THE TOXICOLOGY, KINETICS AND METABOLISM OF PCBs IN FISHES,
WITH SPECIAL REFERENCE TO BLUEFISH, Pomatomus saltatrix
A Report to
U.S. Environemntal Protection Agency
Gulf Breeze Environmental Research Laboratory
Gulf Breeze, Florida
From
Joseph M. O'Connor
New York University Medical Center
A. J. Lanza Laboratories
Tuxedo Park, New York
Su bm i t ed: January 1986
Revised: July, 1986
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EXECUTIVE SUMMflRY
THE TOXICOLOGY, KINETICS OND METftBOLISM OF PCBs IN FISHES
WITH SPECIPL REFERENCE TO BLUEFISH, Pomatomus saltatrix
Polychlorinated biphenyls
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PCB in a fish depends upon the magnitude and frequency of exposure as well as
depuration of the compound, body burdens of PCBs in fishes may vary depending
upon where the exposure occurs. Fishes in contaminated estuarine and coastal
systems may accumulate high body burdens, whereas fishes from less contaminated,
oceanic waters may accumulate lower burdens. Sophisticated models aimed at
predicting PCB body burdens in fishes have been developed for a number of
different aquatic ecosystems.
The greatest amount of data on PCB accumulation, retention and metabolism in
fishes is based upon research with striped bass and rainbow trout. PCB data for
bluefish are limited to several monitoring and survey studies carried out to
ascertain PCB concentrations in commercial, recreational and scientific survay
^
catches of the species. Most such data are available from the states of New York
and New Jersey.
Bluefish from New York and New Jersey waters show wide variation in degree
of PCB contamination. Concentrations in estuarine waters of the Hudson River,
Newark Bay and Raritan Bay vary from below 1.0 part per million to greater than 5
parts per million. Samples taken in Massachusetts have shown PCBs as high as 16
parts per million in the edible flesh. Overall, bluefish from open ocean waters
tend to have lower concentrations of PCBs, while samples taken within estuaries
have higher levels. PCB contamination varies from year to year and site to site,
however, and detecting a clear trend from the sparse data is difficult.
Assuming that the physiology of bluefish and the kinetics of PCBs in
bluefish are similar to those of the striped bass and the rainbow trout,
predictions of bluefish PCB burdens may be made. Such predictions suggest that,
for bluefish feeding upon a PCB-contaminated diet in Atlantic coastal and
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estuarine waters, concentrations of from 0.1 to 11.0 parts per million PCBs may
be expected. Predicted concentrations would be affected most strongly by the
concentrations of PCBs in food organisms, since bluefish spend most of their life
cycle in waters with low concentrations of dissolved PCBs.
Like rainbow trout, bluefish probably do not metabolize PCBs to any
appreciable extent. Reductions in bluefish PCB body burdens that occur from
season to season are due to elimination of parent compound rather than metabolism
of the PCB to more polar metabolites. Although the data are sparse, they would
suggest that PCB burdens in bluefish will consist of PCB congeners with four or
more chlorine substitutions. PCBs with a lesser degree of chlorine substitution
will be eliminated. The data from studies with striped bass, rainbow trout and
other species suggest th'at the congeners likely to accumulate in bluefish will be
those with a high degree of chlorine substitution in ortho, ortho1 positions, and
will not be those identified as having a high potential for toxic effects.
More studies are required to determine patterns of PCB contamination in
bluefish, and to determine the potential toxicity to man from specific groups of
toxic congeners. Such studies are currently underway within a number of state and
federal regulatory agencies.
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THE TOXICOLOGY, KINETICS AND METABOLISM OF PCBs IN FISHES,
WITH SPECIAL REFERENCE TO BLUEFISH, Pomatomus saltatrix
INTRODUCTION
Genera1 Information Concerning Bluefish and "Similar Species"
This report is a summary of the accumulation, metabolism and effects of PCBs
in fishes, emphasizing bluefish (Pomatomus saltatrix) and other, similar, marine
or freshwater fish species. This report provides information on the problem of
PCBs in commercially and recreattonally important fish species, as well as a
scientific background for discussions related to developing, implementing and
enforcing regulations intended to deal with the problems of PCBs in marine
fishes, including bluefish. Any discussion of PCB dynamics in bluefish must be
qualified, in that the data for PCBs in bluefish are entirely monitoring data
describing PCB concentrations in recreational, commercial and scientific survey
catches of the species. We know of no experimental data providing information on
bioaccumulation or pharmacokinetics of PCBs in bluefish.
The bluefish of the world are described as a single species, Pomatomus
*
saltatrix. the only species in the family Pomatomidae. Bluefish occur in most of
the temperate coastal regions of the world, although they have been erroneously
reported as occurring in the Eastern Pacific (Briggs, 1960; Grosslein and
Azarovitz, 1982). Along the eastern coast of the U.S., bluefish occur in
continental shelf waters (Figure 1). Spawning occurs during two distinct periods:
<1) during the spring and suwmer, and (2) in the late fall in waters between the
continental slope and the coast (Breder and Rosen, 1968; Kendall and Waiford,
1979j Grosslein and Azarovitz, 1982). Spawning occurs in the open sea (Norcross
et al., 1974), and juveniles move from the open ocean into coastal waters and
estuaries during the mid- to late-summer months (Bigelow and Schroeder, 1953;
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PCBs in Bluefish
Page 2
Kendall and Walford, 1979). Bluefish populations of the Atlantic coast are highly
dependent upon estuaries as nurseries.
Bluefish are migratory, pelagic predators (Grosslein and flzarovitz, 1982),
acting as secondary or tertiary carnivores in the food web of coastal and
estuarine waters. They feed on a wide variety of fishes and invertebrates
(Bigelow and Schroeder, 1953). Stomach content analysis of young-of the year
bluefish in the Hudson estuary (O'Connor, personal observation) showed that they
feed on amphipods, small crabs and small fish, including young-of-the-year
striped bass (Morone saxatilis) and white perch (M. americana).
Bluefish are the primary recreational fish species in the waters of New York
and New Jersey. About £3,0(919 metric tons were taken in the New York Bight in 1970
(Deuel, 1973), and bluefish ranked first among recreational marine fishes in the
United States (Grosslein and Azarovitz, 1982). The charter-boat fishery for
bluefish is a multi-million dollar annual industry in New Jersey and New York. In
recent years this industry has suffered substantial losses because of reports of
PCB contamination in bluefish, and the subsequent public health advisories
concerning the consumption of bluefish (Belton et al., 1983, 1985).
Since there are so few data regarding the dynamics of PCBs in bluefish, the
inferences to be drawn in this report will be based upon data from other fish
species for which abundant data may be found. Me shall rely heavily upon data
from experiments with striped bass, a species which, like the bluefish,
undertakes migrations in the marine ecosystem, is dependent upon estuaries for
development of the young, and which is a secondary or tertiary carnivore in
marine and estuarine ecosystems (Bigelow and Schroeder, 1953; O'Connor, 1984a).
Like bluefish, striped bass contain large amounts of body lipid; depot fat is
stored intramuscularly as well as in mesenteric fat bodies. PCB exposure is
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PCBs in Bluefish
Page 3
likely to be similar for the two species, especially during the estuarine-
dependent, juvenile stages when their geographic distributions overlap and when
the two species are part of the same summer estuarine food web.
Additional information to be applied to the question of PCB dynamics in
bluefish will be data derived from studies with salmonid fishes; rainbow trout
(Salmo oairdneri) and lake trout (Salvelinus namaycush). Both species are
predators with a high proportion of body fat, and their responses to PCBs are
nell known from experimental and field studies. In fact, more is known about the
kinetics and metabolism of PCBs in rainbow trout than for any other fish species
(Lech and Peterson, 1383).
*
Data about PCBs in other species will be considered as a secondary source;
most PCB studies dealing with marine, freshwater or estuarine fishes other than
those listed above have been monitoring or survey studies, or studies providing
few data on PCB kinetics and dynamics.
This report follows a format in which the physical, physiological and
metabolic processes of PCB accumulation, retention and elimination are addressed
individually, along with a discussion of predictive models that are currently
available for use in determining PCB accumulation in fishes exposed to PCBs. It
is not intended to be exhaustive in detail of PCB accumulation, metabolism or
effects; rather, it is intended as an elucidation of those principles that must
be accounted for in any program assessing the problem of PCBs in fishes in
general and bluefish in particular.
General Facts Regarding Environmental Distribution of. PCBs
Polychlorinated biphenyls (PCBs) are substituted derivatives of the biphenyl
molecule (Figure 2), in which one or more hydrogen atoms have been replaced by
chlorine. PCBs comprise a class of 209 isomers of the chlorobiphenyl molecule
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PCBs in Bluefish
Page 4
each having unique chemical and physical characteristics
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PCBs in Bluefish
Page 5
Mas estimated at 500,000 metric tons (Nisbet and Sarofira, 1972; Hutzinger et al.,
1974). Worldwide PCB production was estimated to be double the U.S. production
(Richardson and Waid, 1982).
Ecological History and Significance of PCBs
Entry of PCBs into the environment occurred as the result of dissipative
uses, as well as from controlled and uncontrolled, closed systems (Hutzinger et
al., 1974; Nisbet and Sarofim, 1972; Richardson and Waid, 1962). Major routes of
entry of PCBs to the environment include vaporization, leaks and disposal of PCB-
contammated fluids, and disposal of PCB-containing products at dumps and in
landfills. 0 flow chart for evaluating various routes of PCB transport in the
global ecosystem is given in Figure 3 (from Nisbet and Sarofim, 1972).
The first report of PCBs in the environment appeared in 1966, when compounds
causing confounding peaks in chromatograms of DDT in fish samples were identified
as polychlorinated biphenyl (Jensen, 1966). Within a short time, the presence of
PCBs in all compartments of the global environment was established, and the
current distribution of PCBs in the environment may be said to be "ubiquitous"
(Risebrough et al., 1968; Koeman and Stasse-Wolthius, 1978; Wasserman et al.,
1979). Although some evidence exists for photodegradation and microbial
Metabolism of some PCB congeners, PCBs will most likely be present and will
recyle in the natural environment for many years (National Academy of Sciences
CNAS3), 1979).
Problems associated with PCBs in the environment are (1) the PCBs are known
to be acutely and chronically toxic to natural populations of animals, and (2)
animals used as food by the human population may serve as a vector for the
transport of PCBs from the environment to man (Hansen et al., 1971; Nimmo et al.,
1971a, 1971b; Hutzinger et al., 1974; Walker, 1976; Mayer et al., 1977; Wasserman
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PCBs in Bluefish
Page 6
et al., 1979; Belton et al., 1983, 1985). Evaluation of PCB transport in the
environment shows that the major route for PCB transport to man is the ingest ion
of finfish and shellfish caught in PCB-contaroinated systems (Nisbet and Sarofin,
1972; Jelinek and Corneiussen, 1976; Swain, 1983; Sloan et al., 1984; Belton et
al., 1985). Worldwide use of PCBs coupled with transport and recycling via the
atmosphere and surface waters has led to the present situation in which PCBs may
be found in virtually any environment, and PCBs may be accumulated in virtually
any species of finfish and shellfish used as food (Wasserman et al., 1979;
National Academy of Sciences (NAS), 1979; Richardson and Maid, 1988).
Despite laboratory evidence describing PCBs as highly toxic at low
«
concentrations (Hansen et al., 1971; Couch and Nimmo, 1974; Mayer et al., 1977;
Califano, 1981), there are few published data showing evidence of ecological
effects due to PCBs in natural systems. However, some studies provide evidence
that PCB effects in natural systems may be subtle and difficult to isolate from
the effects of other environmental contaminants. Mehrle et al. (198S) measured
several parameters of skeletal strength in striped bass from estuarine systems on
the east coast. They related weakness in vertebral columns to ambient
concentrations of PCB in the estuaries; Hudson River bass were found to have the
weakest vertebral columns, whereas bass from other systems were significantly
stronger. PCBs also have been found to induce mixed-function oxidase activity
(MFO; cytochrome P-448/P-450 system) in fishes (fiddison et al., 1978, 1979). In
complex environments subject to discharges of many different pollutants evidence
for increased MFO activity cannot be attributed to PCBs alone. Me know of no data
demonstrating that PCBs in natural environments are the direct cause of chronic
or acute toxicity, and we know of no data showing a relatonship between body
burden of PCBs and lesions in natural populations of animals.
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PCBs in Bluefish
Page 7
Lack of evidence demonstrating environmental or ecological impact by PCBs is
not proof that PCBs are toxicologically "benign." The tendency for aquatic and
terrestrial organisms to accumulate PCBs from environmental media and evidence
relating PCB exposure to definable lesions in animal tests (e.g. Lipsky et al.,
1378; Klaunig et al., 1979) dictates that the question of PCB environmental
impacts continue to be studied in depth. This is necessary so that any adverse
impacts that might occur because PCBs are present in natural environments can be
identified, characterized and, if possible, eliminated (NflS, 1979).
Toxicolooical History and Significance of PCBs
The PCBs are listed as animal carcinogens (IftRC, 1974, 1978), and as
•
hazardous materials, hazardous waste constituents and priority toxic pollutants
by the U.S. EPfl (Sittig, 1985). The tissues affected by PCBs are the skin
(chloracne), the eyes and the liver. PCBs also cause typical lesions of the
thyroid, stomach and lymphoid organs (Klaunig et al., 1979; Sleight, 1983). In
many cases the effects of PCBs on animal tissues are indistinguishable from those
caused by other chlorinated hydrocarbons such as DDT, dibenzodioxins and
dibenzofurans. Certain of the chlorobiphenyls may cause liver tumors in mice and
rats after prolonged exposure (IfiRC, 1974, 1978). Recent data show that the PCBs
function more as cancer promoters than as carcinogens (Kolbye and Carr, 1984),
and controversy still surrounds the interpretation of the original data used to
establish the carcinogenicity of the PCBs (Kimbrough et al., 1975).
Early data on PCB toxicology and pathology were published by Schwartz
(1936), who reported skin lesions and systemic poisoning among workers reported
to have inhaled PCBs. The skin lesion characteristic of PCBs and other
chlorinated hydrocarbons has come to be described as "chloracne." Toxicological
and public health interests in PCBs were increased in 1968 with the occurrence of
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PCBs in Bluefish
Page 8
the "Yusho incident" (Okurnura and Katsuki, 1969), in which more than 1600
Japanese ingested rice oil contaminated with 2,000 parts per million (ug/g; ppm)
of PCBs (Kanechlor 400) from a heat exchanger. Symptoms of Yusho included
chloracne, hyperpigmentation of the skin, eye discharge, weakness, numbness and
disturbances in liver function. Subsequent analysis of samples from Yusho suggest
that the rice oil was contaminated with high concentrations of dibenzofurans as
well as with PCBs (Kuratsune et al., 1976); it would appear that the symptons of
Yusho were the result of exposure to more than a single contaminant, and that
PCBs alone were not responsible for the full range of biological and biochemical
effects observed in Yusho.
