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This is to certify that the

thesis entitled

EFFECTS OF FEEDING FISH COLLECTED DOWNSTREAM
FROM OAK RIDGE RESERVATION ON THE
REPRODUCTIVE PERFORMANCE OF MINK

presented by

Lorin Alyn Lewis

has been accepted towards fulfillment
of the requirements for

M.S. Animal Science

 

 

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November 15, 1995
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EFFECTS OF FEEDING FISH COLLECTED DOWNSTREAM
FROM OAK RIDGE RESERVATION ON THE
REPRODUCTIVE PERFORMANCE OF MINK

By

LorinAlynLewis

A THESIS
Submitted to
Michigan State University
in partial fulfillment of the requirments
for the degree of

MASTER OF SCIENCE

Department of Animal Science

1995

ABSTRACT
EFFECTS OF FEEDING FISH COLLECTED DOWNSTREAM
FROM OAK RIDGE RESERVATION ON THE REPRODUCTIVE
PERFORMANCE OF MINK
By

LorinAlynLewis

Concentrations of radionuclides, metals, and organic compounds in water, sediment, and
biota of the Clinch River and Watts Bar Reservoir on the Oak Ridge Reservation (ORR) have
lead to concern with regard to environmental and human heahh. The objectives of the
present study were to assess the effects of polychlorinated biphenyls and mercury in fish
collected from the reservation on mink reproduction and to provide information to the Clinch
River Environmmtal Restoration Program for evaluating the extent and degree of efl’ects of
ORR operations on wild piscivorous populations. Fish collected fi'om downstream ORR
containing 2.13 ppm total polychlorinated biphenyls (PCBS) and 0.35 ppm mercury were
substituted for marine fish in mink (Musteh yuan) diets at concentrations of 0, 25, 50, and
75%. The experimental diets were fed to adult mink beginning three months prior to breeding
(December) and continuing until the yormg were weaned (June). NO adverse effects on the
reproductive performance of the adult mink or kit survival could be attributed to feeding the

experimental diets.

ACKNOWLEDGEMENTS

I would like to thank my major professor, Dr. Richard Aulerich, for his help and guidance
throughout my graduate studies. I also extend my thanks to the remaining members of my
graduate committee: Drs. Steven Bursian, Karen Chou, and James Render.

I am also greatful for the support and understanding that I received from my collegues at
the MSU Experimental Fur Farm while learning mink husbandry: Christina Bush, Angelo
Napolitano, Jefl‘ Greenley and especially Phil Summer for listening and advising and for
making even the most routine tasks enjoyable. And finally, a sincere thanks is extended to
Ms. Carol Daniel for her friendship, patience and understanding, and for going above and

beyond in helping me with my thesis.

TABLE OF CONTENTS

PAGE

LIST OF TABLES ............................................................................................................. vi
LIST OF APPENDICES ................................................................................................... viii
INTRODUCTION ................................................................................................................ 1
STUDY OBJECTIVES ......................................................................................................... 3
LITERATURE REVIEW ...................................................................................................... 4
OAK RIDGE NATIONAL LABORATORY ................................................................... 4
POLYCHIDRINATED BIPHENYLS ............................................................................. 6
General Introduction. .............................................................................................. 6
Production ofPCBs in the United States ................................................................. 6
Production Process ................................................................................................. 8
Application ............................................................................................................ 8

PCBS in the Ecosystem .......................................................................................... 9

Fate of PCBS. ....................................................................................................... 10
Bioaccumnlation of PCBs .................................................................................... 11
Biodegradation of PCBS ...................................................................................... 13

Toxic Effects ofPCBs in the Ecosystem. .............................................................. l4
Absorption and Metabolism of PCBS .................................................................... 15

MPG Induction by PCB s ..................................................................................... 16

Toxic Equivalency Factors ................................................................................... l7

Toxic and Biological Responses to PCBS ............................................................. 18
Carcinogenicity Of PCBS ..................................................................................... 19

Human Health Effects .......................................................................................... 20
Laboratory Animal Studies .................................................................................. 23

The Effects ofPCBs on Mink .............................................................................. 24
MERCURY .................................................................................................................. 26
General Introduction to Mercury and the Mercury Cycle ...................................... 26
Methylmercury and the Environment. .................................................................... 27

Human Poisoninings ............................................................................................ 28
Mercury Concentrations in the Tennessee River System. ...................................... 29

Effects of Mercury on Wildlife ............................................................................. 30

Effects of Prenatal Exposure to Methylmercury .................................................... 3 1

Tissue Distribution of Mercury .............................................................................. 33
Pathology of Mercury Poisoning .......................................................................... 3 5
MATERIALS AND METHODS ......................................................................................... 37
Fish ............................................................................................................................ 37
Diet Preparation. ........................................................................................................... 38
Experimental Design and Animal Care .......................................................................... 40
Feeding Trial ............................................................................................................... 41
Reproduction ................................................................................................................ 42
Blood Collection ......................................................................................................... 43
Necropsy ........................................................................................................................ 44
Histopathology ............................................................................................................... 46
Chemical Analysis ......................................................................................................... 46
Diet Analysis ................................................................................................................. 47
Statistics ......................................................................................................................... 47
RESULTS ........................................................................................................................... 48
DISCUSSION ..................................................................................................................... 71
SUMMARY ....................................................................................................................... 95
FUTURE STUDIES ........................................................................................................... 97
APPENDICES ................................................................................................................... 99
BIBLIOGRAPHY ............................................................................................................ 103

LIST OF TABLES

TABLE PAGE
1. Composition and nutrient analysis of experimental diets ............................................. 49

2. Concentrations (ppm) of mercury and PCBS in fish used
in mink diets ............................................................................................................ 50

3. Concentrations (ppm) of mercury and PCBS in diets fed
to adult mink ............................................................................................................ 51

4. Mean monthly (December, 1993 through June, 1994) body weights (g)
of adult female and male mink fed diets containing various percentages of
”contaminated" and ”uncontaminated" fish ................................................................. 52

5. Mean body weights (g) at the beginning and end of the trial and body weight
change (g) Of female and male mink fed diets containing various
percentages of "contaminated" or "uncontaminated" fish ........................................... 53

6. Reproductive performance of female mink fed diets containing various percentages of
"contaminated" and "uncontaminated" fish and body
weights and survivability of their kits ........................................................................ 55

7. Reproductive performance of male mink fed diets containing various
percentages of "contaminated" and "uncontaminated" fish ........................................ 57

8. Organ weights (g) of eight adult female mink and two adult male mink fed
diets containing various percentages of "contaminated" and "uncontaminated"
fish .......................................................................................................................... 59

9. Mean organ weights (g) of nine six-week-old kits exposed, via their dams to various
percentages of "contaminated" and "uncontaminated" fish during gestation and
lactation and through consumption of the diets following weaning .......................... 60

10. Mean hematologic values for eight adult female mink fed Diet B or Diet E at
various times during the Study ................................................................................... 62

ll.

12.

13.

14.

15.

16.

Mean serum chemistry values for eight adult female mink fed Diet B or

Diet E at various times during the study ...................................................................... 64

PCB consumption for adult mink fed diets containing various percentages of
Atlantic mackerel and fish from the Clinch River upstream and downstream

from the Oak Ridge Reservatio .................................................................................. 65

Mean concentrations (mg/kg; wet weight) of PCB congeners (identified by
IUPAC number) in adipose tissue of adult mink fed diets containing various

percentages of "contaminated" and "uncontaminated" fish .......................................... 66

Mean concentrations (mg/kg; wet weight) of PCB congeners (identified by
IUPAC number) in the liver of adult mink fed diets containing various

percentages of "contaminated" and "uncontaminated" fish ......................................... 67

Mercury consumption for adult mink fed diets containing various percentages
of Atlantic mackerel and fish fi'om the Clinch River upstream and downstream

of the Oak Ridge Reservation ................................................................................... 69

Mean mercury concentrations (mg/kg; wet weight) in organs from adult mink
fed a diet containing various percentages of "contaminated" and

uncontaminated" fish ............................................................................................... 7 0

LIST OF APPENDICES

APPENDDK PAGE

A. Kilograms and number of species of fish collected from the Oak Ridge
Reservation used in the experimental diets ......................................................... 99

B. Calculations for consumption of feed, polychlorinated biphenyls (PCB s) and

TCDD-EQ ........................................................................................................ 100
C. Calculations for consumption of feed and mercury ............................................. 101

D. IUPAC numbers of PCB congeners found in fish from the Oak Ridge
Reservation ........................................................................................................ 102

INTRODUCTION

Operations and waste disposal activities at the Y-12 plant, the Oak Ridge National
Laboratory (ORNL) and the Oak Ridge Gaseous Diflirsion Plant, all located on the US.
Department of Energy (DOE) Oak Ridge Reservation (ORR) in eastern Tennessee, have
introduced a variety of airborne, liquid and solid wastes into the surrounding environment.
Some of these wastes may affect off-site areas (areas beyond the OR bormdary) by entering
local streams that ultimately drain into the Clinch River and Tennessee River systems.
Concentrations of radionucleotides, metals and organic compounds in water, sediment and
biota of the Clinch River and Watts Bar Reservoir suggest the presence of a variety of
contaminants of possible concern with regard to the health of the environment and human
population The contaminants of particular concern are polychlorinated biphenyls (PCB s) and
mercury (Hg). Thus, the DOE has initiated a comprehensive environmental restoration effort
to eliminate releases of hazardous substances, pollutants and contaminants from the ORR.

Mink (Musteh yisgn) were identified in the Screening Level Risk Assessment for Off-
Site Ecological Efi‘ects in Surface Waters Downstream from the U. S. Department of Energy
Oak Ridge Reservation as a species being at risk (Suter, 1990). Since they also have been
shown to be among the most sensitive, if not the single most sensitive, mammalian species to

PCB toxicity (Aulerich and Ringer, 197 7 ), they were the preferred animal model for this

2

study. Feeding studies conducted by Aulerich a El (1971,1973), Homshaw 91 RI (1983), and
Heaton 91 a1, (1995) have demonstrated the extreme sensitivity of mink to chlorinated
hydrocarbon contaminants, especially PCBS, contained in fish taken fiom the Great Lakes.
Research performed with mink has been instrumental in setting U. S. water quality standards
for PCBS (Aulerich and Bleavins, 1981). Additional studies have shown this species to be
similarly sensitive to other halogenated hydrocarbons, including polybrominated biphenyls
(Aulerich and Ringer, 1979), hexachlorobenzene (Bleavins et aL, 1984), and 2,3,7,8-
tetrachlorobenzo-p-diordn (Hochstein et al, 1988). The toxicosis of mercury has also been
studied extensively in mink ( Aulerich et at, 1974; Wobeser 91 al, 1976). Numerous other
toxicological studies with mink have been reported in the literature and summarized by
Calabrese et a1. (1992).

It is known that fish, a major food item of mink, inhabiting aquatic systems downstream
from the OR contain elevated concentrations of PCBS and mercury. Therefore, the overall
objective of this study was to assess the effects of these environmental contaminants in the
fish on mink. These assessments will provide information for the Clinch River Environmental
Restoration Program for evaluating the adverse effects of ORR operations on wild

piscivorous populations.

STUDY OBJECTIVES
The specific objectives of this study were to:

1. Determine concentrations of polychlorinated biphenyls and mercury contaminants
in fish collected up stream and downstream from the ORR;

2. Determine the reproductive and other physiological efl‘ects in mink fed fish
collected downstream from the ORR;

3. Examine mink for gross and histopathologic alterations of a toxicosis from
contaminants found in fish from ORR which were fed to mink;

4. Determine concentrations of polychlorinated biphenyls and mercury in mink

tissues;

LITERATURE REVIEW

WIDE!

In 1942, the Army Corps of Engineers purchased 92 square miles of land under the guise
(for security reasons) of establishing the Kingston Demolition Range. The area, located in
eastern Tennessee, was given the name Clinton Engineer Works and was intended for the
large-scale production of fissionable isotopes of uranium and plutonium needed for the atomic
bomb. In November, 1942, construction of the headquarters for the Manhattan Engineer
District and nerve center for the wartime atomic energy effort was begun. The first science
facility, Clinton Laboratory (now known as Oak Ridge National Laboratory) was built in
1943. Its purpose was to serve as a pilot plant for the large plutonium-producing reactors
and as a facility for research and development for large scale production of plutonium. The
major Oak Ridge facilities inchrded two uranium production plants (Y -12 and K-25) and a
laboratory (X-lO). Late in 1944, a third plant, the thermal difl‘usion plant, was built to boost
fissionable isotope production. By 1945, Oak Ridge employed a total of 82,000 people.
When World War H ended, the population decreased and the Atomic Energy Commission
began preparing long-range plans for Oak Ridge so that the community could continue to
function as a major scientific center (Thompson, 1973; Kraus, 1976).

The Oak Ridge reactor was the world's second, and was in operation until 1963. The first

batch of irradiated firel slugs was taken from the reactor in 1943 and the first plutonium was

5

stripped from Clinton Laboratory in 1944. On August 6, 1945, the United States dropped the
first atomic bomb of World War II on Hiroshima, Japan. The bomb used enriched uranium
fi'om the Clinton Laboratory. After the war, Clinton Laboratory shifled part of its wartime
effort to peacefirl research. Since the end of World War 11, Oak Ridge has been a center for
numerous projects including: demonstrating the safe production and chemical recovery of
plutonium, developing safe methods for reprocessing nuclear fuels, producing radioisotopes
for scientific research,operator training and development of nuclear power systems such as
breeder and fusion reactors. For over 40 years, Oak Ridge produced nuclear weapons and
has been a world center of study on the efi‘ects of radiation on the environment. Today, Oak
Ridge consists of 24 divisions ( 16 major research divisions and 8 service divisions) located
on 35,300 acres It is a stable and progressive city with a worldwide reputation as a nuclear
center of excellence and it has had a role in virtually every major scientific operation and
activity in the atomic energy program (Thompson, 1973; Krause, 1976).

The many research and production projects conducted at Oak Ridge over the years have
resulted in contamination of the environment of the ORR and adjacent areas with persistent
and potentially toxic compounds, including PCBS and mercury. Because of concern for the
effects these contaminants pose to human health and wild piscivorous populations, this study
was conducted to provide information that could be used in an environmental risk assessment

for the Clinch River Environmental Restoration Program

W5
General Introduction

Polychlorinated biphenyls (PCB s) are molecules that have multiple chlorines attached to
a biphenyl nucleus. Their general formula is C12HxCly where x=0-9 and y=10-x The
biphenyl molecule is made up of two connected rings of six carbon atoms each. Any or all
of the 10 available sites can have chlorine atoms, with 209 difl'erent PCB compounds possible
(Abramowicz, 1990). Approximately 25 congeners account for 50-75% of the total mass of
PCBS formd in the environment (McFarland and Clarke, 1989). The extensive use of PCBS
in many industrial applications is due to the unique physical and chemical properties of these
compounds. PCBS are resistant to acids, bases and other chemical agents, they have
remarkable thermal stability, stability to oxidation and hydrolysis, low solubility in water, low
flammability, high electric resistivity, favorable dielectric constants, and low vapor pressure

at ambient temperature (DeVoogt and Brinkman, 1989).

Production of PCBS in the United States
Commercial production of PCBS began in the United States in 1929. Since that time,
approximately 1.5 million metric tons of PCBS have been produced in 10 countries. The
Swan Chemical Company was the first US. company to produce PCBS. The company was
purchased in the mid 1930s by Monsanto Industrial Chemicals which began producing

various chlorinated biphenyl mixtures under the trade name Aroclor (De Voogt and

Brinkman, 1989).

7

Chlorinated biphenyl mixtures were initially produced for dielectric fluids in transformers.
In time, their utilization in lubricants and heat-transfer systems caused an increase in their
demand. The production and sale of PCBS reached a peak in 1970 but were dramatically
reduced in 1971 when, after much public and scientific concern, Monsanto voluntarily
reduced its PCB production By 1977 the company had ceased all production of PCBS. The
estinnted cumulative production of PCBS in the US. from 1930 to 1975 was 1400 million
pounds. (DeVoogt and Brinkman, 1989).

There are many trade names under which commercial PCB mixtures have been sold. In
the United States and Great Britain, the mixtures were sold under the name of Aroclor. In
other countries they were sold as Clophen (Germany), Fenchlor (Italy), Kanechlor (Japan)
and Phenochlor (France).

