PB87-200176
Health Advisories for 16 Pesticides (Including
Alachlor, Aldicarb, Carbofuran, Chlordane, DBCP
1,2-dichloropropane, 2,4-D, Endrin, Ethylene
Dibromide, Heptachlor/Heptachlor epoxide Lindane
Methoxychlor, Oxamyl, Pentachlorophenol
Toxaphene and 2,4,5-TP
(U.S.) Environmental Protection Agency
Washington, DC
Mar 87
U.S. DEPARTMENT OF COMMERCE
National Technical Information Service
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TECHNICAL REPORT DATA
(Hestt ftfd Imtfuctioat on Iht rrrme be
REPORT NO.
1. RECI»
4. TITLE AND SUBTITLE
Health Advisories for 16 Pesticides
». REPORT DATE
March, 1987
t. PERFORMING ORGANIZATION CODE
. AUTHOR(S)
U.S. Environmental Protection Agency
Office of Drinking Water
I. PERFORMING ORGANIZATION REPORT NO.
PERFORMING ORGANIZATION NAME AND ADDRESS
10. PROGRAM ELEMENT NO.
U.S. Environmental Protection Agency
Office of Drinking Water (WH-550D)
401 M St., S.W.
Washington, D.C. 20460
It. CONTRACT/GRANT NO.
12. SPONSORING AGENCY NAME AND ADDRESS
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCV CODE
15. SUPPLEMENTARY NOTES
16. ABSTRACT
These documents summarize the health effects of 16 pesticides including: alachlor,
aldicarb, carbofuran, chlordane, DBCP, 1,2-dichloropropane, 2,4-D, endrin, ethylene
dibromide, heptachlor/heptachlor epoxide, lindane, methoxychlor, oxamyl, pentachloro-
phenol, toxaphene and 2,4,5-TP. Topics discussed include: General Information and
Properties, Pharcacokinetics, Health Effects in Humans and Animals, Quantification
of Toxicological Effects, Other Criteria Guidance and Standards, Analytical Methods
Treatment Technologies.
7.
KEY WORDS AND DOCUMEN ' ANALYSIS
DESCRIPTORS
b.lOENTIFIER ,/OPEN ENDED TERMS I. COSATI Field/Group
Pesticides
Drinking Water
Health Advisory
Toxicity
U. DISTRIBUTION STATEMENT
Open Distribution
21. NO. OF PAGES
EPA Fw» 2220.1 (*•«. 4.77) P««vi.
REPRODUCED BY
U.S. DEPARTMENTOF COMMERCE
NATIONAL TECHNICAL
INFORMATION SERVICE
SPRNGFELD.VA 22161
22. PRICE
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March 31, 1987
ALACHLOR
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. The chemical 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 bioItgical mechanisms invol 'ed 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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Alachlor March 31, 1987
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This Health Advisory (HA) is based on information presented in the Office
of Pesticide Program's Special Review Position Document 1 (PO 1) and PD 2/3
for Alachlor (U.S. EPA, 1984). Individuals desiring further information on
the toxicological data base or rationale for risk characterization should
consult the PO 1. The PD 1 is available for review at each EPA Regional
Office of Pesticide Programs counterpart or for a fee from the National
Technical Information Service, U.S. Department of Commerce, 5285 Port Royal
Rd., Springfield, VA 22161, PB ft 86118221/AS. The toll-free number is
(800) 336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 15972-60-8
Structural Formula
Alachlor
Synonyms
0 2-Chloro-2',6l-diethyl-N-(methoxymethyl)acetanilide; 2-chloro-
N(2,6-diethylphenyi)-N-(methoxymethyl)acetamide? Lasso®.
Uses
The major use (99%) of alachlor is as a herbicide in pre-emergence
to field corn, soybeans and peanuts.
Properties (Windholz, 1983)
Chemical Formula C^4H20NO2C1
Molecular Weight 269.77
Physical State White crystalline solid at 23°C
Boiling Point
Melting Point 40-41°C
Density
Vapor Pressure
Specific Gravity 1 .133(25/15.6°C)
Water Solubility 240 mg/L
Log Octanol Water/Partition 434
Coefficient
Odor Threshold —
Taste Threshold —
Conversion Factor
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Alachlor March 31, 1987
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Occurrence (U.S. EPA, 1983a)
0 Alachlor had one of the largest production volumes of any pesticide,
130 to 150 million Ibs produced in 1983. Alachlor is applied to the
soil either before or just after the crop has emerged.
0 Alachlor is degraded in the environment by a number of mechanisms.
Alachlor is metabolized rapidly by crops after application. Once in
the soil alachlor is degraded by bacteria both under aerobic and
anerobic conditions. Alachlor is not photodegradeable and does not
hydrolyze under environmental conditions. Alachlor has moderate
mobility in sandy and silty soils and has been demonstrated to migrate
to ground water. Alachlor does not bioaccumulate.
0 Alachlor has been reported to occur in both ground and surface waters.
Limited data have been reported in both Federal and States surveys of
surface water where alachlor was reported to occur at levels of 1
ppb. Based upon the available data, alachlor is believed to have
the potential to contaminate ground and surface water widely.
0 Food does not appear to be a major route of exposure. Residues of
alachlor in food are usually non-detectable. Current EPA standards
for alachlor food residues are limited to levels which when combined,
would result in a maximum daily doses of 0.6 ug/kg. In areas where
alachlor drinking water level in exceed 0.3 ug/L, daily water intake
will exceed this permitted dose, and would be the major source of
alachlor exposure.
III. PHARMACOKINETICS
Absorption
0 Nearly 100% of a single oral dose was absorbed by the gut of male and
female rats (four/sex) (Monsanto, 1983).
0 A dermal absorption study in two Rhesus Monkeys indicated that approxi-
mately 50% of the dermally applied alachlor was absorbed within 24
hours (Monsanto, 1981 a).
Distribution
0 Radioactivity from the administered dose was found in the blood and
in the spleen, liver, kidney and heart, which may be a reflection of
the amount of blood in those organs. In addition, a relatively high
level of radioactivity also was found in the eyes, brain, stomach and
ovaries. These data assume added significance in light of the treat-
ment-related lesions observed in the two year rodent feeding studies
(Monsanto, 1983).
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Alachlor March 31, 1987
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Metabolism
0 A gavage metabolism study in the rat (single dose of 14.0 mg/kg)
indicated that alachlor is metabolized rapidly and eliminated as
conjugates of mercapturic acid, glucuronic acid and sulfate in urine
and feces (37.6 to 45% in urine and 37.0 to 49.3% in feces of females).
Elimination of C(>2 was minimal (Monsanto, 1983).
0 Since metabolism in monkeys may be different than that of man,
the identification of metabolites in urine of monkeys indicted that
only metabolites which contained the diethylaniline (DEA) moiety were
present, while in the human biomonitoring studies, metabolites which
contained the hydroxyethyl ethylaniline (HEEA) moiety were also
present in urine at a level that required attention (i.e., DEArHEEA
was generally 4:1, but in one individual it was 1:2). Hence, all
available data from other animal species (e.g., rat) should be
considered for extrapolation to man (U.S. EPA, 1986a).
Excretion
0 Approximately 89% of a single oral dose, 14 mg/kg, in the rat study
(Monsanto, 1983) was eliminated via the urine during the first 10-day
of the study, with most of the elimination occurring during the first
48 hours (half-life of .2 - 10.6 hours), followed by a slower phase
(half-life of 5 to 16 days). Elimination of CO2 was minimal.
IV. HEALTH EFFECTS
Humans
The Agency is unaware of any human studies that have investigated the
oncogenicity of alachlor. There is one limited epidemiology study
that investigated the ocular status of workers in a plant where alachlor
was manufactured, but found no effects (Coleman and Gaffey, 1980).
Animals
Short-term Exposure
Alajhlor exhibits relatively low acute toxicity by the oral (rat LD50
= 0.93 g/kg), dermal (rabbit 1*050 = 13.3 g/kg) or inhalation (rabbit
LC50 >5.1 ml/1) routes of exposure (Monsanto, 1978a, 1981b). The
technical product has only slight skin and eye irritation potential
after an acute exposure (Monsanto, 1978b, 1984a).
Long-term Exposure
The principal chronic toxic effects other than cancer are hepato-
toxicity and ocular lesions as reported below in the chronic feeding
studies.
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Alachlor March 31, 1987
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0 In a six-month dog feeding study, alachlor was tested at 0, 5.0, 25.0,
50.0 or 75.0 ing/kg/day and showed dose related hepatotoxicity at all
doses (Ahmed et al., 1981). Significantly increased absolute and
relative liver weights were observed at all dose levels for males and
at dose levels of 25 mg/kg/day and above for females. Liver fatty
degeneration and biliary hyperplasia occurred in both sexes at dose
levels of 25 mg/kg/day and greater.
0 In a subsequent one-year dog feeding s-tudy, the NOEL was determined to
to be 1 mg/kg/day based upon hemosiderosis seen in the liver, kidney
and spleen of dogs in the 3 and 10 mg/kg/day groups (Naylor et al.,
1984).
0 A two-year rat feeding study in the Long-Evans strain showed alachlor
to be toxic at all doses tested (0, 14.0, 42.0 or 126.0 mg/kg/day)
(Daly et al., 1981b). The ocular lesion, classified as uveal degenera-
tion syndrome (UDS), is characterized in its mildest form by
free floating iridial and choroidal pigments in the ocular chamber
and pigment deposition on the cornea' and lens. In its most severe
form, the syndrome is characterized by bilateral degeneration of the
iris and diminution in the size of the ocular globe with secondary
total cataract formation. UDS, once established, is an irreversible
condition (Stout et al., 1983b).
0 Two follow-up two-year feeding studies in the same strain of rat were
conducted at 0, 0.5, 2.5 or 15 mg/kg/dsy (Stout et al., 1983a) and at
126 mg/kg/day (Stout et al., 1983b), respectively. At the highest
dose in the first study (Stout et al., 1983a), there was a small
increase in the number of animals exhibiting the initial stage of
UDS, specifically molting of retinal pigmentation. The 2.5 mg/kg/day
dose was considered to be the NOEL for UDS. In the second study
(Stout et al., 1983b), animals exposed to 126 mg/kg/day for different
lengths of time demonstrated that the UDS is an irreversible syndrome.
Reproductive Effects
0 In a three generation reproduction study in rats, alachlor was tested
at 3.0, 10.0 or 30 mg/kg/day and showed a NOEL for renal toxicity
observed in F2 adult males and F3 pups at 10.0 mg/kc.,/day (Schroeder
et al., 1981). The renal toxicity consisted of kidney discoloration,
chronic nephritis and increased absolute kidney weights.
Developmental Effects
0 In a teratology study in the rat (Rodwell and Tracher, 1980), alachlor
was administered by gavage at dose levels of 50, 150 or 400 mg/kg/day.
A maternal and fetotoxic NOEL was established at 150 mg/kg/day in
this study with no teratogenic potential indicated.
0 There are two rabbit teratology studies performed by International
Research and Development Corporation, (IRDC, 1984) that used identical
dose levels of 0, 10, 30 or 60 mg/kg/day. The main difference in
these studies was use of different vehicles in which to suspend the
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Alachlor March 31, 1987
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technical alachlor. The first study used corn oil and the second
used mineral oil. The first study was found to be invalid because of
invalid control data due to the combination of the following factors:
a. The control group lost an average weight of 59 g during the dosing
period (day 6-28 of gestation) and two animals in this group died
due to gavage errors.
b. The incidence of heart anomalies in this control group is high
(8/66 fetuses and 2/10 litters) as compared to the historical
control (2/741 fetuses and 2/118 litters). Also the incidence of
scoliosis in this study is significantly higher than the historical
control.
c. The average weight of the fetuses in this control group is
smaller (27.7 g) than the treatment group (35.7 g in low dose,
28.5 g in mid dose and 29.5 g in high dose) and the historical
control (33.22 g).
d. Congested lungs with red foci at necropsy (indicating the possibility
of gavage error) in more than the two that were reported dead due
to gavage error-.
Prior to insemination, females were randomly assigned to control, and
treatment groups consisting of 16 animals. The test material was
administered by gavage from days 6 through 27 of gestation to pregnant
females. The control group received only 1 ml/kg/day on a comparable
routine.
If one uses the later supplied historical data as an indication of
what the control data should have been, there is a dose-related
maternal loss at 30 and 60 mg/kg. Using the same considerations,
there was an increase in potential teratogenic skeletal (scoliosis)
malformations (historical controls, fetuses 0.54%, litter 3.39%; low-
dose fetuses 1.4%, litter 7.1%; mid-dose, fetuses 4.3%, litter 22.2%;
high-dose, fetuses 3.2%, litter 20.0%). The effects noted in the
high dose may be hindered by the high level of maternal mortality.
There is an increase in 27 presacral vertebrae at all levels tested
as well.
Tne second study used mineral oil as the vehicle. There were eighteen
Dutch Belted rabbits per dose group who were artificially inseminated.
Artificial insemination is not the method of choice for teratology
studies. In spite of the use of mineral oil, there was little evidence
of the laxative-cathartic effects. The mid-dose group exhibited evidence
of increased early resorptions, postimplantation loss and decrease in
total implantations per dam when compared to low of control groups.
The high dose group also had a high preimplantation loss (49%).
There was an increased incidence of the following malformations in
the fetuses of preresacral vertebrae in the high dose, 13th rudiment-
ary rib in all doses, and an increase in major vessel variations in
the high dose group. When one combines the effects seen on the rudi-
mentary and full 13th ribs a dose response increase was seen. There-
fore, the LOAEL for this study is 10 mgA9/day (Monsanto, 1984b).
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Alachlor March 31, 1987
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Mutagenicity
0 A rec-assay conducted at 6 concentrations 120-20,000 ug/plate) in IJ.
subtilis strains M45 and HI 7 showed no evidence of test compound
induced inhibition (Shirasu et al., 1980).
0 A reverse mutation assay conducted at 6 concentrations (10 - 15,000
ug/plate} in jE. coli strain WP2 her and J>. typhimurium strains TA
1538, TA1537, TA1535, TA98 and TA100, with and without S9 metabolic
activation was also negative (Shirasu et al., 1980).
Carcinogenicity
0 Alachlor feeding studies have demonstrated oncogenic effects including
lung tumors in mice and stomach, thyroid, and nasal turbinate tumors
in rats.
0 Female mice of the CD-I strain fed technical grade alachlor in the
diet for 18 months at dosages of 0, 26, 78 or 260 ing/kg/day developed
statistically significant increases (p <0.05) in lung bronchiolar
tumors at the highest dose tested (Daly et al.,198la). The increase
of lung tumors in male mice was not significant at any dose.
0 Three chronic feeding studies were conducted in the Long-Evans strain
of rat with alachlor. In the first study, technical material was
stablized with epichlorohydrin during the first year of the study
and fed to 50 animals/sex at dose levels of 0, 14, 42 or 126 mg/kg/day
(Daly et al., 1981b). During the second year of this study, alachlor
stablized with another intentionally added "inert" was the test
material. Dose-related responses were observed for tumors of the
nasal turbinates of both sexes for the mid and high doses. Also,
statistically significant increases were observed in the incidence
of malignant stomach tumors (described by the authors as neoplasms
pluripotent in ability to form mixed carcinoma-type tumors) in the
high dose group in both sexes (p <0.001). In addition, thyroid
follicular tumors (adenomas plus carcinomas) increased in both sexes
at the high-dosage level with the increase being significant in males
(p <0.001 ).
0 In the second two-year feeding su.udy, throughout which an "inert"
different from epichlorohydrin was used as a stabilizer in the test
ma'terial, three treatment groups of 50 male and 50 female Long-Evans
rats received 0.5, 2.5 and 15 mg/kg/day (Stout et al., 1983a).
0 The nasal epithelial adenoma response was statistically significant
in both sexes of both chronic rat studies (p <0.001) (Daly et al.,
1981b; Stout et al. , 1983a). An increase was noted in the incidence
of thyroid-follicular cell adenomas in males and in a rare stomach
tumor in both sexes in the second study. Brain tumors were observed
in both studies which, although not statistically significant, were
concluded by the registrant as "possibly due to, or secondary to,
treatment with this compound," apparently due to the rarity of this
tumor in Long-Evans rats (Stout et al., 1983a).
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Alachlor March 31, 1987
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Data from a third study which ran concurrently with the Stout et al.,
(T983a) study have recently been submitted to EPA (Stout et al., 1983b),
This study used an additional treatment group, 126 ing/kg/day, that was
exposed to the new technical material (with the new stabilizer in
place of epichlorohydrin). The design of the new study was different
from the previous study because it used a variety of exposure regimens
and had the primary purpose of investigating the nature and reversi-
bility of the ocular lesions (UDS). The biased selection process in
the design of this study limits its usefulness for the quantitative
assessment of carcinogenic potential. However, the results are
useful in the qualitative assessment of the weight-of-the-evidence
for the oncogenicity of the new technical product not stabilized with
epichlorohydrin. This study indicates that the tumor response observed
in the earlier study (Daly et al., I98lb) cannot be explained on the
basis of the presence of epichlorohydrin in the test material and
suggests that a partial lifetime exposure (approximately one-fourth
the lifespan of the animals) can result in a tumor incidence similar
to that of a lifetime exposure.
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) = „ ( *)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
OF = uncertainty factor (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
No duration-specific data are available to derive a One-day HA; therefore,
it is recommended that the Ten-day HA of 0.1 mg/L, calculated below, be
applied for the One-day HA as well.
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Alachlor March 31, 1987
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Ten-day Health Advisory
The Ten-day Health Advisory is derived from the teratogenicity study
in the rabbit reported by Monsanto (1984b). As noted above, there was
an increase in potential teratogenic skeletal (scoliosis malformations) in
both the high and low dose groups. There was an increase in presacral
vertebrae at all levels tested. While a developmental effect may not apply
to a 10-kg child, this is the most sensitive end point to base the Ten-day HA
derivation. The LOAEL therefore, is 10 mg/kg/day.
The Ten-day HA for a 10 kg child is calculated as follows:
(10 mg/kg/day) (10 kg) = 0<1 mg/L (100 u ^
(1,000) (1 L/day)
where:
10 mg/kg = LOAEL (Lowest-Observed-Adverse-Effect-Level), based on the
teratogenic effects in the rabbit exposed to alachlor via
gavage for days 6-27 during gestation.
10 kg = assumed body weight of a 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.
Longer-term Health Advisory
A Longer-term Health Advisory will not be determined for alachlor because
it has been shown to produce carcinogenicity in less than five and one-half
months in rats at the same rate as did the lifetime exposure. It is recommended
that the DWEL, adjusted for a 10-kg child (0.1 mg/L) be used as a conservative
estimate for Longer-term exposure.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to d: inking water and is considered protective of noncar-
cinogenic adverse health effects over a lifeti.ne 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
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Alachlor March 31, 1987
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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 study by Naylor et al., (1984) has been selected to serve as the basis
for the Lifetime HA because a NOEL was determined to be 1 mg/kg/day based on
hemosiderosis seen in liver and spleen of dogs in the higher dose groups.
Using this study, the lifetime HA is derived as follows:
Step 1: Determination the Reference Dose (RfD)
Rfd =. (1 mgAq/day) = 0.01 mg/kg/day
(100)
where:
1 mg/kg/day = NOAEL, based on the absence of hemosiderosis in dogs
exposed to alachlor via feed for 1 year.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.01 ng/kg/day) (70 kg) z 0<35 ^ (350 ^)
(2 L/day)
where:
0.01 mg flag /day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Alachlor may be classified in Group B2: probable human carcinogen. A
Lifetime HA is not recommended. The estimated risk of a 70-kg adult consuming
2 L/day of 350 ug/L alachlor over a lifetime is 4 x 10~2. This data is generate
based on the calculated oncogenic potency, q* = 6.0 x 10 (mg/kg/day) ,
using the multistage model.
Evaluation of Carcinogenic Potential
0 EPA's Carcinogenic Assessment Group (CAG) is currently evaluating
alachlor for carcinogenic risk assessment. However, EPA's Office of
Pesticide Programs (OPP) has performed a risk characterization of the
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Alachlor March 31, 1987
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nasal tumors of alachlor (U.S. EPA, 1984). The OPP assessment for
drinXing water is summarized in the following table.
Table 1 . Assessment of Drinking Water Risks for Alachlor
Exposure Level Upper Limit Estimate of
(ug/L) Excess Lifetime Cancer Risk for;
10 Kg Child 60 Kg Adult
0.15 10-6 10-7 to 10-6
1 .5 10~5 10~6 to 10-5
15.0 10~4 10~5 to 10-4
The Office of Drinking Water uses the 70 kg man as its surrogate.
In these risk calculations, we would not expect to see any significant
change in the degree of calculated risk because of the difference in
the reference man.
Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986b), alachlor is classified in Group B:
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 AKD STANDARDS
0 OPP has estimated additional risks for alachlor (agricultural workers,
and consumers of raw agricultural commodities) (U.S. EPA, 1984).
VII. ANALYTICAL METHODS
0 Determination of alachlor may be accomplished by a liquid-liquid
extraction gas chromatographic procedure (D.S. EPA, 1983b). In this
procedure, a 1-L water sample is spiked with an internal standard and
then extracted with methylene chloride. The extract is concentrated
to 5 mL and the methylene chloride solvent is exchanged for a toluene/
methanol mixture. Separation and identification is by packed column
gas chromatography using a nitrogen selective detector.
0 The method detection limit for alachor is approximately 0.2 ug/L.
If the sample chromatogram contains interfering peaks, the sample
should also be analyzed using an electron-capture detector.
VIII. TREATMENT TECHNOLOGIES
0 Data are available on the removal of alachlor from potable water using
conventional treatment and adsorption. The use of aeration has also
been considered.
0 Available data suggest that conventional water treatment is not
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Alachlor March 31, 19S7
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effective for removing alachlor from drinking water. Baker (1982)
monitored the concentration of alachlor in raw river and in finished
water after alum coagulation, flocculation, sedimentation and filtration.
The concentration range was <0.5 to 5.0 ug/L in the influent and <0.2
to 2.0 ug/L in the effluent. The removal rate was not consistent and
generally less than 50%.
No actual data are available which demonstrate the removal of alachlor
using aeration. However, the estimated Henry's Law Constant (1.94 x
10~4 atm x m^/mole) suggests that this pesticide might be amenable to
such treatment (ESE, 1984).
Limited data suggest that GAC (granular activated charcoal) adsorption
would have limited effectiveness for alachlor. In a laboratory study
(DeFilippi et al., 1980), a waste stream containing 11 mg/L alachlor
was passed, at 1.1 gpm/ft2, through a 3/8 inch diameter, 11-inch
column containing seven grams of (GAC). After 2.6 liters had been
passed through, an effluent concentration of 0.22 mg/L broke through
the column. It was estimated that, for this effluent concentration,
a usage rate of 21.7 lb/1,000 gal would be required.
Laboratory studies with rapid sand filters capped with 16.5 inches of
GAC (Filtrasorb* 300) operated at a filtration rate of 1.2 gpm/ft2
with an empty bed contact time of nine minutes were performed by
Baker (1982). Reported alachlor concentrations ranged from 0.7 to
5.0 mg/L in the raw water and 0.1 to 0.7 mg/L in the finished water.
However, powdered activated carbon in conventional treatment (PAC
dose not reported) resulted in an average concentration reduction of
only 43%.
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Alachlor March 31, 1987
-13-
IX. REFERENCES
Ahmed, F.E., A.S. Tegeris, P.C. Underwood et al.* 1981. Alachlor: Six-month
study in the dog: Testing facility's report no. 7952; sponsor's report
no. PR-80-015. (unpublished study including submitter summary, received
Dec.1, 1981 under EPA Reg. No. 524-316; prepared by Pharmacopathics
Research Labs., Inc., submitted by Monsanto Co., Washington, D.C., CDL:
246229-A and 246293.
Baker, D. 1982. Herbicide contamination in municipal water supplies in
Northwestern Ohio. Final draft report. Prepared for Great Lakes National
Program Office, U.S. EPA, cited in ESE (1984).
Coleman, D.L., and W.R. Gaffey.* 1980. A study of individuals exposed to
alachlor: Ocular examinations for uveitis. Unpublished study received
July 30, 1980 under EPA Reg. No. 524-285; submitted by Monsanto Co.,
Washington, D.C.; CDL: 242943-A.
Daly, I.W., G.K. Hagan, R. Plutnick et al.* 1981a. An eighteen-month chronic
feeding study of alachlor in mice. Project No. 77-1064. Final report.
Unpublished Study received July 1, 1981 under EPA Reg. No. 524-285;
submitted by Monsanto Co., Washington, D.C., CDL: 070168-A, 070169.
Daly I.W., J.B. McCandless, H. Jonassen et al.* 1981b. A chronic feeding
study of alachlor in rats. Project No. 77-2065. Final report. Unpub-
lished study received Jan. 5, 1982 under EPA Reg. No. 524-285, prepared
by Bio-Dynamics, Inc. (BD-77-421, 11/13/81), submitted by Monsanto Co.,
Washington, D.C. CDL: 070586-A, 070587, 8, 9 & 90.
DeFilippi, R.P., V.J. Kyukonis, R.J. Robey and M. Modell. 1980. Supercritical
fluid regeneration of activated carbon for adsorption of pesticides.
U.S. EPA Document EPA-600/2-80-054. U.S. EPA. Research Triangle Park.
IRDC.* 1984. International Research and Development Corporation, Mattawan,
Michigan 49071 (initial study, IRDC Study #401-060, IR-79-022, dated
11/24/80, submitted to the Agency on 1/15/81 and was classified as
Invalid by the Agency in a review dated 6/5/81). The repeat study (IRDC
Study 1401-208) submitted to the Agency on 3/1/84 under Accession #252570
for review was classified as Core-Supplementary-Data ->n 8/29/84.
ESE. 1984. Environmental Science and Engineering. Review of treatability
data for removal of twenty-five synthetic organic chemicals from drinking
water. Prepared for Office of Drinking Water, U.S. EPA.
Monsanto Co.* 1978a. Acute oral-rat, acute dermal rabbit. Submitted by Bio-
Dynamics, Inc., PD-77-433 on June 28, 1978. Unpublished study received
1978; CDL: 241273.
Monsanto Co.* 1978b. Primary eye and primary dermal irritation—rabbit.
Submitted by Bio-Dynamics, Inc., PD-77-433 on March 22, 1978. Unpublished
study received 1978; CDL: 241273.
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Alachlor March 31, 1987
-14-
Monsanto Co.* 1981a. Elimination of 1^C-alachlor in monkeys. Study No.
MA-81-261, prepared by the Univ. of California School of Medicine.
Submitted by Monsanto Co., Washington, D.C., on Nov. 28, 1981. CDL:
070592, 247937.
Monsanto Co.* 1981b. Acute inhalation LDg0—rat. Submitted by Bio-Dynamics,
Inc., PD-1-183 on Dec. 3, 1981. Unpublished study received 1/81; CDL:
Monsanto Co.* 1982. Environmental fate of microencapsulated alachlor: Vol.
I & II. Unpublished study received May 26, 1982 under EPA Reg. No.
524-344, prepared by Monsanto Agricultural Products Co.,Washington,
D.C., CDL 070841. 248053.
Monsanto Co.* 1983. Rat metabolism study. MSL-3198, R.D. 493. Part I and
II. Unpublished study received Oct. 1983 under EPA Reg. No. 524-316;
prepared by Monsanto Agricultural Products Co., submitted by Monsanto
Co., Washington, D.C.; CDL: 251543 and 251544.
Monsanto Co.* I984a. Dermal sensitization—guinea pig. Submitted on March 24,
1984. Unpublished study received 1984; CDL: 252772.
Monsanto Co.* 1984b. Teratology study—rabbit. Submitted on August 29, 1984
Unpublished study recieved 1985; IRDC # 401-208, Accession # 252570.
Naylor, M.W., W.E. Ribelin, D.E. Thake, L.D. Stout and R.M. Folks.* 1984.
Chronic study of Alachlor administered by gelatin capsule to dogs.
Unpublished Study No. 820165, Environmental Health Laboratory, Monsanto
Company, St. Louis, MO for Monsanto Company. Accession No. 25593.
Rodwell, D.E., and E.J. Tracher.* 1980. Teratology study in rats. IRDC No.
401-058; IR-79-020. Unpublished study including submitter summary,
received Oct. 16, 1980 under EPA Reg. No. 524-385; prepared by Interna-
tional Research & Development Corp., Submitted by Monsanto Co., Washington,
D.C.; CDL: 252570.
Stout, L.D. et al.* 1983a. A chronic study of alachlor administered in feed
to Long-Evans rats. EHL #800218, Project # ML-80-186, Report MSL-3282/
3284. Vol. I S II. Unpublished study received Feb. 28, 1984 under U.S.
EPA Reg. No. 524-316, prepared by Monsanto Environmental Health Laboratory
(EHL), submitted by Monsanto Co., Washington, D.C., CDL: 252496-7.
Stout, L.D. et al.* 1983b. A chronic study of alachlor administered in feed
to Long-Evans rats. Project # ML-80-224, Unpublished study received
4/16/84 under EPA Reg. No. 524-316, prepared by Monsanto Environmental
Health Laboratory (EHL), submitted by Monsanto Co., Washington, D.C.,
CDL: 252498.
Schroeder, R.D., G.K. Hogan, M.E. Smock et al.* 1981. A three-generation
reproduction study in rats with alachlor. Project No. 77-2066. Final
report. Unpublished study received July 10, 1981 under EPA Reg. No.
524-285; prepared by Bio-Dynamics, Inc., submitted by Monsanto Co.,
Washington, D.C. CDL: 070177-A.
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Alachlor March 31, 1987
-15-
Shirasu et al.* 1980. Microbial mutagenicity study. Received Feb. 20, 1980
under EPA Reg. No. 524-316. Prepared by Institute of Environmental
Toxicology; Kodira Japan; submitted by Monsanto Co., Washington, D.C.;
CDL: 248053.
U.S. EPA. 1983a. U.S. Environmental Protection Agency. Occurrence of pesti-
cides in drinking Water, food, and air. Office of Drinking Water.
U.S. EPA. 1983b. U.S. Environmental Protection Agency. Method 102. Determi-
nation of alachlor, butachlor, and propachlor in wastewater. Effluent
Guidelines Division, Washington, D.C. 20460.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Alachlor. Special
review position document 1. Office of Pesticide Programs, Office of
Toxic Substances, published December 31, 1984. Washington, D.C. 20460.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Draft health effects
criteria document. Office of Drinking Water.
U.S. EPA. 1986a. Environmental Protection Agency. Office of Pesticide Programs
Alachlor Special Review, Position 2/3. Office of Pesticides and Toxic
Substances. 401 'M1 Street, S.W., Washington, D.C. 20460
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidance for carcinogen
risk assessment. Federal Register. 51(185):33992-34003. September 24.
U.S. FDA. 1984. U.S. Environmental Protection Agency. Surveillance index
for Pesticides. Bureau of Foods.
Windholz, M. 1983. The Merck Index. 10th Edition. Merck and Co., Inc.,
Rahway, N.J., p. 31.
•Confidential Business Information.
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IB
March 31, 1987
ALDICARB
(Sulfoxide and Sulfone)
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. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates b'y
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 stal ad values. Excess cancer risk estimates may also be calculated us^ng
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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17
Aldicarb March 31, 1987
-2-
This Health Advisory (HA) is based on information presented in the Office
of Drinking Hater's draft Health Effects Criteria Document (CD) for Aldicarb
(U.S. EPA, 1985). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological 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 # 86-117751/AS. The toll-free number is
(800) 336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.; 116-06-3
Structural Formula:
CH3 O
1 1
CH3-S-C-CH=NOCN-CH3
I I
CH3 H
2-methyl-2-(methylthio)propionaldehyde 0-methylcarbamoyl oxime
Synonyms; Temik*
Use; Pesticide (nematocide, acaracide)
Properties (U.S. EPA, 1985)
Chemical Formula
Molecular Weight 190.3
Physical State (room temp.) white crystals
Boiling Point decomposes above 100°C
Melting Point 100°C
Density
Vapor Pressure 0.05 torr at 20°C
Specific Gravity 1.195 at 25°C
Water Solubility (' g/L (room temp.)
Taste Threshold (water) —
Odor Threshold (water) —
Odor Threshold (air) odorless to light sulfur smell
Conversion Factor —
Occurrence
0 EPA estimated that aldicarb production ranged from 3.0 to 4.7 million
Ibs per year during 1979-1981. Aldicarb is applied both to the soil
and directly to plants.
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IS
Aldicarb March 31, 1987
-3-
Aldicarb is considered to be moderately persistent as a pesticide.
Aldicarb is metabolized rapidly by plants after application to its
sulfoxide and sulfone. Once in the soil, aldicarb is degraded by
both aerobic and anaerobic bacteria. Aldicarb has a soil half life of
2 to 6 weeks, with residual levels found up to 6 to 12 months later.
Aldicarb in pond water was reported to degrade more rapidly, with a
half life of 5 to 10 days. Aldicarb is expected to hydrolyze slowly
over months or years in most ground and surface waters. Aldicarb and
its sulfoxide and sulfone degradation products do not bind to soil
or sediments and have been shown to migrate extensively in soil.
Aldicarb does not bioaccumulate to any significant extent.
Aldicarb has been reported to occur widely in ground water at levels
in the low ppb range. New York, Florida, Wisconsin and Maine, among
other states, have restricted the use of aldicarb based upon its
potential for ground water contamination. Aldicarb has not been
analyzed for in Agency surveys of drinking water and estimates of
national exposures are unavailable. Because of aldicarb's relatively
rapid degradation rate, it is expected to occur more often in ground
waters than surface waters (U.S. EPA, 1983).
Monitoring of aldicarb residues on foods have found only occasional
low levels of the pesticide and its metabolites (U.S. FDA, 1984).
The Agency has set limits for residues which would result in an adult
receiving a daily dose of 100 ug/kg a day. For drinking water exposures
to exceed this dose, concentrations would need to exceed 50 ug/L.
III. PHARMACOKINETICS
Absorption
0 Aldicarb, as well as its sulfoxide and sulfone metabolites, has been
shown to be absorbed readily and almost completely through the gut
in a variety of mammalian and non-mammalian species (Knaak et al.,
1966; Andrawes et al,, 1967; Dorough and Ivie, 1968; Dorough et al.,
1970; Hicks et al., 1972; Cambon et al., 1979).
8 Dermal absorption of aldicarb has been demonstrated in rabbits (Kuhr
and Dorough, 1976; Martin and Worthing, 1977) and rats (Gaines, 1969),
and would be expected to occur in unprotected humans in manufacturing
and field application settings.
Distribution
0 Aldicarb is distributed widely in the tissues of Holstein cows when
administered in feed (Dorough et al., 1970). Highest residues were
found in the liver. When aldicarb was administered at a lower level,
residues were detected only in the liver.
0 In rats administered aldicarb orally, residues were found in all 13
tissue types analyzed. Hepatic residue levels were similar to those
of many other tissues (Andrawes et al., 1967).
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•10
-*_%./
Aldicarb March 31, 1987
-4-
0 Aldicarb. in a 1:1 molar ratio of the parent compound to the sulfone,
administered orally to laying hens in a single dose or for 21
consecutive days resulted in similar patterns of distribution with
the liver and kidneys as the main target organs (Hicks et al., 1972).
Residues also were present in both the yolks and whites of the eggs
laid by these hens.
Metabolism
0 The metabolism of aldicarb involves both hydrolysis of the carbamate
ester and oxidation of the sulfur to sulfoxide and sulfone derivatives
which have been shown to be active cholinesterase inhibitors (Andrawes,
et al., 1967; Bull et al., 1967).
0 Metabolic end products of aldicarb detected in both the milk and
urine of a cow included the sulfoxides and sulfones of the parent
compound, oxime and nitrile, as well as a number of unknown metab-
olites (Dorough and Ivie, 1968).
Excretion
Elimination of aldicarb and its metabolism products occurs primarily
via the urine as demonstrated in rats (Knaak et al., 1966), cows
(Dorough and Ivie, 1968) and chickens (Hicks et al., 1972).
Excretion of aldicarb via the lungs as C(>2 has been demonstrated
as a minor route in rats (Knaak et al., 1966) and in the milk of
cows (Dorough and Ivie, 1968).
Excretion of aldicarb is relatively rapid with reported 24-hour
elimination values in rats and cows of approximately 80% to 90% of
the administered dose (Knaak et al., 1966; Dorough and Ivie, 1968).
IV. HEALTH EFFECTS
Humans
In two related incidents in 1978 and 1979, ingestion of c icumbers
presumed to contain aldicarb at about 7 to 11 ppm resulted in complain:.s
of diarrhea, abdominal pain, vomiting, nausea, excessive perspiration,
dyspnea, muscle fasciculation, blurred vision, headaches, convulsions
and/or temporary loss of limb function in a total of fourteen residents
of a Nebraska town (CDC, 1979; Goes et al., 1980). Onset of symptoms
occurred within 15 minutes to 2.25 hours and they continued for
approximately 4 to 12 hours.
Industrial exposure by a man bagging aldicarb for one day resulted in
nausea, dizziness, depression, weakness, tightness of chest muscles,
and decreases in plasma and red blood cell cholinesterase activity
(Sexton, 1966). The symptoms lasted more than six hours but the subject
returned to work the following day without symptoms.
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so
Aldicarb March 31, 1987
-5-
In a laboratory study, four adult males orally administered aldicarb
at 0.1 mg/kg experienced a variety of cholinergic symptoms including
malaise, weakness in their limbs, pupil contraction and loss of photo-
reactivity, epigastric cramps, sweating, salivation, nausea, vomiting
and "air hunger" (Raines, 1971). These symptoms did not occur at
0.025 or 0.05 mgA9« Depression of cholinesterase activity occurred
in a dose-dependent manner with values as low as 25% of the control
value measured in two subjects dosed at 0.1 mg/kg.
Fiore et al (1986) studied the effect of chronic exposure to aldicarb-
contaminated groundwater on the human immune function. The study has
been performed on women between the age of 18 to 70. A group of
twenty-three women were exposed to low levels of aldicarb (<61 ppb)
and another group of 27 women were unexposed. The results of this
study suggest a potential association between exposure to aldicarb
and abnormalities in T-cells. However, the statistical analysis of
these data indicates that additional studies are needed before further
^conclusions can be made on the effect of aldicarb on the immune
function.
Animals
Short-term Exposure
NAS (1977) stated that the acute toxicity of aldicarb is probably
the greatest of any widely used pesticide.
Reported oral LDsg values for aldicarb administered to rats in corn or
peanut oil range from about 0.65 to 1 mg/kg (Weiden et al., 1965;
Gaines, 1969). Females appear to be more sensitive than males. The
oral LD5Q in mice is 0.3 to 0.5 mgAg (Black et al., 1973).
Oral LD5Q values for aldicarb were higher when using a vehicle other
than corn or peanut oil. Weil (1973) reported an oral 1.059 of 7.07
mg/kg in rats administered aldicarb as dry granules. Carpenter and
Smyth (1965) reported an 1*050 of 6.2 mg/kg in rats administered aldicarb
in drinking water.
Dermal toxicity also is high with 24-hour LD5Q values of 2.5 and 3
mg/kg reported for female and male rats, respectively (Gaines, 1969)
and 5 mg/kg in rabbits (Weiden et al., 1965).
The principal toxic effect of aldicarb and its sulfoxide and sulfone
metabolites in rats has been shown to be cholinesterase inhibition
(Weil and Carpenter, 1963; Nycum, 1968; Weil, 1969).
Feeding studies of short duration (7 to 15 days) have been conducted
by various authors using aldicarb and/or its sulfone and sulfoxide.
Statistically significant decreases in cholinesterase activity were
observed in rats at dosage levels of 1 mg/kg/day (the approximate
LD5Q in rats) (Nycum and Carpenter, 1970) and at 2.5 mg/kg/day in
chickens (Schlinke, 1970). The latter dosage also resulted in some
lethality in test animals.
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21
Aldicarh March 31, 1987
-6-
0 A NOAEL has been determined for a mixture of aldicarb oxidation
products based on data reported by Mirro et al. (1982) who administered
aldicarb sulfone and sulf oxide in a 1:1 ratio in the drinking water
of young rats for 8 to 29 days. Doses ranged up to 1.67 mg/kg/day
for males and 1.94 mg/kg/day for females. Based on statistically
significant reductions in cholinesterase activity in brain, plasma
and RBC at higher dosage levels, a NOAEL of 0.12 mg/kg/day was
determined.
Long-term Exposure
0 High dosages of aldicarb sulfoxide (0.25 to 1.0 mgAg/day) or aldicarb
sulfone (1.8 to 16.2 mg/kg/day) administered in the diets of rats for
three or six months resulted in decreases in cholinesterase activity
in plasma, RBCs and brain (Weil and Carpenter, 1968a,b). No increases
in mortality or gross or microscopic histopathology were noted in any
group, however. Data derived from the lower dosage levels of this
study have been used by the World Health Organization Committee on
Pesticide Residues (FAO/WHO, 1980) to derive a NOAEL of 0.125 mgAg/day
for aldicarb sulfoxide in the rat. The NOAEL for aldicarb sulfone
alone was 0.6 mg/kg/day.
0 Aldicarb administered for two years in the diets of rats or dogs at
•dosage levels up to 0.1 mg/kg/day resulted in no significant increases
in adverse effects based on a variety of toxicologic endpoints (Weil
and Carpenter, 1965, 1966a). In another two-year study, levels of up
to 0.3 mgAg/day resulted in no adverse effects in rats (Weil, 1975).
Feeding studies using aldicarb sulfoxide at 0.6 mg/kg/day for two
years resulted in an increase in the mortality rates of female rats
(Weil, 1975).
Reproductive Effects
No reproductive effects have been demonstrated to result from the
administration of aldicarb to rats (Weil and Carpenter, 1964, 1974)
Developmental Effects
No teratogenic effects have been demonstrated from the administration
of aldicarb in rabbits (IRDC, 1983) or chickens (Proctor et al., 1976)
No adverse effects on milk production were observed in studies of
lactating cows or rats (Dorough and Ivie, 1968; Dorough et al., 1970).
Statistically significant inhibition of acetylcholinesterase activity
has been demonstrated in the liver, brain and blood of rat fetuses
when their mothers were administered aldicarb by gastric intubation
on day 18 of gestation (Cambon et al. , 1979). These changes were
seen at doses of 0.001 mq/kg and above and were manifested within
five minutes of the administration of 0.1 rag/kg-
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Aldicarb March 31, 1987
-7-
Mutagenicity
0 Aldicarb has not been demonstrated to be conclusively mutagenic in
Ames bacterial assays or in a dominant lethal mutagenicity test in
rats (Ercegovich and Hashed, 1973; Weil and Carpenter, 1974; Godek
et al., 1980).
Carcinogenicity
0 Neither aldicarb nor its sulfoxide or sulfone have been demonstrated
to increase significantly the incidence of tumors in mice or rats in
feeding studies (Weil and Carpenter, 1965; NCI, 1979). Bioassays
with aldicarb in which rats and mice were fed either 2 or 6 ppm in
the diet for 103 weeks revealed no tumors that could be attributed
solely to aldicarb adninistration (NCI, 1979). It was concluded that,
under the conditions of the bioassay, technical grade (99+%) aldicarb
was not carcinogenic to F344 rats or B6C3F1 mice of either sex. A
two-year feeding study reported by Weil and Carpenter (1965) also
produced no statistically significant increase in tumors over controls
when rats were administered aldicarb at equivalent doses of 0.005,
0.025, 0.05 or 0.1 mg/kg bw/day in the diet. Weil (1975) similarly
reported no adverse effects in Greenacres Laboratory Controlled Flora
rats fed aldicarb at 0.3 mg/kg bw/day for 2 years.
0 In the only skin-painting study available to date, Weil and Carpenter
(1966b) found aldicarb to be noncarcinogenic to male C3H/H3J mice
under the conditions of the experiment.
0 Intraperitoneally administered aldicarb did not exhibit transforming
or tumorigenic activity in a host-mediated assay using pregnant
hamsters and nude (athymic) mice (Quarles, et al, 1979),
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 ( ug/L)
(OF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
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Aldicarb March 31 , 1987
-8-
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
The available data suggest that the appearance of cholinergic symptoms
indicative of cholinesterase enzyme inhibition is the most sensitive indicator
of the effects of exposure to aldicarb. Adverse health effects appear to be
related primarily to the depression of cholinesterase activity, as no other
biochemical, morphological, reproductive, mutagenic or carcinogenic effects
have been reported, even after chronic dosing.
Given the nature of the primary toxicity (rapidly reversible cholinesteras*
inhibition) of aldicarb and its oxidative metabolites/degradation products,
it is apparent that the same NOAEL can be used as the basis for the derivation
of acceptable levels over virtually any duration of exposure. In addition,
the Health Advisories calculated in this document are appropriate for use in
circumstances in which the sulfoxide and/or sulfone may be the substance(s)
present in a drinking water sample. Depending upon the analytical method
applied, it may not be possible to characterize specifically the residue(s)
present. By establishing Health Advisories based upon data from valid
studies with the most potent of the three substances, there is greater
assurance that the guidance is protective to human health.
As described above, a NOAEL of 0.125 mg/kg bw/day can be determined from
the Weil and Carpenter (1968b) and Mirro et al., (1982) studies. From this
NOAEL, all HA values can be determined for aldicarb, aldicarb sulfoxide or a
mixture of the sulfoxide and sulfone metabolite (however, if for any reason
one finds that the contaminant is only the sulfone and wants to use a less
conservative value, the NOAEL for the sulfone, 0.6 mg/kg/day, as determined
in the Weil and Carpenter (1986) study, can be used).
One-day Health Advisory
For the 10 kg child:
One-day HA = (0.125 mg/kg/day) (10 kg) = 0.012 mg/L (10 ug/L)
(100) (1 L/day)
where:
0.125 mg/kg/day = NOAEL, based upon lack of significant decreases
in cholinesterase activity in rats.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
(Note: Using the NOAEL for the sulfone alone, the HA value for this metabolite
may also be 0.06 mg/L (60 ug/L) if the sulfone is the only contaminant.)
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Aldicarb March 31, 1987
-9-
Ten-day Health Advisory
Since aldicarb is metabolized and excreted rapidly (>90% in urine alone
in a 24-hour period following a single dose), the One- and Ten-day HA values
would not be expected to differ to any extent. Therefore, the Ten-day HA
will be the same as the One-day HA (10 ug/L).
Longer-term Health Advisory
For the 10 kg child:
Longer-term HA = (0.125 mg/kg/day) (10 kg) = 0.012 mg/L (10 ug/L)
(100) (1 L/day)
where:
0.125 mgAg/dav = NOAEL, based upon lack of significant decreases
in cholinesterase activity in rats.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
(Note; Using the NOAEL for the sulfone alone, the HA value for this metabolite
may also be 0.06 mg/L (60 ug/L) if the sulfone is the only contaminant.)
For the 70 kg adult:
Longer-term HA = (0.125 mg/kg/day) (70 kg) = Q>042 mg/L (40 ug/L)
(100) (2 L/day)
where:
0.125 mg/kg/day = NOAEL, based upon lack of significant decreases
in cholinesterase activity in rats.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
(Note: Using the NOAEL for the sulfone alone, the HA value for this metabolite
may also be 0.21 mg/L (210 ug/L) if the sulfone is the only contaminant.)
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
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Aldicarb March 31, 1987
-10-
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 classifed 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.
As discussed before on page 8, the studies by Weil and Carpenter (1968b)
and Mirro et al. (1982) are used in the following calculations. Both studies
reflected a NOAEL of 0.125 mg/kg/day.
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.125 ing/kg/day) . 0.00125
(100)
where:
0.125 mgAg/day = NOAEL, based upon lack of significant decreases
in cholinesterase activity in rats.
100 = uncertainty factor, chosen in accordance with NAS/ODK
guidelines for use with a NOAEL from an animal study.
(Note; Using the NOAEL of 0.6 mg/kg/day for the sulfone alone the RfD value
for this metabolite may also be 0.006 mg/kg/day.)
Step 2: Determination of the Drinking Water Equivalent Leve] (DWEL)
DWEL = (0.00125 mg/kg/day) (70 kg) . 0>042 mg/L (40 ug/L)
(2 L/day)
where:
0.00125 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
(Note: Using the RfD for sulfone alone, the DWEL for this metabolite may also
be 0.21 mg/L (210 ug/L)).
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Aldicarb March 31, 1987
-11-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA - (0.042 mg/L) (20%) =0.009 mg/L (10 ug/L)
where:
0.42 mg/L « DWEL.
20% = assumed contribution of drinking water to total exposure
to aldicarb.
(Note; Using the DWEL for sulfone alone, the Lifetime HA value for this
metabolite may also be 0.042 mg/L (42 ug/L).
In summary, the Lifetime HA values for aldicarb and its metabolites are
as follows:
aldicarb (parent compound)*: 10 ug/L
aldicarb sulfoxide* : 10 ug/L
aldicarb sulfone** : 10 to 42 ug/L
* The HA values for aldicarb and aldicarb sulfoxide are the same because they
have similar toxicity, and the effects of the parent compound are likely
due to the sulfoxide (and, to a lesser extent, the sulfone).
**The HA value for the sulfone ranges from 10 to 42 ug/L depending on the
presence or absence of other aldicarb/aldicarb sulfoxide residues; only if
the sulfone metabolite is present alone as a contaminant, the HA value of
42 ug/L may be used.
Evaluation of Carcinogenic Potential
0 Since aldicarb has been found to be noncarcinogenic under all
conditions tested, "quantification of carcinogenic risk for lifetime
exposures through drinking water would be inappropriate.
0 The International Agency for Research on Cancer (IARC) has not
classified the carcinogenic potential of aldicarb.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), the Agency has classified aldicarb
in Group E: No evidence of carcinogenicity in humans. This category
is used for agents that show no evidence of carcinogenicity in at
least two adequate animal tests in different species or in both
epidemiologic and animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Academy of Sciences proposed an ADI of 0.001 mg/kg/day
based upon the two-year feeding studies in rats and dogs (NAS, 1977).
NAS reaffirmed this ADI in 1983 (NAS, 1983).
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Aldicarb March 31, 1987
-12-
0 In addition, HAS also derived a chronic suggested-no-adverse-effeet-
level (SNARL) of 7 ug/L, using the studies mentioned above with an
uncertainty factor of 1,000 (HAS, 1977). The SNARL is protective of a
70 kg adult, consuming 2 liters of water per day and for whom drinking
water is assumed to contribute 20 percent of the daily exposure to
aldicarb residues.
0 EPA's Office of Pesticide Programs established an ADI of 0.003
ng/kg/day based upon the data from the six-month rat feeding study
with aldicarb sulfoxide (U.S. EPA, 1981).
0 The FAO/WHO proposed ADIs for aldicarb residues of 0-0.001 mg/kg/day
in 1979 and 0-0.005 mgAg/day in 1982 (FAO/WHO, 1979; 1982).
VI. ANALYTICAL METHODS
0 Analysis of aldicarb is by a high performance liquid chromatographic
procedure used for the determination of N-methyl carbamoyloximes and
N-methylcarbamates in drinking water (U.S. EPA, 1984). In this
method, the water sample is filtered and a 400 uL aliquot is injected
into a reverse phase HPLC column. Separation of compounds is achieved
using gradient elution chromatography. After elution from the HPLC
column, the compounds are hydrolyzed with sodium hydroxide. The
methylamine formed during hydrolysis is reacted with o-phthalaladehyde
(OPA) to form a fluorescent derivative which is detected using a
fluorescence detector. The method detection limit has been estimated
to be approximately 1.3 ug/L for aldicarb.
VIII. TREATMENT TECHNOLOGIES
0 Techniques which have been used to remove aldicarb from water are
carbon adsorption and filtration. Since aldicarb is converted
into aldi'carb sulfoxide and sulfone, all three compounds must be
considered when evaluating the efficiency of any decontamination
technique.