PCB toxicity has been tested in vivo and in vitro using many species,
including several phyla of invertebrates and many vertebrates such as fishes,
birds, rodents, and non-human primates. Epidemiological data are available on the
effects of PCBs on humans in several instances of industrial exposure to PCBs
(Wasserrnan et al., 1979). One of the first indications that PCBs had the
potential to cause severe health effects in mammals was the determination that
reproductive failure among ranch mink fed Great Lakes fish was due to PCBs in the
fish used as food, and that mink were highly susceptible to PCB toxicity
(Hartsbrough, 1965; Ringer, 1983). Subsequent studies in primates by Alien and co-
workers showed that low concentrations of PCBs caused irregular menstrual cycles,
early abortions and stillbirths among Rhesus macaques (Alien et al., 1973, 1974;
Allen and Norback, 1976).
Because of the potential for PCBs to cause health effects in humans, the
U.S. Food and Drug Administration (FDA) between 1969 and 1971 established
temporary tolerance levels for PCBs in food products. Effective April 1, 1981,
the FDA Tolerance Limit for PCBs in foods included milk (1.5 ppm on a fat basis),
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PCBs in Bluefish
Page 9
poultry (3.0 pom fat basis) animal feed (2.0 pprn), packaging materials (10.0 pprn)
and fish and shellfish (2.0 ppm) (Meeting, 1983). Litigation initiated by the
National Fishermen's Association delayed a final ruling on the Tolerance Limit
for PCBs in fish until 1984. 0 5.0 ppm Temporary Tolerance Limit for total PCBs
in fish and shellfish was in force from 1981 through 1984. In 1984 the limit was
reduced to the present value of 2.0 ppm.
The concern over PCB levels in fish and shellfish reflects the fact that
fish are an important link in the food-chain leading to man, and that the
consumption of PCB-contaminated fish is one of the major routes for the transport
of PCBs from the environment to the human population (Hutzinger et al., 1974}
NOS, 1979; Swain, 1983; Belton et al., 1983, 1985). In recent studies of PCBs in
human milk, Schwartz et al. (1983) determined that fish eaters in the Great Lakes
region had higher concentrations of PCBs, even among groups consuming only six to
12 fish meals per year. Although recent investigations have shown that the major
toxicological effects of PCBs are due to specific, individual congeners (Safe,
1984), present regulations regarding the allowable limits of PCBs in foods such
as fish and shellfish are based upon total PCB concentrations (Horn and Skinner,
1985; Belton et al., 1983, 1985).
The Special Problem g_f PCBs in Fishes and Other ftquatic Organisms
The persistence of PCBs in the environment leads, ultimately, to their
transport to and deposition in lakes, rivers, estuaries and oceanic waters. In
addition to domestic and industrial waste water disposal serving as local sources
for PCBs, atmospheric transport assures that surface waters around the globe will
serve as environmental sinks for PCBs and as a source of PCB contamination to all
environmental compartments (Nisbet and Sarofim, 1972; Fuller et al., 1976; NAS
1979; Wasserman et al., 1979). Because PCBs are partially soluble in water, and
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PCBs in Bluefish
Page 10
because they tend to partition to fine participate matter, organic matter and
lipids, PCBs in aquatic systems are available to aquatic biota through several
routes, including bioconcentration from water, accumulation from sedimentary
deposits and transport through the food chain (Hamelink et al., 1971; Hutzinger
et al., 1974; Branson et al., 1975; Pizza and O'Connor, 1983; Rubinstein et al.,
1983, 1984). Once accumulated, PCBs partition to depot lipids where they have a
long half-life. Those organisms serving as food sources for other organisms in
the aquatic food-web may function effectively as vectors for PCB transport in
aquatic systems (Thornann and Connolly, 1984; O'Connor and Pizza, in press, a).
Food chain transport appears to be the major source of PCB contamination for many
*
species of fish, including many that are food resources for the human population
(Thomann and Connolly, 1984; O'Connor, 1984; Rubinstein et al., 1984; O'Connor
and Pizza, in press, a; O'Connor and Huggett, in press).
In highly contaminated aquatic ecosystems, PCBs may accumulate to very high
concentrations in sediments and in fishes. In the Great Lakes, for instance, PCB
concentrations in many commercial and sport fishes may exceed the 2.0 ppn FDfl
Tolerance Limit (Cordle et al., 1982; Schwartz, 1983). In some East coast
estuaries, such as the Hudson River, Raritan Bay, New York Harbor and New Bedford
Harbor, industrial and domestic sources of PCBs have led to the contamination of
many fisheries resources such as eels (Rnauilla rostrata). striped bass, bluefish
and blue crabs (Callinectes sapidus) (Sloan and ftrmstrong, 1982; Belton et al.,
1983, 1985; Weaver, 1984). In several instances public health advisories
concerning the consumption of PCB-contaminated fisheries products have been
issued. In the New York metropolitan area, as well as in New Bedford,
Massachusetts, certain commercial fisheries have been closed or restricted
(Belton et al., 1985; Horn and Skinner, 1985). In 1976 Jelinek and Corneliussen
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PCBs in Bluefish
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reported that "...the occurrence of PCBs Cin the diet] has narrowed to the point
where [fish] are now the primary sources of PCBs Cto humans]." In certain
sections of the country such as the Great Lakes States, metropolitan New York and
New Bedford, evaluations of the potential effects of PCBs in the seafood consumed
by humans have led to serious concern (Swain, 1983; Belton et al., 1985).
Unlike the problem of PCB contamination in foodstuffs such as eggs, milk and
meat, contamination of seafood with PCBs is an ecological problem, rather than a
problem of monitoring contaminated sources of animal feeds. In New Bedford,
Massachusetts, for example, the discharge of PCBs from industrial sites has led
to the contamination of fisheries in and adjacent to New Bedford Harbor. The
question of PCB transport from the Harbor system to the fishing grounds adjacent
to the Harbor is being addressed; however, it would appear that migration of
finfish and shellfish into and out of the Harbor and Buzzards Bay results in PCB
contamination of northern lobster (Homarus americanus), winter flounder
(Pseudopleuronectes afnericanus) and other species. This leads to a lack of
confidence in the suitability for human consumption of fishes that are caught in
the region (Weaver, 1984).
Fishes from the Hudson River, the Hudson estuary, New York Harbor and
adjacent oceanic regions are, likewise, contaminated with PCBs (Nadeau and Davis,
1976; Cahn et al., 1977; Spagnoli and Skinner, 1977; Sherwood et al., 1978;
Stainken and Rollwagen, 1979; flrmstrong and Sloan, 1980, 1982; O'Connor, 1982,
1984a, 1984b; O'Connor et al., 1982; Thomann, 1981; Sloan et al., 1983; Belton et
al., 1983, 1985; Brown et al., 1985; Samuelian et al. in review), A summary of
PCB concentrations in some important fishery resources was provided in O'Connor
et al. (1982) and by O'Connor and Pizza (in press, a). The major source of PCBs
to the Upper Hudson River, and a significant contribution of the pollutant to the
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PCBs in Bluefish
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estuary, was determined to be an industrial discharge (Bopp, 1979; Bopp et al.,
1981, 1984). although that source has been controlled (Horn et al., 1979), there
remain some £00 to 300 metric tons of PCBs still in the process of transport from
upstream sites to New York Harbor and adjacent coastal waters (Schroeder and
Barnes, 1983). In New York City PCBs are still discharged with domestic
wastewater (MacLeod et al., 1981; Mueller et al., 198£). ftlthough PCBs in fishes
from the Hudson estuary have declined since elimination of the major upstream
source (Sloan and flrrnstrong, 1982; Sloan et al., 1983; Brown et al., 1985),
downstream transport and continued discharge of PCBs to the system from
wastewater sources maintain body burdens of PCBs in Hudson River fishes above the
«
FDfl 2.0 pprn Tolerance Limit (Horn and Skinner, 1985; Brown et al., 1985).
PCB contamination of fishes and seafood is not confined to systems with
large PCB inputs, nor is it restricted to species resident in enclosed or semi-
enclosed systems such as rivers, lakes and estuaries. Rather, transport
processes, physical partitioning in the environment and food-chain transport of
PCBs in the global ecosystem have resulted in measureable concentrations of PCBs
in many ecosystems and resources, including the coastal oceans, deep oceans, and
remote areas (Risebrough et al., 1968; Richardson and Maid, 1982; GESOMP, 1984;
Stegeman et al., 1986).
PCB flccurnulation in Fishes
It has long been known that marine organisms, particularly fishes,
concentrated certain elements and compounds in their flesh to concentrations
greater than those in the environment. The phenomenon is referred to as
"bioconcentration" or "bioaccumulation". Such phenomena were described in studies
of radionuclides in marine organisms (Lowman, 1971), and in bioaccumulation
studies of DDT transport in the Flax Pond ecosystem by Woodwell et al. (1967).
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PCBs in Bluefish
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Three different processes may operate when an organism accumulates an
environmental contaminant to concentrations greater than those in the ambient
environment. The terms describing these processes-bioconcentration,
bioaccurnulation and biomagnification-were clarified by Brungs and Mount (1978)
and Macek et al. (1979) as follows:
Bioconcentration ... the process whereby substances enter aquatic
organsisms through the gills or other respiratory epithelia directly from
water;
Bioaccurnu1ation ... the overall accumulation of a chemical substance from
the water, and any other process leading to the accumulation of the
«
substance, including dietary uptake?
B i omaqn i f i cat i on ... a process whereby concentrations of accumulated
materials increase as these materials pass up the food chain through
two or more trophic levels.
In this section of the report we review the current state of knowledge
regarding the bioavailability of PCBs in the environment, and the two major
mechanisms associated with assimilation of PCBs into the body of fishes from
environmental sources; assimilation from water, and assimilation from food.
Bioavailabi Htv of PCBs to Fishes
Critical to an understanding of PCB assimilation by fishes is an
understanding of the extent to which PCBs in various environmental sinks are
"bioavailable"; i.e., exist in a state in which they can enter and be retained by
an organism. PCBs that are dissolved in the water may be completely bioavailable.
That is, if an organism were to irrigate the gills with water containing
dissolved PCBs, or if the organism were to ingest water containing PCBs, they
would be assimilated with high efficiency. PCBs that are associated with
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PCBs in Bluefish
Page 14
sediments, however, are less "bioavailable". Due to their strong tendency to sorb
to participate matter, PCBs associated with either deposited or suspended
sediments would, upon breathing or ingestion, be assimilated with a lower degree
of efficiency, as determined by partitioning between the particulate matter and
the lipids of the organism. PCBs in food items are unavailable to the ingesting
organism unless and until the food organism is consumed, at which point a number
of factors regarding food conversion efficiency, digestive processes and cross-
gut transport phenomena come into play, each with the potential to affect
bioavailability of PCBs in food. Although physical processes dictate that some of
the PCB in sedimentary deposits may become available through the water column,
and that some of the PCB dissolved in water will become adsorbed to sedimentary
material, PCBs in organisms tend to remain stable, and can only become available
after ingestion.
flvailability of PCBs From Water
although PCBs are only "sparingly soluble" in water (Hutzinger et al., 1974;
Haque et al., 1974), PCBs dissolved in the water column may be assumed to be
completely available to fishes by the process of equilibrium partitioning (Pavlou
and Dexter, 1979; McKim and Heath, 1983). The best measure of the direction and
magnitude of equilibrium partitioning for non-polar materials such as PCBs is the
octanol-water partition coefficient, a measure of the tendency for the chemical
(in this case, PCBs) to dissolve in a non-polar solvent (e.g. n-octanol), as
opposed to the highly polar solvent, water (Karickhoff et al., 1979). Octanol-
water partition coefficients are often expressed as log values (Log K ).
although the tissues of fish are not directly equivalent to an organic solvent in
their tendency to accumulate PCB from a water solution, octanol-water partition
coefficient of organic contaminants such as PCBs and the tendency for fishes to
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PCBs in Bluefish
Page 15
bioconcentrate such compounds are directly correlated (Hamelink et al., 1971;
Neely et al., 1974; Spacie and Hamelink, 1982; Mackay and Hughes, 1984).
The bioconcentration factor (BCF) is a measure of the tendency for an
organism to accumulate a substance from the Mater. The BCF for PCBs has been
calculated for many species, and two different estimators of bioconcentration
have been proposed. In the first, BCF is expressed as the concentration of PCS
attained in the tissues of the fish at equilibrium or steady-state, divided by
the concentration of the PCB in exposure Mater (Hamelink et al., 1971). The
second measure of BCF, proposed by Branson (Branson et al., 1975). employed a
kinetic definition, i.e., the BCF Mas stated to be the ratio of the assimilation
rate constant for the compound moving into the fish (k.), divided by the measured
elimination rate constant (k_). Branson's measure Mas directed primarily toward
evaluating the BCF at steady-state, based upon a short-term test (less than 15
days).
Published data for BCF values among fishes exposed to PCBs vary (Spacie and
Hamelink, 1982; Mackay, 1982); hoMever, they generally fall into a narroM range,
between 1 X 10 to 5 X 10 (MAS, 1979; O'Connor and Pizza, in press, a; Mackay
and Hughes, 1984). Values for a number of freshwater and marine species are
presented in Table 1.
As originally proposed, the calculation of BCF Mac used to determine, frow
Mater concentration data, what the probable burden of PCBs might be in fishes
exposed to a contaminated environment. The original experimental work and
evaluation of the technique showed great promise in that the use of the BCF-bas*d
calculations provided a reasonably accurate estimate of the actual PCB body
burden observed in fishes in the environment (usually within a factor of from 3
to 5; Clayton et al., 1977; Pavlou and Dexter, 1979; Mackay, 1982; Shaw and
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PCBs in Bluefish
Page 16
Connell, 1984; O'Connor and Pizza, in press, a). However, as observed by Specie
and Hamelink (198£) and Shaw and Connell (1984), such a range of error is
unsatisfactory when applied to questions of regulation and environmental impact,
in that it is not accurate enough to predict PCB concentrations as being above or
below current FDR Tolerance Limits. O'Connor and Pizza (in press, a), using
accepted BCF values, calculated probable PCB burdens in a number of fishes from
the New York Bight region, including bluefish. They found that in all cases,
observed PCB concentrations in fishes from the Bight region were in excess of
4
that calculated by using a BCF of 1 X 10 (Table 2). The probable reasons for
such discrepancies are:
1) not all the "dissolved" PCB in the aqueous medium is bioavailable;
2) fishes do not retain all the PCB assimilated by bioconcentration
for a long period of time; and
3) fishes accumulate a significant portion of their PCB body burden
from sources other than direct uptake from water (Norstrom et al.,
1976; Shaw and Connell, 1984; Thomann and Connolly, 1984; O'Connor and
Pizza, in press, a).
