Title: Effects on Fish and Wildlife Populations
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Permanent Link: http://ufdc.ufl.edu/WL00001329/00001
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Title: Effects on Fish and Wildlife Populations
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Language: English
Publisher: Elsevier Science Publishers B.V
Spatial Coverage: North America -- United States of America -- Florida
Abstract: Effects on Fish and Wildlife Populations, by Micheal Gilbertson, Commercial Chemicals Branch, Conservation and Protection, Environment Canada, Ottawa, Ontario
General Note: Box 8, Folder 3 ( Vail Conference, 1993 - 1993 ), Item 43
Funding: Digitized by the Legal Technology Institute in the Levin College of Law at the University of Florida.
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Bibliographic ID: WL00001329
Volume ID: VID00001
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b'ngoflh and JIen. i. ) Halagrnate biphenyls. lIphenyls.
JapkIaltear5. dibenmadiazins ndw muete products
0 1989 Elsevier Science Publishers A. V (Biomedical Division)


Effects on fish and wildlife populations


Commercial Chemicals Branch. Conservation and Protection.
Environment Canada. Ottawa. Ontario KA OE7

L' 4.1. Introduction

During the past twenty-five years there has been increasing evidence of outbreaks of
\ disease among fish and wildlife epizooticss) which could not be attributed to conven-
tion;l etiological agents such as microorganisms and parasites. These disease out-
breaks were characterized by reproductive dysfunction and mortality and, in some
cases, led to regional or continental extirpation of populations of certain fish and
wildlife species. In many instances chemicals, and specifically organochlorine chem-
icals, have been demonstrated to be the etiological agents. Though the subject of this
book is concerned with polychlorinated biphenyls, dioxins and furans, it is impossible
to discuss these outbreaks without acknowledging the presence and likely contribu-
tory effects of other organochlorine compounds, notably DDT and dieldrin. Whereas
human incidents of exposure to PCB or dioxin have usually been associated with a
single primary source, such as an explosion causing dermal contact or leakage of a
heat-transfer medium causing contamination of food, fish and wildlife exposures
have frequently been to many compounds from many sources. It has thus been
concomitantly more difficult, than in human epidemiology, to infer causality be-
tween an observed epizootic and a putative causal compoundss.
In the late 1950s and early 1960s, research on British peregrines (Falco peregri-
nus) (Ratcliffe, 1963), sparrow hawks (Accipiter nisus) (Prestt, 1965) and Scottish
golden eagles (4quila chrysaetos) (Lockie and Ratcliffe, 1964) showed declines in
the populations and in reproductive success. Food-chain contamination with dieldrin,
used as a seed dressing and as a sheep dip, seems to have been largely responsible for

severe adult and embryonic mortality (reviewed in Cooke, 1973). A significant
decline in eggshell thickness of these and other raptor species, coincident with the
introduction of organochlorine compounds into agriculture about 1947, was demon-
strated on two continents (Ratcliffe, 1967; Anderson and Hickey, 1972) and later
shown to be caused by DDT and its metabolites (reviewed in Cooke, 1973). In North
America, DDT caused widespread extirpation of populations of peregrine falcons,
bald eagles (Haliaeetus leucocephalus) and ospreys (Pandion haliaetus) due to
eggshell thinning and embryonic mortality (Hickey, 1969). The use of DDT or
related compounds to control biting insects caused mortality of western grebes
(4echmophorus occidentalis) in Clear Lake, California, through food-chain accu-
mulation (Hunt and Bischoff, 1960) and embryonic mortality in lake trout (Salvel-
inus namaycush) in upper New York State (Burdick et al., 1964). Against this
background of early research on effects of organochlorine insecticides on fish and
wildlife populations, Jensen, in 1966, identified the presence of polychlorinated
biphenyls (PCB) in Swedish wildlife and opened up a new chapter of organochlorine
research (Anon, 1966). The presence of PCB, together with other organochlorine
compounds, in marine mammals and in freshwater and marine fish as well as wildlife.
revealed the extent of global contamination with this compound (Holmes et al., 1967;
Holden and Marsden, 1967; Risebrough, 1969; Risebrough et al., 1968; Koeman et
al., 1969; Jensen et al., 1969).
In the early 1970s research burgeoned to document the effects of organochlorine
chemicals on fish and wildlife using the following three distinct approaches: (a) field
observation of effects; (b) analytical chemistry to determine levels; and (c) labora-
tory experimentation to investigate toxicity and modes of action. The synthesis of
these three approaches to make a scientifically defensible case history is a difficult
undertaking. Epizootiology is in many ways much more difficult to undertake than
human epidemiology. Less is known about diseases in animals than in humans and
it is often not clear whether an apparent association is indeed causal or merely
spurious. Background information on natural sources of mortality may not exist and
the effects of human factors, such as guns, traps, fishing nets and automobiles, on
population dynamics may be unestimated. General pollution, habitat destruction and
urban sprawl together with changes in food supply and a host of other man-made
influences can have disastrous effects on populations of wild organisms and may
mask or amplify the effects of organochlorine chemicals. Despite these, and possibly
other, confounding factors, of which readers should be mindful, there is a series of
studies of epizootics in fish and wildlife that are believed to be causally related to the
presence of organochlorine compounds including PCB, dioxins and dibenzofurans.
These epizootics are the subject of this paper. This is not an exhaustive critical
review since the literature is much too large; it is more in the form of a catalogue of
case histories citing reviews and the most recent literature. Nor is it a paper reviewing
laboratory experimentation and levels in the environment except in so far as infor-
mation from these disciplines aids in interpretation of the epizootics.

Michael Gilbertson


Effects on fish and wildlife populations 105

4.2. Case histories

4.2.1. Marine mammals
Chemically-induced epizootics have been suspected in several marine mammal po-
pulations in North America and Europe. Risebrough (1978) reviewed the literature
up toiabout 1977. Several reviews of the levels of organochlorine compounds in
marine mammals have been published in the intervening decade (Gaskin, 1982;
Wagemann and Muir, 1984; Aguilar, 1985) and a recent publication on the decline
of the dolphin (Tursiops truncatus) and porpoise (Phocoena phocoena) in the North
Sea included a detailed consideration of the studies on effects of organochlorine
chemicals on marine mammals and their populations (Kayes, 1985). Pinnipeds. One of the first cases to be investigated was the incidence of
premature parturition in colonies of California sea lions (Zalophus californianus)
(DeLong et al., 1973). Females that aborted had mean DDT and PCB levels in
blubber of 820 ppm and 110 ppm respectively compared with 100 ppm and 17 ppm
in females that carried to full term. The case was, however, inconclusive because
microbiological analysis showed the presence of bacterial and viral agents associated
with abortion in mammals (Gilmartin et al., 1976). The authors discussed the
possibility that organochlorine compounds might have had an immunosuppressive
effect on the sea lion colony. Other preliminary observations on premature parturi-
_ tion and deformities in Pacific coast pinnipeds are cited in Risebrough (1978).
SExtensive work has been undertaken since the early 1970s on declining populations
of common or harbour (Phoca vitulina), ringed (Pusa hispida) and grey (Halichoe-
rus grypus) seals in the Baltic and on harbour seals in the Wadden Sea at the
southern end of the North Sea off Holland, West Germany and Denmark. Baltic
seals exhibit a high incidence of uterine occlusions and stenoses that may be the
result of fetal death and resorption (Helle et al., 1976; Helle, 1980; Bergman et al.,
1981). Blubber levels of DDT and PCB in non-pregnant female ringed seals with
uterine occlusions were significantly higher at 130 ppm and 110 ppm respectively
compared to 100 ppm and 89 ppm in non-pregnant females with a normal uterus.
Pregnant females contained 88 ppm of DDT and 88 ppm PCB (Helle, 1981). These
effects. were thought mainly to be the result of the PCB contamination rather than
the DIT, based on comparisons with the levels of these compounds in California sea
lions and on results of experiments with mink (Mustela vison) (Helle et al., 1976).
Kayes:(1985), in reviewing the ringed seal and organochlorine data, has pointed out
that the higher levels of organochlorine compounds in females with stenoses and
occlusions could be a result rather than a cause of the reproductive failure because
these compounds are removed from the body by pregnancy and lactation. Bergman
and Olsson (1985) reviewed the pathologies of grey and ringed seals recovered from
the Baltic including uterine stenoses and occlusions, intestinal ulcers, adrenocortical
hyperplasia, arteriosclerosis and renal glomerulopathy. Grey seals commonly had


