United States
Environmental Protection
Agency
Office of Research and
Development
Washington, DC 20460
EPA/600/AP-92/0016
August 1992
Workshop Review Draft
Chapter 5.
Reproductive and
Developmental
Toxicity
Review
Draft
(Do Not
Cite or
Quote)
600AP92001<
Notice
This document is a preliminary draft. It has not been formally released by EPA ancI should not
at this stage be construed to represent Agency policy. It is bemg circulated for comment on
its technical accuracy and policy implications.
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DRAKI- EPA/600/AP-92/001C
DO NOT QUOTE OR CITE August 1992
Workshop Review Draft
Chapter 5. Reproductive and Developmental Toxicity
Health Assessment for
2,3,7,8-Tetrachlorodibenzo-p-dioxin (TCDD)
and Related Compounds
NOTICE
THIS DOCUMENT IS A PRELIMINARY DRAFT. It has not been formally released by the U.S.
Environmental Protection Agency and should not at this stage be construed to represent Agency
policy. It is being circulated for comment on its technical accuracy and policy implications.
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, D.C.
Printed on Recycled Paper
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DISCLAIMER
This document is a draft for review purposes only and does not constitute Agency policy.
Mention of trade names or commercial products does not constitute endorsement or recommendation
for use.
Please note that this chapter is a preliminary draft and as such represents work
m progress. The chapter is intended to be the basis for review and discussion at
a peer-review workshop. It will be revised subsequent to the workshop as
suggestions and contributions from the scientific community are incorporated
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CONTENTS
......... iv
Tables [[[
......... v
List of Abbreviations .............................................
........ xii
Authors and Contributors ..........................................
5. REPRODUCTIVE AND DEVELOPMENTAL TOXICITY ......................... 5"1
5.1. INTRODUCTION .................................................. 5"1
5.2. REPRODUCTIVE TOXICITY .......................................... 5'2
5.2.1. Female [[[ 5'2
5.2.2. Male [[[
5.3. DEVELOPMENTAL TOXICITY ....................................... 5'14
5.3.1. Death/Growth/Clinical Signs ...................................... 5-15
5.3.2. Structural Malformations ......................................... £*
5.3.3. Postnatal Effects .............................................. ^
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LIST OF TABLES
5-1 Relationship Between Maternal Toxicity and Fetal Lethality
in Laboratory Mammals Exposed to TCDD During Gestation ................. 5.22
5-2 Developmental Toxicity Following Gestational Exposure to
2,3,7,8-TCDD ..... ^
................................................. 5-27
5-3 TCDD Responsiveness of Palatal Shelves From the Mouse,
Rat and Human in Organ Culture ...............
5-4 Apparent Ah Receptor Binding Affinity and Relative Teratogenic
Potency of Halogenated Aromatic Hydrocarbon Congeners .................. 5.43
5-5 Effects of In Utero and Lactational TCDD Exposure on Indices
of Androgenic Status .....................
5-6 Effects of In Utero and Lactational TCDD Exposure on Indices
of Spermatogenic Function and Reproductive Capability ................ 5.52
5-7 Effects of In Utero and Lactational TCDD Exposure on Indices
of Sexual Behavior and Regulation of LH Secretion in Adulthood ................ 5.57
IV
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ACTH
Ah
AHH
ALT
AST
BDD
BDF
BCF
BGG
bw
cAMP
CDD
cDNA
CDF
CNS
CTL
DCDD
DHT
DMBA
DMSO
DNA
ORE
LIST OF ABBREVIATIONS
Adrenocorticotrophic hormone
Aryl hydrocarbon
Aryl hydrocarbon hydroxylase
L-alanine aminotransferase
L-asparate aminotransferase
Brominated dibenzo-p-dioxin
Brominated dibenzofuran
Bioconcentration factor
Bovine gamma globulin
Body weight
Cyclic 3,5-adenosine monophosphate
Chlorinated dibenzo-p-dioxin
Complementary DNA
Chlorinated dibenzofuran
Central nervous system
Cytotoxic T lymphocyte
2,7-Dichlorodibenzo-p-dioxin
Sa-Dihydrotestosterone
Dimethylbenzanthracene
Dimethyl sulfoxide
Deoxyribonucleic acid
Dioxin-responsive enhancers
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LIST OF ABBREVIATIONS (cont.)
DTG
DTK
ECOD
EGF
EGFR
ER
EROD
EOF
FSH
GC-ECD
GC/MS
GGT
GnRH
GST
HVH
HAH
HCDD
HDL
HxCB
HpCDD
Delayed type hypersensitivity
Delayed-type hypersensitivity
Dose effective for 50% of recipients
7-Ethoxycoumarin-O-deethylase
Epidermal growth factor
Epidermal growth factor receptor
Estrogen receptor
7-Ethoxyresurofin 0-deethylase
Enzyme altered foci
Follicle-stimulating hormone
Gas chromatograph-electron capture detection
Gas chromatograph/mass spectrometer
Gamma glutamyl transpeptidase
Gonadotropin-releasing hormone
Glutathione-S-transferase
Graft versus host
Halogenated aromatic hydrocarbons
Hexachlorodibenzo-p-dioxin
High density lipoprotein
Hexachlorobiphenyl
Heptachlorinated dibenzo-p-dioxin
VI
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LIST OF ABBREVIATIONS (cont.)
HpCDF
HPLC
HRGC/HRMS
HxCDD
HxCDF
Heptachlorinated dibenzofuran
High performance liquid chromatography
High resolution gas chromatography/high resolution mass spectrometry
Hexachlorinated dibenzo-p-dioxin
Hexachlorinated dibenzofuran
I-TEF
LH
LDL
LPL
LOAEL
LOEL
MCDF
MFO
mRNA
MNNG
NADP
NADPH
NK
NOAEL
International TCDD-toxic-equivalency
Dose lethal to 50% of recipients (and all other subscripter dose levels)
Luteinizing hormone
Low density liproprotein
Lipoprotein lipase activity
Lowest-observable-adverse-effect level
Lowest-observed-effect level
6-Methyl-l,3,8-trichlorodibenzoruran
Mixed function oxidase
Messenger RNA
W-methyl-W-nitrosoguanidine
Nicotinamide adenine dinucleotide phosphate
Nicotinamide adenine dinucleotide phosphate (reduced form)
Natural killer
No-observable-adverse-effect level
vn
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NOEL
OCDD
OCDF
PAH
PB-Pk
PCB
OVX
PEL
PGQ
PeCDD
PeCDF
PEPCK
PGT
PHA
PWM
ppm
ppq
ppt
RNA
SAR
SCOT
LIST OF ABBREVIATIONS (cont.)
No-observed-effect level
Octachlorodibenzo-p-dioxin
Octachlorodibenzofuran
Polyaromatic hydrocarbon
Physiologically based pharmacokinetic
Polychlorinated biphenyl
Ovariectomized
Peripheral blood lymphocytes
Quaterphenyl
Pentachlorinated dibenzo-p-dioxin
Pentachlorinated dibenzo-p-dioxin
Phosphopenol pyruvate carboxykinase
Placental glutathione transferase
Phytohemagglutinin
Pokeweed mitogen
Parts per million
Parts per trillion
Ribonucleic acid
Structure-activity relationships
Serum glutamic oxaloacetic transaminase
viu
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LIST OF ABBREVIATIONS (cont.)
SOFT
SRBC
t*
TCAOB
TCB
TCDD
TEF
TGF
tPA
TNF
TNP-LPS
TSH
TTR
UDPGT
URO-D
VLDL
v/v
w/w
Serum glutamic pyruvic transaminase
Sheep erythrocytes (red blood cells)
Half-time
Tetrachloroazoxybenzene
Tetrachlorobiphenyl
Tetrachlorodibenzo-p-dioxin
Toxic equivalency factors
Thyroid growth factor
Tissue plasminogen activator
Tumor necrosis factor
lipopolysaccharide
Thyroid stimulating hormone
Transthyretrin
UDP-glucuronosyltransferases
Uroporphyrinogen decarboxylase
Very low density lipoprotein
Volume per volume
Weight by weight
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AUTHORS AND CONTRIBUTORS
The Office of Health and Environmental Assessment (OHEA) within the Office of Research
and Development was responsible for the preparation of this chapter. The chapter was prepared
through Syracuse Research Corporation under EPA Contract No. 68-CO-0043, Task 20, with Carol
Haynes, Environmental Criteria and Assessment Office in Cincinnati, OH, serving as Project Officer.
During the preparation of this chapter, EPA staff scientists provided reviews of the drafts as
well as coordinating internal and external reviews.
AUTHORS
Richard Peterson
School of Pharmacy
University of Wisconsin
Madison, WI
EPA CHAPTER MANAGER
Gary Kimmel
Office of Health and Environmental Assessment
Washington, DC
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5. REPRODUCTIVE AND DEVELOPMENTAL TOXICITY
5.1. INTRODUCTION
2,3,7,8-TCDD is one of 75 possible CDD congeners. It is one of the most
potent of the CDDs, BDDs, CDFs, BDFs, PCBs, PBBs, and as such serves as the
prototype congener for investigating the toxicity elicited by these classes of
chemicals. Reproductive and developmental toxicity is generally believed to be
caused by the parent compound. There is no evidence that TCDD metabolites are
involved. The toxic potency of TCDD is due to the number and position of
chlorine substitutions on the dibenzo-p-dioxin molecule. CDD congeners with
decreased lateral (2,3,7 and 8) or increased nonlateral chlorine and bromine
substituents are less potent than TCDD (Safe, 1990); however, most of these
congeners will produce toxicity and the pattern of responses within animals of
the same species, strain, sex and age will generally be similar to that of TCDD
(McConnell and Moore, 1979; Poland and Knutson, 1982). PCB congeners with zero
or one ortho chlorines, two para chlorines and at least two meta chlorines can
assume a coplanar conformation sterically similar to TCDD and also produce a
pattern of toxic responses similar to that of TCDD. In contrast, PCB congeners
with two or more ortho chlorines cannot assume a coplanar conformation and do not
resemble TCDD in toxicity (Poland and Knutson, 1982; Safe, 1990).
CDD and CDF congeners chlorinated in the lateral positions, as compared with
those lacking chlorines in the 2,3,7, and 8 positions, are preferentially
bioaccumulated by fish, reptiles, birds, and mammals (Stalling et al., 1983; Cook
et al., 1991; U.S. EPA, 1991). Furthermore, coplanar PCBs and/or monoortho
chlorine-substituted analogs of the coplanar PCBs bioaccumulate in fish,
wildlife, and humans (Tanabe, 1988; Kannan et al., 1988; Mac et al., 1988; Kubiak
et al., 1989; Smith et al., 1990). This is of concern because combined effects
of the lateral-substituted CDD, BDD, CDF, BDF, PCB, and PBB congeners acting
through an Ah receptor mechanism have the potential of decreasing feral fish and
wildlife populations secondary to developmental and reproductive toxicity
(Gilbertson, 1989; Walker and Peterson, 1991; Walker et al., 1991; Cook et al.,
1991). Humans are not exempt from the reproductive and developmental effects of
complex halogenated aromatic hydrocarbon mixtures. Such mixtures which contain
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both TCDD-like congeners and nonTCDD-like congeners have been implicated in
causing reproductive and developmental toxicity in the Yusho and Yu-Cheng
poisoning incidents in Japan and Taiwan (Kuratsune, 1989; Hsu et al., 1985;
Rogan, 1989). Thus, exposure to TCDD-like congeners is a health concern for
humans as well as for domestic animals, fish and wildlife, although the relative
contributions of TCDD- and nonTCDD-like congeners are not known in some exposure
situations.
A mechanism of action which CDD, BDD, CDF, BDF, PCB and PBB congeners
substituted in the lateral positions have in common is that they bind to the Ah
receptor which then binds to a translocating protein that carries the activated
TCDD receptor complex into the nucleus. These activated TCDD receptor complexes
bind to specific sequences of DNA referred to as dioxin-responsive enhancers
(DREs) resulting in alterations in gene transcription. There is evidence that
this Ah receptor mechanism, explained in detail in an earlier chapter, is
involved in the antiestrogenic action of TCDD and in its ability to produce
structural malformations in mice. However, its role in producing other signs of
reproductive and developmental toxicity is less firmly established.
5.2. REPRODUCTIVE TOXICITY
5.2.1. Female
5.2.1.1. REPRODUCTIVE FUNCTION/FERTILITY — TCDD and its approximate
isostereomers have been shown to affect female reproductive end points in a
variety of animal studies. Among the effects reported are reduced fertility,
reduced litter size, and effects on the female gonads and menstrual/estrous
cycle. These studies are reviewed below. other TCDD effects on pregnancy
maintenance, embryo/fetotoxicity, and postnatal development are covered in the
Developmental Toxicity section of this chapter.
The study by Murray et al. (1979) employed a multi-generation approach,
examining the reproductive effects of exposure of male and female rats over three
generations to relatively low levels of TCDD (0, 0.001, 0.01 and 0.1 ^g/kg
bw/day). There was variation in the fertility index in both the control and the
exposed groups, and a lower than desirable number of impregnated animals in the
exposed groups. Even so the results showed exposure-related effects on
fertility, an increased time between first cohabitation and delivery, and a
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decrease in litter size. The effects on fertility and litter size were observed
at 0.1 pg/kg/day in the FQ generation and at 0.01 ^g/kg/day in the Fj and F2
generations. Additionally, in a 13-week exposure to 1-2 ^g/kg/day of TCDD in
nonpregnant female rats, Kociba et al. (1976) reported anovulation and signs of
ovarian dysfunction, as well as suppression of the estrous cycle. However, at
exposures of 0.001-0.01 /ug/kg/day in a 2-year study, Kociba et al. (1978)
reported no effects on the female reproductive system.
Allen and his colleagues reported on the effects of TCDD on reproduction in
the monkey (Allen et al., 1977; Allen et al., 1979; Barsotti et al., 1979;
Schantz et al., 1979). In a series of studies, female rhesus monkeys were fed
50 or 500 ppt TCDD for <9 months. Females exposed to 500 ppt showed obvious
clinical signs of TCDD toxicity and lost weight throughout the study. Five of
the eight monkeys died within 1 year after exposure was initiated. Following 7
months of exposure to 500 ppt TCDD, seven of the eight females were bred to
unexposed males. The remaining monkey showed such severe signs of TCDD toxicity
that she was not bred due to her debilitated state. Of the seven females that
were evaluated for their reproductive capabilities only three were able to
conceive and of these, only one was able to carry her infant to term (Barsotti
et al., 1979). When females exposed to 50 ppt TCDD in the diet were bred at 7
months, two of eight females did not conceive and four of six that did conceive
could not carry their pregnancies to term. As one monkey delivered a stillborn
infant, only one conception resulted in a live birth (Schantz et al., 1979). As
described in an abstracted summary these results at 50 and 500 ppt TCDD are
compared to a group of monkeys given a dietary exposure to PBB (0.3 ppm,
Firemaster FF-1) in which seven of seven exposed females were able to conceive,
five gave birth to live, normal infants and one gave birth to a stillborn infant
(Allen et al., 1979). While the effects at 500 ppt TCDD may be associated with
significant maternal toxicity this would not appear to be the case at the lower
dose. After 50 ppt TCDD there were no overt effects on maternal health, but the
ability to conceive and maintain pregnancy was reduced (Allen et al., 1979).
In a similar series of experiments female rhesus monkeys were fed diets that
contained 0, 5 and 25 ppt TCDD (Bowman et al., 1989b; Schantz and Bowman, 1989).
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Reproductive function was not altered in the 5 ppt group as seven of eight
females mated to unexposed males after 7 months of dietary exposure to TCDD were
able to conceive. Six of these females gave birth to viable infants at term and
one gave birth to a stillborn infant. This was not significantly different from
the results of the control group that was fed a normal diet which contained no
TCDD. All seven of the monkeys in this control group were able to conceive and
give birth to viable infants. The 25 ppt dietary exposure level, however, did
affect reproductive function. Only one of the eight females in this group that
was mated, gave birth to a viable infant. As in the 50 ppt group from earlier
studies there were no serious health problems exhibited by any females exposed
to 0, 5 or 25 ppt TCDD. Therefore, the results in the 25 and 50 ppt groups
suggest that maternal exposure to TCDD, before and during pregnancy can result
in fetomortality without producing overt toxic effects in the mother.
McNulty (1984) examined the effect of a TCDD exposure during the first
trimester of pregnancy (gestational age 25-40 days) in the rhesus monkey. At a
total dose of 1 vg/kg given in nine divided doses, three of four pregnancies
ended in abortion, two of these in animals which demonstrated no maternal
toxicity. At a total dose of 0.2 pg/kg, one of four pregnancies ended in
abortion. This did not appear different from the control population, but the low
number of animals per group did not permit statistical analysis. McNulty (1984)
also administered single 1 /jg/kg doses of TCDD on gestational days 25, 30, 35 or
40. The number of animals per group was limited to three, but it appeared that
the most sensitive periods were the earlier periods, days 25 and 30, and that
both maternal toxicity and fetotoxicity were reduced when TCDD was given on later
gestational days. For all days at which a single 1 pg TCDD/kg dose was given
(gestational day 25, 30, 35 or 40) 10 of 12 pregnancies terminated in abortion.
Thus, of 16 monkeys given 1 fjg TCDD/kg in single or divided doses between days
25 and 40 of pregnancy, there were only three normal births (McNulty, 1984,
1985).
The primary effects on female reproduction appear to be decreased fertility,
inability to maintain pregnancy for the full gestational period, and, in the rat,
decreased litter size. In some studies signs of ovarian dysfunction such as
anovulation and suppression of the estrous cycle have been reported (Kociba et
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al., 1976; Barsotti et al., 1979; Allen et al.f 1979). Unfortunately, the amount
of attention that has been given to the female reproductive system, especially
in the nonpregnant state, has been limited.
5.2.1.2. ALTERATIONS IN HORMONE LEVELS — The potential for TCDD to alter
circulating female hormone levels has been examined, but only to a very limited
extent. In monkeys fed a diet that contained 500 ppt TCDD for S9 months the
length of the menstrual cycle, as well as the intensity and duration of
menstruation were not appreciably affected by TCDD exposure (Barsotti et al.,
1979). However, there was a decrease in serum estradiol and progesterone
concentration in five of the eight exposed monkeys, and in two of these animals
the reduced steroid concentrations were consistent with anovulatory menstrual
cycles. In summary form Allen et al. (1979) described the effects of dietary
exposure of female monkeys to 50 ppt TCDD. After six months of exposure to this
lower dietary level of TCDD there were was no effect on the serum estradiol and
progesterone concentrations of these monkeys. Thus, the presence of these
hormonal alterations is dependent on the level of dietary TCDD exposure.
