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69
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February 1988
Review Draft
APPENDIX C
REPRODUCTIVE AND DEVELOPMENTAL TOXICITY OF 2,3,7,8-TCDD
G. L. Kimmel, Ph.D.
Reproductive Effects Assessment Group
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
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EXECUTIVE SUMMARY
There is not sufficient evidence to link 2,3,7,8-tetrachlorodibenzo-
p_-dioxin (2,3,7,8-TCDD) to human reproductive or developmental toxicity;
however, it has been shown to be a reproductive and developmental toxicant in
animal studies. Among the effects that have been reported are reduced
fertility, litter size, postnatal survival, and offspring body weight, as well
as an increase in structural malformations. Effects on the male and female
gonads and on the female menstrual/estrous cycle have also been reported.
Reproductive and developmental effects have been observed in a variety of
species, indicating that the toxicity is not a species-specific event.
The studies on reproductive function and fertility remain the basis for
establishing the lowest effective (toxic) exposure level. It appears that a
0.01 ug/kg/day exposure is the lowest effect level that can be supported by the
data, although further analysis of the data may provide some support for a
lower effect level of 0.001 ug/kg/day. In addition, a detailed analysis of
studies in the subhuman primate may also provide support for a lower effect
level of 0.001 ug/kg/day. In relation to developmental toxicity, a large
number of studies in a variety of species has demonstrated that 2,3,7,8-TCDD is
a developmental toxicant. Collectively, these studies indicate that long-term,
low-dose exposure is of concern relative to the potential for altering
reproductive function and fertility. The results also demonstrate that acute
and short-term exposures are effective in causing altered development, and
therefore, should also be of concern.
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INTRODUCTION
2,3,7,8-TCDD has been shown to be a reproductive and developmental
toxicant in animal studies. Among the effects that have been reported are
reduced fertility, litter size, postnatal survival, and offspring body weight,
as well as an increase in structural malformations. 2,3,7,8-TCDD also affects
the male and female reproductive systems. Gonadal dysfunction has been
demonstrated in both sexes, and alterations of normal reproductive cycles have
been reported in the female. Although there have been accidental exposures of
humans to mixtures containing 2,3,7,8-TCDD, there is not sufficient evidence
from the case reports and epidemiologic studies that have been carried out to
date to link 2,3,7,8-TCDD to human reproductive or developmental toxicity (U.S..
EPA, 1986a). Much of the information on human exposure is covered in Appendix
D.
In line with the original request for the development of this review, this
appendix covers the effects of 2,3,7,8-TCDD on the integrity of the
reproductive system and fertility and on prenatal and early postnatal
development. The appendix is not inclusive of all studies on these effects.
Rather, it focuses on the key studies and issues that may go into an overall
risk characterization. The present reviewer has tried to present a balanced
view of the studies and the uncertainties inherent in the data and the
analysis. Studies on other congeners of polychlorinated dibenzo-p_-dioxins or
of 2,3,7,8-TCDD as a mixture or contaminant of other agents are not included,
except as support where appropriate. A more comprehensive review of
2,3,7,8-TCDD's toxicity and its relation to risk characterization can be found
in the Health Assessment Document (HAD) for Polychlorinated Dibenzo-p_-Dioxins
-------
(U.S. EPA, 1985).
ANIMAL STUDIES
REPRODUCTIVE FUNCTION/FERTILITY-
The study by Murray et'al. (1979) continues to be the guidepost for
setting standards of exposure relative to reproduction. The study employs a
multigeneration approach and examines the exposure of male and female rats over
three generations to relatively low levels of 2,3,7,8-TCDD (0, 0.001, 0.01, and
0.1 ug 2,3,7,8-TCDD/kg body weight/day). The analysis of the data is made
somewhat difficult by considerable variation in the fertility index in both
control and exposed groups. In addition, the number of impregnated animals in
the exposed groups was lower than desirable (Palmer, 1981). However, there
were effects that cannot be automatically associated with the variation in the
fertility index, including an increased time between first cohabitation and
delivery, a decrease in litter size, a decrease in the gestational survival
index, and a decrease in postnatal body weight. Specifically, Murray et al.
reported statistically significant changes in several of the measured
parameters, and these are outlined in Table 1.
While there is no dispute over the reproductive toxicity seen in this
study, there is some disagreement over the appropriate effect levels. Murray
et al. (1979) indicated that the lowest statistically significant adverse
effect was observed at 0.01 ug/kg/day and that a no-effect level could be
established at 0.001 ug/kg/day. In a reanalysis of this study, however, Nisbet
and Paxton (1982) argued that the analysis of Murray et al. (1979) was limited
by the statistical approach used. Nisbet and Paxton applied a different
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TABLE 1. SUMMARY OF EFFECTS OF 2,3,7,8-TCDD ON REPRODUCTION
Parameter
Fertility
Litter size
Gestation Survival
Postnatal survival
Postnatal body
weight
Generation
fo
fo
fl
fz
fla
fib
f3
f
fib
fs
fla
fib
?3
fla
fib
•fz
ug TCDD/kg/daya
0.001 0.001
_..
. — —
dec
— dec
• — —
— . — .
— dec
dec
._.
-__ ..
dec dec
— dec
dec dec
inc —
— dec
— ' —
— —
— , —
— dec
— dec
0.1
dec
dec
dec
dec
decb
c
b
c
b
c
a (—), unaffected; dec, decreased; inc, increased.
b No liveborn offspring.
c One litter only.
SOURCE: Murray et al., 1979.
-------
statistical approach which included the pooling of data across all generations,
and their reanalysis indicated that 0.001 ug/kg/day (the lowest dose used) was
an effect level and that a no-effect level could not be set. The authors of the
HAD for Polychlorinated Dibenzo-fi-Dioxins (U.S. EPA, 1985) accepted this
argument and used the Nisbet and Paxton reanalysis to establish a lowest
observed adverse effect level (LOAEL) of 0.001 ug/kg/day. However, the HAD
(U.S. EPA, 1985) also noted that the FIFRA Scientific Advisory Panel (SAP) did
not feel that the effects were consistent enough at 0.001 ug/kg/day and that
this would have to be considered a no observed effect level (NOEL). Since this
latter decision by the SAP was made before the Nisbet and Paxton reanalysis, it
is not possible to know if the decision would have been different in light of
the reanalysis. However, the present reviewer feels that effect levels should
not be set on the basis of the Nisbet and Paxton reanalysis. While it-appears
that Nisbet and Paxton's approach for increasing the limited statistical power
of the Murray et al. (1979) study is appropriate statistically, it is difficult
to see the biological rationale for pooling the data. Litters from different
generations (or from subsequent matings within a generation) are not the same.
They have different histories of exposure and each is ti'ed to the effect of the
agent on its parental generation. Thus, as a general rule, pooling of data
from different generations would seem biologically inappropriate. Unless some
specific exception can be identified, it is not clear how pooling can be
biologically justified in this case.
A limited review of a report by Murray et al. (1978), which served as
Exhibit 77 in a 1980 EPA hearing, has raised some questions relative to the
offspring survival. The Murray et al. (1979) paper included the standard
parameters of the Gestation Survival Index and Postnatal Survival Index as
-------
measures of offspring survival (Table 1). These parameters showed significant
changes, but not always in a dose-related fashion. This may be because
offspring viability is examined during discrete periods of offspring
development and the investigators do not report viability over the entire early
postnatal period. Additional data from the 1978 report by Murray et al. has
been summarily reviewed on the basis of overall offspring survival, i.e., not
separating the Gestation Survival Index and Postnatal Survival Index. There
appears to be a general pattern of decreased survival even at 0.001 ug/kg/day,
if one assumes a survival rate of control offspring of 90%. Appropriate
analytical techniques would have to be applied to confirm this. Two points
should be raised regarding the data. The first is that the number of offspring
used in the calculations of Murray et al. (1978) varied considerably among the
control and two exposure groups. How this could affect the parameters is not
entirely clear to this reviewer, but Bailar (1981) spoke to similar issues in
his testimony and noted that he found the data suggestive of an effect at the
0.001 ug/kg/day level. The second point is the survival of the control
population of offspring. In both the fjjj and the f3 litters, survival of the
controls by postnatal day 21 ranged from 70% to 80%. Although viability varies
within any laboratory animal population, this figure seems low and may account
for there not being an established decrease in offspring viability in these two
groups at 0.001 ug/kg/day. A more detailed analysis of this data base may
provide a clearer indication of the potential for decreased postnatal survival
at the 0.001 level.
In addition to the data on offspring survival, the Murray et al. (1978)
report summarizes their observations on renal pathology. When all observed
effects on the kidney (i.e., "slightly dilated" and "dilated") are combined,
-------
there appears to be an increase at the 0.001 and 0.01 ug/kg/day in the fja and
f}k litters. As has been pointed out, it is not entirely appropriate to
combine these two end points, since slight dilation may be due to delayed
development which may be transient in nature. Nevertheless, the kidney is a
recognized end organ for 2,3,7,8-TCDD effects, and the findings of Moore et al.
(1973) in the mouse indicate that the continuous exposure that is found in a
three-generation study may be more likely to lead to the most obvious effects
on kidney development. Research in this particular area is continuing. Abbott
et al. (1987a, b) recently reported that the kidney alterations that occur in
the mouse following a single, prenatal 12 ug/kg dose of 2,3,7,8-TCDD are
consistent with true hydronephrosis.
Allen and his colleagues examined 2,3,7,8-TCDD effects 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 2,3,7,8-TCDD for up to 9 months. Menstrual cycles and serum
steroid levels were examined. Following 7 months of exposure, the females were
bred. Females exposed to 500 ppt showed obvious clinical signs of 2,3,7,8-TCDD
toxicity and lost weight throughout the study. Five of the eight monkeys died
within one year after exposure was initiated. In a summary of the reproductive
function and fertility of these animals, Allen et al. (1979) reported that
although the menstrual cycle and menstruation were normal, there was a decrease
in serum estradiol and progesterone in five of eight monkeys. Only three of
the animals conceived, and only one was able to carry the pregnancy to term.
Females exposed to 50 ppt 2,3,7,8-TCDD in the diet (Schantz et al., 1979)
showed normal menstrual cycles and serum estradiol and progesterone through 6
months of exposure. When they were bred at. 7 months, four of eight females did
-------
not conceive and two of four that did could not carry the pregnancies to term.
Only two conceptions resulted in normal births.
The results of this series of studies could potentially support a lower
LOAEL than that reported by Murray et al. (1979) in the rat (i.e., 0.01
ug/kg/day). The high dose (500 ppt) resulted in considerable maternal toxicity
and reproductive dysfunction, while a comparable exposure level (0.01
ug/kg/day) in the rat did not produce any significant clinical signs in the
parental generation. This could indicate that the rhesus monkey is more
sensitive to 2,3,7,8-TCDD when exposure occurs over long periods and when
reproductive parameters are the critical end points. The low dose (50 ppt) was
reported as resulting in specific reproductive dysfunction in the absence of
maternal toxicity. Since this exposure level is calculated to be approximately
0.002 ug/kg/day, the report suggests that even lower doses are required for
effects on reproductive function than are required in the rat and would support
a lower adverse effect level. Unfortunately, much of the data on the monkeys
has been presented in abstract form or as part of a review, and consequently, a
critical analysis of the data is impossible. There has been some indication
that studies of even lower levels (i.e., 5 and 25 ppt) showed signs of
reproductive toxicity, and the data are now beginning to appear in the
literature. Schantz et al. (1986) reported altered maternal care of offspring
in monkeys exposed to 5 and 25 ppt for 45 to 49 months, and Bowman et al.
(1987a, b) reported on altered maternal-infant interaction and other
reproductive parameters at the Seventh International Symposium on Chlorinated
Dioxins and Related Compounds. These reports are now being evaluated and, if
supported, could significantly affect the adverse effect levels calculated for
reproductive and developmental toxicity.
8
-------
It is important to note that none of these findings establish an
unequivocal effect at the 0.001 ug/kg/day or below level. However, the
evidence is suggestive enough and the uncertainties are great.enough that it
would seem prudent to consider the 0.001 level as highly suspect. ,
DEVELOPMENT
Numerous studies have been done on the developmental toxicity of 2,3,7,8-
TCDD, many of which have been summarized in the HAD for Polychlorinated
Dibenzo-£-Dioxins (U.S. EPA, 1985). Of particular interest to this review are
those studies which present data that may factor into an overall risk
characterization: Courtney and Moore (1971), Giavini et al. (1982a), Khera and
Ruddick (1973), McNulty (1980), Moore et al. (1973), Smith et al. (1976), and
Sparschu et al. (1971).
Developmental toxicity following exposure to 2,3,7,8-TCDD has been
demonstrated in different species, including the chicken, mouse, rat, rabbit,
ferret, and monkey. Thus, developmental toxicity of 2,3,7,8-TCDD does not
appear to be related to a species-specific metabolic or physiological response
to exposure. While specific responses and effective doses do vary among
species and among strains within a species (Courtney and Moore, 1971; Poland
and Glover, 1980), developmental toxicity in response to 2,3,7,8-TCDD exposure
can be expected to occur in all species.
The exposure range at which developmental toxicity first becomes apparent
is 0.125 to 1.0 ug/kg/day when exposure occurs over a major period of
organogenesis. The no-effect level appears to be approximately 0.1 ug/kg/day
in rats, mice, and rabbits. Giavini et al. (1982a) did note an increase in
extra ribs at 0.1 ug/kg/day in the rabbit. However, the number of ribs
9
-------
normally varies between 12 and 13 in rabbits, and it would require a more
detailed analysis of the data to establish this as a true effect level.
Several laboratories have examined the effects of more acute exposures during
organogenesis. Moore et al. (1973) demonstrated that a single oral dose of
1 ug/kg given to mice on gestation day 10 produced hydronephrosis. In an
interesting extension of this finding, they investigated the postnatal
development of the kidney in cross-fostered offspring following prenatal
exposure. The frequency of hydronephrosis seen postnatally was largely
dependent on whether the offspring were nursed by a dam that had also been
prenatally exposed to 2,3,7,8-TCDD. Thus, it appeared that continued exposure
during the lactation period was required to produce or maintain the greatest
effect on kidney development during the postnatal period studied.
Except for the multigeneration studies, which tend not to critically
evaluate many developmental end points, few studies have been carried out on
developmental periods other than the period of organogenesis. Giavini et al.
(1982b) did examine 2,3,7,8-TCDD exposure on gestation days 1-3 in the rat and
reported possible delays in implantation and some effect on fetal weight and
the kidney. There appear to be no studies on exposure during late prenatal
development. As noted above, Moore et al. (1973) examined effects on the
kidney postnatally following prenatal exposure of the dams, and demonstrated
that transfer in the milk is a likely contributing factor to the developmental
toxicity of 2,3,7,8-TCDD. There have also been reports describing the effects
of 2,3,7,8-TCDD exposure on the developing immune system (see Appendix E).
However, carefully designed studies on postnatal exposures or on changes in
postnatal function following prenatal exposure have generally not been carried
out, making it impossible to evaluate the potential effect of 2,3,7,8-TCDD on.
10
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the developing young animal.
