island) and one was a late (last quarter of development) embryo (Alberni Inlet).
Table 2.4 Outcome of artificial incubation of Bald Eagle eggs collected from BritishColumbia, 1992.
Location Treatment No. No. %collected hatched success
Fraser Delta Non-pulp mill 2 2 100
West Vancouver Island Non-pulp mill 6 5 83
(Mean, non-pulp mill) 8 7 88
East Vancouver Island Pulp mill 3b 3 100
Powell River Pulp mill 12 8 67
Alberni Inlet Pulp mill 1 0 0
(Mean, pulp mill) 16 11 69
a pulp mill versus non-pulp mill difference not significant, chi2 = 1.402b 4 eggs were collected, 1 was eliminated as possibly shaken
54
Morphological and histological measurements
No significant differences occurred among sites for mean values of any of the measured
morphological parameters, whether expressed as actual values or as percent yolk-free body
weight. For the 18 chicks measured morphological measurements (mean ± SD) were as
follows: body weight (88 ± 9.4 g), yolk-free body weight (83 ± 8.3 g), liver (1.9 ± 0.29 g), right
kidney (0.70±0.12 g), intestine (2.1±0.27 g), heart (0.56±0.09 g), adrenal glands
(0.04±0.02 g), spleen (0.077±0.025 g), bursa (0. 152±0.039 g), yolk (5.7±2.3 g), thyroid
glands (0.075±0.024 g), dry tibia weight (0.057±0.005 g), tibia length (26.8 ±0.85 mm),
tarsus length (20.5±1.48 mm), wing chord (29.4 ± 2.0 mm). Selected parameters are
compared among sites in Appendix 2.1.
For the tissues examined histologically, variations among individual birds were seen
only for lymphoid organs (Table 2.5). Variations were observed within and between sites in
amount of lymphoid tissue, the number of cells in mitosis, the number of necrotic cells and the
degree of extramedullary hematopoiesis. However, no significant differences among sites
occurred for mean values of any of the measured parameters. The amount of lymphoid tissue
in the spleen was constant among individual birds.
Table 2.5 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 ± 0.84
No. necrotic cells” 90 ± 28 109 ± 28 142 ± 85 105 ± 30
No. cells in mitosisb 29 ± 11 43 ± 5.5 50 ± 3.1 40 ± 14
Spleen No. cells in mitosisb 15 ± 9.2 16 ± 7.5 19 ± 6.7 6 ± 3.9
Degree of E.M.C 1.5 ± 0.71 1.3 ± 0.58 2.2 ± 0.64 1.4 ± 0.55
Thymus Amount of lymphoid tissue’ 3.0 ± 0 3.7 ± 0.58 2.9 ± 3.8 2.0 ± 0.71
No. necrotic cells’ 28 ± 3.5 54 ± 21 66 ± 19 64 ± 18
No. cells in mitosisb 10 ± 8.5 24 ± 22 11 ± 5.4 21 ± 8
a- 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).
55
Biochemical measurements
Mean concentrations of CYP1A were sixfold greater (p <0.05) in chicks from Powell
River compared to west Vancouver Island (Table 2.6). Mean concentrations of a CYP2B-like
protein were two to three-fold higher in livers from Strait of Georgia sites compared to west
Vancouver Island; however, the differences were not significant. Mean EROD activity was
eight-fold higher in east Vancouver Island compared to Fraser delta and mean BROD activity
was nearly nine-fold higher in Powell River than Fraser delta chicks; however, the differences
were not significant, likely in part due to small sample sizes and large variabilities. However,
both hepatic EROD and BROD were significantly induced, if datafor all chicks collected near
pulp mills were pooled compared to non-pulp mills sites (p <0.0005 and p < 0.02,
respectively).
Mean uroporphyrin and Vitamin A levels did not differ significantly among sites,
although liver retinyl palmitate levels were about one-half in chicks from the Fraser delta
compared to west Vancouver Island.
Table 2.6 Measurement of hepatic cytochrome P450 and porphyrin parameters and 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.
56
Concentration-effect relationships
Data from the complete set of 18 Bald Eagle chicks were used to examine relationships
between measured biological parameters and contaminant exposure. The gradient of exposure
from lowest to highest was 16-fold for 2,3,7,8-TCDD and 80-fold for 2,3,7,8-TCDF.
Regression analysis was performed using both normal and log-transformed chemical residue
data; results are presented in Table 2.7 for each parameter based on which form of the residue
data gave the best fit (greatest r2 value) to the regression curve.
Highly significant positive regressions were found between hepatic CYP1A and most of
the individual PCDD, PCDF and PCB compounds in yolk sacs; however, the best fits were
with log 2,3,7,8-TCDF and 2,3,7,8-TCDD (Table 2.7, Figure 2.4). No significant regressions
were found between a CYP2B-like protein and yolk sac concentrations of any of the chemical
parameters measured. For EROD, the best r2 value was with 2,3,7,8-TCDD, while the
strongest regression for BROD was found with log 2,3,7,8-TCDF. Hepatic urophorphyrin also
showed a significant positive regression on 2,3,7,8-TCDD, log-2,3,7,8-TCDF and log-TEQs.
Hepatic retinyl-palmitate levels showed a weakly significant positive regression with log-PCB
126, but not with any other chemical parameters. The hepatic cytochrome P450 and porphyrin
parameters all regressed more strongly with either 2,3,7, 8-TCDD or log 2,3,7, 8-TCDF than
with TCDD-TEQ5 estimated using three different TEFs (Table 2.8).
Among the morphological parameters, a weakly significant positive regression was
found between yolk-free body weight and log PCB 126. Yolk sac weight negatively regressed
with both total PCBs and log TEQs. A weakly significant positive regression was also
determined 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)
E
E020
0uJ
(B)
0 500 1,000 1 .500 2,000 2,500 3,000 3,500
2378-TCDD (nglkg, lipid basis)
()1
0)EC
202
0
I I I I
2,3,7,8-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 *
500
40
1,000 1 500 2,000 2,500
2378-TCDD (nglkg, lipid basis)
3.000 3,500
r2 = 0.721
*
1,000 10,000
58
Tab
le2.
7C
once
ntra
tion-
effe
ctre
latio
nshi
psbe
twee
nbi
oche
mic
alan
dm
orph
olog
ical
mea
sure
men
tw
ithch
lori
nate
dhy
droc
arbo
nle
vels
inyo
lksa
csof
Bal
dE
agle
chic
ks.
1T
EQ
s,ac
cord
ing
toA
hlbo
rg(1
994)
NS
-not
sign
ific
ant
ci,
Par
amet
er2,
3,7,
8-T
CD
DL
og2,
3,7,
8-T
CD
FL
ogPC
B12
6T
otal
PCB
sL
ogTE
Qsw
HO
1
NSl
ope
r2p
r2p
r2p
r2p
r2p
CY
P1A
14(+
)0.
850
<0.
0001
0.88
7<
0.00
010.
371
<0.0
30.
576
<0.0
02
0.72
8<
0.00
05
CY
P2B
14(+
)0.
082
NS
0.11
4N
S0.
057
NS
0.25
5N
S0.
136
NS
ER
OD
18(+
)0.
748
<0.
0005
0.70
8<
0.00
050.
297
<0
.02
0.58
8<
0.00
010.
633
<0.
0005
BR
OD
13(+
)0.
601
<0.
002
0.72
1<
0.00
050.
396
<0.0
30.
346
<0.
006
0.54
9<
0.00
4
Uro
porp
hyri
ns17
(+)
0.31
6<
0.0
20.
298
<0
.02
0.19
4N
S0.
122
NS
0.23
2<
0.0
5
Ret
inyl
-18
(+)
0.05
9N
S0.
028
NS
0.26
<0.0
30.
097
NS
0.20
2N
Spa
lmit
ate
liver
Yol
k-fr
eebo
dy18
(+)
0.03
2N
S0.
042
NS
0.24
7<
0.0
40.
023
NS
0.07
2N
Sw
eigh
t
Yol
ksa
c18
(-)0.
034
NS
0.04
5N
S0.
098
NS
0.26
9<
0.0
30.
128
NS
Thy
mic
18(+
)0.
191
NS
0.09
1N
S0.
058
NS
0.09
NS
0.25
0<
0.0
4ly
mph
oid
tiss
ue
Table 2.8 Comparison of regression (r2) values of some hepatic biochemical parameters onTEQs derived from three sets of toxic equivalence factors (TEFs).
Toxic Equivalent Factors
Parameter TCDD/F’ Safe2 CEH3 WHO4
P450 1A 0.887 0.687 0.759 0.805
EROD 0.748 0.529 0.607 0.633
BROD 0.601 0.427 0.515 0.549
Uroporphyrin 0.316 0.107 0.162 0.232
‘Best 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. 1994
Discussion
Bald Eagle chicks collected from nests near pulp mills were exposed to elevated
concentrations of potent embryotoxic PCDD and PCDF congeners, compared to chicks from
reference nests. Symptoms of TCDD-like exposure, such as have been observed in field
studies of fish-eating birds (Hoffman et at. 1986; 1987; Kubiak et at. 1989; Bosveld et at.
1994; Van den Berg et at. 1994b; Elliott et at. 1989a; Bellward et at. 1990; Hart et at. 1991;
Sanderson et at. 1994a; Whitehead et at. 1992b), were not found in Bald Eagle chicks.
Laboratory hatching success did not differ between eggs from pulp mill versus reference sites.
However, hepatic CYP1A levels were significantly higher in eagle chicks from pulp mill sites
and regressed positively on yolk sac concentrations of 2,3,7,8-TCDD and 2,3,7,8-TCDF.
Induction of CYP1A can be linked primarily to PCDDs and PCDFs acquired by the female
parent from local sources, as breeding Bald Eagles on the Pacific coast are year round residents
(Hancock 1964). Yolk sacs contained high concentrations of the toxic non-ortho PCBs, 126
and 77, although regressions with biochemical and morphological parameters were weak and
inconsistent compared to TCDD and TCDF. Concentrations of total PCBs and other
organochiorines in eagle yolk sacs also varied little among sites.
60
Laboratory hatching success
Except for one edematous chick, no signs were apparent in either the hatched eaglets or
in failed eggs of GLEMEDS (Great Lakes embryo mortality, edema, and deformities
syndrome) (Gilbertson et at. 1991), such as reported for fish-eating birds in the Great Lakes
and elsewhere (Hoffman et at. 1986; 1987; Kubiak et at. 1989; Bosveld et a!. 1994; Van den
Berg et at. 1994b; Elliott et at. 1989a; Bellward et at. 1990; Hart et a!. 1991; Sanderson et a!.
1994a; 1994b; Whitehead et a!. 1992b; White and Seginak 1994), which is similar to the toxic
syndrome caused by TCDD in chicken embryos. In embryos of other avian species, such as
ring-necked pheasants (Phasianus cotchicus), mortality is the most sensitive response to TCDD
exposure and the symptoms seen in chickens at lower doses are not observed (Nosek et at.
1993). However, there were no significant differences in laboratory hatching success of eagle
eggs among sites or between pulp mill and non-pulp mill areas. The overall artificial hatching
success of 78.3 % was comparable to the average of 75 % (range 62 - 87 %) reported for wild
and captive Bald Eagles from a number of studies (Stalmaster 1987). The absence of
deformities and other GLEMEDS symptoms in Bald Eagle chicks from this study is likely 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 total
PCB levels than any of the fresh eggs from the Strait of Georgia.
Patterns and trends of PCDD, PCDF and PCB contaminants in yotk sacs
Local pulp mill and chiorophenol inputs account for the particular pattern and elevated
levels of 2,3,7,8-substituted PCDDs and PCDFs in Bald Eagles and other wildlife from the
Strait of Georgia (Elliott et at. 1989a; Whitehead et at. 1990; 1992b), compared to similar
samples from other North American and European sites (Van den Berg et at. 1994b; Yamashita
et at. 1993; Hebert et at. 1994). In particular, Bald Eagle yolk sacs contained high
concentrations of 2,3,7, 8-TCDF, which is reported elsewhere at only nominal levels in wildlife
samples. High TCDF levels such as in the eagle yolk sacs from Powell River reflect exposure
to prey items contaminated by local pulp mill discharges (Harding and Pomeroy 1990).
61
Elevated TCDF levels have also been reported in tissues of common mergansers (Mergus
merganser) and herring gulls breeding near a bleach kraft pulp mill in Quebec (Champoux
1993). Assuming that 2,3,7,8-TCDF should be cleared quickly from the body (Braune et al.
1989; Norstrom et al. 1976), the presence of this chemical in eggs likely results, therefore,
from recent exposure and direct yolk deposition of contaminated lipids as suggested previously
for herons (Elliott et al. 1 989a). Accumulation of TCDF in eagle tissues is probably not linked
to the low absolute EROD activity found in Bald Eagle chicks (Table 6); a recent study
compared EROD induction with in vitro capability to metabolize PCB 77, and concluded that
low EROD activity does not reflect reduced capability to metabolize typical CYP1A substrates,
such as PCB 77 or 2,3,7,8-TCDF (Murk et al. 1994).
Recent exposure and direct shunting of dietary lipids to the yolk may also explain the
presence of non-2,3,7,8 substituted PCDDs and PCDFs in eagle yolk sacs. Fish are able to
metabolize most compounds of this type (Sjim et al. 1989), leading to low levels in the diet of
fish-eating species; birds are also likely capable of further metabolizing them. The presence of
elevated levels of 1,2,3,4,6,7,8-HpCDD and OCDD in the yolk sac from River Road in the
Fraser River delta is consistent with reports of high concentrations of those contaminants in
sediments from near the nest site (Tuominen and Sekela 1992). Elevated levels of higher
chlorinated dioxins in Fraser estuarine sediments are indicative of the intensive past use of
chlorophenol wood preservatives at industrial sites in the Fraser delta (Drinnen et al. 1991).
In contrast to the well-defined local point sources of PCDDs and PCDFs, the uniformity
among sites in concentrations of PCBs and other organochiorines in eagle yolk sacs reflects the
importance of diffuse atmospheric inputs for those compounds (Elliott et al. 1989b). The
geographically uniform PCB congener pattern contrasts with the finding of significant
differences in the percent contribution of certain congeners in great blue herons between
Crofton and Vancouver in 1987 (Elliott et al. 1989a). Because of their restricted seasonal
movement and diet, herons appear to be better indicators of local PCB contamination than
eagles.
62
Biochemical responses
The results of this study confirm for another avian wildlife species the value of CYP1A
induction, particularly as measured by western blotting, as a sensitive marker of exposure to
TCDD-like compounds. Absolute EROD activities in these embryonic Bald Eagle microsomes
were low, although the overall degree of induction from lowest to highest exposure groups,
from six to eight fold, was the same as that observed for other species such as cormorants and
herons (Sanderson et at. l994a; Whitehead et al. 1992b). Interspecific variation of this type is
not surprising as there is increasing evidence that cytochrome P450 isoforms vary substantially
even among closely related species (Yamashita et at. 1992).
Absolute BROD activity was about five-fold higher than EROD in livers of Bald Eagle
chicks, while differences in rates from least to most contaminated individuals was similar for
the two activities. As with EROD, the best 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 activity
in phenobarbital-induced rats (Burke et at. 1994). However, Rattner et al. (1993) recently
reported that, while phenobarbital treatment of black-crowned night-heron embryos caused a
2,000-fold increase in a CYP2B-like protein, there was only a threefold increase in BROD
activity. In contrast, 3-methylcholanthrene treatment increased BROD six to fourteen-fold.
Based on that work and other recent reports (Yamashita et at. 1992), isoforms cross-reactive
with putative fish CYP2B and rat CYP2B are present in at least some groups of birds, but the
substrate specificities may be quite different. The results suggest the presence of a CYP2B
isoform in Bald Eagles. Although Bosveld and Van den Berg (1994) in a recent review
concluded that there is no evidence of chlorinated hydrocarbon-inducible CYP2B isoforms in
birds which cross-react with mammalian CYP2B antibodies, further experiments using purified
CYP enzymes and antibodies are required for a better understanding of substrate specificities.
Uroporphyrin levels in chicks from the various sites were similar. Although there was
a significant concentration-effect relationship between uroporphyrin levels and both 2,3,7,8-
TCDD and -TCDF, this finding must be treated cautiously as normal uroporphyrin levels in
avian livers range from 5-25 pmol/g (Fox et al. 1988). PCBs have been reported to cause
63
accumulation of porphyrins in chick embryo hepatic cell cultures (Kennedy et al. 1995) and in
liver and other tissues of adult birds of common laboratory species (Elliott et al. 1990), but not
apparently in captive predatory birds (Elliott et al. 1991). In previous field studies, hepatic
porphyrins were elevated in adult herring gulls from more polluted areas of the Great Lakes
(Fox et al. 1988), but not in great blue heron embryos exposed to elevated PCDDs and PCDFs
(Beliward et al. 1990).
Plasma and liver retinoid concentrations and the molar ratios of retinol to retinyl
palmitate did not differ among sites, although a weakly significant positive relationship between
hepatic retinyl-palmitate and PCB 126 (34-345) was found. In contrast, laboratory data for rats
report that PCDDs, PCDFs and PCBs caused depletion of liver retinoid stores (Chen et al.