0 Granular activated carbon (GAC) has been used in two studies of aldicarb
removal from contaminated water (Union Carbide, 1979; ESE 1984). Both
studies utilized home water treatment units rather than large scale
water treatment systems. Union Carbide tested the Hytest Model HF-1
water softener in which the ion exchange ion was replaced with 38.5
Ib Filtrasorb * 400 (Calgon GAC). The unit was operated at a flow rate
of 3 gal/min. Water spiked with 200 ppb or 1000 ppb of a mixture of
aldicarb, aldicarb sulfoxide and aldicarb sulfone in a 10:45:45 ratio
was treated. Under these conditions, the total aldicarb residue
level was reduced by 99% to 1 ppb for the treatment of 13,500 gallons
of water with 200 ppb of residues and 41,500 gallons with 1000 ppb
total residues. No breakthrough of aldicarb occurred. When the
study was terminated, the carbon had adsorbed 9 mg aldicarb residue
per gram. This value can be compared with an equilibrium loading
value of 21 mg per gram of carbon at 166 determined using 200 ppb
aldicarb residues. In the second study, ESE (1984) did a field
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28
Aldicarb March 31, 1987
-13-
Btudy in Suffolk County, NY. Nineteen units using type CW 12 x 40
mesh carbon were tested. After 38 months of use, breakthrough of
aldicarb occurred to levels over 7 ug/L in eight units tested.
The range of usage values can be attributed to the fact that the
natural well samples contained a variety of adsorbable substances
in addition to aldicarb.
Chlorination also appears to offer the potential for aldicarb removal
(Union Carbide, 1979). The company reported that 1.0 ppm free chlorine
caused a shift in the ratio of aldicarb, its sulfoxide and its sulfone
so that all residues were converted to the sulfoxide within five
minutes of chlorine exposure. Normal conversion of aldicarb to
aldicarb sulfone did not appear to be affected. On standing, the
sulfoxide and sulfone decomposed. The decomposition products were
not identified. However, should these be non-toxic, then chlorination
could be feasible as an aldicarb removal technique.
Aeration or air stripping which is commonly used to remove synthetic
organic chemicals is not a good technique for the removal of aldicarb
(ESE, 1984). This is because aldicarb has a low Henry's Law Constant
(2.32 x 10~4 atm).
-------
29
Aldicarb March 31, 1987
-14-
IX. REFERENCES
Andrawes, N.R., H.W. Dorough and D.A. Lindquist. 1967. Degradation and
elimination of Temik in rats. J. Econ. Entomol. 60(4):979-987.
Black, A.L., Y.C. Chiu, M.A.R. Fahmy and T.R. Fukuto. 1973. Selective
toxicity of N-sulfenylated derivatives of insecticidal nethylcar-
bamate esters. J. Agr. Food Chen. 21:747-751.
Bull, D.L., D.A. Lindquist and J.R. Coppedge. 1967. Metabolism of 2-
methyl-2-(methylthio)propionaldehyde 0-(methyl carbamoyl) oxime
(Temik, DC-21149) in insects. J. Agr. Food Chem. 15(4):610-616.
Cambon, C., C. Declume and R. Derache. 1979. Effect of the insecticidal
carbamate derivatives (carbofuran, primicarb, aldicarb) in the activity
of acetylcholinesterase in tissues from pregnant rats and fetuses.
Toxicol. Appl. Pharmacol. 49:203-208.
Carpenter, C.P. and H.F. Smyth. 1965. Recapitulation of pharraacodynamic
and acute toxicity studies on Temik. Mellon Institute Report No. 28-78.
EPA Pesticide Petition No. 9F0798.
CDC (Centers for Disease Control). 1979. Epidemiologic notes and reports:
Suspected carbamate intoxications — Nebraska. Morbid. Mortal. Week.
Rep. 28:133-134.
Dorough, H.W., R.B. Davis and G.W. Ivie. 1970. Fate of Temik-carbon-14
in lactating cows during a 14-day feeding period. J. Agr. Food Chem.
18(1):135-143.
Dorough, H.W. and G.W. Ivie. 1968. Temik-S35 metabolism in a lactating
cow. J. Agr. Food Chem. 16(3):460-464.
Ercegovich, C.D. and K.A. Rashid. 1973. Mutagenesis induced in mutant
strains of Salmonella typhimurium by pesticides. Abstracts of Papers.
Am. Chem. Soc. p. 43.
ESE. 1984. Environmental Science and Engineering. Review of treat-
ability data for removal of twenty-five synthetic organic chemicals
from drinking water. Prepared for EPA's Office of Drinking Water.
FAO/WHO. 1979, 1980 and 1982. References not available.
Fiore, M.C., H.A. Anderson, R. Hong, R. Golubjatnikov, J.E. Seiser,
D. Nordstrom, L. Hanrahan and D. Belluck. 1986. Chronic Exposure
to Aldicarb—Contaminated Groundwater and Human Immune Function.
Enw. Res. 41: 633-645.
Gaines, T.B. 1969. The acute toxicity of pesticides. Toxicol. Appl.
Pharmacol. 14:515-534.
Godek, E.S., M.C. Dolak, R.W. Naismith and R.J. Matthews. 1980. Ames
Salmonella/Microsome Plate Test. Unpublished report by Pharmakon
Laboratories. Submitted to Union Carbide June 20, 1980.
-------
30
Aldicarb March 31, 1987
-15-
Goes, E.H., E.P. Savage, G. Gibbons, M. Aaronson, S.A. Ford and H.W.
Wheeler. 1980. Suspected foodborne carbamate pesticide intoxications
associated with ingestion of hydroponic cucumbers. Am. J. Epidemiol.
111:254-259.
Haines, R.G. 1971. Ingestion of aldicarb by human volunteers: A
controlled study of the effect of aldicarb on man. Union Carbide
Corp., Unpublished report with addendum (A-D), Feb. 11, 1971, 32
pages.
Hicks, B.W., H.W. Dorough and H.M. Mehendale. 1972. Metabolism of aldi-
carb pesticide in laying hens. J. Agr. Food Chem. 20(1):151-156.
IRDC. 1983. International Research and Development Corporation. 1983.
Teratology study in rabbits. Union Carbide Corporation.
Knaak, J.B., M.J. Tallant and L.J. Sullivan. 1966. The metabolism of 2-
methyl-2-(methylthio) propionaldehyde 0-(methyl carbamoyl) oxime in
the rat. J. Agr. Food Chen. 14(6):573-578.
Kuhr, R.J. and H.W. Dorough. 1976. Carbamate Insecticides: Chemistry,
Biochemistry, and Toxicology. CRC Press, Inc., Cleveland, OH. pp. 2-6.
103-112, 187-190, 211-213, 219-220.
Martin, H. and C.R. Worthing, Ed. 1977. Pesticide Manual. British Crop
Protection Council, Worcestershire, England, p. 6.
Mirro, E.J., L.R. DePass and F.R. Frank. 1982. Aldicarb sulfone: aldicarb
sulfoxide twenty-nine-day water inclusion study in rats. Mellon
Inst. Rep. No. 45-18.
NAS. 1977. National Academy of Sciences. Drinking Water and Health
Volume 1. National Academy Press. Washington, D.C. pp. 635-643.
NAS. 1983. National Academy of Sciences. Drinking Water and Health
Volume 5. National Academy Press. Washington, D.C. pp. 10-12.
NCI. 1979. National Cancer Institute. Bioassay of aldicarb for possible
carcinogenicity. National Institutes of Health. U.S. Public Health
Service. U.S. Department of Health, Education and Welfare.
Washington, D.C. NCI-CG-TR-136.
Nycum, J.S. 1968. Toxicity studies on Temik and related carbamates.
Mellon Institute, unpublished report 31-48, 5 pages.
Nycum, J.S., and C. Carpenter. 1970. Summary with respect to Guideline
PR70-15. Mellon Institute Report No. 31-48. EPA Pesticide Petition
No. 9F0798.
Proctor, N.H., A.D. Moscioni and J.E. Casida. 1976. Chicken embryo NAD
levels lowered by teratogenic organophosphorus and methylcarbamate
insecticides. Biochem. Phannacol. 25:757-762.
-------
31
Aldicarb March 31, 1987
-16-
Quarles, J.M., M.W. Sega, C.K. Schenley and W. Lijinsky. 1979. Trans-
formation of hamster fetal cells by nitrosated pesticides in a
transplacental assay. Cancer Res. 39:4525-4533.
Schlinke, J.C. 1970. Toxicologic effects of five soil nematocides in
chickens. J. Am. Vet. Med. Assoc. 31:119-121.
Sexton, W.F. 1966. Report on aldicarb. EPA Pesticide Petition No.
9F0798, Section C.
Union Carbide. 1979. Union Carbide Agricultural Products Company. Temik •
aldicarb pesticide. Removal of residues from water. Research and
Development Department.
U.S. EPA. 1981. U.S. Environmental Protection Agency. 40 CFR 180.
Tolerances and exemptions from tolerances for pesticide chemicals in or
on agricultural commodities: aldicarb. Federal Register 46 (224): 57047.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of pesti-
cides in drinking water, food, and air. Office of Drinking Water.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Method 531. Meas-
urement of N-methyl carbamoyloximes and N-methylcarbamates in drinking
water by direct aqueous injection HPLC with post column derivatization.
Enviromental Monitoring and Support Laboratory, Cincinnati, Ohio 45268.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Draft health effects
criteria document for aldicarb. Criteria and Standards Division.
Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51(185)33992-34003.
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index for
Pesticides. Bureau of Foods.
Weiden, M.H.J., H.H. Moorefield and L.K. Payne. 1965. o-(Methyl carbamoyl)
oximes: A new class of carlamate insecticides-acaracides. J. Econ.
Entomol. 58:154-155.
Weil, C.S. 1969. Purified and technical Temik. Results of feeding in
the diets of rats for one week. Mellon Institute, unpublished report
32-11, 6 pages.
Weil, C.S. 1973. Aldicarb, Seven-day inclusion in diet of dogs. Carnegie-
Mellon Institute of Research, unpublished report 36-33, 4 pages.
Weil, C.S. 1975. Mellon Institute Report No. 35-72, Section C. EPA
Pesticide Petition No. 3F1414.
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Aldicarb March 31, 1987
-17-
Weil, C.S. and C.P. Carpenter. 1963. Results of three months of inclusion
of Compound 21149 in the diet of rats. Mellon Institute, unpublished
report 26-47, 13 pages.
Weil, C.S. and C.P. Carpenter. 1964. Results of a three-generation
reproduction study on rats fed Compound 21149 in their diet. Mellon
Institute Report No. 27-158. EPA Pesticide Petition No. 9F0798.
Weil, C.S. and C.P. Carpenter. 1965. Two year feeding of Compound 21149
in the diet of rats. Mellon Institute, unpublished report 28-123, 40
pages.
Weil, C.S. and C.P. Carpenter. 1966a. Two year feeding of Compound
21149 in the diet of dogs. Mellon Institute, unpublished report
29-5, 22 pages.
Weil, C.S. and C.P. Carpenter. 1966b. Skin painting in mice. No
reference available.
Weil, C.S. and C.P. Carpenter. 1968a. Temik sulfoxide. Results of
feeding in the diet of rats for six months and dogs for three months.
Mellon Institute Report No. 31-141. EPA Pesticide Petition No. 9F0798.
Weil, C.S. and C.P. Carpenter. 1968b. Temik sulfone. Results of feeding
in the diet of rats for six months and dogs for three months. Mellon
Institute Report No. 31-142. EPA Pesticide Petition No. 9F0798.
Weil, C.S. and C.P. Carpenter. 1974. Aldicarb. Inclusion in the diet
of rats for three generations and a dominant lethal mutagenesis test.
Carnegie-Mellon Institute of Research. Unpublished report 37-90,
46 pages.
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33 March 31, 1987
CARBOFURAN
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. The chemical 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 th£.t
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.
-------
Caroofuran
March 31, 1987
-2-
This Health Advisory is based on information presented in the Office of
Drinking Water's draft Health Effects Criteria Document (CD) for Carbofjran
(U.S. EPA, 1985a). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological 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 f86-118007/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1553-66-2
Structural Formula
2,3-dihydro-2,2-dimethyl-7-benzof aranyl-N-methylcarbamate
Synonyms
0 Furadan®, Curaterr®.
Uses
0 Pesticide (insecticide, acaricide, nematocide)
Properties (Windholz, 1983; Kuhr and Dorough, 1976; Midwest Research
Institute, 1976; Cook, 1973)
Chemical Formula
Molecular Weight
Physical State (roo:a temp.)
Boiling Po: nt
Melting Point
Densi ty
Vapor Pressure
Water Solubility
Octanol/Water Partition Coefficient
Taste Threshold (water)
Odor Threshold (water)
Odor Threshold (air)
Conversion Factor
C12H15N03
221.26
white, crystalline solid
153° to 154°C
2x10-5 mm Hg (33°C)
1.1x10-4 mm Hg (50°C)
700 ng/1 (25°C)
— (odorless to slightly
phenolic)
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35
Carbofuran March 31, 1987
-3-
Oceurrence
• Carbofuran has a large production volume; EPA estimated that more
than 10 Billion Ibs were produced in 1980. Carbofuran is applied to
the soil and directly to plants. Because of its water solubility
(700 ppm), Carbofuran is taken up by the plants both from the soil
and from leaves.
0 Carbofuran is degraded in the environment by a number of mechanisms.
Carbofuran is metabolized rapidly by plants after application and,
once in the soil, is degraded over 2 to 3 months. Repeated applications
of carbofuran do not result in an accumulation of residues. Carbofuran
is expected to be stable in most surface and ground waters; however,
significant hydrolysis may occur in alkaline waters. Carbofuran does
not bind to soil or sediments and has been shown to migrate extensively
in soil. Carbofuran does not bi©accumulate.
0 Carbofuran has been reported to occur in ground water by the States
and other sources. Carbofuran has not been monitored in past Agency
surveys of drinking water. Based upon carbofuran1s physical and
chemical properties, carbofuran has a potential for contaminating
both ground and surface water (U.S. EPA, 1983).
0 Monitoring of carbofuran residues in or on foods has yielded only
occasional low levels of the parent compound and its metabolites
(U.S. FDA, 1984).
III. PHARMACOKINETICS
Absorption
0 Carbofuran administered to female mice by gavage was absorbed rapidly
(Ahdaya et al., 1981); approximately 51% after 15 minutes and 67%
after 60 minutes. Ahdaya and Guthrie (1982) presented evidence that
stomach absorption was about 28% of total absorption.
0 Dermal absorption also has been shown to be rapid in female mice; 33%
was absorbed after five minutes, 76% after 60 minutes, and 95% 8 hours
after application (Shah et al., 1981).
Distribution
0 Distribution of orally administered carbofuran one hour after dosing
in mice whose stomachs had been ligated included similar levels in
the liver and blood (about 1%) with the highest levels observed in
the urine and carcass (about 5%) (Ahdaya and Guthrie, 1982).
• Dermal application of carbofuran to mice resulted in a one-hour
distribution pattern of approximately 1% total in the liver, fat
and blood, about J3% total in urine, carbon dioxide and feces, and
more than 66% in'the remaining carcass. At eight hours, approximately
16% had been distributed to various organs and tissues, 6% remained
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36
Carbofuran March 31, 1987
-4-
in the gastrointestinal tract and about 73% was recovered in
excretory products (Shah et al., 1981).
Metabolism
Metabolism of carbofuran in plants, insects, rats, and mice appears
to consist of hydroxylation and/or oxidation reactions resulting
in the formation of carbofuran phenol, 3-hydroxycarbofuran, 3-
hydroxycarbofuran-7-phenol, 3-ketofuran, and/or 3-ketofuran-7-
phenol (Dorough, 1968; Metcalf et al., 1968).
Hydrolysis is a significant pathway for carbofuran to be metabolized
in mammals, but is considered minor in insects and plants.
Excretion
Elimination of carbofuran has been shown to be rapid with approximately
72% of a single orally administered dose excreted in the urine of
rats within 24 hours and a total of about 92% after 120 hours (Dorough,
1968). Total fecal excretion was about 3%.
Some pulmonary excretion of carbofuran was shown by Ahdaya et al.
(1981) who reported that after 60 minutes, 6% and 24% of an orally
administered dose were recovered in exhaled carbon dioxide and urine,
respectively, of mice.
Dermal administration to female mice resulted in higher levels of
fecal excretion of carbofuran with two-thirds of the residues
recovered from feces and one-third from urine. Fecal recovery
accounted for approximately one half of the total administered
dose (Shah et al., 1981).
IV. HEALTH EFFECTS
Humans
In a controlled experiment, carbofuran was administered orally to
healthy males at two subjects per dose level. The subjects were
observed for 24 hours after dosing. Ho symptoms were observed at
0.05 mgAg (FMC, 1977). At 0.10 mg/kg, symptoms included headache
and, possible lightheadedness; at 0.25 mg/kg, symptoms of acetylchol-
inesterase depression were observed, including salivation, diaphoresis,
abdominal pain, drowsiness, dizziness, anxiety and vomiting. The
0.05 mgAg dose level was defined as the NOAEL in this study.
Several cases of adverse effects in applicators and fonnulators using
carbofuran have been reported (Tobin, 1970). Symptoms included mild
and reversible symptoms of acetylcholinesterase depression such as
malaise, hyperhydrosis, lightheadedness, nausea, blurring of vision,
hypersalivation and vomiting. Symptoms which may occur with more
severe poisoning include chest tightness, muscular twitching,
convulsions and coma.
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37
Carbofuran March 31, 1987
-5-
Animals
Short-term Exposure
The acute toxic effects, including lethality, resulting from exposure
to carbofuran are attributed to rapid inhibition of acetylcholinesterase
activity.
Acute oral LD5Q values in mammals have been reported as 2.0 mgAg in
mice (Fahmey et al., 1970) and 6.4 to 14.1 mgAg in rats (MRI,
1976). In dogs, 20% lethality was observed at a dosage level of
18.85 mgAg (MRI, 1976).
The dermal toxicity of carbofuran applied in organic solvent also is
relatively high with an LD50 of 14.7 mgAg reported for rabbits (MRI,
1976). For dry granule applications, however, 1.050 values were
higher than 10,000 mgAg*
Dose-related inhibition of cholinesterase activity in the blood,
liver and brain of pregnant rats and their fetuses was demonstrated
by Cambon et al. (1979). Carbofuran was administered at 0.05, 0.25
or 2.5 mgA9 on d*v 18 of gestation. In the high-dose group, toxic
signs appeared within five minutes; 8/32 dams died within 30 minutes;
and acetylcholinesterase activity was reduced in all maternal and
fetal tissues sampled one hour after dosing. At the lower dosage
levels, inhibition was found in some tissues at one hour. This study
defines a LOAEL of 0.05 mgAg for a single dose based on inhibition
of maternal and fetal blood acetylcholinesterase and maternal liver
acetylcholinesterase.
Long-term Exposure
Dietary administration of carbofuran to rats at 0.49 or 1.18 mgAg/day
for 180 days did not result in dose-related or clearly demonstrable
effects on liver enzymes (Rotaru et al., 1981).
There was no indication of cumulative or delayed adverse effects on
mortality, food consumption, reproduction or development of young in
two strains of wild mice fed carbofuran in their diets at 19.6 and
12.2 mgAg/day, respectively, for eight months (Wolfe and Escher, 1980)
The highest NOAEL that can be defined from this study is 19.6
In a one-year feeding study in beagle dogs which were exposed to dosage
levels of 0, 0.25, 0.50 and 12.5 mgAg/day, FMC (1983) reported no
biologically significant adverse effects on various biochemical,
hematological or clinical parameters at 0.25 or 0.50 mgfkg/day. At
12.5 mgAg/day, there was marked depression of plasma and erythrocyte
cholinesterase levels in both sexes, testicular degeneration and some
aspermia in males, and uterine hyperplasia and hydrometria in females.
A NOAEL of 0.50 mg/kg/day was identified for dogs based on the results
of this study.
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38
Carbofaran March 31, 1987
-6-
In a two-year study by FMC (1980a), rats were administered carbofuran
in the diet at dosage levels of 0,. 0.5, 1.0 or 5 mg/kg/day. No
adverse effects were observed on body weight, food consumption,
behavior, ophthalmoscopy, hematology, biochemistry, urinalysis or
histopathology. At the highest dosage, slight decreases in mean body
weight were observed in males; there also was an inhibition of plasma,
RBC and brain cholinesterase levels in both sexes. The NOAEL for
this study was determined to be 1.0 mgAg/day.
In a similar two-year study in mice, FMC (1980b) reported that dietary
administration of carbofuran at dosage levels of 3, 18.8 or 75 mg/kg/day
resulted in no observable adverse effects on food consumption, behavior,
hematology, biochemistry, urinalysis or histopathology. At the
highest dose, there was a temporary decrease in body weight. At the
two highest doses, reductions in brain cholinesterase levels were
observed. This study defines a NOAEL of 3 mgAg/day.
Reproductive Effects
In beagles fed carbofuran for a year at dosage levels of 0.25, 0.50
or 12.5 mg/kg/day, aspermia in males was observed at the two highest
dosage levels (FMC, 1983). The effect was not statistically signifi-
cant at the 0.5 mg/kg/dose. At the highest dosage level, testicular
degeneration was observed in males in addition to uterine hyperplasia
and hydrometria in females.
In a three-generation study in which rats were fed carbofuran at 1.0
or 5.0 mgA9/day* no adverse effects were observed on female or male
fertility, length of gestation, litter size or growth, or pup viability
(FMC, 1980c). At the high dose, however, the survival of the first
litter in all three generations was slightly lower by day 4 of lactation.
The NOAEL for reproductive effects was determined to be 1.0 mg/kg/day.
Developmental Effects
In rats fed carbofuran at dosage levels of 1.0, 2.9, 5.8, 7.7 or 9.7
•gAg/day (FMC, 1980d) or at 1, 3 or 8 mgAg/day (FMC, 1981a) on days
6 through 19 of gestation, there were no observable clinical signs of
toxicity or adverse effects on pup survival or visceral or skeletal
development. Maternal body weight gains were reduced at the 2.9 to
9.7 mgAg/day dosage levels in the first study and ir the 3 and 8
mgAg/day dosage groups in the second study. A NOAEL of 1.0 mgAg/day
was determined from these studies.
No adverse effects were observed on the 28 or 800 day survival rates
of mice whose mothers had been fed carbofuran at 0.01 or 0.50 mg/kg/day
throughout gestation (Barnett et al., 1980).
When rabbits were administered carbofuran at 0.12, 0.5 or 2.0 mg/kg/day
by gavage on days 6 to 18 of gestation, no terata were observed in
the offspring (FMC, 1981b). There also were no decreases in the
numbers of fetuses or litters or observable developmental or
genetic abnormalities. Dams at the highest dosage level experienced
a 20% reduction in weight gain during the treatment period.
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39
Carbofuran March 31, 1987
-7-
Mutagenicity
• In data from six authors (U.S. EPA, 1985a), Ames bacterial test
results have been negative except in one study, Moriya et al. (1983)
in which carbofuran applied at up to 10 mg/plate with rat liver 5-9
activation was mutagenic in Salmonella typhimurium strains TA98 and
TA1538.
0 Results of mutagenicity tests of carbofuran in each of several
other test organisms were negative except for CHO V79 cells in
which Wojciechowski et al. (1982) reported positive results at an
unspecified dosage of carbofuran without, but not with, rat liver
S-9 activation. '
Carcinogenicity
0 There was no evidence of carcinogenicity in a two-year dietary
study in which rats were administered carbofuran at dosage levels
of 0.5, 1.0 or 5 mg/kg/day (FMC, 1980a).
0 Similarly, there was no evidence of carcinogenicity in a two-year
dietary study in which mice were administered carbofuran at dosage
levels of 3, 18.8, or 75 mg/kg/day (FMC, 1980b).
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 noncarcihogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
where:
HA = (NOAEL or LOAEL) x (BW) = /L ( /L)
(UF) x ( L/day)
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
OF » uncertainty factor (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).
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Carbofuran 40 March 31, 1987
-8-
One-day Health Advisory
The study by FMC Corporation (1977) has been selected to serve as the
basis for calculation of the One-day health advisory (HA) for children. Hie
NOAEL observed in this study is 0.05 mg/kg, based on the absence of signs and
symptoms of acetylcholinesterase inhibition following oral, single dose
exposure to one of several levels of carbofuran in humans.
The One-day HA for the 10-kg child is calculated as follows:
One-day HA - (0.05 ing/kg/day) (10 kg) a 0.05 /L (50 /L)
(10) (1 L/day)
where:
0.05 mg/kg/day = NOAEL, based on absence of signs and symptoms of
acetylcholinesterase inhibition in humans exposed
to a single oral dose of carbofuran.
10 kg = assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from a human study.
1 L/day = assumed daily water consumption of a child.
Ten-day Health Advisory
No studies involving short-term exposure suitable for calculations of a
Ten-day HA were found in the literature. Because of the rapidly reversible
toxic effects of low doses of carbofuran and the absence of evidence of
cumulative toxicity, the One-day HA for the child can also serve as the
Ten-day HA (50 ug/L). This value is also identical to the Longer-term HA.
Longer-term Health Advisory
The one-year feeding study in dogs by FMC (1983) has been selected to
serve as the basis for the Longer-term Health Advisory. This study identified
0.50 mg/kg/day as the NOAEL, based on statistically but not biologically
significant plasma chulinesterase depression and testicular degeneration in
males. Other chronic studies involving rats (FMC, 1980a) and mice (FMC, 1980b)
defined NOAELs at higher levels (1.0 and 3.0 mg/kg/day, respectively). Because
of the nature of this chemical (a cholinesterase inhibitor), the acute human
study by FMC (1977) was also taken in consideration when computing this HA.
For a 10-kg child:
Longer-term HA = (0.50 mg/kg/day) (10 kg) = 0.05 mg/L (50 ug/L)
(100) (1 L/day)
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41
Carbofuran March 31, 1987
-9-
where:
0.5 ng/kg/day - NOAEL, based upon absence of acetylcholinesterase
inhibition and testicular degeneration in beagle dogs
exposed to carbofuran via the diet for one year.
10 kg * assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day - assumed daily water consumption of a child.
For a 70-kg adult:
Longer-term HA = (0.50 mg/kg/day) (70 kg) = 0.18 mg/1 (180 ug/L)
(100) (2 L/day) * y/
where:
0.5 mg/kg/day - NOAEL, based upon absence of acetylcholinesterase
inhibition and testicular degeneration in beagle dogs
exposed to carbofuran via the diet for one year.
70 kg = assumed body weight of an adult.
100 « uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day « assumed daily water consumption of an adult.
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 » 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
(DHEL) 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 classifed as a
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42
Carbofuran March 31, 1987
-10-
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 Lifetime Health Advisory for the 70-kg adult also has been
determined from the one-year study in dogs (FMC, 1983) as described above.
The Lifetime Health Advisory is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
where :
RfD - ' »g9ay , 0.005 mg/kg/day
0.5 mg/kg/day » NOAEL, based on absence of acetylcholinesterase
inhibition and testicular degeneration in beagle dogs
exposed to carbofuran via the diet for one year.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DHEL - (0.005 mgAg) (70 kg) = 0.18 /L (18Q /L)
(2 L/day)
Where:
0.005 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day - assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA - 0.18 mg/L x 20% = 0.036 mg/L (36 ug/L)
where:
0.018 mg/L - DWEL.
20% = assumed relative source contribution
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Carbofuran *•* March 31, 1987
-11-
Evaluation of Carcinogenic Potential
0 Data from studies conducted to evaluate the carcinogenic potential of
carbofuran do not identify any compound-related increases in tumor
incidences in either rats or nice. Therefore, quantification of
carcinogenic risk for lifetime exposures through drinking water would
be inappropriate.
0 The International Agency for Research on Cancer (IARC) has not
evaluated the carcinogenic potential of carbofuran.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), carbofuran is classified
carbofuran in Group E: Mo evidence of carcinogenic!ty in humans.
This category is used for agents that show no evidence of carcinogenic!ty
in at least two adequate animal tests in different species or in both
epidemiologic and animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The threshold limit value (TLV) on a time weighted average basis is
0.1 ng/ni3 (ACGIH, 1980).
0 NAS (1983) stated that insufficient information was available to
permit calculation of a suggested no-adverse-response level (SNARL)
or to assess the possibility of chronic exposure hazards. Most toxi-
cological data on carbofuran, however, are classified as Confidential
Business Information under the Federal Insecticide, Fungicide and
Rodenticide Act (FIFRA) and were not available to NAS for their
evaluation.
0 The U.S. EPA, Office of Pesticide Programs (OPP) has established
tolerances for carbofuran in or on raw agricultural commodities
(40 CFR 180.254). These tolerances are based on an Acceptable Daily
Intake (ADI) of 0.005 mg/kg/day. This ADI is also referred to as the
Reference Dose (RfD). It is also calculated as presented on page 10
of this HA.
0 WHO calculated an ADI of 0.01 mg/kg/day for carbofuran (Vettorazzi
and Van den Hurk, 1985).
VII. ANALYTICAL METHODS
0 Analysis of carbofuran is by a high performance liquid chromatographic
procedure used for the determination of N-methyl carbamoyloximes and
N-methylcarbamates in drinking water (U.S. EPA, 1984). In this
method, the water sample is filtered and a 400 uL aliquot is injected
into a reverse phase HPLC column. Separation of compounds is achieved
using gradient elution chromatography. After elution from the HPLC
column, the compounds are hydrolyzed with sodium hydroxide. The
methylamine formed during hydrolysis is reacted with o-phthalaladehyde
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44
Carbofuran March 31, 1987
-12-
(OPA) to form a fluorescent derivative which is detected using a
fluorescence detector. The method detection limit has been estimated
to be approximately 0.9 ug/L for carbofuran.
VIII. TREATMENT TECHHOLOGIES
0 Treatment techniques which may be effective in removing carbofuran
from drinking water include: adsorption on granular activated carbon
(GAC) or powdered activated carbon (PAC), reverse osmosis (RO) and
oxidation by ozone or ozone/ultraviolet. Only limited performance
data are available for carbofuran; however, the physical properties
and structure of the compound as well as information provided by the
manufacturer suggest that these treatment methods may be effective.
0 Carbofuran is expected to be amenable to activated carbon adsorption
due to its molecular configuration and water solubility. Troxler, et
al. (1980) reported a full-scale GAC plant efficiency on removing
carbofuran from wastewater as 99.9% from an initial concentration of
2250 mg/L and a carbon loading of 0.09 g carbofuran/g of carbon.
Operating parameters were as follows: carbon usage 207 lb/1000 gal of
treated water, and an empty bed contact time of 292 minutes. A
field study of in-home carbon adsorption units also showed GAC to be
effective (U.S. EPA, 1985b).
0 Reverse osmosis using polyamine membrane may be a feasible technology
for the removal of carbofuran from drinking water. Chian et al.
(1975) examined the use of RO in the rejection of several pesticides,
not including carbofuran. They reported that both polyethyleneamine
acetate and cross-linked polyethyleneamine membranes performed excellently.
However, the former was less effective on the more polar pesticides.
Because carbofuran exhibits some polarity, extrapolation from these data
leads to the hypothesis that the cross-linked membrane would be better
suited for use. When RO is used for the remocal of pesticides such as
carbofuran, attention must be given to the disposal of the reject stream
which may contain high concentrations of the chemical.
0 Ozone and/or ozone/ultraviolet oxidation may be feasible technology for
reducing concentrations of carbofuran in drinking water. Although no
data on carbofuran oxidation have been published Wilkinson et al.
(1978) reported on a bench scale study of ozone/ultraviolet reduction
of Baygon • (propoxur), another carbamate. It was reported that a
removal of 99.9% of the chemical was achieved in 30 minutes from
49 mg/L, using 20 mg/L O3 and OV light at 1.32 W/L. The authors
suggested that the process would be suitable for other pesticides as
well.
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45
Carbofuran March 31, 1987
-13-
IX. REFERENCES
ACGIH. 1980. American Conference of Govenmental Hygienists. Documentation
of the threshold limit values. 4th ed., pp.67-68.
Ahdaya, S. and F.E. Guthrie. 1982. Stomach absorption of intubated insecticides
in fasted mice. Toxicology 22:311-317.
Ahdaya, S.M., F.J. Monroe and F.E. Guthrie. 1981. Absorption and distribution
of intubated insecticides in fasted mice. Pestic. Biochem. Physiol.
16:38-46.
Barnett, J.B., J.M. Spyker-Cranmer, D.L. Avery and A.M. Hoberman. 1980.
Immunocompetence over the life span of mice exposed in utero to carbofuran
or diazinon: 1. Changes in serum immunoglobulin concentrations. J.
Environ. Pathol. Toxicol. 4:53-63.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.254. July 1, 1985.
pp. 299-300.
Cambon, C., D. Declume and R. Derache. 1979. Effect of the insecticical
carbamate derivatives (carbofuran, pirimicarb, aldicarb) on the activity
of acetylcholinesterase in tissues from pregnant rats and fetuses.
Toxicol. Appl. Pharmacol. 49:203-208.
Chian, E.S.K, H.N. Bruce and H.H.P. Fang. 1975. Removal of pesticides by
reverse osmosis. Environ. Science Technol. 9(1):
Cook, R.F. 1973. Carbofuran. In: Sharma, J. and G. Zweig, eds. Analytical
methods for pesticides and plant growth regulators. Vol. VII. New
York: Academic Press, pp. 187-210.
Dorough, H.W. 1968. Metabolism of Furadan (NIA-10242) in rats and houseflies.
J. Agr. Food Chem. 16:319-325.
Fahmey, M.A.H., T.R. Fukudo, R.O. Myer and R.B. March. 1970. The selective
toxicity of new N-phosphorothiocarbamate esters. J. Agr. Food Chem.
18:793-796.
FMC. 1977. FMC Corporation, Agricultural Chemical Group. Industrial
hygiene studies, final report. MRI Project No. 4?30-B. EPA Accession
No. 241303.
FMC. 1980a. FMC Corporation, Agricultural Chemical Group. 1980a. Two-year
dietary toxicity and carcinogenicity study in rats. Carbofuran Technical
Report No. 130.51. EPA Accession No. 244491.
FMC. 1980b. FMC Corporation, Agricultural Chemical Group. 1980b. Two-year
dietary toxicity and carcinogenicity study in mice. Carbofuran Technical
Report No. Act 150.52. EPA Accession No. 244489.
FMC. 1980c. FMC Corporation, Agricultural Chemical Group. 1980c. Three gener-
ation reproduction study in rats. Carbofuran Technical Report No. Act
131.53. EPA Accession No. 244490.
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'16
Carbofuran March 31, 1987
-14-
FMC. 1980d. FMC Corporation, Agricultural Chemical Group. 1980d. Pilot tera-
tology study in the rat with carbofuran in the diet. Study No. FMC A-80
443/IRDC 167-116. EPA Accession No. 244389.
FMC. 1981a. FMC Corporation, Agricultural Chemical Group. 1981a. Teratology
and postnatal study in the rat with carbofuran dietary administration.
Study No. FMC A80-444/IRDC 167-154. EPA Accession No. 244388.
FMC. 1981b. FMC Corporation, Agricultural Chemical Group. 1981b. Teratology
study in the rabbit with carbofuran. Study No. FMC A80-452/IRDC 167-156.
EPA Accession No. 1245268.
FMC. 1983. FMC Corporation, Agricultural Chemical Group. 1983. One-year
chronic oral toxicity study in beagle dogs with carbofuran. Study No.
FMC A81-605/ ToxiGenics 410-0715. EPA Accession No. 250740-250744.
Kuhr, R.J., and H.W. Dorough. 1976. Carbamate insecticides: Chemistry,
biochemistry, and toxicology. Chemical Rubber Company Press, Inc.,
Cleveland, OH.
Metcalf R.L., T.R. Fukuto, C. Collins et al. 1968. Metabolism of 2,2-
dimethyl-2,3-dihydrobenzofuranyl-7-N-methylcarbamate (Furadan) in plants,
insects and mammals. J. Agr. Food Chem. 16:300-311.
MRI. 1976. Midwest Research Institute. Substitute chemical program: initial
scientific and minieconomic review of carbofuran. Washington, D.C.:
U.S. Environmental Protection Agency. Contract No. EPA 68-01-2448.
EPA 54/1-76-009.
Moriya, M., T. Ohta, K. Watanabe, T. Miyazawa, K. Kato and Y. Shirasu. 1983.
Further mutagenicity studies on pesticides in bacterial reversion assay
systems. Mutat. Res. 116:185-216.
NAS. 1983. National Academy of Sciences. Drinking water and health. Volume 5.
Safe Drinking Water Committee. National Academy Press. Washington, DC
pp. 12-15.
Rotaru G., S. Constantinescu, G. Filipescu and E. Ratea. 1981. Experimental
research on chronic poisoning by carbofuran. Med. Lav. 5: 399-403.
Shah, P.V., R.J. Monroe and F.E. Guthrie. 1981. Comparative rates of dermal
ptnetration of insecticides in mice. Toxicol. Appl. Pharmacol.
59:414-423.
Tobin, J.S. 1970. Carbofuran: a new carbamate insecticide. J. Occup. Med.
12:16-19.
Troxler, W.L., C.S. Parmele and D.A. Barton. 1980. Survey of industrial
applications of aqueous-phase activated carbon adsorption for controls
of pollutants from manufacture of organic compounds. Prepared by Hydro-
science for U.S. EPA.
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Carbofuran March 31, 1987
-15-
Tudcer, R.K. and D.G. Crabtree. 1970. Handbook of toxicity of pesticides
to wildlife. U.S. Bureau Sport Fish. Wildl. Resour. Publ. 84. 131 pp.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of pesti-
cides in drinking water, food, and air. Office of Drinking Water.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Method 531. Meas-
urement of N-methyl carbamoyloximes and N-methylcarbarates in drinking
water by direct aqueous injection HPLC with post column derivatization.
Environental Monitoring and Support laboratory, Cincinnati, Ohio 45268.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Draft health effects
criteria document for carbofuran. Criteria and Standards Division,
Office of Drinking Water.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Draft Technologies
and costs for the removal of synthetic organic chemicals from potable
water supplies. Office of Drinking Water, Science and Technology Branch.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51 (185):33992-34003.
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance index for
pesticides. Bureau of Foods.
Vettorazzi, G. and G.W. Van den Hurk. 1985. Pesticides Reference Index, JMPR
1961-1984, p. 10.
Wilkinson, R.R., G.L Kelso and F.C. Hopkins. 1978. State-of-the-art report
pesticide disposal research. U.S. EPA, MERL, Cincinnati, Ohio.
EPA 600/2-78-163.
Windholz, M., ed. 1983. The Merck Index. An encyclopedia of chemicals and
drugs. 10th ed. Rahway, NJ: Merck & Col, Inc., p. 250.
Wojciechcwski, J.P., P. Kaur and P.S. Sabharwal. 1982. Induction of ouabain
resistance in V-79 cells by four carbamate pesticides. Environ. Res.
29:148-53.
Wolfe, J.L., and R.J. Esher. 1980. Toxicity of carbofuran and lindane to the
old field mouse (Peromyscus polionotus) and the cotton mouse (P. gossypinus),
Bull. Environ. Contain. Toxicol. 24:894-902.
*u, C.C., G.M. Booth, D.J. Hansen and J.R. Larsen. 1974. Fate of carbofuran
in a model ecosystem. J. Agr. Food Chem. 22:431-434.
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48
March 31, 1987
CHLORDANE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. IMTRODUCTION
Die 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 cot 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. The chemical 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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Chlordana
March 31, 1987
-2-
This Health.Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for Heptachlor,
Heptachlor Epoxlde and Chlordana (U.S. EPA, 1985a). The HA and CD fornate
are siailar for easy reference. Individuals desiring further information on
the toxicological data base or rationale for risk characterization should
consult the CD. Ihe CD is available for review at each BPA Regional Office
of Drinking Water counterpart (e.g.. Water Supply Branch or Drinking Water
Branch), or for a fee fro* the National Technical Information Service,
D.S. Department of Commerce, 5285 Port Royal Rd., Springfield, VA 22161,
PB I 86-117991/AS. The toll-free number is (800) 336-4700* in the Washington,
D.C. areai (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 57-74-9
Structural Formula
Synon;
Dichlorodenei 1, 2,4, 5,6, 7, 8, B-octachlor-2, 3,3a, 4,7,7a-hexahydro-
4, 7-methano-1H-indenei 1,2,4,5,6,7,8, 8-octachloro-4, 7-methano-
3a,4,7,7a-tetrahydroindane; Octachlor*i Velsicol 1068*j Toxichlor*;
Dowclor*
Uses
• Broad spectrum insecticide currently used for termite control.
Properties (U.S. EPA, 1985a)
Chemical Formula
Molecular Weight
Physical State (room temp.)
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
409.76
vi-coua amber liquid
106-107*C (cis isomer)
104-105*C (trans isomer)
1 z 10-5 an Eg at 25*C
9 ug/L at 25»C (tech. grade)
5.16
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50
Chlordane March 31, 1987
-3-
Occurrence
• Chlordane has a large production volume; EPA estimated that more
than 10 million Ibs were produced for this purpose in 1980. Chlordane
is generally applied to the soil by subsurface injection.
• Chlordane is degraded poorly in the environment. Chlordane does
photodegrade but since it is applied by soil injection this is not a
significant removal mechanism. Chlordane is hydrolysed poorly and
does not undergo significant biodegradation. Chlordane is reported
to have a half life in soil of 4 years. Residues in soils may persist
for 14 years. Once in the ground, Chlordane rapidly binds onto soils
and migrates very slowly. Chlordane has the potential for bio-
accumulation.
• Chlordane has been reported to occur in both ground and surface water
at low levels, 0.01 to 0.001 ug/L. The highest levels have been
reported for Hawaii which uses large amounts of chlordane. Other
data have been reported by States and other sources. Based upon
chlordane's use as a soil injected insecticide and its persistence, it
is believed to have the potential to contaminate ground water,
particularly when it is applied over or near existing wells. Chlor-
dane has been found in low levels in food and air. The current
information is insufficient to indicate which is the major route of
exposure for chlordane.
0 Harrington et al. (1978) reported that a section of the public water
system of Chattanooga, Tennessee, supporting 105 people in 42 houses,
was contaminated with chlordane on March 24, 1976. Chlordane concen-
trations in the tap water of affected houses ranged from less than
0.1 to 92,500 ppb. In 23 houses, the concentration exceeded 100 ppb;
11 of these had concentrations greater than 1,000 ppb.
III. PHARMACOKINETICS
Absorption
0 Evidence that chlordane is absorbed from the gastrointestinal tract
is derived from reports of systemic toxicity and excretion data
following oral exposure to the compound.
0 Data reported in two case studies of children (Aldrich and Holmes,
1969; Curley and Garrettson, 1969) indicate that ingested chlordane
was absorbed into the bloodstream. In one of these children, a blood
level of 2.71 mg/L was measured less than three hours after ingestion
(Curley and Garrettson, 1969).
• Urinary excretion data indicate at least 2 to 8.5* absorption of
chlordane by rats and 33* by rabbits when orally administered (Barnett
and Dorough, 1974).
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51
Chlordane March 31, 1987
-4-
Distribution
Chlordane and its major metabolite, oxychlordane, appear to be distrib-
uted preferentially to and stored in adipose tissue.
In the case of a 20-month-old boy who drank an unknown quantity of
Chlordane, total residues in adipose tissue were calculated to be
5*9, 36 and 65 mg at 3 hours, one day and eight days after ingestion,
respectively (Curley and Garrettson, 1969).
Chlordane and oxychlordane have been measured at various levels in
adipose tissue analyzed during autopsy. Biros and Enos (1973) reported
that 21 of 27 samples from the general population were positive at a
mean concentration of 0.14 ppm.
In male rats dosed with chlordane for 56 days in the diet (5 »gA9 ***)»
residues were distributed in muscle, brain, kidney and liver (<1 ppm
each) but at 14.73 ppm in fat. In females dosed at 25 mgAg in the
diet, residues of the trans-isomer still were detected in all tissues
examined 56 days after treatment (Barnett and Dorough, 1974).
Ambrose et al. (1953) found that the perirenal fat of male rats
contained 43, 41 and 81 ppm of chlordane residues following feeding
of a diet containing chlordane for 5, 148 and 407 days, respectively.
The fat of female rats contained approximately twice the values for
males.
In rabbits orally administered daily trans-chlordane dose of 14.3
mgA9 for ten weeks, two weeks after treatment, low levels were
detected in kidney, liver, heart, lung, spleen, testes and brain
(<1% each) (Poonawalla and Korte, 1971). Higher levels were fou«d in
adipose tissue and muscle (about 4.1% and 5.7%, respectively).
Metabolism
Oxychlordane is presumed to be the major metabolite of chlordane. In
tissue distribution studies, levels of oxychlordane generally were
comparable to, or higher than, those of chlordane itself (Polen et al.,
1971; Poonawalla and Korte, 1971; Street and Blau, 1972; Barnett and
Dorough, 1974; Balba and Saha, 1978).
Street and Blau (1972) have proposed a metabolic pathway for chlordane
based on in vitro studies with rat liver homogenates. Chlordane is
dehydrogenated to 1,2-dichlorochlordene with subsequent epoxidation
to oxychlordane. The trans isomer is converted to oxychlordane at a
sevenfold greater rate than is the cis isoaer.
Data presented by Tashiro and Matsumura (1978) indicate that very
little interspecies difference was f—*nd between rat and human
during in vitro metabolism of cis- and trans-chlordane.
Some biotransformation of chlordane takes place in the gastrointestinal
tract since various chlordane metabolites were found in the feces of
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52
Chlordane March 31, 1987
-5-
rats and a rabbit after oral administration of chlordane (Barnett and
Dorough, 1974). In rats orally administered oxychlordane, however,
only unchanged oxychlordane was excreted in the feces.
The urine of rats administered chlordane by diet included the same
metabolites as found in the feces, plus oxychlordane (Barnett and
Dorough, 1974). The urine of a rabbit contained a higher percentage
of the conjugated hydroxylated metabolites than did the urine of rats.
Excretion
Excretion of orally administered chlordane is relatively slow (days
to weeks) and occurs via feces and urine.
Clearance of ingested chlordane from serum also is relatively slow
with a biological half-life of 88 days estimated in the case of a four-
year-old girl (Aldrich and Holmes, 1969) and a serum half-life of 21
days in the case of a 20-month-old boy (Curley and Garrettson, 1969).
In the urine of a four-year-old girl, levels of chlordane declined
rapidly during the first three days (1.93 to 0.05 ppm) but rose to
0.13 ppm by 35 days, presumably due to the release of stored chlordane
(Aldrich and Holmes, 1969). Fecal levels also declined rapidly
during the first three days and no chlordane was detected in the
feces one or two months later.
In a survey conducted by Strassman and Kutz (1977) in Arkansas and
Mississippi in 1973 and 1974, 54.4% of human milk samples contained
oxychlordane at trace levels or higher. The mean level of quantifiable
residues (45.6% of samples) was 0.012 ppm. Thus, lactation is a route
of excretion in females.
In rats orally administered chlordane, seven-day elimination rates of
86% and 66% (for cis- and trans-isomers, respectively) and 90% have
been reported (Tashiro and Matsumura, 1977; Barnett and Dorough, 1974).
In the latter study, urinary excretion was reported as 2% for males
and 6% for females. After 56 days of dietary administration to rats,
fecal elimination ranged from 70% to 80% with increasing dietary
levels (1 to 25 mg/kg diet) (Barnett and Dorough, 1974).
In rabbits orally administered chlordane, reported levels of urinary
excretion of chlordane and/or its metabolites ranged from 18% to 49%
of the administered dose and fecal levels ranged from 22.7% to 49%
(Stohlman et al., 1950; Poonawalla and Korte, 1971; Barnett and
Dorough, 1974; Nye and Dorough, 1976; Balba and Saha, 1978). Higher
levels of urinary excretion generally were associated with longer
periods of chlordane administration.
Chlordane was not detected in the expired air of rats intratracheally
administered chlordane (Nye and Dorough, 1976).
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Chlordane
53
March 31, 1987
-6-
IV. HEALTH EFFECTS
Humans
In clinical case studies of acute or chronic exposure to chlordane,
the effects most frequently observed are central nervous system (CHS)
effects and blood dyscrasias (U.S. EPA, 1985a). Heuroblastoma has
been reported in association with chlordane/heptachlor exposure.
Reported CMS effects include irritability, salivation, labored
respiration, muscle tremors, convulsions, deep depression and death.
Ingestion of chlordane resulted in similar CMS effects in the cases
of a 32-year-old woman who ingested 104 mg/kg, an 18-year-old woman
who ingested about 10 mg/kg, a 15-month-old infant who ingested about
11.1 mgAg and a four-year-old child who ingested about 0.15 mg/kg
(Lensky and Evans, 1952; Dadey and Rammer, 1953; Derbes et al.,
1955; Aldrich and Holmes, 1969).
Blood dyscrasias have been associated with dermal or inhalational
exposure to chlordane at unspecified dose levels. These reports have
included cases of aplastic anemia (Klemmer et al., 1977; Infante
et al., 1978), refractory megaloblastic anemia (Furie and Trubowitz,
1976), acute stem cell leukemia, acute lymphobiastic leukemia and
acute nyelomonocytic leukemia (Infante et al., 1978).
A total of 14 cases of neuroblastoma have been reported in children
with pre- and/or post-natal exposure to chlordane and heptachlor
'T^fante et al., 1978). Exposure was via inhalation and/or dermal
contact but levels could not be estimated.
In an epidemologic study of white males employed for more than three
months in the production of chlordane and heptachlor, Wang and McMahon
(1979a) reported a significant increase in cerebrovascular disease.
Animals
Short-term Exposure
Acute oral WSQ values for chlordane vary with the purity of the test
compound. In rats, reported values range from 83 mgAg for pure cis-
chlordane (Podowski -.t al., 1979) to 560 mgAg for chlordane of
unspecified purity (Ambrose et al., 1953). Values for technical
grade chlordane fall within an intermediate range.
Symptoms of acute intoxication include CNS stimulation, as evidenced
by irritability, tremors and convulsions (Stohlman et al., 1950;
Boyd and Taylor, 1969; Hyde and Falkenberg. 1976). Boyd and Taylor
(1969) described a wide range of CNS disturbances, including phonation,
piloerection, tremors and convulsions alternating with lethargy,
diarrhea and food and water refusal. Necropsy of rats revealed
vascular congestion, nephritis, hepatitis and decreased organ weight.
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Chlordane 54 March 31, 1987
-7-
• Chlordane was »ore toxic when administered orally to rats and rabbits
in Tween-20 than in olive oil, as evidenced by a greater incidence of
•ortality occurring at shorter periods following treatment (Stohlman
et al., 1950).
• In a 42-day study by NCI (1977), maximum tolerated doses of Chlordane
were established at 400 and 800 »gAg diet for female and male rats,
respectively, and at 80 and 40 mg/kg diet for female and male mice,
respectively.
Long-term Exposure
• In the genital tissue of male rats fed technical grade Chlordane at
a dosage level of 19.5 mgAg/day for 90 days, Shain et al. (1977)
demonstrated increased nuclear androgen receptor site content and
decreased RNA, DNA and ventral prostate protein content.
• In rats administered Chlordane by gavage at dosage levels of 6.25 to
200 mgAg/day for 15 days, Ambrose et al. (1953a) observed slight
changes in the livers (intracytoplasmic bodies) of animals at all
dose levels and severe effects, including death, at 50 mg/kg/day and
above.
0 In a two-year dietary study in rats. Ingle (1952) demonstrated dose-
related adverse effects ranging from minor liver damage at 10 mg/kg
diet to a high incidence of mortality, eye and nose hemorrhaging and
severe histopathologic damage to the liver, kidney, heart, lung,
adrenal, myocardium and spleen at 300 mg/kg diet. No adverse effects
were noted at 5 mg/kg diet.
0 In an NCI (1977) bioassay, rats treated with chlordane at 120.8 to
407.0 mgAg diet for 80 weeks had increased mortality rates, tremors,
clinical signs of toxicity and reduced mean body weight.
0 In studies designed to assess the carcinogenicity of chlordane in
mice, such effects as increased mean liver weight, decreased mean
body weight and increased mortality were observed at dietary levels
ranging from 25 to 63.8 mgA9 diet (IRDC, 1973; NCI, 1977). At
5 mgAg diet in the IRDC (1973) study, the only observed effects were
increased mean liver weights and hepatocytomegaly in females.
0 Based on a two-year feeding study in dogs by Nazeter 11967, as cited
in Vettorazzi, 1975), a NOAEL of 0.075 mgAg/day (3 mgAg diet) was
established. Ingestion of chlordane at 15 or 30 mgAg diet resulted
in increased liver weight and histologic changes.
• F-344 rats (80/sex/group) were fed technical chlordane at dietary
levels of 0, 1, 5 or 25 ppm for 130 weeks (Yonemura et al., 1983b).
