Environmental Health Perspectives Volume
103, Supplement 4, May 1995
[Citation
in PubMed] [Related
Articles]
A Review of Factors Affecting Productivity of Bald Eagles in the Great
Lakes Region: Implications for Recovery
William W. Bowerman,1 John P. Giesy,1 David
A. Best,2 and Vincent J. Kramer1
1Department of Fisheries and Wildlife, Pesticide Research
Center, and Institute for Environmental Toxicology, Michigan State University,
East Lansing, Michigan; 2USDI-Fish and Wildlife Service, East
Lansing, Michigan
Abstract
The bald eagle (Haliaeetus leucocephalus) population in North
America declined greatly after World War II due primarily to the eggshell
thinning effects of p,p'-DDE, a biodegradation product of DDT. After
the banning of DDT in the United States and Canada during the early 1970s,
the bald eagle population started to increase. However, this population
recovery has not been uniform. Eagles nesting along the shorelines of the
North American Great Lakes and rivers open to spawning runs of anadromous
fishes from the Great Lakes still exhibit impaired reproduction. We have
explored both ecological and toxicological factors that would limit reproduction
of bald eagles in the Great Lakes region. Based on our studies, the most
critical factors influencing eagle populations are concentrations of environmental
toxicants. While there might be some continuing effects of DDE, total PCBs
and most importantly 2,3,7,8-tetrachlordibenzo-p-dioxin equivalents
(TCDD-EQ) in fishes from the Great Lakes and rivers open to spawning runs
of anadromous fishes from the Great Lakes currently represent a significant
hazard to bald eagles living along these shorelines or near these rivers
and are most likely related to the impaired reproduction in bald eagles
living there. -- Environ Health Perspect 103(Suppl 4):00-00 (1995)
Key words: bald eagle, DDT/DDE, Great Lakes, Haliaeetus leucocephalus,
reproduction, PCBs, TCDD-EQ
This paper was presented at the Conference on Environmentally
Induced Alterations in Development: A Focus on Wildlife held 10-12 December
1993 in Racine, Wisconsin.
This paper is a summary of a number of studies that were
funded by Consumers Power Company, U.S. Forest Service, U.S. Fish and Wildlife
Service, U.S. National Park Service, Michigan, Ohio, and Wisconsin Departments
of Natural Resources, Ontario Ministry of Natural Resources, Canadian Wildlife
Service, EARTHWATCH, and the Great Lakes Protection Fund. We thank E. Addison,
A. Bath, J. Bruce, K. Doran, R. Eckstein, J. Edde, R. Ennis, E. Evans,
L. Grim, T. Grubb, J. Hammill, J. Hendricksen, J. Holt, P. Hunter, L. Kallemeyn,
T. Kubiak, E. Lindquist, J. Mathisen, M. Meyer, J. Papp, S. Postupalsky,
J. Robinson, M. Shieldcastle, C. Sindelar, J. Weinrich, T. Weise, C. Weseloh,
L. Williams, and the many agency employees, student interns, and volunteers
who have assisted in this research. We thank E. Evans, M. Gilbertson, J.
Newsted, R. Rolland, M. Shieldcastle, and an anonymous reviewer for comments
on previous drafts of this manuscript.
Address all correspondence to Dr. W. W. Bowerman, Eagle
Environmental, Inc., 6154 Columbia Street, Haslett, MI 48840. Telephone/Fax
(517) 339-0105.
Introduction
The numbers of bald eagles (Haliaeetus leucocephalus) in North
America declined greatly after World War II due primarily to the eggshell
thinning effects of p,p´-DDE, a biodegradation product of DDT
(1-3). After the banning of DDT in the United States and Canada during
the early 1970s, the numbers of bald eagles increased (3) (Figure
1). This population recovery has not been uniform however. Eagles nesting
along the shorelines of the North American Great Lakes and rivers open to
spawning runs of anadromous fishes from the Great Lakes still exhibit impaired
reproduction (4) (Figure 2). It is also apparent that adult mortality
in eagles nesting near the Great Lakes is greater than expected. To understand
the implications of impaired reproduction and greater mortality of adults
on the recovery of this species within the Great Lakes region, we will discuss
factors related to bald eagle population dynamics.
Figure 1. Numbers
of occupied breeding areas of bald eagles nesting within the Great Lakes
region for the period 1977 to 1993.
Figure 2. Productivity
of bald eagles within the Great Lakes region for the period 1977 to 1993.
Study Area
Our studies have focused on 10 subpopulations of bald eagles within the
Great Lakes region (Figure 3). These areas were defined as: the area within
8.0 km of the United States' and Canadian shorelines of the Great Lakes
and rivers open to Great Lakes fish runs, hereafter referred to as anadromous
accessible rivers, along a) Lake Superior, b) Lake Michigan,
c) Lake Huron, and d) Lake Erie; areas in Michigan greater
than 8.0 km from the shorelines of the Great Lakes and not along anadromous
accessible rivers in e) the lower peninsula, f) the eastern
upper peninsula east of U.S. Highway 41, and g) the western upper
peninsula west of U.S. Highway 41; and h) the Chippewa National Forest,
i) the Superior National Forest outside of the Boundary Waters Canoe
Area, and j) Voyageurs National Park in Minnesota (Figure 3).
Figure 3. Ten
subpopulations used for comparison of PCB and p,p´-DDE concentrations
in plasma of nestling bald eagles in the Midwest. Subpopulations were: within
8.0 km of Lakes a) Superior, b) Michigan, c) Huron,
and d) Erie; interior areas within e) the northern lower,
f) eastern upper, and g) western upper peninsulas of Michigan;
and h) the Chippewa and i) Superior National Forests excluding
the Boundary Waters Canoe Area, and j) Voyageurs National Park, Minnesota.
Only data from survey areas that complied with accepted reproductive
survey techniques were utilized. Since only breeding area occupancy and
not breeding area productivity in the Boundary Waters Canoe Area was known,
we excluded this area from the Superior National Forest subpopulation. Due
to the use of nonstandard classification that resulted in an overestimation
of bald eagle productivity relative to other regions in the Great Lakes,
data on reproduction in interior Wisconsin were not included in the analyses.
Also, nests found within the past 4 years on the Ontario shorelines of Lakes
Superior and Huron were not included in the analysis since only productive
nests were accurately reported, and the actual number of unsuccessful nests
were not known.
Factors Affecting Eagle Populations
Although many potential factors could affect bald eagle reproductive
success or productivity, the three primary factors currently influencing
bald eagle productivity in the Great Lakes region are habitat availability,
degree of human disturbance to the nesting eagles, and the concentrations
of environmental contaminants in the prey of the nesting eagles. Territories
unoccupied by another breeding pair must include sufficient nest, perch,
and roost trees, foraging areas, and a prey base in sufficient quantities
to successfully raise young to fledging (5). Human disturbances must
be of a type, degree, amount, and timing not to cause nest abandonment by
an individual breeding pair of eagles (5). Environmental contaminants,
primarily chlorinated hydrocarbons such as p,p´-DDE and PCBs,
must be below concentrations associated with impaired productivity, egg
lethality, or teratogenicity to produce a viable population (6-8).
