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EFFECTS OF 3,3’,4,4’-TETRACHLOROBIPHENYL ON REPRODUCTIVE
PERFORMANCE AND GAMETE FERTILIZING ABILITY IN MICE

BY

Jeng-Fang Huang

A THESIS

Submitted to
Michigan State University
in partial fulfillment of the requirements
for the degree of

MASTER OF SCIENCE

Department of Animal Science
1995

Dr. Karen Chou

ABSTRACT

EFFECTS OF 3,3’,4,4’-TETRACHLOROBIPHENYL ON REPRODUCTIVE
PERFORMANCE AND GAMETE FERTILIZING ABILITY IN MICE

By

Jeng-Fang Huang

The objectives of this study were to investigate the reproductive effects of
3,3’,4,4’-tetrachlorobiphenyl (TCB) on mature female mice (F-O) and the effects of
perinatal and continuous TCB exposure on reproductive performance and gamete
fertilizing ability of the offspring (F-l). F-O mice were treated with 0, 3 ppm, and 30
ppm TCB two weeks before breeding. Treatments continued through mating, gestation,
and lactation. F-l mice were given the same diet as their dams throughout this study.

After 2 weeks of treatment, both fecundity and offspring survival in F-O females
decreased in the 30 ppm-treated group. Egg fertilizing ability in vitro in F-l females
decreased in the 3 ppm and 30 ppm groups. This decrease was associated with an
increase in degenerated eggs. All offspring from 3 ppm and 30 ppm-treated F-l females
died within 4 days of birth. Increased testis weights were observed in the 30 ppm-
treated F-l weanlings, but not in pubertal and mature mice. Results of sperm motion
analysis for F-l males were the same among all treatment groups, except for decreased
sperm motility observed in 9-week-old 3O ppm-treated males 2.5 hours after collection.
The percentage of eggs fertilized in vitro by sperm from l9-week-old 3O ppm-treated F -1
males decreased. Increased body weights were observed in offspring of 17-week-old 30
ppm-treated F-l males. Reproductive performance and gamete fertilizing ability were

impaired by TCB exposure.

ACKNOWLEDGMENTS

The completion of this thesis would not have been possible without the help of
many individuals. I am very grateful for those who have helped to design or conduct the
research.

I would like to express my deepest gratitude to Dr. Karen Chou for not only her
advice in scientific thinking, but also her financial support. It has been a privilege to
work with such a distinguished and respected mentor. I greatly appreciate Dr. Thomas
Voice and his group for their help in the feed and tissue analysis. Drs. Robert Cook and
Steven Bursian deserve many thanks for their enthusiastic guidance in interpreting the
data.

Special thanks are due to Amy Barber for her help in taking care of the animals.
I appreciated the learning environment provided by my fellow graduate students Li Chen,
James Kuo, Debbie Powell, Rochelle Inglis, Chang-Yi Lin, Barbara Salem, and Kathy
Li. I am very thankful to my wife Shirley who gave me help whenever I needed it. I
am also indebted to my son Ryan and my parents who taught me how to be tough and
happy in my life.

This study was supported by NIEHS Superfund Grant P42 ESO4911.

TABLE OF CONTENTS

LIST OF TABLES ............................................................................... vi
LIST OF FIGURES .............................................................................. viii
LIST OF ABBREVIATIONS ......................................................................... ix
INTRODUCTION .................................................................................. 1
LITERATURE REVIEW .............................................................................. 6
Physical and Chemical Properties of PCBs ............................................. 6
Absorption7
Distribution ................................................................................. 8
Metabolism ................................................................................. 9
Elimination and Excretion .............................................................. 10
Human Exposures and Clinical Investigations ...................................... 12
General PCB Toxicity in Laboratory Animals ....................................... 14
Wasting Syndrome ................................................................... 14

Dermal Toxicity ...................................................................... 15
Immunotoxicity ....................................................................... 15
Hepatotoxicity ....................................................................... l6

Neurotoxicity ........................................................................ 18

Mutagenic and Carcinogenic Effects ........................................ 19

Reproductive Toxicity ............................................................... 19

Effects on Gonad Weights ............................................ 20

Effects on Steroid Hormones ......................................... 22

Effects on Gamete Production ........................................ 25

Effects on Gamete Fertilizing Ability .............................. 26

Teratogenic and Developmental Toxicity ................................... 27

TCB Toxicity ......................................................................... 29
HYPOTHESIS ......................................................................................... 32
MATERIALS AND METHODS ................................................................... 33
Phase I ..................................................................................... 33
Animals ............................................................................... 33

Chemicals ............................................................................ 33

iv

Feed Preparation ..................................................................... 33

Treatments ........................................................................... 34

Sample Collection .................................................................... 35

Sample Extraction and Analysis ............................................... 35

Phase II ........................................................................................ 36
Animals ................................................................................ 36

Chemicals ............................................................................ 36

Feed Preparation ..................................................................... 36

Treatments ............................................................................. 36

Parameters Measured ................................................................ 37

Feed Intake .................................................................. 37

Body Weight ................................................................. 37

Organ Weight ................................................................ 37

Fecundity, Litter Size, Sex Ratio, and Survival Index .............. 37

In Vitro Fertilization ........................................................ 38

Sperm Concentration and Motion Analysis ............................. 39

Statistical Analysis ................................................................... 40

RESULTS .............................................................................................. 41
Phase I .......................................................................................... 41

Phase II ......................................................................................... 47
DISCUSSION ......................................................................................... 69
Phase I .......................................................................................... 69

Phase II ......................................................................................... 73

Feed Intake and Body Weight ..................................................... 73
Hepatotoxicity ........................................................................ 75

Thymic Atrophy ...................................................................... 76

Female Reproductive Toxicity ..................................................... 77

Male Reproductive Toxicity ........................................................ 81

Sexual Differentiation ............................................................... 86
CONCLUSIONS ...................................................................................... 88
APPENDIX I ........................................................................................... 89
APPENDIX 11 ......................................................................................... 90
BIBLIOGRAPHY ..................................................................................... 91

Table 1.

Table 2.

Table 3.

Table 4.

Table 5.

Table 6.

Table 7.
Table 8.

Table 9.

Table 10.

Table 1 1.

LIST OF TABLES

TCB concentrations in treatment feed ............................................ 43

Concentrations of TCB equivalents in tissues and organs of
non-pregnant and pregnant mice ................................................... 44

Concentrations of TCB equivalents in adipose tissue and organs
of non-pregnant and lactating mice ............................................... 45

Concentrations of TCB equivalents in adipose tissue and organs
of male offspring at 3 weeks of age ............................................ 46

Effects of TCB on feed intake and body weights of F-O female mice ...... 48

Effects of TCB on organ weights and relative organ weights

of F-O females ........................................................................ 49
Effects of TCB on reproductive performance of ED female mice ......... 50
Effects of TCB on body weights of F-l mice .............................. 51

Effects of TCB on body weights, organ weights, and relative organ
weights of F-l females between 5 and 7 weeks of age ........................ 52

Fertilizing ability of eggs from F-l females exposed
to TCB in utero through postnatal development ............................ 54

Effects of TCB on reproductive performance of F-l females
at 7 weeks of age ..................................................................... 55

vi

Table 12.

Table 13.

Table 14.

Table 15.

Table 16.

Table 17.

Table 18.

Table 19.

Table 20.

Table 21.

Table 22.

Effects TCB on body weights, organ weights, and relative organ
weights of F-l males at 3 weeks of age ......................................... 56

Effects TCB on body weights, organ weights, and relative organ
weights of F -1 males at 9 weeks of age ......................................... 57

Effects of TCB on body weights, organ weights, and relative organ
weights of F- 1 males at 19 weeks of age ........................................ 58

Effects of TCB on reproductive performance of F-l males
at 7 weeks of age ..................................................................... 59

Effects of TCB on reproductive performance of F-l males
at 17 weeks of age ................................................................... 61

Effects of TCB on reproductive performance of F-l males
at 60 weeks of age ................................................................... 64

Effects of TCB on body weights of offspring sired by F-l
males at 7 weeks of age ............................................................. 60

Effects of TCB on body weights of offspring sired by F-l
males at 17 weeks of age ........................................................... 62

Effects of TCB on body weights of offspring sired by F-l
males at 60 weeks of age ........................................................... 65

Effects of TCB on fertilization rates in vitro and sperm
motion analysis for F-l males at 9 weeks of age ........................... 66

Effects of TCB on fertilization rates in vitro and sperm
motion analysis for F-l males at 19 weeks of age ......................... 67

vii

LIST OF FIGURES

Figure 1. Motility of sperm taken from 9-week-old F-l males ............................. 68

viii

ALH Dis.

BW
4-CB
6-CB
DHT
HCB
HCG
i.p.
3MC
PB
PCBs
PsCB
PMSG
TCB
TCDD

TEF

VLDL

LIST OF ABBREVIATIONS

amplitude of lateral head displacement
body weight

2 ,2 ’ ,4 ,4 ’ -tetrachlorobiphenyl

2 ,2’ ,4 ,4’ ,5 ,5 ’ -hexachlorobiphenyl
dihydrotestosterone

3,3’,4,4’,5 ,5 ’-hexachlorobiphenyl
human chorionic gonadotropin
intraperitoneal

3-methylcholanthrene

phenobarbital

polychlorinated biphenyls

3 ,3 ’ ,4,4’ ,S-pentachlorobiphenyl
pregnant mare’s serum gonadotropin
3,3’,4,4’-tetrachlorobiphenyl
2,3,7,8-tetrachlorodibenzo—p-dioxin
toxic equivalent factor

toxic equivalent

very low density lipoprotein

ix

INTRODUCTION

Polychlorinated biphenyls (PCBs) are a family of congeners with different
numbers of chlorines binding covalently to biphenyl rings. Chlorination of biphenyl
rings can produce 209 possible chlorobiphenyls, each substituted with 1-10 chlorine
atoms (Sawhney, 1986). This group of chemicals was first synthesized by Schmidt and
Schulz in 1881 and was commercially produced beginning in 1929 (Tanabe, 1988).
PCBs have been frequently used in industry as heat transfer fluids, hydraulic fluids,
lubricants, transformers, capacitors, plasticizers, and petroleum additives due to their
chemical and physical stability and versatility (Richardson and Waid, 1982; Sawhney,
1986; Safe et al., 1987). Polychlorinated biphenyls are resistant to acids and bases,
compatible with organic materials, resistant to oxidation/reduction, nonflammable,
nonconducting, and heat-resistant. However, improper disposal practices and accidental
leakages from industrial facilities have led to environmental contamination (Tanabe,
1988; Safe, 1994). In addition to the flow of ground and surface water, the dynamic
activity of the atmosphere accounts for much movement in the ecosphcrc (Atlas et a1. ,
1986; Hansen, 1987). PCBs have, therefore, been detected in the atmosphere of polar
areas (Atlas et al., 1986).

Exposure to PCBs has been linked to public health risks on numerous occasions.

2

Accidental human poisoning from PCB contaminated rice oil occurred in Japan in 1968,
exposing over 1000 individuals (Rogan, 1982; Kashimoto and Miyata, 1986). Over 20
deaths were reported among these individuals within 5 years of contamination. A similar
poisoning outbreak occurred in Taiwan in 1979 (Chen et al. , 1981; Rogan et a1. , 1988).
Within 4 years of the accident, 24 deaths were reported among 2060 poisoned patients
(Chen and Hsu, 1989). Approximately half of the deaths were from hepatoma, liver
cirrhosis, or other liver diseases accompanied by hepatomegaly (Chen and Hsu, 1989).
Other clinical symptoms observed included mucocutaneous pigmentation, acneform
eruptions, increased eye discharge, swelling of the upper eyelids, abnormal menstrual
cycles, dark-brown pigmented babies, and stillbirths (Hirayama, 1976; Kikuchi and
Masuda, 1976; Chen and Hsu, 1989).

Toxic responses to PCBs in laboratory animals have been extensively reviewed
(Fishbein, 1974; Hansen, 1987; Parkinson and Safe, 1987). Hepatomegaly is a common
syndrome observed in PCB-treated animals (Villeneuve and Grant, 1971; Sanders et al. ,
1974; Clarke et al., 1984). Other types of liver damage include centralobular
degeneration, centralobular liver cell atrophy, focal necrosis, cytoplasmic vacuolization,
and bile-duct proliferation (Curley et al. , 1973; Burse et a1. , 1974). In addition, wasting
syndrome (Garthoff er a1. , 1977), hepatic and gastric intestinal hyperplasia (Ward, 1985),
lymphoid involution (McConnell, 1985), porphyria (Safe, 1984), skin disorders (Barsotti
et al., 1976), carcinogenesis (Hansen, 1987), teratogenesis (Simmons et al., 1984), and
endocrine and reproductive dysfunction (Allen et al. , 1979; Golub et at. , 1991) have all

been observed.

3
Reproductive disturbances caused by PCB exposure have been identified in many

species. Female dysfunctions include increased and reduced germ cell numbers in mice
(Ronnback, 1991; Ronnback and de Rooij, 1994), low frequencies of implanted embryos
and embryotoxicity in mice (Orberg and Kihlstrom, 1973; d’Argy et al., 1987),
decreased litter sizes in rats (Brezner et al., 1984), and prolonged estrous cycles in both
mice and rats (Orberg and Kihlstrom, 1973; Brezner et al., 1984). High incidences of
abortion were also evident in Aroclor 1248 exposed non-human primates (Barsotti et al. ,
1976). Linzey (1988) reported reduced ovary, uterus, and body weights in F-l mice
exposed to Aroclor 1254 in utero and through lactation. Male reproductive abnormalities
have also been reported. Increased relative testis weights (mg testis/100 g bw) have been
observed in mice exposed to 2,2’,4,4’,5,5’-hexachlorobiphenyl (6—CB) perinatally
(Johansson, 1987). Sanders et a1. (1977) reported reduced testicular spermatozoa
concentrations in Aroclor 1254-treated mice. Decreased number and percentage of
normal fertilized eggs and eggs at the 2-cell, 4-cell, and blastocyst stages were observed
in non-treated females mated to males exposed to Aroclor 1254 through milk on days 1,
3, 5, 7, and 9 postpartum (Sager et al., 1991). Altered hepatic microsomal testosterone
hydroxylase activities were observed in adult male and female rats which were
administered Aroclor 1254 postnatally (Haake-McMillan and Safe, 1991).

PCB congeners vary in toxicity depending on their chemical structure.
Quantitative structure-activity relationships (QSARs) have been investigated both in vivo
and in vitro (Leece et al., 1985). In vivo QSARs have been determined by ED50 values

for cytochrome P-448-dependent monooxygenases induction, body weight loss, as well

4

as thymic atrophy in immature rats. In vitra QSARs have been developed from their
potencies as inducers of cytochrome P-448-dependent monooxygenases and relative Ah
receptor binding affinities in rat hepatoma H-4-II-E cells. A high rank order correlation
between in viva QSARs data and the in vitra QSARs has been observed (Leece et al.,
1985).

The three coplanar congeners, 3,3’,4,4’-tetrachlorobiphenyl (TCB), 3,3’,4,4’,5-
pentachlorobiphenyl (P5CB), and 3,3’,4,4’,5,5’-hexachlorobiphenyl (HCB), are
approximate isostereomers of the highly toxic 2,3,7,8-tetrachlorodibenzo-p—dioxin
(TCDD), and are capable of eliciting similar toxicity (Safe, 1984). Based on the hepatic
microsomal monooxygenases inducing potency or Ah receptor binding affinity, TCB,
P5CB, and HCB are classified as Group I, where other PCB congeners are classified as
Group II. Group I congeners are more potent than Group II congeners as inducers of
cytochrome P-448-dependent monooxygenases activity in cultured rat hepatoma H-4-II-E
cells (Sawyer and Safe, 1982). On the basis of Ah receptor binding affinity, Group I
congeners are more competitive than Group II congeners in displacing TCDD from rat
liver Ah receptors (Bandiera et al., 1982; Parkinson and Safe, 1987). Many of the
adverse effects elicited by PCBs appear to be dependent on the presence of Ah receptors
in target organs (Poland and Knutson, 1982; Nebert, 1989). The Ah receptor-ligand
complex regulates the transcription of two P4501 genes, cyplal and Cypl a2 (Nebert,
1989). The receptor appears to have a recognition site to which PCB isosteric
compounds with a configuration similar to TCDD can bind (Poland and knutson, 1982).

High reproductive and developmental toxicities of these approximate isostereomers of

5
TCDD-TCB, PSCB, and HCB are attributed, in part, to the presence of Ah-receptor in

embryonic tissue and CypIaI mRNA in ovarian cells (Dencker and Pratt, 1981; Safe,

1984; Nebert, 1989).