Michael Gilbertson

uterine tumours and changes in the integumentary system including thin epidermis,
hyperkeratosis, cystic dilatation of hair follicles and deformation and fractures of the
claws. The authors believed that the pathologies were part of a disease complex
caused by exogenous toxins such as PCB. Recent evidence shows that Baltic ringed,
grey and common seals are now close to extinction (Olsson, 1986). Mention should
also be made of the high levels of DDT and PCB (both about 65 ppm) found in adult
Saimaa ringed seal (Phoca hispida saimensis); a freshwater subspecies that lives in
Lake Saimaa in southern Finland (Helle et al., 1983).
Research has been carried out on the status of the harbour seal in the Wadden Sea
(Reijnders, 1976; Summers et al., 1978; van Haaften, 1982). The Wadden Sea is an
extensive area of the North Sea off the coast of Holland, Germany (Niedersachsen
and Schleswig-Holstein) and Denmark. The area is separated from the open sea by
a series of barrier islands and contains extensive tidal flats that are ideal habitat for
harbour seals as well as many other species of wildlife. In the decade between 1965
and 1975 the population off the Dutch coast decreased from about 1450 to 550
animals despite cessation of hunting. Similarly in Niedersachsen the population
declined from about 2200 in 1960 to 1000 in 1974. Populations in Schleswig-Holstein
and Denmark remained stable. Juvenile mortality in the Dutch population was
higher than in Schleswig-Holstein (Reijnders, 1978; Drescher, 1978). Based on the
results of organochlorine analyses, Reijnders (1980) concluded that the decrease in
reproductive success of the Dutch harbour seal population was strongly correlated
with elevated concentrations of PCB in seal tissues.
Further support for the role of PCB involvement comes from experimental work
undertaken on common seal reproduction (Reijnders, 1986). Two groups of 12 female
common seals were fed diets of fish obtained from different sources. One source of
fish was from the Dutch Wadden Sea and the second source was from the north-east
Atlantic. Three males that were fed Atlantic fish were alternated between the groups
during the breeding season. Only 4 of the 12 females on the Wadden Sea diet were
pregnant compared with 10 of the 12 females on the Atlantic fish diet. Based on the
variation in circulating levels of hormones in the non-pregnant females it appeared
that the effect occurred past ovulation, at the time of implantation.
An anomalous observation should be recorded. In contrast to the findings in the
Baltic, no stenoses or occlusions were found in any of the Wadden Sea harbour seals
even though Reijnders (1980) noted that levels of PCB in the Schleswig-Holstein
population of harbour seals were comparable with the levels in Baltic seals. The term
'PCB' may be adequate for some purposes of describing levels of this family of
compounds in various environmental materials including biological tissues. It is
probably inadequate to assist in the interpretation of biological observations on
highly contaminated populations from widely different locations because the isomer-
specific composition of the PCB and associated furans is unlikely to be the same
(Bowes et al., 1975). The experimental findings (Darnerud et al., 1986) that the
highly potent 3,3',4,4'-tetrachlorobiphenyl congener is selectively concentrated in

Effects on fish and wildlife populations 107

the mammalian fetus provides a possible feasible explanation of the anomaly. Thus
when a more complete analysis of the Baltic and Wadden Sea tissues has been
undertaken, the reasons for this anomaly may become apparent.
I* Puget Sound, Washington State, USA, long-term research has been undertaken
on the harbour seal population not only to document population changes but also to
investigate the possible role of contaminants (Calambokidis et al., 1985). Generally,
the harbour seal population in Puget Sound has increased under protection. A high
incidence of premature births and neonatal mortality was found at some sites but
this did not appear to be related to the occurrence of organochlorine contaminants.
Two anomalies that appeared to be possibly linked to PCB or related compounds
were indistinct pelage, particularly in subadults, and umbilical lesions. Similar urn
biblical lesions had been noted in the Dutch Wadden Sea population (Reijnders et al.,
1982) and the possibility of changes in pelt quality has been raised by Holden (1975). Cetaceans. There are several populations of cetaceans that have accumu-
lated high concentrations of organochlorine compounds, particularly PCB (Wage-
mane and Muir, 1984); however, there are very few case histories where epizootics
or population changes have been demonstrated to be related to these compounds.
Cetaceans are difficult species for which to obtain accurate population estimates or
on which to undertake experimentation. Some of the evidence therefore seems to be
anecdotal or only partially systematized though some comprehensive studies have
been undertaken.
The harbour porpoise populations in Puget Sound (Calambokidis et al., 1985), the
Dutcl :Wadden Sea (Verwey and Wolff, 1982s) and the Baltic (Otterlind, 1976,
Almquist, 1982) have been regionally extirpated. The declines in each case were
suspected to be caused by high levels of PCBs, though high levels of DDT were also
sometimes present. PCB and DDT levels in the blubber of harbour porpoises from
the Wadden Sea were 88 ppm and 41 ppm respectively (Koeman ct al., 1972) and
from the Baltic Sea 114 ppm and 38 ppm (Huschenbeth, 1977). High levels (79 ppm
and 103 ppm DDT in males versus 47 ppm PCB and 39 ppm DDT in females) were
found ib harbour porpoises in the Bay of Fundy (Gaskin, 1982; Gaskin et al., 1983);
however,; no adverse biological effects seem to have been noted though there are
preliminary indications that the proportion of pregnant females may have declined
from 45% in 1950-1955 to 22% in 1970-1980 (Gaskin et al., 1984).
Declines in killer whales (Orcinus orca) in Washington State have been associated
with high levels of organochlorine compounds in blubber (250 ppm PCB and 640
ppm DDT) (Calambokidis et al., 1985). There has also been a decline in the popu-
lation of bottlenose dolphins in the Wadden Sea (Verwey and Wolff, 1982b) and it is
suspected that the reason for their decline is the same as for the harbour porpoise.
It would seem that the presence of 190 ppm PCB and 94 ppm DDT in blubber of
pilot whales (Globicephala melaena) from the French Coast would warrant investi-
gation of their effects on the population (Alzieu and Duguy, 1979).

Michael Gilbertson

Detailed investigations are being undertaken on the decline of the population of
beluga whales (Delphinapterus leucas) in the St-Lawrence Estuary (Bcland and
Martineau, 1986). Estimates of the population of St-Lawrence belugas indicate
about 5000 animals in the 1880s (Reeves and Mitchell, 1984). Commercial hunting
ceased in the 1950s; however, the population, despite protection, has declined to
about 350-500 animals, likely through excess mortality at an early age (Beland and
Martineau, 1986). Detailed autopsies on 15 belugas revealed two with perforated
gastric ulcers and six with tumours including two splenic fibromas. Other pathologies
included a bladder cancer, and a dissecting aneurism of the pulmonary artery
(Martineau ct al., in press). High levels of organochlorine compounds were found in
the blubber (about 250 ppm PCB and 100 ppm DDT in adult males versus about 50
ppm PCB and 10 ppm DDT in adult females). One juvenile female contained over
500 ppm PCB and over 90 ppm .DDT (Martineau et al., 1987), which can be
compared with the levels of 800 ppm PCB and 827 ppm DDT found in a dead
juvenile female found in 1971 (Sergeant, 1980). The presence of mirex in the belugas
indicates that Great Lakes sources of organochlorine chemicals may be significant.
The etiology, however, may be further complicated by the presence of polynuclear
aromatic hydrocarbons from the Saguenay Fjord and possible toxicologicalinterac-
tions between these two families of compounds.

4.2.2. Freshwater mammals
Several studies have been undertaken to investigate whether regional declines in
mink and river otters (Lutra canadensis) might be related to organochlorine com-
pounds. Field research on these animals is extremely difficult because of their
solitary and secretive habits. Interpretation has, however, been aided by the large
body of experimental toxicology on mink which has been extensively reviewed
(Aulerich and Ringer, 1977; Ringer, 1982; Gilbertson, 1988). This experimental
evidence was developed to determine the cause of high kit mortality in Michigan and
Ontario ranch mink operations (Hartsough, 1965) in which Great Lakes fish were
being incorporated into the feed. PCBs are much more toxic to mink than DDT or
dieldrin and mortality is associated with a liver or muscle concentration of about 5
ppm PCB (Aulerich and Ringer, 1977).
Trapping data have been used to infer trends in estimates of populations of otter
and mink in several parts of the United States. Declines in the otter population of
the lower Columbia River, Oregon, (Henny ct al., 1981) were associated with high
PCB levels in liver (geometric mean: 9.3 ppm in males, 3.5 ppm in females). The
apparent decline in the mink population in the lower Columbia River paralleled the
apparent decline in the rest of Oregon state, based on trapping records. PCB levels
in livers of mink from the lower Columbia River were about I ppm, a level which has
been associated with kit mortality (Platonow and Karstad, 1973). The possible
association of these PCB levels with power generation on the Columbia River should
be investigated. A similar analysis of otter trapping data for Washington State was