1992). In field studies, such as with herring gulls in the Great Lakes, yolk retinoids varied
among colonies and the molar ratio of retinol to retinyl palmitate correlated positively with
TEQs in eggs (Spear et al. 1990). Van den Berg et al also reported a non-significant reduction
in hepatic retinyl palmitate in cormorants from a contaminated relative to a reference site in the
Netherlands.
Morphological and histological parameters
Morphological and histological measurements did not differ among sites. However, as
with retinyl-palmitate and PCB 126, a number of weakly significant exposure-response
relationships occurred, at the p < 0.05 level (Table 2.8), which are likely not meaningful
biologically. For example, yolk-free body weight appeared to increase with PCB 126 levels in
yolk sacs; this contrasts to data from a number of field studies which report statistically
significant negative relationships between PCDDs/PCDFs or PCBs and embryonic weight and
other morphological characteristics (Hoffman et al. 1986; 1987; Van den Berg et al. 1994b;
Hart et al. 1991; Sanderson et al. 1994a). The negative relationship between PCBs and yolk
weight is consistent with similar findings for cormorants from British Columbia (Sanderson et
al. 1994b), but contrasts with reports of a positive relationship for cormorants from the
Netherlands (Van den Berg et al. l994b). The positive relationship between density of thymic
64
lymphoid tissue and log 2,3,7,8-TCDD is in contrast to reports from a number of laboratory
studies that TCDD and related compounds cause atrophy of the thymus with depletion of
lymphocytes (Elliott et al. 1990; Nikolaides et al. 1988). Therefore, it is likely that these
findings in Bald Eagles are spurious in nature due in part to the relatively small sample size and
large number of variables analyzed.
Comparison of toxic equivalents
As reported previously for great blue heron chicks in the Strait of Georgia (Bellward et
al. 1990; Sanderson et al. 1994a), regression of the biochemical endpoints against 2,3,7,8-
TCDD or 2,3,7, 8-TCDF produced the best coefficients of determination (r2). This contrasts to
data for other avian species and locations, where non-ortho and mono-ortho PCBs or TEQs,
commonly using Safe’s (1990) TEFs, provided the best statistical fit to CYP1A parameters
(Van den Berg et al. 1994b; Sanderson et al. 1994a; Rattner et al. 1993). However, exposure
to PCDDs and PCDFs relative to PCBs was low in all of those studies, whereas the reverse
was true for Bald Eagles. Data on fish-eating birds in the Great Lakes region (Kubiak et al.
1989; Yamashita et al. 1993) and in the Rhine estuary of The Netherlands (Van de Berg et al.
1994b, Bosveld and Van den Berg 1994) indicated that PCB congeners, in particular PCB 126
(34-345) and PCB 118 (245-34), were the major contributors to TCDD-like toxicity. The
relative contribution of the major Ah-receptor active congeners in yolk sacs of Bald Eagles is
compared among sites and to common terms from the Netherlands in Figure 2.5. Total
TEQ5WHO and the pattern of contributors was similar between Bald Eagles from west Vancouver
Island and the common terms; however, PCDDs and PCDFs made a much greater contribution
in the Bald Eagles from the Strait of Georgia and the Fraser Delta.
Further comparison of avian laboratory data on relative toxicity of PCBs to PCDDs and
PCDFs suggests that TEFs derived from mammals such as Safe’s (1990) TEF’s tend to
overestimate the toxicity of the both the mono-ortho and non-ortho PCBs, in all avian species
with the possible exception of the chicken ((Brunstrom 1990; Brunstrom and Anderson 1988;
65
14000
xCl)Cl)(U
0
Cl)CwI-
Figure 2.5 - The contribution of various chlorinated hydrocarbon groups to the sumof TCDD toxic equivalents (TEQ) in Bald Eagle yolksacs from coastal British
Columbia, 1992 (N values and variances are in the tables), compared to values forcommon terns from the Netherlands. Toxic equivalents factors for PCDDs/PCDFs
from Safe (1990) and for PCBs from Ahlborg et al. [1994].
Kennedy et al. 1994; Bosveld et al. 1992). In Table 2.8, three sets of TEFs were compared;
biochemical parameters in Bald Eagle livers were regressed against yolk sac concentrations of
either TCDD or TCDF and TEQs using the different TEFs. The WHO-TEFs, which give
lower weighting to the mono-ortho PCBs, produced r2 values which were closest to those
determined using the individual contaminants. These results suggest that in Bald Eagle chicks,
PCBs are relatively less toxic than TCDD for the endpoints measured.
A number of fish-eating bird studies concluded that embryonic CYP1A induction is a
sensitive biomarker for other deleterious Ah-receptor mediated responses (Hoffman et al. 1987;
12
10
8
6
4
2
0
Li mono-o-PCBs1
-
-. LJ non-o-PCBs
PCDFs
El other-PCDDs1
•TCDD
.
Bald Eagle
\ \
I— Common Tern
66
Bosveld and Van den Berg 1994; Bosveld et al. 1994; Sanderson et al. 1994a; Rattner et a!.
1994). In Bald Eagle chicks from west Vancouver Island, low EROD activity and low levels
of the CYP1A cross-reactive protein indicate background exposure to TCDD-like compounds.
On a lipid weight basis, TEQ5WHO in yolk sacs were about 6,000 ng/kg. Converting this result
to a whole egg, wet weight basis, (dividing by a mean factor of 60, based on comparison for
Bald Eagles of a yolk sac and whole egg analyzed from the same nest), mean TEQ5PCDD,PCDFS
were about 15 ng/kg in west Vancouver Island eggs. If we include the PCB contribution,
TEQsWHO in the west Vancouver Island reference area were about 100 ng/kg. This is a
suggested no-observed-effect-level (NOEL) in Bald Eagle eggs, using CYP1A as a marker.
Likewise, levels of the CYP1A cross-reactive protein were significantly higher at Powell
River, where mean TEQ5WHO in yolk sacs, on a lipid weight basis, were about 12,600 ng/kg, or
about 210 ng/kg, on a wet weight basis in the whole egg. This is suggested as a lowest
observed-effect-level (LOEL).
In conclusion, Bald Eagle chicks collected near pulp mills were exposed to elevated
concentrations of PCDDs and PCDFs which correlated with induction of a hepatic CYP1A
cross-reactive protein. Levels of PCBs and other organochiorines did not vary among sites and
were less important in the CYP1A induction.
67
Acknowledgements
Many people contributed their time to the success of this project. I would especially
thank I. Moul and G. Compton for assistance in the field. C. Kuehier suggested the incubation
conditions. M.S. Bhatti and A. Roble assisted with dissecting and initial processing of embryos.
Dr. H. Philibert undertook the histology at the University of Saskatchewan, Western College of
Veterinary Medicine. M. Simon, M. Mulvihill and A. Idrissi performed the chemical analysis.
W. Ko prepared the microsomes. F. Maisonneurve, G. Sans-Cartier and K. Williams are
thanked for their technical assistance with the biochemistry, which was performed in the
laboratory of S. Trudeau (NWRC). A. Lorenzen performed the CYP1A assay. B. Woodin
performed the CYP2B assay. J. Smith provided advice on the statistics. S. Bucknell typed the
tables and P. Whitehead assisted with drafting figures. The research was supported by the
Canadian Wildlife Service and th Wildlife Toxicology Fund of Environment Canada and by the
National Science and Engineering Research Council of Canada.
68
Appendix 2.1 Selected morphological measurements in Bald Eagle chicks collected 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 + 10.6
Relative 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 + 2.26
Tibia length (mm) 26.5 + 0.53 27.5 + 1.16 26.5 + 0.69 26.8 + 1.09
NOTE: No significant differences were detected among locations for body or yolk-free body,liver, kidney, intestine, heart, adrenal, yolk, tibia and thyroid weights; tibia, tarsus, culman orwing lengths.
69
CHAPTER 3
BIOACCUMULATION OF CHLORINATED HYDROCARBONS ANDMERCURY IN EGGS AND PREY OF BALD EAGLES
The purpose of the bioaccumulation study was to measure chlorinated hydrocarbon
levels, particularly for PCDDs and PCDFs, in eagle eggs in order to determine spatial and
temporal patterns and trends, and to relate the levels to critical concentrations in their food
using a simple model. At issue was the determination of site specific concentrations of PCDDs
and PCDFs in representative sentinel food items, such as forage fish andfish-eating birds, that
would not adversely affect Bald Eagles. The development of guidelines for chlorinated
hydrocarbon levels in dietary items of eagles should have broader applicability in other North
American jurisdictions.
Materials and Methods
Sample collection
From 1990 to 1992, a total of 32 Bald Eagle eggs were collected at six sites on the
south coast of British Columbia (Figure 3.1). Four treatment areas were selected based on
proximity of eagle breeding sites to industrial pollutant sources. The lower Fraser valley near
Vancouver is a heavily urbanized and industrialized area that receives wastes from numerous
local and upstream pulp, paper and lumber mills and wood treatment operations. The Crofton
and Powell River areas each receive effluent inputs from local kraft pulp mills. Nanaimo is an
urbanized area with a large kraft mill and other wood milling and yarding operations. The
main reference site was an area of northern Johnstone Strait, with little industrial activity other
than lumber yarding. Three single eggs were also obtained from 1) Clayoquot Sound on the
west coast of Vancouver Island in an area where lumber cutting is the only industrial activity 2)
lower Alberni Inlet, a bleached-kraft pulp mill is at the head of the inlet 3) Langara Island in
the Queen Charlotte’s archipelago, remote from any industrial activity.
70
•1 C C.) I I I
Suitable nests were located by ground, boat and aerial surveys, during which nests were
scored numerically to estimate access, suitability for climbing and land tenure. In 1990 and
1991, in an effort to obtain fresh eggs, collections were made during the first two weeks of
April in the lower Fraser valley and the Strait of Georgia, during the first week of May on the
west coast of Vancouver Island and during the third week in May in Johnstone Strait. Normally
a single fresh egg was collected and only from nests with at least two eggs, except at Stillwater
Bay in 1992, when both eggs were taken. The two eggs collected in 1994 were addled, they
were retrieved from nests in June or July during blood sampling of nestlings. To encourage
continued incubation of the remaining egg and thus to minimize the impact of collection, time
near the nest and in the nest tree was minimized. Eggs collected in 1990 and 1991 were
refrigerated until the contents were removed and placed into chemically-cleaned
(acetone/hexane) glass jars with aluminum foil lid-liners and then frozen. The eggs collected in
1992 were initially incubated as part of another study (Chapter 2); the failed eggs from this
study were removed from the incubator and then treated the same as eggs from other years.
Frozen eggs were shipped to the CWS National Wildlife Research Centre (NWRC) in Ottawa.
Chemical analyses
Whole eggs were homogenized and prepared for analysis at NWRC. Aliquots for
organochlorine pesticides and PCBs were analyzed according to methods described in Norstrom
et al. (1988) and outlined in Chapter 1, except that total PCB levels are reported as the sum of
28 congener peaks (24 listed in Figure 3.2, plus trace amounts of PCBs 137, 195, 200 and
206). Eggs collected in 1990 and 1991 were analyzed for PCDDs/PCDFs by low resolution
GC/MS using a Hewlett-Packard 5987B machine with a 30 m DB-5 capillary GC column
(Norstrom and Simon 1991); the method is described in Chapter 1. PCDD/PCDF and non
ortho PCB analyses of eggs collected in 1992 were carried out on a VG Autospec high
resolution mass spectrometer linked to a HP 5890 Series II data system according to methods
described by Letcher et al. (in press), also as outlined in Chapter 1. Mercury was analyzed at
the NWRC by cold vapour atomic absorption according to methods described by Scheuhammer
72
& Bond (1991), and methyl mercury was extracted as described in Callum and Ferguson
(1981).
Eggshell thickness measurement
Eggshells were air-dried in the laboratory for two weeks or more. Using a ball
micrometer, shell thickness was measured at the equator of the shell, including the membrane;
five readings were made and averaged.
Statistical treatment
For each location, data were combined for all years in order to give a larger sample
size. Chemical residue data were transformed to common logarithms and geometric means and
95 % confidence intervals determined. Data were also converted to common logarithms and
SAS routines used to perform a one-way analysis of variance followed by Tukey’s multiple
comparison procedure (MCP) to determine significant differences in mean residue levels among
sites. For determination of statistical differences among sites for percent PCB congeners, an
arcsine transformation was used, followed by ANOVA and Tukey’s MCP. Unless otherwise
indicated, a significance level of p <0.05 was applied to all statistical tests. Patterns of all
chlorinated hydrocarbons and other the major PCB congeners as percent total PCBs were
analyzed using principle components analysis (PCA) in SAS. As for the other statistical
analyses, residue concentrations were transformed to common logarithms, while for the percent
PCB congener contributions, an arcsine transformation was used.
Toxic equivalents (TEQs) were estimated using standard toxic equivalent factors for
PCDDs and PCDFs as suggested in Safe (1990), except that for the mono-ortho and non-ortho
PCBs, the World Health Organization toxic equivalents (WHO-TEFs, Ahlborg et al. 1994).
Bioaccumulation model
In order to relate PCDD and PCDF levels in Bald Eagle eggs to their diet, a simple
bioaccumulation model was used (modified after U.S. Environmental Protection Agency 1993
and US Fish and Wildlife Service, 1994). The model assumes: 1) breeding Bald Eagles are
year-round residents and, therefore, acquire most of their contaminant burden from local
73
sources 2) levels in eagle eggs are in equilibrium with those in the female’s diet. The model
has the form:
BEE = BMF [F1(X1) + F2(X2) ... + FN(XN)]
BEE = Contaminant concentration in Bald Eagle egg
BMF = Biomagnification factor for a given contaminant
F1 = Fraction of item one in diet
X1 = Contaminant concentration in item one
FN = Fraction of the Nth item
XN = Contaminant concentration in the Nth item
As input, we used data on PCDD and PCDF levels in avian and fish prey from near
pulp mills and at reference sites on the British Columbia coast, summarized in Tables 3.1 and
3.2. Estimates of Bald Eagle diet composition were taken from Knight et al. (1990), Vermeer
et al. (1989) and Watson et al. (1991). The eagle diet was divided into components, which
varied among sites based on availability of contaminants data: 1) fish-eating waterfowl (grebes,
cormorants, herons and mergansers) 2) non-fish-eating waterfowl (invertebrate and plant-eating
waterfowl) 3) omnivorous gulls 4) non-salmonid fish 5) salmonid fish. Biomagnification factors
determined in Lake Ontario Herring gulls relative to forage fish (Braune and Norstrom, 1989)
were used: 2,3,7,8-TCDD (21), 1,2,3,7,8-PnCDD (10), 1,2,3,6,7,8-HxCDD (16), 2,3,7,8-
TCDF (1.4, estimated), 2,3,4,7 ,8-PnCDF (4.5). The biomagnification factors for herring gulls
were similar to those estimated for great blue herons to forage fish at Crofton, 25 and 10
respectively for 2,3,7,8-TCDD and 1 ,2,3,6,7,8-HxCDD (Elliott et al. 1989a). Where only egg
or liver data was available for a species, the inter-tissue ratios in Braune and Norstrom (1989)
were used to convert to whole body concentrations.
74
Tab
le3.
1M
ean
PC
DD
/PC
DF
leve
ls(n
g/kg
,w
etw
eigh
t)in
fish
colle
cted
near
thre
epu
lpm
ills
onth
eSt
rait
ofG
eorg
ia,
Bri
tish
Col
umbi
a.
Are
aS
peci
esN
*T
issu
eC
olle
ctio
n%
2,3,
7,8
1237
8T
otal
Tot
al2,
3,7,
823
467
Dat
a
peri
odli
pid
TC
DD
PnC
DD
HxC
DD
HpC
DD
TC
DF
PnC
DF
Sou
rce
Nan
aiin
oE
ngli
shS
ole
4/4
Fil
let
Jan-
Feb
.19
902.
91.
01.
54.
5<
1.0
13<
0.0
51
Nan
aim
oE
ngli
shS
ole
4/4
Liv
erJa
n-F
eb.
1990
8.5
2.6
6.6
441.
658
1.7
1
Nan
aim
oC
hino
okS
alm
onF
ille
tJa
n-F
eb.
1990
6.5
<1.
0<
1.0
4.2
<1.
047
<1.
02
Cro
fton
Eng
lish
Sol
e2/
2F
ille
tJa
n-F
eb.
1990
2.1
1.0
2.0
4.0
1.0
9<
0.0
51
Cro
fton
Eng
lish
Sol
e2/
2L
iver
Jan-
Feb
.19
9010
5.0
8.0
64.
<3.
089
2.0
1
Cro
fton
Arr
owto
oth
Flo
unde
r1/
1F
ille
tJa
n-F
eb.
1990
3.4
<0.
5<
1.5
<1.0
<1.5
7<
0.5
1
Cro
fton
Arr
owto
oth
Flo
unde
rL
iver
Jan-
Feb
.19
903.
63.
5<
2.5
9.0
<1.
563
<1
.01
Cro
fton
Roc
kF
ish
Fil
let
Jan-
Feb
.19
901.
81.
7<
1.0
9.1
<1.
017
<1
.02
Pow
ell
Riv
erE
ngli
shS
ole
3/3
Fil
let
Jan-
Feb
.19
901.
4<
0.5
<1.0
1.5
<1
.015
<0.5
1
Pow
ell
Riv
erE
ngli
shS
ole
3/3
Liv
erJa
n-F
eb.