The Aroclor mixtures produced in the United States had a chlorine content of
21,32,42,48,54,60 or 61% by weight. The mbdures were designated by a four digit number,
the first two digits, "12" (except for Aroclor 1016), represented the 12 carbons of the
biphenyl skeleton The second two digits indicated the percentage of chlorine in the mixture.
Thus, Aroclor 1254 had 12 carbons and contained 54% chlorine. Since July, 1979, through
authority of the Toxic Substance Control Act (TSCA), the manufacture, importation,
distribution and fiuther processing of PCBS in the US. has been banned (Hooper et 31.,

1990)

8

Production Process

The production of PCBS involved chlorination of biphenyl and separation and
purification of chlorinated biphenyl fractions For the production of PCBS, a chlorinator was
charged with proper quantities of biphenyl and a catalyst, such as ferric chloride or iron
filings. Anhydrous chlorine was allowed to flow through the mixture and the charge was
circulated using a pump. The mixture was heated above the melting point of the biphenyl and
the hydrogen chloride produced by the chlorination process was discharged The chlorination
process typically took 12-36 hours. The crude Aroclor products were then treated with a
0.3% alkali mixture and purified to remove the catalyst, color, and hydrogen chloride by
vacuum distillation Commercial PCBS were produced as technical grade liquids and as liquid
mixtures ready for application (DeVoogt and Brinkman, 1989). The resulting products
ranged from light oily fluids (di, tri, and tetra-chlorobiphenyls) to heavy, honey-like oils
(penta-chlorobiphenyl) to greases and waxes (more highly chlorinated biphenyls)
(Abramowicz, 1990). During certain chemical production processes, accidental production
of PCBS has been recognized. It is estimated that 50 tons of PCBS are produced annually as
by-products of industry in the United States (Callahan e]; n.1,, 1984; DeVoogt and Brinkman,

1989)

Applications
Depending on their application, the uses of PCBS have been divided into two categories:
open and closed systems. Open-ended systems are those from which PCBS can not be

re—collected. Open systems include use of PCBS in plasticizers, carbonless copy paper,

9

hrbricants, inks, laminating agents, paints, adhesives, waxes, additives in cement and plaster,
casting agents, sealing liquids, fire retardants, immersion oils and pesticides. The use of
PCBS in open systems can lead to environmmtal contamination and it was for this reason that
most cormtries decided to terminate the open-ended use of PCBS during the years 1971
through 1973. In 1973, the use of PCBS was restricted to four closed systems. The closed
systems included the use of PCBS in cooling liquids in transformers, dielectric liquids in
capacitors, heat-conducting fluids in heat-exchangers and fire- or heat-resistant corrosion-free
hydraulic fluids in mining equipment and vacuum pumps (DeVoogt and Brinkman, 1989).

At present, some closed systems in which PCBS are used are not considered closed as these
wstems are known to leak to some extent (UNEP, 1985), and thus, PCBS from many sources

may lead to environmental contamination (DeVoogt and Brinkman, 1989).

PCBS in the Ecosystem

In spite of the ban on PCB production in this country, the persistence of PCBS in the
environment and their concentration in the biological food chain have caused much concern.
The extensive application of these chemically and thermally stable compounds has resulted
in widespread contamination. It is estimated that several million pounds have been released

into the environment (Abramowicn 1990).
Polychlorinated biphenyls are toxic, mutagenic and teratogenic agents with
bioaccurmrlation and bioconcartration ability. They are a global environmental health hazard
as they have a detrimental impact on nearly every member of the biota, appearing in the

tissues of most living creatures as well as in the air, water, soils and sediments over most of

9

lubricants, inks, laminating agents, paints, adhesives, waxes, additives in cement and plaster,
casting agents, sealing liquids, fire retardants, immersion oils and pesticides. The use of
PCBS in open systems can lead to environmental contamination and it was for this reason that
most countries decided to terminate the open-ended use of PCBS during the years 1971
through 1973. In 1973, the use of PCBS was restricted to four closed systems. The closed
systems included the use of PCBS in cooling liquids in transformers, dielectric liquids in
capacitors, heat-conducting fluids inn heat-exchangers and fire- or heat-resistant corrosion-free
hydraulic fluids inn mining equipment and vacuum pumps (DeVoogt and Brinkman, 1989).

At present, some closed systems in which PCBS are used are not considered closed as these
syaans are known to leak to some extennt (UNEP, 1985), and thus, PCBS from many sources

may lead to environmental contamination (DeVoogt and Brinnkman, 1989).

PCBS inn the Ecosystem

In spite of the ban on PCB production in this country, the persistence of PCBS inn the
environment and their concentration inn the biological food chain have caused rmnch concern.
The extensive application of these chennically and thermally stable componmds has resulted
in widespread contaminnation. It is estimated that several million pounds have been released

irnto the environment (Abramowicz, 1990).
Polychlorinated biphenyls are toxic, mutagenic and teratogenic agents with
bioaccurnulation and bioconcantration ability. They are a global environmental health hazard
as they have a detrimental impact on nearly every member of the biota, appearinng in the

tissues of most living creatures as well as in the air, water, soils and sediments over most of

10

the earth, and are thus of ecological and medical concern ( Hooper et al, 1990). The
environmental burden in the air, water, sediment, soils and biota was estimated by the
National Research Connncil (1979) to be 82 million kg. The Great Lakes have received
considerable attention with regards to PCB contamination because of substantial food-chain
concentrations Although the ban of PCB production may eventually result in a decrease to
trivial levels in systems such as the Great Lakes, differences in mobility, biodegrability, waste
storage and continued use of PCBS suggests that we may have to cope with this pollutant for
many years to come.

The chemical and physical properties that make PCBS so commercially desirable also
make them serious environmental pollutants Their stability, along with the case at which they
are taken up by livinng organisnns and accumulated at high levels inn the food chain has resulted
inn environmental problems Their nonpolar nature makes them a good insulating coolant in
electrical equipment because of the high dielectric constant, but it is their nonpolarity that also
causes them to be highly lipophilic and to bioaccumulate. The chlorinne in PCBS is
electronegative and stabilizes the biphenyl molecule. This results inn a heat-resistant fluid with
long- term stability, making them ideal for oils and hydraulic fluids but also recalcitrant to
physical and biological degredation Highly substituted biphenyls are less volatile, less soluble

in water, and more chemically stable than the lesser substituted congeners.

Fate of PCBS
The highest biological activities in a water cohmnn are at the air-water and water-sediment

boundaries. Accumulation of PCBS at these surfaces mediates their bioaccurnulation by

l 1
bacteria and plankton, resulting inn the introduction of chlorinated biphenyls into the food web
(Young et. al,l977).

Once PCBS enter a freshwater ecosystem, they have numerous abiotic interactions with
the ecosystem. Because PCBS are highly nonpolar, they tend to accunnnlate and localize in
sediments (Y 01mg e. aL, 1977; Steen et. al, 1978). They can be removed from sediments
by beinng transported downstream or by diflirsional loss to the water column PCBS in the
water cohmnn can be removed by transport downstream and are subject to photodegredation
and volatility losses (Baxter and Sutherland, 1984). Biotic innteractions occur inn fieshwater
and because of the nonpolar nature of PCBS, they accurrnnlate in the lipids of aquatic biota.
The lipophilic nature of PCBS and resistance to breakdown, allow them to biomagnify,

(Sanders and Chandler, 1972) impacting organisms at every level of the food chain.

Bioaccumulation of PCBS

There are three physiochemical properties that control PCB bioaccunnnlation; the
degree of chlorinnation, water solubility and PCB stereochennistry. The higher chlorinnated
penta-and hexachloro biphenyl isomers accumulate to the greatest extent and thus have a
higher bioaccumulation rate. The higher chlorinated PCBS also have half-lives in organisms
that are related exponentially to their degree of chlorination (Kahnaz and Kalmaz, 1979). The
bioaccmnulation of PCBS from water by aquatic organisms is correlated with the lipophilicity

of PCBS (Veith er a1, 1980).
Bioaccumulation is also affected by the ability of PCBS to pass through biological

membranes Therefore, there is an optimal steric configuration for PCB bioaccurrmlation.

12

The stereochemistny of PCB molecules affects the strength of adsorption of PCBS to
membrane surfaces. The most nonplanar molecules are the most strongly absorbed while
those PCBS with planar aromatic rirngs are weakly absorbed (Veith et al, 1980).

Biota can acquire PCBS fi'om three sectors of the environment; atmosphere, water and
food Terrestrial organisnns acquire PCBS by absorption of PCBS in the atmosphere through
the hang walls, absorption of PCBS in the atmosphere through the epidermis, and absorption
of PCBS derived fi'om food and/or water through the gastrointestinal tract. Aquatic
organisms acquire PCBS by absorption of PCBS in the water through the gills, absorption of
PCBS in the water through the epidermis and consumption of contaminnated food.

The mainn route of uptake of PCBS by aquatic organisms is via absorption through the
gills Since the gills represent the active membrane surface for water exchange (Phillips, 1980).
Once absorbed through the gills, PCBS are partitioned into the blood and then transported
from the blood to the tissues (Kenaga, 1975). The assimilation of PCBS from ingested food
occurs by partition across the lipoprotein membranes lining the gut into the bloodstream
(Walker, 1975). The PCBS associated with benthic organisms are derived from contaminnated
sediment, and are directly related to the sediment PCB concentration (Ninnmo er al, 1971).

Inert organochlorines with high octanol/water partition coeficients (Kow) can
biomagnify ( the concentration of the chemical in the organism reaches a level that exceeds
that inn the diet of the organism). This results inn food-chain accunnnlation, inn which the
concentration of a chennical increases with every step up the food chainn (Clark :1 fl, 1988;
Connelly and Pederson, 1988). PCBS are transported to the inntestinal wall inn association with

lipid molecules. They separate from the lipid and diffuse as a single molecule through the

l3

intestinal wall. At the other side of the intestinal wall, the chennical is reassociated with
lipoproteins and resynthesized triglycerides (Vetter 91 al, 1985).

Biomagnification occurs as a result of food ingestion. Gobas et a]. (1993) conducted
studies with humans and found that the concentration of PCBS in the blood of mothers was
four times higher than that in the cord blood of the fetuses However, on a lipid-weight basis,
PCB concentrations in the mothers and the fetuses were approximately equal because lipid
concentrations inn the cord blood were three to four times lower than that in maternal blood
(Fomon 5 al, 1970; Gobas e n.1,, 1989). After birth, no biomagnification occurred but when
the infants were exposed to the mother’s milk, the PCB lipid-based concentration in the

infants rose to exceed those in the milk (biomagnification).

Biodegradation of PCBS

Numerous aerobic and anaerobic bacteria have been identified as being capable of
PCB degradation ( Sayler et. al, 197 8; Klages and Linngens, 1980). Aerobes oxidatively
attack PCBS, breaking open the carbon ring and destroying the compound Anaerobes leave
the biphenyl ring intact while removing the chlorines. Anaerobic dechlorination degrades
highly chlorinated compounds irnto less chlorinnated derivatives. The two types of bacteria can
work together to biologically destroy all PCB mixtures (Abramowicz, 1990). Certain fungi

may also aerobically degrade PCBS to lower chlorirnated compounds.
Besides microbial degradation, some PCBS may be susceptible to photochemical reactions
or biochemical degradation. Most PCBS that accumulate inn aquatic sediments are shielded

from photolysis (Larsson, 1984). Photochemical dechlorinnation may yield products that are

14

more readily degraded by microorganisms than the original compormds or it can also lead to

the formation of toxic polychlorirnated dibenzofinrans (Baxter and Sutherland, 1984).

Toxic Effects of PCBS in the Ecosystem
Toxic effects of PCB contamination to the aquatic biota appear to be sublethal and
dnronic . Inn wildlife, physiological and developmental effects are the most sensitive endpoinnts
for PCB toxicity. Growth retardation, immune system suppression ( Friedman and Sklan,
1989), elevated rates of disease, wasting syndrome, subcutaneous, pericardial and peritoneal
edena, hepatic porphyria, congenital malformations (Fox fl 3]., 1991;Gilbertson et al, 1991),
altered hormone, retinnol and vitamin A concentrations (Government of Canada, 1991;
McFariand and Clarke, 1989), innpaired calcium metabolism, thyroid alterations ( Jefliies and
French, 1972; Hurst et a1, 1974) and behavioral changes ( McArthur et a1, 1983) have all
been documented
Epidenniological data for the Great Lakes suggest hmnans and wildlife may exhibit subtle,
chronic effects due to PCB exposures. Wildlife, uner humans, do not recognize or avoid
contanninated food supplies and thus may receive greater dietary exposure to contaminants.
Predators high on the food chain, such as fish-eating birds and manuals, are exposed to
greater concentrations of PCBS than animals lower in the food chain due to biomagnification
and bioaccnunulat'non Small birds and mammals have high metabolic rates, eat more per unit
body weight per day and are thus exposed to greater concentrations of PCBS in m at a

faster rate than larger species.

15

Absorption and Metabolism of PCBS

PCBS are readily absorbed by passive difi'ussion from the gastrointestinal tract, and
transported by blood to all tissues with little to no elimination. In°nial tissue distribution is
proportional to the rate of blood flow to the tissues and the tissue volume. Initially the
highest concentrations are seen inn highly perfinsed tissues such as the liver and large volume
muscles. Eventually equih’brium is reached for all tissues, the concentration at equihhrium
being determined by the lipid content of that tissue. Thus, there are higher concentrations of
PCBS inn adipose tissue and lower concentrations inn blood and liver. When the PCB
concentration inn the liver is reduced by metabolism and excretion, more PCBs will partition
from the blood to the liver to reestablish the tissue/blood ratio. Additional PCBS will partition
from all other tissues into the blood to reestablish tissue/bood ratios. PCBs that are cleared
fiom the liver are cleared from all other tissues as well, and those PCBs not metabolized will
concentrate inn the adipose tissue (Lutz e]; 3]., 1977).

It is possible that PCBS can be excreted through the small intestine wall and excreted in
the feces Williams at al (1965) found that dieldrin innjected inntravenously into rats with the
bile duct cannulated excreted the organochlorine compound into the feces, suggesting
excretion from the gut wall This suggests that PCBs may be partially excreted in an
unchanged state into the gastroinntestinnal tract.

Many foreign compounds such as drugs, food additives, pesticides, and industrial
chemicals are metabolized to more polar derivatives (Parke, 1968). PCBs must be
transformed innto more polar metabolites for excretion. Because commercial mixtures of

PCBs are complex, it is diflicult to obtainn information on the quantitative and qualitative

16

aspects of PCB metabolism.

Yamamoto et a]. (1973) studied the metabolism of PCBS and found that a major
metabolite of 2,4,3',4'-tetrachlorobiphenyl was a 5-hydroxyderivative and they attributed the
acute toxicity of 2,4,3',4'-tetrachlorobiphenyl to the production of the phenolic metabolite
within the body. Although no evidence, other than for 2,5,2',5'-tetrachlorobiphenyl, has been
found for the formation of epoxides during the metabolism of PCBs, Brodie et al (1971)
reported that halogenobenzenes injected irntraperitoneally innto rats induced massive necrosis
of the centrolobular regions of the liver. They suggested that the epoxide produced in
aromatic hydroxylation as a labile irntermediate could be responsible for the hepatic necrosis.

This hypothesis could also extend to metabolically-induced PCB toxicity.

MFO Induction by PCBS

The mixed-function oxidase (MFO) enzymes are major components of the biological
defense of livinng organisms againnst chemical stresses in the environment. These enzymes
work by adding oxygen to lipophilic, endogenous and foreign compounds, catalyzing their
biotransformation to more water sohrble and readily excreted products. The cytochrome
P-450 system is capable of hydroxylating, epoxidating and dealkylating xenobiotics. Because
MFO enzymes are present in a wide variety of organisms, including humans, and since they
play a central role in detoxification, they are good non-specific biomarkers of exposure to
xenobiotic chemicals (Ionnides et. a1, 1984). Exposure to significant quantities of mixed
inducers such as PCBs can result inn innduction of isozymes which may activate other

contamirnants (Ionnides et a1, 1984) resulting in the formation of free radicals. Free radicals

17

may damage cells or organ systems and /or alter the rate and patterns of normal biosynthesis
and metabolism of essential biomolecules such as retinoids (Parke, 1968) and steroid
hormones (Wood et a1, 1983) resulting in secondary effects on growth, reproduction and
disease susceptibility.

Though usually beneficial, biotransformation can lead to reactive inntermediates that are
more toxic than the parent compound leadinng to chemical-induced toxicities irncludinng
rnutagenesis, carcinnogenesis, teratogenesis, and neurotoxicity. Many carcinogenic xenobiotics
depend on their conversion by cytochrome P-450 to the carcinogenic metabolite (Conney,
1982)

Animals having a "normal" aromatic hydrocarbon receptor (Ah receptor) are more
responsive to the inductive effects of various polycych aromatic hydrocarbons. When the
receptor is defective, the animals are non-responsive to the inductive effects of aromatic

hydrocarbons (Poland e al, 1979).

Toxic Equivalency Factors
Polyhalogenated aromatic hydrocarbons (PHAHs) are a group of lipophilic and
chemically stable environmental contaminants and innclude the polychlorinnated dibenzo-p-
dioxins (PCDDs), polychlorinated drhenzofurans (PCDFs) and polychlorinated biphenyls
(PCBS). The most toxic lnnown PHAH is 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) which
birnds to the aromatic hydrocarbon receptor (Ah receptor) through which its toxic effects are
proposed to be mediated Some of these efl‘ects innclude reproductive failure, teratogenesis,

carcinogenesis and imrrnmotoxicity (Ahlborg gt 3]., 1994).