Body weight, food consumption and water intake were monitored at
regular intervals. Clinical laboratory studies were performed and
organ weights were measured on eight animals/sex/group at weeks 26
and 52, and on all survivors at week 130. Gross and microscopic
pathology were performed on all tissues. Daily dose levels of 0.045,
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55
Chlordane March 31, 1987
-8-
0.229 and 1.175 mgAg for the 1, 5 and 25 ppir treatment groups,
respectively, were calculated from food consumption and body weight
data. Ho effects were observed for hematology, clinical chemistry
and urinalysis endpoints, and no treatment-related effects were
reported for body weight and mortality. Hepatocellular necrosis was
observed in 3, 13, 11 and 27 males (64/group) in the 0, 1, 5 and 25
ppn groups, respectively. The increased incidence was statistically
significant for all treatment groups. Liver adenomas were found in
the high-dose males. The only significant effect in females was
hepatocellular swelling in the 25 ppm group.
• Increased liver-to-body weight ratios were reported for male and
female mice fed Chlordane for 2 years at 0.76 ppm (0.09 ngfkg/day),
the lowest dose administered (Yonemura et al., 1983a). Liver necrosis
was observed at 0.43 and 1.1 mgAg/day for males only.
Reproductive Effects
0 Fertility was reduced significantly (by about 50%) in female mice
intraperitoneally injected with Chlordane at 25 mg/k? once a week for
3 weeks (Welch et al., 1971).
Developmental Effects
0 Ingle (1952) observed no fetotoxic or teratogenic effects in rats
born to dams fed Chlordane at 5 to 300 mg/kg diet in a two-year study.
Pups nursed by dams ingesting Chlordane at 150 and 300 mg/kg diet,
however, developed dose-related symptoms of toxicity.
Mutagenicity
0 Negative results for mutagenicity of Chlordane were reported for nine
strains of Salmonella typhimurium and two strains of Bacillus subtilis
for reverse mutation with or without metabolic activation (Probst,
et al., 1981; Gentile et al., 1982), in rat, mouse and hamster primary
hepatocyte cultures for unscheduled DHA synthesis (Maslansky and
Williams, 1981; Probst et al., 1981) and in mice for the dominant
lethal assay (Arnold et al., 1977). For details please refer to
support document (U.S. EPA, 1985a).
• Positive results were obtained in Sactharoayces cerevisiae for mitotic
gene conversion with, but not without, metabolic activation (Blevins
and Sholes, 1978) and in maize for reverse mutation (Gentile et al.,
1982).
Carci nogeni ci ty
• The major target organ for carcinogenic effects in mice is the liver.
• A re-evaluation of an IRDC (1973) study by Epstein (1976) indicates
that Chlordane at dietary levels of 25 and 50 mgAg diet for 18 months
resulted in very high incidences of hepatic carcinoma.
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Chlordane March 31, 1987
-9-
Becker and Sell (1979) found that C57BL/6N nice, a strain that does
not develop spontaneous liver lesions, developed primary hepatocellular
carcinomas during chronic exposure to a chlordaneiheptachlor (90:10)
mixture at 25 and 50 mgAg diet levels.
MCI (1977) also found a highly significant dose-related increase in
the incidence of hepatocellular carcinoma in mice exposed to chronic
dietary levels of chlordane ranging from 29.9-63.8 mg/kg diet.
Chlordane was not a hepatic carcinogen in Osborne-Mendel rats (NCI,
1977).
P-344 rats (80/sex/group) were fed technical chlordane at dietary
levels of 0, 1, 5 or 25 ppm for 130 weeks (Yonemura et al., 1983b).
Body weight, food consumption and water intake were monitored at
regular intervals. Clinical laboratory studies were performed and
organ weights were measured on eight animals/sex/group at weeks 26
and 52, and on all survivors at week 130. Gross and microscopic
pathology were performed on all tissues. Daily dose levels of 0.045,
0.229 and 1.175 mgAg for the 1,5 and 25 ppm treatment groups,
respectively, were calculated from food consumption and body weight
data. No effects were observed for hematology, clinical chemistry
and urinalysis end points, and no treatment-related effects were
reported for body weight and mortality. Hepatocellular necrosis was
observed in 3, 13, 11 and 27 males (64/group) in the 0, 1, 5 and 25
ppm grdaps, respectively. The increased incidence was statistically
significant for all treatment groups. Liver adenomas were found in
the high-dose males. The only significant effect in females was
hepatocellular swelling in the 25 ppm group.
Increased liver-to-body weight ratios were reported for male and
female mice fed chlordane for 2 years at 0.76 ppm (0.09 mg/kg/day),
the lowest dose administered (Yonemura et al., 1983). Liver necrosis
was observed at 0.43 and 1.1 mg/kg/day for males only.
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 tollowing formula:
HA = (NOAEL or LOAEL) x (BW) = /L ( /L)
(OF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mgA? bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
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Chlordane >' March 31, 1987
-10-
DF « uncertainty factor (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
Satisfactory dose-response data are not available from which a One-day
health advisory (HA) can be derived. Therefore, it is recommended that the
Ten-day HA of 0.06 »g/L be used as a conservative estimate for a One-day
exposure.
Ten-day Health Advisory
A Ten-day HA for chlordane is calculated from the Ambrose et al. (1953)
study in rats. .The toxic effects resulting from daily gastric intubation of
doses of 6.25, 12.5, 25.0, 50.0, 100.0 or 200 mo/kg chlordane in rats for 15
days were histological changes in the liver of the treated animals at all
dose levels and central nervous system effects at higher dose levels. Only
minimal histopathological changes characterized by the presence of abnormal
intracytoplasmic bodies of various diameters were evident at the lowest dose
level (6.25 mg/kg). It is recognized that histologic changes such as intra-
cytoplasmic inclusion bodies in the liver of animals at various dosage levels
may not-represent a true adverse effect; however, it does reflect a minimum
effect of chlordane in animals. Ambrose et al. (1953) also pointed out CMS
effects in animals followed by death of a few animals at higher doses of 50,
100 or 200 mg/kg/day of chlordane. The study of Den Tonkelaar and Van Esch
(1974) also provides the dose response for technical chlordane administered
in the diet for 14 days to groups of six male Hi star rats. Significantly
elevated activities of aniline hydroxylase and aminopyrine demethylase occurred
at a chlordane concentration of 10 mgAg; increases in hexabarbital oxidase
activity occurred at 20 mg/kg. A slight increase was observed at the 5 mgAg
dose level. The results of this study and those of the Ambrose et al. (1953)
study suggest that the effect level of chlordane is between 5-6 ng/kg/day.
Therefore, using 6.25 mg/kg as a Lowest-Observed-Adverse-Effect-Level (LOAEL),
the Ten-day HA is derived as follows:
Ten-day HA - (6.25 mgAg/day) (10 kg) = 0.0625 mg/L (63 ug/L)
(1,000)) (1 L/day)
Where:
6.25 mgAg/day = LOAEL based on study by Ambrose et al. (1953).
10 kg « assumed body weight of a 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.
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Chlordane 58 March 31, 19B7a
-11-
Longer-term Health AdviBory
There are insufficient lexicological data available to calculate a Longer-
tern HA for chlordane. It is recommended that the DWEL adjusted for a 10-kg
child (0.05 ug/L) be used as a conservative estimate for a longer-term exposure.
The National Research Council Report (KRC, 1982), "An Assessment of the
Health Risks of Seven Pesticides Used for Termite Control" was considered for
the derivation of a Longer-term HA for chlordane. However, the review of the
limited human studies with long-term exposure did not reveal any consistent
or significant detrimental effect that might be considered for health advisory
level for chlordane. Details of these human studies are given below.
Princi and Spurbeck (1951) evaluated 34 persons engaged in the manufac-
ture of insecticides, including chlordane (exposed through skin contact and
inhalation for 11-36 months). Physical examinations, chest xrays, urinary
dilution and concentration tests, routine urinalysis, hemoglobin measurements,
sedimentation rate, and urinary porphyrin determinations failed to suggest
any abnormalities in the men. The authors concluded that no adverse effects
were detected in men working in a plant with air concentrations of chlorinated
hydrocarbons as high as 10 mg/m^. Authors did not specify that exposure was
exclusive to chlordane and, therefore, this study was considered inappropriate
for a longer-term health advisory for chlordane.
Alvarez and Hyman (1953) reported a clinical and laboratory study of 24
men 21-49 years old who were exposed to chlordane for 2 months to 5 years
while working in a plant where it was manufactured. Each man was given a
complete examination, including blood chemistry and urine studies. None of
the 24 men had evidence of abnormalities in liver, kidneys, skin, nervous
system and blood-forming organs. However, the authors had observed in seven
men slight fibrotic changes in the apices of the lungs; one person with a
diabetic condition and two more with hypertension in chlordane-exposed workers.
These observations (even though not attributed to chlordane) and United
numbers of subjects in this study did not justify its consideration for a
Longer-term health advisory level.
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
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59
Chlordane March 31, 1987
-12-
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 classifed 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 study by Wazeter (1967, as reported by Vettorazzi, 1975) was considered
the most appropriate to derive the DWEL. However, the results of the recent
chronic rat dietary study by Yonemura et al. (1983) are available for the
derivation of the DWEL. In this study, F344 rats were fed technical chlordane
at dietary levels of 0, 1, 5 or 25 ppm for 130 weeks. Clinical laboratory
studies were performed and organ weights were measured on eight animals/sex/
group at weeks 26 and 52, and on all survivors at week 130. Gross and micro-
scopic pathology were performed on all tissues. Daily dose levels of 0.045,
0.229 and 1.175 mgAg for the 1, 5 and 25 ppm treatment groups, respectively,
were calculated from food consumption and body weight data. No effects were
observed for hematology, clinical chemistry and urinalysis endpoints, and no
treatment-related effects were reported for body weight and mortality.
Hepatocellular necrosis was observed in 3, 13, 11 and 27 males (64/group) in
the 0, 1, 5 and 25 ppm groups ,. respectively . The increased incidence was
statistically significant for all treatment groups. Liver adenomas were
found in the high-dose males. The only significant effect in females was
hepatocellular swelling in the 25 ppm group. The LOAEL of 1 ppm diet (0.045
•gAg/day) was identified based on liver necrosis in male rats. Using this
LQAEL, the DWEL is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD » (°'04^T">0t0ffiday)- * 0.000045 mgAg/day (0.05 ugAg/day)
where:
0.045 mgAg/day « LOAEL based on liver necrosis in male rats.
1,000 « uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Do termination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.05 ugAgyday) (70 kg) . 0.0017 mg/L {2 ug/L)
(2 L/day)
where:
0.05 ugAg/day « RfD.
70 kg - assumed body weight of an adult.
2 L/day * assumed daily water consumption of an adult.
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Chlordane March 31, 1987
-13-
Chlordane is classified in Group B: Probable human carcinogen. The
estimated excess cancer risk associated with lifetime exposure to drinking
water containing chlordane at 2 ug/L is approximately 1 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
• The U.S. EPA (1987) derived a human carcinogenic potency factor,
qj*t of 1.3 (ing/kg/day)~1 for chlordane. This derivation was based
on the geometric mean of four potency estimates which were based on
the incidence of hepatocellular carcinoma in male and female CD-1
mice (IRDC, 1973) and male and female B6C3F1 mice (NCI, 1977). This
estimate supersedes the potency of 1.61 (mg/kg/day)-1 previously
calculated by the U.S. EPA (1980).
0 The concentrations in drinking water corresponding to increased
lifetime risk levels of 10"4, 10~5 and 10~6 for a 70 kg human who
consumes 2 L/day are calculated to be 2.7, 0.27 and 0.027 ug/L,
respectively.
e Cancer risk estimates (95% upper limit) with other models are presented
for comparison with that derived with the multistage. For example,
one excess cancer per 1,000,000 (10~&) is associated with exposure to
chlordane in drinking water at levels of 50 ug/L (probit), 2 ug/L
(logit) and 0.03 ug/L (Weibull).
0 IARC (1979) has classified chlordane in Category 3: Inadequate
evidence of carcinogenicity.
• Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), chlordane is 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 AliD STANDARDS
• The Federal Water Pollution Control Administration (1968) set a permiss-
ible surface water criterion for public water supplies at 0.003 mg/L
for chlordane. The criterion for fish and other aquatic life based
on an LCso of 0.002 mg/L for chlordane would be very low; therefore, it
was recommended that this compound not be used near a marine environment.
The Water Quality Criterion for farmstead use was 0.003 mg/L for
chlordane.
0 In 1980, EPA estimated a range of excess cancer risks for lifetime
exposure to chlordane when developing ambient water quality criteria
(U.S. EPA, 1980). This range was 4.6 ng/L» 0.46 ngA and 0.046 ng/L,
respectively, for risks of 10-5, 10-6 and lO'7, assuming consumption
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61
Chlordane March 31, 1987
-14-
of 2 liters of water and 6.5 grams of contaminated fish per day by a
70-kg adult.
• FAO/WHO (1978) recommended a maximum acceptable daily intake (ADI)
value of 1 mg/kg bw for chlordane.
• The MAS—National Research Council (NRC, 1982) has recommended an
interim guideline of 5 ug/m3 for airborne chlordane in military
housing.
• WHO (1984) has recommended a drinking water guideline of 0.3 mg/1 for
chlordane.
VII. ANALYTICAL METHODS
4 Determination of chlordane is by a liquid-liquid extraction gas chrora-
tographic procedure (U.S. EPA, 1978; Standard Methods, 1985). Spe-
cifically, the procedure involves the use of 15% methylene chloride in
hexane for sample extraction, followed by drying with anhydrous
sodium sulfate, concentration of the extract and identification by
gas chromatography. Detection and measurement is accomplished by
electron capture, microcoulometric or electrolytic conductivity gas
chroma tography. Identification nay be corroborated through the use
of two unlike columns or by gas chromatography-mass spectroscopy
(GC-MS). The method sensitivity is 0.001 to 0.010 ug/L for single
component pesticides and 0.050 to 1.0 ug/L for multiple component
pesticides when analyzing a 1-liter sample with the electron capture
detector.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies which are capable of removing chlordane from
drinking water include adsorption by granular activated carbon (GAC)
and powdered activated carbon (PAC) and aeration. The only treatment
system for which performance data are available is carbon adsorption.
Further studies are required to determine the effectiveness of air
stripping systems.
0 Dobbs and Cohen (19
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62
Chlordane March 31, 1987
-15-
an emulsion containing 5 mg chlordane/L were passed through 12 gm of
activated carbon. Additional operating parameters, such as carbon
emulsifying agent and contact time were not reported.
The AC system in U.S. EPA's Hazardous Materials Spills Treatment
Trailer was used to treat 104,000 gal of pesticide-contaminated water
containing chlordane (U.S. EPA, 1985b). Water analysis showed
13 ug/L of chlordane in the contaminated water. 97.3% chlordane
removal was achieved at a contact time of 17 minutes.
The Henry's Law Constant is a good predictive tool for forecasting the
relative amenability of any chemical to treatment by air stripping.
McCarty et al. (1979) estimated that a Henry's Law Constant of
1 x 10-3 atm-m3/Bole is the cutoff point below which treatment by
aeration would not be practical. Based on reported solubility data
of 9 ug/L at 20°C (for the gamma isomer) and a vapor pressure of
1 x 10-5 nun Hg at 20 °C, Edwards estimated a Henry's Law Constant for
chlordane of 6 x 10~4 atm-m3/">ole (U.S. EPA, 1984b) . This suggests
that chlordane is not amenable to aeration. These differences
indicate that further investigations are required to determine the
actual performance data of air stripping treatment in the removal of
chlordane .
Treatment technologies for the removal of chlordane from drinking
water have not been extensively evaluated (except on an experimental
level). Whichever individual or combination of technologies is
ultimately selected for chlordane reduction must be based on a case-
by-case technical evaluation and an assessment of the economics
involved.
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Chlordane March 31, 1987
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IX. REFERENCES
Aldrich, F.D., and J.H. Holmes. 1969. Acute chlordane intoxication in a child.
Arch. Environ. Health. 19:129.
Alvarez, H.C., and S. Hyman. 1953. Absence of toxic manifestations in workers
exposed to chlordane. A.N.A. Arch. Ind. Hyg. Occup. Med. 8:480-483.
Ambrose, A.M., H.E. Christensen, D.J. Robbins and L.J. Rather. 1953. Toxi-
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7:797-210.
Arnold, D.W., 6.L. Kennedy, J"., M.L. Keplinger, J.C. Calandra and C.J. Calo.
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Atlas, E.R. Foster, and C.S. Ginn. 1982. Air-sea exchange of high molecular
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Balba, H.M., and J.G. Saha. 1978. Studies on the distribution, excretion
and metabolism of alpha and gamma isomers of (24C) chlordane in rabbits.
J. Environ. Sci. Health. B13<3):211-233.
Barnett, J.R., and H.H. Dorough. 1974. Metabolism of chlordane in rats.
J. Agric. Food Chem. 22:612-619.
Becker, F.F., and S. Sell. 1979. Alpha-fetoprotein levels and hepatic*
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Biros, F.J., and H.F. Enos. 1973. Oxychlordane residues in human adipose
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Blevins, R.D., and T. E. Sholes. 1978. Response of HeLa cells to selected
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Boyd, E.M., and F.I. Taylor. 1969. The acute oral toxicity of chlordane in
albino rats. Ind. Med. ^8:42.
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EPA 440/4-79-029b.
Cur ley, A., and L.K. Ga^ettson. 1969. Acute chlordane poisoning. Arch.
Environ. Health. 18:211-215.
Dadey, J.L., and A.G. Rammer. 1953. Chlordane intoxication. J. Am. Med.
ASSOC. 153:723.
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64
Chlordane March 31, 1987
-17-
Den Tonkelaar, E.M., and G.J. Van Esch. 1974. Mo effect levels of organo-
chlorine pesticides based on induction of Bicrosoaal liver enzymes in
•hort-term toxicity experiments. Toxicology. 2:371.
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plants. Scand. J. Work Environ. Health. 7(Suppl. 4):140-146.
Dobbs, R.A., and J.M. Cohen. 1980. Carbon adsorption isotherms for toxic
organics, Office of Research and Development. EPA 600/8-80-023.
Epstein, S.S. 1976. Carcinogenicity of heptachlor and chlordane. Sci.
Total Environ. 6:103.
ESE. 1982. Review of organic contaminants in ODW data base for summary of
all available treatment techniques, compound chlordane. February 1982,
EPA 68-01-6494. Office of Drinking Water.
FAD/WHO. 1978. Food and Agricultural Organization/World Health Organization.
FAO Plant Production and Protection Paper 10 Rev. Pesticides Residues
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FDA. 1980a. Food and Drug Administration. Compliance program report of
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Washington, D.C.
FDA. 1980b. Food and Drug Administration. Compliance program report of
findings. FY 77 total diet studies — Infants and toddlers (7320.74).
Food and Drug Administration, U.S. Department of Health, Education and
Welfare, Washington, D.C.
FDA. 1982a. Food and Drug Administration. Compliance program report of
findings. FY 79 total diet studies — Adult (7305.002). Food and Drug
Administration, U.S. Department of Health and Human Services.
Washington, D.C.
FDA. 1982b. Food and Drug Administration. Compliance program report of
findings. FY 79 total diet studies — Infants and toddlers (7305.002).
Food and Drug Administration, U.S. Department of Health and Human
Services. Washington, D.C.
FWPCA. 1968. Federal Water Pollution Control Administration. Water quality
criteria: Report of the National Technical Advisory Committee to the
Secretary of the Interior. U.S. GPO, Washington, D.C.
Furie, B., and S. Trubowitz. 1976. Insecticides and blood dyscrasias:
Chlordane exposure and self-limited refractory megaloblastic anemia.
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65
Chlordane March 31, 1987
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Gentile, J.M., G.J. Gentile, J. Bultman, R. Sechriest, E.D. Wagner and M.J.
Plewa. 1982. An evaluation of the genotoxic properties of insecticides
following p ant and animal activation. Mutat. Res. 101(1):19-29.
Harrington, J.M., E.L. Baker, D.S. Folland, J.W. Saucer and S.H. Sandifer.
1978. Chlordane contamination of a municipal water system. Environ. Res.
15:155-159.
Hyde, K.M., and R.L. Falkenberg. 1976. Neuroelectrical disturbance as indi-
cator of chronic chlordane toxicity. Toxicol. Appl. Pharmacol. 37:499.
I ARC. 1979. International Agency for Research on Cancer. IARC monograph on
the evaluation of the carcinogenic risk of chemicals to humans. Volume 20.
Infante, P.P., S.S. Epstein and W.A. Newton, Jr. 1978. Blood dyscrasias
and childhood tumors and exposure to chlordane and heptachior. Scand.
J. Work Environ. Health. 4:137-150.
Ingle, L. 1952. Chronic oral toxicity of chlordane to rats. Arch. Inc.
Hyg. Occup. Med. 6:357.
IRDC. 1973. International Research and Development Corporation. Unpublished
report to Velsicol Chemical Corporation, eighteen month oral carcinogenic
study in mice, December 14. (Cited in Epstein, 1976)
Klemmer, K.W., A.M. Budy, W. Takahasdhi and T.J. Haley. 1977. Human tissue
distribution of cyclodiene pesticides Hawaii 1964-1973. Clin. Toxicol.
Kutz, F.W., A.R. Yobs and H.S.C. Yang.. . 1976. National pesticide monitoring
programs. In: Air Pollution from Pesticides and Agriculture Processes,
R.E. Lee, Ed. CRC Press, Cleveland, OH. pp. 95-136.
Lensky, P., and M. Evans. 1952. Human poisoning by chlordane. J. Am. Med.
Assoc. 149:1394.
Mabey, W.R., J.H. Smith, R.P. Podoll et al. 1981. Aquatic fate process
data for organic priority pollutants. Monitoring Data Support Division.
Office of Water Regulations and Standards. Washington, D.C.
EPA 440/4-81-014.
Maslansky, C.J., and G.M. Williams. 1981. Evidence for an epi gene tic mode
of action in organochlorine pesticide hepatocarcinogenicity: A lack of
genotoxicity in rat, mouse and hamster hepatocytes. J. Toxicol. Environ.
Health. 8(1-2):121:130.
McCarty, P.L., K.H. Sutherland, J. Graydon and M. Reinhard. 1979. Volatile
organic contaminants removal by air stripping. Presented at seminar on
controlling organics in drinking water, 99th Annual National AHHA Con-
ference, San Francisco, CA.
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Chlordane March 31, 1987
-19-
NCI. 1977. National Cancer Institute. Bioassay of chlordane for possible
carcinogenicity. NCI Carcinogenesis Tech. Rep. Ser. No. 8. 117 p.
[Also publ. as DHEW Publication No. (NIH) 77-B08]
NRC. 1982. National Research Council. An assessment of the health risks of
seven pesticides used for termite control. Prepared for Dept. Navy,
Washington, D.C. NTIS PB 83-136374.
Nye, D.E., and H.W. Dorough. 1976. Fate of insecticides administered endo-
tracheally to rats. Bull. Environ. Contarn. Toxicol. 15:291.
Podowski, A.A., B.C. Banerjee, A. Feroz, M.A. Dudek, R.L. Willey and M.A.Q.
Khan. 1979. Photolysis of heptachlor and cis-chlordane and toxicity of
their photoisomers to animals. Arch. Environ. Contain. Toxicol. 8(5):
509-518.
Polen, P.B., M. Nester and J. Benzinger. 1971. Characterization of oxychlor-
dane, animal metabolite of chlordane. Bull. Environ. Contain. Toxicol.
5:521.
Poonawalla, M.H., and F. Korte. 1971. Metabolism of trans -chlordane-14C and
isolation and identification of its metabolites from the urine of rabbits.
J. Agric. Food Chem. 19(3):467-470.
Princi, F., and G.H. Spurbeck. 1951. A study of workers exposed to the
insecticide chlordane, aldrin, dieldrin. A.M.A. Arch. Ind. Hyg. Occup.
Ned. 3:64-72.
Probst, G.S., R.E. McMahon, L.E. Hill, C.Z. Thompson, J.K. Epp and S.B. Neal.
1981. Chemically-induced unscheduled DNA synthesis in primary rat
hepatocyte cultures: A comparison with bacterial mutagenicity using 218
compounds. Environ. Mutagen. 3(1):11-32.
Shain, S.A., J.C. Shaeffer and R.W. Boesel. 1977. The effect of chronic
ingestion of selected pesticides upon rat ventral prostate homeostasis.
Toxicol. Appl. Pharmacol. 40(1): 115-130.
Sovocool, G.W., and R.G. Lewis. 1975. The identification of trace levels of
organic pollutants in human tissues: Compounds related to chlordane/hep-
tachlor exposure. Trace Subst. Environ. Health. 9:265.
Standard Methods. 1985. Method 509A. Organochlorine Pesticides. In:
Standard methods for the examination of water and wastewater, 16th
Edition, APHA, AWWA, HPCF.
Stohlman, E.F., H.S. Thorp and M.F. Smith. 1950. Toxic action of chlordane.
Arch. Ind. Hyg. 1:13.
Strassman, S.C., and F.W. Kutz. 1977. Insecticide residues in human milk
from Arkansas and Mississippi, 1973-1974. Pestic. Monitor. J. 10:130-133.
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Chlordane March 31, 1987
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Street, J.E., and S.E. Blau. 1972. Oxychlordane : Accumulation in rat
adipose tissue on feeding chlordane isomers or technical chlordane.
J. Agric. Food Chem. 20:395-397.
Tabak, H.H., S.A. Quave, C.I. Mashni and E.F. Barth. 1981. Biodegradability
studies with organic priority pollutant compounds. J. Water Pollut.
Control Fed. 53:1503-1518.
Tashiro, S., and F. Matsumura. 1977. Metabolic routes of cis- and trans-
chlordane in rats. J. Agric. Food Chem. 25:872-880.
Tashiro, S., and F. Matsumura. 1978. Metabolism of trans-nonachlor and
related chlordane components in rat and nan. Arch. Environ. Contain.
Toxicol. 7(1): 11 3-127.
U.S. EPA. 1975a. U.S. Environmental Protection Agency. Analytical report:
New Orleans water supply study. Region VI, U.S. EPA. EPA 906/9-75-003.
U.S. EPA. 1975b. U.S. Environmental Protection Agency. Preliminary assess-
ment of suspected carcinogens in drinking water. Office of Toxic
Substances, U.S. EPA, Washington, D.C.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for organo-
chlorine pesticides in drinking water. In: Methods for Organochlorine
Pesticides and Chlorophenoxy Acid Herbicides in Drinking Water and Raw
Source Water, Interim. July.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water
quality criteria for chlordane. Environmental Criteria and Assessment
Office, Cincinnati, OH. EPA 440/5-80-027. NTIS PB 81-117384.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of
pesticides in drinking water, food and air. Office of Drinking Water.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Review of treat-
ability data for removal of 25 synthetic organic chemicals from drinking
water. Prepared by ESE for U.S. EPA, Office of Drinking Water, March.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Draft health
effects criteria document for chlordane. Office of Drinki.ig Water.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Technologies and
costs for the removal of synthetic organic chemicals from potable water
supplies. Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185): 33992-34003. September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Drinking water
criteria document for heptachlor, heptachlor epoxide and chlordane.
Environmental Criteria and Assessment Office, Cincinnati, OH.
ECAO-CIN-406.
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Chlordane March 31, 1987
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Vettorazzi, A.G. 1975. lexicological decisions and recommendations resulting
from the safety assessment of pesticide residues in food. Crit. Rev.
Toxicol. 4:125-183.
Wang, H.H., and B. MacMahon. 1979a. Mortality of workers employed in the
manufacture of chlordane and heptachlor. J. Occup. Med. 21:745-748.
Wang, H.R., and B. MacMahon. 1979b. Mortality of pesticide workers.
J. Occup. Med. 21:741-744.
Wazeter, F.X. 1967. Unpublished report. (Cited in Vettorazzi, 1975)
Welch, R.M., W. Levin, R. Kuntzman, M. Jacobson and A.M. Conney. 1971.
Effect of halogenated hydrocarbon insecticides on the metabolism and
uterotropic action of estrogen in rats and mice. Toxicol. Appl. Pharmacol.
19:234-246.
WHO. 1984. World Health Organization. Guidelines for drinking water
quality. Volume 1. - Recommendations. EFP/82-39.
Yonemura, T., F. Takamura and Y. Takahashi. 1983a. Two-year feeding/oncogenic
study in mice. (Unpublished study — EPA Pesticide Accession Nos. 254665,
251815)
Yonemura, T., F. Takamura and Y. Takahashi. 1983b. Thirty-month chronic
toxicity and tumorigenicity test in rats by chlordane technical. (Unpub-
lished study — EPA Pesticide Accession No. 252267)
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March 31, 1987
1, 2-DIBROMO-3-CHLOROPROPANE (DBCP)
Health Advisory
Office of Drinking Hater
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Hater (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. The chemical 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, Heibull, 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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1,2-Dibromo-3-Chloropropane (DBCP)
70
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March 31, 1987
This Health Advisory (HA) is based on information presented in the Office
of Drinking Hater's Health Effects Criteria Document (CD) for 1,2-Dibromo-3-
chloropropane (U.S. EPA, 1985a). The HA and CD formats are similar for easy
reference. Individuals desiring further information on the toxicological
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 # 86-118064/AS.
The toll-free number is (800) 336-4700; in the Washington, D.C. area: (703)
487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 96-12-8
Structural Formula
H H H
I I I
H-C-C-C-H
I I I
BrBrCl
Synonyms
0 DBCP, Nemafume, Fumazone, Nemagon.
Dses
0 Nematocidal fumigant.
Properties (U.S. EPA, 1985a)
Molecular Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C3H5Br2Cl
236.36
Technical - light yellow to brown
liquid with pungent odor
Pure - colorless, clear liquid
196°C
0.8 Torr at 21°C
2.08 at 20°C
1,230 mg/L
2.43
0.01 mg/L
0.01 mg/L
1 ppm = 9.67 mg/m3
1 mg/m3 = 0.103 ppm
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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Occurrence
0 Dibromochloropropane (DBCP) is a nematocide which up until 1977 was used
widely on more than 40 crops. Production volume in 1977 is estimated
to have been 30 million Ibs. Between 1977 and 1979, the Agency
cancelled all uses of DBCP except for use on pineapples in Hawaii.
Current production is estimated to be 300,000 Ibs per year.
0 DBCP is regarded as a highly persistent and mobile pesticide. The
major route of removal of DBCP from soil is by volatilization. DBCP
is decomposed slowly in soil both by microbial action and by
hydrolysis. DBCP has been shown to remain in soils for more than
2 years. DBCP has been shown to migrate in soil and has been reported
as a contaminant in ground water. OBCP is expected to be removed
from surface water by volatilization. There is no available
information on DBCP's potential for bioaccumulation.
0 DBCP has not been included in Federal and State monitoring surveys of
ground water and only limited data on its occurrence are available.
A survey of drinking water wells near locations where DBCP had been
used within the last 2 years found levels in the low ug/L. DBCP has
been detected in non-drinking water wells at levels up to 20 ug/L.
DBCP also has been identified in one surface water supply at less
than 1 ug/L. DBCP has been identified as a contaminant in vegetables
grown in soils treated with DBCP. DBCP also has been reported as a
low level contaminant in air. The available data are insufficient to
show whether drinking water is the major route of exposure for DBCP.
Because of the cancellation of all DBCP uses outside of Hawaii,
occurrences of DBCP are expected to decline with time (U.S. EPA, 1983).
III. PHARMACOKINETICS
Absorption
0 Quantitative information pertaining to the absorption of DBCP from
the GI tract, by the lungs or by the skin of laboratory animals was
not located in the available literature (U.S. EPA, 1985a). However,
the Kato et al. (1979) study discussed under Excretion shows a high
absorption potential by the oral route.
0 Gingell (1984b) compared DBCP levels in portal blood from rats orally
treated with DBCP at single doses of 0.1 or 1.0 mg/kg- Compared to
the low dose, there were an initial rapid spike for DBCP in portal
blood and a DBCP level in blood 30 times higher at 10 minutes after
treatment and 10 times higher at 20 minutes after treatment with the
high dose.
0 Gingell (1984b) also reported experiments indicating longer retention
of DBCP in the gastrointestinal tract using oral dosing with corn oil
instead of water as the vehicle.
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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Distribution
0 Ruddick and Newsome (1979) studied the distribution of DBCP in 25
pregnant Wistar rats that had been administered a dose of 25
in corn oil by gavage on days 6 through 15 of gestation. Peak levels
in all tissues examined occurred at 3 to 6 hours after the last dose.
One levels in most tissues declined to below detection levels at 12
hours. Highest levels of DBCP were found in abdominal adipose tissue,
which accumulated the pesticide for up to 6 hours after the last
dose. The level in adipose tissue declined appreciably over the next
6 hours, but relatively high levels were still detected at 24 hours.
After adipose tissue, the tissues that achieved the highest concentra-
tions of DBCP were, in descending order, lung, heart and brain. Low
levels were detected in the fetus, indicating transplacental transfer.
0 The distribution and macromolecular binding of 13-14cj-DBCP (94%
radiochemical purity), administered orally in olive oil to 9-week-old
male Wistar rats at doses of 20, 50, 100, 200 or 400 mgA9» were
studied by Kato et al. (1980). Whole-body autoradiography, at 6
hours following administration of 20 or 200 mg 14C-DBCP/kg bw, indi-
cated that highest levels of radioactivity were in the liver and
renal cortex. No difference in the pattern of distribution for the
two doses was observed. In different groups of rats receiving oral
doses of 14C-DBCP at 20 to 400 mg/kg, the concentration of DBCP residue,
total radiocarbon, and total bound radiocarbon in plasma, blood
cells, liver, kidney and testes at 6 hours all increased with dose.
By 24 hours, the levels of DBCP had declined to £0.32 ppm for all
tissues in each dose group except for adipose tissue in rats given
dosages higher than 20 mg/kg« DBCP itself accounted for only 4 to
6.5% of the total radioactivity in the kidney, liver and testes
(target organs of DBCP toxicity) at 6 hours. Approximately 40 to 70%
of the radiocarbon was bound to tissue (kidney, liver and testes)
macromolecules with all doses at 6 and 24 hours after dosing. Thus,
DBCP was metabolized rapidly to a reactive species that binds with
tissue macromolecules. Binding sites were not saturated at the doses
studied.
0 Gingell (1984a,b) reported liver, kidney and forestomach as organs
with highest levels of 14C in rats orally treated with 14c-DBCP.
At one day following treatment, 14C levels in liver and kidney were
equivalent to those in fat.
Metabolism
0 Urinary metabolites found in rats were mercapturic acid conjugates,
^chlorolactic acid,/?-bromolactic acid and 2-bromoacrylic acid. Pro-
posed metabolic pathways involve intermediates including epihalohydrins
and other reactive epoxides and 2-bromoacrolein (U.S. EPA, 1985a).
0 DBCP administration depleted the glutathione content of the liver and
kidney of rats and mice (Kato et al., 1980; Kluwe et al., 1981). The
depletion of glutathione was dose-related and coincided in time with
the dose-related increased binding of radioactive metabolites of
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HMO
/o
1,2-Dibromo-3-Chloropropane (OBCP) March 31, 1987
-5-
14C-DBCP to liver and kidneys, which resulted in toxicity (Kato
et al., 1980).
9 It has been proposed that intermediate and end metabolites of OBCP,
such as epichlorohydrin,tf -chlorohydrin and oxalic acid, account for
the wide variety of toxic effects (Jones et al., 1979). However,
Gtngell et al. (1983, 1984a, 1984b, 1985) and Beatty et al. (1983)
reported that little, if any, DBCP is metabolically converted to
epichlorohydrin and dominant metabolites of epichlorohydrin are hardly
detectable in urine of animals treated with DBCP. Kluwe et al. (1983)
demonstrated that DBCP, epichlorohydrin and *-chlorohydrin may produce
toxic effects in testes, epididymis and forestomach by similar
mechanisms. The renal effects of DBCP were unlike those of oxalic
acid, epichlorohydrin and*-chlorohydrin.
Excretion
Kato et al. (1979) administered 14c-DBCP orally to male Wistar rats
at doses of 20 to 400 mg/kg bw and observed that nearly 85% of the
radioactivity was eliminated in the urine, bile and expired air. No
unchanged DBCP was eliminated in the urine and only traces of the
eliminated label in the expired air were unchanged DBCP. Urinary and
fecal excretion accounted for 51.4 and 22%, respectively, of the
radioactivity after 14 days, while 17.6% was excreted in the expired
air after 48 hours. Biliary excretion accounted for 22.7% of the
dose in 24 hours. The urine was the predominant route of elimination
for metabolites of DBCP.
IV. HEALTH EFFECTS
Humans
0 No case studies of acute exposure to DBCP were found in the available
literature (U.S. EPA, 1985a). No association between DBCP exposure
and cancer (discussed under Carcinogenicity) or renal effects in
humans has been documented.
0 Reproductive effects studies are discussed under Reproductive Effects.
Animals
Short-term Exposure
0 Acute oral LDjQ values include 170, 410 and 440 mg/kg for rats, mice
and rabbits, respectively (U.S. EPA, 1985a).
0 An acute lethal oral dose (400 mg/kg bw) to rats resulted in necrosis
of hepatocytes and degeneration of renal tubules (Kato et al., 1980);
a lower dose (100 mg/kg bw) resulted in reduced spennatogenesis
(Reznik and Sprinchan, 1975). Oral doses of 40 or 50 mg/kg/day
administered for 4 or 5 days resulted in decreased body weight.
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impaired renal function, necrotic lesions of liver and kidney and
degeneration of testes and epididymis in rats (Kluwe, 1981; Saito-
Suzuki et al., 1982).
• Acute inhalation exposures of 483-6,475 mg/m3 for up to 7 hours
resulted in such effects as scarring of kidney tissue, pulmonary
irritation, liver damage, CNS depression and death in rats (Kodama
and Dunlap, 1956, Torkelson et al., 1961). Continuous inhalation of
29 to 97 mg/m3 DBCP for 14 days resulted in atrophy of seminiferous
tubules, necrotic germ cells of the testes, necrosis of proximal
tubules in kidneys and necrotic lesions in lung epithelial tissue in
rats (Saegusa et al., 1982).
0 The dose-response for single subcutaneous injection of DBCP in adult
male Fischer 344 rats was described by Kluwe et al. (1981). The
no-effect level was 20 rag/kg* At i40 mgAg bw, such effects as
reduced body weight, impaired kidney function, degeneration of prox-
imal renal tubule epithelium, hepatocellular necrosis and decreased
spermatogenesis were observed. Kluwe et al. (1981) found that DBCP
induced similar effects in rats whether given by gavage or subcu-
taneous injection, although the former route was slightly less toxic,
at a dose of 40 mg/kg/day for 4 days.
0 Kluwe et al. (1985) compared toxic responses of 6-day-old and 25-day-
old male Fischer 344 rats to a single subcutaneous injection of DBCP.
Six-day-old rats were more sensitive to DBCP toxicity as shown by
reduced survival, renal dysfunction and renal and hepatic necrosis
with doses i80 mg/kg. Doses i20 mg/kg, the lowest dose given,
reduced body and gonadal weight gains and caused hypospermatogenesis
or seminiferous tubular atrophy in rats exposed when 6 days old.
Doses *160 mgAg were needed to produce residual toxic effects in
rats treated when 25 days old.
Long-term Exposure
0 Dietary administration of DBCP to rats for 90 days resulted in
increased kidney weights at i2 mg/kg/day, reduced body weight gain
at 15 mg/kg/day, increased liver weight and ruffled fur at i45
ngAg/day, and muscular weakness and increased mortality at 135
mg/kg/day. The no-effect level was 0.5 mg/kg/day (Torkelson et al.,
1961).
0 Effects of chronic exposure to DBCP by gavage in a carcinogenic!ty
bioassay included dose-related increased mortality and a high incidence
of toxic tubular nephropathy in mice and rats (NCI, 1977). These
effects were observed for TWA doses of 10.7 to 146 mgAg bw/day.
0 Lifetime treatment with DBCP in the diet resulted in stomach nodules
in male and female Han/ICR Swiss mice with doses of 0.28 mgAg/day
and higher and kidney lesions in female Charles River CD rats and
reduced body weight and organ weight changes in male Charles River
rats given 2 mg/kg/day (Hazelton, 1977, 1978).
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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0 Consumption of DBCP in drinking water by male Sprague-Dawley rats for
64 days induced renal lesions at 100 and 200 ppm levels but not at
5 and 50 ppm levels. However, increases in protein and glucose levels
and the specific gravity of urine found with 100 and 200 ppm levels
were not assessed at the lower levels. Exposure levels in this study
were estimated as 9.38, 5.21, 3.17 and 0.37 mgAg (Heindel et al.,
1983).
0 Johnston et al. (1986) administered DBCP at 0.02, 0.2, 2 and 20
mgAg/day in drinking water to male and female Sprague-Dawley rats
for 60 days before mating and throughout mating, gestation and the
first 5 days of lactation in a one-generation reproduction study.
Treatment-related effects were not evident at doses below 20 mg/kg/day
where reduced body weights in the parents and 4-day old pups occurred.
0 Commonly observed effects of subchronic inhalation exposure in male
animals were testicular atrophy and reduced spermatogenesis (Torkelson,
et al., 1961; Rao et al., 1982, 1983). Lesions of the upper respi-
ratory system and nasal cavity occurred in rats and in mice (Resnik,
et al., 1980 a,b. NTP, 1982). Reduced body weight gain, increased
mortality and histopathologic changes in kidney tubules, liver,
testes and adrenal cortex of rats have been observed at high exposure
levels (Torkelson et al., 1961; Rao et al., 1983). At inhalation
. exposures of >1 ppm (9.7 mg/m3) 6 or 7 hours/day, 5 days/week for
up to 14 weeks in rats and mice, dose-related effects included
decreased body weight, increased liver weight and focal histopatho-
logic changes in the testes, renal tubules, lung and nasal cavities,
and increased mortality (Torkelson et al., 1961; NTP, 1982). A
no-effect level for mortality, clinical chemistry, hematology, body
weight, organ weights and testicular effects was reported for rats
and rabbits as 0.1 ppm (0.97 mg/m3) 6 hours/day, 5 days/week for 14
weeks (Rao et al., 1982, 1983). Effects of chronic inhalation
exposure to DBCP in a carcinogenicity bioassay included dose-related
decreased mean body weight, increased mortality, increased incidences
of toxic tubular nephropathy and histopathologic lesions of the nasal
cavity and stomach in rats and mice at concentrations of 5.8 and
29 mg/m3 (NTP, 1982).
Reproductive Effects
0 A reported effect of DBCP in humans is reduced spermatogenesis in
chemical plant workers and agricultural workers (U.S. EPA, 1985a).
Recovery of normal sperm counts occurs when DBCP exposure ceases; the
amount of time required depends upon the intensity and duration of
the exposure.
0 Potashnik and Abelovich (1985) found no chromosome aberrations in men
who had suppressed spermatogenesis from occupational exposure to DBCP.
Nor were there increases in abortions and malformations regarding
their offspring.
0 EHA (1985a) made an epidemiological investigation'of the relationship
between DBCP contamination in drinking water and reproduction (birth
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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rate, btrth weight, sex ratio, birth injury and birth defects) between
1978 and 1982 in Fresno County, California. Negative results were
concluded. Exposures to 97.8% of the 45,914 mothers evaluated were
to 3 ppb DBCP and lower*
In the previously mentioned reproduction study by Johnston et al.
(1986), treatment-related effects on reproduction and pathology were
not found except for reduced pup weight in the 20 mg/kg group.
Inhalation exposure of female Sprague-Dawley rats to levels as high
as 97 mg/m3 or 1 4 weeks before mating with unexposed males had no
effect on reproduction or fetal development. Males were exposed
similarly and mated with unexposed females during 14 weeks of treat-
ment and 27 weeks of recovery after treatment. Exposure to 0.97
mg/m3 was ineffective, but exposure to 9.7 mg/m3 decreased repro-
ductive success when ma tings were done up to 5 weeks post-treatment;
however, ma tings at 27 weeks after exposure did not affect reproduc-
tion (Rao et al., 1983). The dose response pattern was similar in a
comparable study with male New Zealand rabbits (Rao et al., 1982).
Foote et al. (1986a,b) assessed the reproductive effects of DBCP in
male Dutch rabbits (six per group) given 0, 0.94, 1.88, 3.75, 7.5 or
15 mg DBCPAg body weight in drinking water 5 days/week for 10 weeks.
Body and organ weights and survival were unaffected except for decreased
testis weights at 15 mg/kg » a dose which also increased FSH and reduced
sperm production. Mean seminiferous tubular diameter was decreased
with 7.5 and 15 mg/kg. Sperm morphology was the most sensitive
indicator of toxicity with 1 .88 mg/kg and higher being observable
effect levels. The authors concluded 0.94 mg/k
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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and increases in LH and FSH occurred. At 5 mgAg* body and testis
weights were decreased. There was no observable effect at 1 mgAg.
• Liu (1985) found that treatment of male rats every other day during
the first 20 days of age with 1, 5, 10 or 20 mg DBCP/kg by subcutaneous
injection resulted in no observable gonadotropic effect at 1 mg/kg»
although seminal vesicle/body weight ratios were reduced, and reduc-
tions in testis/body weight ratios and serum androgen levels at higher
levels. Testicular lesions were induced with 10 and 20 mgAg.
Developmental Effects
0 Ruddick and Newsome (1979) found no teratogenic effects in fetuses
of pregnant Wistar rats treated with DBCP by gavage at 12.5, 25 or
50 mgAg/day on days 6 through 15 of gestation. The 50 mgAg dose
was fatal to embryos and toxic to dams, and the 25 mgAg dose reduced
body weight in dams.
Mutagenicity
0 Technical grade DBCP was mutagenic in S_. typhimurium strains TA1535,
TA1530, TA100 and TA98 and in E. coli, with and without metabolic
activation (Rosenkranz, 1975; Prival et al., 1977; Stolzenberg and
Hine, 1979; Moriya et al., 1983; Traul et al., 1985; Ratpan and
Plauman, 1985; Ohta et al., 1984). Some of the mutagenic potential
of DBCP has been attributed to epichlorohydrin, a contaminant of
technical grade DBCP (Biles et al. (1978), who reported technical and
purified grades of DBCP as equally mutagenic in TA100 with metabolic
activation]. Negative results were obtained with £. typhimurium
strains TA1537 and TA1538 (Rosenkranz, 1975; Moriya et al., 1983;
Ratpan and Plauman, 1985, who also reported A-98 as negative). DBCP
was positive in the recessive lethal assay, in a genetic crossing
over assay, and for chromosome breakage in I), melanogaster (Kale and
Baum, 1982; Inoue et al., 1982, and Zimmering, 1983). Dominant
lethal assays were positive in rats (Teramoto et al., 1980; Saito-
Suzuki et al., 1982; Rao et al., 1979, 1983) but negative in mice
(Teramoto et al., 1980; Generoso et al., 1985). Positive results
were obtained in a sister chromatid exchange study in cultured Chinese
hamster cells (Tezuka et al., 1980), for chromosome aberrations in
rats treated in vivo (Kapp, 1979) and for unscheduled DNA synthesis
in germ cells of prepubertal mice treated ±n_ vivo (Lee and Suzuki,
1979). Russell (1985) reported DBCP as negative in the mouse specific
locus test.
Carcinogenicity
0 DBCP has been studied for carcinogenic!ty in mice and rats by oral
and inhalation routes and in mice by dermal application. An NCI
(1977) bioassay reported highly significant dose-related increased
incidences in Osborne-Mendel rats of squamous-cell carcinoma of the
forestomach in males and females and mammary adenocarcinoma in
females receiving chronic gavage time-weighted-average (TWA) doses
of 10.7 and 20.7 mgAg bw/day. Significant dose-related increased
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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incidences of squamous-cell carcinoma of forestomach of male and
female B6C3Fi mice were found for chronic gavage TWA doses of 78.6-
149.3 mg/kg bw/day. In a chronic dietary carcinogenicity bioassay in
rats conducted by Hazelton Laboratories (1977) and also reported by
U.S. EPA (1979a,b,c), high-dose (2.0 mgAg bw/day) male and female rats
had significantly increased incidences of carcinoma of the renal
tubules and squamous-cell carcinoma of the stomach. High incidences
of stomach squamous-cell carcinoma were observed in high-dose male
and female mice as well. An NTP (1982) bioassay showed dose-related
increased incidences of nasal cavity tumors in male and female F344
rats and B6C3Fi mice receiving chronic inhalation exposures to DBCP
at concentrations of 5.8 and 29 mg/m3, 6 hours/day, 5 days/week.
The mice also had treatment-related increased incidences of pulmonary
tumors. DBCP was positive as a tumor initiator in the skin of Ha/ICR
Swiss mice but negative as a whole carcinogen for skin (Van Duuren
et al., 1979). In the whole carcinogen assay, however, the incidence
of distant tumors of lung and stomach were significantly increased
over controls.
0 EHA (1986b) did not find an association between incidences of gastric
cancer and leukemia and DBCP contamination in drinking water in Fresno
County, California. Census tract data show the range of average DBCP
levels to be 0.0041-5.7543 ppb DBCP with 14* of the tracts showing
levels >1 ppb. Other organ sites were not assessed. This study
contrasts an earlier, similar study by Jackson et al. (1982) which
indicated a tentative association between gastric cancer and leukemia
and DBCP exposure in drinking water in Fresno County, California.
0 Hearn et al. (1984) did not find an association between cancer
induction and DBCP exposure in a cohort of 550 Dow Chemical Co.
employees potentially exposed to DBCP during its production from
1957 to 1975. Exposure levels were not estimated.
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) = „ ( „,
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW - assumed body weight of a child (10 kg) or
an adult (70 kg).
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UF - uncertainty factor (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).
Organoleptic Considerations
The taste and odor threshold levels of 0.01 mg/L are lower than the
One-day and Ten-day Health Advisories.
One-day Health Advisory
As investigated by Kluwe (1981), subcutaneous injections of DBCP at
40 mgAg/day in adult rats for 4 days resulted in approximately equivalent
toxic effects as did oral doses of 40 mgAg/day for 4 days. When the dose-
response of single subcutaneous injections of DBCP in rats was defined,
40 mgAg/day resulted in cytoplasmic vacuolization of renal tubule epithelium
and impaired renal function, as evidenced by increased urinary excretion of
proteins and ions. At 20 rag A 9 bw subcutaneously, no toxic effects were
observed and this dose was considered to be a NOAEL. However, Kluwe et al.
(1985) found a single subcutaneous injection of 20 mg/kg effective in reduc-
ing body weight gain and in producing gonadotoxic effects when given to
6-day-old rats. Therefore, 20 mgAg i-s concluded to be a LOAEL. Taking this
subcutaneous LOAEL as equivalent to the acute oral LOAEL, the 20 mgAg bw
dose of DBCP can be used to calculate the One-day HA for a 10 kg child as
follows:
One-day HA = (20 mgAg/day) (10 kg) „ 0.2 mg/L (200 ug/L)
(1,000) (1 L/day)
where:
20 mgAg/day = LOAEL for body weight and gonadotoxicity in 6-day old
rats given DBCP.
10 kg = assumed body weight of a 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.
Ten-day Health Advisory
The 90-day study by Torkelson et al. (1961) is selected for the Ten-day
HA calculation since the dietary administration used closely approximates
drinking water exposure, a wide dose range showing effect and no-effect
levels was evaluated, and the NOAEL is not considered to be unreasonable,
taking into account doses and questions regarding NOAELs in other studies
described herein which could possibly serve as the basis. The recent and, as
yet, unpublished study by Heindel et al. (1983), where male rats were given
DBCP in drinking water for 64 days, is not used because female rats were not
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 198.7
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used In the study and the complete dose response for kidney toxicity (increases
in glucose and protein levels in urine of rats given 5.21 and 9.38 mgAg/day
doses; these levels were not assessed at 3.1 and 0.38 mgAg/day doses) was
not measured to provide a NOAEL. The reproduction study in rats given DBCP
in drinking water by Johnston et al. (1981) showed no reproductive or other
pathological effects at doses of 2 mgAg/day or less, but kidney effects
found by Heindel et al. (1983) were not assessed by Johnston et al. (1981).