1990
5.7
3.0
8.0
987.
413
79.
01
Pow
ell
Riv
erC
hino
okS
alm
onF
ille
tJa
n-F
eb.
1990
4.3
2.2
2.4
2.5
9.4
621.
32
Dat
aS
ourc
e:
1-
Har
ding
&P
omer
oy,
1990
2-
Env
iron
men
tC
anad
a,un
publ
.da
ta
*-
Num
ber
anal
yzed
IN
umbe
rco
llec
ted
Lii
Table 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 PnCDF
Fish eating birds
Western Grebe Port Alberni 1989 liver
1992 bm*
Nanaimo 1992 bm
Powell River 1992 bm
Alert Bay 1992 bm
Double-crested Fraser Delta 1990 eggsConuorant
Fraser Delta 1992 eggs
Crofton 1990 eggs
Crofton 1992 eggs
Nanaimo 1992 eggs
Great Blue Fraser Delta 1990 eggsHeron
Fraser Delta 1992 eggs
Crofton 1990 eggs
Crofton 1992 eggs
Rhinoceros Johnstone 1990 eggsAuldet Strait
10 4.7 8
7 4.2 26
8 4.8 8
10 5.6 10
8 6.4 42
10 5.6 10
5 6.3 102
5 6.1 19
12 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 2
ND ND 11 2
2
3
3
2
Invertebrate feeders
Bufflehead Fraser Delta 1990 bm
Crofton 1990 bm
Alert Bay 1992 bm
Surf Scoter Fraser Delta 1989 bm
Port Alberni 1989 liver
Crofton 1990 bm
Nanaimo 1992 bm
Powell River 1992 bm
Alert Bay 1992 bm
Glaucous- Nanaimo 1992 eggswinged Gull
1 3.4 18
2 2.1 9.4
11 3.7 <1
10 4.1 <1.1
5 3.1 24
10 2.8 5.1
6 2.5 0.5
8 3.1 <1.6
7 1.6 <2
10 8.8 <1.0
<1 17
<2 29
<3 <3
<1.9 <3.1
22 30
<0.5 3.3
0.7 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 1
11 123 13 1
3 40 1 1
7.3 7.4 1.3 2
<8.2 12 <1.8 2
<22 3.2 <2.8 2
<1.0 <1.0 1.0 2
5 7.3 117 385 249
8 3.8 25 66 64
3 4.6 2.9 <2.9 8.6
5 7.1 4.4 <2.6 21
2 3.1 <1.6 <3.2 <3.5
6 4.5 25 42 64
14 11 ND ND ND 2
42 65 ND 2 11 2
25 47 ND 1 7 2
22 38 ND ND ND 2
45 57 5 15 16 2
10 11 ND 4 1
223 229 3 10 32
40 47 1 5 7
3 3 1 27 2
*bm- breast muscle
ND - not detected (detection limit = 0.5 ng/kg)References: 1 - Whiteheact et al. 1990; 2 - Elliott et al., 1995b; 3 - Whitehead et al. 1992.
76
Results
PCDDs and PCDFs
Major PCDD contaminants were 1,2,3,6,7,8-HxCDD > 1,2,7,8-PnCDD > 2,3,7,8-
TCDD, except in the lower Fraser valley, where 2,3,7,8-TCDD was the greater than the other
two compounds (Table 3.3). All eggs contained detectable levels of the three major PCDD
congeners. Lesser concentrations of 1,2,3,4,6,7,8-HpCDD and OCDD were found in most
eggs. The only PCDFs consistently detected in Bald Eagle eggs were 2,3,7,8-TCDF and
2,3,4,7,8-PnCDF. Eggs from Johnstone Strait contained significantly less 2,3,7,8-TCDD than
did eggs from other sites. Concentrations of 1,2,3,7,8-PnCDD were significantly higher in
eggs from Crofton than either the lower Fraser valley or Johnstone Strait. Concentrations of
1,2,3,6,7,8-HxCDD and 2,3,4,7,8-PnCDF were significantly lower in Fraser valley eggs than
from the pulp mill sites, but did not differ significantly from Johnstone Strait.
Organochiorines
Quantifiable residues of DDE, DDD, trans-nonachlor, cis-nonachlor oxychlordane, cis
chiordane, heptachlor epoxide, dieldrin, mirex, fi-HCH and HCB were found in all eggs
analyzed (Table 3.4). DDT was found in the majority of the eggs at low levels, generally <
0.01 mg/kg. Photomirex was detected in 65 % of the eggs; where present concentrations were
about 50 % of mirex concentrations. Organochiorine levels were generally highest in eggs from
Powell River, although differences were significant in only one case: trans-nonachlor was
significantly higher at Powell River and Nanaimo than the Fraser valley. Mean concentrations
of cis-chlordane were significantly higher in eggs from Johnstone Strait than other sites except
Powell River and were also significantly lower at Crofton than all other sites.
Mercury
Highest mean concentrations of total mercury were in eggs from Johnston Strait and the
Fraser valley and were significantly higher than those from Nanaimo and Crofton, but not
Powell River (Table 3.5). Methyl-mercury was also determined in the eleven eggs from 1990
and constituted an average of 88 % (SD = 11, range 73 - 100%) of the total mercury present in
those Bald Eagle eggs.
77
Tab
le3.
3P
olyc
lilo
iina
ted
dibe
nzod
ioxi
n(P
CD
D)
and
poly
chio
rina
ted
dibe
nzof
uran
(PC
DF
)re
sidu
ele
vels
(wet
wei
ght
basi
s)in
Bal
dE
agle
eggs
from
Bri
tish
Col
umbi
a,19
90-
1992
.
PC
DD
and
PC
DF
Lev
els
(ng/
kg)
(geo
met
ric
mea
nan
d95
%co
nfid
ence
inte
rval
)
Nes
t(M
apN
o.*)
Yea
r%
liyid
%m
oist
ure
2378
1237
812
3678
1237
8923
7823
478
(mea
n±
SD)
TC
DD
PnC
DD
HxC
DD
HxC
DD
TC
DF
PnC
DF
Low
erF
rase
rV
alle
yB
runs
wic
kPt
.(1
)19
905.
182
.542
3742
423
13A
nnac
isIs
.(2
)19
905.
982
.758
5511
25
112
12C
haha
lis
Flat
s(3
)19
906.
183
.158
5255
289
14Is
land
20(4
)19
915.
382
.351
717
ND
16N
DC
heam
Isla
nd(5
)19
914.
982
.023
615
ND
732
Aga
ssiz
Bri
dge
(6)
1991
4115
181
1310
Mea
n5.
682
.444
220
233
21.
339
25a
±0.6
±0.5
(30-
63)
(7-5
7)(1
4-76
)(0
.5-3
.8)
(14-
105)
(1.3
-22)
Nan
aim
oC
anox
y(7
)19
903.
085
.759
109
198
529
16L
eask
(8)
1990
4.7
83.0
6399
250
936
18C
anso
(9)
1990
4.2
84.1
8211
634
612
2931
Jack
Pt.
(10)
1990
4.5
84.2
7913
322
76
4920
Nor
thw
est
Bay
(11)
1992
4.9
81.2
1422
371
185
Mau
deIs
land
(12)
1991
4.5
83.2
7010
417
32
119
35So
uthe
yIs
land
(13)
1991
4.2
83.3
2829
795
6512
Jack
Poi
nt(1
0)19
924.
72
2M
ean
4.4
83.4
452
66
134”
34
143
2
±0.6
±1.
4(2
6-78
)(3
5-12
2)(6
8-26
4)(1
.5-7
.6)
(26-
70)
(8-2
7)C
roft
onR
.Pr
ingl
e(1
4)19
904.
482
.710
421
137
410
1627
Sout
hey
(15)
1990
5.9
80.0
110
149
310
726
34C
roft
on(1
6)19
916.
082
.851
108
173
160
22M
ean
5.4
81.8
842
150”
270L
3229
22
7”
±0.
9±
1.6
(29-
243)
(65-
346)
(99-
742)
(0.1
-6)
(6-1
54)
(16-
47)
Tab
le3.
3co
nt...
PCD
Dan
dPC
DF
Lev
els
(ng/
kg)
(geo
met
ric
mea
nan
d95
%co
nfid
ence
inte
rval
)
Nes
t(M
apN
o.*)
Yea
r%
liyid
%m
oist
ure
2378
1237
812
3678
1237
8923
7823
478
(mea
n±
SD)
TC
DD
PnC
DD
HxC
DD
HxC
DD
TC
DF
PnC
DF
Pow
ell
Riv
erK
elly
Pt.
(17)
1990
6.1
80.0
9812
924
47
5927
Con
vent
(18)
1990
5.7
82.4
8812
837
215
9737
Lun
d(1
9)19
915.
082
.141
5918
67
110
24P
owel
lR
iver
(20)
1992
3.7
81.9
3247
116
218
5S
tili
wat
er(A
)(2
1)19
925.
879
.678
104
143
316
648
Sti
liw
ater
(B)
(21)
1992
6.7
79.3
8110
614
63
168
50G
rise
Poi
nt(2
2)19
926.
179
.610
1880
1M
ean
5.6
80.7
49ä
71b,
c17
0b4
85
27b
±1.0
±1.4
(23-
105)
(37-
138)
(103
-261
)(2
-9)
(49-
147)
(15-
49)
John
ston
eS
trai
tP
lum
per
5(2
3)19
914.
882
.822
3972
558
10P
lum
per
8(2
4)19
914.
383
.312
2816
73
397
Pear
ce3
(23)
1991
6.0
80.6
3264
111
580
3Pe
arce
5(2
6)19
914.
483
.311
4373
237
5H
arbi
edon
Isla
nd(2
7)19
912.
885
.415
3379
ND
6810
Swan
son
Isla
nd(2
8)19
914.
185
.410
2652
129
5O
wl
Isla
nd(2
9)19
913.
983
.311
2543
1M
ean
4.3
84.1
iSh
35
78
’1.
3947
±1.
0±
4.1
(10-
22)
(26-
48)
(51-
118)
(0.5
-3.4
)(3
3-66
)(6
-11)
Poc
ahon
tas
Pt(3
0)19
922.
488
.117
5438
17
4B
erry
man
Pt(3
1)19
925.
981
.410
167
ND
74
Lan
gara
Is.*
*19
946.
580
.12
53
ND
52
Map
Nos
.30
-A
lbem
iIn
let,
31-
Cla
yoqu
otSo
und.
**
-Q
ueen
Cha
rlot
teIs
land
sab
,c-
mea
nsth
atdo
not
shar
eth
esa
me
lett
erar
esi
gnif
icia
ntly
diff
eren
t(p
0.05
)
Tab
le3.
4O
rgan
ochi
orin
ean
dPC
Bre
sidu
ele
vels
(mg/
kg,
wet
wei
ght)
inB
ald
Eag
leeg
gsfr
omth
eB
ritis
hC
olum
bia
coas
t,19
90-1
992,
expr
esse
das
geom
etri
cm
eans
and
95%
conf
iden
cein
terv
als
(ran
gein
brac
kets
).
Loc
atio
nN
Tot
alPC
Bs
DD
ED
DD
tran
s-ci
s-ox
y-di
eldr
inm
irex
beta
-H
CB
nona
chio
rno
nach
ior
chio
rdan
e11
CR
Low
er6
26
8ab
2j7
a0
.05
80.
142k
0.0
30.
037a
0.03
7a0
.00
9o.o
osa
0.02
5aF
rase
r1.
49-4
.81.
07-4
.41
0.04
1-0.
083
0.09
8-.2
040.
020-
0.04
40.
023-
0.05
90.
019-
0.07
30.
002-
0.04
50.
001-
0.02
40.
017-
0.03
9V
alle
y(1
.08-
6.21
)(.
90-4
.14)
(0.0
30-0
.075
)(0
.082
-0.2
34)
(0.0
18-0
.049
)(0
.022
-0.0
82)
(0.0
20-0
.091
)(0
.001
-0.0
38)
(0.0
01-0
.032
)(0
.016
-0.0
32)
Nan
aim
o8
4ab
3.13
a0.o
52
0245
ab
00
4.7ab
0.04
2a0.
042a
o.o
lsa
0.02
2a0.0
12
3.28
-6.7
91.
65-5
.92
0.03
5-0.
079
0.18
6-0.
323
0.03
6-0.
060
(0.0
29-0
.062
)0.
029-
0.06
10.
005-
0.04
70.
011-
0.04
20.
004-
0.03
8(1
.80-
7.14
)(.
672-
8.52
)(0
.023
-0.1
01)
(0.1
48-0
.432
)(0
.027
-0.0
76)
(0.0
14-0
.062
)(0
.018
-0.0
64)
(0.0
01-0
.030
)(0
.004
-0.0
53)
(0.0
01-0
.033
)
Cro
fton
2.77
a0.
039a
•162
b0
03
6ab
0.03
3a0.
03k
0.01
6k0.
018k
o.o
iz2.
23-1
0.2
1.48
-5.1
70.
014-
0.11
00.
121-
0.21
70.
020-
0.06
20.
020-
0.55
00.
014-
0.06
50.
006-
0.04
20.
008-
0.04
00.
008-
0.02
0(3
.47-
6.38
)(2
.07-
3.26
)(0
.024
-0.5
0)(0
.143
-0.1
79)
(0.0
28-0
.044
)(0
.027
-0.0
41)
(0.0
21-0
.038
)(0
.010
-0.0
23)
(0.0
13-0
.024
)(0
.010
-0.0
15)
Pow
ell
75
.08
33a
0.0
71
0.32
k’0066
b0046
a0.
044a
0028
a0025
ao
.olz
aR
iver
3.88
-6.6
51.
64-
6.66
0.03
5-0.
143
0.23
4-0.
438
0.04
6-0.
094
0.03
4-0.
062
0.03
1-0.
064
0.01
6-0.
048
0.01
3-0.
047
0.00
3-0.
046
(3.3
2-6.
96)
(1.3
1-8.
70)
(0.0
29-0
.215
)(0
.192
-0.4
78)
(0.0
41-0
.109
)(0
.030
-0.0
75)
(0.0
21-0
.065
)(0
.015
-0.0
63)
(0.0
07-0
.052
)(0
.001
-0.0
31)
John
ston
e7
2.5&
’229
a0.0
47”
o.0
z40031
a0.o
20
.01
30.
024a
Str
ait
1.78
-3.6
41.
46-3
.58
0.02
4-0.
058
0.14
4-0.
301
0.03
1-0.
079
0.00
5-0.
117
0.02
0-0.
048
0.01
4-0.
028
0.00
3-0.
052
0.01
6-0.
036
(1.7
0-5.
34)
(1.2
2-5.
95)
(0.0
26-0
.112
)(0
.142
-0.4
53)
(0.0
33-0
.119
)(0
.001
-0.0
81)
(0.0
17-0
.071
)(0
.014
-0.0
44)
(0.0
01-0
.046
)(0
.015
-0.0
55)
Alb
erni
14.
475.
140.
028
0.15
10.
028
0.02
40.
038
0.01
90.
024
0.00
1In
let
Cla
yoqu
ot1
3.86
5.12
0.04
40.
293
0.03
70.
047
0.06
0.03
80.
026
0.01
9S
ound
Lan
gara
11.
912.
970.
034
0.21
50.
039
0.09
50.
024
0.06
20.
061
0.05
1Is
land
a,b
-m
eans
that
dono
tsh
are
the
sam
ele
tter
are
sign
ific
iant
lydi
ffer
ent
(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 0294b
0.186 - 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 analyzed
Polychiorinated biphenyls
Mean sum-PCB concentrations were significantly different only between Powell River,
where they were higher, and Johnstone Strait (Table 3.2). There were a number of significant
differences in mean concentrations of individual PCB congeners: PCBs 170 (2345-234), 171
(2346-234), 182 (2345-246), 201 (2356-2345) and 203 (23456-245) were significantly higher in
eggs from Crofton, Nanaimo and Powell River than Johnstone Strait; PCBs 180 (2345-245),
183 (2346-245) and 194 (2345-2345) were significantly higher at Crofton, Nanaimo and Powell
River than both Johnstone Strait and the Fraser valley; PCBs 153 (245-245) and 128 (234-234)
were significantly higher only at Powell River compared to Johnstone Strait and the Fraser
valley and PCB 138 (234-245) was significantly higher at Powell River than Johnstone Strait.
The percent contribution of individual congeners was determined and compared among
sites (Figure 3.2). The major peaks were 153, 138, 180, 182, 118 (245-34) and 99 (245-24),
which together contributed 64 % of the total PCBs present in all eggs. There were a number of
statistically significant differences among sites in percent contribution of individuals PCBs.
Percent contribution of a number of the lower chlorinated congeners, including PCBs 66 (24-
34), 101 (245-25), 99, 87 (234-25), 118 and 105 (234-34), was significantly higher at both
Johnstone Strait and the lower Fraser Valley than the other three sites. In addition, among
those compounds, percent contribution of PCBs 99 and 118 were significantly higher at Powell
River than at Crofton. In contrast, the percent contribution of a number of the higher
81
chlorinated congeners, PCBs 183, 180, 170, 203 and 194, was significantly lower in eggs from
the lower Fraser Valley and Johnstone Strait than Nanaimo, Crofton and Powell River. This
geographic trend of differences in the PCB pattern was supported by results of principle
components analysis. Principle components analysis of the PCB pattern was carried out using
only congeners, 170, 180 and 194, which are indicative of Aroclor 1260 (Mullin et al. 1984),
PCBs 99 and 118, indicative of Aroclor 1254, and PCB 66, considered indicative of Aroclor
1242. Two significant principle components were apparent which explained 90 % of the total
variance among individual egg analyses. The first component (PC 1) explained 75 % and the
second component (PC2) explained 15 %. As shown in Figure 3.3, the Johnstone Strait and
Fraser Valley eggs clump separately from the other locations, although there is some overlap,
particularly of some samples from Powell River.