18

Much of the toxicity caused by PCBs has been attributed to specific congeners that
resenble TCDD and it is believed that they exert a number of common toxic responses similar
to TCDD because of the common mechanism of binding to the Ah receptor (Giesy et al,
1994a). Environmental sannples of dioxin-like compounds usually exist as a complex mixture
of congeners, therefore in order to simplifiy risk assessment and regulatory control, the
concept of toxic equivalents (TEQS) has been introduced. Toxic equivalency factors (TEFs)
are estimates of the relative potency of inndividual congeners expressed relative to TCDD
(Safe, 1990; Giesy et al, 1994b). TEFs are based on efl‘ects such as lethality, deformities, or
enzyme induction TEFs can be used to calculate concentrations of TEQs that are contributed
by individual congeners. The T'EFs are summed and expressed as a total equivalent
concentration of TCDD. The TEQ is determined by multiplying individual congener
concentrations with their corresponding TEFs. Evaluation of congeners contributinng to the
TEQ reveals that congeners 77, 156, 105, 157 and 114 ( IUPAC congener identification
mariners) are the dominatinng congeners inn the lesser-chlorinated Aroclor mixtures (Ahlborg

a an, 1994).

Toxic and Biological Responses to PCBs
PCB mixtures and inndividual congeners elicit toxic and biologic responses in organisnns
which innclude: l. a wasting syndrome (progressive weight loss which is not related to
decreased food consunnption); 2. skin disorders ( acneform eruptions or chloracne, alopecia,
edema, hyperkeratosis and blepharitis due to hypertrophy of the Meibonnian glands); 3.

hyperplasia of the epithelial lining of the extrahepatic bile duct, the gall bladder and urirnary

l9

tract; 4. lynnphoid invohrtion ( thynnic and splenic atrophy with humoral and/ or cell-mediated
immunosuppression and/or associated bone marrow and haematologic dyscrasias); 5.
hepatomegaly and liver damage( necrosis, hemorrhage and irntrahepatic bile duct hyperplasia);
6. porphyria( disordered porphyrin metabolism of the cutanea tarda type); 7. endocrinne and
reproductive disfimction (altered plasnna concentrations of steroid and thyroid hormones with
menstrual irregularities, reduced conception rate, early abortion, excessive menstrual and
postconceptional haemorrhage, anovolution inn females, and testicular atrophy and decreased
spermatogenesis inn males); 8. teratogenesis ( cleft palate and kidney malformations); 9.
carcinnogenesis (Safe 91 al, 1982).

Numerous studies have shown that PCBS administered as a Single dose are less toxic
than the same amount administered over a long period of time. There is usually a latent
period of time between the time of enqrosure and the onset of signs of toxicity. Some
disorders are manifested after several months of exposure, such as porphyria, while other
clinical signs may occur withinn days of exposure, such as thymic atrophy. Because PCBS
have such a long half-life, the synptoms of chronic toxicity can develop even after exposure

has ceased.

Carcinogenicity of PCBs
It appears that virtually all PCB congeners are stable and not readily converted by
biotransformation enzymes into reactive inntermediates that could potentially cause damage
to DNA. Although there is some evidence that some PCBs can covalently bind to DNA and

cause genotoxic effects inn some in yitm SyStems (Morales and Matthews, 1979), short-term

20

administration of PCBs to mice failed to initiate carcinogenesis in the skin (Oesterle and
Deni, 1984; Hayes g at, 1985). However, mice and rats continuously exposed to PCBS for
two years or more developed preneoplastic livers (Ito et al, 1973; Nishizumi, 1976), gastric
lesions (Morgan e 3]., 1981), and hepatocellular and gastric carcinnomas. Therefore, initiation
may occur with long-term exposure to PCBs.

Because PCBs are potent innducers of hepatic nnicrosomal cytochrome P-450 isozymes,
hepatocytes that were previously induced by PCBS may be more susceptible to initiation by
carcinogens that require microsomal activation for genotoxicity . Promotion by PCBS is
dose-dependent and there is a threshold dose below which promotion of preneoplastic liver
lesions is not observed. This threshold may be above what most animals and humans

encounter in the environment (Morales and Matthews, 1979).

Human Health Efl‘ects

There are three major scenerios in which humans have been exposed to PCBs: 1. workers
who produced or utilized PCBs; 2. accidental exposure; and 3. environmental exposure
through contaminated food, air or water (Safe, 1994). PCBs enter the human body in an
occupational setting through the body surface by direct contact (dermal exposure) or through
the respiratory tract. The reported effects of PCBS on occupationally- exposed humans
include denatological conditions, liver damage, irnduction of hepatic monoxygenase enzymes
and pulmonary dysfimction (Warshaw 91 al, 1979). These irndividuals also have relatively
high levels of PCBs in their serum or adipose tissues and irncreased serum activities of hepatic

enzymes and serum concentration of lipids. Women exposed to PCBs typically have given

21

birth to children with low birthweights (Hara, 1985; Lawton :1 a1., 1985; Takamatsu et a1,
1985). It has also been reported that there is a correlation between serum PCB concentration
and concentrations of PCBS inn milk which is available to nursing infants. Many of these
responses we'e reversible once exposure to PCBs ceased, the serum concentrations of PCBS
decreased Though no overall increases inn cancer-related mortality have been correlated with
occupational exposure to PCBS, inncreased incidences of specific cancers have been reported
(Brown, 1987). It is unnlikely that environmental uptake of PCBS results in significant human
health effects, however, individuals who consume large amounts of fish from PCB-
contamirnated waters may be exposed to high levels of PCBS which are reflected in elevated
serum PCB concentrations (Kreiss et 31., 1981).

Although PCBS were produced in large quantities since the 1930s, it was not until 1968,
when a major human poisoning (" Yusho") occurred in Japan that interest inn their toxicity
was aroused. In 1968, 1600 inndividuals in southwestern Japan sufl‘ered toxic effects afier
consuming rice oil contaminnated with a commercial PCB industrial fluid, Kanechlor 400.
Eleven years later a second large-scale human poisoning (" Yu-Cheng") occurred in Taiwan
due to contaminated rice oil PCBs were used as a heat conductor in the process of heating
the rice oil. The leakage of PCB from a heating pipe to the rice oil resulted in the
contamination (Safe, 1987).

Though the most characteristic symptoms of PCB inntoxication in humans included dermal
problems such as chloracne, patients reported a broad spectrum of effects including
headaches stomach aches, numbness of the extremeties, coughing, bronchial disorders, and

joirnt pains Children who were poisoned in the 1968 accident had retarded growth rates and

22

abnormal tooth development. Newborns were undersized and exhibited systemic
pigmentation (Urabe and Koda, 1976; Urabe and Asahi, 1984). In the initial stage of
intoxication, nonspecific symptoms such as fatigue and weight loss were observed With
time, promirnannt features arch as swellirng of the upper eyelids, cheeseer discharge from the
eyes, tennporary failing of the eyesight, acneform eruptions, and blackenirng of the pores, nails,
and dermal pigmentation were observed ( Urabe and Koda, 1976; Urabe and Asahi, 1984).
There are many non-dermal manifestations of PCB poisoning including nervous, endocrinne,
respiratory, hematologic, hepatic, metabolic, bone and joint disorders and effects on fetal and
infinnt life( Harada, 1976; Hirayama, 1976;1washito et a1, 1977 and Ohnishi and Arakawa,
1977)

PCB-contaminated fish may be part of the human food chainn. Because lakes are stocked
and sport fishing is popular, people consume these fish despite guidelines on the amount that
should be eaten due to contamination. Evidence suggests that mammals that eat these fish
are at risk of physiological and behavioral changes attributed to these chemicals. Jacobson
91 al (1984) studied the behavioral differences between children of mothers who ate Great
Lakes fisln versus those who did not It was fornnd that the offspring of mothers who ate two
to three Lake Michigan fish meals per month for at least six years, had a lower birth weight,
smaller head circumference, shorter gestational age, and less neuromuscular activity than
oflSprinng born to mothers who ate little to no fish from the Great Lakes (Feinn et a1, 1984).
The exposed babies also had poorer lability states, a greater amonmt of startle, weak reflexes
and were more worrisome than non-exposed infants. At four years of age, these babies had

lower verbal and memory scale scores and refinsal to cooperate was common (Jacobson et

23

an,1990)

Laboratory Animal Studies

PCBs produce a wide variety of biological effects in experimental aninnals These irnclude
enzyme induction and innhibition, decreased reproductive efliciency, changes in plasma lipid
concentrations, decreased imnnnrnocompetence, dermatological effects and changes inn liver
morphology which include hepatic porphyria, and liver tumor production in rodents (Neal,
1985)

Most species of animals administered an acute dose of PCBs will display a "wasting
syndrome" characterized by progressive body weight loss followed by weakness, debilitation
and death. The dramatic loss of weiglnt is due only in part to feed refusal The acute toxicity
of PCBs inn domestic mammals appears to decrease as the percent chlorination increases,
whereas toxicity increases with increasing percent chlorination for mallards, pheasants,
bobwhites and Japanese quail while the opposite is true for the chicken. Birds Show a
depressed growth rate, rufllcd feathes, decreased egg production and hatchability, embryonic
death, structural deformaties, weakness and death. For a given species, the female is often
more susceptible than the male to the toxic effects of PCBs and this susceptibility usually
decreases with age. This is because males have a higher drug-metabolizing capability
(Parkinson and Safe, 1987).

There are marked differences in the sensitivity of various species of animals to the toxic
efects of PCBs. Various strains of mice differ in their susceptibility to PCB intoxication.

Species also differ qualitatively in their response to PCBs. Rabbits are more sensitive to

24

PCBs than rats inn regards to fetotoxic and reproductive effects (Villeneuve et a1, 1971) and
nnink are more sensitive than rats or birds (Aulerich et al, 1973). Rats adnninistered PCBS
develop diarrhea, diminnished exploratory behavior, decreased response to pairn stimuli,
adipsia, oliguria, anorexia, erythema of the limbs, ataxia, coma and death. Reproductive
effects irnclude decreased number of females that give birth, decreased mating performance,
reduced litter size, and increased newbom mortality (Kimbrough et al, 1978). Avian species
administered PCBs displayed tremor, ataxia, ruffling and loss of feathers and fluid
accnmnrlation inn the abdominal and thoracic cavities. Chickens display a specific edematous
disorder called hydropericardium. Rhesus monkeys are sensitive to the acnegenic effects.
However, the most consistent symptom of halogenated aromatic hydrocarbon intoxication
in all species is thyan atrophy, one of the most sensitive responses to PCB exposure

(Parkinson and Safe, 1987).

The Effects of PCBs on Mink

PCBs are persistent environmental contaminants that continue to pose a potential risk
to hurmns and wildlife even though their production has been banned since 1977. Within the
last few decades, noticable declines in wild minnk populations in Sweden (Gerell, 1967) and
throughout the Great Lakes basin (Wren et al, 1986) have been reported. These declines
have been attributed to PCB contamination of species consumed by mink.

In the mid 19605, mink farmers inn the Great Lakes region reported decreased litter sizes
in mink fed fish fi'om the Great Lakes and its tributaries (Hartsough, 1965). With the

introduction of echo salnnon innto the Great Lakes and the eventual use of this species for

25

feedinng mink, high inncidences of reproductive failure in mink were reported by fur farmers
(Aulerich and Ringer, 1977). Newborn rnnink kit mortality as high as 80% was observed,
although the adult mink appeared rmafl‘ected. It was originally thought that the reproduction
problems in the mink associated with feedirng them Great Lakes fish were due to pesticide
contamination of the fish. However, mink feeding studies conducted at Michigan State
University revealed that the concentration of PCBS in the fish was directly related to the
degree of reproductive impairment inn the mirnk.

These and other studies (Platonow and Karstad, 1973; Bleavins er a1, 1982,1984; Wren,
1991) have shown that mink are among the most sensitive mannnals, if not the sirngle most
sensitive, to PCBs. Mink have also been shown to be highly sensitive to other halogenated
hydrocarbon contaminants inncluding dioxins (Hochsteinn et al, 1988), hexachlorobenzene
(Bleavins 91 al, 1982,1984) and polybronninated biphenyls (Aulerich 91 fl, 1986). Since
mink are camivores that occupy a top position in the food chain, they are exposed to higher
concentrations of metabolized forms of contaminants, such as PCB, than species lower in the
food chainn. Thus, they have become a perferred Species for studying the effects of these
contaminants in animals (See Calabrese et al, 1992 for a review of the use of mink as an
aninnal model).

The clinical signs and lesions or alterations observed inn minnk exposed to PCBS include,
decreased feed consumption (Aulerich 91 al, 1985), progressive body weight loss (Bleavins
a al, 1980; Aulerich a a1, 1985), bloody stools (Bleavins et al, 1980; Aulerich et al, 1986)
and lethargy (Bleavinns et al, 1980). Clinical examination of PCB-intoxicated mink have

denonstr'ated an irncrease inn several organ weights (Heaton, 1992), fatty liver (Heaton, 1992),

26
henorrhagic gastric ulcers (Kimbrough et al, 197 8) ln'dney degeneration (Kimbrough et al,
1978) and induction of enzyme activity (Aulerich et al, 1985).
The effects on reproduction inchrde a high incidence of embryo toxicity (Bleavins er al,
1980) and therefore a decrease in the number of females whelping, smaller litter sizes
(Bleavinns e a]., 1980), decreased kit birth and four-week body weights (Aulerich and Ringer,

1980; Bleavinns et al, 1980) and higher kit mortality (Aulerich 91 al, 1973; Heaton, 1992).

General Introduction to Mercury and the Mercury Cycle

Mercury (Hg) is a rare element in the earth's crust. In elemental form inn liquid state,
it is relatively nontoxic. In general, inorganic mercurials are not significant problems inn
environmental contamination In nature, mercury is distributed by a complex cycle involving
the atmosphere, hydrosphere arnd lithosphere. Mercury is released into the environment from
degase'ng of the earth's crust through volcanic gases into the atmosphere or by evaporation
fi'om waters. Mercury vapor is converted innto soluble forms of mercury and is returned to
the lithosphere by sedimentation from water and precipitates from the atmosphere. Once
depositied inn sediments, mercury rapidly and strongly binnds to those components that have
sulfur-containing organic and inorganic particles or iron and manganese oxides. Very little
mercury is founnd inn the environnmet in the unbonmd form. Mercury from industrial discharges
is mainly in the inorganic form. Once released into the environment, elemental mercury

becomes available for potential methylation by certain classes of organisms present inn the soil.

27
Inorganic mercury is nnethylated to mono-and dimethylmercury componmds (Williams, 1981).
Dimethylmercury is highly volatile and lipophilic and decomposes to the highly toxic and
stable methylmercury at acidic pH levels Methylation is a normal biologic process which may
occur inn anaerobic ecosystems which are associated in an inndustrialized society with polluted

water S.

Methylmercury and the Environment
In regard to the environment the mairn concerns are with the organic mercurials. There
are two major classes of organic mercurials, the aryl compounds and the alkyl compounds.
It is the latter group which poses the greatest threat to animals as the alkyl compounds are
highly toxic and innflict a wide range of damage fiom congenital mental retardation to
chromosome abnormalities while the aryl mercurials are rapidly metabolized to innorganic
mercury (Williams, 1981).

Methyhnercury is probably the most lethal compound of mercury. Although
methylrnercury can be converted to inorganic mercury, the rate of decompos'nion is slow. It
is completely reabsorbed when excreted in the bile and urinary excretion is low. The strong
carbon-mercury bond is not readily dissociated and the toxic effects are attributed to the
action of the intact molecule (Williams, 1981). Methyhnercury is rapidly cleared fiom water
starting with the uptake inn small organisms such as plankton and reachirng its greatest
concentrations inn large predatory fish Methyhnercury strongly binds to muscle and
accunnrlates with increased muscle mass.

The toxicity and target organs for methylmercury vary with different animal species. In

28

man and other primates, the central nervous system serves as the target. The fetal brain of
primates is critical if methyhnercury is inngested durirng pregnancy, while inn lower mannnals
the peripheral nervous system may be afi‘ected. Methyhnercury is easily absorbed by the
body, either in the gastrointestinal system, respiratory tract or through the skirn.
Methyhnercury readily crosses biological membranes and thus the brain is an easy target for
methyhnercury poisoninng as methyhnercury easily passes through the blood brain barrier after
being rapidly taken up by erythrocytes and distributed to all tissues and organs of the body.
Methylmercury decrmses the number of neurons inn the cerebelhnm causing permanent damage
due to the high amnity mercury has for the sulfinr in the sulflrydryl groups in the cell
membrane proteirns Mercury affects physiological finnctions in which proteirns are involved.
Neurological symptoms of methyhnercury toxicosis in humans occur from one week to
several months after exposure. Signs include: numbness of the lips, mouth, hands and feet,
ataxia, visual disturbances and difficulty in speaking. Moderate cases have displayed difficulty
in hearinng, tunnel vision and partial paralysis. With inncreased exposure, there are mental
changes, invohmtary movements, loss of vision, complete paralysis , coma and death

(Clarkston, 1983).