In summary, studies on the developmental toxicity of 2,3,7,8-TCDD have
clearly demonstrated that it is a developmental toxicant in a wide variety of
species at relatively low doses. These studies have also shown that extended
periods of exposure are not necessary for developmental toxicity to result.
Thus, there is a need for concern over acute or short-term exposure to 2,3,7,8-
TCDD. A composite review and summary of the studies on the developmental
toxicity of 2,3,7,8-TCDD is limited by factors such as the simultaneous
occurrence of maternal toxicity, the use of different animal strains and
exposure routes, and in many studies the small number of animals per treatment
group that are included in the final data analysis. In addition, there are
differences in study designs and approaches to data analysis which must be
considered in comparing the studies. These factors do not alter the finding
that 2,3,7,8-TCDD is a developmental toxicant at very low exposure levels.
However, they can potentially affect the final assessment of exposure levels
that can be considered toxic.
MALE AND FEMALE REPRODUCTIVE SYSTEM
Specific components of the male and female reproductive systems are
affected by 2,3,7,8-TCDD exposure. In the male, exposures above 1 ug/kg/day
resulted in evidence of testicular atrophy with destruction of the seminiferous
tubules and spermatogenic cells (Kociba et al., 1976; Norback and Allen, 1973;
McConnell et al., 1978). However, following exposures of 0.001 to 0.1
ug/kg/day over a 2-year period, Kociba et al. (1978) reported that the male
reproductive organs appeared to be unaffected, relative to the controls. In a
study of offspring of male mice exposed to 0.16 to 2.4 ug/kg/day of 2,3,7,8-
11
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TCDD combined with higher'doses of 2,4-dichlorophenoxyacetic acid and
2,4,5-trichlorophenoxyacetic acid, there did not appear to be any effect on
fetal or neonatal development or viability (Lamb et al., 1981). In the female,
exposure to 1 to 2 ug/kg/day for 13 weeks resulted in changes in estrous
cycles, anovulation, and signs of ovarian dysfunction (Kociba et al., 1976).
At exposures of 0.001 to 0.01 ug/kg/day in a 2-year study, Kociba et al. (1978)
reported no effects on the female reproductive system. At 0.1 ug/kg/day, a
decrease in uterine changes such as endometrial hyperplasia were reported. As
noted previously in the Allen et al. studies, female monkeys exposed to 500 ppt
2,3,7,8-TCDD were reported to exhibit changes in their serum steroid levels.
SUMMARY
2,3,7,8-TCDD has been shown to be a reproductive and developmental
toxicant in animal studies. Among the effects that have been reported are
reduced fertility, litter size, postnatal survival, and offspring body weight,
as well as an increase in structural malformations. Effects on the male and
female gonads and on the female menstrual/estrous cycle have also been
reported. The effects have been observed in a variety of species, indicating
that the toxicity is not a species-specific event and can be expected to occur
in all species, including the human.
The studies on reproductive function and fertility remain the basis for
establishing the lowest effective (toxic) exposure level. There is some
disagreement, centered on the appropriate approach for data analysis, over the
effect level based on the Murray et al. (1979) study. Based on the current
information from this study, a 0.01 ug/kg/day level is the lowest effect level
12
-------
that can be supported by the data. However, there is enough suggestive
evidence to indicate a real potential for an effect at 0.001 ug/kg/day.
Further analysis of this study and of the subhuman primate studies by Allen
and his colleagues may provide support for a lower effect level of at least
0.001 ug/kg/day.
In relation to developmental toxicity, a large number of studies in a
variety of species demonstrated that 2,3,7,8-TCDD is a developmental toxicant.
Although a longer term exposure appears to cause effects at slightly lower
doses, acute and short-term exposures are effective in causing altered
development, and therefore, should also be of concern. When exposure occurs
during the prenatal period of organogenesis, the lowest effect level is in the
range of 0.125 to 1.0 ug/kg/day and the no-effect level is approximately 0.1
ug/kg/day. Studies focusing on other periods of development are limited, but
they do indicate that exposure at any time during prenatal and early postnatal
life must be considered a potential threat to normal development.
2,3,7,8-TCDD also affects the male and female reproductive systems.
Unfortunately, the amount of attention that has been given, to these areas of
investigation has not been as great as that given to reproductive function and
development. Gonadal dysfunction has been demonstrated in both sexes, and
alterations of normal reproductive cycles have been demonstrated in the female.
Continuing investigations of the effect of 2,3,7,8-TCDD and related agents on
reproductive physiology and cellular events should increase our understanding
of the potential effect of 2,3,7,8-TCDD on the male and female reproductive
systems.
The uncertainties that arise from this data base are many, largely because
the area of reproductive and developmental toxicology covers a wide range of
13
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potential exposure-response scenarios, and because the reproductive and
developing systems present constantly changing targets with which a toxicant
may interact. In the case of 2,3,7,8-TCDD exposure, many of these
uncertainties have been discussed in this review and should be taken into
account in carrying out the final risk characterization. The laboratory data
establishes 0.01 ug/kg/day as a lowest observed adverse effect level (LOAEL)
when exposure is chronic. Standard approaches for applying uncertainty/
modifying factors and determining a reference dose (RfD) have been used (U.S.
EPA, 1986b; 1987). An unequivocal no observed adverse effect level (NOAEL) can
be questioned. A 1000-fold uncertainty factor (UF) can be applied to account
for variation in sensitivity within the human population, uncertainty in
extrapolating animal data to the human situation, and the use of a LOAEL
instead of a NOAEL in calculating the reference dose. An additional modifying
factor seems unnecessary, since the uncertainty factor accounts for many of the
concerns of this reviewer, i.e., the true LOAEL/NOAEL and the potential for the
pregnant woman and her offspring to be more sensitive than the average healthy
adult. An exception to this position would arise if an actual effect level
much below 0.001 ug/kg/day was established, and as noted above, suggestive
evidence is accumulating that would support a lower NOAEL/LOAEL. The
calculation of a reference dose (RfD) for reproductive and developmental
toxicity, based on the current data and literature base, is as follows:
RfD = LOAEL/UF
= (0.01 ug/kg/day)/1000
= 1 x 10~5 ug/kg day..-.
14
-------
This does not account for potential or actual human exposures which would have
to be factored into the final risk characterization.
Future investigations should be encouraged to more clearly define the
substantial data base that already exists. In the area of reproductive
function and fertility, it would be helpful if a more critical review of the
data of the three-generation rat study and the monkey studies was carried out
to determine if a lower effective dose can be established. Additionally, a
carefully designed multigeneration study could address some of the limitations
of previous studies and fill certain data gaps. However, multigeneration
studies are not easily designed, executed, or evaluated, and this step should
only be taken when it is obvious that these data are necessary. In the area of
developmental toxicology, the most pressing needs seem to be the evaluation of
toxicity during periods of development that have not been adequately assessed,
i.e., the late prenatal and early postnatal periods. With regard to the
postnatal period, the potential for childhood exposure from breast feeding and
from other environmental sources (e.g., ingestion of soil) seems considerable,
and it has been suggested that the child may be particularly sensitive to
exposure to polychlorinated dibenzo-p_-dioxins. As relates to reproductive and
developmental toxicity in general, a greater effort should be directed at
identifying the effects of polychlorinated dibenzo-£-dioxins on hormonal
regulation and normal cellular and tissue functions. There is evidence that
2,3,7,8-TCDD influences steroid metabolism and may be associated with steroid
action at the cellular receptor level. Pratt and his colleagues (Dencker and
Pratt, 1981; Pratt et al., 1984) have also shown a correlation in various mouse
strains between the susceptibility to induction of cleft palate and the
occurrence of the 2,3,7,8-TCDD receptor, and have proposed that 2,3,7,8-TCDD
15
-------
exerts its teratogenic effect on the palate directly through this receptor.
A considerable amount of literature has developed on the mechanisms of cellular
interaction and action aspects of 2,3,7,8-TCDD, as well as on the
structure-activity relationships of 2,3,7,8-TCDD with other dioxins and related
compounds (recent reviews include: Safe, 1986; Silbergeld and Mattison, 1987).
Efforts should be made to incorporate this information and to evaluate
2,3,7,8-TCDD within the context of the larger family of related agents. As
additional information becomes available, this review will be updated and
reconsidered relative to its appropriateness in defining critical studies and
issues related to the reproductive and developmental toxicity of 2,3,7,8-TCDD.
16
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Pharmacol. 48:(2)A180.
Schantz, S.L.; Laughlin, N.K.; Van Valkenberg, H.C.; Bowman, R.E. (1986)
Maternal care by rhesus monkeys of infant monkeys exposed to either lead
or 2,3,7,8-tetrachlorodibenzo-p_-dioxin. Neurotoxicology 7:637-650.
Silbergeld, E.K.; Mattison, D.R. (1987) Experimental and clinical studies on
the reproductive toxicology of 2,3,7,8-tetrachlorodibenzo-p_-dioxin. Am.
J. Ind. Med. 11:131-144.
Smith, F.A.; Schwetz, B.A.; Nitschke, K.D. (1976) Teratogenicity of 2,3,7,8-
tetrachlorodibenzo-p_-dioxin in CF-1'Mice. Toxicol. Appl. Pharmacol.
38:517-523.
Sparschu, G.L.; Dunn, F.L.; Rowe, V.K. (1971) Study of the teratogenicity of
2,3,7,8-tetrachlorodibenzo-p_-dioxin in the rat. Food Cosmet. Toxicol.
9:405-412.
U.S. Environmental Protection Agency (EPA) (1985) Health assessment document
for polychlorinated dibenzo-p_-dioxins. Office of Health and Environmental
Assessment. EPA/600/8-84/014F. NTIS PB86-122546/AS.
U.S. Environmental Protection Agency (EPA) (1986a) Teratology and
reproduction studies with TCDD. In: Pitot, H., ed. The report of the
"dioxin" Update Committee. Office of Pesticides and Toxic Substances,
Washington, DC.
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U.S. Environmental Protection Agency (EPA) (1986b)
assessment of suspect developmental toxicants.
51:34028-34040.
Guidelines for the health
Federal Register
U.S. Environmental Protection Agency (EPA) (1987) Reference dose (RfD):
description and use in health risk assessments. Integrated Risk
Information System (IRIS): Appendix A. Online. Intra-Agency Reference
Dose Workgroup, Office of Health and Environmental Assessment.
Environmental Criteria and Assessment Office, Cincinnati, OH.
EPA/600/8-86/032a.
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February 1988
Review Draft
APPENDIX D
EPIDEMIOLOGIC DATA ON REPRODUCTION AND
EXPOSURE TO 2,3,7,8-TCDD:
ITS USEFULNESS IN QUANTITATIVE RISK ASSESSMENT
Sherry G. Selevan, Ph.D.
Reproductive Effects Assessment Group
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
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INTRODUCTION '
The following is a discussion of epidemiologic data on the reproductive
effects of 2,3,7,8-tetrachlorodibenzo-p_-dioxin (2,3,7,8-TCDD) and their
usefulness in qualitative and quantitative risk assessment. Several factors
affect the usefulness of these studies. First, and probably most important, is
the assessment of exposure. Other important factors include the biologic
plausibility of the effects of exposure, given the time frames of exposure and
its association with relevant outcomes, and the difficulty of attributing the
effects of combined exposure to polychlorinated dibenzo-p_-dioxins (hereafter
referred to as "dioxin"), a contaminant in herbicides or pesticides.
Due to the narrow focus of this report, that is, the use of existing
reproductive data in the risk assessment of dioxin, certain restrictions will
be made on the data discussed: The usefulness of data for risk assessment
depends upon the manner in which the probabilities of exposure for the study
members are determined. The quality of exposure data may range from very
indirect data to detailed industrial hygiene or environmental monitoring.
These indirect data are less useful and result from assumptions that
individuals were exposed due to his/her presence in a potentially exposed
region/plant during a specific time period. More useful data would include
measurement of actual levels of dioxin in air, soil, or water with the most
useful data describing the individuals' levels of exposure. Studies with only
indirect, assumed exposure data contribute little to a risk assessment because
of limited confidence in the ability to determine whether a given individual
was actually exposed. Misclassification of exposure will inevitably occur,
typically resulting in a reduction of the estimate of risk (if such a risk
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truly exists). Studies with more detailed exposure data will contribute more
specific information to such a risk assessment.
ROUTES OF EXPOSURE
Humans may be exposed to 2,3,7,8-TCDD through several routes: dermal,
ingestion, and inhalation. The rates of absorption through these routes vary,
as do the important routes for each individual. The routes for each individual
can be affected by such characteristics as location during spraying (e.g.,
Vietnam/Oregon forests), diet (e.g., consumption of contaminated water/fish),
and work practices for those occupationally exposed. As these characteristics
vary, so will each person's effective dose. Table 4 (U.S. EPA, 1988) describes
the absorption fraction for dioxin from several routes: soil ingestion - 0.3,
dermal exposure to soil - 0.005, vapor inhalation - 0.75, fat ingestion from
dairy products or beef - 0.68, dust inhalation - 0.27, fish ingestion - 0.68,
and surface water ingestion - 0.5. For example, an individual exposed to equal
amounts of 2,3,7,8-TCDD from different routes may have radically different
internal doses (e.g., during the ICMESA plant explosion in Seveso, Italy, an
individual would have a different potential exposure than.she/he would have if
exposed to soil dust later on). Consequently, each individual's actual dose is
difficult to estimate accurately in an epidemiologic study. Another potential
source of 2,3,7,8-TCDD exposure of potential reproductive importance may be the
infant's exposure through human breast milk (Rappe, 1985; Schecter et al.,
1987; van den Berg et al., 1986). ,
The epidemiologic literature, while limited at this point in time, seems
to cover two broad categories: In occupational settings, paternal effects have
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been examined (e.g., aerial sprayers in Vietnam, Vietnam veterans, agricultural
sprayers, and employees of manufacturers of 2,4,5-trichlorophenoxyacetic acid
[2,4,5-T]). Birth defects and fetal loss have been examined through
environmental exposures (e.g., aerial spraying of 2,4,5-T in Oregon, New
Zealand, and Vietnam, and the ICMESA plant accident in Seveso, Italy). In
these settings, separation of maternal and paternal effects is very difficult.
The only exception may be in such settings as the plant accident in Seveso
where the woman may already be pregnant and In utero exposures are the ones of
importance. None of these exposures are "clean;" that is, of dioxih alone.
All the human exposures to dioxin have been as a contaminant of some other
agent (e.g., 2,4,5-T), and typically, only sketchy qualitative and quantitative
exposure data were available. Therefore, it is difficult to separate out the
effects of dioxin from the agent it contaminates. Only some very general
conclusions may be drawn from the relative toxicities of the associated
compounds.
A number of reviews have discussed the human data in detail (Kimbrough et
a!., 1984; Constable and Hatch, 1985; Friedman, 1984; Hatch, 1984; Hatch and
Stein, 1986; U.S. EPA, 1985). These efforts will not be repeated here, but the
discussion will be limited to several key areas of study and certain studies of
special interest to help define the limits of our knowledge on the human
reproductive effects of dioxin.
