March 31, 1987
820K87120
2,3f 7,8-TETRACHLORODIBENZO-p-DIOXIN
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. "ThireTJeTnlcal concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin
March 31, 1987
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This Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for 2,3,7,8-tetra-
chlorodibenzo-p-dioxin (U.S. EPA, 1985a). The HA and CD formats are similar
for easy reference. 'Individuals desiring further information on the toxico-
logical data base or rationale for risk characterization should consult the
CD. The CD is available for review at each EPA Regional Office of Drinking
Water counterpart (e.g., Water Supply Branch or Drinking Water Branch), or
for a fee from the National Technical Information Service, U.S. Department of
Commerce, 5285 Port Royal Rd., Springfield, VA 22161, PB #86-117983/AS. The
toll-free number is (800) 336-4700; in the Washington, D.C. area: (703) 487-4650c
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1746-01-6
Structural Formula
Synonyms
0 Dioxin; TCDBD; TCDD; 2,3,7,8-tetrachlorodibenzodioxin, 2,3,7,8-tetra-
chlorodibenzo-1,4-dioxin; 2,3,7,8-TCDD.
Uses
0 There are no commercial uses for TCDD. (U.S. EPA, 1985a).
Properties (U.S. "EPA, 1985a)
Molecular Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Vapor Pressure
Water Solubility
Log Octanol/Water Partition
Coefficient
Odor Threshold
Taste Threshold
Conversion Factor
C12H4C1402
321.9
colorless solid, needle shape
303 - 305°C
3.5 x 10-9 inn, Hg* at 30.1°C
7.9 x 10-3 ug/L**
1.4 x 106
not available
not available
* Cheng et al. (1983-1984). Converted from 4.68 x 10~7 pascals.
**Adams and Blaine (1985).
Occurrence
0 TCDD is a synthetic chemical which has no natural sources. TCDD is
not produced directly but is formed as a by-product in the manufac-
ture of a number of chlorinated phenolic compounds. It can also be
present in fly ash and flue gases of incinerators.
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0 TCDD is extremely resistant to degradation once adsorbed onto soil
with a reported half-life of 10-12 years. TCDD has a very low water
solubility and binds readily to soil. TCDD has been shown to migrate
very slowly in soil. TCDD also has been demonstrated to bioaccumulate
in fish and mammals.
8 TCDD has not been included in drinking water surveys. Given its
limited solubility, it is not expected to occur at detectable levels
in either ground or surface water. TCDD has been reported to occur
at low levels in some surface waters where it is probably bound to
suspended materials. TCDD has been found in a number of freshwater
fish at levels ranging from 1-695 ng/kg- TCDD also has been reported
to occur at low levels in rice treated with phenolic herbicides and
in the fat of animals that grazed on pasture treated with phenolic
herbicides. Due to TCDD's physical characteristics, diet is expected
to be a greater route of exposure than drinking water; however, the
available data are insufficient to evaluate the actual levels of
either route (U.S. EPA, 1984a).
III. PHARMACOKINETICS
Absorption
Gavage treatment with single or repeated doses of 2,3,7,8-TCDD in oil
has resulted in absorption of approximately 50% of the dose
(unspecified) administered to guinea pigs (Nolan et al., 1979) and
approximately 70-83% of the dose administered to rats (1 or 50 ugAg)
(Rose et al., 1976; Piper et al., 1973) or to hamsters (650 ug/kg)
(Olson et al., 1980a). Absorption of a single oral dose of 1.14 ng
3H-2,3,7,8-TCDD/kg in corn oil by a male volunteer has been estimated
to be 88.5% (Poiger and Schlatter, 1986).
Dietary administration of 0.5 or 1.4 ug 2,3,7,8-TCDD/kg/day for 42
days resulted in somewhat reduced gastrointestinal absorption by rats
(approximately 50-60% of the administered dose was absorbed) (Fries
and Marrow, 1975).
Percutaneous absorption of 2,3,7,8-TCDD (26 ng) has been estimated in
rats to be approximately 40% of the absorption of an equivalent dose
orally administered (Poiger and Schlatter, 1980).
Inhalation absorption of 2,3,7,8-TCDD has not been studied (U.S. EPA,
1985a).