LITERATURE REVIEW

BMW

Polychlorinated biphenyls (PCBs) have been identified in diverse environmental
matrices worldwide. PCBs are produced by the chlorination of biphenyl rings and their
names reflect the percentage of chlorine content. For example, Aroclor 1221, a
commercial product produced by Monsanto Chemical Company (U.S.A.), contains 21%
chlorine by weight. Similar commercial PCB mixtures have been produced by other
manufacturers. These include the Clophens (Bayer, Germany), Phenoclors and Pyralenes
(Prodelec, France), Fenclors (Caffaro, Italy), Kanechlors (Kanegafuchi, Japan), Soval
(Soval, U.S.S.R.), and Delor (Chemko, Czechoslovakia) (Safe, 1984; Safe, 1994).
Because of their chemical stability, heat resistance, and miscibility with organic
compounds, PCBs have been used as organic diluents, plasticizers, pesticide extenders,
adhesives, dust-reducing agents, cutting oils, flame retardants, heat transfer fluids,
dielectric fluids for transformers and capacitors, hydraulic lubricants, and sealants. Some
of the uses of PCBs have resulted in their direct introduction into the environment,
however, a significant portion of the environmental burden of these compounds has
resulted from careless disposal practices, accidents, leakages, and mishandling at
chemical waste disposal sites (Safe, 1994).

PCB congeners with zero or one artha chlorine, two para chlorines, and at least

6

7

two meta chlorines can assume a coplanar conformation sterically similar to that of
TCDD, therefore their toxicity is similar to TCDD. In contrast, PCB congeners with
two or more artha chlorines cannot assume a coplanar conformation and do not resemble
TCDD in toxicity (Poland and Knutson, 1982; Safe, 1990). For those non-artha and
mono-artha PCBs, dose dependent and structure dependent correlations were observed
between aryl hydrocarbon hydroxylase (AHH) induction and several toxic responses such
as thymic atrophy, reduced body weight gain, and teratogenicity in laboratory animals
(Safe, 1990).

Alumnae

Mammals can take up PCBs by three mechanisms (Shaw and Connell, 1990).
They are:

1. absorption of PCBs from the atmosphere through the lungs.

2. absorption of PCBs from the atmosphere through the epidermis.

3. oral intake.

The atmospheric route through the lungs and epidermis is of little significance due
to the low concentrations of PCBs in the atmosphere. Gastrointestinal absorption of
commercial PCB mixtures and individual congeners has been investigated extensively in
laboratory animals (Matthews and Anderson, 1975; Gage and Holm, 1976; Tanabe et al.,
1981). A complex mixture of Kanechlors 300, 400, 500, and 600 (1:1:1:1 by weight)
in corn oil was administered orally to immature Wistar rats. The rats were dosed with
3 mg daily for 5 days. During this 5-day period, more than 84% of the total dose was

absorbed from the gastrointestinal tract. Gas chromatograph-mass spectrometer analysis

8

of the PCBs in the fecal material demonstrated that the absorption efficiency of individual
PCBs varied from 6696-9696 (Tanabe et al., 1981). The major structural determinants
which govern absorption efficiencies in the intestinal wall are molecular size and
lipophilic properties. It is generally expected that more small and highly lipophilic
molecules will be more easily absorbed through the cell wall than large lipophobic
molecules. However, the absorption efficiency of PCBs decreased as the number of
chlorine atoms increased (Tanabe et al., 1981). This suggests that the differences in
lipophilic properties among individual chlorobiphenyls plays a minor role in absorption.
12' I 'l |°

Exposure of target tissues is, to a great extent, closely related to chemical
solubility and lipophilicity. Water solubility tends to decrease with increasing degrees
of chlorination, as does the extent and rate of biotransformation and elimination (Phillips,
1986). Thus, highly chlorinated congeners tend to be more persistent and have greater
potential for bioaccumulation (Takagi et al., 1986; Morrissey and Schwetz, 1989).

Once absorbed, PCBs are rapidly transported to all tissues by blood. The initial
distribution of PCBs to tissues is determined by biophysical factors such as tissue
volume, lipid content, tissue/blood partition ratios, binding to proteins, and perfusion rate
(Maliwal and Guthrie, 1982; Matthews and Dedrick, 1984; Safe, 1989). Liver and
muscle are the primary early depots for several reasons. The liver is highly perfused and
has an affinity for these compounds; while muscle has the largest volume of all body
tissues (Matthews and Anderson, 1975; Matthews and Dedrick, 1984). However, since

PCBs are highly lipid-soluble compounds, they have a high affinity for lipid-rich tissues,

9

particularly adipose tissue. Simultaneously with initial distribution to the liver and
muscles, redistribution to adipose tissue also occurs. These compounds are eventually
concentrated in adipose tissue and skin with a dynamic equilibrium developing between
all body tissues to maintain this concentration (Burse et al., 1974; Matthews and
Anderson, 1975; Lutz at al., 1977). At equilibrium, a change in the PCB concentration
of any tissue will result in a corresponding change in all tissues.
Mmhnlism

Metabolism and excretion are assumed to be important modes of detoxification
of PCBs. Biotransformation of PCBs to hydroxylated metabolites by the microsomal
cytochrome P-450 system is the critical event that determines the biological half-lives of
these widespread environmental contaminants. The factors that determine the rate of
biotransformation of PCBs are: the number of chlorines on the biphenyl nucleus, the
position Of these chlorines, and the animal species, age, and sex (Sipes and Schnellmann,
1987).

When gender and age were compared, enzyme induction was more pronounced
in male rats than in females and increased with age (Chen and DuBois, 1973;
McConnell, 1985). PCBs are generally metabolized to hydroxylated products either by
direct oxygen insertion or via arene oxide intermediates into phenolic metabolites (Safe,
1984; Yoshimura et al., 1987; Safe, 1989; Safe 1994). Two adjacent unsubstituted
carbon atoms are important to the metabolism of PCBs because their presence facilitates
the formation of arene oxides (Matthews and Tuey, 1980). Dechlorination can occur

during further metabolism of the arene oxides (Hansen, 1987). The reactive arene-oxides

10

formed during biotransformation can bind covalently to tissue macromolecules or
conjugate with glutathione. Therefore, the more readily metabolized congeners may pose
a greater threat of toxicity than those that are more persistent, but slowly metabolized
(Safe, 1984; Lutz and Dedrick, 1987).

1 Ex 1 n

In this review, excretion refers to removal of PCBs and their metabolites via feces
and urine. Elimination, on the other hand, is the process through which PCBs are
removed through gestation, parturition, and lactation. PCBs are passed from the mother
to the fetus and carried in the offspring (Matthews and Dedrick, 1984).

PCBs are excreted to an appreciable extent only after they are metabolized and
conjugated to form polar molecules. Derivatives of glutathione conjugates are major
excretion products of PCBs as are glucuronides of the hydroxylated products (Matthews
and Dedrick, 1984; Sipes and Schnellmann, 1987). They can also be metabolized by
conjugation with cysteine and methionine, and then excreted in bile. Intestinal flora
metabolize PCBs to stable methyl sulfides and methyl sulphones, which are mostly
reabsorbed (Hansen, 1987). Methyl sulfide and methyl sulphone metabolites of PCBs
are emerging as important contributors to food chain toxicities (Brandt et al., 1982;
Darnerud et al., 1986; Hansen, 1987).

The majority of elimination is carried out through substances with high volume
and lipid content, e. g. milk. Elimination by these routes is of great concern due to the
possible threat to offspring (Matthews and Dedrick, 1984). Lower concentrations of

PCBs are detected in umbilicus tissue, umbilicus blood and amniotic fluid than in

11
maternal blood. This is evidence of blood-placental barriers (Safe, 1989). In addition,

the fetus is relatively lean compared with maternal tissues. Therefore, PCBs are not
concentrated in fetal tissues.

The partition of PCBs is associated with physiological status. The creation of a
new lipid depot with the onset of lactation results in passive movement of PCBs from
blood to milk and a corresponding movement of PCBs from all tissues to blood in order
to maintain their respective tissue/blood ratios (Matthews and Dedrick, 1984). Due to
the placental barriers to PCBs and lipophilic properties of PCBs, transplacental exposure
to PCBs is considerably less than exposure of the offspring through lactation (Takagi et
al., 1976; Bleavins et al., 1984; Takagi et al., 1986). Radioactivity was measured in
mice treated with one dose ‘of 1‘C-6-CB (50 mg/kg bw) two weeks before mating
(Gallenberg and Vodicnik, 1987). Almost all of the maternal body burden on day 19 of
gestation was transferred to the offspring after 20 days of lactation (Gallenberg and
Vodicnik, 1987). In addition, pregnant rats were treated orally with 1.33 mg 1‘C-
Kanechlor—400/kg bw once a week from day 8 of gestation to weaning. The
concentrations of PCBs in the suckling rats increased gradually for 10 days after birth
and then decreased gradually (Takagi et al., 1976). The tissues measured included fat,
muscle, bone, adrenal gland, blood, spleen, kidney, heart, and brain. The concentrations
of PCBs in these tissues of suckling rats from day 1 through day 45 postpartum were
higher than those in nursing dams sacrificed on day 23 of lactation.

When young adult rats were injected intravenously with a single dose of l‘C-TCB

at 0.6 mg/kg bw (Abdel-Hamid et al., 1981), cumulative excretion of radioactivity in

12

feces and urine reached 50% of the dose 1.5 days after administration (Abdel-Hamid et
al., 1981). Tissues samples were analyzed from 1 hour to 7 days after administration.
TCB metabolites persisted in the blood over the 7-day period because concentrations
were higher in this tissue than in liver, muscle, skin, and adipose tissue. Most of the
radioactivity in rat tissues was parent TCB. Similar results were observed by Mizutani
et al. (1977). The half-life of TCB in the carcass was 0.9 days when mature mice were
fed 10 ppm TCB ad libitum for 20 days, and then given control feed. The TCB residue

in the carcasses was analyzed daily for 4 days after TCB exposure.

 

Widespread environmental PCB contamination was first brought to the attention
of the scientific community in the 1960s when PCBs were detected in fish, fish-spawn,
eagles, and humans in Sweden (Anon, 1966). In 1968, accidental PCB contamination
of rice oil occurred in Japan resulting in the poisoning of over 1000 individuals, which
has been referred to as Yosho incident (Kashimoto and Miyata, 1986). Chloracne is a
refractory follicular dermatitis characteristic of PCB poisoning in humans. This was
observed in some Yosho patients 3 years after exposure to PCB-contaminated rice oil
(Fishbein, 1974). A similar poisoning outbreak (Yu-Cheng) occurred in Taiwan in 1979
(Chen et al. , 1981). Yu-Cheng syndrome was characterized by acne-form skin eruptions,
dark brown pigmentation of skin and nails, and hypersecretion of the Meibomian glands.
Children born between 1978 and 1985 to Yu-Cheng mothers who were exposed to PCB-
contaminated rice oil in 1978 and 1979 were investigated. The heights of Yu-Cheng

children were 3.1 cm smaller than those in the control group. These children had less

13

total lean mass and soft tissue mass as compared to the matched controls. In addition,
those musculoskeletal changes were only seen in the first born Yu-Cheng children after
maternal ingestion of the contaminated rice oil, but not in subsequent siblings (Guo et
al., 1994).

Due to concern about the impactof industrial environmental contamination on
nursing infants, PCBs concentration in human milk from the general population in
industrialized regions have been compared with concentration in non-industrialized
regions. In industrialized countries, PCB concentrations in human milk were 120-760
ppb, while those in the non-industrialized countries were 15-58 ppb (Schecter et al.,
1994). Similar results in New Zealand were reported by Bates et al. (1994) who
compared PCB concentrations in human milk from urban and rural areas. Age was also
associated with increasing concentrations of PCBs in milk (Bates et al., 1994). The net
absorption of PCBs in nursing infants was measured over 12 days. Over 95 % of the
PCBs in the milk was absorbed through the digestive tract (McLachlan, 1993). Some
PCB congeners were also detected in both follicular and sperm fluids (Schlebusch et al.,
1989).

All of the PCB congeners have been assigned a toxic equivalent factor (TEF)
value, which is the fractional toxicity of the congener relative to a standard toxin,
TCDD. Toxic equivalents (TEq) are the sum of the concentrations of PCB congeners
multiplied by their respective TEFs. Coplanar and mono-artha-substituted congeners are
the major sources of TEq in the blood of US. adults (Schecter et al., 1994). Although

coplanar PCBs are more toxic than other congeners, mono-artha-substituted congeners

14

contribute more TEq than coplanar congeners (Schecter et al. , 1994). The authors
suggested that in the general population of the United States, PCBs contribute more of
the TEq in human tissue than do dioxins and furans. This is probably the case in other
industrialized countries as well. There is increasing empirical evidence that human
exposure to low levels of PCBs in utera and through nursing may result in subtle
behavioral and developmental deficits (Jacobson et al., 1983). Occupational

neurobehavioral dysfunction was observed in firemen exposed to PCBs in transformer

fires (Kilburn et al., 1989).

 

The adverse effects of PCB exposure in laboratory animals have been widely
reviewed (Fishbein, 1974; Hansen, 1987; Parkinson and Safe, 1987). They are species,
age, and sex dependent.

Wasting Syndrome

‘Wasting syndrome refers to progressive weight loss which is not totally related
to decreased feed consumption. It appears to be due primarily to decreased feed
efficiency, not consumption (Garthoff et al. , 1977). At lethal doses, the primary clinical
sign in most species is a progressive loss of body weight followed by weakness,
debilitation, and finally death. A steady loss of body weight was observed within the
first 2 weeks of treatment when marmoset monkeys were orally dosed with 3 mg TCB/kg
bw twice a week. Over a period of 18 weeks, body weight of the animals decreased by

34% (van den Verg et al., 1988).

15
Dermal Toxicity

Nonhuman primates exhibit dermal lesions after exposure to low levels of PCBs.
Alopecia, subcutaneous edema, and acne were observed in rhesus monkeys exposed to
2.5 ppm Aroclor 1248 for 6 months (Barsotti et al., 1976). Atrophy and cystic dilation
were observed in marmoset monkeys orally treated with 1 mg TCB/kg bw or 3 mg
TCB/kg bw twice/week for 18-23 weeks (van den Berg et al., 1988). Hyperplasia and
hyperkeratosis of the epidermal and follicular epithelium were observed in rabbits after
38 days of dermal exposure to Phenoclor DP6, Clophen A 60 or Aroclor 1260 with a
total of 27 applications of 118 mg/day (5 applications/week) (Vos and Beems, 1971).

Immunotoxicity

Thymic atrophy is a consistent result of TCDD and PCB exposure in all animals
so far examined (Vos and Beems, 1973; Faith and Moore, 1977; Clarke et al., 1984; van
den Berg et al., 1988; McConnell, 1989; Smialowicz et al., 1989). The reduction in
size is microscopically observed almost entirely as a loss of cortical lymphocytes
(McConnell et al., 1978; McConnell and Moore, 1979; van den Berg et al., 1988). In
mice receiving a single oral dose of 340 pg TCDD/kg bw, the thymus weight after 30
days was only 0.1 of the thymus weight of control animals (McConnell et al., 1978).
This decrease in thymus weight is often accompanied by necrotic debris in the medulla
(McConnell, 1985). Rats were treated with 5 pg TCDD/kg bw either on days 0, 7, and
14 postnatally or on day 18 of gestation and days 0, 7, and 14 postnatally. Thymus/body
weight ratios were found to be suppressed up to 39 days of age. in those treated only

postnatally. In those exposed to TCDD during gestation and postnatally, the thymic

16
effect was observed up to 145 days of age (Faith and Moore, 1977). Increased mortality

was observed in young mallards co-exposed to PCBs and duck hepatitis virus (Friend and
Trainer, 1970). Mortality increased markedly over the control value when ducklings

were exposed to duck hepatitis virus after 10 days of 25, 50, or 100 ppm Aroclor 1254

treatment.
Hepatotoxicity

Liver enlargement was observed in PCB-exposed rats (Curley et al., 1973;
Garthoff et al., 1977; Clarke et al., 1984), mice (Sanders et al., 1974), rabbits
(Villeneuve and Grant, 1971), and mink (Aulerich et al., 1985). Total liver lipid
content, including cholesterol, phospholipid, and neutral lipid, increased in rats treated
with 50 ppm or 500 ppm Aroclor 1254 for 3 weeks (Garthoff et al., 1977). Total
protein also increased with the 500 ppm treatment, while the relative total protein (mg/g
wet weight) decreased with the same treatment. The reduced form of nicotinamide
adenine dinucleotide phosphate (NADPH) is necessary for biosynthetic reactions, e. g.
cholesterol, lipid, and some amino acid synthesis (Voet and Voet, 1990). The
administration of 200 ppm Aroclor 1254 to rats for 3 and 14 days increased the activity
of the major NADPH-generating enzymes: malic enzyme, glucose-6-phosphate
dehydrogenase, and 6-phosphogluconate dehydrogenase. The activity of these enzymes
in the liver increased approximately 2-fold above control activity (Hitomi et al. , 1993).
PCB administration also causes a marked increase in serum cholesterol through the
stimulation of hepatic cholesterogenesis (Kato and Yoshida, 1980). The rate-limiting

enzyme of cholesterol synthesis, 3-hydroxy-3-methylglutaryl coenzyme A reductase,

17

increased in rats fed 1000 ppm Aroclor 1248. Light microscopic examination of the
livers of rats fed Aroclors 1242 and 1016 for 6 months showed enlarged hepatocytes with
vacuolated cytoplasm and inclusions in the cytoplasm (Burse et al., 1974). Bile-duct
proliferation surrounded by fibrosis was also observed in 21-day—old weanling rats from
dams given 50 mg Aroclor 1254/kg bw/day orally on days 7-15 of gestation (Curley et
al., 1973). Administration of 118 mg of Phenoclor DP6, Clophen A 60, or Aroclor
1260 (5 times per week, for 38 days) caused centrolobular degeneration, centrolobular
liver cell atrophy, focal necrosis, and cytoplasmic hyalin degeneration in rabbits (V 08 and
Beems, 1971).