Currently, we feel that the most important factor controlling bald eagle
reproduction along the shorelines of the Great Lakes where eagles currently
nest is the influence of environmental contaminants. We have shown that
concentrations of p,p´-DDE and PCBs in both abandoned eggs
and plasma of nestling eagles are correlated with impaired reproductive
potential of eagles along the shorelines of Lakes Superior, Michigan, Huron,
and Erie, as well as at Voyageurs National Park (4) (Figures 4,5).
Furthermore, current concentrations of both PCBs and p,p´-DDE
in eggs of bald eagles are sufficiently great, based on controlled laboratory
studies, to cause adverse effects in birds (9). While eggshell thinning
due to p,p´-DDE may still be influencing eagle reproduction
in the Great Lakes region, we have shown that the negative correlation between
productivity and PCBs in bald eagle eggs has become stronger and more statistically
significant than the relationship between productivity and p,p´-DDE
(9). The occurrence of teratogenic effects in nestlings, which are
similar to those that are known to be caused by dioxinlike coplanar compounds
including polychlorinated dibenzofurans (PCDFs), polychlorinated dibenzodioxins
(PCDDs), and some PCB congeners, indicates that these compounds are the
most likely causative agents (10). These effects also occur in other
avian species exposed to relatively great concentrations of TCDD-EQ in controlled
laboratory studies (10). We have shown that concentrations of mercury
(Hg) are not correlated with bald eagle productivity (11) and are
less than the no observable adverse effect concentration (NOAEC) predicted
from controlled laboratory studies with birds.
Figure 4. Relationship
between overall productivity, 1977 to 1993, and geometric mean concentrations
of p,p´-DDE (ng/g ww) in plasma of 10 subpopulations of nestling
bald eagles in the upper Midwest.
Figure 5. Relationship
between overall productivity, 1977 to 1993, and geometric mean concentrations
of total PCBs (ng/g ww) in plasma of nine subpopulations of nestling bald
eagles in the upper Midwest.
Availability of physical habitat does not seem to be limiting expansion
of the bald eagle population along the upper Great Lakes shorelines. While
bald eagles are restricted from some areas due to human disturbance or physical
structure of the habitat, there are still areas deemed to be suitable nesting
habitat, which are currently unoccupied by bald eagles. This is especially
true of the northern forested regions that are less populated by humans.
Habitat along Lake Erie is relatively scarce and may become a limiting factor
in the near future as the populations become reestablished. The management
strategy of the Ohio Department of Natural Resources and the Ontario Ministry
of Natural Resources includes control of potential human disturbance near
nests early in the nesting period along Lake Erie and transformation of
otherwise marginal habitat into good habitat.
Ecological Hazard Assessment
To numerically determine if environmental contaminants in fish prey of
the bald eagle could be influencing productivity, we conducted an ecological
hazard assessment that examined concentrations of organochlorine pesticides,
PCBs, TCDD-EQ, and Hg in fishes from areas above and below barrier dams
along three Great Lakes tributaries--the Au Sable, Manistee, and Muskegon
Rivers in Michigan (Figure 6). A hazard index (HI) of individual organochlorine
pesticides, PCBs, TCDD-EQ, and Hg to the bald eagle or surrogate species
[e.g., wood duck (Aix sponsa) and herring gull (Larus argentatus)]
was determined by a toxic units approach. One toxic unit was defined as
the quotient of the concentration in the diet divided by the dietary NOAEC
(Equation 1):
[1]
When the HI for an adverse effect was greater than one (1 toxic
unit), the concentration in the diet was expected to be sufficiently great
to equal the threshold concentrations to elicit a statistically significant
response. The lowest observable adverse effect concentration (LOAEC) is
approximately 10-fold greater than the NOAEC. One would not expect to see
population-level effects at an HI of 1.0, but, depending on the slope of
the dose response relationship, values of 10 to 20 are related to population-level
effects. We used a weighted average exposure to each chemical of interest,
based on the relative proportions of each species of fish in the eagles'
diets. The relative proportions of each species of fish in the diet were
determined from visual observations of the prey taken by eagles and from
an analysis of the prey remains in or around the nests of eagles in the
various areas.
Figure 6. Map
of Michigan showing the locations of the Au Sable, Manistee, and Muskegon
Rivers.
Dose-response relationships for different end points in the bald eagle
were used when available (Table 1). However, since this is a threatened
or endangered species in many areas, it is difficult if not impossible to
conduct controlled laboratory experiments or make field collections. Thus,
it is often necessary to use the results of studies with surrogate species.
We have tried to choose results from species that were similar to bald eagles
or that are known to have similar sensitivities to compounds for which there
is information for bald eagles. To verify the hazard assessments, they were
reconciled with the current distributions of bald eagles and their exposure
to the various toxicants. We have not applied any uncertainty factors in
the hazard assessment.
Biomagnification factors (BMF) were used to estimate the magnification
of toxicants from fishes to bald eagle eggs. Where possible, we calculated
BMFs from measurements of the toxicants in fishes and bald eagle eggs in
a region. However, since it was not always possible to obtain empirical
values, we also used BMF values from the literature (9-12) (Table
1). Since there were not enough samples to test for significant differences
in concentrations among species within or among rivers, predicted concentrations
of toxicants in bald eagle eggs were calculated from mean concentrations
representative of the concentrations observed in the fish populations both
below and above dams that drain into the Great Lakes. Two BMFs were calculated,
one for the inland population (more than 8 km from the Great Lakes shoreline)
and one for bald eagles living along the shoreline. We selected an average
BMF that allowed us to use a single threshold level to determine the toxic
units in fish. We then calculated NOAEC of individual organochlorine pesticides,
PCBs, Hg, and TCDD-EQ in fishes based on the relative dietary intake of
each species of fish eaten by bald eagles (Table 1).
Results of Hazard Assessment
Dieldrin. Whereas dieldrin is known to be toxic to birds
and is suspected of having caused population-level effects (13,14),
it is not likely that current concentrations of dieldrin in fishes of the
Great Lakes present a significant hazard to bald eagles. The NOAEC used
in the hazard assessment is conservative and based on the regression of
Wiemeyer et al. (6), who relates dieldrin concentrations in eggs
to productivities of individual pairs of bald eagles (Table 1). However,
the authors of that study state that "while dieldrin concentrations
greater than 1.0 µg/g fresh ww in eggs cannot be ruled out as having
an effect on reproduction, the major effect of dieldrin was related to adult
survival" (6). The apparent association of dieldrin concentrations
with productivity of bald eagles is most likely an artifact due to cocorrelation
of the concentrations of dieldrin with those of total PCBs and the DDT complex
(6). When considered with the dose-response
relationships obtained in controlled laboratory studies of other avian species
(15), we conclude that the correlation is most likely spurious and
not indicative of actual toxicity of dieldrin at the concentrations observed.