STUDIES OF VIETNAM VETERANS
Two studies have examined the reproductive effects of 2,3,7,8-TCDD and
defoliants in U.S. Vietnam veterans. The primary exposure was to Agent Orange
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which consists of defoliants 2,4-dichlorophenoxyacetic acid (2,4-D) and
2,4,5-T, the latter contaminated with 2,3,7,8-TCDD. (Agents Purple, Pink, and
Green were primarily used prior to July 1965.) These two studies are the Ranch
Hand study (Lathrop et al., 1984) and the Center for Disease Control's (CDC)
study of birth defects (Erickson et al., 1984). A third study examines the
birth defects in children born to Australian Vietnam veterans (Donovan et al.,
1983).
The Ranch Hand study (Lathrop et al., 1984) compared airmen who had flown
in the aerial spraying of defoliants (in fixed-wing aircraft) to those who
belonged to flying organizations responsible for transporting cargo. Other
types of spraying (helicopters and backpacks) were not included in the Ranch
Hand operations. Hhile the data in this report have not yet been verified
(through birth registration or medical records), preliminary analyses have
examined various measures of fertility and reproductive success (through
pregnancy outcomes, sperm count, and morphology). The researchers are
currently validating the interview data and will subsequently (they state) do
more detailed analyses. Exposure to 2,3,7,8-TCDD was estimated by calculating
the amount of 2,3,7,8-TCDD (in herbicides) used during each airman's tour and
dividing this number by the total number of airmen with the same
responsibilities during the subject's tour.
The Ranch Hand study individually matched comparisons to each exposed man,
put in replacements for refusals, and did not maintain the matches in the
analyses. Due to these procedures, determination of the total study population
and response rate is difficult; 1,174 Ranch Handers were included in the
analysis of reproductive-data, as were 1,531 comparisons. Both the man and his
current or former spouse were interviewed on reproductive history; research has
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shown such interview data on reproductive events.to have greater validity when
collected from the wife. However, the researchers combined pregnancies from
both spouses (but not necessarily reported by both), so it is impossible to
examine the women's data separately. No differences were observed in early or
late fetal loss, induced abortion, or live births in a crude analysis of the
two groups. More detailed analyses examined enlisted men and officers
separately, thus reducing the power of the analyses. Some important
independent risk factors were not adjusted for in the analysis (e.g., prior
fetal loss in analyses of current fetal loss). It appears that multiple
pregnancies per family group were all analyzed; however, there was no
discussion of the problems associated with the analysis of nonindependent
events. A statistically significant excess was found for birth defects,
controlling for parental age and for maternal smoking and drinking; however,
the reproductive outcome of "birth defects" is probably the one which most
needs validation against medical records to assure accurate classification. No
differences were observed in sperm count or morphology in the two groups. The
more detailed analysis planned for the future could result in useful
information since the potential for misclassification of exposure in this study
appears to be less than for the other veterans studies.
Another major study of U.S. Vietnam veterans was a case-referent study of
birth defects done by the CDC (Erickson et al., 1984). The cases and referents
were drawn from births occurring in metropolitan Atlanta from 1968 through
1980. The 7,133 eligible cases consisted of live and still births with serious
or major defects identified using the 8th revision of the International
Classification of Diseases (ICD-8). The referent group was selected from
metropolitan Atlanta live births during the same years; the referents were
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frequency-matched on race, year of birth, and hospital of birth. Power
calculations suggested that a referent group of 3,000 was sufficient; above
that no substantial increase in power would be found for "all defects." The
authors also stated that power for odds ratios greater than 3.0 for specific
birth defects also would not be substantially affected by using more
comparisons. To allow for a 70% response rate, 4,246 referent births were
selected. Seventy percent of the women responded, but only 56% of the men did.
The potential for exposure for the CDC study was based on three
definitions: First, a man was considered exposed if he had served in any
capacity at any time in Vietnam before the conception of the infant. Second,
each veteran was queried about his exposure to Agent Orange. Third, an
Exposure Opportunity Index (EOI) score was (subjectively) developed by a panel
of specialists who evaluated the veterans' duties, location, and time spent in
Vietnam. Veteran status was examined to determine whether any military service
might be associated with the 96 birth defects examined. Veteran status was not
associated and subsequently dropped from further analyses.
These data were analyzed a number of ways, but for all, the data were
stratified on the three variables used for the frequency matching: (1) without
potentially confounding characteristics; (2) with key characteristics
identified a priori (maternal age, education, and alcohol consumption, and the
presence of birth defects in close (first-degree) relatives of the child
studied; and (3) with a. posteriori testing of 108 -other characteristics. In
addition, certain groupings of the 96 birth defect categories, thought to be
related, were also examined. A limited number of elevated findings were
reported out of almost 400 analyses. Spina bifida was associated with both
levels of the EOI indexes; cleft lip without cleft palate was associated with
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veteran status and high EOI; specified anomalies of nails was reported in
Vietnam veterans; and "other neoplasms" were associated with high EOI. Due to
the large number of analyses and hypotheses tested, one would expect
approximately 20 significant associations (elevated or decreased odds ratios)
with a significance level of p = 0.05 by chance alone. Only 11 statistically
significant findings were reported, less than that expected by chance. The
authors appropriately concluded that "... the data collected contain no
evidence to support the position that Vietnam veterans have had a greater risk
than other men for fathering babies with all types of serious structural birth
defects combined."
The Australian government examined the association between military
service in Vietnam and birth defects in their offspring (Donovan et al., 1983).
A case-referent study compared births occurring from 1966 through 1979, of
8,517 children with defects recognizable at birth to an equal number of live
born children without birth defects. The referents were matched by time of
birth and maternal age. Exposure was defined as presence of, the father in
Vietnam; no specific index of exposure was developed, but the investigators
stated that such exposure was probably low for the Australian troops in
Vietnam. No difference lwas observed in exposure patterns between cases and
referents. " ,
For veteran studies, the exposures are to the fathers, typically many
years before the pregnancy under study was conceived. These studies have used
different methods to assign veterans into different exposure groups, based on
service in Vietnam (Erickson et al., 1984; Donovan et al., 1983), matching of
troop movement with aerial spraying (Erickson et al.,' 1984), or participation
in the Ranch Hand operations (Lathrop et al., 1984). At the time of these
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studies, a suitably sensitive assay for 2,3,7,8-TCDD in serum did not exist.
The CDC recently reported on a new assay they developed in 1986; this assay
yielded serum 2,3,7,8-TCDD levels which correlated well (r = 0.98) with
2,3,7,8-TCDD levels in the adipose tissue of the same individuals (CDC, 1987).
Paired samples collected from the same individuals in 1982 and 1987 were used
to derive the half-life of the 2,3,7,8-TCDD body burden (6 to 10 years). In
this preliminary report, CDC compared the assayed levels of 2,3,7,8-TCDD to the
exposure categories they had established using the EOI and interview data: no
association was found between the three methods of scoring for potential
exposure and the serum levels observed. Furthermore, in their examination of
the sera of Vietnam era veterans (444 Vietnam veterans versus 75 non-Vietnam
veterans), the median sera levels did not differ for these two subgroups.
These data raise a number of issues in the consideration of this body of
research:
(1) The relatively long half-life broadens the range of potential
mechanisms that could occur if an association exists between paternal
2,3,7,8-TCDD exposure and birth defects. Friedman (1984) and Hatch and
Stein (1986) have suggested that such an exposure would cause birth
defects in offspring through gene or chromosomal mutation of
spermatogonial stem cells (premeiotic effects). However, the long half-
life observed means that postmeiotic effects on sperm are also possible.
The long half-life also lends more support to Friedman's suggestion of
maternal/fetal exposures from 2,3,7,8-TCDD in seminal fluid.
(2) These data suggest that a great deal of misclassification of exposure
occurred in the assignment of veteran's to their exposure categories for
the CDC study. Plus use of this assay would improve exposure estimation
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in the other studies.
(3) Finally, the levels and similarity of median 2,3,7,8-TCDD sera values
in both Vietnam and non-Vietnam veterans raises questions about the
magnitude of exposure of Vietnam veterans. Other data suggest that the
exposures of Vietnam veterans were less than originally thought: the
median levels of 2,3,7,8-TCDD in both groups are below those found in the
adipose tissue of the control group of a study in eastern Missouri
(Patterson et al., 1986).
OCCUPATIONAL STUDIES
Occupational studies tend to be more useful in the evaluation of a
potentially toxic exposure in risk assessment. This is especially true for
qualitative risk assessment, since one can be fairly certain that exposure to
the worker did occur. If good historical industrial hygiene data are
available, these can also be of use in quantitative risk assessment. As
described in the Vietnam veterans studies, the exposures evaluated here are
primarily to male workers and the studies have been restricted to examination
of paternal effects; however, wives may have been exposed indirectly, through
handling their husband's clothes, etc. Unlike the veterans studies, the
exposures may be occurring concurrently with conception, thus increasing the
potential exposure- level at the time of the pregnancy.
Smith et al. (1982) identified 616 male chemical applicators from a list
maintained by New Zealand's agricultural Chemicals Board and 531 comparison
workers at small agricultural contracting companies. A total of 89% of the
chemical applicators and 83% of the contractual workers responded to a mailed
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questionnaire requesting occupational histories from the men and reproductive
histories from their wives. Exposure to 2,4,5-T was defined as having sprayed
this pesticide during the year of the pregnancy studied or during the year
preceding that pregnancy. Group 1 (not exposed to any spraying) consisted of
392 pregnancies; group 2 (sprayed other chemicals than 2,4,5-T) had 109
pregnancies; and group 3 (sprayed 2,4,5-T) had 473 pregnancies. Comparisons of
group 3 to group 1 found no statistically significant differences in fetal loss
or birth defects. Other risk factors did not appear to be controlled in the
analysis of these data: for example, the analysis of miscarriages did not
appear to control for maternal age or occurrence of prior fetal loss. Multiple
pregnancies per family unit were studied; however, the authors did not address
the problem of nonindependent events. Additionally, the power of this study
was very low for birth defects.
Dow Chemical studied 930 male workers potentially exposed to dioxins,
through work with chlorophenol processes, for at least one month from January
1939 through December 1975 (Townsend et al.', 1982). The reproductive
experience of their wives, obtained by interview, was compared to the wives of
an equal number of unexposed male employees. For the 930 exposed workers, 586
wives were identified and 370 responded (63%); for the comparison group, 559
wives were identified and 345 responded (62%). Exposure potentials from low to
high were assigned to jobs using historic surface contamination data by an
industrial hygienist familiar with the history of the facility. A pregnancy
was considered exposed if the employee had worked in an exposed area for at
least one month at any time prior to conception. Thus, this study, in its
analysis, has similar problems to those described above for the study of
Vietnam veterans concerning the time delay between exposure and conception.
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The outcomes examined included miscarriages (< 20 weeks gestation),
stillbirths, and birth defects identified (or potentially .identified) before
the child's first birthday. The relationships of these outcomes were compared
to several classifications of exposure: any dioxin, only 2,3,7,8-TCDD,
moderate and higher levels of 2,3,7,8-TCDD only, other dioxins only. The "any
dioxin" and "only 2,3,7,8-TCDD" categories were done as broad groups, plus
split into two groups each, based on duration of exposure (< 12 months, or > 12
months) during the entire employment period preceding conception. Over two
thousand pregnancies were in the "non-dioxin" group, while the "dioxin" group
consisted of 737 pregnancies. The "non-dioxin" group included pregnancies
occurring before exposure for exposed workers. No differences were found in
the two groups. The power was low for the examination of birth defects. As
noted above, the assumption of any exposure greater than one month at any time
prior to conception could obscure a true effect in this population. A
reanalysis of these data, looking at possible associations with paternal
exposure during the 3 to 4 months preceding conception might give more insight
into potential paternal effects associated with this exposure.
In a clinical, cross-sectional study, reproductive characteristics in
current (N = 131) and retired (N = 161) men who had worked with 2,4,5-T were
selected for comparison to workers without exposure (N = 133). Of the workers
selected, 55% responded. Historic exposure data were not available, and job
titles were not sufficient to estimate exposure; therefore, interview exposure
history was used to group workers into a probable exposure category.
Contamination of the plant with 2,4,5-T occurred in 1949. For the purpose of
comparing reproductive histories, reported chloracne was used to distinguish
groups. No differences were found for fetal loss or birth defects between the
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workers with chloracne (N = 107/number of pregnancies = 235) and those without
(N = 91/number of pregnancies = 203) during or after 1948 (when 2,4,5-T was
first handled at the plant under study). Differences were observed in reports
of decreased libido and difficulty with erection or ejaculation. Due to the
very indirect nature of exposure assessment, the probability of
misclassification of the exposure category is great.
AGRICULTURAL AND FORESTRY SPRAYING
A number of studies have examined the relationship between reproductive
effects and aerial spraying of pesticides (Field and Kerr, 1979; Hanify et al.,
1981; Nelson et al., 1979; Thomas, 1980). In these studies, exposures could be
to either or both parents. These are all ecologic studies, in which exposure
of study members is assumed by their presence in an area (e.g., by their
residence), with a specified likelihood of exposure to a given agent. These
studies are heir to the "ecologic fallacy" in that individuals defined by
residence or some other factor as being in a certain exposure category, may not
in fact belong in that category. Thus, such studies are of very limited value
in either qualitative or quantitative risk assessments; therefore, only brief
discussions of published reports of these follow.
The earliest of these studies (Nelson et al., 1979) examined the
association between cleft lip and/or cleft palate (CL/P) in live births
occurring during the 32-year period beginning in 1943 and the spraying of
2,4,5-T in Arkansas. Approximately 1,200 cases of CL/P Were identified using
both birth certificate data and records of the Crippled Children's Services of
Arkansas Social and Rehabilitative Services. 2,4,5-T was primarily used on
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rice; therefore exposure was determined by estimating the proportion of rice
acreage to total acreage in each county from data supplied by the Arkansas
State Plant Board (1970-1974). (The exposure definition does not address the
use of 2,4,5-T in forestry.) The 75 counties were then divided into categories
of low, medium, and high potential for 2,4,5-T exposure. An increase was
observed in facial clefts over time, which the authors suggested were related
to improved recording and case ascertainment and not to an association with
2,4,5-T spraying.
A letter to the editor of Lancet at approximately the same time (Field and
Kerr, 1979) discussed the relationship between 2,4,5-T and 2,3,7,8-TCDD and
neural-tube defects in New South Wales, Australia. Only limited data were
presented in this letter: in New South Wales, annual rates for neural-tube
defects (anencephaly and meningomyelocele) were compared to annual usage
figures for 2,4,5-T for all of Australia. Seasonal variation of 2,4,5-T usage
was obtained by questionnaire of local governments from the years 1965-1976.
2,4,5-T usage increased from 90 tonnes (1 tonne = 1,000 kg) in 1965 to 482
tonnes in 1976. The authors reported a linear correlation between the previous
years' 2,4,5-T usage and neural-tube defects. In addition, a seasonal pattern
was noted, with the highest rates in conceptions occurring during the summer
months (December, January, February). The authors also noted that in the
Northern Hemisphere, the highest rates in conceptions also occurred during the
summer months. In addition to the limitations present in ecologic studies, no
comparison rates of neural-tube defects were reported in communities without
such exposure, nor were high versus low exposure communities compared.