Shiverick and Muther (1983) reported that there was no change in circulating
levels of estradiol in the rat after exposure to 1 pg/kg/day on gestation days
4-15. Taking all of these results together, the effect of TCDD exposure on
circulating female hormone levels may depend both on species and level of
exposure. It appears that any significant effect is only seen at relatively high
exposure levels, but very little research has been done and the studies to-date
have not been designed to examine alterations in female hormones specifically and
carefully.
5.2.1.3. ANTIESTROGENIC ACTION
5.2.1.3.1. In Vivo — Estrogens are necessary for normal uterine
development and for maintenance of the adult uterus. The cyclic production of
estrogens partially regulates the cyclic production of FSH and LH that results
in the estrous cycling of female mammals. In addition, estrogens are necessary
for the maintenance of pregnancy. Any effect that causes a decrease in
circulating or target cell estrogen levels can alter normal hormonal balance and
action.
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Early experimental results in rats and monkeys indicated that TCDD may have
an antiestrogenic action. Following administration of 1 /jg TCDD/kg/day to rats
for 13 weeks Kociba et al. (1976) reported morphologic changes in the ovaries and
uterus that were interpreted as being due to a suppression or inhibition of the
estrous cycle. Rhesus monkeys exposed to 500 ppt of TCDD in the diet for 6
months developed hormonal irregularities in their estrous cycles that were
associated with reduced conception rates as well as a high incidence of early
spontaneous abortions (Allen et al., 1977; Barsotti et al., 1979).
In rhesus monkeys the severity of the TCDD-associated reproductive
alterations was correlated with decreased plasma levels of estrogen and
progesterone (Barsotti et al., 1979). Thus, one possible mechanism for these
effects would be increased metabolism of estrogen and progesterone due to
induction by TCDD of hepatic microsomal enzymes and/or a decrease in the rate at
which these steroids are synthesized. On the other hand, serum concentrations
of 17p-estradiol are not significantly affected when TCDD is administered to
pregnant rats (Shiverick and Muther, 1983). Thus, an alternative mechanism for
TCDD-associated reproductive dysfunction could involve effects of TCDD on gonadal
tissue itself such as a decrease in its responsiveness to estrogen. In support
of this latter mechanism the administration of TCDD to CD-I mice results in a
decreased number of cytosolic and nuclear estrogen receptors in hepatocytes and
uterine cells. While TCDD treatment induces hepatic cytochrome P-450 levels in
these animals, it has no effect on serum concentrations of 17p-estradiol (DeVito
et al., 1992). This indicates that the antiestrogenic effect of TCDD in CD-I
mice is not caused by a decrease in circulating levels of estrogen.
Effects of estrogen on the uterus include a cyclic increase in uterine
weight, increased activity of the enzyme peroxidase, and an increase in the
tissue concentration of progesterone receptors. Antiestrogenic effects of TCDD
administration to female rats include decreased uterine weight, decreased uterine
peroxidase activity, and a decrease in the tissue concentration of progesterone
receptors (Safe et al., 1991). In addition, when TCDD and 17p-estradiol are co-
administered to the same female rat, the antiestrogenic action of TCDD diminishes
or prevents 17p-estradiol-induced increases in uterine weight, peroxidase
activity, progesterone receptor concentration, and expression of EGF receptor
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mRNA (Astroff et al., 1990; Safe et al., 1991). Similarly in mice, TCDD
administration decreases uterine weight and antagonizes the ability of 170-
estradiol to increase uterine weight (Gallo et al., 1986).
The ability of TCDD to antagonize the effects of exogenously administered
estrogen in the rat is dependent on the age of the animal. In 21-day-old rats
TCDD does not affect l?p-estradiol-induced increases in uterine weight or
progesterone receptor concentration. On the other hand, in 28-day-old intact
rats and 70-day-old ovariectomized rats both of these 17p-estradiol-mediated
responses are attenuated by TCDD (Safe et al., 1991). Previously, it had been
reported that TCDD administration does not alter the dose-dependent increase in
uterine weight due to exogenously administered estrone in sexually immature rats
(Shiverick and Muther, 1982). The later work by Safe et al. (1991) suggest that
this apparent lack of an antiestrogenic effect of TCDD may have been due to the
young age of the rats used.
5.2.1.3.2. In Vitro — Both TCDD and progesterone can affect a decrease
in the nuclear estrogen receptor concentration in rat uterine strips. However,
the effect of progesterone is inhibited by actinomycin D, cycloheximide and
puromycin, whereas the effect of TCDD is inhibited only by actinomycin D. The
reasons that the TCDD-induced decrease in nuclear estrogen receptors is blocked
by a transcription inhibitor, but not by protein synthesis inhibitors are not
understood. However, this result indicates that TCDD and progesterone decrease
the nuclear estrogen receptor concentration by different mechanisms (Romkes and
Safe, 1988). In addition, the antiestrogenic actions of TCDD can be demonstrated
in cell culture and two prominent mechanisms could potentially be involved. They
are (1) increased metabolism of estrogen due to Ah receptor mediated enzyme
induction, and (2) a down regulation of estrogen receptors within the target
cell.
In MCF-7 cells, which are estrogen responsive cells derived from a human
breast adenocarcinoma; antiestrogenic effects caused by the addition of TCDD to
the culture medium include a reduction of the 17p-estradiol-induced secretion of
a 160 kDa protein, 52 kDa protein, and a 34 kDa protein (Biegel and Safe, 1990).
These last two proteins are believed to be procathepsin D and cat heps in D
respectively. In addition, treatment of MCF-7 cells with TCDD suppresses the
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170-estradiol enhanced secretion of tPA, and inhibits estrogen dependent post-
confluent cell proliferation (Gierthy et al., 1987; Gierthy and Lincoln, 1988).
Thus, cultured MCF-7 cells have several estrogen-dependent responses that are
inhibited by TCDD; this characteristic makes them a useful model system for
studying the antiestrogenic actions of the compound.
In cultured MCF-7 cells TCDD treatment induces aryl hydrocarbon hydroxylase
(AHH) activity, the hallmark response of Ah receptor binding, and increases
hydroxylation of 17p-estradiol at the C-2, c-4, C-6a, and C-15a positions (Spink
et al., 1990). It turns out that the particular cytochrome P-450 that catalyzes
the C-2, C-15a and c-6a hydroxylations of 17p-estradiol is cytochrome P-450IA1
which is identical to AHH (Spink et al., 1992). TCDD treatment also results in
reduced levels of occupied nuclear estrogen receptors (Harris et al., 1990).
These results indicate, in MCF-7 cells, that the antiestrogenic effect of TCDD
could result from (1) an increased metabolism of estrogens due to Ah receptor
mediated enzyme induction, and/or (2) a decreased number of estrogen receptors
in the nucleus. Safe's group has published TCDD-concentration response
information for both the TCDD-induced decrease in occupied nuclear estrogen
receptors (Harris et al., 1989), and the induction of AHH and EROD activities in
MCF-7 cells (Harris et al., 1990). in addition, they have reported that TCDD
causes a decreased number of cytosolic and nuclear estrogen receptors in Hepa
Iclc7 cells which are a mouse hepatoma cell line (Zacharewski et al., 1991).
Independent analysis of the data suggests that the EC50 values for these effects
are not dissimilar enough to distinguish between the proposed mechanisms.
Instead, it appears as though TCDD induces the enzymes AHH and EROD over the same
concentration range that it causes a decreased concentration of occupied nuclear
estrogen receptors in MCF-7 cells. In Hepa Iclc7 cells the lowest concentration
used was 10 pM. While exposure to 10 PM TCDD resulted in a statistically
significant down regulation of estrogen receptors, Israel and Whitlock (1983)
reported that this concentration is the approximate EC50 for the induction of
cytochrome P-450IA1 mRNA and enzyme activity in these cells. Therefore, in Hepa
Iclc7 cells, as well as in MCF-7 cells it would appear that the TCDD concentra-
tions required to produce enzyme induction and reduction in occupied nuclear
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estrogen receptor levels are not dissimilar enough to distinguish between the two
proposed mechanisms.
More recently Safe's group has used an analog of TCDD, MCDF, that inhibits
the 17p-estradiol-induced secretion of the 34, 52 and 160 kDa proteins and down
regulates estrogen receptors in MCF-7 cells. These effects occurred at
concentrations of MCDF for which there is no detectable induction of EROD
activity (Zacharewski et al., 1992). In addition, it has been stated that the
down regulation of estrogen receptors in Hepa Iclc7 cells can be detected as
early as 1 hour after exposure of the cell cultures to 10 nM TCDD (Zacharewski
et al., 1991). This time is slightly less than the 2 hours required for Israel
and Whitlock (1983) to detect an increase in cytochrome P-450IA1 mRNA levels
after exposure of Hepa Iclc7 cells to 10 pM TCDD. After exposure of Hepa Iclc7
cells to a maximally inducing concentration of 1 nM TCDD; however, there are
significant increases in the cellular concentration of cytochrome P-450IA1 mRNA
after 1 hour, whereas the induction of aryl hydrocarbon hydroxylase activity
takes slightly longer (Israel and Whitlock, 1983).
Gierthy et al. (1987) reported that exposure of MCF-7 cells to 1 nM TCDD
caused suppression of the 17p-estradiol-induced secretion of tPA. This effect
of TCDD, however, occurred in the absence of any measurable decrease in the whole
cell concentration of estrogen receptors. While Gierthy-s group pretreated their
cultures with serum free medium, this was done to reduce cell proliferation and
maximize the cellular content of estrogen receptors. The disparity between this
I .
result of Gierthy et al. (1987) which suggests no effect of TCDD on the estrogen
receptor content of MCF-7 cells, and the results of Safe's group to the contrary
in this same cell line, remains largely unexplained. Overall it appears as
though no obvious distinction between the two proposed mechanisms can be made at
the present time. Therefore, it seems that the antiestrogenic effect of TCDD
results from both an increased metabolism of estrogen and a decreased number of
estrogen receptors. It is important to note that TCDD dops not compete with
radiolabeled estrogens or progesterone for binding to estrogen or progesterone
receptors, and that these steroids do not bind to the Ah receptor or compete with
radiolabeled TCDD for binding (Romkes et al., 1987; Romkes and safe, 1988).
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5.2.1.3.3. Evidence for an Ah Receptor Mechanism
5.2.1.3.3.1. Ah Receptor Mutants - while the precise cellular mechanism
by which TCDD produces its antiestrogenic effect is subject to a discordance
between two primary schools of thought, there is agreement that the response is
mediated by the Ah receptor. Thus, the antiestrogenic effects of TCDD in
cultured cells appear to involve an Ah receptor-mediated alteration in the
transcription of genes. This is indicated by studies using wild-type Hepa Iclc7
cells and mutant Hepa Iclc? cells in culture (Zacharewski et al., 1991). m
wild-type cells TCDD reduces the number of nuclear estrogen receptors and this
response can be inhibited by cycloheximide and actinomycin D. However, in class
1 mutants which have relatively low Ah receptor levels, TCDD has only a small
effect. Similarly, in class 2 mutants which have a defect in the accumulation
of transcriptionally-active nuclear Ah receptors, there was no effect of TCDD on
the number of nuclear estrogen receptors. Taken together, these results indicate
that the down regulation of estrogen receptors in Hepa Iclc? cells involves an
Ah receptor mediated effect on gene transcription. As previously noted TCDD
induces cytochrome P-4501A1 mRNA transcription and enzyme activity in Hepa Iclc?
cells (Israel and Whitlock, 1983). This effect is also Ah receptor mediated
(Nebert and Gielen, 1972).
5.2.1.3.3.2. Structure Activity Relationships In Vivo — Relative
potencies of halogenated aromatic hydrocarbon congeners as inhibitors of uterine
peroxidase activity in the rat are similar to their relative Ah receptor binding
affinities (Astroff and Safe, 1990). Only limited relative potency information
is available for the reduction of hepatic and uterine estrogen receptor
concentrations per se, by these substances in rats. TCDD and 1,2,3,7,8-PeCDD
both exhibit high affinity for the Ah receptor. At an 80 fig/kg dose of either
of these two substances, hepatic estrogen receptor concentrations are reduced 42%
and 41%, whereas uterine estrogen receptor concentrations are reduced 53% and 49%
by TCDD and 1,2,3,7,8-PeCDD respectively. On the other hand, 1,3,7,8-TCDD and
1,2,4,7,8-PeCDD bind less avidly to the Ah receptor. At a 400 ^g/kg dose of
either of these two substances, hepatic estrogen receptor concentrations are
reduced 36% and 40%, whereas uterine estrogen receptor concentrations are reduced
21% and 24% by 1,3,7,8-TCDD and 1,2,4,7,8-PeCDD respectively (Romkes et al.,
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1987). As the potency of these congeners for reducing estrogen receptor
concentrations correlates with their Ah receptor binding affinities these in vivo
results provide evidence that the antiestrogenic effect of TCDD is mediated by
the Ah receptor.
5.2.1.3.3.3. Genetic Evidence — Consistent with the interpretation based
on structure activity relationships there is a greater reduction in the number
of hepatic estrogen receptors when AhbAhb C57BL/6 mice are exposed to TCDD than
when AhdAhd DBA/2 mice are similarly exposed (Lin et al., 1991). To date,
however, the antiestrogenic effects have not been studied in the progeny of test
crosses between AhbAhb and AhdAhd mouse strains that respectively produce Ah
receptors with high or low binding affinity for TCDD. Therefore, the potential
segregation of the antieetrogenic effects of TCDD with the Ah locus has not been
verified by the results of genetic crosses.
5.2.1.3.3.4. Structure Activity Relationships In Vitro — The Ah receptor
is detectable in MCF-7 cells, and AHH as well as EROD activities are both
inducible in these cells (Harris et al., 1989). The relative abilities of TCDD
and other CDD, CDF and PCB congeners to suppress 17p-estradiol-induced secretion
of tPA by MCF-7 cells are consistent with the structure activity relationship for
other Ah receptor mediated responses (Gierthy et al., 1987). In addition, the
rank order of potency for several Ah receptor agonists in reducing nuclear
estrogen receptors in MCF-7 cells is TCDD > 2,3,4,7,8-PeCDD > 2,3,7,8-TCDF >
1,2,3,7,9-PeCDD > 1,3,6,8-TCDF (Harris et al., 1990). The rank order of potency
for these substances is consistent with their relative activities as Ah receptor
agonists. These results in vitro support a role for the Ah receptor in the
antiestrogenic actions of TCDD.
5.2.2. Male
5.2.2.1. REPRODUCTIVE FUNCTION/FERTILITY — TCDD and related compounds
decrease testis and accessory sex organ weights, cause abnormal testicular
morphology, decrease spermatogenesis, and reduce fertility when given to adult
animals in doses sufficient to reduce feed intake and/or body weight. Certain
of these effects have been reported in chickens, rhesus monkeys, rats, guinea
pigs, and mice treated with overtly toxic doses of TCDD, TCDD-like congeners, or
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toxic fat that was discovered later to contain TCDD (Allen and Lalich, 1962;
Allen and Carstens, 1967; Khera and Roddick, 1973; Kociba et al., 1976; Van
Miller et al., 1977; McConnell et al., 1978; Moore et al., 1985; Chahoud et al.,
1989; Morrisey and Schwetz, 1989). in testes of these different species, TCDD
effects on spermatogenesis are characterized by loss of germ cells, the
appearance of degenerating spermatocytes and mature spermatozoa within the lumens
of seminiferous tubules, and a reduction in the number of tubules containing
mature spermatozoa (Allen and Lalich, 1962; Allen and Carstens, 1967; McConnell
et al., 1978; Chahoud et al., 1989). The lowest cumulative dose of TCDD to
decrease spermatogenesis in the rat was 65 pg/kg administered over 13 weeks
(Kociba et al., 1976). At this dose body weights and feed consumption of the
rats were also significantly depressed. Thus, suppression of spermatogenesis is
not a highly sensitive effect when TCDD is administered to post-weanling animals.
5.2.2.2. ALTERATIONS IN HORMONE LEVELS — Effects of TCDD on the male
reproductive system are believed to be due in part to an androgenic deficiency.
This deficiency is characterized in adult rats by decreased plasma testosterone
and DHT concentrations, unaltered plasma LH concentrations, and unchanged plasma
clearance of androgens and LH (Moore et al., 1985, 1989; Mebus et al., 1987;
Moore and Peterson, 1988; Bookstaff et al., 1990a). The ED50 of TCDD for
producing this effect in adult male rats is 15 ^g/kg, and it can be detected
within 1 day of treatment. As described in the following sections, the cause of
the androgenic deficiency is decreased testicular responsiveness to LH and
increased pituitary responsiveness to feedback inhibition by androgens and
estrogens (Moore et al., 1989, 1991; Bookstaff et al., 1990a,b; Kleeman et al.,
1990).
5.2.2.3. TARGET ORGAN RESPONSIVENESS
5.2.2.3.1. Inhibition of Testicular Steroidogenesis. Testicular
steroidogenesis occurs within Leydig cells and is regulated primarily by plasma
LH concentrations (Payne et al., 1985; Hall, 1988). Binding of LH to the LH
receptor causes cAMP and possibly other second messengers to be formed (Cooke et
al., 1989). In response, cholesterol is rapidly transported to the initial
enzyme in the testosterone biosynthetic pathway, a cholesterol side chain
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cleavage enzyme, which is a cytochrome P-450 (cytochrome P-450^), located on
the inner side of the inner mitochondrial membrane that converts cholesterol to
pregnenolone. The mobilization of free cholesterol rather than its conversion
to pregnenolone and other metabolites is generally considered to be the rate
limiting step in testicular steroidogenesis. TCDD inhibits testosterone
biosynthesis, predominantly if not exclusively by inhibiting the mobilization of
free cholesterol which acts as a substrate for cytochrome P-450SCC (Moore et al.,
1991). Thus, in the testes of TCDD-treated rats, cholesterol is provided to the
cytochrome P-450SCC enzyme at too slow a rate to maintain androgenic homeostasis,
even when the plasma LH concentration characteristic of "normal" androgen levels
is present.
5.2.2.3.2. Altered Regulation of Pituitary LH Secretion. In TCDD-treated
male rats the expected increase in plasma LH concentration that would facilitate
testicular compensation for the decreased plasma androgens does not occur (Moore
et al., 1989). The failure of the plasma LH concentration to rise appropriately
is not caused by an increase in the plasma clearance of LH or by a decrease in
the maximal rate of pituitary LH synthesis or secretion (Bookstaff et al., 1990a;
1990b). Rather, TCDD alters the feedback regulation of LH secretion in male rats
by increasing the potency of testosterone and its metabolites (DHT and 170-
estradiol) as inhibitors of LH secretion. The ED^Q of TCDD for enhancing the
testosterone mediated inhibition of LH secretion is the same as its ED^Q for
causing the androgenic deficiency (15 pg/kg). Also, both responses are detected
within 1 day of TCDD dosing and are fully developed after 7 days. Decreased
plasma androgen concentrations normally result in compensatory increases in the
number of pituitary GnRH receptors, and the responsiveness of the pituitary to
GnRH. TCDD treatment prevents the increases in GnRH receptor number and
responsiveness that would be expected in the light of the decreased plasma
androgen concentrations (Bookstaff et al., 1990b). The pituitary is thus a
target organ for TCDD because its responsiveness to hormones secreted by the
testis (testosterone) and hypothalamus (GnRH) is altered by TCDD.