The cytochrome P-450 system is a group of membrane-bound enzymes located
in the endoplasmic reticulum. Three coplanar congeners, TCB, P,CB and HCB, bind
to the Ah-receptor, resulting in induction of cytochrome P-448-dependent hepatic
microsomal monooxygenases (MC type). This binding elicits TCDD-like toxic effects
(Lucier and McDaniel, 1979; Poland and Knutson, 1982; Nebert, 1989). 2,2’,4,4’-
Tetrachlorobiphenyl (4-CB) and 6-CB, not ligands of the Ah-receptor, induce cytochrome
P-450-dependent hepatic microsomal monooxygenases (PB type). This group of
congeners has generally been proven to be less toxic than the coplanar PCB congeners
(Lucier and McDaniel, 1979; Poland and Knutson, 1982; Parkinson et al., 1983; Neal,
1985). Most congeners tested as well as Aroclor 1254 which is a commercial mixture,
appear to be a combination of the MC and PB types: in that they induce both
cytochromes P-448 and P-450 (Parkinson and Safe, 1981; Parkinson et al., 1983).

Porphyria is a hepatic disorder distinguished by altered porphyrin metabolism

18

(Safe, 1984). The porphyrin biosynthesis involves several enzyme-catalyzed steps in
which 6—aminolevulinic acid is ultimately converted into heme, an integral component of
. several enzymes, including the cytochrome P-450-dependent monooxygenases (V oet and
Voet, 1990). The mechanisms leading to hepatic porphyria appear to be induction of 6-
aminolevulinic acid synthetase (5-ALAS)(Safe, 1984) and inhibition of uroporphyrinogen
decarboxylase (Neal, 1985). Uroporphyrinogen decarboxylase is responsible for the
stepwise decarboxylation of uroporphyrinogen to caproporphyrinogen (Voet and Voet,.
1990). Together, induction of 6—ALAS and inhibition of uroporphyrinogen decarboxylase
lead to increased hepatic and urinary porphyrins characterized by increased carboxyl
groups (Safe, 1984; Neal, 1985).
Neurotoxicity

Prenatal exposure to TCB can also influence neurobehavioral functioning of mice
during adulthood (Tilson et al., 1979). Catecholamine levels have been shown to be
altered in various brain regions of mink, doves, and rats exposed to Aroclor 1254 as well
as to some individual congeners (Hansen, 1987). Decreased striatal dopamine
concentration and receptor binding sites in the caudate nucleus were observed in mice at
1 year of age when those mice were treated with TCB in utera (Agrawal et al., 1981).
Impaired development of cliff avoidance reflexive behavior, swimming ability, and open
field activity were found in offspring of PCB-exposed female rats when compared
controls (Pantaleoni et al. , 1988).

Feed containing PCBs was fed to adult female mink for 1 month prior to breeding

and through parturition (Aulerich er al. , 1985). Elevated hypothalamus concentrations

19

of both norepinephrine and dopamine were observed in the offspring, which could affect
neuroendocrine control of reproduction.
Mutagenic and Carcinogenic Effects

PCBs are generally considered to be promoters rather than initiators (Hansen,
1987). The differential carcinogenicity of PCB congeners is dependent on a number of
factors including the sex, strain, and species of the test animals (Safe, 1984).
Experimental data show that commercial PCBs cause liver damage which leads to
preneoplastic changes, and hepatocellular carcinomas. The results also indicate that these
lesions are only observed after long-term exposures to relatively high doses of the
chemicals.

Reproductive Toxicity

Reproductive dysfunction as a result of PCB intoxication was first reported in
women exposed to PCBs in 1968 in Japan. Dysmenorrhea and irregularities of the
menstrual cycle were noted (Hirayama, 1976). Reproductive toxicity of PCBs received
extensive experimental examination when early kit mortality was observed in domestic
mink that were exposed to PCBs (Ringer et al., 1972). Later studies revealed a
correlation between early mink kit mortality and feed containing Great Lakes fish. In
the past 20 years, PCB reproductive toxicity has been investigated in a large number of
studies on humans, laboratory animals, and wildlife. Postnatal weight gain,
neurobehavioral changes, and skin hyperpigmentation are the most sensitive end points
of reproductive toxicity. A lowest observable adverse effect level (LOAEL) of 0.25 mg

Aroclor 1254/kg bw/day was identified for rodents based on postnatal weight gain and

20
behavioral changes (Overmann et al. , 1987; Golub et al. , 1991).

Other toxic effects of PCBs on reproduction are changes in gonad weights, steroid
hormone concentrations, and gamete production and fertilizing ability. These adverse
effects are reviewed in more detail below.

(I ei h

PCBs have been reported to affect the weights of female and male reproductive
organs. Increases in testis weights occurred in the 10-12-week-old offspring of mice
gavaged with 20 mg 6—CB/kg bw every 3 days from day 13 of gestation through day 24
postpartum. This increase was not observed in mice gavaged with 40 mg 6-CB/kg bw
or 160 mg Clophen A 50/kg bw every 2 days for 5 weeks, beginning at 5 weeks of age
(Johansson, 1987). Nor were changes in testis weights observed when adult rats were
treated with 500 ppm Aroclor 1254 in feed for 3 weeks (Garthoff et al., 1977) or when
adult mice were treated with 200 or 1000 ppm Aroclor 1254 in feed for 2 weeks (Sanders
et al., 1977). A decrease in testicular weight was noted in cockerels treated with 500
ppm Aroclor 1254 in the feed for 9 weeks from day of hatching (Platonow and Funnell,
1971). Decreased seminal vesicle weights were observed in white-footed mice treated
with 200 ppm Aroclor 1254 in the feed for 4 weeks (Merson and Kirkpatrick, 1975).

Based on the observed changes in testis weights, animals exposed to PCBs
perinatally are more susceptible to toxic effects than those animals treated during and
after puberty. This may be related to the hormonal imbalance of the prenatal period
when organogenesis occurs, thus impairing the development of reproductive organs.

Gellert (197 8) investigated the uterotrophic activity of Aroclors 1221 , 1242, 1254,

21

and 1260. Significant uterine growth 24 hours after treatment in weanling rats was only
observed with a single subcutaneous injection of 1 mg Aroclor 1221/kg bw. This was
not observed when rats were treated with Aroclors 1242, 1254, and 1260. Increased
uterus weights were also observed with a single intraperitoneal (i.p.) injection of 8
mg/rat of Aroclors 1221, 1232, or 1248. In agreement with Gellert (1978), the highly
chlorinated Aroclors 1254 and 1260 did not affect uterine weights (Beobichon and
MacKenzie; 1974). The increase in uterine weights appears to be associated with
increased uterine glycogen content (Bitman and Cecil, 1970). Glycogen content in the
uterus, an indicator of estrogenic activity, was measured in immature rats 18 hours after
a single subcutaneous injection of 8 mg of Aroclor/rat (approximately 200 mg/kg bw).
Increased glycogen content in uterus was detected in Aroclors 1221, 1232, 1242, and
1248, but not in Aroclors 1254, 1260, 1262, and 1268 (Bitman and Cecil, 1970).

According to the previous studies, exposure to the lesser chlorinated PCB
mixtures is correlated with estrogenic activity than exposure to the more highly
chlorinated mixtures. The lesser chlorinated PCBs are more rapidly metabolized and
eliminated than higher chlorinated PCBs (Burse et al., 1974; Safe, 1984). Compounds
containing a phenol ring appear to bind to estrogen receptors. The phenol ring mimics
the estrogen A ring and apparently promotes this binding (Duax and Griffin, 1985). The
PCB metabolites that result from the rapid metabolism of lesser chlorinated PCBs contain
hydroxyl groups. These create a structure that promotes binding to estrogen receptors
and, therefore, cause an increase in estrogenic activity.

All the rats in the above studies were treated with a single injection at the age of

22

23-26 days which resulted in increased uterine weights. However, when adult mice were
treated daily with 10 ppm Aroclor 1254 for up to 9 months, decreased uterine weights
in the 8- and 12-week-old offspring were observed (Linzey, 1988). Exposure period
appeared to be the primary contributor to differences in uterine weights. As with PCB
effects on testis weights, multiple exposures during gestation cause impairment in the
development of reproductive organs of offspring.
Effigy, on Steroid Hormones

During the prenatal period, testosterone is essential for the differentiation of male
internal genital ducts. Dihydrotestosterone (DHT), a metabolite of testosterone, is
responsible for differentiation of male external genitalia. During the early postnatal
period, testosterone is converted to estrogen in the brain, which is critical for
differentiation of the brain into the male type (Harris et al., 1964; Hadley, 1992;
Peterson et a1. , 1993). Exposure to foreign compounds may alter neonatal androgen
imprinting by changing the availability of steroid hormones during the critical
developmental period, or by direct action of the xenobiotic on the hypothalamus and
pituitary of neonates. The mechanism for decreasing plasma testosterone may involve
induction of detoxification enzymes which also metabolize steroids. Altered hepatic
microsomal testosterone hydroxylase activities were observed in adult male and female
rats treated intraperitoneally with a single dose of approximately 32 mg Aroclor 1254/kg
bw at 4 days postpartum (Haake-McMillan and Safe, 1991). Decreased plasma
testosterone was observed in the 5- and lO-week-old male offspring of rats gavaged with

3 mg TCB/kg bw daily from day 6 through 18 of gestation (Vincent et al., 1992). There

23

were, however, no significant changes in the plasma concentration of testosterone in the
10 to 12-week-old mouse offspring. Dams were gavaged with 20 mg 6-CB/kg bw every
3 days from day 13 of gestation through day 24 postpartum. The same results were
observed in mice gavaged with 40 mg 6—CB/kg bw or 160 mg Clophen A 50 every 2
days for 5 weeks, beginning from 5 weeks of age (Johansson, 1987). The treatments had
no effect on the synthesis of testosterone in testicular interstitial cells in vitra (Johansson,
1987).

Decreased testosterone concentrations were observed in the offspring of mice
exposed to TCB on days 6 to 18 of gestation, but not in offspring of mice exposed to 6-
CB from day 13 of gestation to day 24 after birth. In addition to treatment regimens,
the results of these two studies may be due to the nature of the congeners; TCB is a 3-
methylcholanthrene (3MC) type inducer while 6-CB is a phenobarbital (PB) type inducer.
Decreased testosterone concentrations were not detected in animals exposed to PCBs
during puberty or after puberty. This is supported by unaffected hepatic microsomal
testosterone hydroxylase activity in postpubertal rats treated with a single i.p. injection
of 32 mg Aroclor 1254/kg bw (Haake-McMillan and Safe, 1991). Although sample sizes
in the Johansson study are larger than those in the Vincent et al. study, standard
deviations in the former are larger than those in the latter. The large standard deviations
in the former study could explain the lack of effect of treatment on testosterone
concentrations.

Progesterone is imperative for the maintenance of pregnancy. Changes in

progesterone concentration have been documented in PCB-treated female rhesus monkeys

24
(Macaca mulatta) and mink (Allen et al., 1979; Aulerich et al., 1985). A decrease in

peak serum progesterone concentrations during the menstrual cycle was observed in adult
female rhesus monkeys treated with 2.5 ppm (approximately 100 pg/kg bw/day) and 5.0
ppm Aroclor 1248 for 4 months (Allen et al., 1979). However, when a daily oral dose
of 0, 5, 20, 40, or 80 pg Aroclor 1254/kg bw was delivered in gelatin capsules for 2
years, serum progesterone concentrations in PCB-treated female adult rhesus monkeys
were the same as those in the control group (Truelove et al. , 1990). Although the lowest
dose in the Allen et al. study is only marginally higher than the highest dose in the
Truelove et al. study, this may reflect the dose where treatment affects progesterone
concentrations. In addition, the different results may be due to the nature of the
mixtures, exposure routes, and length of treatment. Progesterone reduction is associated
with high incidences of abortion in Aroclor 1248—exposed rhesus monkeys (Barsotti et
al., 1976). In adult female mink fed 2.5 ppm Aroclor 1254 beginning 1 month prior to
breeding through parturition, decreased plasma progesterone concentration was noted.
However, treatment with 0.1 ppm HCB resulted in an increase in progesterone
concentrations (Aulerich et al., 1985).

Decreased progesterone concentrations are probably caused by increased
hydroxylation of progesterone. Although both Aroclor 1254 and HCB caused increased
hepatic microsomal monooxygenases activity, decreased progesterone concentrations were
only observed in Aroclor 1254-treated mink. These results may be due to the nature of
the PCB fed; HCB is a 3-methylcholanthrene (3MC) type inducer while Aroclor 1254

is a mixed type inducer. Aulerich et al. (1985) suggest that the adrenal glands are the

25

source of the elevated plasma progesterone. This is supported by the increase in adrenal
weights of the HCB-treated mink. Symptoms of debilitation such as lethargy and
weakness were observed in these mink. This is probably associated with the increase in
adrenal weights.

Although a dose-dependent in vitra breakdown of estradiol was observed in the
livers of American kestrels treated with Aroclor 1254 or Aroclor 1262 for 5 months
(Lincer and Peakall, 1970), changes in plasma estradiol were not observed in mammals
exposed to PCBs. When adult female mink were fed any of the following treatments 1
month prior to breeding through parturition, plasma 176-estradiol was not affected: 0.5
ppm HCB, 5 ppm 6-CB, 5 ppm 2,2’,3,3’,6,6’-hexachlorobiphenyl, or 2.5 ppm Aroclor
1254 (Aulerich et al., 1985). Nor were any changes observed in circulating estrogen
concentrations in the offspring of female rats gavaged with 3 mg TCB/kg bw/day for 13
days, beginning on day 6 of gestation (Vincent at al., 1992).

ff 11 me Pr i n

A reduction in testicular spermatozoa concentrations (sperm/mg testis) was
observed in adult mice treated with 50 or 200 ppm Aroclor 1254 in the feed for 2 weeks
after maturation (Sanders et al. , 1977). The number of sperm per testis was also reduced
in adult mice with a dose of 400 ppm Aroclor 1254 in the feed for 2 weeks, but not with
a dose of 100 or 200 ppm (Sanders and Kirkpatrick, 1975).

Decreased sperm production is accompanied by morphological changes in the
testis. Derangement of seminiferous lobules and hyperplasia of lobule walls have been

reported in Atlantic cod treated with 5 or 10 ppm Aroclor 1254 in the feed for 5.5

26

months (Freeman et al., 1982). However, 5 daily i.p. injections of Aroclor 1254 at 125-
1000 mg/kg bw in male mice caused no sperm head abnormalities when examined 5
weeks later (Topham, 1980).

A decrease in the number of oocytes and follicles was observed in the ovaries of
4-week-old C57/BL female mice exposed in utera via a single maternal i.p. injection of
15 mg TCB/kg bw (Ronnback, 1991). In contrast, an increased number of oocytes and
follicles has been observed in the 8-week-old female offspring of dams exposed to 9 or
15 mg TCB/kg bw weekly. Treatment began 2 weeks prior to breeding and continued
through gestation and lactation. No change of the number was observed when offspring
were at 4 weeks of age. (Ronnback and de Rooij, 1994).

ff n F ilizin A ili

To determine effects of early postnatal exposure to PCBs on gamete quality, rats
received 8, 32, or 64 mg Aroclor 1254/kg bw on days 1, 3, 5, 7, and 9 of lactation
(Sager et al., 1983). At 130 days of age, male offspring were mated with non-exposed
females. A decrease in the number of implanted eggs and live fetuses, and an increased
embryo resorption rate were noted. The authors speculated that this was caused by the
hypoandrogenic effect of PCBs. In follow-up studies, a decrease in the percentage of
normal fertilized eggs and eggs at the 2—cell, 4—cell, and blastocyst stages were observed
(Sager et al., 1987, 1991). In another study, when lactating mice were injected with 50
mg Clophen A 60/kg bw weekly for 3 weeks, with the first injection on the day of
parturition, no effect on fertility was observed when exposed male offspring were mated

to non-exposed mature females (Kihlstrom et al., 1975). Although the dose (50 mg

27
Clophen A 60/kg bw) in this study was significantly higher than the LOAEL (8 mg

Aroclor 1254/kg bw) in the Sager et al. studies, no effect on fertility was observed. The
similarity in chlorination of Aroclor 1254 (54%) and Clophen A 60 (60%) suggests a
similar toxic potency, but the results of this study do not appear to support this.
However, the effects observed by Sager et a1 . could represent the relatively high
exposure frequency of the Aroclor 1254 when compared to the Clophen A 60 treatment,
rather than a difference in toxic potency between the two mixtures. In addition to a high
exposure frequency, the rats in the Sager et al. studies received 4 treatments during the
first week postpartum. This is a critical period for sexual differentiation of the brain and
for the initiation of spermatogenesis. In contrast, the Kihlstrom et al. animals received
only 2 treatments during this period.