Foote et al. (1986a,b) concluded that the lowest dose (0.94 mgAg) used in
their reproduction study in rabbits could be considered a no-effect level,
but they also concluded that the actual no-effect level could have been
lower. In Torkelson et al. (1961), a dietary NOAEL of 5 mgAg diet in rats
was determined. At the next highest dietary level (20 mgAg diet), female
rats had significantly increased kidney-to-body weight ratios. The NOAEL of
5 mgAg diet can be used to calculate the Ten-day HA as follows:
Transformed dose (d) for rats:
d - 5 mgAg diet x 0.10 kg diet/kg bw/day
« 0.50 mgAg bw/day
For a child:
Ten-day HA - (0.50 mgAg/day) (10 kg) = 0.050 mg/L (50 ug/L)
(100) (1 L/day) * y/
where:
0.5 mgAg/day = NOAEL for kidney effects in rats given DBCP in the
diet for 90 days.
0.10 = assumed proportion of body weight ingested per day by
a young growing rat (Mitruka et al., 1976).
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day » assumed daily water consumption of a child.
For comparison with the Ten-day HA, an HA is estimated from 0.4 ppro
(3.9 mg/m3), the estimated exposure level in chemical production workers
(Whorton et al., 1977) which resulted in reduced sperm counts, as follows:
Transformed dose (d) can be calculated as follows:
d = 0.9 mg/m3) (10 m3/dav) (5/7) (0.5) . 0.20 mg/kg ^/^
(70 kg)
where:
3.9 mg/m3 « LOAEL for reduced sperm counts/
10 m3/day » assumed human breathing volume in an 8-hour workday.
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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• Using the 95% upper limits, risk estimates with other models are
presented for comparison with that derived with the multistage.
For example, an excess cancer risk of one in 1,000,000 (10~6) is
associated with DBCP levels in drinking water of 50 ug/L (probit),
2 ug/L (logit) and 0.2 ug/L (Weibull). While recognized as statisti-
cally alternative approaches, the range of risks described by using
any of these modeling approaches has little biological significance
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.
• The IARC (1979) categorized DBCP as a 2B carcinogen, i.e., sufficient
evidence of carcinogenic!ty in animals, inadequate evidence in humans.
• Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), DBCP is 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.
VT. OTHER CRITERIA, GUIDANCE AMD STANDARDS
0 The National Academy of Sciences (NAS, 1986) used the forestomach
tumor data in male Osborne-Mendel rats in the NCI (1977) carcinogenic!ty
bioassay and the multistage model to calculate estimated human lifetime
risk of 7.8 x 10~6 and upper 95% confidence estimate of lifetime cancer
risk of 9.9 x 10-6 from daily consumption of 1 L of water containing
DBCP at a level of 1 ug/L. The NAS (1986) did not have the data from
Hazelton (1977, 1978) for review.
0 Earlier, the National Academy of Sciences (NAS, 1982) did not calculate
SNARLs (Suggested-No-Adverse-Response-Levels) for DBCP on the grounds
that contaminants of DBCP might be responsible for the observed
adverse effects in humans and animals.
0 The Office of Drinking Water (ODW) of the U.S. EPA has given some
guidance in this area (Cotruvo and Melone, 1983). The upper limit
excess lifetime cancer risk associated with 50 ng/L of DBCP is 9.0 x
10-6, assuming a consumption of 2 L of contaminated vnter per day and
an average adult body weight of 60 kg. The toxicity-based drinking
water concentration was 50 ng/L.
0 NIOSH (1978) has recommended that a ceiling of 10 ppb (0.1 mg/m3) be
set for occupational exposure to DBCP.
0 In 1977, the Occupational Safety and Health Administration (OSHA)
proposed to set a permissible exposure limit for DBCP at 1 ppb (0.01
mg/m3) for an 8-hour TWA exposure and a mean ceiling of 10 ppb (0.1
mg/m3) for any 15-minute period during the workshift (OSHA, 1977).
These proposed limits were based on the view of OSHA that the lowest
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
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level of DBCP detectable by industrial sampling and analysis methods
and the lowest level capable of being achieved is 1 ppb (0.01 mg/m3).
• In 1979, the U.S. EPA (U.S. EPA, 1979c) suspended the registration of
pesticide products containing DBCP.
• The U.S. EPA (1985b) recently issued an intent to cancel all registra-
tions for pesticide products containing DBCP.
0 The proposed RMCL by the U.S. EPA Office of Drinking water is zero
(U.S. EPA, 1985c).
VII. ANALYTICAL METHODS
* Analysis of dibromochloropropane is by a purge-and-trap gas chromato-
graphic procedure used for the determination of volatile organohalides
in drinking water (U.S. EPA, 1985d). This method calls for the bubbling
of an inert gas through the sample and trapping dibromochloropropane
on an adsorbant material. The adsorbant material is heated to drive
off the dibromochloropropane onto a gas chromatographic column. This
method is applicable to the measurement of dibromochloropropane over
a concentration range of 0.3 to 1500 ug/L. Confirmatory analysis for
dibromochloropropane is by mass spectrometry (U.S. EPA, 1985e). The
detection limit for confirmation by mass spectrometry is 0.2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Dobbs and Cohen (1980) reported that the adsorption capacities of 18,
6.0 and 2.0 mg of DBCP per mg of granular activated carbon (GAC) at
initial concentrations of 0.1, 0.01 and 0.001 mg/L DBCP, respectively.
0 Environmental Science and Engineering (ESE, 1984), in laboratory studies,
used the Dynamic Mini Column Adsorption Technique (DMCAT) to study DBCP
adsorption. Deionized water spiked with DBCP at approximately 100 ug/L
or 50 ug/L was passed through a 2.1 mm diameter column filled with 50 mg
GAC (reactivated Filtrasorb* 300). The data obtained from this study
were used to predict carbon usage rates in lbs/1,000 gallons: 0.18 and
0.105 for influent concentrations of 93 and 51 ug/L, respectively.
0 The Henry's Law Constant for DBCP has been reported to be 1.26 x 10-4
atm x m3/mole at 20°C (Selleck et al., 1983). This value suggests
that high air-to-water ratios or packing heights will be needed to
remove DBCP. A pilot air stripping study conducted by Selleck et al.
(1983) used a 13-foot column (cross sectional area 3.32 ft2) packed
with 2-inch polypropylene Intalox saddles. The study included runs
at a variety of treatment conditions. Up to 98% removal was achieved
at 19.2°C and an air-to-water ratio of 600. Thus air stripping could
be applied to the removal of DBCP from water.
0 Air stripping has been found to be an effective, simple and relatively
inexpensive process for removing many volatile organics from water.
However, this process transfers the contaminant directly to the air
stream, and consideration must be given to the overall environmental
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
-16-
occurrence, fate, route' of exposure and various hazards associated
with the chemical.
Aeration and carbon adsorption for the removal of DBCP from water are
available and have been reported to be effective. Selection of indi-
vidual or combinations of technologies to achieve DBCP reduction must
be based on a case-by-case technical evaluation, and an assessment of
the economics involved.
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IX. REFERENCES
Amann, R.P., and W.E. Berndtsen. 1986. Assessment of procedures for screening
agents for effects on male reproduction. Effects of dibromochloropropane
(DBCP) in the rat. In press.
Beatty, P.W., R.L. Mueller and A.C. Page. 1983. In vitro metabolism of ^C
epichlorohydrin in hepatic and extrahepatic microsomes from F-344 rats.
The Toxicologist. 3:5.
Biles, R.W., T.H. Connor, N.M. Trieff and M.S. Legator. 1978. The influence
of contaminants on the mutagenic activity of dibromochloropropane (DBCP).
J. Environ. Pathol. Toxicol. 2(2):301-312.
Burlinson, N.E., L.A. Lee and D.H. Rosenblatt. 1982. Kinetics and products
of hydrolysis of 1,2-dibromo-3-chloropropane. Environ. Sci. Technol.
16(9):627-632.
Burns, L.H., D.M. Cline and R.R. Lassiter. 1981. Exposure analysis modeling
system (EXAMS). Prepared by Environmental Research Laboratory. Office
of Research and Development, U.S. EPA, Athens, GA.
Cotruvo, J.A., and J.H. Melone. 1983. Personal communication to Charles G.
Clark, Director of Health, State of Hawaii. August 2.
Dobbs, R.A., and J.M. Cohen. 1980. Carbon adsorption isotherms for toxic
organics. Report No. EPA-600/8-80-023. U.S. EPA. Office of Research
and Development, MERL, Cinicinnati, OH.
EHA. 1986a. Environmental Health Associates, Inc. An epidemiologic investi-
gation of the relationship between DBCP contamination in drinking water
and reproductive effects in Fresno County. Unpublished report submitted
to Shell Oil Company.
EHA. 1986b. Environmental Health Associates, Inc. Final Report: Examination
of the possible relationship between DBCP water contamination and leukemia
and gastric cancer in Fresno County, California. Submitted to Shell Oil
Company. Unpublished.
ESE. 1984. Environmental Science and Engineering. Review of treatability
data for removal of twenty-five synthetic organic chemicals from drinkinc
water. Prepared for U.S. EPA. Office of Drinking Water, Washington, DC.
Foote, R.H., E.C. Schermerhorn and M.E. Simkin. 1986a. Measurement of semen
quality, fertility, and reproductive hormones to assess dibromochloro-
propane (DBCP) effects in live rabbits. Fund. Appl. Toxicol. 6:628-637.
Foote, R.H., W.E. Berndtson and T.R. Rounsaville. 1986b. Use of quantitative
testicular histology to assess the effect of dibromochloropropane (DBCP)
on reproduction in rabbits. Fund. Appl. Toxicol. 6:638-647.
Generoso, W.M., K.T. Cain and L.A. Hughes. 1985. Tests for dominant lethal
effects of 1,2-dibromo-3-chloropropane (DBCP) in male and female mice.
Mutat. Res. 156:103-108.
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1,2-Dibromo-3-Chloropropane (DBCP) March 31, 1987
-18-
Gingell, R., H. Mitschke, P.W. Beatty and A.C. Page. 1983. Disposition and
metabolism of 14C epichlorohydrin. The Toxicologist. 3:5.
Gingell, R., and A.C. Page. 1984a. Biochemistry of halogenated three-carbon
compounds: An overview and discussion of proprietary and published
information on the disposition and metabolism of epichlorohydrin and
1,2-dibromo-3-chloropropane. Tech. Info. Record Mo. WRC-822.
Gingell, R., and A.C. Page. 1984b. Biochemistry of halogenated three-carbon
compounds; pharmacokinetic disposition of 1,2-dibromo-3-chloropropane in
rats after oral administration in water or corn oil. Tech. Info. Record
No. WRC-891.
Gingell, R. et al. 1985. Evidence that epichlorohydrin (ECH) is not a
metabolite of 1,2-dibromo-3-chloropropane (DBCP) in the rat. The Toxi-
cologist. 5:77.
Hazelton Laboratories America, Inc. 1977. 104-Week dietary study in rats,
1,2-dibromo-3-chloropropane (DBCP). Final Report. Unpublished report
submitted to Dow Chemical Co., Midland, MI. Oct. 29, 1977.
Hazelton Laboratories America, Inc. 1978. 78-Week toxicity and carcinogenic!ty
study in mice. Final Report. Project No. 174-125. Unpublished report
submitted to Dow Chemical Co. Nov. 3, 1978.
Hearn, S., M.G. Ott, R.C. Kolesor and R.R. Cook. 1984. Mortality experience
of employees with occupational exposure to DBCP. Arch. Environ. Hlth.
39:49-55.
Ueindel, J.J., J.V. Bruckner and E. Steinberger. 1983. A protocol for the
determination of the no-effect level of 1,2-dibromo-3-chloropropane
(DBCP) on the qualitative morphological integrity of the testicular
seminiferous epithelium. Submitted to the U.S. EPA, Office of Drinking
Water.
IARC. 1979. International Agency for Recearch on Cancer. IARC Monographs
on the Evaluation of the Carcinogenic Risk to Humans. 1,2-dibromo-3-
chloropropane. WHO, IARC, Lyon, France. 20:83-96.
Inoue, T., J. Miyazawa, N. Tanahashi, M. Moriya and Y. Shirasu. 1982.
Induction of sex-linked recessive lethal mutations in ProsPIhlla melano-
gaster males by gaseous 1,2-dibromo-3-chloropropane (DBCP). Mutat. Res.
105:89-94.
Jackson, R.J., C.J. Greene, J.T. Thomas, E.L. Murphy and J. Kaldor. 1982.
Literature review on the toxicological aspects of DBCP and an epidemic-
logical comparison of patterns of DBCP drinking water contamination with
mortality rates from selected cancers in Fresno County, California,
1970-1979. California Department of Health Services. Unpublished.
Johnston, R.V., D.C. Mensik, H.W. Taylor, G.C. Jersey and F.K. Dietz. 1986.
A single-generation drinking water reproduction study of 1,2-dibromo-3-
chloropropane in Sprague-Dawley rats. Bull. Environ. Contam. Toxicol.
In press.
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87
1,2-Dibromo-3-Chloropropane (DBCP) February 12, 1987
-19-
Jones, A.R., G. Fakhouri and P. Gadiel. 1979. The metabolism of the soil
fumigant DBCP in the rat. Experientia. 35:1432-1434.
Kale, P.G., and J.W. Baum. 1982. Genetic effects of 1,2-dibromo-3-chloro-
propane in Prosophila. Environ. Mutagen. 4(6):681-688.
Kapp, R.W., Jr. 1979. Mutagenicity of 1,2-dibromo-3-chloropropane (DBCP): ^n
vivo cytogenetics studies in the rat. Toxicol. Appl. Pharmacol. 48:A46.
Kato, Y., K. Sato, S. Maki, O. Matano and S. Goto. 1979. Metabolic fate of
1,2-<3ibromo-3-chloropropane (DBCP) in rats. J. Pestic. Sci. 4:195-203.
Kato, Y., K. Sato, T. Harada, S. Maki, O. Matano and S. Goto. 1980. Metabolic
fate of DBCP in rats. III. Correlation between macroraolecular binding
of DBCP-metabolite and pathogenicity of necrosis. J. Pestic. Sci.
5(1):81-88.
Kluwe, W.M. 1981. Acute toxicity of 1,2-dibromo-3-chloropropane in the F344
male rat. 1. Dose-response relationships and differences in routes of
exposure. Toxicol. Appl. Pharmacol. 59:71-83.
Kluwe, W.M., R. McNish, K. Smithson and J.B. Hook. 1981. Depletion by
1,2-dibromoraethane, 1,2-dibromo-3-chloropropane, tris(2,3-dibromopropyl)-
phosphate, and hexachloro-1,3-butadiene of reduced nonprotein sulfhydryl
groups in target and nontarget organs. Biochem. Pharmacol. 30(16)2265-
2271.
Kluwe, W.M., B.N. Gupta and J.C. Lamp IV. 1983. The comparative effects of
1,2-dibromo-3-chloropropane (DBCP) and its metabolites, 3-chloro-T,2-
propaneoxide (epichlorohydrin), 3-chloro-1,2-propanediol (alphachloro-
hydrin), and oxalic acid, on the urogenital system of male rats.
Toxicol. Appl. Pharmacol. 70(1):67-86.
Kluwe, W.M., H. Weber, A. Greenwell, and F. Harrington. 1985. Initial and
residual toxicity following acute exposure of developing male rats to
dibromochloropropane. Toxicol. Appl. Pharmacol. 79:54-68.
Kodama, J.K., and M.K. Dunlap. 1956. Toxicity of 1,2-dibromo-3-chloropropane.
Abst. No. 1459. Fed. Proc. 15:448.
Lee, I.P., and K. Suzuki. 1979. Induction of unscheduled DN\ synthesis in
mouse germ cells following 1,2-dibromo-3-chloropropane (DBCP) exposure.
Mutat. Res. 68:169-179.
Liu, E.M.K. 1985. Reproductive function of adult male rats following neonatal
exposure to 1,2-K3ibromo-3-chloropropane. The Toxicologist. 5:120.
Mabey, W.R., J.H. Smith, R.T. Podoll et al. 1981. Aquatic fate process
data for organic priority pollutants. EPA-440/4-81-014.
Moody, D.E., G.A. Clawson, C.H. Woo and E. Smuckler. 1982a. Cellular
distribution of cytochrome P-450 loss in rats of different ages treated
with alkyl halides. Toxicol. Appl. Pharmacol. 66(2):278-289.
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1,2-Dibromo-3-Chloropropane (OBCP) February 12, 1987
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Moody, D.E., G.A. Clawson and E.A. Smuckler. 1982b. The integrity of liver
protein synthesis in male rats treated with 1,2-dibromo-3-chloropropane.
Toxicol. Lett. 12(2-3):101-108.
Moriya, M., T. Ohta, K. Watanabe, T. Miyazawa, K. Kato and Y. Shirasu. 1983.
Further mutagenicity studies on pesticides in bacterial reversion assay
systems. Mutat. Res. 116(3-4):185-216.
MAS. 1982. National Academy of Sciences. Drinking Water and Health.
Volume 4. National Academy Press, Washington, D.C. pp. 209-214.
NAS. 1986. National Academy of Sciences. Drinking Water and Health. Volume 6.
National Academy Press, Washington, D.C. pp. 315-326.
NCI. 1977. National Cancer Institute. Bioassay of dibromochloropropane for
possible carcinogenicity. NCI Carcinogenesis Tech. Rep. Ser. No. 28.
93 pp. NTIS PB 277-472.
NIOSH. 1978. National Institute for Occupational Safety and Health. Criteria
for a recommended standard ... Occupational exposure to dibromochloropro-
pane (DBCP). NIOSH 78-115.
NTP. 1982. National Toxicology Program. Carcinogenesis bioassay of
1,2-dibromo-3-chloropropane (CAS No. 96-12-8) in F344 rats and B6C3F!
mice (inhalation study). NTP Technical Report No. 81-21. 173 pp.
[Also publ. as DHHS (NIH) 82-1762]
Ohta, T. et al. 1984. The SOS function-inducing activity of chemical
mutagens of Echerichia coli. Mutat. Res. 131:101-109.
OS HA. 1977. Occupational Safety and Health Administration. Occupational
exposure to 1,2-dibromochloropropane (DBCP). Proposed standard, hearing.
Federal Register. 42(21 0):57266-57283.
Potashnik, G., and D. Abelovich. 1985. Chromosomal analysis and health
status of children conceived to men during or following dibromochloro-
propane-induced spermatogenic suppression. Andrologia. 17:291-296.
Prival, M.J., E.G. McCoy, B. Gutter and H.S. Rosenkranz. 1977. Tris(2,3-
dibromophosphate) mutagenicity of a widely used flame retardant.
Science. 195:76-78.
Rao, K.S., F.J. Murray, A.A. Crawford et al. 1979. Effects of inhaled
1,2-dibromo-3-chloropropane (DBCP) on the semen of rabbits and the
fertility of male and female rats. Toxicol. Appl. Pharmacol. 48:A121.
Rao, K.S., J'.D. Burek, F.J. Murray et al. 1982. Toxicologic and reproduc-
tive effects of inhaled 1,2-dibromo-3-chloropropane in male rabbits.
Fund. Appl. Toxicol. 2(5):241-151.
Rao, K.S., J.D. Burek, F.J. Murray et al. 1983. Toxicologic and reproduc-
tive effects of inhaled 1,2-dibromo-3-chloropropane in rats. Fund.
Appl. Toxicol. 3(2):104-110.
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l,2-Dibrono-3-Chlorcpropane (DBCP) February 12, 1987
-21-
Ratpan, F., and H. Plauman. 1985. Mutagenicity of halogenated three-carbon
conpounds and their methylated derivatives. Environ. Mutagen. 7:15.
Reel, J.R., R. Wblkowski-Tyl, A.D. Lawton and J.C. Lamb. 1984. Dibromo-
chloroprcpane: Reproduction and fertility assessment in EO-1 mice when
administered by gavage. NTP-84-263. September 11.
Reznik, G., S.F. Stinson and J.M. Ward. 1980a. , Respiratory pathology in
rats and mice after inhalation of l,2-dibromo-3-chloropropane or
l,2^3ibronomethane for 13 weeks. Arch. Toxicol. 46(3-4) :233-240.
Reznik, G., H. Reznik-Schuller, J.M. Ward and S.F. Stinson. 1980b.
Morphology of nasal-cavity tumors in rats after chronic inhalation of
l,2-dibrono-3-chloropropane. Br. J. Cancer. 42:772-781.
Reznik, Y.B., and G.K. Sprinchan. 1975. Experimental data on the gonadotoxic
effect of Nemagon. Gig. Sanit. 101-102. (Translation)
Rosenkranz, H.S. 1975. Genetic activity of l,2-dibromo-3-chloropropane, a
widely used fumigant. Bull. Environ. Contain. Toxicol. 14(1):8-12.
Ruddick, J.A., and W.H. Newsome. 1979. A teratogenicity and tissue distri-
bution study on dibranochloropropane in the rat. Bull. Environ. Contain.
Toxicol. 21:483-487.
Russell, W.L. 1985. For some chemicals, genetic risks based on tests other
than germ-cell nutagenicity in the whole mammal may be exaggerated.
Environ. Mutagen. 7:78.
Saegusa, J., H. Hasegawa and K. Kawai. 1982. Toxicity of l,2-dibromo-3-
chloropropane (EBCP): 1. Histopathological examination of male rats
exposed to DBCP vapor. Ind. Health. 20(4): 315-323.
Saito-Suzuki, R., S. Teramoto and Y. Shirasu. 1982. Dominant lethal studies
in rats with l,2-dibromo-3-chloroprcpane and its structurally related
compounds. Mutat. Res. 101(4):321-327.
Selleck, R.E., F.H. Pearson, V. Diyamandoglu and Z.G. Ungun. 1983. Application
of air stripping technology for the removal of DBCP residues in community
and industrial water supplies. Report to Occidential Chemical Company,
Lathrop, Louisiana.
Stolzenberg, S.J., and C.H. Hine. 1979. Mutagenicity of halogenated and
oxygenated three-carbon compounds. J. Toxicol. Environ. Health.
5(6):1149-1158.
Suzuki, K., and I.P. Lee. 1981. Induction of aryl hydrocarbon hydroxylase
and epoxide hydrolase in rat liver, kidney, testis, prostate glands, and
stomach by a potent nematocide, l,2-dibromo-3-chloropropane. Toxicol.
Appl. Pharmacol. 58(1):151-155.
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l,2-Dibromo-3-Chloropropane (DBCP) February 12, 1987
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Teramoto, S., R. Saito, H. Aoyama and Y. Shirasu. 1980. Dominant lethal
nutation induced in male rats by l,2-dibromo-3-chloropropane (CBCP).
Mutat. Res. 77(l):71-78.
Tezuka, H., N. Ando, R. Suzuki, M. Terahata, M. Moriya and Y. Shirasu.
1980. Sister-chromatid exchanges and chromosomal aberrations in cultured
Chinese hamster cells treated with pesticides positive in micrcbial
reversion assays. Mutat. Res. 78(2):177-191.
Tofilon, P.J., R.P. Clement and W.N. Piper. 1980. Inhibition of the bio-
synthesis of rat testicular heme by l,2-dibromo-3-chloropropane. Biochem.
Phannacol. 29(19):2563-2566.
Torkelson, T.R., S.E. Sadek and V.K. Rowe. 1961. Toxicologic investigations
of l,2-dibromo-3-chloropropane. Toxicol. Appl. Phamacol. 3:545-559.
Traul, K.A., R.H. McKee and R.D. Phillips. 1985. The genetic toxicology of
1,2-dibrono-3-chloropropane, 1,2-dibromo-3-chloro-2-methylpropane/ and
l,2,2-tribromo-2-methylpropane. Environ. Mutagen. 7:17-18.
U.S. EPA. 1979a. U.S. Environmental Protection Agency. Carcinogen Assessment
Group's Re-evaluation of DBCP Risks Incorporating Recent Chronic Testing
Data. U.S. EPA, CAG. Unpublished report dated June 17, 1979.
U.S. EPA. 1979b. U.S. Environmental Protection Agency. Direct testimony of
Dr. Roy Albert dated September 5, 1979. FIFRA Docket No. 485.
U.S. EPA. 1979c. U.S. Environmental Protection Agency. Dibromochloropropane
(DBCP); Suspension order and notice of intent to cancel. Federal Register.
44(219):65135-65179. November 9.
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cides in drinking water, food, and air. Office of Drinking Water.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Drinking water
criteria document for l,2-Dibromo-3-chloropropane (DBCP). Office of
Drinking Water. Washington, D.C. ECAO-CIN-410. April.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Dibromochloropropane;
Intent to cancel registrat'ons of pesticide products containing dibromo-
chloropropane (DBCP). Federal Register. 50(6):1122-1130. January 9.
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drinking water regulations; Synthetic organic chemicals, inorganic chemicals
and microorganisms; Proposed rule. Federal Register. 50(219) :46934-47022.
November 13.
U.S. EPA. 1985d. U.S. Environmental Protection Agency. Method 502.1.
Volatile halogenated organic. compounds in water by purge and trap gas
chromatography. Environmental Monitoring and Support Laboratory.
Cincinnati, Ohio 45268. June.
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l,2-Dibromo-3-Chloroprqpane (DBCP) " ' February 12, 1987
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U.S. EPA. 1985e. U.S. Environmental Protection Agency. Method 524.1.
Volatile organic compounds in water by purge and trap gas chroma tography/
mass spectrometry. Environmental Monitoring and Support Laboratory.
Cincinnati, Ohio 45268. June.
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September 24.
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Pesticides. Bureau of Foods.
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icity of halogenated olefinic and aliphatic hydrocarbons in mice. J. Natl.
Cancer Inst. 63(6):1433-1439.
Vterren, D.W., et al. 1984. Effects of l,2-dibronD-3-chloropropane on male
reproductive function in the rat. Biol. Reprod. 31:454-463.
Whorton, D., R.M. Krauss, S. Marshall and T.H. Milby. 1977. Infertility in
male pesticide workers. Lancet. 2:1259-1261.
Zimmering, S. 1983. l,2-Dibromo-3-chloropropane (DBCP) is positive for sex-
linked recessive lethals, heritable translocations and chromosome loss
in Drosophila. Mutat. Res. 119(3-4):287-288.
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March 31, 1987
1,2-DICHLOROPROPANE
Health Advisory
Office of Drinking Hater
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. The chemical 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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1,2-Dichlorpropane March 31, 1987
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This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for 1,2-Dichloro-
propane (U.S. EPA, 1985a). The HA and CD formats are similar for easy
reference. Individuals desiring further information on the toxicological
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 #86-117850/AS.
The toll-free number is (800) 336-4700; in the Washington, D.C. area: (703)
487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 78-87-5
Structural Formula
Propylene dichlonde, 1,2-DCP
0 1,2-Dichloropropane has been used as a solvent for oils and fats,
dry cleaning and degreasing operations, and as a component of soil
fumigants.
Properties C)
Specific Gravity 1.15
Water Solubility 2,700 mg/L
Log Octanol/Water Partition 2.28
Coefficient
Odor Threshold (air) 420 mg/m3
Occurrence
0 1,2-Dichloropropane (1,2-DCP) is a volatile synthetic compound with
no natural sources. The major release of 1,2-DCP to the environment
will be from its use as a soil fumigant (U.S. EPA, 1983).
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0.1
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1,2-Dichlorpropane March 31, 1987
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0 1,2-DCP is expected to be a persistent and mobile compound in soil.
The major route of removal of 1,2-DCP from soil and surface waters is
by volatilization (U.S. EPA, 1979). 1,2-DCP has been shown to be
stable in some soil for years (Roberts, 1976). 1,2-DCP has been shown
to migrate in soil and has been reported as a contaminant in ground
water (Cohen, 1983). 1,2-DCP is expected to be removed from surface
water by volatilization. 1,2-DCP also has been shown to biodegrade
in water over a number of weeks. There is no available information on
1,2-DCP's potential for bioaccumulation.
0 1,2-DCP has been identified as a contaminant of both ground and surface
water. 1,2-DCP has been surveyed for in the Ground Water Supply Survey
and has been found in approximately 1-2% of rural wells at levels
around 1 ug/L. Local monitoring has found levels as high as 1,200
ug/L in shallow wells near sites where 1,2-DCP has been used as a soil
fUmigant (Cohen, 1983). 1,2-DCP also has been reported in the Delaware
river at levels of 20-30 ug/L (U.S. EPA, 1983). 1,2-DCP has been
identified as a contaminant in fish. 1,2-DCP also has been reported
in urban air at low levels, approximately 100 ppt. The available
data are insufficient to show whether drinking water is the major
route of exposure for 1,2-DCP.
III. PHARMACOKINETICS
Absorption
0 The results of various studies suggest that the absorption of 1,2-DCP
is approximately 90 percent of the orally administered dose (Hutson
et al., 1971).
Distribution
0 Although no specific data were located which quantified the distribu-
tion of 1,2-DCP in animals, approximately 0.5% of a radioactive dose
was recovered from the gut of animals within 96 hours, 1 .6% was
recovered from skin and 3.6% was detected in the carcass (Hutson
et al., 1971).
Metabolism
0 The metabolic end products of 1,2-DCP are predominantly N-acetyl-S-
(2-hydroxypropyl) cysteine and B-chloroacetate (Jones and Gibson,
1980).
Excretion
1,2-DCP was eliminated rapidly by rats dosed orally with 4 mgAg (Hutson
et al., 1971). Approximately 80% to 90% of the radioactivity was
excreted in the urine, feces and expired air of rats following dosing.
Urinary and fecal excretion accounted for 53% and 6%, respectively,
of the radioactivity recovered in the expired air of rats as carbon
dioxide; 23% was recovered as other volatile radioactivity.
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or:
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1,2-Dichloropropane March 31, 1987
-4-
IV. HEALTH EFFECTS
Humans
0 Data on the toxicity of this compound to humans are limited to a
single case of acute poisoning, reported only as an abstract (Larcan
et al., 1977). Centre- and mediolobular hepatic necrosis were observed
in a man who died 36 hours after ingesting approximately 50 ml of
a cleansing substance. The toxic material was found to contain 1,2-DCP,
but it is unclear whether it contained other compounds as well.
Animals
Short-term Exposure
0 The acute LD50 values for 1,2-dichloropropane are given below:
Route Species LD«;n Value Reference
Inhalation Rat 9224 mg/m3 Smyth et al. (1969)
Oral Rat 2200 mgAg Smyth et al. (1969)
Rat 2200 mgAg Ekshtat et al. (1975)
Dermal Rabbit 10,200 mgAg Smyth et al. (1969)
0 The acute toxicity of 1,2-DCP was determined following single oral
administration to fasted dogs of unspecified age and sex (Wright and
Schaffer, 1932). Doses of approximately 250 to 350 mgAg in the dogs
produced gastrointestinal irritation without any histologic changes
in the kidney. A dose of 580 mgAg produced swelling of the epithelial
cells of the kidney tubules and fatty infiltration in the convoluted
tubules. At an approximate dose of 5800 mgAg* there was a lack
of coordination and partial narcosis followed by death in one dog.
Necropsy revealed congestion in the lungs, kidney, and bladder,
hemorrhage in the stomach and respiratory tract, and fatty degeneration
of the liver and kidneys.
0 Heppel et al. (1946) reported no apparent signs of toxicity following
single 7-hour inhalation exposures to 6900 mg/m3 in rats, rabbits and
guinea pigs.
0 -nhalation exposure for 1 hour at 10,400 mg/m3 showed evidence of
slight visceral congestion, fatty liver and kidneys, liver glycogen
storage and marked necrosis of the adrenals (Heppel et al., 1946).
0 Drew et al. (1978) measured SCOT, SGPT, glucose-6-phosphatase and
ornithine carbamyl transferase enzymes in the serum of male rats
following a single 4-hour inhalation exposure to 1,2-DCP at a concen-
tration of 4620 mg/m3. A significant increase in enzyme activities
was observed for SGOT, SGPT and ornithine carbamyl transferase at
24 and 48 hours.
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, 2-Dicloropropane ^
-5-
Longer-term Exposure
0 Heppel et al., (1948) reported the results of multiple inhalation
exposures to 1,2-DCP by rats, mice, guinea pigs and rabbits to daily
7-hour exposure periods in the concentration range of 4400 mg/m3 to
10,400 mg/n>3. The concentration of 10,400 mg/m3 produced lethality in
over 50 percent of the animals. Gross and histopathological findings
included liver abnormalities such as visceral congestion, fatty
degeneration, extensive coagulation and necrosis in multilobular
areas. Renal tubular necrosis and fibrosis, splenic hemosiderosis,
pulmonary congestion, bronchitis, pneumonia and fatty degeneration of
the heart were observed among animals exposed to all concentrations.
0 The effects of 1,2-DCP on the functional state of the liver in rats
has been studied by Kurysheva and Ekshtat (1975). The groups of
animals were given daily oral doses of 1,2-DCP at 14.5 mgAg or
360 mgAg for 30 days. Levels of serum cholesterol, betalipoprotein
and gamma globulin increased after the 10th day following the daily
administration of both doses. By the 20th day of dosing, serum
cholinesterase activity was inhibited, whereas the fructose-1-mono-
phosphate aldolase, SGPT and SCOT activities were increased; after
30 days of dosing, only SGPT activity was inhibited.
0 Ekshtat et al., (1975) orally administered 1,2-DCP to rats at daily
doses of 8.8, 44 or 220 rag/kg for 20 days. The animals were reported
to have had disturbances in protein formation and hepatic enzyme and
lipid metabolism.
0 NTP (1983) dosed groups of female F344 rats and B6C3F1 male and
female mice with 1,2-DCP (0, 125 or 250 mg/kg/day) in corn oil by
gavage (5 days/week) for about 2 years (103 weeks). Groups of male
Fischer 344 rats were administered 1,2-DCP at 0, 62 or 125 mg/kg/day
in the same manner. Observations included survival, body weight,
overt signs of toxicity and gross and histological appearance of a
wide range of organs and tissues. In rats, survival was decreased
only among the females of the 250 mg/kg group. An increased incidence
of liver lesions (focal and centrilobular necrosis) and decreased
mean body weight also were observed in this group. At the 125 mg/kg
dose level, survival among rats was unaffected, but males had decreased
mean body weights and females had increased incidences of mammary
gland hyperplasia. No effects were observed in male rats given
62 mg/kg•
0 In mice (NTP, 1983), there was a decrease in survival rates among
females receiving both 125 and 250 mgAg 1,2-DCP. This was attributed,
in part, to an increased incidence of severe infections of the
respiratory tract for both low- and high-dose groups. The only other
non-neoplastic effects in mice were increased incidences of liver
lesions (hepatomegaly and focal and centrilobular necrosis) in males
receiving 125 or 250 mgAg*
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1,2-Dichloropropane March 31, 1987
-6-
Mutaqenicity
• A positive dose-related mutagenic response at concentrations of 10,
20 or 50 mg/plate 1,2-DCP was observed in Salmonella typhimurium
strains TA 1535 and TA 100 (DeLorenzo et al., 1977). No increase
in mutagenicity was seen following the addition of the S-9 liver
microsomal fraction.
Carcinogenicity
0 The NTP (1983) chronic gavage study (discussed under longer-term
exposure) is the only adequately designed carcinogenicity study
available. The results show that 1,2-DCP may be carcinogenic for
mice as indicated by dose-related increased incidences of hepato-
cellular adenomas in male and female mice. The incidences of
hepatocellular carcinomas were increased (not significantly) in males
and in females. Evidence of carcinogenicity in rats was equivocal.
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) = ~ ( u
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a csild
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
There are insufficient toxicological data available in the published
scientific literature to derive a One-day HA.
Ten-day Health Advisory
Three studies in animals have been considered for use in the calculation
of a Ten-day HA; these are Kurysheva and Ikshtat (1975); Ekshtat et al. (1975);
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1,2-Dichloropropane «~>O March 31, 1987
-7-
and NTP (1983). However, these studies lack some relevant toxicological data
necessary for use in this calculation. Kurysheva and Ekshtat (1975) reported
the effect of 1,2-DCP on the functional state of the liver of rats. The
groups of animals were given daily doses of 1,2-DCP at 14.5 mg or 360 mgAg
orally for 30 days. ' Increases in the serum levels of cholesterol, lipoprotein,
and gamma-globulin were noted after the tenth day following the daily dosing.
By day 20 of dosing, serum cholinesterase was inhibited, whereas fructose-1-
monophosphate aldolase, SGPT and SGOT enzyme activities were increased; after
30 days of dosing, only SGPT was inhibited. Other information such as strain,
number of animals, weight and age, as well as which doses caused what effects
were not reported.
The study by Ekshtat et al. (1975) is selected as the basis for the
Ten-day HA. The authors reported the results of orally administered 1,2-DCP
at dose levels of 8.8, 44 or 220 mgAg for 20 days. The investigators observed
disturbances in the animals' protein formation, hepatic enzyme levels and
lipid metabolism. The NAS (1979) in a request from the Office of Drinking
Nater provided a 7-day Suggested-No-Adverse-Response-Level (SNARL) for 1,2-DCP
based on the Ekshtat et al. (1975) study in rats. The following formula was
used to derive a 7-day level for a 70 kg adult consuming 2 liters water/day.
The NAS SNARL can be used as an interim Ten-day HA as well. The HA is derived
as follows:
For a 10 kg child, the level of 1,2-DCP would be:
Ten-day HA = (8.8 mgAq/day) (10 kg) - Q.088 mg/L = 0.090 mg/L
(1,000) (1 L/day)
or 90 ug/L
where:
8.8 mgA9/day = minimal effect level from the subacute ingestion
studies in rats.
10 kg = assumed weight of a 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 water consumption by a child.
Longer-term Health Advisory
There are no satisfactory toxicological data available from which to
calculate a Longer-term Health Advisory.
Lifetime^ Heaj.th Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of
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89
1,2-Dichloropropane March 31, 1987
-8-
noncarcinogenic adverse health effects over a lifetime exposure. The Life-
time 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
estimate 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 classifica-
tion scheme of carcinogenic potential (U.S. EPA, 1986), then caution should
be exercised in assessing the risks associated with lifetime exposure to this
chemical.
Only one chronic ingestion study (NTP, 1983) has been carried out for
1,2-DCP. This study was designed primarily to investigate carcinogenic
effects. This study may provide some data on non-carcinogenic effects which
may be considered for a Lifetime HA in absence of other chronic animal studies.
However, it should be noted that the NTP (1983) study is recently audited and
there are some changes as a result of this audit. These changes are being
evaluated before its consideration for a Lifetime HA for 1,2-DCP.
Evaluation of Carcinogenic Potential
0 The dose-response data for hepatocellular adenoma and carcinoma in
B6C2F1 mice (NTP, 1983) are used for a quantitative assessment of
cancer risk from exposeure to 1,2-DCP. Based on these data and using
a linearized multistage model, a carcinogenic potency factor (q-j*) for
humans of 6.33 x 10~2 (mg/kg/day)~1 was calculated from the data for
male mice and a q^ of 2.25 x 10~2 (mg/kg/day)~1 was calculated from
the data for female mice. The higher of the two values is the appropriate
basis for the estimation of cancer risk levels. The doses corresponding
to increased lifetime excess cancer risks for a 70 kg human of 10~4,
10-5 and 10-6 are 1.11 x 10-1, 1.11 x 10~2 and 1.11 x 10~3 mg/day,
respectively. Assuming a water consumption level of 2 liters per day,
the corresponding concentrations of 1,2-DCP in drinking water at 5.6 x
10-2, 5.6 x 10"3 and 5.6 x 10~4 mg/L, respectively. However, it should
be noted that these risk assessments for 1,2-DCP are based on the
results of a carcinogenicity bioassay in animals reported in the NTP
(1983) draft report. An audit has been completed and minor changes
noted in the audit will be incorporated in the near future.
0 Cancer risk estimates (95% upper limit) with other models are presented
for comparison with that derived with the multistage. For example,
one excess cancer per 1,000,000 (10"6) is associated with exposure to
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100
1,2-Dichloropropane March 31, 1987
-9-
1,2-DCP in drinking water at levels of 0.5 ng/L (Probit), 0.002 mg/L
(Logit), and 0.0002 mg/L (Weibull).
0 IARC has not assessed 1,2-DCP for its carcinogenic potential.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1966), 1,2-DCP is classified in Group C:
Possible human carcinogen. This category is for agents with limited
evidence of carcinogenic!ty in animals in the absence of human data.
(However, the Carcinogen Assessment Group upgraded the classification
of 1,2-DCP (to Group B2) on February 26, 1987. The final decision
be incorporated in the near future.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The ACGIH (1983) has adopted a TLV of 75 ppm ( 350 mg/m3) and STEL of
110 ppm { 500 mg/m3) for 1,2-DCP in workroom air. The TLV
represents a TWA concentration for an 8-hour day or 40-hour workweek.
The TLV and STEL are based primarily on the data of Heppel et al.
(1946, 1948).
0 The U.S. EPA (1980) concluded that data regarding the toxicity of
1,2-DCP were insufficient for the derivation of an ambient water
quality criterion for- the protection of human health.
VII. ANALYTICAL METHODS
0 Analysis of 1,2-DCP is by a purge-and-trap gas chromatographic procedure
used for the determination of volatile organohalides in drinking water
(U.S. EPA, 1985b). This method calls for the bubbling of an inert
gas through the sample and trapping 1,2-DCP on an adsorbant material.
The adsorbant material is heated to drive off the 1,2-DCP onto a gas
chromatographic column. The applicable concentration range for this
method has not been determined. Confirmatory analysis for 1,2-DCP
is by mass spectrometry (U.S. EPA, 1985). The detection limit for
confirmation by mass spectrometry is 0.2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies which have been shown to be effective in
removing 1,2-DCP from drinking water are adsorption on granular
activated carbon (GAG) and ion exchange. Other methods which are
expected to be effective for removal of 1,2-DCP from water are air
stripping and boiling.
0 Granular activated carbon (GAG) and powdered activated carbon (PAG)
have been tested for their effectiveness in removing 1,2-dichloropro-
pane. Dobbs and Cohen (1980) developed adsorption isotherms for DCP.
They reported that Filtrasorb* 300 carbon exhibited adsorption capa-
cities of 5.9 mg of. DCP per gram of carbon at equilibrium concentration
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101
1,2-Dichloropropane March 31, 1987
-10-
of 1.0 mg/L and 1.5 mg of DCP per gram of carbon at equilibrium
concentration of 0.1 mg/L.
Isotherm studies using Filtrasorb* 400 carbon reported carbon loadings
of 240 mg DCP/gm of carbon and 480 mg DCP/gm of carbon at equilibrium
concentrations of 100 mg/L and 1,000 mg/L, respectively. No usage or
loading rates were available (U.S. EPA, 1985d).
Removal of DCP by air stripping is expected to be effective. Several
methods of air stripping have been tested. Air stripping in a column
packed with 1/4" Ceramic Intalox Saddle, proved to be effective in
removing chloroform with a Henry's Law Constant of 3.4 x 10~3 atm-mV
mole and 1,2-dichloroethane with a Henry's Law Constant of 1.1 x 10-3
atm-m3/mole. Although no actual performance data have been provided
for removing DCP by this treatment system, its Henry's Law Constant of
2 x 10-3 atm-m3/mole is an indication that this chemical is amenable to
air stripping (Love & Eilers, 1982; Singley and Bilello, 1981; McCarty
and Sutherland, 1979).
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102
1,2-Dichloropropane March 31, 1987
-11-
IX. REFERENCES
ACGIH. 1983. American Conference of Governmental Industrial Hygienists.
Threshold limit values for chemical substances and physical agents in
the work environment with intended changes for 1983-1984. Cincinnati,
OH. p. 30.
Cohen, D.B., D. Gilmore, C. Fischer and G.W. Bowes. 1983. 1,2-Dichloropropane
and 1,3-dichloropropane. Prepared for State of California, Water Resources
Control Board, Sacramento, CA.
DeLorenzo, P., S. Degl'Innocenti, A. Ruocco, L. Silengo and R. Cortese.
1977. Mutagenicity of pesticides containing 1,3-dichloropropane.
Cancer Res. 37:1915-1917.
Dobbs, R.A., and J.M. Cohen. 1980. Carbon adsorption isotherms for toxic
organics. U.S. EPA, Contract No. EPA-600/8-80-023.
Drew, R.T., J.M. Patel and F.N. Lin. 1978. Changes in serum enzymes in rats
after inhalation of organic solvents singly and in combination. Toxicol.
Appl. Pharmacol. 45:809-819.
Ekshtat, B. Ya., N.G. Kurysheva, V.N. Fedyanina and M.N. Pavlenko. 1975.
Study of the cumulative properties of substances at different levels of
activity. Uch. Zap.-Mosk. Nauchno-Issled. Inst. Gig. 22:46-48.
Heppel, L.A., P.A. Neal, B. Highman and V.T. Potterfield. 1946. Toxicology
of 1,2-dichloropropane. I. Studies on effects of daily inhalations.
J. Ind. Hyg. Toxicol. 28:1-8.
Heppel, L.A., B. Highman and E.Y. Peake. 1948. Toxicology of 1,2-dichloro-
propane. IV. Effects of repeated exposures to a low concentration of
the vapor. J. Ind. Hyg. Toxicol. 30:189-191.
Hutson, D.H., J.A. Moss and B.A. Pickering. 1971. Excretion and retention
of components of the soil fumigant D-D and their metabolites in the rat.
Food Cosmet. Toxicol. 9(5):677-680.
Hwang, S.T., and P. Fahrenthold. 1980. Treatability of the organic priority
pollutants by steam stripping. The American Institute of Chemical
Engineers.
Jones, A.R., and J. Gibson. 1980. 1,2-Dichloropropane: Metabolism and fate
in the rat. Xenobiotica. 10:835-846.
Kurysheva, N.G., and B.Y. Ekshtat. 1975. Effect of 1,3-dichloropropylene
and 1,2-dichloropropane on the functional state of the liver in animal
experiments. Uch. Zap.-Mosk. Nauchno-Issled. Inst. Gig. 22:89-92.
(CA 86:115725).
Larcan, A., H. Lambert, M.C. Kaprevok and B. Gustin. 1977. Acute poisoning
induced by dichloropropane. Acta. Pharmacol. Toxicol. 41:330. (Abstr.)
-------
1,2-Dichloropropane j-^"^ March 31, 1987
-12-
Love, O.T., Jr., and R.G. Eilers. 1982. Treatment of drinking water containing
trichloroethylene and related industrial solvents. AWWA.
McCarty, P.L., and K.H. Sutherland. 1979. Volatile organic contaminants
removal by air stripping. Paper presented at the Seminar on Controlling
Organics in Drinking Hater, American Water Works Annual Conference, San
Francisco, CA.
MAS. 1979. National Academy of Sciences. Emergency Response Report on 1,2-
Dichloropropane.
NTP. 1983. National Toxicology Program. NTP technical report on the carcino-
genicity bioassay of 1,2-dichloropropane (CAS No. 78-87-5) in F344/N
rats and B6C3FJ mice (gavage study). May. NIH Publ. No. 83-2519. Draft.
Final Technical Report in Preparation (Management Status Report, 6/10/86).
Perry, R.H., and C.H. Chilton. 1973. Chemical Engineers Handbook, 5th
edition, McGraw Hill Book Company, New York.
Roberts, T.R., and G. Stoydin. Degradation of (Z) and (E) 1,3-Dichloropropane
and 1,2-dichloropropane in soil. Pestic. Sci. 7:325-335.
Singley, J.E., and L.J. Bilello. 1981. Advances in the development of design
criteria for packed column aeration. Environmental Science and Engi-
neering, Inc.
Smyth, H.F., Jr., C.P. Carpenter, C.S. Weil, U.C. Pozzani, J.A. Striegel and
J.S. Nycum. 1969. Range-finding toxicity data, VII. Am. Ind. Hyg.
Assoc. J. 30(5):470-476.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Water related environ-
mental fate of 129 Office of Water Planning and Standards. EPA-440/4-79-
-029.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality
criteria for dichloropropanes/ propenes. Environmental Criteria and
Assessment Office, Cincinnati, OH. EPA 440/5-80-043. NTIS PB81-117541.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Philadelphia Geographic
Area Multimedia Pollutant Survey, Integrated Environmental Management
Di\ision. Washington, DC.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Draft drinking
water criteria document for 1,2-dichloropropane. Office of Drinking
Water.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Method 502.1,
Volatile halogenated organic compounds in water by purge and trap gas
chromatography. Environmental Monitoring and Support Laboratory, Cin-
cinnati, OH. 45268.
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104
1,2-Dichloropropane March 31, 1987
-13-
U.S. EPA. 1985c. U.S. Environmental Protection Agency. Method 524.1,
Volatile organic compounds in water by purge and trap gas chromato-
graphy/mass spectrometry. Environmental Monitoring and Support Labora-
tory* Cincinnati, OH. 45266.
U.S. EPA. 1985d. U.S. Environmental Protection Agency. Treatment techniques
available for removal of 1,2-dichloropropane. (Draft) Science and
Technology Branch, Criteria and Standards Division, Office of Drinking
Hater.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 51(185):33992-34003.
September 24.
Wright, W.H., and J.M. Schaffer. 1932. Critical antihelminthic tests of
chlorinated alkyl hydrocarbons and a correlation between the antihel-
minthic chemical structure and physical properties. Am. J. Hyg.
16:325-428.
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March 31, 1987
2 , 4-DICHLOROPHENOXYACETIC ACID
Health Advisory
Office of Drinking Hater
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. The chemical 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 ris . 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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2,4-Dichlorophenoxyacetic Acid March 31, 1987
-2-
This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for 2,4-Dichloro-
phenoxyacetic Acid (U.S. EPA, 1985a). The HA and CD formats are similar for
easy reference. Individuals desiring further information on the toxicological
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 #86-117884/AS.
The toll-free number is (800) 336-4700; in the Washington, D.C. area: (703)
487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 94-75-7
Structural Formula
r~\
-( 0 V-O-C
:H2-C-OH
Cl
Synonyms
0 Amidox, Amoxone, Aqua Kleen, 2,4-D
Uses
0 Herbicide on wheat, corn, rangeland/pasture, sorghum, barley and lawns.
Properties (Weast, 1980; Weed Science Society of America, 1979; Sigworth, 1965)
Chemical Formula CgHgp3Cl2
Molecular Weight 221
Physical State White crystalline powder
Boiling Point —
Melting Point 138°C
Density —
Vapor Pressure —
Water Solubility 540 mg/L
Log OctanoI/Water Partition —
Coefficient
Taste and Odor Threshold (water) 3.13 mg/L
Conversion Factor ~
Occurrence
0 2,4-Dichlorophenoxyacetic acid (2,4-D) is a systemic herbicide widely
used to control broadleaf weeds. It has a large production volume,
estimated to be between 53 and 65 million Ibs in 1982 and is used
directly and in the form of various salts and esters. 2,4-D is used
on wheat, corn, rangeland/pasture, sorghum and barley.
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107
2,4-Dichlorophenoxyacetic Acid March 31, 1987
-3-
2,4-D is degraded in the environment and is not considered to be a
persistent compound. It is metabolized by plants with half lives
of 1-3 weeks, is degraded readily by soil bacteria and undergoes
hydrolysis under environmental conditions. 2,4-D is reported to
have a half life of from 1-6 weeks in soil. Degradation in surface
waters is more variable with half lives ranging from a few days to
several months. Once in the soil, 2,4-D and some of its salts and
esters have been demonstrated to migrate. 2,4-D does not tend to
accumulate in soils and is reported not to bioaccumulate in plants
and animals.
2,4-D has been included in a number of national and regional surveys.
2,4-D has been detected in only a small number of drinking water
supplies. Reported levels of contamination have been below 0.5 ppb,
with most levels below 0.1 ppb. Contamination has occurred more
frequently in surface waters than ground waters. Contamination of
surface waters appears to be the result of surface water runoff from
agricultural usage. The Agency has received no report that a drinking
water supply has exceeded the MCL of 100 ppb.
2,4-D has been reported to occur in some foods in surveys taken in
the early 1970's. More recent surveys have failed to find detectable
levels of 2,4-D. Although large numbers of tolerances exist on food
crops, the available data are insufficient to determine whether food
or water is the greater source of exposure for 2,4-D.
III. PHARMACOKINETICS
Absorption
2,4-D is absorbed almost completely after ingestion. Khanna and Fang
(1966) reported that 93 to 96% of an oral dose of 3 to 30 mg/kg of
l^c-2,4-D (acid) to rats was excreted almost entirely in urine within
24 hours of dosing.
Distribution
2,4-D acid is distributed into blood, liver, kidney, heart, lungs and
spleen with lower levels occurring in muscle and brain. Peak concentra-
tions of 1 ^-2,4-0 were reached between six and eight hours at a dose
level of 1 mg/kg by gavage, with no detectable radioactivity after 24
hours (Khanna and Fang, 1966).