Diamond Shamrock (1985) noted greater oral absorption of 2,3,7,8-
TCDD in animals given contaminated soil containing oil than without
oil.
Distribution
In the Poiger and Schlatter (1986) study, concentrations of 3.0 and
2.8 ppt of 3H-2,3,7,8-TCDD were detected in adipose tissue 10 and 69
days, respectively, after treatment.
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0 Tissue distribution following oral or intraperitoneal (i.p.) admini-
stration of 2,3,7,8-TCDD to rats appears to be preferentially to the
liver and adipose tissue (Fries and Marrow, 1975; Rose et al., 1976;
Van Miller et al., 1976; Kociba et al., 1978). Other tissues showed
substantially lower concentrations of 2,3,7,8-TCDD. Soon after treat-
ment, the liver may have concentrations about three (Kociba et al.,
1978a) to five (Rose et al., 1976) times that in adipose tissue.
It was suggested that male rats accumulate 2,3,7,8-TCDD in the liver
more efficiently than female rats (Fries and Marrow, 1975). Tissue
distribution in mice (Manara et al., 1982) and hamsters (Olson
et al., 1980a) seems to be similar to that in rats.
0 Monkeys, however, appear to accumulate 2,3,7,8-TCDD preferentially
in adipose tissue to a greater extent than in the liver (Van Miller
et al., 1976; McNulty et al., 1982). Two years after a single oral
dose to a monkey, adipose tissue contained 100 ppt and the liver 15
ppt 2,3,7,8-TCDD (McNulty et al., 1982). Prolonged tissue retention
of the compound was thus demonstrated. Tissue distribution in guinea
pigs appears to be similar to that in monkeys (Gasiewicz and Neal,
1979; Nolan et al., 1979) since tissue levels in adipose tissue
exceed those in the liver.
0 Evidence that 2,3,7,8-TCDD accumulates in the adipose tissue of
exposed humans was presented by Young et al. (1983) who reported
levels of 3 to 99 ppt in the adipose tissue.of armed forces veterans
claiming health problems related to Agent Orange.
0 Fetal distribution of 2,3,7,8-TCDD has been studied in rats (Moore
et al., 1976) and mice (Nau and Bass, 1981; Nau et al., 1982).
Levels of 2,3,7,8-TCDD were low in rat fetuses on gestation days 14
and 18 of gestation and appeared to be evenly distributed in all fetal
tissues. On day 21 of gestation, the fetal liver showed a marked
affinity for 2,3,7,8-TCDD (Moore et al., 1976). 2,3,7,8-TCDD was
distributed to the fetuses of mice following oral, i.p. or subcu-
taneous (s.c.) administration (Nau et al., 1982). Maximum fetal
concentrations occurred on days 9 and 10 of gestation; lower fetal
concentrations were observed on gestation days 11 through 18, coinci-
dent with placentation. The fetal liver had less affinity for the
compound than did the maternal liver.
0 Ryan et al. (1985) reported 2,3,7,8-TCDD levels of 5-10 ppt in
adipose tissue samples from humans taken at autopsy across Canada.
Higher levels of other dioxins were also found.
Metabolism
0 In an early metabolism study, Vinopal and Casida (1973) reported that
in vivo or in vitro studies with mice showed that polar metabolites
of 2,3,7,8-TCDD were not produced by this species. In rats, however,
hydroxylation and conjugation with glucuronide and sulfate have been
demonstrated (Poiger and Schlatter, 1979; Poiger et al., 1982;
Olson et al., 1983). Glucuronide conjugates tended to predominate
in the bile (Poiger and Schlatter, 1979) and sulfate conjugates were
located in the urine (Olson et al., 1983).
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Poiger and Schlatter (1979) stated that metabolism of 2,3,7,8-TCDD
proceeds slowly in the liver. Neal et al. (1982) demonstrated
that the rate of hepatic metabolism was enhanced by activated cyto-
chrome P-450 raono-oxygenase. It was suggested that metabolism of
2,3,7,8-TCDD proceeds by the formation of reactive epoxide intermedi-
ates (Poland and Glover, 1979). Dechlorination also was demonstrated
by Olson et al. (1983) and Sawahata et al. (1982), who identified
tri- and dichlorodibenzo-p_-dioxins as metabolites in in vitro rat
hepatocyte systems. From the bile of dogs, six major metabolites
have been identified (Poiger et al., 1982); hydroxylated conjugates
of tetra-, tri- and dichlorodibenzo-p_-dioxin predominated.