In female mice, 10 weeks of daily doses as low as 1.25 mg Clophen A 60/kg bw
reduced frequencies of implanted embryos. Treatment began 9 weeks prior to breeding
and continued for 8-10 days of gestation (Orberg and Kihlstrom, 1973). Similar results
were also observed in mice treated with individual congeners. A daily dose of
approximately 25.0 mg 2,4’,5-trichlorobiphenyl or 6-CB/kg bw, administered from day
1 to day 7 of gestation, resulted in decreased frequencies of implanted embryos (Orberg,
1978a).

Teratogenic and Developmental Toxicity

Commercial PCB mixtures have low teratogenic potency in laboratory animals.

PCBs are thought to exert their teratogenic effects via similar mechanisms as TCDD, a

widely recognized teratogen (Golub et al., 1991). Individual coplanar congeners are

28

structurally similar to TCDD and thus are the most teratogenic of all the congeners
(d’Argy et al., 1987; Morrissey and Schwetz, 1989; Peterson et al., 1993). Most
structure-related overt teratogenic effects of PCBs are consistent with an Ah receptor-
mediated mechanism. Nevertheless, one exception is 2,2’,3,3’,4,4’-hexachlorobiphenyl,
which, although having a very weak binding affinity for the Ah receptor, causes the same
pattern of teratogenic effects in mice as does TCDD (Marks and Staples, 1980).
Coplanar congeners, more potent teratogens than other congeners, are present in low
amounts in frequently tested commercial mixtures. This may account for the teratogenic
potency of mixtures. Up to 65 % malformations were seen in mice when dams were
gavaged with 8 and 16 mg HCB/kg bw on days 6 through 15 of gestation (Marks et al.,
1981). Cleft palate and hydronephrosis were the dominant malformations detected.
Teratogenicity of TCB was also investigated in mice (Simmons et al., 1984; d’Argy et
al., 1987). The ED50 for cleft palate in mice is a single i.p. injection of approximately
100 mg TCB/kg bw (d’Argy et al., 1987).

Sex ratios (cf/9) of neonates decreased when pregnant rats were treated with a
single dose of 3 mg TCB/kg bw on day 6 to day 18 of gestation (exact day not indicated)
(Simmons et al., 1984). This change appeared to be correlated with the tendency of
TCB metabolites to accumulate in fetal tissue, as demonstrated with the following study.
Three days after pregnant rats were administered “C-labeled 4-chloro-, 4,4’-dichloro—,
3,3’,4,4’-tetrachloro-, 2,2’,3,3’,6,6’-hexachloro-, or 2,2’,4,4’,5,5’-hexachlorobiphenyl
on day 18 of gestation, fetal blood levels of TCB radioactivity were higher than those of

the other congeners (Lucier et al. , 1978).

29

PCBs have caused developmental toxicity in many laboratory animals. The signs
of toxicity are similar to those that have been observed in humans (Hemminki and
Vineir, 1985). These signs include neonatal death, growth retardation, hyperkeratosis,
atrophy and cystic dilation of skin, and neurobehavioral dysfunction. In addition to these
indications of toxicity, structural malformations have been observed in laboratory animals
(Morrissey and Schwetz, 1989; Peterson et al., 1993). Impaired reproductive
performance and neurobehavioral effects are the most frequently observed functional
alterations. Although decreased postnatal growth has been observed in most cases,
increased postnatal growth rates were observed in mice treated daily with 0.05 mg
2,4’,5-trichlorobiphenyl or 6-CB pre- and postnatally (Orberg, 1978b).

Decreased litter sizes at weaning were observed when wild-caught and laboratory-
raised mice were treated with 10 ppm Aroclor 1254. It was suggested that exposure to
PCB-contaminated foods in the field would contribute to a potential decline in the wild
population of wild white-footed mice by reducing the number of young mice entering the
breeding population (Linzey, 1987). When adult female mink were fed 0.5 ppm HCB
beginning 1 month prior to breeding through parturition, all offspring died within 60 days
of birth (Aulerich et al., 1985).

3,3’,4,4’-Tetrachlorobiphenyl (TCB) Toxicity

The toxicity of TCB, P,CB, and HCB are frequently discussed together because
they are all approximate isostereomers of TCDD. Cytochrome P-450 induced by either
of these congeners or TCDD produces similar spectral maximal absorbances as

cytochrome P-450 induced by 3-methylcholanthrene (3MC). These three congeners are,

30

therefore, known as MC-type congeners. The mechanism of toxicity of these coplanar
congeners and TCDD is via Ah receptors in the cytosol. After TCDD or coplanar
congeners bind to the Ah receptor, a receptor-ligand complex translocates into the
nucleus. Following the interactions between this complex and genomic recognition sites,
transcription and translation of the specific genes that encode for cytochrome P-448
activity are initiated (Sipes and Gandolfi, 1991). Since there is a high correlation
between Ah receptor binding affinity and toxicity, a toxic equivalency factor (TEF)
approach has been adopted by regulatory agencies for the risk assessment of
polychlorinated dibenzodioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and
relevant compounds (Safe, 1994). All individual PCB congeners have been assigned a
TEF value, which is the fractional toxicity of the congener relative to TCDD. Based on
the mean of TEFs of over 10 responses elicited by each congener, the range of TEFs
among those mono-ortha coplanar congeners is between 0.001 and 0.00005. The TEFs
for TCB, P,CB, HCB are 0.01, 0.1, and 0.05, respectively (Safe, 1994).

These 3 coplanar congeners are generally deemed highly toxic, with TCB being
the least toxic of the three. After a single i.p. injection, ED50$ of PSCB, HCB, and TCB
for body weight loss were 3.25, 15.1, and 730 pmol/kg bw, respectively (Leece er al.,
1985). This pattern was reflected by the induction of benzo[a]pyrene hydroxylase (BP
hydroxylase), a MC-type hepatic microsomal enzyme. The increasing order of induction
potencies of BP hydroxylase was TCB < < HCB s PSCB (Yoshimura et al. ,1979). The
low toxicity of TCB when compared to the other 2 coplanar congeners corresponds with

its relatively rapid biotransformation in viva (Abdel-Hamid et al., 1981; Mizutani et al.

31
1977).

The 2 TCB metabolites in rat feces, 5-hydroxy-TCB and 4-hydroxy-3,3’,4’,5-
tetrachlorobiphenyl were considered less toxic than TCB due to their inability to cause
liver enlargement, thymus atrophy, or induction of BP hydroxylase (Yoshimura et al.,
1987). Similar results were observed in the comparison between the toxicity of PsCB
and its metabolite 4’-hydroxy-P,CB (Koga et al. , 1990). Chicken embryo mortality was
not affected by a dose of 2.0 mg hydroxylated metabolites of TCB/kg egg (Klasson
Wehler et al. , 1990). In contrast, 75 % mortality was reported for chicken embryos
injected with 0.02 mg TCB/kg egg in ava. These studies demonstrate the considerable

toxicity of TCB when compared to the inability of TCB metabolites to elicit toxic effects.

HYPOTHESIS

am;

The first hypothesis is that the TCB tissue concentrations will be relatively higher
in the low-dose group than the tissue concentrations in the high-dose group. This is
derived from the fact that high concentrations of PCBs can induce hepatic microsomal
monooxygenases to a greater extent than low concentrations.

The second hypothesis is that accumulation of TCB is higher in pregnant mice
than in non—pregnant mice. This is based on the observation that the PCB-induced
increase in hepatic microsomal monooxygenases was less in pregnant mice than in virgin
animals.

M

The hypothesis is that exposure to TCB will impair both reproductive performance
and gamete integrity in mice. PCBs induce hepatic microsomal monooxygenases that can
metabolize steroid hormones which are essential for prenatal and early postnatal
imprinting and development of accessory sex organs. The imbalance of steroid hormones

can elicit adverse effects on organogenesis and gametogenes is.

32

MATERIALS AND METHODS

mal
Animak
Twenty four C57BL/6J female mice at 15 weeks of age were used in this phase.
Mice were housed in a containment room at University Laboratory Animal Resources,
Michigan State University. Room temperature was maintained at 70—72°F with 40-60%
humidity and 12 hour light/12 hour dark cycle.
Chemicals
UL—“C-TCB (Lot # 026F9204; 50 pCi/50 pl toluene) was purchased from Sigma
Chem. Co., St. Louis, MO. Specific activity of l‘C-TCB was 37.1 mCi/mmol and
chemical purity exceeded 98%. TCB (Lots # 70075 and 10162; purity>99.4%) was
purchased from AccuStandard Inc. , New Haven, CT. Liquid scintillation cocktail was
purchased from Beckman Instruments, Inc., Fullerton, CA. Acetonitrile (HPLC grade)

was purchased from EM Science, Gibbstown, NJ. Acetone was purchased from J. T.

Baker Inc., Phillipsburg, NJ.
Feed Preparation
In this feeding trial, 1‘C-TCB was used to measure placental transfer of TCB as

well as distribution of TCB in tissues. Two types of feed were prepared: “C-TCB feed
(“C-TCB and TCB) and TCB feed.

33

34
“W

Feed contained 0, 3 ppm, or 30 ppm total TCB. Purchased l“C-TCB (50 pCi/50
pl) was diluted by a factor of 20 with toluene to make a 50 pCi/1 ml stock solution.
Acetone was used as a carrier to disperse l‘C-TCB and TCB in ground mouse chow 5015
(PMI Feeds, Inc., St. Louis, MO). After 9.52 pCi l‘C-TCB in 190.5 pl toluene and
TCB were dissolved in 60 ml of acetone, 1 kg ground mouse chow was added and
mixed. The specific radioactivity was 3.17 and 0.317 pCi/mg TCB in 3 ppm and 30
ppm “C-TCB feed, respectively. The feed was tumbled for 1 hour and air dried in a
hood at room temperature for 24 hours, before being stored in glass jars in a 04°C
refrigerator until use. The control feed consisted of ground mouse chow mixed with
acetone and toluene and similarly tumbled and dried. Feed samples were analyzed by
the laboratory of Dr. Thomas Voice, Department of Civil and Environmental
Engineering, Michigan State University, at the onset of this study.

193.1524

Feed containing TCB only was prepared using the same procedure described

above but without l‘C-TCB. The feed was prepared monthly.
Treatments

Before mating, 8 female C57BL/6J mice (F—O) at 15 weeks of age were randomly
assigned to each treatment and provided with “C-TCB feed ad libitum for 2 weeks.
Each female was then paired with one non-treated mature C57BL/6J male for 10 days
between 7:30 PM and 7:30 AM, during which time feed was withheld to avoid exposing

males to treatment feed. The presence of the vaginal plug was defined as day 1 of

35

gestation. F—O females continued on the “C-TCB treatment until parturition and were
then switched to a TCB diet for the next 21 days. This study was terminated 21 days
postpartum when the offspring were weaned.
Sample Collection

On day 19 of gestation, liver, thymus, uterus, abdominal adipose tissue, placenta,
and fetus samples were collected and weighed. These same tissues, except uterus,
placenta, and fetus, were also collected on day 21 of lactation. One animal from each
treatment group was used for sample collection. One non-pregnant mouse from each
treatment was sacrificed concurrently with the animals on day 19 of gestation and day
21 of lactation. The liver, thymus, and abdominal adipose tissue of 3-week-old male
offspring were collected, weighed and stored for subsequent analysis.

Sample Extraction and Analysis

Each sample was prepared for TCB and radioactivity analysis by disruption in 10
ml acetonitrile with a Dismembrator Sonic (sonicator model 550, Fisher Scientific,
Pittsburgh, PA) at 20% power for 10 minutes. The sample was then centrifuged at 1500
RPM for 10 minutes. Ten ml liquid scintillation cocktail was added to 5 ml supernatant
for the measurement of radioactivity with a liquid scintillation counter (Model 1500 Tri-
Carb, Packard Instrument Company, Meriden, CT). The concentrations of TCB
equivalents in the tissues and organs were calculated from the radioactivity measured in
the tissues and organs and the specific activity in feed (Appendices I and II). The term

TCB equivalents refers to TCB and its metabolites.

36
M
Animals
Ninety C57BL/6J female mice at 13 weeks of age were used in this phase. Non-
treated B6D2-Fl males were used as sperm donors for the in vitra fertilization assay.
Mice were raised in the same facility as in Phase I.
Chemicals
TCB was purchased from AccuStandard Inc., New Haven, CT. Pregnant mare’s
serum gonadotropin (PMSG), human chorionic gonadotropin (HCG), and all other
chemicals were purchased from Sigma Chemical Co., St. Louis, MO.
Feed Preparation
The method of feed preparation was the same as described in Phase I.
Treatments
Female C57BL/6] mice (F-O) were fed Purina lab chow containing 0, 3 ppm, and
30 ppm TCB for 2 weeks beginning at 13 weeks of age. Each female was then paired
with one non-treated mature C57BL/6J male for 10 days, between 7:30 PM and 7:30
AM, during which time feed was withheld to avoid exposing males to treatment feed.
The treatment for F—O females continued through mating, gestation, and lactation. The
offspring (F-l) were given the same feed as their dams received throughout the study.
Females (F—l) were superovulated at 5, 6, and 7 weeks of age and egg fertilizing ability
was assessed by in vitra fertilization (IVF) assay. At 7 weeks of age, 10 F-l females in
each treatment group were bred with non-treated mature B6D2-F1 males for 10 days.

Ten treated F-l males in each treatment group, aged 7 and 17 weeks, were used to breed

37
non-treated mature B6D2-F1 females for 10 days. After breeding, sperm from these

males were collected and fertilizing ability was assessed by IVF assay. Determination
of sperm concentration and sperm motion analysis were performed on sperm from 9-
week-old F-l males hourly for 4.5 hours after sperm collection with the CellSoft
computer-assisted digital image analysis system (CRYO Resources Inc., New York,
1986).

Parameters Measured

W Feed intake was recorded for F-0 females before mating, during
gestation, and during lactation. The amount of feed added every day was recorded.
Actual feed consumption was calculated every 3 days by subtracting the amount of feed
left in the feed jars from the total feed provided over the 3-day period.

W Body weights of F—O females were recorded at the initiation of the
study (day 1), before breeding (day 14), on day 19 of gestation (day 33), and on day 21
of lactation (day 54). When the offspring were 4 days, 1 week, and 2 weeks of age,
body weights were recorded by litters. The average weight of offspring in a litter was
used for statistical analysis. After weaning, individual weights were recorded weekly.

Might Liver and thymus weights of F-O females were recorded for animals
on day 19 of gestation and on day 21 of lactation. Liver and thymus weights of El
females were recorded at 5, 6, and 7 weeks of age. Liver, thymus, and testis weights
of F-l males were determined at 3, 9, and 19 weeks of age.

WWW Fecundity. litter sizes, sex
ratios, and 4-day and 21-day survival indices were recorded for both F-0 and F-l

38

generations. Sex of pups was determined by anogenital distances at day 4 of age or at
death if they died before day 4 of age. If pups died before sex could be determined, they
were not included. A 4—day survival index is defined as 100 x [number of pups viable
at day 4 of age]/[number of viable pups born]. A 21—day survival index is defined as 100
x [number of pups viable at day 21 of age]/[number of pups retained at day 4 of age]
(Thomas, 1991).

W Treated F-l females were superovulated with 8 IU PMSG, and
then 8 [U HCG 48-50 hours later. Thirteen hours after HCG injection, the females were
killed by cervical dislocation. Oviducts and distal portions of the uteri were excised.
The eggs were collected by gentle teasing from the oviducts and incubated 5-15 minutes
prior to insemination. Epididymal sperm from non-treated B6D2-F1 males were
collected and incubated for 1.5 hours before insemination. Eggs in each petri dish,
collected from one female, were inseminated with 50 pl sperm suspension (approximately
6 x 10’ sperm) and incubated for 24 hours. Incubation occurred in modified Tyrodes
medium at 37°C and 5 % CO2 after collection and insemination. Twenty-four hours after
insemination, 50 pl of a 35 pM solution of bisBenzimide (Hoechst 33258) were added
to the petri dish. It was incubated for another 30 minutes, and then observed under a
Nikon Optiphot microscope equipped with a 100w mercury bulb, 365/10 nm excitation
filter, 400 nm dichroic mirror, and 400 nm barrier filter. Eggs at the 2-cell stage or
eggs containing a second polar body and two pronuclei were considered fertilized.
Degenerated eggs were considered non-fertilized.

The IVF assay was also used to evaluate sperm fertilizing ability of treated males.

39

The method was the same as described above. Epididymal sperm from treated males at
9 and 19 weeks of age were collected and used to inseminate eggs from non-treated
B6D2-F1 mice.