Metabolism
The data indicate that 2,4-D does not undergo biotransformation to
any great extent. Of five men who ingested 5 mg/kg of 2,4-D, four
excreted between 4.8 and 27.1% of the administered dose as conjugated
2,4-D. The rest of the 2,4-D excreted (82%) was unchanged (Sauerhoff
et al., 1977).
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108
2,4-Dichlorophenoxyacetic Acid March 31, 1987
-4-
Excretion
Fedorova and Belova (1974) reported that, following oral administration
of 14C-2,4-D to rats at a level of 0.05 mg/lcg/ 92.1% of the admini-
stered dose was excreted in the urine within 3 days, while 6.1% of
the radioactivity was detected in the feces in this time period.
IV. HEALTH EFFECTS
Humans
A male agricultural student who ingested at least 6 g of a commercial
herbicide preparation of the dimethyl amine salt of 2,4-D (50% by
weight) died after vomiting and convulsions. Pathological examina-
tion showed degenerative ganglion cell changes in the brain (Nielson
et al., 1965).
Occupational exposure to 2,4-D (along with other chemicals such as
2,4,5-TP and 2,4,5-T) resulted in reduced nerve conduction velocities
(Singer et al., 1982).
Case-controlled epidemiological studies of populations in Scandinavian
countries exposed to the phenoxy herbicides (as well as other chemicals
and contaminants) indicate excess risk of the development of soft-
tissue sarcomas and malignant lymphomas (Bardell et al., 1981).
Animals
Short-term Exposure
Acute oral LD5QS in the range of approximately 350-500 mg/kg of 2,4-D
acid have been reported for rats, mice and guinea pigs. There does
not appear to be significant differences in toxicity between the free
acid and the various salt and ester derivatives. LD5QS in the range
of 300 to 1000 mg/kg have been reported for 2,4-D compounds (U.S. EPA,
1985a).
Hill and Carlisle (1947) determined oral LD50s of 666, 375, 800 and
1000 mg/kg for 2,4-D sciium salt in rats, mice, rabbits and guinea pigs,
respectively; the maximum doses in these species not causing death
were 333, 125, 200 and 333 mg/kg, respectively.
Drill and Hiratzka (1953) reported an LDso of 100 mgAg in dogs with
pathologic changes of gastrointestinal mucosa irritation, moderate
hepatic necrosis and mild renal tubular degeneration.
Long-term Exposure
In a 90-day feeding study by Hazelton Laboratories (1983), doses of
5, 15 or 45 mg/kg bw/day to rats resulted in significant reductions
in blood indices at all doses; liver enzyme activities were reduced
at higher doses; kidney toxicity also was evident at higher doses.
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2,4-Dichlorophenoxyacetic Acid -*-<-^ March 31, 1987
-5-
These included increased homogeneity and altered tinctorial properties
of the cytoplasm and fine vacuolization of the cytoplasm in the renal
cortex. Other effects of higher doses included gastrointestinal irri-
tation and mild liver effects (e.g., cloudiness, swelling, increased
weights), as well as mortality and other characteristic overt signs
of toxicity. A NOAEL of 1 mgAg bw/day was identified.
Reproductive Effects
0 Increased preweanling mortality and weight loss were observed in the
offspring of rats that were exposed to 1500 ppm levels (approximately
75 mgAg bw) of 2,4-D in the diet in a 3-generation reproduction
study, but adverse effects on litter size or fertility were not
observed. No adverse effects were reported at lower doses (100 or
500 ppm) (Hansen et al., 1971).
0 Another reproduction study using 2,4-D acid in Fischer rats at doses
of 5, 20 or 80 mgAg/day in the diet indicated a maternal and fetotoxic
NOAEL of 5 mgAg/day. Effects at the next higher dose (20 mgAg/day)
included a decrease in maternal body weight and a reduced pup weight
(U.S. EPA, 1986b).
Developmental Effects
0 The teratogenic and embryotoxic effects of 2,4-D and several deriva-
tives of 2,4-D have been investigated in several species including
mice, rats and hamsters. Overall, 2,4-D and its derivatives appear
to be embryotoxic but only weakly teratogenic or nonteratogenic.
Oral doses (expressed as 2,4-D) of 124 mgAg/day in CD-1 mice (days
7-15 of gestation, Courtney, 1977), 75-125.5 mgAg/day in various
strains of rats (days 6-15 of gestation, Schwetz et al., 1971; Unger
et al., 1981; Khera and McKinley, 1972) and 40-100 mgAg/day in Golden
Syrian hamsters (days 6-10 of gestation, Collins and Williams, 1971)
produced fetotoxic effects (as evidenced by decreased fetal weights
and/or increased fetal mortality) or malformations (cleft palate and
other skeletal malformations (cited in U.S. EPA, 1985a).
0 Schwetz et al. (1971) indicated a NOAEL of 25 mgAg/day in rats for
2,4-D and its propylene glycol butyl ester (PGBE) and isooctyl esters.
These authors classified all of the anomalies as embryot.xic or feto-
toxic effects rather than as teratogenic responses because none of
these anomalies adversely affected either fetal or neonatal development.
0 Another study in Fischer 344 rats using 2,4-D acid at maternal doses
of 8, 25 or 75 mgAg/day reported a maternal NOAEL of 75 mgAg/day
and a fetotoxic NOAEL of 25 mgAg/day (U.S. EPA, 1986b).
Mutagenicity
0 2,4-D was not mutagenic in the Salmonella typhimurium reversion assay
using strains 1535 and 1538, at concentrations of 0.3 to 0.8 mg/mL,
without metabolic activation (Zetterberg et al., 1977).
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2,4-Dichlorophenoxyacetic Acid March 31, 1987
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• 2,4-D was shown to cause a dose-dependent increase in gene conversion
and cellular toxicity when tested in the Saccaromyces cerevisiae
assay at low pH with concentrations of 0.1 to 0.6 mg/mL and without
metabolic activation (Zetterberg et al., 1977). At neutral pH,
neither effect was observed in this system.
Carcinogenicity
0 Available data from laboratory animals have not provided a sufficient
demonstration of carcinogenicity of 2,4-D although increased tumor
production of a non-specific nature is suggested (U.S. EPA, 1985a).
0 The Agency is currently reviewing the results of an oncogenicity
study conducted in rats to make a final determination on its
significance (U.S. EPA, 1986b).
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) . /L( /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL * No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BH - assumed body weight of a child (10 kg) or
an adult (70 kg).
UF « uncertainty factor (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
A One-day HA can be calculated from the tolerated single dose for mice
(125 mg 2.4-D sodium saltAg bw» 2,4-0 equivalent to about 114 mgAg bw) from
the Hill and Carlisle (1947) study using an uncertainty factor of 1,000.
This factor represents two 10-fold factors for both intra- and interspecies
variability in the toxicity of a chemical when specific data are lacking and
an additional 10-fold factor because the tolerated single dose is assumed to
have caused unreported adverse effects and is, therefore, considered a LOAEL
rather than a NOAEL (Hill and Carlisle, 1947).
For a 10 kg child, the One-day HA is calculated as follows:
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2,4-Dichlorophenoxyacetic Acid March 31, 1987
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One-day HA- - (114 mqAg/day) (10 kg) , , , /L (1 100 /L)
(1,000) (1 L/day) y/
where:
114 mgAg/day « tolerated dose in mice (assumed to be LOAEL).
10 kg « assumed body weight of a 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.
This HA is equivalent to 1.1 mg/day or 0.1 mgAg bw/day.
Ten-day Health Advisory
The Rowe and Hymas (1954) report is used to estimate the Ten-day Health
Advisory. They administered 2,4-D in the diet at 0, 100, 300 or 1000 ppm to
groups of five young female rats for 114 days. If it is assumed that young
rats consume 10% of their body weight as food per day, the corresponding
daily doses would be 0, 10, 30, and 100 mgAg bw/day. No effects were found
at 10 or 30 mgAg bw/day, but 100 mgAg bw/day produced "excessive mortality"
with depressed growth rate, slightly increased liver weights, and slight
cloudiness and swelling of the liver. Rats exposed to higher levels of 2,4-D
in the diet (3,000 and 5,000 ppm) were not evaluated because they refused
food and consequently lost weight. Both of the above Dow Chemical Company
studies used small groups of animals and were not reported in detail, but
multiple dose levels were tested and a number of toxicity indices were
evaluated.
Using the same assumptions as in the One-day HA calculation, a Ten-day
Health Advisory is calculated as follows:
Ten-day HA (child) * (30 mgAg/day) (10 kg) , 0>30 mg/L (30o Ug/L)
(1,000) (1 L/day)
where:
30 mgAg/day * NOAEL.
10 kg * assumed body weight of a child.
1,000 =« uncertain!ty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study
with deficiencies in the study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
A Longer-term HA has not been calculated due to the lack of appropriate
data.
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2,4-Dichlorophenoxyacetic Acid March 31, 1987
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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 1 0%
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.
A Lifetime Health Advisory has been developed for 2,4-D based on an
interim report on a 90-day experiment with rats by Hazelton Laboratories
(1983). In this study, a NOAEL of 1.0 mg/kg "as established for blood, renal
and hepatic effects. An uncertainty factor of 1,000 should be used in these
calculations, representing a 10-fold factor for both intra- and interspecies
variability to the toxicity of a chemical when specific data are lacking and
an additional 10-fold factor because the results are from a subchronic study.
However, 100-fold uncertainty factor is used to calculate a tentative Lifetime
HA since the preliminary report suggests that 1 mg/kg 'may be NOAEL at the end
of the 2-year study. If at the end of the 2-year experiment there is no
change in the NOAEL, an uncertainty factor of 100 can be applied to calculate
the HA. Based on currently available data, however, a Lifetime HA for a 70 kg
man can be calculated as follows:
Step 1: Determination of a Reference Dose (RfD)
RfD = d mgAg/day) = 0.01 mgAg/day
(100) "'
where:
1 mg/kg/day = NOAEL, based on absence of blood, renal and hepatic
effects in rats.
100 = Uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
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2,4-Dichlorpphenoxyacetic Acid March 31, 1987
-9-
Step 2: Determination of a Drinking Water Equivalent Level (DWEL)
DWEL « .(0.01 mgAq/day) (70 kg) B 0.350 /L (350 /L)
(2 L/day)
where:
0.01 mgAg/day - RfD
70 kg * assumed body weight of an adult.
2 L/day * assumed daily water consumption of an adult.
Step 3: Determination of a Lifetime Health Advisory
Lifetime HA » (0.350 mg/L) (20%) * 0.070 mg/L (70 ug/L)
where:
0.350 mg/L « DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 IARC (1982) has classified 2,4-D into Group 3, indicating its inability
to assess carcinogenic potential to humans.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), 2,4-D may be classified in Group D:
Not classified. This category is for agents with inadequate animal
evidence of carcinogenicity.
VI. OTHER CRITERIA, ^UIDANCE AND STANDARDS
0 The interim primary drinking water standard for 2,4-D is 0.1 mg/L
(Federal Register, 1975).
0 The National Academy of Sciences has suggested an acceptable level of
0.09 mg/L for 2,4-D in drinking water, assuming that 20% of exposure
is attributable to drinking water (HAS, 1977). This level was calcu-
lated from a NOEL from the Hansen et al. (1971) 2-year feeding study
with dogs.
0 The American Conference of Governmental Industrial Hygienists (ACGIH)
currently recommends an 8-hour time-weighted average, threshold limit
value (TWA-TLV) of 10 mg/m^ for occupational exposure to 2,4-D (ACGIH,
1980). ACGIH also recommends a short-term exposure level (STEL) of
20 mg/m3 for any 1 5-jninute exposure period. These recommendations
are intended to protect against local and systemic effects by inhalation
and are derived from unspecified ingestion studies.
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2,4-Dichlorophenoxyacetic Acid March 31, 1987
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0 Established tolerances for residues of 2,4-D are 1 ppm from applica-
tion of its dimethylamine salt for water hyacinth control in slow
moving aquatic media (e.g.. Western United States irrigation ditch
banks) and in fish and shellfish (U.S. EPA, 1982).
0 An Acceptable Daily Intake (ADI) of 2,4-D for -man has been recommended
as <0.03 mgAg by the Joint Meeting of the FAO Working Party of
Experts on Pesticide Residues and the WHO Expert Committee on Pesticide
Residues (WHO, 1977), after considering published experimental data
and national tolerances established by several countries.
0 The World Health Organization has recommended a value of 0.1 mg/L
in drinking water for 2,4-D (WHO, 1984).
VII. ANALYTICAL METHODS
0 Determination of 2,4-D is by a liquid-liquid extraction gas chromato-
graphic procedure (U.S. EPA, 1978; Standard Methods, 1985). Specifi-
cally, the procedure involves the extraction of chlorophenoxy acids
and their esters from an acidified water sample with ethyl ether.
The esters are hydrolyzed to acids and extraneous organic material is
removed by a solvent wash. The acids are converted to methyl esters
which are extracted from the aqueous phase. Separation and identifi-
cation of the esters is made by gas chromatography. Detection and
measurement is accomplished by an electron capture, microcoulometric
or electrolytic conductivity detector. Identification may be corrobo-
rated through the use of two unlike columns. The detection limit is
dependent on the sample size and instrumention used. Typically, using
a 1-L.sample and a gas chromatograph with an electron capture detector
results in an approximate detection limit of 50 ng/L for 2,4-D.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies which are capable of removing 2,4-D from
drinking water include adsorption by granular (GAC) or powdered
activated carbon (PAC) and reverse osmosis (RO).
0 Aly and Faust (1965) developed adsorption isotherms for several 2,4-D
compounds and 2,4-DCP in drinking water. They reported -hat the
activated carbon Aqua Nuchar exhibited adsorptive capacities of 0.118
mg, 0.032 mg and 0.009 mg of 2,4-D per gm carbon at equilibrium
concentrations of 1,000 ug/L, 100 ug/L and 10 ug/L, respectively.
The results indicate that the sodium salt of 2,4-D is much less
easily adsorbed than the 2,4-D esters. Another bench-scale study
conducted at the Agricultural University of Wazeningen, The Netherlands,
investigated the use of flocculated PAC in water treatment for several
compounds, including 2,4,-D (U.S. EPA, 1985b). The results of the
experiments revealed that 2,4-D adsorption on the flocculated carbon
was higher than the non-flocculated carbon.
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2,4-Dichlorpphenoxyacetic Acid March 31, 1987
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Edwards and Schubert (1974) evaluated the selectivity of cellulose
acetate RO membrane for several derivatives of 2,4-D in aqueous solu-
tion. All tests were performed in batches with RO membranes from
four different manufacturers. The results showed a range of removal
of 1 to 65% from an initial 2,4-D (sodium salt) concentration of 50
mg/L. Further investigations are required to verify the removal
efficiencies of RO treatment of 2,4-D in water.
Conventional treatment, such as coagulation/filtration, has been
tested for the removal of certain SOCs, including 2,4-D (Aly and
Faust, 1965). The results of the study indicated that conventional
treatment consisting of coagulation/filtration is not effective for
2,4-D removal*
Treatment technologies for the removal of 2,4-D from drinking water
have not been extensively evaluated (except on an experimental level).
An evaluation of some of the physical and/or chemical properties of
2,4-D indicates that the following techniques would be candidates for
further investigation: powdered activated carbon (PAC) adsorption
and reverse osmosis (RO). Individual or combinations of technologies
for 2,4-D reduction must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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2,4-Dichlorophenoxyacetic Acid March 31, 1987
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IX. REFERENCES
Aly, O.M., and S.D. Faust. 1965. Removal of 2,4-dichlorophenoxyacetic acid
derivatives from natural waters. JAWWA. 57:221-230.
ACGIH. 1980. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 4th ed., with supplements through 1981. Cincinnati, OH. pp. 117-118.
Collins, T.F.X., and C.H. Williams. 1971. Teratogenic studies with 2,4,5-T
and 2,4-D in the hamster. Bull. Environ. Contain. Toxicol. 6(6) :559-567.
Courtney, K.D. 1977. Prenatal effects of herbicides: Evaluation by the
prenatal development index. Arch. Environ. Contarn. Toxicol. 6:33-46.
Drill, V., and T. Hiratzka. 1953. Toxicity of 2,4-dichlorophenoxyacetic
acid and 2,4,5-trichlorophenoxyacetic acid in dogs. AMA Arch. Ind. Hyg.
Occup. Med. 7:61-67.
Edwards, V.H., and P.F. Schubert. 1974. Removal of 2,4-D and other persistent
organic molecules from water supplies by reverse osmosis. JAWWA.
13:610-616.
Federal Register. 1975. Vol. 40, 0. 59566.
Fedorova, L.M., and R.S. Belova. 1974. Incorporation of 2,4-D into animal
organs. Paths and dynamics of its excretion. Gig. Sanit. 2:105-107.
(Translation for U.S. EPA by Literature Research Company TR-79-1000)
Hansen, W.H., M.L. Quaife, R.T. Habermann and O.G. Fitzhugh. 1971. Chronic
toxicity of 2,4,-dichlorophenoxyacetic acid in rats and dogs. Toxicol.
Appl. Pharmacol. 20(1):122-129.
Hardell, L., M. Eriksson, P. Lenner and E. Lundgren. 1981. Malignant lymphoma
and exposure to chemicals especially organic solvents, chlorophenols and
phenoxy acids. A case control study. Br. J. Cancer. 43:169-176.
Hazelton Laboratories. 1983. Document Accession Number 251473. U.S. EPA,
Office of Pesticides Programs, Washington, D.C.
Hill, E.V., an-1 H. Carlisle. 1947. Toxicity of 2,4-dichlorophenoxyacetic
acid for experimental animals. J. Ind. Hyg. Toxicol. 29:85-95.
IARC. 1982. International Agency for Cancer Research. IARC monographs
on the evaluation of the carcinogenic risk of chemicals to humans.
Supplement 4, Lyon, France.
Khanna, S., and S.C. Fang. 1966. Metabolism of C~14 labeled 2,4-dichloro-
phenoxyacetic acid in rats. J. Agric. Food Chem. 14(5):500-503.
Khera, K.S., and W.P. McKinley. 1972. Pre- and post-natal studies on 2,4,5-
trichlorophenoxyacetic acid and 2,4-dichlorophenoxyacetic acid and their
derivatives in rats. Toxicol. Appl. Pharmacol. 22:14-28.
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117
2,4-Dichlorophenoxyacetic Acid March 31, 1987
-13-
NAS. 1977. National Academy of Sciences .'Drinking Water and Health.
Volume 1. National Academy Press. Washington, D.C.
Nielson, K., B. Kaempe and J. Jensen-Holm. 1965. Fatal poisoning in man by
2,4-D; Determination of the agent in forensic materials. Acta. Pharmacol.
Toxicol. 22:224-234.
Rowe, V.K., and T.A. Hymas. 1954. Summary of toxicological information on
2,4-D and 2,4,5-type herbicides and an evaluation of the hazards to
livestock associated with their use. Am. J. Vet. Res. 15:622-629.
Sauerhoff, M.W., W.H. Braun, G.E. Blau and P.J. Gehring. 1977. The fate of
2,4-dichlorophenoxyacetic acid (2,4-D) following oral administration to
man. Toxicology. 8(1):3-11.
Schwetz, B., G.L. Sparschu and P.J. Gehring. 1971. The effect of 2,4-D and
esters of 2,4-D on rat embryonal, fetal and neonatal growth and develop-
ment. Food Cosmet. Toxicol. 9:801-817.
Sigworth, E. 1965. Identification and removal of herbicides and pesticides.
J.A.W.W.A. 55:1016-1022.
Singer, R., M. Noses, J. Valciukas, R. Lilis and I.J. Selikoff. 1982. Nerve
conduction velocity studies of workers employed in the manufacture of
phenoxy herbicides. Environ. Res. 29:297-311.
Standard Methods. 1985. Method 509B, Chlorinated Phenoxy Acid Herbicides.
Standard Methods for the Examination of Water and Wastewater, 16th Edition,
APHA, AWWA, WPCF.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for chloro-
phenoxy acid herbicides in drinking water. In: Methods for Organochlorine
Pesticides and Chlorophenoxy Acid Herbicides in Drinking Water and Raw
Source Water, Interim. July.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Tolerances and
exemptions from tolerances for pesticide chemicals in or on raw agricul-
tural commodities. Food Drug Cosmetic Law Reporter, 40 CFR 180.142.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Drinking water
criteria document for 2,4-dichlorophenoxyacetic acid (2,4-'j). Final
draft. Office of Drinking Water. March, 1985.
U.S. EPA. 1985b. U. S. Environmental Protection Agency. Draft technologies
and costs for the removal of synthetic organic chemicals from potable
water supplies. Office of Drinking Water.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 50(185):33992-34003.
September 24.
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118
2,4-Dichlorophenoxyacetic Acid March 31, 1987
-14-
U.S. EPA. 1986b. U. S. Environmental Protection Agency. Pesticide Pact
Sheet No. 94. 2,4-D. July, 1986. Office of Pesticides and Toxic
Substances, Office of Pesticide Programs.
Heast, R.C., ed. 1980. CRC Handbook of Chemistry and Physics, 61st e<3.
Chemical Rubber Co., Cleveland, OH. p. C482.
Heed Science Society of America. 1979. Herbicide Handbook, 4th ed.
Champaign, IL. pp. 129-135.
WHO. 1977. World Health Organization. Pesticide Residues in Food. WHO
Tech. Rep. Serv. No. 592.
WHO. 1984. World Health Organization. Guidelines for Orinking-Water Quality.
Volume 1. Geneva, Switzerland, pp. 72-73.
Zetterberg, G., L. Busk, R. Elovson, I. Starec-Nordenhammer and H. Ryttman.
1977. The influence of pH on the effects of 2,4-D on Saccharomyces
eerevislae and Salmonella typhimurium. Mutat. Res. 42:3-18.
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March 31, 1987
ENDRIN
Health Advisory
Office of Drinking Hater
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. The chemical 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 Jian 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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Endrin March 31, 1967
-2-
This Health Advisory (HA) is based on information presented in the
Office of Drinking Water's Health Effects Criteria Document (CD) for endrin
(U.S. EPA, 1985a). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological 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 I 86-117967/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS Mo. 72-20-8
H
Structural Formula
H
Synonyms
• 1,2,3,4,10,10-Hexachloro-6,7-epoxy-1,4,4a, 5,6,7,8,8a-octa-hydro-1,4-
endo, endo-5,8-dimethanonapthalene
Uses
• Organochlorine cyclodiene pesticide once widely used in the U.S.
• EPA cancelled the use of endrin for a number of uses and registration
for new uses of endrin was denied (Federal Register, 1979).
• Endrin is registered presently only for the control of cutworms,
grasshoppers and moles.
Properties (U.S. EPA, 1985a)
Cheeical Formula C 1288^6°
Molecular Weight 380.93
Physical State solid
Boiling Point
Melting Point 245°C
Density
Vapor Pressure 2.7 x 10-7 mm Hg (25°C)
Water Solubility 0.24 mg/L (25°C)
Octanol/Water Partition 2.18 x 105
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor —
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Endrin March 31, 1987
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Occurrence
Endrin is considered to be a persistent compound. Endrin is bio-
degraded poorly. Once in the ground, endrin rapidly binds onto soils
and migrates slowly. Endrin has the potential for bioaccumulation
(U.S. EPA, 1983).
Endrin has been included in a number of national and regional surveys
of drinking water supplies. Endrin has not been detected in any of
the surveys. Endrin has been detected in a few surface waters. The
highest level reported was 0.008 ug/L.
Endrin has been reported to occur at very low levels in food and air.
However, the available data are insufficient to evaluate exposures from
these routes or to determine if drinking water is a significant
source of exposure.
Because endrin is no longer commercially used, future trends are
expected to be lower than current data indicate.
III. PHARMACOKINETICS
Absorption
Rates of absorption by the oral, dermal and inhalational routes have
not been documented. Absorption has been demonstrated by the
detection of residue levels in animals and humans following exposure
(U.S. EPA, 1985a).
Distribution
Endrin is distributed (fat, liver, brain, kidneys) and metabolized
rapidly in mammals. The time of sample collection is important
since endrin residues decline rapidly after cessation of exposure
(U.S. EPA, 1985a).
Both wild and domestic birds, however, store endrin in various body
tissues, especially fat (Terriere et al., 1959; Reichel et al.,
1969).
Metabolism
The metabolic pathway for endrin in mammals is complex and species-
dependent. In all species, the unsubstituted methylene bridge (C12)
is attacked preferentially to form mostly anti- and lesser amounts of
syn-12-hydroxyendrin. The syn-isomer is oxidized quickly by micro-
soraal mono-oxygenases to produce 12-ketoendrin, which is considered
to be the major toxicant. Glucuronide and sulfate conjugates of the
anti-isomer are formed (Hutson, 1981; U.S. EPA, 1985a).
To a smaller extent, hydroxylation at the 3-position also occurs, and
the epoxide functional group probably is hydrated (U.S. EPA, 1985a).
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Endrin March 31, 1987
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0 The rapid metabolism of endrin has been explained in terns of the steric
influence of the epoxide anion or C-12 hydroxylation in promoting
anti-C-12-hydroxylation. The bulky hexachlorinated fragment inhibits
attack at C-3 and C-4 (U.S. EPA, 1985a).
Excretion
Endrin is eliminated rapidly in both humans and in animals. A half-
life of 1 to 2 days in blood serum was estimated for humans (U.S. EPA,
1985a).
Anti-12-hydroxyendrin as the glucoronide has been detected in both the
feces and urine of endrin workers (Baldwin and Hutson, 1980), but
12-ketoendrin was not detected (Hutson, 1981). Analysis of D-glucuric
acid in urine is a useful test for endrin exposure (Vrij-Standhardt
et al., 1979).
In rats, over 50 percent of endrin metabolites are eliminated in the
bile within 1 day as the glucaronides of anti-12-hydroxyendrin (Hutson
et al., 1975). In rabbits, the metabolites are conjugated with
sulfate and excreted in the urine (Bedford et al., 1975b). This
behavior is consistent with molecular weight thresholds for biliary
excretion, which are 325 +50 in the rat and 475 ^50 in the rabbit
(Hirom et al., 1972).
IV. HEALTH EFFECTS
Humans
Exposure to endrin may cause sudden convulsions which may occurr
from 30 minutes to 10 hours after exposure. Headache, dizziness,
sleepiness, weakness and loss of appetite may be present for 2 to 4
weeks following this exposure.
A number of deaths have occurred from swallowing endrin. In less
severe cases of endrin poisoning, the complaints include headache,
dizziness, abdominal discomfort, nausea, vomiting, insomnia, agitation
and mental confusion (U.S. DHHS, 1978).
Electroencephalograms (EEGs) show dysrrh;. thmic changes which frequently
precede convulsions; withdrawal from exposure usually results in a
normal electroencephalogram within 1 to 6 months (U.S. DHHS, 1978).
Several incidents of endrin poisoning from contaminated flour have
been reported. In Hales, bread made from flour contaminated with
endrin during shipment in a railway car resulted in 59 poisoning
cases with no deaths in 1956 (Davies and Lewis, 1956). The bread
contained endrin at up to 150 mg/kg bread; the smallest dosage level
to elicit serious effects was calculated to be 0.2 mg/kg bw. Incidents
also have occurred in Doha, Qatar and Hofuf, Saudi Arabia (Weeks,
1967; Curley et al., 1970).
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Endrin JL<-o March 31, 1987
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Mo illnesses were noted when 1% endrin was applied at 544-634 kg/acre
as an emulsifiable concentrate for mouse control at a calculated
dermal dose of 0.28 mg/kg/day in combination with a calculated
respiratory exposure of 0.0011 mg/kg/day (Wolfe et al., 1963).
Concentrations of endrin in the blood of 45 operators from an endrin
processing plant were determined at least once a year from 1964 to
1968 (Jager, 1970). The threshold level of endrin in the blood below
which no sign or symptoms of intoxication were seen was 0.050-0.100
ug/ml. The half-life of endrin in the blood, and thus in the body,
was estimated to be approximately 24 hours. Medical files and routine
medical examinations revealed no abnormalities other than those that
would be expected in any group of 233 long-term workers (4 to 13.3
years' exposure). Determinations of alkaline phosphatase, SCOT,
SGPT, LDH, total serum proteins and the spectra of serum proteins
did not show any changes that could be correlated with the level or
duration of exposure to these insecticides for these parameters. In
all cases of intoxication characterized by typical EEC changes, EEC
patterns returned to normal.
Cases of fatal endrin poisoning have been reported from intentional
and accidental ingestion. Tewari and Sharma (1978) reported 11 fatal
poisonings; the time periods from administration of the pesticide
(route not known in seven cases) to death ranged from 1 to 6 months.
Endrin ingestion with milk or alcohol appeared to increase toxicity
as death occurred within an hour or two. Increased toxicity was
attributed by the authors to more rapid absorption through the GI
tract.
Animals
Short-term Exposure
The acute oral LDso of endrin given to mammals by gavage ranges from
2.3 mg/kg to 43.4 mgAg (U.S. EPA, 1985a).
Revzin (1968) reported an increase in the amplitude of the EEC and a
tendency toward spiking after 7 daily doses of endrin at 0.2 mg/kg in
rats. No effects were noted after 1 or 2 days' exposure at the same
dose level in monkeys.
Speck and Maaske (1958) reported EEG changes and occasional convulsions
after 1 week of daily oral doses of 3.5 mg/kg bw in rats. No effects
were reported when the rats were dosed with 0.8 and 1.7 mg/kg bw.
Bedford et al. (1975a) determined the acute oral LD50 values (based
on 10-day mortality) for three metabolites of endrin which have been
identified in mammals. Each metabolite was more toxic than the parent
pesticide. 1 2-Ketoendrin and sjqi-1 2-hydroxyendrin were about 5 times
more toxic in male rats, and anti -1 2-hydroxyendrin 2 times more toxic
than endrin itself in male rats. In females, 1 2-Ketoendrin was 5 times
and syn-1 2-hydroxyendrin 2 times more toxic than endrin.
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Long-term Exposure
• In an NCI (1979) study both mice and rats (fifty animals of each sex
constituted a treatment group of rats and mice) were exposed chronically
to endrin. The mice were administered a time-weighted-average (TWA)
concentration in the diet of 1.6 or 3.2 mg/kg/day, while the rats
received 3 or 6 ppm. Neither mortality nor body weights were affected
by either dose. According to the investigators, a variety of clinical
signs (alopecia, diarrhea, epistaris, rough hair coats, etc.) were
observed in the exposed rats. However, these findings and interpre-
tations were questioned by another reviewer (Reuber, 1979). These
have been explained in the support document (U.S. EPA, 1985a).
0 Deichmann et al. (1970) administered endrin to rats at concentrations
of 2, 6 or 12 mg/kg/day in the diet for up to 37 months. There was
no significant effect on mean body weight or weight gain in endrin-
treated rats. Signs of toxicity observed during the course of the
experiment were limited to episodes of tremors and clonic convulsions
with "outcries". These signs were dose-related; however, no further
details were provided. Histologic changes in the livers of rats fed
endrin (2, 6 or 12 ppm) were similar to those receiving the control
diet with the exception of a moderate increase in the incidence of
centrilobular cloudy swelling. jThere was also an increase in cloudy
swelling of the renal tubular epithelium. Even though the authors
stated that the effects were not dose-related, the presence of centri-
lobular swellings and cloudy swellings of the renal tubular epithelium
are suspect.
0 Nelson et al. (1956) exposed adult Sprague-Dawley rats to endrin at
1, 5, 25, 50 and 100 mg/kg/day in the diet for 16 weeks. A dose-
dependent increase in alkaline phosphatase levels was observed, while
body weights in all exposed groups decreased in comparison with
controls after 16 weeks. All rats receiving 100 ppm endrin died
within the first two weeks of exposure. Rats exposed to 25, 50 or
100 ppm manifested convulsive spasms.
• Beagle dogs (4/group) were exposed to endrin at 1, 3 or 4 ppm in the
diet for 18.7 months. Body weight gains were depressed in the 4 ppm
but not in the 1 or 3 ppm groups. Kidney and heart weights were
significantly greater in the 3 ppm but not in the 1 ppm group. Based
upon increases in kidney and heart weights, the NOAEL for chronic
exposure of dogs is determined to be 0.045 mgAg bw/day (Treon and
Cleveland, 1955).
0 Rats (20 males and 20 females/group) were exposed to endrin at 1, 5,
25, 50 and 100 ppm in the diet for 2 years. The average length of
survival was decreased significantly in females exposed to 25 ppm or
greater and males exposed to 50 ppm or greater. Diffuse degeneration
of the brain, liver, kidneys and adrenals was reported in animals
that died during exposure. Based upon liver weight change, the NOAEL
was determined to be 1 ppm (0.05 mgAg/ bw assuming daily food intake
is 5% of body weight) (Treon and Cleveland, 1955).
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Reproductive Effects
0 Mo information was found in the available literature on the repro-
ductive effects of endrin.
Developmental Effects
0 Endrin administered by oral gavage to Golden Syrian hamsters on days
5 to 14 of gestation resulted in maternal lethality at doses of 1.5
»9A9/
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Endrin March 31, 1987
-8-
1979). The results were negative in all of these studies. Treon and
Cleveland (1955) also failed to note any increase in tumorigenesis in
dogs exposed to endrin up to 18.7 months at the maximum tolerated
dose. Details of various investigations have been given in the support
document (U.S. EPA, 1985a).
The only positive carcinogenic effects of endrin were reported by
Reuber (1978, 1979). Reuber's criteria appear to differ from those
of other investigators (U.S. EPA, 1985a).
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 - (MOAEL or LOAEL) x (BW) , Bg/L ( ug/L)
(UF) x ( L/day)
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 factor (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
The study by Revzin (1968) is selected as the basis for the One-day HA.
In this study, Revzin reported alterations ia the EEC of squirrel monkeys
after 7 daily doses of 0.2 mgAg endrin. No effects were noted, however, at
this dose level for shorter exposures. Thus, 0.2 mgAg can be considered a
NOAEL for a one-day exposure. If this study were considered suitable for the
development of a One-day HA, it would be derived as below. This study is
supported by Davis and Lewis (1959) and Hayes (1963).
One-day HA = *°'2 »9Ag/day) (10 kg) = 0.02 mg/L
(100) (1 L/day)
where:
0.2 mgAg/day * NOAEL based on absence of EEC changes in squirrel
monkeys after one-day exposure.
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Endrin March 31, 1987
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10 kg - assumed body weight of a child.
100 - uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day « assumed daily water consumption of a child.
Based upon data from Davies and Lewis (1956) in which the human response
to ingestion of bread contaminated with 150 ppm endrin was reported, Hayes
(1963) estimated that the dosage necessary to produce a single convulsion in
man is about 0.25 mg/kg« If this estimate were considered suitable for the
development of a One-day HA, it would be derived thusly:
One-day HA - (0.25 mg /kg/day) (10 kg) . 0.025 nq/L
(TOO) (1 L/day) y/
where:
0.25 mg/kg/day - minimum-effect level for convulsions in humans.
10 kg = assumed body weight of a child.
100 » uncertainty factor, chosen in accordance with NAS/ODW
guidelines for 'use with a LOAEL from a human study.
1 L/day « Assumed daily water consumption of a child.
It is recommended that a concentration of 0.02 mg/L for a child be
accepted as the One-day HA for endrin. The derivations from the human data
are based upon rather severe effects and dosages are estimated rather than
actual measurements. The HAs based upon the Hayes (1963) estimates, however,
are only slightly greater than the ones developed using the Revzin (1968)
study and, thus, provides some support for the recommended values.
Ten-day Health Advisory
In the teratology studies by Kavlock et al. (1981) decreases in maternal
weights were reported for rats dosed orally for 14 consecutive days with 0.3
but not 0.15 mgAg bw endrin. If this study were considered suitable for the
development of a Ten-day HA, it would be derived thusly:
Ten-day HA = (0'15 nqAg/day) (10 kg) = 0.015 mg/L
(100) (1 L/day) y
where:
0.15 mg/kg/day « NOAEL for short-term effects in exposed animals.
10 kg - assumed body weight of a child.
100 » uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day » assumed daily water consumption of a child.
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Endrin March 31, 1987
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The study by Kavlock et al. (1981) is appropriate to calculate the
Ten-day HA. In this study behavioral effects in offspring of rats treated
for 14 consecutive days with 0.15 but not 0.075 ng/kg endrin were reported.
Using a NOAEL of 0.075 ng/kg/day, the Ten-day HA is calculated as follows:
Ten-day HA « (0.075 mgAg/day) <10 kg) . Q.0075 mq/L
(100) (1 L/day) *'
where:
0.075 ng/kg/day « NOAEL based on absence of behavioral changes in
offspring of exposed rats.
10 kg « assumed body weight of a child.
100 - uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day « assumed daily water consumption of a child.
Nelson et al. (1956) reported that body weights of rats exposed 13 weeks
to 5 ppm endrin but not 1 ppm in the diet decreased relative to controls. If
this study were considered suitable for the development of a Ten-day HA, it
would be derived thusly:
Ten-day HA = (0.05 mg/kg bw/day) (10 kg) = .005 mq/L
(100) (1 L/day)
where:
0.05 mg/kg/day = NOAEL for body weight changes in rats based upon
1 ppm in the diet.
10 kg » assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
It is recommended that a concentration of 0.005 mg/L for a child, based
on Nelson et al. (1956), be accepted as the Ten-day HA for endrin. Depressed
body weight is considered to be an adequate indication of detrimental effect.
Behavioral effects in offspring of rats administered similar doses provide
additional support for this HA.
Longer-term Health Advisory
Treon and Cleveland (1955) exposed dogs for up to 18.7 months to 1 , 3 or
4 mgAg/day endrin in the diet. Increases in heart and kidney weight were
noted at 3 and 4 mgAg/day in diet, but not at 1 mg/kg/day. Based on measured
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Endrin March 31, 1987
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food intake, the daily dose for the 1 mgAg/day group varied from 0.045-0.12
ng/kg bw.
The Longer-term HAs are calculated as follows:
For a child;
longer-term HA - ( 0.045 ^g/k^dayM 10 kg) . 0.0045 ng/L (4<5 ug/L)
where:
0.045 mgAg/day = NOAEL based on absence of heart and kidney weight
changes in dogs.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
For an adult;
Longer-term HA = (0.045 mgAg/day) <70 kg) = 0.016 mg/L (16 /L)
(100) (2 L/day)
where:
0.045 mgAg/day - NOAEL based on absence of heart and kidney weight
changes in dogs.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
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.
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Endrin March 31, 1987
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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 classifed 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 chronic study in dogs by Treon and Cleveland (1955) is also used to
calculate the Lifetime HA.
Using the NOAEL of 0.045 mgAg/day, the Lifetime Health Advisory is
calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD « (0.045 mg/kq/day) . 0.000045
(100) (10)
where:
0.045 mg/kg/day = NOAEL based on absence of effects on heart and
kidney weight changes in dogs.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
10 « uncertainty factor, appropriate for question
regarding dietary intake (discrepancy between
published and unpublished studies).
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL « (0*000045 agAg/day) (70 kg) - Q.0016 mg/L (1.6 ug/L)
(2 L/day)
where:
0.000045 mg/kg/day = RfD.
70 kg * assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = 0.0016 mg/L x 0.20 = 0.00032 mg/L (0.32 ug/L)
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Endrin March 31, 1987
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where:
0.0016 mg/L - DWEL.
0.20 - assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Assessment of the evidence for carcinogenicity of endrin in either
humans or animals suggests that no potential exists. As a result, a
quantitative risk estimate for cancer induction was not derived.
0 IARC has not evaluated the carcinogenic potential of endrin.
0 Applying the criteria in the guideline for assessment of carcinogenic
risk (U.S. EPA, 1986), endrin is classified in Group £: Mo evidence
of carcinogenicity in at least two adequate animal tests or in both
epidemiologic and animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA (1975) has set an interim standard for endrin in finished
drinking water of 0.0002 mg/L or 0.2 ug/L.
0 The U.S. EPA (1980) proposed an ambient water criterion for endrin of
0.001 mg/L or 1 ug/L.
e The World Health Organization (PAD/WHO, 1973) established as a guide-
line a maximum intake of 2 ug/kg/day.
VII. ANALYTICAL METHODS
0 Determination of endrin is by a liquid-liquid extraction gas chromato-
graphic procedure (U.S. EPA, 1978; Standard Methods, 1985). Specifi-
cally, the procedure involves the use of 15% methylene chloride in
hexane for sample extraction, followed by drying with anhydrous sodium
sulfate, concentration of the extract and identification by gas
chromatography. Detection and measurement is accomplished by electron
capture, microcoulometric or electrolytic conductivity gas chromato-
graphy. Identification may be corroborated through the use of two
unlike columns or by gas chromatography-mass spectroscopy (GC-MS).
The method sensitivity is 0.001 to 0.010 ug/L for single component
pesticides and 0.050 to 1.0 ug/L for multiple component pesticides
when analyzing a 1-liter sample with the electron capture detector.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies which are capable of removing endrin from
drinking water include adsorption by activated carbon, both granular
and powdered (GAC and PAC, respectively), air stripping, reverse
osmosis (RO) and coagulation/filtration.
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Endrin JLO/O
March 31, 1987
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Dobbs and Cohen (1980) developed adsorption isotherms for a number
of organics, including endrin. Based upon the isotherm data, they
reported that activated carbon exhibited absorptive capacities of
106 mg, 17 mg and 2.7 ng of endrin at initial endrin concentrations
of 100 ug/L, 10 ug/L and 1 ug/L, respectively.
One study was undertaken to evaluate a number of water treatment
processes by PAC for pesticide removal (U.S. EPA, 1985b). PAC was
examined by conducting test runs with initial concentrations (1 to
10 ug/L) of pesticide in distilled and river water. The distilled
water was spiked with the required concentration of endrin, PAC was
added and mixed with the water. The river water was used in a pilot
plant where it was mixed with PAC. Based upon the reported results,
PAC treatment appears to be an effective technology for the removal
of endrin.
A RO pilot plant in Miami, Florida, was evaluated for the removal of
certain organic chemicals, including endrin. The RO unit showed 90+%
removal of endrin from an initial concentration of 1 ug/L.
A study pilot plant was used to test the effectiveness of conventional
treatment in removing endrin. In this study, the treatment scheme
consisted of the addition of alum, flocculation, sedimentation and
sand filtration. The results indicated that alum coagulation removed
about 35% of the endrin, no matter what the initial concentration was.
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J. Agric. Food Chem. 7:502-504.
Tewari, S.N., and I.C. Sharma. 1978. Study of the distribution of chlorinated
organic pesticides in different autopsy materials of human poisoning cases
using TLC and UV spectrophotometric techniques. Chem. Era. 14:215-218.
Treon, J.F., and F.P. Cleveland. 1955. Toxicity of certain chlorinated
hydrocarbon insecticides for laboratory animals, with special reference
to aldrin and dieldrin. J. Agric. Food Chej. 3:402-408.
U.S. DHHS. 1978. U.S.'Department of Health and Human Services. Occu-
pational health guideline for endrin. In; Occupational Health
Guidelines for Chemical Hazards, F.W. Mackison, R.S. Stricoff and L.J.
Partridge, Jr., Eds. DHHS (NIOSH) Publ. No. 81-123.
U.S. EPA. 1975. U.S. Environmental Protection Agency. National interim
primary drinking water regulations. Federal Register. 40(248):59566-
59588. December 24.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for
organochlorine pesticides in drinking water. In: Methods for Organ-
ochlorine Pesticides and Chlorophenoxy Acid Herbicides in Drinking Water
and Raw Source Water, Interim, July.
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136
Endrin March 31, 1987
-18-
U.S. EPA. 1980. U.S.* Environmental Protection Agency. Ambient water
quality criteria for endrin. Environmental Criteria and Assessment
Office, Cincinnati, OH. EPA 440/5-80-047.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of
pesticides in drinking water, food and air. Office of Drinking Water.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Health Effects
Criteria Document for Endrin. Criteria and Standards Division. Office
of Drinking Water, Washington, D.C.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Drinking Water
Criteria Document for Endrin. Criteria and Standards Division. Office
of Drinking Water, Washington, D.C.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Draft technologies
and costs for the removal of synthetic organic chemicals from potable
water. Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 51 (185):33992-34003.
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index for
Pesticides. Bureau of Foods.
Vrij-Standhardt, W.G., J.J.T.W.A. Strik, C.F. Ottevanger and N.J. Van Sittert.
1979. Urinary D-glucaric acid and urinary total porphyrin excretion in
workers exposed to endrin. In; Chemical Porphyria in Man, J.J.T.W.A.
Strik and J.H. Koeman, Eds. Esevier/North Holland Biomedical Press,
Mew York, pp. 113-121.
Weeks, D.E. 1967. Endrin food-poisoning. A report on four outbreaks caused
by two separate shipments of endrin-contaminated flour. Bull. WHO.
37:499-512.
Williams, G.M. 1980. Classification of genotoxic and epigenetic hepatocar-
cinogens using liver culture assays. Ann. M.Y. Acad. Sci. 349:273-282.
Witherup, S., K.L. Stemmer, P. Taylor and P. Bietsch. 1970. The incidence
of neoplasms in two strains of mice sustained on diets containing endrin.
Kettering Lab., Univ. of Cincinnati, Cincinnati, OH.
Wolfe, H.R., W.F. Durham and J.F. Armstrong. 1963. Health hazards of the
pesticides endrin and dieldrin. Arch. Environ. Health. 6:458-464.
Zabik, M.J., R.D. Schuetz, W.L. Burton and B.E. Pape. 1971. Photochemistry
of bioactive compounds: Studies of a major photolytic product of endrin.
J. Agric. Food Chem. 19:308-313.
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March 31, 1987
137
ETHYLENE DIBROMIDE
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. The chemical 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 AS 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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Ethylene Dibromide
March 31, 1987
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This Health Advisory is based upon information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for Ethylene
Dibromide (U.S. EPA, 1985a). The HA and CD formats are similar for easy
reference. Individuals desiring further information on the toxicological
data base or rationale for risk characterization should consult the CO. 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 # 86-118247/AS.
The toll free number is (800) 336-4700; in Washington, D.C. area: (703)
487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 106-93-4
Structural Formula
H H
I I
Br-C-C-Br
I I
H H
1,2-Dibromoethane
Synonyc
0 EDB, glycoldibromide, ethylene bromide, Dowfume , Pestmaster*,
Soilbrome*
Uses
0 Lead scavenger in gasoline. Pesticide-fumigant for soil, grain and
fruit (all uses cancelled).
Properties (Stenger, 1978)
Chemical formula
Molecular weight
Physical state (room temp)
Boiling point
Melting point
Density
Vapor pressure
Water solubility
Octanol/water partition
coefficient
Taste threshold in water
Odor threshold in water
Odor threshold in air
a; Back-calculated from the solubility (Lyman, 1982)
BrCH2CH2Br
187.87 (Weast, 1980)
colorless, clear liquid
131.4°C
9.9°C
2.1792 g/ml
11 mm Hg at 20°C (Verschueren, 1983)
4310 mg/L (Verschueren, 1983)
135a
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Ethylene Dibromide '/"Xl March 31, 1987
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Occurrence (U.S. EPA, 1983)
0 Ethylene dibromide (EDB) is a fumigant which, until 1983, was used
widely on more than 40 crops. Production volume in 1983 is estimated
to have been 280 million Ibs; however, the vast majority of EDB
produced was used in gasoline, where it served as an anti-knock
ingredient. Agricultural usage of EDB in 1983 was estimated to be
20 million Ibs. Most of the EDB was used as a soil fumigant for the
control of nematodes, with smaller amounts used as a fumigant of grain
and fruit. In 1983, EPA cancelled all major uses of EDB.
0 EDB is regarded as a highly persistent and mobile pesticide. The major
route of removal of EDB from soil is by volatilization. In the absence
of volatilization, EDB is decomposed slowly in soil by microbial
action with a biodegradation half life of less than 18 weeks. EDB
slowly hydrolyzes, with a half life in sterile water of more than 6
years. EDB has been shown to migrate in soil and has been reported
as a contaminant in ground water. Because of EDB's potential for
volatilization, it is expected to occur more often in ground water
than surface water. There is no available information on EDB's
potential for bioaccumulaion.
0 EDB has not been included in Federal and State monitoring surveys of
ground water; only limited data on its occurrence are available.
However, surveys of wells, including some public water supplies, near
sites where EDB has been used as a soil fumagant have found levels of
contamination in the ug/L and lower range. EDB has not been identified
in surface water supplies. EDB has been identified as a contaminant
in.a number of foods at the ppb to ppm level. The residues are due
to the fumigation of vegetables and grains during shipping or storage.
EDB also has been reported as a wide-spread contaminant in air in the
low ppt range. Atmospheric levels of EDB are believed to result
from the incomplete combustion of gasoline containing EDB. The avail-
able data are insufficient to show whether drinking water is a major
route of exposure for EDB. Because of the cancellation of the majority
of EDB uses, occurrence of EDB in ground water and food is expected
to decline with time.
III. PHARMACOKINETICS
Absorption
0 Uptake of EDB readily occurs in rats following exposure by inhalation
(Watanabe et al., 1978), oral intubation (van Bladeren et al., 1980;
Plotnick et al., 1979) and dermal application (Jakobson et al.,
1982).
0 Quantitative absorption data are not available. It may be inferred
that uptake from the GI tract in rats is extensive since urinary
excretion accounted for 73 percent of an orally administered dose of
15 mg/k9 14C-EDB (Plotnick et al., 1979).
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Ethylene Dibromide March 31, 1987
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Distribution
0 A tissue distribution study in guinea pigs was undertaken by Plotnick
and Conner (1976) because of the close similarities in metabolic path-
ways between guinea pigs and humans* Following a single i.p. injection
of 30 mgA9 14C-EDB, the highest concentrations were detected in the
liver, kidney and stomach.
e Plotnick et al. (1979) treated rats with a single oral dose of 15
mg/kg 14C-EDB in corn oil. Tissue analysis revealed the highest
concentrations to be in the liver, kidney and spleen.
Metabolism
0 Studies with rats have provided evidence that two pathways of meta-
bolic bioactivation exist for EDB, each producing a reactive metabolite
capable of eliciting toxic effects (van Bladeren et al., 1980, 1981).
An oxidative pathway predominates over the conjugative pathway by a
4:1 ratio.
0 The reactive metabolite produced by the oxidative pathway, 2-bromoacet-
aldehyde, is important in cell macromolecule binding and associated
histopathological changes such as liver damage (Nachtomi, 1981; Shih
and Hill, 1981).
0 The conjugative pathway (principally,glutathione) is more closely
associated with DNA binding and outagenesis (Hill et al., 1978;
van Bladeren et al., 1980, 1981). 5-(2-Bromoethyl)-glutathione or
the resulting episulfonium ion is believed responsible for these
effects (Livesey and Anders, 1979).
Excretion
In the rat, orally administered EDB is excreted primarily in the
urine as mercapturic acid derivatives (Jones and Edwards, 1968).
Unchanged EDB apparently is not excreted in the urine, although it
may be eliminated in small quantities in expired air.
Rates of urinary excretion of radioactivity following inhalation of
14c-EDB indicated a half-life for elimination in the range of 5.1 to
5.6 hours (Watanabe et al., 1978).
Patterns of elimination by the guinea pig are similar to thos ? of
the rat (Plotnick and Conner, 1976).
IV. HEALTH EFFECTS
Humans
0 The available data from case history reports indicate that EDB may be
lethal to humans after a single oral dose of 65 mg/kg (Olmstead,
1960) and that local and systemic reactions can result from direct
dermal contact (Pflesser, 1938).
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Ethylene Dibromide JL/ll March 31, 1987
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0 Several morbidity studies on EDB-exposed workers have focused on the
evaluation of adverse effects on fertility. Hie results are equivocal,
showing only a slight indication of reduced fertility in two studies
(Griffith et al., 1978; Wong et al., 1979), but no indication of
reduced fertility (Levine, 1981) or impaired spermatogenesis (Ter Haar,
1978, 1981) in other studies. The available epidemiologic evidence is
not adequate to establish or deny that EDB affects human reproductive
function.
0 Mortality studies conducted on workers exposed to EDB are inconclusive
with respect to death by specific target organ effects (Ott et al.,
1980; Turner, 1976, 1977). Because of limitations of scope and
design, these epidemiologic studies are not considered to provide
definitive results.