undertaken to investigate whether a decline had occurred around Puget Sound
(Calambokidis et al., 1985). The authors concluded that there was no evidence for
such a decline in river otters in the counties bordering Puget Sound.
Levels of PCB which could be associated with reproductive impairment have been
detected in livers of mink sampled from Maryland (1.4 ppm in females and 1.5 ppm
in males on a fresh weight basis) (O'Shea et al.. 1980). These samples were taken
from a rural area with no obvious potential source of PCB pollution. These mink
may have become contaminated with PCB which was deposited in these remote areas
after long-range atmospheric transport.
Based on trapping data, there is preliminary evidence that the river otter has
declined in parts of Wisconsin (Kohn and Ashbrenner, 1984) and Michigan (Kubiak,
personal communication) contiguous with Lake Michigan. A wild mink found dead
in a Green Bay, Wl, marsh had a PCB concentration in the liver of 5.7 ppm (Kubiak,
personal communication). This level of PCB is similar to levels found in mink killed
experimentally and is consistent with PCB as the cause of death of the wild mink. It
seems likely that levels of PCB are so high in certain areas close to Lake Michigan
that neither otters nor mink can survive or reproduce successfully. Further field
investigations are needed to document this situation.
SA preliminary survey in southern Ontario of levels of organochlorine compounds
Sin mink trapped near Lake Erie, Lake Ontario and the St-Lawrence River indicated
That levels of PCB were high enough to cause reproductive impairment and even
mortality (Proulx et al., 1985).
Recent analytical data on mink and otters trapped in New York State indicate
elevated levels of PCB contamination in samples from the Hudson River valley and
from the Lake Ontario region (Foley et al., 1987). Livers of Hudson River mink
contained about 20 ppm of PCB (95% confidence limits, 10-40 ppm) on a lipid
basis, those from Lake Ontario mink contained about 14 ppm (95% confidence limits
4-44 ppm). Livers of otters from the Hudson River contained about 50 ppm (95%
confidence limits 30-100 ppm). No otters now seem to live near Lake Ontario, but
the cause of their absence is not known at this time (Foley, personal communication).
These results suggest that populations of piscivorous mammals in these two locations
in New York State are at risk. In addition, the data on levels of PCB in otters and
mink from remote areas such as the Adirondacks indicated that some individuals
were at risk and supported the idea that PCB could be deposited in remote locations
from long-range atmospheric transport.
The decline of the European otter (Lutra lutra) has recently been reviewed in
detail (Mason and Macdonald, 1986). Much of the work has been based on extensive
field observations of the presence or absence of characteristic scats (spraints) or
footprints in soft mud. The authors noted that comparable extensive studies have not
been undertaken on the North American river otter. European otters have recently
declined in much of their former range. However, populations persist throughout
Ireland and much of Scotland, in northern Scandinavia and a few remote regions of

Michael Gilbertson

the European continent (Reuther and Festetics, 1980). There is no doubt that habitat
destruction mainly related to agricultural 'improvements' has significantly contri-
buted to the declines but, in addition, organochlorine chemicals are implicated as
important factors.
The Swedish otter population declined steadily from 1950 to 1970, particularly in
southern Sweden (Erlinge, 1980). The declines could not be explained by such
factors as acidification, water regulation, disturbance, fishing, hunting or traffic
accidents (Olsson et al., 1981). Analysis of samples of muscle showed that otters
from southern Sweden had higher levels (180 ppm) of PCB in extractable fat than
in samples from northern Sweden (50 ppm) or from Norway (20 ppm). Based on the
results of mink reproductive studies (Jensen et al., 1977) it was concluded that the
otter population in southern Sweden is at significant risk and that even populations
in northern Sweden are at the threshold where incipient reproductive impairment
may be evident. There is no suggestion that otters in Norway are at risk or have
Chanin and Jefferies (1978), based on British records of otter hunting, reported a
marked decline in hunting success throughout England and southern Scotland start-
ing in 1957. They argued that there had been an extensive decline in the otter
population that coincided with the widespread introduction of dieldrin for use in
agriculture. The failure of the otter to increase after the 1975 cessation of dieldrin
use cast doubt on this hypothesis, particularly in the light of the recovery of British
birds of prey. Recent data on levels of organochlorine compounds in British otters
(Mason et al., 1986) have confirmed that levels of PCB and dieldrin are generally
higher in areas associated with declines. In remote parts of Britain, such as the
Orkney Isles and northern Scotland, most samples had undetectable or trace levels
of PCB and dieldrin. In Wales, where the population seems to be stable, levels are
elevated (5.3 ppm PCB, 6.8 ppm dieldrin in liver on a fresh weight basis) and this
population may be sufficiently contaminated to be experiencing incipient reproduc-
tive impairment. The 'fragmentary' population in East Anglia had the highest levels
(130 ppm PCB and 32 ppm dieldrin). Two muscle samples from the Devon and
Cornwall population, which is at the 'edge of the current range', had intermediate
levels (67 ppm PCB and 13 ppm dieldrin). In experimental feeding studies (Aulerich
and Ringer, 1970), mink were found to be able to tolerate a level of 2.5 ppm dieldrin
for about 6 months but showed nearly 100% mortality within nine months. Possibly
mobilization of lipid reserves associated with the natural stress of the onset of winter
resulted in the release of toxic levels of dieldrin. It is possible that the decline of the
British otter was initiated by dieldrin causing adult mortality but that the added
presence of high levels of PCB has resulted in continuing regional extirpation. Recent
post-mortem observations (Mason, personal communication) suggest that British
otters have pathological signs resembling those seen in Baltic Sea seals (Bergmann
and Olsson, 1985) exposed to high levels of organochlorine compounds.

Effects on fish and wildlife populations

Effects on fish and wildlife populations

Michael Gilbertson

4.2.3. Fish
Fish have proved to be extremely difficult organisms on which to investigate pollu-
tant-associated epizootiology. There are several recent reviews of this topic (Kraybill
et al., 1977; Dawe et al., 1976; Sindermann, 1983; Sindermann et al., 1980). The
concept of stress has become a central idea in fisheries pathobiology and much of the
research has been oriented to investigating sources of stress and the ways in which
aquatic animals respond to stress. Pollutants may have a direct effect on organisms
through cytotoxic, mutagenic or teratogenic action or have indirect interactive ef-
fects with infectious microorganisms through activation of pathogens, immuno-
suppression or alteration of the integumentary system (Sindermann, 1983). Table
4.1 is a compilation of various outbreaks of disease in fish populations that were
believed to be related directly or indirectly to PCB or other related compounds.
Sindermann (1979) set out six epizootiological criteria that were needed to make a
firm association between a disease in fish and an environmental pollutant. These
criteria, which are essentially analogous to the criteria used by epidemiologists
(Susset, 1986), included:
(1) knowledge of the history of occurrence of the disease in a particular species in
thd geographic area of concern;
(2) knowledge of the history of occurrence and levels of particular pollutants in that
(J\ area;
(3) a review of the biology, life history, and occurrence of the disease in other areas.
in other species, and under different environmental conditions;
(4) an intensive baseline survey of the current disease and pollution situation, with
attention to statistical reliability of sampling;
(5) laboratory and field experimentation with the principal objective of reproducing
the disease by exposure to known levels of contaminants; and
(6) resurveys of the disease and pollution levels over several years, looking for
changes or trends.
He noted that these requirements had been fully satisfied for few if any of the fish
diseases and concluded that, at that time, no causal relationship between a disease
and a specific contaminant could be definitively demonstrated and that the evidence
for associations was largely circumstantial. Sindermann et al. (1980) recommended
that, in monitoring the effects of pollutants on marine organisms, an epidemiological
approach should be employed together with concurrent experimental work.
During the 1980s considerable advances have occurred in fisheries-pollutant etiol-
ogy and a few case histories are beginning to be compiled that meet many of the
above criteria. Some of the instances of tumor epizootics appear to be related to the
presence of polynuclear aromatic hydrocarbons or other carcinogens such as carba-
zoles rather than to organochlorine compounds. These include the outbreak of hepa-
tocelhlar carcinoma in Pleuronectidae in Puget Sound (Roubal and Malins, 1985)
and o skin and lip tumors in brown bullheads (Ictalurus nebulosus) in the Buffalo
River NY, (Black, 1983) and skin and liver tumors in brown bullheads from the

Outbreaks of disease in fish thought to have been associated with PCB or related compounds (after
Sindermann t at., 1980)

Location (year) Species Description of Putative cause Reference

Northeast Irish Sea, Plaice Lymphocystis. PCB dumping into Perkins
U.K. (1971) (Plerwontcts epidermal ulcers, the Irish Sea et al., 1972
ploattss), fin damage

Northeast Irish Sea. Flounder Lymphocystis. Disputed the claim Shelton and
U.K. (1972) (Plalichlhys epidermal ulcers, that PCB might Wilson, 1973
flessus) fin damage be involved

Pales Verdes, CA, Dover sole Fin erosion (30%) Chlorinated Sherwood and
U.S.A. (Microusomus hydrocarbons. Mearns. 1977
(1972-1976) pacficus) DDT and PCB

San Francisco Bay. Starry flounder Embryonic PCB Spies et al., 1985
CA. U.S.A. (Plaoichrhys mortality