Culture Medium Modified Tyrodes media contained 99.23 mM sodium chloride,
2.68 mM potassium chloride, 1.80 mM calcium chloride, 0.36 mM sodium phosphate
monobasic anhydrous, 0.49 mM magnesium chloride hexahydrate, 25.0 mM sodium
bicarbonate, 52.0 mM sodium lactate, 0.25 mM sodium pyruvate, 5.56 mM D-glucose,
100 IU/ml sodium penicillin G., and 100 IU/ml streptomycin sulphate.

3mm Conoontgation and Motion Aoalysis Twenty pl of the sperm suspension
(approximately 2.4 x 10’ sperm/ml) were placed on a CellSoft 20 pm chamber and
analyzed with the CellSoft computer-assisted digital image analysis system. A minimum
of 100 sperm cells were analyzed to obtain measurements of concentration, motility,
velocity, linearity, mean amplitude of lateral head (ALH) displacement, and beat/cross
frequency. Motility is the percentage of sperm that travel more than 20 pm/sec.
Velocity is the average distance (pm) traveled by motile sperm in one second. Linearity
is a measure of whether or not the sperm move in a straight line. It is calculated by
dividing the length of the straight line distance by the actual track distance. This fraction
is then multipled by 10. Linearity, therefore, ranges from 0 to 10. The straighter the
actual cell track, the higher the linearity value such that a value of 10 indicates a
perfectly straight line and 0 indicates a circular track. ALH displacement is a measure
of the displacement of the sperm head from a computer-calculated curval mean of its

track. The maximum ALH displacement is measured from the curval mean for every

4O

cycle of the cell’s track, then multipled by 2. For the individual cell data, the mean
ALH displacement is the mean of the maximum ALH displacement for the cycles of the
track. For the summary data, the mean ALH displacement is an average of the mean
ALH displacement for each cell. High ALH displacement appears to be linked with the
process of capacitation. The beat/cross frequency (Hz) is the number of beats (or
crosses) per second. Every time the sperm cell crosses the computer-calculated curval
mean, the computer counts that crossing as one beat.
Statistical Analysis

Fecundity, survival indices, fertilization rate, degeneration rate, and sperm
motility were evaluated with the Chi square test. Body weight, organ weight, feed
intake, and ALH displacement were initially tested with One Way Analysis of Variance
(ANOVA). Log transformation was performed for the discrete quantitative parameters
and ratios (i.e., litter size, number of eggs ovulated, sperm concentration, beat/cross
frequency, sperm linearity, sex ratios, and relative organ weights) before ANOVA.
Relative organ weights refer to organ weights as percentages of body weights. Those
data passing homogeneity of variance and normality tests were analyzed with the Student-
Newman-Keuls multiple pairwise comparison. The Kruskal-Wallis Test was performed
to evaluate the non-Gaussian data. These analyses were performed with Sigmastat" (Kuo

et al., 1992).

RESULTS

Lime!

Results of feed analysis are shown in Table 1. TCB concentrations were
approximately 80% of expected concentrations in Phase I and over 90% in Phase II.

After 5 weeks of exposure, TCB equivalents in the liver, adipose tissue, uterus,
and fetus from the 30 ppm group were 3-6 fold higher than the levels in the 3 ppm
group, except for the uterus in the non-pregnant groups (Table 2). Concentrations of
TCB equivalents were similar in the uterus, placenta, and fetus in the 30 ppm group
(Table 2). No placenta data was available in the 3 ppm group.

After 8 weeks of I‘C—TCB treatment, TCB equivalents were still accumulating in
the liver and adipose tissue of non-pregnant mice, in both the low and high dose groups
(Table 3). TCB equivalents in the thymus of lactating mice were higher than those in
the non-pregnant mice, regardless of dose (Table 3). TCB equivalents in the thymus of
30 ppm- treated lactating mice were over 40 times higher than equivalents in the 3 ppm-
treated lactating mice (Table 3).

In the 3 ppm-treated group, approximately 15 % of the concentrations of TCB
equivalents that had accumulated in the liver and adipose tissue before parturition were
detected after termination of lactation (Tables 2 and 3). However, in the 30 ppm-treated

group, approximately 56% and 4% , respectively, of the concentrations of TCB

41

42

equivalents that had accumulated in the adipose tissue and liver before parturition were
observed after termination of lactation (Tables 2 and 3).

After weaning, radioactivity was only detected in the liver and adipose tissue, not
in the thymus of the offspring (Table 4). Similar concentrations of TCB equivalents were
detected in the adipose tissue and liver of the 3 ppm-treated offspring. However, in the
30 ppm group, concentrations of TCB equivalents in the adipose tissue were 14.6 times
higher than concentrations in the liver (Fable 4). Although the pups were nursed by the
same dam in the 30 ppm group, large variations in the concentrations of TCB equivalents

were observed (Table 4).

43

 

 

 

 

Table 1. TCB concentrations in treatment feed.
Treatments
Control 3 ppm 30 ppm
Concentration (ppm)
Phase I‘ 0 2.36 :l: 0.27 23.49 i 1.18
(n=2) (n=3) (n=3)
Phase II" 0 3.16 :l: 0.24 27.59 i: 0.70
(n=3) (n=3) (n=3)
a The TCB concentrations in feed were calculated from the specific activity in

feed and the radioactivity measured in the feed with a liquid scintillation counter
(Appendices I and II).

b TCB concentrations in Phase II were analyzed by gas chromatography.

44

Table 2. Concentrations of TCB equivalents in tissues and organs of non-pregnant
and pregnant mice.

 

 

 

 

 

 

Organ/ Concentration' Treatments
Tissue
Non-pregnantb Pregnant"
3 . 30 3 30
PM“ PM“ PM“ PPm°
Adipose
Tissue ppm 0.88 4.86 0.99 3.26
Liver ppm 0.94 3.23 0.92 5.71
Thymus ng TCB/
Thymus 4.4 25.2 2.5 15.8
Uterus ppm 0.29 0.21 0.12 0.54
Placenta ppm NAd NA -‘ 0.61
Fetus ppm NA NA 0.14 :l: 0.03 0.47 :I: 0.13
(n=8) (n=21)
a The term equivalents refers to TCB and its metabolites. The concentrations of

TCB equivalents in organs and tissues were calculated from the specific activity
in feed and the radioactivity measured in the organs and tissues with a liquid
scintillation counter (Appendices I and II).

b Both non-pregnant and pregnant mice were treated with l‘C-TCB feed for 5
weeks.

0 Values are the means of data from 2 pregnant mice in the 30 ppm group. Table
values from all other groups reflect a sample size of one.

(1 Not applicable.

e Missing data.

45

Table 3. Concentrations of TCB equivalents in adipose tissue and organs of non-
pregnant and lactating mice.

 

 

 

 

 

 

Organ/ Concentration' Treatments
Tissue

Non-pregnantb Lactating"

3 ppm 30 ppm 3 ppm 30 ppm
Adipose Tissue ppm 1.26 8.71 0.14 1.83
Liver ppm 2.29 5.98 0. 15 0.24
Thymus ng TCB/

Thymus 0 132.1 4.8 205.2
a The term equivalents refers to TCB and its metabolites. The concentrations of

TCB equivalents in organs and tissues were calculated from the specific activity
in feed and the radioactivity measured in the organs and tissues with a liquid
scintillation counter (Appendices I and II).

b Non-pregnant mice were treated with l‘C-TCB feed for 8 weeks. lactating mice
were treated with l‘C-TCB feed for 5 weeks before parturition, and TCB feed for
3 weeks after parturition. Table values reflect a sample size of one.

46

Table 4. Concentrations of TCB equivalents in adipose tissue and organs of male
offspring at 3 weeks of age.

 

 

 

 

Organ/ Concentration‘l Treatments
Tissue

3 ppm 30 ppm

(n=4) (n=5)
Adipose Tissue ppm 0.15 :t 0.03 8.15 :i: 7.84
Liver ppm 0.10 :I: 0.01 0.56 :l: 0.14
Thymus ng TCB/

Thymus 0 0
a The term equivalents refers to TCB and its metabolites. The concentrations of

TCB equivalents in organs and tissues were calculated from the specific activity
in feed and the radioactivity measured in the organs and tissues with a liquid
scintillation counter (Appendices I and II). The calculated concentrations of TCB
equivalents were transferred through the placenta and milk from the dams which
were treated with 1‘C-TCB feed for 5 weeks before parturition, and then TCB
feed for 3 weeks after parturition.

47
Biased

After the first two weeks of treatment, there were no significant differences in
body weights in the F-0 mice, although mice in the 30 ppm group consumed more feed
than those in the 3 ppm group (Table 5). Feed intake during gestation days 1 - 16 was
also higher in the 30 ppm group than in the 3 ppm group, although there were still no
significant differences in body weights between the 2 groups. On day 21 of lactation,
however, the 3 ppm TCB-treated mice consumed more feed than the 30 ppm-treated
mice, although they weighed less than the 30 ppm-treated mice (Table 5). On day 19
of gestation and on day 21 of lactation, relative liver weights of F-0 mice were higher
in the 30 ppm treatment group than in the 3 ppm and the control groups (Table 6).
These differences were not apparent in non-pregnant females. TCB treatments had no
significant effect on thymus weights in F-O females (Table 6).

Fecundity in F-0 females was 80%, 71%, and 47% in the control, 3 ppm, and
30 ppm groups, respectively (Table 7). The treatments had no effect on litter sizes and
sex ratios. Four-day and 21-day survival indices showed a trend of dose-response
relationships. Decreased survival was observed in the 30 ppm group, in which most of
the mortality occurred before 4 days of age (Table 7). There were no significant
differences in body weights of F-l offspring (Table 8). At 5 and 6 weeks of age, F-l
females in the 30 ppm group showed an increase in liver weights, but not in relative liver
weights (Table 9). A decrease in thymus weights was observed in the 30 ppm-treated
F-l females between 5 and 7 weeks of age (Table 9). However, relative thymus weights

decreased in the 30 ppm-treated F-l females only at 5 weeks of age.

Table 5.

Effects of TCB on feed intake and body weights of F-0 female mice.

 

 

 

Treatments

Control 3 ppm 30 ppm
Body Weight (g)c 21.6 :1: 0.8 21.7 :1: 0.7 21.8 :I: 1.0
(13 Weeks of Age) (11: 18) (n=24) (n=46)
Feed Intake (g/day)‘l 4.7 :I: 0.9‘”- 4.4 d: 0.8‘I 5.2 :1: 0.6”
(13-15 Weeks of Age) (n=15) (n=15) (n=15)
Body Weight (g)° 21.7 :I: 0.9 21.6 :t 0.5 22.2 i 1.1
(15 Weeks of Age) (n=18) (n=24) (n=46)
Feed Intake (g/day)e 4.8 i 1.2 4.6 :l: 1.4 5.4 :I: 1.0
During Gestation (n=7) (n=7) (n=4)
Body Weight (g) 36.9 :i: 2.1 37.3 i 1.2 35.6 :I: 2.9
Day 19 of Gestation (n=3) (n=3) (n=3)
Feed Intake (g/day)f 8.8 :I: 2.3‘ 10.3 :I: 1.6‘ 7.7 :I: 2.3"
During Lactation (n=9) (n=8) (n=10)
Body Weight (g) 28.5 :I: 1.6” 26.2 :1: 0.1‘ 29.3 :1: 1.6b
Day 21 of lactation (n=3) (n=3) (n=3)

 

ab

"'5

(P <0.05).

Different superscripts represent significant differences within the same row

Females were treated with TCB 2 weeks beginning at 13 weeks of age. Each
female was then paired with one non-treated mature male for 10 days.

Values represent mean feed intake :1: SD during the two weeks before mating.
Values represent mean feed intake i SD from day 1 to 16 of gestation.

Values represent mean feed intake :I: SD from day 1 to 21 of lactation.

49

 

 

 

 

Table 6. Effects of TCB on organ weights and relative organ weights of F-O
females.
Weightc Treatments
Control 3 ppm 30 ppm
(n=3) (n=3) (n=3)
N on-pregnant _ (18 Weeks of Age)
LW (g) 1.33 :I: 0.36 1.08 :1: 0.04 1.47 i 0.55
LW/BWd X 100 5.73 :l: 1.24 5.03 :I: 0.11 6.62 :l: 2.65
TMW (g) 24.4 :1: 8.4 32.4 :1: 5.3 36.8 i 9.8
TMW/BW‘I x 100 0.11 :1: 0.04 0.15 :I: 0.03 0.16 i 0.04
Day 19 of Gestation
LW (g) 1.69 i 0.07 1.81 :l: 0.12 1.82 :l: 0.13
LW/BWe x 100 4.59 :I: 0.09' 4.84 :I: 0.23' 4.95 :I: 0.07"
TMW (mg) 19.0 :t 9.2 25.6 :I: 7.0 18.0 :I: 1.4
TMW/BW‘ x 100 0.06 :I: 0.03 0.07 :I: 0.02 0.05 :I: 0.003
Day 21 of Lactation
LW (g) 2.06 i: 0.16‘” 1.99 d: 0.11‘ 2.37 :l: 0.20b
LW/BW° x 100 7.22 :1: 0.34' 7.60 :t 0.42‘I 8.10 :l: 0.32”
TMW (mg) 34.4 :t 2.3 29.0 :1: 5.9 31.7 :I: 9.6
TMW/BW‘ x 100 0.12 :t 0.01 0.11 i 0.02 0.11 :l: 0.03
ab Different superscripts represent significant differences within the same row

C

(P<0.05).

BW, LW, and TMW represent body, liver, and thymus weights, respectively.
Values represent mean organ weights or organ weights as percentages of body
weights :1: SD.

Body weights were 23.1 :1; 1.2, 21.5 :1: 0.5, 22.2 :t 0.5 g in control, 3 ppm, 30
ppm group, respectively.

Body weights were based on the data in Table 5.

50

 

 

 

 

Table 7. Effects of TCB on reproductive performance of F-0 female mice.
Treatments
Control 3 ppm 30 ppm
Fecundity (%)c 80“ 71" 47"
Litter Size 8.3 :I: 2.0 ‘ 8.3 :I: 1.3 8.2 :1: 1.2
(n=12) (n=16) (n=20)
Sex Ratio (9/6) 1.11 :I: 0.51 1.09 :I: 0.72 1.05 :l: 1.00
(n=12) (n=16) (n=20)
4-Day Survival 47.4 :I: 34.7‘I 41.6 i 39.1‘l 30.9 :I: 30.9"
Indexd (n = 12) (n = 16) (n =20)
21-Day Survival 96.3 :1: 11.1‘ 88.0 :1: 31.6‘ 75.8 :I: 37.7"
Index° (n =9) (11 = 10) (n = 13)
ab Different superscripts represent significant differences within the same row
(P<0.05).
c Values are expressed as percentages of females mated which became pregnant.
(I Four-day survival index = 100 X [number of pups viable at day 4 of age]/ [number
of viable pups born].
e Twenty-one-day survival index=100 x [number of pups viable at day 21 of

age]/[number of pups retained at day 4 of age]. The litters in which all pups died
before day 4 of age were not included.

51

 

 

 

Table 8. Effects of TCB on body weights of El mice.

Age' Treatments

Control 3 ppm 30 ppm
BW (g)

Day 4 1.8 i 0.3 1.8 :l: 0.4 1.7 :I: 0.3
(n=9) (n=10) (n=l3)

Week 1 3.3 i 0.7 3.6 :t 0.4 3.1 :l: 0.6
(n=9) (n=9) (n=12)

Week 2 6.9 :t 1.4 7.3 :l: 0.6 6.5 :I: 1.5
(n=9) (n=9) (n=12)

Week 3 7.9 :I: 1.2 8.4 :I: 0.7 7.6 :i: 2.4
(n=45) (n=51) (n=37)

Week 4 14.5 :1; 1.3 14.7 :i: 2.8 14.9 :I: 3.5
(n=42) (n=51) (n=32)

Week 5 19.3 :I: 2.4 19.0 i 3.2 19.0 :I: 3.6
(n=32) (n=47) (n=29)

week 6 20.0 :I: 2.5 20.1 :I: 3.2 20.6 :t 3.1
(n=41) (n=49) (n=27)

Week 7 21.7 i 3.1 21.2 :t 3.5 21.3 :I: 3.3
(n=33) (n=45) (n=21)

 

a Body weights on day 4, week 1, and week 2 represent the means :1: SD of litters.
Values for week 3 through week 7 are means :1: SD of individual weights.

52

Table 9. Effects of TCB on body weights, organ weights, and relative organ
weights of F-l females between 5 and 7 weeks of age.