If the plasma LH concentrations in TCDD-treated rats did increase
appropriately in response to decreased plasma androgens, it is expected that
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plasma androgens would return to normal levels (Kleeman et al., 1990). This is
because the testes of TCDD-treated rats are capable of synthesizing more
testosterone than is needed to maintain androgen concentrations in the
physiological range, although this would require significantly elevated levels
of LH in TCDD-treated rats. The fact that there is a testicular reserve capacity
to provide for sufficient amounts of androgen synthesis; even when compromised,
underscores the importance of the effects of TCDD on pituitary LH secretion in
producing the effects of TCDD on plasma androgen concentrations.
5.2.2.3.3. Differential Responsiveness of Androgen Target Organs. The
dose-related reductions in plasma testosterone and DHT concentrations in intact
adult rats are accompanied by similar dose-related reductions in seminal vesicle
and ventral prostate weights (Moore et al., 1985). In contrast, TCDD has no
effect on accessory sex organ weights (or plasma androgen concentrations) in
castrated adult rats implanted with either testosterone- or DHT-containing
capsules (Moore and Peterson, 1988; Bookstaff et al., 1990a; 1990b). As the
trophic responsiveness of the seminal vesicles and ventral prostate to
testosterone and DHT are unaffected by postpubertal TCDD treatment, it follows
that TCDD can increase responsiveness of the pituitary to these androgens without
affecting the responsiveness of the accessory sex organs.
5.2.2.4. SUMMARY — In conclusion, although the androgenic deficiency is
an early-occurring effect following exposure of adult male rats to TCDD, has an
ED50 in the nonlethal range, and is far more severe in TCDD-treated animals than
in pair-fed controls, it is only detected at overtly toxic doses of TCDD that
reduce feed intake and body weight. Similarly, effects on male reproductive
function and fertility assessed in animals exposed as adults to TCDD are elicited
only by overtly toxic doses. Thus, the male reproductive system is relatively
insensitive to TCDD toxicity when exposure occurs in adulthood. Male reproduc-
tive toxicity induced by perinatal and lactational exposure to lower doses of
TCDD will be described in Section 5.3.3.
5.3. DEVELOPMENTAL TOXICITY
The manifestations of developmental toxicity have been divided into three
categories for convenience in assessing the data base with respect to an
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Ah-receptor mediated response. These categories include: death/growth/clinical
signs, structural malformations, and functional alterations. Exposure related
effects on death/growth/clinical signs are described for fish, birds, laboratory
mammals, and humans along with structure activity results that are consistent
with, but do not prove, an Ah-receptor mediated mechanism. Structural
malformations, particularly cleft palate formation and hydronephrosis in mice,
provide the most convincing evidence of an Ah receptor-mediated response.
However, postnatal functional alterations, some of which may be irreversible, are
more sensitive.
5.3.1. Death/Growth/Clinical Signs
5.3.1.1. FISH — Early life stages of fish appear to be more sensitive to
TCDD-induced mortality than adults. This is suggested by the LD50 of TCDD in
rainbow trout sac fry (0.4 A/g/kg egg weight) being 25 times less than that in
juvenile rainbow trout (10 pg/kg body weight) (Walker and Peterson, 1991; Kleeman
et al., 1988). The significance of this finding is that early life stage
mortality caused by high concentrations of TCDD-like congeners in fish eggs may
pose the greatest risk to feral fish populations (Walker and Peterson, 1991; Cook
et al., 1991). Cooper (1989) reviewed the developmental toxicity of CDDs and
CDFs in fish and Cook et al. (1991) discussed components of an aquatic ecological
risk assessment for TCDD in fish. The reader is referred to this literature for
more in depth coverage than will be presented here.
TCDD is directly toxic to early life stages of fish. This has been
demonstrated for Japanese medaka, pike, rainbow trout, and lake trout exposed as
fertilized eggs to graded concentrations of waterborne TCDD. In these species
TCDD causes an overt toxicity syndrome characterized by edema, hemorrhages and
arrested growth and development culminating in death (Helder, 1980, 1981; Wisk
and Cooper, 1990a; Spitsbergen et al., 1991; Walker et al., 1991; Walker and
Peterson, 1991). Histopathologic evaluation of lake trout embryos and sacfry has
shown this syndrome to be essentially identical to that of blue sac disease
(Helder, 1981; Spitsbergen et al., 1991). Following egg exposure to TCDD, signs
of toxicity are not detected in medaka until after the liver rudiment forms (Wisk
and Cooper, 1990a) and in lake trout toxicity is first detected -1 week prior to
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hatching but becomes fully manifest during the sac fry stage (Spitsbergen et al.,
1991; Walker et al., 1991). Among all fish species investigated thus far, lake
trout are the most sensitive to TCDD developmental toxicity. Following exposure
of fertilized lake trout eggs to graded waterborne concentrations of TCDD, the
NOAEL for sac fry mortality is 34 pg TCDD/g egg, the LOAEL is 55 pg TCDD/g egg,
and the egg TCDD concentration that causes 50% mortality above control at swim
up (LD50) is 65 pg TCDD/g egg (Walker et al., 1991). Thus, TCDD is a potent
developmental toxicant in fish and the effect is not secondary to maternal
toxicity.
The Ah receptor has not been identified in early life stages of fish;
however, it is assumed to be present because PCBs induce hepatic cytochrome
P-450IA1 in lake trout and brook trout embryos and fry (Binder and Stegeman,
1983; Binder and Lech, 1984). The Ah receptor has been identified in adult
rainbow trout liver (Heilmann et al., 1988) and in a rainbow trout hepatoma cell
line (Lorenzen and Okey, 1990). CDD and CDF congeners that are approximate
isostereomers of TCDD produce essentially the same pattern of toxic responses as
TCDD in early life stages of medaka and rainbow trout suggesting that they may
act through a common mechanism (Wisk and Cooper, 1990b; Walker and Peterson,
1991). Also in rainbow trout their potencies relative to TCDD (i.e., TEFs) for
causing early life stage mortality (TCDD LD50/congener LD50) are in the same
range as those proposed for human health risk assessment based on a diverse
spectrum of acute and subchronic toxicity tests in mammalian species (Safe, 1990;
Walker and Peterson, 1991). However, for the coplanar PCBs and monoortAo
chlorinated analogs of the coplanar PCBs, TEFs based on early life stage
mortality in rainbow trout are 1/14 to 1/80 less (Walker and Peterson, 1991) than
the TEFs proposed for risk assessment (Safe, 1990).
5.3.1.2. BIRDS — Bird embryos are also more sensitive to TCDD toxicity
than adults. The LD5Q of TCDD in the chicken embryo (0.25 pg/kg egg weight) is
100-200 times less than the TCDD dose that causes mortality in adult chickens
(25-50 M9/kg body weight) (Greig et al., 1973; Allred and Strange, 1977). The
LD5Q of TCDD injected into fertilized ring-necked pheasant eggs (1.1-1.8 pg/kg
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egg weight) is 14-23 times less than the TCDD dose that causes 75% mortality in
ring-necked hen pheasants (25 ^g/kg body weight) (Nosek et al., 1989, 1991).
Among bird species, most developmental toxicity research has been done on
chickens. Injection of TCDD or its approximate isostereomers into fertilized
chicken eggs causes a toxicity syndrome in the embryo characterized by
pericardial and subcutaneous edema, liver lesions, inhibition of lymphoid
development in the thymus and bursa of Fabricius, microophthalmia, beak
deformities, cardiovascular malformations, and mortality (Cheung et al., 1981;
Brunstrom and Darnerud, 1983; Rifkind et al., 1985; Brunstrom and Lund, 1988;
Brunstrom and Andersson, 1988; Nikolaidis et al., 1988a,b). On the other hand,
injection of a coplanar PCB into fertilized turkey eggs at a dose high enough to
cause microopthalmia, beak deformities, and embryo mortality did not produce
liver lesions, edema or thymic hypoplasia, all hallmark signs of TCDD toxicity
in the chicken embryo (Brunstrom and Lund, 1988). This disparity in signs of
TCDD embryotoxicity among bird species is not unique to the turkey and chicken.
In fertilized eggs of ring-necked pheasants and eastern bluebirds injection of
TCDD produces embryo mortality, but all of the other signs of toxicity seen in
t
the chicken embryo are absent, including cardiovascular malformations (Martin et
al., 1989; Nosek et al., 1989). Thus, in bird embryos the signs of toxicity
elicited by TCDD and its approximate isostereomers are highly species-dependent;
the only toxic effect common to all bird species is embryomortality.
There is evidence in chicken embryos that the Ah receptor may be involved
in producing developmental toxicity. The Ah receptor has been detected in chicken
embryos (Denison et al., 1986; Brunstrom and Lund, 1988) and the rank order
potency of PCB congeners for producing chicken embryo mortality: 3,3',4,4',5-PCB
> 3,3',4,4'-TCB > 3,3',4,4',5,5'-HCB > 2,3,3',4,4'-PCB > 2,3,4,4',5-PCB with
2,2',4,5'-TCB, 2,2',4,4',5,5'-HCB and 2,2•,3,3',6,6'-HCB being inactive, is
similar to that for a classic Ah receptor mediated response in the chicken
embryo, cytochrome P-450IA1 induction (Rifkind et al., 1985; Brunstrom and
Andersson, 1988; Brunstrom, 1989). However, while induction of cytochrome
P-450IA1 and toxicity may both be part of a pleiotropic response linked to the
Ah receptor, they are not otherwise causally related. This is demonstrated by
the nonsteroidal anti-inflammatory drug, benoxoprofen, suppressing 3,3',4,4'-TCB
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induced toxicity in the chicken embryo without altering its ability to induce
microsomal enzyme activity (Rifkind and Muschick, 1983). Also for 3,3',4,4'-TCB,
3,3',4,4',5,5'-HCB and TCDD there is a marked dissociation of the dose response
relationship for lethality and enzyme induction in the chicken embryo (Rifkind
et al., 1985).
A decreased activity of URO-D and an increased accumulation of uroporphyrins
are effects that are readily produced by exposure of cultured chicken embryo
liver cells to TCDD, 3,3•,4,4'-TCB and other PCBs (Sinclair et al., 1984; Marks,
1985; Lambrecht et al., 1988). Coplanar PCB congeners are more potent inhibitors
of URO-D activity in cultured chicken embryo liver cells than are noncoplanar PCB
congeners (Sassa et al., 1986), suggesting an Ah receptor mediated mechanism.
Unlike the results in cultured cells; however, a lethal dose of TCDD (6 nmol/egg)
does not affect URO-D activity or cause an increased accumulation of uropor-
phyrins in chicken embryos (Rifkind et al., 1985). Thus, TCDD-induced lethality
in chicken embryos is not associated with effects of TCDD on URO-D activity, even
though a decrease in URO-D activity might be expected to occur if a sufficient
dose of TCDD could be reached without being lethal.
The chicken embryo heart is a target organ for TCDD and other halogenated
aromatic hydrocarbons that act by an Ah receptor mechanism. The classic sign of
chick embryo toxicity involving the heart is pericardial edema. However, TCDD
has other effects on the chick embryo heart that are less well known. These
include its ability to produce cardiovascular malformations and to increase
cardiac release of arachidonic acid metabolites. When fertilized chicken eggs
are injected with graded doses of TCDD cardiovascular malformations are produced
including ventricular septal defects, aortic arch anomalies, and conotruncal
malformations. Approximately 1 pmol TCDD/egg causes cardiovascular malformations
in 50% of treated embryos versus 26-29% of control embryos (Cheung et al., 1981).
The cardiovascular malformation response may be unique to the chicken embryo
because in fertilized ring-necked pheasant and eastern bluebird eggs injected
with TCDD the incidence of such malformations is not increased (Nosek et al.,
1989; Martin et al., 1989).
In the chicken embryo heart arachidonic acid metabolism is stimulated by
TCDD resulting in increased formation of prostaglandins (Quilley and Rifkind,
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1986). Dose response relationships for the release of immunoreactive PGE2,
and TxB2 from chick embryonic heart are biphasic with an apparent maximally
effective dose of 100 pmol TCDD/egg. When the egg TCDD dose is - further
increased, release of these prostaglandins tends to decline towards levels in
control hearts. Biphasic dose response curves for cardiac PGE2 release also were
obtained with 3,3',4,4'-TCB and 3,3',4,4',5,5'-HCB (Quilley and Rifkind, 1986).
The thymus and bursa of Fabricius are other TCDD target organs in the
chicken embryo. TCDD, 3,3',4,4'-TCB and 3,3',4,4'-TCAOB cause dose-related
decreases in lymphoid development of both of these immune system organs
(Nikolaidis et al., 1988a,b, 1990). Cultured thymus anlage from chick embryos
are 100 times more sensitive to TCDDs inhibitory effect on lymphoid development
than cultured thymus anlage from turkey and duck embryos (Nikolaidis et al.,
1988a). This suggests that the reason thymic atrophy was not seen in turkey
embryos at egg doses of 3,3',4,4'-TCB that were overtly toxic (Brunstrom and
Lund, 1988) was not because the turkey embryo thymus was incapable of responding
to 3,3',4,4'-TCB. Rather, turkey embryos appear to be more sensitive to the
lethal than immunotoxic effect of this coplanar PCB.
Within the same bird species the signs of developmental toxicity elicited
by TCDD and its approximate isostereomers are similar. In the chicken embryo
TCDD, 3,3',4,4',5-PCB, 3,3',4,4'-TCB, and 3,3',4,4',5,5'-HCB all cause
pericardial and subcutaneous edema, liver lesions, microopthalmia, beak
deformities, and mortality, and TCDD, 3,3',4,4'-TCB and 3,3',4,4'-TCAOB inhibit
lymphoid development (Cheung et al., 1981; Brunstrom and Andersson, 1988;
Nikolaidis et al., 1988a,b). In pheasant embryos an altogether different pattern
of responses is seen. Nevertheless the TCDD-like congeners injected into
fertilized pheasant eggs, TCDD and 3,3',4,4'-TCB, produce the same pheasant
embryo-specific pattern. This pattern consists of embryo mortality in the
absence of edema, liver lesions, thymic hypoplasia, and structural malformations
(Brunstrom and Reutergardh, 1986; Nosek et al., 1989).
The lethal potency of TCDD and its approximate isostereomers in embryos of
different bird species varies widely. The chicken embryo is an outlier in that
it is by far the most sensitive of all bird species to TCDD. Turkey, ring-necked
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pheasant, mallard duck, domestic duck, domestic goose, golden-eye, herring gull,
black-headed gull and eastern bluebird embryos are considerably less sensitive
to the embryo lethal effect of TCDD and TCDD-like congeners (Brunstrom and Lund,
1988; Brunstrom and Reutergardh, 1986; Martin et al., 1989; Elliott et al., 1989;
Nosek et al., 1989). TCDD is 4-7 times more potent in causing embryo mortality
in chicken than pheasant embryos, and 3,3',4,4'-TCB is 20-100 times more potent
in chicken than turkey embryos (Allred and Strange, 1977; Brunstrom and Lund,
1988; Nosek et al., 1989). In chicken embryos an egg dose of 3,3•,4,4'-TCB of
4 /jg/kg increased embryomortality whereas an egg dose of 100 /jg/kg of the same
coplanar PCB had no embryotoxic effect in pheasants and mallard ducks and a dose
of 1000 pg/kg egg had no effect on embryomortality in domestic ducks, domestic
geese, golden eyes, herring gulls and black-headed gulls {Brunstrom, 1988;
Brunstrom and Reutergardh, 1986). In contrast to the above species differences,
the potency of 3,3',4,4'-TCB in causing embryomortality among different strains
of chickens is quite similar with the LD50 in six different strains varying
<4-fold (Brunstrom, 1988).
Graded doses of TCDD have been administered to fertilized eastern bluebird
and ring-necked pheasant eggs for the purpose of determining a LOAEL and NOAEL
for embryotoxicity. Mortality was the most sensitive embryotoxic effect in both
species. For eastern bluebirds, the LOAEL was 10,000 pg TCDD/g egg and the NOAEL
was 1000 pg TCDD/g egg (Martin et al. 1989). For ring-necked pheasants, the
LOAEL was 1000 pg TCDD/g egg and the NOAEL was 100 pg TCDD/g egg (Nosek et al.,
1989). In contrast, for chickens, the LD5Q for embryomortality is 250 pg TCDD/g
egg (Allred and Strange, 1977).
5.3.1.3. LABORATORY MAMMALS — when exposed to TCDD during adulthood
laboratory mammals display wide differences in the LD5Q of TCDD. It is
interesting to note, however, that when exposure occurs during prenatal
development, the potency of TCDD tends to be more similar across species. The
LD5Q of TCDD in adult hamsters, 1157-5051 j/g/kg, makes adult hamsters three
orders of magnitude more resistant to TCDD-induced lethality than are adult
guinea pigs (Olson et al., 1980; Henck et al., 1981). Yet, a maternal dose of
18 ^g TCDD/kg can increase the incidence of prenatal mortality in the hamster
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embryo/fetus. Since this dose is only 12 fold larger than the dose, 1.5 /ig
TCDD/kg, that increases the incidence of prenatal mortality in the guinea pig,
the hamster embryo/fetus approaches other rodent species in its sensitivity to
TCDD-induced lethality (Ol«on and McGarrigle, 1990; 1991). Thus, the magnitude
of the species differences in lethal potency of TCDD is affected by the timing
of TCDD exposure during the life history of the animal.
Exposure to TCDD during pregnancy causes prenatal mortality in the monkey,
guinea pig, rabbit, rat, hamster, and mouse (Table 5-1). Given a particular
dosage regimen the response is dose related and there appear to be species and/or
strain differences in susceptibilty to TCDD induced prenatal mortality. The rank
order of susceptibility from the most sensitive to least sensitive species would
appear to be monkey - guinea pig > rabbit = rat - hamster > mouse. However, an
important caveat roust be applied to the information presented in Table 5-1. This
is that the time period during which exposure of the embryo/fetus to TCDD occurs
is just as important a determinant of prenatal mortality as is the dose of TCDD
administered. This point will be illustrated in the text that follows when
prenatal mortality is described for different strains of mice.