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 identified
according to their IUPAC number.
25
20
5
0p.. p. .. p..
PCB Congener Number
82
Concentrations of six non-ortho PCB congeners were determined in eight eggs collected
in 1992 (Table 3.6) and in two eggs collected in 1994. The pattern was consistent in the 1992
samples and the 1994 sample from Langara Island, with 126 (345-34)> 77 (34-34) > 169
(345-345) > 81 (345-4) > 37 (34-4). However, in the 1994 sample from Herrling Island, 77
> 126 > 81 > 169 > 37. Linear regressions were determined between concentration of
PCBs 126 and 77 and sum-PCBs for the ten eggs in Table 3.3, in order to estimate values in
the whole data set for estimation of TCDD toxic equivalents (TEQs):
PCB 126 (ng/kg) = 156 [sum-PCBs (mg/kg)] + 78, r2=0.634, p<O.Ol
PCB 77 (ng/kg) 69 [sum-PCBs (mg/kg)] + 85, r2 = 0.505, p <0.04
PCBs
66
j99V 118
L Lower Fraser Valley C Crofton J Johnstone Strait
N Nanaimo P Powell River
Figure 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.
TPCBs
170180
194
z
-3 -2.5 -2 -1.5 -1 -0.5 0
PRIN2PCB 99
0.5 1 1.5
PCB 66
83
Table 3.6Non-ortho PCBs in Bald Eagle eggs (ng/kg, wet weight) collected from British Columbia, 1992.
Nest (Map No.*) PCB-37 PCB-81 PCB-77 PCB-126 PCB-169 PCB-189
Northwest Bay (13) 6.2 24 146 323 65 22
Jack Point (10) 31 52 349 709 131 41
Powell River (20) 13 42 207 544 131 35
Stiliwater A (21) 26 105 720 1354 285 84
Stillwater B (21) 24 107 684 1326 277 97
Grise Point (22) 52 46 387 547 121 41
Pocohantas Point (30) 15 51 459 685 114 35
Berrryman Point (31) 23 74 691 754 135 39
Herrling Is. 27 96 576 314 47 5
Langara Is. 5 32 310 585 203 1
* 10-13 - Nanaimo,
600
20-22 - Powell River, 30 - Albemi Inlet, 31 - Clayoquot Sound.
500
—400c,)
0)
U)CwI— 200
çcc
lIE non-O-PCBs
mono-O-PCBs
LI PCDFs
other-PCDDs
mTCDD
ii100
0
0’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 et
al. 1994.
84
Table 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 ± 0.024
Nanaimo 1990-92 5 0.587 ± 0.035
Crofton 1990-91 3 0.583 ± 0.024
Powell River 1990-92 5 0.590 ± 0.038
Johnstone Strait 1991 6 0.569 ± 0.036
Percent 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 prey
An example output from the model is shown in Table 3.8, based on
Crofton, 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 from
In the absence of
The putative diet
TCDD toxic equivalents (TEQs)
Highest mean TEQs0were in eggs from Crofton, followed by Powell River, both of
which were significantly greater than Johnstone Strait. The relative contribution of PCDDs and
PCDFs to total TEQs0,64 %, was also highest at Crofton and was lowest, 47 %, in the
lower Fraser Valley eggs, as shown in Figure 3.4.
Eggshell thickness results
Neither mean eggshell thickness nor the percentage difference from the pre-1946
average for the Pacific North West of 0.6088 mm differed significantly among sites
(Table 3.7). There were no significant regressions between eggshell thickness and DDE or
other organochlorines.
85
Tab
le3.
8A
sim
ulat
ion
ofPC
DD
and
PC
DF
leve
lsin
Bal
dE
agle
eggs
atC
roft
on,
1990
,ba
sed
onco
ncen
trat
ions
inth
edi
et.
Con
tam
inan
tco
ncen
trat
ion
indi
etar
yite
ms
(ng!
kg,
wet
wei
ght)
(Fra
ctio
nof
that
item
insi
mul
ated
diet
)
Bir
ds(0
.475
)F
ish
(0.5
25)
Con
tam
inan
tco
ncen
trat
ion
in
bald
eagl
eeg
gs
Che
mic
alB
MF’
Non
-fis
hG
ulls
Her
ons
Cor
mor
ants
Non
-Sa
lmon
ids
Cal
cula
ted
Mea
sure
dM
ean
eatin
gbir
ds
2(0
.25)
(0.0
5)(0
.025
)sa
lmon
ids
3(0
.125
)V
alue
Val
ue
(0.1
5)(0
.4)
2378
-TC
DD
217
310
030
21
117
107
1237
8-Pn
CD
D10
15
220
462
0.5
8218
0
1236
78-H
xCD
D16
117
230
7017
2.5
284
342
2378
-TC
DF
1.4
350.
510
0.5
3715
3121
2347
8-P
nCD
F4.
52
135
120.
50.
58
31
1B
MF
-bi
omag
nifi
cati
onfa
ctor
2B
uffl
ehea
dan
dS
urf
Scot
er
Sole
,fl
ound
er,
rock
fish
consisted 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 predicted
2,3,7,8-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 and
PCDFs were near the detection limit; thus, a small difference in forage fish concentrations
would translate to a large error in the estimated BMF. The other major source of error is the
putative eagle diet, particularly the relative importance of fish and non-fish-eating birds.
The example in Table 3.8 approximates an average coastal Bald Eagle diet; however,
individual eagles or sub-populations can prey on greater amounts of fish-eating birds. For
example, Knight et at. (1990) reported that western grebes, which can accumulate extremely
high PCDD/PCDF levels (Table 3.2), were the main prey item of Bald Eagles in the Puget
Sound area. Eagles nesting near great blue heron colonies may also prey on chicks and adults
(Norman et al. 1989). Figure 3.5 shows how 2,3,7,8-TCDD concentrations would increase in
eagle eggs with an increasing fish-eating bird diet. Feeding on fish-eating birds may account
for extremely high liver levels of PCDDs, PCDFs and other chlorinated hydrocarbons of adult
eagles found dead or dying near Powell River and other areas of the Strait of Georgia
(Chapter 1).
87
5*Crofton, 1993
C) ±Crofton 1990
)K
0 10 20 30 40 50 60 70 80 90
Percent fish-eating birds in diet
Figure 3.5 Concentration of 2,3 ,7,8-TCDD predicted in Bald Eagle eggs based on the percent offish-eating birds in the diet. Prediction is based on a bioaccumulation model described in the text and
the simulation is based on data from Crofton, British Columbia.
88
Discussion
The data presented in this chapter show that Bald Eagle eggs collected near bleached
kraft pulp mills in the Strait of Georgia contained higher levels of 2,3,7,8-TCDD and -TCDF
when compared to other locations on the British Columbia coast. Total PCB levels were also
highest in eggs from the Strait of Georgia, reflecting greater industrialization. Concentrations
of organochiorine pesticides, including DDE, in eagle eggs were relatively consistent among
sites. Total-mercury levels were significantly higher in eggs from the Fraser Valley and
Johnstone Strait than the Strait of Georgia.
Patterns and sources of PCDDs/PCDFs
The formation of 2,3,7,8-TCDD and 2,3,7,8-TCDF during molecular chlorine bleaching
of wood pulp is a well known phenomenon (Kuehi et at. 1987; Luthe et al. 1990). By 1991,
all pulp mills studied here had implemented bleaching technology changes designed to minimize
TCDD/TCDF formation (Table 3.9), which has resulted in declining PCDD levels, particularly
of 2,3,7 ,8-TCDD, in sediments and biota near the mills (Whitehead et at. 1992; Elliott et a!.
1995). Concentrations of 2,3,7,8-TCDF in eagle eggs from Nanaimo and Powell River were
still elevated in 1992, suggesting that efforts to reduce TCDF contamination have been less
successful. In birds, 2,3,7,8-TCDF is quite quickly cleared from the body (Braune and
Norstrom 1989; Van den Berg et at. 1994). Other studies of wild birds have reported low
2,3,7 ,8-TCDF concentrations from the Great Lakes (Hebert et at. 1994; Ankley et at. 1993)
and Europe (Van Den Berg et at. 1987). However, elevated TCDF levels have been reported
in fish, invertebrates, and waterfowl near both riverine and marine pulp mills (Mah et a!.
1989; Harding and Pomeroy, 1990; Table 3.1; Champoux 1993). Osprey eggs collected from
nest locations downstream of pulp mills in the British Columbia interior contained 2,3,7,8-
TCDF levels up to 68 ng/kg (Whitehead et at. 1993). The high TCDF levels in eggs of eagles
and ospreys likely reflect a combination of recent exposure and direct yolk deposition of
contaminated dietary lipids, as suggested previously for great blue herons (Elliott et at. 1989a).
89
Until 1989, up to several million kg of chiorophenolic compounds were used annually
by the British Columbia forest industry, particularly on the coast, to prevent sap staining of
undried lumber. Although HxCDDs and HpCDDs predominate as dioxin contaminants in
chlorophenol mixtures, HxCDDs are further produced in large amounts when chlorophenol
contaminated woodchips are pulped (Luthe et al. 1990). Monitoring chip supplies for
chiorophenols, followed by a regulatory ban on their use as anti-sapstains, produced significant
HxCDD reductions in effluents and foodchains at the Crofton mill site (Whitehead et al. 1992).
A reduction in PCDD levels in eagle eggs is apparent, particularly between 1990 and 1992 at
Jack Point near Nanaimo.
In Fraser valley eagle eggs low HxCDD : TCDD ratios are consistent with lower
HxCDD concentrations in sediments and biota downstream of Fraser river pulp mills, the
putative sources of PCDDs and PCDFs at that site (Mah et al. 1989; Whitehead et al. 1993;
Harfenist et al. 1995). Due to the cooler, dryer climate of the British Columbia interior, lesser
amounts of chlorophenol antisapstain agents were use by lumber operations on the upper Fraser
and Thompson Rivers. Osprey eggs collected in 1991 from nests located downstream of the
pulp mill on the Thompson River at Kamloops had mean values of 47:3:22 ng/kg,
TCDD:PnCDD:HxCDD (Whitehead et al. 1993). In contrast, some osprey eggs contained
very high levels of 1 ,2,3,4,6,7,8-HpCDD and OCDD, indicative of direct chlorophenolic
inputs, rather than via pulp milling of contaminated wood chips.
Although there are no pulp or large saw mills on northern Johnstone Strait (only log
sorting facilities), PCDD/PCDF levels in eagles were relatively high. A non-kraft pulp mill
located to the west at Port Alice reported non-detectable PCDD/PCDF levels in effluents
(Anonymous 1994), and only trace amounts, 4 ng/kg of 2,3,7,8-TCDF, in crab hepatapancreas
from near the mill site (Harding and Pomeroy 1990). The PCCD/PCDF pattern in Johnstone
Strait eagle eggs is similar to the Strait of Georgia, which is the most likely source; however,
the exposure route is not clear. Acquisition of contaminants during seasonal southern
90
movements is unlikely as resident Bald Eagles on the Pacific coast remain on territory for most
of the year (Frenzel et al. 1989). Residents may leave breeding territories periodically during
the fall and winter to feed at salmon spawning sites; however, Pacific salmon, even from near
pulp mill sites, contained low PCDD/PCDF levels, with the exception of some 2,3,7,8-TCDF.
Eagle eggs from the west coast of Vancouver Island also had low PCDD/PCDF levels (Table
3.3, Chapter 2), probably indicating that they had not dispersed to more contaminated sites.
Long range transport is unlikely as a major vector, as pulp mill pollution is relatively localized
even within the Strait of Georgia (Elliott et al. 1989a; Harding and Pomeroy, 1990). There is,
however, an estuarine surface flow out of the Georgia Strait through Johnstone Strait (Thomson
1981), which may conceivably transport some suspended sediment-bound PCDDs and PCDFs.
A sediment sample from Louchborough Inlet, a fjord off of central Johnston Strait, was
reported to have levels of higher chlorinated PCDDs comparable to those near industrial sites in
the Fraser delta (Harding 1990). However, eagle prey species, such as western grebes and surf
scoters collected from Johnstone Strait in mid-March 1992, timed to obtain birds which had
wintered on site, had very low PCDD/PCDF levels, while samples of the same species
collected near pulp mills showed the typical pulp mill PCDD/PCDF signature. Johnstone Strait
Bald Eagles may still be exposed to contaminants from the Strait of Georgia by feeding on
waterfowl during spring migration along the coast towards their northern breeding grounds.
Rhinoceros auklets, large numbers of which breed in northern Johnston Strait, contained low
PCDD levels, although the mean 2,3,7,8-TCDF concentration was quite high and could
partially account for this compound in Johnston Strait eagle eggs.
The pattern of HxCDD > PnCDD > TCDD in Strait of Georgia wildlife differs from that
reported at other locations such as the Great Lakes (Ankley et al. 1993), interior rivers of
British Columbia (Whitehead et al. 1993) and elsewhere in North America (Elliott et al.
1995a). Hebert et al. (1994) used principal components analysis to show that Strait of Georgia
blue heron eggs clustered separately from Great Lakes herring gulls and other biota, based on
higher PnCDD and HxCDD concentrations, attributed to chlorophenol sources. However, a
91
sample of common merganser eggs from downstream of a pulp mill in Quebec had a pattern,
24:28:40 ng/kg TCDD:PnCDD:HxCDD, similar to that observed in British Columbia, perhaps
indicating a chlorophenol and a pulp mill source (Champoux 1993). Baltic Sea Common Murre
(Uria aalge) eggs contained 27:45:59 mg/kg TCDD: PnCDD: HxCDD (wet weight, re
calculated based on 17 % lipid in common murre eggs (Noble and Elliott 1986; Cederberg et
al. 1991), similar to the Strait of Georgia pattern. Grey Heron (Ardea cineria) livers from the
Netherlands also had a pattern somewhat similar to the Strait of Georgia, which was attributed
mainly to chlorophenols (Van den Berg et al. 1987).
European wildlife samples, at least from The Netherlands, appear to have higher
2,3,4,7,8-PnCDF concentrations (Bosveld et at. 1994; Van den Berg et a!. 1994b) compared to
those from North America (Elliott et at. 1989a; Hebert et at. 1994). This compound is a
known contaminant in PCB mixtures (Van den Berg et al. 1985), which would explain its
association with areas of PCB contamination (Hebert et a!. 1994) and its tendency to correlate
closely with PCB congeners in eggs (Elliott et at. 1989). Bosveld et at. (1994) suggested that
higher PCB levels in European wildlife samples explained the elevated 2,3,7,8-PnCDF levels;
they determined that lipid-normalized PCB concentrations in Common Tern (Sterna hirundo)
yolksacs from the Rhine-Meuse estuary were two to three-fold higher than in fish-eating bird
eggs from industrialized areas of the Great Lakes. However, direct comparison of lipid-
normalized whole egg to yolksac concentrations may overestimate concentrations in yollcsacs.
For example, in Bald Eagles, concentrations of chlorinated hydrocarbons were three-fold higher
on a lipid weight basis in a single yolksac compared to the sibling whole egg. On a wet weight
basis, total PCB levels in Great Cormorant (Phatacrocorax carbo) eggs from the contaminated
Biesbosch colony in the Netherlands (Van Hattum et at. 1993 cited in Bosveld and Van den
Berg, 1994) were similar, about 23 mg/kg, to those in double-crested cormorants from highly
contaminated Hamilton Harbour in the Great Lakes (Bishop et al. 1992). Therefore,
differences in PCB formulations or other sources may account for higher PnCDF levels in
European wildlife samples, rather than higher PCB levels.
92
Patterns and sources of organochiorines and mercury
The uniformity in OC residues indicates similar dietary exposure among most
individuals. The few eggs with distinctly lower organochiorines are probably individual eagles
feeding on larger amounts of fish, non-fish-eating birds or mammals. Based on OC patterns in
seabird eggs, Elliott et al. (1989) concluded that atmospheric sources were dominant over a
wide area of the British Columbia coast. However, local sources can still pre-empt the
influence of atmospheric input: DDE levels in heron eggs were significantly higher in colonies
from the Fraser delta (0.49 mg/kg), an area of intensive farming, than non-agricultural
locations (0.11 mg/kg) (Elliott et al. 1989; Whitehead, 1989). In fact, the mean DDE level in
two eagle eggs collected within the Fraser delta, 3.86 mg/kg, is significantly higher than the
four eggs from upstream of the main agricultural areas, 1.63 mg/kg DDE. High DDE levels
continue to be reported in wildlife from areas of former high DDT use, such as orchards (Blus
et al. 1987; Elliott et al. 1994).