Thus, it was concluded that the current concentrations of dieldrin in fishes
above the dams are well below the concentration that would be expected to
cause any adverse effects. Dieldrin concentrations below the dams are slightly
greater than the NOAEC, but are probably not sufficiently elevated to be
causing any population-level adverse effects (Figure 7). Dieldrin is probably
not currently considered to be the critical contaminant limiting the distribution
or productivity of bald eagles in Michigan.
Figure 7. Comparison
between hazard indices (HI), corrected for relative proportion of fish in
the diet, for fish from above and below barrier dams along three Great Lakes
tributaries to Lakes Michigan and Huron for critical contaminants related
to depressed reproduction in bald eagles.
DDE. The effects of p,p´-DDE on bald eagle
reproduction have been correlated in field studies (6). However,
the concentrations of DDE and total PCBs were significantly cocorrelated,
and separation of the effects of DDE from co-occurring toxicants such as
PCBs is difficult (3,6). In laboratory studies, DDE has been linked
to eggshell thinning in several species of birds (16-23). Therefore,
we have compared the NOAEC predicted from a regression analysis of the effects
of DDE on eagles in the field (6) with values from similar species
and to the results of controlled laboratory studies where cocorrelation
did not confound the analysis.
As concentrations of DDE in the environment have declined, populations
of bald eagles have increased (1,2). Although populations of bald
eagles seem to be doing well at the interior locations in Michigan where
they do not eat fish from the Great Lakes or anadromous accessible rivers,
there is some reason for concern; it has been projected that the concentrations
of DDE need to decrease to almost zero before there is no predicted adverse
effect (3). Current concentrations of DDE in several populations
of raptors may still be limiting reproductive success (24).
Currently, concentrations of DDE in bald eagle eggs collected from two
of four breeding areas along the shores of Lakes Michigan and Huron exceed
the 15.0 µg/g of p,p´-DDE associated with a 75% reduction
in productivity (6). However, the currently observed concentrations
of DDE in prey from these areas are not greatly in excess of the NOAEC.
This supports the results of the hazard assessment, which indicates there
should still be some effects of DDE on the reproduction of bald eagles due
to eggshell thinning (Figure 7).
Total PCBs. Concentrations of PCBs in the food and eggs
of birds of the Great Lakes region have been suggested as a major causative
agent for the observed adverse effects on productivity of fish-eating birds
(25). Total concentrations of PCBs in the eggs of bald eagles have
been inversely correlated with productivity (4,6,26-31). PCBs have
also been identified as a major cause of birth defects in the white-tailed
sea eagle (Haliaeetus albicilla) in Europe (32). It has been
difficult to demonstrate a cause-effect relationship between concentrations
of PCBs in bald eagle eggs and impairment of productivity because the concentrations
of PCBs are always cocorrelated with the concentrations of other organochlorine
toxicants, such as the DDT complex (3,6,28). However, as the concentrations
of DDE in bald eagle eggs have declined, egg mortality due to eggshell thinning
has decreased. Current concentrations of DDE are less than that thought
to be necessary to cause a critical degree of eggshell thinning. As the
concentrations of DDE have decreased, the negative correlation between productivity
and concentrations of DDE has become poorer (r2=0.63),
but the negative correlation between productivity and concentrations of
PCBs in bald eagle eggs in the Great Lakes region has become stronger and
more statistically significant (r2=0.80) (9). When
the effects of DDE (primarily on eggshell thinning) are removed statistically,
there is still a significant inverse relationship between the concentrations
of other chemicals (primarily total PCBs) and productivity of bald eagles
(3,6,9). These other chemicals are thought to be responsible for
most of the currently observed adverse effects.
The threshold egg concentration of PCBs to maintain healthy bald eagle
productivity (>1.0 young per occupied nest), based on analysis of samples
from Michigan and Ohio, has been estimated to be approximately 6.0 mg PCB/kg
ww (ppm) (4). This value is similar to the NOAEC of 4.0 mg PCB/kg,
that has been suggested based on regression analysis of samples throughout
continental North America (6). Determination of the critical concentration
for effects in eggs from regression analyses in field studies is limited
by the statistical influence of cocorrelation with other compounds and the
slope of the dose-response relationship (33,34). Ideally, these values
for NOAEC estimated from regression analysis should be compared to the results
from studies under more controlled conditions. There are no controlled studies
of PCBs with bald eagles and few with other raptors (25). Laboratory
studies with a number of species of birds have demonstrated that PCB exposure
can result in effects on the survival of bird embryos (6,35) and
result in population-level effects (25). The NOAEC used in our assessment
was similar to the concentration for threshold effects in chicken eggs (36).
Chronic exposure to 5 mg/kg of Aroclor 1254 in the diet had no effect on
the productivity of chickens (37). In fact, concentrations of Aroclor
1254 as great as 40 mg/kg, in the diet of white leghorn chickens did not
affect productivity (38). Deformities were observed in white leghorn
chickens when the concentration of Aroclor 1254 reached 10 mg/kg, ww, in
the yolk (39). Therefore, the concentration of 4.0 mg PCB/kg, ww,
in eggs is a reasonable estimate of the concentration that causes effects
in bald eagle eggs. The NOAEC (4.0 mg/kg, ww; Table 1) used in our hazard
assessment was derived from the regression given by Wiemeyer et al. (6,31)
(Table 1). The value that we have selected for our hazard assessment is
the same as that used by Kubiak and Best (9) but is 10-fold greater
than the value of 0.4 mg PCB/kg, ww, in bald eagle eggs, as suggested by
Ludwig et al. (40). We have based our hazard assessment on reproductive
effects, but it should be remembered that survival of the adults is also
an important parameter in determining the success of bald eagle populations
(41). It is possible that toxicants such as PCBs may affect adults
in subtle ways at concentrations less than those required to affect egg
survival.
The results of the hazard assessment indicate that current concentrations
of total PCBs in fishes of the three rivers (Figure 6) upstream of the barrier
dams should not be having adverse effects on bald eagles but that anadromous
fishes below the lowest dams would present a significant hazard to bald
eagles living near rivers below the dams and along the Great Lakes shoreline
(Figure 7). This hazard assessment predicts the observed productivities
of bald eagles in the two areas. Our field monitoring indicates that the
productivity of bald eagles upstream of the dams was greater than 1.0 young
per occupied nest and indicated a healthy bald eagle population. Bald eagles
along the shorelines of the Great Lakes or along anadromous-accessible rivers
had productivities of approximately 0.67 young per occupied nest, which
is less than the 1.0 necessary for a healthy population and the 0.7 required
for a stable population (4,9,25-27). Total concentrations of weathered
PCBs in addled bald eagle eggs of 83 and 98 mg PCB/kg, ww, have been measured
for Lakes Michigan and Huron, respectively (9). These concentrations
are approximately 20 times greater than the NOAEC used in our hazard assessment
and indicate that exposure of these populations to total PCBs is causing
adverse, population-level effects.