Although metabolite profiles are consistent with an arene oxide
intermediate, the covalent interaction of 2,3,7,8-TCDD with cellular
macromolecules is minimal.
Excretion
0 When the excretion data are plotted send-logarithmically, a straight
line results, suggesting that elimination of 2,3,7,8-TCDD is a first-
order phenomenon, especially in rats. Excretion in the guinea pig
may be a zero-order process (Gasiewicz and Neal, 1979). The half-life
for body elimination varied considerably with estimated ranges of 10
to 15 days in the hamster (Olson et al., 1980a), the species least
sensitive to the toxic effects of 2,3,7,8-TCDD, 11 to 24 days in the
mouse (Gasiewicz et al., 1983a,b), 17 to 31 days in the rat (Piper,
et al., 1973; Allen et al., 1975; Rose et al., 1976) and 22 to 30
days in the guinea pig (Gasiewicz and Neal, 1979; Nolan et al.,
1979). One strain of mice, DBA/2J, had a half-life for elimination
of approximately 24 days, about twice as long as in other strains
tested by Gasiewicz et al. (1983a,b). These authors also noted that
this strain of mice had a greater tendency to accumulate 2,3,7,8-TCDD
in adipose tissue than did other strains and that this phenomenon
probably resulted in slower body elimination. Half-lives for body
elimination of 2,3,7,8-TCDD have not been calculated for the monkey,
but it was suggested that the tendency of this species to accumulate
2,3,7,8-TCDD in adipose tissue may also result in slow body elimination
(Van Miller et al., 1976).
0 Recently, Olson and Bittner (1983) examined the elimination of 2,3,7,8-
TCDD in rats over a longer period than in the studies previously
summarized and determined that biphasic elimination occurred. They
estimated a half-life of approximately 7 days for the initial rapid
phase and a half-life of approximately 75 days for the slower phase,
probably related to release from stores of body fat. McNulty et al.
(1982) estimated the half-life for elimination from the fat of monkeys
to be approximately 1 year.
8 In the Poiger and Schlatter (1986) study, 11.5% of the 3H-TCDD was
excreted in feces during the first three days after treatment, and
no 3u activity was found in urine. These investigators estimated an
elimination half-life of 4.95 years for the 3n-TCDD.
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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0 The fecal route seems to be the major pathway for the elimination of
2,3,7,8-TCDD-derived radioactivity in rats (Piper et al., 1973;
Allen et al., 1975; Rose et al., 1976; Van Miller et al., 1976),
guinea pigs (Gasiewicz and Neal, 1979) and mice (Gasiewicz et al.,
1983a,b). Urinary excretion played less of a role in these species,
accounting for <1 to 28% of total excreted radioactivity while fecal
excretion accounted for 72 to >99% of the eliminated radioactivity.
Urinary excretion accounted for a more substantial proportion of body
elimination in hamsters (41% as compared with 59% by feces) (Olson,
et al., 1980a) and that strain of mice (DBA/2J) which preferentially
accumulated 2,3,7,8-TCDD in body fat (Gasiewicz et al., 1983a,b).
0 The failurb to detect metabolites of 2,3,7,8-TCDD in liver and fat
(Olson et al., 1983) indicates that elimination of the metabolites
occurs rapidly and that the rate of elimination is governed primarily
by the rate of hepatic metabolism.
IV. HEALTH EFFECTS
Humans
Either acute or chronic exposure to 2,3,7,8-TCDD (usually in combi-
nation with other substances) may result in chloracne, altered liver
function, hematological lesions, porphyria cutanea tarda, hyperpig-
mentation, hirsutism and neural degeneration in the extremities (U.S.
EPA, 1985a). Stevens (1981) has estimated that the minimum cumulative
toxic dose of 2,3,7,8-TCDD in humans is 0.1 ug/kg.
Rowe (1968) has described experiments showing a dose-response for
-chloracne in humans acutely exposed to topical applications of
2,3,7,8-TCDD.