0 A recent study by NIOSH (1986) on the semen quality of 46 papaya workers
exposed to an average of 88 ppb (8-hour time weighted average) for
approximately five years reflected a decrease in sperm quality (i.e.,
mobility, shape) when compared to data from 43 unexposed men.
Animals
Short-term Exposure
• Single-dose oral LD5Q's of 146 and 117 mg EDBAg bw (in olive oil)
have been determined for male and female rats, respectively (Rowe
et al., 1952). Oral LD50's of 420 mgAg (female mice), 55 mgAg
(female rabbits) and 110 mgAg (both sexes of guinea pigs) were deter-
mined in the same study.
0 In a series of studies on the hepatic effects of EDB, single oral
doses of 75 to 120 mgAg produced hepatomegaly, centrilobular necrosis,
increased levels of liver lipids and serum enzymes, and evidence of
DNA damage and repair (Nachtomi et al., 1968; Nachtomi and Alumot,
1972; Broda et al., 1976; Nachtomi and Farber, 1978; Nachtomi and
Sarma, 1977).
0 Administration of EDB by gavage at a dose of 10 mgAg bw/day for 12
days produced significantly elevated levels of serum glutamic pyruvic
transaminase (GPT) and sorbi,tal dehydrogenase (SDH) in rats, but
dietary administration of 10-20 mgAg/day for 18 days did not result
in a significant increase in liver weight, DNA content or thynidine
incorporation (Nachtomi, 1980).
Long-term Exposure
0 Data on the non-neoplastic effects of chronic oral exposure to EDB
are available from the NCI (1978) carcinogenesis bioassay in which
Osborne-Mendel rats of both sexes were exposed by gavage to time-
weighted average (TWA) doses of 0 or approximately 28 mgAg/day for
49 to 61 weeks and B6C3F] mice of both sexes were exposed to 0, 44 or
77 mgAg/day (approximately) for 53 weeks respectively. Treatment-
related non-neoplastic effects were found in the forestomach (hyper-
keratosis and acanthosis in the high-dose male rats, male and female
mice, and in the low- and high-dose female rats), liver (peliosis
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Ethylene Dibromide March 31, 1987
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hepatitis or inflammation in the low-* and high-dose male and female
rats), adrenal cortex (degeneration in the low- and high-dose male
rats and high-dose female rats) and testes (atrophy in the low- and
high-dose rats and high-dose mice).
0 Chronic inhalation exposure to EDB at concentrations of 0, 1, 10 or
40 ppm (6 hours/day, 5 days/week for 78-106 weeks) produced increased
mortality in male and female F344 rats at the high dose and increased
mortality in female B6C3FJ mice at the low and high doses (31/50 and
43/50, respectively, vs. 10/50 in controls) (NTP, 1982). Treatment-
related non-neoplastic lesions occurred in the respiratory system
(epithelial hyperplasia, sguamous metaplasia or inflammation of the
nasal cavity, bronchus or lung) in the low- and high-dose rats and
mice of both sexes, liver (necrosis) in the high-dose rats of both
sexes, kidneys (toxic nephropathy) in low- and high-dose male rats
and high-dose female rats, testis (degeneration and atrophy) in low-
and high-dose rats, and adrenal cortex (degeneration) and retina
(atrophy) of low- and high-dose female rats.
0 In another chronic inhalation study, mortality was increased and the
incidence of atrophy of the spleen was elevated significantly (6/48
vs. 0/48 in controls) in groups of 48 Sprague-Dawley rats that were
exposed to 20 ppm EDB (7 hours/day, 4 days/week for 18 months), but
testicular atrophy was not found and the nasal cavity was not examined
(Wong et al., 1982).
Reproductive Effects
0 Dietary administration of EDB to bulls at an average daily dose of
2 mgAg» beginning at 4 days of age to 14 to 16 months of age or from
14 to 16 months of age, produced reversible antispermatogenic effects
without other evidence of toxicity (Amir and Volcani, 1965). Anti-
spermatogenic effects were evident as early as 2 weeks after initiation
of treatment.
0 A high percentage of abnormal spermatozoa also was produced in bulls
after 10 oral doses of 4 mgAg EDB given on alternate days (Amir and
Ben-David,1973; Amir and Lavon, 1976), or given as a single peri-
testicular injection of 110 to 120 mg (Amir and Ben-David, 1973) or
270 mgAg (Amir et al., 1979).
0 Reversible "ntispermatogenic effect of EDB has also been produced in
rams following 12 consecutive daily subcutaneous injections of 7.3 to
13.5 mgAg (acute systemic toxicity was evident at 16.9 mgAg/day)
(EUack and Hrudka, 1979) and by single peri testicular injections of
250 or 430 mgAg (Amir et al., 1983) but not by chronic oral admini-
stration (dose not reported) (Amir and Ben-David, 1973).
0 When compared with bulls and rams, rats may be relatively resistant
to the spermicidal actions of EDB, since adverse effects on fertility
are produced only at levels of exposure associated with systemic
toxicity (Short et al., 1979; Amir et al., 1983). Nonetheless,
testicular atrophy has been demonstrated in both rats (NCI, 1978;
Wong et al., 1982; NTP, 1982) and mice (NCI, 1978) chronically
exposed to EDB.
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Ethylene Dibromide March 31, 1987
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Reproductive Effects
0 The teratogenic potential of BOB has been evaluated in rats and mice
that were exposed to BOB by inhalation, or rats exposed by i.p.
injection. The inhalation studies (Short et al., 1976, 1979) showed
that nearly continuous inhalation exposure (23 hours/day) to 20 ppm
EDB vapor on days 6 to 15 of gestation produced skeletal anomalies in
both species with reduced maternal food consumption. Similar exposure
to higher concentrations of EDB (32 to 80 ppm) produced more pronounced
skeletal anomalies and dose-related maternal and fetal toxicity. The
anomalies were attributed to the toxicity, rather than a true terato-
genic response.
* Intraperitoneal injection of EDB on days 1 through 15 of gestation
reportedly did not produce fetotoxicity or external, gross or visceral
abnormalities in rats at a dose (55 mgAg) that produced changes in
maternal organ weights (but not in body weight) (Hardin et al., 1981).
Mutagenicity
0 In bacterial systems, EDB caused both reverse mutations (Barber et al.,
1981; Moriya et al., 1983) and forward mutations (Brem et al., 1974;
Principe et al., 1981).
0 Tan and Hsie (1981) used a Chinese hamster ovary cell system that
detects forward mutations to evaluate the mutagenicity of EDB. A
dose-related increase in mutation frequency was detected both in the
presence and absence of S-9, but at a higher concentration in the
absence of S-9.
0 Clive (1973) reported that EDB induced mutations in mouse lymphoma
cells in the absence of S-9. In a later study, Clive et al. (1979)
reported that inclusion of S-9 increased the toxicity of EDB to the
cells by 10 fold, and that EDB was more mutagenic under these
conditions.
0 EDB was negative in the dominant lethal assay in both rats and mice
(Epstein et al., 1972; Teramoto et al., 1980).
0 Tezuka et al. (1980) reported sister chromatid exchanges and chromo-
somal aberrations that increased in a linear manner with dose in
cultured Chinese hamster V79 cells exposed to EDB.
Carcinoqenicity
0 EDB has been demonstrated to be a potent carcinogen in rats and mice.
When administered by gavage to Osborne-Mendel rats at TWA doses of
approximately 27 to 29 mgAg /day for 49 or 61 weeks, EDB produced sig-
nificantly increased incidences of squamous cell carcinomas of the
forestomach of both sexes, hemangiosarcomas of the circulatory system
in males and hepatocellular carcinomas and liver neoplastic nodules
(NCI, 1978). Similar administration to B6C3F-| mice at average doses
of approximately 44 or 77 mg/kg/day for 53 weeks induced squamous
cell carcinomas of the forestomach and alveolar/bronchiolar adenomas
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Ethylene Dibromide March 31, 1987
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in both sexes. Forestomach carcinomas developed in 55 to 90% of the
treated rats and mice (none were found in vehicle controls); occur-
rence of these tumors was dose-related.
0 High mortality and early onset of tumors prompted an interim discon-
tinuation of dosing in the rats, periodic adjustment of doses in the
mice, and early termination of both the rat and mice studies (NCI,
1978).
0 Inhalation exposure to EDB at concentrations of 10 or 40 ppm, 6
hours/day, 5 days/week for 78 to 103 weeks produced significantly
increased tumor incidences of nasal cavity tumors (particularly
adenocarcinomas, carcinomas and adenomatous polyps) in F344 rats of
both sexes, alveolar/bronchiolar carcinomas or adenomas in female
F344 and B6C3F-| mice of both sexes and nasal cavity tumors (particu-
larly carcinomas) in female B6C3F-, mice (NTP, 1982). Significantly
increased incidences of circulatory system hemangiosarcomas (male and
female rats, female mice), pituitary adenomas (male and female rats),
tunica vaginalis mesotheliomas (female mice) and subcutaneous fibro-
sarcomas (female mice) also were found in the NTP study.
0 The results of another chronic inhalation study in which Sprague-
Dawley rats were exposed to 20 ppm 1,2-dibromoethane for 7 hours/day,
5 days/week for 72 weeks are consistent with those of the NTP (1982)
bioassay (Wong et. al., 1982). In this study with Sprague-Dawley
rats, significantly increased incidences of hemangiosarcomas (males
and females), adrenal tumors (males and females), subcutaneous
mesenchymal tumors (males) and mammary gland tumors (females) were
induced; the nasal cavity was not, however, examined in this study.
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 { Ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet-Level
in mg/k9 bw/day.
BW = assumed body weight of protected individual
(10 kg for a child and 70 kg for an adult).
UF « uncertainty factor (10, 100 or 1,000), in
accordance with NAS/OOW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
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Ethylene Dibroraide ^o March 31, 1987
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One-day Health Advisory
Data are not available to calculate a One-day HA. It is recommended that
the Ten-day HA of 8 ug/L be used as a One-day HA.
Ten-day. Health Advisory
The study by ElJack and Hrudka (1979) has been chosen to serve as the
basis for the Ten-day Health Advisory. In this study, 18 rams were given BOB
at 7.8, 9.6 or 13.3 mg/kg/day subcutaneously for 12 consecutive days. Substantial
effects on the testis were noted. These effects included reduction in motility
and an increase in the number of morphologically abnormal and degenerating
sperm. The severity of the response was dose-dependent. A NOAEL could not be
identified in the study, but the lowest dose can be considered a LOAEL, since
the changes observed in the parameters measured ranged from 10 to 15% below
control levels.
The Ten-day HA is calculated as follows:
Ten-day HA = (7.8 mgAg/day) (10 kg) =3.6 ug/L
(1,000) (10) (1 L/day)
where:
7.8 mg/kg/day « LOAEL in rams for reproductive effects.
10 kg = assumed body weight of a child.
1,000 - uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
10 = uncertainty factor, considered appropriate to accom-
modate for possibility that the human is closer in
sensitivity to the bull than to the ram.*
1 L/day = assumed daily water consumption of a child.
* A series of studies in bulls has been published in which a total of 24
animals were treated orally for periods ranging from 20 days to approxi-
mately two years (see Reproductive/Teratogenic Effects section). Only one
dose level (4 mg/k? every other day) was given in these experiments. While
the data from these studies are not considered suitable for quantitative
risk assessment, they do show that the bull is more sensitive to the effects
of EDB than is the ram. Since no evidence exists to conclude that the
human male is closer in sensitivity to the bull, to the ram, or even to the
rodent species, an added uncertainty factor of 10 has been utilized to
account for the possibility that the human is as sensitive as the bull.
Longer-term Health Advisory
No adequate data were available for use in calculating a Longer-term
Health Advisory. In any case, exposure over the longer-term would not be
recommended due to the potential carcinogenic risk associated with exposure
to EDB.
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Ethylene Dibromide ~ March 31, 1987
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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.
EDB has been shown to be a potent mutagen and carcinogen. A Lifetime
Health Advisory was not calculated due to the potential carcinogenic risk
posed by exposure to EDB.
Evaluation of Carcinogenic Potential
0 EDB has been shown to be a potent carcinogen as well as a mutagen.
These properties must be considered when developing and implementing
any strategy addressing contamination of drinking water by this
chemical. While a One-day and Ten-day Health Advisory can be calcu-
lated for exposure based upon non-carcinogenic end-points of toxicity,
it is important to be aware of the potential attendant carcinogenic
risk at these levels. It is not unusual to expect that by the time a
contamination incident i j noted and verified, the users of that
drinking water source/supply may actually ha1 e been exposed to the
chemical for an extended period of time, perhaps a year or longer.
For that reason, estimated excess cancer risks associated with exposure
to 8 ug/L EDB over a ten-day period have been developed. In addition,
concentrations of EDB in drinking water which equate to a risk rate
of 10-6 over several exposure durations is identified (Kimm and Anderson,
1985). All risks are projected for the 10 kg child, drinking 1 liter
of water per day. They are as follows:
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Ethylene Dibromide X'lT March 31, 1987
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Exposure Duration Estimated Excess Cancer Risk at 6 ug/L
10 days 1.4 x 10-4
Concentration (ug/L) equal to 10-6 risk
10 days 0.06
1 year 0.02*
2 years 0.0006*
* Estimates are valid only if dose/kg body weight remains constant with any
change in body weight and water intake over time of exposure.
0 EPA's Carcinogen Assessment Group also has calculated estimated excess
lifetime cancer risks over a 70-year lifespan for a 70 kg adult
drinking 2 liters of water per day. These estimates, reflecting the
upper 95% confidence limit, are 0.05, 0.005 and 0.0005 ug/L for a
risk of 10~4, 10-5 an(j 1 o~6, respectively.
0 All of the above calculations are based on the NCI (1978) studies in
mice and rats. The oncogenic response was significant in both sexes
of both species with the effect most noted in the male rat. Five
mathematical models were considered in the risk assessment of EDB;
however, the multistage model was found to be the most appropriate for
use in the above cancer risk estimations.
0 IARC has classified ethylene dibromide in Group 2B: Sufficient evidence
of carcinogenic!ty in animals (IARC, 1982).
0 Applying the criteria described in EPA's guidelines for the assessment
of carcinogenic risk (U.S. EPA, 1986), ethylene dibromide is 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.
OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The use of EDB as a soil fumigant has been suspended, and use of this
substance in fumigation of citrus fruits has been limited (Federal
Register, 1983a). Allowable residues for EDB on food products are
900 ppb for raw grain for human consumption, 150 ppb for flour and
30 ppb for ready-to-eat products.
0 The Occupational Safety and Health Administration (OSHA) (Federal
Register, 1983b) has lowered the 8-hour TWA exposure from 20 ppm to
0.1 ppm in workroom air. The short-term exposure is 0.5 ppm over 15
minutes.
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Ethylene Dibromide March 31, 1987
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0 The American Conference of Governmental Industrial Hygienists (ACGIH,
1984) states that all exposures should be carefully controlled, but
does not provide a suggested TWA limit.
VII. ANALYTICAL METHODS
0 Analysis of ethylene dibromide is by a purge-and-trap gas chroma to-
graphic procedure used for the determination of volatile organohalides
in drinking water (U.S. EPA 1985b). This method calls for the bubbling
of an inert gas through the sample and trapping ethylene dibromide on
an ads or bant material. The adsorbant material is heated to drive off
the ethylene dibromide onto a gas chromatographic column. This
method is applicable to the measurement of ethylene dibromide over a
concentration range of 0.3 to 1500 ug/L. Confirmatory analysis for
ethylene dibromide is by mass spectrometry (U.S. EPA, 1985c). The
detection limit for confirmation by mass spectrometry is 0.4 ug/L.
VIII. TREATMENT
e
Aeration, boiling and adsorption have been considered as possible
treatment techniques for the removal of ethylene dibromide (EDB) from
drinking water.
Tests conducted using the Dynamic Mini Column Adsorption Technique
(DMCAT), a rapid evaluation method, suggests that adsorption onto gran-
ulated activated carbon (GAG) is likely to be a successful treatment
technique (ESE, 1983). DMCAT runs were conducted with deionized
water spiked at approximately 100 ppb or approximately 50 ppb. An
initial run at 95 ppb EDB demonstrated that 1,415 mLs could be passed
with no EDB detected in column effluent (<0.01 ppb). Other runs
determined breakthrough volumes: 0.051 ppb were detected in column
effluent after the passage of 2,040 mL at 45.4 ppb; 0.009 ppb after
the passage of 1,265 mL at 89.7 ppb; 6.3 ppb after the passage of
3,000 mL at 95.8 ppb «0.1 ppb was detected at 1,400 mL, no intervening
values were reported). These data were used to estimate carbon usage
rates. To maintain an effluent concentration of 0.10 ppb, 0.15 to
0.21 lb/1,000 gal would be required for influent concentrations of 45
to 96 ppb.
ESE (1983) also conducted pilot studies of EDB removal using air
stripping. Well water spiked with EDB at 100 ppb was treated in a
system consisting of four 1.5-foot diameter columns operated in
series to give a total height of 50 feet. The columns were packed
with 1-inch polypropylene Intalox saddles. The trial included runs
at various liquid loading rates, air-to-water ratios and packing
heights. Better EDB removals were achieved at higher air-to-water
ratios and additional packing height. A packing depth of greater
than 40 feet at an air-to-water ratio of above 30 would be required
to achieve over 95% removal for an 8,340 Ib/hr/ft2 liquid loading
rate. In another study, contaminated well water (<1 ppb EDB) was
passed through a 1.2-foot diameter, 15-foot column containing No. 2
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Ethylene Dibromide March 31, 1987
-13-
Tripacks packing (U.S. EPA, 1985d). Greater removals were achieved at
higher air-to-water ratios. The highest rate of removal (81.3) was
achieved at an air-to-water ratio of 150 and a liquid loading rate of
15 gpm/ft2.
Air stripping transfers EDB directly to the air, thus air pollution is
a potential disadvantage of this technique. Frink (1985) estimated
that air stripping 1 ppb from water would yield a concentration of
1.20 ppb in the exiting air. This was not thought to pose a significant
additional health hazard because the concentration, before dilution,
is 1/1Oth of the OSHA standard. It also was indicated that exposure
to EDB from this source is likely to be much lower than from exposure
to unleaded gasoline hydrocarbon vapors. EDB concentrations of up to
19.8 ppb have been reported in automobile exhaust fumes.
Isaacson et al. (1984) demonstrated that EDB may be removed easily
from water by boiling. No EDB was detected in water samples initially
containing 0.1 to 5 ppb following a minute or less of boiling in an
open vessel. This suggests that boiling could be used to remove EDB
from drinking water in an' emergency situation. However, the authors
demonstrated that EDB was not degraded during heating. Thus, the
potential health hazard due to EDB inhalation should be evaluated.
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Ethylene Dibromide March 31, 1987
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IX. REFERENCES
ACGZH. 1984. American Conference of Governmental Industrial Hygienists.
TLVs: Threshold limit values for chemical substances and physical agents
in the work environment with intended changes for 1983-84. pp. 42-43.
Amir, D., and E. Ben-David. 1973. The pattern of structural changes induced
in bull spermatozoa by oral or injected ethylene dibromide (EDB). Ann.
Biol. Anim. Biochem. Biophys. 13(2):165-170.
Amir, D., and V. Lavon. 1976. Changes in total nitrogen, lipoproteins and
amino acids in epididymal and ejaculated spermatozoa of bulls treated
orally with ethylene dibromide. J. Reprod. Pert. 47(1):73-76.
Amir, D., and R. Volcani. 1965. Effect of dietary ethylene dibromide on
bull semen. Nature. 206:99-100.
Amir, D., J.C. Nicolle and H. Courot. 1979. Changes induced to bull
spermatids and testicular spermatozoa by a single peritesticular injec-
tion of ethylene dibromide. Int. J. Androl. 2(2):162-170.
Amir, D., B.L. Gledhill, D.L. Garner, J.C. Nicolle and A. Tadmor. 1983.
Spermiogenic, epididymal and spermatozoal damage induced by a peri-
testicular injection of ethylene dibromide to rams. Anim. Reprod. Sci.
6(1):35-50.
Barber, E.D., W.H. Donish and K.R. Mueller. 1981. A procedure for the
quantitative measurement of the mutagenicity of volatile liquids in the
Ames Salmonella/microsome assay. Mutat. Res. 90(1):31-48.
Brem, H., A.B. Stein and H.S. Rosenkranz. 1974. The mutagenicity and
DNA-modifying effect of haloalkanes. Cancer Res. 34:2576-2579.
Broda, H., E. Nachtomi and E. Alumot. 1976. Differences in liver morphology
between rats and chicks treated with ethylene dibromide. Gen. Pharmacol.
7:345-348.
Clive, D. 1973. Recent development with the L5178Y TK heterozygote mutagen
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Clive, D., K.O. Johrson, J.F.S. Spector, A.G. Bastor. and M.M.M. Brown. 1979.
Validation and characterization of the L5178Y/TK+/- mouse lymphoma
mutagen assay system. Mutation Res. 59:61-108.
El Jack, A.H., and P, Hrudka. 1979. Pattern and dynamics of teratospermia
induced in rams by parenteral treatment with ethylene dibromide.
J. Ultrastruct. Res. 67(2):124-134.
Epstein, S.S., E. Arnold, J. Andres, W. Bass and Y. Bishop. 1972. Detection
of chemical mutagens by the dominant lethal assay in the mouse. Toxicol.
and Appl. Pharmacol. 23:288-325.
-------
Ethylene Dibromide •*-**-*- March 31, 1987
-15-
ESE. 1983. Environmental Science and Engineering. Evaluation of the treata-
bility of ethylene dibroroide and dibromochloropropane by activated
carbon and packed column air stripping. ESE No. 81-227-280. For U.S.
EPA, Office of Drinking Water, STB.
ESE. 1984. Environmental Science and Engineering. Review of treatability data
for removal of twenty-five synthetic organic chemicals from drinking
water. For U.S. EPA, Office of Drinking Water.
Federal Register. 1983a. Ethylene dibromide; Decision and emergency order
suspending registrations of pesticide products containing ethylene
dibromide for use as a soil fumigant. U.S. EPA, Washington, DC.
48(197):46228-46248.
Federal Register. 1983b. Occupational exposure to ethylene dibromide.
Dept. of Labor, Washington, DC. 48(196):45956-46003. 29 CFR part 1910.
Frink, C.R. 1985. EDB: 1. Well treatment. Connecticut Academy of Science
and Engineering. Hartford, CT. Response to inquiry from Environment
Committee, Connecticut General Assembly.
Griffith, J., R. Heath and F. Davids. 1978. Spermatogenesis in agricultural
workers potentially exposed to ethylene dibromide (EDB). An interim
report by the Epidemiologic Studies Program, Human Effects Monitoring
Branch, Technical Services Division, OPP, OTS, EPA. June 8.
Hardin, B.D., G.P. Bond, M.R. Sikov, F.D. Andrew, R.P. Bellies and R.W.
Niemeier. 1981. Testing of selected workplace chemicals for teratogenic
potential. Scand. J. Work Environ. Health. 7(4):66-75.
Hill, D.L., T-W. Shin, T.P. Johnston and R.F. Struck. 1978. Macromolecular
binding and metabolism of the carcinogen 1,2-dibromoethane. Cancer Res.
38(8):2438-2442.
IARC. 1982. International Agency for Research on Cancer. IARC Monographs
on the evaluation of the carcinogenic risk of chemicals to man.
Supplement 4.
Isaacson, P.J., L. Hankin and C.R. Frink. 1984. Boiling drinking water to
remove EDB.
Jakobson, I., J.E. Wahlbert, B. Holmbery and G. Johansson. 1982. Uptake via
the blood and elimination of 20 organic solvents following epicutaneous
exposure of anesthetized guinea pigs. Toxicol. Appl. Pharmacol.
63(2):181-187.
Jones, A.R., and K. Edwards. 1968. The comparative metabolism of ethylene
dime thanesulphonate and ethylene dibromide. Experientia. 24:1100-1101.
Kimm, V.J., and E.L. Anderson. 1985. Memorandum to William N. Hedeman, Jr.
Ethylene dibromide: Interim 1- and 10-day Health Advisories for Drinking
Water. June 5.
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Ethylene Dibromide J-v^/w March 31, 1987
-16-
Levine, R.J. 1981. The reproductive experience of workers exposed to ethylene
dibromide at Ethyl Corporation, Magnolia, Arkansas. Chemical Industry
Institute of Toxicology, manuscript submitted to Occupational Safety and
Health Standards Board, Sacramento, CA, November 9.
Livesey, J.C., and M.W. Anders. 1979. In vitro metabolism of 1,2-dihalo-
ethanes to ethylene. Drug Metab. Dispos. 7(4):199-203.
Lyman, W.J. 1982. Octanol Water Partition Coefficient. In; Handbook of
Chemical Property Estimation Methods. McGraw-Hill, Inc., New York, NY.
pp. 1-1 to 1-54.
Moriya, M., T. Ohta, K. Watanabe, T. Miyazawa, K. Kato and Y. Shirasu.
1983. Further mutagenicity studies on pesticides in bacterial reversion
assay systems. Mutat. Res. 116(3-4):185-216.
Nachtomi, E. 1980. Modulation of the mitotic action of ethylene dibromide.
Chera. Biol. Interact. 32:311-319.
Nachtomi, E. 1981. Role of diethyldithiocarbamate in ethylene dibromide
metabolism and covalent binding. Toxicol. Appl. Pharmacol. 57(2):247-253.
Nachtomi, E., and E. Alumot. 1972. Comparison of ethylene dibromide and
carbon tetrachloride toxicity in rats and chicks: Blood and liver
levels; lipid peroxidation. Exp. Mol. Pathol. 16(1):71-78.
Nachtomi, E., and E. Farber. 1978. Ethylene dibromide as a mitogen for
liver. Lab Invest. 38(3):279-283.
Nachtomi, E., and D. Sarma. 1977. Repair of rat liver DNA in vivo damaged
by ethylene dibromide. Biochem. Pharmacol. 26:1941-1945.
Nachtomi, E., E. Alumot and A. Bondi. 1968. Biochemical changes in organs
of chicks and rats poisoned with ethylene dibromide and carbon tetra-
chloride. Isr. J. Chem. 6:803-811.
NCI. 1978. National Cancer Institute. Bioassay of 1,2-dibromomethane for
possible carcinogenicity. NCI Carcinogenicity Tech. Rep. Ser. No. 86.
PB-288-428. 64 pp. [Also publ. as CHHS (NIH) 78-1336.]
NIOSH. 1986. N .tional Institute for Occupational Safety and Health. Semen
Study of Papaya Workers Exposed To Ethylene Dibromide. In Press.
NTP. 1982. National Toxicology Program. Carcinogenesis Bioassay of
1,2-Dibromomethane in F344 rats and B6C3F! Mice. (Inhalation Study).
NTP-80-28. NIH Pub. No. 82-1766.
Olmstead, E.V. 1960. Pathological changes in ethylene dibromide poisoning.
Am. Med. Assoc. Arch. Ind. Health. 21:45-49.
Ott, M.G., B.C. Scharnweber and R.R. Langner. 1980. The mortality experience
of 161 employees exposed to ethylene dibromide in two production units.
Br. J. Ind. Med. 37:163-168.
-------
Ethylene Dibromide -*.«-" J March 31, 1987
-17-
Pflesser, G. 1938. Skin-damaging effect of ethylene dibromide — A constit-
uent of the liquid from remote water gauges. Arch. Gewerbepathol.
Gewerbehyg. 8:591-600. [Cited in NIOSH, 1977].
Plotnick, H.B., and W.L. Conner. 1976. Tissue distribution of 14c-labeled
ethylene dibromide in the guinea pig. Res. Commun. Chem. Path. Pharmacol.
13(2):251-258.
Plotnick, H.B., W.W. Weigel, D.E. Richards and K.L. Cheever. 1979. The
effect of dietary disulfiram on tissue distribution and excretion of
14c-1,2-dibromomethane in the rat. Res. Commun. Chem. Pathol. Pharmacol.
26(3):535-545.
Principe, P., E. Dogliotti, M. Bignami, et al. 1981. Mutagenicity of
chemicals of industrial and agricultural relevance in Salmonella, Strep-
tomyces, and Aspergillus. J. Sci. Food Agric. 32(8):826-832.
Rowe, V.K., H.C. Spencer, D.D. McCollister, R.L. Hollingsworth and E.M.
Adams. 1952. Toxicity of ethylene dibromide determined on experimental
animals. Ind. Hyg. Occup. Med. 6:158-173.
Shih, T.-W., and D.L. Hill. 1981. Metabolic activation of 1,2-dibromoethane
by glutathione transferase and by microsomal mixed function oxidase:
Further evidence for formation of two reactive metabolites. Res. Commun.
Chem. Pathol. Pharmacol. 33(3):449-461.
Short, R.D., J.L. Minor, B. Ferguson, T. Unger and C.C. Lee. 1976. Toxicol-
ogy studies of selected chemicals. Task I: The developmental toxicity
of ethylene dibromide inhaled by rats and mice during organogenesis.
U.S. EPA 560/6-76-018. NTIS PB-256 659. 15 pp.
Short, R.D., J.M. Winston, C.B. Hong, J.L. Minor, C.C. Lee and J. Seifter.
1979. Effects of ethylene dibromide on reproduction in male and female
rats. Toxicol. Appl. Pharmacol. 49(1):97-1 05.
Stenger, V.A. 1983. Bromine Compounds. In; Kirk-Othmer Encyclopedia of
Chemical Technology, 3rd ed., Vol. 4, M. Grayson and D. Eckroth, eds.
John Wiley and Sons, Inc., New York, NY. pp. 243-263.
Tan, E., and \.W. Hsie. 1981. Mutagenicity and cytotoxicity of haloethanes
as studied in the CHO/HGPRT system. Mutat. Res. 90:183-191.
Teramoto, S., R. Saito, H. Aoyama and Y. Shirasu. 1980. Dominant lethal
mutation induced in male rats by 1,2-dibromo-3-chloropropane (DBCP).
Mutat. Res. 77(1):71-78.
Ter Haar, G. 1978. Comments on: EPA's Rebuttal Presumption Against Regis-
tration and Continued Registration of Pesticide Products Containing EDB.
Ethyl Corporation, January 23, 1978. Rebuttal Document No. 48 (30000/25).
Ter Haar, G. 1981. Comments on: EPA's Preliminary Notice of Determination
Concluding the RPAR on EDB. Ethyl Corporation, February 25, 1981.
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Ethylene Dibromide J-O'l March 31, 1987
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Tezuka, H., N. Ando, R. Suzuki, M. Terahata, M. Moriya and Y. Shirasu. 1980.
Sister-chromatid exchanges and chromosomal aberrations in cultured
Chinese hamster cells treated with pesticides positive in microbial
reversion assays. Mutat. Res. 78(2): 177-191.
Turner, D. 1976. Appendix II. Dibromoethane — A Survey of Amlwch Records.
(See Ter Haar, 1978).
Turner, D. 1977. A mortality survey on employees at ethylene dibromide
plant. The Associated Octel Company Limited. (See Ter Haar, 1978).
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cides in drinking water, food, and air. Office of Drinking Hater.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Health effects
criteria document for ethylene dibromide (ED6). Criteria and Standards
Division, Office of Drinking Water. Washington, DC.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Method 502.1. Volatile
halogenated organic compounds in water by purge and trap gas chromatography.
Environmental Monitoring and Support Laboratory, Cincinnati, Ohio 45268,
June 1985.
U.S. EPA. 1985c. Method 524.1. Volatile organic compounds in water by purge
and trap gas chromatography/mass spectrometry. Environmental Monitoring
and Support Laboratory, Cincinnati, Ohio 45268, June 1985.
U.S. EPA. 1985d. Technologies and costs for the removal of synthetic organic
chemicals from potable water supplies. First Draft. Science and Tech-
nology Branch, CSD, ODW. U.S. EPA, Washington, DC.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51(185):33992-34003.
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance index for
pesticides. Bureau of Foods.
van Bladeren, P.J., D.D. Breimer, G.M.T. Rotteveel-Smijs, et al. 1980. The
role of glutathione conjuga-ion in the mutagenicity of 1,2-dibromoethane.
Biochem. Pharmacol. 29(21):2975-2982.
van Bladeren, P.J., D.D. Breimer, J.A.T.C.M. Van Huijgevoort, N.P.E. Vermeulen
and A. Van der Gen. 1981. The metabolic formation of n-acetyl-S-2-
hydroxyl-L-cysteine from tetradeutero-1,2-dibromoethane. Relative
importance of oxidation and glutathione conjugation in vivo. Biochem.
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2nd ed. Van Nostrand Reinbold Co., New York, NY. pp. 635-636.
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Fate of inhaled ethylene dibromide in rats. Toxicol. Appl. Pharmacol.
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Ethylene Dibromide o March 31, 1987
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Heast, R.C., Ed. 1980*. Handbook of Chemistry and Physics, 61st ed. CRC
Press, Inc., Boca Raton, FL. p. C309.
Wong, O., H.M.D. Utidjian and V.S. Karten. 1979. Retrospective evaluation
of reproductive performance of workers exposed to ethylene dibromide
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63(2):155-165.
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March 31, 1987
HEPTACHLOR AND HEPTACHLOR EPOXIDE
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. The chemical 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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Heptachlor and Heptachlor Epoxide
March 31, 1987
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Ihis Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Docuaent (CD) for Heptachlor,
Beptachlor Epoxide and Chlordane (U.S. EPA, 1985a). The HA and CD format*
are similar for easy reference. Individuals desiring further information on
the toxicological 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 Rational Technical Information Service, D.S.
Department of Commerce, 5285 Port Royal Rd., Springfield, VA 22161, PB i
86-117991/AS. The toll free number is (800) 336-4700) in the Washington, D.C.
areas (703) 487-4650.
II. GENERAL INFORMATION AMD PROPERTIES
CAS Mo. 76-44-8
Structural Formula
A. Heptachlor
Synonyms
Use
3-Chlorochlordenei 3, 4, 5, 6, 7, 8,8a-heptachlorodicyclopentadiene;
1* 4» 5, 6, 7, 8,8-hept«chloro-3a, 4,7,7a-tetrahydro-4,7-endomethanoindene,
Insecticide
Properties
Chemical Formula
Moleealar Weight
Physical State (room temp.)
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
373.32
white, crystalline solid
135-145*C (at 1-1.5 ma Hg)
93»C
3 jc 10-4 mm Hg (at 25-C)
0.056 mg/L (at 25*C)
3.87
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Heptachlor and Heptachlor Epoxide March 31, 1987
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I. Heptachlor Epoxide
CAS Mo. 1024-57-3
Structural Formula
Cl
Cl
Synonyms
• 1,4,5,6,7,8,8.-Heptachloro-2,3-epoxy-3a,4,7,7a-tetrahydro-4,7-methanoindan
Use
• Insecticide
Properties
Chemical Formula CioHsClTO
Molecular Weight 389.32
Physical State (room temp.) solid
Boiling Point
Melting Point 160-161.5*C (99.5% pure)
Density —
Vapor Pressure 3 x 10'4 mm Eg (at 25"C)
Mater Solubility 0.35 mg/L (at 25«C)
Log Octanol/Water Partition 2.65, 4.43, 5.40 (by 3 methods)
Coefficient
Taste Threshold —
Odor Threshold —
Conversion Factor —
Occurrence
• Heptachlor is an insecticide which in the past has been used on corn,
alfalfa, hay and vegetables* and as a terniticide. During the mid-
70s,
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Heptachlor and Heptachlor Epoxide March 31, 1987
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• Heptachlor epoxide, but not heptachlor itself, is a common low level
contaminant in food. Heptachlor has been detected in air at very low
levels, approximately 1 ppt. However, the available data are insuf-
ficient to evaluate exposures from these areas or to determine if
drinking water is a significant route of exposure.
III. PHARMACOKINETICS
Absorption
0 Heptachlor was absorbed rapidly from the gastrointestinal tract of rats
following intragastric administration as evidenced by its detection in
blood within one hour after dosing (Mizyukova and Kurchatov, 1970).
Distribution
0 In female rats, intragastrically administered heptachlor was detected
in blood, liver, kidney and adipose tissue within one hour. After
four hours, heptachlor epoxide was detected in blood, liver and fat,
persisting in adipose tissue for 3 to 6 months (Mizyukova and Kurchatov,
1970). With dietary administration of heptachlor to rats for two
months or to dogs by capsule for 12 to 18 months, Radomski and Oavidow
(1953) reported similar tissue distribution. Heptachlor epoxide
levels in the fat of female rats, however, were about 5 to 10 times
higher than those in male rats. Retention in adipose tissue was 6
and 8 weeks for male and female rats respectively.
0 Heptachlor epoxide has been detected in tissue samples from 77 autopsies
performed from 1966 to 1968 at 1 to 32 ppb per whole tissue, with
highest concentrations in bone marrow and liver (Klemmer et al., 1977).
0 Heptachlor epoxide has been detected in human adipose tissue in surveys
conducted in Great Britain (Abbott et al., 1972; 1981), Brazil (Wasser-
mann et al., 1972), Japan (Curley et al., 1973), Israel (Wassermann
et al., 1974), Texas (Burns, 1974), Louisiana (Greer et al., 1980)
and the United States (Kutz et al., 1979; Sovocool and Lewis, 1975).
0 Evidence of transplacental transfer of heptachlor or heptachlor
epoxide in humans (levrls of 0.01-0.3 mg/kg in fat; 0.001 mg/L in
blood) comes from a study by Cur ley et al. (1969), who detected
heptachlor epoxide in adipose tissue, brain, adrenals, lungs, heart,
liver, kidney and spleen of ten stillborn babies and two babies who
died soon after birth and in 27 of 30 samples of cord blood from
healthy neonates.
Metabolism
0 Metabolism of heptachlor to heptachlor epoxide ir± vitro was similar
using rat and human liver microsomal preparations. A major species
difference was that four times more heptachlor epoxide was formed in
the rat system than in the human system (Tashiro and Matsumura, 1978).
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Heptachlor and Heptachlor Epoxide March 31, 1987
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The major fecal netabolites of orally administered heptachlor in rats
include heptachlor epoxide,. 1-hydroxychlordene, and 1-hydroxy-2,3-
epoxychlordene (Tashiro and Matsunura, 1978).
Excretion
The major route of heptachlor elimination by rats is via the feces,
amounting up to 50% of the administered oral dose over 10 days
(Tashiro and Matsumura, 1978). Urinary excretion of metabolites
amounted to <5% of the dose.
The only information available on human excretion of heptachlor are
reports of heptachlor epoxide detected in miIX of lactating women
(Kroger, 1972; Ritcey et al., 1972; Savage et al., 1973; Bakken and
Seip, 1976; Polishuk et al., 1977; Strassman and Xutz, 1977; Takahashi
et al., 1981).
IV. HEALTH EFFECTS
Humans
Clinical case studies of acute exposure (via ingestion, dermal or
inhalation routes) to chlordane containing heptachlor document a
pattern of CHS effects similar to that found in animals (e.g.,
irritability, salivation, labored respiration, muscle tremors,
convulsions, etc.HDadey and Kammer, 1953; Derbes et al., 1955).
Several blood dyscrasias (e.g., anemias and leukemias) are associated
with inhalation and dermal exposure of humans to heptachlor (Furie
and Trubowitz, 1976; Klemmer et al., 1977; Infante et al., 1978).
*
Wang and McMahon (1979) reported a non-significant increased incidence
of lung cancer and a statistically significant increased incidence of
cerebrovascular disease in a cohort of 1403 white male workers
employed for 73 months in the production of chlordane and heptachlor.
Animals
Short-term Exposure
Symptoms of acute intoxication from heptachlor or heptachlor epoxide
include tremors, convulsions, paralysis and hypothermia (Hrdina
et al., 1974; Yamaguchi et al., 1980).
Acute oral LDso values in rats for heptachlor range from 40 mg/kg for
a commercial formulation (Ben-Dyke et al., 1970) to 162 mg/kg for
technical grade heptachlor (Gaines et al., 1960).
The acute oral LDso value for heptachlor epoxide in rats ranged from
46.5 to 60 mg/k? (HAS, 1977; Sperling and Ewinike, 1969; Podowski
et al., 1979).
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Heptachlor and Heptachlor Epoxide March 31, 1987
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A single, acute oral dose of 60 mgAg heptachlor in rats was associ-
ated with increased levels of serum GPT and serum aldolase, and
moderate to severe histological liver damage (Krampl, 1971).
Evidence of significant liver damage and altered liver function was
reported in rats maintained on diets containing heptachlor at 7 to 12
mgAg bw/day for up to 14 days (Xrampl, 1971) and 10 mg/kg diet for
5 to 7 days (Enan et al., 1982).
A dose-related significant induction of liver microsomal enzymes, at
dietary levels of heptachlor of 2 to 50 mg/kg diet for 14 days, was
observed in rats (Den Tonkelaar and Van Esch, 1974).
Long-term Exposure
At dietary levels of 10 mgAg of heptachlor or heptachlor epoxide in
mice for 2 years, Reuber (1977a) diagnosed hepatic vein thrombosis
and cirrhosis of the liver from slides of the Davis (1965) study.
In the JRDC (1973) study, reviewed by Epstein (1976), a 75% heptachlor
epoxide and 25% heptachlor mixture was fed to mice for 18 months;
females and males had dose-related significantly increased mean liver
weights and hepatocytomegaly at 1, 5 and 10 mgAg diet.
Jolley et al. (1966) found dose-related increased mortality in rats
fed 5 to 12.5 mgAg diet levels of a 75% heptachlor and 25% heptachlor
epoxide mixture for 2 years.
Witherup et al. (1955) found non-neoplastic lesions in rats at
dietary levels ±7.0 mgAg diet of heptachlor for 110 weeks. Treated
males had dose-related increased liver weights at levels of 3 to 10
mgAg diet.
Dose-related liver weight increases, hepatocytomegaly and hepatic
cell vacuolization were observed in rats maintained for 108 weeks on
diets containing heptachlor epoxide at 0.5-10 mgAg diet (Witherup
et al., 1959).
Dose-related changes in clinical measurements related to liver
function and microscopic changes in liver were noted in dogs admini-
stered hepte rhlor epoxide in the diet &t dose levels of 3, 5, 7 and
10 mgA9/day for 2 years (U.S. EPA, 1971; IRDC, 1973).
Beagle dogs from 23 to 27 weeks of age were given diets containing
0, 0.5, 2.5, 5 or 7.5 mg/kg/da.y of heptachlor epoxide for 60 weeks.
Results included liver weight to body weight ratios which were
significantly increased in a treatment-related fashion. Effects were
noted for both males and females at the dose of 0.5 ppm. No NOEL was
determined from this study (U.S. EPA, 1958, Kettering Laboratory).
Reproductive Effects
No information was found in the available literature on the reproductive
effects of heptachlor or heptachlor epoxide.
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Developmental Effects
No information was found in the available literature on the develop-
mental effects of heptachlor or heptachlor epoxide.
Mutagenicity
Heptachlor has been tested for mutagenicity in a number of systems.
Negative results have been obtained in bacterial systems (Moriya
et al., 1983; Probst et al., 1981; Gentile et al., 1982; Shirasu
et al., 1976), in mitotic gene conversion (Gentile et al., 1982),
in the recessive lethal assay in fruit flies (Benes and Sram, 1969),
in assays -for unscheduled DNA synthesis in rat, mouse and hamster
primary hepatocyte cultures (Probst et al., 1981; Maslansky and
Williams, 1981), and for the dominant lethal assay in mice (Arnold
et al., 1977).
Positive results were reported for unscheduled DNA synthesis in
transformed human fibroblasts with S-9 activation (Ahmed et al.,
1977) and in the dominant lethal assay in rats (Cerey et al., 1973).
Heptachlor epoxide was negative in bacterial systems (Moriya et al.,
1983; Marshall et al., 1976), in the recessive lethal assay in fruit
flies (Benes and Sram, 1969) and in the dominant lethal assay in mice
(Arnold et al., 1977).
Heptachlor epoxide was positive for unscheduled DNA synthesis in
human fibroblasts in the presence of S-9 (Ahmed et al., 1977).
Carcinogenicity
In a National Cancer Institute bioassay (NCI, 1977), heptachlor was
tested for possible* carcinogenicity in male and female mice and rats.
Male B6C3Fj mice received heptachlor at dietary concentrations of 0,
6.1 and 13.8 mg/kg diet and female B6C3Fi mice received diets contain-
ing 0, 9.0 and 18.0 mg/k9 diet, both for 80 weeks. The incidence of
hepatocellular carcinomas was statistically significant in the high-
dose males, while a highly significant dose-related trend also was
observed between high- and low-dose females. Heptachlor was not
carcinogenic in male and female Osborne-Mendel rats similarly treated
with concentrations of 25.7 to 77.9 mg/kg diet.
Re-analysis of the study results reported by Witherup et al. (1955),
indicate that administration of heptachlor to male and female CF rats
at dietary levels of 1.5 to 10.0 ppm (mg/kg diet) for 110 weeks
resulted in a statistically significant increase in malignant and any
tumors in multiple sites in some female test groups (Epstein, 1976).
Significantly increased incidences of hepatic carcinoma were determined
by Reuber and Williams (Epstein, 1976) upon re-analysis of histologic
slides from the Witherup et al. (1959) study. Witherup administered
heptachlor epoxide to male and female CFN rats at doses of 0, 0.5,
2.5, 5*0, 7.5 and 10.0 mg/kg diet for 108 weeks. Die incidences were
significantly different from controls for female rats at the 5 and
10 mg/kg dietary concentrations (Epstein, 1976).
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• Histological re-examination of the slides from the Davis (1965) study
resulted in a conclusion of significantly increased incidence of
hepatocellular carcinoma in C3H mice receiving 10 mg/kg diet of
heptachlor epoxide for 728 days (Reuber, 1977b).
• Histological re-examination of the slides of the IRDC (1973) study
resulted in a conclusion of significantly increased incidence of
hepatocellular carcinoma in CD-1 mice administered a 75:25 mixture of
heptachlor epoxide:heptachlor in the diet at concentrations of 1.0,
5.0 or 10.0 ppm (mgA9 diet) for 18 months (Reuber, 1977b).
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 ( ug/L)
(OF) x ( L/day)
where:
NOAEL or LOAEL - No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW m assumed body weight of a child (10 kg) or
an adult (70 kg).
UF - uncertainty factor (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
There are insufficient toxicological data available to derive a One-day
HA for heptachlor or heptachlor epox/de. The Ten-day HA, however, would be
protective for a One-day exposure period for heptachlor of 0.01 mg/L.
Ten-day Health Advisory
A Ten-day HA for heptachlor can be derived from a study conducted by
Enan et al. (1982) in which rats were administered heptachlor at 1.0 mg/kg/day
(10 ppm) in the feed for 14 days. Exposure resulted in evidence of liver
damage and altered liver function: increased blood urea, increased blood
glucose, decreased liver glycogen content, and increased acid and alkaline
phosphatase levels when compared with controls. Using 1.0 mg/kg/day as the
LOAEL, the Ten-day HA for the 10 kg child is calculated as follows:
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164
-9-
. U9/L)
where:
1.0 «g/kg/day - LQAEL based on liver effects in rats.
10 kg » Assumed body weight of a child*
100 « 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.
No data are available from which to derive a Ten-day HA for heptachlor
epoxide .
Longer-term Health Advisory
There are insufficient toxicological data available to derive a Longer-
term HA for heptachlor or heptachlor epoxide. The DWEL of 0.0015 mg/L adjusted
for a 10-kg child is recommended as a conservative estimate for a longer-term
exposure .
Lifetime Health Advisory for Heptachlor
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). Prom 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 multip. ication 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 classifed 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 study by Witherup et al. (1955) is the most appropriate from which
to derive the DWEL. Investigators studied the effects of heptachlor on
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groups of 20 male and 20 female CF rats. The compound was administered at
dietary concentrations of 0, 1.5, 3, 5, 7 or 10 ppm (10 mg/kg/dose) of
heptachlor. Mortality among test groups was not dose-related. Loss of body
weight, decreased food consumption and increased liver weights were seen
among treated males. Lesions in the liver were limited to 7 ppm and above
and were characteristic of chlorinated hydrocarbons, i.e., hepatocellular
swelling, homogeneity of the cytoplasm and peripheral arrangements of the
cytoplasmic granuled of cells of the central zone of the liver lobules.
The NOEL for increased liver to body weight for males only was 3 ppm and LEL
was 5 ppm. [Note: A re-analysis of the Witherup et al. (1955) dietary study
on the toxicity of heptachlor to rats (by the OPP, RfD Work Group, 1987)
indicated that the NOEL of 3 ppm (0.15 mg/kg/day) for increased liver to body
weight for male rats was the most appropriate for a Lifetime Health Advisory
for heptachlor.] Using this NOEL, the DWEL is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (0'15 mgAg/day) . Q.0005
(300)
where:
0.15 mg/kg/day (3 ppm) * NOEL based on absence of increased liver
to body weight for male rats.
300 « uncertainty factor, chosen in accordance with
NAS/ODW guidelines for use with a NOAEL from
an animal study (also RfD meeting, April 16,
1987).
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - <0*0005 mgAg/day) (70 kg) , 0.0175 mgA (17.5 ug/L)
(2 L/day)
where:
0.0005 mg/kg/day « RfD.
70 kg • assumed weight of an adult.
2 L/day « assumed daily water consumption of an adult.
Heptachlor is classified as Group B: Probable human carcinogen. The
estimated excess cancer risk associated with lifetime exposure to drinking
water containing heptachlor at 17.5 ug/L is approximately 3 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.
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Heptachlor and Heptachlor Epoxide March 31, 1987
Lifetime Health Advisory for Heptachlor Epoxide
Two studies in dogs are the most appropriate from which to derive the
DWEL. In the 60-week dog feeding study (U.S. EPA, 1958) beagle dogs from 23
to 27 weeks of age were divided into five groups (three females and two males)
and were given diets containing 0, 0.5, 2.5, 5 or 7.5 ppm of heptachlor
epoxide. Results included liver weight to body weight ratios which were
significantly increased in a treatment-related fashion. Effects were noted
for both males and females at the 0.5 ppm (0.0125 mg/kg/day) dose level of
heptachlor epoxide. No NOEL was determined for the study. In another two-
generation reproduction study in dogs (U.S. EPA, 1971) animals were administered
diets containing various dose levels of heptachlor epoxide. Die dose levels
were 0, 1, 3, 5, 7 or 10 ppm of heptachlor epoxide in the diet. This study
was designed to investigate reproduction parameters associated with heptachlor
epoxide administration. The OPP and the RfD Work Group considered that the
former study in dogs, 60-week dog feeding study providing the LEL of 0.5 ppm
(0.0125 mg/kg/day) is the most appropriate for the derivation of the DWEL.
Using this LEL, the DWEL is derived as follows:
Step Is Determination of the Reference Dose (RfD)
RfD « (0.0125 mg/kg/day) » 0.000013 mgAg/day
(1,000)
Where:
0.0125 mg/kg/day - Low Effect Level (LEL).
1,000 - uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of £he Drinking Water Equivalent Level (DWEL)
DWEL = (0.000013 mgAg/day) (70 kg) = Q.00044 mg/L (0.4 ug/L)
(2 L/day)
Where:
0.000013 mgAg/day « RfD.
70 kg « assumed weight of an adult.
2 L/day » assumed daily water consuption of an adult.
Heptachlor epoxide is classified in Group B: Probable human carcinogen.
The estimated excess cancer risk associated with lifetime exposure to drinking
water containing heptachlor epoxide at 0.4 ug/L is approximately 2 x 10~3.
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 consid-
erable uncertainty as to the accuracy of risks calculated by this methodology.
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Heptachlor and Heptachlor Epoxide March 31, 1987
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Evaluation of Carcinogenic Potential
• The U.S. EPA (1987) derived a human carcinogenic potency factor, qi*,
of 4.5 (mgAg/day)-1 for heptachlor. This derivation was based on
the geometric mean of four potency estimates which were based on the
incidence of hepatocellular carcinoma in male and female CH3 mice
(Davis, 1965, as diagnosed by Reuber, 1977b) and Bale and female
B6C3F-) mice (MCI, 1977). this estimate supersedes the potency of
3.37 (mg/kg/day)-1 previously calculated by the U.S. EPA (1980). The
concentrations in drinking water corresponding to increased lifetime
risk levels of 10-4, 10-5 and 10-6 for a 70 kg human consuming 2 L/day
are 7.6, 0.76 and 0.076 ug/L, respectively (U.S. EPA, 1987).