The subject of bioaccumulation of PCBs from dietary sources will be addressed
in a subsequent section of this report.
Os noted at the beginning of thi» report, the PCBs are a family of 2<99
compounds. Based upon differences in physical and chemical characteristics, each
isomer may have different modes of behavior in the environment (Mullin et al.,
1984; Oliver and Niimi, 1985). Since different PCBs have different solubilities
and octanol-water partition coefficients, it may be expected that different
isomers would be bioconcentrated from the environment with different levels of
efficiency.
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PCBs in Bluefish
Page 17
Early studies with different commercial mixtures of PCBs showed that the
more highly chlorinated PCBs had a tendency for greater bioconcentration from
water (Metcalf et al., 1975; Mayer et al., 1977). Other authors have reasoned
that PCBs with greater numbers of chlorines (i.e. up to 6 or 7) should be
bioaccumulated in fish to a greater extent, primarily as the result of increased
lipophilicity of such molecules (Mackay, 1982; Mackay and Hughes, 1984). Asa
result of such partitioning, the distribution of PCB congeners in natural
populations of fishes may resemble PCB mixtures more similar to the higher
chlorinated industrial preparations (e.g., Aroclor 1254; Oroclor 1260) than
commercial mixtures containing a mixture of congeners with fewer chlorines.
However, the differential accumulation of lower- and higher-chlorinated PCB
isomers has been difficult to demonstrate in natural populations of fishes.
Karickhoff (1979), Spacie and Hamelink (1982) and Niimi and Oliver (1983) have
noted that the congener distributions of PCBs in the bodies of feral fishes are
deter mi ried more by the kinetics of elimination than by assimilation. The reason
for this is that differences in partitioning between water and tissue for
different PCB congeners are so small as to be trivial, whereas differences in the
structure of PCB congeners with identical K 's may be sufficient to lead to
measureable differences in rates of elimination or metabolism (Bruggeman et al.,
1981; Spacie and Hamelink, 1982; Shaw and Connell, 1984; Oliver and Niiwi, 1985).
Availability of PCBs from Food
Some of the earliest research on PCB accumulation in fishes was directed
toward defining bioconcentration (Hansen et al., 1971; Hamelink et al., 1971).
However, it was well known at the time that PCBs and compounds with similar
physical-chemical characteristics were assimilable from the food (Johansson tt
al., 1972) and were transferred from predator to prey in the food-chain (Isaacs,
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PCBs in Bluefish
Page 18
1973; Lieb et al., 1974; Krzeminski et al., 1977; Young, 1984).
Whether fishes accumulate PCBs primarily from water or primarily from the
food may be academic since the final outcome of the contaminant uptake process is
the same, regardless of the source of the contaminant (Pizza, 1983; O'Connor and
Pizza, in press, b). However, from the point of view of environmental fate and
transport, ecosystem modeling and regulatory decision-making, the distinction is
quite important. If, on the one hand, PCBs in fishes derive primarily fro» the
water column, efforts to understand the fate of PCBs in the environment and in
fishes can be simplified and directed at straightforward problems of aqueous
transport, partitioning and bioconcentration (Mackay and Hughes, 1984; Shaw and
Connell, 1984). Models predictive of bioconcentration may be based upon basic
environmental parameters such as water concentration, exposure frequency and gill
transport (Califano, 1981; Califano et al., 1982; Brown et al., 1982; McKin and
Heath, 1983; Mancini, 1983). fin aggressive program designed to limit discharge of
PCB to suface waters nay be implemented as a means for solving the contamination
problem (Hetling et al., 1979; Horn and Skinner, 1985). If, on the other hand,
PCB transport occurs primarily via the diet, PCB sinks in the sediments and the
biota serve as the primary sources for maintaining body burdens in fishes
(Rubinstein et al., 1983, 1984; O'Connor and Pizza, in press, a; Connolly and
Uinfield, 1984), and control of PCB contamination through regulation of waste and
wastewater discharges may be much more difficult. More important, if PCBs are
accumulated primarily from dietary sources, fishes will retain higher body
burdens than might occur from water exposure alone, especially if the species in
question feed primarily upon benthic organisms.
Studies of food-chain transport of PCBs in fishes were first conducted in
1973 and 1974 (Metcalf et al., 1975), although the potential for PCBs to b*
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PCBs in Bluefish
Page 19
"magnified" in food chains was noted as early as 1966 and 1968 (Jensen, 1966;
Risebrough et al., 1968; Duke et al., 1971). Initial food chain studies (Metcalf
et al., 1975) demonstrated both the persistence and transport potential of PCBs
in aquatic food chains. In a subsequent food-chain study, Scura and Theilacker
(1977) attempted to discern the relative importance of food and water uptake of
PCBs; they concluded that there was no evidence of a "food-chain-phenomenon" in a
three-tiered laboratory model ecosystem, but that PCB transport was due to
equilibrium partitioning. Essentially the same conclusion was reached by Clayton
et al. (1977) and Pavlou and Dexter (1979) based upon data from field monitoring
studies in Puget Sound.
*
Food-chain studies in natural and model ecosystems, however, could not
provide the degree of resolution needed to ascertain whether dietary PCB
transport was an important phenomenon. Since all PCB transport may be assumed to
occur due to equilibrium partitioning between source and tissue regardless of the
pathway, investigators had to tolerate analytical limitations in the
determination of which PCBs derived from which sources.
Beginning in the early 1970's, investigators employed radiotracers in the
analysis of PCB transport from food to fishes. Hansen et al. (1976) reported an
efficient dietary uptake of the components of firoclor 1242 in channel catfish
(Ictalurus punctatus). as well as differential retention of PCB congeners. Mayer
et al. (1977) reported on the magnitude of dietary uptake of PCB by channel
catfish. They showed that the dietary accumulation of PCB by catfish increased
with the degree of chlorination and that higher chlorinated congeners in ftroclor
1260 accumulated to levels two times that of the congeners in Aroclor 1232.
Mitchell et al. (1977) reported efficient and rapid transport of dietary 14-
C PCB in the codfish (Gadus morhua). PCBs were detectable in all tissues of th»
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PCBs in Bluefish
Page 20
codfish within two hours after administration of a single dose. Following long
periods of exposure to a PCB-contaminated diet, Sangalang et al. (1981) detected
high concentrations of ftroclor 1254 in testes and livers of codfish.
O'Connor (1982) reported widely varying PCB concentrations in larval striped
bass taken at different locations in the Hudson estuary. Investigations by
Califano (1981) and by West in et al. (1985) showed that PCBs in early, non-
feeding larvae were determined by PCBs passed from the female in the yolk of the
egg. West in et al. (1985) determined that feeding larval stages assimilated PCBs
from food with high efficiency, but only if the larvae had low concentrations of
PCBs to begin with. Larval body burdens of PCBs increased in proportion to the
amount of the contaminant in the food source.
Califano (1981) performed comparative studies of PCB accumulation from food
14
and water in young-of-the-year striped bass by using C labelled firoclor 1254.
He showed that PCB uptake from food and water was important, but that uptake from
food accounted for more than half the body burden accumulated during 48-hour
exposures. Based upon experimentally-determined bioaccumulation factors (BftF) for
young-of-the-year striped bass, Pizza and O'Connor (1983) estimated that between
55% and 83* of the PCB burden of Hudson River resident striped bass derived from
dietary sources.
Pizza (1983 and unpublished data) studied PCB accumulation fro* food in
striped bass, spot (Leiostomus xanthurus). white perch and winter flounder
(Pseudop1euronectes americana) and determined the following:
1) PCBs are accumulated from the food with an efficiency of 85* to 95 *»
2) Dietary PCBs accumulate rapidly to high tissue concentrations; and
3) The relative contribution of dietary PCBs to body burdens in all the*e
species ranged from about 50* to more than 80% in the environment.
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PCBs in Bluefish
Page 21
It is important to note, however, that high assimilation efficiency does not
always lead to the accumulation of high body burdens. Among species with little
body fat (e.g. flounder), PCB accumulations are generally low; the Major
proportion of the dietary dose may be eliminated during a short period of time (6
to IS hours) following transport across the wall of the gut (Pizza, 1983;
O'Connor and Pizza, in press, b).
Essentially the same conclusions have been reached in studies with
freshwater fishes. Following acute dietary exposure of rainbow trout and yellow
perch (Perca flavescens) to PCBs, concentrations increased rapidly in all tissues
(Guiney and Peterson, 1980). Pizza (1983) calculated the assimilation
efficiencies associated with these experiments to be from 80 to 90%. When yellow
14.
perch were given a single, oral dose of C-labelled 2,5,2', 5'-
tetrachlorobiphenyl, they retained about 85% of the 8(90 ng administered; 15% of
the total body burden was determined to be in the muscle tissue (Guiney and
Peterson, 1980).
Niirai and Oliver (1983) fed rainbow trout mixtures of 80 PCB congeners in a
single dose and determined assimilation efficiencies of from 62% to 85%; they
detected no trend in assimilation efficiency among the congeners, a fact
consistent with the notion that bioaccumulation of PCBs is more dependent upon
elimination rate constant than upon congener-specific efficiency of assimilation
(O'Connor and Pizza, in press, a).
flmong rainbow trout reared on PCB-contaminated diets, Hilton et al. (1983)
showed that contaminant accumulation was in direct proportion to dietary
exposure, and did not appear to reach a "steady-state". Evaluation of the study
by Lieb et al. (1974) in which rainbow trout were fed a PCB-contaminated diet for
32 weeks shows an apparent approach to steady state, at least in term of
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PCBs in Bluefish
Page 22
concentration. Lech and Peterson (1983) point out, however, that when growth is
factored into experiment, there occured a continual increase in the total mass of
PCB accumulated by the fish. Most important, these studies show that PCBs in
rainbow trout have a very long half-life, in excess of 200 days (Niimi and
Oliver, 1983). Mayer et al. (1977) exposed coho salmon (Onchorhvnchus kisutch).
to Oroclor 1254 in the diet for periods up to 250 days. Their data showed that
PCB accumulation in coho salmon was in direct proportion to dietary
concentrations, and that the approach to "steady-state" required long periods of
time O200 days).
Virtually all attempts to relate PCB accumulation in natural populations of
*
fishes have arrived at the same conclusions: (1) the major source of PCBs to fish
may be found in the diet; and (2) the primary determinants of the ultimate burden
to be found in a given species of fish are the mass accumulated per dose (meal)
and the inherent rate of PCB elimination for the species (Norstron et al., 1976;
Weininger, 1978; Thomann and St. John, 1979; Thomann, 1981; Jensen et al., 1982;
Pizza and O'Connor, 1983; 0*Connor, 1984a, 19846; Thomann and Connolly, 1984;
O'Connor and Pizza, in press, b). In several instances where authors have
concluded that fishes acuumulate PCBs directly from water, we have found that
insufficient data have been collected with which to evaluate the dietary route of
PCB accumulation (e.g. Macek et al., 1979; Brown et al., 1985), or that the
criteria applied to a satisfactory prediction of PCB burden* were so broad as to
accept predictions + 50* or more (Branson et al., 1975; Clayton et al., 1977;
Pavlou and Dexter, 1979).
Perhaps the most comprehensive evaluation of PCB transport to fishes from
the environment was carried out by Thomann and Connolly (1984). By using data
from the study of food-webs in the Lake Michigan ecosystem (see also Weininger,
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PCBs in Bluefish
Page 23
1978} Thomarm, 1981; Connolly and Winfield, 1984), Thomann and Connolly (1984)
concluded that PCB transport in the Lake Michigan food-web followed energy flow
from smaller organisms in lower trophic levels to the lake trout (Figure 4). Lake
trout accumulated as much as 90* of their body burden of PCB from dietary
sources.
The only study in which actual dietary doses of PCBs were evaluated
simultaneously with body burden was reported by 0'Connor <1984a>. In that study,
striped bass flesh and stomach contents were measured for PCB content, and
regression analysis was used to establish the relationship between body burden
and daily dose of PCB in the food. O'Connor <1984a) established that striped bass
from the New York harbor region ingested a daily ration equivalent to about 5% of
body weight per day, and that measurement of the PCB mass in samples of food
enabled the calculation of mass of PCB ingested per day per fish (Table 3).
Coefficients of determination for the regression of PCB body burden on the daily
dose of PCB taken in with the food were 0.67 and 0.65 for bass from samples taken
at weehawken, New Jersey and at Canal Street in Manhattan (O'Connor, 1984a).
O'Connor concluded <1984a; p. 157):
"PCB body burdens in. ..striped bass are maintained by the consumption of a
PCB-contaminated diet. The source of PCB to the prey is from both the
water and the sediments. Once ingested, the PCB in prey organisms is
assimilated into the striped bass with high efficiency... and plateau
levels are achieved rapidly..."
Overall, it is apparent from the literature that PCBs in fishes are
accumulated from two sources, direct water uptake and food-chain transport. In
different environments one or the other of these processes may dominate,
depending upon the concentrations of PCB in the water column. In the Upper
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PCBs in Bluefish
Page £4
Hudson, for example, where dissolved and suspended PCB concentrations in the
Mater may be very high (Schroeder and Barnes, 1983; Sloan et al., 1984; Brown et
al., 1985), fishes may accumulate a large proportion of their PCB by direct
uptake from water. In environments such as the open ocean where there is little
suspended material and PCB concentrations in the water are exceedingly small, the
proportion of the body burden deriving from water uptake is reduced. Under such
conditions, the primary route for PCB accumulation would be via the food-chain.
Such phenomena have been tested experimentally by Rubinstein et al. (1983, 1984)
in model ecosystems.
Bluefish
*
Considering the problem of PCB accumulation in bluefish, it is most likely
that the primary source of PCBs is the food chain, and that the processes
involved in PCB transport to bluefish in coastal waters are essentially the same
as for striped bass in the New York harbor region, and for lake trout in the open
waters of Lake Michigan, fls noted by O'Connor and Pizza (in press, a) for striped
bass, the calculation of PCB concentrations in bluefish from water concentration
data results in estimates that are much lower than the values observed (Table 2).