The toxic effects of chloracne from exposure to 2,3,7,8-TCDD may
persist for many years, though other effects noted in various
individuals are apparently reversible after a short period. Epidemio-
logical studies have failed to demonstrate a convincing connection
between 2,3,7,8-TCDD exposure and spontaneous abortions or malfor-
mations in humans. Some evidence of cytogenetic damage has been
reported in humans exposed to chemicals contaminated with 2,3,7,8-TCDD,
but negative results have also been reported; exposures were not
quantitated and the other chemicals cannot be ruled out as causative
agents (U.S. EPA, 1985a).
Swedish case-control studies provide limited evidence for the carcino-
genicity of phenoxy acids or chlorophenols or both in humans. However,
with respect to the dioxin impurities contained within them, the evidence
for the human carcinogenicity for 2,3,7,8-TCDD based on epidemiologic
studies is only suggestive because of the difficulty of evaluating
the risk of 2,3,7,8-TCDD exposure in the presence of the confounding
effects of phenoxy acids and/or chlorophenol (U.S. EPA, 1985a).
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2, 3, 7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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Animals
Short-term Exposure
0 There are wide variations in species sensitivity to the acute toxicity
of 2, 3, 7, 8-TCDD. LDsgs range from 0.6 ug/kg for the male guinea pig
to >5,000 ugAg bw for the male hamster (Schwetz et al., 1973; Olson,
et al., 1980b; Henck et al., 1981). The toxic manifestations seem
to be the same whether the compound is given as a single oral dose or
as a limited number of multiple treatments, with death occurring from
5 to 45 days post-treatment. Lethal exposures result in weight loss,
often described as "wasting away" and thymic atrophy. In some species,
particularly rats and mice, extensive liver damage is observed (Gupta
et al., 1973). In general, no specific cause of death has been
identified, although extensive hemorrhaging has been implicated in
mice (Vos et al., 1974).
0 In rats, single high doses (200 ug/kg) produce liver necrosis (Jones
and Butler, 1974), while lower doses (5 and 25 ug/kg) result in fatty
changes in the liver and proliferation of the endoplasmic reticulum
(Fowler et al., 1973). Other effects seen in some species include
induction of microsomal enzymes, degeneration of plasma membranes
with loss of ATPase activity, a decreased ability to excrete some
xenobiotics in the bile, porphyria, altered gastrointestinal absorption
of some nutrients and decreased blood cellularity (U.S. EPA, 1985a).
Turner and Collins' (1983) found treatment-related liver lesions in
guinea pigs given single gavage doses of 2, 3, 7, 8-TCDD at 0.1 ugAg
and higher.
8 2, 3, 7, 8-TCDD is an immunotoxin in laboratory animals, predominantly
affecting cell-mediated immunity. Hypers ens itivity, adverse effeofes-—
on the thymus and increased sensitivity to antigens have demonstrated
the immunotoxic potential of 2, 3, 7, 8-TCDD. Weanling rodents show
greater susceptibility to immune effects compared to adults (U.S.
EPA, 1985a).
Long-term Exposure
0 In rats and mice, the liver appears to be the most sensitive organ
following chronic or subchronic exposure. Hepatotoxicity develops
following a long induction period and the changes persist for long
periods following the termination of exposure (King and Roesler,
1974; Goldstein et al., 1982).
0 Liver lesions as well as other toxic signs were observed in the
following studies in rats and mice. In the subchronic studies, the
NOAEL of 0.01 ugAg/day (Kociba et al., 1976) and 0.5 ugAg/week
(NTP, 1980) have been reported for rats. A NOAEL of 2 ugAg/week was
identified for female mice and a LOAEL of 1 ugAg/week for male mice
in the NTP (1980) subchronic study. A NOAEL of 0.001 ugAg bw/day, a
LOAEL of 0.01 ugAg/day, and an effect level of 0.1 ugAg/day have
been reported for rats following chronic dietary exposure (Kociba
et al., 1978a,b, 1979; NTP, 1980). Toth et al. (1978, 1979) observed
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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toxic effects in mice at doses as low as 0.007 ug/kg/week given for
one year by gavage. Gavage dosing for two years led to toxic hepatitis
at a NOAEL of 0.05 ug/kg/week and a LOAEL of 0.5 ugAg/week in rats,
a LOAEL of 0.5 ug/kg/week in male mice, and a LOAEL of 2.0 ugAg/week
in female mice (NTP, 1980).