 

 

 

Age Treatments
Control 3 ppm 30 ppm
BW (g)
Week 5 17.8 i 0.5 (n=7) 18.3 :1: 1.5 (n=9) 18.3 :I: 1.8 (n=8)
Week 6 18.1 :I: 0.7 (n=6) 18.9 :I: 0.2 (n=3) 18.4 :t 0.1 (n=2)
Week 7 19.6 :t 0.7 (n=7) 19.8 :I: 1.0 (n=10) 18.9 j: 1.2 (n=4)
Relative Liver Weight (LW/BW x 100)°
(LW (g))‘
Week 5 6.26 :1: 0.53 6.09 :I: 0.47 7.05 :I: 0.42
(1.11 :1; 0.08‘) (1.11 i 0.14‘) (1.30 :I: 0.17")
Week 6 6.00 :I: 0.23 5.84 :I: 0.33 7.04 :l: 0.37
(1.09 i 0.07‘) (1.10 :I: 0.07‘) (1.30 :I: 0.08")
Week 7 5.75 :i: 0.60 6.06 :I: 0.38 6.35 i 0.39
(1.13 :I: 0.14) (1.20 :I: 0.12) (1.26 :I: 0.15)
Relative Thymus Weight (TMW/BW x 100)°
(TMW 011g»d
Week 5 0.44 i 0.88‘" 0.46 :l: 0.05‘ 0.36 :t 0.03"
(77.7 :1: 12.0‘) (81.0 :I: 7.0‘) (60.8 :I: 5.9")
Week 6 0.43 :I: 0.08 0.42 :l: 0.09 0.30 :t 0.02
(77.5 :1: 13.7') (80.1 :I: 16.9‘") (55.1 :1; 2.3")
Week 7 0.39 :t 0.08 0.37 :I: 0.08 0.30 :I: 0.06
(75.2 j: 11.7‘) (67.8 i 9.8‘") (55.8 :1: 9.4")

 

ab Different superscripts represent significant differences within the same row
(P < 0.05). .

c BW, LW, and TMW represent body, liver, and thymus weights, respectively.
Values represent mean of organ weights as percentages of body weights :I: SD.

(1 The numbers in the parentheses under relative organ weights represent organ
weights :1: SD.

53

Decreases in fertilizing ability of F-l eggs in vitra were observed in the 3 ppm
and 30 ppm groups when compared to the control group (Table 10). Fertilization rates
were 68.4%, 47.6%, and 45.6% for the control, 3 ppm, and 30 ppm groups,
respectively. The percentage of degenerated eggs in the 3 ppm and 30 ppm groups was
higher than that in the control group. The number of eggs superovulated by exogenous
PMSG and HCG were the same among all groups (Table 10). F-1 females at 7 weeks
of age showed similar fecundity among all groups when bred with non-treated males
(Table 11). Nonetheless, all offspring from 3 ppm and 30 ppm-treated F-l females died
within 4 days of birth (Table 11).

Testes in treated F-l males at 3 weeks of age were heavier in mice treated with
30 ppm TCB than for mice in the control group (Table 12). However, relative testis
weights were higher in both 3 ppm and 30 ppm groups than in the control group. Liver
weights, relative liver weights, thymus weights, and relative thymus weights were similar
among all groups (Table 12). At 9 and 19 weeks of age, liver weights and relative liver
weights were higher in the 30 ppm group than in the other 2 groups. However, thymus
and testis weights were the same among all 3 groups (Tables 13 and 14). Treated F-l
males, at 7 and 17 weeks of age, were bred with non-treated females. Fecundity, litter
sizes, sex ratios, and survival of the offspring were the same among all treatment groups
(Tables 15 and 16). The 4—day survival index decreased by 20% in the 3 ppm and 30

ppm groups when treated males at 60 weeks of age were paired with non-treated females

(Table 17).

54

Table 10. Fertilizing ability of eggs from F-l females exposed to TCB in utera

through postnatal development.

 

 

 

Treatments

Control 3 ppm 30 ppm
No. Females Tested 15 16 12
No. Exp. 6 6 6
Eggs Ovulated
Per Animal (No.) 40 j; 10 29 j; 21 34 j; 22
Fertilization
In Vitra (%) 68.4 :1: 15.1‘ 47.6 :I: 25.9" 45.6 :I: 18.7"
Degeneration
Rate (%) 6.0 :I: 4.8‘l 15.6 :I: 20.2" 13.4 :I: 7.9"

 

ab Different superscripts represent significant differences within the same row

(P < 0.05) '

55

Table 11. Effects of TCB on reproductive performance of F -1 females at 7 weeks

 

 

 

 

of age.
Treatments
Control 3 ppm 30 ppm
No. Females Mated 10 10 10
Fecundity (%) 50 . 40 70
Litter Size 7.0 :t 2.3 5.3 :I: 1.7 7.6 i 1.9
(n=5) (n=4) (n=7)
4-Day Survival 46.4 :1; 43.0' 0" 0"
Indexc (n =5) (n =4) (n =7)
ab Different superscripts represent significant differences within the same row
(P<0.05).
c Four-day survival index = 100 x [number of pups viable at day 4 of age]/[number

of viable pups born].

56

Table 12. Effects of TCB on body weights, organ weights, and relative organ

weights of F-l males at 3 weeks of age.

 

 

 

 

Weightc Treatments
Control 3 ppm 30 ppm
(n=7) (n=7) (n=8)
BW (g) 8.4 :1: 0.5 8.7 :I: 0.5 8.5 :l: 0.7
LW (g) 0.35 :I: 0.08 0.31 :1: 0.05 0.34 :t 0.04
LW/BW X 100 4.2 :l: 0.8 3.6 :I: 0.5 4.0 :I: 0.2
TMW (mg) 53.2 :t 5.0 53.3 :I: 9.8 49.4 :t 13.4
TMW/BW X 100 0.64 i 0.09 0.62 :l: 0.13 0.58 :t 0.14
'I'TW (mg) 32.9 d: 3.3‘ 40.6 :1: 5.7‘" 43.2 :1: 7.0"
TTW/BW X 100 0.39 :I: 0.03‘I 0.47 i 0.05" 0.51 :t 0.10"
ab Different superscripts represent significant differences within the same row

(P < 0.05).

BW, LW, TMW, TTW represent body, liver, thymus, and testis weights,
respectively. Values represent mean organ weights or organ weights as
percentages of body weights :I: SD.

57

Table 13. Effects of TCB on body weights, organ weights, and relative organ

weights of F-l males at 9 weeks of age.

 

 

 

 

Weightc Treatment
Control 3 ppm 30 ppm
(n=10) (n=10) (n=10)
BW (g) 26.8 :I: 2.2 23.5 i 4.3 24.7 :1: 2.5
LW (g) 1.25 :I: 0.16‘ 1.26 :l: 0.28‘ 1.48 :I: 0.20"
LW/BW X 100 5.05 :I: 0.50' 5.33 :I: 0.45'I 5.99 :i: 0.46"
TMW (mg) 37.5 :1: 13.1 32.5 :t 7.2 32.6 :I: 5.0
TMW/BW X 100 0.12 i- 0.02 0.12 :I: 0.01 0.12 i 0.01
TTW (mg) 194.3 :I: 9.9 192.7 :I: 10.8 197.6 i 24.0
TTW/BW X 100 0.79 i 0.06 0.85 :i: 0.22 0.80 :t: 0.08
ab Different superscripts represent significant differences within the same row

(P<0.05).

BW, LW, TMW, TTW represent body, liver, thymus, and testis weights,
respectively. Values represent mean organ weights or organ weights as
percentages of body weights :1: SD.

58

Table 14. Effects TCB on body weights, organ weights, and relative organ weights
F-l males at 19 weeks of age.

 

 

 

Weight° Treatment
Control 3 ppm 30 ppm
(n=6) (n=6) (n=6)
BW (g) 30.9 :1: 2.1 _ 29.4 :I: 3.5 31.3 :1: 2.0
LW (g) 1.32 i 0.14' 1.32 :I: 0.14‘ 1.67 :I: 0.16"
LW/BW X 100 4.27 :1: 0.31' 4.50 :I: 0.21‘l 5.35 :1: 0.31"
TMW (mg) 51.0 :1: 11.8 38.0 i 11.0 41.6 :1: 12.0
TMW/BW X 100 0.16 :I: 0.03. 0.13 :I: 0.03 0.14 21:0.04
TTW (mg) 219.4 :1; 17.5 223.5 :1: 17.8 232.4 :I: 14.6
TTW/BW X 100 0.71 :I: 0.05 0.77 :I: 0.08 0.74 :I: 0.04

 

ab Different superscripts represent significant differences within the same row
(P<0.05).

c BW, LW, TMW, TTW represent body, liver, thymus, and testis weights,
respectively. Values represent mean organ weights or organ weights as
percentages of body weights :1: SD.

Table 15.

Effects of TCB on reproductive performance of F-1 males at 7 weeks of

 

 

 

age.
Treatments

Control 3 ppm 30 PM“
No. Males Mated‘ 10 10 10
Fecundity (%) 90 80 90
Litter Size 8.9 :1: 2.4 9.8 :I: 2.4 9.3 :I: 2.3

(n=9) (n=8) (n=9)
Sex Ratio ((3/9) 2.0 :1; 2.3 1.4 :1; 0.8 1.3 :I: 1.0

(n=9) (n=8) (n=9)
4-Day Survival 100 100 100
Index" (n=9) (n=8) (n=9)
21—Day Survival 100 100 100
Index° (n=9) (n=8) (n=9)

 

One non-treated female was mated with one treated male for 10 days, from 7 :30

PM to 7:30 AM. TCB treatment was withheld during breeding.

of viable pups born].

Four-day survival index = 100 x [number of pups viable at day 4 of age]/ [number

Twenty-one-day survival index= 100 X [number of pups viable at day 21 of

age]/[number of pups retained at day 4 of age]. The litters in which all pups died
before day 4 of age were not included.

Table 16.

Effects of TCB on reproductive performance of F-l males at 17 weeks of

 

 

 

age.
Treatments

Control 3 ppm 30 ppm
No. Males Matedc 6 6 6
Fecundity (%) 67 100 100
Litter Size 8.8 i 1.3 9.0 :1: 0.9 8.8 i 1.2

(n=4) (n=6) (n=6)
Sex Ratio (cf/9) 1.4 i 1.0 2.4 :1: 2.8 2.1 :I: 2.9

(n=4) (n=6) (n=6)
4-Day Survival 94.7 :I: 6.1 100 100
Index‘1 (n=4) (n=6) (n=6)
21-Day Survival 100 96.3 :t 9.1 100
Index‘ (n=4) (n=6) (n=6)

 

ab Different superscripts represent significant differences within the same row

(P <0.05).

c One non-treated female was mated with 1 treated male for 10 days, from 7:30 PM

to 7:30 AM. TCB treatment was withheld during breeding.

d Four-day survival index = 100 x [number of pups viable at day 4 of age]/[number
of viable pups born].
e Twenty—one—day survival index= 100 x [number of pups viable at day 21 of

age]/ [number of pups retained at day 4 of age].

pups died before day 4 of age were not included.

The litters in which all

61
Table 17. Effects of TCB on reproductive performance of F—l males at 60 weeks of

 

 

 

 

age.
Treatments
Control 3 ppm 30 ppm
No. Males Matedc 5 4 5
Fecundity (%) 25 50 33
Litter Size 9.0 :1: 1.0 8.0 i 1.3 9.2 :t: 2.2
(n=3) (n=6) (n=5)
Sex Ratio (6/9) 0.74 :1; 0.13 1.54 :I: 0.63 1.23 :t 0.70
(n=3) (n=6) (n=5)
4-Day Survival Index“ 100‘ 75.8 :1; 28.3" 78.5 :1; 20.9"
(n=3) (n=6) (n=5)
21-Day Survival 100 100 100
Indexe (n=3) (n=6) (n=5)
ab Different superscripts represent significant differences within the same row
(P<0.05).

0 Three non-treated females were mated with 1 treated male for 90 days. TCB
treatment was withheld during breeding.

d Four-day survival index = 100 X [number of pups viable at day 4 of age]/[number
of viable pups born].

e Twenty-one—day survival index=100 X [number of pups viable at day 21 of
age]/[number of pups retained at day 4 of age]. The litters in which all
pups died before day 4 of age were not counted.

62
Body weights of offspring sired by treated 7- and 60-week-old F-l males were not

affected (Tables 18 and 20). However, body weights of offspring sired by the 30 ppm-
treated 17-week-old F-l males were higher than the control (Table 19).

At 9 and 19 weeks of age, epididymal sperm was collected from treated F-l
males. Sperm concentration, velocity, linearity, ALH displacement, and beat/cross
frequency were similar among all groups (Tables 21 and 22). Percentages of eggs
fertilized by the sperm from 30 ppm-treated males was lower than the percentages for
control group at 19 weeks of age, but not at 9 weeks of age (Tables 21 and 22). At 2.5
hours after sperm collection, decreased sperm motility was observed in 9-week-old 30
ppm-treated males. This was not observed at 0.25, 1.5, 3.5, and 4.5 hours after

collection (Figure 1).

63
Table 18. Effects TCB on body weights of offspring sired by F-l males at 7 weeks

 

 

 

of age.
Age" Treatment
Control 3 ppm 30 ppm
BW (3)
Day 4 2.4 i 0.3 (n=9) 2.4 :1: 0.3 (n=8) 2.3 :I: 0.2 (n=9)
Week 1 4.3 :1: 0.4 (n=9) 4.3 :1: 0.4 (n=8) 4.2 :1; 0.3 (n=9)
Week 2 8.2 i 0.9 (n=9) 7.9 :I: 0.9 (n=8) 7.9 :1: 0.5 (n=9)

Week 3 (d‘) 10.5 :I: 1.0 (n=46) 10.3 :I: 1.0 (n=42) 10.2 :1; 0.9 (n=47)
(9) 10.5 :I: 1.1 (n=23) 10.1 :t 1.0 (n=29) 10.0 :I: 1.1 (n=36)

a Body weights on day 4, week 1, and week 2 represent the means :1: SD of litters.
Values for week 3 are means :1: SD of individual weights.

64

 

 

 

Table 19. Effects of TCB on body weights of offspring sired by F—l males at 17
weeks of age.
Age" Treatments
Control 3 ppm 30 ppm
BW (3)

Day 4 2.0 :I: 0.3 (n=4) 2.1 :t 0.1 (n=6) 2.5 :1: 0.4 (n=6)
Week 1 3.7 :1: 0.4 (n=4) 4.0 :t 0.2 (n=6) 4.4 :I: 0.5 (n=6)
Week 2 7.6 :I: 0.7 (n=4) 7.8 :t: 0.9 (n=6) 8.0 :1; 0.7 (n=6)
Week 3 (d) 9.8 i 1.6" (n=17) 10.5 :I: 1.1‘" (n=30) 10.7 :1; 1.0" (n=27)

(9)

9.4 :I: 1.2" (n=16)

10.2 :1: 0.9" (n=22)

10.5 :I: 1.0b (n=25)

Body weights on day 4, week 1, and week 2 represent the means :1; SD of litters.

Values for week 3 are means :1: SD of individual weights.

65
Table 20. Effects of TCB on body weights of offspring sired by F-l males at 60

 

 

 

weeks of age.
Age" Treatments"
Control 3 ppm 30 1’1““
BW (3)
Week 1 4.0 :I: 0.3 (n=3) 3.5 :I: 0.6 (n=6) 3.5 :I: 1.0 (n=4)
Week 2 7.8 i 0.7 (n=3) 7.4 :1: 1.3 (n=6) 8.0 :I: 1.5 (n=4)

Week 3 (d) 10.2 :1: 1.2 (n=12) 10.3 :I: 1.3 (n=20) 10.5 :I: 1.6 (n=17)
(9) 10.2 :I: 1.2 (n=16) 10.0 :1: 1.9 (n=12) 9.7 :1: 1.1 (n=13)

a Three non-treated females were mated with 1 treated male for 90 days. TCB
treatment was withheld during breeding.

b Body weights on week 1 and week 2 represent the means :1: SD of litters. Values
for week 3 are means j: SD of individual weights.

66

 

 

 

 

Table 21. Effects of TCB on fertilization rates in vitra and sperm motion analysis
for F-l males at 9 weeks of age.
Parameters" Treatments
Control 3 ppm 30 ppm
(n=10) (n=10) (n=10)
Concentration
(x 10"/ml) 12.4 :1: 3.1 12.7 :1: 5.8 12.1 :1: 3.8
Motility (%)" 70.9 :1: 10.2 69.2 :1: 11.7 72.9 :1: 7.0
Velocity (pm/sec)c 246.9 :1: 59.9 210.8 :1: 92.9 260.2 i 67.3
Linearity“ 5.4 i 0.8 4.6 :1: 1.8 5.6 :1: 0.9
Amplitude of Lateral
Head Displacement (pm)° 7.3 i 1.9 5.9 :1; 2.8 7.6 :1: 3.4
Beat/Cross Freq. (Hz)f 12.7 :1: 1.3 11.2 :1: 4.3 12.6 :1: 1.6
Fertilization
In Vitra (%) 66.1 :1: 27.4 68.9 :1: 28.1 63.3 :1: 21.6
a In vitra fertilization and sperm motion analysis were performed 1.5 hours after
sperm collection.
b Percentage of sperm that travel more than 20 pm/sec.
c Average distance traveled by motile sperm.

(1 A calculation of the length of the straight line distance divided by the track
distance covered by sperm, multipled by 10. This is expressed on a scale of 0
to 10; 10 indicates a perfectly straight line and 0 indicates a circular track.

e A measure of the displacement of the sperm head from a computer-calculated
curval mean of its track.
f The number of beats (or crosses) per second. Every time the sperm cell crosses

the computer-calculated curval mean, the computer counts that crossing as one
beat.