Human Poisonings
In the 1950's and 60's, poisoninng of humans and wilder inn Japan, Iraq and Sweden as
well as the high concentrations of mercury found inn freshwater fish in Canada, the northern
U. S. and Scandinavia (Joslinn, 1994) lead to much concern over mercury in the environment

(Nelson et. al, 1971). A factory close to the Minamata River in Japan using innorganic

29

mercury salts as catalysts, released an emuent containing mercuric chloride into the river.
This was transformed innto methyhnercury, which concentrated in the fissures of fish.

Some fish contained up to 20 ppm methylmercury. Widespread fatalities were reported
among the people of the nearby fishing villages that consumed the fish. Other epidemics
involvinng mercury, inncluding Niigata (1946-1965) and Iraq (1971-1972) also involved
sigrnificant mortality and morbidity. In the latter case, humans misused methyl and
ethylmercury fungicides. Farmers used the fimgicide treated grain for homemade bread

instead of for planting (WHO, 1976: Tsubaki and Irukayama, 1977).

Mercury Concentrations in the Tennessee River System

In recent decades, portions of the Tennessee River System have received major
industrial discharges of mercury including the Clinch River and Watts Bar Reservoir
downstream from Oak Ridge, Tennessee. While mercury concentrations in fish immediately
downstream from these sources have been found to be elevated, current mercury
concentrations have returned to acceptable levels (<0.5 ppm; Dycus, 1986). Piscivorous fish
from 26 U. S. states had mean fish flesh mercury concentrations that exceeded the widely used
criteion from advisories conceminng fish consumption by pregnant women (0.5 ppm; Wiener
and Stokes, 1990) and the recommended level of consumption by the general public (1.0
ppm) and thus a fish consumption advisory was published (Clean Water Frund, 1992). The
states with the fish containing the highest concentrations of mercury are Mirnnesota,

Wisconsin, Michigan and New York.

30

Effects of Mercury on Wildlife

Mercury has a great capacity to accunnnlate inn organisms occupying the upper trophic
level of food webs (Wren, 1987). Piscivorous mammals such as river otter (W
and mink (Museum can serve as sensitive indicator species of the adverse effects of
conuaminants inn the environment as they are top carnivores in aquatic food webs, relying on
fish for a large percentage of their diet. Mercury-contaminated fish in inland waters have been
reported and are often considered to be responsible for the decline in wild minnk and otter
populations (Wren, 1985; Mason et al, 1986). Mercury concentrations have been reported
in fish fiom many locations irncludinng remote unpopulated areas (Johnson 91 al, 1986).

Mercury has been reported inn the tissues of numerous piscivorous mammals and birds
(Eisler, 1987). The lethality of methylmercury to wilder has been documented numerous
times (Borg fl al, 1966; Aulerich g al, 1974; Wobeser 91 al, 1976). O'Connor and Nielson
(1981) reported that 2 ppm methyhnercury inn the diets of river otter caused death in two out
of three otters within 213 days. Mink are sensitive to dietary methylmercury, with fatalities
reported at 1 ppm inn the diet for two months. They are however, more tolerant of inorganic
mercury, where it is documented that 10 ppm in the diet for five months caused no adverse
effects The inorganic form of mercury is more readily excreted while the organic form easily
penetrates the brain which could account for the differences inn their toxicity (Eyle et 31.,
1970). Aulerich e a]. (1974) noted 5 ppm methyhnercury in the diet was lethal to adult mink
within one month. Accordinng to Kirk (1971), mirnk can be raised succussfirlly on diets that
contain up to 0.5 ppm mercury fiom contaminated fish, however, 1 ppm caused fatafity withinn

two months. Total consumption of 18 mg methylmercury caused death inn female mink

3 1
(Aulerich Q al, 1974) which compares with results reported by Hanko et a1 (1970) inn which
a total of 20 mg mercury caused death in female ferrets fed a diet that contained 5.0 ppm
methyhnercury.

The nervous syuem appears to be the target for methylmercury poisoninng in mink The
primary action of mercury is neurotoxicity and its severity is directly related to the amount
of mercury consumed. The reduction inn neurological function in carnivores is thought to
threaten their suvival inn the wild Mercury irnduced behavioral and reproductive efl‘ects have
been noted in mammalian and avian species in laboratory studies (Borg et al, 1969; Spyker
gt 3]., 1972; Khera, 1973). The clinical signs that have been observed in mink that have died
due to mercury exposure include: inncoordination, loss of balance, anorexia, loss of weight,
ataxia, paralysis, tremors, convulsions, high pitched vocalizations and death (Aulerich :1 al,
1974). These signs are similar to those reported for rats and cats (Takeuchi, 1970), pigs
(Piper 91 a1, 1971) and ferrets (Hanko et al, 1970) exposed to mercury. Wobeser et a].
(1976) observed merked posterior ataxia, shufling gait and rear leg ”splaying" in mink fed
1.8 ppm mercury in the diet. Symptoms of mercury inntoxication inn mink appear after a latent
period which varies inversely with the mercury concentration of the diet (Hanko et al, 1970;

Wobeser et a1, 1976;Au1erich et. al, 1974).

Effects of Prenatal Exposure to Methylmercury
Methylmercury is a recognized embryotoxic and teratogenic compound Human
evidence irndicated fetotoxicity of methylmercury at exposure levels innducing only slight and

reversible maternal toxicity (Marsh 91 a1, 1981). Animal studies confirm that serious brain

32

damage could be produced by prenatal exposure to methylmercury in offspring. Mercury
affects prenatal neuronal development which can lead to mental innpairment, behavioral
disorders, paralysis, retardation or death. Severely affected infants from the Minamata
outbreak had gross impairment of motor and mental development. In studies with rodents,
prenatal exposure to mercury caused cleft palate and other teratogenic malformations.

Mercury readily crosses the placenta where fetal blood concentrations are often higher
than those formd in maternal blood Fetal uptake of elemental mercury in rats has been shown
to be 10 to 40 times higher after exposure to inorganic mercury salts. After exposure to
alkyhnercnnic connpounds, fetal concentrations of mercury were twice those fonmd in maternal
tissues, and methylmercury concentrations in fetal red blood cells were 30% higher than inn
maternal red cells, which would enhance fetal exposure to mercury (Goyer, 1991).

Postmortem observations of humans that died of mercury poisoning in Japan indicated
that damage was generalized throughout the brain in cases of prenatal exposure in contrast
to adult exposure where focal lesions were predonninant. These prenatal cases inndicated a
disturbance of development inn the cytoarchitecture of the brainn and the brainn size was
dinninislned due to nerronal damage and inhibition of cell division during the critical stages of
formation of the central nervous system (Takeuchi, 1970).

There are very few laboratory investigations concerning the effects of methylmercury
poisonirng on the central nervous system during pregnancy. Most of what is known about the
effects on fetal survival is fi'om the Mirnamata outbreak involving humans inn Japan. It has
been observed that when the female's intake of mercury is large and she becomes ill, prenancy

does not occur. When the dosage is smaller, pregnancy occurs but the fetus is aborted

33
spontaneously or is stillborn. An even smaller dosage permits conception and live birth, but
the baby may suffer fi'om congenital diseases involving neurological frmction and mental
deficiency. In a study with pregnnant mice, 0.1 ppm methylmercury dicyanidannide was
injected innto the mice on day 10 of pregnancy and produced a high frequency of resorbed
litters and an increased percentage of dead fetuses (Harada, 197 8).

Wren (1987) found no significant differences in the average birth weight of mink kits
between treatment groups receiving 1 ppm PCB, 1 ppm methylmercury, or a combination of
the PCB and mercury. At three and five weeks of age, the average weight of the kits in the
group receivirng 1 ppm PCB and 1 ppm methylmercury inn combinration, was significantly
lower than the average kit weights inn the other treatment groups. At a level of 0.5 ppm PCB
and 0. 5 ppm methylmercury in combinnation, there were no significant differences in the
average weights of kits at three or five weeks.

Enqneriments have slnown that methyhnercury can pass the placental barrier and result in
even higher concentrations inn the fetus than in the mother. Infants born to mothers in
Minamata Bay, Japan, showed a syndrome consistirng of cerebral paresis, ataxia, mental
retardation, dysarthria and hypersalivation. Effects incurred prenatally, however, may not
become apparent until the nervous system has matured. High mercury concentrations have

been documented in fish eating human populations inn Greenland, Canada and Alaska.

Tissue Distribution of Mercury
Studies on mustelids have shown higher concentrations of mercury in the liver than in

the kidney (Wobeser and Swift, 1976; Kucera, 1983; Wren 91 al, 1986). Ropek and Neely

34

(1993) found that mercury concentrations inn the liver and kidney tissues were statistically
higher inn males than inn fenules. This contradicts the results of a study by Wren et al (1986)
who reported no differences in the mean tissue mercury concentration between the sexes.

Mink fed 5 ppm methylmercury had higher mercury tissue concentrations, even when
receiving the diet for a significantly shorter period of time when compared to mink fed twice
the concentration (10 ppm) of supplemental mercuric chloride for a longer period of time
(Aulerich 91 al, 1974). The concentration and distribution of mercury in the tissues of the
mink fed methylmercury differed considerably from those of minnk fed mercuric chloride.
Organic mercury tends to accunnnlate readily in the brainn while innorganic mercury does not.
Thus, the concentrations ofmercury are higher in the brairn than in the liver and kidney tissues
in organic mercury poisoning when compared to inorganic mercury poisoning.

Wobeser :1; II]. (1976) conducted a study with mirnk feeding them fish contaminated with
mercury at a concentration of 0.44 ppm and found that intoxication did not occur within the
experimental peiod In another study, Wobeser et a1 (1976) fed 1.1, 1.8, 4.8, 8.3 and 15.0
ppm dietary methylmercury to mink for 93 days. In general, the mercury concentrations in
the tissues of the mink that died were similar, despite difl‘erences in the mercury content of
the diet anndtimeto death Mercury concentrationsinthebrain and muscle tissue were similar
and lower than those inn liver and kidney of the same animal The mean concentrations of
mercury (ppm) in the tissues of mink that died were: brain 11.9, muscle 16.0, kidney 23.1,
and liver 24.3 ppm. It appears that in some regions, wild mink may be exposed to far greater
concentrations than 15.0 ppm inn their diet. Wobeser et al (1976) conducted postmortem

analysis on wild mink from Saskatchewan, Canada and found fur, muscle and liver mercury

35

concentrations of 34.9, 15.2 and 58.2 ppm, respectively. Wren et a1 (1986) fed mink diets
containing 1.0 ppm methylmercury and mortality and clinical sigrns were observed on day 73

and the average liver mercury concentration of the mink that died was 44.1 ppm

Pathology of Mercury Poisoning

Necrospy of nnink that have died from methyhnercury poisonirng have shown hemorrhagic
and congested hmgs, enlarged hearts, pale livers and kidneys, local superficial, hemorrhagic
gastric ulcers and enlarged and mottled spleens. Hyperennia, fatty and hydropic degeneration
of the liver and kidneys, splenic and glomerular amyloidosis and splenic giant cells were also
observed (Aulerich et a1, 1974). Studies with methylmercury have shown a variety of
neuronal innsults, including alterations in protein, DNA and RNA biosynthesis, changes in
phospholipid/phosphoprotein metabolites, abnormatlities in mitochondrial firnction and
perturbations in membrane permeability. Methylmercury also damages nnicrotubuli and has

caused pronninant changes in the dorsal root ganglia and peripheral nerves (Fox, 1990).
Like histopathological results from humans who died in the Minamata outbreak,
pathological changes inn experimental animals are also found primarily inn the cerebelhnm,
calcarinne cortec, and forsal root ganglia. Wobeser et al (1976) fed 1.1 to 15.0 ppm
methylmercury chloride to mink for 93 days and found that in all treatnnent groups the
characteristic lesions were essentially the same, though varyirng in irntensity, and consisted of
neuronal necrosis of the occipital cortex for all the mink with axonal degeneration. An
interference with proteinn synthesis in the nerve cells can occur as a result of methyhnercury

irntoxication Ultrastr'ucture examinnation revealed accumulation of lysosomes, membraneous

36

degenerations, degranulation and destruction of the rough endosplasnnic reticulum and
cytoplasnnic coagulation The disintegration of the endoplasnnic reticulum and degranulation
of ribosomes in nen'ons suggests alterations in both RNA and prtoeinn metabolism in the nerve
cells. Axonal degeneration and disirntegration of myelin inn peripheral nerves has also been
observed (Fox, 1990).

In animals exposed in mm to methylmercury, loss of nerve cells along with
cytoarchitectural abnormalities can be irnduced. Small hemorrhages in the cortex and white
matter of the brain have been observed with methylmercury poisoninng. Examination of the
brainn revealed neuronal destruction, disruption of neuronal migration and incomplete
cerebellar granule cell layer formation Besides non-specific cytological changes such as
lysosomal accunnnlation, disintegration of endoplasmic reticulum, and cytoplasmic
degeneration, there have been reports of incomplete myelination of axons, abnormal

formation of myelin sheaths and abnormal synaptic development (Fox, 1990).

MATERIALS AND METHODS

FISH

Approximately 719 kg of whole Atlantic mackerel were obtained fi'om Boston Feed
Supply, (Natick, ME) for use as an uncontaminated species for the control mink diet. These
fish were stored frozen in a walk-in outdoor freezer at -6.7° C. Approximately 392 kg of
various species of fish were collected from the Clinch River upstream of the U. S. Department
of Energy (DOE), Oak Ridge Reservation (ORR) to serve as an "uncontaminated" (control)
source of fish for feedirng to mink. Seven hundred fourty two kg of fish believed to contairn
various environmental contaminants, particularly PCBs and mercury, were collected from the
Watts Bar Reservoir-Clinch River sytem downstream from the ORR for feeding to mink All
fish were collected by gill netting and kept frozen at the Oak Ridge National Laboratory until
shipped to Michigan State University (MSU) by overnight Federal Express in plastic bags
withinn "secmity sealed" coolers. A chairn of custody form accompanying the fish was signed
and dated when the fish were received at the MSU Experimental Fur Farm

The first shipment of fish fiom Oak Ridge National Laboratory arrived at the
Experimental Fur Farm on August 25, 1993. The final shipment of fish was received on
November 30, 1993. Upon delivery, the bags of fish were weighed and the weights

recorded. The fish were placed in heavy plastic bags, sealed and labelled with a tag

37

38
identifyinng the bag number, collection site, date, project identification number, and the
marchers initials. This procedure was repeated a second time so that all fish were sealed
within two heavy plastic bags, each bag containing its own identification tag. The bags were
stored in a dedicated section of a walk-in fieezer at the Experimental Fur Farm at -6.7° C
until needed for diet preparation. The up stream fish were kept separated fi'om the
downstream fish.

Upon acquisition of all the fish fiom Oak Ridge National Laboratory, both upstream and
downstream fish were taken from the freezer, removed from the sealed bags, and sorted by
species, making sure that the up stream fish remained in one area of the facility and the
downstream fish in a different area. For each species, the fish were cournted and a total
weight by species was recorded. The species were identified by Dr. Thomas Koons,

Professor, MSU Department of Fisheries and Wildlife (Appendix A).

W

The minnk diets were prepared using the equipment at the MSU Experimental Fur Farm
The experimental diets were formulated to meet the nutrient requirements of mink
(NRC,1982). The fish portion of the two control diets containned 75% "rmcontaminated"
Atlantic mackerel ( Diet A) or 75% "uncontaminated" fish collected upstream from OR
(Diet B). The treatment groups diets contained 75% fish with either 25,50, or 75%
"contaminated" fish from downstream ORR and the remaining portion consisting of 50, 25
or 0% Atlantic mackerel (Diets C,D,E ).