0 DeCaprio et al. (1986) fed 2,3,7,8-TCDD in the diet for 90 days to
male and female Hartley guinea pigs, and found SCAELs of 0.12 and
0.61 (male) and 0.68 (female) mg/kg/day; decreased body weight gain,
increased relative liver weights, decreased relative thymus weights,
and hepatocellular cytoplasmic inclusion bodies at 4.90 (males) and
4.86 (females) mg/kg/day; and, mortality and other mentioned effects
at 26 (males) and 31 (females)
Reproductive Effects
0 Adverse effects of 2,3,7,8-TCDD on reproduction in rats exposed
through the diet were observed by Murray et al. (1979) and are
detailed under Lifetime Health Advisory.
Developmental Effects
0 2,3,7,8-TCDD has been demonstrated to be teratogenic in mice. The most
common malformations observed are cleft palate and kidney anomalies;
however, other malformations have'been observed occasionally. With an
effect level of 1 ugAg/day, 2,3,7,8-TCDD is the most potent teratogen
known. At higher doses, 2,3,7,8-TCDD has a marked fetotoxic effect,
as measured by decreased fetal weight and increased fetal toxicity.
Hemorrhagic GI tract has been associated with 2,3,7,8-TCDD fetal
toxicity .(U.S. EPA, 1985a).
0 Poland and Glover (1980) produced evidence that responsiveness of
mice to cleft palate from 2,3,7,8-TCDD treatment is related to the
presence of Ah receptor.
0 In rats, it has also been consistently observed that 2,3,7,8-TCDD
produces fetotoxic responses. In this species, the most common fetal
anomalies observed were edema, hemorrhage and malformation of the
kidney with effects observed at doses of *0.01 ugAg/day. In
addition, there is some evidence that 2,3,7,8-TCDD can induce micro-
somal enzymes in the fetus exposed in utero, and this induction is
accompanied by damage to the fine structure of the liver cell; however,
other reports indicate that enzyme induction occurs only after birth
following exposure to 2,3,7,8-TCDD through the mother's milk. As in
mice, hemorrhagic GI tracts have been observed in rat fetuses exposed
in utero to 2,3,7,8-TCDD (U.S. EPA, 1985a).
0 Rabbits and monkeys are also susceptible to the fetotoxic effects of
2,3,7,8-TCDD; however, the studies of these species have been too
limited to clearly demonstrate a teratogenic response or define a
threshold dose for fetotoxicity (U.S. EPA, 1985a).
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Mutagenicity
° In vivo and in vitro mutagenicity tests have produced inconclusive
evidence as to the mutagenicity of TCDD (U.S. EPA, 1985a).
0 Early reports indicated that 2,3,7,8-TCDD was mutagenic in S_. typhi-
murium strain TA1532 (Hussain et al., 1972; Seiler, 1973); however,
later attempts to confirm these results have been unsuccessful (Nebert,
et al., 1976; McCann, 1978; Gilbert et al., 1980; Geiger and Neal,
1981). 2,3,7,8-TCDD has been reported to be mutagenic to 12. coli in
vitro (Hussain et al., 1972) and to S_. cerevisiae in vitro, and in a
host-mediated assay (Bronzetti et al., 1983). Covalent interactions
with nucleic acids are minimal if they occur at all (Kondorosi et al.,
1973; Poland and Glover, 1979). Only marginal effects have been
observed on the incidence of chromosomal aberrations in vivo (Green
and Moreland, 1975). A test for unscheduled DNA synthesis in cultured
male rat hepatocytes was negative (Althaus et al., 1982). Loprieno
et al. (1982) reported 2,3,7,8-TCDD as clastogenic in mice in vivo,
negative for cytogenetic effects in vivo, and negative for unscheduled
DNA synthesis in a human cell live in vitro. Hay (1983) reported
2,3,7,8-TCDD as mutagenic in the baby hamster kidney cell transfor-
mation assay.
Carcinogenici ty
0 Several bioassays have demonstrated this compound to be a potent
carcinogen in rats and mice (Kbciba et al., 1978a; Toth et al., 1979;
NTP, 1980). Adenomas or carcinomas of the thyroid, hepatocellular
carcinomas, carcinomas of the tongue and hard palate, and adenomas of
the adrenal gland have been induced in rats and mice.