If water were the only source of PCB to bluefish in Atlantic coastal waters, one
would expect concentrations of PCBs in bluefish to remain at or below 1.0 ppm.
However, both bluefish and the food items upon which they prey in the estuary and
in the ocean are contaminated with PCBs at concentrations between 1.0 and 20.0
pprn (Belton et al., 1983, 1985).
Assuming the dietary requirements of bluefish to be approximately the sane
as striped bass (i.e. about 5X of body weight per day), and a food resource
contaminated with PCBs at concentrations between 1.0 and 5.0 ppm, application of
dietary mass transport models and pharmacokinetics results in the prediction that
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PCBs in Bluefish
Page £5
bluefish body burdens would range from 2.0 to 10 ppm (wet-weight basis),
depending upon the age of the specimen sampled. During winter months, when
bluefish move offshore into waters less contaminated with PCB, one would expect
lower overall body burdens; however, the approach of body burdens to plateau (or
"steady-state") as the result of dietary exposure is so rapid that maximum body
burdens could be expected to be reached as soon as the population migrated back
to coastal waters and encountered prey contaminated with high concentrations of
PCBs (Pizza and O'Connor, 1983).
Pharmacokinetics of. PCB accumulation in Fishes
PCBs are assimilated into fish from water by processes which follow first
order kinetics (Branson et al., 1975; McKim and Heath, 1983; Mackay and Hughes,
1984). That is, a constant proportion of the PCBs in the water to which the fish
are exposed is transported across the gill surface into the blood and distributed
to the tissues. The mechanism for cross-gill transport has not been defined, but
is predictable based upon equilibrium partitioning using the concepts of
thermodynamic mass-transport (Thomann, 1981). It has been suggested that the
transport of PCBs across the gills of fishes is neither a diffusional process nor
active transport, but is best described as "ligand-assisted-diffusion," in which
large molecules (probably lipoproteins) in the gill tissue sorb or bind the PCBs.
Once in contact with the blood on the internal side of the membrane, PCBs sorb to
or dissolve in blood lipoproteins, and are transported to the tissues in
proportion to the blood supply of each tissue (Califano, 1981).
PCBs in the food of fishes are instantaneously incorporated into the body
burden of the fish (i.e. zero-order pprocess). They are not, however,
assimilated instantaneously into the various tissues; as with the process of
transport from water to the body of the fish, partitioning from the gut to th»
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PCBs in Bluefish
Page £6
tissues follows first-order processes (Bruggeman et al., 1981; Pizza and
O'Connor, 1983).
Pizza and O'Connor (1983) defined the kinetics of PCB assimilation from food
into striped bass in a series of experiments that involved both single- and
multiple dosing of fish with known quantities of PCBs. They found that the
dynamics of the PCBs conformed to pharmacokinetic models developed for drugs by
Goldstein et al. (1974). Transport of PCBs from the site of absorption (the gut)
to the tissues occurred in two phases over a period of 120 hours. The first
phase, lasting about 24 hours, showed a rapid loss of PCB from the gut coupled
with an increase in the quantity of PCB in the remaining tissues. The loss of
PCBs from the gut was equivalent to the rate of assimilation into the remaining
tissues, and was defined by
log M = log M - k t/2.30,
o a
where K is the quant itiy at the absorption site at time zero, M is the quantity
remaining at time t and k is the assimilation rate constant obtained from the
slope of the regression of log unabsorbed dose in the gut over time (Figure 5).
Pizza and O'Connor (1983) noted that elimination of PCBs from striped bats
began as soon as PCBs were transported from the gut to the tissues. This suggests
that not all the PCBs assimilated remained within the body of the organism, even
for a compound with a high degree of persistence. They defined the elimination
rate constant k for PCBs, as
log X = log XQ - kg t/2.30,
where X0 is the quantity of compound in the body of the fish at time 0, X is the
quantity present at time t, and kg is the elimination rate constant.
In reality, fishes are not exposed to single doses of PCBs in the
environment. In contaminated regions fishes are exposed to PCBs at varying
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PCBs in Bluefish
Page £7
concentrations in the food (O'Connor, 1984a) and in the water (Mancini, 1983). By
applying pharmacokinetic principles to multiple doses of PCBs in food, Pizza and
O'Connor (1983; O'Connor and Pizza, in press, a) were able to determine the rate
at which a long-term PCB burden would accumulate in fishes, and to estimate the
burden as it accumulated over time, and as the organism increased in size.
Similar approaches have been used by Thomann (1981) and by Thomann and Connolly
(1984) in their studies of contaminant accumulation in lake trout from Lake
Michigan.
These results show that fishes, once exposed to PCBs in the food, accumulate
PCB rapidly and achieve a "plateau" concentration of contaminant quickly. For
striped bass in the Hudson, 90* of plateau was reached within 8 doses; assuming
fish that feed twice per day, a fish entering a new, contaminated environment
will have reached pleateau PCB concentrations within 4 to 6 days (O'Connor and
Pizza, in press, a).
Pharmacokinetic, or mass-transport, concepts have been applied in several
models aimed at predicting PCB concentrations and burdens in fishes from
contaminated environments. The most important factors determining the body burden
were: (1) the dose of PCB given to the fish per unit time, and (2) the rate
constant for elimination of the PCB from the fish. It has also been established
that accumulation of PCBs in fishes differs according to the physical-chemical
characteristics of individual PCB congeners, and the extent to which individual
congeners are metabolized, transformed or eliminated by fishes (Hansen et al.,
1976; Shaw and Connell, 1980; Matsuo, 1980; Bruggernan et al., 1981; Califano,
1981; Niimi and Oliver, 1983; Smith et al., in press). Separating the
accumulation process into two segments (assimilation and retention), most
chlorobiophenyl congeners are assimilated in roughly equal proportions from
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PCBs in Bluefish
Page 28
environmental media, whereas the elimination process follows kinetics that differ
substantially for different congeners (Hutzinger et al, 1972; Bruggeman et al.,
1981; Niimi and Oliver, 1983). It has been suggested that for PCBs partitioning
from the environment to fishes, whether from water or from food, partition
4 6
coefficients of PCB congeners are sufficiently alike (18 to 10 ) to be
unimportant as a determinants of the mass of each congener assimilated. However,
congener-specific differences in solubility, 1ipophi1icity, macromolecular
binding and physical structure are sufficiently great to lead to measureable
differences in the metabolic and transport processes which determine elimination
from the body of the fish (Bruggeman et al., 1981; Niirni and Oliver, 1983; Smith
et al., in press).
Tissue Disposition and Elimination of PCBs in Fishes
PCBs accumulate in the order from greatest to least concentration as
follows: nervous tissue > liver > gonad > muscle > kidney (Mitchell et al., 1977;
Buiney and Peterson, 1980; Stein et al., 1984; Califano, 1981; O'Connor and
Pizza, in press, b). Pharmacokinetic studies aimed at determining the transport
of PCBs from the site of uptake to the tissues are few; O'Connor and Pizza (in
press, b; Pizza, 1983) determined that residues of ftroclor 1254 were measureable
in all tissues of striped bass within 6 hours after exposure, and that rates of
increase of PCB concentration were different among different tissues. For
example, in single-dose studies, PCBs in muscle, heart and spleen of striped bass
increased during the first 12 to 24 hours after exposure and subsequently
declined as the contaminant was either removed from the body or distributed to
other tissues. In the liver, however, PCBs continued to increase for 24 to 48
hours before a decline was measureable. O'Connor and Pizza (in press, b) were
14
also able to detect a translocation of C-labelled PCB residues from liver
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PCBs in Bluefish
Page £9
tissue to gall bladder and bile that Mas related to PCB elimination (see also
Melancon and Lech, 1976).
Essentially the same pattern of tissue disposition was observed in multiple
exposure studies with striped bass (Table 4), except concentrations in all
tissues increased in proportion to the exposure concentration (Pizza, 1983;
O'Connor, 1984b). Interestingly, there occurred a rapid loss of a portion of each
PCB dose amounting to about 40% to 50% (as mass), whereas between 50% and 60% of
the mass of each dietary dose was retained. O'Connor and Pizza (in press, b)
speculated that PCB disposition in fishes proceeded in two phases. In the first
phase, a fraction of the.assimilated PCBs may be described as "labile," subject
to the sort of rapid elimination seen in laboratory pharmacokinetic studies
(Bruggeman et al., 1981; Pizza and O'Connor, 1983; Lech and Peterson, 1983; McKim
and Heath, 1983). fl second, "stable" fraction becomes stored in tissues or in
depot fat. The stable fraction shows longer elimination half-lives O 200 days)
similar to those observed in elimination studies conducted after long-term
exposure, or with specimens caught from highly contaminated environments (Hansen
et al., 1971; Nisbet and Sarofim, 1972; Metcalf et al., 1975; Mayer et al.,
1977; Niimi and Oliver, 1983). O'Connor and Pizza (in press, b) suggested that
the proportion of the PCB dose likely to enter the stable compartment is
proportional to the body lipid concentration of the species in question. Thus,
species such as striped bass, lake trout or blufish, all of which have a high
concentration of body lipid, may accumulate PCBs to high concentrations, whereas
species with low body fat (e.g. flounder, codfish, etc.) generally show low PCB
burdens (Lieb et al., 1974).
Possible routes for PCB elimination from fishes include diffusion across the
gill to the water and removal via the hepatic pathway. Loss of PCBs across the
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PCBs in Bluefish
Page 30
surface of the gill has been demonstrated in both striped bass and in rainbow
trout (Califano, 1981; Suiney et al., 1977). For such a phenomenon to occur,
however, the organism must be in a medium in which the dissolved PCS in the water
is very low, favoring a diffusional exchange from the gill to the water. Given
the high lipid solubility of PCBs, it is unlikely that such a pathway will
operate for fishes in any situation other than in laboratory exposures where body
burdens may be very high. However, the formation through metabolism of any water-
soluble PCB metabolites (Melancon and Lech, 1976; Stein et al., 1984) may result
in PCB metabolite removal via the gill.
The roost likely route for PCB elimination is via the hepatic pathway; i.e.
«
partitioning to liver tissue from the blood, solubilization in bile fluids and
excretion with the bile to the intestine (Pizza, 1983; O'Connor and Pizza, (in
press, b). In their study of PCB kinetics in individual striped bass tissues,
O'Connor and Pizza (in press, b) determined that the k for PCBs was essentially
the same for all tissues; that is, tissues such as muscle, liver, spleen, etc.,
released PCBs in constant proportion to the PCB mass in the tissue. Since the
liver contained about four times the mass of PCB in other tissues, the greatest
•ass of PCBs was being removed from liver tissue and being transported into bile
for eventual elimination in the feces.
PCB Metabolism in Fishes
Although PCBs are persistent in the environment, their susceptibility to
degradation has been well documented. Hutzinger et al. (1974) and Baxter and
Sutherland (1984) described the photodegradability of PCBs in the atmosphere, and
many workers have demonstrated the potential for microbial populations to either
metabolize or transform PCBs (Furukawa and Matsumura, 1976; Tucker et al., 1975}
Reichardt et al., 1981; Suflita et al., 1983). Recently, Brown et al. (1984)
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PCBs in Bluefish
Page 31
provided evidence of PCB degradation by natural populations of aerobic and
anaerobic bacteria in the upper Hudson River. Their studies showed that bacteria
cause reductive dech1orination of various PCB congeners in anaerobic systems
which facilitated later ring-opening and mineralization of PCBs by aerobic
microorganisms. Brown et al. (1984) speculated that bacterial metabolism of
selected PCB congeners may be one of the major mechanisms, along with selective
volatilization, whereby industrial mixtures of PCBs become transformed to consist
primarily of higher chlorinated congeners with a high proportion of 0,0'- Cl
substitutions. Such congeners are generally recognized as the least hazardous of
the PCB congeners, and the process of bacterial degradation may, in fact, be a
process for PCB detoxification in the environment
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PCBs in Bluefish
Page 32
rate, fis noted by Lech and Bend (1980):
"Several classes of compounds, including some polychlorinated biphenyls,
are metabolized slowly, and their disposition in fish may not be
influenced to any great extent by biotransforrnation. "
Our review of the literature has identified few papers dealing with
metabolite formation and identification of PCB metabolites in fishes. Like
mammals, however, fishes have been shown to possess the hepatic mixed-function
oxidase (MFO) system necessary for metabolism of PCBs to a variety of conjugated
metabolites (Addison et al., 1978, 1979; Forlin and Lidman, 1981; Forlin et al.,
1984). Hutzinger et al. .(1972) studied the metabolism of four PCBs in rainbow
trout as well as in pigeons and rats. Rats and pigeons produced identifiable
hydroxy-PCB metabolites, but no evidence for metabolism was found in the rainbow
trout. Hutzinger et al. (1972) found no evidence for reductive dech1orination of
PCBs in any of the species studied.
Melancon and Lech (1976) isolated a polar metabolite of £,£',5,5'-
tetrachlorobiphenyl in the bile of rainbow trout exposed to the PCB in water. The
metabolite was identified as a glutathione conjugate of 4-hydroxy-2,2',5,5'-
tetrachlorobiphenyl (see Hesse et al., 1978; Shimada et al., 1981). Similar
results were found for PCBs in English sole (Parophrvs vetulus) (Stein et al.,
1984); aqueous-soluble radioactivity deriving from apparent metabolism of PCBs
was detected in the bile of sole, and was shown to be a glutathione conjugate.
However, metabolism of PCB in the English sole proceeded at a very slow rate;
more than 98* of the PCB-derived radioactivity recovered by Stein et al. (1984)
was in the form of parent PCB compounds.
In general, it may be concluded that although fishes contain the enzyme
systems required for the metabolism of PCBs, such metabolism proceeds very
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PCBs in Bluefish
Page 33
slowly, on the order of 1% of the rates for mammals, flssuming that the same
structural and steric phenomena affect PCB metabolism in fishes, we would
speculate that some metabolism of the lower-chlorinated PCB congeners (from 2 to
4 chlorines) would occur. However, metabolism of higher chlorinated PCB classes
would be virtually zero. In ecological systems where PCB concentrations are
substantial, kinetic processes of elimination of parent PCBs would be far more
important in the removal of PCBs from the body of fishes than would the process
of Metabolism (O'Connor and Pizza, in press, a).
accumulation and Disposition of PCB Congeners in. Fish
It has been apparent since the earliest studies of PCB distribution in
natural environments that fishes accumulated groups of PCB congeners that were
not identical with those found in pollutional sources (Risebrough et al., 1968;
Nisbet and Sarofim, 1972; Hutzinger et al., 1974; Nadeau and Davis, 1976). PCB
burdens in fishes and shellfish comprise higher-chlorinated congeners (> 4
chlorine molecules) rather than the lower chlorinated congeners more abundant in
PCB discharges (flrmstrong and Sloan, 1980; Sloan and Armstrong, 1982; O'Connor et
al., 1982).