It is important to note that the concept of a critical time period for
exposure makes the analysis of lethality data in the embryo/fetus qualitatively
different from that which might be applied to similar data in adult animals. For
example, a common dosing regimen used in mice, rats and rabbits (Table 5-1) is
to administer 10 cumulative doses of TCDD to the pregnant dam on days -6-15 of
gestation. This dosing regimen is presumably, expected to cover the critical
period resulting in what might be the maximal possible incidence of prenatal
mortality. In nearly all species of adult laboratory mammals however, single
lethal doses of TCDD would be expected to produce a similar delayed onset death
regardless of the age of the adult animal. Susceptibility to TCDD-induced
prenatal mortality, in contrast, may be greatly dependent on the age of the
embryo/fetus. In this case, multiple doses of TCDD that cover this critical
period might result in prenatal mortality, whereas a single dose might miss the
critical time and not result in prenatal mortality.
The following paragraphs will illustrate a type of analysis using an index
of cumulative maternal dose similar to the type of analysis that might be applied
5-21 08/06/92
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DRAFT—DO NOT QUOTE OR CITE
TABLE 5-1 I
Relationship Between Maternal Toxicity and Prenatal Mortality in Laboratory
Mammals Exposed to TCDD During Gestation*
Species/Strain
Monkey/ rhesus
Guinea pig/Hartley
Rabbit/New Zealand
Rat/Wistar
Rat/Sprague-Dawley
lamster/Golden
Syrian
Mouse/CD -1
Daily TCDD
Dose
(M9/ kg/day)
Of
0.1
0.25
0.5
1
Of
0.125
0.25
0.5
1
1
2
4
Of
0.03
0.125
0.5
2
8
Oj
25
50
100
200
400
Cumulative TCDD
Dose
Otg/kg)
Od
0.2
1
5
Oe
0.15
1.5
0
1
2.5
5
10
0
1.25
2.5
5
10
10
20
40
0
0.3
1.25
5
20
80
Oh
1.5
3
6
18
0
250
500
1000
2000
4000
Overt Maternal
Toxicity0
+
+
+
•f
+
+
t
+
+
+
+
+
+
-
+
+
Percent
Prenatal
Mortality0
25
25
81
100
+
7
12
42
22
100
DO O) CT
i*»«-p>jo.eo>o 10 o
ro in o
25
21
&
958
1008
58
7
6
13
H
87
97
Reference
McNulty, 1984
Olson and
McGarrigle, 1991
Giavini et al.,
1982
Khera and
Ruddick, 1973
Sparschu et a I.,
1971
Olson and
McGarrigle, 1991
Courtney, 1976
Source: Couture et al. 1990
Decreased body weight gain or marked edema compared to vehicle dosed control*. A (+) or (-) indicates
the presence or absence of an effect, respectively.
Percentage of absorptions plus late gestational deaths relative to all Implantations
.(-) is given it indicates the presence or absence of an effect, respectively.
TCDD administered in a single or divided doses between gestational days 20 and 40.
Single dose of TCDD administered on gestational day 14.
TCDD administered daily on days 6-15 of gestation.
^Significant at p<0.05
. Single dose of TCDD administered on gestational day 7 or 9.
'TCDD administered daily on days 7-16 of gestation.
A (+) or
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08/06/92
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DRAFT—DO NOT QUOTE OR CITE
to lethality data resulting from multiple dosing of adult animals. After
presenting the results of applying this type of analysis to prenatal mortality
data from different species, the caveat of critical time dependence will be
applied to the data obtained by using different strains of mice. This will
illustrate the importance of considering dosage regimen when evaluating prenatal
mortality data that is available in the literature. In this case a difference
of one gestational day might be critically important. It turns out that the form
of analysis using cumulative maternal dose may give the greatest possible degree
of species variation. As such different species may actually be more similar
with respect to susceptibility to prenatal mortality than would be apparent by
the results of this type of an analysis.
Using the cumulative dose data that is given in Table 5-1 there appears to
be a 10- to 20-fold difference in the fetolethal potency of TCDD when the
monkey/guinea pig is compared to the rabbit/rat/hamster. In the CD-I mouse
administered cumulative doses of TCDD on gestational days 7-16, not including day
6, it appears to require a daily dose of 200 pg TCDD/kg to significantly increase
prenatal mortality. Given a -5.5 day half-life of TCDD in the pregnant dam
(Weber and Birnbaum, 1985), the pregnant CD-I mouse would be exposed to a maximal
accumulated dose of -1200 pg TCDD/kg by the lowest dosage regimen that
significantly increased prenatal mortality. Therefore, by using the index of
cumulative dose the CD-I mouse would appear to be -1200 fold less sensitive than
the monkey/guinea pig for TCDD-induced prenatal mortality. However in NMRI mice
administered TCDD only on day 6 of gestation, prenatal mortality begins to
increase after a single dose of 45 /jg TCDD/kg (Neubert and Dillman, 1972). The
NMRI embryo/fetus is less susceptible to TCDD-induced prenatal mortality when the
TCDD is administered on later gestational days up to day 15. Thus, there appears
to be only about a 45-fold difference between the monkey/guinea pig and the NMRI
mouse when the NMRI embryo/fetus is exposed specifically on day 6. In C57BL/6
mice prenatal mortality is significantly increased after a single maternal dose
of 24 fjg TCDD/kg given on gestational day 6 (Couture et al., 1990b). This mouse
strain therefore, is about 20 to 30 fold less sensitive to TCDD-induced prenatal
mortality than is the monkey/guinea pig when exposed specifically on day 6. As
with the NMRI mouse there was little or no increase in prenatal mortality for the
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DRAFT—DO NOT QUOTE OR CITE
C57BL/6 stain when the TCDD was administered to the pregnant dam on gestational
days 8, 10, 12 or 14.
The concept of a critical window for TCDD-induced lethality in the
embryo/fetus suggests an explanation for the apparent insensitivity of the CD-I
embryo/fetus exposed to cumulative doses of TCDD. It could very well be that the
critical window for prenatal mortality in the mouse occurs approximately on
gestational day 6. If the embryo/fetus is not exposed to TCDD on gestational day
6, much larger doses of TCDD are required to produce prenatal mortality. Given
that exposure of the pregnant CD-I dams did not begin until gestational day 7,
this interpretation is consistent with the ability of a single 24 pg TCDD/kg dose
to increase the incidence of prenatal mortality when administered to pregnant
C57BL/6 mice on gestational day 6, but not when administered on gestational days
8, 10, 12 or 14 (Couture et al., 1990b). Similarly, Neubert and Dillman (1972)
found that the largest increase in prenatal mortality occurred when a single dose
of TCDD was given on day six compared to when the TCDD dose was administered on
one of the days 7-15. In addition, this would suggest that the CD-I embryo/fetus
does not have quite the relative insensitivity to the lethal effects of TCDD,
compared to the embryo/fetus of other species that would be indicated by using
cumulative maternal dose as the index of exposure.
It should be noted that the concept of a critical window for prenatal
mortality could potentially alter all of the species comparisons made previously
that were based on the cumulative maternal doses shown in Table 5-1. If this
turned out to be the case, then the true differences between species with respect
to their susceptibility to TCDD-induced prenatal mortality could be substantially
less than those indicated by using the cumulative maternal dose. This of course,
would involve a comparison between species using only single doses of TCDD given
during the critical time period for each species. At the present time it does
not seem possible to make such a comparison from the information available in the
literature.
Similar to fish and birds, the mammalian embryo/fetus is more sensitive to
the lethal action of TCDD than the adult. The maternal dose of TCDD that causes
58% fetal mortality in hamsters is 64-280 times less than the LD5Q of TCDD in
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DRAFT—DO NOT QUOTE OR CITE
adult hamsters (Olson »t al., 1980; Henck et al., 1981; Olson et al., 1990). In
Sprague-Dawley rats the cumulative maternal dose of TCDD that causes 41% prenatal
mortality is 5-10 time* less than the approximate LD50 of TCDD in adult rats of
the same strain (Sparschu et al., 1971; Seefeld et al., 1984). In rhesus
monkeys, the cumulative maternal TCDD dose that causes 81% prenatal mortality is
6 and 25 times less, respectively, than the lowest TCDD dose reported to cause
mortality in 1-year-old and adult rhesus monkeys (McNulty, 1977, 1985; Seefeld
et al., 1979).
A general finding in all nonprimate laboratory mammals, with the possible
exception of the hamster, is that TCDD-induced prenatal mortality is most
commonly associated with maternal toxicity that is not severe enough to result
in maternal lethality. This is seen in Table 5-1 for the guinea pig, rabbit, rat
and mouse. In each species the dose response relationship for maternal toxicity,
indicated by decreased maternal weight gain and/or marked subcutaneous edema of
the dam, is essentially the same as that for increased prenatal mortality. What
this means is that there may be an association between the fetolethal effect of
TCDD and maternal toxicity in all of these species. Even in the hamster where
maternal toxicity is far less severe, hematological alterations in the dam (Olson
and McGarrigle, 1991), could contribute to prenatal mortality.
In rhesus monkeys, on the other hand, the association between prenatal
mortality and maternal toxicity is not as easy to make. Only small numbers of
monkeys have been studied to date. However, the results following dietary
exposure to 25 ppt TCDD (Bowman et al., 1989b; Schantz and Bowman, 1989) and 50
ppt TCDD (Allen et al., 1977; Allen et al., 1979; Barsotti et al., 1979; Schantz
et al., 1979) before and during pregnancy suggest that TCDD-induced prenatal
mortality can occur in monkeys in the absence of overt toxic effects on the
mother. In four monkeys given a total cumulative dose of TCDD in nine divided
doses during the first trimester of pregnancy, McNulty (1984) observed that three
animals could not carry their pregnancies to term. Two of these abortions
occurred in monkeys that exhibited no overt signs of maternal toxicity, while the
third occurred in an overtly affected animal. Given the results of these
studies, extrapolation from which is limited by the small number of monkeys
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DRAFT—DO NOT QUOTE OR CITE
used, it would appear that there is no association between prenatal mortality and
maternal toxicity in the monkey even though such an association appears to exist
in other mammalian species. Indeed, the studies suggest that prenatal mortality
would not be an uncommon occurrence in monkeys at some exposure levels, even when
the mother is not overtly affected.
In guinea pigs and monkeys, minimal doses of TCDD that are lethal to the
embryo/fetus can in some instances produce no overt toxic effects on the mother.
In some cases however, these same doses of TCDD can produce a delayed onset
mortality of the dam (Table 5-2). In guinea pigs this is illustrated by the
lowest dose of TCDD that significantly increases prenatal mortality, 1.5 /ug/kg,
being lethal to one of 4 dams (Olson and McGarrigle, 1991). In rhesus monkeys
exposed to a total cumulative TCDD dose of 1 pg/kg, 14 of 16 pregnancies were
terminated by prenatal mortality, and 20 to 147 days after aborting 8 of 14
females showed signs of maternal toxicity and 3 of these 8 monkeys died (McNulty,
1984; 1985). Nevertheless, in most laboratory mammals, minimal doses of TCDD
that produce statistically significant increases in prenatal mortality cause a
much higher incidence of mortality to the embryo/fetus than to the dam. In fact,
treatment of pregnant rats, rabbits, hamsters and mice with minimal doses of TCDD
that result in prenatal mortality does not increase mortality of the dams at all
(Table 5-2).
Gestational exposure to TCDD produces a characteristic pattern of fetotoxic
responses in most laboratory mammals consisting of thymic hypoplasia, hematologic
alterations, subcutaneous edema, decreased fetal growth and prenatal mortality.
In addition to these common fetotoxic effects are other effects of TCDD that are
highly species-specific. Examples of the latter are cleft palate formation in
the mouse and intestinal hemorrhage in the rat. Table 5-2 shows those maternal
and fetal toxic responses that are produced by gestational exposure to TCDD in
various species of laboratory mammals. In the mouse, hydronephrosis is the most
sensitive sign of prenatal toxicity, followed by cleft palate formation and
atrophy of the thymus at higher doses, and by subcutaneous edema and mortality
at maternally toxic doses (Couture et al., 1990a; Courtney 1976; Courtney and
Moore, 1971; Neubert and Dillman, 1972). In the rat, TCDD prenatal toxicity is
manifested by intestinal hemorrhage, subcutaneous edema, decreased fetal growth
5-26 08/06/92
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Ul
I
ro
O
oo
o
a\
TABLE 5-2
Developmental Toxicity Following Gestations I Exposure to 2,3,7,8-TCOD8
Species/Strain
Mice/C57BL/6N
Nice/C57BL/6N
Nice/C57BL/6N
Nice/C57BL/6N
Mice/C57BL/6N
Mice/C57BL/6J
Mice/C57BL/6J
Mice/C57BL/6J
Hice/NMR
Mice/CF-1
Mice/DBA
Mice/DBA
Mice/CO-1
Daily Dose
0, 1 or 3 fig/k9
0. 12. 17 or 22 ng/kg
0, 3 or 12 fig/ kg
0 or 3 ftg/kg
0, 6, 9. 12, 15 or 18
(tg/kg
0 or 3 fig/kg
(subcutaneous)
20 fig/kg
0. 0.5, 1, 2 or 4 M/kg
0.3. 3, 4, 5 or 9
-------
TABLE 5-2 (cont.)
Species/Strain
Mice/CD-1
Rats/CD
Rats/Sprague-Daw I ey
Ra ts/Sprague-Daw I ey
Rats/Wistar
Guinea pigs/Hartley
Hamsters/Golden
Syrian
Rabbits/New Zealand
Rhesus monkeys
Monkeys/ rhesus
Dai ly Doseb
0, 25, 50, 100, 200 or
400 jig/kg
0 or 0.5 /tg/kg
(subcutaneous)
0, 0.125, 0.5 or 2 /tg/kg
0.03, 0.125, 0.5, 2 or 8
Jig/kg
0. 0.125. 0.25, 0.5, 1,
2, 4, 8 or 16 jig/kg
0, 0.15 or 1.5 /ig/kg
0, 1.5, 3, 6 or 18 ng/kg
0, 0.1, 0.25. 0.5 or 1
*9/kg
0, 5, 25, 50 or 500 ppt
7 months before and
during pregnancy
v, «.t , , , , or 5e
*g/kg
Treatment
Days
6-15
6-15
0-2
6, 15
5-14
14
7, 9
6-15
chronic
20-40
Sacrifice
Day0
17
20
20
21
21
58
15
28
"
" ~
Maternal Effectsd
increase in liver-to-body
weight ratio
none
decrease in weight gain
decrease in weight gain;
toxicity
toxicity
increase in mortality;
toxicity
increase in liver- to- body
weight ratio
decrease in weight gain;
toxicity
increase in mortality;
toxicity
increase in mortality;
toxicity
Embryo/Fetal Effectsd
increase in cleft palate,
hydronephrosis and fetal
mortality
increase in kidney anomaly
decrease in fetal body
weight
increase in fetal mortality,
resorptions, edema and
gastrointestinal hemorrhage
increase in fetal mortality,
edema and gastrointestinal
hemorrhage; decrease in
fetal weight
increase in fetal mortality
increase in fetal mortality,
hydronephrosis and renal
congestion; decrease in
increase in fetal mortality
and resorptions; extra ribs
increase in fetal mortality
increase in fetal mortality
==^==^==
Reference
Courtney, 1976
Courtney and
Moore, 1971
Giavini et al..
1982a
Sparschu et al.,
1971
Khera and
Roddick, 1973
Olson and
McGarrigle, 1990
Olson and
McGarrigle, 1990
Giavini et al.
1982b
Allen et al.,
1979; Bowman et
al., 1989
McNulty, 1985
D
£
"9
H
O
0
z
O
*3
I
w
8
O
M
n
B
tO
00
O
00
Source: Couture et al., 1990s
Oral exposure unless otherwise noted
°AU days adjusted to reflect plug day -- gestation day 0
Effects reported are only those that were statistically significant.
^Cumulative dose divided into nine oral closes administered between days 20 and 40 of gestation; two to four monkeys/dose.
Three animals given single oral dose, on either gestation days 25, 30, 35 or 40; 12 monkeys total.
NR = Not reported
\o
(O
-------
DRAFT—DO NOT QUOTE OR CITE
and mortality (Sparschu et al., 1971; Khera and Ruddick, 1973). Structural
abnormalities do occur in the rat but only at relatively large doses (Couture et
al., 1990a). In the hamster fetus, hydronephrosis and renal congestion are the
most sensitive effects, followed by subcutaneous edema and mortality at
fetolethal doses (Olson and McGarrigle, 1991). In the rabbit, an increased
incidence of extra ribs and prenatal mortality is found (Giavini et al., 1982),
while in the guinea pig and rhesus monkey prenatal mortality is seen (Olson and
McGarrigle, 1991; McNulty, 1984).
5.3.1.4 Structure Activity Relationships in Laboratory Mammals
The structure activity relationship for developmental toxicity in laboratory
mammals is generally similar to that for Ah receptor binding. Gestational
treatment of rats with CDD congeners that do not bind the Ah receptor, 2-MCDD,
2,7-DCDD, 2,3-DCDD or 1,2,3,4-TCDD, do not cause TCDD-like fetotoxic effects
(Khera and Ruddick, 1973). On the other hand, hexachlorodibenzo-p-dioxin, which
has intrinsic Ah receptor activity, produces fetotoxic responses in rats that are
essentially identical to those of TCDD (Schwetz et al., 1973). Similarly, when
administered to pregnant rhesus monkeys or CD-I mice PCB congeners that act by
an Ah receptor-mediated mechanism, 3, 3 • ,4,4 • -TCB and 3, 3 • ,4,4 • , 5, 5 --HCB cause the
same type of fetotoxic effects as TCDD. In contrast, 4,4'-DCB, 3,3',5,5'-TCB,
2,2',4,4',5,5'-HCB, 2,2 • ,4,4•6,6'-HCB and 2,2•,3,3',5,5•-HCB, which have
essentially no or very weak affinity for the Ah receptor, do not produce a TCDD-
like pattern of prenatal toxicity in mice (Marks and Staples, 1980; Marks et al.,
1981; 1989; McNulty, 1985). Thus, most structure activity results for overt
fetotoxic effects of the halogenated aromatic hydrocarbons are consistent with
an Ah receptor-mediated mechanism. Nevertheless, one finding which stands out
as being inconsistent is that 2,2' , 3,3 • ,4,4' -HCB which has very weak if any
affinity for binding to the Ah receptor causes the same pattern of fetotoxic
effects in mice as TCDD (Marks and Staples, 1980).