TCDD-EQ. The polychlorinated diaromatic hydrocarbons that
can attain a planar configuration and cause effects similar to those of
2,3,7,8-TCDD have been demonstrated through both laboratory and field studies
(8,24,42) to be the current critical factors that could cause the
effects observed in most wildlife populations, especially the deformities
of bald eaglets (9). TCDD-EQ can be contributed by a number of compounds,
including the PCDDs, PCDFs, and planar PCBs (3). In the Great Lakes, PCB
congeners contribute a great proportion of the TCDD-EQ and may be responsible
for the observed toxicity. For these reasons, we conducted a hazard assessment
of the potential effects of TCDD-EQ on the bald eagle populations living
along the three rivers (Figure 6).
Because bald eagles are a threatened or endangered species, there have
been no controlled laboratory studies of the effects of TCDD-EQ on bald
eagles. Similarly, there have been few field studies that have correlated
concentrations of TCDD-EQ in the diet or eggs of bald eagles with observed
effects. We have therefore derived a range of values for the LOAEC and NOAEC
based on the studies of the effects of TCDD and dioxinlike compounds on
surrogate species to calculate a HI. Published LOAEC values were in the
range of 10 ng TCDD-EQ/kg, ww, in avian eggs (25). The LC50
(concentration to be lethal to 50% of the eggs exposed) for the toxicity
of PCB congener #126 as determined by egg injection studies of the eggs
of the American kestrel (Falco sparverius) is between 40 and 70 µg/kg,
ww (43). The relative toxicity of PCB #126 to that of 2,3,7,8-TCDD
is approximately 0.015 for avian species (44,45). Application of
this factor to the LC50 for PCB #126 in American kestrels results
in a LC50 of between 0.6 and 1.0 µg TCDD-EQ/kg, ww, in
the egg. The ratio between the LOAEC value and LC50 of TCDD in
white leghorn chicken eggs is approximately 100 (46). Application
of this ratio to the LC50 for lethality of American kestrel eggs
results in an LOAEC value for TCDD of between 6 and 10 ng/kg, ww, in egg.
The LOAEC value for TCDD, based on lethality has been reported to be 10
pg/g, ww, for the chicken embryo (47). If this value is divided by
a 10-fold application (safety) factor to extrapolate from the LOAEC to the
NOAEC for lethality, a value of 1 ng 2,3,7,8-TCDD/kg in the egg is derived.
This concentration injected into chicken eggs resulted in 6 to 15% deformities.
The LC50 for wood ducks has been reported to be approximately
70 ng TCDD-EQ/kg, ww, in their eggs (48). If this value is divided
by an application factor of 10, the estimated NOAEC for eggs is approximately
7 ng TCDD-EQ/kg. Alternatively, based on an LOAEC value of 21 pg TCDD/g
for the effects of TCDD-EQ on wood ducks under field conditions (48),
an NOAEC value of 2.1 pg/g TCDD-EQ/g in the egg can be estimated by using
the standard 10 times application factor. Based on the above information,
we chose a value of 7 ng/kg, ww, in egg as the LOAEC/NOAEC to be used in
the hazard assessment. While on the conservative side, the value selected
for the NOAEC is near the median for the NOAEC values calculated from the
literature on the toxicity of TCDD to avian species.
The NOAEC value used in our HI is similar, but not identical, to those
suggested by other workers. Our value is approximately 16-fold less than
that derived by the U.S. EPA in their guidance document for hazard assessments
of the effects of 2,3,7,8-TCDD on wildlife, but is in the range of values
predicted from studies of other species (42). The U.S. EPA determined
that a concentration of 6 ng TCDD/kg in fish would be associated with little
hazard to fish-eating birds, based on assumptions about the proportions
of fish in the diet and the BMF values and the NOAEC value of 100 ng 2,3,7,8-TCDD/kg
in ring-necked pheasant (Phasianus colchicus) eggs (49). Based
on the BMF value of 19 that we have used in our study, this would be equivalent
to approximately 114 ng/kg in the eggs of bald eagles. In a hazard assessment
of the effects of TCDD-EQ on bald eagles, Kubiak and Best (9) used
an NOAEC value of 20 ng TCDD-EQ/kg in the egg that was estimated from the
effects of 2,3,7,8-TCDD on the white leghorn chicken (50). This value
is three times greater than the value we have used in our assessment. The
NOAEC determined for wood ducks under field conditions is approximately
3-fold less than our value. However, in their field study White and Setinak
(48) did not measure the concentrations of other compounds such as
PCBs, which would likely contribute to the total TCDD-EQ. Thus, it would
be expected that their value would be an underestimate of the NOAEC (overestimate
of the toxicity of the measured TCDD-EQ). A dietary NOAEC value of 1.5 ng
TCDD-EQ/kg in the egg has been suggested to protect sensitive avian species
(40). By predicting the NOAEC in eggs using the BMF of 19, this would
correspond to a value of 28.5 ng/kg in the egg. The analysis of the potential
range of NOAEC values, based on literature values and assumptions, yields
a range of NOAEC values from 1 to 114 ng/kg in the egg. The value we have
used in our hazard assessment is greater than the value derived from a simple
application of the results with the chicken, a very sensitive species, but
less than values based on the pheasant, which seems to be one of the more
tolerant species. The value we have chosen to use is similar to that predicted
for the kestrel and similar to that derived for several other species. Implicit
in our choice of an NOAEC value is the assumption that bald eagles are more
sensitive to the effects of TCDD-EQ than pheasants but are less sensitive
than white leghorn chickens. The value we have chosen is approximately in
the middle of the range of NOAEC values observed in bird eggs: 10 times
less than the more tolerant species and 10 times greater than the least
tolerant. Thus, our value can be assumed to be conservative and protective
of most species, and there would seem to be no need to apply a safety factor
to our NOAEC value to protect eagles.
The uncertainty in the BMF for accumulation of TCDD-EQ from fish to the
eggs of bald eagles is not as great as that for estimates of the NOAEC.