Another letter to the editor of Lancet the next year (Thomas, 1980)
discussed a comparison of selected birth defects (cleft lip, cleft palate,
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spina bifida, anencephalus, and cystic kidney disease) and use of 2,4,5-T in
Hungary. The birth defects were identified using Hungary's malformation
registry, which records malformations recognized at up to one year of age. In
Hungary, the use of 2,4,5-T rose from 46 tonnes in 1969 to over 1,200 tonnes in
1975. The rates of some birth defects (spina bifida and anencephalus)
decreased over the time period examined, while others (cleft lip, cleft palate,
and cystic kidney disease) remained essentially unchanged. In a comparison of
this report with the New South Wales letter from Field and Kerr (1979), Thomas
noted that: (1) the population examined in Australia was spread over a much
greater geographic area (thus reducing the probability for exposure), (2) over
half of the population in Hungary lived in rural areas, while only 15% of the
population in Australia did, and (3) 24.6% of the population in Hungary was
actively involved in forestry and agriculture as compared with only 7.4% of the
Australian population. While these data suggest no association between the
defects and 2,4,5-T, this report also did not attempt to compare exposed
communities with less exposed (or unexposed) groups.
Hanify et al. (1981) compared rates of diagnosed birth defects in
stillbirths (> 28 weeks gestation) and live births in Northland, New Zealand,
to densities of aerial 2,4,5-T spray application (1960-1977).' The reproductive
events were identified through seven regional hospital records: 3.7,751 births
which included 436 stillbirths, 264 neonatal deaths, and 510 with recorded
birth defects. The location, data, and quantity of 2,4,5-T sprayed was
obtained from company records, and monthly estimates were made for each of.the
seven regions. Environmental levels were estimated modeling both the new
applications plus that fraction of the previous applications thought to still
be present. 2,4,5-T was not sprayed from 1959-1965; thus, for this time
• i-''
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period, exposure was considered to be zero. The zero time period (1960-1965)
was compared to years thought to be representative of "spraying years"
(1972-1976). Associations were found for all birth defects and for talipes
under certain assumptions of environmental persistence of 2,4,5-T. The authors
did not discuss the possibility of secular trends in the data, due to other
medical or environmental factors, or differences in the
identification/recording of malformations over time.
OTHER ENVIRONMENTAL EXPOSURES -- SEVESO, ITALY
This category of "other environmental exposures" includes unplanned
exposures such as exposures from accidental emissions from industrial
facilities. With an immediately recognizable environmental incident, exposure
of interest for reproduction could occur for both parents, or just for the
mother, if she is already pregnant at the time of first exposure (unless there
could be in utero exposure to dioxin in seminal fluid). In 1976, during the
production of trichlorophenol at the ICMESA plant in Seveso, Italy, a runaway
reaction resulted in an explosion that ultimately contaminated 700 acres in the
surrounding community. Environmental levels of 2,3,7,8-TCDD were determined
using wipe tests, evaluating toxic effects in small animals, and analyzing
grass samples. Approximately 2 weeks later, over 200 families were evacuated
from high contamination areas. Exposures were sufficiently high for chloracne
to be observed in this environmentally exposed population. Several reports
(Bisanti et al., 1979; Homberger et al., 1979; Pocchiari, 1980; Pocchiari et
al., 1980; Reggiani, 1978; Rehder et al., 1978; Tuchmann- Duplessis, 1977)
reported on comparisons of four potentially affected communities (Seveso, Meda,
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Cesano, and Desio) to nearby, unexposed commun'ities. Changes in fetal loss
rates occurring during the last quarter of 1976 and first quarter of 1977 were
found in both exposed and unexposed communities. An estimated 150 women were
in the first trimester of pregnancy at the time of the accident (Rehder et al.,
1978); of these, 125 women wished therapeutic abortions by October 1976.
Therapeutic abortions were approved for 30. Another estimate
(Tuchmann-Duplessis, 1980) reports a total of 108 (50 in 1976 and 58 in 1977)
therapeutic abortions in the four affected communities. Several reports
(Bisanti et al., 1979; Pocchiari, 1980; Pocchiari et al., 1980) suggested that
a large number of women obtained unapproved, and therefore not reported,
therapeutic abortions. This supposition was supported by a steep decrease in
birth rates in the first 6 months of 1977, primarily observed in the exposed
communities. (All communities had had decreases over time; however, the
decrease in exposed communities at this key time was much larger.) Marked
increases in the number of birth defects were noted in 1977. These have been
attributed to changes in reporting. Prior to this time, only certain birth
defects were required to be reported to Italian Health Officers (Reggiani,
1980). In addition to the limitations due to legal requirements, there were
other reasons for the underreporting of malformations: "Traditionally, Italian
physicians have under-reported congenital malformations because of their severe
negative social implications" (Tuchmann-Duplessis, 1980a, b). Although
physicians and midwives were encouraged to do more complete recording, Reggiani
(1978) concluded that the malformation data were "missing" for 1976 and
"incomplete" for the first trimester of 1977. In general, the researchers
cited above felt that, in spite of the problems associated with the data,
adverse reproductive effects did not occur in this population. However, the
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data do not appear to be complete enough to make definitive conclusions. In
addition, these reports have the same problems concerning exposure assessment
as those described previously in the Agricultural/Forestry Spraying section.
STUDIES OF THE VIETNAMESE
While studies of the Vietnamese have not been published in Western
journals, a review has been presented by Constable and Hatch (1985). Two types
of studies have been executed: In the South, the reproductive experience of
exposed and unexposed couples was compared; in this case, effects of maternal
and paternal exposures cannot be separated. In the North, reproductive
experience was compared for families of men who had served in the South (thus,
assuming paternal exposure) to families of men remaining in the North (and
therefore assuming no exposure of the father). Exposures for these studies
have been defined by residence, using historical data on spraying and/or
determined by the evidence of destruction of vegetation. Thus, these studies
have the same limitations for use in qualitative and quantitative risk
assessment as discussed previously.
Only one study (Lang et al., 1983 described in Constable and Hatch, 1985)
has attempted to determine types of exposure (e.g., exposed during spraying
versus the more indirect exposure due to dust or diet) and assign exposure
scores. In this study of veterans in the North, exposure was restricted to the
father at some time prior to conception, thus resulting in the same problems
encountered in the American and Australian studies of veterans.
17
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SUMMARY
All three Vietnam veterans studies examined the occurrence of birth
defects. The Ranch Hand study found an excess for "all birth defects" (these
were not validated); the CDC study found a difference in exposure for cases
with spina bifida, cleft lip without palate, specified anomalies of the nails,
and "other neoplasms"; the third study (Donovan et.al., 1983) found no
differences. No associations with other reproductive outcomes were found in
the Ranch Hand study (the only study of the three that looked at the other
outcomes). No differences were found in reproductive outcome in the three
workplace studies (Moses et al., 1984; Smith et a!., 1982; Townsend et al.,
1982), however, Moses et al. did report increased impotence and sexual
dysfunction in 2,4,5-T workers with chloracne (assumed to result from 2,4,5-T
and/or 2,3,7,8-TCDD). Several of the ecologic studies (Nelson et al., 1979;
Field and Kerr, 1979; Hanify et al., 1981) reported associations with some
birth defects and volume of 2,4,5-T application; these data were contested by
Thomas (1980). Birth defects in Seveso increased after the plant accident,
potentially due to changes in reporting (Reggiarii, 1980). As presented in this
paragraph, one might conclude that 2,3,7,8-TCDD/2,4,5-T is related to adverse
reproductive outcomes, but as discussed previously, limitations in the design
of the studies neither allow nor rule out such an interpretation.
Due to the very nature of;the exposures to 2,3,7,8-TCDD, obtaining
suitable data for either a qualitative or quantitative risk assessment is
difficult, if not impossible. First, the exposures are primarily in general
environmental settings (wartime exposures are, for these purposes, being
classified as environmental) where it is difficult to define whether an
18
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individual is even exposed and nearly impossible.to quantitate that exposure.
In occupational settings,, both qualitative and quantitative exposure data may
exist, but small population sizes limit the power of the studies. A nationwide
study planned by the National Institute for Occupational Safety and Health
(NIOSH) on their Dioxin Registry could overcome the problems of limited power.
All of the studies described above contained a major limitation:
imprecise definitions of exposure of the subjects. In many studies, the
exposure was assumed by place of residence or presence in a particular area.
Even where more descriptive exposure data are available, such as the CDC and
Ranch Hand studies, misclas.sification may be great (CDC, 1987). Other studies,
such as the ecologic studies, compared the same area over time rather than to a
comparison area (e.g., Field and Kerr, 1979; Thomas, 1980); thus, differences
in these studies may be due to differences in secular trends rather than to
changes in potential exposure to 2,3,7,8-TCDD. The evidence from all of these
studies, therefore, is open to question.
The reports with more detailed exposure information include those
examining paternal exposure. Kimmel has reviewed the animal literature in
Appendix C and reported that 2,3,7,8-TCDD resulted in testicular atrophy and
reduced sperm count, but .a dominant lethal study found no effects (Kimmel,
1988). The animal data are too limited, at this time, to suggest the presence
or absence of an effect in the offspring in studies with human male exposure.
The large differences in time frame between male exposure and conception in the
epidemiologic studies probably reduces the chance of identifying any paternally
mediated effect that is potentially occurring. However, new data on the half-
life of 2,3,7,8-TCDD in humans (CDC, 1987) suggests some exposure may be
occurring many years later, depending on the level of initial exposure.
19
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Reanalyses of occupational studies, with stricter exposure definitions, might
yield useful information.
The only studies that address the possibility of female exposure are the
ecologic/environmental studies and the studies of Seveso, In his review of the
animal'literature, Kimmel (1988) found evidence of reproductive and
developmental effects of dioxin. The human studies have less informative
exposure data, and so are less useful in drawing conclusions concerning the
reproductive effects of 2,3,7,8-TCDD. Additionally, in both types of studies,
males may also be exposed, thus clouding the interpretation of such data.
20
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Bisanti, L.; Bonetti, F.; Caramaschi, F.; Del Corno, G.; Favaretti, C.;
Giambelluca, S.; Marni, E.; Montesarchio, E.; Puccinelli, V.; Remotti, G.;
Volpato, C.; Zambrelli, E. (1979) Experience of the accident of Seveso.
Proceedings of the 6th European Teratology Conference, Akademiai Kiado.
Pub.
Centers for Disease Control (CDC). (1987) Serum dioxin in Vietnam-era
veterans - preliminary report. Morbidity and Mortality Weekly Report.
36(28):470-475.
Constable, J.D.; Hatch, M.C. (1985) Reproductive effects of herbicide
exposure in Vietnam: recent studies by the Vietnamese and others.
Teratogenesis Carcinog. Mutagen. 5(4):231-250.
Donovan, J.W.; Adena, M.A.; Rose, G.; Batistutta, D. (1983) Case-control
study of congenital anomalies and Vietnam service. Canberra: Australian
Government Publishing Services. (As described in Hatch and Stein, 1986).
Erickson, J.D.; Mulinare, J.; McClain, P.W.; Fitch, T.G.; James, L.M.;
McClearn, A.B.; Adams, M.J. (1984) Vietnam veterans' risks for fathering
babies with birth defects. JAMA 252(7):903-912.
Field, B.; Kerr, C. (1979) Herbicide use and incidence of neural-tube
defects. Lancet 1(8130):1341-1342.
Friedman, J.M. (1984) Does agent orange cause birth defects? , Teratology
29:193-221.
Hanify, J.A.; Metcalf, P.; Nobbs, C.L.; Worsley, R.J. (1981) Aerial spraying
of 2,4,5-T and human birth malformations: an epidemiological investiga-
tion. Science 212:349-351.
Hatch, M.C. (1984) Reproductive effects of the dioxins. In: Lowrance, W.W.,
ed. Public health risks of the dioxins. Proceedings of a symposium held
at The Rockefeller University in New York City, Oct. 19-20, 1983. Los
Altos, CA: William Kaufmann.
Hatch, M.C.; Stein, Z.A. (1986) Agent orange and risks to reproduction: the
limits of epidemiology. Teratogenesis Carcinog. Mutagen. 6(3):185-202.
Homberger, E.; Reggiani, G.; Sambeth, J.; Wipf, H. .(1979) The Seveso
accident: its nature, extent and consequences. Report from Givaudan
Research Company Ltd. and F. Hoffmann-La Roche & Co. Ltd.
Kimbrough, R.D.; Falk, H.; Stehr, P.; Fries, G. (1984) Health implications of
2,3,7,8-tetrachlorodibenzodioxin (TCDD) contamination of residential soil.
J. Toxicol. Environ. Health 14(l):47-93.
21
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Kimmel, G.L. (1988) Reproductive and developmental toxicity of 2,3,7,8-TCDD.
In: A cancer risk-specific dose estimate for 2,3,7,8-TCDD, Appendix C.
External Review Draft. U.S. Environmental Protection Agency, Washington,
DC.
Lathrop, 6.D.; Wolfe, W.H.; Albanese, R.A.; Moynahan, P.M. (1984) An
epidemiologic investigation of health effects in air force personnel
following exposure to herbicides: baseline morbidity study results.
USAF School of Aerospace Medicine, Brooks Air Force Base, Texas.
Moses, M.; Lillis, R.; Crow, K.D.; Thornton, J.; Fischbein, A.; Anderson, H.A.;
Selikoff, I.J. (1984) Health status of workers with past exposure to
2,3,7,8-tetrachlorodibenzo-j)-dioxin in the manufacture of 2,4,5-
trichlorophenoxyacetic acid: comparison of findings with and without
chloracne. Am. J. Ind. Med. 5:161-182.
Nelson, C.J.; Holson, J.F.; Green, H.G.; Gaylor, D.W. (1979) Retrospective
study of the relationship between agricultural use of 2,4,5-T and cleft
palate occurrence in Arkansas. Teratology 19:377-384.
Patterson, D.G.; Hoffman, R.E.; Needham, L.L.; Roberts, D.W.; Bagby, J.R.;
Pirkle, J.L.; Falk, H.; Sampson, E.J.; Houk, V.N. (1986) 2,3,7,8-Tetra-
chlorodibenzo-p-dioxin levels in adipose tissues of exposed and control
persons in Missouri: an interim report. JAMA 256(19):2683-2686.
Pocchiari, F. (1980) Accidental TCDD contamination in Seveso (Italy):
epidemiological aspects. FIFRA Docket No. 415, Exhibit 1469. U.S.
Environmental Protection Agency, Washington, DC.
Pocchiari, F.; Silano, V.; Zampieri, A. (1980) Human health effects from
accidental release of TCDD at Seveso (Italy). FIFRA Docket No. 415,
Exhibit 1470. U.S. Environmental Protection Agency, Washington, DC.
Rappe, C. (1985) Problems in analysis of PCDDs and PCDFs and presence of
these compounds in human milk'. , Consultation on organohalogen compounds in
human milk and related hazards. Geneva: World Health Organization.