5.3.1.5. HUMANS — In the Yusho and Yu-Cheng poisoning episodes
developmental toxicity was reported in babies born to affected mothers who
consumed rice oil contaminated with PCBs, CDFs and PCQs (Hsu et al. 1985;
Yamashita and Hayashi, 1985; Kuratsune, ,1989; Rogan, 1989). In these incidents
it is essentially impossible to determine the contribution of TCDD-like versus
5-29 08/06/92
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DRAFT—DO NOT QUOTE OR CITE
nonTCDD-like congeners to the fetal/neonatal toxicity. Nevertheless, high
perinatal mortality was observed among hyperpigmented infant, born to affected
Yu-Cheng women who themselves did not experience incr.a..d mortality (Hsu et al.,
1985). Thus, in humans the developing embryo/fetus may be more sensitive than
the intoxicated mother to mortality caused by halogenated aromatic hydrocarbons.
in most cases, women who had affected children in the Yusho and Yu-Cheng
episodes had chloracne themselves (Rogan, 1982). B..ed on this evidence Rogan
suggested that "exposure to amounts insufficient to produce some effect on the
mother probably lessens the chance of fetopathy considerably- (Rogan, 1982).
in support of this interpretation overt signs of h.log.n.t.d aromatic hydrocarbon
toxicity were not observed in infants born to apparently unaffected mothers in
the Seveso, Italy, and Times Beach, Missouri, TCDD incidents (Reggiani, 1989;
Hoffman and Stehr-Green, 1989).
In laboratory mammals the studies summarized previously in Table 5-1 have
indicated an apparent association between prenatal mortality and maternal
toxicity in nonprimate species. However, some TCDD exposed rhesus monkeys were
not able to carry their pregnancies to term even in the absence of any overt
signs of maternal toxicity. This result in monkey, indicates that the
relationship between maternal toxicity and any prenatal toxic effect, on the
human embryo/fetus must be cautiously defined. More data may be required to
determine whether or not there is any association between overt maternal toxicity
and embryo/fetal toxicity in humans.
Effects of chemical exposure on normal development of the human fetus can
have four outcomes depending on the dose and time during ge.tation when exposure
occurs: fetal death, growth retardation, structural malformations and organ
system dysfunction, in the Yusho and/or Yu-Cheng incident all of these outcomes
were found (Yamashita and Hayashi, 1985; Kurat.une, 1989; Rogan, 1989).
Increased prenatal mortality and low birth weight .uggeeting fetal growth
retardation were observed in affected Yusho and Yu-Ch.ng women (Wong and Hwang
1981; Law et al., 1981; Yama.hita and Hayashi, 1985; H.u et al., 1985; Miller,
1985; Lan et al., 1989; Rogan et al., 1988). A .twctwral malformation, rocker
bottom heel, was observed in Yu.ho infant. (Yamaahita and Haya.hi, 1985). organ
dysfunction involving the CNS that was characterised by delay, in attaining
5"30 08/06/92
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DRAFT—DO NOT QUOTE OR CITE
developmental milestones and neurobehavioral abnormalities was reported in
Yu-Cheng children exposed transplacentally (Rogan et al., 1988; Hsu et al.,
1991).
Organs and tissues that originate from embryonic ectoderm are well known
targets for toxicity following exposure to TCDD-like halogenated aromatic
hydrocarbons. For example, treatment of monkeys with TCDD results in effects
involving the skin, meibomian glands and nails (Allen et al., 1977). Similarly,
a hallmark sign of fetal/neonatal toxicity in the Yusho and Yu-Cheng episodes is
an ectodermal dysplasia syndrome. It is characterized by hyperpigmentation of
the skin and mucous membranes, hyperpigmentation and deformation of finger and
toe nails, hypersecretion of the meibomian glands, conjunctivitis, gingival
hyperplasia, presence of erupted teeth in newborn infants, and altered eruption
of permanent teeth, missing permanent teeth and abnormally shaped tooth roots
(Taki et al., 1969; Yamaguchi et al., 1971; Funatsu et al., 1971; Wong and Hwang,
1981; Hsu et al; 1985; Yamashita and Hayashi, 1985; Rogan et al., 1988;
Kuratsune, 1989; Rogan, 1989; Lan et al., 1989). Additional effects on human
infants that are not related to ectoderm, but resemble effects that have been
observed following TCDD exposure in adult monkeys such as subcutaneous edema of
the face and eyelids were also reported (Allen et al., 1977; Moore et al., 1979;
Law et al., 1981; Yamashita and Hayashi, 1985; Rogan et al., 1988). Also, larger
and wider fontanels, and abnormal lung auscultation were found in the human
infants (Law et al., 1981; Yamashita and Hayashi, 1985; Rogan et al., 1988). The
similarities between these effects in human infants with those in adult monkeys
exposed to TCDD suggest that the effects in human infants exposed during the
Yusho and Yu-Cheng incidents may be caused by exposure to TCDD-like congeners.
This possibility is important given the fact that the affected human infants were
exposed to a complex mixture of substances that included TCDD-like congeners.
While chloracne is the most often cited effect of TCDD exposure involving
the skin in adult humans, has an animal correlate in the hairless mouse, and can
be studied by using a mouse teratoma cell line in tissue culture (Poland and
Knutson, 1982), it has rarely been recognized that the nervous system, like the
skin, is derived from embryonic ectoderm (Balinsky, 1970). As will be described
in Section 5.3.3.2, neurobehavioral effects occur following transplacental and
5-31 08/06/92
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DRAFT—DO NOT QUOTE OR CITE
neonatal exposure to TCDD-like congeners in mice, as well as transplacental
exposure to TCDD, itself in monkeys. In addition, some of the Yu-Cheng children
that were exposed transplacentally to PCBs, PCDFs and PCQs have affected a
clinical impression of developmental or psychomotor delay including impairment
of intellectual development (Rogan et al., 1988; Hsu et al., 1991). It is
possible to speculate that effects of TCDD-like congeners on the only internal
organ derived from ectoderm, the nervous system, are responsible for some of the
neurobehavioral effects observed in these children. Additional research is
required however, to characterize and elucidate the mechanisms by which TCDD
affects the nervous system.
5.3.2. Structural Malformations. Developmental effects consisting of cleft
palate, hydronephrosis and thymic hypoplasia are produced in mice following in
utero exposure to halogenated dibenzo-p-dioxin, dibenzofuran, biphenyl and
naphthalene congeners, which bind stereospecifically to the Ah receptor (Weber
et al., 1985; Birnbaum et al., 1987a,b, 1991). Of these effects in the mouse,
cleft palate is less responsive than hydronephrosis, as the latter is induced in
the absence of cleft palate (Couture et al., 1990b). Both responses can be
induced at TCDD doses that are not otherwise overtly toxic (Couture et al.,
1990a). The potency of TCDD for producing teratogenesis in the mouse is clearly
evident when one considers that only 0.0005% of a maternally administered dose
reaches the fetal palatal shelves or urinary tract. More specifically, a
maternal TCDD dose of 30 pg/kg results in 1.5 pg TCDD/mg in the palatal shelves
and 1 pg TCDD/mg in the kidneys 3 days after dosing (Abbott et al., 1989).
Susceptibility to the developmental actions of TCDD in mice depends on two
factors: genotype of the fetus and stage of development at the time of exposure.
The Ah receptor is thought to mediate the developmental effects of TCDD (Poland
and Knutson, 1982). Mouse strains that produce Ah receptors with relatively high
affinity for TCDD respond to lower doses of TCDD than mouse strains that produce
relatively low affinity Ah receptors (Poland and Glover, 1980; Hassoun et al.,
1984a). Thus, one genetically encoded parameter that determines the responsive-
ness of different mouse strains is the Ah receptor protein itself.
The differences that exist between mouse strains with respect to develop-
mental responsiveness to these chemicals are not absolute, as all strains
5-32 08/06/92
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DRAFT—DO NOT QUOTE OR CITE
including those with Ah receptors of relatively low affinity, respond when
exposed to sufficiently large doses during the critical period of organogenesis
(Birnbaum, 1991). In the mouse, the peak times of fetal sensitivity vary
slightly depending on which developmental effect is used as the endpoint.
However, exposure between days 6 and 15 of gestation will produce teratogenesis
(Couture et al., 1990a,b).
In inbred strains of mice the developmental response, characterized by
altered cellular proliferation, metaplasia and modified terminal differentiation
of epithelial tissues (Poland and Knutson, 1982), is extremely organ-specific
occurring only in the palate, kidney and thymus (Birnbaum, 1991). Pharmaco-
kinetic differences are not responsible for this high degree of tissue
specificity, and Ah receptors are not found exclusively in the affected organs
(Carlstedt-Duke et al., 1979; Gasiewicz et al., 1983). Therefore, other factors
intrinsic to the palate, kidney and thymus appear to play a role along with the
Ah receptors in these tissues in producing the structural malformations. For
certain developmental effects the time at which exposure occurs is important as
there may be a critical period during which the toxicant must be present in order
to produce the effect. This critical period can be different for different
organs and tissues.
Between mammalian species differences exist with respect to susceptibility
to the developmental effects of TCDD. While genetic differences between species
or strains might affect absorption, biotransformation and/or elimination of TCDD
by the maternal system and its absorption across the placenta, such species
differences do not account for the lack of cleft palate formation in species
other than mice (Birnbaum, 1991). Rather, the species differences in suscepti-
bility to cleft palate formation appear due to differences in the interaction
between TCDD and the developing palatal shelves themselves. This is demonstrated
by the occurrence of similar responses when palatal shelves from different
species are exposed to TCDD in organ culture (Abbott et al., 1989; Abbott and
Birnbaum, 1990a, 1991). The key difference is that much higher concentrations
of TCDD are required to elicit essentially the same palatal response that is seen
in the mouse in other species (Table 5-3).
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TABLE 5-3
TCDD Responsiveness of Palatal Shelves From the Mouse,
Rat and Human in Organ Culture8
Species
Mouse
Ratb
Humanc
Molar Concentration of TCDD
Prevention of the Epithelium to
Mesenchyme Transformation Process
LOEL
IxlCT13
IxlO-10
sxicr11
EC100
SxlO'11
IxlO'8
ixicr8
Cytotoxicity
1X1CT10
IxlCT7
IxlO'7
aSource: Birnbaum, 1991
At the highest concentration tested, 60% of the palatal shelves
failed to undergo programmed cell death.
cOne of four shelves responded by failing to undergo programmed cell
death at SxlO'11 M.
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With respect to the occurrence of similar developmental effects in mammalian
species other than the mouse, no other species develops cleft palate except at
maternal doses that are fetotoxic and maternally toxic (Couture et al., 1990a;
Birnbaum, 1991). In mice and hamsters hydronephrosis can be elicited at TCDD
doses that are neither fetotoxic nor maternally toxic (Olson and McGarrigle,
1991), whereas thymic hypoplasia is a fetal response to TCDD observed in
virtually all laboratory mammalian species that have been tested (Vos and Moore,
1974). Studies in humans have not clearly identified an association between TCDD
exposure and structural malformations (Fara and Del Corno, 1985; Mastroiacovo et
al., 1988; Stockbauer et al., 1988; Reggiani, 1989).
5.3.2.1. CLEFT PALATE
5.3.2.1.1. Characterization of TCDD Effect. Palatal shelves in the mouse
originate as outgrowths of the maxillary process. Eventually they come to lie
vertically within the oral cavity on both sides of the tongue. In order to form
the barrier between the oral and nasal cavities, the shelves in the mouse must
reorient themselves from a ventromedial (vertical) direction to a horizontal
direction. Once they come together horizontally, their medial aspects bring
apposing epithelia into close contact (Coleman, 1965; Greene and Pratt, 1976).
At this stage, the apposing medial edge epithelia of the separate palatal shelves
each consist of an outer layer of periderm that overlays a strata of cuboidal
shaped basal cells. These basal cells, in turn, rest on top of a continuous
basal lamina. There is a sloughing of the outer periderm cells followed by the
formation of junctions between the newly apposing basal epithelial cells. The
midline seam so formed consists of the two layers of basal cells, all of which
remain viable, even though the outer periderm cells die and slough away. As the
palatal shelf continues to grow, the bilayer seam, which itself grows at a slower
rate, turns into a single layer of cells, and then breaks up into small islands
of cells. Eventually, the basal lamina disappears, and the elongating former
basal cells within the small islands extend filopodia into the adjacent
connective tissue. During this process the former basal cells lose epithelial
characteristics and gain fibroblast-like features. Essentially, the medial edge
epithelium is an ectoderm that retains the ability to transform into mesenchymal
cells. Upon completion of this epithelial to mesenchyme transformation, the once
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separate and apposing palatal shelves are fused so that a single continuous
tissue is formed (Fitchett and Hay, 1989; Shuler et al., 1991).
Cleft palate can result from a failure of the shelves to grow and come
together, or a failure of the shelves to fuse once they are in close apposition
(Pratt et al., 1985). TCDD and other Ah receptor agonists are unusual inducers
of cleft palate because the shelves grow and make contact, but the subsequent
process involving the epithelial to mesenchyme transformation does not occur.
Therefore, a cleft is formed as the palatal shelves continue to grow without
fusing. When TCDD is administered to pregnant mice on gestational days 6-12, the
incidence of cleft palate formation increases with time. However, day 12 is a
critical window, after which the incidence of cleft palate formation decreases.
No cleft palates are formed when TCDD is administered on day 14 (Couture et al.,
1990b).
Palatal shelves of the mouse, rat and human can be removed from the fetus
and placed into organ culture. Under these conditions, when the separate shelves
are placed in an apposing condition in vitro, sloughing periderm cells are
trapped within the seam. Thus, due to the presence of these trapped dead cells,
the fusion process was characterized as a programmed cell death (Coleman, 1965;
Greene and Pratt, 1976; Pratt et al., 1984). However, the newer model, which
involves transformation of the basal epithelial cells into mesenchyme rather than
their death, is believed to be valid under explant conditions in vitro, as well
as in vivo (Fitchett and Hay, 1989). When exposed to TCDD as explants in vitro
the palatal shelves of the mouse, rat and human all respond to TCDD in a similar
way by not completing the fusion process (Abbott et al., 1989; Abbott and
Birnbaum, 1989, 1990a, 1991). The epithelial to mesenchyme transformation of the
basal epithelial cells does not occur, and instead there is a differentiation
into a stratified squamous epithelium such that these cells resemble the squamous
keratinizing oral cells within the tissue (Birnbaum and Abbott, personal
communication).
Table 5-3 shows the lowest TCDD concentration which prevents the epithelial
to mesenchyme transformation process in isolated palatal shelves (LOEL), TCDD
concentration that produces a 100% maximal response (ECj00), and lowest
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concentration of TCDD that produces cytotoxicity. Palatal shelves of rats and
humans respond to TCDD in a manner identical to the mouse; however, higher
concentrations of TCDD are required to prevent the epithelial to mesenchyme
transformation process. The relative insensitivity of rat palatal shelves may
explain the lack of cleft palate when fetal rats are exposed to nonmaternally
toxic doses of TCDD. Sensitivity of human palatal shelves to TCDD in vitro is
similar to the rat. This suggests that exposure to maternally toxic and
fetotoxic doses of TCDD would be required to cause cleft palate formation in
humans.
A disruption in the normal spatial and temporal expression of EGF, TGF-a,
TGF-pl and TGF-02 correlates with altered proliferation and differentiation in
the medial region of the developing palate resulting in a palatal cleft. Thus,
the abnormal proliferation and differentiation of TCDD-exposed medial cells may
be related to reduced expression of EGF and TGF-a. Also, decreased levels of
immunohistochemically detectable TGF-pl could contribute to the continued
proliferation and altered differentiation of medial cells (Abbott and Birnbaum,
1990b).
5.3.2.1.2. Evidence for an Ah Receptor Mechanism
5.3.2.1.2.1. Genetic — When wild-type C57BL/6 (AhbAhb) mice are crossed
with DBA/2 (AhdAhd) mice that contain a mutation at the Ah locus, all of the
heterozygous, B6D2F1 progeny (AhbAh(') resemble the wild-type parent in that AHH
activity is inducible by TCDD and other halogenated aromatic hydrocarbons (Nebert
and Gielen, 1972). Test crosses between the B6D2F1 progeny and each original
parent strain, and other B6D2F1 progeny mice demonstrate that in the C57BL/6 and
DBA/2 strains susceptibility to AHH induction segregates as a simple dominant
trait in the backcross and F2 progeny. Thus, the trait of AHH induction is
expressed in progeny that contain the AhbAhb and AhbAhd genotypes, but is not
expressed in the AhdAhd progeny from these crosses. Certain other effects of
TCDD, such as its binding affinity for the hepatic Ah receptor (Okey et al.,
1979), thymic atrophy (Poland and Glover, 1980), hepatic porphyria (Jones and
Sweeney, 1980) and immunosuppressive effects (Vecchi et al., 1983; Nagarkatti et
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al., 1984) have been shown in similar genetic crosses and test crosses to
segregate with the Ah locus that permits AHH induction. Thus, for these effects
of TCDD, genetic evidence demonstrates an involvement of the Ah locus (Poland and
Knutson, 1982).
Neberf s group was the first to relate developmental toxicity to the Ah
locus in mice (Lambert and Nebert, 1977; Shum et al., 1979). Subsequently,
Poland and Glover (1980) administered a single 30 /jg TCDD/kg dose to pregnant
mice on gestational day 10. It was found that there was a 54% incidence of cleft
palate in homozygous C57BL/6 (AhbAhb) fetuses, a 13% incidence in heterozygous
B6D2F1 (C57BL/6 and DBA/2 hybrid, AhbAhd) fetuses and only a 2% incidence in
homozygous DBA/2 (AhdAhd) fetuses. This pattern of inheritance in which the
incidence of developmental toxicity in the heterozygous Fl generation is
intermediate between that of the homozygous parental strains is consistent with
the autosomal dominant pattern of inheritance described for AHH inducibility and
the Ah locus (Nebert and Gielen, 1972), even if dominance is incomplete in the
case of developmental toxicity. However, the pattern of inheritance for
developmental toxicity described when Poland and Glover (1980) crossed C57BL/6
and DBA/2 mice is not proof positive that the Ah locus is the genetic locus that
controls susceptibility to TCDD-induced developmental toxicity in these mouse
strains.
To provide such proof it is necessary to show genetic linkage between the
susceptibility for developmental toxicity and the Ah locus. The standard of
proof would be that developmental toxicity and a particular allele at the Ah
locus must always segregate together in genetic crosses, because if the loci are
the same there can be no recombination between the loci. This is generally
accomplished by demonstrating cosegregation between the two loci not only in
crosses between the two homozygous parental strains, which in and of itself is
insufficient proof of genetic linkage, but also in test crosses or back crosses
between the heterozygous Fl hybrids with each homozygous parental strain.