Approximately 358 kg of the frozen Atlantic mackerel were thawed and ground

39

through a 3/8 inch face plate and mixed inn a paddle mixer for 15 minutes. The ground
mackeel was then removed fiom the mixer and the process was repeated for a second batch
of approximately 358 kg of mackerel The two batches of mackerel were then mixed together
so that a homogeneous mixture was obtained. Thirteen samples (500g each) of the mixture
were placed in whirlpac bags, labelled with the diet code, date, project identification number,
and researchers initials and frozen in a chest freezer dedicated to the project. Both the
up stream and downstream fish from OR were ground separately and mixed inn a paddle
mixer and 13 samples (500g each) of each were frozen for subsequent analysis as described
above for the Atlantic mackerel

The diets were prepared by blending the appropriate quantity of the prescribed fish with
the appropriate quantities of the other components of the mink diets (eggs, liver, vitamin and
nnineral premix, d-biotin, and cereal) in a paddle mixer for 15 to 20 minnutes. Thirteen samples
(500g each) of each diet were placed in whirlpac bags and stored frozen inn a chest fieezer
dedicated to the study for subsequent analyses. All samples were labelled with the diet code,
project identification number, date, and researchers initials. The prepared mink diets were
placed in plastic buckets lined with plastic bags. Each bucket contairned a three-day supply
of feed for the 10 animals on a particular treatment (approximately 6.8 kg). The bags were
sealed with a twist-tie and an identification card with the diet, diet color code, date, project
identification number, and researchers initials was placed in the bucket and the buckets were
sealed with plastic lids. The buckets were stored in a dedicated area of a walk-in freezer at

-6.7° C rmtil needed for feedinng.

40

W

On December 2, 1993, after a two-week acclimation period, 50 standard dark mirnk
(Mustela vison) were randomly assigned to the five treatment groups. Each treatment group
consisted of two males that had previously sired litters and eight females. Care was taken
so that littermates were not placed within the same treatment group in an attempt to reduce
any genetic predisposition to heavy metal or PCB toxicity.

The mink were housed inrdividnally in wire cages (76 cm LX61 cm W x 47 cm H) with
attached nest boxes ( 38 cm LX30.5 cm W X30.5 cm H) and bedded with wood shavings
(Pestell Agni-Products, Ontario, Canada). Prior to whelpirng, the female's neebox was bedded
with aspen wood shavings to prevent the kits from being exposed to terpirnes in the wood
shavirngs, which are toxic to young kits, and "wood wool” excelsior (American Excelsior
Conpany, Arlington, TX). A false floor (1/ " wire mesh) was fitted innto the cage to prevent
newbonn kits fi'om falling through the 1X1 1/2 innch wire mesh cage floor. A wooden nestbox
divider was fitted irnto the nestbox to prevent the young kits from crawling out of the nestbox
into the cage. Feed and water were provided ad flhmnm throughout the study. The minnk
were previously immnmized agairnst canine distennper, virus enteritis, hemorrhagic pneumonia,
and botulism (Biocom-DP; United Vaccines Inc., Madison, WI)

The animals were inndividually identified, each having an identification card with the
project identification number, mink number, and color-coded diet letter above the individual
nnink's cage. A color-coded feed tag identifying the diet was also fastened to the lid of each

cage for ease in identifying the appropriate diet during feeding.

41

The animals were observed daily and any behavioral changes or clinical signs of toxicosis
were recorded Any adult mink that lost 30% of their original body weight were euthanized
with carbon dioxide gas (C02). A necropsy was performed by a veterinary pathologist (Dr.

J .A Render) on all mink that were euthanized or died before the end of the trial.

EEEDJNCLTBIAL
All animals were acclimated over two weeks to the test facility prior to the start of the

feeding trial which began on December 2, 1993 and ended in mid-June 1994. During the
peiod of acclimation, the mink were fed a basal diet in excess of what they would consume
each day and were weighed twice. After the acclimation period they were weighed at
monthly intervals.

A three-day supply (1 container) of each diet was removed from the freezer and allowed
to thaw overnnight at room temperature. Approximately 0.23 kg of feed was placed on a wire
feed grid on top of each cage. The remainninng unused feed was stored in a walk-in cooler
(1.7 ° C). Ifall the feed inn a bucket was not used within a three-day period, it was discarded.
Each morning, the previous day's feed was scraped ofi‘ the feed grid and discarded before
providing flesh feed.

Because some species of fish contain the enzyme thiaminase which hydrolyzes thiamine
resulting in Chastek's paralysis, a thiamine supplement was given to the mink. Each day, 0.4
mg of thiaminne was dissolved in 50 ml of water and mixed into 1 kg of a basal farm diet
(containing thiaminase-flee fish). Each mink was fed approximately 20 g of the thiamine-

supplemented feed daily before the treatment diet was provided.

42

REPRODUCTION.

Mating of females to the males withinn their respective treatment group began March 1,
1994 and ended March 22, 1994. Females were given the opportunity to mate every fourth
day until a confirmed mating (presence of motile sperm in a vaginal aspiration) was obtained.
Once a confirmed mating was obtainned, the female was given the opportnmity for additional
matings the day following the initial mating and/or eight days later.

During mating attempts, males were locked out of their nestboxes and the females were
introduced into the male's cage. If no evidence of mating was observed within the first 15
nninutes, the female was retmned to her cage and given a check mark on her breeding records
for that day. If mating appeared to be occurrinng, the pair was left alone until they separated
The female was then taken into the laboratory where a pipet, containinng a small amount of
warm saline, was inserted into her vagina. An aspiration was taken and placed on a glass
slide and examined under a microscope. If motile sperm were found, the female was
considered bred and the male's identification number was written on her breedinng chart for
that day. The male was given an "X" for the day on his chart to indicate that he had produced
motile sperm in a mating. If no sperm or non-motile sperm were found inn the aspiration, the
female was either given the opporturnity to mate with the other male in her treatment group
that day or given the opportunity to mate the following day. Mating attempts were continued
throughout the breeding season nmtil at least two confirmed matings were obtained for each
female.

All nestboxes were checked daily during the whelpirng season for newborn kits. Newborn

kits were sexed, counted, and weighed at birth. Any dead (stillborn) kits were sexed,

43

weighed, and placed in whirlpac bags and flow for firture analysis. The mother's body
weight was also recorded at whelping. The dam and all surviving kits were counted and
weighed again at three and six weeks of age.

At three weeks of age, the females that whelped and their kits were fed the appropriate
diet, which had been watered down, on a feed plate placed on the bottom of the cage in front
of the nestbox entrance. The diet was mixed with water so that it's consistency would
encourage the fits to begin consumirng "solid" feed. This feed was fed over the next three
weeks with the amount of added water gradually reduced until the fits were consunning the
solid diet at six weeks of age. Once the kits were old enough to venture out of the nestbox
and begin consuminng solid feed (three weeks of age), the nestbox divider was taken out to

promote easy access from the nestbox to the cage.

W

Blood samples were collected from the eight adult female mirnk fed Diets B and E (7 5%
upstream fisln and 75% downstream fish) during the acclimation period (November, 1993) and
in February and J1me, 1994 for measurement of various hematologic parameters. All females
were anesthetized with 0.4 mg Ketaset (ketamine hydrochloride; F ort Dodge Laboratory,
Inc., Fort Dodge, IA) innjected intrannrscularly innto the hind leg. Four ml of blood were
collected for a serum biochemistry profile in Serum Separation Tubes and 0.5 ml blood was
collected for hematologic measurements in Microtainner Tubes coated with EDTA

The seum biochemical analyses and calculations were performed by the MSU Veterinary

Clinical Pathology Laboratory with an Abbott Spectrum Analyzer (Abbott Laboratories,

44

Dallas, TX) to determine calcium (Ca), chloride (Cl), iron (Fe), phosphorus (P), potassium
(K), magnesium (Mg), sodium (Na), carbon dioxide (CO2), anion gap, total proteinn, albumin,
globulin, albunnin globulin ratio (A/G), creatininne, alkaline phosphatase (Alla Phos.), aspartate
annino transferase (AST), creatine finase (CK), gamma glutamyl transpeptidase (GGTP),
sorbitol dehydrogenase, cholesterol, glucose, blood urea nitrogen (BUN), and osmolality.

A Technicon H1 system (Technicon Diagnostic Systems Division, Tarrytown, NY) was
used for the determination of the red blood cell (RBC) count, white blood cell (WBC) count,
hemoglobin (HGB), hematocrit (HCT), mean corpuscular volume (MCV), mean corpuscular
hemoglobinn (MCH), mean corpuscular hemoglobinn concentration (MCHC), and total platelets

(PLT).

NECRQBSX

Two dates for necropsy were selected so that all the fits would be at least six weeks of
age and not over four days older than the youngest fit on the day of the necropsy. On June
8, 1994, 17 females and 24 fits were examined. On June 15, the remaining females, their
fits, and the breeder males were examined.

The females were anesthetized with Ketaset (ketamine hydrochloride). After losing
consciousness, the females were weighed on an electronic balance and their weights recorded.
Their backs were shaved using electric clippers and 5g of hair were collected from each minnk
for mercury analysis. The hair was placed inn a labelled whirlpac bag and frozen for
subsequent analysis. Twenty ml of blood were collected from each adult mink via heart

purncture usirng a 5 cc syringe and 18 gauge 1 1/4 inch needle. Tenn ml of blood were placed

45

in each of two labelled test tubes The sennm was separated by centrifugation. Enough blood
to fill two nnicrohematocrit tubes was taken from each adult via toe clip. For the females fed
Diets B and E, an additional four ml of blood were collected via heart puncture using a 5 cc
syringe and 18 gauge 1 1/4 inch needle for a biochemical profile and various hematologic
measurements which were conducted by the MSU Veterinary Clinical Pathology Laboratory.

After the blood samples were obtained, the females and fits were placed in a tightly
sealed wooden box and euthanized by an overdose of carbon dioxide gas. The brain, liver,
fidneys, spleen, hangs, heart, ovaries, thyroids, adrenals, and a sample of adipose tissue were
collected from the adults at necropsy. All tissues, except for adipose tissue, were trimmed and
weighed. The weights of all organs were recorded. Portions of the liver, spleen, fidneys,
hmgs, ovaries, and brain were placed in a 10% neutral-bufi'ered formalin solution inn labelled
glass jars for subsequent histopathologic examination The rermining portion of these tissues,
along with the adipose tissue, was wrapped inndividually in labelled tin foil, frozen in liquid
nitrogen, and stored inn an ultra-cold freezer at -57 ° C until shipped to Oak Ridge National
Laboratory for analyses The liver, fidneys, and spleen were taken fi'om the fits and weighed.
A portion of each tissue was then frozen in liquid nitrogen and stored in the ultra-cold freezer.
Additionally, a sanrple of each tissue was placed in formalin for histolopathologic examination
as previously described for the adults. All carcasses of kits were wrapped in labelled tin foil,
frozen in liquid nitrogen, and Stored in the ultra-cold freezer until shipped to Oak Ridge
National Laboratory for analyses. The reproductive tracts of the females that did not whelp
were excised and their ovaries and uteri were examirned for corpura lutea and implantation

Sites under a dissecting microscope.

46
111W

Afier fixation, the organs collected for histolopathologic examination were trimmed,
processed accordinng to routine histologic procedures, sectioned at Sum, stained with
hematoxylin and eosin and examined usirng a light microscope. The blood collected for
analyses was centrifinged for 6 minutes at 1200 rpm The serum was collected and placed in
labelled plastic centrifirge tubes and stored in the ultra-cold freezer at -57 ° C for subsequent
analysis by Oak Ridge National Laboratory. The nricrohematocrit tubes were centrifuged
for 6 nninutes using a nnicrocapillary centrifirge and the percent packed cell volume determined
with a nnicrocapillary reader and recorded

The frozen tissue samples, carcasses, hair and serrun samples were released to Dr.
Richard Halbrook at the terminnation of the trial The tissues stored in formalin for
histopathological examination were submitted to Dr. James A Render, a veterinnary

pathologist in the Department of Pathology, Michigan State University.

CHEMICALANALXSIS

Tern subsamples of each diet (A,B,C,D, and E) as well as subsamples of each of the
homogenous mixtures of Atlantic mackerel, and fish from upstream and downstream fi'om
ORR were shipped frozen on dry ice in a sealed cooler via Federal Express overnight mail to

Oak Ridge National Laboratory for total PCB and mercury analyses.

47
W

Samples of each diet (A,B,C,D, and B) were submitted to Litchfied Analytical Services,
Litchfield, MI for analysis of nutritive content (proximate analysis). The diets were shipped

frozen on dry ice in a Styrofoam container.

STATISTICS

Data were analyzed using statistical software (SAS Institute Inc., 1987). Statistical
treatment of the data to determinne the treatment means and standard errors was by the
General Linear Models procedure. Treatment effect was determined by one-way analysis of
variance (ANOVA). Comparisons among treatments were analyzed using Tukey or Schefi‘e's

tests The significant differences between means were based on p_<_ 0.05.

RESULTS

Major dietary components and the nutrient analysis of the mink diets are shown inn Table
1. Crude protein, fiber and fat did not differ between the diets by more than 3.5 percent.
Moisture, protein, ash and crude fiber were relatively constant inn all diets. Fat content was
highest in the Atlantic mackerel diet and decreased with inncreasinng percentages of
downstream fish Thus, Diet C had the highest percent fat and Diet E had the lowest percent
fat. The percent TDN (total digestible nutrients) also followed this pattern. The reason Diet
B had 108-130 ppm more zinc than any other diet is unknown.

The raw fish used in the Study contained from 0.03 to 0.35 ppm mercury and 005-2. 13
ppm PCBs (Table 2 ). AS shown in Table 3 , total dietary mercury concentrations ranged

from 0.02-0.22 ppm wet weight and total dietary PCB concentrations ranged fi'om 0.04-1.86

PPm

W

Mean monthly body weights are summarized inn Table 4. Adult mean body weights and
body weight changes are summarized in Table 5. No statistical differences inn innitial or final
body weights were noted among the treatment groups. Females in each treatment group lost
weight, on average, as the trial progressed. Female mink fed the diets with the highest PCB

concentrations (D and E) had the greatest body weight losses (5%). This trend, however,

48

  
 

  

Table 1 .

42

T

Composition and nutrient analysis of experimental diets

    

    

 

  

 

Diet A Diet 8 Diet C Diet D Diet E
Ingredients
Ocean fish, ‘161 75 -- 50 25 --
Upstream fish, %2 -- 75' -- -- --
Downstream fish, %3 -- -- 25 50 75
Cereal, %‘ 18.5 18.5 18.5 18.5 18.5
Eggs, % 3 3 3 3 3
Beef liver, % 3 3 3 3 3
d-biotin, mg/kg‘ 0.05 0.05 0.05 0.05 0.05
Vitamin/mineral premix, %° 0.5 0.5 0.5 0.5 0.5
Nutrientjnalxsiflmtnaeighn’

Dry matter. % 43.68 42.00 43.74 43.50 40.92
Fat, % 10.37 7.64 9.10 8.00 6.85
Crude protein, % 19.85 19.32 20.07 19.68 18.90
Crude fiber, % 0.90 1.09 0.92 1.04 0.96
Total digestible nutrients, % 42.25 36.89 40.82 38.55 35.36
Calcium. % 0.68 1.51 0.95 1.39 1.42
Phosphorus, % 0.54 0.77 0.64 0.83 0.80
Potassium, % 0.41 0.35 0.41 0.39 0.36
Magnesium, % 0.07 0.08 0.08 0.09 0.08
Sodium, % 0.2 0.204 0.217 0.21 0.198
Iron, ppm 148 189 169 188 175
Manganese, ppm 54 66 65 60 61
Copper, ppm 14 16 18 15 17
Zinc. Ppm 104 234 119 126 113
Ash, % 3.76 5.42 4.16 5.39 5.39

 

Supply, Natick, MA
Oak Ridge, TN

Oak Ridge, TN

Lansing, MI

 

‘ XK-40 Mink Food, XK Mink Foods Inc., Plymouth, WI
5 Sigma Chemical Co., St. Louis, MO
" MSU Swine Vitamin Trace Mineral Premix, Michigan State University, East

7 Proximate analysis of diets by Litchfield Analytical Services, Litchfield, Ml

 

' Atlantic mackerel, GMF Brand Ocean Fresh Fish Product, Boston Feed
2 Fish from upstream of the U.S. Dept. of Energy Oak Ridge Reservation,

3 Fish from downstream of the U.S. Dept. of Energy Oak Ridge Reservation,

 

 

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54

was not observed for the male mink, where mean body weights inncreased during the trial for
all diets except Diet B. The greatest male body weight gains were observed in Diets D and

E which contained the highest concentrations of PCBs.

ADIILLMORIALUX

On February 2, 1994, two months after the feeding trial began, a male mink fed Diet A
had blood in his feces which continued sporatically throughout the remairnder of the trial In
mid-March, the second male fed Diet A, had blood in his feces. This was followed by
vomiting and feed refusal The male died two days later. In June, a female fed Diet C was
erthanized with ketamirne hydrochloride afier refusing to eat for several days, being lethargic
and passing tarry feces. A cyst was found on her ovary and her gallbladder was enlarged
and a gallstone was also found Another female fed diet C had dystocia. Examination of
the reproductive tract showed 10 dead fetuses which had no visible abnormalities. The
following day the female was euthanized after passing tarry feces and being lethargic and

refusing feed.