Hudson River Atlantic tomcod Hepatocellular PCB Smith et al.. 1979;
estuary. NY, (Microladus carcinoma Klauda et al.. 1981
U.S.A. (1977-78) toncod) (25%)

Hudson River. Striped bass Impaired vertebral PCB induced Mehrle et al.. 1982;
Potomac River, (Morone mechanical changes in Hickey and Young.
Nanticoke River. saxadllis) properties: early collagen: mineral 1984
U.S.A. warning of likely ratio caused by
breakage and impaired
damage. Vitamin C
Pugheadedness, metabolism
lordosis. scoliosis

Gullmar Fjord. Atlantic hagrish Hepatocellular Decline due to Falkmer et at..
Sweden (Myxint carcinoma decline legislative 1976. 1977
(1972-76) glutinosa) from .8% to restriction on
0.6% PCB

Puget sound, WA. English sole Fin erosion. PCB Wellings et al..
U.S.A. (1974-76) (Paraphrys hepatocellular 1976; McCainet
velutus) carcinoma at.. 1977
Starry flounder


Effects on fish and wildlife populations

TABLE 4t (continued)

Location (year) Species Description of Putative cause Reference

.alifax Nova Atlantic cod Fatty degeneration PCB Freeman et al.,

of hepatocytes

Baltic. Sweden Atlantic salmon 'Fatty livers'.
(1965-70) (Selmo saler) embryonic

Baltic between
Germany and
Denmark (?)

Travemunde. Baltic.
Germany (1979)

Lower Rline,
t3 (1976)

Lake Michigan. MI



Rainbow trout

Lake trout

vitrum virreum)

Decrease in
viability of hatch

Decrease in viable

retardation, liver
and kidney
haemoglobin and
blood glucose

embryonic and
fry mortality

embryonic and
larval mortality

PCB and possibly
other undetected

PCB and DDT and
probably other

and polycyclic

PCB, DDE and
other undetected


Johansson et al.

Von Westernhagen
et al., 1981

Hansen et al.. 1985

Pels et al.. 1980

Scotia. Canada

Willford et al..
1981:Macet al..

Organochlorine Hokanson and
compounds with Lien, 1986

Black River, OH (Baumann et al., 1982) and of gonadal tumors in goldfishxcarp
hybrids (Carassius auratus x Cyprinus carpio) (Black et al., 1981).
Two teams of research scientists, working on Great Lakes fish, have begun to relate
various pathologies to organochlorine pollutants. In an exemplary series of epizo-
otiolotical and laboratory experimental studies, Sonstegard and Leatherland (re-
viewed 1984) investigated the etiology of thyroid hyperplasia in coho salmon (Oncor-

(Godus morhua)

TABLE 4.1 (continued)

Location (year) Species Description of Putative cause Reference

Lake Erie. Coho salmon Thyroid PCB Sonstegard and
Lake Michigan. (Onorhynchus hyperplasia. Mirex Leatherland.
Lake Ontario, klsutch) growth DDT and dieldrin 1976:
Ontario. Canada inhibition, liver Leatherland and
(1974-76) enlargement and Sonstelard.
enzyme 1982a: Moccia
induction, lipid at al.. 1977

Lake Ontario. White sucker Lip papilloma Viruses mediated Sonstegard. 1977
Ontario. Canada (Catostomus (30%) by PAH
commersont) perchlorinated
and heavy metals

Lake Erie. Goldfish x carp Gonadal tumor PCB and/or DDT Sonstegard. 1977
Lake Michigan. hybrids;
Lake Huron, (Ceresslus
Lake Ontario auraus x
Cyprinus carpio)

hynchus kisutch) and chinook salmon (Oncorhynchus tshawytscha) that were stocked
in the Great Lakes. Iodine did not appear to be involved since, in interlake studies,
there was no relationship between the degree of thyroid hyperplasia and the iodine
content of the lakes, and blood serum levels of thyroxine and triiodothyronine were
high (Moccia et al., 1981). Coho salmon, from various Great Lakes, fed to coho
salmon and to rats (Leatherland and Sonstegard, 1982 a,b) elicited thyroid hyperpla-
sia showing the presence of a substances) with goitrogenic activity. The degree of
thyroid hyperplasia in the rats was correlated with the levels of organochlorine
compounds in the flesh of the coho salmon. Certain organochlorine compounds are
known to be goitrogenic. However feeding PCB and/or mirex to rainbow trout
(Salmo gairdneri) (Leatherland and Sonstegard, 1979) and coho salmon (Leather-
land and Sonstegard, 1978) failed to elicit thyroid hyperplasia despite high dosage
rates. The identity of the Great Lakes goitrogens has yet to be established. Other
lesions, with a possible organochlorine etiology, in coho salmon from some of the
Great Lakes included growth inhibition, liver enlargement and dysfunction of the
lipid mobilization and osmoregulatory system (Sonstegard and Leatherland, 1984).

Michael Gilbertson

Effects on fish and wildlife populations

Michael Gilbertson

The second series of studies (Stauffer, 1979; Willford ct al., 1981) was undertaken
on Lake Michigan lake trout to investigate the reproductive failure of stocks that
had been introduced to rehabilitate the fishery after the depredations of the sea
lamprey (Petromyzon marinus). In an extensive coordinated study (Mac et al.,
1985), artificially fertilized eggs from a hatchery broodstock and from Lake Michi-
gan, Lake Huron and Lake Superior feral stocks were artificially incubated in Lake
Michigan. Lake Huron and Lake Superior waters and in well water. These experi-
ments were undertaken to investigate whether the reproductive failure was caused
by an intrinsic or extrinsic factor acting on the eggs or whether there was an
interactive effect. It was concluded that eggs from Lake Michigan had an intrinsic
factor that influenced hatching success and fry survival and that incubation of eggs
in Lake Michigan water had a further effect on hatchability indicating the presence
of an extrinsic factor. Comparison of the levels of DDE (1.3 ppm) and PCB (2.7
ppm) in the Lake Michigan eggs with levels in other studies of the influence of these
compounds on hatchability suggested that these compounds alone could not account
for the observed embryonic mortality. It was concluded that there was probably
another embryotoxic factors) in the lake; however, this factor has not yet been

4.2.4. Birds
The effects of PCB, dioxins and furans on birds have been the subject of extensive
experimental work (reviewed in Peakall, 1987). In poultry operations, these com-
pounds hhave caused outbreaks of chick edema disease characterized by high mortal-
ity, edema, liver necrosis and porphyria and been responsible for substantial econ-
omic losses (reviewed in Firestone, 1973).
Extensive field studies have been undertaken on a variety of wild bird species to
investigate whether PCB, dioxins and furans have caused reproductive dysfunction
or other lesions. As with other fish and wildlife studies there has been a continuing
problem in attempts to separate the effects of DDE from the effects of other organ-
ochlorine compounds, particularly since levels of organochlorine compounds tend to
be co-correlated. In fact, the literature on the effects of organochlorine compounds
on wild birds is dominated by papers on the eggshell thinning phenomenon caused
by DDT and particularly its metabolite DDE (Cooke, 1973). Falconiformes and
pelicaniformes are particularly susceptible (Anderson and Hickey, 1972) and dec-
lines and" recoveries of populations of North American peregrines, Connecticut and
Long Island ospreys and South Carolina and California brown pelicans (Pelecanus
occidentalis) have been related to variations in the levels of DDE (reviewed in
Newton, 1979). As already mentioned, another pesticide, dieldrin, was responsible
for the 1960s population declines of the Scottish golden eagle (Lockie et al., 1969),
British peregrine falcon and sparrow hawk, though DDE contributed to low reprod-
uctive success through changes in eggshell thickness (reviewed in Newton, 1979).
Organochlorine compounds were implicated in the 1965-1966 mortality of white-