• Cancer risk estimates (95% upper limit) with other models are presented
for comparison with that derived with the multistage. For example,
one excess cancer per 1,000,000 (10-6) is associated with exposure to
heptachlor epoxide at levels pf <0.0001 ug/L (probit), <0.00001 ug/L
(logit) and <0.0001 ug/L (Heibull).
0 The U.S. EPA (1987) derived a human carcinogenic potency factor, qi*r
of 9.1 (mg/kg/day)"1 for heptachlor epoxide. This derivation was
based on the geometric mean of four potency estimates which were
based on the incidence of hepatocellular carcinoma in male and female
CH3 mice (Davis, 1965, as diagnosed by Reuber, 1977b) and male and
female CD-1 mice (IRDC, 1973). this estimate supersedes the potency
of 5.786 (mg/kg/day)-1 previously calculated by the U.S. EPA. The
concentrations in water corresponding to increased lifetime risk
levels of ID'4, 10"5 and 10~6 for a 70 kg human consuming 2 L/day
are 3.8, 0.38 and 0.038 ug/L, respectively (U.S. EPA, 1987).
0 The HAS (1977) determined 0.119 ug/L for heptachlor as the water
concentration corresponding to an increased lifetime risk of cancer
of 10-5. HAS (1977) categorizes heptachlor epoxide as a suspect animal
carcinogen, but noted that there are insufficient data to permit a
statistical extrapolation of- risk.
• IARC (1979) classified heptachlor as Group 3: inadequate evidence
of carcinogenicity in humans and limited evidence of carcinogenicity
in animals. The IARC (1979) position on heptachlor epoxide is that
there is limited evidence that heptachlor epoxide is carcinogenic in
experimental animals.
• Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), heptachlor and heptachlor epoxide
is 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 In 1980, EPA estimated a range of excess cancer risks for lifetime
exposure to heptachlor when developing ambient water quality criteria
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Heptachlor and Heptachlor Epoxide March 31, 1987
-13-
(U.S. EPA, 1980). This range was 2.78 ng/L/ 0.28 ng/L and 0.028 ng/L,
respectively, for risks of 10-5, 10-6 and 10-7, assuming consumption
of 2 liters of water and 6.5 grains of contaminated fish per day by
a 70 kg adult.
• FAO/WHO recommended an ADI value of 0.5 ug/kg bw for heptachlor or
heptachlor epoxide. This recommendation was established by the Joint
FAO/WHO Expert Committee on Food Additives (FAO/WHO, 1978).
• A guideline value of 0.1 ug/L in drinking water also was recommended
by the WHO (1984), based upon this level as one percent of the ADI.
• The American Conference of Governmental Industrial Hygienists (ACGIH,
1983) has adopted TWA-TLVs of 0.5 mg/m3 for heptachlor in workroom
air.
0 It should be noted that an estimated concentration for detection by
taste and odor in water for heptachlor.was 0.02 mg/L (Sigworth, 1965).
VII. ANALYTICAL METHODS
0 Determination of heptachlor is by a liquid-liquid extraction gas
chromatographic procedure (U.S. EPA, 1978; Standard Methods, 1985).
Specifically, the procedure involves the use of 15% methylene chloride
in hexane for sample extraction, followed by drying with anhydrous
sodium sulfate, concentration of the extract and identification by
gas chromatography. Detection and measurement is accomplished by
electron capture, microcoulometric or electrolytic conductivity gas
chromatography. Identification may be corroborated through the use
of two unlike columns or by gas chromatography-mass spectroscopy
(GC-MS). The method sensitivity is 0.001 to 0.010 ug/L for single
component pesticides and 0.050 to 1.0 ug/L for multiple component
pesticides when analyzing a 1-liter sample with the electron capture
detector.
nil. TREATMENT TECHNOLOGIES
0 Treatment technologies which are capable of removing heptachlor from
drinking water include adsorption by granular activated carbon (GAG)
and ozone (03) or ozone/ultraviolet oxidation (O3/UV).
• Dobbs and Cohen (1980) developed adsorption isotherms for a number of
organic chemicals in drinking water, including heptachlor. Based on
the isotherm data, they reported that the activated carbon Filtrasorb®
300 exhibited adsorptive capacities of 45 mg, 18 mg and 8 mg of
heptachlor per gm of carbon at equilibrium concentrations of 100
ug/L, 10 ug/L, and 1 ug/L, respectively.
• The GAC system in -U.S. EPA's Hazardous Materials Spills Treatment
Trailer was used to treat 104,000 gal of pesticide-contaminated water
containing heptachlor. Water analysis showed 6.1 ug/L of heptachlor
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Heptachlor and Heptachlor Epoxide March 31, 1987
-14-
in the contaminated water. Hinety-nine percent heptachlor removal
was achieved at a contact time of 17 minutes (U.S. EPA, 1985b).
Hansen (1977) reported on the efficiency of GAC used in Mount Clements
water treatment plant to remove synthetic organic chemicals from the
raw water source. Heptachlor epoxide was detected in the raw water
at concentrations of 220 ng/L. Ihe GAC column reportedly was capable
of removing 99.9+ percent (below its detectable limit) of the heptachlor
epoxide.
Gilbert (as referenced in U.S. EPA, 1985b) summarized the results pre-
sented by a number of different researchers on the ability of ozone
to remove several SOCs from drinking water, including heptachlor.
The results indicate that greater than 99% of the heptachlor was
removed by ozone oxidation, while heptachlor epoxide was only partially
removed (i.e., 26%) at an applied ozone dose of 17 mg/L.
Treatment technologies for the removal of heptachlor from drinking
water have not been extensively evaluated (except on an experimental
level). An evaluation of some of the physical and/or chemical
properties of heptachlor indicates that the following techniques
would be candidates for further investigation: adsorption by granular
activated carbon and ozone oxidation. Whichever individual or combi-
nations of technologies for heptachlor reduction are used, it must be
based on a case-by-case technical evaluation, and an assessment of
the economics involved.
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IX. REFERENCES
Abbott, D.C., G.B. Collins and R. Goulding. 1972. Organochloride pesticide
residues in human fat in the United Kingdom 1969-1971. Br. Med. J.
2:553-556.
Abbott, D.C., G.B. Collins, R. Goulding and R.A. Hoodless. 1981. Organo-
chlorine pesticide residues in human fat in the United Kingdom 1976-1977.
Br. Med. J. 283(6304):1425-1428.
ACGIH. 1983. American Conference of Governmental Industrial Hygienists.
TLVs: Threshold limit values for chemical substances and physical agents
in the work environment with intended changes for 1983-1984. Cincinnati,
OH. pp. 14, 21.
Ahmed, F.E., R.W. Hart and N.J. Lewis. 1977. Pesticide-induced DNA damage
and its repair in cultured human cells. Mutat. Res. 42:161-174.
Arnold, D.W., G.L. Kennedy, Jr., M.L. Keplinger, J.C. Calandra and C.J. Calo.
1977. Dominant lethal studies with technical chlordane, HCS-3260 and
heptachlor:heptachlor epoxide. J. Tbxicol. Environ. Health. 2:547-555.
Bakken, A.F., and M. Seip. 1976. Insecticides in human breast milk. Acta.
Paediatr. Scand. 65:535-539.
Ben-Dyke, R., D.M. Sanderson and D.N. Noakes. 1970. Acute toxicity data
for pesticides. Wildl. Rev. Pestic. Control. 9:119.
Benes, V., and R. Sram. 1969. Mutagenic activity of some pesticides in
Drosophila melanogaster. Ind. Med. 38:50-52.
Burns, J.E. 1974. Organochlorine pesticide and polychlorinated biphenyl
residues in biopsied human adipose tissue - Texas, 1969-1972. Pestic.
Monitor. J. 7:122.
Cerey, K., V. Izakovic and J. Ruttkay-Nedecka. 1973. Effects of heptachlor
on dominant lethality and bone marrow in rats. Mutat. Res. 21:26.
Curley. A., M.F. Copeland and R.K. Kimbrough. 1969. Chlorinated hydrocarbon
insecticides in organs of stillborn and blood of newborn babies. Arch.
Environ. Health. 19:628-632.
Curley. A., V.H. Burse, R.W. Jennings, E.C. Villaneuva, L. Tomatis and
K. Akazaki. 1973. Chlorinated hydrocarbon pesticides and related
compounds in adipose tissue from people of Japan. Nature. 242:338-340.
Dadey, J.L., and A.G. Rammer. 1953. Chlordane intoxication. J. Am. Med.
Assoc. 153:723.
Davis, H.J. 1965. Pathology report on mice fed aldrin, dieldrin, hepta-
chlor or heptachlor epoxide for two years. Internal FDA memorandum to
Dr. A.J. Lehman, July 19. (Cited in Epstein, 1976)
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Heptachlor and Heptachlor Epoxide March 31, 1987
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Den Tonkelaar, E.M., and G.J. Van Each. 1974. No-effect levels of organo-
chlorine pesticides based on induction of microsomal liver enzymes in
short-term toxicity experiments. Toxicology. 2:371.
Derbes, V.J., J.H. Dent, W.H. Forrest and M.F. Johnson. 1955. Fatal
chlordane poisoning. JAMA. 158:1367-1369.
Dobbs, R.A., and J.M. Cohen. 1980. Carbon adsorption isotherms for toxic
organics. Office of Research and Development. EPA 600/8-80-023.
En an, E.E., A.H. El-Sebae and O.H. Enan. 1982. Effects of some chlorinated
hydrocarbon insecticides on liver function in white rats. Meded. Fac.
Landbouwwet., Rijksuniv. Gent. 47(1):447-457.
Epstein, S.S. 1976. Carcinogenicity of heptachlor and chlordane. Sci. Total
Environ. 6:103.
FAO/HHO. 1978. Food and Agricultural Organization/World Health Organization.
FAO Plant Production and Protection Paper 10 Rev. Pesticides Residues in
Food - 1977. Rep. Joint Meet. FAO Panel of Experts on Pesticide Residues
and Environment and the WHO Expert Committee on Pesticide Residues, Rome.
FDA. 1980a. Food and Drug Administration. Compliance program report of
findings. FY77 total diet studies — Adult (7320.73). Food and Drug
Administration, U.S. Department of Health, Education and Welfare,
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FDA. 1980b. Food and Drug Administration. Compliance program report of
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FDA. 1982b. Food and Drug Administration. Compliance program report of
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172
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177
LINDANE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Hater (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. The chemical 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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Lindane March 31, 1987
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This Health Advisory (HA) is based upon infornation presented in the
Office of Drinking Water's Draft Health Effects Criteria Document (CD) for
Lindane (U.S. EPA, 1985a). The HA and CD formats are similar for easy
reference. Individuals desiring further information on the toxicological
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 • 86-117819/AS.
The toll free number is (800) 336-4700; in the Washington, D.C. area:
(703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 58-89-9
Structural Formula
a
Synonyms
• Gamma-hexachlorocyclohexane
Gamma-benzene hexachloride
Kwell
Uses
• Lindane has been used in the control of various wood-inhabiting
beetles, seed treatment, and pharmaceutical preparations (1% lotion,
cream or shampoo) as a scabicide and pediculocide
Properties (U.S. EPA, 1985a)
Chemical Formula CgHgClg
Molecular Weight 290.85
Physical State white crystals
Boiling Point 323.4°C
Melting Point 112.5»C
Density 1.85
Vapor Pressure (0.094-3.3) x 10~4 mm Hg (20°C)
Water Solubility 7.3-7.9 mg/L (25°C)
Log Octanol/Water Partition 3.61-3.72
Coefficient
Taste Threshold
Odor Threshold —
Conversion Factor —-
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Lindane 179
March 31, 1987
-3-
Occurrence
Lindane is imported into the U.S. Import levels are confidential,
but in the late 1970s, less than one million Ibs were imported.
Lindane is degraded poorly in the environment. Lindane is hydrolyzed
poorly and undergoes biodegradation slowly. Soil half lives are
reported to be on the order of TOO days. Lindane is relatively immobile
in soil and migrates slowly. However, lindane has a slight vapor
pressure and does volatilize from soil. Once in air, lindane photo-
degrades. Lindane has been reported to bi©accumulate; however, its
potential is limited since it can be metabolized by plants and animals.
Lindane has not been found in large amounts in drinking water. Only
1 ground water sample out of 71 in the Rural Water Survey reported a
measurable level of lindane: 0.006 ug/L. No water system has
reported exceeding the interim drinking water standard of 4 ug/L.
Lindane has been found in a few non-drinking water surface and ground
waters in areas near its agricultural use. Level up to 0.5 ug/L have
been reported. Lindane has been found in low levels in food and air.
The current information is insufficient to indicate which is the major
route of exposure for lindane.
III. PHARMACOKINETICS
Absorption
Fasted IRC rats absorbed 70.7 percent of an intragastrically admini-
stered dose of 1 mgAg lindane 60 minutes after treatment (Ahdaya
et al., 1981).
Albro and Thomas (1974) estimated 95-99 percent absorption of technical
grade lindane within 4 days following single oral doses. Variations
of dosage rates .from 30-120 mgAg had no influence on the proportion
absorbed.
Human studies of topically applied pharmaceutical preparations con-
taining 0.3-1.0 percent lindane (Ginsburg et al., 1977; Hosier et al.,
1979; Lange et al., 1981) showed ready absorption. Peak blood levels
were obtained within 6 hours.
Distribution
Technical grade lindane preferentially accumulated in the fatty
tissue of albino rats when fed at 2.5 mgAg bw in the diet (Chand and
Ramachandran, 1980). Accumulation in the brain also has been reported
(Lakshmanan et al., 1979).
Extensive accumulation of lindane occurs in the milk of exposed women
(Siddigui et al., 1981). Lindane also has been shown to enter the
fetus through the placenta (Poradovsky et al., 1977).
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Lindane March 31, 1987
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Metabolism
Metabolism of lindane in humans entails dehydrochlorination to form
cyclohexene derivatives and various chlorinated phenols by way of
either oxicative or nonoxidative pathways (U.S. EPA, 1985a).
Fitzloff et al. (1982) reported that human liver microsomes converted
lindane to hexachlorocyclohexene, 1,3,4,5,6-pentachlorocyclohexene,
2,4,6-trichlorophenol, 2,3,4,6-tetrachlorophenol and pentachlorobenzene.
Engst et al. (1979) observed that lindane was metabolized to tri- and
pentachlorophenols when inhaled by humans.
The half-life of radioactive lindane in rats was 3 to 5 days (Engst,
et al., 1979). Kujawa et al. (1977) administered lindane orally to
rats (8 mg/kg bw) after which they studied the nature of the metabo-
lites in urine, liver and blood. The major products found in urine
were pentachlorophenol, 2,3,4,6- and 2,3,5,6-tetrachlorophenol and
2,4,6-trichlorophenol. Metabolites in the blood were the same as
those found in urine. In the liver, 2,3,4,5,6-pentachlorobenzene and
pentachlorcyclohexene were found in addition to the tetrachloro-
phenols. The kidneys contained considerably higher levels of the
pentachlorocyclohexene than did the liver. Pentachlorocyclohexene
also was detectable in the spleen, heart and brain. No metabolites
were found in the adrenals.
Lindane has been shown to induce increases in levels of xenobiotic
metabolizing enzymes in the liver in several studies (Lowy et al.,
1977; Plass et al., 1981; RCC, 1983).
Excretion
Even after prolonged administration, lindane is eliminated completely
from the body soon after application is terminated. Frawley and
Fitzhugh (1949) demonstrated that, in rat fatty tissue, a lindane
concentration of 102 mgfkg (102 ppm) dropped to zero 1 week after
administration of lindane was discontinued. Lehman (1952a,b) demon-
strated that a concentration of 281 mg/kg (281 ppm) in the fatty
tissue was eliminated completely within 2 weeks. Kitamura et al.
(1970) fed rats a diet containing 10 mg/kg of gamma lindane over a
20-day pef'od. One day after return to a normal diet, no residue
could be detected in the body.
Very little lindane is excreted unaltered. Laug (1948) detected only
about 4% gamma lindane in the urine of rats fed lindane in the diet
(dosage unspecified). No reports of unaltered gamma lindane excretion
following intraperitoneal injection have been located.
Glutathione, glucuronide and sulfate conjugates of lindane metabolites
have been reported (Chadwick et al., 1978; Kurihara et al., 1979).
The metabolites and conjugation products are excreted mostly in the
urine. Excretion in milk also occurs.
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Lindane 1.SJL March 31, 1987
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IV. HEALTH EFFECTS
Humans
0 Case reports indicate that the acute effects of lindane resulting
from either excessive dermal or oral intake include functional
alterations in the nervous system in the form of seizures and uncon-
trollable eye movements. The effects appear to be reversible, with
full recovery within 1 year of exposure.
0 Lindane appears to have a definite inhibitory effect on white blood
cells (lymphocytes) in vitro. In a study conducted by Roux et al.
(1979), 10~4 M lindane sharply inhibited protein, DNA and RNA synthesis
in cultured lymphocytes, either in the unstimulated, phytohemagglutinin
(PHA)-stimulated or in the lymphoblast state. Lindane treatment
resulted in sharply inhibited PHA-induced mitogenesis in the exposed
lymphocytes.
0 The only reported effects of lindane on the blood cells have been
equivocal (including possible anemia) (U.S. EPA, 1985a).
Animals
Short-term Exposure
0 Lindane has higher acute toxicity than other chlorinated hydrocarbons
because it is absorbed rapidly. Clinical symptoms are apparent soon
after exposure (Lehman, 1951). Its high water solubility and rapid
rate of absorption explain the narrow range between its NOAEL and
lethal doses as compared with wider ranges in similar compounds, such
as DDT (Gunther et al., 1968; Martin, 1971).
0 The single dose oral LD£Q has been shown to vary from a high of 1000
mg/kg bw in mice (Wolfe and Esher, 1980) to 840 mgAg in adult humans
(Engst et al., 1979), 400 mg/kg in pigeons (Blakley, 1982), 180 mgAg
in children (Engst et al., 1979), 125 mgAg in rats (Farkas et al.,
1976) and 60 mgAg in rabbits (Desi et al., 1978).
0 Muller et al. (1981) reported a decrease in motor conduction velocity
in the tail nerve of Wistar rats fed gami^a lindane in the diet for
30 days at doses of 25.4, but not at 12.3 or 1.3 mgAg bw.
0 Desi (1974) measured behavioral endpoints in Wistar rats (8 animals
per group) exposed to lindane up to 40 days at daily intakes of 2.5,
5, 10 and 50 mgAg bw. After approximately 2 weeks of exposure, maze
running times and numbers of errors were increased significantly
at dosages * 5 mgAg* The number of lever presses in an operant
conditioning test (Skinner Box) was increased significantly even at
the 2.5 mgAg dose level, indicating an effect upon irritability.
0 Muller et al. (1981) studied the electroneurophysiological effects
of lindane when fed in the diet to groups of IS-Wistar rats for 30
days. A delay in conduction velocity was observed in animals fed a
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Lindane 1.82 March 31, 1987
-6-
daily dose of 25 mgAg but not 12 or 1.3 mgAg bw. me lindane
metabolite gamma-pentachlorocyclohexene caused a conduction delay when
administered at concentrations of 38-782 mgAg bw.
0 Desi et al. (1978) studied the response of rabbits to Salmonella typhi
vaccine following treatment with lindane at 1.5-12 ag/kg bw given
orally 5 times /week for 5-6 weeks and compared the immunologic behavior
with normal, untreated animals. Six animals were used in each group.
The treated rabbits displayed a dose-related decrease in immunologic
titers, indicating inmunosuppressive effects. Similar results were
reported by Dewan et al. (1980) who found that male and female
albino rats fed lindane (6.25 or 25 mgAg in olive oil on alternate
days for 35 days) displayed immunosuppressive behavior when challenged
with £. typhi and j». paratyphi antigens. Again, the effects were
dose-dependent.
Long-term Exposure
0 In the RCC (1983) study, both male and female rats of the KFM-HAN
(outbred) SPF strain were fed 99.85% pure lindane in the diet at
levels of 0, 0.2, 0.8, 4, 20 and 100 ppm for 84 consecutive days.
Liver hypertrophy, kidney tubular degeneration, hyaline droplets,
tubular casts, tubular distension, interstitial nephritis and
basophilic tubules were seen at the 20 and 100 ppm levels. Effects
were rare and very mild when noted at 4 ppm.
0 Fitzhugh et al. (1950) exposed 10 Wistar rats of each sex per dosage
group to gamma lindane at levels of 5, 10, 50, 100, 400, 800 or 1600
mgA9 in the diet for 2 years or longer. An increase in liver weight
and a the slight degree of kidney and liver damage were noted at 100
in the diet but not at 50
Wolfe and Esher (1980) exposed two strains of wild mice to 200 ppm
lindane in the diet for 8 months with no reported effects on food
consumption, growth rate, mortality, reproduction or behavior. Heisse
and Herbs t (1977) exposed SPF mice to 12.5, 25 or 50 mgAg lindane in
the diet for 80 weeks and reported no fine structural hepatocellular
alterations .
In a study conducted by Fitzhugh et al. and reported by Lehman (1965)
dogs were fed (2 animals /sex/group) 0 or 15 mgAg lindane in the diet
for 63 weeks. No differences were noted for food consumption, hemato-
logical or histopathological parameters. Rivett et al. (1978), fed
beagles (4 dogs /sex/group) 0, 25, 50 or 100 mgAg lindane in the diet
for 2 years. The daily intake of lindane based on measured food
consumption was 0.83, 1.60 or 2.92 mgAg bw, respectively. No effects
were reported for the 25 and 50 ppm groups. At 100 ppm, serum alkaline
phosphatase was increased significantly and the livers were dark,
friable and greatly enlarged.
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Lindane JLS3 March 31, 1987
-7-
Reproductive Effects
0 Palmer et al. (1978b) reported no effects of lindane on reproductive
function.or on the incidence of malformation following dietary admini-
stration of 0, 25, Saor 100 ppm (1.25, 2.5 or 5 mgAg bw) lindane.
0 Mo effects were observed in pregnant rabbits fed lindane on days
6 through 18 of gestation at levels equivalent to 5, 10 and 15 mgAg
bw and to pregnant CFY rats fed the same doses of lindane on gestation
days 6 through 16 (Palmer et al., 1978a).
Developmental Effects
0 Contrary to the results of the above studies, Dzierzawski (1977)
reported a 2- to 20-fold increase in resorbed fetuses in hamsters
treated with 20 or 40 mgAg lindane on day 8 of pregnancy. Similar
results were obtained in rats treated with 50 or 100 mgAg on day 9
of pregnancy and 40 mgAg doses on days 6, 8 and 10, and in rabbits
treated with 40 or 60 mgAg on day 9. While the three reports
presented above indicate that there is no evidence of reproductive
or teratogenic effects on mammals at lower doses, the report by
Dzierzawski (1977) suggests that further studies may be necessary
before a final conclusion is reached.
0 In a study in which female Wistar rats were dosed orally with lindane
at levels ranging from 6.25-25 mgAg bw from days 6 through 15 of
gestation, Khera et al. (1979) observed no statistically significant
changes in numbers of dead or resorbed fetuses, nor did they observe
any type of birth defects in the offspring.
Mutagenicity
0 The evidence of the mutagenic activity of lindane is equivocal. Only
one study indicated a weak mutagenic effect of lindane at a dose of
50 mgAg in mice (Rohrborn, 1977). Another study indicated a positive
dominant lethal mutation in male Swiss mice fed approximately 65 mgAg
bw technical grade lindane for 4 to 8 months (Lakked et al., 1982).
These cases, however, appear to be an exception as the majority of
similar studies indicate negative results (Benes and Sram, 1969;
Ahmed et al., 1977; Rohiborn, 1977; Probst et al., 1981).
Carci nogeni ci ty
0 NCI (1977) reported no significant increases in the incidence of liver
cancer in male or female B6C3F1 rats fed up to 472 ppm (74 mgAg bw)
in the diet for 80 weeks. Reuber (1979), however, reevaluated the
results and reexamined tissue sections from the same study and
concluded that the incidence of tumors was increased in the treated
animals. Since he gave no indication as to why he considered the
original NCI interpretation of the tissues questionable or how the
tissues were reexamined, it is difficult to draw conclusions from his
review.
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Wndane March 31, 1987
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0 In the study by NCI (1977), both male and female B6C3F1 nice were
exposed to lindane in the diet at either 80 or 160 ppm (10.4 or
20.8 mgAg bw). A significant increase in liver tumor incidence was
reported only for low-dose males. Because of the high spontaneous
incidence (20.8%) of hepatocellular carcinoma in B6C3F-| male mice and
because the incidence among high-dose males was not increased signifi-
cantly, NCI (1977) concluded that the occurrence of these tumors in
these mice could not be related conclusively to the administration of
lindane under the conditions of this bioassay. On the other hand,
the incidence of hepatocellular carcinomas in low-dose males, while
not showing a significant increase compared with matched controls,
did exhibit a highly statistically significant increase compared with
pooled controls. As was the case with the data resulting from the
rat study, Reuber (1979) reported a different interpretation of the
results of the same experiment.
0 Thorpe and Walker (1973) exposed 30/sex/group CF1 mice to gamma
lindane at 400 ppm in the diet (52 mgAg bw) for up to 110 weeks.
Liver tumors developed in exposed males and females (P <0.001). This
study was compromised by the low percentage of exposed mice surviving
110 weeks (3% of females and 17% of males).
0 Goto et al. (1972) reported liver tumors in 5 of 10 IRC-JCL male mice
fed gamma lindane at 600 ngfkg/day in diet.
0 Hanada et al. (1973) reported that 1 of 3 surviving female mice and
3 of 4 surviving male mice developed liver tumors after 37 to 38 weeks
or exposure to 600 mg/kg/day in diet lindane compared with none in
controls.
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) = Bg/L ( „,
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mgAg bw/day.
BW m assumed body weight of a child (10 kg) or
an adult (70 kg).
OF » uncertainty factor (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).
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Lindane March 31, 1987
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One-day Health Advisory
There are insufficient toxicological data in the scientific literature
to derive a One-day HA. The Ten-day HA of 1.2 mg/L is recommended as a
conservative estimate for a One-day exposure.
Ten-day Health Advisory
The electroneurophysiological effects of lindane on Wistar rats were
studied by Muller et al. (1981). Fifteen rats were fed daily doses of 1.3,
12.3 or 25.4 ng/kg bw in the diet for 30 days. Nerve conduction delay was
observed in the animals fed a daily dose of 25.4 mg/lcg but not 12.3 or
1.3 mgAg- A NOAEL of 12.3 mg/kg/day was identified. The Ten-day HA for
a 10 kg child is calculated as follows:
Ten-day HA - <12-3 mg/kg/day) (10 kg) =1.2 mg/L or 1200 ug/L
(100) (1 L/day)
where:
12.3 mg/kg/day - NOAEL based on absence of nerve conductance delay
in rats.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
Hale and female rats of the KFM-HAN (outbred) SPF strain were fed pure
lindane at dietary levels of 0, 0.2, 0.8, 4, 20 or 100 mg/kg/day for 84
consecutive days (RCC, 1983). Liver hypertrophy, kidney tubular degeneration,
hyaline droplets, tubular casts, tubular distension, interstitial nephritis and
basophilic tubules were observed in the 20 and 100 ppm groups. Effects were
rare and very mild when noted at 4 ppm. The NOAEL was considered to be 4 ppm
in this study. Based upon measured food consumption, the daily intake of
lindane at 4 ppj in thcv diet was 0.29 mg/kg in males and 0.33 mg/kg in females.
Using 0.33 mg/kg as the NOAEL, the Longer-term HA is calculated as follows:
For a child:
Longer-term HA = <°'33 mqAq/day) (10 kg) . 0.033 mg/L or 33 ug/L
(100) (1 L/day)
where:
0.33 mgAg/day = NOAEL based on absence of liver hypertrophy in rats.
10 kg - assumed body weight of a child.
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Lindane J-bb March 3V, 1987
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100 « uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day - assumed daily water consumption of a child.
For an adult:
Longer-term HA - (0.33 mgAg/day) (70 kg) « 0.12 «g/L or 120 ug/L
(100) (2 L/day)
where:
0.33 mgAg/day « NOAEL based on absence of liver hypertrophy in rats.
70 kg « assumed body weight of an adult.
100 « uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
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). Prom 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
vc.lue of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classifed 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 study by RCC (1983) has been selected as the basis for calculating
a Lifetime HA. Four longer-term studies were identified as potential candi-
dates for the determination of the RfD. Collectively, they describe doses
spanning the toxic threshold, thus allowing a maximum NOAEL to be defined.
They include the chronic study of Fitzhugh et al. (1950), the chronic study
in rats for 80 weeks (NCI, 1977), the chronic dog study by Rivett et al.
(1978) and the 12-week feeding study using rats by RCC (1983). The study by
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Lindane March 31, 1987
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RCC (1983) is the nost appropriate from which to derive the Lifetime HA. The
reasons for selecting the study by RCC (1983) for the Lifetime HA have been
delineated in the Drinking Water Criteria Document for Lindane (U.S. EPA,
1985a). Male and female rats were fed pure lindane at dietary levels of 0,
0.2, 0.8, 4, 20 or 100 ppm for 84 consecutive days. Various adverse effects
as noted earlier were observed in the 20 and 100 ppm groups. Effects were
rare and mild at 4 ppm. From these results a NOAEL of 0.33 mg/kg/day was
identified.
Using this NOAEL, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - 10.33^mgykg/day) - 0.0003 mgAg/day
where:
0.33 mgAg/day « NOAEL based on absence of liver hypertrophy in rats.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study
of less-than-lifetime duration.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0003 mgAg/day)(70 kg) = 10 /L
(2 L/day) y/
where:
0.0003 mgAg/day - RfD.
70 kg = assumed body weight of an adult.
2 L/day » assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = 10 mg/L x 20* = 0.002 mg/L (2 ug/L)
where:
10 mg/L = DWEL.
20* « assumed relative source contribution.
Evaluation of Carcinogenic Potential
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogen risk (U.S. EPA, 1986), lindane appears to fall somewhere
between Group B2: Probable Human Carcinogen and Group C: Possible
Human Carcinogen. The Group B category is for agents for which there
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Undane 1.88 March 31, 1987
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is inadequate evidence from human studies and sufficient evidence
from animal studies, while Group C is for agents with limited evidence
of carcinogenic! ty in animals in the absence of human data. However,
the Office of Pesticide Programs recently has decided to treat lindane
as a Group C carcinogen (U.S. EPA, 1985b).
Risk estimates were calculated by EPA's Carcinogen Assessment Group
(U.S. EPA, 1980) and the National Academy of Sciences (HAS, 1977)
based on the oncogenic effects observed in the liver of CF1 mice fed
lindane in the diet (Thorpe and Walker, 1973). The estimated levels
that would result in increased lifetime risks of 10-4, 10-5 and 10-6
are given below:
Excess Lifetime Cancer Risk (ug/L)
10-5 10-6
CAG 2.65 0.265 0.0265
NAS 5.5 0.55 0.055
0 An overall IARC (1982) classification for lindane is group 3, indi-
cating that carcinogen! city cannot be determined.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 An MCL of 0.004 mg/L or 4 ug/L for lindane in drinking water was
promulgated in 1975 as an interim primary standard by EPA (Federal
Register, 1975).
0 The World Health Organization (WHO, 1984) has recommended a drinking
water criterion of 3 ug/L for lindane.
0 It should be noted that an estimated concentration for detection by
taste and odor in water was 12.0 mg/L (Sigworth, 1965).
VII. ANALYSIS
Determination of lindane is by a liquid-liquid extraction gas chromato-
graphic procedure (U.S. EPA, 1978; Standard Methods, 1985). Specific-
ally, the procedure involves the use of 15% nethylene chloride in
hexane for sample extraction, followed by drying with anhydrous
sodium sulfate, concentration of the extract and identification by
gas chromatography. Detection and measurement is accomplished by
electron capture, micro-coulometric or electrolytic conductivity gas
chromatography. Identification may be corroborated through the use
of two unlike columns or by gas chromatography-mass spectroscopy
(GC-MS). The method sensitivity is 0.001 to 0.010 ug/L for single
component pesticides and 0.050 to 1.0 ug/L for multiple component
pesticides when analyzing a 1-liter sample with the electron capture
detector.
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Lindane 'ICO March 31, 1987
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HII. TREATMENT TECHNOLOGIES
0 Treatment technologies which are capable of removing lindane from
drinking water are adsorption or granular activated carbon (GAC),
reverse osmosis (RO) and oxidation. Granular activated carbon columns
(GAC) have been tested for their effectiveness in removing lindane.
A pilot-scale column was tested on lake water which was spiked with
50 ug/L of lindane. Three different carbons were tested and reportedly
produced lindane removal efficiencies of 99.9, 94.8 and 99.8%.
0 A treatment plant in Mount Clemens, Michigan has used GAC to remove
pesticides including lindane from the source water. The columns
proved to be 100% effective in reducing lindane from an initial
concentration of 5 ng/L (U.S. EPA, 1978).
0 One bench-scale study evaluated the performance of RO cellulose
acetate membrane in the removal of insecticides, including lindane.
Water containing different concentrations of lindane (0.683 mg/L,
50 mg/L and 500 mg/L) was fed to the Rd membranes. Removal
efficiencies of 52, 84 and 73%, respectively, were reported (U.S.
EPA, 1978). A pilot-scale plant was field tested in Miami, Florida,
for the removal of SOC, including lindane. The RO process removed
40 percent of the lindane at initial concentrations of 40 ug/L (U.S.
EPA, 1978).
0 Oxidation by ozone (03) has been tested primarily at bench-scale for
the removal of SOC from drinking water. A number of researchers
presented on the ability of ozone to remove several SOCs from water,
including lindane. Lindane initial concentration varied from 0.05 to
0.1 mg/L. Lindane was not removed appreciably (0 to 10%) at low
levels of ozone does, i.e., 0.4 to 11 mg/L. However, when the ozone
dose was increased to 149 mg/L, lindane was completely removed from the
source water. The high ozone dose might make this treatment technique
economically unfeasible for the removal of lindane.
0 Other treatment technologies, such as reverse osmosis and oxidation
have not been extensively evaluated (except on an experimental level).
An evaluation of some of the physical and/or chemical properties of
lindane indicates that these methods wovld be candidates for further
investigation.
0 Selection of individual or combinations of technologies for lindane
reduction must be based on a case-by-case technical evaluation, and
an assessment of the economics involved.
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Lindane JLSO March 31, 1987
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water supplies. (Draft). Science and Technology Branch, CSD, ODW,
Washington, DC.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 51(185):33992-34003.
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index for
Pesticides, Bureau of Foods.
Weisse, I., and M. Herbst. 1977. Carcinogenicity study of lindane in the
mouse. Toxicol. 7:233-238.
WHO. 1984. World Health Organization. 1984. Guidelines for Drinking Water
Quality. Volume I. Recommendations. WHO, Geneva, p. 6.
Wolfe, J.L., and R.J. Esher. 1980. Toxicity of carbofuran and lindane
to the old field mouse (Peramyscus polionotus) ani the cotton mouse
(P_. gossypinus). Bull. Environ. Contain. Toxicol. 24(6):894-902.
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March 31, 1987
METHOXYCHLOR
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 duratidns. 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. The chemical 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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Methoxychlor -*-»-", March 31, 1987
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This Health Advisory is based on information presented in the Office of
Drinking Water's Health Effects Criteria Document (CD) for Methoxychlor (U.S.
EPA, 1985a). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological 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 I 86-117876/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 72-43-5
Structural Formula
Synonyms
2,2-Bis(4-methoxyphenyl)-1,1,1-Trichl or oe thane
Malate* "
2,2-Di-2-anisyl-1,1,1-trichloroethane
DMDT
Methoxy-DDT
Dses
• Methoxychlor has been used as as an insecticide (mosquito larvae and
horseflies) (Windholz, 1976), In dairy barns (Hawley, 1977) and is
registered for 87 crops (HAS, 1977).
Properties (U.S. EPA, 1985a)
Chemical Formula CeH15Ci3°2
Molecular Weight 346.65
Physical State pale yellow crystalline solid
Boiling Point
Melting Point 78-78.2°
Density
Vapor Pressure —
Water Solubility 0.1 mg/L (25»C) (Richardson and
Miller, I960)
Log Octanol/Water Partition 3.05 (Coats et al., 1979)
Coefficient 3.31 (Kapoor et al., 1973)
3.68 (Kapoor et al., 1970)
4.30 (Fujita, 1979)
Taste Threshold ~
Odor Threshold --
Conversion Factor
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•"I CV-<
Methoxychlor JL*-' « March 31, 1987
-3-
Occurrence
Methoxychlor production is confidential but was estimated to be
approximately 3 million Ibs in 1982.
Methoxychlor is degraded poorly in the environment and is considered
to be persistent in soil. Soil half lives are reported to be greater
than 6 months. Due to methoxychlor's very low water solubility and
high water-to-soil partition coefficient, the chemical is immobile in
soil and migrates slowly, if at all. Methoxychlor has the potential
to bi©accumulate.
Methoxychlor has not been found in large amounts in drinking water.
Only 1 ground water sample out of 71 in the Rural Water Survey
reported a measureable level of methoxychlor (0.09 ug/L). No water
system has reported exceeding the interim MCL of 100 ug/L. Methoxy-
chlor has been found in a few non-drinking water surface and ground
waters in areas near its agricultural use. Levels up to 50 ug/L
have been reported. Methoxychlor has been found in low levels in
food. The current information is insufficient to indicate which is
the major route of exposure for methoxychlor.
Ill. PHARMACOKINETICS
Absorption
Quantitative data on the absorption of methoxychlor by experimental
animals were not located. Absorption of methoxychlor through the
gastrointestinal tract and skin can be inferred from methoxychlor's
demonstrated systemic toxicity to animals when administered by these
routes (U.S. EPA, 1985a) and from excretion data (see Excretion
section).
Distribution
Rats fed methoxychlor in the diet did not accumulate or store this
insecticide to a significant extent in their fat or other tissues.
Methoxychlor was not detected in the livers or body fat of adult rats
or. in the livers, brains or carcasses of weanlings of any of the four
generations fed these diets.
The feeding of 20 mg/kg methoxychlor in the diet to male weanling
Wistar rats for 350 days did not result in detectable levels of
methoxychlor in their fat, livers, hearts or brains (Villeneuve
et al., 1972).
With higher concentrations of methoxychlor in the diet, low levels of
methoxychlor were detected in the perineal fat. Male and female
weanling rats were fed technical grade methoxychlor in the diet for 4
to 18 weeks (Kunze et al., 1950). No methoxychlor was detected in the
fat of rats fed 25 ppm methoxychlor; however, at 100 ppm and 500 ppm,
detectable levels were found in the fat in the 4th and 9th weeks of
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Methoxychlor -»-«-'•-; March 31, 1987
-4-
fceding, respectively. Two weeks after the treated rats were
transferred to the control diet, methoxychlor could no longer be
detected in their fat.
Metabolism
0 The major metabolites in feces and urine of female Swiss mice given
3H-ring-labeled or 14C-ring-labeled methoxychlor in a 4:1 mixture of
olive oil and acetone as a single oral dose of 50 mg/kg were identified
by thin layer chromatography as the monophenol [2-(p-methoxyphenoD-
2-(p-hydroxyphenyl)-1,1,1-trichloroethane] and bisphenol [2-2-bis-
(p-hydroxyphenyl)!, 1,1-trichloroethane] resulting from 0-demethylation
of methoxychlor and as 2, 2-bis(p-hydroxyphenyl)-1,1-dichloroethylene,
the dehydrochlorination product of the bisphenol (Kapoor et al.,
1970). Other metabolites present in significant quantities were
bis(£-hydroxyphenyl)-acetic acid and £,£'-dihydroxybenzophenone.
Methoxychlor itself was apparently not dehydrochlorinated because
2, 2-bis(£-methoxyphenyl )-1,1-dichloroethylene was not detected.
0 In vitro studies performed with hepatic microsomes from a variety of
species indicate that methoxychlor is 0-demethylated by the microsomal
mixed function oxidase system. Incubations of hepatic microsomes
from rats or mice with radioactively labeled methoxychlor and an
KADPH-generating system resulted in the production of the monCphenol
and bisphenol metabolites and in the evolution of formaldehyde (Kapoor
et al., 1970; Bulger et al., 1978; Coats et al., 1979). The evolution
of formaldehyde was inhibited by hexobarbital and SKF-525A, indicating
that the MFO system was involved in the O-demethylation of methoxychlor
(Bulger et al., 1978).
Excretion
Kapoor et al. (1970) reported that, within 24 hours, female Swiss
mice excreted in urine and feces 98.3% of the orally administered
radioactivity from ^jj-ring-labeled methoxychlor administered at 50
Weikel (1957) studied the fate of C14-labeled methoxychlor injected
intravenously into adult male rats at 3 mg/kg. There was rapid
disappearance of C14 -methoxychlor from the blood and concomitant
rapid accumulation of radioactivity in the liver. Approximately 50%
of the administered radioactivity was excreted in the feces and 5 to
10\ was excreted in the urine within 4 days; the majority of excretion
occurred within the first 24 hours.
IV. HEALTH EFFECTS
Humans
Stein (1968) orally administered methoxychlor to volunteers at levels
of 0.5, 1.0 or 2.0 mg/kg/day for six weeks. No adverse effects were
reported for routine biochemical and hematologic parameters, such as
SCOT, SGPT and alkaline phosphatase.
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Methoxychlor ^ March 31, 1987
-5-
Animals
Short-term Exposure
0 Methoxychlor has a low order of toxicity: in rats, acute oral LDso's
of methoxychlor in lipophilic vehicles have been estimated at
approximately 6 g/kg bw (Smith et al., 1946; Hodge et al., 1950;
Lehman, 1951). An LD5Q value for mice was reported as 2.0+0.5 g/kg
(Coulston and Serrone, 1969). A 50% mortality incidence was not
observed at the highest doses tested for monkeys (2.5 g/kg) or
hamsters (2.0 g/kg) (Coulston and Serrone, 1969; Cabral et al.,
(1979).
0 Symptoms of toxicity include CNS depression, progressive weakness,
diarrhea and death within 36 to 48 hours (Smith et al., 1946;
Lehman, 1951).
Single, oral doses of methoxychlor have been reported to produce
changes in hepatic glycogen metabolism in rats, such as decreased
lactate and glycogen phosphorylase and increased glucose-6-phosphatase
(Morgan and Hickenbottom, 1979). These effects were observed in a
group of nude Holtzman rats after a single oral dose of methoxychlor
at 640 mg/kg in corn oil. Rats were given 0, 10, 40, 160 or 640 mg
methoxychlor/kg orally in corn oil. Animals, were sacrificed 24 hrs
after dosing.
0 Lillie et al., (1947) administered single doses of methoxychlor at
2 to 8 gAg to rats. Only one animal died. Histopathological exam-
ination of this animal revealed several isolated hepatocytes .in
various stages of coagulative necrosis and fatty degeneration in the
liver, kidney and heart muscle. Fatty degeneration of isolated
hepatic and renal cells, focal interstitial nephritis, small foci of
interstitial myocarditis and pulmonary interstitial and perivascular
infiltration were observed in the remaining animals.
0 Loss of body weight and growth retardation were the most frequent
observations in studies of the oral toxicity of methoxychlor in
laboratory animals (Hodge et al., 1950; Tuliner and Edgcomb, 1962;
Shain et al., 1977) after short-term exposures of up to 45 days.
These effects were attributed to food refusal in pair-fed ontrol
experiments (Hodge et al., 1950; Tullner and Edgcomb, 1962) rather
than to methoxychlor toxicity.
Long-term Exposure
0 The only treatment-related observation of noncarcinogenic toxic
effects in the HCI (1978) bioassay was a dose-related decrease in
body weight between the treated animals (448 and 845 ppm for male
rats, 750 and 1385 ppm for female rats). Weight differences between
treated and control animals disappeared during the post-exposure
observation period.
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200
ne-cnoxycnior March 31, 1987
-6-
0 Lehman (1965) reported the results of a feeding study in rats follow-
ing administration of 0, 10, 25, 100, 200, 500 and 2000 ngAg diet of
nethoxychlor for 2 years. Growth retardation was observed at 200 ppm
and above in animals. No histological damage attributable to the
chemical was noted in animals over a 2-year period. A no-effect
level of 100 ppm (5 ngAg) was identified in the study.
Reproductive Effects
0 No information was found in the available literature on the repro-
ductive effects of methoxychlor.
Developmental Effects
0 Khera et al., (1978) demonstrated that treatment-related effects on
the rat fetus (wavy ribs) were present only at doses of methoxychlor
that were toxic to dams. This abnormality was considered to be the
result of disturbed maturation of the fetus rather than an indication
of the teratogenic potential of methoxychlor.
Mutagenicity
° In vitro mutagenicity assays of methoxychlor using the bacteria E_.
Coli and £>. typhimurium and the yeast'jJ. cerevisiae were negative
both in the presence and absence of a metabolic activation system
from rat liver (Ashwood-Smith et al., 1972; Simmon, 1979). Other
short-term assays of genotoxicity, e.g., unscheduled DNA synthesis,
recessive lethal assay in D^. melanogaster, and induction of DNA
damage in DNA repair-deficient strains of E_. coli and !J. subtilis,
also were negative.
0 Neither the study of Grant et al. (1976) suggesting the possible
presence of the mutagenic contaminant 3,6,11,14-tetramethoxydibenzo-
(g,p)chrysene in methoxychlor nor the weakly positive transformation
response of methoxychlor in cultured BALB/3T3 cells demonstrated by
Dunkel et al. (1981) provide convincing evidence of geneotoxic
potential for the compound.
Carci nogeni ci ty
0 Methoxychlor has been tested for carci nogeni city in a number of
studies using both rats and mice (Deichmann et al., 1967; Hodge
et al., 1952; Radomnski et al., 1965; NCI, 1978). Statistically
significant increases in tumor incidences were not observed in any
of these studies. Although Reuber (1978, 1979a,b, 1980), after
reevaluation of the data, asserts that methoxychlor is carcinogenic,
the conclusion of both NCI (1978) and IARC (1979) is that methoxychlor
is not an animal carcinogen.
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Methoxychlor 201 March 31, 1987
-7-
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories
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202
Methoxychlor March 31, 1987
-8-
Ten-day Health Advisory
The study in humans by Stein (1968) which identifies a NOAEL of 2.0
»gAg/<3ay is used for driving a Ten-day HA. Using the NOAEL of 2.0
the Ten-day HA is calculated as follows:
Ten-day HA - (2.0 ag/kq/day) (10 kg) . 2.0 /L <2,000 ug/L)
(10)(1 L/day) y y/
where:
2.0 mgAg/day - NOAEL in adults.
10 kg «= assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from a human study.
1 L/day « assumed daily water consumption of a child.
Longer-term Health Advisory
Insufficient toxicological data are available to derive a Longer-term
Health. Advisory. The DWEL, adjusted for a 10 kg child, of 0.5 mg/L is
recommended as a conservative estimate for a Longer-term exposure.
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 classifed 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.
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Methoxychlor -' March 31, 1987
-9-
The NCI (1978) bioassay in which the 845 ppm (42.3 mg/kg/day) dietary
level which produced growth retardation in male rats, in addition to statisti-
cally insignificant, but dose-related histological changes in the spleens of
treated animals, is considered to be the lowest-observed-adverse-effect level
(LOAEL) in this study. The 100 ppm (5 mg/kg/day) dietary level producing no
growth retardation in rats (Lehman, 1965) represents the highest NOAEL for
lifetime methoxychlor exposure. Thus, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (5.0 mg/kg/day) = 0.05 mg/kg/day
where:
5.0 mg/kg/day = NOAEL based on the absence of growth retardation in
rats.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DVIEL)
DWEL = (0.05 mg/kg/day) (70 kg) = K7 Bg/L (1 700 ug/L)
(2 L/day)
where:
0.05 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA » 1.7 mg/L x 20% = 0.34 mg/L (340 ug/L)
where:
1.7 mg/L = DWEL.
20% = assumed relative source contribution from water.
In addition, it should be noted that these values exceed the solubility
of methoxychlor (0.10-0.12 mg/L at 25°C) in water, as reported in the published
literature.
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Methoxychlor March 31, 1987
-10-
Evaluation of Carcinogenic Potential
• Methoxychlor has been tested for carcinogenicity in a number of
studies using both rats and mice (Hodge et al., 1952; Radomnski et al.,
1965; Deichmann et al., 1967; NCI, 1978). Statistically significant
increases in tumor incidences were not observed in any of these
studies. Although Reuber (1978, 1979a,b, 1980), after re-evaluation
of the data, asserts that methoxychlor is carcinogenic, the conclusion
of both NCI (1978) and IARC (1979) is that methoxychlor is not an
animal carcinogen.
0 IARC has not evaluated methoxychlor for its carcinogenic potential.
e Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), methoxychlor is classified in
Group D: Not classified. This category is for agents with inadequate
animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 A maximum contaminant level (MCL) of 0.1 mg/L for methoxychlor in
drinking water was promulgated in 1975 as an interim standard by the
U.S. EPA (Federal Register, 1975a).
0 The same maximum contaminant level (0.1 mg/L) has been established
for bottled water by the FDA (Federal Register, 1975b).
0 NAS (1977) has suggested a SNARL of 0.700 mg/L for methoxychlor in
drinking water, assuming that 20% of the total daily intake comes
from this source or 3.5 mg/L, assuming that 100% of the total
daily intake comes from this source.
0 WHO has recommended a drinking water criterion of 30 ug/L for
methoxychlor (WHO, 1984).
VII. ANALYTICAL METHODS
0 Determination of methoxychlor is by a liqud-3-liquid extraction gas
chromatographic procedure (U.S, EPA, 1978; Stancard Methods, 1985).
Specifically, the procedure involves the use of 15% methylene chloride
in hexane for sample extraction, followed by drying with anhydrous
sodium sulfate, concentration of the extract and identification by
gas chromatography. Detection and measurement is accomplished by
electron capture, microcpulometric or electrolytic conductivity gas
chromatography. Identification may be corroborated through the use
of two unlike columns or by gas chromatography-mass spectroscopy
(GC-MS). The method sensitivity is 0.001 to 0.010 ug/L for single
component pesticides and 0.050 to 1.0 ug/L for multiple component
pesticides when analyzing a 1-liter sample with the electron capture
detector.
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Methoxychlor '205 March 31, 1987
-11-
VIII. TREATMENT TECHNOLOGIES
• Conventional treatment, granular activated carbon adsorption and
reverse osmosis have been examined as treatment techniques for
the removal of nethoxychlor from potable water.
0 The conventional treatment methods examined include coagulation/
filtration and water softening (Steiner and Singley, 1979). Jar
testing procedures were used to evaluate methoxychlor removal from
water containing 1, 5 or 10 mg/L methoxychlor. Coagulation was
carried out using alum or ferric sulfate at pH 6 or pH 4.5, re-
spectively. After mixing and settling, samples were filtered. The
reduction in methoxychlor concentration ranged from 74% to 97%.
Greater percentage reductions were obtained for the higher initial
concentrations. However, the 5 mg/L and 10 mg/L test solutions were
reportedly cloudy, indicating that the solubility of methoxychlor may
have been exceeded. Thus, at the above levels, some reduction could
be due to phase separation (U.S. EPA, 1985b).
c Additional jar testina evaluated softening as a method for methoxychlor
treatment (Steiner and Singley, 1979). Water samples were spiked with
nwthoxychlor at 1, 5 or 10 mg/L. Prior to spiking, hardness was
adjusted by the addition of calcium or calcium and magnesium. These
samples then were softened by a cold lime-soda process at pH 9.5 and
10.5 (Ca-hardened) or pH 11.0 and 11.3 (Ca-Mg hardened). Percentage
removal achieved by softening ranged from 48 to 97%. In general,
the percent removal increased with increasing initial methoxychlor
concentration. Higher removals also were obtained at higher pH
values; it was postulated that this reflected adsorption onto
precipitated Mg(OH)2.
0 In a laboratory study (Steiner and Singley, 1979), water containing
1, 5 or 10 mg/L methoxychlor was passed through a granular activated
carbon (GAC, Calgon's Filtrasorb® 400) column (19mm diameter by 265mm
long). A 250 ml sample was passed through the column with a volu-
metric loading of 0.5 gpm/ft^. No methoxychlor was detected in the
column effluent.