WON

Reproductive performance of the females is shown in Table 6 . Of the 40 females bred,
31 whelped The average number of confirmed matings per female inn the treatment groups
ranged from 2.0-3.0. Of the nine females which did not whelp, four were fed Diet C. The
only group in which all the females whelped was Diet E. The average gestation lengths

across the treatment groups was 44.3-47.5 days. Females fed Diet C had the shortest mean

55

 

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56

gestation period while females fed Diet D had the longest gestation period. The percentage
of ln'ts born alive ranged from 77-95%. Females fed Diets C and A had the fewest kits born
alive. Average litter sizes at birth ranged from 4.3-7 .8 across the treatments. Females fed

Diet C had the largest litters while females fed Diet E had the smallest litters.

W

Reproductive performance of the adult male mink is shown in Table 7 . All males were
proven breeders with those males fed Diets A or D had 100% successful matings. One male
fed Diet B had a lower percentage of successful matings (77%), but this percentage was not

considered to be abnormal

KILMQRIALHX

Differences were seen in the proportion of mortality that occurred during birth, three and
six weeks of age ( Table 6). Except for kits of females fed Diet E, which had no mortalities
after birth, the mortality levels were highest from birth to three weeks of age. After three

weeks of age the only kits to die were those of females fed Diet B.

W
Kits whelped by females fed Diet C had the lowest birth weights, while those whelped

by females fed Diets D and E had the greatest birth weights (Table 6 ). At three weeks of
age, kits fi'om Diets A and C had the greatest body weights while those ldts fi'om Diets B and

E had lower average body weights At six weeks of age, the kits from females fed Diet C had

Table 7. Reproductive performance of male mink fed diets containing

various percentages of 'contaminated" and “uncontaminated“

 

 

Total no. No. matings 96
Male attempted with motile successful
Diet no. matings sperm matings

r A 1 1 O 1 0 100
2 12 12 100 l

B 3 1 O 9 9O

4 1 3 1O 77

C 5 1 O 9 9O

6 9 9 100

D 7 9 9 100

8 7 7 100

E 9 1 3 1 2 92

1O 1 3 1 2 92
_

 

 

 

58

the greatest body weights while the kits whelped by females fed Diet E had the lowest body

weights.

AWE

Mean adult organ weights are summarized in Table 8 . No statistical differences were
noted among the females in the treatment groups for brain, liver, kidneys, spleen, lungs, heart,
ovary , thyroids, and adrenals A significant difference was observed among adult male liver
weights The liver weights of the males ranged from 46.4-95.3g with Diet A males exhibiting

the highest liver weights and Diet C males, the lowest.

W

Thirty-five kits were euthanized at six weeks of age and subjected to necropsy. No
necropsies were performed on kits from Diet D. Kit organ weights taken at necropsy are
summarized in Table 9 . There were no statistical difl‘erences noted between the treatment

groups for liver, kidney or spleen weights.

HECEQRSX

Gross examination of the kits and the adult mink at necrospy revealed no abnormalities.

W
Histopathlogical examination of the male from Diet A that died revealed that the mucosa

of the urinary bladder was ulcerated with an intense neutrophil infiltration of the submucosa

59

 

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61

and fibrin exudation A marked necrotin'ng vasculitis with vascular thrombosis and fibroblast
proliferation was present. Numerous bacterial colonies were found on the luminal surface.
The kidneys had petechial and ecchymotic hemorrhages, and the urethra had petechial
hemorrhage. The cause for the clinical signs appeared to be due to a severe
fibrinosuppurative and focally hemorrhagic bacterial cystitis.

The majority of the adults examined from all treatment groups had a prominence of Ito
cells in the liver. Mild hepatocellular microvesicular cytoplasmic vacuolation was observed
in several adults fed Diets A,B,C and D. Lymphocytes were present in the portal areas of the
liver fiom an adult fed Diet A and one fed Diet B. Mild microvesicular vacuolation of renal
tubular cells was observed in adults fed Diets C and E. Splenic amyloidosis was observed in
three adult mink each fed a difi‘erent diet ( Diets C,D and E). Megakaryocytes were also
present in the spleen of a majority of the adults fed Diets C,D, and E. No histopathological

alterations were present in the liver, kidney or spleen fi'om kits in each treatment group.

W

No significant effects on hematologic parameters were present between the females fed
Diets B and B (Table 10 ). The white blood cell (WBC) counts for both treatment groups
during the month of February increased significantly compared with the WBC count noted

in November during acclimation and at the termination of the trial

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W

No significant differences were noted among the females fed Diet B or Diet E for serum

chemistry parameters for any of the three collections during the trial ( Table l 1).

W
The cummulative PCB dose ranged from 1.19 mg/mink to 55.24 mg/ mink with the dose

for males ranging from 0.0024 mg/kg/day to 0.124 mg/kg/day and the dose for females
ranging from 0.0047 mg/kg/day to 0.23 mg/kg/day (Table 12; Appendix B). Concentrations
of total PCBs in livers of adult mink ranged from 0.02 pg PCB/g liver (wet wt.) for those fed
Diets B and C to 7.25 pg PCB/g liver for those fed Diet E. Mink fed Diet A had livers with
total PCB concentrations averaging 0.03 pg/ g liver (wet wt. ). The concentration of total
PCB in the liver of mink fed the diets containing downstream fish increased in a dose-
dependent manner.

Concentrations of total PCBs in fat of mink ranged from 0.25 pg PCB/g fat (wet wt.) to
106.7 pg PCB/g fat in the mink fed Diet E. Females fed Diet B had the second highest
concentration of PCBs in their fat at 56.6 ug PCB/g fat. The concentration of total PCB in
the fat of mink fed the diets containing downstream fish increased in a dose-dependent
manner. The fat samples of the mink contained numerically higher concentrations of PCBs
than the liver samples The concentrations of various PCB congeners detected in the liver and

fat tissue of the mink are presented in Tables 13 and 14.

64

 

 

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68

0; \Iilaiis 0 uli 1L lsa'l .lbsl ala’t as” UL
IISSHE
The cumulative mercury dose ranged from 0.594 mg/ mink to 6.53 mg/mink while the
mercury dose for the males ranged from 0.001 mg/kg/day to 0.015 mg/kg/day and for the
females ranged from 0.002 mg/kg/day to 0.027 mg/kg/day (Table 15; Appendix C).
Concentrations of total Hg in the kidneys of adult mink ranged from 0.79 pg Hg/ g kidney
(wet wt.) for mink fed Diet A to 4.23 pg Hg/g kidney for mink fed Diet E ( Table 16 ).
Concentrations of total mercury in the heart of adult mink ranged fiom 3.44 pg Hg/ g heart
(wet wt.) for mink fed Diet A to 17.82 pg Hg/g heart for mink fed Diet E. Concentrations
of total mercury in the liver of adult mink ranged from 0.38 pg Hg/g liver (wet wt.) for mink
fed Diet A to 3.27 pg Hg/g liver for mink fed Diet E.

The concentrations of mercury in all tissues of those mink fed the diets containing
downstream fish (Diets C,D, and E) increased in a dose-dependant manner. The mercury
concentrations in tissues of mink fed Diet B were higher than those of mink fed Diet A .
The concentrations of mercury in the tissues of mink fed the upstream fish were higher than
those formd in the tissues of mink fed the downstream fish. The heart contained numerically

higher concentrations of mercury than the kidney or liver for every diet group.

69

 

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[n1

DISCUSSION

The primary goal of this study was to assess the effects of consumption of environmental
contaminants contained in fish collected from the ORR on mink as a method of evaluating the
extent and degree of adverse effects of ORR operations on wild piscivorous populations.
Attempts were made to sinmlate the exposure of mink to an environmentally-contaminated
diet. Mink are opportunistic predators. While a great proportion of their diet is comprised
of fish (30%), they will also eat amphibians, reptiles, birds and other mammals (Heaton 91 a1,
1995). Even though mink may consume a variety of prey species, concentrations of
compounds like PCBs and organochlorine pesticides in fish correlate well with
concentrations found in mink tissues. Thus, it can be concluded that fish comprise a
significant vector of exposure of wild mink to environmental contaminants. In the present
audy, fish comprised an abnormally large percentage (75%) of the diet and this may in part
accomt for some of the reproductive problems that were observed. The concentration of fish
in the diets did not follow the standard operating procedure followed by the MSU
Experimental Fur Farm, but was specified by the biologists at Oak Ridge National
Laboratory. Commercial mink diets typically contain less than 50% fish (Aulerich, personal
communication).

In addition to PCBs, the fish collected from the Oak Ridge Reservation undoubtedly

contained other organochlorine contaminants and variable concentrations of mercury. The

71

72

fish were only analyzed for PCBs and mercury as these were the primary contaminants of
interest to ORR and thus were the focus of this research. Therefore, it is possible that effects
atlnhrted to total PCBs and mercury in this study my be due in part to the presence of other
contaminants in the fish. Because the fish collected from upstream of the ORR ( originally
considered to be "uncontaminated") were found to contain PCB concentrations near that of
the downstream "contaminated" fish (1.69 vs. 2.13 ppm, rewectively) all the diets that
contained fish from ORR were considered as treatment groups. The diet comprised of
upstream fish (Diet B) was not considered as a control when calculating statistical differences.
The fish collected from upstream of the ORR contained considerably less mercury than those
fidr collected downstream from the ORR (0.07 and 0.35 ppm respectively). Thus, operations
at ORR appear to be contributing to the contamination of the river system with mercury.
Many r0searchers have documented a "wasting syndrome" or marked reduction in body
weights of mink associated with halogenated hydrocarbon intoxication (Aulerich 91 al, 1987;
Hochaein g, at, 1988). Other species, including rhesus monkeys (Barsotti 91 al, 1976) and
rats (Courtney 91. fl, 197 8) have also exhibited a marked reduction in body weights when
exposed to PCBs. This weight loss could be due in part to the dose-dependent decrease in
food conmmption observed in animals exposed to PCBs. No statistical difl‘erences in initial
or final body weights were noted among the mink in the treatment groups in this study.
Although feed consumption was not measured in this study, emperical observation suggested
that the mink fed the higher concentrations of PCBs did not decrease their daily food intake
during the study period Heaton (1992) did not observe the "wasting syndrome" in mink fed

PCB-contaminated fish, though she did observe other clinical signs of toxicity. Perhaps the

73

wasting syndrome was not observed because of the greater caloric value obtained from
feeding the mumally high percentage of fish in the mink diets. This may also partly explain
why no significant body weight losses were observed in the present study . The body weight
loss observed for Diet D and E females (5%) is not unusual for lactating females because of
the greater energy demand for milk production. Also, Diet E females had, on average, more
kits to nurse than females in the other treatment groups, thus placing an even greater demand
on the female's energy reserves that would normally allow her to gain or maintain body
weight.

Previous studies of the efl‘ects of PCBs on mink by Aulerich gt a]. ( 1971, 1973, 1986)
Bleavins 91 al ( 1980), and Wren ( 1987) showed that PCBs have a detrimental effect on
adult mink srnvivability. Numerous investigations have documented the toxicity of PCBs to
adult mink. Bleavins 01 a]. (1980) demonstrated the sensitivity of adult mink to PCBs. In
general, the mean survival time of mink fed Aroclor 1242 was inversely related to the
concentration of PCB. Adult mink fed 20 ppm Aroclor 1242 experienced 100% mortality
within nine months while 10 ppm caused 66.7% mortality within nine months. All animals
that died were subjected to necropsy which revealed emaciation characterized by almost
complete absence of body fat. This "wasting syndrome" is commonly associated with
halogenated hydrocarbon poisoning and has been documented for mink (Aulerich and Ringer,
1979; Aulerich g al, 1985). This body weight loss is related only in part to decreased food
intake and the cause of death is unknown.

In another study by Aulerich and Ringer (197 7 ), deaths occurred earlier and mortality

was greater for mink receiving metabolized Aroclor 1254 than for those mink fed the same

74

concentration of a technical grade Aroclor 1254. At 75 ppm, 80% of the mink receiving the
metabolized form died within 28 days, while 50% of those receiving the technical grade died
within the 28 days Aulerich a al,(l985) found that even greater mortality occurred in adult
mink fed considerably lower concentrations of congeners having chlorine atoms in three of
the four lateral positions in the aromatic ring system

Mink have been shown to be more sensitive to PCB-contaminated fish canning by-
products (heads, fins, tails, viscera and belly fat) than to the whole raw fish (Aulerich and
Ringer, 1970). The higher rate of mortality fi'om consumption of comparable quantities of
the canning by-products was attributed to the higher fat content of the by-products. PCBs
are stored in the fatty tissue and thus the mink were exposed to higher concentraitons of
PCBs in the fatty by-products.

Several studies have indicated that female mink may be more susceptible to PCB
contamination than male mink Although no differences were found in mortality rates
between males and females fed diets supplemented with 10, 20 or 40 ppm Aroclor 1245,
Aroclor 1242 fed at 5 ppm or Aroclor 1016 fed at 20 ppm caused mortality limited to female
mink. Male mink have also been shown to survive longer than females when fed low
concentrations (0.64-3.57 ppm) of Aroclor 1254 (Platonow and Karstad, 1973). Parkinson
and Safe (1987) suggest that females are more susceptible to PCB toxicity than males because
males have a higher drug metabolizing activity and the toxicity of PCBs to mink is inversely
related to their drug metabolizing acitivity.

The clinical signs that have bear observed in mink that have died during exposure to

PCBs include listlesness, nervousness, bloody stools, and anorexia (Bleavins 91. al, 1980;

75

Aulerich 5 al, 1986). In the present study, mortality was limited to one adult male mink fed
Diet A while two female mink were euthanized that were fed Diet C. Althought two of the
mink were "of feed" for two days before death and passed tarry stools, none of the mink
displayed the other signs commonly observed with the "wasting syndrome" associated with
PCB intoxication. In mink, tarry stools is a common clinical sign associated with anorexia.
In this study, mortality from arch a low level of exposure to PCBs would not be expected to
occur. The male mink that died had a fibrinosuppurative and focally hemorrhagic bacterial
cystitis and one female that was euthanized had conglications associated with a gallstone and
had a cyst on her ovary. The other female that was euthanized had complicationsdue to
dystocia but no necropsy was performed.

The clinical signs that have been observed in mink that have died due to mercury exposure
include incoordination, loss of balance, anorexia, loss of weight, ataxia, paralysis, tremors and
high pitched vocalintions (Aulerich 0 al, 1974). No clinical signs of mercury exposure were
observed in the current study. Mortality of mink would not be expected to occur with the
low concentration of mercury present in the experimental diets when compared to those
concentrations causing death in mink in other studies. For example, Aulerich et al (1974)
found that 5 ppm methylmercury was lethal to adult mink within one month and that total
consumption of 18 mg methylmercury caused death in female mink, which compares to
results reported by Hanko 91. al (1970) who found that consumption of 20 mg mercury
caused death in female ferrets fed a diet containing 5.0 ppm methyhnercury. The
cummulative mercury consumption for the adult mink in the current study ranged fi'om 0.594

mg/mink to 6.534 mg/mink (Table 15; Appendix C). Thus, the cumnmlative mercury

76

consumption for even the highest treatment group in the current study was one-third that
reported by Aulerich and coworkers and death would not be expected to occur in the adult
mink on this trial

Previous studies of the efl‘ects of PCBs on mink by Aulerich 91 a1. (l97l,1973,1986);

Bleavins gt a]. (1980) and Wren (1987) showed that PCBs have not only a detrimental effect
on adult mink survivability, but also on mink reproduction.
Subchronic consumption of PCBs can cause reproductive failure in female mink, although
some females exposed to PCBs ovulate and implant fertilized eggs. Dietary concentrations
as low as 2 ppm Aroclor 1254 impaired mink reproduction when fed for eight months.
Complete fetotoxicity for Aroclors 1242 or 1254 occurred at less than 5 P131“, with 50 percent
lethality of the adult mink being 8.6 and 6.65 ppm respectively (Aulerich :1 al, 1981).
Bleavins 91 a1. (1980) reported that female mink fed Aroclor 1016 had fewer full term
pregnancies and those that whelped, had kits with lower birth weights and lower four week
body weights than the control kits. Kit mortality reaching 80% occurred when female mink
were fed 15% grormd, whole, raw coho sahnon from Lake Michigan containing 15 ppm total
PCBs and no kits survived longer than 24 hours (Aulerich et al, 1973).

In the present study, there were several reproductive parameters that were statistically
significant between one or more treatment groups. Although statistically significant,
differences in parameters which inchrded kit birth weight, kit three-week body weight, and
fit mortality at birth may not be relavent as they fall within a range of values considered to
be"normal" for mink. Results fiom studies by Aulerich and Ringer (1977) and Bleavins at al

(1980) have demonstrated that mink are among the most sensitive species to PCBs.