tailed sea eagles (Haliaeetus albicilla) in the archipelago of Stockholm in the Baltic
(anon, 1966; Jensen et al., 1969, 1972). The mean levels of DDT and PCB in the
brain were 100 ppm and 47 ppm respectively on a fresh weight basis (1900 ppm
DDT and 910 ppm PCB on a lipid basis). More recently, organochlorine compounds
have been implicated in the 1981-1982 mortality of great horned owls (Bubo
virginianus) from near the Hudson River (357 ppm PCB in brain) and near Lake
Ontario (117 ppm DDE, 2.85 ppm dieldrin, 65.4 ppm PCB in brain) (Stone and
Okoniewski, 1983).
Over 15 000 guillemots (Uria aalge) died in the Irish Sea in the autumn of 1969.
This abnormal mortality, which became known as 'the sea bird wreck', was thought
to have been related to the high levels of PCB found in the livers of dead guillemots.
After a thorough investigation (Holdgate, 1971) it was concluded that, because the
total quantity of PCB in apparently healthy guillemots was about the same as the
quantity in the dead guillemots, PCB could not be a cause of death. Subsequent,
more extensive analyses showed that the total amounts of PCB (arithmetic mean 5.5
mg) in guillemots that died were in fact about twice as high as the amounts (2.9 mg)
in healthy guillemots shot from the same area. Thus it was concluded that, indeed,
PCB was implicated in the deaths of the guillemots, though this was likely a secon-
dary cause that exacerbated the primary natural stresses of moulting, storms and
perhaps food availability (Parslow and Jefferies, 1973).
Much the same reasoning was used in concluding that high levels of PCB (25-39
mg) were a secondary contributory cause of mortality of gannets (Sula bassana) in
1972 in the Irish Sea (Parslow et al., 1973). There is the possibility that a parallel
case occurred in common murres (guillemots) in 1969 near the breeding ground off
Oregon, USA, when an estimated 51 000 birds died (Scott et al., 1975).
One of the first indications that PCB, dioxins or furans might be exerting a subtle
effect in wild bird populations was the finding of abnormal chicks in a colony of
common (Sterna hirundo) and roseate terns (Sterna dougalhii) at the end of Long
Island Sound in 1969 and 1970 (Hayes and Risebrough, 1972). The abnormalities,
which included eye, bill and foot deformities in newly hatched chicks and feather
loss in juveniles, resembled those produced experimentally by dioxins and PCB
(reviewed in Firestone, 1973).
There have been several epizootics among a variety of colonial fish-eating birds in
the Great Lakes basin since about 1960 (reviewed in Gilbertson, 1988; Fox and
Weseloh, 1987). Some of these, particularly eggshell thinning and feminization of
male embryos, were probably primarily caused by DDT and its metabolites. The
search for the causes of other lesions, including embryotoxicity, teratogenicity, goitre,
porphyria and liver enzyme induction, have posed difficult problems for etiological
research but there is a growing body of evidence implicating PCB, dioxins and furans
and possibly other polyhalogenated hydrocarbons (see Table 4.2).
Two outbreaks of chick edema disease have been documented in colonial fish-
eating bird populations in the Great Lakes basin. Serious reproductive failure in

115 116


Effects on fis and wildlife populations

Michael Gilbertsor

Epizootics and associated phenomena in Great Lakes fish-eating birds thought to have been caused by PCB,
dioxins and furans

Location (year of Species Description of Putative cause Reference
observations) disease

Lake Ontario Herring gull Embryonic 2,3.7,8-TCDD and Gilbertson and
(1966-1976) (Lanrs mortality, possibly other Fox. 1977;
artentatus) congenital chick edema Gilbertson, 1983
abnormalities, active
subcutaneous compounds
edema, growth
liver necrosis.
aberrant nesting Fox et at.. 1978

Lake Ontario. Black-crowned Congenital PCB Gilbertson et al.,
Lake Erie. night-heron anomalies: 1976
Detroit River (Nyeticorox crossed and
(1971-1974) n)yclcorx); twisted bills.
Ring-billed gull slipped tendons.
(Larus supernumerary
delaowrensis); toes at
Common tern tarsometatarsal
(Sterna joint, heart
hirundo): deformity
Caspian tern
Herring gull

Lake Huron. Herring gull Hepatic aryl Correlation with Ellenton et al..
LakeOntario(1981) (Laru srgenratus) hydrocarbon 2.3.7,8-TCDD 1985
induction in

Icunw, wi ri.

TABLE 4.2 (continued)

Location (year of Species Description of Putative cause Reference
observations) disease

Great Lakes. Herring pull Decrease in storage 2.3.7,8-TCDD- Spear et at.. 1986
Lake Superior. (Larun of vitamin A in induced increased
Lake Michigan. *rgrenrte ) livers of adults rate of vitamin A
Lake Ontario (1982) metabolism

Great Lakes. Herring gull Thyroid goitre in PCB and other Moccia et al.. 1986
Lake Ontario, (Larus adults goitrogenic
Lake Erie. argenatus) chlorine
Lake Huron. compounds.
Lake Superior. including DDT
Lake Michigan and dieldrin

Lake Michigan. Herring gull Porphyria in liven Correlation with Fox et al.. 1988
Lake Huron (Larun of adults total dioxins
Lake Ontario argentatus) substituted in
(1980-1985) 2,3.7.8 positions
and with PCB
but not with

Green Bay. Lake Foster's tern Embryonic Possibly PCB and/ Hoffman et al..
Michigan (Sternafortrt) mortality, or 2.3,7.8-TCDD 1987
(1973-present) growth and other dioxins
hepatic aryl

Green Bay. Lake Herring gull Congenital Toxic Substances
Michigan(1983) (Larui abnormalities. Task Force, 1983
rgentMats); crossed bills
Common tern
Double crested
Virginia rail

Effects on fish and wildlife populations 119

TABLE 4.2 (continued)

Location (year of Species Description of Putative cause Reference
observations) disease

Green Bay.
Lake Michigan

4/ Southern Ontario


Forster's tern
(Sterna forsteri)

Ring-billed gull

Extrinsic factors:

Spasmodic muscle
tremors, weak
and emaciated.
hepatic fatty
hemorrhage'.4', and

Kubiak et al., 1989

PCB in conjunction Sileo etal., 1977
with DDE and
possibility of
furns and

Lake Ontario herring gulls (Larus argentatus) was noted in 1966 (reviewed in
Gilbertson, 1983). Extensive field studies on various colonial fish-eating birds showed
that there was a high incidence of embryonic mortality, congenital abnormalities and
lossof eggs. In artificial incubation studies undertaken in 1974, herring gull embryos
an' chicks from Lake Ontario colonies exhibited signs of chick edema disease and
subsequent analysis revealed the presence of over 1 ppb of 2,3,7,8-tetrachlorodi-
bento-p-dioxin (Gilbertson, 1983). It seems likely that the presence of such a high
level of this potent chick edema factor was the major contributory factor to this
outbreak of chick edema disease. However, further isomer-specific analysis of deep-
frozen herring gull eggs may reveal the presence of other chick edema active com-
pounds. Reproductive success improved about 1976 and has been normal since about
1977 (Mineau et al., 1984).
The second outbreak has been occurring in Green Bay, WI, since about 1973
(reviewed in Kubiak, 1987). Impaired reproduction in Forster's terns (Sterna for-
steri) was investigated using techniques developed for the research on Lake Ontario
herring gulls. Isomer-specific analysis revealed the presence of a variety of chick

Michael Gilbertson

edema active compounds. The potencies of many chick edema active compounds
relative to 2,3,7,8-TCDD have been evaluated (Bradlaw et al., 1980: Bandiera et al.,
1984; Sawyer and Safe, 1982; Mason et al., 1986). Thus transformation of the data
on levels of the chick edema active compounds found in the Forster's tern eggs to an
equivalent value relative to 2,3,7,8-TCDD yielded a measure of the overall toxicity
and of the relative contribution of each chick edema active compound present. Over
90% of the overall 2,3,7,8-TCDD equivalents was contributed by 2,3,4.3',4'-pen-
tachlorobiphenyl and by 3,4,5,3',4'-pentachlorobiphenyl. In relation to all the case
histories reviewed in this paper this piece of research provides a landmark since, for
the first time, it relates the relative contributions of the presence of a variety of
compounds with a specific mode of action to an ongoing epizootic. This is a com-
mendable achievement in collaborative epizootiology and analytical chemistry.
Application of the Great Lakes protocol to eggs from a black-crowned night heron
colony in San Francisco Bay National Wildlife Refuge revealed (Hoffman et al..
1986) that chicks at hatching weighed 15% less than chicks from a colony at Patuxent
Wildlife Research Centre. The embryo weights were inversely correlated with the
PCB content of the eggs.
Organochlorine residue analysis usually reveals the presence of multiple contami-
nants, and the interpretation of the residue levels in relation to reproductive effects
or other lesions has proved extremely difficult, particularly when the compounds are
co-correlated. An excellent case has been made for the effect of both DDE and PCB
in the reproduction of two congeneric species; the hold eugle in North America and
the white-tailed sea eagle in Europe and Greenland (Helander et al., 1982). The
authors combined data on productivity and the concentration of organochlorine and
mercury compounds of twelve populations (six of each species) and concluded that a
combination of both DDE and PCB was the main factor affecting reproductive
success. Highest levels of contamination and lowest productivity were seen in bald
eagle populations near Lake Superior and in white-tailed sea eagle populations
around the Baltic and in Schleswig-Holstein, West Germany. Results of studies on
bald eagles in the lower Columbia River (Robert Anthony, personal communication)
demonstrate reduced productivity (0.04 young/occupied nest) and elevated levels of
DDE (9.3 ppm) and PCB (12.9 ppm) in eggs and are consistent with the interpreta-
tion of Hellander et al. (1982). Bald eagle reintroductions are being undertaken in
upper New York State and near Long Point, Lake Erie, to replace populations
extirpated, presumably by organochlorine pollutants. Bald eagle predation on con-
taminated Lake Superior herring gulls is hindering reestablishment of the population
in the Apostle Islands through mortality of adults and effects on reproduction
(Kozie, 1986).
A similar role of both DDE and PCB was found in relation to breeding success of
sparrow hawks in northern England and southern Scotland between 1971 and 1974
(Newton and Bogan, 1978). While high levels of DDE were found to be significantly
correlated with eggshell thinning, egg breakage, reduced hatching success and ad-

Effects on fish and wildlife populations

Michael Gilbertson

dling (no embryonic development), high levels of PCB were found to be related not
only to reduced hatching but more particularly to addling. This implied an embry-
otoxic role of high levels of PCB and was consistent with experimental findings of
the effects of PCB on the avian embryo (reviewed in Peakall, 1987). Further field
work to 1980 and further statistical analysis, however, did not support a relationship
between PCB and brood size (Newton et al., 1986).