After the DDT-related compounds, chiordanes were present at the highest concentrations
in eagle eggs. Among chiordanes, trans-nonachlor was consistently the dominant component,
constituting a mean of 67 % (SD =5.3, range 51-77 %) of the total. Oxychiordane, considered
to be the most stable metabolite (Nomeir & Hajjar 1987), made a mean contribution of 13 %
(SD =5, range 0.2-27 %). Some authors have suggested that a high ratio of trans-nonachlor to
oxychlordane levels in tissues shows a lower specific capacity to metabolize chlorinated
hydrocarbon compounds (Kawano et al. 1986; Yamashita et al. 1993).
The concentrations of chlordane-related and heptachlor epoxide compounds found here
are similar to those reported in addled Bald Eagle eggs collected in the early 1980s from a
variety of United States locations (Wiemeyer et al. 1993). Concentrations of mirex and
dieldrin were somewhat higher in those U.S. Bald Eagle eggs collected a decade earlier than in
the fresh eggs from the British Columbia coast in 1992. Mean DDE and PCB levels were
about three-fold higher in eagle eggs from the lower Columbia River than the lower Fraser
River (Anthony et al. 1993). Dietary differences may partly account for this; eagles in the
93
lower Columbia reportedly ate more Western Grebes (Watson et al. 1991), which tend to have
high levels of chlorinated hydrocarbons (Table 3.7), while Fraser estuary eagles ate a large
proportion of Glaucous-winged Gulls which tend to have low organochiorine levels (Table 3.7),
probably because in that area they consume mainly garbage (Vermeer et al. 1989). Differences
in organochiorine levels in estuarine biota also reflect differences in agricultural and industrial
development of the respective watersheds. Areas of intensive agriculture, particularly fruit
orchards are more prevalent in the Columbia basis and account for high DDT (Rinella et al.
1993). Hydroelectric development is much greater on the Columbia river and likely accounts
for higher PCB concentrations, evident in Osprey eggs collected in the upper reaches of each
watershed (Whitehead et al. 1993).
Higher mercury levels in Bald Eagle eggs from the Fraser estuary are consistent with
data in herons from that site (Elliott et al. 1989a), and with Fimreite et al. ‘s (1971) findings of
higher mercury in aquatic versus coastal marine fish. Elevated mercury levels in fish-eating
birds were associated with industrial, including pulp mill, sources by Fimreite et al. Based on
the levels in eagle eggs, any mercury discharges from Crofton and Nanaimo pulp mills have
not had a lasting impact in local food chains. Highest mercury levels were in the Johnstone
Strait eagle eggs. A great proportion of fish in the diet may explain higher mercury levels in
Johnstone Strait and the lower Fraser Valley, as suggested below to account for the PCB
pattern at those sites.
Polychiorinated bihenyls
Mean total PCBs in Bald Eagle eggs were highest near the three pulp mill sites, which
contrasts with data on great blue herons, in which highest PCBs were from colonies in the
Fraser delta near Vancouver (Elliott et al. 1989a; Whitehead 1989). However, the PCB
concentration in the single Bald Eagle egg from an industrial site in the Fraser delta, 6.21, was
in the same range as the eggs from near the pulp mill sites; other Fraser valley Bald Eagle eggs
were from agricultural or woodland locations and PCB levels were 50 % lower. The PCB
pattern in great blue herons varied significantly among sites which was attributed to local
94
differences in Aroclor inputs (Elliott et al. 1989a). Variability in PCB congener patterns in
wildlife in the the Green Bay area were also attributed to different industrial Aroclor sources
(Ankley et a!. 1993). However, in British grey herons, Boumphrey et al. (1993) considered
dietary differences as the best explanation for individual variation in PCB patterns. This may
also apply in Bald Eagles given the consistent differences in the PCB pattern between Johnstone
Strait and lower Fraser valley eggs compared to those from the Strait of Georgia sites. Total
PCB levels were also lower in the Johnstone Strait and lower Fraser valley eggs. The most
likely explanation is of more fish in the diet of Fraser and Johnstone Strait eagles and thus
greater exposure to the lower chlorinated PCBs. Higher total mercury levels at those two sites
are also consistent with more fish in their diet. The PCA results can be used to support this
explanation; however, alternatively the differences among sites may also be explained by
differing local Aroclor inputs. Fraser delta eagle eggs, like Great Blue Heron eggs, contain
more of PCB 66, indicative of Aroclor 1242 input, while the pulp mill areas, including
Crofton, generally contain more Aroclor 1260 peaks, again similar to Great Blue Herons
(Elliott et al. 1989a). A preponderance of lower chlorinated PCBs in the Johnstone Strait area
may be indicative of greater atmospheric sources over local industrial inputs (Eisenreich et al.
1981).
The single egg from the lower Fraser analyzed for non-ortho PCBs , Herrling Island,
also had a lower ratio of PCBs 126:77 than eagle eggs from other areas, also suggesting higher
consumption of fish which have low capacity to metabolize PCBs (Brown 1994). The ratio at
most sites of non-ortho PCBs 126:77 was 2:1, except in the eggs from Alberni Inlet and
Clayoqot Sound, where the ratio is closer to 1:1, and the egg from Herrling Island in the
Fraser valley, where the ratio was 1:2. Although the ratios vary somewhat, the other non
ortho PCB levels such as 169, are consistently less than either 77 or 126 in Bald Eagle eggs.
Bosveld and Van Den Berg (1994) suggested that lower levels of PCB 77 in adult tissues
compared to egg were caused by reduced metabolic capability in embryos. Levels of more
95
rapidly metabolized compounds such as PCB 77 may also be higher in eagle eggs as a result of
direct deposition of dietary lipids to egg yolk, as suggested above for 2,3,7,8-TCDF.
Comparison of total PCB levels to those in the literature is confounded by changes in
methodology. Total PCB numbers in Wiemeyer et al. (1993) were probably determined as
Aroclor estimates based on the analytical references. Determination of total PCBs based for
example on Aroclor 1254:1260 overestimate total PCBs, based on the sum of congeners, by
about two-fold (Turle et al. 1991).
Toxicological significance of PCDD and PCDF levels
The bioaccumulation model was developed in order to estimate critical concentrations of
2,3,7,8-TCDD and other contaminants in forage fish (eg. sculpin, perch and flounder species)
or fish-eating birds (herons, cormorants, waterfowl), components of the foodchain which are
more easily monitored than eagles. Levels in the monitoring species should indicate a degree
of foodchain contaminant which should result in accumulation in bald eagle eggs less than the
suggested NOEL from Chapter 2.
Using the same BMIF of 21, the average 2,3 ,7,8-TCDD concentration in forage fish
consumed by great blue herons in 1990 at Crofton would have been about 5 ng/kg. With the
postulated eagle diet in Table 3.8, TEQ5PCDD,PCDF. in eagle eggs were calculated as 193 ng/kg
versus the measured value of 248 ng/kg. If an average value of 115 ng/kg TEQs0 for non
ortho and mono-ortho PCB contribution at Crofton, 1990 is included, the total TEQs0were
308 and 355 ng/kg, calculated and measured respectively, both of which exceed the LOEL (210
ng/kg), determined for Bald Eagle embryos (Chapter 2). If the data from Crofton, 1992, are
used the estimated mean value of 1 ng/kg in forage fish gives a calculated TEQWIIO value in
eagle eggs of 194 ng/kg (79 TEQsPCDD,PCDFS + 115 TEQsPCBS), less than the LOEL, but still
greater than the NOEL of 100 ng/kg. Therefore, assuming that both the ratios of other
PCDDs/PCDFs and PCB levels remain constant, a maximum value of 0.5 ng/kg 2378-TCDD
in forage fish is suggested as site-specific dietary concentration in the Strait of Georgia, to
avoid adverse toxic effects of TCDD-like chemicals in Bald Eagle populations. The
corresponding concentration of 2,3,7,8-TCDD in Great Blue Heron eggs, to avoid TCDD
96
toxicity in both herons and top predators, such as the Bald Eagle, in the Crofton area is 10
ng/kg. At other areas in the Strait of Georgia, given that ratios of PCDDs, PCDFs and PCBs
are similar, a value of 10 ng/kg in double-crested or pelagic cormorants, would also indicate
that levels in local foodchains should not cause toxicity in Bald Eagles, given a typical diet, as
shown in Table 3.8. The utility of colonial waterbirds as sentinel species for monitoring of
toxic contaminants has been demonstrated in many studies (Gilbertson et al. 1987). Given that
the embryonic life stage appears to be the most sensitive to TCDD-like effects (Peterson et al.
1993) and that the NOEL from Chapter 2 was derived using a very sensitive endpoint, CYP1A
induction, then these critical values suggested for prey items, should provide a reasonable
margin of safety.
The above values would be effective in areas with contaminant profiles which are
similar to the Strait of Georgia. However, as shown in Figure 3.3, in Common Tern eggs,
PCDDs made only a minor contribution to the 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 den
Berg 1994). There are few published data on PCDD and PCDF levels in Bald Eagle eggs.
Mean 2,3 ,7,8-TCDD levels in live fresh Bald Eagle eggs collected in 1985-87 from the lower
Columbia river, were 32 ng/kg, less than those found in eagle eggs near pulp mills on the Strait
of Georgia. However, total PCB levels were 12.7 mg/kg, more than two-fold higher than the
highest mean concentrations in Table 3.4. Thus, TEQs110 in Bald Eagle eggs from the Lower
Columbia River would be dominated by the PCB contribution.
Other studies have reported high PCB levels in Bald Eagle egg and plasma samples;
however, because of correlations with DDE, no clear statistical relationships between PCBs and
productivity were determined (Wiemeyer et al. 1984; 1993; Bowerman 1993, Dystra 1994;
Welch 1994). Recent studies of PCB toxicity in other avian species have focused on the non
ortho PCBs, particularly 126 and 77, and certain mono-oilho PCBs, such as 118 and 105,
which are partial Ah-receptor agonists and thus cause TCDD-like toxicity in laboratory animals
(Safe, 1990) and apparently in wildlife (Kubiak et al. 1989; Bosveld et al. 1994; Sanderson et
97
at. 1994b). However, the data on Bald Eagle chicks reported in Chapter 4 suggests that PCB
congeners are less potent relative to PCDDs and PCDFs in Ah-receptor mediated biomarker
responses, such as CYP1A induction. Nevertheless, total PCB levels up to 119 mg/kg have
been reported in recent years in adled Bald Eagle eggs from the Great Lakes region (Bowerman
et al, 1994); that egg would have contained about 18,500 ng/kg of PCB 126 using the
regression from the Result section above. PCB concentrations of that degree may partly
account for the poor productivity and reports of deformed young in the Great Lakes region.
Although the data are not shown here, the same modelling approach can be used to
determine total PCB concentrations in foraging fish and a sentinel fish-eating bird, which would
result in a PCB contribution to TEQ5 in eagle eggs less than the NOEL of 100 ng/kg. Using
the BMF for PCBs of 30 from Braune and Norstrom (1980), assuming constant ratios of non
ortho and mono-ortho PCBs to total PCBs, for Crofton (assuming TEQSDD/PCDFS = 79 ng/kg)
site-specific values of 0.01 mg/kg in forage fish and 0.3 mg/kg in fish-eating bird eggs are
suggested, would be necessary to achieve TEQs0 less than the NOEL of 100 ng/kg in Bald
Eagle eggs. This value for forage fish is much lower than 0.2-0.4 mg/kg suggested by Harris
et at. (1993) to produce a NOEL in Forster’s Tern eggs in Green Bay, Michigan. However,
eagles feed at a higher trophic level than terns; therefore, a lower target level in forage fish
would be required to avoid accumulation of toxic levels in eagles. However, at most sites the
contribution of PCDDs and PCDFs to TEQs is considerably less than at Crofton and is
probably in the order of 25 ng/kg, in which cases a higher PCB contribution could be tolerated.
Application of this or more sophisticated models to sites with both lower PCDDs and PCDFs
and a comprehensive dataset on PCBs would enable determination of better guidelines for
PCBs.
Toxicological signjficance of organochtorine and mercury levels
Wiemeyer et at. (1993) determined that DDE was the chemical contaminant most
associated with reduced breeding success of Bald Eagles in the United States during the period
1969 - 1984. Production of young began to decrease at DDE levels > 3.6 mg/kg, and further
decreased at > 6.3 mg/kg. DDE levels of 16 mg/kg were associated with fifteen percent
98
eggshell thinning, a threshold related to population declines in other raptors (Noble et at.
1993). Wiemeyer et at. (1993) also found a highly significant relationship (r = 0.912, p <
0.0001) between DDE and shell thickness in a large sample of Bald Eagle eggs from the United
States. Mean DDE levels in the eggs in Table 2.2 were all less than 16 mg/kg, although 31 %
(11/35) contained > 3.6 mg/kg and nine percent (3/35) had > 6.3 mg/kg. Although mean
eggshell thickness was less than the pre-1946 mean at all sites, there was no significant
relationship between DDE and eggshell thickness, likely because of the narrow range of DDE
concentrations.
Although quantitative data are limited, there were no reports of widescale declines of
coastal eagle populations in British Columbia, as occured in other areas of North America
during the organochlorine era. However, Vermeer et al. (1989) reported an increase in Bald
Eagles nests in the southern Strait of Georgia between the mid-1970s and the late 1980s. They
attributed eagle population growth to increased prey populations, particularly glaucous-winged
gulls, populations of which had increased due to greater availability of human refuse.
However, in the 1970s, DDE and other organochlorines were also likely much higher in Strait
of Georgia eagle eggs. In Great Blue Heron eggs from a Fraser delta colony, DDE declined
from a mean of 2.0 mg/kg in 1977 to 0.42 mg/kg in 1990 (Whitehead 1989; Canadian Wildlife
Service, unpublished data). In Pelagic and Double-crested Cormorant eggs from Mandarte
Island in the southern Strait of Georgia, DDE decreased by factors of five and ten respectively
between the early 1970s and the late 1980s (Elliott et at. 1989a). Organochlorine levels in
Bald Eagle eggs are currently about ten-fold higher than in those of marine and fish-eating
birds from the Pacific coast (Elliott et at 1989a; 1989b). If the ten-fold difference was constant
over time, then during the late 1970s mean DDE levels in Bald Eagle eggs from the Fraser
delta would have been about 25 mg/kg, high enough to cause nest failures and reduced
productivity. It is probable, therefore, that the population increase reported by Vermeer et at.
(1989) was partly due to declining DDE levels. In the Okanagan valley of interior British
Columbia, Bald Eagles declined as a breeding species between the 1930s and 1970s (Cannings
et at. (1987). Although habitat loss was likely a factor, the extremely high DDE levels in
99
Okanagan valley foodchains (Elliott et al. 1994) probably continue to impact Bald Eagle
reproduction in that area.
None of the Bald Eagle eggs analyzed in this study had mercury levels > 0.5 mg/kg
(wet weight), determined by Wiemeyer et al. (1993) to be associated with effects on
productivity.
In conclusion, Bald Eagle eggs collected in the Strait of Georgia contained elevated
levels of PCDDs and PCDFs; the pattern was similar to that measured in other components of
food chain and indicative of both bleached kraft pulp mill and chiorophenol sources. Relatively
high PCDDs and PCDFs in a supposed reference area in northern Johnstone Strait probably
resulted from feeding on waterbirds migrating north from the Strait of Georgia. Recommended
site specific concentrations of 2,3,7 ,8-TCDD are 0.5 ng/kg in forage fish and 10 ng/kg in
sentinel fish-eating bird eggs in the Strait of Georgia are suggested to avoid accumulation of
potentially harmful levels in Bald Eagle eggs. Likewise, total PCB concentrations of 0.01
mg/kg in forage fish and 0.3 mg/kg in fish-eating bird eggs are suggested as maximum
concentrations to prevent accumulation of potentially harmful PCB levels in Bald Eagle
populations.
Acknowledgements
Ian Moul, George Compton, Andre Breault, Dave Dunbar and Ray Caton assisted with
collection of eggs. Mary Simon did the PCDD/PCDF and non-ortho PCB analysis, while
Henry Won did the organochlorine pesticide analysis. John Smith provided statistical advice.
Funding was provided by the Canadian Wildlife Service and the British Columbia Ministry of
Environment.
100
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CHAPTER 4
INFLUENCE OF CONTAMINANTS AND FOOD SUPPLYON BALD EAGLE PRODUCTIVITY
The results of the previous chapters showed that Bald Eagle populations in the Strait of
Georgia were exposed to elevated levels of PCDDs and PCDFs relative to reference
populations. Eggs collected in 1990 and 1991, particularly near Crofton, had higher
PCDD/PCDF levels and modelling showed that theoretically eagles which preyed on a larger
component offish-eating birds in the diet would acquire a substantial TCDD body burden.
Among dead eagles examined between 1988 and 1993, about 20 % of a sample of 19 adults,
found during the breeding season in the Strait of Georgia, contained TEQ5WHO > 1,000 ng/kg in
livers. Thus, some component of the breeding population may be affected each year by
chlorinated hydrocarbon toxicity. Eggs collected near pulp mills in 1992 and incubated in the
laboratoiy did not exhibit significant effects on hatchability and most morphological and
physiological endpoints, although a hepatic CYP1A cross-reactive protein was induced. For the
work described in this chapter, I measured breeding success of Bald Eagles near three pulp
mills in the Strait of Georgia, at two areas of the Fraser delta, and at reference sites on the
west coast of Vancouver Island, in northern Johnstone Strait and in the Queen Charlotte
Islands. The objective of the study was to determine occupancy of breeding territories, measure
nest success and compare the results to chlorinated hydrocarbon levels in nestling plasma
samples.