In the Hudson River and Hudson-Raritan estuary, it was found that fish
samples taken farther from industrial PCB sources contained a higher proportion
of PCBs resembling flroclor 1254 than flroclor 1221 or flroclor 1016, even though
the major sources of PCBs to the system were flroclor 1221 and 1016 (Bopp, 1979;
flrmstrong and Sloan, 1980, 1982; Brown et al., 1985). Bopp et al. (1981) proposed
that the abundance of flroclor 1242 and 1254 downstream from PCB discharges was
related to selective retention of higher chlorinated PCB congeners on sediments
subject to bed-load transport; the lower chlorinated congeners present in flroclor
1016 were gradually lost by transport out of the system in the dissolved form and
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PCBs in Bluefish
Page 34
by volatilization to the atmosphere. The results of modelling studies by Thomann
(1981) and pharmacokinetic studies by Pizza and O'Connor (1983; see also
O'Connor, 1984b) would suggest that although fishes may accumulate lower- and
higher-chlorinated congeners in roughly equal proportions from the environment,
higher rates of elimination for mono-, di-, and trichlorobiphenyls would lead to
the presence in fishes of PCB body burdens chromatigraphically similar to flroclor
1254, rather than flroclor 1016. Laboratory studies by other workers substantiate
these hypotheses (Bruggernan et al., 1981; Niimi and Oliver, 1983).
Time-series analysis of PCBs in striped bass and other species from the
Hudson River (Sloan et al., 1983, 1984; Brown et al., 1985) have demonstrated
that as the mass of PCB input to the Lower Hudson was reduced, PCB concentrations
in fishes not only declined, but also showed a change in the proportion of lower
chlorinated isomers relative to higher chlorinated isorners. Since
downstream sources provide lower-chlorinated congeners to fishes in the New York
Harbor region (MacLeod et al., 1981; O'Connor and Pizza, unpublished data), we
conclude that reductions in mono-, di-, and trichlorobiphenyls seen in fishes
from the Hudson River and estuary are due primarily to selective elimination of
these congeners and the retention of congeners with four or more chlorine
molecules.
Apart from their value as indicators of selective elimination of PCB
congeners, detailed analysis of PCB body burdens in fishes also provides insight
into the potential toxic response associated with the consumption of contaminated
fish by humans. It has been found that PCB congeners with fewer chlorine
nolecules are less toxic than congeners with a greater degree of chlorination.
Thus, as body burdens of PCBs in fish change due to selective elimination of
lower chlorinated congeners, one might expect potential toxicity to increase.
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PCBs in Bluefish
Page 35
Such a conclusion is not fully warranted, however, since structural factors
related to degree and position of chlorine substitution in PCBs play a strong
role in defining toxicity (Goldstein et al., 1977). Safe (1984) has discussed
such factors and has concluded that PCB congeners that are "approximate iso-
stereomers" of 2,3,7,8-tetrachlorodibenzodioxin (Figure 6) have the greatest
potential to exert toxic effects at the metabolic level (as measured by MFO or
ftHH induction). The rather crude analysis of PCB as "firoclor 1016" or "ftroclor
1254" provides no information in this regard, and the assessment of potential
toxic effects of PCBs based upon Oroclor analysis is probably not warranted.
Recently developed techniques allow the isolation and identification of PCB
congeners in environmental samples. Beginning with Ballschmitter and Zell (1980),
standard classification of chlorobiphenyl congeners was established. Mullin et
al. (1984) reported the synthesis and chromatographic properties of all 289
potential PCB congeners and data on the concentrations and mass of PCB congeners
in environmental samples are now emerging from several laboratories (Mullin et
al., 1983; Bush et al., 1983; Smith et al., in press; Samuelian et al.,
manuscript in review).
Smith et al. (in press) analysed samples of sediment, fish and fish food
organisms fro* the Great Lakes for 72 PCB congeners and determined that the most
toxic congeners were either absent or present in very low quantities. They
concluded from their analysis that "...estimates of toxic exposure based on total
PCB values may be unreliable...," due primarily to variation in the partitioning
of PCB congeners in the water column-sediment-fauna ecosystem under study.
Samuelian et al. (manuscript in review) identified 47 PCB congeners from liver
and fle«h of fttlantic tomcod (Microuadus tomcod) from the East River, New York,
as well as from shrimp (Cranoon septemsoinosa) used as food by tomcod. Comparison
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PCB5 in Bluefish
Page 36
of PCB congeners in tomcod with those in food organisms showed sustantial
differences; Cranqon contained greater quantities of lower chlorinated PCB
congeners (di- and trichlorobiphenyls), whereas tomcod contained no
dichlorobiphenyls, and trichlorobiphenyls were present as a minor constituent of
the PCB burden. When Samuelian et al. compared the profile of congeners in the
fish to a profile of congeners from a mixed standard of Aroclor 1016 and 1S54,
they concluded that environmental samples of PCB should not be quantified on the
basis of flroclors since body burden profiles differed significantly from Aroclor
standards.
From the toxicological perspective, Clarke et al. (1986) applied cluster
*
analysis to PCB congeners identified as having potential biological effects based
upon their ability to induce monooxygenase enzymes in mammalian systems. Their
technique holds promise as an effective means of evaluating the PCB burden of an
environmental fish sample for potential toxicity by isolating those components
most likely to influence the health of the consumer.
Evaluation of Data on PCBs iri Bluefish
Despite the value of the bluefish fishery and the fact that bluefish are the
species most sought by recreational fishermen on the Atlantic coast, there are
relatively few data on chemical contamination of the species. A full summary of
bluefish PCB data is presented in a recent data report submitted to Congress
(Anon., 1986). The New York State Department of Environmental Conservation (1981)
reported PCB values for bluefish from a number of sites, including the estuarine
portions of the Hudson River near Peekskill, New York Harbor, the Atlantic and
Long Island Sound coasts of Long Island and open Atlantic waters. Samples taken
in the estuarine system (Peekskill and New York Harbor) had higher PCB
concentrations than samples from outside the harbor system (Table 5). However,
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PCBs in Bluefish
Page 37
one sample taken at Orient Point on the eastern end of Long Island had PCB
concentrations of 3.6 ppm, substantially higher than the values measured in
bluefish from either Peekskill (3.1 ppra) and in New York Harbor (2.1 ppm). fllong
the coast of Long Island, PCB concentrations in bluefish ranged from a low of
0.46 ppm in Long Island Sound, to 0.94 ppm at Cold Spring, and 1.15 ppm at a site
in the Eastern Sound. In contrast to the usual trend where older specimens
contain the higher PCB concentrations, young-of-the-year bluefish from the
Peekskill area had PCB concentrations of 3.1 ppm, whereas older specimens taken
in the open ocean and in Long Island Sound ranged between 350 and 600 mm total
length and had much lower PCB levels (Table 5).
BeIton et al. (1983) reported PCB concentrations in bluefish from New Jersey
waters of the Hudson River and along the fltlantic Coast. Bluefish from the Hudson
River contained 3.44 ppm PCB in 1975 and 1976, while specimens sampled in 1981
had a PCB concentration of 1.78 ppm. Bluefish samples obtained from offshore
sites contained from 0.67 to 1.44 ppm total PCBs. In all, the samples analysed by
the New Jersey Dept. of Environmental Protection (Belton et al., 1983) confirmed
the data reported from NYSDEC, even though sample sizes reported by Belton from
the Hudson estuary were small (n = 4, n = 2 for 1975-76 and 1981, respectively).
fls part of a study to determine toxic hazards to recreational urban
fishermen, Belton et al. (1985) again sampled bluefish from the Hudson River and
Newark Bay region for PCBs. For samples taken in 1982, total PCB concentrations
were 3.29 ppra (n = 5), while in 1983, samples from several sites ranged from 1.51
to 5.44 ppra (n, for the most part, =1). Belton et al. (1985) concluded that PCB
levels in blufish from the Hudson River to New York Bay were likely to exceed 4.0
ppm, and that PCB levels in blufish taken fron the Newark Bay complex were likely
to exceed 2.0 ppm; a public health advisory has been published with regard to the
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PCBs in Bluefish
Page 38
consumption of Bluefish from New Jersey waters (Figure 7, from Belton et al.,
1985).
Hypotheses and Speculations Regarding the Dynamics of PCBs in Bluefish on the
Atlantic Coast of. North America
Although the data are restricted in quantity and quality, it is possible to
propose hypotheses regarding the dynamics of PCBs in blufish populations of the
Western North Atlantic, and to speculate as to sources of PCBs to the bluefish
population and the future course of PCB contamination in bluefish. In large part
these hypotheses and speculations are based upon data from modeling studies with
striped bass and lake trout, and upon pharmacokinetic studies of the behavior of
*
PCBs in striped bass and rainbow trout (Thomann, 1981; Jensen et al., 198S; Pizza
and O'Connor, 1983; Thomann and Connolly, 1984; Connolly and Winfield, 1984;
O'Connor, 19S4a; O'Connor and Pizza, 1985, in press, b). Although these
speculations are made with full knowledge that the data are insufficient, the
relative constancy of PCB dynamics among fish species studied suggests that th(~
concepts and trends put forth will be accurate, although actual levels of PCB
contamination in the bluefish population will be the final determinant of the
time-frame involved.
1. Sources of PCB Contamination in Bluefish
PCB contamination is worldwide, mediated by atmospheric transport, surface
water flow patterns and the transport through the environment of dissolved and
particle-associated PCBs. Due to high concentrations of PCBs in many estuarine
systems and transport of PCB-contaminated estuarine water to coastal oceans, PCB
concentrations in near-coastal waters will be higher than in waters from more
remote ocean areas. Estuarine source of PCBs to coastal waters will influence PCB
concentrations in bluefish in two ways: (1) by causing the direct uptake of PCBs
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PCBs in Bluefish
Page 39
by bluefish exposed to the contaminant dissolved in water; and (2) by causing the
contamination of bluefish prey. O'Connor and Pizza (in press, a), in their paper
on sources of PCBs in marine fishes showed that, for open ocean waters where
dissolved PCB concentrations are low, most of the PCB burden will be accumulated
via the food chain. They predicted that for striped bass more than 70% of the PCB
burden was the result of dietary uptake. Thomann and Connolly (1984), working
with the analogous system of lake trout in Lake Michigan, concluded that more
than 90* of the PCB in lake trout derived from dietary uptake.
PCB concentrations in water, sediments and biota generally show a gradient
from onshore to offshore sites, with the highest concentrations occurring in
estuaries and in near-coastal waters. O'Connor et al. (1982) showed a gradient in
PCB concentration among striped bass from New York waters that decreased with
distance from New York Harbor. It may be expected, therefore, that PCB
concentrations in bluefish will be lower the greater the distance from the coast,
and especially in relation to the distance from New York harbor. Conversely, it
may be predicted that, as bluefish migrate from shelf waters toward the coast
during the spring months, body burdens of PCBs will increase as the fish ingest
food more highly contaminated with PCBs.
2. Concentrations of PCBs in Bluefish
We hypothesize that bluefish, like striped bass and lake trout, will derive
most of their PCB body burden from the diet. Lacking data on dietary
requirements, growth, metabolism and other factors necessary for construction of
an accurate model (Thomann and Connolly, 1984; O'Connor and Pizza, in press, a),
only crude estimates can be made as to what body burdens may be accumulated. ft
means for making such an estimate may be derived from the food-chain studies
conducted on striped bass in New York Harbor, as well as pharmacokinetic studies
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PCBs in Bluefish
Page 40
of PCB assimilation. In those studies it was determined first, that the BflF for
dietary PCBs in striped bass was about 0.75 (Pizza and O'Connor, 1983), and
second, that the relationship between daily dose of PCBs to striped bass and the
body burden was equal to about 2 X log of the dose U.98 and 1.67 for data sets
from Weehawken and Canal Street, respectively; O'Connor, 1984a).
Let us assume, then, that a bluefish weighing 1.5 kg (slightly more than 3
Ib) resembles a striped bass in that: (1) it has similar PCB kinetics; (2) it has
similar rates of metabolism; and (3) it consumes approximately 5* of its body
weight per day in food (from O'Connor, 1984a). Under such conditions, a bluefish
feeding on a contaminated food resource will reach plateau burdens of PCB after a
few days of exposure (Pizza and O'Connor, 1983). The plateau burden, as
micrograms of PCB, may be approximated as:
Log B = 2.0 log D - 1.0 (from O'Connor, 1984a)
where B is the PCB burden and D is the daily dose of PCB. Concentration was
estimated as burden divided by fish weight, or B/1500. PCB concentrations in
bluefish prey range from less than 0.5 ppm to more than 4.0 ppm total PCBs
(NYSDEC, 1981 and Belton et al., 1985). Using the formula above, PCB
concentrations in the adult bluefish of 1,500 g weight may be estimated for
reasonable PCB doses as follows:
PCB in food (uq/q) PCB in bluefish (uq/q)
0. 5 0.3£
1.0 1.27
2.0 5.09
3.0 11.44
These calculated values, ranging from 0.32 to more than 11 ppm PCB in
bluefish have a precision of + 35X, and may be considered accurate only for
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PCBs in Bluefish
Page 41
bluefish deriving their PCBs from dietary sources. Interestingly, they cover the
full range of PCB concentrations seen in bluefish from the New York-New Jersey
metropolitan region. The data are nearly useless, however, in testing the
predictive power of the relationships in the literature. Fortunately we have an
instance (NYSDEC, 1981) in which Ptlantic menhaden and bluefish were collected
frow New York Harbor within two weeks of one another, making it reasonable to
assume that the bluefish in the harbor (average length 590 mm) had been foraging
on the menhaden (average length 243 mm). Given a measured PCB concentration in
menhaden of 1.34 ppm, and an average weight of bluefish of 2,465 g, we would
estimate a concentration' from these data of 4.07 ppm total PCB in the bluefish.
In fact, the observed range for the bluefish sample was from 0.11 to 5.77 ppm
total PCB, with a mean of 2.33 ppm. The calculated value of 4.07 falls within the
estimated range of precision of the prediction (+ 35X) noted by O'Connor <1984a).