77

However, in these studies, the dietary PCB concentrations were greater than those in the
current study. Aulerich and Ringer ( 1977) found that reproductive failure occurred in female
mink fid only 2 ppm Aroclor 1254 and at concentrations even as low as 5 ppm, death of the
adults occurred within nine months. Another study by Aulerich and Ringer (1980) showed
that levels as high as 25 ppm Aroclor 1016 fed for 18 months did not adversely affect
reproduction, although growth and survival of the newborn kits was suboptirmrm. The
concentration of PCB (Aroclor 1016) fed to mink in the latter study which did not cause
reproductive impairment may have differed from those concentrations of PCBs (Aroclor
1254) fed to mink which caused reproductive impairment because of differences in absorption
or greater metabolism or excretion rates betwear certain Aroclors or because of the different
concentrations of the various congeners in the Aroclor mixtures. Platonow and Karstad
(1973) and Homshaw g; a]. (1983) conchrded that reproductive impairment can occur in mink
at lower concentrations when the PCBs have been first metabolized by another species and
then fed to mink as in the current study.

It does not appear that the concentrations of PCBs or mercury in this study directly
affected embryo implantation or maintenance of pregnancy since all females fed Diet E (which
had the highest concentrations of both PCBs and mercury) whelped ( Table 6). A
significamly higher nunber of females fed diet C did not whelp compared to the other dietary
groups, but it cannot be concluded that PCBs or mercury were the cause of the impairment
because there were only eight females per dietary treatment. One female was badly bitten by
an aggressive male early in the breeding season and refirsed to mate, a second female was

formd to have a cyst on her ovary at the time of necropsy. A third female had several

78

implantation scars but lost the kits before whelping, while a fourth female in the group had
no implantation scars at necrospy even though she had several confirmed matings. The
number of kits whelped per female is important as it accounts for the proportion of females
which either firiled to inmlant or lost fetuses through early resorption. The number of kits per
female that whelped between the groups in this study was not significantly different. Thus,
it could be concluded that in this study the PCB and mercury concentrations were not
suflicient to cause impairment of ovulation or implantation.

Studies by Wheeler (1971), and Ahamed et a1 (1978) have shown that PCBs can have
an adverse effects on the reproductive performance of males. Male beagle dogs fed 10 or
100 ppm of Aroclor 1254 for two years showed diffuse hyperplasia of interstitial cells of the
testes and aspermatogenesis (Wheeler, 1971). Chickens fed PCBs during the maturation
period showed reduced testis weights (Platonow and Funnel], 1971) and cooks fed a diet
contamining PCBs exhibited decreased sperm volume and concentration (Ahamed 91 a1,
1978). Male quail administered a C14 labelled PCB and killed 24 hours after administration
showed very poor uptake of PCBs into the seminiferous epithelium (Biessman, 1981). This
and other investigations by Berlin 91 a]. (1975) and Brandt ( 1975) showed that there is poor
uptake of PCBs into the testes which may be due in part to poor blood-flow through the
testes. The reproductive performance of the male mink in this experiment was not impaired
by any of the treatments ( Table 7). This agrees with the results of other studies which
danonstrated that reproductive functions of female mink were much more sensitive to PCBs
than of males (Bleavins 91 a1, 1980). Because the males used in this experiment were fed

PCBs afier sexual maturity , the sexual organs may not have been affected by the PCBs in the

79

diet as might be expected if the PCBs were administered during sexual development. As
stated above, the testes have been found to have poor uptake of PCBs, and with the low
concentrations of PCBs in the experimental diets of this study, no detrimental effects on
spermatogenesis would be expected.

Sexual difi‘erences in whole-body clearance and tissue distribution of intestinally-absorbed
methyhnercury have been demonstrated in rats, mice, and humans (Magos 91 a1, 1981).
Female mice seem to retain considerably less mercury than do male mice. Therefore, the male
mice were more susceptible to the toxicity of methylmercury than female mice. Although the
mechanism is unknown, testosterone levels were suggested as a possible modifying factor as
the sex difl‘erence observed in the adult animals (Hirayama and Yasutake, 1985).

The concentration of mercury in the present study did not impair the reproductive
performance of the male mink This agrees with the results of studies by Wren (1987) in
which no effects were observed in the mating performance of male mink fed diets containing
either 1.0 ppm methylmercury, a combination of 1.0 ppm PCB and 1.0 ppm methyhnercury
or 0.5 ppm methyhnercury.

Numerous studies have shown that there is a dose-dependent decrease in body weights
of ofl‘spring prenatally exposed to PCBs (Heaton, 1992). Research conducted in the early
19808 has shown that babies born to women who had eaten Lake Michigan fish containing
PCBs were lighter at birth by 169-190 g and had smaller head circumferences than babies
born to women who had not eaten Lake Michigan fish (Jacobson et a1, 1984). Infants
exposed in m to PCBs by affected mothers in the Japanese Yusho outbreak tended to be

small for their gestational age (Hirayama, 1976). Evidence of reduced birth size in humans

80

is consistent with experimental studies with rhesus monkeys. Ingestion of PCBs by female
rhesus monkeys was associated with lower birth weights among livebom offspring, even for
those conceived more than 12 months after maternal ingestion of PCBs had ceased (Barsotti
91 a1, 1976). Birth weights of mink fits whelped by dams fed Aroclor 1016 averaged
numerically less than those of fits whelped by dams fed a control diet, however, the weights
of the kits were not statistically different (Bleavins, 1980). Lower pup birth weights have
also been reported when adult female rats were fed a diet containing Kanechlor 300, 500 or
Aroclor 1254.

In the current study, fits born to mothers fed Diets D and E had significantly (ps0.05)
higher birth weights than those fits born to mothers fed diets having lower PCB
concentrations ( Table 6). This contrasts with the results of the studies reported above.
However, the average birth weights for kits in all treatment groups in the current study fell
within the normal range for ranch mink (based on MSU Experimental Fur Farm records). This
suggests that the PCB concentrations in the current study caused no detrimental in Men:
effects through placental transfer. It is possible that because, on average, fewer kits were
born to those females fed Diet E, that this could account for the heavier birth weights
observed in the Diet E fits.

The results of this study may support the observation that PCBs are transferred through
the mother‘s milk Although kits fi'om Diet E were the heaviest at birth, by three weeks of
age the kits which were exposed to the higher concentrations of PCBs were lighter than
those kits fiom the other treatment groups. This suggests that possibly milk quality or

quantitiy was adversly affected If PCB loading had been through placental transfer, the fits

81

would have probably been smaller at birth. The decreased growth rate of the Diet E fits
could also be because there were more kits per female for diet E females than for any other
treatment group. If the mothers had poor milk quantity, the greater number of kits would
only conpormd the problem and result in slower growth rates. Aulerich and Ringer (197 7)
noted impaired growth and excessive mortality in fits that were born with normal birth
weights and were nursed by PCB-exposed females.

In the currart study, kits born to mothers fed Diets D and E, which contained the highest
concentrations of PCBs, had the greatest birth weights but gained body weight at a lesser
rate to six weeks of age than did those fits born to mothers fed the lower concentrations of
PCBs Comparing kit body weights from birth to six weeks of age suggests that PCB transfer
by lactation may have played a role in the reduction of body weight gains in fits from Diets
D and E that were exposed to higher concentrations of PCBs during lactation. Those kits
hour to mothers fed Diet A (75% Atlantic mackerel) had statistically higher (p50.05) three
week bodyweights than those kits born to mothers fed Diet B (75% upstream fish) that were
the lightest at three weeks of age ( Table 6). This was the only statistically significant
difl‘erence in kit growth from birth to six weeks of age. The reason for the greater three-
week kit body weights in the Diet A kits is unknown.

Many studies of various species have documented the passage of PCBs from mother to
fetus through transplacental transfer. However, most offspring, including mink, receive a
considerably greater quantity of PCBs via lactation. Thus, milk is a more important source
of PCB contamination to mink kits than placental transfer. Kit body weights may vary as

mich as 10% or more fi'om average depending on the mother’s physical condition, diet, and

82

the genes she inherited for milk production (Leonard, 1966). Since the genetic traits for milk
production cannot be accounted for, it could be that fit growth rates may have been affected
by the did of the females during lactation. The dids contained an Imusually high percentage
of fish (75%), which could have resulted in marginal nutritional deficiencies that, in
combination with the higher concentration of PCBs in Dids D and E could have contributed
to the slower growth rates observed in did D and E fits.

Perinatal mink mortality was studied by Martino and Villar (1990). Of 2122 kits
(standard dark mink) born, 62 were stillborn and 548 of those born alive died within the first
four weeks of life (25.8% mortality). Death resulted from many causes (septicemia,
starvation and hypothermia were the most common). The highest percentage of fits (61.9%)
died within the first week of life, with the highest mortahty attributed to starvation occurring
in litters of nine or more fits (46.8%) or seven to nine fits (26.6%). The number of stillborn
kits was also greater in those litters having seven or more kits. Most stillborn fits had below
average birth weights. Undersized newborn fits may have developmental deficits and be
more susceptible to cold and less able to compete for food.

In the current study, the percentage of fits born alive (77-95%) fell within the normal
range observed by Martino and Villar (1990). Although it is possrble that the higher fit
mortality observed in four of the five treatment groups from birth to three weeks could
suggest that the kits may have obtained higher quantities of PCBs from birth to weaning than
during fetal developement, all kits born alive to mothers fed Did E survived to six weeks of
age. Therefore, it is thought that the concentrations of PCBs in the dids were not high

enough to cause an abnormal increase in fit mortality. The females fed Dids A and C had

83

significantly (ps0.05) higher kit mortality at birth than those females fed the dids containing
the highest concentrations of PCBs and mercury. It should be noted that Did A contained
only one species of fish (Atlantic mackerel) and it is possible that 7 5% of this single
ingredient resulted in an unbalanced did causing negative effects on kit survivability at birth.

The concentrations of mercury in the dids of the adult mink in the current study did not
appear to afl‘ect pregnancy or the development of the fetal central nervous system as all
females fed the did with the highest concentration of mercury whelped and none of the fits
displayed the characteristic "cerebral palsy" type symptoms indicative of mercury poisoning.
The brains of the fits were not taken for histopathological examination or mercury analysis
during necrospy, therefore, no comparisons could be made between treatment groups. Wren
(1987) observ ed no difference in fit survival rates when mink were exposed to either 1 ppm
PCB or 1 ppm mdhylrnercury singly, but when the treatment groups received both chemicals
simultaneously, a synergistic effect was noted resulting in a significant reduction (35.8%) in
fit survivability. The form of mercury plays a significant role in the toxicological result.
When mercury was fed in the form of mercuric chloride to adult mink at 10 ppm for five
months, the number of fits whelped and alive at four weeks of age was comparable to the
control (Aulerich 91 al, 1974).

In the current study, though there were noticeable differences in the survival rates of the
kits, it could not be concluded that this was the result of in_ mm exposure to mercury, as
fits from mothers fed Did E containing the highest mercury concentration (0.22ppm) had the
greatest survival rate when compared to the other treatment groups which contained lower

concentrations of mercury.

84

There were also no detrimental effects on growth and body weights of mink fits
observed in this study ( Table 6). When comparing mercury concentrations in the currdrt
study with previous studies, it does not appear that the concentrations of mercury in the
treatment dids was high enough to cause a noticeable decrease in growth rates in kits
exposedinutem as even thehighest concentration ofmercuryinDidE was lessthanhalfthat
used in the study by Wren. (1987). Although fits born to mothers fed Did E tended to have
a slower growth rate, it is believed that this may have been due to the larger average litter
sizes.

Heaton (1992) observed significant increases in the weights of adult female mink liver,
spleen, lungs, fidneys, thyroid glands and adrenal glands when mink had been fed PCB-
contaminated fish. These results are similar to results reported by Aulerich d 31 (1987) who
observed increased liver, adrenal gland, and fidney weights in female mink fed a did
containing 3,4,5, 3',4',5'-hexachlorobiphenyl (HCB). However, in the latter study, no
increases in the weights of spleen, hmgs or thyroid glands were observed, demonstrating that
mink exhibit a sensitivity and variablility between mdabolized forms of PCB and individual
congeners (Aulerich d aL, 1987). An increase in the weight of the liver has proven to be a
sensitive , but not exclusive, indicator of PCB toxicity in many species. Enlarged livers
observed in mink fed Aroclor 1254 are consistent with results of other studies in which
commercial PCB mixtures were fed to mink (Aulerich and Ringer, 1977), rats, mice (Orberg
and Imdberg, 1974), swine (Hansdr d 31., 1975), monkeys (Allen 91 31., 1974) and rabbits

(Koller and link], 1973 ).

85

In the current study, no differences were noted in organ weights among the treatment
groups for the females (see Table 8). This, along with the fact that there were no differences
in body weights among the females in the various groups, supports the hypothesis that the
concentrations of PCBs in this study were not great enough to cause an increase in organ
weights, decrease in food consumption and/or "wasting syndrome" associated with PCB
intoxication. Although liver weights of males fed Did A were twice those of males fed Did
C, (see Table 8), because there were only two males per did, this difference was not
considered significant due to the high variability in individual liver weights. In the current
study, there were also no differences observed in fit organ weights between the didary
groups.

Unfortunately, very little research has been done on the effects of mercury on organ
weights. However, mercury has been documented to effect organ weights in laboratory
studies with mink Five ppm methylmercury fed to mink caused a significant increase in the
weights of the heart and fidneys when compared to control animals (Aulerich :1 a1, 1974).
In the current study, no significant increases in adult or fit organ weights were
observed When compared to the concentrations of mercury in the study by Aulerich 91 a1
( 1974), the mercury concentrations in the current study were minute and therefore, would
not be expected to increase organ weights.
Numerous studies have documented the harmful effects of PCBs on a varidy of
mammalian species. The symptoms associated with administration of a toxic dose of a
cormnercial PCB mixture (Aroclor 1242) in rats consist of diarrhea, diminished exploratory

behavior, adipsia, anorexia, erythema of limbs, ataxia, coma and death (Kimbrough 91 0.,

86

1978). In mammals, the pathology involves: follicular pyodermatitis (chloracne), liver
atrophy and necrosis, and an increase in the activity of drug mdabolizing enzymes. Mink fed
PCBs showed depressed plasma progesterone concentrations, elevated hepatomicrosomal
cytochrome P450 concentrations, increased benzo (a) pyrene hydroxylase activites and
enhanced cerebral and depressed nridbrain dopamine concentrations (Aulerich :1; a1, 1985).

Symptoms of PCB poisoning in birds consist of tremors, ataxia, milling and loss of
feathers, enlarged livers, edema of subcutaneous tissues and fluid accunnrlation in the
abdominal and thoracic cavities These symptoms characterize the disease designated as chick
edema disease (Kimbrough :1 a1, 1978). The physiopathology of PCB toxicosis in poultry
includes: swelling and hemorrhages of the fidneys, centrilobular liver degeneration,
microsplenia and depressed body growth.

In contrast to acute toxicity of PCBs, which is of lower order when the substances are
administered as a single dose, subacute toxicity is of greater concern. Rhesus monkeys fed
dids containing 25 ppm Aroclor 1248 for two months developed facial edema, alopecia,
acne, anemia, hypoproteinenria, bone marrow atrophy and severe hypertrophic gastritis.
However, no clinical signs of PCB intoxication were observed in any of the females or fits
in this study. Two male mink passed blood in their feces for a short period of time, and one
of these males died shortly afier, but the cause of death was not related to PCB exposure.

Mink fed Aroclor 1254 have shown pathological lesions that included mild splenomegaly,
increased megakaryocytes and gastrointestinal tract hemorrhage and have frequently shown
ascites and pancreatomegaly ( Platonow and Karsted, 1973; Aulerich 91 al, 1985). The

difl‘erences in the pathological effects seen with PCB intoxication may be due in part to the

87

interaction of the various PCB isomers which comprise technical grade Aroclors but which
may not be present in those PCB mixtures formd in the environment (Gillette 91 al, 1987).
PCB induced sfin lesions inchrded hyperplasia and hyperkeratosis of epidermal and follicular
epithelium in adult female New Zealand rabbits. Histopathology of the rabbit livers revealed
centrolobular degeneration, liver cell atrophy, focal necrosis and cytoplasmic hyalin
degeneration PCB-induced kidney lesions inchrded hydropic degeneration of the convoluted
tubules and tubular dilation. In rats administered a single toxic dose of Aroclor 1242, all
organs appeared normal except the liver and fidneys. Histopathology revealed large discrde
sudanophilic vacuoles in hepatocytes and scattered foci of tubular epithelial cells present in
the fidneys (Kimbrough et al, 1978). Kimbrough and coworkers also reported ulceration of
gastric and duodenal mucosa in rats after a single oral dose of Aroclor 1254 or 1260.
Numerous studies have concluded that the primary targd organ of orally administered
PCBs is the liver in mammals (Hansen 91 aL, 1975; Gillette 91 a1, 1987; Heaton, 1992).
Heaton (1992) reported the first mink fit to show teratogenesis that may be due to exposure
to PCBs. The mother of this fit had the highest concentration of PCBs in her liver (10.6
mg/kg) of the adults on trial. Mink fed 20 or 40% Saginaw Bay carp (1.53 and 2.56 ppm
PCB, respectively) had hepatic lipidosis, marked congestion and moderate lymphocytic
infiltration in their livers Rats, mice, mink, rabbits and rhesus monkeys have all demonstrated
liver hypertrophy due to expomre to PCBs. Focal liver necrosis has also been observed.
Sherman rats fed 100 ppm Aroclor 1242 or 100 ppm Aroclor 1016 for six months showed
eviddrce of hepatic lipid accumulation, enlarged liver cells, inclusions in a number of livers,

and hemorrhage and necrosis.