The BMF values of Giesy et al. (10) for the accumulation of PCDD
and PCDF from Great Lakes fishes to the eggs of fish-eating colonial water
birds indicate that the BMF value was approximately 21. We have used the
consensus BMF value of 19 reported for the accumulation of TCDD-EQ from
fish to bald eagle eggs (9). If this biomagnification factor is applied,
a value of approximately 0.37 ng TCDD-EQ/kg, ww, is obtained for the dietary
NOAEC value (Table 1). Even if the NOAEC value were more like one of the
two extreme values, it would not change the conclusions made about the relative
hazard of fish consumption above or below the dams.
The greatest uncertainty in predicting the concentration of TCDD-EQ likely
to be deposited in eggs of bald eagles from eating fish is due to the relative
proportion of fish in the diet. We derived weighted average dietary content
of fishes in the diet that were based on measurements of the relative proportion
of each species of fish in the diet at each of the locations for which an
HI was calculated. The predicted concentration of TCDD-EQ in the eggs can
be underestimated if the eagles ate a great number of other fish-eating
birds such as gulls because there is an additional trophic magnification
step. Bald eagles can also take less contaminated mammals in the diet which
results in an overestimate of exposure. We did not correct for either of
these eventualities.
Concentrations of TCDD-EQ in eggs of bald eagles as great as 1650 ng
TCDD-EQ/kg, ww, have been measured in eggs of bald eagles living on the
shoreline of Lake Huron (9). This is approximately 236 times greater
than the NOAEC values that we used in our hazard assessment, 16.5 times
greater than the NOAEC values for pheasants (49), and approximately
165 times greater than the NOAEC values in white leghorn chicken eggs (50).
Since these values were determined with the H4IIE assay used in our studies
(25), the values are directly comparable to those reported here.
Thus, current concentrations of TCDD-EQ are sufficient to cause the observed
reduction in productivity of bald eagles living along the shores of the
Great Lakes or anadromous-accessible rivers (25). This substantiates
the hazard assessment conducted for consumption of fishes that TCDD-EQ is
the primary cause of the observed adverse effects in populations of bald
eagles that consume fishes from the Great Lakes (10) (Figure 7).
The observed exceedance of 230 for these bald eagle eggs is about 10 times
greater than would be predicted from total concentrations of PCBs. This
is due to weathering and trophic-level enrichment of the TCDD-EQ relative
to total concentrations of PCBs. It can be concluded that, at this time,
TCDD-EQ is the critical contaminant in the eggs of bald eagles along the
Great Lakes and that the greatest proportion of the TCDD-EQ is contributed
by the non-ortho-substituted PCBs (25).
Mercury. It is difficult to establish an NOAEC value for
the adverse effects of mercury on bald eagles. A theoretical NOAEC value
of 0.5 mg Hg/kg in the eggs of bald eagles has been proposed (6).
This value was derived from a study in which mallards (Anas platyrhynchos)
were fed a dietary dose of 0.5 mg Hg/kg (51). The concentration of
p,p´-DDE contained in eggs of bald eagles from studies reported
by Wiemeyer et al. (6), in which mercury concentrations were above
0.5 ppm, were greater than the p,p´-DDE concentration associated
with greater than a 50% decline in productivity. Thus, it would be difficult
to ascribe the observed effects to mercury alone. No effects of Hg have
been observed on reproduction of the white-tailed sea eagle, a species similar
to the bald eagle (52). A theoretical concentration for effect in
eggs was given as 1.0 mg Hg/kg in eggs, although no direct link to adverse
effects was noted (52). When concentrations of Hg in feathers of
white-tailed sea eagles of the Baltic ranged from 40 to 65 mg Hg/kg, eggs
from these areas were observed to seldom hatch (53). It should be
noted that no organochlorine pesticide analysis had been completed at the
time of publication for those data. It is likely that the observed effects
on hatchability of white-tailed sea eagle eggs were due to the effects of
organochlorine compounds. Subsequent reports refute the Hg/reproduction
theory of Berg et al. (53) and link white-tailed sea eagle reproductive
problems primarily to p,p´-DDE and PCBs (52,54). The
effects of Hg on wild populations of nesting bald eagles is difficult to
assess because there are nearly always organochlorine compounds present
(55). This is also true in the Great Lakes Basin where p,p´-DDE
and PCBs have been correlated with reproductive effects in bald eagles (4).
Mercury in fishes upstream of the dams represents a greater hazard to
bald eagles than at the downstream locations. The HIs, based on a conservative
estimate of the NOAEC value, were not very great (Figure 7). Hg is not the
most critical contaminant in the fish of these river systems, but it is
currently greater than the NOAEC for bald eagles if they ate, exclusively,
several of the species of fish studied, with yellow perch (Perca flavescens)
and walleye (Stizostedion vitreum) having the greatest Hg concentrations
and presumably posing the greatest risk to eagles. However, the relative
proportion of these two fish species in the diet of bald eagles along the
three streams was small (<3% of total diet). Because these fishes are
not a large part of the diet of bald eagles, it is not likely that current
concentrations of Hg are the cause of any population-level effects.
Implications for Continued Recovery
Models used to predict bald eagle population dynamics predict that decreases
in productivity are less critical than increases in adult mortality for
declines in population (43). This is unique to long-lived species
with delayed adult maturity. Modeling efforts have failed to use the scenario
of both increased adult mortality and depressed productivity occurring simultaneously
in a region with significant immigration. This appears to be the scenario
that is currently occurring among eagles who nest along the shorelines of
the Great Lakes and anadromous-accessible rivers. Eagles nesting along the
Great Lakes continue to show chronic effects of organochlorine pollutants,
primarily p,p´-DDE and PCBs. These effects include impaired
productivity, adult mortality, and teratogenicity (4,26,56). Eagles
nesting in interior regions are less contaminated and exhibit generally
greater productivity. In areas of great density such as the Chippewa National
Forest density-dependent declines in reproductive success have been noted.
These interior areas are the source of relatively uncontaminated bald eagles
that are supplying the "population sink" along the Great Lakes.
The Great Lakes have been the last area where bald eagle recovery has occurred;
however, the potential for occupancy along the shorelines and anadromous-accessible
rivers is great, based on the availability of unoccupied nesting habitat.
In Michigan, as the numbers of breeding areas have increased from 85 breeding
areas in 1977 to 246 in 1993, the percentage of Great Lakes shoreline and
anadramous-accessible river breeding areas has increased from 14 (n=12)
to 35% (n=86) during the same time period. It is important that essential
habitat be protected in the interior areas and managed to control human
activities near breeding areas during critical periods of eagle nesting.
This is necessary to maintain the greater productivity of these interior
areas so that the influence of the Great Lakes "population sink"
will not jeopardize the regional recovery of this species.
The importance of a vulnerable, relatively uncontaminated forage base
for bald eagles during the breeding season is imperative for successful
reproduction. Effects of environmental contaminants on bald eagle productivity
are well known (4,6,11,26,30), but other aspects of habitat requirements
need to be considered. Management techniques that control populations of
prey species used by bald eagles need to take into account the effect that
increases or decreases in utilization of con- taminated species will have
on the bald eagle's reproductive success. The need to maintain populations
of primarily warmwater fish in interior foraging areas for inland eagles
in the Midwest is imperative for maintaining the continuing recovery of
this species.