Unfortunately we have no real data with which to determine the actual
relationship between food organism PCB content and the concentration of PCBs in
bluefish. Such data are sorely needed, and plans for their collection should be
included in any prograo designed to obtain further information on PCB
contamination of coastal bluefish populations, fls shown in the data from O'Connor
(1984a), PCB concentrations vary widely at different sites even within a confined
environment such as New York Harbor, and the only way to obtain the proper data
is to carry out simultaneous sampling of bluefish, bluefish stomach contents and
forage fish, all from the same site.
3. Persistence of PCBs in Bluefish
Based upon a wealth of data from striped bass, rainbow trout, lake trout and
other species, it is known that PCBs in fish flesh are not permanent; that is,
even though assimilated into fish tissues and into depot fat, PCBs may be
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PCBs in Bluefish
Page 42
removed, gradually, from the body of a fish, as determined through bi-modal
elimination kinetics. As seen in the striped bass (O'Connor and Pizza, in press,
b) a high proportion (40 to 50X) of the PCBs assimilated from a dietary dose are
lost rapidly, whereas the remainder appear to partition to "storage areas" in
tissues, tissue lipids and depot fat. These storage areas retain PCBs for a
longer period of time, with a half-life for elimination on the order of 100 to
200 days.
For many species of fish from the Hudson River, PCB elimination may proceed
at fairly rapid rates once major sources are controlled. Sloan and his co-workers
(Armstrong and Sloan, I960; Sloan and Armstrong, 1982; Sloan et al., 1983; Brown
*
et al., 1985) showed a rapid decline in PCB concentration in Hudson River fish
between 1978 and 1984. The calculated half-times for such declines were rapid,
far in excess of those estimated for the Hudson system in earlier work by Thornann
and St. John (1979). It would appear that the declines observed in PCB
concentrations in Hudson River fish have halted, having reached a quasi-steady-
state imposed by the presence of PCBs in sediments throughout the system (Sloan
et al., 1983, 1984).
4. PCB Congener Distribution in Bluefish
None of the data available to us at this time provide information on the PCB
congener distribution in bluefish; all data from NYDEC and NJDEP are available to
us only as total PCBs or as ftroclors, with no apportionment among congeners or
chlorinated classes. In this regard, one can only speculate that congener
distribution in bluefish is similar to that found for other species and in
bluefish forage organisms. From the data of Smith et al. (in press) and Samuelian
et al. (in review) we would predict the presence of 40 to 50 PCB congeners in
bluefish from ocean waters, with the bulk of the congeners representing tetra-,
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PCBs in Bluefish
Page 43
penta- and hexachlorobiphenyl congeners with a high degree of chlorine
substitution in the o, o' positions. Based upon the few environmental data
available, it would be most unlikely to find in bluefish high concentrations of
PCB congeners known to be particularly hazardous or toxic.
-------
REFERENCES
flddison, R., M. Zinck and D. Willis. 1978. Induction of hepatic mixed-
function oxidase (MFC) enzymes in trout (Salvelinus fontinalis) by feeding
flroclor 1254 or methylcholanthrene. Como. Biochem. Physiol. 61(C): 323-
385.
flddison, R., M. Zinck, D. Willis and D. Darrow. 1979. Induction of hepatic
mixed-function oxidase in trout by polychlorinated biphenyls and
butylated monochlorodiphenyl esters. Toxicol. flppl. Pharmacol. 49: 245-
248.
flllen, J., L. flbraharnson and D. Norback. 1973. Biological effects of
polychlorinated biphenyls and triphenyls on sub-human primates. Environ.
Res. 6: 344-354
flllen, J., D. Norback and I.C. Hsu. 1974. Tissue modifications in monkeys as
related to absorption, distribution and excretion of polychlorinated
biphenyls. flrch. Environ. Contarn. Toxicol. 2: 86-94
fillen, J., and D. Norback. 1976. Pathobiological responses of primates to
polychlorinated biphenyl exposure. Proc. Nat'1. Conf. on Polychlorinated
biphenyls. EPfi-560?6-75-004. pp. 43-49.
flnon. 1986. Report on the 1984-1986 federal survey of PCBs in Atlantic Coast
bluefish. NOflfi Data Report 179 pp.
flrmstrong, R. and R. Sloan. 1980. Trends in levels of several known chemical
contaminants in fish from New York State waters. Bur. Env. Prot. NY
State Dept. Environmental Cons., fllbany, New York. 77 pp.
flrmstrong, R. and R. Sloan. 1982. PCB patterns in Hudson River fish. I.
Resident/Freshwater species. Proc. Hudson River Env. Soc., Hyde Park, New
York.
Ballschmitter, K., and M. Zell. 1980. Onalysis of polychlorinated biphenyls
-------
Bigelow, H., and W. Schroeder. 1953. Fishes of The Sulf of Maine. Fish.
Bull. 74: 1-577
Bopp, R. 1979. The geochemistry of polychlorinated biphenyls in the Hudson
River. Ph.D. Diss. Columbia Univ. Lament-Doherty Geological
Observatory. Palisades, New York. 191 pp.
Bopp, R., H. Simpson, C. Olsen and N. Kostyk. 1981. Polychlorinated biohenyls
in sediments of the tidal Hudson River, New York. Environ. Sci. Technol.
15: £10-216.
Bopp, R., H. Simpson, B. Deck and N. Kostyk. 1984. The persistence of PCS
congeners in sediments of the lower Hudson. Northeastern Env. Sci. 3: 180-
184.
Branson, D., G. Blau, H. Alexander and W. Neely. 1975. Bioconcentration of
2,£',4,4'-tetrachlorobiphenyl in rainbow trout as measured by an
accelerated test. Trans, flmer. Fish. Soc. 104: 785-792.
Breder, C., and D. Rosen. 1966. Modes of Reproduction in Fishes. American
Museum of Natural History, New York. Natural History Press. 941 pp.
Briggs, J. I960. Fishes of world-wide (circumpolar) distribution. Copeia
I960: 171-180.
Brown, J., R. Wagner, D. Bedard, M. Brennan, J. Carnahan, and R. May. 1984.
PCS transformations in upper Hudson sediments. Northeast. Env. Sci. 3:
167-179.
Brown, M., J. McLaughlin, J. O'Connor and K. Wyman. 1982. A mathematical
model of PCB bioaccumulation in plankton. Ecol. Modelling 15: £9—47.
Brown, M., M. Werner, R. Sloan and K. Simpson. 1985. Recent trends
in the distribution of polychlorinated biphenyls in the Hudson River
system. Environ. Sci. Technol. 19: 656.
Bruggernan, W., L. Marton, D. Kooiman and 0. Hutzinger. 1981. Accumulation and
elimination kinetics of di-, tri-, and tetrachlorobiphenyls by goldfish
after dietary and aqueous exposure. Chemosphere 10: 811-832.
Brungs, W. and D. Mount. 1978. Introduction to a discussion of the use of
aquatic toxicity tests for evaluation of the effects of toxic substances.
In: J.Cairns, K. Dickson and A. Maki (eds.) Estimating the Hazard of
Chemical Substances t^o Aquatic Life. Philadelphia, ASTM Press, pp. 15-26
Bush, B., J. Snow and S. Conner. 1983. High resolution gas chromatographic
analysis of non-polar chlorinated hydrocarbons in human milk. J. Assoc.
Off. final. Chern. 66: £48-255.
Cahn, P., J. Foehrenbach and W. Guggino. 1977. PCB levels in certain organs
of some feral fish from New York State. In: F. Vernberg, A. Calabrese, F.
Thurberg and W. Vernberg (eds.). Physiological Responses of Marine Biota
t£ Pollutants. New York, Academic Press. pp. 51-61.
-------
Califano, R. J. 1981. Occurnulat ion and distribution of DOlychlorinated
biphenyls
-------
Fuller, B., J. Gordon and M. Kornreich. 1976. Environmental Assessment of
PCBs in the Atmosphere. MTR-7210, Rev. 1. The Mitre Corp. McLean, Va.
Furukawa, K. 1982. Microbial degradation of polychlorinated biphenyls: .
In: A. Chakrabarty (ed. ) Biodegradation and Detoxification of
Environmenta1 Pollutants. CRC Press, Boca Raton Fl. pp. 33-57.
Furukawa, K., and F. Matsurnura. 1976. Microbial metabolism of polychlorinated
biphenyls: Studies on the relative degradability of polychlorinated
biphenyls components by Plkiliqenes sp. J. Agric. Food Chem. 24: 251-£56.
Gage, J., and S. Holm. 1976. The influence of molecular structure on the
retention and excretion of polychlorinated biphenyls by the mouse.
Toxicol. Appl. Pharmacol. 36: 555-560.
GESAMP (Joint Group of Experts on the Scientific Aspects of Marine Pollution)
1982. The Health of the Oceans. UNEP Regional Seas Repts. and Studies
No. 16. 108 pp.
Goldstein, A. L. Aronow and S. Kalman. 1974. Principles of Drug Action; The
Basis of Pharmacology. New York. J. Wiley and Sons. 854 pp.
Goldstein, J., P. Hickman, H. Bergman, J. McKinney and M. Walker. 1977.
Separation of pure polychlorinated biphenyl isomers into two types of of
inducers on the basis of induction of cytochrome P-450 or P-448. Chem.
Biol. Interact. 17: 69-87.
Grosslein, M., and T. Azarovitz. 1982. Fish Distribution. New York Bight
Atlas, Monograph 15. Albany, New York. New York Sea Grant Institute.
182 pp.
Guiney, P., and R. Peterson. 1980. Distribution and elimination of
polychlorinated biphenyl after acute dietary exposure in yellow perch and
rainbow trout. Arch. Env. Contain. Toxicol. 9:667-674.
Guiney, P., R. Peterson, M. Melancon and J. Lech. 1977. The distribution and
elimination of 2,2',5,5'-<14-C)-tetrachlorobiphenyl in rainbow trout (Salmo
gairdneri). Toxicol. Appl. Pharmacol. 39: 329-338.
Hamelink, J., R. Wybrant and R. Ball. 1971. A proposal: exchange equilibria
control the degree chlorinated hydrocarbons are biologically magnified in
lentic environments. Trans. Amer. Fish. Soc. 100: 207-214.
Hansen, D., P. Parrish, J. Lowe, A. Wilson, Jr. and P. Wilson. 1971. Chronic
toxicity and uptake and retention of.Aroclor 1254 in two estuarine fishes
Bull. Env. Contarn. Toxicol. 6:113-119.
Hansen, D., P. Parrish and J. Forester. 1974. Aroclor 1016: Toxicity and
uptake by estuarine animals. Env. Res. 7: 363-373.
Hansen, L., W. Wiekhorst and J. Simon. 1976. Effects of dietary Aroclor 1242
on channel catfish (Ictalurus punctatus) and selective accumulation of
PCB components. J. Fish. Res. Bd. Canada 33: 1343-1352.
-------
Hague, R., D. Schmedding, and V. Freed. 1974. ftqueous solubility, adsorption
and vapor behavior of polychlorinated biphenyl flroclor 1254. Environ.
Sci. Technol. 8: 139-142.
Hartsbrough, 6. 1965. Great Lakes fish now susoect as mink food, flrner. Fur
Breeder 38: 25-27.
Hesse, S., M. Metzger and J. Wolff. 1978. Rctivation of (14-C)chlorobiphenyls
to protein-binding metabolites by rat liver microsomes. Chem. Biol.
Interact. 20: 355-365.
Hetling, L., T. Tofflernire, E. Horn, R. Thonas and R. Mt. Pleasant. 1979. The
Hudson River PCB problem: Management alternatives. ftnn. N. Y. Ocad. Sci.
3£0: 630-650.
Hilton, J., P. Hodson, H. Braun, J. Leatherland and S. Slinger. 1983.
Contaminant accumulation and physiological response in rainbow trout
(Salmo gairdneri) reared on naturally contaminated diets. Can. J. Fish.
fiquat. Sci. 40: 1987-1994.
Hoeting, «. 1983. FDO regulation of PCB in food. In: F. D'Itri and M.
Kamrin (eds). PCBs; Human and Environmental Hazards. ftnn ftrbor. Onn
flrbor Science Press, pp. 393-408.
Horn, E., L. Hetling and T. Tofflemire. 1979. The problem of PCBs in the
Hudson River system. finn. N.Y. Bead. Sci. 320: 591-609.
Horn, E., and L. Skinner. 1985. Final Environmental Impact Statement for
Policy on Contaminants in fish. Div. Fish and Wildlife, NY State Dept. of
Env. Cons., fllbany, New York. 150 pp.
Hubbard, H.L. 1964. Chlorinated biphenyl and related compounds. In:
Encyclopedia of Chemical Technology. 2 ed. 5: 289-298.
Hutzinger, 0., D. Nash, S. Safe, fl. DeFreitas, R.Norstrom, D. Wildish and V.
Zitko. 1972. Polychlorinated biphenyls: Metabolic behavior of pure
isomers in pigeons, rats and brook trout. Science 178: 312-314.
Hutzinger, 0., S. Safe and V. Zitko. 1974. The Chemistry of PCBs. CRC Press,
Cleveland, Ohio. 269 pp.
International Agency for Research on Cancer (IORC). 1974. IflRC Monograph on
the carcinogenic risk of chemicals to humans. 7:241.
International flgency for Research on Cancer (IftRC). 1978. IflRC Monograph on
the carcinogenic risk of chemicals to humans 18:43.
Interstate Electronics Corporation. (IEC) 1979. Environmental Impact Statement
for
New York Dredged material ocean disposal site designation. Rept. to U.S.
EPft, fipril 1981. IEC, flnaheim, CO. 148 pp.
-------
Isaacs, J. 1973. Potential trophic biomasses and trace substance
concentrations in unstructured marine foodwebs. Mar. Biol. 22: 97-104.
Jelinek, C. and P. Corneliussen. 1976. Levels of PCBs in the U.S. food
supply. Proc. Nat'1 Conf. on PCBs. EPfi-560/6-75-004. pp. 147-154.
Jensen, ft., S. Spigarelli and M. Thommes. 1982. PCB uptake by five species of
fish in Lake Michigan and Cayuga Lake, New York. Can. J. Fish. ftquat.
Sci. 39: 780-709.
Jensen, S. 1966. Report of a new chemical hazard. New Sci. 32: 612
Johansson, N., ft. Larsson and K. Lewander. 1972. Metabolic effects of PCB
(polychlorinated biphenyls) on the brown trout (Sa1mo trutta). Comp. gen.
Pharrnacol. 1972: 310-314.
Karickhoff, S., D. Brown, and T. Scott. 1979. Sorption of hydroohobic
pollutants on natural sediments. Water Res. 13: 241-248.
Kato, S., J. McKinney and H. Matthews. 1980. Metabolism of symmetrical
hexachlorobiphenyls isomers in the rat. Toxicol. flppl. Pharmacol. 53: 389-
398.