Reggiani, G. (1978) Medical problems raised by the TCDD contamination in
Seveso, Italy. Arch. Toxicol. 40:161-188.
Reggiani, G. (1980) Direct testimony before the U.S. Environmental Protection
Agency. FIFRA Docket No. 415, Exhibit 861. U.S. EPA, Washington, DC.
Rehder, H.; Sanchioni, L.; Cefis, F.; Gropp, A. (1978) Pathological-
embryological investigations in cases of abortion related to the Seveso
accident. Journal of Swiss Medicine 108(42):1617-1625.
Schecter, A.; Ryan, J.J.; Constable, J.D. (1987) Polychlorinated dibenzo-£-
dioxin and poiychlorinated dibenzofuran levels in human breast milk from
Vietnam compared with cow's milk and human breast milk from the North
American continent. Chemosphere 16(8/9):2003-2016.
22
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Smith, A.H.; Fisher, D.O.; Pearce, N.; Chapman, C.J. (1982) Congenital
defects and miscarriages among New Zealand 2,4,5-T sprayers. Arch.
Environ. Health 37(4):197-200.
Thomas, H.F. (1980) 2,4,5-T use and congenital malformation rates in Hungary.
Lancet ii:214-5.
Townsend, J.C.; Bodner, K.M.; Van Peenen, P.P.; Olsen, R.D.; Cook, R.R. (1982)
Survey of reproductive events of wives of employees exposed to chlorinated
dioxins. Am. J. Epidemiol. 115(5):695-713.
Tuchmann-Duplessis, H. (1977) Embryo problems posed by the Seveso accident.
Le Concours Medical No. 44.
Tuchmann-Duplessis, H. (1980a) Direct testimony before the U.S. Environmental
Protection Agency. FIFRA Docket No. 415, Exhibit 864. U.S. EPA,
Washington, DC.
Tuchmann-Duplessis, H. (1980b) Tables in direct testimony before the U.S.
Environmental Protection Agency. FIFRA Docket No. 415, Exhibit 864a.
U.S. EPA, Washington, DC.
U.S. Environmental Protection Agency. (1985) Health assessment document for
polychlorinated dibenzo-p_-dioxins. Office of Health and Environmental
Assessment, Washington, DC. EPA/600/8-84/014F. NTIS PB86-122546/AS.
U.S. Environmental Protection Agency. (1988) Estimating exposures to 2,3,7,8-
TCDD. Exposure Assessment Group, Office of Health and Environmental
Assessment, Washington, DC. External Review Draft. EPA/600/6-88/005A.
van den Berg, M.; van der Wielen, F.W.M.; Olle, K.; Van Boxtel, C.J. (1986)
The presence of PCDDs and PCDFs in human breast milk from the Netherlands.
Chemosphere 15(6):693-706.
23
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March 1988
Review Draft
APPENDIX E
IMMUNOTOXICITY OF 2,3,7,8-TCDD: REVIEW, ISSUES, AND UNCERTAINTIES
Babasaheb R. Sonawane, Ph.D.
Reproductive Effects Assessment Group
Office of Health and Environmental Assessment
Ralph J. Smialowicz, Ph.D.
Health Effects Research Laboratory
Office of Health Research
Robert W. Luebke, Ph.D.
Health Effects Research Laboratory
Office of Health Research
Office o.f Research and Development
U.S. Environmental Protection Agency
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EXECUTIVE SUMMARY - - . ,
2,3,7,8-Tetrachlorodibenzo-jp_-dioxin (2,3,7,8-TCDD) is a potent immuno-
suppressant in the laboratory animal species studied; however, the immunologi-
cal effects are apparent at exposure levels that also produce other discernible
pathological lesions and reproductive/developmental effects. There are no
unequivocal cases of significant immune function alterations in humans following
exposure to 2,3,7,8-TCDD. The cellular and molecular mechanism(s) of 2,3,7,8-
TCDD-induced immunotoxicity is unknown. Significant data gaps and uncertainties
exist to prevent an immunotoxicity-based health hazard evaluation.
INTRODUCTION
This document discusses the relevant scientific literature on the immuno-
toxicity of 2,3,7,8-tetrachlorodibenzo-j3-dioxin (2,3,7,8-TCDD) in laboratory
animals and humans. The document does not provide a comprehensive literature
review of the 2,3,7,8-TCDD-induced immune effects; however, attempts were made
to identify and discuss critical studies and issues, strengths and weaknesses
of the data, and significant uncertainties associated with the evaluation and
interpretation of the data. Furthermore', significant data gaps in knowledge
are recognized as they may relate to potential risk to the immune system of
humans upon exposure to 2,3,7,8-TCDD.
The immunotoxic effects of 2,3,7,8-TCDD have been studied by numerous
investigators for over a decade. Several recent reviews have been published
that provide a very good overview of the animal and human data dealing with
the immune alterations following exposure to 2,3,7,8-TCDD (Dean and Lauer, 1984;
Dean and Kimbrough, 1986; Thomas and Faith, 1985). The U.S. Environmental
1
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Protection Agency also reviewed the immunoloyical effects of 2,3,7,8-TCDD in
the Health Assessment Document for Polychlorinated Dibenzo-j3-Dioxins (U.S. EPA,
1985). The present review updates the immunotoxicity literature on 2,3,7,8-
TCDD and presents issues of concern and uncertainties in risk assessment.
ANIMAL STUDIES
Early experimental work with 2,3,7,8-TCDD revealed that the thymus is a
target organ of toxicity. All animal species that have been studied have con-
sistently displayed involution of the thymus and loss of cortical thymocytes
following acute or chronic exposure to 2,3,7,8-TCDD, as well as lymphocyte
depletion of T-cell areas of the lymph nodes and the spleen (Harris et al.,
1973; Zinkl et al., 1973; Gupta et al., 1973; Vos et al., 1973; 1974; Luster et
al., 1979; 1982). While these changes in the lymphoid tissues were similar to
those produced by glucocorticosteroids, adrenalectomy failed to prevent 2,3,7,8-
TCDD-induced thymic atrophy or hepatotoxicity (van Logten et al., 1980). The
administration of thymosin to mice exposed to 2,3,7,8-TCDD by postnatal maternal
treatment did not affect thymus atrophy or decrease mitogen responses (Vos et
al., 1978). Thyroidectomy, however, was found to protect rats from the immuno-
suppressive effects of 2,3,7,8-TCDD (Pazdernik and Rozman, 1984). Vos and
co-workers (1973) reported the effects of 2,3,7,8-TCDD on the immune system of
guinea pigs, rats, and mice. Several tests were employed in these three species,
and the authors concluded that the effect of 2,3,7,8-TCDD was primarily on the
cell-mediated immune function of these animals. Guinea pigs treated with
1 ug/kg of 2,3,7,8-TCDD per week either died or were moribund after four doses.
At this dose and also at lower doses (0.008, 0.04, and 0.2 ug/kg), guinea pigs
showed severe loss of body weight, depletion of the lymphoid organs, particularly
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the thymus, and lymphopenia. At sublethal doses, the effects were primarily on
the thymus. The weights of other lymphoid organs, such as the spleen, cervical
lymph nodes, and popliteal lymph nodes, were not affected. Adrenals showed a
slight enlargement at the 0.2 ug/kg dose level. Serum cortisol and corticoste-
rone values were not influenced by the 2,3,7,8-TCDD treatment. No microscopic
lesions were apparent in any of the lymphoid organs upon dosing with up to
0.2 ug/kg of 2,3,7,8-TCDD per week other than atrophy of the thymic cortex (Vos
et al., 1973).
In selected experiments, the treated and control animals were sensitized
to tetanus toxoid or killed Mycobacterium tuberculosis. Seven days after the
tetanus toxoid injection, there was a small but significant increase in the
serum, antitoxin levels at 0.008 and 0.04 ug/kg dose levels. The animals were •
challenged by a second dose of tetanus toxoid 2 weeks after'the first challenge
to evaluate the secondary response. One and 2 weeks after the second dose, the
guinea pigs treated with 0.2 ug/kg of 2,3,7,8-TCDD per week had significantly
lower tetanus antitoxin titers. The animals challenged with IM. tuberculosis
were tested for their skin reactivity 12 and 19 days later by intradermal
tuberculin injections. Both skin thickness and the diameter of the reaction
site were significantly decreased, the former being significant at the
0.04 ug/kg dose level (Z.inkl et al., 1973; Vos et al., 1973).
Female albino rats were treated orally with 0.2, 1, and 5 ug/kg of
2,3,7,8-TCDD once a week for 6 weeks. The body weights and thymus weights were
reduced at the highest dose level and adrenal weights were reduced at the ,1 and
5 ug/kg dose level. The relative thymus weight (organ/body weight ratio) was
approximately half that of the control animals at the 5 ug/kg dose level,
whereas at the same dose level, the relative spleen weights were significantly
increased. The total peripheral blood leukocyte count and lymphocyte numbers
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did not show a significant 2,3,7,8-TCDD-related effect as observed in guinea
pigs. Skin reactivity of rats challenged to tuberculin was not affected at any
2,3,7,8-TCDD dose levels (Zinkl et al., 1973; Vos et al., 1973).
Graft versus host (GVH) activity of mice was evaluated after 0.2, 1, 5,
and 25 ug/kg of 2,3,7,8-TCDD treatment per week for 4 weeks. The absolute
and relative thymus weights were decreased in groups treated with 5 ug/kg of
2,3,7,8-TCDD. Dose-related reduction in GVH activity was observed (Vos et al.,
1973).
Mice treated orally with 2,3,7,8-TCDD were challenged with either Salmonel-
la bern or Herpesvirus suis (Thigpen et al., 1975). In two separate experiments
the animals were exposed to 2,3,7,8-TCDD levels ranging from 0.5 to 20 ug/kg
once every week for 4 weeks. The animals were challenged with the infectious
agents one day after the fourth dose of 2,3,7,8-TCDD. Treatment of mice with
2,3,7,8-TCDD reduced the time of death after the bacterial challenge at the
5 ug/kg dose level whereas the increased mortality was obvious at 1 ug/kg. The
challenge with H. suis influenced neither the period of time to death nor the
mortality rate.
Vos and co-workers (1978) reported the response of young mice .to
Escherichia coli endotoxin after 2,3,7,8-TCDD treatment. There was a dose-
related increase in mortality, both with respect to 2,3,7,8-TCDD and the endo-
toxin. While the administration of 250 ug of endotoxin produced no mortality
in the control group, as little as 10 ug of endotoxin produced death in mice
treated with 15 or 50 ug 2,3,7,8-TCDD. A similar increase in endotoxin sensi-
tivity was reported when mice were treated with a single oral dose of 100 ug/kg
of 2,3,7,8-TCDD and challenged with 20 ug of endotoxin. In these experiments,
a 2,3,7,8-TCDD dose-related decrease in body, thymus, and spleen weights was
observed.
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The increased susceptibility to S_. bern reported in mice exposed to
2,3,7,8-TCDD (Thigpen et al., 1975) has been postulated to be due to an in-
creased sensitivity of 2,3,7,8-TCDD-exposed mice to bacterial endotoxin (Vos
et al., 1978), although variable effects have been reported in host resistance
following 2,3,7,8-TCDD exposure (Dean and Lauer, 1984; Vos et.al., 1978; Luster
et al., 1980). Other studies have shown a correlation between increased mortal-
ity from bacterial or parasitic infections and decreased serum complement levels
(White et al., 1986) or decreased B-cell mediated responses (Tucker et al., 1986)
in 2,3,7,8-TCDD-exposed mice, respectively.
Clark et al., (1981) have shown that low doses of 2,3,7,8-TCDD (4 ng/kg to
0.4 ug/kg) administered intraperitoneally once a week for 4 weeks suppress the
generation of cytotoxic T-cells by lymph node and spleen cells in male C57BL/6
mice but have little effect on delayed hypersensitivity, antibody responses, or
thymic cellularity. Suppressor cells capable of blocking the cytotoxic T-cell
response were found in the thymus of 2,3,7,8-TCDO-treated mice following cumula-
tive doses of 2,3,7,8-TCDD as low as 4 ng/kg. The immunosuppressive effects
observed at 4 ng/kg were significantly less than the 4 ug/kg required to induce
thymic atrophy. In 1983, the same investigators (Clark et al., 1983) reported
that susceptibility to 2,3,7,8-TCDD-induced immunosuppression in mice is strain-
dependent, and occurs at doses that have little effect on hepatic microsomal
enzymes. Clark et al. also reported that other types of haloaromatics can
induce similar immunosuppression provided they possess sufficient binding
affinity for the genetically controlled 2,3,7,8-TCDD-receptor protein. These
observations suggest a receptor-dependent mechanism for the stimulation of
suppressor T-cell activity by haloaromatic hydrocarbons.
Examination of T-cell-mediated responses in 2,3,7,8-TCDD-exposed animals
has generally demonstrated a correlation between thymic atrophy and impaired
-------
cell-mediated immunity. Exposure of mice, rats, and/or guinea pigs to
2,3,7,8-TCDD has been reported to result in depressed delayed hypersensitivity
(Vos et al., 1973; Faith and Moore, 1977; Clark et al., 1981), prolonged allo-
graft rejection (Vos and Moore, 1974), depressed GVH response (Vos and Moore
1974), decreased responses to T-cell mitogens and/or allogeneic cells in vitro
(Dean and Lauer, 1984; Vos and Moore, 1974; Luster et al., 1980), and decreased
generation of cytotoxic T lymphocytes (CTL) (Clark et al., 1981, 1983).
Depressed CTL activity in adult mice following exposure to 2,3,7,8-TCDD has
been reported to be associated with an increase in suppressor T-cells, but not
associated with a reduction in the frequency of CTL precursors (Clark et al.,
1981; 1983; Nagarkatti et al., 1984).
Nonspecific immune responses affected by natural killer cells (Dean and
Lauer, 1984; Mantovani et al., 1980) and jnacrophages (Dean and Lauer, 1984; Vos
et al., 1978; Mantovani et al., 1980) have not been observed to be affected by
2,3,7,8-TCDD exposure. Both of these cell types possess tumoricidal ,
bactericidal, and virucidal activity.
2,3,7,8-TCDD has also been reported to affect bone marrow and humoral
immune responses of experimental animals. Exposure of mice to 2,3,7,8-TCDD
resulted in myelotoxicity and suppression of bone marrow progenitor cells
(Tucker et al., 1986; Luster et al., 1985; Chastain and Pazdernik, 1985).
Studies in mice exposed to 0, 1.0, 5.0, or 15 ug/kg body weight of 2,3,7,8-TCDD
pre- and postnatally by maternal dosing indicated that both 5 and lb ug/kg
dosage groups had a significant reduction in bone marrow cellularity, (colony-
forming unit-spleen (CFU-S) or pluripotent stem cells and colony-forming unit-
granulocyte/macrophage (CFU-GM). He'matology profiles and blood smears
revealed a normocytic anemia in; these mice (Luster et al., 1980). Bone marrow
toxicity correlated with depressed immunologic and host-resistance responses.