4.3. Discussion

The above evidence of chemically-induced epizootics is an important part of the
rationale for maintaining a strong international stand on the prohibition of manufac-
ture, use or disposal of certain persistent organochlorine compounds. There seems to
be a pervasive feeling among the electrical and agricultural industries and consumer
advocates that organochlorine chemicals do not pose as serious a hazard to human
health as was initially thought and that regulatory controls should be relaxed (McGraw,
1983; Tschirley, 1986; ACSH, 1985). The epizootiological evidence of chemically-
induced declines of fish and wildlife populations, characterized by reproductive
failure, morphological lesions and even adult mortality, is thus likely to become
increasingly important in maintenance of controls. Thus there is a priority require-
ment for epizootiological research and for the production of scientifically defensible
evidence of the highest quality. A second reason for producing this evidence is that
organochlorine production, including PCB, is increasing in third-world countries and
thus the toxicological implications for their fish and wildlife populations are predict-
One of the recurrent themes in these case histories is that all relate to the effects
of contamination of aquatic food webs. The case histories almost all come from the
following four enclosed locations characterized by low rates of water exchange: the
Baltic, the Wadden Sea, Puget Sound and the Great Lakes Basin including the
St-Lawrence Estuary. There are no case histories that come from the Mediterranean
Sea, from the Southern Hemisphere or Eastern Europe though it is still a moot point
whether this is because there are no epizootics or because there are no scientists
working on these aspects of species that are likely to be affected.
A second recurrent theme is the species and congeners at risk. Toothed marine
mammals were clearly more at risk than planktivorous mammals, and several species
such as the common seal, common porpoise and beluga whale have holarctic distri-
bution, making toxicological interpretations relatively more direct. Similarly pisci-
vorous freshwater mammals such as mink and otter were more at risk than herbivo-
rous mammals such as beaver or racoon. Extensive work has been undertaken to
document the effects on piscivorous bird species at risk such as the herring gull and
Forster's tern and congeneric species such as the bald eagle in North America and
the white-tailed sea eagle in Greenland and Europe. The case histories tend to

suggest that epizootiological studies of fish are much more difficult to undertake and
that it is difficult to identify species at risk.
It is disappointing to see how few case histories have been successfully completed
in which an observed lesion or population decline has been reliably related with an
inferred cause. The epizootiological criteria set out by Sindermann (1979) for fish
epizootics are analogous to the criteria used in human epidemiology to infer causality
(Susser, 1986). These various lines of research have been used to relate avian eggshell
thinning with DDE (reviewed in Newton, 1979). However, their systematic use to
plan and execute epizootiological and etiological research and to discuss case his-
tories is not evident in the scientific literature on other organochlorine compounds
for other classes of vertebrates.
The logistical difficulties in undertaking epizootiological research on fish may be
the reason that fisheries biologists have produced a disproportionate amount of
laboratory experimental work compared with bird and mammal biologists. The latter
tend to start with an observation of an anomaly in the field and, while documenting
the geographic and temporal variation, initiate etiological research on the likely
cause. Fisheries biologists tend to start with the observation of a chemical in the
environment, initiate experimental work to investigate the effects on laboratory
species and then to see whether the same phenomena occur in the environment. This
approach, however, has not produced very satisfactory case histories. The highly
successful work by Malins and co-workers on Pleuronectidae in Puget Sound, by
Sonstegard and co-workers on Salmonidae in the Lower Great Lakes and by Mac
and co-workers on Salmonidae in Lake Michigan are notable exceptions. The results
of these studies is that these teams of researchers are close to solving some of the
complex toxicological riddles posed by discharges of chemicals to the environment,
It seems that only through collaborative research undertaken by field biologists,
analytical chemists and laboratory biologists and biochemists can a scientifically
defensible case history be compiled on which rational regulatory decisions can be
One of the recurrent problems in the interpretation of the case histories is the
exact identity and quantitation of the toxic agents) present. In only one case
(Kubiak et al., 1989) was there an exhaustive isomer-specific analysis of PCB,
dioxins and furans and thus this is the only case in which a relatively comprehensive
explanation of the epizootic can be given. Similar isomer-specific analyses should be
undertaken retrospectively on Lake Ontario herring gulls, Baltic ringed and common
seals, Wadden Sea common seals and St-Lawrence beluga whales to see whether the
presence of specific compounds can better explain the observed pathologies. Clearly,
the appellation'PCB' or 'T, dioxins' will not suffice to interpret epizootics.
Finally there is a need to apply bioassay techniques to investigate total activity as
'2,3,7,8-TCDD equivalents' in a sample and the relative contribution of identifiable
components. Several bioassay techniques have been developed, including aryl hydro-
carbon hydroxylase activity, ethoxyresorufin O-deethylase induction and in vitro


Effects on fish and wildlife populations

keratinization and their relationship to whole body responses such as weight loss and
thymic atrophy (reviewed in Safe, 1987). Only through this means will it be possible
to account for overall toxicity and possible interactions of these groups of chemicals
with a common mode of action, and ensure that other as yet unidentified compounds,
with the same mode of action, do not go undetected.


I should like to acknowledge the following for their help in directing me to the
various case histories reviewed in this paper: R.F. Addison, R. Anthony, P. Bdland,
V. Cairns, J.E. Elliott, R.F. Foley, G.A. Fox, C.J. Henny, K.E.F. Hokanson, A.V.
Holden, T.J. Kubiak, K. Marshall, C.F. Mason, G.B. McCullough, C. Moise, D.B.
Peakall, G. Proulx, D.E. Sergeant, H. Shear, J.D. Somers, J.F. Uthe, D.V. Weseloh,
W.A. Willford, V. Zitko. I am grateful to Dr. J.A. Buccini for giving me the freedom
from administrative duties to write this paper and to Jeannette Lacroix for word


ACSH (1985) Report by the American Council on Science and Health. 47 Maple St.. Summit. NJ 07901.
Aguilar. A. (1985) Residue Rev. 95. 91 14.
Almkvist. L. (1982) ICES.C.M. 1982/N:16.
Alricu.C. and Duguy. R.(1979) Oceanog. Acta 2. 107-120.
Anderson. D.W. and Hickey. J.J. (1972) Proc. 15th Int. Ornithol. Congr. 514-540.
Anon (1966) New Sci. 612.
Aulerich. R.J. and Ringer. R.K. (1970) Am. Fur. Breeder 43, 10-11.
Aulerich. R.J. and Ringer. R.K. (1977) Arch. Environ. Contam. Toxicol. 6.279-292.
Bandiera. S.. Sawyer. T.. Romkes, M.. Zmudzka. B.. Safe, L., Mason. G.. Keys, B. and Safe. S. (1984)
Toxicology 32. 131-144.
Baumann. P.C.. Smith, WD. and Ribick. M. (1982) in Polynuclear Aromatic Hydrocarbons: Physical and
Biological Fate (Cook, M., Dennis. A.J. and Fisher, G.L.. eds.), pp. 93-102. Battelle Press. Columbus,
Bland. P. and Martineau. D. (1986) Interface. Nov/Dec, 20-25.
Bergman. A. and Olsson. M. (1985) Finnish Game Res. 44, 47-62.
Bergman. A.. Olsson. M. and Reutergaard, L. (1981) ICES. CM. 1981/N:10.
Black. J.J. (1983) J. Great Lakes Res. 9. 326-334.
Black. J.J.. Dymerski, P.P. and Zapisek. W.F. (1981) ASTM. STP 737, 215-225.
Bowes, G.W.. Mulvihill, M.J., Simoncit, B.R.T., Burlingame, A.L. and Risebrough, R.W. (1975) Nature
Bradlaw. J.A.. Garhoff. L.H. and Hurley, N.E. (1980) Food Cosnet. Toxicol. 18, 627-635.
Burdick. G.E.. Harris, E.J.. Dean H.J.. Walker. T.M., Skea, J. and Colby. D. (1964) Trans. Am. Fish: Soc.