Kendall, fl., and L. Walford. 1979. Sources and distribution of bluefish,
Pornatomus saltatrix. larvae and juveniles off the east coast of the United
States. Fish. Bull. 77: 213-227.
Kimbrough, R., R. Squire, R. Linder, J. Strandberg, R. Montali and V. Burse.
1975. Induction of liver tumors in Sherman strain female rats by
polychlorinated biphenyl Rroclor 1260. J. Nat'1. Cancer Inst. 55: 1453-
1459.
Klaunig, J., M. Lipsky , B. Trump and D. Hinton. 1979. Biochemical and
ultrastructural changes in teleost liver following subacute exposure
to PCB. J. Environ. Pathol. Toxicol. 2: 953-963.
Krzeminski, S., J. Gilbert and J. Ritts. 1977. ft pharmacokinetic model for
predicting pesticide residues in fish, flrch. Env. Contain. Toxicol. 5: 157-
166.
Koeman, J., and M. Stasse-Wolthius. 1978. Environmental toxicology of
chlorinated hydrocarbon compounds in the marine environment of europe.
Comm. Eur. Comm. EUR 5814 en. 137 pp.
Kolbye, ft., and C. Carr. 1984. The evaluation of the carcinogenicity of
environmental substances. Reg. Toxicol. Pharmacol. 4: 350-354.
Kuratsune, M., Y. Masuda and J. Nagayama. 1976. Some of the recent findings
concerning Yusho. Proc. National Conf. of Polychlorinated Biphenyls. EPfl-
560/6-75-004. pp. 14-29.
-------
Lech, J., and J. Bend. 1980. Relationship between biotransformation and the
toxicity and fate of xenobiotic chemicals in fish. Env. Health Persp. 34:
115-131.
Lech, J., and R. Peterson. 1983. Biotransformation and persistence of
polychlorinated biphenyls (PCBs) in fish. In: F. D'Itri and M. Kamrin
(eds.). PCBs; Human and Environmental Hazards, finn ftrbor, ftnn flrbor
Science Press. pp. 187-202.
Lee, G. and R. Jones. 1977. fin assessment of the environmental significance
and chemical contaminants present in dredged sediments dumped in the New
York Bight. Rept. to U.S. firmy Engineers, New York District, NY. &£ pp.
Lieb, fi. , D. Bills and 0. Sinhuber. 1974. ficcumulation of dietary
polychlorinated biphenyls (firoclor 1254) by rainbow trout (Salmo
gairdneri). J. figr. Food. Chem. 22: 638-642.
Lipsky, M., J. Klaunig and D. Hinton. 1978. Comparison of acute response tp
polychlorinated biphenyl in liver of the rat and channel catfish: ft
biochemical and morphological study. J. Toxicol. Environ. Health 4: 107-
121.
*
Lowman, F., T. Rice and F. Richards. 1971. ficcumulation and redistribution of
radionuclides by marine organisms. In: ft. Seymour (ed.) Radioactivity
rn the Marine Environment. Washington, D.C. National ftcademy Press. pp.
161-199.
Macek, K., S. Petrocelli and B. Sleight, III. 1979. Considerations in
assessing the potential for, and significance of, biomagnification of
chemical residues in aquatic food chains. In: L. Marxins and R. Kirnmerle
(eds). flguatic Toxicology. ftSTM STP 667. Philadelphia, fimer. Soc. for
Testing Materials. pp. 251-268.
Mackay, D. 1982. Correlation of bioconcentration factors. Env. Sci. Technol.
16: 274-278.
Mackay, D., and fi. Hughes. 1984. Three-parameter equation describing the
uptake of organic compounds by fish. Environ. Sci. Technol. 18j 439-441.
MacLeod, W., L. Ramos, ft. Friedman, D. Burrows, P. Prohaska, D. Fisher and D.
Brown. 1981. Onalysis of residual chlorinated hydrocarbons and arowatic
hydrocarbons and related compounds in selected sources, sinks and biota of
New York Bight. NOfift Tech. Memo. OMPft-6. NOfifi/OMPfl,Boulder, CO. 128 pp.
Mancini, J. 1983. fi method for calculating effects, on aquatic organisms, of
time-varying concentrations. Water Res. 17: 1355-1362.
Matsuo, M. 1980. ft therrnodynamic interpretation of bioaccumulation of ftroclor
1254 (PCB) in fish. Chemosphere 9: 671-675.
Mauck, W., P. Mehrle and F. Mayer. 1978. Effect of the polychlorinated
biphenyl ftroclor 1254 on growth, survival and bone development in brook
trout (Salveil mis fontinalis). J. Fish. Res. Bd. Canada 35: 1084-1088.
-------
Mayer, F.L., P. Mehrle, and H.0. Sanders. 1977. Residue dynamics and
biological effects of polychlorinated biphenyls in aquatic organisms.
flrch. Env. Contam. Toxicol. 5: 501-511.
McKirn, J., and E. Heath. 1983. Dose determinations for waterborne 2, 5,2', 5'-
(14-C)-tetrachlorobiphenyl in two species of trout (Salmo gairdneri and
Salvelinus fontinalis): A mass balance approach. Toxicol. Appl.
Pharmacol. 68: 177-187.
Mehrle , P., T. Haines, S. Hamilton, L. Ludke, F. Mayer and M. Ribick. 198£.
Relationship between body contaminants and bone development in east-coast
striped bass. Trans, Arner. Fish. Soc. Ill: 231-241.
Melancon, M., and J. Lech. 1976. Isolation and identification of a polar
metabolite of tetrachlorobiphenyl from bile of rainbow trout exposed to
14-C tetrachlorobiphenyl. Bull. Env. Contam. Toxicol. 15: 181-187.
Metcalf, R., J. Sanborn, P. Lu and D. Nye. 1975. Laboratory model ecosystem
studies of the degradation and fate of radiolabelled tri-, tetra-, and
pentachlorobiphenyl compared with DDE. Arch. Env. Contam. Toxicol. 3: 151-
Mitchell, A., P. Plack and I. Thompson. 1977. Relative concentrations of 14-C
DDT and of two polychlorinated biphenyls in the lipids of cod tissue after
a single oral dose. Arch. Env. Contam. Toxicol. 6: 525-532.
Monsanto Corp. 1978. Polychlorinated biphenyls (PCBs): A reoort on uses,
environmental health effects and disposal. The Monsanto Corp., St. Louis.
Mo. 29 pp.
Moriarty, F. 1975. Exposures and Residues. In: F. Moriarty
-------
Nau-Ritter, S. 1980. The dynamics of PCB transfers among marine
phytoplankton, clay particles and water. MS Thesis, State Univ. of NY at
Stony Brook. Marine Sciences Res. Center. 117 pp.
Neely, W., D. Branson and G. Blau. 1974. Partition coefficients to measure
bioconcentration potential of organic chemicals in fish. Environ. Sci.
Technol. 8: 1113-1115.
Nelson, N. 1972. PCBs-Environrnental Impact. Env. Res. 5: 249-362
New York State Department of Environmental Conservation
-------
O'Connor, J., and J. Pizza, in press, a. Pharmacokinetic model for the
accumulation of PCBs in Marine Fish. In: Oceanic Processes in Marine
Pollution, Vol. 1. Capuzzo, J., and D. Kester (eds. ) Biological Processes
and Wastes in the Ocean. Krieger Pub.
O'Connor, J., and J. Pizza, in press, b. PCB dynamics in Hudson River
striped bass. III. Tissue disposition and routes for elimination.
Estuaries.
O'Connor, J., J. Klotz and T. Kneip. 1982. Sources, sinks and distributionof
organic contaminants in the New York Bight ecosystem. In: 6. Mayer
(ed.) Ecological Stress and the New York Bight. Estuarine Research
Federation, Charleston, 5.C. pp. 631-653.
Okumura, M., and S. Katsuki. 1969. Clinical observation on Yusho (Chloro-
biphenyl poisoning). Fukuoka Act a Med. 60: 440-446.
Oliver, B., and A. Niimi. 1985. Bioconcentration factors of some halogenated
organics for rainbow trout: Limitations in their use for prediction of
environmental residues. Environ. Sci. Technol. 19: 842-849.
Pavlou, S., and R. Dexter.- 1979. Distribution of polychlorinated biphenyls
(PCBs) in estuarine systems. Testing the concept of equilibrium
partitioning. Environ. Sci. Technol. 13:65-71.
Pequegnat, W., B. James, E. Kennedy, P. Fredericks, R. Fay and F. Hubbard.
1980. Application of the biotal ocean monitor system to a study of the
impacts of ocean dumping of dredged material in the New York Bight.
TerEco Report to U.S. Army Engineers, NY District. 61 pp.
Pizza, J. 1983. Pharmacokinetics and distribution of dietary PCBs in Hudson
River striped bass, Morone saxatilis. Ph.D. Diss. New York University,
New York. 109 pp.
Pizza, J. and J. O'Connor. 1983. PCB dynamics in Hudson River striped bass.
II. Accumulation from dietary sources. Aquatic Toxicol. 3: 313-327.
Preston, B., and J. Allen. 1980. 2,2',5,5'-tetrachlorobiphenyl: Isolation and
identification of metabolites generated by rat liver microsomes. Drug
Metab. Disp. 8: 197.
Reichardt, P., B. Chadwick, M. Cole, B. Robertson and D. Button. 1981.
Kinetic study of biodegradation of biphenyl and its monochlorinated
analogues by a mixed marine microbial community. Env Sci. Technol. 15: 75-
79.
Richardson, B., and J. S. Waid. 1982. Polychlorinated biphenyls (PCBs): On
Australian viewpoint on a global problem. Search 13: 17-25.
Ringer, R. 1983. Toxicology of PCBs in mink and ferrets. Ins F. D'Itri and
M. Kamrin (eds). PCBs; Human and Environmental Hazards, Ann flrbor. Ann
Arbor Science Press, pp. 227-240.
-------
Risebrough, R., P. Reichle, S. Herman, D. Peakall and M. Kirven. 1968.
Polychlorinated biphenyls in the global ecosystem. Nature 220: 1098-1102.
Rubinstein, N., E. Lores and N. Gregory. 1983. Pccumulation of PCBs, mercury
and cadmium by Nereis virens. Mercenaria mercenaria and Palaemonetes puoio
from contaminated harbor sediments, Oquatic Toxicology 3:249-260.
Rubinstein, N., N. Gregory and W. Gilliam. 1984. Dietary accumulation of PCBs
from a contaminated sediment source by a demersal fish species, Leiostomus
xanthurus. Aquatic Toxicol. 5: 331-342.
Safe, S. 1984. Polychlorinated biphenyls (PCBs) and polybrominated biphenyls
(PBBs): Biochemistry, toxicology and mechanism of action. CRC Crit. rev.
Toxicol. 13: 319-393.
Samuelian, J., M. Moese and J. O'Connor. MS in review. PCB congeners in the
tomcod (Microoadus tomcod) from the East River, NY. NYU Inst. Env. Med.
Tuxedo, NY.
Sangalang, G. H. Freeman and R. Crowell. 1981. Testicular abnormalities in
Cod (Gadus morhua) fed ftroclor 1254. Orch. Env. Contain. Toxcol. 10: 617-
626.
Schroeder, R., and C. Barnes. 1983. Trends in polychlorinated biphenyl
concentrations in Hudson River water five years after elimination of point
sources. U.S. Geol. Survey Water Res. Invest. Rept. 83-4026. 28 pp.
Schwartz, L. 1936. Dermatitis from synthetic resins and waxes, flm. J. Pub.
Health. 26: 586-592.
Schwartz, P., S. Jacobson, G. Fein, J. Jacobson and H. Price. 1983. Lake
Michigan fish consumption as a source of polychlorinated biphenyls in
human cord serum, maternal serum and milk. flm. J. Public Health 73: 293-
296.
Scura, E., and G. Theilacker. 1977. Transfer of the chlorinated hydrocarbon
PCB in a laboratory marine food chain. Mar. Biol. 40: 317-325.
Shaw, G., and D. Connell. 1980. Relationships between steric factors and
bioconcentration of polychlorinated biphenyls (PCBs) by the sea mullet
(Muuil cephalus Linnaeus). Chemosphere 9: 731-743.
Shaw, G., and D. Connell. 1984. Physicochemical properties controlling
polychlorinated biphenyl (PCB) concentrations in aquatic organisms. Env.
Sci. Technol. 18: 18-23.
Sherwood, M., 0. Mearns, D. Young, B. McCain, R. Murchelano, G. Olexander, T.
Heesen and Tsu-Kai. 1978. A comparison of trace contaminants in diseased
fish from three areas. Rept. to NOftfl/MESfi New York Bight Project.
Southern California Coastal Water Research Project, El Segundo, Cfl. 116
pp.
-------
Shimada, T., Y. Imai and R. Sato. 1981. Covalent binding of polychlorinated
biphenyls to proteins by reconstituted monooxygenase system containing
cytochroroe P-450. Chem. Biol. Interact. 38: £9-44.
Sittig, M. 1985. Handbook of Toxic and Hazardous Chemicals and Carcinogens.
Park Ridge, N. J. Noyes Press. 9519 po.
Sleight, S.D. 1983. Pathologic effect of PCBs in mammals. In: F. D»Itri and
M. Kamrin (eds). PCBs; Human and Environmental Hazards, flnn flrbor, flnn
flrbor Science Press, pp. £15-£26
Sloan, R., and R. flrrnstrong. 198£. PCB patterns in Hudson River fish. II.
Migrant/Marine species. Proc. Hudson River Env. Soc. Hyde Park, N.Y. 53
pp.
Sloan, R., K. Simpson, R. Schroeder and C. Barnes. 1983. Temporal trends
toward stability of Hudson River PCB contamination. Bull. Env. Contam
Toxicol. 31: 377-385.
Sloan, R., M. Brown, R. Brandt, and C. Barnes. 1984. Hudson River PCB
relationships between resident fish, water and sediment. Northeast. Env.
Sci. 3:138-15£.
Smith, V., J. Spurr, J. Filkins and J. Jones. In press. Organochlorine
contaminants of wintering ducks foraging on Detroit River sediments. J.
Great Lakes Res. 1985.
Spacie, fl., and J. Hamelink. 1982. Alternative models for describing the
bioconcentration of organics in fish. Environm. Toxicol. Cheni. 1: 309-
3£0.
Spagnoli, J., and L. Skinner. 1977. PCBs in fish from selected waters of New
York State. Pestic. Mon. Journal. 11: 69-87.