67

 

 

 

Table 22. Effects of TCB on fertilization rates in vitra and sperm motion analysis
for F-l males at 19 weeks of age.
Parameters" Treatments
Control 3 ppm 30 ppm
(n=6) (n=6) (n=6)
Concentration
(x lO‘lml) 13.6 :1: 2.8 11.6 :1: 3.3 11.0 :1: 2.0
Motility (%)“ 63.5 :1: 9.5 60.6 :1: 13.2 66.5 i 10.3
Velocity (pm/sec)e 208.9 :I: 35.4 190.8 :1: 24.0 200.9 :1: 11.8
Linearity‘ 4.9 :1: 0.2 4.8 :t 1.2 4.7 :1: 0.5
Amplitude of Lateral
Head Displacement (pm)‘ 5.7 :1: 1.5 5.5 i 1.1 5.5 :l: 1.9
Beat/Cross Freq. (Hz)" 11.5 :1: 1.9 12.7 :1: 2.3 12.6 :1: 1.0
Fertilization
In Vitra (%) 57.9 :1: 12.6" 57.2 :1: 19.7" 46.7 :1: 11.5"

 

ab Different superscripts represent significant differences within the same row.

c In vitra fertilization and sperm motion analysis were performed 1.5 hours after
sperm collection.

d Percentage of sperm that travel more than 20 pm/sec.

e Average distance traveled by motile sperm.

f A calculation of the length of the straight line distance divided by the track
distance covered by sperm, multipled by 10. This is expressed on a scale of 0
to 10; 10 indicates a perfectly straight line and 0 indicates a circular track.

g A measure of the displacement of the sperm head from a computer-calculated
curval mean of its track.

h The number of beats (or crosses) per second. Every time the sperm cell crosses

the computer-calculated curval mean, the computer counts that crossing as one
beat.

Motility (%)

 

 

 

 

 

 

 

 

 

 

 

 

 

T j l j I

15 90 1 50 21 0 270
Time (min)

Figure l. Motility of sperm taken from 9-week-old
F-l males. Females (F-O) were treated with 0, 3,
and 30‘ ppm TCB 2 weeks prior to breeding.
Treatments continued through mating, gestation,
and lactation. Ofi‘spring (F-l) were given the
’ same diet as their dams received. Each value
reflects the mean of data from 7 mice. Different
superscripts represent significant differences

1 (P<0.05).

 

 

-o- Control
-a— 3 ppm
+30 ppm

 

 

DISCUSSION

Ritual

Results of this study demonstrated. that accumulation of TCB in liver and adipose
tissue is not arithmetically proportional to exposure concentrations. After 5 weeks of
exposure, concentrations of TCB equivalents in the liver, adipose tissue, uterus, and fetus
of mice in the 30 ppm group were only 3-6 times higher than concentrations in mice in
the 3 ppm group. Greater induction of hepatic microsomal monooxygenases has been
observed in pregnant rabbits treated with 10 mg Aroclor 1254/kg bw than in those treated
with 1.0 mg/kg (Villeneuve and Grant, 1971). Thus, it is possible that in the present
study, hepatic microsomal monooxygenases in the 30 ppm-treated mice were also induced
to a greater extent than in 3 ppm-treated mice, resulting in an increase in metabolism and
elimination of TCB in the high dose group.

After 5 weeks, the concentrations of TCB equivalents in the liver of pregnant
mice treated with 30 ppm TCB in the feed was 1.8 times higher than those in non-
pregnant mice. This may be due, in part, to an increase in the amount of fat in the liver
during pregnancy, causing a redistribution of TCB and its metabolites from adipose
tissues to liver. This hypothesis is also evident in the low concentrations of TCB
equivalents in the adipose tissue of pregnant mice, which was 2/3 of that in non-pregnant

mice. Increases in deposits of body fat cause dilution of TCB concentrations and may

69

70

also contribute to lower concentrations of TCB equivalents in pregnant mice than in the
non-pregnant mice. Nonetheless, when two groups of mice were exposed to a single
injection of 100 mg l“C-6-CB/kg bw, and one group was then mated while one group
remained virgin, radioactivity in the kidney, adipose tissue, and liver of pregnant mice
was higher than that in the virgins sacrificed concurrently (Vodicnik and Lech, 1980).
The differences in adipose tissue radioactivity in the two studies could be attributed to
different administration doses and different congeners. Induction of hepatic microsomal
ethoxycoumarin-O—deethylase (ECOD) by 4-CB was less in pregnant mice than in virgins
over a 4-day period (V odicnik, 1986). A slower rate of elimination of 4-CB equivalents
was observed in pregnant mice than in virgin mice.

Results in the present study also indicated that dosage affected distribution of TCB
in the liver and adipose tissue. The concentrations of TCB equivalents in the liver and
adipose tissue of the 3 ppm-treated mice were the same in pregnant and non-pregnant
mice. Any differences in body burden between pregnant and non-pregnant mice can not
be explained, since PCBs in the fetus only aeoount for approximately 0.003 % of maternal
body burden of pregnant rats (Takagi et al., 1976).

In this study, concentrations of TCB equivalents, calculated from radioactivity,
were similar in the uteri, placentas, and fetuses of treated mice. Darnerud et al. (1986)
reported that TCB metabolites, but no parent TCB, were detected in the fetus when
pregnant mice were gavaged with 25 mg TCB/kg bw on day 13 of gestation and
sacrificed on day 17 of gestation. Therefore, it is likely that the radioactivity detected

in the fetal tissue in the current study is from TCB metabolites.

71

Results from this study also point to redistribution of TCB equivalents to the
thymus from other tissues during lactation in the 30 ppm group. Although “C—TCB feed
was changed to TCB feed at parturition, radioactivity in the thymus increased during
lactation. However, only background radioactivity values were detected in the thymus
of the 3 ppm-treated mice. Although TCB redistribution into the thymus was observed
in the high dose group during lactation, the mechanisms by which this occurs is
unknown.

The redistribution of radioactivity in pregnant and lactating mice is likely to be
associated with the increase of very low density lipoprotein (VLDL), the primary carrier
of 6—CB in viva which occurs during gestation (Spindler-Vomachka and Vodicnik, 1984).
VLDL is a major substrate for mammary gland lipoprotein lipase which is elevated
during late gestation and lactation (Spindler-Vomachka and Vodicnik, 1984). More than
70% of circulating 6-CB was associated with VLDL during late gestation in female mice
treated with 6-CB two weeks prior to mating (Gallenberg and Vodicnik, 1987).
Continuous exposure to varying doses of TCB results in differences in TCB redistribution
between organs and tissues. In the 3 ppm-treated group, approximately 85% of the
concentrations of TCB equivalents disappeared from both liver and adipose tissue during
lactation. However, in the 30 ppm group, 96% disappeared from the liver and only 44%
disappeared from the adipose tissue. It is possible that circulating lipoproteins in the
blood were "saturated" with TCB equivalents in the 30 ppm group, which decreased the
redistribution of TCB from adipose tissue to liver. This phenomenon,"saturable

kinetics", was observed in the fetal compartment when dams were pretreated with TCB

72
(12.5-50 mg/kg BW) and followed 4 hours later by l‘C-TCB (10 mg/kg) (Darnerud et

al., 1986). Radioactivity in fetal brain, liver, thymus, skin, and muscle in the
pretreatment groups was less than half of that observed in the group without TCB
pretreatment (Darnerud et a1 . , 1986).

Different doses affect the distribution of TCB in the liver and adipose tissue of
offspring. In the offspring of the 30 ppm treatment group, concentrations of TCB
equivalents in the adipose tissue were 14.6 times higher than those in the liver.
However, concentrations of TCB equivalents in the adipose tissue were only 1.5 times
higher than those in the liver of 3 ppm-exposed offspring. In addition, although the pups
were nursed by the same dam in the 30 ppm group, large variations in concentrations of
TCB equivalents were observed. These results are due to large variations of offspring
body weights and fat deposits in the 30 ppm-exposed offspring. Variations of offspring
body weights were smaller in the 3 ppm group than those in the 30 ppm group.

This preliminary study indicates that accumulation of TCB equivalents in the
adipose tissue and liver of mice is not arithmetically proportional to exposure. This is
probably due to the greater induction of hepatic microsomal monooxygenases in 30 ppm-
treated mice than in the 3 ppm-treated mice, resulting in an increase in metabolism and
excretion of TCB. A similar accumulation of TCB equivalents between 3 ppm-treated
non-pregnant and pregnant mice suggests similar activities of hepatic microsomal
monooxygenases under differing physiological conditions. Different doses result in

variations of TCB redistribution during late gestation and lactation.

73
13m

Feed Intake and Body Weight

Based on feed intake, F-0 females in the 30 ppm treatment group consumed
approximately 7.5, 5.9, and 8.3 mg TCB/kg bw/day before breeding, from day 1 to day
16 of gestation, and during lactation, respectively. Body weights were not significantly
different among F-0 treatment groups before breeding and during gestation. During
lactation, body weights in the 30 ppm group were higher than those in the 3 ppm group.
However, body weights in the control group were not different from TCB treated groups.
As Yoshimura et al. (1979) reported, body weight gains in weanling rats over a 4-day
period were not affected by a single i.p. injection of 50 mg TCB/kg bw. In immature
male rats, ED50 for body weight loss caused by TCB was approximately 213 mg/kg bw
in a single i.p. injection (Leece et al., 1985). Reduced body weight gains were reported
by Sanders et al. (1974) in rats after 2 weeks of 62.5 or 1000 ppm of Aroclor 1254
exposure through the feed. These doses are much higher than the high dose tested in this
study which was 7.5 mg/kg bw/day prior to breeding. High feed intake observed in the
3 ppm-treated mice during lactation was in all likelihood a compensation response to the
low feed intake before breeding and during gestation.

Changes in body weights of offspring treated with PCBs perinatally were
dependent on exposure period, dose, and on the congener or commercial mixture. F-l
body weights at weaning in the 3 ppm group were numerically greater than those in the
30 ppm group. The increase in F-l weaning weights in the 3 ppm group could be the

result of increase in milk production of the dams which had high feed intake. As Gellert

74
and Wilson (1979) reported, no effects on body weights of female rat offspring were

observed when dams were gavaged with 30 mg Aroclors 1242 and 1260/kg bw daily on
days 14-20 of gestation. Offspring body weights, however, increased when dams were
treated with Aroclor 1221 (Gellert and Wilson, 1979). On the other hand, decreased
body weights of young mice were observed when parents were dosed with 10 mg Aroclor
1254/kg bw throughout gestation and lactation (Linzey, 1988), although these litters did
not differ in body weights at birth. Similar regimens using different PCBs and doses
produced diverse results. The relative growth rates of young male mice from dams
administered orally 0.005 mg 2,4’,5-trichlorobiphenyl or 6-CB/day, from day 5 of
gestation to day 22 of lactation, were higher than those of male offspring in the control
group (Orberg, 1978b). The differences in offspring body weights at weaning are mostly
diminished after maturity. For example, decreased body weight gains during lactation
followed by increased body weight gains after weaning were observed in rat offspring
when dams were treated on days 1, 3, 5, 7, and 9 postpartum with a dose of 32 or 64
mg Aroclor 1254/kg bw (Sager et al., 1991). Weight gains were similar among
treatment groups after those offspring reached 5 months of age.

In the present study, the body weights of offspring of treated F-l males are
dependent on the exposure period. No differences were observed in body weights of
offspring born to non-treated females bred with 7-week-old 30 ppm-treated F-l males.
However, body weights of offspring from non-treated females bred with 17-week-old 30
ppm-treated F-l males increased compared to control. These increases occurred at 3

weeks of age and were independent of the body weights of treated F-l males and dams

75

at weaning. Since only sperm from the treated males could have contributed to the
increases in body weights,TCB exposure must have altered some characteristic of pre-
ejaculated sperm.
Hepatotoxicity

Liver enlargement has been observed by others in TCB-treated rats (Yoshimura
et al. , 1979; Clarke et al., 1984). Increased numbers of peroxisomes and lipid droplets,
proliferated smooth endoplasmic reticulum, abnormal mitochondria (Harris and
Bradshaw, 1984; MacLellan et al., 1994), and distended cisternae of the rough
endoplasmic reticulum (Harris and Bradshaw, 1984) are the most overt signs of toxicity
in the livers of rats treated with TCB. Total liver lipid content, including cholesterol,
phospholipid and neutral lipid, increased in mature rats exposed to 50 or 500 ppm
Aroclor 1254 for 3 weeks (Garthoff et al., 1977). Total protein was also increased by
the 500 ppm treatment, while the relative total protein (mg/g wet weight) decreased. The
increases in lipid content were associated with increases in the activity of 3-hydroxy-3-
methylglutaryl coenzyme A reductase, the rate-limiting enzyme of cholesterol synthesis,
as well as increases in three NADPH-generating enzymes: malic enzyme, glucose-6-
phosphate dehydrogenase, and 6-phosphogluconate dehydrogenase (Kato and Yoshida,
1980; Hitomi et al., 1993). NADPH is essential for biosynthesis of cholesterol, lipid,
and several amino acids (Voet and Voet, 1990).

This study demonstrated that the effects of TCB on liver weights depends on the
physiological condition and age of the animal. Increases in liver weights were apparent

in F-0 pregnant and lactating mice, but not in non-pregnant mice. large variations in

76

liver weights of non-pregnant mice may contribute to the lack of statistical significance
of the data. Notably, F—l females treated with 30 ppm TCB showed increases in liver
weights at 5 and 6 weeks of age, but not in relative liver weights. At 7 weeks of age,
while there was an increase in liver weights and relative liver weights in the 30 ppm
group, the differences were not statistically significant.

Differences in exposure period and inducibility of hepatic microsomal enzymes
may have contributed to the differences in offspring liver weights before and after
puberty. Both liver weights and relative liver weights increased in the 30 ppm group in
F-l males at both 9 and 19 weeks of age, but not at 3 weeks of age. A significant
induction of uridine diphosphoglucuronic acid glucuronyltransferase (UDPGT), an
important microsomal glucuronide conjugation enzyme, was observed in the TCB treated
rats (3 mg/kg bw/day prenatally) at 3 weeks of age (Harris and Bradshaw, 1984).
However, this induction diminished after 3 weeks of age. In addition, a higher-fold
induction of hepatic microsomal monooxygenases above the control level was observed
in the weanlings than in the adults (Chen and DuBois, 1973). This high induction of
hepatic microsomal enzymes enhances biotransformation and excretion of TCB in males
at 3 weeks of age.

Thymic Atrophy

Thymic atrophy is a consistently observed effect of TCDD and PCBs in all
animals thus far examined (Faith and Moore, 1977 ; McConnell, 1989; Smialowicz et al. ,
1989). Microscopically, the reduction in size is reflected almost entirely as a loss of

cortical lymphocytes (McConnell et al., 1978). Although Clarke et al. (1984) reported

77
that 5 mg TCB/kg bw/day for 3 weeks caused thymic atrophy in young female rats, the

present study showed no such changes in F-0 female mice after 5 weeks of treatment.
The difference between the two observations may be explained by species variation or
the differences in maturity of the test animals. Rats in the study by Clarke et al.
weighed only half of their mature body weight, while the F-0 females in the present
study had reached their mature body weight. Decreases in thymus/body weight ratios
in the offspring of dams exposed to TCDD during gestation and lactation were more
apparent than those only exposed to TCDD postnatally (Faith and Moore, 1977). At 5,
6 and 7 weeks of age, F-l females treated with 30 ppm TCB showed decreases in thymus
weights, however, TCB did not induce thymic atrophy in F-l males at 3, 9, and 19
weeks of age in the present study. Gender appears to be the primary factor contributing
to the differences of thymic effects between the F-1 males and females.
Female Reproductive Toxicity

In our study, fecundity in 30 ppm—exposed females (F-0) was reduced by 40%,
while the fecundity in treated F-l females was the same as in the control group. The
decrease in the F-0 females could be the result of many reproductive dysfunctions from
mating to fertilization, to embryo and fetal development. Existing information points to
a decrease in mating of PCB-treated females. In rats, the percent of females mated,
judged by vaginal plugs, decreased by 40% when those females were dosed with 7.5 mg
Aroclor 1242/kg bw/day for 36 weeks (Jonsson et al., 1976). Similar results have been
reported by Brezner et al. (1984) who stated that sexual receptivity was decreased by

20% in rats exposed orally to 30 mg Aroclor 1254/kg bw for 1 month. This sexual

78

receptivity was determined by the presence of sperm in the vaginal smear on the day
after mating. Small sample sizes of F-l females may, in part, have prevented the
detection of differences in fecundity among treatment groups.