Michael Gilbertson

Calambokidis. J., Speich, S.M., Pcard, J., Steiger. G.H. and Cubbage, J.C. (1985) NOAA Tech. Memo.
Chanin. P.R.F. and Jefferies. D.J. (1978) Biol. J. Linn. Soc. 10, 305-328.
Cooke, A.S. (1973) Environ. Pollut. 4.85-152.
Darnerud. P.O.. Brandt. I.. Klasson-Wehler. E., Bergman. A., dArgy. R., Dencker. L. and Sperber. G.O.
(1986) Xenobiotica 16. 295-306.
Dawe. C.J., Scarpelli, D.G. and Wellings. S.R. (Eds.) (1976) Prog. Exp. Tumor Res. Vol. 20, 438pp. Karger.
New York.
Drescher. H.E. (1978) ICES. CM. 1978/N:12.
DeLong. R.L. Gilmartin. W.G. and Simpson J.G. (1973) Science 181. 1168-1170.
Ellenton. J.A., Brownlee, L.J. and Hollebone. B.R. (1985) Environ. Toxicol. Chem. 4.615-622.
Erlinger.S. (1980) in Der Fischotter in Europa (Reuther, C. and Festetics. A., eds.). pp. 103-106.Oderhaus
und Gottingen.
Falkmer. S., Emdin. S.O.. Osthebrg Y., Mattisson, A., Johansson Sjobeck. M.-L. and Fange. R. (1976) Prog.
Exp. Tumor Res. 20, 217-250.
. Falkmer. S.. Marklund, S.. Maltson, P.E. and Rappe, C. (1977) Ann. N.Y. Acad. Sci. 298. 342-355.
Firestone. D. (1973) Environ. Health Perspect. 5, 59-66.
Foley. R.I.. Jackling. S.J., Sloan. R.J. and Brown. M. (198) J. Environ. Toxicol. Chem. 7, 363-374.
Fox. G.A.. Kennedy. S.W. Norstrom. R.J. and Wigficld. L. (1988) J. Environ. Toxicol. Chem. 7. 831-839.
Fox. G.A. and Weseloh. D.V. (1987) in The Value of Birds (Diamond. A.W. and Filion, F.. eds.). ICBP Tech.
Pub.. No. 6.
Fox. G.A.. Oilman. A.P.. Peakall. D.B. and Anderka. F.W (1978) J. Wildt. Manage. 42. 477-483.
Freeman, H.C., Sangalong. G.B., Uthe. J.F., Garside. E.T. and Daye. P.G. (1983) Arch. Environ. Contam.
Toxicol. 12, 627-632.
Gaskin. D.E. (1982) The Ecology of Whales and Dolphins, pp. 393-433. Heinemann. London.
Gaskin. D.E., Frank. R. and Holdrinet. M. (1983) Arch. Environ. Contam. Toxicol. 12.211-219.
Gaskin. D.E., Smith. G.J.D., Watson, A.P., Yasui, W.Y. and Yurick. D.B. (1984) Int. Whaling Comm.
(Special Issue 6). 135-148.
Gilbertson. M. (1983) Chemosphere 12. 357-370.
Gilbertson, M. (1988) in Toxic Contaminants and Ecosystem Health: A Great Lakes Focus (Evans. M.S.,
ed.), Adv. Environ. Sci. Technol.. Wiley, New York.
Gilbertson. M. and Fox. G.A. (1977) Environ. Pollut. 12,211-216.
Gilbertson. M.. Morris. R.D. and Hunter, R.A. (1976) Auk 93, 434-442.
Gilmartin, W.G.. DeLong. R.L.. Smith. A.W., Sweeney, J.C., DeLappe. B.W. Risebrough. R.W. Griner.
L.A., Dailey. M.D. and Peakall, D.B. (1976) J. Wildl. Dis. 12, 104-115.
Hansen, P.-D., von Westernhagen. H. and Rosenthal, H. (1985) Mar. Environ. Res. 15. 59-76.
Hartsough. G.R. (1965) Am. Fur Breeder. 38.25-27.
Hayes. H. and Risebrough. R.W. (1972) Auk 89. 19-35.
Helander. B., Olsson, M. and Reutergaardh, L. (1982) Holarctic Ecol. 5. 349-366.
Helle. E. (1980) Ann. Zool. Fennici 17. 147-158.
Helle. E. (1981) ICES. C.M. 1981/E:37.
Ilelle. F... Olsson. M. and Jensen, S. (1976) Ambio 5, 261 -263.
Hellc. R.. Hyvarinen. H., Pyysalo. H. and Wickstrom. K. (1983) Mar. Poll. Bull. 14. 256-260.
Henny. C.J., Blus. L.J., Gregory, S.V. and Stafford. C.J. (1981) in Worldwide Furbearer Conf. Proc.
(Chapman. J.A. and Pursley. D.. eds.). pp. 1763-1780. Frostburg. Maryland.
Hickey. C.R. and Young. B.H. (1984) N.Y. Fish Game J. 31, 104-108.
Hickey. J.J. (ed.)(1969) Peregrine Falcon Populations. Unix Wisconsin Press. WI. 596pp.
Hoffman. D.J., Rattner. B.A., Bunck, C.M. and Krynitsky, A. (1986) J. Toxicol. Environ. Health 19.
Hoffman. D.J.. Rattner. B.A.. Sileo. L.. Docherty. D. and Kubiak. T.J. (1987) Environ. Res. 42. 176-184.

125 126

Effects on fish and wildlife populations

Hokanson. K.E.F. and Lien. G.J. (1986) USEPA Internal Report, Duluth. MN 55804. USA.
Holden, A.V. (1975) Rapp. P.-v. Riun. Cons. Int. Explor. Mer. 169.353-361.
Holden. A.V. and Marsden. K. (1967) Nature 216, 1274-1276.
Holdgate. M.W. (ed.) (1971) The Seabird Wreck of 1969 in the Irish Sea. Nat. Environ. Res. Council.
London. 17pp.
Holmes. D.C.. Simmons, J.H. and Tatton, J. O'G (1967) Nature 216. 227-229.
Hunt. E.G. and Bischoff. A.I. (1960) Calif. Fish Game 46,91-106.
HIuschenbeth. E. (1977) Inf. Fischwirtsch 24. 162-164.
Jensen. S., Johnels. A.G., Olsson. M. and Otterlind. G. (1969) Nature 224. 247-250.
Jensen. S.. Johnels. A.G., Olsson. M. and Westermark, T. (1972) Proc. XV. Int. Ornithol. Congr. pp.
455-465. Brill. Leiden.
Jensen. S.. Kihlstrom. J.E.. Olsson. M.. Lundberg. C. and Orberg. J. (1977) Ambio 6.239.
Johansson. N.. Jensen. S. and Olsson. M. (1970) in PCB Conference I. pp. 58-68. Stockholm.
Kayes. R.J. (1985) The Decline of Porpoises and Dolphins in the Southern North Sea: A Current Status
Report. Report RR-14. Political Ecology Research Group Ltd.. Oxford. U.K., 109 pp.
Klauda. R.J., Peck. T.H. and Rice. G.K. (1981) Bull. Environ. Contain. Toxicol. 27, 829-835.
Koeman. J.H.. Ten Noever de Brauw. M.C. and DeVos. R.H. (1969) Nature 221. 1126-1128.
Koeman. J.H.. Peters. W.H.M.. Smit. C.J.. Tjioe. P.S. and Dc Goeij. J.J.M. (1972) TNO Nieuws 17,
Kozie. K.D. (1986) M.Sci. Thesis. Univ. of Wisconsin.
Kraybill. H.F.. Dawe, C.J., Harsbarger. J.C. and Tardiff, R.G. (eds.)(1977) Ann. N.Y. Acad. Sci. Vol. 298.
New York. New York. 604pp.
Kubiak. T.J., Harris. H.J.. Smith. L.M., Schwartz, T.R.. Stalling. D.L., Trick. J.A., Silco. L., Docherty,
D.E. and Erdman. T.C. (1989) Arch. Environ. Contam. Toxicol.. in press.
Leatherland. J.F. and Sonstegard, R.A. (1978) J. Fish Res. Bd. Can. 35. 1285-1289.
Leatherland. J.F. and Sonstegard. R.A. (1979) J. Fish Dis. 2.43-48.
Lectherland. J.F. and Sonstegard. R.A. (1982a) Comp. Biochem. Physiol. 72C, 91-99.
Leatherland. J.F. and Sonstegard. R.A. (1982b) Bull. Environ. Contam. Toxicol. 29.341-346.
Leatherland. J.F. and Sonstegard. R.A. (1984) Adv. Environ. Sci. Technol. 16, 115-149.
Lockie. J.D. and Ratcliffe, D.A. (1964) Brit. Birds 57. 89-101.
Lockie.J.D.. Ratcliffe. D.A. and Balharry. R. (1969) J. Appl. Ecol. 6.381-389.
Mac. M.J.. Edsall. C.D. and Seelye. J.G. (1985) J. Great Lakes Res. 11.520-529.
Martineau. D.. Bland. P.. Desjardins, C. and Lagact. A. (1987) Arch. Environ. Contam. Toxicol. 16.
Martineau. D.. Lagace. A., Bdland. P., Higgings, R., Armstrong. D. and Shugart. L.R. (in press) J. Comp.
Mason, C.F. and Macdonald. S.M. (1986) Otters: Ecology and Conservation. Cambridge Univ. Press.
Cambridge. 229pp.
Mason. C.F.. Ford. T.C. and Last. N.I. (1986) Bull. Environ. Contain. Toxicol. 36,656-661.
Mason. G., Farrell. K.. Keys. B.. Piskorska-Pliszczynska, J., Safe. L. and Safe. S. (1986) Toxicology 41.
McCa^n. B.D.. Pierce. K.V. Wellings. S.R. and Miller. B.S. (1977) Bull. Environ. Contam. Toxicol. 18. 1-2.
McGraw. M.G. (1983) Electrical World. February 1983: 49-72.
Mehrle. P.M.. Haines. T.A., Hamilton, S., Ludke, J.L.. Mayer, F.L. and Ribick. M.A. (1982) Trans. Am.
Fish Soc. 111, 231-241.
Mineau. P.. Fox, G.A., Norstrom, R.J., Weseloh, D.V.. Hallett, D.J. and Ellenton, J.A. (1984) in Toxic
Contaminants in the Great Lakes (Nriagu, J.O. and Simmons, M.S. eds.). pp. 425-452, Wiley, New
Moccia. R.D.. Leatherland. J.F. and Sonstegard. R.A. (1977) Science 198.425-426.
Moccia, R.D., Leatherland. J.F. and Sonstegard. R.A. (1981) Cancer Res 4,2200-2210.