It has been stated previously (first paragraph of this section), that
certain effects of TCDD are well known to segregate with the Ah locus due to the
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results of appropriate crosses and back crosses between responsive and nonrespon-
sive mouse strains and their hybrid Fl progeny. With this standard of proof in
mind, the evidence that specifically links developmental toxicity with the Ah
locus will now be described. It is intended that this information be provided
with a considerable degree of detail. This is so that the reader can indepen-
dently determine whether or not the standard of proof has been satisfied by the
evidence available.
In order to strengthen their conclusion based on the results of simple
crosses between C57BL/6 and DBA/2 mice Poland and Glover (1980) planned to
perform a backcross between the hybrid B6D2F1 and DBA/2. However, the low
incidence of cleft palate in B6D2F1 mice would have required characterizing and
phenotyping a prohibitively large number of fetuses. Alternatively, the
backcross between B6D2F1 and C57BL/6 was considered in which Ah Ah and Ah Ah
progeny would have been distinguished by the amount of high affinity specific
binding for TCDD in fetal liver. In this case however, overlap between
individual mice would have made the results uncertain in some of the progeny.
Therefore, it was not possible to obtain satisfactory results from either
backcross.
Instead Poland and Glover (1980) examined the incidence of cleft palate in
10 inbred strains of mice; 5 strains with high affinity Ah receptors and 5
strains with low affinity receptors. In the five latter strains, there was only
a 0-3% incidence of cleft palate formation, whereas four of the five strains with
high affinity Ah receptors developed a >50% incidence. The one strain with high
affinity Ah receptors that did not follow the pattern, CBA strain, is also
resistant to cleft palate formation induced by glucocorticoids. Overall, these
results indicate that cleft palate formation probably segregates with the Ah
locus.
The incidence of cleft palate formation was studied in fetuses from a cross
between C57BL/6 and AKR/NBom mice administered 3,3' ,4,4'-TCAOB on gestational day
12 (Hassoun et al., 1984b). While C57BL/6 mice are responsive for AHH induction
and cleft palate formation, AKR mice are less responsive, requiring higher doses
for both effects. In a manner unlike the result of a cross between C57BL/6 and
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DBA/2, the incidence of cleft palate formation in the B6AKF1 progeny was <2%
showing that nonresponsivenesa segregates as the dominant trait when C57BL/6 mice
are crossed with AKR mice. Similarly, cleft palate formation was virtually
absent in the progeny of a backcross between AKR/NBom and B6AKF1 demonstrating
dominance of the noninducible trait. While Ah phenotyping of the backcross
progeny was not performed in this particular study, Robinson et al. (1974) had
previously evaluated segregation of the Ah locus in backcrosses between C57BL/6
and AKR/N mice. They found in these two strains that noninducibility for AHH
activity segregates as the dominant trait. Thus, inducibility for cleft palate
formation and AHH activity both segregate as dominant traits when C57BL/6 mice
are crossed with DBA/2, but noninducibility is dominant for both traits when
C57BL/6 mice are crossed with AKR/N. These results are consistent with the
interpretation that cleft palate induction probably segregates with the Ah locus.
Like Poland and Glover (1980), Hassoun et al. (1984a) were unable to
determine whether or not cleft palate formation segregates with the Ah locus in
C57BL/6 and DBA/2 mice by performing simple backcrosses. Instead, they evaluated
co-segregation of the Ah locus and 2,3,7,8-TCDF induced cleft palate formation
using a series of recombinant strains called BXD mice. These strains are fixed
recombinants produced from an original cross between the two parental strains
C57BL/6J and DBA/2J. Hybrid B6D2F1 mice were crossed to produce F2 progeny and
these were strictly inbred by sister and brother matings into several parallel
strains. The mice used in this study were from the F42 and F58 generations of
inbreeding. It was found that the incidence of TCDF-induced cleft palate
formation after matings within eight different BXD strains with high affinity Ah
receptors is >85%. After similar matings with eight different BXD strains with
low affinity Ah receptors, the incidence of TCDF-induced cleft palate formation
is <2%. These results of Hassoun et al. (1984a) corroborate those of Poland and
Glover (1980) and provide the best evidence currently available that cleft palate
formation segregates with the Ah locus. Thus, the Ah locus and the Ah receptor
are involved in the formation of palatal clefts that are induced by TCDD-like
congeners.
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As additional evidence, stereospecific, high affinity Ah receptors can be
isolated from cytosol fractions prepared from embryonic palatal shelves. These
receptors are present in palatal shelves of Ah"Ah , C57BL/6 fetuses, but are not
detectable in similar tissue from Ah^Ah1*, AKR/J fetuses (Dencker and Pratt,
1981). However, the significance of this finding may be mitigated to some extent
by the following observation. In cytosols prepared from homogenates of whole
embryo/fetal tissue (minus head, limbs, tail and viscera), the concentration of
specific binding TCDD receptors is 256 fmol/mg protein in C57BL/6 mice compared
to a concentration of 21 fmol/mg protein in the less responsive DBA/2 strain, 15
fmol/mg protein in the less responsive AKR/J strain and 19 fmol/mg protein in the
less responsive SWR/J strain. However, when embryonic tissue is cultured, the
differences between the strains in receptor number are less pronounced, and in
the receptors isolated from cultured embryonic cells of different strains, there
•»
is only about a 2-fold difference in the relative binding affinity for JH-TCDD.
The mechanistic reasons for the diminished degree of difference between
responsive and less responsive mouse strains during embryonic cell culture are
not known (Harper et al., 1991).
The possible influence of maternal toxicity on cleft palate formation was
evaluated by performing reciprocal blastocyst transfer experiments using the high
affinity Ah receptor-NMRI and lower affinity Ah receptor-DBA strains of mice
(D'Argy et al., 1984). After administration of 30 pg TCDD/kg or 8 mg TCAOB/kg
to pregnant dams on gestational day 12, 75-100% of all NMRI fetuses develop cleft
palates. This is true whether the fetuses remain within the uterus of their
natural mother or are transferred into the uterus of a DBA mouse. Under the same
conditions, none of the 24 DBA fetuses transferred into an NMRI mother develop
a cleft palate, even though 89% of their NMRI litter mates are affected. Thus,
these results, along with the presence of Ah receptors in palatal shelves and
responsiveness of palatal shelves in organ culture to TCDD, indicate that cleft
palate formation in mice is due to a direct effect of TCDD on the palatal shelf
itself, and is not secondary to maternal toxicity.
5.3.2.1.2.2. Structure Activity — As TCDD induced cleft palate formation
and hydronephrosis in mice appears to be mediated by the Ah receptor, structure-
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activity requirements based on Ah receptor binding characteristics should predict
the relative potencies of different agonists for producing cleft palate and
hydronephrosis. Of the halogenated aromatic hydrocarbons TCDD has the greatest
affinity for binding to the Ah receptor and it is the most potent teratogen in
inbred mouse strains. Table 5-4 shows the relative potencies for cleft palate
induction and hydronephrosis in C57BL/6 mice for a number of TCDD-like congeners.
As TCDD is the most potent, it is assigned a value of 1.000. When examined by
probit analysis the dose response curve of each congener, compared to all of the
others, did not deviate from parallelism. Therefore, the relative potencies of
the congeners are valid for any given incidence of cleft palate formation or
hydronephrosis. The main finding, however, is that the rank order potency of the
various congeners for producing these two developmental effects is generally
similar to that for binding to the Ah receptor (Table 5-4), with the notable
exception that the apparent binding affinities for the brominated dibenzofurans
have not yet been reported.
Other ligands for the Ah receptor that cause cleft palate formation in
C57BL/6 mice at nonmaternally toxic doses include 3,3',4,4'-TCAOB (Hassoun et
al., 1984a), 3,3',4,4'-tetrachlorobiphenyl (Marks etal., 1989), 3,3',4,4',5,5'-
hexachlorobiphenyl (Marks et al., 1981) and a mixture that contained 1,2,3,4,6,7-
and 2,3,4,5,6,7-hexabromonaphthalenes (Miller and Birnbaum, 1986).
Also consistent with the structure-activity relationships for binding to the
Ah receptor, a number of hexachlorobiphenyls do not induce cleft palate
formation. These congeners either lack sufficient lateral substitution or are
substituted in such a manner that they cannot achieve a planar conformation.
Included in this category are the diortAo and tetraortho chlorine-substituted
2,2',3,3',5,5'-, 2,2',3,31,6,6'-, 2,2',4,4•5,5•- and 2,2',4,4',6,6'-hexachloro-
biphenyls (Marks and Staples, 1980). In addition, it is consistent with the
structure-activity relationships that the monoortho chlorine-substituted
2,3,4,5,3',4'-HCB is a weak teratogen. Its potency relative to that of TCDD
varies from 3xlO~5 to 9xlO'5 for cleft palate formation, AHH induction and
hydronephrosis (Table 5-4) (Kannan et al., 1988).
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TABLE 5-4
Apparent Ah receptor Binding Affinity and Relative Teratogenic
Potency of Halogenated Aromatic Hydrocarbon Congeners*
Congener
2,3,7, 8-TCDD
2,3,7, 8-TBDD
2,3,7,8-TBDF
2,3,4,7,8-PeCDF
2,3,7,8-TCDF
1,2,3,7,8-PeCDF
1,2,3,4,7,8-HxCDF
2,3,4,7,8-PeBDF
1,2,3,7,8-PeBDF
2,3,4,5,3' ,4'-HxCB
Apparent
Binding
Affinity
EC50b'C
(mol/L)
l.OxlO'8
1.5xlO'9
l.SxlO'8
4.1xlO'8
7.4xlO*8
2.3xlO*7
S.OxlO^6
Relative Potency
(ED^Q TCDD/ED^Q Congener)
Cleft Palated
1.000
0.235
0.100
0.095
0.049
0.026
0.010
0.005
0.004
0.0000287
Hydronephrosis^
1.000
0.444
0.333
0.057
0.021
0.074
0.049
0.009
0.018
0.0000894
"Source: Weber et al., 1985; Birnbaum et al., 1987a,b, 1991 and
Safe, 1990
"Determined for Ah receptor binding in H-4-IIE rat hepatoma cells
using 3H-TCDD as the radioactive ligand.
cn
cBlank spaces in this column indicate that no
reported for the congener in H-4-IIE rat hepatoma cells.
Determined in C57BL/6 mice.
value has been
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A result that would not be expected according to the structure activity
relationships for binding to the Ah receptor is that the diortAo chlorine-
substituted 2,2',3,3',4,4'-hexachlorobiphenyl causes cleft palate formation and
hydronephrosis in mice (Marks and staples, 1980). However, another diortho
chlorine-substituted PCB congener, 2,2•,4,4•,5,5'-hexachlorobiphenyl, can also
cause hydronephrosis and is a very weak inducer of EROD activity (Biegel et al.,
1989; Morrissey et al., 1992). It is consistent with the interpretation that
2,2',4,4',5,5'-hexachlorobiphenyl is a partial Ah receptor agonist, that it can
competitively displace TCDD from the murine hepatic cytosolic receptor and, at
large enough doses, can inhibit TCDD-induced cleft palate formation and
immunotoxicity in C57BL/6 mice (Biegel et al., 1989; Morrissey et al., 1992).
These results suggest that PCB congeners do not have to be in a strictly planar
configuration to cause teratogenesis.
5.3.2.1.3. Species Differences. Cleft palate is induced in rats only at
maternally toxic TCDD doses that are associated with a high incidence of fetal
lethality. Schwetz et al. (1973) reported an increased incidence of cleft palate
after maternal administration of 100 pg hexachlorodibenzo-p-dioxin/kg/day on
days 6-15 of gestation to Sprague-Dawley rats. Couture et al. (1989) also
observed an increased incidence of cleft palate formation after a single dose of
300 pig/kg of 2,3,4,7,8-pentachlorodibenzofuran given to Fisher 344 rats.
Similarly, cleft palate can be produced in fetal hamsters following maternally
toxic and fetotoxic doses of TCDD (Olson et al., 1990).
In monkeys, bifid uvula (Zingeser, 1979) and bony defects in the hard palate
(McNulty, 1985) were reported, but there were no corresponding soft tissue
defects or clefts of the secondary palate. Cleft palates have not been reported
in human fetuses of mothers accidentally exposed to TCDD or mixtures of PCBs and
CDFs (Fara and Del Corno, 1985; Mastroiacovo et al., 1988; Stockbauer et al.,
1988; Rogan, 1989). Thus, sensitivity of the palate in mice to TCDD is unique.
In other species, including humans, other forms of fetal toxicity occur at doses
lower than those required for cleft palate formation.
5.3.2.2. HYDRONEPHROSIS
5.3.2.2.1. Characterization of TCDD Effect. Hydronephrosis is the most
sensitive developmental response elicited by TCDD in mice. It is produced by
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maternal doses of TCDD too low to cause palatal clefting and is characterized as
a progressive hydronephrosis preferentially occurring in the right kidney which
can be accompanied by hydroureter and/or abnormal nephron development (Courtney
and Moore, 1971; Moore et al., 1973; Birnbaum et al., 1985; Weber et al.f 1985;
Abbott et al., 1987a; 1988b). Hyperplasia of the ureteric lumenal epithelium
results in ureteric obstruction. Therefore, the TCDD-induced kidney malformation
in the mouse is a true hydronephrosis in that blockage of urine flow results in
back pressure damaging or destroying the renal papilla (Abbott et al., 1987a).
When dissected on gestational day 12 from control embryos, isolated ureters
exposed to 1x10"^ M TCDD in vitro display evidence of epithelial cell hyper-
plasia (Abbott and Birnbaum, 1990c). This is significant in that it shows that
the hydronephrosis response is due to a direct effect of TCDD on the ureteric
epithelium. Embryonic cell proliferation within the ureter may be regulated by
the actions of growth factors, including EGF (Abbott and Birnbaum, 1990c). In
control ureteric epithelia the expression of EGF receptors decreases with
advancing development, whereas after TCDD exposure the rate of ^H-thymidine
incorporation and EGF receptor number do not decline. Therefore, in TCDD-treated
mice there is a correlation between excessive proliferation of ureteric
epithelial cells and increased expression of EGF receptors.
Other effects of TCDD on the developing kidney involve changes in the
extracellular matrix components and basal lamina (Abbott et al., 1987b). In
TCDD exposed fetal kidneys extracellular matrix fibers are of a diameter
consistent with Type III collagen similar to such fibers in unexposed fetal
kidneys. However, the abundance of these Type III collagen fibers is reduced by
TCDD treatment. In the developing kidney these collagen fibers are associated
with undifferentiated mesenchymal cells. Similarly, the expression of
fibronectin, which is also associated with undifferentiated mesenchymal cells is
decreased by TCDD exposure. In the glomerular basement membrane the distribution
of laminin and Type IV collagen is altered by TCDD exposure. These changes in
the glomerular basement membrane may affect the functional integrity of the
filtration barrier, and could exacerbate the hydronephrosis and hydroureter. The
proteins within the extracellular matrix and basal lamina that are altered by
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TCDD exposure, laminin, fibronectin and collagen are considered markers of a
commitment to differentiate into epithelial structures. In the mouse em-
bryo/fetus TCDD exposure also blocks differentiation within the epithelium of the
developing palate.
5.3.2.2.2. Evidence for an Ah Receptor Mechanism
5.3.2.2.2.1. Genetic — With respect to involvement of the Ah locus in
TCDD-induced hydronephrosis very few genetic studies have been done. Prior to
discovery of the Ah locus, however, Courtney and Moore (1971) reported a 62%
incidence of hydronephrosis in C57BL/6 mice exposed to a maternal TCDD dose of
3 /^g/kg/day on days 6-15 of gestation, whereas the incidence in similarly exposed
DBA/2 mice was only 26%. More recently, Silkworth et al. (1989) reported that
when TCDD is administered on gestational days 6-15 the incidence of hydro-
nephrosis is dose related. As the maternal dose of TCDD is increased from 0.5-4
pg/kg/day the incidence of hydronephrosis in C57BL/6 mice increases from 31-92%,
whereas in DBA/2 mice the incidence varies from 5-37% over the same dose range.
In DBA/2 mice the incidence of hydronephrosis increases to 60% when the largest
dose of TCDD administered is doubled to 8 pg/kg/day (but does not reach the 92%
level seen in C57BL/6 mice at 4 pg TCDD/kg). Thus, the incidence of hydronephro-
sis is higher in the mouse strain that produces high affinity Ah receptors
(C57BL/6) compared to that strain (DBA/2) which produces Ah receptors having
lower ligand binding affinity (Okey et al., 1989). The largest dose of TCDD used
in these experiments resulted in hydronephrosis of the fetus without affecting
the mean body weight or body weight gain of the dam. In the BXD strains (Hassoun
et al., 1984a) the incidence of 2,3,7,8-TCDF-induced hydronephrosis is 34-48% in
eight strains with high affinity Ah receptors and 3-4% in eight strains with low
affinity Ah receptors. These results obtained in the BXD stains of mice provide
the best evidence currently available of an association between the ability of
TCDD-like congeners to induce hydronephrosis and the wild-type Ahb allele. Thus,
the Ah locus and the Ah receptor are involved in the hydronephrosis that is
induced by TCDD-like congeners.
5.3.2.2.2.2. Structure Activity — The rank order of potencies for various
halogenated aromatic hydrocarbon congeners to cause hydronephrosis in mice is
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consistent with the structure-activity requirements for binding to the Ah
receptor (Table 5-4). This provides further evidence that the Ah receptor
mediates the effects of these TCDD-like congeners on the developing mouse kidney.
5.3.2.2.3. Species Differences. Hydronephrosis has been reported after
administration of low maternal doses of TCDD to rats and hamsters. Possibly due
to the small numbers of fetuses examined; the observed incidences of hydro-
nephrosis in rats after exposure to cumulative maternal doses >5 /jg TCDD/kg have
not been statistically significant (Courtney and Moore, 1971; Giavini et. al.,
1983). On the other hand, following a 1.5 \iq TCDD/kg dose administered on
gestational days 7 and 9, the incidence of hydronephrosis in hamster fetuses was
11% and 4.2% respectively. This is in contrast to an incidence of <1% in control
hamster fetuses. Accordingly, in hamsters hydronephrosis is one of the most
sensitive indicators of prenatal toxicity (Olson and McGarrigle, 1991).
5.3.3. Postnatal Effects
5.3.3.1. MALE REPRODUCTIVE SYSTEM OF RATS — Since TCDD can decrease
plasma androgen concentrations and be transferred from mother to young in utero
and during lactation (Moore et al., 1976; Van den Berg et al., 1987), it is
expected to have a great impact on the male reproductive system during early
development (Mably et al., 1991). Testosterone and/or its metabolite DHT are
essential prenatally and/or early postnatally for imprinting and development of
accessory sex organs (Chung and Raymond, 1976; Rajfer and Coffey, 1979; Coffey,
1988) and for initiation of spermatogenesis (Steinberger and Steinberger, i989).