Most previous studies of contaminants in Bald Eagles (for example Wiemeyer et al.
1993) used addled eggs, because of concern that collection offresh eggs would impact already
poor reproduction. My initial studies on the coastal BC eagle breeding population (Chapters 2
and 3) used eggs collected during incubation; however, this resulted in an unacceptable level of
nest abandonments, even when only two egg clutches were sampled and a single egg removed.
Some researchers had previously used blood samples of nestling eaglets to obtain a more
102
randomized sample for contaminant analysis (Henny et a!. 1981; Frenzel 1985), an approach
that has been used more frequently in recent years (Anthony et a!. 1993; Bowerman 1993;
Dykstra 1994, Welch 1994). Blood sampling has the further advantage of not eliminating nests
from productivity estimates from an area, and also permits determination of a direct
relationship between contaminant levels in chicks and 5-year average productivity for the
territory in which they were produced. Because of development of advanced high resolution
gas chromatography/mass spectrometry (GC/MS) analytical techniques, beginning in 1993, the
NWRC lab was able to quantify PCDDs, PCDFs and non-ortho PCBs in nestling Bald Eagle
plasma samples.
Materials and Methods
Productivity
Survey routes were flown in exposed or ‘treatment’ study areas, selected on the basis of
eagle nest density near industrial pollutant sources: Crofton, Nanaimo and Powell River (pulp
mills) and the lower Fraser valley (mixed industrial sources) (Figure 4.1). Reference or control
sites were based on concentrations of nesting eagles remote from industrial point sources:
Barkley and Clayoquot Sounds, northern Johnstone Strait and the Queen Charlotte Islands.
Bald Eagle breeding success was estimated in each area by a standard two-flight
approach (Fraser et al., 1983) using rotary aircraft (Bell jet/long ranger or Aerospatial Astar).
A minimum of two observers were used. The first survey took place during incubation to
determine the number of eagle pairs attempting to breed. Timing of this flight varied from late
March in the Fraser delta to mid-May in the Queen Charlotte Islands. The second flight was
timed to count nestlings at 5-8 weeks of age and took place between late May and early July.
Mean productivity at each study area was determined by dividing the total number of young
produced by the number of occupied breeding territories, as described in Postupalsky (1974).
103
Figure 4.1 Locations of Bald Eagle productivity survey routes and blood collections. At Langara Island,the survey circumscribed the coastline of the island.
Ce
Ce)CC
=ci
eC
104
Prey deliveries
Prey deliveries were observed during 1995 at five nests in the Fraser delta and at four
nests in Barkley Sound, using methods described in Dykstra (1994). From a blind, dawn to
dusk observations were made using a 20-60X spotting scope. Prey deliveries and other nestling
and adult behaviours were recorded. Observers were switched every four to eight hours. In
the Fraser delta five nests were observed for five days each. In Barldey Sound, three nests
were observed for five complete days each, while one nest was watched for part of each day
and therefore was not included in statistical analyses.
Sample collection
Nests suitable for collecting were located by ground, boat and aerial surveys, when they
were scored to estimate ground access and suitability for climbing; land tenure was also
considered. Samples were collected when nestlings were 5-9 weeks old. Collections were
made during the first week of June in the Fraser Valley and the Strait of Georgia, during late
June or early July on the west coast of Vancouver Island, Powell River and Langara Island
(Figure 4.2).
Nests were accessed by a professional tree climber. Nestlings were lowered to the
ground in a soft bag, weighed and aged by measuring the length of the eighth primary feather
(Bortolotti 1984). Up to 24 ml of blood was drawn from the brachial vein (12 ml per wing)
using a 12 ml sterile disposable syringe and a 21 gauge needle. Blood was transferred
immediately to heperinized vacutainers and stored on ice. Samples were centrifuged within six
hours of collection and plasma transferred to chemically cleaned (acetone/hexane) glass vials
with teflon liners and then frozen.
Chemical analysis
Frozen plasma samples were shipped to the CWS National Wildlife Research Centre
(NWRC) for analysis in the laboratory of Dr. R.J. Norstrom. For organochlorines, the
samples (1 ml of each) were first deproteinized with 0.5 ml of methanol containing aldrin as an
internal standard (Smrek et al. 1981). The plasma was then extracted with hexane and
105
Dio
xin
Fis
her
y
LXZJ
Clo
sure
Are
a
•B
ald
Eag
lenea
tsi
tes
surv
eyed
So
uth
eyI.
‘<—
Win
ch
els
ea
I.
.
\rvia
ude
I.
Pow
ell
Riv
erP
ulp
Mill
Cro
fto
nP
ulp
Mill
VA
NC
OU
VE
R
05
10km
II
I
centrifuged. The hexane extracts were passed through sodium sulphate, evaporated to 1 ml or
less and separated into three fractions with hexane and methylene chloride on a florisil column.
Analyses were performed by GC-electron capture detector with capillary-column separation on
a Hewlett Packard 7673A. PCBs were quantitated as the sum of 33 major congener peaks.
Quality assurance procedures included the simultaneous analysis of 6 diluted Herring Gull egg
pool reference material samples (Tune et al 1991).
Plasma samples (1.98 - 12.94 gram samples) were simultaneously analyzed for PCDDs,
PCDFs and non-ortho PCBs as follows: isotopically labelled internal standards (‘3C12-
PCDDsIPCDFsInon-ortho PCBs) were added to the plasma, and allowed to equilibrate for 30
minutes. Saturated aqueous animonium sulphate and absolute ethanol were added to the spiked
plasma, and the samples were then extracted four times with hexane. The hexane layers were
combined, filtered through anhydrous sodium sulphate and the volume reduced for clean-up
with by gel penneation chromatography (GPC) (Norstrom et a!. 1986). Lipids and biogenic
materials were removed by GPC and alumina column clean-up. Separation of PCDDs, PCDFs
and non-ortho PCBs from other contaminants was achieved using a carbon/fibre column
(Norstrom and Simon 1991); further separation of PCDDs and PCDFs from the non-ortho
PCBs was done with florisil column chromatography. Quantitation was performed with a VG
Autospec double-focusing high resolution mass spectrometer linked to a HP 5890 Series II high
resolution gas chromatograph. Recoveries of13C12-PCDDsIPCDFs/non-ortho PCBs were
calculated by comparing the integrated areas of the labelled internal standards and the areas of
the recovery standards in the samples to the areas of these compounds measured in an external
standard mixture, analyzed along with the samples. Results were accepted when recoveries of
13C12-PCDDs/PCDFsInon-ortho PCBs were between 70% and 120%. For a few Bald Eagle
plasma samples, the internal standard recoveries were <70%, due to losses during lipid
extraction.
Lipid was determined by combining 1-2 ml of sample with 4 ml of hexane in a
centrifuge tube, which was then extracted with an Ultra-Turrax homogenizer for 2 minutes.
The contents of the tube were then centrifuged to separate the hexane and plasma layers,
107
similar to the method of Mes (1987). The hexane was then passed through sodium sulphate to
remove any moisture. This process was repeated twice more and the sodium sulphate washed
with hexane after the final extract. The three hexane extracts were combined on a pre-weighed
aluminum dish, the hexane was then evaporated and the dish re-weighed to determine the
amount of lipid. Lipid was then calculated on the basis of grams per ml plasma.
Statistical analyses
The SYSTAT software package was used for statistical analyses of all data. Wet weight
chemical residue data were transformed to common logarithms and geometric means and 95 %
confidence intervals were calculated with the data grouped by collection site. The majority of
chlorinated hydrocarbons tested were significantly correlated with percent plasma lipid (Table
4.1). DDE was only weakly correlated with plasma lipid, while the higher chlorinated PCDDs
and PCDFs were not significantly correlated. There was also a significant interaction between
plasma lipid and sampling location. Therefore, for testing of differences among locations, all
of the contaminants which correlated significantly with plasma lipids, were further transformed
using an analysis of covariance (ANCOVA) to account for the effect of variation in plasma
lipid content among individuals and locations (Hebert and Keenlyside, 1995). Differences
among locations were then determined using Bonferroni’s test. In a few cases, percent plasma
lipids were three to ten-fold greater than the mean of the other samples at that site; those
samples were fatty in appearence and the nest contained fresh, partly eaten prey remains,
indicating that the chick was sampled during or immediately after feeding. Those ‘outliers’
were not removed from the data, rather, it was assumed that they were corrected by the
ANCOVA.
Productivity measures were compared among locations with a one-way analysis of
variance (ANOVA); significant differences were determined using Tukey’s multiple comparison
procedure (MCP). Data were also compared on the basis of a pulp mill versus non-pulp mill
grouping and significant differences identified using Student’s t-test. At each pulp mill site,
108
Tab
le4.
1C
orre
lati
onM
atri
x(r
valu
e)fo
rper
cent
plas
ma
lipid
and
sele
cted
chlo
rina
ted
hydr
ocar
bon
inba
ldea
gle
nest
ling
sfr
omB
ritis
hC
olum
bia,
1993
—94
Lip
id
OC
DD
HpC
DD
HX
CD
D
PnC
DD
TC
DD
OC
DF
HxC
PnC
DF
TC
DF
DD
E
HL
B
CM
irex
1—n
on
ach
lor
SUM
—P
CB
s
PC
B—
99
PC
B—
118
PCB
—15
3
PCB
—18
0
PC
B—
37
PC
B—
77
PCB
—12
6
PCB
—16
9
TE
Qa
9911
815
315
037
0.93
40.
942
0.94
20.
942
0.99
50.
993
0.99
5
0.94
20.
238
—0.
046
—0.
082
—0.
051
—0.
170
0.88
80.
442
0.94
90.
393
0.98
20.
328
7712
616
9T
EQ
sP
rodu
ctiv
ity
0.94
30.
867
0.95
8—
0.07
1
—0.
046
—0.
081
—0.
015
—0.
083
—0.
050
—0.
095
—0.
005
—0.
073
0.84
00.
779
0.89
3—
0.08
6
0.90
80.
842
0.95
2—
0.07
0
0.95
40.
870
0.98
6—
0.07
4
0.90
60.
990
—0.
114
0.90
00.
986
—0.
137
0.33
00.
343
—0.
094
0.93
30.
977
0.09
4
0.96
40.
987
0.07
4
0.92
70.
004
—0.
080
tran
sS
um
Lip
idO
CD
DH
pCD
DH
XC
DD
PnC
DD
TC
DD
OC
DF
HxC
LE
PnC
DF
TC
DF
DD
EH
CB
Mir
exn
on
aclo
r—
PC
Bs
—0.
025
—0.
008
0.77
10.
872
0.96
1—
0.00
40.
032
0.94
20.
931
0.56
90,
868
0,89
80.
969
0.96
10.
978
0.97
30.
964
0.83
5—
0.03
5—
0.05
8—
0.03
50.
134
0.79
1—
0.03
1—
0.07
7—
0.04
6—
0.07
7—
0.07
4—
0.06
0—
0.03
4—
0.04
0—
0.02
7—
0.03
3
0.04
1—
0.04
7—
0.02
80.
179
0.58
7—
0.02
9—
0.08
5—
0.05
9—
0.10
1—
0.08
7—
0.06
2—
0.04
0—
0.04
3—
0.02
7—
0.03
7
0.96
20.
896
0.08
10.
066
0.84
10.
886
0.47
60.
634
0.78
40.
851
0.87
00.
830
0.85
10.
857
0.96
4—
0.00
70.
056
0.88
40.
943
0.53
40.
744
0.86
20.
923
0.93
80.
909
0.92
50.
926
—0.
005
0.04
00.
927
0.96
70.
580
0.01
00.
904
0.98
30.
986
0.97
70.
983
0.98
0
0.91
9
—0.
001
—0.
008
0.85
2
0.90
8
0.94
3
0.26
1—
0.02
1—
0,00
8—
0,06
5—
0.05
5—
0.07
4—
0.03
4—
0.02
5—
0.03
5—
0.02
8—
0.03
0—
0.02
30.
183
—0.
031
—0.
051
—0.
065
—0.
019
0.18
2
0.05
70.
007
—0.
101
—0.
042
—0.
045
—0.
018
0.00
5—
0.00
60.
009
0.00
40.
004
0.05
4
0.94
10.
528
0.85
40.
866
0.93
80.
931
0.94
30.
942
0.93
6
0.53
20.
799
0.87
20.
956
0.95
50.
953
0.95
50.
949
0.68
10.
624
0.65
40.
629
0.62
50.
618
0.61
8
0.89
70.
849
0.84
40.
864
0.85
50.
845
0.04
80.
032
0.04
30.
050
0.13
5
0.91
70.
364
0.94
90.
444
0.61
80.
091
0.83
00.
135
0.95
00.
200
0.98
70.
316
0.94
3
0.92
5
0.56
5
0.79
2
0.90
9
0.95
9
0.94
8
0.99
4
0.95
8
0.92
5
0.60
8
0.85
3
0.94
0
0.97
3
0.92
90.
959
0.01
4
0.86
00.
962
0.03
8
0,54
40.
600
—0.
349
0.86
50.
840
—0.
128
0.92
50.
932
—0.
142
0.89
40.
987
—0.
122
0.99
40.
990
0.99
80.
996
0.30
00.
963
0.97
20.
895
0.99
0—
0.11
9
0.99
80.
996
0.98
40.
278
0.95
90.
973
0.89
90.
985
—0.
104
0.99
10.
989
0.28
40.
965
0.97
50.
901
0.99
0—
0.10
3
0.29
90.
289
0.96
9
0.39
40.
961
0.39
4
0.97
8
0.97
0
0.32
4
0.98
3
productivity at nests adjacent to dioxin fishery closure areas was compared to nests adjacent to
areas outside the closure area, using a one-way ANOVA. We treated the closure areas as an
indication of the area impacted directly by PCDD and PCDF contaminants in the respective
pulp mill effluents. Mean 3-year productivity at individual nests was also compared to
contaminant levels in nestling blood samples from each nest using regression analysis. Unless
stated otherwise, a value of p < 0.05 was considered statistically significant in all analyses.
TCDD-toxic equivalents (TEQs) were calculated using the toxic equivalency factors
(TEF5) proposed by Ahlborg et al. (1994) and referred to here as the WHO (World Health
Organization) TEFs.
Results
Productivity
Mean three-year productivity was highest at nests in the Fraser valley and delta and
comparable along south-east Vancouver Island (Table 4.2). The number of young/occupied
territory was lower (significantly compared to the lower Fraser valley) at nests around Powell
River and at Langara Island. Lowest productivity was in Clayoquot Sound, Johnstone Strait
and South Moresby.
Productivity of eagle nests located along the shoreline adjacent to the dioxin fishery
closures in the Crofton area was significantly lower than at nests located outside the closure
area (Figures 4.2 and 4.3). There were no significant differences in productivity at nests
adjacent to the closure areas compared to those outside the closure areas at both Nanaimo and
Powell River. However, the four eagle nests closest to the mill on the north side (Powell River
nest and three Gibson’s Beach Park nests) produced only two chicks between 1992 and 1994 in
nine nesting attempts. In contrast the next five nests to the north (three Scuttle Bay nests, Kees
Bay and Lund) during the same time frame produced 21 chicks in 15 nesting attempts. This
difference, was not statistically significant, however, likely due to small sample sizes.
110
Table 4.2 Nest success and production of young for Bald EaglesBritish Columbia coast (1992-94).
at nine study areas on the
a,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 produced
1992 19 19 100 27 1.41993 22 18 82 27 1.21994 21 18 86 29 1.4
Lower Fraser Valley
Fraser Delta
South-east VancouverIsland
Powell River
Baridey Sound
Clayoquot Sound
Johnstone Strait
South Moresby
Langara Island
Mean 89 1.3a
1993 9 7 78 12 1.31994 11 9 82 14 1.3
Mean 86 l.3
1991 19 11 58 17 0.901992 30 19 63 30 1.001993 34 22 65 35 1.001994 42 27 64 43 1.00
Mean 63 0. 97ab
1992 24 14 58 18 0.751993 37 25 68 36 0.971994 36 21 58 33 0.92
Mean 61 0.88&
1992 36 16 44 21 0.581993 35 20 57 26 0.741994 30 8 27 12 0.4
Mean 43 0.57c
1992 23 12 52 12 0.521993 43 10 23 14 0.331994 35 2 57 3 0.09
Mean 27 0.31d
1991 6 2 33 2 0.331992 26 10 39 12 0.461993 34 3 8.8 4 0.121994 31 13 42 14 0.45
Mean
1994
1994
19
22
5
13
31
26
59
6
16
0.34’
0.32’
0.73
111
1.2
4-,Cl)ci)z-c,a)0.8aDC.)C)o 0.6
D0>-oZ 0.2
0
Outside
: Inside
0
Figure 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. Sample
sizes were: Powell River, N=20 inside and N=26 outside; Nanaimo, N =15 inside and N =13 outside;Crofton, N=9 inside and N=8 outside.
No significant correlations occurred between productivity and any of the PCDD, PCDF
or PCB compounds measured or with TEQs (Figure 4.4a). For the organochiorine pesticides,
log-DDE in nestling plasma regressed weakly with 3-year average productivity for each
corresponding territory (r2 = 0.128, p < 0.011, Figure 4.4b).