The fact that concentrations of PCBs and DDT remain at concentrations
that are still associated with lesser average productivities presents continuing
management issues, even though production of these compounds has ceased
in North America, and concentrations of most halogenated hydrocarbons in
the prey of eagles are decreasing in the Great Lakes Region (25).
Current concentrations of both PCBs, p,p´-DDE and TCDD-EQ are
sufficiently great to cause adverse effects in nesting bald eagles feeding
on the Great Lakes food web (6,10,25,57) (Figure 7). Our results
verify that poor productivity of eagles is inversely correlated with exposure
to PCBs, TCDD-EQ, and p,p´-DDE but not with mercury (11).
Furthermore, we have observed congenital deformities in bald eagle nestlings
(56). Developmental deformities have been observed in the populations
in which the greatest concentrations of PCBs have been found in the blood
of nestling eagles. The results of laboratory and field studies indicate
that the lethality of and deformities in embryos of colonial, fish-eating
water birds of the Great Lakes are due to the toxic effects of multiple
compounds, primarily TCDD-EQ contributed by coplanar PCBs, PCDDs, and PCDFs.
These compounds express their effects through a common mode of action, the
Ah receptor (25). The concentration of total PCBs and TCDD-EQ (58),
converted from congener specific data, in two addled bald eagle eggs collected
near Lakes Michigan and Huron were 83 and 98 µg/g total PCBs and 21,369
and 30,894 pg/g as TCDD-EQ, respectively (58). Currently, TCDD-EQ,
contributed primarily from coplanar PCBs, seems to be the critical toxicant
limiting bald eagle reproduction. Concentrations of TCDD-EQ in bald eagle
eggs exceed known effect levels in poultry experiments, either by total
PCB concentration or by conversion of individual PCB congeners (9,37,58).
Our results suggest that exposure of bald eagles to Great Lakes fishes
should be minimized. It would be inappropriate to use hacking programs to
reestablish populations of eagles or improve their genetic diversity along
the Great Lakes shoreline, especially for Lakes Erie and Ontario, with the
current concentrations of organochlorine contaminants in their potential
prey. Management practices that increase the potential exposure of eagles
to chlorinated hydrocarbons in Great Lakes fishes, e.g., passage of fishes
around dams on tributaries to Lakes Michigan, Huron, and Erie, could have
adverse effects on productivity of bald eagles in regions that currently
produce sufficient numbers to act as a source of eagles to colonize other
areas. Only by maintaining a readily available, relatively uncontaminated
food source for eagles during the breeding season can we continue to experience
the population recovery of this species in the Midwest.
Summary
By far the most important factor in controlling current bald eagle reproduction
along the shorelines of the Great Lakes is the influence of environmental
contaminants. We have shown that p,p´-DDE and TCDD-EQ from
PCBs are correlated with impaired reproductive success of eagles along the
shorelines of Lakes Superior, Michigan, Huron, and Erie, as well as at Voyageurs
National Park. We have shown further that concentrations of mercury are
not correlated with bald eagle productivity in the Great Lakes region.
The results of the hazard assessment indicate that current concentrations
of DDE, total PCBs, and TCDD-EQ in fishes continue to have adverse effects
on bald eagles. living above the dams on the three rivers should not have
adverse effects on bald eagles. While there might be some effects of DDE
on bald eagle productivity, it is not at this time the critical contaminant.
Concentrations of both total PCBs and TCDD-EQ in fishes below the dams currently
represent a significant hazard to bald eagles living along the Great Lakes
shoreline or on rivers below the downstream-most dams. Of these two measures
of contamination, currently TCDD-EQ is the more critical. Even though the
majority of the TCDD-EQ found in Great Lakes fishes is contributed by the
planar PCB congeners, there are additional sources of TCDD-EQ that result
in more TCDD-EQ than would be expected from technical Aroclors alone. Weathering
of Aroclor mixtures results in an enrichment of the non-ortho-substituted
congeners and results in a PCB mixture in both fishes and bald eagle eggs
that contains more TCDD-EQ than would be expected in the original Aroclor
technical mixtures. Our findings indicate that the known toxicants, total
PCBs and TCDD-EQ, are both occurring at sufficient concentrations in fishes
and in bald eagle eggs to explain the poorer productivity observed in eagles
that nest along the shorelines of the Great Lakes and along anadromous-accessible
rivers, without the need to invoke other causes such as weather, food availability,
or other, as yet undefined, contaminants. The assimilative capacity of the
Great Lakes has been exceeded, and no additional loadings of compounds that
can contribute to the total concentrations of TCDD-EQ should be allowed.
The results of the hazard assessment are supported by the observed productivities
of bald eagles in the upstream and downstream areas. Bald eagle populations
throughout the Midwest have experienced a steady increase in breeding pairs
throughout 1977 to 1993. However, productivity has not been uniform throughout
the study area. Bald eagles nesting along the Great Lakes shoreline and
at Voyageurs National Park were significantly less productive than those
from interior areas of Michigan and the Chippewa and Superior National Forests
in Minnesota.
Availability of physical habitat does not seem to be limiting expansion
of the bald eagle population along the upper Great Lakes shorelines. Bald
eagles are restricted from some areas due to human disturbance or physical
structure of the habitat. There are still areas deemed to be suitable nesting
habitat, which are currently unoccupied by bald eagles. This is especially
true of the northern forested regions that are less populated by humans.
Suitable habitat along Lake Erie is scarce and may be a limiting factor
in the near future.
As the bald eagle continues to reoccupy areas where they were extirpated
during the 1950s and 1960s, differential effects of productivity could become
even more pronounced. Density-dependent factors will continue to cause eagles
from the more interior areas, where more eagles are fledged than is necessary
to maintain a stable age distribution, to reoccupy the Great Lakes shorelines.
This is already occurring, as the Great Lakes subpopulation has the greatest
growth rate in terms of numbers of new breeding areas established. Additional
investigation into the dynamics of these populations is needed to monitor
the recovery of this species and to compare areas with exposure to greater
concentrations of organochlorine compounds with more pristine areas. The
effect of differential adult turnover along the Great Lakes shoreline also
needs to be understood before a population model of the region can be produced
and verified. Although the number of bald eagles in the Great Lakes Basin
and adjacent areas has continued to increase as the effects of p,p´-DDE
have subsided, the carrying capacity of the region is still uncertain.
REFERENCES
1. Grier JW. Ban of DDT and subsequent recovery of reproduction
in bald eagles. Science 218:1232-1235 (1982).