In addition, aromatization of testosterone to 17/3-estradiol within the CNS is
required perinatally for the imprinting of typical adult male patterns of
reproductive behavior (Gorski, 1974) and LH secretion (Barraclough, 1980). Thus,
normal development of male reproductive organs and imprinting of typical adult
sexual behavior patterns require sufficient testosterone be secreted by the fetal
and neonatal testis at critical times in early development before and shortly
after birth (MacLusky and Naftolin, 1981; Wilson et al., 1981).
5.3.3.1.1. Perinatal Androgen Deficiency. To determine if in utero and
lactational exposure to TCDD produces a perinatal androgenic deficiency, Mably
et al. (1991, 1992a) dosed pregnant rats with 1.0 /jg TCDD/kg on day 15 of
gestation. Plasma testosterone concentrations were greater in control male than
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in control female fetuses on days 17-21 of gestation, particularly during the
prenatal testosterone surge (days 17-19). On days 18-21 of gestation, TCDD
exposure reduced the magnitude of this sex-based difference. Postnatally, plasma
testosterone concentrations peaked 2 hours after birth in control males, whereas
in TCDD-exposed males, the peak did not occur until 4 hours after birth and was
only half as large. Thus, in male rats perinatal exposure to TCDD can produce
both prenatal and early postnatal androgenic deficiencies.
5.3.3.1.2. Overt Toxicity Assessment. To determine how the male
reproductive system is affected by in utero and lactational TCDD exposure, Mably
et al. (1991, 1992a,b,c) treated pregnant rats with a single oral dose of TCDD
(0.064, 0.16, 0.4 or 1.0 pg/kg) or vehicle on day 15 of gestation (day 0 - sperm
positive). Day 15 was chosen because most organogenesis in the fetus is complete
by this time and the hypothalamic/pituitary/testis axis is just beginning to
function (Warren et al., 1975; 19S4; Aubert et al., 1985). The pups were weaned
21 days after birth. The consequences of this single, maternal TCDD exposure for
the male offspring were characterized at various stages of postnatal sexual
development.
Mably et al. (1992a) found that TCDD treatment had no effect on daily feed
intake during pregnancy and the first 10 days after delivery, nor did it have an
effect on the body weight of dams on day 20 of gestation or on days 1, 7, 14 or
21 postpartum. Treating dams with graded doses of TCDD on day 15 of gestation
had no effect on gestation index, length of gestation or litter size. Except for
an 8% decrease at the highest maternal dose, TCDD had no effect on live birth
index. Neither the 4-day nor 21-day survival index was significantly affected
by TCDD. In all dosage groups, the number of dead offspring was equally
distributed between males and females and of the females that failed to deliver
litters, none were pregnant. Signs of overt toxicity among the offspring were
limited to the above mentioned 8% decrease in live birth index (highest dose
only), initial 10-15% decreases in body weight (two highest doses) and initial
10-20% decreases in feed intake (measured for males only, two highest doses).
The latter two effects disappeared by early adulthood, after which the body
weights of the maternally exposed and nonexposed rats were similar. No male or
female offspring with gross external malformations were found.
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5.3.3.1.3. Androgenic status. Androgenic status of the male offspring
which includes such parameters as plasma androgen concentrations and androgen-
dependent structures and functions, was reduced by a single maternal TCDD dose
as low as 0.16 jug/kg. Anogenital distance, which is dependent on both
circulating androgen concentrations and androgenic responsiveness (Neumann et
al., 1970), was reduced in 1- and 4-day-old male pups, even when slight decreases
in body length were considered. Testis descent, an androgen-mediated develop-
mental event that normally occurs in rats between 20 and 25 days of age (Rajfer
and Walsh, 1977), was delayed £1.7 days.
For accessory sex organs of an adult male rat to grow normally and respond
fully to androgens, there is a critical period which starts before birth and
lasts until sexual maturity during which adequate concentrations of androgens are
necessary (Desjardins and Jones, 1970; Chung and Ferland-Raymond, 1975; Chung and
Raymond, 1976; Rajfer and Coffey, 1979; Coffey, 1988). To determine if perinatal
TCDD exposure affects postnatal growth of the accessory sex organs, one rat from
each litter was sacrificed at 32, 49, 63 and 120 days of age, corresponding to
juvenile, pubertal, postpubertal and mature stages of sexual development,
respectively. At each developmental stage dose-related decreases in seminal
vesicle and ventral prostate weights were found. These decreases could not be
explained by decreases in body weight.
There were trends (though not statistically significant) for plasma
testosterone and DHT concentrations to be decreased at these times, while plasma
LH concentrations were generally unaffected. An exception was a 95% decrease in
plasma LH concentration on postnatal day 32 caused by a maternal TCDD dose of 1.0
pg/kg. The lowest maternal TCDD dose to affect a parameter of androgenic status
was the lowest dose tested - 0.064 pg/kg. This dose resulted in a significantly
depressed ventral prostate weight at 32 days of age. The reductions in seminal
vesicle and ventral prostate weights may be due to modest reductions in plasma
androgen concentrations and/or androgen responsiveness caused by incomplete
perinatal imprinting of the accessory sex organs (Mably et al., 1992a).
Collectively, these results demonstrate that in utero and lactational TCDD
exposure decreases androgenic status of male rats from the fetal stage into
adulthood. Table 5-5 summarizes these effects (Mably et al., 1991, 1992a).
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TABLE 5-5
Effects of In Utero and Lactational TCDO Exposure on Indices of Androgem'c Status8
Index
Anogenital distance
Time to test is descent
Plasma testosterone concentration
Plasma 5a-dihydrotestosterone
concentration
Plasma LH concentration
Seminal vesicle weight
Ventral prostate weight
Lowest Effective Maternal Dose
(*g TCOD/kg)D
0.16 (days 1 and 4)
0.16
NS
NS
1.0 (day 32)
0.16 (days 32 and 63)
0.064 (day 32)
Maximum Effect6
21X decrease (day 1)
1.7 day delay
69X decrease (day 32)
59X decrease (day 49)
95X decrease (day 32)
56X decrease (day 49)
60X decrease (day 32)
"Source: Mably et al. 1991 and 1992a
The lowest dose of TCDD (given on day 15 of gestation) that caused a significant (p<0.05) effect in
the male offspring and the day or days at which this dose caused such an effect are shown.
cThe magnitude of the greatest change seen in response to maternal dosing with 1.0 fig TCDD/kg and the
day at which this effect was seen are shown.
NS = not statistically significant
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5.3.3.1.4. Spermatogenesis. Mably et al. (1991, 1992c) found that
decreased spermatogenesis was among the most sensitive responses of the male rat
reproductive system to perinatal TCDD exposure. Testis and epididymis weights
and indices of spermatogenesis were determined on postnatal days 32, 49, 63 and
120. Perinatal TCDD exposure caused dose-related decreases in testis and
epididymis weights. Weights of the caudal portion of the epididymis where mature
sperm are stored prior to ejaculation were decreased the most, by -45%. The
number of sperm per cauda epididymis was decreased by 75% and 65% on days 63 and
120, respectively, and appeared to be the most sensitive effect of perinatal TCDD
exposure on the male reproductive system. Daily sperm production was decreased
by £43% at puberty, day 49, but the decrease was less at sexual maturity, day
120. Seminiferous tubule diameter was decreased at all four developmental
stages. Each effect of TCDD was dose-related and in all cases a significant
decrease was seen in response to the lowest maternal TCDD dose tested, 0.064
/jg/kg, during at least one stage of sexual development. In general, the
magnitude of the decreases recovered with time, though not completely, suggesting
that perinatal TCDD exposure delays sexual maturation. These results are
summarized in Table 5-6 (Mably et al., 1991, 1992c).
Severe preweaning and/or post-weaning undernutrition can affect the
reproductive system of adult male rodents, including decreased spermatogenesis
(Ghafoorunissa, 1980; Jean-Faucher et al., 1982a,b; Glass et al., 1986).
However, reductions in sex organ weights, epididymal sperm reserves and
spermatogenesis occurred at the two lowest maternal TCDD doses, neither of which
reduced feed intake or body weight of the male offspring. Only at the highest
TCDD doses did modest decreases in feed consumption and body weight occur that
could contribute to these reproductive system effects (Mably et al., 1992a,c).
Thus, undernutrition cannot account for the decreases in spermatogenesis observed
at the lower maternal doses of TCDD.
Since FSH and testosterone are essential for quantitatively normal
spermatogenesis (Steinberger and Steinberger, 1989), an alternative explanation
for the decreases in daily sperm production is a decrease in FSH and/or
testosterone levels. In rats, the duration of spermatogenesis is 58 days (Blazak
et al., 1985; Amann, 1986; Working and Hurtt, 1987), so the decreases in plasma
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TABLE 5-6
Effects of In Utero and Lactational TCOD Exposure on Indices of
Spermatogenic Function and Reproductive Capability"
Index
Test is weight
Epididymis weight
Cauda epididymis weight
Sperm per cauda epididymis
Daily sperm production rate
Seminiferous tubule diameter
Plasma FSH concentration
Leptotene spermatocyte: Sertoli cell
ratio
Sperm mot ili ty; percentage abnormal
sperm
Fertility
Gestation index; litter size; live
birth index; pup survival
Lowest Effective Maternal Dose
<#g TCDD/kg)B
0.40 (days 32)
0.064 (days 49, 120)
0.064 (days 63, 120)
0.064 (days 63, 120)
0.064 (days 63, 120)
0.064 (day 32, 49, 120)
0.40 (day 32)
NS
NS
NS
NS
Maximum Effect0
17X decrease (day 32)
35X decrease (day 32)
53X decrease (day 63)
75X decrease (day 63)
43X decrease (day 49)
1SX decrease (day 32)
1SX decrease (day 32)
no dose-related effects
no dose-related effects
22X decrease (day 70)
no dose-related effects
"Source: Mably et al. 1991 and 1992c
The lowest dose of TCOD (given on day 15 of gestation) that caused a significant (p<0.05) effect in
the male offspring and the day or days at which this dose caused such an effect are shown.
The magnitude of the greatest change seen in response to maternal dosing with 1.0 pg TCDD/kg and the
day at which this effect was seen are shown.
NS = not statistically significant
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FSH concentrations in 32-day-jbld male offspring could contribute to the
reductions of spermatogenesis when the rats were 49 and 63 days of age. However,
the modest depressant effect of perinatal TCDD exposure on plasma FSH concentra-
tions was transitory; no effect was found on plasma FSH levels when the offspring
were 49, 63 and 120 days old. It was concluded that reduced spermatogenesis in
120-day-old male rats, perinatally exposed to TCDD, is not due to decreases in
plasma FSH levels when the animals were 49-120 days of age (Mably et al., 1992c).
Plasma testosterone concentrations in the same rats were reduced £69% by
perinatal TCDD exposure, yet intratesticular testosterone concentrations must be
reduced by at least 80% in rats before spermatogenesis is impaired (Zirkin et
al., 1989). Based on the magnitude of the reductions in plasma androgen
concentrations, it was concluded that corresponding reductions in testicular
testosterone production in perinatal TCDD-exposed offspring would probably not
be severe enough to impair spermatogenesis (Mably et al., 1992a,c).
In normal rats, daily sperm production does not reach a maximum until
100-125 days of age (Robb et al., 1978), but in rats perinatally exposed to TCDD
it takes longer for sperm production to reach the adult level. Furthermore,
length of the delay is directly related to maternal TCDD dose (Mably et al.,
1992c), and if the dose is high enough, the reduction in spermatogenesis may be
permanent. This is suggested by a maternal TCDD dose of 1.0 pg/kg decreasing
daily sperm production in male rat offspring that are 300 days of age (Moore et
al., 1992). Since the mechanism by which perinatal TCDD exposure decreases
spermatogenesis in adulthood is unknown, it is unclear whether the irreversible
effect at the largest maternal dose, 1 pg/kg, is caused by the same mechanism as
that at smaller maternal doses from which the male offspring may eventually
recover.
A key observation for postulating mechanisms by which perinatal TCDD
exposure reduces spermatogenesis in adulthood is the finding that the ratio of
leptotene spermatocytes per Sertoli cell in the testes of 49-, 63- and 120-day-
old rats is not affected by in utero and lactational TCDD exposure even though
daily sperm production is reduced (Mably et al., 1992c). Since Sertoli cells
provide spermatogenic cells with functional and structural support (Bardin et
al., 1988) and the upper limit of daily sperm production in adult rats is
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directly dependent on the number of Sertoli cells per testis (Russell and
Peterson, 1984), three possible mechanisms for the decrease in daily sperm
production may be involved. TCDD could increase the degeneration of cells
intermediate in development between leptotene spermatocytes and terminal stage
spermatids (the cell type used to calculate daily sperm production); decrease
post-leptotene spermatocyte cell division (meiosis); and/or decrease the number
of Sertoli cells per testis (Orth et al., 1988). Elucidating the mechanism by
which perinatal TCDD exposure decreases spermatogenesis is important because it
is one of the most sensitive responses of the male reproductive system to TCDD.
5.3.3.1.5. Epididymis. The epididymis has two functions: in proximal
regions, spermatozoa mature gaining the capacity for motility and fertility,
whereas in distal regions mature sperm are stored before ejaculation (Robaire and
Hermo, 1989). Mably et al. (1991, 1992c) found that motility and morphology of
sperm taken from the cauda epididymis on postnatal days 63 and 120 were
unaffected by perinatal TCDD exposure. Thus, no effect of TCDD on epididymal
function was detected. The dose-dependent reduction in epididymis and cauda
epididymis weights in postpubertal rats, 63 and 120 days old, can be accounted
for, in part, by decreased sperm production. However, in immature males, 32 and
49 days of age, where sperm are not present in the epididymis, the decrease in
weights of epididymal tissue cannot be explained by effects on sperm production.
Since epididymal growth is androgen dependent, a TCDD-induced androgenic
deficiency and/or decrease in androgen responsiveness of the epididymis, could
account for decreased size of the organ (Setty and Jehan, 1977; Dhar and Setty,
1990).
5.3.3.1.6. Reproductive Capability. To assess reproductive capability,
male rats born to dams given TCDD (0.064, 0.16, 0.40 or 1.0 pg/kg) or vehicle on
day 15 of gestation were mated with control virgin females when the males were
-70 and 120 days of age (Mably et al., 1991, 1992c). Fertility index of the
males is defined as number of males impregnating females divided by number of
males mated. The two highest maternal TCDD doses decreased fertility index of
the male offspring by 11% and 22%, respectively. However, these decreases were
not statistically significant, and at lower doses, the fertility index was not
reduced. Gestation index, defined as the percentage of control dams mated with
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TCDD-exposed males that delivered at least one live offspring was also not
affected by perinatal TCDD exposure. With respect to progeny of these mating*,
there was no effect on litter size, live birth index, or 21-day survival index.
When perinatal TCDD-exposed males were mated again at 120 days of age, there was
no effect on any of these same parameters. Thus, despite pronounced reductions
in cauda epididymal sperm reserves, when the TCDD-treated males were mated,
perinatal TCDD exposure had little or no effect on fertility of male rats or on
survival and growth of their offspring. These results are summarized in Table
5-6 (Mably et al., 1991, 1992c).
Since rats produce and ejaculate 10 times more sperm than are necessary for
normal fertility and litter size (Aafjes et al., 1980; Amann, 1982), the absence
of a reduction in fertility of male rats exposed perinatally to TCDD is not
inconsistent with the substantial reductions in testicular spermatogenesis and
epididymal sperm reserves. In contrast, reproductive efficiency in human males
is very low; number of sperm per ejaculate being close to that required for
fertility (Working, 1988). Thus, a percent reduction in daily sperm production
in humans, similar in magnitude to that observed in rats (Mably et al., 1991,
1992c) may reduce fertility in men.
5.3.3.1.7. Sexual Differentiation of the CNS. Sexual differentiation of
the CNS is dependent on the presence of androgens during early development. In
rats the critical period of sexual differentiation extends from late fetal life
through the first week of postnatal life (MacLusky and Naftolin, 1981). In the
absence of adequate circulating levels of testicular androgen during this time,
adult rats display high levels of feminine sexual behavior (e.g., lordosis), low
levels of masculine sexual behavior and a cyclic (i.e., feminine) pattern of LH
secretion (Gorski, 1974; Barraclough, 1980). In contrast, perinatal androgen
exposure of rats will result in the masculinization of sexually dimorphic neural
parameters including reproductive behaviors, regulation of LH secretion and
several morphological indices (Raisman and Field, 1973; Gorski et al., 1978).
The mechanism by which androgens cause sexual differentiation of the CNS is not
completely understood. In the rat, it appears that 170-estradiol, formed by the
aromatization of testosterone within the CNS, is one of the principal active
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steroids responsible for mediating sexual differentiation (McEwen, 1978);
however, androgens are also involved.
5.3.3.1.7.1. Demasculxnization of Sexual Behavior — Mably et al. (1991,
1992b) assessed sexually dimorphic functions in male rats born to dams given
graded doses of TCDD or vehicle on day 15 of gestation. Masculine sexual
behavior was assessed in male offspring at 60, 75 and 115 days of age by placing
a male rat in a cage with a receptive control female and observing the first
ejaculatory series and subsequent post-ejaculatory interval (Table 5-7). The
number of mounts and intromissions (mounts with vaginal penetration) before
ejaculation were increased by a maternal TCDD dose of 1.0 jug/kg. The same males
exhibited 12- and 11-fold increases in mount and intromission latencies,
respectively, and a 2-fold increase in ejaculation latency. All latency effects
were dose-related and significant at a maternal TCDD dose as low as 0.064 fjg/kg
(intromission latency) and 0.16 ^g/kg (mount and ejaculation latencies).
Copulatory rates (number of mounts + intromissions/time from first mount to
ejaculation) were decreased to less than 43% of the control rate. This effect
on copulatory rates was dose-related, and a statistically significant effect was
observed at maternal TCDD doses as low as 0.16 ^g/kg. Post-ejaculatory intervals
were increased 35% above the control interval and a statistically significant
effect was observed at maternal doses of TCDD as low as 0.40 ng/kg. Collective-
ly, these results demonstrate that perinatal TCDD exposure demasculinizes sexual
behavior.