0 Significant (P <0.05) neoplastic effects were evident at dietary
levels of 0.01 and 0.1 ug/kg/day but not at 0.001 ug/kg/day in the
two-year study with Sprague-Dawley rats by Kociba et al. (1978). In
Osborne-Mendel rats given 2,3,7,8-TCDD in corn oil:acetone twice
weekly for total weekly doses of 0.01, 0.05 and 0.5 ugAg/week for
two years, significant (P <0.05) tumor increases were thyroid in mid-
and high-dose males and liver in high-dose males (NTP, 1980). In the
NTP O980) study in which B6C3F1 mice were dosed like the rats except
that females received 0.04, 0.2 and 2.0 ug/kg/week, significant
(P <0.05) tumor increases were in liver in high-dose males and females
and thyroid in high-dose males.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = mg/L ( u /L)
(UF) x ( L/day)
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mgAg bw/day.
BW * assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty ;actor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
_____ L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
Turner and Collins (1983) administered single oral doses of 2,3,7,8-TCDD
at 0.1, 0.5, 2.5, 12.5 or 20 ugAg in aqueous methyl cellulose to groups of
4 to 7 female guinea pigs. Survivors were killed 42 days after dosing and
examined for histopathologic changes in the liver. Four of the 7 animals in
the highest dose group and 1 of 5 in the 12.5 ug/kg group died before the end
of the observation period. Mild histopathologic changes, including steatosis
(fatty change), focal necrosis and cytoplasmic degeneration were noted in
animals from all treated groups, but not in controls. The authors indicated
that quantitative differences among the dosage groups were not detectable by
light microscopy.
A LOAEL of 0.1 ug/kg can be derived from the study of Turner and Collins
(1983) for calculating a One-day HA, using an uncertainty factor (UF) of 1,000
for an animal LOAEL. This UF consists of two 10-fold factors to account for
both intra- and interspecies variability to the toxicity of this chemical in
the absence of chemical-specific data, and an additional 10-fold factor
because the HA is based on a LOAEL and not a NOAEL.
For a 10-kg child consuming 1 L of drinking water per day, the One-day
HA is calculated as follows:
One-day HA - (0.1 ugA
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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Ten-day Health Advisory
Because of the demonstrated sensitivity of the guinea pig to acute
toxicity of 2,3,7,8-TCDD, the Ten-day HA is derived by dividing the One-day
HA by ten. The Ten-day HA is, therefore, 0.0001 ug/L.
Longer-term Health Advisory
The three-generation reproduction study in rats by Murray et al. (1979)
has been selected because the animals in this study were administered 2,3,7,8-
TCDD by diet on a daily basis for an appropriate duration as opposed to the
gavage method of treatment used in other studies considered and because the
adverse effect was on reproduction. Comparison with the other studies in
which different treatment protocols were used suggests that the dose of
0.001 ugAg/day, concluded by the U.S. EPA as a LOAEL for adverse effects
on the pups and dams in the Murray et al. (1979) study would be protective
against the toxic effects found in the other studies. Although DeCaprio
et al. (1986) found NOAELs of 0.61 and 0.68 ng/kg/day in their 90-day guinea
pig study, this dose is slightly below the LOAEL of 0.001 ugAg/day
(1 ng/kg/day) in another species which, in turn, is below the LOAEL of
4.86 ng/kg/day in the DeCaprio et al. (1986) study.
Using an uncertainty factor of 1,000 for an animal LOAEL (i.e., 10-fold
for intra- and 10-fold for interspecies variability to the toxicity of a
chemical in the absence of specific data, and an additional 10-fold factor
because the estimate is based on a LOAEL rather than a NOAEL), a Longer-term
HA can be calculated from the LOAEL of 0.001 ugAg/day concluded for the
Murray et al. (1979) study.
For a 10-kg child consuming 1 L of drinking water each day, the Longer-
term HA is calculated as follows:
Longer-term HA » (0.001 ugAg/day) (10 kg) = 0.00001 ug/L
(1,000) (1 L/day)
where:
0.001 ugAg/day - LOAEL from study by Murray et al. (1979).