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The hematopoietic stem cells have a limited renewal capacity, and damage to
these cells can induce a permanent decrease in their pro!iferative capacity;
however, this limited evidence of bone marrow toxicity, as observed in mice,
is difficult to evaluate due to the lack of additional significant studies in
other species or humans.
The antibody response of 2,3,7,8-TCDD-exposed mice, as measured by the
plaque-forming cell (PFC) assay, was suppressed following immunization with
T-cell dependent and/or T-cell independent antigens (Tucker et al., 1986;
Vecchi et al., 1983; Holsapple et al., 1986). Furthermore, these effects on
B-cell-mediated responses were observed to occur at 2,3,7,8-TCDD doses below
those that cause thyrnic atrophy (van Logten et al., 1980; Tucker et al., 1986;
Luster et al., 1985; Vecchi et al., 1983; Holsapple et al., 1986).
Susceptibility to suppression of humoral immune responses and 2,3,7,8-TCDD
induction of the aryl hydrocarbon hydroxylase (AHH) system were found to corre-
late (Luster et al., 1980; Tucker et al., 1986; Chastain and Pazdernik, 1985;
Vecchi et al., 1983; Holsapple et al., 1986), as does thymic atrophy (Poland and
Glover, 1980), T-cel1-mediated responses (Nagarkatti et al., 1984), and serum
complement levels (White et al., 1986). A good correlation between the degree
of AHH inducibility and immunosuppression was observed in Fj crosses of
"responsive" (Ahb/Ahb) and "nonresponsive" (AhdAhd) mice (Nagarkatti et al.,
1984; Vecchi et al., 1983). These results, taken together, indicate that there
is a strong association between the presence of the Ah receptor and the induc-
tion of immune effects following 2,3,7,8-TCDD exposure in experimental animals,
and susceptibility to 2,3,7,8-TCDD-induced immune effects in the mouse may be
under genetic influences.
The immunological reactivity of young or/suckling rats and mice has been
evaluated using a variety of experimental protocols .employing prenatal or post-
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natal exposure of mothers or a continued prenatal through postnatal exposure
with 2,3,7,8-TCDD. One such study reported by Vos and Moore (1974) is summa-
rized below. Pregnant rats were treated orally with 1 or 5 ug/kg body weight
of 2,3,7,8-TCDD on day 11 and 18 of gestation. The weights of 1-day-old pups
from mothers treated with 5 ug/kg dose level were significantly reduced. At
this dose level, reduced weights of the thymus and spleen were also observed.
Most of the pups in the high-treatment group died within 25 days after birth.
The newborn animals from mothers given a 1 ug/kg dose level were further
treated with the same dose of 2,3,7,8-TCDD administered to mothers on day 4,
11, and 18. Additional groups of suckling rats were exposed to 2,3,7,8-TCDD
postnatally by oral treatment of mothers with 5 ug/kg. Reduction in body,
thymus, spleen, and adrenal weights were observed in different groups at 25
days of age. Splenic lymphocytes of rats exposed postnatally to the 5 ug/kg
level showed a decrease in phytohemagglutinin (PHA)-induced DNA synthesis.
This effect was not observed in animals exposed pre- and postnatally to 1
ug/kg of 2,3,7,8-TCDD. Thymocytes cultured from postnatally exposed male rats
to 5 ug/kg of 2,3,7,8-TCDD showed a significant reduction of thymidine incor-
poration in the presence of PHA. DNA synthesis in response to concanavaltn A
(Con-A), however, was not reduced in thymocytes. The authors (Vos and Moore,
1974) reported the insensitivity of 4-month-old mice to 2,3,7,8-TCDD treatment.
One-month-old mice treated with four weekly doses of 25 ug/kg of 2,3,7,8-TCDD
showed a decreased responsiveness of their splenic lymphocytes to PHA, whereas
5-month-old animals failed to show this effect after six weekly doses.
Similar effects were observed on GVH activity of spleen cells from 25-day-
old pre- and/or postnatally exposed rats. Only those animals treated with 5
ug/kg postnatally showed a significant decrease in GVH activity of spleen cells,
Reaction times of heterologous skin grafts was prolonged in rats exposed in
-------
utero and in mice exposed pre- and postnatally to 2,3,7,8-TCDD (Vos and Moore,
1974).
In summary, exposure of mice, rats, and/or guinea pigs (as described
earlier) to 2,3,7,8-TCDD during the perinatal period resulted in thymic atrophy
(Faith and Moore, 1977; Vos and Moore, 1974; Luster et al., 1980; Moore and
Faith, 1976'; Faith et al., 1978). These animals exhibited a "wasting syndrome"
that is similar to that seen in neonatally thymectomized animals that have been
treated with corticosteroids (Thomas and Faith, 1985). 2,3,7,8-TCDD adminis-
tered during immune ontogenesis has been observed to affect immune responses
more dramatically than when administered to adults. Generally, thymic atrophy,
suppressed T-cell mediated responses, and bone marrow toxicity have been report-
ed to be more profound following pre- and/or postnatal exposure to 2,3,7,8-TCDD
than with adult exposure (Dean and Lauer, 1984; Dean and Kimbrough, 1986;
Thomas and Faith, 1985; Luster et al., 1979; 1980; Faith and Moore, 1977; Vos
and Moore, 1974). These results suggest that the developing immune system is
more susceptible to 2,3,7,8-TCDD-induced alterations and, consequently, that
the very young may be at a higher risk than adults to the immunotoxic effects
of 2,3,7,8-TCDD.
Exposure of animals to 2,3,7,8-TCDD has been shown to decrease the respon-
siveness of lymphocytes to various mitogens in culture. For example, rabbits
exposed for 8 weeks to 2,3,7,8-TCDD at 0.01 to 10 ug/kg/week and challenged
with tetanus toxoid had a decreased PHA response at the highest dose (Sharma
et al., 1979). Sharma et al. (1979) also reported that immunologic effects of
a single dose of 2,3,7,8-TCDD in mice were reversible during an 8-week period
while the hepatic lesions persisted. Mice were treated with a single oral
dose of 10 ug/kg of 2,3,7,8-TCDD, and selected animals were sacrificed at 2, 4,
and 8 weeks after this dosing. The spontaneous increase of DNA synthesis in
-------
splenic cultures and a decreased responsiveness of splenic lymphocytes to phy-
tohemagglutinin and pokeweed mitogen was apparent at 2 weeks after the treat-
ment. The effects persisted up to 4 weeks after the administration of
2,3,7,8-TCDD, but were not noticed when the spleens were obtained from animals
at 8 weeks after the 2,3,7,8-TCDD dosing. Lymphocyte depletion in the thymus
and reduced thymus weights showed a partial recovery at this time.
The influence of direct addition of 2,3,7,8-TCDD to mouse-splenic cultures
was reported by Sharma and Gehring (1979). 2,3,7,8-TCDD decreased the unstimu-
1 ated DNA synthesis in lymphocytes at concentrations as low as 10~9 M; however,
no effects were observed in mitogen-stimulated cultures. Vos and Moore (1974)
reported that the addition of 2,3,7,8-TCDD up to 0.01 ug/0.5 mL in culture
medium did not alter the DNA synthesis in mouse spleen or rat thymus cells,
either with or without the presence of phytohemagglutinin or Con-A. Luster and
co-workers (1979) exposed the spleens from mice to 2,3,7,8-TCDD in dimethylsul-
foxide. DNA, RNA, and protein synthesis in the spleens were inhibited at
2,3,7,8-TCDD concentrations of 10~7 M. The ability of lymphocytes to bind with
mitogens was not influenced by 2,3,7,8-TCDD.
The mechanism(s) for 2,3,7,8-TCDD-induced immunosuppression is not fully
known. However, genetic and structure-activity relationship studies have
provided evidence that 2,3,7,8-TCDD-induced immunosuppression in mice is asso-
ciated with the presence of the Ah locus and is mediated through a 2,3,7,8-TCDD
cytosol receptor (Nagarkatti et al., 1984; White et al., 1986; Tucker et al.,
1986; Luster et al., 1985; Vecchi et al., 1983; Holsapple et al., 1986; Poland •
and Glover, 1980; Dencker et al., 1985; Kouri and Ratrie, 1975). Some evidence
suggests that thymic epithelial cells may be the principal target for 2,3,7,8-
TCDD-induced immunotoxicity (Clark et al., 1983; Greenlee et al., 1985). Bind-
ing of 2,3,7,8-TCDD to receptors in the thymus may promote altered T-cell matu-
10
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ration and differentiation, and it may be the molecular basis for the observed
thymic atrophy and immunotoxicity. Other work, however, suggests that hemato-
poietic stem cells and B-cells may also be targets for 2,3,7,8-TCDD-mediated
immune effects. For example, it has recently been shown that 2,3,7,8-TCDD selec-
tively inhibits the differentiation of B-cells into antibody-secreting cells in
vitro (Tucker et al., 1986) and inhibits bone marrow stem cell colony growth in
vivo and in vitro (Luster et al., 1985). The binding of 2,3,7,8-TCDD to recep-
tors on lymphocytes, thymus epithelial cells, and/or hematopoietic precursor
cells may cause alterations in maturation and differentiation that may result
in the immune alterations observed in animals following in vivo exposure to
2,3,7,8-TCDD, Further work is clearly needed to investigate the mechanism(s)
for 2,3,7,8-TCDD-induced immune effects in order to make better estimates of
the potential risks associated with exposure of humans to 2,3,7,8-TCDD.
HUMAN STUDIES '..'-.
For a variety of reasons, humans have been exposed to 2,3,7,8-TCDD in the
environment. The immune function has been examined in individuals with pro-
bable or known exposure to 2,3,7,8-TCDD using assays designed to evaluate the
component parts of the immune system. Several accounts of immune functions in
humans exposed to 2,3,7,8-TCDD have been referred to in summary-type articles
reviewed in the preparation of this paper. Some of these reports have not
been published in the scientific literature or are anecdotal. However, none
of these reports, in the opinion of the reviewers (Dean and Lauer, 1984; Dean
and Kimbrough, 1986; Marshall, 1986), presented convincing evidence for altered
immune function in the exposed populations. In one study, no abnormalities in
measured immune parameters were observed in military personnel who had been
11
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involved in the spraying of 2,3,7,8-TCDD-contaminated Agent Orange during
operation Ranch Hand in Vietnam (conversation between J. Silva, CHHS, Test and
Evaluation Activity, Bethesda, MD, and Dr. Ralph Smialowicz, U.S. Environmental
Protection Agency, February 6, 1987).
Attempts have been made to study immune functions in populations that have
been exposed to 2,3,7,8-TCDD. All of these attempts have been complicated by
technical difficulties in both design and execution, in spite of efforts by
the researchers to control for all possible variables. Three such studies are
summarized.
In July of 1976, an uncontrolled chemical reaction at a chemical plant in
Seveso, Italy, resulted in the release of an estimated 300 g of 2,3,7,8-TCDD
mainly into an uninhabited area. However, a significant number of people were
potentially exposed, and thus, a major health effects surveillance effort was
initiated. The results of this study, including immunologic assessment, have
been published (Homberger et al., 1979; Reggiani, 1980). A group of 45
2,3,7,8-TCDD-exposed children, 20 of whom had chloracne, and 44 children without
2,3,7,8-TCDD exposure were evaluated immunologically every 4 months for approxi-
mately 1.5 years. These studies revealed no differences between the two groups
in serum immunoglobulin or complement levels or in the ability of their T- and
B-cells to respond to mitogens in vitro. Critical evaluation of these data are
not possible, however, since quantitative data were not presented. It should
be noted that a review (Tognoni and Bonaccorsi, 1982) of the Seveso incident,
published 5 years after the fact, reported increased serum complement hemolytic
activity and significantly higher lymphocyte proliferative,responses in exposed
subjects. However, no quantitative data were presented, the patient population
was not identified per se, and no reference was made as to how or when these
data were collected. Thus, critical review and interpretation of the reported
12
-------
findings are both impossible and unadvisable.
The town of Times Beach, Missouri, was contaminated with 2,3,7,8-TCDD in
1972 and 1973 when waste from a chemical plant was sprayed on roadways to
control dust. Soil samples were tested for 2,3,7,8-TCDD 10 years later, and
levels of contamination were so great that the entire town was purchased under
the provisions of the Superfund law (Powell, 1984). Selected residents were
subsequently classified as having had high or low risk of 2,3,7,8-TCDD exposure
and were evaluated immunologically (Knutsen, 1984). Tests of delayed type
hypersensitivity (DTH) to a standard battery of skin test antigens, lymphocyte
blastogenic responses to mitogens, and T-cell subset analysis revealed no sig-
nificant differences between the high- and low-risk groups. However, there was
a tendency among members of the high-risk group to respond to fewer skin test
antigens and to have slightly different T-cell subset profiles than low-risk
group members. Lymphocytes from children'in the high-risk group also had a
decreased proliferative response to tetanus toxoid compared with those from
children in the low-risk group. It should be noted, however, that in a prelim-
inary study of an unspecified population of Missouri residents exposed to
2,3,7,8-TCDD, no differences were detected in T-cell subset profiles, skin test
responses, or lymphocyte proliferation responses (Stehr et al., 1986).
Residents of the Quail Run Mobile Home Park in Gray Summit, Missouri, were
exposed for various lengths of time to 2,3,7,8-TCDD-contaminated soil. Contami-
nation was the result of dust control efforts using waste oil containing chemi-
cal, sludge from the same source as in Times Beach. Soil sampling studies
revealed levels of 2,3,7,8-TCDD ranging from 39 to 2,200 parts per billion.
Immunologic testing was performed on residents who had lived in the park for
at least 6 months (N = 154), and their results were compared to residents
(N = 155) of other mobile parks in areas where no evidence of 2,3,7,8-TCDD
13
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contamination was found by soil sampling (Hoffman et al., 1986). DTH to skin
test antigens, antigen- and mitogen-stimulated lymphocyte proliferation
responses, lymphocyte subset analysis, cytotoxic T-lymphocyte responses, and
serum immunoglobulin levels were evaluated in both populations, Some members
of the high-risk group did not respond to skin test antigens (anergy) and, of
those that did respond, positive reactions were obtained for fewer antigens
than in the unexposed group (relative anergy). The exposed group had an in-
creased frequency of anergy (11.8% vs. 1.1%) and relative anergy (35.3% vs.
11.8%). It must be pointed out that there were significant technical problems
with the interpretation of the skin test responses and, as a'result, nearly 50%
of the data had to be discarded. Several skin test readers, with various
amounts of experience, were employed in this investigation. In addition, two
of the four readers who had received special training in the interpretation of
DTH skin tests recorded anergy in 15% or 40% of the control group, a rate 75 or
200 times the expected rate. Although allowances were made for'this by the
investigators, the DTH data in this study are thus questionable. The mean
ratio of T helper/inducer cells to T suppressor/cytotoxic cells was similar in
both groups, although there was a nonsignificant proportion of exposed group
members with a T4/T8 ratio less than 1.0 (8.1% vs. 6.4%). The report likewise
states that 12.6% of the exposed group versus 8.5% of the low-risk group had
abnormal in vitro T-cell functions, although examination of the tabular data
provides no evidence whatsoever for a difference between levels of immunocom-
petence in the two groups. The authors concluded that long-term exposure to
2,3,7,8-TCDD is associated with depressed cell-mediated immunity, although the
effects have not resulted in an excess of clinical illness. Individuals from
this initial study (Hoffman et al,, 1986) have been re-evaluated, and the
results of this follow-up study, as reported by Evans et al. (1987) at the
14
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International Conference on Dioxin, failed to corroborate the report of anergy
in the 2,3,7,8-TCDD exposed cohort of the initial study (Hoffman et al., 1986).