2. Postupalsky S. The bald eagles return. Nat Hist 87:62-63
(1985).
3. Colborn T. Epidemiology of Great Lakes bald eagles.
J Toxicol Environ Health 33:395-453 (1991).
4. Best DA, Bowerman WW, Kubiak TJ, Winterstein SR, Postupalsky
S, Shieldcastle M. Reproductive impairment of bald eagles along the Great
Lakes shorelines of Michigan and Ohio. In: Raptor Conservation Today, (Meyburg
BU, Chancellor RD, eds). East Sussex, U.K.:World Working Group on Birds
of Prey and The Pica Press, 1994; 697-702.
5. Stalmaster MV. The Bald Eagle. New York:Universe Books,
1987.
6. Wiemeyer SN, Lamont TJ, Bunck CM, Sindelar CR, Gramlich
FJ, Fraser JD, Byrd MA. Organochlorine pesticide, polychlorobiphenyl, and
mercury residues in bald eagle eggs--1969-1979--and their relationships
to shell thinning and reproduction. Arch Environ Contam Toxicol 13:529-549
(1984).
7. Gilbertson M, Kubiak TJ, Ludwig JP, Fox GA. Great Lakes
embryo mortality, edema, and deformities syndrome (GLEMEDS) in colonial
fish-eating birds: similarity to chick-edema disease. J Toxicol Environ
Health 33:455-520 (1991).
8. Kubiak TJ, Harris HJ, Smith LM, Schwartz T, Stalling
DL, Trick JA, Sileo L, Docherty D, Erdman TC. Microcontaminants and reproductive
impairment of the Forster's tern in Green Bay, Lake Michigan--1983. Arch
Environ Contam Toxicol 18:706-727 (1989).
9. Kubiak TJ, Best DA. Wildlife risks associated with passage
of contaminated, anadromous fish at Federal Energy Regulatory Commission
licensed dams in Michigan. Manuscript Report. East Lansing, MI:U.S. Fish
and Wildlife Service,1993.
10. Giesy JP, Ludwig JP, Tillitt DE. Deformities in birds
of the Great Lakes region: assigning causality. Environ Sci Technol 28:128-135
(1994).
11. Bowerman WW, Evans ED, Giesy JP, Postupalsky S. Effects
of mercury and selenium on bald eagle reproduction in the Great Lakes Basin.
Arch Environ Contam Toxicol 27:294-298 (1994).
12. Braune B, Norstrom R. Dynamics of organochlorine compounds
in herring gulls: II. Tissue distribution and bioaccumulation in Lake Ontario
gulls. Environ Toxicol Chem 8:957-968 (1989).
13. Davison KL, Sell JL. DDT thins shells of eggs from
mallard ducks maintained on ad libitum or controlled-feeding regimes.
Arch Environ Contam Toxicol 2:222-232 (1974).
14. Blus L, Cromartie E, McNease L, Joanen T. Brown pelican:
population status, reproductive success, and organochlorine residues in
Louisiana, 1971-1976. Bull Environ Contam Toxicol 22:128-135 (1979).
15. Tucker RK, Crabtree DG. Handbook of toxicity of pesticides
to wildlife. Resource publication 84. Washington:United States. Bureau of
Sport Fisheries and Wildlife, 1970.
16. Heath RG, Spann JW, Kreitzer JF, Vance C. Marked DDT
impairment of mallard reproduction in controlled studies. Nature 224:47-48
(1972).
17. Longcore JR, Samson FB, Whittendale TW Jr. DDT thins
eggshells and lowers reproductive success of captive black ducks. Bull Environ
Contam Toxicol 6:485-490 (1971).
18. Wiemeyer SN, Porter RD. DDE thins eggshells of captive
American kestrels. Nature 227:737-738 (1970).
19. McClain MAR, Hall LC. DDE thins screech owl eggshells.
Bull Environ Contam Toxicol 8:65-68 (1972).
20. Peakall DB, Lincer JL, Risebrough RW, Pritchard JB,
Kinter WB. DDE-induced egg-shell thinning: structural and physiological
effects in three species. Comp Gen Pharmacol 4:305-313 (1973).
21. Lincer JL. DDE-induced eggshell-thinning in the American
kestrel: a comparison of the field situation and laboratory results. J Appl
Ecol 12:781-793 (1975).
22. Newton I. Population Ecology of Raptors. Vermillion,
SD:Buteo Books, 1979.
23. Mendenhall VM, Klass EE, McClain MAR. Breeding success
of barn owls (Tyto alba) fed low diets of DDE and dieldrin. Arch
Environ Contam Toxicol 12:235-240 (1983).
24. Stendsell RC, Gilmer DS, Coon NA, Swingford DM. Organochlorine
and mercury residues in Swainson's and ferruginous hawk eggs collected in
North and South Dakota, 1974-1979. Environ Monit Assess 10:37-41 (1988).
25. Giesy JP, Ludwig JP, Tillitt DE. Dioxins, dibenzofurans,
PCBs and similar chlorinated, diaromatic hydrocarbons in and their effects
on birds: wildlife biomonitoring for hazards of complex environmental mixtures
in the Laurentian Great Lakes. In: Dioxins and Health (Schecter A, ed).
New York:Plenum Press, 1994; 254-307.
26. Bowerman WW, Best DA, Kubiak TJ, Postupalsky S, Shieldcastle
MC, Giesy JP. The influence of environmental contaminants on bald eagle
populations in the Laurentian Great Lakes, North America. In: Raptor Conservation
Today, (Meyburg BU Chancellor RD, eds). East Sussex, U.K.: World Working
Group on Birds of Prey and The Pica Press, 1994;703-707.
27. Sprunt A IV, Robertson WB Jr, Postupalsky S, Hensel
RJ, Knoder CE, Ligas FJ. Comparative productivity of six bald eagle populations.
Trans N Am Wildl Nat Resour Conf 38:96-106 (1973).
28. Postupalsky S. Toxic chemicals and declining bald eagles
and cormorants in Ontario. Manuscript Report 20. Ottawa: Canadian Wildlife
Service, 1971.
29. Nisbet ICT. Organochlorines, reproductive impairment
and declines in bald eagle (Haliaeetus leucocephalus) populations:
mechanisms and dose-response relationships. In: Raptors in the Modern World
(Meyburg B-U, Chancellor RD, eds). Berlin: World Working Group on Birds
of Prey, 1989;483-489.
30. Kozie KD, Anderson RK. Productivity, diet, and environmental
contaminants in bald eagles nesting near the Wisconsin shoreline of Lake
Superior. Arch Environ Contam Toxicol 20:41-48 (1991).
31. Wiemeyer SN, Bunck CM, Stafford CJ. Environmental contaminants
in bald eagle eggs--1980-84--and further interpretations of relationships
to productivity and shell thickness. Arch Environ Contam Toxicol 24:213-227
(1993).