Moccia. R.D., Fox. G.A. and Britton, A. (1986) J. Wildly. Dis. 22,60-70.
Newton, I. (1979) Population Ecology of Raptors, Buteo Books, Vermillion, S. Dakota. 399pp.
Newton. I. and Bogan, J. (1978) J. Apple. Ecol. 15, 105-116.
Newton, I., Bogan, J.A. and Rothery. P. (1986) J. Apple. Ecol. 23, 461-478.
Olsson. M. (1986) Fauna Flora 81. 155-156.
Olsson. M.. Reutergaardh. L. and Sandegren. F. (1981) Sveriges Natur 6/8. 234-240.
O'Shea. T.J., Kaiser, T.E., Askins, G.R. and Chapman. J.A. (1981) in Worldwide Furbearer Conf. Proc.
(Chapman, J.A. and Pursley, D., eds.), pp. 1746-1752. Frostburl, Maryland.
Otterlind, G. (1976) ICES. C.M. 1976/N:16.
Parslow, J.L.F. and Jefferies, D.J. (1973) Environ. Poll. 5, 87-101.
Parslow. J.L.F., Jefferies. D.J. and Hanson, H.M. (1973) Mar. Poll. Bull. 4.41-43.
Peakall, D.B. (1986) In PCBs and the Environment (Waid, J.S.. ed.). Vol. 2. pp. 31-47. CRC Press. Boca
Raton. FL.
Perkins, E.J., Gilchrist, J.R.S. and Abbott, O.J. (1972) Nature 238. 101-103.
Platonow, N.S. and Karstad, L.H. (1973) Can. J. Comp. Med. 37. 391-400.
Poels. C.L.M., van der Gaag. M.A. and van de Kerkhoff. J.F.J. (1980) Water Res. 14, 1029-1035.
Prest, I. (1965) Bird Study 12. 196-221.
Proulx. G., Weseloh. D.VC.. Elliott, J.E., Teeple. S., Anghern, P.A.M. and Mineau. P. (1985) Unpublished
manuscript, Canadian Wildlife Service.
Ratcliffe. D.A. (1963) Bird Study 10. 56-90.
Ratcliffe. D.A. (1967) Nature 215. 208-210.
Reeves. R.R. and Mitchell. E. (1984) Naturaliste Can. 11. 63-121.
Reijnders, P.J.H. (1976) Neth. J. Sea Res. 10, 223-235.
Reijnders. P.J.H. (1978) Neth. J. Sea Res. 12. 164-179.
Reijnders. P.J.H. (1980) Neth. J. Sea Res. 14, 30-65.
Reijnders. P.J.H. (1986) Nature 324.456-457.
Reijnders. P.J.H.. Clausen. B., van Haaften. J.L. and van der Kamp. J. (1982) in Marine Mammals of the
Wadden Sea (Reijnders. P.J.H. and Wolff. WJ.. eds.), pp. 7/33-37, Balkema, Rotterdam.
Reulher, C. and Festetics. A. (eds.) (1980) Der Fischotter in Europa. Oderhaus und Gottingen. 288pp.
Ringer, R.K. (1982) in Hazardous Waste Disposal: Assessing the Problem. (Highland. J.H., ed.), pp.
227-240. Ann Arbor Sci. Publi., Woburn, MA.
Risebrough. R.W. (1969) in Chemical Fallout (Miller. M.W. and Berg. G.G.. eds.). pp. 5-23. Thomas.
Springfield. IL.
Risebrough. R.W (1978) Mar. Mamm. Comm. Washington. DC NTIS PB-290. 64pp.
Risebrough. R.W.. Rieche. P.. Herman. S.G., Peakall. D.B. and Kirven. M.N. (1968) Nature 220.1098-1102.
Roubal. W.T. and Malins. D.C. (1985) Aquat. Toxicol. 6, 87-103.
Safe, S. (1987) Chemosphere 16,791-802.
Sawyer. T. and Safe. S. (1982) Toxicol. Lett. 13, 87-94.
Scott. J.M., Wiens, J.A. and Claeys. R.R. (1975) J. Wildl. Manage. 39. 310-320.
Sergeant. D.E. (1980) ICES. C.M. 1980/E:55.
Shelton. R.G.J. and Wilson. K.W. (1973) Aquaculture 2.395-410.
Sherwood, M.J. and Mearns. A.J. (1977) Ann. N.Y. Acad. Sci. 298.177 -189.
Silco. L.. Karstad, L.. Frank. R., Holdrinet. M.V.H., Addison. E. and Braun. H.E. (1977) J. Wildl. Dis. 13.
Sinderman, C.J. (1979) Fish. Bull. 76,717-749.
Sindermann. C.J. (1983) Rapp. P.-v. Reun. Cons. int. Explor. Mer. 182. 37-43.
Sindermann. C.J.. Bang, F.B.. Christensen. N.O.. Dethlefson. V.. Harshbarger. J.C.. Mitchell, J.R. and
Mulcahy. M.F. (1980) Rapp. P.-v. Rdun. Cons. int. Explor. Mer. 179, 135-151.
Smith. C.E.. Peck, T.H.. Klauda. R.J. and McLaren. J.B. (1979) J. Fish. Dis. 2,313-319.
Sonstegard. R.A. (1977) Ann. N.Y. Acad. Sci. 298. 261-269.

Michael Gilbertson

,. ,.ul. loxiul I'harmacol. 6. 116- 128.
loui: Substances Task Force (1983) Final Report on the Toxic Substances Task Force on the Lower Fox
River Systese' Dept. Nat. Res.. U.S. Geol. Surv. U.S. Fish Wildl. Serv., U.S. E.P.A.. U. Wisc.
Water Chem. lMadison, cooperating agencies).
Tschirley. F.H. (1986) Sci. Am. 254. 29-35.
van Haaften. J.L. (1982) in Marine Mammals of the Wadden Sea (Reijnders. V.J.H. and Wolff. W.J., eds.),
pp. 7/15-32. Balkema. Rotterdam.
Verwey. J. and Wolff. W.J. (1982a) in Marine Mammals of the Wadden Sea (Reijnders. P.J.H. and Wolff,
W.J.. eds.). pp. 7/51-58. Balkema. Rotterdam.
Verwey J. and Wolff. W.J. (1982b) in Marine Mammals of the Wadden Sea (Reijnders. P.J.H. and Wolff,
W.J.. eds.). pp. 7/59-64. Balkema. Rotterdam.
Von Westernhagen. H.. Rosenthal. H., Dethlefsen. V., Ernst, W. Harms. U. and Hansen, P.-D. (1981) Aquat.
Toxicol. 1.85-99.
Wagemann, R. and Muir. D.C.G. (1984) Can. Tech. Rep. Fish. Aquat. Sci. 1279. Winnipeg, Manitoba.
Wellings. S.R.. Alpers, C.E.. McCain. B.8. and Miller. B.S. (1976) J. Fish. Res. Bd. Can. 33. 2577- 2586.
Willford. WA.. Bergstedt. R.A.. Berlin. W.H.. Foster, N.R., Hesselberg. R.J., Mac. M.J., Passino, D.R.M.,
Re irrt. R.E. and Rotliers. D.V. (1981) U.S. Fish and Wildl. Serv. Tech. Papers 105, Washington. DC.

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