UNCERTAINTIES, DATA GAPS, AND RESEARCH NEEDS
The immune system, unlike many organ systems, is self-renewing from a
pool of pluripotent stem cells. Functional cells are generally end-stage
cells that have a limited lifetime and are replaced on a regular basis.
Therefore, unless the store of stem cells is destroyed or unless there is a
permanent blockade of cellular differentiation, acute chemical exposure is
unlikely to cause long-term suppression of the host-defense system. However,
the uncertainty in the premise of self-limited chemical immunotoxicity resides
in the unknown effects of chronic or repeated acute, exposure to immunotoxic
agents on the regenerative capacity of the immune system.
Although a great amount of time and effort has gone into evaluating the
immunotoxic effects of 2,3,7,8-TCDD in experimental animals, there,are a
number of uncertainties that remain to be resolved. These include but are
not limited to the following: (1) the lack of a strong association between
2,3,7,8-TCDD exposure and decreased host resistance, (2) the different suscep-
tibility of species and strains to 2,3,7,8-TCDD-induced immunosuppression, (3)
the transient and reversible nature of 2,3,7,8-TCDD-induced immune effects,
(4) the apparent but not proven increased susceptibility of very young animals
to 2,3,7,8-TCDD-induced immune effects, (5) the uncertainty of the mechanism(s)
of 2,3,7,8-TCDD-induced immune effects, and (6) the lack of evidence between
observed immunological effects in animals and the questionable immune altera-
tions reported in human populations inadvertantly exposed to 2,3,7,8-TCDD.
These areas of uncertainty are briefly discussed.
15
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Inconsistent results have been reported regarding the susceptibility of
2,3,7,8-TCDD-exposed animals to tumor development and infection (Dean and
Lauer, 1984; Faith and Moore, 1977; Luster et al., 1980; Thigpen et al., 1975;
White et al., 1986; Tucker et al., 1986). Despite the fact that exposure to
2,3,7,8-TCDD results in depressed T- and B-cell responses, further research is
necessary to define the conditions under which host resistance is adversely
affected by 2,3,7,8-TCDD.
In general, the experimental animal data suggest that species differences
exist in 2,3,7,8-TCDD sensitivity. With the exception of the earliest work,
the mouse has been the predominant species for studying the immunotoxic effects
of 2,3,7,8-TCDD. This is most probably due to the fact that more is known
about the immune system of the mouse than any other animal species, as well as
the fact that there are more validated methods available for examining the
immune system in this species. Nevertheless, extensive interspecies comparisons
among several animal species are warranted considering their pharmacokinetic
differences. These studies are needed not only to substantiate and corroborate
the results of studies with the mouse but also to provide the framework to
extrapolate the potential for 2,3,7,8-TCDD to affect the human immune system.
Interspecies studies are also necessary in order to determine if an association
exists between the Ah locus and 2,3,7,8-TCDD-induced immune effects in species
other than the mouse. This includes extension of in vitro studies which have
demonstrated the presence and inducibility of AHH in human lymphoid tissue
(Kouri and Ratrie, 1975). Hopefully, these studies will provide useful infor-
mation about the role that the Ah locus may play in the susceptibility of the
human population to 2,3,7,8-TCDD.
The immune effects that have been observed following 2,3,7,8-TCDD exposure
have, in all cases where it has been examined, returned to normal levels over
16
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a period of time after the cessation of exposure to 2,3,7,8-TCDD. This is
probably a result of the plastic nature of the immune system, >as wel 1 as the
fact that chronic dosing with 2,3,7,8-TCDD has not been performed. It is not
clear whether and how repeated doses might affect the immune, system or whether
short-term exposure could result in irreversible effects. Low-dose chronic
studies in animals are needed to determine whether such exposures not only
produce immune alterations but also to determine if long-lasting impairment is
produced. Low-dose chronic studies are important to extrapolate data from the
high dosages used under experimental situations because they are most likely
to mimic the form of human exposure to 2,3,7,8-TCDD in the environment.
Exposure to 2,3,7,8-TCDD during immune system development has been shown
to affect the immune system of experimental animals (Faith and Moore, 1977;
Vos and Moore, 1974; Luster et al., 1980; Faith et al., 1978). The effects
produced following perinatal exposure have been reported to be more profound
than those produced following adult exposure, although in some cases a high
degree of fetal toxicity has been reported (Vos and Moore, 1974). While this
may be true, no attempt has been made to test this hypothesis by making a
direct comparison between perinatal and adult exposure to 2,3,,7,8-TCDD using
identical exposure regimens (i.e., B- and T-cell function, host resistance
models, natural killer cell activity, etc.). Work is warranted in this area
in order to provide evidence of an increased risk of the developing immune
system to 2,3,7,8-TCDD exposure. This is necessary so that an informed judg-
ment can be made as to the potential increased relative risk to infants and
children exposed to 2,3,7,8-TCDD.
There is still a great deal of uncertainty about the mechanism(s) by which
2,3,7,8-TCDD causes immune alterations in animals. Work with other animal
species and in vitro work with human tissues will hopefully provide new insights
17
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In this area. A clearer understanding of the mechanisms of 2,3,7,8-TCDD-
induced immunosuppression in animals will be invaluable for extrapolation of
the potential health risk in humans.
Laboratory studies have clearly demonstrated the immunotoxic effects of
2,3,7,8-TCDD in a variety of animal models. Extrapolation of these data to
predict effects of human 2,3,7,8-TCDD exposure are complicated at best, since
actual exposure levels, route of exposure, and even comparability of human and
animal susceptibility to the toxic effects of 2,3,7,8-TCDD are unknown. Results
of rodent studies suggest that the immune responses of children should be more
sensitive to the immunotoxic effects of 2,3,7,8-TCDD than that of adults (Faith
and Moore, 1978). However, immune function was followed in children from the
most heavily contaminated area of Seveso for 18 months after the accident and
appeared to be normal (Homberger et al., 1979; Reggiani, 1980). Furthermore,
baseline data for the immune system of children is not readily available, and
the normal response in children of various ages is not well defined. Further-
more, altered immunocompetence has been reported in 2,3,7,8-TCDD-exposed
residents of Missouri (Knutsen 1984; Hoffman et al., 1986), although the differ-
ences between mean values for control and exposed populations were not statis-
tically significant. Suppression was not detected in functional parameters;
rather, trends in the distribution of exposed group members into "normal" and
"abnormal" response categories were cited as indicative of immune dysfunction.
While these trends may or may not be related to 2,3,7,8-TCDD exposure, there has
been no report of an increase in clinical illness attributable to suppressed
immune function. Thus, it appears to be that there are no unequivocal cases of
significant immunotoxicity in humans following 2,3,7,8-TCDD exposure (Dean and
Lauer, 1984; Dean and Kimbrough, 1986; Marshall, 1986; Evans et al., 1987).
18
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The available animal data reviewed above suggest that 2,3,7,8-TCDD-induced
thymus atrophy and immune alterations may result from direct actions on peri-
pheral lymphocytes or progenitor lymphoid cells in the bone marrow and through
altered differentiation of intrathymic precursor cells, specifically by a direct
action on thymic epithelium. The induction of T-cells, cytotoxic for tumor
target cells, was found to be impaired in mouse studies conducted by Clark et
al. (1981) following total exposure to a 2,3,7,8-TCDD dose of 4 ng/kg over a
4-week period. These dosages are below those levels significantly altering
other cell-mediated immune responses (Clark et al., 1981); however, these
results are unconfirmed and questionable. For example, the kinetics of CTL
suppression was not defined in the Clark et al. studies (1981, 1983), and the
effect of 2,3,7,8-TCDD on CTL-mediated tumor resistance remains to be resolved.
Additional concern is raised about Clark's findings (Clark et al., 1981) by
researchers at the Chemical Industry Institute of Toxicology, (CUT) who claim
that the CTL response was not suppressed at a dose of 4 ng/kg (conversation
between Jack Dean, CUT, Research Triangle Park, NC, and Bob Sonawane, U.S.
Environmental Protection Agency, February 17, 1987). The £059 (dose producing
50% maximal response) for the induction of thymic atrophy in sensitive mouse
strains is approximately 10 umol/kg (Poland and Glover, 1980) and for antibody
plaque-forming cell (PFC) suppression varies by 30-fold (1 ug/kg to 30 ug/kg)
between C57BL/6 and DBA/2 mice (Dean and Lauer, 1984). Furthermore, the issue
of host-resistance effects following exposure to 2,3,7,8-TCDD is unresolved.
In summary, it may be inappropriate to derive an immunotoxicity-based
hazard assessment for 2,3,7,8-TCDD from mostly acute and/or subchronic type
studies. A three-generation reproductive study by Murray et al. (1979) demon-
strated the critical noncancer end point for adverse effects at 1 ng/kg/day
compared to questionable immunosuppressive effects observed in mice by Clark
19
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et al. (1981) at 4 ng/kg by parental administration. It seems that the level
of concern identified in either of these studies or in chronic bioassays for
carcinogenicity is essentially the same. The magnitude of differences rein-
forces a common conclusion that the biological significance of any immunological
effects observed in laboratory animals is not adequately established to support
its use as the critical end point in hazard evaluation of 2,3,7,8-TCDD exposure
to humans.
20
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Toxicol. Appl. Pharmacol. 72:169-176.
Pazdernik, T.L.; Rozman, K.K. (1984) Effect of thyroidectomy and thyroxine
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36:695-703.
Poland, A.; Glover, E. (1980) 2,3,7,8-tetrachlorodibenzo-£-dioxin: segre-
gation of toxicity with the Ah locus. Mol. Pharmacol. 17:86-94.
Powell, R.L. (1984) Dioxin in Missouri: 1971-1983. Bull. Environ. Contam.
Toxicol. 33:648-654.
Reggiani, G. (1980) Acute human exposure to TCDD in Seveso, Italy.
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"toxicologic effects of a single exposure of 2,3,7,8-tetrachlorodibenzo-£-
dioxin in mice. Trace Substances Environ. Health 12:473.
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23
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Tucker, A.N.; Vore, S.J.; Luster, M.I. (1986) Suppression of B cell differen-
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24
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June 1988
Review Draft
APPENDIX F
RATIONALE FOR A HORMONE-LIKE MECHANISM
OF 2,3,7,8-TCDD FOR USE IN RISK ASSESSMENT
Michael A. Gallo
Professor and Chief, Division of Toxicology
Department of Environmental and Commmunity Medicine
University of Medicine and Dentistry of New Jersey
Robert Wood Johnson Medical School
675 Hoes Lane
Piscataway, NJ 08854-5635
Office of Health and Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
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The mechanisms of action of 2,3,7,8-tetrachlorodibenzo-|)-dioxin (2,3,7,8-
TCDD) have been intensely investigated since the pioneering work of Kimmig and
Schultz (1957) in elucidating the chloracnegen in the chlorophenol processes.
The most complete and thought-provoking treatise on the subject is that of
Poland and Knutson (1982). These authors and others in Poland's laboratory
drew upon their work and others to present a unified hypothesis to account for
the varied responses in animals exposed to 2,3,7,8-TCDD. In essence, this
hypothesis, which is partly based on the classic receptor theories invoked for
steroid action, suggests that (1) there is a cytosolic receptor for
arylhydrocarbons (the Ah receptor) that binds several compounds and then
translocates to the nucleus,' and (2) there is a second stage to the toxic
reaction(s) that is related to, but not congruent with, the induction of
cytochrome Pi-450. Activation of this receptor leads to a cascade of reactions
culminating with the association of the receptor-2,3,7,8-TCDD complex with
nuclear DNA. This association leads to the synthesis of a specific microsomal
protein designated as cytochrome Pj-450. To date, the chemical that binds this
putative receptor with the greatest avidity is 2,3,7,8-TCDD. However, many
other halogenated and non-halogenated compounds also bind to this cytosolic
protein (Nebert et a!., 1972). The xenobiotics with the greatest affinity are
those that are planar, with two phenyl rings and contain substitutions in the
lateral positions. Several investigative teams have examined the structure
activity relationships (SARs) among the polyhalogenated biphenyls (PCBs, PBBs),
polychlorinated dibenzo-j)-dioxins (PCDDs), and polychlorinated dibenzofurans
(PCDFs) (Knutson and Poland, 1980). Safe and his co-workers have synthesized
several of the highly active PCBs and have demonstrated remarkable SARs for
several biological end points (Mason et al., 1987). Excellent reviews on the
1
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SAR for PCDDs, PCDFs, and PCBs can be found in recent issues of the Annual
Reviews in Pharmacology and Toxicology (Vols. 22 [Poland and Knutson, 1982] and
26 [Safe, 1986], respectively). The hypothesis also states that ". . .there is
a second stage to the toxic reactions of 2,3,7,8-TCDD that are related to, but
not congruent with, the induction of cytochrome Pj-450." This portion of the
hypothesis is supported by the reports of Poland and his co-workers with XB
cells in culture (Knutson and Poland, 1980), Safe and co-workers' findings of
PCB inhibition of 2,3,7,8-TCDD toxicity (Haake et al., 1987), and the Umbreit
et al. report of apparent maximal induction of arylhydrocarbon hydroxylase (the
major system affected by cytochrome Pj-450) (Umbreit et al., 1987) by complex
mixtures of polyaromatic hydrocarbons and very low bioavailability of
PCDDs/PCDFs as measured by tissue levels and signs of toxicity.
At this point in time, it appears clear that 2,3,7,8-TCDD is working
through the proposed receptor mechanism for the first phase of its activity
(i.e., binding to a putative cytosolic receptor with subsequent induction of
P-450). However, all of the biological effects of this molecule cannot be
explained by simple receptor binding and induction of cytochrome PI-450. The
recent evidence from several laboratories has expanded on the initial studies
of Poland and Knutson (1982), Neal et al. (1982), and Barsotti et al. (1979) to
show quite dramatically that 2,3,7,8-TCDD markedly affects the interaction of
steroids with their respective receptors (Romkes et al., 1987; Gallo et al.,
1986) and 2,3,7,8-TCDD alters the number of Epidermal Growth Factor (EGF)
receptors in susceptible cell lines (Matsumura et al., 1984). Molloy et al.
(1987) recently reported the alteration of specific epidermal keratins in the
HRJ/S strain of mice after treatment with 2,3,7,8-TCDD. This finding is
especially relevant to the database since it was in this strain of mice that
-------
Poland and Knutson (1982) reported the model for chloracne. Matsumura et al.