112
A2.5 -
. 2- ••ci
• .1A4D 0
) AALI.C
1- LZJs.c>
AA
I I 1111111 11111111
0.1 1 10 100
BTEQs - WHO (ng/kg wet weight)
2.5-
L
ci)
ci)> U
0U)
E’i- .0
+1 A2 A0
0— 111111 ‘‘I 1111111 1111111
1 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 Langara
Island.
113
Mean percent lipid in plasma regressed positively on mean productivity among sites
(Figure 4.5).
1.4
1.2
0a) I
a)0.8DC.)0oO.6c3)C
0>-
0.2
0
Figure 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 deliveries
At surveyed nests in the Fraser Delta, mean daily prey deliveries were greater at, 3.5 than
at Baridey Sound nests, 2.4 deliveries per day; however the difference was not statistically
significant, likely due in part to small sample sizes in this pilot study.
r2=O.423
U
0.01 log (% plasma lipid) 0.1
114
PCDD and PCDF levels in plasma
Plasma samples from 52 Bald Eagle chicks were analyzed for PCDD and PCDF levels
(Table 4.3). The pattern in plasma near pulp mill sites was generally 2,3,7, 8-TCDF >
l,2,3,6,7,8-HxCDD > 1,2,3,7,8-PnCDD > OCDD > 2,3,7,8-TCDD. At other sites, OCDD
was often comparable or greater than 1 ,2,3,6,7,8-HxCDD, while in the Fraser delta, OCDD
and 1,2,3,4,6,7,8-HpCDD were the dominant congeners. Most samples also contained
detectable amounts of 2,3 ,4,7,8-PnCDF.
Because of the significant interaction with plasma lipid content, selected PCDDs and
PCDFs are further presented as lipid-adjusted log-normalized mean values (Figure 4.6). Mean
plasma TCDD concentrations were significantly higher at Powell River and East Vancouver
Island than other sites. Mean concentrations of PnCDD, HxCDD and TCDF were also highest
near the pulp mill sites at Powell River and along east Vancouver Island; however, the
differences were not consistently significant from the Fraser Delta and Johnstone Strait.
Highest mean levels of HpCDD and OCDD occurred in samples from the Fraser Delta,
although the mean was not significantly different from east Vancouver Island.
PCBs in plasma
Highest concentrations of total PCBs were in samples from Powell River and east
Vancouver Island (Table 4.4), which on a lipid-adjusted basis were significantly greater than
Clayoquot Sound and the Fraser valley (data not shown). Mean concentrations of individual
PCB congeners generally followed the geographical pattern of the total PCBs; for example,
highest concentrations of PCBs 153 (245-245) and 105 (234-34) were also at Powell River and
east Vancouver Island and were significantly different from Clayoquot Sound and the lower
Fraser Valley.
115
Tab
le4.
3P
CD
D/P
CD
Fle
vels
,ge
omet
ric
mea
nsan
d95
%co
nfid
ence
inte
rval
(ng/
kg,
wet
wei
ght)
inbl
ood
plas
ma
ofB
ald
Eag
lech
icks
from
the
coas
tof
Bri
tish
Col
umbi
a,19
93-9
4.
Loc
atio
nN
2378
1237
812
3678
1234
678
OC
DD
2,3,
7,8
1237
823
478
2346
78O
CD
FT
CD
DP
nCD
DH
xCD
DH
pCD
DT
CD
FP
nCD
FP
nCD
FH
xCD
F
Fra
ser
Del
ta5
0.07
0.23
0.45
1.7
2.4
0.11
0.07
0.11
0.16
0.13
0.04
-0.3
1.0
9-.5
90.
08-2
.50.
12-2
50.
19-3
10.
02-0
.69
0.01
-0.4
10.
04-0
.28
0.04
-0.6
10.
07-0
.25
Low
erF
rase
rV
alle
y5
0.05
0.14
0.07
0.09
0.30
0.19
0.02
0.02
0.12
0.08
0.04
-0.0
6.1
-.19
0.02
-0.3
20.
04-0
.23
0.19
-0.4
70.
15-0
.25
0.01
-0.0
40.
01-0
.04
0.04
-0.3
50.
04-0
.17
Eas
tVan
couv
erls
land
110.
330.
623
1.2
0.31
1.1
2.8
0.15
0.21
0.16
0.16
0.21
-0.5
3.3
7-1.
10.
59-2
.40.
20-0
.48
0.7-
1.6
2.0-
3.8
0.10
-0.2
30.
13-0
.35
0.12
-0.3
50.
07-0
.38
Pow
ell
Riv
er10
0.37
0.90
2.2
0.13
0.56
4.54
0.12
0.27
0.09
0.05
0.17
-0.8
0.5
1-1.
61.
2-3.
90.
07-0
.23
0.40
-0.7
72.
69-7
.66
0.07
-0.2
10.
13-0
.55
0.06
-0.1
40.
03-0
.09
Bar
kley
Soun
d8
0.01
0.02
0.07
0.04
0.57
0.24
0.16
0.16
0.06
0.05
0.01
-0.0
3.0
1-.0
50.
03-0
.17
0.03
-0.0
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-Ua)ci,
-U
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U)
U)
2
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a
b 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. Means
that do not share the same lower case letter are significantly different (p <0.05).
1 2378- PnCDD 23478- P, CD F
1 23678-HCDD OCDD
2.5 a- -
117
Table 4.4 Organochiorine levels, geometric means and 95% confidence interval (tg/kg,wet weight) in blood plasma of Bald Eagle chicks from the coast of BritishColumbia, 1993-94.
Location N Total DDE trans- Oxychiordane Dieldrin Mirex HCBPCBs nonachlor
Fraser 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 0.1-0.8
Lower 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 0.2-0.5
East 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 0.2-0.5
Powell 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 0.3-1.0
Barkley 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 0.2-0.6
Clayoquot 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 * * * 0.1-0.7
Johnstone 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 0.2-0.5
Langara 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 0.3-2.2
ND - Not detected, minimum detection limit 0.01-0.05 nglkg, wet weight.*
- values all the same
In Bald Eagle plasma samples the general pattern of non-ortho PCB congeners was: 77
(34-34) 37 (34-4) > 126 (345-34) > 169 (345-345) > 81(345-4) (Table 4.5). Highest
lipid-adjusted mean concentrations of individual congeners were generally in samples from
Powell River or east Vancouver Island, although the highest mean concentrations of PCB 169
were from Langara Island (Figure 4.7).
Organochiorines in plasma.
Highest mean organochlorine pesticide levels were in samples from the Strait of Georgia
region, including the Fraser Delta and from Langara Island (Table 4.4). Most lipid-adjusted
plasma OC levels did not differ significantly among sites. Mean oxychiordane levels were
significantly greater at Langara Island than either Johnstone Strait or the lower Fraser Valley.
118
PCB 126 PCB 169
Figure 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 not
share the same lower case letter are significantly different (p <0.05).
PCB 37 PCB 77
. ‘,. . c,& q• c \G3
—.
119
Table 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 189
Fraser 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 0.05-0.37
Lower 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 0.04-0.22
East 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 0.38-0.74
Powell 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 0.23-1.49
Barkley 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 0.11-0.46
Clayoquot 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 0.07-0.38
Johnstone 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 0.05-0.75
Langara 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 0.03-0.27
ND - Not detected, minimum detection limit 0.01-0.05 ng/kg, wet weight.
Mean mirex concentrations were significantly greater at Langara Island and Powell
River than the lower Fraser Valley.
Discussion
Higher concentrations of chlorinated hydrocarbons in Bald Eagle nestlings from the
Strait of Georgia were not associated with significant effects on breeding success at most sites.
With the exception of a sample of nests near Crofton, mean 3-year productivity at study sites
around the strait, particularly the estuary of the Fraser river, was substantially higher than the
0.7 young/occupied territory, necessary to sustain an eagle population (Sprunt et al. 1973). In
contrast, eagle productivity at the more remote reference sites was generally less than 0.7. Only
at Langara Island at the north end of the Queen Charlotte archipelago, an area of high
biological productivity, was eagle breeding success comparable to the Fraser delta and the Strait
120
of Georgia. Using nestling plasma lipid content as a marker of body condition, food supply is
likely the main factor limiting eagle productivity on the British Columbia coast. However, low
productivity at a sample of eagle nests adjacent to the dioxin fishery-closure zone at Crofton is
probably not caused by differences in food availability.
The geographic pattern of PCDDs and PCDFs in plasma is similar to that found in eagle
eggs and is discussed in detail in Chapter 3. Essentially, elevated levels of TCDD, PnCDD,
HxCDD and TCDF are associated with pulp mill sources. Elevated HpCDD and OCDD in the
Fraser delta samples likely reflect heavy past use of chlorophenolic wood preservatives in that
area, and some contribution from combustion sources.
There are few published data on PCDD and PCDF levels in avian plasma. Blood
samples of osprey nestlings taken in 1992 downstream of a bleached-kraft pulp mill on the
Thompson River, in the interior of British Columbia, did not contain any lower chlorinated
dioxins and furans (minimum detection limit = 0.5 ng/kg, wet weight); sample sizes were
small, however, averaging about 3.6 ml of plasma. OCDD and HpCDD (0.1 - 1.0 ng/kg)
were detected in most osprey samples (Norstrom and Simon 1994). Osprey eggs from the
same sites in 1991 contained relatively high concentrations of TCDD, TCDF, HpCDD and
OCDD (Whitehead et al. 1993).
In bald eagles, five of the six non-ortho compounds displayed a good correlation with
plasma lipid content, while PCB 37 was only weakly correlated with plasma lipid. Ratios of
PCB 37 relative to other congeners were high in eagle plasma compared to eggs or liver. High
ratios of PCB 37 to other non-ortho PCB congeners were also found in osprey samples
(Norstrom and Simon 1994) This suggests that PCB 37 may bind with plasma proteins.
Corraborative data on PCDDs, PCDFs or non-ortho PCBs in avian blood samples from other
studies is unavailable. However, studies of human subjects have shown that, although absolute
levels on a lipid weight basis were much lower than those found in the eagle samples, OCDD
was the major congener present (Papke et a!. 1990). In humans, blood:adipose tissue ratios are
highest for OCDD compared to other PCDDs and PCDFs (Schechter et al. 1990). As we found
121
with eagles, OCDD did not partition with lipid in human blood; it is believed to bind primarily
to serum protein components (Patterson et al. 1989).
Published data on total PCBs and DDE in avian plasma samples is available from a
number of studies. Mean concentrations of PCBs and DDE in plasma of nestling Bald Eagles
from the lower Columbia River, 1984-86 were 0.04 and 0.05 mg/kg, wet weight, respectively,
(Anthony et al. 1993); those levels were comparable to eagle plasma samples from Powell
River and east Vancouver Island nests. Meanwhile, PCB and DDE levels in eggs were about
three-fold higher in eagle eggs from the lower Columbia compared to the Strait of Georgia
(Anthony et al. 1993; Chapter 2). However, plasma lipid levels were not reported for the
lower Columbia; therefore, the low levels of PCBs and DDE in those samples may reflect low
plasma lipid levels.
Geometric mean levels of DDE and PCBs (wet weight) in eagle plasma samples
collected between 1987 and 1993 from less contaminated areas of the Great Lakes were
comparable to samples from our reference sites: DDE, 3-12 ng/kg and total PCBs 5-34 ng/kg
(Bowerman 1993; Dykstra 1994). Levels of DDE in eaglets from most shoreline areas of the
Great Lakes, 20-25 ng/kg, were comparable to data for the Strait of Georgia and Langara
Island. Eaglets from Lake Michigan had somewhat higher levels, 35 ng/kg, DDE, than other
sites. Mean levels of total PCBs in nestling eagle blood samples from the Great Lakes
shoreline were two-fold (Lake Superior) to four-fold (Lake Erie) higher than Strait of Georgia
samples. Maine eagle blood samples, 1991-1992, particularly from estuarine sites, had up to
150 ng/kg DDE and 1,250 ng/kg total PCBs (Welch 1994). However, plasma lipid data were
also not reported for either the Great Lakes or Maine samples. The potential influence of
geographic variation in plasma lipids on contaminant levels is particularly relevant for some
Great Lakes samples, as Dykstra (1994) determined that low food availability was the main
cause of poor breeding success at the Lake Superior nests, compared to those inland. This was
reflected in lower rates of prey delivery to nests, greater time spent away from the nests by
adults and increased time spent by nestlings sleeping and resting. Concentrations of DDE, but
not PCBs, in nestling plasma samples from Lake Superior also regressed negatively on mean 5-
122
year productivity at the respective territories, indicating that DDE may still have been a factor
contributing to low productivity.
Low eagle productivity at certain areas of the British Columbia coast, such as Barkley
and Clayoquot Sounds, Johnstone Strait and South Moresby may also be caused by low food
availability. Mean plasma lipids were significantly lower in nestlings from those sites,
indicating chicks in poorer body condition. The significant association among sites between
productivity and mean percent plasma lipids also suggests that in productive areas, chicks are
fed more regularly, are in better body condition and are more likely to survive to fledging.
Breast muscle of eagle chicks found dead at inland nests near Lake Superior had higher mean
fat content than those found at shoreline nests (Kozie and Anderson 1991). The pilot study on
prey deliveries failed to show a significant difference between samples of nests in the Fraser
Delta and Barkley Sound, although there were significant differences between those sites in
both mean 3-productivity and percent plasma lipids in nestlings. However, because of logistical
difficulties in observing nests at more remote areas of the coast, where productivity is
particularly low, observations in Barkley Sound were made at nests which tended to be more
accessible and to have higher productivity.
Food supply during breeding is a major factor affecting avian productivity, including
raptors (Newton, 1980; Gardarsson and Einarsson 1994). In addition to Dykstra’s (1994) study
of eagles, Shutt (1994) related breeding failure and poor body condition of both herring gull
chicks and adults to lack of food at Lake Superior breeding colonies. Prey availability was
critical to productivity of white-tailed sea eagles (Helander 1985), European sparrowhawks
(Accipiter nisus) (Newton et al. 1986) and ospreys (Van Daele and Van Daele 1982). A
minimum food supply was required for successful breeding of wedge-tailed eagles (Aquila
audax) in Australia, while Hansen’s (1987) experiment showed that Bald Eagle nesting and
fledging success could be increased by providing additional food.
Bald Eagle breeding densities in Saskatchewan were related to availability of key prey
species, which correlated with primary productivity (Dzus and Gerrard 1993). Fish eating
birds, particularly gulls, are important prey species to north west eagles (Knight et al. 1990).
123
On the west coast of Vancouver Island, colony sizes and breeding success were lower for gulls
and cormorants (Vermeer et al. 1992) than the Strait of Georgia with its more stable food
regime (Vermeer et a!. 1989). The steep fjord-like topography of the shoreline and the islands
of the west coast of Vancouver Island, Johnstone Strait and Moresby Island also limits prey
availability and foraging opportunities, compared to the beaches and tidal mudflats of the Strait
of Georgia, which harbour abundant bird populations (Vermeer 1983). Food concentrated
along the highly productive La Perouse Bank, to the west of Barkley Sound is beyond the reach
of Bald Eagles. Langara Island is the only site outside the Georgia basin with relatively high
eagle productivity. This island lies at the bottom of the Alaska gyre, an area of summer
upwelling (Thomson 1981), which creates high marine productivity, evident by a rich fauna of
salmonids, seabirds and cetaceans.
Low eagle productivity in Barldey and Clayoquot Sounds and Johnstone Strait is
characterized by a high incidence of failed nesting attempts. Many nests had incubating adults
during the activity flight, but were empty during the productivity flight. Without nest
observations throughout the breeding cycle, we cannot determine at what stage those attempts
failed, although some nests certainly failed during incubation, as we often observed nests with
abandoned eggs during the later flight. A high incidence of nest failures, indicated by the
‘fledging ratio’ (young per successful nest/young per occupied nest) has been suggested as a
criteria for contaminant impact on an eagle population (Colborn 1991). The fledging ratio was
as high as 11 in bad years in Clayoquot Sound, where, at least PCDD/PCDF levels are lower.
High rates of nest failure in those areas is probably caused by the presence during nest
initiation in March and April of abundant food resources, such as Pacific Herring spawn
(Clupea harengus) (Hay et a!. 1992) and wintering waterbird prey (Vermeer and Morgan
1992), which are not available in May and June and are not adequately replaced by other food
items.
With the present data, it is difficult to determine why eagle productivity is low in the
Crofton area. In contrast to Clayoquot Sound and other areas, eagle nesting near Crofton
should not be food stressed. A number of the Crofton area nests are situated on small islands
124
(Shoal and Willy Islands), virtually in the estuary of the Chemainus River. Numerous
waterbirds, including flocks of several hundred White-winged Scoters (Melanitta fusca), feeding
on the abundant shellfish, are present during the breeding season. Eagle productivity is also
high in the area immediately to the north, where major habitat differences are not apparent.
It is conceivable that in the immediate past, PCDD and PCDF exposure at Crofton and
also possibly Powell River, Nanaimo and other pulp mill sites affected bald eagle reproduction.