Mortality of offspring mostly occurred within 4 days postpartum. Litter sizes of
treated F-0 females and F—l offspring were similar among all 3 groups at parturition.
These results reflect the observations by Rands et al. (1982) that the mortality during
parturition was the same in TCB treated rats as in controls. Four-day and 21-day
survival indices only decreased in the 30 ppm group in F-0 females, while all pups born
to TCB treated F-l females of both dose groups died within 4 days of age. In rats, a
high incidence of neonatal mortality was also observed after prenatal exposure to TCB
at 3 mg/kg bw/day from day 6 through day 18 of gestation (Rands et al., 1982; Harris
and Bradshaw, 1984). The reasons for the mortality of the progeny are likely associated
with the behavioral changes of dams (Pantaleoni et al., 1988), or direct toxicity caused
by TCB and its metabolites (Darnerud et al., 1986), or some combination of these
factors. In addition, there is a low percentage of fat in the young which is one of the
major tissues where PCBs accumulate. This heightens the amount of absorbed PCBs
distributed to target organs and tissues in the body.

Postnatal mortality is also related to toxicity of TCB and its metabolites in
embryos and fetuses. Accumulations of 2-hydroxy-TCB and methylsulphonyl-
tetrachlorobiphenyl, two TCB metabolites, in fetal tissue and uterine fluid were observed
after a single intravenous injection of 3.5 mg TCB/kg bw in pregnant mice (Darnerud

et al., 1986). Hemorrhage in intestinal mucosa of fetuses was observed when pregnant

79
rats were treated with 3 mg TCB/kg bw on days 6—18 of gestation (Rands et al., 1982).

Embryonic death in mice increased after a single oral dose of 25 mg TCB/kg bw on day
11 of gestation (d’Argy et al. , 1987). Since the parent TCB was not found in fetal tissue
(Darnerud et al., 1986), the metabolites of TCB play a very important role in
embryotoxicity.

The in vitra fertilization assay showed that egg fertilizing ability decreased in F-l
females exposed to 3 ppm and 30 ppm TCB. The poor egg quality was also
demonstrated by the high percentage of degenerated eggs after collection and culture.
Similar results were observed in a study in which sperm and eggs were incubated in
medium containing PCBs. After 1.5 hours of incubation, sperm were added to the petri
dish where superovulated eggs were treated with the following PCBs: TCB, Aroclor
1221, and Aroclor 1268. Fertilization rates decreased at a concentration of 1 pg/ml for
these PCBs tested (Kholkute et al., 1994).

In the present study, it is suggested that impaired egg quality is associated with
decreased estrogen levels in the follicles. Preovulatory follicular growth is dependent on
the interaction of estradiol, FSH, and LH (Richards, 1980). The ability of healthy eggs
with germinal vesicles to resume meiosis in vitra is highly correlated with the
concentration of estradiol in antral fluid (McNatty et al., 1979). Theca interna, an
androgen biosynthesizing site in the ovary in all species, is differentiated from stroma
cells. The synthesized androgens are then transferred to the basement membrane of the
follicle and on to the granulosas where they are aromatized to estrogens (Peters, 1979;

Erickson et al. , 1985; Ryan, 1988). Characteristic changes in the ovarian stroma cells,

80

including a spindling of ovarian cells accompanied by looseness of cellular arrangement
and reduction of follicle numbers, were observed in 150 ppm Aroclor 1242-treated rats
(Jonsson et a1 . , 1976). The stroma changes may affect differentiation of theca interna
and decrease androgen and estrogen biosynthesis. In addition, metabolism and
elimination of estrogen can be increased by hepatic microsomal monooxygenases induced
by PCBs or other chemicals (e.g., phenobarbital) (Risebrough and Brodine, 1970; Chen
et al. , 1993).

Environmental estrogens, including PCBs, have received much attention lately
(Hileman, 1994; Stone, 1994). Several commercial PCB mixtures and congeners related
to estrogenic activity have been investigated (Bitman and Cecil, 1970; Ecobichon and
MacKenzie, 1974; Gellert, 1978; Korach et al., 1987). In general, the lesser chlorinated
PCBs have higher estrogenic activity than the more highly chlorinated PCBs, based on
the glycogen response of the immature rat uterus. It has been suggested that uterotropic
activity of PCBs is dependent upon the formation of metabolites (Ecobichon and
MacKenzie, 1974). The lesser chlorinated PCBs can be more easily biotransformed to
hydroxylated metabolites than the more highly chlorinated PCBs. In addition, for the
hydroxylated congeners tested, those compounds with strong affinities to estrogen
receptors possess either single or multiple artha-chlorine substitutions (Korach et al. ,
1987). No artha-substituted TCB metabolites were detected in rats (Yoshimura and
Yamamoto, 1974; Yoshimura et al., 1987; Koga et al., 1989) and mice (Darnerud et al. ,
1986; Klasson Wehler et al., 1989). Although TCB metabolites show relatively little

estrogenic activity, the role they play in reproductive and developmental toxicity can not

8 1
be ignored.

In the present study, eggs superovulated by exogenous PMSG and HCG were
similar in all groups. Similar results were observed by Brezner et al. (1984) in that the
number of ovulations was not affected in mature rats exposed to Aroclor 1254 at a dose
of 10 mg/kg bw/day for 1 month. In contrast, both decreases and increases in germ cells
have been reported in TCB-exposed mouse offspring (Ronnback, 1991; Ronnback and
de Rooij, 1994). The different results appear to be dependent on exposure regimens in
the 2 studies. F-O females in the Ronnback (1991) study were only injected
intraperitoneally on day 13 of gestation with 1.5-15 mg TCB/kg bw. However, in the
study by Ronnback and de Rooij (1994), F0 females were gavaged with TCB at 9 or 15
mg TCB/kg bw weekly for 2 weeks prior to mating. The treatment continued through
gestation and lactation for a total of 7 doses. The increases in germ cells in F—l females
were associated with the inhibition of atresia (Ronnback and de Rooij, 1994).

Male Reproductive Toxicity

In this study, testis weights increased in F-l males at 3 weeks of age, but not at
9 and 19 weeks of age. Similar results were also observed in albino mice (Johansson,
1987). Increases in relative testis weights (mg/ 100 g bw) were noted in mice treated
perinatally with 6-CB at 200 or 400 mg/kg bw doses, but not in mice treated in the
pubertal period (Johansson, 1987). Testis weights were not affected in studies where
adult rats were exposed to 500 ppm Aroclor 1254 (Garthoff et al. , 1977) or in adult mice
treated with 200 or 1000 ppm Aroclor 1254 (Sanders et al., 1974, 1977). In contrast,

increased testis weights of offspring were observed in pubertal rats when dams were

82
treated with 32 mg Aroclor 1254/kg bw on days 1, 3, 5, 7, and 9 of lactation (Sager et

al. , 1983).

Existing knowledge suggests two possible mechanisms for these effects. One is
the development of a blood-testis barrier which may play a role in decreasing testis
toxicity in adults. The blood-testis barrier develops at the time of puberty and just prior
to the onset of spermatogenesis (Sever and Hessol, 1985; Thomas, 1991). In rats, the
barriers between pairs of Sertoli cells form at approximately 18 days of age (Gilula et
al. , 1976). A second possible mechanism is the induction of hepatic microsomal
enzymes. Induction of N-demethylase in treated weanling rats was 3.5-4.2 times that of
the control group, while treated adults had induction of 2.4-2.8 times that of control
group. This pattern also exists with 0-ethyl 0-(4-nitrophenyl)phenyl phosphonothioate
(EPN) detoxification activity (Chen and DuBois, 1973). Uridine diphosphoglucuronic
acid glucuronyltransferase (UDPGT), one of the major enzymes in the second phase of
biotransformation, converts both exogenous and endogenous compounds to polar and
water-soluble compounds (Sipes and Gandolfi, 1991). When pregnant rats were orally
administered 3 mg TCB/kg bw daily on days 8 - 18 of gestation, the hepatic UDPGT
activity in the offspring increased at 3 weeks of age, but diminished by 8 weeks of age
(Lucier and McDaniel, 1979). The increased induction of hepatic enzymes may enhance
the metabolism of testosterone in weanlings. The enlargement of the testis may have
been compensation for the enhanced testosterone metabolism by PCB-induced hepatic
microsomal enzymes (Risebrough and Brodine, 1970; Haake-McMillan and Safe, 1991).

Testicular toxicity may result in impairment of spermatogenesis. In this study,

83

epididymal sperm concentration was similar among the treatment groups at both 9 and
19 weeks of age. To the contrary, Sanders et al. (1977) reported that 200 ppm Aroclor
1254 in the feed of adult albino male mice significantly reduced the number of sperm
cells per milligram of testis. A similar effect of Aroclor 1254 was observed in adult
white-footed male mice (Sanders and Kirkpatrick, 1975). Exposure to 400 ppm Aroclor
1254 for 2 weeks reduced the number of sperm per testis. There was no effect on the
number of sperm cells at 100 or 200 ppm (Sanders and Kirkpatrick, 1975). Compared
with those observations, the doses which affected sperm production were much higher
than the high dose in the present study. The sites used to measure the number of sperm
may have led to the diversity of results. The authors measured the number of sperm in
the testis (Sanders and Kirkpatrick, 1975; Sanders et al. , 1977), however, in the present
study, the epididymal sperm concentration was measured. Although enlarged testes were
observed at 3 weeks of age, epididymal sperm concentration was not affected when mice
were at 9 and 19 weeks.

Decreases in egg fertilizing ability in vitra were observed with sperm from 30
ppm-treated F-l males at 19 weeks of age, but not at 9 weeks of age. In another study,
a decrease in the normal number of embryos was observed as a result of the mating of
non-treated female rats with treated males dosed with 8 mg Aroclor 1254/kg bw on days
1, 3, 5, 7, and 9 of lactation (Sager et al., 1987, 1991). Based on the existing
information, there are two possible mechanisms associated with this decreased sperm
quality. The first one is testicular oxidative stress. Two of the antioxidant enzymes,

superoxide dismutase (SOD) and catalase, decreased in mature rats after a single i.p.

34
injection of 100 mg Clophen A 50/kg bw (Peltola et al., 1994). In addition, testicular

mRNAs for NADPH-generating enzymes decreased by 23-32% in rats fed 200 ppm
Aroclor 1254 for 3 days (Hitomi et al., 1993). NADPH, generated primarily by the
reactions catalyzed by malic enzyme (ME), glucose-6-phosphate dehydrogenase (G6PD),
and 6-phosphogluconate dehydrogenase (6PGD) (V oet and Voet, 1990), is necessary for
the reduction of glutathione disulfide to glutathione. Secondly, administration of TCB
throughout gestation and lactation is likely to disrupt hormonal balance during postnatal
development, thus affecting spermatogenesis. Decreases in plasma testosterone were
observed in 5- and lO-week-old male offspring of dams gavaged with 3 mg TCB/kg
bw/day on day 6 through 18 of gestation (Vincent et al., 1992).

Stem cells differentiate into type A spermatogonia during early postpartum. The
number of type A spermatogonia continues to increase until 35 days of age, and
decreases thereafter in mice (Foster, 1989; Vergouwen et al., 1993). Rapidly dividing
and differentiating spermatogonia are the cells most susceptible to chemical toxicity
during spermatogenesis (Meistrich, 1986). Sensitivity of stem cells to toxic materials in
this critical period explains, in part, the affected sperm quality in this study and in those
studies by Sager et al. (1987 and 1991). Exposure during the critical postnatal period
(approximately one week) may affect subsequent reproductive performance. In the Sager
studies, lactating dams were treated orally on days 1, 3, 5, 7, and 9 postpartum.
Reduced incidences of implantation were observed in non-treated females mated with
treated males. The lowest observable adverse effect level was the maternal oral dose of

8 mg Aroclor 1254/kg bw every 2 days during days 1-9 of lactation. However, no effect

85
on fertility was observed by Kihlstrom et al. (1975) when male mice offspring exposed

to PCBs during suckling were mated to non-treated females. In that study, those
offspring were exposed to PCBs only from dams injected with 50 mg Clophen A 60/kg
bw subcutaneously on the day of parturition and repeated weekly for three successive
weeks (Kihlstrom et al., 1975). The frequency of administration during the critical
postnatal period in the Sager studies is higher than that in the study by Kihlstrom et al.
(1975). This may account for their different results.

Parameters assessed by sperm motion analysis were similar among all treatment
groups, except for the decreased motility observed in the 30 ppm-treated males at 2.5
hours after collection. These parameters included velocity, linearity, ALH displacement,
beat/cross frequency, and motility. The decreased sperm motility was probably
associated with inhibition of ATPase inhibition (Ia Rocca and Carlson, 1979). A
significant decrease in total ATPase activity of liver, kidney, and brain tissue was
observed in rats treated with 25 mg Aroclor 1242/kg bw for 7 days (Ia Rocca and
Carlson, 1979). It is well established that ATPase causes the hydrolysis of ATP to ADP
with the release of energy. This is critical for sperm motility. Fecundity and litter sizes
were the same in non-treated females bred with treated F-l males at 7 and 17 weeks of
age. The 21-day survival index was decreased in the offspring of treated F-l males at
60 weeks of age in the 3 ppm and 30 ppm groups. Sperm from the treated males was
the only factor resulting in decreased survival of offspring. It is, therefore, hypothesized
that TCB exposure must have altered some characteristic of sperm. This conclusion

should be taken conservatively due to the small sample sizes in this observation. Only

86

1-2 males in each treatment group (4-5 males/group) at 60 weeks of age were fertile after
3 months of breeding with non-treated females.
Sexual Differentiation

Androgenic deficiency caused by exposure to TCDD and PCBs during prenatal
and early postnatal periods can impair male reproductive function by disrupting the
development of sex organs and impairing sexual differentiation of the central nervous
system. Hormones regulate sexual differentiation in the critical periods during both
prenatal and early postnatal stages (George and Wilson, 1988; Hadley 1992).
Testosterone and DHT are responsible for the differentiation of male internal genital
ducts and adult secondary sex characteristics, and external genitalia, respectively
(Hadley, 1992). When testosterone in the brain is transformed into estradiol by
aromatase, the estradiol controls the differentiation of male hypothalamus (Hadley 1992).

Sex ratios were similar throughout this study in both treated males and females.
In contrast, the sex ratios (6/ 9) of neonates decreased when pregnant rats were treated
by gavage with a single dose of 3 mg TCB/kg bw from day 6 to day 18 of gestation (the
exact day was not reported) (Simmons et al. , 1984). The high dose used in the present
study was approximately 7.5 and 5.9 mg TCB/kg bw/day before breeding and from day
1 to day 16 of gestation, respectively. It is likely that the induction of hepatic
microsomal monooxygenases biotransformed the highly toxic TCB into less toxic
metabolites when dams were treated continuously in the present study. However, in the
study by Simmons et al., TCB was administered during the critical period when

organogenesis occurred. Large variations of sex ratios were observed in all groups

87

which made it difficult to establish a statistical difference in sex ratios among treatment
groups. A decrease in testosterone concentrations was observed in the fetus on days 18
through 21, and shortly after birth when dams were treated with 1.0 pg TCDD/kg bw
on day 15 of gestation (Mably et al., 1992). Anogenital distances, measured 1-5 days
after parturition, were shorter in the TCDD exposed neonates than those in the controls.
Thus, TCDD and PCBs may cause feminization by disrupting the hormonal balance,

especially that of testosterone and DHT.

CONCLUSIONS

Liver weights increased in the 30 ppm-treated lactating F-0 mice and in all F-l
mice except for the weanling males. Exposure period and induction of hepatic
microsomal enzymes may explain the differences of liver weights between F-l male
weanlings and adults. Thymus weights only decreased in the F—1 females between 5 and
7 weeks of age. In this study, thymic effects were shown to be age and sex dependent.

Fecundity only decreased in the 30 ppm-treated F-0 group, but not in the F-1
mice. Unaffected fecundity in F-l mice can be attributed to small sample sizes.
Survival of the offspring of treated F-0 and F-l females decreased. This increased
mortality is probably associated with either the behavioral changes of dams or direct
toxicity caused by TCB and its metabolites, or any combination of the two factors.
Increased testis weights were only observed in the F-1 weanlings, not in the pubertal and
mature mice. Development of blood-testis junction and hepatic microsomal enzymes may
contribute to this result. Decreased gamete fertilizing ability was observed in both
treated F-l males and females.

Reproductive performance and gamete integrity were impaired by TCB exposure.
Since testosterone, DHT, and estrogen are critical for organogenesis and gametogenesis,
these effects are in all likelihood predominantly associated with the hormonal imbalance

caused by TCB treatment.

88

APPENDIX

APPENDIX I

Appendix I. Specific activity of l‘C-TCB in feed

1 kg of 3 ppm TCB fgog
Specific activity of l‘C-TCB
= “C-TCB/total TCB

= 9.52 pCi/3 mg TCB

= 3.17 pCi/mg TCB

1 kg of 30 pom TCB foeo

Specific activity of l‘C-TCB= I‘C-TCB/total TCB
= 9.52 pCi/30 mg TCB

= 0.317 pCi/mg TCB

89

APPENDIX II

Appendix II. Calculation of concentrations of TCB equivalents in the sample.

DPM

= CPM/efficiency of liquid scintillation counter for 1“C
= CPM/0.90

Radioactivity (pCi/g)
= 2{[(DPM-background DPM)/(2.22 X 10")]/sample weight}

Concentrations of TCB equivalents
= Radioactivity/specific activity of l‘C-TCB in feed

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