32. Helander B. (1983) Reproduction of the white-tailed
sea eagle Haliaeetus albicilla (L.) in Sweden, in relation to food
and residue levels of organochlorine and mercury compounds in the eggs.
PhD dissertation. Stockholm:University of Stockholm, 1983.
33. Blus LJ. DDE in bird's eggs: comparison of two methods
for estimating critical levels. Wilson Bull 96:268-276 (1984).
34. Shirazi MA, Bennett RS, Lowrie L. An approach to environmental
risk assessment using avian toxicity tests. Arch Environ Contam Toxicol
16:263-271 (1988).
35. Britton WM, Huston TM. Influence of polychlorinated
biphenyls in the laying hen. Poult Sci 52:1620-1624 (1973).
36. Scott ML, Zimmerman JR, Marinsky S, Mullenhoff PA,
Rumsey GL, Rice RW. Effects of PCBs, DDT, and mercury compounds upon egg
production, hatchability and shell quality in chickens and Japanese Quail.
Poult Sci 54:350-368 (1975).
37. Platonow NS, Reinhart BS. The effects of polychlorinated
biphenyls (Aroclor 1254) on chicken egg production, fertility and hatchability.
Can J Comp Med 37:341-346 (1973).
38. Scott ML. Effects of PCBs, DDT, and mercury compounds
in chickens and Japanese Quail. Fed Proc 36:1888-1893 (1977).
39. Tumasonis CF, Bush B, Baker FD. PCB levels in egg yolks
associated with embryonic mortality and deformity of hatched chicks. Arch
Environ Contam Toxicol 1:312-324 (1973).
40. Ludwig JP, Giesy JP, Summer CL, Bowerman WW, Heaton
SN, Aulerich RJ, Bursian S, Auman HJ, Jones PD, Williams LL, Tillitt DE,
Gilbertson M. A comparison of water quality criteria in the Great Lakes
Basin based on human or wildlife health. J Gr Lakes Res 19:789-807 (1993).
41. Grier JW. Modeling approaches to bald eagle population
dynamics. Wildl Soc Bull 8:315-322 (1980).
42. Cook PM, Erickson RJ, Spehar RL, Bradbury SP, Ankley
GT. Interim report on data and methods for assessment of 2,3,7,8-tetrachlordibenzo-p-dioxin
risks to aquatic life and associated wildlife. Report No. 600/R-93/055.
Washington:U.S. Environmental Protection Agency, 1993.
43. Hoffman DJ, Rice CP, Kubiak TJ. (1994) PCBs and dioxins
in birds. In: Interpreting Environmental Contaminants in Animal Tissue (Beyer
WN, Heinz GH, eds). Chelsea, MI:Lewis Publishers (in press).
44. Jones PD, Giesy JP, Newsted JL, Verbrugge DA, Beaver
DL, Ankley GT, Tillitt DE, Lodge KB. 2,3,7,8-Tetrachlordibenzo-p-dioxin
equivalents in tissues of birds at Green Bay, Wisconsin, USA. Arch Environ
Toxicol Safety 24:345-354 (1993).
45. De Vito MJ, Maier WE, Diliberto JJ. Comparative ability
of various PCBs, PCDFs and TCDD to induce cytochrome P4501A1 activity following
4 weeks treatment. Fundam Appl Toxicol 20:125-130 (1993).
46. Henshel DS. 1993 LD50 and teratogenicity
studies of the effects of TCDD on chicken embryos. In: Abstracts of 14th
Annual Meeting, Society of Environmental Toxicology and Chemistry, Houston,
TX, November 1993. Pensacola, FL:Society of Environmental Toxicology and
Chemistry, 1993.
47. Henshel DS, Hehn BM, Vo MT, Steeves JD. A short-term
test for dioxin teratogenicity using chicken embryos. In: Environmental
Toxicology and Risk Assessment, Vol 2, ASTM-STP 1216 (Gorsuch JW, Dwyer
FJ, Ingersoll CG, La Point TW, eds). Philadelphia, PA:American Society for
Testing and Materials, 1993.
48. White DT, Setinak JT. Dioxins and furans linked to
reproductive impairment in wood ducks. J Wildl Manage 58:100-106 (1994).
49. Nosek JA, Craven SR, Sullivan JR, Hurley SS, Peterson
RE. Toxicity and reproductive effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin
in ring-necked pheasant hens. J Toxicol Environ Health 35:87-198 (1992).
50. Verret MJ. Investigation of the toxic and teratogenic
effects of halogenated dibenzo-p-dioxins and dibenzofurans in the
developing chicken embryo. Memorandum Report. Washington:U.S. Food and Drug
Administration, 1976.
51. Heinz GH. Methylmercury: reproductive and behavioral
effects on three generations of mallard ducks. J Wildl Manage 43:394-401
(1979).
52. Helander B, Olsson M, Reutergašrdh L. Residue
levels of organochlorine and mercury compounds in unhatched eggs and the
relationship to breeding success in White-tailed sea eagles Haliaeetus
albicilla in Sweden. Holarct Ecol 5:349-366 (1982).
53. Berg W, Johnels A, Sjostrand B, Westermark T. Mercury
content in feathers of Swedish birds from the past 100 years. Oikos 17:71-83
(1966).
54. Koivusaari J, Nuuja I, Palokangus R, Finnlund M. Relationships
between productivity, eggshell thickness and pollutant contents of addled
eggs in the population of white-tailed eagles Haliaetus albicilla L.
in Finland during 1969-1978. Environ Pollut (Series A) 23:41-52 (1980).
55. Frenzel RW, Anthony RJ. Relationship of diets and environmental
contaminants in wintering bald eagles. J Wildl Manage 53:792-802 (1989).
56. Bowerman WW, Kubiak TJ, Holt JB, Evans DB, Eckstein
RJ, Sindelar CR. Observed abnormalities in mandibles of nestling bald eagles.
Bull Environ Contam Toxicol 53:450-457 (1994).
57. Safe S. Polychlorinated biphenyls (PCBs) and polybrominated
biphenyls (PBBs): biochemistry, toxicology, and mechanism of action. CRC
Crit Rev Toxicol 13:319-393 (1984).
58. Bowerman WW, Best DA, Evans ED, Postupalsky S, Martell
MS, Kozie KD, Welch RL, Scheel RH, Durling KF, Rogers JC, Kubiak TJ, Tillitt
DE, Schwartz TR, Jones PD, Giesy JP. PCB concentrations in plasma of nestling
bald eagles from the Great Lakes Basin, North America. In: Proceedings of
10th International Conference on Organohalogen Compounds Vol 4 (Fiedler
H, Huttzinger O, eds). Bayreuth, Germany, September 1991. Bayreuth, Germany:Ecoinforma
Press, 1991; 212-216.
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