Since perinatal exposure to a maternal TCDD dose of 1.0 ng/kg has no effect
on the open field locomotor activity of adult male rats (Schantz et al., 1991),
the increased mount, intromission and ejaculation latencies appear to be specific
for these masculine sexual behaviors, not secondary to a depressant effect of
TCDD on motor activity. Postpubertal plasma testosterone and DHT concentrations
in litter mates of the rats evaluated for masculine sexual behavior were as low
as 56% and 62%, respectively, of control (Mably et al., 1991, 1992a). However,
plasma testosterone concentrations which were only 33% of control are still
sufficient to masculinize sexual behavior of adult male rats (Demassa et al.,
1977). Therefore, the modest reductions in adult plasma androgen concentrations
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TABLE 5-7
Effects of In Utero and Lactations I TCDD Exposure on Indices of Sexual
Behavior and Regulation of LH Secretion in Adulthood0
Index
Lowest Effective Maternal Dose
(jig TCDD/kg)D
Maximum Effect0
MASCULINE SEXUAL BEHAVIOR**
Mount latency
Intromission latency
Ejaculatory latency
Number of mounts
Number of intromissions
Copulatory rate (mounts plus
i nt romi ss i ons/mi nute
Post-ejaculatory interval
FEMININE SEXUAL BEHAVIOR6
Lordosis quotient
Lordosis intensity score
0.16
0.064
0.16
0.064
1.0
0.16
0.40
1200X increase
1100X increase
97% increase
130% increase
38% increase
43% decrease
35% increase
0.16
0.40
300% increase
50% increase
REGULATION OF LH SECRETION
LH surge
0.40
460% increase"
aSource: Mably et al., 1991 and 1992b
bThe lowest dose of TCDD (given on day 15 of gestation) that caused a significant (p<0.05) effect in
the male offspring is shown.
°The magnitude of the greatest change seen in response to maternal dosing with 1.0 ng TCDD/kg is shown
(average of three trials for masculine behavior and two for feminine.
Measured when the rats were -60, 75 and 115 days of age.
8Feminine sexual behavior was measured following castration, estrogen priming and progesterone
administration. The rats were 170-184 days old.
Number of times lordosis was displayed in reponse to a mount, divided by the number of times each rat
was mounted, times 100.
°Since control males do not secrete LH in response to progesterone, this percentage was calculated by
comparing peak plasma LH concentrations in TCDD exposed rats with plasma LH concentrations in control
males at the same time after progesterone was administered.
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following perinatal TCDD exposure were not of sufficient magnitude to
demasculinize sexual behavior.
Reductions in perinatal androgenic stimulation can inhibit penile
development and subsequent sensitivity to sexual stimulation in adulthood
(Nadler, 1969; Sodersten and Hansen, 1978). Therefore, the demasculinization of
•exual behavior could, to some extent, be secondary to decreased androgen-
dependent penile development. However, perinatal TCDD exposure had no effect on
gross appearance of the rat penis. In addition, TCDD-exposed males exhibited
deficits in such masculine sexual behaviors as mount latency and post-ejaculatory
interval which do not depend on stimulation of the penis for expression (Sach»
and Barfeld, 1976). Thus, while some effects of TCDD, such as decreased
copulatory rate and prolonged latency until ejaculation, could be due to reduced
sensitivity of the penis to sexual stimulation, the 12-fold increase in mount
latency and increase in post-ejaculatory interval cannot be explained by this
mechanism.
5.3.3.1.7.2. Feminization of Sexual Behavior — Mably et al. (1991, 1992b)
determined if the potential of adult male rats to display feminine sexual
behavior was altered by perinatal TCDD exposure. Male offspring of dams treated
on day 15 of gestation with various doses of TCDD up to 1 pg/kg or vehicle were
castrated at -120 days of age and beginning at -160 days of age were injected
weekly for 3 weeks with 17/J-estradiol benzoate, followed 42 hours later by
progesterone. Four to 6 hours after the progesterone injection on weeks 2 and
3, the male was placed in a cage with a sexually excited control stud male. The
frequency of lordesis in response to being mounted by the stud male was increased
from 18% (control) to 54% by the highest maternal TCDD dose, 1.0 ^9/kg (Table 5-
7). Lordosis intensity scored afte% Hardy and DeBold (1972) as a (1) for light
lordosis, (2) for moderate lordosis and (3) for a full spinal dorsiflexion was
increased in male rats by perinatal TCDD exposure. loth effects on lordosis
behavior in males were dose-related and significant at maternal TCDD doses as low
as 0.16 pg/kg (increased lordotic frequency) and 0.40 A/g/kg (increased lordotic
intensity). Together they indicate a feminization of sexual behavior in these
animals. Although severe undefnutrition from 5-45 days after birth potentiates
the display of lordosis behavior in adult male rats (Forsberg et al., 1985) the
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increased frequency of lordotic behavior was seen at a maternal TCDD dose, 0.16
pg/kg, which had no effect on feed intake or body weight. It was concluded that
perinatal TCDD exposure feminize sexual behavior in adult male rats independent
of undernutrition.
5.3.3.1.7.3. Feminization of LH Secretion Regulation — The effect of
perinatal TCDD exposure on regulation of LH secretion by ovarian steroids was
determined in male offspring at -270 days of age. There is normally a distinct
sexual dimorphism to this response. In rats castrated as adults, estrogen-primed
females greatly increase their plasma LH concentrations when injected with
progesterone, whereas similarly treated males fail to respond (Taleisnik et al.,
1969). Progesterone had little effect on plasma LH concentrations in estrogen-
primed control males, but significant increases were seen in males exposed to
maternal TCDD doses as low as 0.40 jug/kg. Thus, perinatal TCDD exposure
increases pituitary and/or hypothalamic responsiveness of male rata to ovarian
steroids in adulthood indicating that regulation of LH secretion is permanently
feminized. Table 5-7 summarizes sexual behavior and LH secretion results (Mably
•t al., 1991, 1992b).
5.3.3.1.7.4. Comparison to Other Ah-Receptor Mediated Responses — The
induction of hepatic P-4501A1 and its associated EROD activity are extremely
sensitive Ah receptor-mediated responses to TCDD exposure. Yet in 120-day-old
male rats that had been exposed to TCDD perinatally, alterations in sexual
behavior, LH secretion and spermatogenesis were observed when induction of
hepatic EROD activity could no longer be detected (Mably et al., 1991,
1992a,b,c). These results suggest that TCDD affects sexual behavior, gonado-
trophic function and spermatogenesis when virtually no TCDD remains in the body,
and that the demasculinization and feminization of sexual behavior, feminization
of LH secretion and reduced spermatogenesis caused by in utero and lactational
exposure to TCDD may be irreversible (Mably et al., 1992b,c).
5.3.3.1.7.5. Possible Mechanisms and Significance — The most plausible
explanation for the demasculinization of sexual behavior and feminization of
sexual behavior and LH secretion is that perinatal exposure to TCDD impairs
sexual differentiation of the CNS. Neither undernutrition, altered locomotor
activity, reduced sensitivity of the penis to sexual stimulation nor modest
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reductions in adult plasma androgen concentrations of the male offspring can
account for this effect (Mably et al., 1992b). On the other hand, exposure of
the developing brain to testosterone, conversion of testosterone into 1713-eetra-
diol within the brain, and events initiated by the binding of 17fl-estradiol to
its receptor are all critical for sexual differentiation of the CNS and have the
potential to be modulated by TCDD. if TCDD interferes with any of these
processes during late gestation and/or early neonatal life it could irreversibly
demasculinize and feminize sexual behavior (Hart, 1972; McEwen et al., 1977;
Whalen and Olsen, 1981, and feminize the regulation of LH secretion (Gogan et
al., 1980, 1981) in male rats in adulthood.
Treatment of dams on day 15 of gestation with 1.0 ^g TCDD/kg significantly
decreases plasma testosterone concentrations in male rat fetuses on days 18 and
20 of gestation and in male rat pups 2 hours postpartum (Mably et al., 1992a).
Thus, the ability of maternal TCDD exposure to reduce prenatal and early
postnatal plasma testosterone concentrations can account, in part, for the
impaired sexual differentiation of male rats exposed perinatally to TCDD. other
mechanisms which may potentially contribute to the TCDD-induced impairment in CNS
sexual differentiation are: a decrease in the formation of 17/J-estradiol from
testosterone within the CNS that is independent of the decrease in plasma
testosterone concentrations and/or a reduction in responsiveness of the CNS to
estrogen during the critical period of sexual differentiation. The latter
mechanism is consistent with the Ah receptor-mediated anti-estrogen action of
TCDD described above for rat and mouse uterus and for estrogen responsive MCF-7
and Hepa Iclc7 cells.
In utero and/or lactational exposure to TCDD may cause similar effects in
other animal species, including nonhuman primates (Pomerantz et al., 1986;
Thorton and Goy, 1986; Goy et al., 1988), in which sexual differentiation is
under androgenic control. In humans social factors account for much of the
variation in sexually dimorphic behavior; however, there is evidence that
prenatal androgenization influences both the sexual differentiation of such
behavior and brain hypothalamic structure (Erhardt and Meyer-Bahlburg, 1981;
Hines, 1982; Levay, 1991).
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5.3.3.1.8. Relative Sensitivity. The male reproductive system in rats is
-100 times more susceptible to TCDD toxicity when exposure occurs perinatally
(ED5Q 0.16 pg/kg) rather than in adulthood (ED50 15 /ug/kg). To illustrate this
sensitivity, a single maternal TCDD dose as low as 0.064 /jg/kg given on day 15
of gestation significantly decreases epididymis and cauda epididymis weights,
cauda epididymal sperm numbers and daily sperm production in male offspring at
various stages of sexual development. Decreases in ventral prostate weights in
32-day-old male offspring and in older males increases in the number of mounts
preceding ejaculation and increases in intromission latency are also produced by
maternal TCDD doses as low as 0.064 pg/kg. The 0.064 /jg TCDD/kg dose is not
maternally toxic and produces no signs of overt toxicity in male or female
offspring. Other effects of perinatal exposure on the male reproductive system
were detected at a maternal TCDD dose of 0.16 pg/kg or higher (Mably et al.,
1991, 1992a,b,c).
In adult rats, the most sensitive toxic responses to TCDD have been observed
following long term, low level exposure. In a 3-generation reproduction study,
Murray et al. (1979) reported that dietary administration of TCDD at doses as low
as 0.01 /ug/kg/day significantly affected reproductive capacity in female rats
with no effects seen at 0.001 /jg/kg/day (NOAEL). The same NOAEL was found in a
2-year chronic toxicity and oncogenicity study in which an increased incidence
of certain types of neoplasms was altered among rats given TCDD doses of 0.01 or
0.1 pg/kg/day (Kociba et al., 1978). Based on the pharmacokinetics of TCDD in
the rat (Rose et al., 1976), the steady-state body burden of TCDD in these rats
that were chronically dosed (>90 days) with either 0.01 or 0.001 jug TCDD/kg/day
is approximately 0.29 jug/kg (LOAEL) and 0.029 jug/kg (NOAEL), respectively. Yet,
Nably et al. (1991, 1992a,b,c) found that a single TCDD dose of 0.064 pg/kg given
on day 15 of gestation produces a number of statistically significant effects on
the reproductive system of male rat offspring. Since 0.064 pg TCDD/kg was the
lowest dose tested, a NOAEL for developmental male reproductive toxicity, defined
as the lowest dose used that had no statistically significant effect, could not
be determined by Mably et al. (1991, 1992a,b,c). It is concluded that
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developmental effects on spermatogenesis occur at a maternal TCDD dose that is
lower than any previously shown to produce toxicity in rats.
5.3.3.2. NEUROBEHAVIOR — since differentiated tissues derived from
ectoderm, namely, skin, conjunctiva, nails and teeth are sites of action of
halogenated aromatic hydrocarbons in transplacentally exposed human infants,
another highly differentiated tissue derived from ectoderm, the CNS, should be
considered a potential site of TCDD action. In support of this possibility
sexual differentiation of the CNS of adult male rats is irreversibly altered in
a dose-related fashion by perinatal exposure to TCDD (Mably et al., 1991, 1992b).
As will be shown below, the CNS of mice transplacentally exposed to 3,3',4,4'-
TCB, monkeys perinatally exposed to TCDD and children transplacentally exposed
to a mixture of PCBs, CDFs and PCQs in the Yu-Cheng incident is also affected.
Thus, functional CNS alterations, which may or may not be irreversible, are
observed following perinatal exposure to halogenated aromatic hydrocarbons. Ah
receptors have been identified in brain (Carlstedt-Duke et al., 1979) but may be
associated with glial cells rather than neurons (Silbergeld, 1992). Following
administration of 14C-TCDD in the rat the highest concentrations of TCDD derived-
14
C are found in the hypothalamus and pituitary. Much lower concentrations are
found in the cerebral cortex and cerebellum (Pohjanvirta et al., 1990). No
specific information with respect to the presence of Ah receptors at these sites
is available. Ah receptors appear to be absent in the human frontal cortex
(Silbergeld, 1992).
5.3.3.2.1. Mice. CD-I mice exposed transplacentally to 3,3',4,4'-TCB at
a maternal oral dose of 32 mg/kg administered on days 10-16 of gestation
exhibited neurobehavioral, neuropathological and neurochemical alterations in
adulthood (Tilson et al., 1979; Chou et al., 1979; Agrawal et al., 1981). The
neurobehavioral effects consisted of circling, head bobbing, hyperactivity,
impaired forelimb grip strength, impaired ability to traverse a wire rod,
impaired visual placement responding and impaired learning of a one-way avoidance
task {Tilson et al., 1979). The brain pathology in adult mice exhibiting this
syndrome consisted, in part, of alterations in synapses of the nucleus accumbens
(Chou et al., 1979). This suggested that in utero exposure to 3,3',4,4'-TCB may
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interfere with synaptogenesis of dopaminergic systems. In support of this
possibility, Agrawal et al. (1981) found that adult mice transplacentally exposed
to 3,3',4,4'-TCB had decreased dopamine levels and decreased dopamine receptor
binding in the corpus striatum both of which were associated with elevated levels
of motor activity. It was concluded that transplacental exposure to 3,3',4,4'-
TCB in mice may permanently alter development of striatal synapses in the brain.
Eriksson et al. (1988) examined the neurobehavioral effects of 3,3',4,4'-TCB
in NMRI mice exposed to a single oral dose of 0.41 or 41 mg/kg on postnatal day
10. Following sacrifice of the mice on day 17 muscarinic receptor concentrations
in the brain were significantly decreased, at both dose levels. This effect was
shown to occur in the hippocampus but not in the cortex. More recently (Eriksson
et al., 1991), NMRI mice were exposed to the same two doses of 3,3',4,4'-TCB
similarly administered on postnatal day 10. At 4 months of age, the effects of
the PCB on locomotor activity were assessed. At both dose levels, abnormal
activity patterns were exhibited in that the treated mice were significantly less
active than controls at the onset of testing, but were more active than controls
at the end of the test period. This pattern of effects can be interpreted as a
failure to habituate to the test apparatus. In contrast to the previous results
with CD-I mice, circling or head bobbing activities were not observed in these
animals. Upon sacrifice after the activity testing was complete, a small but
statistically significant decrease in the muscarinic receptor concentration of
the hippocampus was found in animals from the high dose group. These results
suggest that the neurochemical effects of 3,3' ,4,4'-TCB are complex and could
potentially involve cholinergeric as well as dopaminergic systems in the brain.
The main problem in applying the above results to TCDD is that the mechanism
by which 3,3',4,4'-TCB produces these neurobehavioral effects are not known. The
parent compound acting through an Ah receptor might be involved and/or a
neurotoxic metabolite of 3,3',4,4'-TCB could be the causative agent. Until the
mechanism is resolved, dose-response studies are conducted, and other TCDD-like
congeners are evaluated for their ability to produce the effect, the findings
cannot be extrapolated with confidence to TCDD. However, 3,3',4,4'-TCB is an Ah
receptor agonist and other known developmental effects of this congener are
mediated by an interaction of the parent compound with Ah receptors.
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5.3.3.2.2. Honkeys. Schantz and Bowman (1989) and Bowman et al. (l9S9a)
have conducted a series of studies on the long-term behavioral effects of
perinatal TCDD exposure in monkeys. Because these- were the first studies to
evaluate the behavioral teratology of TCDD, monkeys exposed to TCDD via the
mother during gestation and lactation were screened on a broad selection of
behavioral tests at various stages of development (Bowman et al., 1989a). At the
doses studied (5 or 25 ppt in the maternal diet?, TCDD did not affect reflex
development, visual exploration, locomotor activity or fine motor control in any
consistent manner (Bowman et al., 1989b). However, the perinatal TCDD exposure
did produce a specific, replicable deficit in cognitive function (Schantz and
Bowman, 1989). TCDD-exposed offspring were impaired on object learning, but were
unimpaired on spatial learning. TCDD exposure also produced changes in the
social interactions of mother-infant dyads (Schantz et al., 1986). TCDD-exposed
infants spent more time in close physical contact with their mothers. The
pattern of effects was similar to that seen in lead-exposed infants and suggested
that mothers were providing increased care to the TOTD-exposed infants (Schantz
et al,, 1986).
5.3.3.2.3. Humans. The intellectual and behavioral development of
Yu-Cheng children transplacentally exposed to PCBs, CDFs and PCQs was studied
through 1985 by Rogan et al. (1988). In Yu-Cheng children, matched to unexposed
children of similar age, area of residence, and socioeconomic status, there was
a clinical impression of developmental or psychomotor delay in 12 (10%) Yu-Cheng
children compared with 3(3%) control children, and of a speech problem in 8 (7%)
Yu-Cheng children versus 3 (3%) control children. Also except for verbal IQ on
the Wechsler Intelligence Scale for Children, Yu-Cheng children scored lower than
control children on three developmental and cognitive tests (Rogan et al., 1988).
Neurobehavioral data on Yu-Cheng children obtained after 1985, shows that the
intellectual development of these children continues to lag somewhat behind that
of matched control children (Hsu et al., 1991). Also Yu-Cheng children are rated
by their parents and teachers to have a higher activity level, more health, habit
and behavioral problems, and to have a temperamental clustering closer to that
of a "difficult child" (Hsu et al., 1991). it is concluded that in humans
transplacental exposure to halogenated aromatic hydrocarbons can affect CNS
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function postnatally. However, which congeners, TCDD-like versus nonTCDD-like,
are responsible for the neurotoxicity is unknown.
Further research on the mechanism of these postnatal neurobehavioral
effects, dose response relationships, and reversibility of the alterations is
needed before the role of TCDD-like congeners versus nonTCDD-like congeners in
causing this toxicity can be understood. Mechanisms that respond uniquely to
TCDD-like congeners may not necessarily be involved since three lightly
chlorinated, ortho-substituted PCB congeners, 2,4,4'-TCB, 2,2',4,4'-TCB and
2,2',5,5'-TCB, have been detected in monkey brain following dietary exposure to
Aroclor 1016 and appear to be responsible for decreasing dopamine concentrations
in the caudate, putamen, substantia nigra and hypothalamus of these animals
(Seegal et al., 1990). These nonplanar PCB congeners are believed to cause these
effects by acting through a mechanism that does not involve the Ah receptor. On
the other hand, the results presented for mice and monkeys suggest that TCDD-like
congeners could be involved in producing the observed postnatal neurobehavioral
effects in humans.
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