(1984) studied the role of 2,3,7,8-TCDD and EGF receptors, while several
laboratories (Safe, 1986; Gierthy et al., 1987; Umbreit and Gallo, 1988;
Goldstein et al., 1987) have been pursuing the interactions of 2,3,7,8-TCDD
with steroids, primarily estrogen-sensitive tissues. The interactions with
glucocorticoids has been studied extensively by several laboratories (Luster et
al., 1984; Sunihara et al., 1987). In general, the response can be summarized
as a decrease in the number of available cytosolic receptors for EGF or the
steroids without a decrease in the affinity for the respective ligand. This
phenomenon is termed "down-regulation" of the cytosolic receptor. The
measurement of receptor binding is biochemical. The estimate of affinity and
binding site number is by extrapolation of the response curve(s) by Scatchard
analysis (1949). The strengths of this analysis are obvious, but the
weaknesses are difficult to reconcile. The major weakness is the lack of
direct binding information; this does not allow segregation of the receptors by
tertiary structure nor does it completely account for nonspecific binding. The
second weakness of ligand binding experiments is the inability of the analysis
to shed any light on the reason for the down-regulation. It must be emphasized
at this point that none of the steroid receptor research has demonstrated an
antagonism between 2,3,7,8-TCDD and the endogenous steroid for the respective
steroid receptor, nor has any competitive binding of steroids by the Ah
receptor been demonstrated. However, the steroid receptors are products of a
supergene family which is responsible for the protein synthesis of all these
receptors (Nebert et al., 1972), and the Ah receptor has many of the structural
and functional characteristics of the steroid receptors (Poellinger et al.,
1986).
-------
To better understand the role of 2,3,7,8-TCDD in cellular function (or
dysfunction), one must look to the results of the laboratories working on the
mechanisms of action of 2,3,7,8-TCDD at the molecular level. The major groups
involved in this research are Poland, Nebert's group at the National Institutes
of Health, and Whitlock's laboratory at Stanford University. As stated above,
Poland established the role of the Ah receptor in some of the actions of
2,3,7,8-TCDD. Whitlock (1987) has summarized his data and that of other
investigators regarding the regulation of the cytochrome P-450 gene family,
along with the data supporting the hypothesis that the Ah locus is part of a
super gene family responsible for the metabolism of xenobiotics and endogenous
compounds. Recent work in this laboratory has also elucidated a region on DNA,
which is sensitive to the 2,3,7,8-TCDD-cytosolic receptor complex, upstream
from the cytochrome Pj-450 gene (Neuhold et al., 1986). These findings are
critical in light of the findings of the 2,3,7,8-TCDD- responsive gene
expression enhancer system (region) described by Whitlock (Jones et al:, 1986).
Hence, the two laboratories have defined the regulatory mechanisms by which
2,3,7,8-TCDD controls gene expression of the cytochrome Pj-450 (Whitlock,
1987). The significance of these findings for 2,3,7,8-TCDD risk analysis is
the congruency between gene regulation for the Ah receptor, the glucocorticoid
receptor, and the estrogen receptor (Becker et al., 1986). The importance of
these findings cannot be underestimated. There is direct analogy with the
steroid receptor mechanisms and the control of the steroid receptor messenger
RNA (Yamamoto, 1985). The role of the estrogen receptor (ER), and other
steroid receptors, is understood to a greater extent than the Ah receptor
probably because of the greater emphasis on the physiology of steroids. The
analogy between the receptor complexes and DNA leads to the obvious comparison
-------
of effects of 2,3,7,8-TCDD and steroids. Many of the changes seen in animals
after dosing with 2,3,7,8-TCDD mimic estrogen or antiestrogen effects. Umbreit
and Gallo (1988) reviewed these findings, which are presented in Table 1.
Kociba et al. (1978) demonstrated the hepatocarcinogenic effect of orally
administered 2,3,7,8-TCDD, but in the same study there was a marked
dose-dependent decrease in tumors of the mammary glands and uteri which are
estrogen-sensitive organs. These are highly significant observations which
have been pursued by some laboratories. If 2,3,7,8-TCDD is acting through
hormonal (estrogen) mechanisms, then alteration of ovarian function, exogenous
estrogens, or antiestrogens should modify the response(s) to 2,3,7,8-TCDD.
Recent results have shown that 2,3,7,8-TCDD effects can be overridden by
exogenous estradiol (Gallo et al., 1986) and the down-regulation of the
estrogen receptor is also antagonized by estradiol (Romkes et al., 1987). The
significance of these findings are amplified if one couples the reports of
regulation of the EGF receptor by estrogens (Mukku and Stance!, 1985; Madhukar
et al., 1984) along with the consistent observation that the lowest doses in
the lifetime bioassays of 2,3,7,8-TCDD decrease tumor yield in rodent livers
but do not affect the backround levels of breast or uterine tumors (Kociba et
al., 1978). 2,3,7,8-TCDD inhibition of tumor growth at low doses and
enhancement at higher levels (in the bioassays) is supported by the recent
report of a marked decrease in tumorigenesis in the two-stage liver model at
the lowest dose of 2,3,7,8-TCDD after diethylnitrosamine (DEN) initiation
(Pitot et al., 1987). These findings are consistent with the hypothesis that
2,3,7,8-TCDD may be working through an endocrine-sensitive mechanism to yield
its toxic effects. If one accepts this premise, then it is reasonable to
assume that the actions of 2,3,7,8-TCDD can be explained using a
-------
TABLE 1. ASSOCIATION OF EXTROGENS WITH 2,3,7,8-TCDD TOXICITY
Many of the toxic effects of 2,3,7,8-TCDD are similar to effects of
estrogens in non-2,3,7,8-TCDD treated animals.
1. Some effects of 2,3,7,8-TCDD resemble effects of elevated
estrogens.
2. Other effect of 2,3,7,8-TCDD resemble antiestrogenic effects
[most antiestrogens have estrogenic effects at different doses].
3. Some 2,3,7,8-TCDD effects are not straightforward estrogenic or
antiestrogenic effects. For some of these, an influence of
estrogen on the effect is known.
4. Other signs of 2,3,7,8-TCDD toxicity may be related to cholesterol
mobilization for estrogen synthesis.
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TABLE 1. (continued)
2,3,7,8-TCDD EFFECTS
Fat loss 1,4
Wasting 1,4
Changes in serum lipids: 1,3,4
increased cholesterol
increased LDL, VLDL
Anorexia 1
Hypophagia 1
Hypoinsulinemia 1
Altered serum fatty acids 1,4
Hypoglycemia 1,4
Lowered Q£ consumption
Immunosuppression 1
Thymic involution 1
Decreased thymic
cellularity
Hirsutism 1
Chloracne 3
Skin keratinization 1
Membrane damage 1,4
Stimulates differentiation 1
in certain epithelial
cells
Lowered T4 in serum 3
Increased:
thyroid weight 3
serum TSH
T4 excretion as
glucuronide
Lower serum testosterone 3
Uterine suppression 2
Reproductive failure 1,2,3
Blockage of E2 uterotrophism 1
Terata 1,2,3
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TABLE 1. (continued)
Lowered serum corticoids 3,4
Blockage of ACTH stimulation of corticosteroid synthesis 3,4
Downregulates:
E6F receptor 1
Prolactin receptor 1
Glucocorticoid receptor 1
LDL receptor 1
Estrogen receptor 1
Ascites 1,4
Hepatocyte membrane damage 1,4
Hepatocyte membrane cAMP reduced
Enzyme inductions 3
AHH (EROD, P-448, P-450c, and d)
Ornithine decarboxylase
UDP-GTs
ALA synthetase
Others
Anemia
Porphyria cutanea tarda 1
Iron accumulation in liver
Altered iron transport in gut
8
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physiologically based model. The physiological implications of an endocrine
mechanism can explain many of the responses seen in animals after exposure to
2,3,7,8-TCDD (see Table 1) since these responses are similar to hyper- and
hypo-hormonal states (O'Malley and Buller, 1977; Potter et al., 1986; Jones et
al., 1986; Gustafsson et al., 1987). As stated above, the analogy to the
estrogen system is arguably the strongest to 2,3,7,8-TCDD effects (both
hyperplastic and dysplastic responses in endocrine sensitive organs), and the
similarity between the cytosolic receptors and their stabilization by molybdate
(Denison et al., 1986), activation of a nuclear site, and the anti-2,3,7,8-TCDD
effect of estradiol strengthens the analogy. The effects of estrogens are
widespread throughout the body. Some of these effects may not be receptor-
mediated, but the majority of the effects are directly attributable to receptor
binding. The toxic effects of estrogens have recently been summarized (Umbreit
and Gallo, 1988) and include thymic involution and decreased response to septic
challenge (Luster et al., 1984; Grossman, 1984), a wasting syndrome
characterized by weight loss, hirsutism, and epidermal lesions. As a note,
there are recent reports of estrogens enhancing or causing cachexia and wasting
which are major effects of 2,3,7,8-TCDD seen in intoxicated animals. The role
of estrogens as immunosuppressives is not well understood, but it is
hypothesized that the putative suppressant is either an excess of circulating
estradiol or perhaps an excess of trophic hormones. Estrogens also play a role
in the action of other hormones and trophic factors such as EGF (Kirk!and et
al., 1981; Gardner et al., 1978; Mukku, 1984; Mukku and Stance!, 1985; Hsueh et
al., 1981; Gonzales et al., 1984; Dickson and Lippman, 1987). These findings
lead to the conclusion that the multiple effects of TCDD could be mediated
through an endocrine mechanism. The weakness of this assumption is that
-------
2,3,7,8-TCDD causes effects that appear similar to both hyperestrogenemia and
hypoestrogenemia. It has been hypothesized that this apparent paradox is the
result of 2,3,7,8-TCDD or the ligand complex preventing the endogenous
substrate from interacting "correctly" with both the active site and a
secondary binding site (Umbreit and Gallo, 1988; Umbreit et al *, 1988).
Pleiotropism is not an uncommon finding with molecules such as hormones or
in this case 2,3,7,8-TCDD. One has only to review the early experiments on
multistage mouse skin carcinogenesis of 2,3,7,8-TCDD to see that in some cases
it inhibited tumor formation by PAH initiators (DiGiovanni et al., 1977; Berry
et al., 1978). It must be emphasized that the responses in multistage models
are dependent on time, sequence of administration, dose, and species. Hence,
inhibition under some conditions might have been predictable. This is
juxtaposed to the two-stage liver model (Pitot et al., 1980) in which it has
been shown that orally administered 2,3,7,8-TCDD enhances the tumorigenic
action of DEN. However, in subsequent experiments at lower doses of 2,3,7,8-
TCDD, a parabolic dose-response curve has been reported in the DEN/2,3,7,8-
TCDD initiation-promotion protocol (Pitot et al., 1987). This paradoxical
effect is not well understood, but it does not appear to be solely the function
of enhanced metabolism, or Ah receptor binding (Mason et al., 1987). Perhaps
it is the result of alteration of EGF receptors at low doses (Madhukar et al.,
1984) which displays a commonality with several steroid hormone receptors.
The importance of these findings to the approximation of human and animal
health risks from exposure to PCDDs and related molecules cannot be overstated.
Mathematical modeling of physiological phenomena, especially those related to
receptor function, is conducted using the Michaelis-Menten equation (1913) as
modified by Clark for the "classical" receptor model (1933). The weight of
10
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evidence for the most prevalent 2,3,7,8-TCDD effects falls into the category of
the receptor model (Poland and Knutson, 1982). The recent finding that the
hepatocarcinogenesis is related to estrogen levels or to the presence of
functional ovaries (Goldstein et al., 1987), and that DEN hepatocarcinogenesis,
in partially hepatectomized rats, is first inhibited then promoted by 2,3,7,8-
TCDD (Pitot et al., 1980, 1987) indicate that 2,3,7,8-TCDD is not causing its
myriad of effects in liver by a simple one-step event such as binding to the Ah
receptor and subsequent induction of cytochrome Pj-450. However, operationally
2,3,7,8-TCDD is a potent hepatocarcinogen in some species and strains of
rodents. ,
Risk modeling for carcinogenic xenobiotics has recently been segregated
into three classes or types of models: physiologically based pharmacokinetic
(PBPK) models in which the body is considered to be a small group of
physiological compartments (Hoel et al., 1983; Krewski et al., 1986; Bischoff,
1987); biologically motivated models of carcinogenesis (BMMC) in which the
carcinogenic process is considered to occur through a series of linked
reactions that result from two or more molecular events followed by cellular
amplification by "promoter" molecules (Moolgavkar, 1986; Thorslund et al.,
1987; Krewski et al., 1987); and the linearized multistage model (IMS) of
Armitage-Doll as modified by Crump and Howe (1984) in which it is assumed that
a sequence of mutational events occur within a single cell leading to the
neoplastic change (Armitage, 1985).
The model that appears to accommodate most of the critical components from
the biological data base on 2,3,7,8-TCDD is the BMMC model, which is generally
referred to as the Moolgavkar-Venzon-Knudson (M-V-K) model (Moolgavkar and
Venzon, 1979; Moolgavkar and Knudson, 1981). This model allows for several of
11
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the concepts of initiation-promotion-progression, along with the growth-
stimulating role of endogenous substrates such as hormones (Moolgavkar, 1986).
Incorporation of some of the factors necessary for the PBPK model can also be
done using the M-V-K model as modified, or, more correctly, expanded by
Thorslund et al. (1987). These expansions of the M-V-K model give the risk
assessor a powerful tool for looking at cancer risk mechanistically.
This option is not available with the IMS model as originally proposed.
The use of the IMS model may not be appropriate for the 2,3,7,8-TCDD data set
since this model assumes an initiating event, such as a point mutation, to
start the process. However, the IMS model can be accommodated if one
hypothesizes that the initiating event: (1) is the result of an indirect
action of 2,3,7,8-TCDD through modification of exogenous or endogenous
compounds, (2) that a population of initiated cells exists, or (3) 2,3,7,8-
TCDD leads to focal necrosis which serves as a mitogenic stimuli.
Recent reports have shown that 2,3,7,8-TCDD and other promoters in liver
enhance stimulation of DNA synthesis in situ, and stimulate repair of
0-6-methylguanine in liver DNA (Busser and Lutz, 1987; Den Engelse et al.,
1986). Lutz et al. (1984) presented a scheme for promoter potency based on
stimulation of DNA synthesis and the assumption that cell division is a
prerequisite for several stages in the carcinogenesis process. These reports
indicate that 2,3,7,8-TCDD can act as a complete-indirect-carcinogen, including
promoter activity, despite the lack of DNA binding or direct mutagenesis. The
sum of all these findings, along with the myriad of toxic responses, suggests a
model for 2,3,7,8-TCDD carcinogenesis in rodent liver as shown in Figure 1.
This model can account for the dose-response data in the bioassays and the
multistage promotion experiments, as well as allow for incorporation into
12
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existing risk models, and the scheme is not incongruent with the reports of
decreased tumor formation in some tissues. If pathway (A) can be verified by
demonstration of reactive intermediates after 2,3,7,8-TCDD treatment, then the
LMS model, slightly modified, can be used. The preponderance of evidence at
the moment supports a mechanistic model(s) which is at variance with the LMS
model. However, Figure 1 presents possibilities that are not mutually
exclusive for the existing models. The scheme also presents several testable
hypotheses which should be examined.
13
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15
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