* 10 kg • assumed weight of child
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
By substituting 70-kg body weight and daily consumption of 2L of water
for the adult in the above equation, the Longer-term HA for the 70-kg adult
becomes 0.000035 ug/L.
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Lifetime Health Advisory * *
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic .potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The EPA has developed for comparison with cancer-based criteria, a pre-
sumed safe daily intake level based on noncarcinogenic effects as indicated
in U.S. EPA (1984b). For consistency, the rationale used by EPA for the
calculation of this value in U.S. EPA (1984b) is used here for the DWEL
calculation. The rationale as presented in U.S. EPA (1984b) is as follows:
2,3,7,8-TCDD displays an unusually high degree of reproductive
toxicity. It is teratogenic, fetotoxic and reduces fertility. In a
3-generation reproductive study, Murray et al. (1979) reported a
reduction in fertility after daily dosing at 0.1 or 0.01 ug 2,3,7,8-
TCDD/kg in the FI and F2 generations of Sprague-Dawley rats. Although
Murray et al. (1979) considered the lowest dose tested, 0.001 ug/kg»
to be a no-observed-effect level (NOEL), a re-evaluation of these data
by Nisbet and Paxton (1982), using different statistical methods,
indicated that there was a reduction in the gestation index, decreased
fetal weight, increased liver to body weight ratio, and increased
incidence of dilated renal pelvis at the 0.001 ugAg dose. The
reevaluated data would suggest that equivocal adverse effects were
seen at the lowest dose (0.001 ug/kg/day) and that this dose should,
therefore, represent a lowest-observed-adverse-effect level (LOAEL).
Schantz et al. (1979) found reductions in fertility and various other
toxic effects in rhesus monkeys fed a 50 ppt 2,3,7,8-TCDD diet for
20 months. This corresponds to a calculated daily dose of 0.0015 ug
2,3,7,8-TCDD/kg/day. These results suggest that monkeys may be
somewhat more sensitive than rats, since the effects in monkeys were
more severe and not equivocal. Since the data from the limited study by
Schantz et al. (1979) are supportive of the findings by Murray et al.
(1979) it seems reasonable to determine an ADI based on the LOAEL.
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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From these results, a LOAEL of 0.001 ug/kg was identified. Using this
LOAEL, the DWEL is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.001 ug/kg/day) = , x 10-6 ug/kg/day
(1,000)
where:
0.001 ug/kg/day = LOAEL.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (1 * 10-6 ug/kg/day) (70 kg) = 0.000035 ug/L
(2 L/ day)
where:
1 x 10-6 ug/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/ day = assumed daily water consumption of an adult.
2,3,7,8-TCDD is placed in Group B: Probable human carcinogen. The
estimated excess cancer risk associated with lifetime exposure to drinking
water co'ntaining 2,3,7,8-TCDD at 3.5 x 1 0~5 ug/L is approximately 2 x 10~4.
This estimate represents the upper 95% confidence limit from extrapolations
prepared by EPA's Carcinogen Assessment Group using the linearized, multistage
model. The actual risk is unlikely to exceed this value, but there is
considerable uncertainty as to the accuracy of risks calculated by this
methodology.
Evaluation of Carcinogenic Potential
0 Cancer potency estimates were derived using the multistage model and
the tumor data on female rats in the chronic feeding study by Kociba
et al. (1978a) (U.S. EPA, 1985a,b).
0 The 95% upper-limit carcinogenic potency factor for humans, q-)*, is
1.56 x 105 (mg/kg/day)-1. For a 70 kg human drinking 2 L water/day,
the water concentration should be 2.2 x 10~6 ug/L in order to keep
the upper-limit individual lifetime cancer risk at 10-5. Water
concentrations corresponding to excess cancer risk of 10-4 and 10-6
are, therefore, 2.2 x 10-5 and 2.2 x 10~7 ug/L, respectively.
0 Maximum likelihood estimates as well as 95% upper limits of cancer
risks by the multistage model have been calculated (U.S. EPA, 1985b).
For example, at 1 x 10~3 ng/kg/day or 0.035 ng/L cancer risk estimates
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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are 1.1 x 1CT4 (MLE) and 1.5 x 1CT4 (UL) and at 1 x 10~2 ngAg/day
cancer risk estimates are 1.1 x 10~3 (MLE) and 1.5 x 10~3 (UL).