Health effects in Great Blue Herons at Crofton were attributed to PCDD and PCDF exposure in
the late 1 980s (Elliott et al. 1989a; Sanderson et al. 1994a). Based on the extrapolation in
Figure 4.8, levels of 2,3,7,8-TCDD and other chemicals would likely have been even higher in
eagles than herons. PCDDs and PCDFs in eagle eggs collected in 1990 and 1991 and on a
lipid-adjusted basis in the one eagle plasma sample from Crofton were comparable to those
from Powell River, yet a reduction in mean productivity in the dioxin fishery closure area was
not found. However, for pragmatic reasons, fishery closures from persistent pollutants such as
dioxins must be defined over broad areas, even though there are wide gradients in
contamination within the zones (Harding and Pomeroy 1990). For example, higher PCDD and
PCDF concentrations were consistently found in invertebrates collected to the north than to the
south of the Powell River mill (Dwernychuck et al. 1994). This corresponds, perhaps
coincidentally, with poor productivity at the four eagle nests immediately north of that mill. At
Crofton, eagle productivity was also particularly poor at Shoal and Willy Islands, the nests
closest to the Crofton mill; those nests have often been active, but have rarely produced chicks.
Adult eagles, presumed to be from nests near the pulp mills, have been observed to forage in
the heron colonies at Crofton and Powell River (Norman et a!.; C. Burton, person. comm.),
which would cause very high PCDD exposure (Chapter 3).
However, by 1991 when the first eagle productivity surveys were done, PCDD and
PCDF concentrations in fish eating birds at Crofton had decreased by an order of magnitude
from the high levels of the late 1980s (Whitehead et a!. 1 992a; Figure 4.8). The rapid decline
of PCDDs/PCDFs in fish-eating birds was ascribed to their feeding primarily on small fish,
including many young-of-the-year age classes, in which reductions in local contaminant inputs
125
would be more quickly apparent. Sample sizes are small, nevertheless, mean PCDD/PCDF
levels in eagle eggs decreased between 1990 and 1992 at Crofton and Nanaimo, although
possibly at a slower rate than in herons and cormorants. As larger animals feeding at a higher
trophic level, clearence of TCDD and other compounds may occur more slowly in eagles.
300
4-.-c
ci)
4-.ci
-
150
UU
I— 100coI-.C)
C” 50
0
Figure 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-TCDD
ratio of eagles:herons, 1990-1992.
Assuming that poor productivity at Crofton is contaminant-related, it is also conceivable
that some adult eagles suffer chronic reproductive impairment due to past high PCDD/PCDF
exposure in ovo or during early growth and development. Rats and monkeys, of both sexes,
dosed with < 1 ug/kg of TCDD display abnormal reproductive function in laboratory studies
(Peterson et a!. 1993). For example, rhesus monkeys fed 25 ppt of TCDD, showed significant
extrapolated
Ii Heron
-*- Eagle
•Eagle
Cormorant
1987 1988 1989 1990 1991 1992 1993
126
reproductive impairment, but no apparent health problems (Bowman et al. 1989). Male rats
exposed both in utero and lactationally to as little as 0.064 ug/kg TCDD via maternal dosing
had damaged reproductive systems (Mably et al. 1992); however, fertility was not affected.
Mably et al. speculated that the high critical sperm volume of the rat would mitigate against
reduced fertility; other animals, for example man, which have a lower critical sperm volume
could be more affected. Although similar studies have not been done in birds, extrapolation
from the mammalian models implies that Bald Eagles hatched and raised in the Crofton area,
particularly during the period of highest PCDD/PCDF contamination, may also appear
externally normal, but have reduced capability to reproduce.
The potential for wildlife exposure to other chlorinated compounds of pulp mill origin
has received little attention. Although no samples were analyzed from the Crofton area,
waterfowl breast muscle tissues collected from 1990 to 1992 near various pulp mills on the
British Columbia coast, including Nanaimo and Powell River, contained from 0.5 to 5 ag/kg
pentachiorophenol and traces (<1.0-3.3 tg/kg) of di- and tetrachloroquaiacols (Canadian
Wildlife Service 1994). Those compounds are considered indicative of bleached-kraft pulp mill
contamination of receiving water, sediments and biota (Dwernychuck et al. 1994). Release of
organochlorines (AOX) in pulp mill effluents has decreased significantly since the installation of
secondary treatment systems at all British Columbia coastal pulp mills (see Table 3.9). Studies
of fish collected from both bleached-kraft and non-kraft pulp mills in eastern Canada have also
reported the presence of an unidentified factor(s) present in effluents of both mill types that
induce CYP 1A and affect reproductive hormone levels (Carey et al. 1992). Presence of that
factor was independent of either chlorine bleaching or secondary treatment. However, both
chiorophenols and chloroquaiacols and the unidentified factor appear to be cleared fairly rapidly
in fish, ie. within two weeks; therefore, it seems unlikely that Bald Eagles would accumulate
significant amounts of this class of chemicals.
Alternatively, the low productivity measured in nests adjacent to the dioxin closure area
at Crofton may be explained as either a sampling artifact or the result of ecological factors that
have not been identified. Because of the cost of helicopter surveys and the difficulty in locating
127
nests, the sample may not be representative of the area, implying that some productive nests
were not surveyed each year. However, the probability of overlooking a significant number of
productive versus unsuccessful nests in the Crofton area should be no different than in other
areas. Although the Crofton area is surveyed at the end of the flight, after only 1.5 hours,
observer fatigue should not be a factor. Because of the history of contamination, the Crofton
area likely receives greater attention. Quality of nesting habitat near Crofton appears
comparable or better than most areas of the survey route; there are large numbers of suitable
nest trees in relatively undisturbed areas and only limited activity.
Currently, I am unable to determine the cause for poor eagle productivity at nests
adjacent to the dioxin fishery closure area at Crofton. It is probably not caused by low food
supply. It may be caused by other ecological factors which we have failed to identify;
however, the effect of contaminants whether from past or ongoing exposure cannot be ruled
out. Further intensive work in this area is necessary to confirm the results and investigate
causes.
My conclusions agree with those of Dykstra (1994) that the role of food supply needs to
be factored into any studies of the effects of contaminants or other habitat quality variable in
studies of Bald Eagles. Measurement of plasma lipids may provide a useful surrogate for
energetic status of eagle nestlings. Further work is required to determine the causes of the
apparent low productivity in the Crofton area.
Acknowledgments
A special thanks to Ian Moul and George Compton for all of their support and
assistance in the field. Chris Coker and Brenda Li-Pak-Tong are thanked for their field work on
the prey deliveries. Ron McLaughlin (MacMillan-Bloedel) and Ken Stenerson (Scott Paper) are
also thanked for personal and corporate financial support with helicopter surveys. Working in
the laboratory of Dr. Ross Norstrom, Mary Simon did the PCDD/PCDF and non-ortho PCB
analysis; Henry Won did the organochiorine and plasma lipid analyses.
128
App
endi
x4-
1.Pr
oduc
tivity
,%
lipid
and
sele
cted
chlo
rina
ted
hydr
ocar
bon
resi
due
leve
lsin
plas
ma
ofin
divi
dual
Bal
dE
agle
chic
ksco
llect
edfr
omth
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ast
ofB
ritis
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olum
bia,
1993
-94
Res
idue
leve
ls(w
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sis)
Loc
atio
n,N
est
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duct
.1
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378-
1236
78-
OC
DD
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478-
PCB
PCB
PCB
PCB
PCB
DD
ET
otal
(chi
cks/
%W
HO
TC
DD
PnC
DD
HxC
DD
TC
DF
PnC
DF
-77
-126
-118
-105
-153
PCB
sac
tive
terr
.)(n
g/kg
)(w
etw
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t)(p
g/kg
)(w
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Fra
ser
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elta
Ala
ksen
10.
106
2.39
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ND
ND
26.8
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2.9
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210.
660.
570.
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880.
225
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2.83
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GENERAL SUMMARY AN]) CONCLUSIONS
The overall purpose of this research was to investigate the toxic hazard posed by
chlorinated hydrocarbon contaminants to Bald Eagle populations breeding and wintering in the
Strait of Georgia area of British Columbia. The research tested a general hypothesis that as
top predators in marine and estuarine systems, Bald Eagles would bioaccumulate high levels of
chlorinated hydrocarbons. Consequent to high exposure and as ensuing hypotheses, both
survival and reproduction would be adversely affected. These hypotheses were tested by a
number of field and laboratory studies.
Adult exposure and mortality study
The investigation began by collecting samples from the large number of Bald Eagles
found dead or dying each year in British Columbia. Many sick birds and carcasses are turned
in by concerned members of the public or individuals seeking taxidermy permits. Of 484
eagles examined in this study, 59 found between 1988 and 1993 were selected for
organochlorine analysis. Of those birds 5% had liver residue levels of DDE and chlordane
related compounds diagnostic of acute toxicity. Even this percentage is surprising and the long
term persistence of OC pesticides and continued input from atmospheric sources and migratory
birds is indicated. These findings reinforce the need for vigilance in both the enforcement of
current regulations and scrutiny of new commercial chemicals.
Of 19 Bald Eagles further analyzed for PCDDs, PCDFs and non-ortho PCBs, livers of
four birds (21 %) contained TCDD-toxic equivalents (TEQsWHO) > 1 ,000 ng/kg. Birds with
high PCDD and PCDF levels were found in the vicinity of bleached-kraft pulp mills. Most
bird with elevated chlorinated hydrocarbon levels were in poor body condition indicating lipid
and contaminant mobilization. Based on high TCDD/TCDF ratios in at least three eagles,
hepatic CYP1A enzymes were likely induced
131
Study of biological effects in eagle chicks
In order to assess embryotoxic effects of chlorinated hydrocarbons in Bald Eagles, eggs
were collected within an exposure gradient and incubated in the laboratory. Yolk sacs of
chicks collected near bleached-kraft pulp mills contained higher concentrations of PCDDs and
PCDFs, although there were no significant effects on hatching success or morphological
endpoints. Hepatic CYP1A levels were induced in chicks from pulp mill sites and correlated
significantly with 2,3,7,8-TCDD, 2,3,7,8-TCDF and TEQ5WHO in yolk sacs. TEQsWHO
associated with CYP1A induction and converted to a whole egg wet weigh basis, 210 ng/kg,
were suggested as a LOEL for the Bald Eagle; TEQsWHO associated with background CYP1A
levels were suggested as a NOEL for the Bald Eagle, 100 ng/kg.
These findings suggest that the Bald Eagle embryo is perhaps an order of magnitude less
sensitive to TCDD-like toxicity than the chicken embryo. The LD50 for the chicken embryo is
about 250 ng/kg (Alired and Strange 1977; Janz 1995), similar to the 210 ng/kg TEQ5WHO
measured in eagle eggs without apparent effects on hatching success or histological,
morphological and some biochemical endpoints. At 100 ng/kg TEQ5WHO in eagles, no
significant CYP1A induction occurred, while two-fold AHH induction was measured at 10
ng/kg injected into chicken eggs. With regard to CYP1A induction, Bald Eagles appear
somewhat more sensitive than Great Blue Herons and Double-crested Corinorants. In heron
chicks, EROD activity was significantly induced (six-fold) at about 440 ng/kg, but not at 250
ng/kg TEQsWHO (Sanderson et at. 1992a). In cormorant chicks, significant eight-fold EROD
induction occurred at 550 ng/kg but not at 217 ng/kg TEQsWHO (Sanderson et at. 1992b).
Bioaccumulation study
For this study, fresh Bald Eagle eggs were collected at a variety of locations on the
British Columbia coast, representing different chlorinated hydrocarbon exposure scenarios. A
data base of contaminant levels in Bald Eagle prey items, principally from pulp mill sites in the
Strait of Georgia, was compiled using existing data. A simple model was used to examine the
relationships between contaminant levels in Bald Eagles and their foodchain. The model
accurately predicted 2,3,7, 8-TCDD levels in Bald Eagle eggs and was reasonably accurate for
132
other compounds. The model was used to estimate 2,3,7,8-TCDD and TEQWHO levels in
forage fish and sentinel fish-eating bird species (herons, cormorants, grebes, mergansers),
which would be protective of Bald Eagles consuming an average diet. The NOELs and LOELs
generated in the above embryotoxicity study were used as critical values in eagle eggs.
Concentrations of 0.5 ng/kg in forage fish and 10 ng/kg in fish-eating birds were suggested as
site specific guidelines for the Strait of Georgia. The same approach was used to derive similar
values for total PCBs, suggested to be 0.01 ng/kg in forage fish and 0.3 ng/kg in fish-eating
birds.
Productivity study
The research described in the previous studies addressed acute toxicity of adult birds
and determination of critical levels in eggs, associated with embryotoxicity. During the fourth
part of this work, Bald Eagle breeding success was measured for up to three years at eight sites
on the British Columbia coast. Because of annual variability, assessment of breeding success in
Bald Eagles requires a minimum of three years data. Studies elsewhere showed that
reproduction in birds of prey is a critical endpoint affected by chlorinated hydrocarbons in birds
of prey (Newton 1979). In order, to relate productivity of individual nests to contaminant
exposure, blood samples were taken from nestlings, to minimize the impact of sample
collection.
Bald Eagle productivity was highest overall at nests in the lower Fraser River valley and
delta, while at four of five reference areas, selected for their remoteness from direct industrial
input of pollutants, productivity was less than the level of 0.75 young/occupied nest considered
necessary to sustain an eagle population. Only at Langara Island, an area of very high
biological productivity, was eagle breeding success comparable to the Fraser valley and most
Strait of Georgia sites. At the reference locations, low breeding success is likely due to low
food availability, particularly during chick rearing. This was supported by finding of
significantly lower nestling plasma lipid content at those sites and a significant positive
regression between mean nestling plasma lipid levels and mean productivity among sites.
Despite higher plasma levels of PCDDs and PCDFs, Bald Eagle productivity was relatively
133
high at nests near two pulp mill areas on the Strait of Georgia (Nanaimo, Powell River); at
those sites, no significant differences in mean productivity occurred at nests adjacent to
PCDD/PCDF fishery closure areas compared to nests outside of the closure area. However,
productivity was significantly lower at nests inside the fishery closure area at one site, Crofton,
than outside the dioxin closure.
Low breeding success around Crofton likely is not due to low food availability; the area
is rich in marine life. Data from biomonitoring studies of fish-eating birds showed that PCDD
and PCDF levels in local food chains fell dramatically between 1989 and 1992, subsequent to
modifications to the bleaching process employed by the mill and a ban on chlorophenolic anti
sapstain usage. Alternative hypotheses to explain the low eagle productivity in the area
include: first, the presence of a substance released in the mill effluents, that has contaminated
local food chains and is either embryotoxic or capable of affecting parental breeding behaviour.
Second, some eagle pairs may be reproductively impaired as a result of past exposure in ovo or
during early development of the reproductive system, to elevated levels of 2,3,7, 8-TCDD and
related chemicals. This last hypothesis requires further study and testing.
In conclusion, during the recent past reproduction of Bald Eagles in the Strait of
Georgia was probably reduced by exposure to significant chlorinated hydrocarbon levels,
particularly DDE. Increases in nest occupancy reported for the southern Gulf Island between
the early 1 970s and late 1 980s is typical of the population recoveries documented in many areas
of North America and attributed to declining environmental DDE contamination. During the
1 980s and at least until the early 1 990s, eagles breeding and wintering near bleached-kraft pulp
mills on the British Columbia coast were exposed to relatively high levels of PCDDs and
PCDFs. At Crofton, the effects of this pollution may be continuing, although the mechanism is
obscure. At other areas of the British Columbia coast, Bald Eagle breeding success appears to
be influenced mainly by food supply.
The effects of chlorinated hydrocarbons on Bald Eagle populations have to be
considered in the context of multiple stresses, both chemical and otherwise, on survival and
reproduction. Lead poisoning from ingestion of spent shot is a major cause of death for British
134
Columbia Bald Eagles; many eagles have also been sublethally poisoned, with probable
consequences for longterm health and survival. In some areas, such as the Lower Fraser
Valley, pesticide poisoning is a major cause of mortality. Bald eagles are also vulnerable to
loss and disturbance of nest sites. Given these factors, and the growing human population of
the Georgia Basin, maintenance of a healthy eagle population will require ongoing vigilance.
Finally, although the Bald Eagle has some merits as a sentinel species of pollutant exposure and
effects, it may be more cost-effective to monitor colonial fish-eating birds.
Future Directions
Ecotoxicological work on Bald Eagles should further investigate the low reproductive
rate measured at Crofton. All nests in the area from Cowichan Bay to Thetis Island should be
located. A sample of nests including those nearest the mill, should be intensively observed to
determine breeding behaviour and the timing of nest failures. Toxicological hypotheses can be
tested by trapping adult eagles on their breeding territories to obtain blood samples for
contaminant analyses and measurement of reproductive and thyroid hormones. Similar studies
are required at a reference site, such as Barkley Sound, and also possibly at another pulp mill
site, either Nanaimo or Powell River, depending on available funding.
Laboratory research using in vitro cell cultures of primary hepatocytes from eagles or
other raptors would provide data on sensitivity of raptors compared to more commonly studied
laboratory species and sentinel species such as Herring Gulls. Alternatively, in vivo
comparative toxicology with American Kestrels, particularly of TCDD effects on reproductive
endpoints would be valuable.
Further ecological work on the role of food supply in breeding success of British
Columbia eagle populations should also be undertaken. A long term monitoring study of Bald
Eagle reproduction at one or more of the remote sites could provide valuable information on
fluctuations in coastal productivity, and the influence of large scale processes such as global
warming.
135
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