88

The most important hepatic efl‘eds of PCB poisoning included increased weight, fatty
degeneration, hyalin degeneration, and necrosis. Increased liver weights, due to the
proliferation of smooth surfaced membranes of the endoplasmic reticulum were found by
Nishizruni (1970) in mice and monkeys and by Norback and Allen (1970) in rats.

Changes in the liver morphology observed using light microscopy are most commonly
observed in the centrolobular zones. The distribution of the mdabolizing system in the liver,
resulting in a higher concentration of the ultimate toxicant in the centrolobular region,
accormts for the occurrance and frequency of centrolobular toxicity. The centrolobular
hepatocytes are larger, contain more smooth endoplasmic reticulum and have higher
concentrations of cytochrome P450 and associated enzymes that mdabolize and activate
xenobiotics than perilobular hepatocytes. Following exposure to commercial PCB mixtures,
mink exhibited high incidences of centrolobular fatty changes in hepatocytes, hemosiderosis
of Kupfi‘er cells and neutrophil reactions (Bergman 91 al, 1992). Planar PCB congeners
caused centrolobular fatty changes in the livers of mink They are most toxic due to their
strong binding affinity to the cytosolic arormtic hydrocarbon (Ah) receptor protein. No cases
of centrolobular fatty changes were noted when mink were treated with non-planar PCB
congeners, which are considered less toxic because of their lower binding afinity to the Ah
receptor protein. Occurrence of liver changes observed in mink was not due exclusively to
effects caused by congeners regarded to be most toxic, but their frequency and severity were
due to the combined effects of the different fi'actions presart in the commercial PCB mixtures.
The fatty changes observed in the liver due to PCB exposure suggest an excess accunmlation

of triglycerides within the hepatocytes which may result from disturbances in any of the events

89
in the sequence fi'om fatty acid entry to lipoprotein exit.

Hepatocellular carcinomas have been obseerved in rats fed 100 ppm Aroclor 1260 in
the did for 21 months (Kimbrough, 1973) or 500 ppm Aroclor 1254 for six months.
Adenofibrosis, a focal proliferation of glandular epithelium fornring ducts surrounded by
extensive fibrosis, occurred with hepatocellular carcinomas found in rat livers.

Even though numerous studies have documented pathological changes in the organs of
animals exposed to PCBs, no toxic changes were observed in the livers or any of the other
organs colleded at necropsy fiom either the adult mink fed the contaminated dids or their
kits in this study. The low concentrations of PCBs in the current study were probably not
suficient to cause an accurmrlation of lipid within the hepatocytes which leads to the
congestion and inflammatory response commonly observed in mink fed dids containing
PCBs.

Even though numerous studies document severe pathological changes in the central
nervous system of both humans and animals exposed to mercury in 11191.0, no changes were
observed at necropsy or on histopathological examination that would reveal mercury
intoxication in the mink in the present study. It is assumed, therefore, that the concentrations
of mercury in the current dudy were probably not suflicient to cause affects to the central or
periopheral nervous system with in mm exposure. As stated earlier, when comparing the
mercury concentration of the dids in the current study and the respective tissue
concentrations with those studies of other investigators who reported pathological lesions and
death, pathological alterations would not be expected with didary mercury concentrations

of only 0.02 to 0.22 ppm

90

At ldhal exposure levels of PCBs, hematological changes appear to be directly related
to lesions in the bone marrow in all species of animals studied In acute studies (<30 days),
thrombocytopenia and lymphopenia have been reported in several species of lab animals.
Increased erythrocyte counts have been reported by McConnell (1985) in acute studies, but
they may have bedr due to dehydration. In chronic studies, mild to moderate anemias were
the most consistent hematological finding. Some studies showed decreases in leukocytes
while others showed leukocytosis. These difl"erences may be due to the presence of
secondary infections (McConnell, 1985).

Generalized effects on hematological paramders in mink due to PCB intoxication include
increased blood concentration of thyroxine and alanine amirrotransferase (ALAT) and
decreases in progesterone, alkaline phosphatase (ALP), serum bile acids (BA), fi'uctosamine,
and cholesterol (Edqvist 91 al, 1992). Blood serum changes are very complex with PCB
intoxication. Changes in serum conrponents may reflect lesions of the hepatocytes. Since
anatomic pathology varies bdween species, serum chenristry values vary to reflect lesions
(McConnell, 1985 ).

Studies by Edqvist d a]. (1992) revealed significantly different biochemical paramders
in preganant mink fed PCBs when compared to the pregnant control females. The most
fi'equently altered paramders were increases in serrrm ALAT, ALP, BA, and fi'uctosamirre.
The increase in ALAT activ'n:y was due to inflammatory cell reactions in the liver due to
disturbed hepatic cell integrity or hepatic necrosis. An increase in serum ALP activity is often

seen with bile duct obstruction or irrtrahepatic cholestasis. Seasonal changes of serum ALP

91

activity occur in mink with the lowest activity recorded during the winter and increases
occurring during the latter part of gestation reaching a ten-fold increase in the summer
(Edqvist d 3]., 1992). Decreases in serum ALP may be related to mahrourishment or stress
and have been observed in anorexic humans and foxes (Edqvist 91 a1 ., 1992). Most
disturbances of hepatocelhrlar integrity are accompanied by intrahepatic cholestasis (Edqvist
d d, 1992). Rhesus monkeys and mink fed PCBs have displayed decreased serum
cholesterol, which can cause lowered serum BA Elevated serum ghrtamate dehydrogenase
(GLDH) activity has been recorded in mink fed PCBs. This enzyme is considered to be
exclusively located in the mitochondria of hepatic cells. Elevated concentrations of GLDH
are due to necrosis of hepatic cells (Edqvist 91 a1, 1992).

Treatment of animals with difl‘erent chemicals, inchrding PCBs, has been shown to
increase the activities of biotransformation enzymes and has been described for cytochrome
P450-dependent monooxygcnases Serum ALAT, ALP, aspartate aminotransferase (ASAT)
and GLDH are not traditional induction enzymes but represent intracellular enzymes which
through disturbed cell integrity, leak from the cells with an increased activity in peripheral
blood (Edqvist :1 a1, 1992).

Values of all hematologic paramders measured for the treated females fed Dids B and
E in the present study were within the normal range for ranch-bred female mink (Table 10;
Karnedy 1935; Kubin and Mason, 1948; Rotenberg and Jorgenson, 1971). The increase in
the white blood cell (WBC) cormt for both treatments that occurred during the month of
February could be attributed to the increased stress on the animals due to freezing

tenrperatures.

92

Curley d a]. ( 1971) and Weigel and Smith( 1974) found that PCBs with a higher
number of chlorirres per molecule are rdained in tissues for longer periods of time than those
with lower percent chlorination. Goldstein d 31, (197 5) suggested that preferential rdention
of the higher chlorinated PCB congeners might explain the impairment on reproduction in
mink observed with Aroclor 1254 but not with the lesser chlorinated PCBs.

Several researchers (Kimbrough, 1973; Burse 91 a1, 1974; Biessrnan, 1981) have shown
that PCBs tend to accunnrlate in higher concentrations in the liver and adipose tissue of
experimental animals Continued didary exposure to commercial PCB mixtures in mammals
results in their storage in adipose tissue and over an extended period of time, high levels may
be attained (Burse 91 a1, 1974). Rats fed 100 ppm Aroclor 1242 or Aroclor 1016 for six
months had the highest PCB concentrations in the adipose tissue, where steady state was
approached in two months and reached in four months after exposure was discontinued
(Burse e1 :1, 1974). Alter a 10 month feeding period, the concentration of both PCBs was
about the same in the liver of rats that were subjeded to necropsy, but in rats sampled four
months after exposure ceased, ahnost twice as much Aroclor 1016 as Aroclor 1242 was
found in the liver. The levels in the adipose tissue revealed that after two months, no
appreciable increases were observed in the adipose tissue for either PCB in the rat. Following
a six month recovery period, the residue levels in adipose tissue of the rats were 21.8% of
those observed alter the six month exposure period to Aroclor 1242. Following a five month
withdrawal period, the residue concentrations in the adipose tissue were 11.8% of those

observed after the six month exposure period to Aroclor 1016 (Burse 91 al, 1974).

93

Male Sherman rats fed didary concentrations of 500 ppm Aroclor 1254 for six months
had pronounced lipid accunnrlation in the liver which persisted for a 10 month period
following the discontinuation of exposure to PCBs At 10 months, high concentrations of the
higher chlorinated biphenyl isomers were still present in adipose and liver tissue (Kimbrough.,
1973 ). When Sherman rats were allowed to recover for 16 months following didary
exposrre to 100 ppm Aroclor 1254 for six months, a concentration of 4.4 mg/kg of PCB was
still present in the liver and 152 mg/kg in the adipose tissue (Kimbrough 91 a1, 1975).

In the current study , PCB liver and adipose tissue residues were directly proportional
to the didary PCB concentration. Adult mink fed Did B or did E had significantly (ps0.05)
higher concentrations of PCBs in their liver and adipose tissues than mink fed the other
treatment dids.

The concentrations of mercury in the tissues of fits was not analyzed in the current
study. The adult mink consumed from 0.594 mg to 6.53 mg (see Tablel6) total mercury
while on trial and the highest adult mercury tissue concentrations were found in the heart and
ranged from 3.44 ppm to 17.82 ppm wd weight. Because the fits received mercury through
the placarta, it is possible that they could have accunnrlated concentrations higher than those
found in the adults at necropsy.

The adult mink fed Dids D and E had significantly (ps0.05) higher mercury
concentrations in their liver, heart and fidney tissues than mink fed the other treatment dids.
This was to be expected since these mink were fed the two dids containing the highest
concentrations of mercury. However, unlike Wobeser and Swift (1976), Kucera ( 1983) and

Wren 91 a1. (1986) who found higher mercury concentrations in the liver of mustelids

94
conqrared to the kidney, analyses in this study showed higher mercury concentrations in the
fidneys than in the livers of the adult mink in all treatment groups. When comparing the
concentrations of mercury in the dids of mink in the current study with previously discussed
studies, it is clear that even the highest concentration of mercury in Did E (0.22 ppm) which
produced tissue concentrations in the fidney, heart and liver of 4.23, 17.8 amd 3.2 mg/kg,
respectively, was very small when compared to the concentrations of mercury used in the
dids and formd in the tissues by other investigators who observed pathological alterations and

death in mink.

SUMMARY

Environmentally altered PCBs and mercury, as well as other contaminants in fish
collected from the ORR and fed to mink three months prior to breeding did not significantly
impair mink reproduction All females consuming Did E ( 7 5% downstream fish) having the
highest concentrations of both PCBs (1.86 ppm) and mercury (0.22 ppm) whelped. The kits
were born with above average birth weights and had below normal mortality rates and
average growth rates through weaning. No teratogenic efl‘ects were observed. The normal
birth weight of fits is eight to ten grams and females in all treatment groups whelped fits
averaging normal to above normal birth weights. Those fits born to females fed Did A
(Atlantic mackerel) and Did C ( 25% downstream fish) had significantly higher mortality
rates at birth and three weeks of age than kits whelped by dams fed Did E (75% downstream
fish). A possible explanation could be that the dids contained too high a concentration of
a single ingredient (7 5% Atlantic mackerel) and that this in some way afl‘ected fit
survivability.

A statistically significant number of females fed Did C did not whelp, however the
reasons for the reproductive failures were not attributed to mercury or PCB concentrations.
There were no apparent detrimental afl‘ects on implantation or gestation in any of the
treatment groups All females that whelped had within treatment group gestations averaging

44 to 47 days There were also no apparent affects on male reproduction or spermatogenesis,

95

96

as all the males had viable sperm and all but one male had over 90% successful matings.

The mean mercury concentrations were highest in the adult heart and lowest in the liver.
The mercury and total PCB tissue concentrations showed a positive correlation with didary
mercury and PCB concentrations. The highest PCB concentrations were found in the fit of
the adult mink Though concentrations of both PCB and mercury were found in several organ
tissues of the adult mink, no histopathological alterations were observed in any organs.

Neither the adult males nor females displayed the commonly observed "wasting syndrome"
associated with PCB exposure and their body weights difl‘ered only slightly between the start
and termination of the trial

No significant differences were noted in hematological or serum chemistry values for the
females fed Did B or Did E. The increase in white blood cells in both treatment groups
during February was thought to be due to increased stress due to colder winter

temperatures.

FUTURE STUDIES

Future studies which may involve feeding conatminated fish to mink should consider
increasing the number of individuals per treatment group, thus compensating for unexpected
losses or individuals that may refuse to mate. It may also be expedient to feed the fits on
trial to 12 weeks of age in order to ddermine long-term effects on behavior, grth and
development. For example, when mink kits from this study, which were not necrop sied, were
weighed two months after the trial's termination and compared to 1mtreated mink kits, the
treated fits tended to be smaller in body size and lighter in body weight. It was also noted
that these fits tended to be more hyperactive than untreated fits. In the current study, the
fits ate the contaminated feed for approximately two weeks, and because of their small size,
they ingested very little of the feed and were thus exposed to only minute quantities of the
PCBs and mercury through solid feed. If allowed to eat the experimental dids for several
weeks, effects on reproductive maturity or on the developing nervous system that could not
be seen in the current study might be observed which would be expected in nature if the only
food source available was contaminated. Another important consideration for firture studies
would be to reduce the total didary percentage of fish, as some of the detrimental efl‘ects
observed in the current study may have been associated with the abnormally high percentage
of one ingredient (fish) in the did. Also, more than one species of "clean" fish in a control

did may prove to be beneficial as detrimental effects on fit survivability at birth were seen

97

98

in the control group which contained 7 5% Atlantic mackerel.

 

APPENDICES

 

 

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100

APPENDIX B

Calculations for consumption of feed, polychlorinated biphenyls (PCB s) and TCDD-EQ.

 

A. Average daily feed consumption
(g/mink/day) 150 g
B. Cumnmlative feed consumption { }
(s/mink) E (g/mink/dRY) X ((183%)
C. Cummulative PCB dose
(mg/mink) C=B x dietary PCB concentration
D. Daily average PCB dose D== C / # days
(mg PCB/mink/day)

E. Adjusted daily average PCB dose
(mg/kg body weight/ day) E= D / X weight

l. 150g is an average vahle for mink (Leonard, 1965).

101

APPENDIX C

Calculations for conmtion of feed and merfl .

A Average daily feed consumption
(g/mink/day) 150 g
B. Cumnmlative feed consumption { }
(g/mka) 3 (g/mink/day) X (daYS)
C. Cummulative mercury dose
(mg/mink) C=B x dietary mercury concentration
D. Daily average mercury dose D= C / # days
(mg mercury/mink/day)

E. Adjusted daily average mercury dose
(mg/kg body weight/ day) E= D / X weight

1. 150g is an average value for mink (Leonard, 1965).

102

APPENDIX D

IUPAC numbers of PCB congeners found in fish fi'om the Oak Ridge Reservation

Number Congener

77 3,3',4,4‘-Tetrachlorobiphenyl

81 3,4,4',5-Tetrachlorobiphenyl

101 2,2',4,5,5'-Pentachlorobiphenyl
118 2,3',4,4',5-Pentachlorobiphenyl

126 3,3',4,4',5-Pentachlorobiphenyl
128 2,2',3,3',4,4'-Hexachlorobiphenyl
138 2,2',3,'4,4',5'—Hexachlorobipheny1
153 2,2’,4,4',5,5'-Hexachlorobiphenyl
156 2,3,3',4,4',5-Hexachlorobiphenyl
167 2,3',4,4',5,5'-Hexachlorobiphenyl
170 2,2',3,3',4,4',5-Heptachlorobiphenyl
171 2,2',3,3',4,4',6-Heptachlorobiphenyl
180 2,2’,3,4,4',5,5'-Heptachlorobiphenyl
183 2,2',3,4,4',5',6-Heptachlorobiphenyl
189 2,3,3',4,4',5,5'-Heptachlorobiphenyl
194 2,2',3,3',4,4',5,5'-Octachlorobiphenyl
195 2,2',3,3',4,4‘,5,6-Octachlorobiphenyl
198 2,2',3,3',4,5,5',6-Octachlorobiphenyl

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