0 The EPA's Carcinogen Assessment Group has estimated cancer risks with
other models besides the multistage (U.S. EPA, 1985b). As an example,
1 x 10~3 ngAg/day lifetime exposure was associated with additional
risks (95% upper confidence limit) of 1.5 x 10"4 by the multistage
and one-hit, 2.9 x 1CT3 by the Weibull, and 7.5 x 1CT8 by the log-
probit, using the Kociba analysis of the data. While recognized as
statistically alternative approaches, the range of risks described by
using any of these modeling approaches has little biological signifi-
cance unless data can be used to support the selection of one model
over another. In the interest of consistency of approach and in
providing an upper bound on the potential cancer risk, the EPA has
recommended use of the linearized multistage approach.
8 The IARC (1981) classified TCDD as a 2B chemical (sufficient animal
evidence; inadequate human evidence) for carcinogenicity.
8 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), 2,3,7,8-TCDD maybe classified in
Group B2: Probable human carcinogen. This category is for agents for
which there is inadequate evidence from human studies and sufficient
evidence from animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 For 2,3,7,8-TCDD, the U.S. EPA has established criteria of 1.3 x 1CT7,
1.3 x 10~° or 1.3 x 10~9 ug/L in ambient waters, based on an assumed
daily consumption of 6.5 g of contaminated fish and shellfish and 2 L
of drinking water (U.S. EPA, 1984b). Under these conditions, 94.2%
of the total exposure would result from the consumption of aquatic
organisms. The recommended levels correspond to estimated human
lifetime excess cancer risks of 10~5, 10~° or 1CT7, respectively.
These values are considerably lower than the HAs for drinking water,
reflecting the high bicaccumulation potential of this compound in
aquatic species.
8 The FDA advises that fish containing >50 ppt of 2,3,7,8-TCDD should
not be consumed and those containing >25 ppt, but <50 ppt, should not
be consumed more than twice a month (FDA, 1983). This is reflected
in a Canadian limit of 20 ppt in the Lake Ontario commercial fish
imported into the United States (NKCC, 1981).
0 An ADI of 1CT4 ug/kg bw/day has been proposed previously for 2,3,7,8-
TCDD by the National Academy of Sciences Safe Drinking Water Committee
(NAS, 1977). This ADI was based on a 13-week rat feeding study by
Kociba et al. (1976) and was proposed before convincing evidence for
the carcinogenicity of 2,3,7,8-TCDD had accumulated.
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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VII. ANALYTICAL METHODS
0 Determination of dioxin is by a gas chromatographic/mass spectrometer
(GC-MS) method (Method 613. U.S. EPA, 1984c). In this method, a one
liter sample is spiked with an internal standard of a labeled dioxin
and extracted with methylene chloride using a separatory funnel. The
methylene chloride extract is exchanged to hexane during concentration
to a volume pf approximately 1 mL. The extract is then analyzed by
capillary column GC/MS to separate and measure dioxin. The method
detection limit is dependent upon the nature of interferences, but it
is estimated to be about 0.02 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Because of its high toxicity and low potential for occurrence in
drinking water, very little information is available on the removal
of dioxins from drinking water. Granular activated carbon adsorption
is likely to be the most reasonable treatment approach and the small
amount of empirical evidence available bears this out.
0 While looking for a method to concentrate polychlorinated dibenzo-p-
dioxins and dibenzofurans, scientists from the U.S. Fish and Wildlife
Service's fish-pesticide research laboratory in Columbia, Missouri,
found that TCDD is extremely difficult to recover from GAC once it
has been adsorbed (Chemical Engineering and News, 1977). Subsequent
. pilot-scale tests of carbon adsorption of Agent Orange [50-50 mixture
of the acid esters of 2,4,5-T and 2,4-dichlorophenoxyacetic acid
(2,4-D)] reduced an initial concentration of 10 mg/L dioxin in the
herbicide to a final concentration of less than 0.1 mg/L. Details of
the adsorption test were not reported by the authors. Based on these
data and the reported low water solubility of 0.2 ug/L dioxin in
water (Bollen and Norris, 1979), it appears that GAC adsorption of
dioxin from water is potentially feasible.
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2,3,7,8-Tetrachlorodibenzo-p-Dioxin March 31, 1987
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