Spies, R., J. Felton, and L. Dillard. 1984. Hepatic mixed-function oxidases
in California flatfishes are increased in contaminated environments and by
oil and PCB ingest ion. Mar. Biol. 70: 117-1£7.
Stainken, D., and J. Rollwagen. 1979. PCB residues in bivalves and sedinents
of Raritan Bay. Bull. Env. Centam. Toxicol. £3: 690-697.
Stallings, D., and F. Mayer. 197£. Toxicity of PCBs to fish and environmental
residues. Env. Health Perspect. li 159-164.
Stegeman, J., P. Kloepper-Sams and J. Farington. 1986. Monooxygenase induction
and chlorobiphenyls in the deep sea fish Coryphaenoides armatus. Science
£31: 1£87-1£89.
Stein, J., T. Horn and U. Varanasi. 1984. Simultaneous exposure of English
sole
-------
Subcommittee on Health Effects of PCBs and PBBs, U.S. Dept. of Health,
Education and Welfare. 1976. Final Report. 193 pp. + appendices.
Suflita, J., J. Robinson and J. Tiedje. 1983. Kinetics of microbial
dehalogenation of haloaromatic substrates in methanogenic environments.
flppl. Env. Microbiol. 45: 1466-1473.
Swain, W.R. 1983. fin overview of the scientific basis for concern with
polychlorinated biphenyls in the great lakes. In: F. D'Itri and M. Kamrin
(eds.) PCBs; Human and Environmental Hazards. ftnn ftrbor, ftnn flrbor
Press, pp. 11-48.
Thornann, R. 1981. Equilibrium model of fate of microcontaminants in diverse
aquatic food chains. Can. J. Fish, fiquatic Sci. 38: 280-296.
Thomann, R., and J. St. John. 1979. The fate of PCBs in the hudson River
ecosystem. finn. N.Y. ficad. Sci. 320: 610-629.
Thomann, R and J. Connolly. 1984. Model of PCBs in the Lake Michigan Lake
trout food chain. Environ. Sci. Technol. 18: 65-71.
Tucker, E., V. Saeger and 0. Hicks. 1975. Activated sludge primary
degradation of polychlorinated biphenyls. Bull. Env. Contain. Toxicol. 14:
705-712.
Walker, C.R. 1976. The occurrence of PCB in the National fish and wildlife
monitoring program. In: Proceedings of the National Conference on
Polychlorinated Biphenyls. EPft-560/6-75-004. pp.161-176.
Wassermai, M., D. Wasserman, S. Cucos and H. Miller. 1979. World PCBs map:
Storage and effects in man and his biologic environment in the 1970's.
flnn. NY flcad. Sci. 320: 69-124.
Weaver, G. 1984. PCB contamination in and around New Bedford, Mass. Env. Sci.
Technol. 18: 22fl-27ft.
Weininger, D. 1978. accumulation of PCBs by lake trout in Lake Michigan.
Ph.D. Diss. Univ. Wisconsin-Madison. 232 pp.
West in, D., C. Olnney and B. Rogers. 1985. Effects of parental and dietary
organochlorines on survival and growth of striped bass larvae. Trans. A*.
Fish. Soc. 114: 125-136.
Woodwell, G., C. Wurster and P. Isaacsson. 1967. DDT residues in an east
coast estuary: ft case of biological concentration of a persistent
pesticide. Science 156: 821-824.
Young, D. 1984. Methods of evaluating pollutant biomagnification in marine
ecosystems. In: H. White (ed.) Concepts in Marine Pollution
Measurements. Maryland Sea Grant Press, College Park, pp. 261-278.
-------
Table 1. Bioconcentration of various fVoclors in fishes. Bioconcentration factor calculated as the
concentration in the fish divided by the concentration in the uatar.
Organism
Channel Catfish
Uctalurus punctatus)
Bluegill sunfish
(Lepoeis nacrochirus^
Brook trout (fry)
(Salve linus fontinalis)
Spot
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Table 2. Calculation of estimated PCS body burdens in fishes based uoon
equilibrium partitioning. The data used are from the New York Bight
and adjacent marine waters.
Minimum value Maximum value
Water Column PCB Concentration
(ng/1) (Note 1)
Particulate/Dissolved Ratio
(Note 2)
Dissolved (available) PCB (ng/1)
Bioconcentration Factor
(Note 3)
0.67
6.7
40
0.67
27
10,000
Expected Concentration in Fish
(ug/g wet weight)
Observed Concentrations (ug/g)
(Note 4)
0.07
0.27
Striped bass
Winter flounder
Mackerel
Bluefish
American eel
Tautog
0.6 - 3.8
0.1
0.5 - 0.7
0.7 - 3.&
0.5 - 0.8
0.6
Note 1. Concentrations from Lee and Jones (1978), IEC (1979), Pequegnat et al.
(1980) and MacLeod et al. (1981)
Note 2. Various authors suggest particulate/dissolved ratios from 0.0 to 1.0.
The value of 0.67 was arrived at based upon data from Brown et al.
(1982), Nau-Ritter (1980) and Pavlou and Dexter (1979).
Note 3. The value of 10,000 was based upon BCF data ranging from 16,000 to
61,000 for various species. Assuming some portion of the BCF was from
feeding and water ingestion we concluded 10,000 to be a reasonable BCF
approximation.
Note 4. Observed concentration data from O'Connor et al., 1982; NYSDEC, 1981;
NJ DEP, 1982.
-------
Table 3. Height of fish, weight of stomach contents and calculation of PCB dose for a sample of age class 1+
striped bass collected at Canal Street, Manhattan. The ratio of stooach content to weight of fish
was used to calculate a daily food ration and an expected rate of ingest ion of PCB with the food.
The calculated doses were based upon «ean ration and nean PCB content of the food. The regression of
PCS burden
8.6
8.5
10.3
10.9
12.5
11.4
11.9
12.6
14.3
16.7
21.5
12.1
21.3
19.4
20.1
25.9
23.1
25.5
24.0
26.2
33.0
WEIGHT OF
FOOD
0.68
0.96
0.71
0.87
1.35
1.41
1.70
1.45
1.50
1.37
1.06
2.07
1.49
1.44
1.68
2.71
2.13
2.12
1.37
3.33
2.81
RATIO FOOD/
FISH MT.
0.02
0.03
0.02
0.02
0.03
0.03
0.03
0.03
0.02
0.02
0.03
0.02
0.02
0.02
0.03
0.02
0.02
0.03
0.01
0.03
0.02
PCB FISH
-------
Table 4. Distribution of 14-C-labelled flroclor 1254 aaong tissues and organs of young-of the year striped
bass measured 48 hours after administration of 1, 2, and 3 doses of PCB in the food. Each dose
was 387 ng PCB. Data are presented as the mean of 5 fish C+ standard error of the mean).
Doses
Given
1
Cn = 5)
Percent
retained
ug PCB/g
of
burden
Cdry)
Percent of cum-
ulative dose
2
3
Cn « 5>
Percent
retained
ug PCB/g
Percent
ulative
Cn = 5)
Percent
retained
ug PCB/g
Percent
ulative
of
burden
Cdry)
Of CUfli-
dose
of
burden
Cdry)
of cum-
dose
Gill
2.47
C0.38)
0.33
CO. 06)
1.92
CO. 42)
2.44
0.53
1.61
C0.26)
2.10
C0.22)
0.74
CO. 07)
1.25
CO. 18)
Liver and
6allbl«dder
5.
1.
CO.
4.
6.
2.
3.
6.
4.
3.
CO.
94
66)
51
17)
45
43)
12
88)
98
23)
89
33)
15
34)
47
58)
63
31)
Alimentary
Tract
5.
0.
CO.
4.
CO.
5.
I.
3.
CO.
6.
1.
3.
CO.
35
54
06)
00
25)
64
10
66
34)
48
73
83
71)
Spleen and
Heart
0.57
CO. 08)
0.34
CO. 06) '
0.42
CO. 04)
0.58
CO. 11)
0.95
CO. 13)
0.36
CO. 04)
0.56
CO. 04)
0.79
CO. 04)
0.34
CO. 04)
Head
28.54
Cl.OO)
0.41
C0.04)
21.70
C1.82)
30.11
C1.12)
0.69
CO. 15)
19.48
C1.40)
27.61
CO. 41)
1.01
CO. 08)
16.25
C0.82)
Epaxi-al
Carcass Muscle
57. 14
C1.38)
0.32 0.26
C0.03) CO. 04)
46.14
C5.40)
55.11
C1.90)
0.54 0.53
CO. 09) CO. 03)
36.21
C4.92)
57.09
C1.32)
0.87 0.85
C0.07) C0.07)
33.60
C2.08)
Whole
Fish
100
0.37
C0.04)
76.24
C6.26)
100
0.63
CO. 11)
65.23
C6.90)
100
0.98
CO. 08)
58.91
C3.28)
-------
Table 5. PCB concentrations determined in bluefish, from the waters of New
York, New Jersey and Massachusetts, 1979 through 1983.
Location of sampling
Dates
Number of fish
PCB Concentration
Peekskill, NY (a)
New York Harbor
Fire Island, NY (a)
Cold Soring Hr. , NY (a)
Eastern L.I. Sound (a)
Herod Pt., NY (a)
Orient Pt., NY (a)
Great South Bay (a)
Hudson River, NJ (b)
Hudson River, NJ (b)
Newark Bay, NJ (b)
Raritan River, NJ data from Belton et al., 1983; (c) data from Belton
et al., 1985; (d) data from Weaver, 1984
-------
Figure 1. The distribution of bluefish on the Atlantic Coast of the U. S
Solid circles indicate the yield from trawl catches oerformed
by the National Marine Fisheries Service. Lightly hatched area
shows the general spawning area during the summer months, and
the strongly hatched area shows areas of concentrated summer
spawning. From Grosslein and Qzarovitz, 1982.
-------
\flV' MASS o
X
/• y
General spawning - summer
r'
<•>'/"<* ' '
>'••;*# concentrated spawning summer
;-^r c *e
< 5 Ibs.
6 -20 Ibs
21 -100 Ibs.
101 -1000 Ibs.
-------
Figure 2. The structure of the biohenyl molecule, with ortho-, rneta-
and para- positions labelled for the primary and secondary
rings.
-------
meta ortho
ortho meta
para <4
4) para
meta ortho
ortho meta
FIGURE 2. The structure of biphenyl.
-------
Figure 3. Schematic diagram of transport pathways for PCBs in the
environment, with pathways from various manufacturing and
applications processes to environmental media labelled. Note
that the primary receptor for PCBs from all processes is water
(W), whereas the least common transport end point is
destruction (D). fill transport pathways leading to air (ft) have
the potential for PCB transport to the water via surface
runoff, wet fallout and dry fallout. Diagram from Nisbet and
Sarofirn, 1972.
-------
n-burnIng
itock of PCI
containing product*
(quantity unknown)
o
Reservoir
Flow Path
-{> Route Into the air (A),
water (V), or terrestrial
(T) environment;
(D) - destroyed.
Adapted from Hlsbet, I. C. T. and A. F. Saroflv, "Rates and
Routes of Transport of PCBs In the Environment," Environmental
llf.ilth Perspectives fxp. 1. 21-38, 1972.
-------
Figure 4. Schematic diagram of the transoort of PCBs in a typical food
chain showing the relationship of water uptake, assimi lat ion
from the food and the effects of metabolism. The schematic was
prepared in conjunction with a second oortion of the model
(blow the dotted line) describing transoort of PCBs in the
physical compartments of the ecosystem. From Thornann, 1981.
-------
SPECIFIC
CONSUMPTION
(q/q d)
RESPIRATION
(d")
x
o
UJ
AGE
TOXICANT
CONCENTRATION
IN PHYTOPLANKTON
(D
Predation
TOXICANT
CONCENTRATION
IN ZOOPLANKTON
(/*g/g) (2)
>'
Predat/on
TOXICANT
CONCENTRATION
IN SMALL FISH
(Mq/q) (3)
Predation
TOXICANT
CONCENTRATION
IN LARGE FISH
'AVAILABLE" (DISSOLVED) CHEMICAL WATER CONCENTRATION (Mg/D
PHYSOL-CHEMICAL
MODEL OF
PARTICIPATE AND DISSOLVED
CONCENTRATIONS
-------
Figure 5. Removal of PCBs (ftroclor 1254) from the gut of striped bass
dosed with radiolabelled compound and sampled at intervals for
5 days. PCBs are recorded as the percentage of the dose
administered to the fish at time zero. Although more than 90%
of the dose had been lost from the gut within 24 hours, the
whole body samples showed that the majority of the dose had
been distributed from the out to the tissues. From Pizza and
0'Connor, 1983.
-------
O)
iT>
O
O
.0
o
10
O)
o
O)
Q_
Phase 1
95% C.I. for Ka= 0.0850 to 0.1212 hr
-1
"I-
Phase 2
*
t-
I I
24
48 72
Time (hr.)
... .1.
96 120
-------
Figure 6. approximate isostereomers of 2,3,7,8-tetrachlorodibenzo-o-
dioxin (TCDD) as halogenated biphenyls, halogenated
azobenzenes and halogenated dibenzofurans. In all cases the
molecular size, shape and plananty are sufficiently similar to
TCDD to lead to the conclusion that the compounds should have
similar biochemical effects. From Safe, 1384.
-------
3,3,4,4,5 - Pentahalobiphenyl 3,3,4,4,5,5 - Hexahalobiphenyl
Cl
Cl
Cl
N
\\
N
Cl
3,3',4,4' - Tetrahalobiphenyl 3,3*. 4,4'- Tetrachloroazo benzene
Cl
Cl
Cl
Cl
2,3,7,8 - Tetrachlorodlbenzo-p-dioxin 2,3,7,8-Tetrachlorodlbenzofuran
-------
Figure 7. Fishing advisory areas in the vicinity of MetroDdlitan New
York and New Jersey. The advisory from the State of New Jersey
warns against consuming fish from coastal marine waters due to
their high PCB concentrations. In 1986 New York State banned
all possession (recreational and commercial) of strioed bass in
all marine waters of the state due to high PCB concentrations.
Figure from Belton et al., 1983.
-------
New York
New Jersey
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FISHING ADVISORY AREA
DUE TO PCB's IN FISH TISSUE
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Hudson River
Upper New York Bay
Newark Bay
Tidal Passaic River
Tidal Hackensack River
Arthur Kill
Kill Van Kull
Tidal Raritan River
Raritan Bay
Sandy Hook Bay
Lower New York Bay
STRIPED BASS and BLUEFISH
advisory includes Offshore Waters
for Northern Costal Area.
Ijjjjjjjjjjjl; AMERICAN EEL advisory includes
all waterways statewide.
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