0 In a pilot study, groundwater spiked with methoxychlor and two other
pesticides was passed through a system that included a reverse
osmosis unit, prefilter and two GAC beds (Regunathan et al., i983).
The influent concentration was 1000 ug/L methoxychlor. Greater than
90% removal was achieved with the reverse osmosis unit. The overall
removal was 99-100%.
0 Treatment technologies for the removal of methoxychlor from water
are available and have been reported to be effective. Selection o'f
individual or combinations of technologies to achieve methoxychlor
reduction must be based on a case-by-case technical evaluation, and
an assessment of the economics involved.
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J206
Methoxychlor ' March 31, 1987
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IX. REFERENCES
Adams, M., F.B. Coon and C.E. Poling. 1974. Insecticides in the tissues of
four generations of rats fed different dietary fats containing a mixture
of chlorinated hydrocarbon insecticides. J. Agric. Food Chero.
22(1>:69-75.
Ashwood-Smith, M.J., J. Trevino and R. Ring. 1972. Mutagenicity of
dichlorovos. Nature (London). 240:418-420.
Bulger, W.H., R.M. Miccitelli and D. Kupfer. 1978. Studies on the in vivo
and in vitro estrogenic activities of methoxychlor and its metabolites:
Role of hepatic monooxygenase in methoxychlor activation. Biochem.
Pharmacol. 27(20):2417-2424.
Cabral, J.R.P., F. Raitano, T.O. Mollner, S. Bronczyk and P. Shubik. 1979.
Acute toxicity of pesticides in hamsters. Toxicol. Appl. Pharmacol.
48(1):A192.
Coats, J.R., R.L.. Metcalf, I.P. Kapoor, L. Chio and P.A. Boyle. 1979.
Physical-chemical and biological degradation studies on DDT analogues
with altered aliphatic moieties. J. Agric. Food Chem. 27(5):1016-1022.
Coulston, F. and D.M. Serrone. 1969. The comparative approach to the role
of nonhuman primates in evaluation of drug toxicity in man: A review.
Ann. NY Acad. Sci. 162:681-704.
Deichmann, W.B., M. Keplinger, F. Sala and E. Glass. 1967. Synergism among
oral carcinogens IV. The simultaneous feedings of four tumorigens to
rats. Toxicol. Appl. Pharmacol. 11:88-103.
Dunkel, V.C., R.J. Pienta, A. Sivak and K.A. Traul. 1981. Comparative
neoplastic transformation responses of BALB/3T3 cells, Syrian hamster
embryo cells and Rauscher murine leukemia virus-infected Fischer 344 rat
embryo cells to chemical carcinogens. J. Natl. Cancer Inst. 67:1303-1315.
Federal Register. 1975. National Interim Primary Drinking Water Regulations.
U.S. EPA. 40(248):59566-59588.
Fujita, T. 1979. Kyoto Univ. Unpublished results. In: Substitu<-it
Cjnstants for Correlation Analysis in Chemistry and Biology, C. Hant.cn
and A.J. Leo, Ed. Wiley Interscience Publ. John Wiley and Sons, Inc.
NY. p. 289.
Grant, E.L., R.H. Mitchell, P.R. West, L. Mazuch and M.J. Ashwood-Smith.
1976. Mutagenicity and putative carcinogenicity tests of several
polycyclic aromatic compounds associated with impurities of the
insecticide methoxychlor. Mutat. Res. 40(3):225-228.
Hawley, G.C., Ed. 1977. The Condensed Chemical Dictionary, 9th ed. Van
Nostrand Reinhold Co., NY. p. 556.
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Methoxyclor ^rcti 31 , 1987
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-13-
Hodge, H.C., E.A. Maynard, J.F. Thomas, H.J. Blanchet, Jr., W.G. Wilt, Jr.
and K.E. Mason. 1950. Short-term oral toxicity tests of methoxychlor
(2,2-di<£-methoxyphenyl)-1,1,1-trichloroethane) in rats and dogs.
J. Pharmacol. Exph. The rap. 99:140-148.
Hodge, B.C., E.A. Maynard and H.J. Blanchet, Jr. 1952. Chronic oral toxicity
tests of nethoxychlor [2.2-di-(£-methoxyphenyl)-1, 1 , 1-trichloroethane]
in rats and dogs. J. Pharmacol. Exp. Ther. 104:60-66.
IARC. 1979. (International Agency for Research on Cancer). IARC monographs
on the evaluation of the carcinogenic risk of chemicals to humans. Some
halogenated hydrocarbons. WHO, IARC, Lyon, France. Volume 20.
Kapoor, I. P., R.L. Metcalf, A.S. Hirwe, J.R. Coats and M.S. Khalsa. 1973.
Structure activity correlations of biodegradability of DDT analogs. J.
Agric. Food Chem. 21 <2): 310-315.
Kapoor, I. P., R.L. Metcalf, R.F. Kystrom and G.K. Sangha. 1970. Comparative
metabolism of methoxychlor, methiochlor and DDT in mouse, insects and in
a model ecosystem. J. Agric. Food Chem. 18:1145-1152.
Khera, K.S., C. Whalen and G. Trivett. 1978. Teratogenici ty studies on
linuron, malathion and methoxychlor in rats. Toxicol. Appl . Pharmacol.
45<2):435.
Kunze, P.M., E.P. Laug and C.S. Prickett. 1950. The storage of methoxychlor
in the fat of the rat. Proc. Soc. Exp. Biol. Med. 75:415-416.
Lehman, A.J. 1951. Chemicals in foods: A report to the Association of Food
and Drug Officials on current developments. Part II. Pesticides.
Assoc. Food Drug Off. 15:123-133.
Lehman, A.J. 1965. Summaries of Pesticide Toxicity. (FDA - Unpublished
study) .
Lillie, R.D., M.I. Smith and E.F. Stohlman. 1947. Pathologic action of DDT
and certain of its analogs and derivatives. Arch. Path. 43:127-142.
(CA 41:5967e)
Morgan, J.M. and J.P. Hickenbottom. 1979. Relative sensitivities of various
biochemical, toxicological and pathological techniques in dei.onstratj.ng
sublethal lesions in the rat following oral administration of low levels
of methoxychlor. Toxicol. Appl. Pharmacol. 45(1):237.
HAS. 1977. National Academy of Sciences. Drinking water and health.
Volume 1. Washington, DC.
NCI. 1978. National Cancer Institute. Bioassay of methoxychlor for possible
carcinogeni city. NCI-CG-TR-35. Carcinogenesis Program, p. 91.
Radmonski, J.L., W.B. Deichmann, W.E. MacDonald and E.M. Glass. 1965.
Synergism among oral carcinogens. I. Results of the simultaneous feeding
of four tumorigens to rats. Toxicol. Appl. Pharmacol. 7(5) : 652-656.
-------
2GS
Methoxychlor March 31, 1987
-14-
Regunathan, P., W.H. Beaunan, and K.G. Kreusch, 1983. Efficiency of point
of use treatment. JAWWA. 42-49.
Reuber, M.D. 1978. Carincomas and other lesions of the liver in mice
ingesting organochlorine pesticides. Clin. Toxicol. 13(2):231-256.
Reuber, M.D. 1979a. Interstitial cell carcinomas of the testis of BALB/c male
•ice ingesting methoxychlor. J. Cancer Res. Clin. Oncol. 92(2):173-179.
Reuber, M.D. 1979b. Carcinomas of the liver in Osborne-Mendel rats ingesting
methoxychlor. Life Sci. 24(15):1367-1371.
Reuber, M.D. 1980. Carcinogenicity and toxicity of nethoxychlor. Environ.
Health Perspect. 36:205-219.
Richardson, L.T., and D.M. Miller. 1960. Fungitoxicity of chlorinated
hydrocarbon insecticides in relation to water solubility and vapor
pressure. Can. J. Bot. 38:163-175.
Shain, S.A., J.C. Schaeffer and R.W. Boesel. 1977. The effect of chronic
ingestion of selected pesticides upon rat ventral prostate homeostasis.
Toxicol. Appl. Pharmacol. 40(1):115-130.
Simmons, V.F. 1979. In vitro microbiological mutagenicity and unscheduled
DHA synthesis studies of 18 pesticides. EPA 600/1-79-041.
Smith, M.I., H. Bauer, E.F. Stohlman and R.D. Lillie. 1946. The pharmacologic
action of certain analogues and derivatives of DDT. J. Pharmacol. Exptl.
Therap. 88:359-365.
Standard Methods. 1985. Method 509A. Organochlorine Pesticides. In:
Standard Methods for the Examination of Water and Wastewater. 16th
Edition, APHA, AWWA, WPCF.
Stein, A.A. 1968. Comparative methoxychlor toxicity in dogs, swine, rats,
monkey and man. Ind. Med. Surg. 37:540-541.
Steiner, J., and J.E. Singley, 1979. Methoxychlor removal from potable water.
JAWWA. 284-286.
Tullner, W.W., and J.H. Edgcomb. 1962. Cystic tubular nephropathy and
decrease in testicular weight in rats following oral methoxychlor
treatment. J. Pharmacol. Exph. Therap. 138:126-130.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for organ-
ochlorine pesticides in drinking water. In: Methods for Organochlorine
Pesticides and Chlorphenoxy Acid Herbicides in Drinking Water and Raw
Source Water. Interim.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of
pesticides in drinking water, food, and air. Office of Drinking Water.
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Methoxychlor 209
March 31, 1987
-15-
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Draft health
document for methoxychlor. Office of Drinking Water.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Technologies and
costs for the removal of synthetic organic chemicals from potable water
(draft). Science and Technology Branch, C&SD, ODW.
Washington, D.C.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 51(185):33992-34003*
September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index for
Pesticides. Bureau of Foods.
Villeneuve, D.C., D.L. Grant and W.E.J. Phillips. 1972. Modification of
pentobarbital sleeping times in rats following chronic PCB ingestion.
Bull. Environ. Contain. Toxicol. 7(5): 264-269.
Weikel, J.H., Jr. 1957. The metabolism of methoxychlor (1,1,1-trichloro-
2,2-bis(p-methoxyphenyl)ethane. I. The role of the liver and biliary
excretion in the rat. Arch. Intern. Pharmacodyn. 110:423-432.
WHO. 1984. World Health Organization. Guidelines for Drinking Water Quality.
Volume I. Recommendations. WHO, Geneva, p. 6.
Windholz, M., ed. 1976. The Merck Index, 9th ed. Merck and Co., Inc.,
Rahway, HJ. pp. 5865-5866.
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March 31, 1987
210
OXAMYL
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. The chemical 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 ri ,k 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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Oxamyl <«v-»-l March 31, 1987
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No criteria document is available for Oxamyl at this time.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 23135-22-0
Structural Formula
CH3 0 0
V M II
N-C-C=N-0-C-NH- CH3
CH3 SCH3
Synonyms
0 Vydate, DPX-1410, Methyl-JJ1, Nl-dimethyl-N-[(methyl carbamoyl)oxy]-
1 -thiooxamimidate.
Uses
0 Pesticide (insecticide, nematocide).
Properties (Reinhardt, 1971; Windholz, 1983)
Chemical Formula C7H1 3N3°3S
Molecular Weight 219.3
Physical State Off-white, crystalline powder
Boiling Point —
Melting Point 100-102°C crystalline form changes;
108-110°C melts
Specific" Gravity 0.97
Density —
Vapor Pressure 2.3x10-4mm Hg at 25°C
Water Solubility 280g/L water (25°C)
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Oxamyl is an insecticide and nematocide used on a variety of fruit
and vegetable crops, including potatoes, peanuts, soybeans and cotton.
EPA estimated that oxamyl production in 1980 ranged from 500,000 to
750,000 Ibs. Oxamyl is applied both to the soil and directly to plants.
0 Oxamyl is considered to be non-persistent as a pesticide.
0 Oxamyl is metabolized rapidly by plants after application; once in the
soil, it is degraded by both aerobic and anaerobic bacteria. Oxamyl
is hydrolyzed rapidly in neutral and alkaline soils and more slowly
in acid soils. Oxamyl has a soil half life of 1 to 5 weeks, with
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Oxamyl ^ March 31, 1987
/Oj_<0
-3-
residual levels found up to 6 to 12 months later. Oxamyl in river
water was reported to degrade more rapidly, with a half life of 1 to
2 days. Oxamyl does not bind to soil or sediments and has been shown
to migrate in soil. Oxamyl does not bioaccumulate to any great extent.
0 Oxamyl has been reported to .occur in ground water at levels in the
low ppb range in California, New York and Rhode Island. The range in
areas of agricultural use was 5;-65 ppb. The 85th centile for non-
zero samples in the STORET data base from 541 ground-water stations
was 10 ug/liter. Oxamyl levels have not been analyzed in past
Agency surveys of drinking water; estimates of national exposures are
not available. Because of oxamyl's relatively rapid degradation
rate, it is expected to occur more frequently in ground waters than
surface waters. No information on oxamyl in food or air has been
identified (U.S. EPA, 1983; U.S. FDA, 1984; STORET database).
II. PHARMACOKINETICS
Absorption
0 An estimated 48 to 61% of a dose of 1.0 mg 14C-oxamyl administered
to rats in 2 mL peanut oil by intragastric intubation was absorbed
in 72 hours based upon recovery in the urine (Harvey and Han, 1978).
Distribution
0 Seventy-two hours after intragastric intubation of 1 mg 14C-oxamyl in
2 mL peanut oil to rats, low levels of radioactivity were detected
throughout the body, but mainly in the hide (skin/hair) (7 to 12%),
carcass (4 to 60%), Gl-tract and blood. About 50% of the radioactivity
found in the hide, blood and carcass was incorporated into proteins
(Harvey and Han, 1978).
Metabolism
0 The metabolic end products of oxamyl are methyl-N-hydroxy-N',N'-
dimethyl-1-thioxoaminidate (DMTO), methyl-N-hydroxy-N'-methyl-1-
thioxoaminidate (MTO), N,N-dimethyl oxamic acid (DMOA) and N-methyl-
oxamic acid (MOA) (Harvey and Han, 1978).
Excretion
Intragastric intubation of 1 mg 14C-oxamyl in 2 mL peanut oil to
rats by Harvey and Han (1978) resulted in excretion of most of the
radioactivity in 72 hours in urine and feces (68 to 72%). No radio-
activity was found in the expired air (<0.3 %). Total recovery was
about 91%. Most of the radioactivity excreted was found in the urine
(48 to 61%) with smaller amounts in feces (6 to 23%).
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Oxamyl «;> 1 o March 31, 1987
-4-
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of oxamyl in humans.
Animals
Short-term Exposure
0 The oral LD50 for Oxamyl (tested as 90% active ingredient) in fasted
rats was 4.0 mg/kg for males and 2.8 mg/kg for females. In non-fasted
males, the 1*050 is 5.4 mg/kg (Reinhardt, 1971). Clinical signs were
heavy breathing, fasciculations, salivation and lacrimation.
0 Male rats administered oxamyl at 2.4 mgAg (90%+ technical) by gavage,
five times per week for two weeks, exhibited typical anticholinesterase
symptoms such as fasciculations and salivation (Reinhardt, 1971).
No deaths occurred. No apparent cumulative toxicity was seen.
0 Oxamyl (4.86 mgAg) administered by intragastric intubation as an
aqueous solution to male rats resulted in a 40% decrease in cholin-
esterase activity of whole blood after five minutes, with a maximum
effect after four hours (58%). After 24 hours, the activity was
normal (Schmoyer, et al. 1970).
Long-term Exposure
0 In a study on beagle dogs (4/sex/dose) fed a diet containing oxamyl
(95% technical) at 0, 50, 100 or 150 ppm (0, 1.25, 2.5 or 3.75 mgAg
bw/day) for two years, animals fed 150 ppm had higher levels of
alkaline phosphatase activity in whole blood; male animals of this
group had higher cholesterol values (Sherman et al., 1972). Hemoglobin
content, hematocrit and erythrocyte counts in the blood of animals
fed the highest dose were somewhat lower than those of the controls.
Whole blood cholinesterase activity measured at various intervals was
not significantly different from that of the control group. There
were no differences in the weight gain, urinalysis or organ weights
of animals in all experimental groups compared to the control animals.
The 1OAEL identified from this study was 2.5 mgAg/day.
0 Rats fed a diet containing oxamyl (95% technical) at 0, 50, 100 or
150 ppm (0, 2.5, 5.0 or 7.5 mgAg bw/day) for two years resulted in
lower body weight curves for animals fed 100 and 150 ppra oxamyl
compared to control animals throughout the experiment (p <0.05)
(Sherman et al., 1972). At 50 ppm there was a slight drop in body
weight which was not statistically significant. Average cholinesterase
activity of female rats receiving 150 ppm oxamyl was 19.3% lower than
that of the controls (p <0.05) after four days of feeding and 33.3%
lower than that of the males (p <0.05) after eight days of feeding
but at no other time. In the animals fed 150 ppm oxamyl, relative
weights of the heart, testes and adrenals were decreased in males; in
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Oxamyl 21.4 March 31, 1987
-5-
females, the relative weights of the brain, heart, lungs and adrenals
were increased. In females, most of these organs showed similar
effects at 100 ppm also. Histopathological changes were not observed
in animals fed the highest dose of oxamyl (-150 ppm). The NOAEL
identified in this study was 2.5 mgAg/day.
0 Mice fed diets containing oxamyl (97.1% active ingredient) at dose
levels of 0, 25, 50 or 75 ppm (0. 3.75, 7.5 or 11.25 mgAg/day) for
two years showed a statistically significant decrease in weight gain
at 50 or 75 ppm (p <0.05). No body weight changes were seen in mice
fed 25 ppm oxamyl (Kennedy, 1986). No significant hematological or
pathological changes were seen at any dose level tested. The NOAEL
from this study was 3.75 mgAg/day.
Neurotoxicity
0 Adult hens receiving single oral doses of oxamyl at 20 or 40 mg/kg bw
followed by intramuscular injections of 0;5 mg/kg atropine were
observed for 28 days (Lee and Zapp, 1970). The animals showed marked
symptoms of cholinesterase inhibition, but recovery was complete
after 12 hours. Immediately after administration of oxamyl, the
animals showed sudden depression, lethargy, ruffled feathers, slight
respiratory difficulty, ataxia and incoordination. Respiratory signs
disappeared within 30 minutes, but depression and nervous signs
continued for 12 hours. Animals recovered completely by twelve hours
after dosing. No signs of delayed neurotoxicity were observed.
Reproductive Effects
0 In a three-generation, six litter (two litters per generation) repro-
duction study in rats fed oxamyl (95% technical) at 0, 50, 100 or 150
ppm (0, 2.5, 5.0 or 7.5 mg/kg bw/day) for 90 days, the litter size,
viability and lactation indices and weanling body weights were lower
at the two higher doses (100 and 150 ppm) throughout the study (Sherman
and Zapp, 1971). No effects on the fertility or gestation indices
were seen at any dose level. Relative kidney weights of the pups of
the F3B generation were increased slightly at 150 ppm; relative
weights of the testes were increased at 100 and 150 ppm. There were no
histopathological changes observed. The NOAEL determined from these
data was 2.5 mg/kg/day.
Developmental Effects
0 In a study conducted to evaluate the embryotoxic and teratogenic
potential of oxamyl, pregnant rats were fed oxamyl at concentrations
of 0, 50, 100, 150 or 300 ppm (0, 2.5, 5.0, 7.5 or 15 mgAg bw/day)
on days 6 through 15 of gestation (Haskell Laboratory, 1571). There
was a dose-related decrease in the maternal body weight and food
consumption rates in animals fed 100, 150 or 300 ppm. There were no
effects on the number of implantation sites, resorptions and live
fetuses, or on embryonal development, fetal anomalies or gross changes
in tissues and organs. The NOAEL identified from these data was
2.5
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Oxamyl *Oj_. » March 31, 1987
-6-
0 In New Zealand white rabbits administered oxamyl in 1 mL of distilled
water on days 6 through 19 of gestation at dose levels of 0, 1, 2 or
4 tag/kg/day, significantly lower mean body weights were observed in
animals fed the two higher doses (2 and 4 mg/kg/day) (Snyder, 1980).
Slightly lower mean ovarian and uterine weights with and without
fetuses also were noted in these groups. Gross pathological observa-
tions did not reveal any treatment-related changes. Pregnancy rate,
number of corpora lutea and implantation efficiences were comparable
between controls and all treatment groups. There were slightly higher
incidences of resorptions in the mid- and high-dose groups compared
to the control group; fetal viability was slightly lower in the high-
dose group. Fetal mean body weight and length were comparable to the
control groups. From these data, a maternal NOAEL of 1 mg/Tng/day was
identified.
Mutagenicity
0 Oxamyl (94% active ingredient) showed no mutagenic activity in a
rec-assay using two strains of Bacillus subtilis and in reverse mutation
tests using five strains of Salmonella typhimurium and Escherichia
coli W?2 her, with or without a liver activation system. A host-
mediated assay in mice using Salmonella typhimurium G-46 also was
negative (Shirasu et al., 1976).
Carcinogenicity
0 Two-year feeding studies with oxamyl at dose levels of 0, 50, 100 or
150 ppm (0, 2.5, 5.0 or 7.5 mg/kg bw/day in rats (Sherman et al.,
1972) and in mice at dose levels of 0, 25, 50 or 75 ppm (0, 3.75, 7.5
or 14.25 mg/kg/day (Kennedy, 1986) did not result in a significantly
increased incidence of neoplastic lesions.
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 ( ug/L)
(UF) x ( L/day)
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 factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
-------
.
Oxaroyl March 31, 1987
-7-
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No suitable studies are available to calculate a One-day HA. It is
recommended that the Lifetime HA, 175 ug/L, be used, which would be
protective for this duration of exposure.
Ten-day Health Advisory
No studies of design or duration strictly appropriate for calculation
of a Ten-day HA were located. Therefore, the Lifetime HA (175 ug/L) may be
applied; it should be protective for exposures of shorter-than-lifetime
duration.
A teratogenicity study in rabbits (Snyder, 1980} was reviewed. Dosing
at 1, 2, and 4 mg/kg/day was carried out from day 6 through day 19 of
gestation. There was a decrease in maternal body weight but no significant
teratogenic effects at 2 and.4 mg/kg/day; 1 mg/kg/day was identified as
a no-adverse-effect level (NOAEL) in this study. Calculations based on
this study, if done, would yield results generally similar to the
Lifetime HA that may be used here in lieu of a Ten-day HA.
Longer-term Health Advisory
There are no appropriate studies available for the derivation of a
Longer-term HA. The Lifetime HA (175 ug/L) may be used for the Longer-term HA.
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
estimate of a daily exposure to the human populatior. 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.
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Oxamyl A--L t March 31, 1987
-8-
The Lifetime Health Advisory may be calculated from the two year chronic
feeding study in rats fed oxamyl at 0, 50, 100 and 150 ppm levels (0, 2.5, 5.0
or 7.5 rag/kg bw/day) (Sherman et al., 1972). In this study, 100 and 150 oxamyl
in the diet led to significantly lower body weight curves compared to controls
(p <0.05); 50 ppm (2.5 mg/kg bw/day) did not show any effects. The KOAEL of
2.5 mg/kg/day in this study is supported by their two year dog study which
also gave a NOAEL of 2.5 mg/kg/day.
Using 2.5 mg/kg/day as the NOAEL, a Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (2.5 mg/kg/day) __ 0.025 mg/kg/day
(100)
where:
2.5 mg/kg/day = NOAEL, based on absence of depression of weight gain
or other sign of toxicity.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.025 mg/kg/day) (70 kg) = 0.875 Bg/L (875 ug/L)
(2 L/day)
where:
0.025 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health / dvisory
Lifetime HA = (0.875 mg/L) (20%) = 0.175 mg/'L (175 ug/L)
where:
0.875 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic potential has been seen following long-
term dietary exposure in rats and mice.
-------
21.8
Oxamyl "^ March 31, 1987
-9-
Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986), oxamyl may be classified
in Group E: Evidence of non-carcinogenicity for humans. This group
in for agents that show no evidence of carcinogenic!ty in at least two
adequate animal tests in different species or in both epidemiologic
and animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Acceptable daily intake of 0.03 mg/kg bw/day has been calculated by
WHO (1985) using the 2-year dog feeding study.
0 US EPA Office of Pesticide Programs calculated an ADI for oxamyl of
0.025 mg/kg/day, based on a 2-year rat tolerance (40 CFR 180.303).
VII. ANALYTICAL METHODS
0 Oxamyl is analyzed by a high performance liquid chromatographic
procedure used for the determination of N-methyl carbamoyloximes and
N-methylcarbamates in drinking water (U.S. EPA, 1984). In this
method, the water sample is filtered and a 400 uL aliquot is injected
into a reverse phase HPLC column. Separation of compounds is achieved
using gradient elution chromatography. After elution from the HPLC
column, the compounds are hydrolyzed with sodium hydroxide. The
methyl amine formed during hydrolysis is reacted with o-phthalaldehyde
(OPA) to form a fluorescent derivative which is detected using a
fluorescence detector. The detection limit for this method has been
estimated to be approximately 1.6 ug/L for oxamyl.
VIII. TREATMENT TECHNOLOGIES
0 No data are available on the removal of oxarcyl from drinking water
(ESE, 1984).
0 Using solubility and vapor pressure data, the Henry's Law Constant for
oxamyl has been estimated to be 2.37 x 10~7 atm x m3/mole (ESE, 1984).
This value suggests that aeration is >.ot likely to be a suitable
water treatment technique for removal of oxamyl.
0 Adsorption of oxamyl by montmorillonite clay has been demonstrated
(Bansal, 1983); adsorption mechanisms were thought to include covalent
bonding, coordination, protonation, hydrogen bonding and van der Waals
forces. The demonstrated adsorption of oxamyl by clay suggests that
adsorption may be a suitable technique for the removal of oxamyl from
water (ESE, 1984). However, further studies are needed to confirm
the effectiveness of adsorption techniques and to define the optimal
conditions for use.
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Oxamyl i^jL.- March 31, 1987
-10-
Selection of individual or combinations of technologies to attempt
oxamyl reduction must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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Oxamyl March 31, 1987
-11-
IX. REFERENCES
Bansal, O.P. 1983. Adsorption of oxamyl and dimecron in montmorillonite
suspensions. Soil Sci. Soc. Am. J. 47:877-882.
ESE. 1984. Environmental Science and Engineering, Inc. Review of treatability
data for removal of twenty-five synthetic organic chemicals from drinking
water. ' U.S. EPA. Office of Drinking Water.
Federal Register. 1986. U.S. Environmental Protection Agency. Guidelines
for carcinogen risk assessment. 51(185):33992-34003. September 24, 1986.
Harvey, J., Jr., and C.Y. Han. 1978. Metabolism of oxamyl and selected
metabolites in the rat. Agric. Fd. Chem. 26:902-910.
Haskell Laboratory. 1971. Teratogenic study in rats with S-methyl-1-dimethyl
carbamoyl-N-[(methylcarbamoyl)oxy] thioformimidate (IND-1410). Report
No. 5-71, MR No. 1358. EPA Accession No. 66909.
Kennedy, G.L., Jr. 1986. Chronic toxicity, reproductive and teratogenic
studies with oxamyl. 7:106-118.
Lee, K.P., and Zapp, J.A., Jr. 1970. Oral ALD and delayed paralysis test
(white Leghorn chickens). Haskell Laboratory for Toxicology and Industrial
Medicine. Report No. 234-70, MR No. 581. EPA Accession No. 66893.
Reinhardt, C.F. 1971. Toxicological information on DPX-1410. Haskell
Laboratory for Toxicology and Industrial Medicine. EPA Accession No.
113391.
Schmoyer, L.A., N.W. Henry and J.A. Zapp, Jr. 1970. IND-1410 and cholin-
esterase activity. Haskell Laboratory. Report No. 18-70, MR No. 581.
EPA Accession No. 66907.
Sherman, H., and J.A. Zapp, Jr. 1971. Three-generation reproductive study in
rats with 1-(dimethylcarbamoyl)-N-(methylcarbamoyloxy) thioformidic acid,
methyl ester (IND-1410). Haskell Laboratory. Report No. 315-71.
MR No. 1203. Accession No. 66912.
Sherman, H., J.R. Barnes, and J.G. Aftosmis. 1972. Long-term feeding study
in rats and dogs with 'l-(dimethylcarbamoyl)-N-(methylcarbamoyloxy)
thioformidic acid, methyl ester (IND-1410). Final report. MRP No.
MR-1203. Haskell Laboratory Report No. 37-72. EPA Accession No. 83352.
Shirasu, Y., M. Moritani and K. Watanabe. 1976. Oxamyl mutagenicity study
using bacteria. Institute of Environmental Toxicology, Toxicity Dept.
EPA Accession No. 40594.
Snyder, F.G. 1980. Teratology study in rabbits — Oxamyl. Final report.
Hazelton Laboratory. MR No. 3724-001, HLO-0801-80. EPA Accession No.
63009.
STORET database. U.S. Environmental Protection Agency, Washington, DC.
-------
221
Oxamyl March 31, 1987
-12-
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of pesti-
cides in drinking water, food and air. Office of Drinking Water.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Method 531. Measure-
ment of N-methyl carbamoyloximes and N-methylcarbamates in drinking
water by direct aqueous injection HPLC with post column derivatization.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Final guidelines for
carcinogen risk assessment. Federal Register. 51(185):33992-34003.
September 24, 1986.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index for
Pesticides. Bureau of Foods .
WHO. 1985. World Health Organization. Oxamyl. Joint Meeting on Pesticide
Residues, 34-35.
Windholz, M., ed. 1983. The Merck index, 10th edition. Rahway, NJ:
Merck & Co., Inc., page 992.
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March 31, 1987
PENTACHLOROPHENOL
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. The chemical 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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Pentachlorophenol *->*- 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 pentachlorophenol
(U.S. EPA, 1985a). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological 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.f Springfield, VA 22161, PB # 86- 118015/AS. The toll-free number is
(800) 336-4700; in the Washington, D.C. area: (703) 487-4650.
GENERAL INFORMATION AND PROPERTIES
CAS No. 87-86-5
Structural Formula
Synonyms
0 PCP; Pentachlorohydroxybenzene.
Uses
0 Wood preservative, herbicide, antimicrobial agent, disinfectant,
mossicide and defoliant.
Properties (U.S. EPA, 1985a)
Molecular Formula CsClsOH
Physical State White to light yellow beads,
powder, or crystals
Molecular Weight 266.34
Boiling Point 309-310°C
Melting Point 191°C (anhydrous)
Density
"Vapor Pressure 0.00011 mmHg at 20°C
Water Solubility 14 mg/L water at 20°C
Log Octanol/Water Partition 5.86
Coefficient
Specific Gravity 1.978 at 22°C
Odor Threshold (water) i,600 ug/L
Taste Threshold (water 30 ug/L
Conversion Factor
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Pentachlorophenol ££4 March 31, 1987
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Occurrence
0 Pentachlorophenol (PCP) production was 35 million Ibs in 1985.
0 PCP is very persistent in some soils with half lives of up to 5 years
reported. PCP has been shown to photodecompose and under certain
conditions to be degraded by soil bacteria. While PCP is thought
to bind tightly to soil, migration has been shown to occur in neutral to
alkaline soils. PCP has been shown to photodegrade in a few days
in surface waters.
0 PCP has been identified at low levels in ground and surface waters.
The occurrence of PCP in water is reported to be in the low ppb range.
In one Federal survey of surface drinking water supplies, PCP was
reported to occur in the low ppb range in 2 out of 105 systems tested.
PCP has been reported to occur at low levels in food. No information
on PCP levels in air were identified. There is insufficient information
to evaluate the relative levels of exposure of PCP in water, food
and air (U.S. EPA, 1983).
0 Pentachlorophenol is the preservative in plywood treated with Cellon,
and this plywood has been used to cover distribution reservoirs.
III. PHARMACOKINETICS
Absorption
0 The available data indicate that the biological handling of PCP is
similar across mammalian species. Pentachlorophenol is absorbed
readily following oral, dermal or inhalation exposure (U.S. EPA,
1985a).
0 Meennan et al. (1983) examined the uptake of PCP and sodium penta-
chlorophenol by male Wistar rats (100-120 g) after ad libitum exposures
for 1 week in the diet (350 ppm) or drinking water (1.4 mM or 320 mg/L,
sodium Pentachlorophenol only). The investigators noted wide diurnal
variations in plasma levels of PCP associated with changes in feeding
activity, with the highest plasma concentrations occurring during the
night. Based on an analysis of plasma levels during a 24-hour period
as we J 1 as toxicokinetic parameters obtained from a separate study
using intravenous (i.v.) administration,- the investigators calculated
that virtually all of the administered PCP was absorbed from drinking
water.
0 Braun and Sauerhoff (1976) and Braun et al. (1977) compared the
toxicokinetics of PCP in Rhesus monkeys and Sprague-Dawley rats.
Groups of six rats, three of each sex, received a single gavage dose
of 14C-PCP at 10 or 100 mg/kg bw in 1 ml corn oil. Three male and
three female monkeys, Macaca mulatta, weighing 3.3 to 4.9 kg received
a single dose of 10 mg/kg bw by gavage in 10 ml corn oil. In both
species, PCP was absorbed rapidly with peak plasma levels occurring
in 12 to 24 hours in the monkeys and 4 to 6 hours in the rats.
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Pentachlorophenol March 31, 1987
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Absorption rate constants were not determined in rats. The average
absorption half-time was 2.7 hours in monkeys.
0 Braun et al. (1978) compared the absorption of PCP in man with the
values previously determined for monkeys and rats. Four healthy male
volunteers ingested a single dose of PCP at 0.1 mg/kg bw (vehicle not
specified). The average half-time for absorption was found to be
1.3 ± 0.4 hours.
Distribution
0 Once absorbed, PCP is distributed throughout the body, accumulating
in the liver, kidneys, brain, spleen and fat (Braun et al., 1977;
Grimm et al., 1981; Jakobson and Yllner, 1971). The PCP in the
plasma is highly protein-bound, which greatly reduces the tissue/plasma
concentration ratios and the renal clearance rate (Braun et al., 1977).
Metabolism
0 Pentachlorophenol apparently is not metabolized readily, since a
large portion of the administered dose is excreted unchanged by all
species tested. The major metabolic reactions of PCP are conjugation
to form the glucuronide and oxidative dechlorination to form tetra-
chlorohydroquinone (U.S. EPA, 1985a).
Excretion
Braun et al. (1978) orally administered PCP.at 0.1 mg/kg bw to four
male volunteers. They reported a plasma half-life of 30.2 hours.
Within 168 hours, 74% of the administered dose had been eliminated in
the urine as PCP and 12% as pentachlorophenol-glucuronide. An addi-
tional 4% was excreted in the feces as PCP and pentachlorophenol-
glucuronide combined.
Based on a single dose study, Braun et al. (1978) predicted that
steady-state levels would be reached in man within 8-9 days during
chronic exposure. Under these conditions, maximum blood levels would
be only about twice those observed following a single dose. Other
investigators (Casarett et al., 1969; Begley et al., 1977) have
reported half-lives for elimination of about 10 hours following an
acute exposure, which is consistei t with the value obtained by Braun
et al. (1978); however, after chronic exposure, half-lives of about
20 days were reported. Pentachlorophenol may, therefore, have a greater
potential for accumulation than the acute studies would indicate.
The major route of elimination is in the urine with fe^es as a minor
route. Only trace amounts of metabolites have been detected in
expired air. Biliary excretion occurs; however, extensive entero-
hepatic recirculation prevents this from being a major factor in the
elimination of PCP. Elimination is generally biphasic, with an
initial rapid phase, followed by a period of much less rapid
elimination. This pattern has been observed in rats (Braun et al.,
1977) and man (Bevenue et al., 1967), but not in subhuman primates
{Braun and Sauerhoff, 1976).
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Pentachlorophenol Si£6 March 31, 1987
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IV. HEALTH EFFECTS
Humans
0 Human exposure to PCP results in local irritation, systemic effects
and, in a limited number of people, an allergic response (Dow Chemical
Co., 1969). Pentachlorophenol poisoning is characterized by profuse
sweating, often accompanied by fever, weight loss and gastrointestinal
complaints (Gordon, 1956; Bergner et a^., 1965; Chapman and Robson,
1965). Liver and kidney involvement have been indicated in cases of
fatal poisoning (Robson et al., 1969; Armstrong et al., 1969).
0 Epidemiological studies have revealed effects following occupational
exposure to PCP. One group of subjects (wood treatment workers and
farmers/pest control operators in Hawaii) had a significantly increased
incidence of low-grade infections or inflammations (Klemmer et al.,
1980). Kidney function was depressed in wood treatment workers in-
Hawaii during chronic exposure, but these effects were at least
partially reversible (Begley et al., 1977). Gilbert et al. (1983)
indicated no adverse effects in wood treatment workers in Hawaii.
Animals
Short-term Exposure
0 Acute exposure of experimental mammals to pentachlorophenol results
in an initial rise in body temperature and respiration rate (U.S. EPA,
1985a). Respiration then becomes slower and dyspneic as coma develops.
Death is characterized by cardiac and muscular collapse with terminal
asphyxial convulsions. An immediate and pronounced rigor mortis often
is noted. These observations have been noted in studies where oral
LD^gs ranging from 27—>300 mg/kg bw have been reported, with no species
being noticeably more susceptible than any other. The lower LDjg
values tend to be found in the older literature and may reflect a high
degree of contamination by chlorinated dibenzo—p_-dioxins and dibenzo-
furans, although different dosing vehicles used in these various studies
could also have been influential.
0 Nishimura et al. (1982) found increased liver/body weight ratios in
male Wistar rats after single oral doses of sodium pentachlorophenate
at levels greater than 10 mgAg- The authors described the doses as
Pentachlorophenol.
Long-term Exposure
0 Johnson et al. (1973) fed PCP by diet at levels of 3, 10, or 30
mgAg/day to Sprague-Dawley rats for 90 days. Increased liver and
kidney weights were induced at all doses with a technical grade
containing high levels of dioxins (1,980 ppm OCDD, 19 ppm HCDD),
whereas increased liver and kidney weights were not evident at the
3 mg/kg/day feeding level with either a purified grade containing
no dioxins or an improved technical grade containing 30 ppm OCDD and
1 ppm HCDD.
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Pentachlorophenol /C,C < March 31, 1987
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0 Schwetz et al. (1978) fed commercial PCP (Dowicide E-7 containing
1 ppm HCDD and 15 ppm OCDD) in the diet at levels of 3, 10 or 30
mg/kg/day to male and female Sprague-Dawley rats for 2 years. Pigmen-
tation in liver and Kidneys was found with the two highest feeding
levels, and Schwetz et al. (1978) concluded that adverse effects were
not observed at 3 and 10 mg/kg/day in males and 3 mg/kg/day in females.
0 Oral doses of purified PCP at levels of 5, 10 and 15 mg/kg/day were
given to pigs for 30 days. Liver weights were increased in the 10
and 15 mg/kg/day groups (Greichus et al., 1979).
Reproductive Effects
0 Pentachlorophenol (Dowicide E-7) in the diet had no effect on repro-
ductive function and fetal development at 3 mg/kg/day in a one-gene-
ration reproduction study in Sprague-Dawley rats (Schwetz et al.,
1978). A feeding level of 30 mg/kg/day adversely affected reproduction
and fetal development.
Developmental Effects
0 Administration of commercial (88% pure) and purified (98% pure)
grades of PCP by gavage to pregnant Sprague-Dawley rats during days
6 through 15 of gestation did not result in teratogenic effects
(Schwetz and Gehring, 1973; Schwetz et al.,. 1974). The authors
concluded that 5 mg/kg/day was a no-observable-effeet-level (NOEL)
for fetotoxicity with the commercial grade and that an increase
in delayed skull ossification was evident at 5 mg/kg/day with the
purified grade. Effect levels were also 15, 30 and 50 mg/kg/day.
0 The conclusion in the U.S. EPA Position Document 4 (U.S. EPA, 1984a)
is that the results presented above did not establish a fetotoxicity
NOEL for either grade of pentachlorophenol and that 3 mg/kg/day
could be considered a provisional NOEL for fetotoxicity with penta-
chlorophenol.
Mutagenicity
0 Pentachlorophenol was negative for mutagenicity in Salmonella
typhinurium, Escherichia coli, Serratia marcescens, and Drosophila
melancgaster (U.S. EPA, 1985a). P«ntachlorophenol was positive for
forward mutation and intragenic recombination and negative for
intergenic recombination in Saccharomyces cerevisiae {Fahrig, 1974;
Fahrig et al., 1978). Pentachlorophenol was reported as positive
in the Bacillus subtilis rec assay (Shirasu, 1976; Waters et al.,
1982), in the mouse spot test (I'ahrig et al., 1978), and in cultured
human lymphocytes (Fahrig, 1974). Positive results in these studies
were reported as "slight" or "weak."
Carcinogenicity
0 Data currently are available from two oral studies in which the
carcinogenicity of PCP has been assessed in mice and rats (BRL, 1968;
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Pentachlorophenol <22S March 31. 1987
Schwetz et al, 1978). Pentachlorophenol was not found to be carcino-
genic in either of these studies, even though doses that produced
mild signs of toxicity were used. Catilina et al. (1981) also found
no evidence of carcinogenicity in Wistar rats following subcutaneous
administration; however, the dose level and duration of exposure were
limited in this study. Boutwell and Bosch (1959) also reported that
PCP is not a promoter of DMBA-induced skin carcinogenesis in Sutter
mice.
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) = _ /L ( _ /L)
(UF) x ( _ L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 1 00 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).
Taste and Odor Thresholds
The 30 ug/L taste threshold noted previously under Properties is below
the HAs based on toxicity data. The odor threshold noted is 1,600 ug/L.
One-day Health Advisory
Using the NOAEL of 10 mg/kg bw/day based on liver effects derived from
the study of Nishimura et al. (1982), which was selected in that it is a
single treatment study inducing effects in liver which is a target organ in
other studies, and an uncertainty factor (UF) of 100 for a NOAEL derived from
an animal study, a One-day HA is calculated as follows:
For a 10-kg child, the one-day HA is calculated as follows:
One-day HA = (1° mgAg/day) (10 kg) = •, <0 mg/L (j 000 ug/L)
(100) (1 L/day)
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Pentachlorophenol £*w March 31, 1987
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where:
10 mgAg/day « NOAEL for liver effects.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = daily water consumption of a child.
Ten-day Health Advisory
Appropriate studies for calculation of a Ten-day HA are not available.
It is recommended that the Longer-term HA for a 10-kg child of 0.3 mg/L be
used.
Longer-term Health Advisory
The NOAEL of 3 mg/kg/day for renal and hepatic effects identified in the
studies by Johnson et al. (1973) and for reproductive effects in the study
by Schwetz et al. (1978) is appropriate for calculating a Longer-term HA.
Using an uncertainty factor of 100 for a NOAEL derived from an animal study,
a Longer-term HA is calculated for a 10-kg child as follows:
Longer-term HA = (3 mg/kg/day) (10 kg) = 0>30 mg/L (300 ug/L)
(100) (1 L/day)
where:
3 mg/kg/day = NOAEL for renal, hepatic and reproductive effects.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
The Longer-term HA for the 70-kg adult is calculated as follows:
Longer-term HA = (3 mg/kg/day) (70 kg) = U05 /L (1 050 ug/L)
(100) (2 L/day)
where:
3 mg/kg/day = NOAEL for renal, hepatic and reproductive effects.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
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Pentachlorophenol 230 March 31, 1987
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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 24-month feeding study by Schwetz et al. (1978) is the most appropriate
from which to derive the Lifetime Health Advisory. The NOAEL of 3 mg/kg/day is
the same as that in the study of reproductive effects by Schwetz et al. (1978).
Using this NOAEL, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (3 mgAg/o'ay) = 0.03 mq/kq/day
(100)
where:
3 mg/kg/day = NOAEL for renal, hepatic and reproductive effects.
1 00 = uncertainty factor, chosen in accordance with NAS/ODW
guideliras for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.03 mg /kg/day) (70 kg) = , 05 /L (1 050 ug/L)
(2 L/day)
where:
0.03 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Pentachlorophenol "-*J_i_ March 31, 1987
-10-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = 1.05 mg/L x 20% « 0.22 mg/L (220 ug/L)
where:
1.05 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Since the available carcinogenicity studies on pentachlorophenol are
negative, water concentrations associated with an excess cancer risk
of 10-4, 10-5, and 10-6 were not calculated.
0 The IARC (1982) categorized pentachlorophenol as a Group 3 chemical
for carcinogenesis (inadequate data in humans and animals).
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), pentachlorophenol may be classified
in Group D: Not classified. This category is for agents with
inadequate animal evidence of carcinogenicity.
0 It is noted that the oncogenic substances hexachlorodibenzo-p-dioxin
and hexachlorobenzene have been detected as contaminants in technical
grade pentachlorophenol products (U.S. EPA, 1984a), and analysis for
pentachlorophenol in drinking water should also include analysis for
these substances.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Pentachlorophenol has been registered as a wood preservative since
1948, and it is currently registered for agricultural use as a seed
treatment for beans, alfalfa, clover, lespedeza and vetch. However,
the U.S. EPA (1984a) has issued a notice of Rebuttable Presumption
Against Registration (RPAR) for pesticide products containing penta-
chlorophenol (43 FR 48443-48477; 46 FR 13020) because of its feto-
toxicity and carcinogenic contaminants.
0 The ACGIH (1980, 1983) has established a threshold limit value (TLV)
of 0.5 mg/m^ and a short-term exposure level (STEL) of 1.5 mg/m^ for
dermal exposure.
n U.S. EPA (1980) reconmended an ambient water quality criterion for
pentachlo^opnenol of either 1.01 mg/L based on its toxicity or 0.030
mg/L based on its organoleptic properties.
0 The NAS (1977) calculated a safe drinking water level for a 70 kg
adult of 0.021 mg/L (21 ug/L) using the NOEL of 3 mg/kg/day in the
90-day toxicity study with purified pentachlorophenol in rats by
Johnson et al. (1973), a 1,000-fold uncertainty factor, and a 20%
relative source contribution.
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Pentachlorophenol 232 March 31, 1987
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The NAS (1986) stated that the basis used by the HAS (1977) would
yield a suggested-no-adverse response level (SNARL) of 6 ug/L for a
10 kg child. Using a 1,000-fold uncertainty factor and 20% residue
source contribution, NAS (1986) calculated for a 70 kg adult a SNARL
of 7 ug/L for technical pentachlorophenol using a NOEL of 1 mg/kg/day
and a SNARL for commercial pentachlorophenol of 9 ug/L using a NOEL
of 1.25
The WHO (1984) has recommended a drinking water criterion of 10 ug/L
for pentachlorophenol.
The proposed RMCL by the U.S. EPA Office of Drinking Water is 0.22 mg/L
(U.S. EPA, 1985b).
VII. ANALYTICAL METHODS
Determination of pentachlorophenol is by a liquid-liquid extraction
gas chromatographic procedure (U.S. EPA, 1984b). Specifically, the
procedure involves acidification of a 1-L sample followed by extraction
with methylene chloride using a separatory funnel. The methylene
chloride extract is dried and exchanged to 2-propanol during concen-
tration to a volume of approximately 10 mL. The compounds in the
extract are separated by gas chromatography and the phenols are
measured with a flame ionization detector. The method detection limit
is about 7.4 ug/L using this detector. Identification may be corrobo-
rated through the use of two unlike columns or using a gas chroma to-
graphy/mass spectrometer procedure (U.S. EPA, 1984c).
VIII. TREATMENT TECHNOLOGIES
Available data on pentachlorophenol (PCP) removal from water pertains
predominantly to adsorption techniques. The use of aeration also has
been considered.
Dobbs and Cohen's (1980) laboratory study reported the adsorption
capacities of Filtrasorb® 300 granular activated carbon (GAC) at
neutral pH for a wide range of toxic organic compounds. The adsorp-
tion capacity of 150 mg/g, reported for pentachlorophenol, was in the
top 50% of the 128 values reported, which ranged from 11,300 mg/g to
6.8 x 10~5 mg/g. Adsorption isotherms also were prepared for penta-
chlorophenol at pH 3, 7 and 9. These data indicated that adsorption
decreased with increasing pH.
Activated carbon has been used on a pilot-scale to remove PCP from
wood preservative effluent (Jank and Fowlie, 1980). The carbon
adsorption system consisted of three 10 mm diameter columns each
containing 6.8 kg of Calgon's Filtrasorb* 400 GAC which filled the
columns to 2.1 m depth. Feedwater to the column was prefiltered in
a column containing 11 kg of granular anthracite filtering media.
In different phases of the pilot study, wastewater was either passed
directly through the adsorption system or was subjected to an activated
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Pentachlorophenol •"-Oo March 31, 1987
-12-
sludge treatment prior to carbon adsorption. Activated sludge treat-
ment reduced the average PCP concentration from 8.4 mg/L to 3.6 mg/L;
this was reduced further to 3.4 mg/L after filtration and 0.03 mg/L
after GAC treatment. No additional PCP breakthrough occurred after
42 days of operation. When activated carbon was used to treat the
waste directly, the average PCP concentration was reduced from 0.44
mg/L to 0.02 mg/L. Adsorption onto GAC also was used to reduce the
PCP concentration in contaminated groundwater from 10 mg/L to less
than 1 ug/L (ESE, 1984). The water was passed through three 7-foot
diameter columns containing 1,230 pounds of 18 x 40 mesh GAC; total
contact time was 26 minutes. The GAC columns were part of a mobile
treatment unit, in which samples were pretreated using three mixed-
media (anthracite/sand) filters to remove suspended solids. In other,
separate incidents, the same unit was used to reduce PCP concentrations
in water from 400 ug/L to <0.075 ug/L and from 400 ug/L to <1 ug/L;
contact times were 30 to 40 minutes and 26 minutes, respectively.
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Pentachlorpphenol J2*3-t March 31, 1987
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IX. REFERENCES
ACGIH. 1980. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances In workroom
air. 4th ed. with supplements through 1981. Cincinnati, OH.
ACGIH. 1983. American Conference of Governmental Industrial Hygienists.
TLVs. Threshold limit values for chemical substances and physical agents
in the work environment with intended changes for 1983. Cincinnati, OH.
Armstrong, R.W., E.R. Eichner, D.E. Klein, et al. 1969. Pentachlorophenol
poisoning in a nursery for newborn infants. II. Epidemiologic and
toxicologic studies. J. Pediatr. 75(2):317-325.
Begley, J., E.L. Reichert, M.N. Rashad and H.W. Klemmer. 1977. Association
between renal function tests and pentachlorophenol exposure. Clin.
Toxicol. 11:97-106.
Bergner, H.f P. Constantinidis and J.H. Martin. 1965. Industrial penta-
chlorophenol poisoning in Winnipeg. Can. Med. Assoc. J. 92:448-451.
Bevenue, A., J. Wilson, L.J. Casarett and H.W. Klemmer. 1967. A survey of
pentachlorophenol content in human urine. Bull. Environ. Contain.
Toxicol. 2:319.
Boutwell, R.K., and K.K. Bosch. 1959. The tumor-promoting action of phenol
and related compounds for mouse skin. Cancer Res. 19:413-424.
Braun, W.H., and M.W. Sauerhoff. 1976. The pharmacokinetic profile of
pentachlorophenol in monkeys. Toxicol. Appl. Pharmacol. 38:525.
Braun, W.H., J.D. Young, G.E. 31au and P.J. Gehring. 1977. The pharmaco-
kinetics and metabolism of pentachlorophenol in rats. Toxicol. Appl.
Pharmacol. 41:395.
Braun, W.H., G.E. Blau and M.B. Chenoweth. 1978. The metabolism/pharmaco-
kinetics of pentachlorophenol in man, and a comparison with the rat and
monkey model, Toxicol. Appl. Pharmacol. 45:278.
BRL. 1968. Bionetics Research Laboratories. Evaluation of the carcino-
genic, teratogenic and mutagenic activities of selected pe jticides and
industrial chemicals. Volume 1. Carcinogenic study, Publ. No. NCI-
DCCP-CG-1973-1-1 . Prepared by BRL, Bethesda, MD, under Contract Nos.
PH 43-64-57 and PH 43-67-735. National Cancer Institute, Bethesda, MD.
(NTIS PB-223-159). p. 393.
Casarett, L.J., A. Bevenue, W.L. Yauger, Jr. and S.A. Whalen. 1969. Obser-
vations on pentachlorophenol in human blood and urine. Am. Ind. Hyg.
Assoc. J. 30:360-366.
Catilina, P., A. Chamoux, M.J. Catilina and J. Champeix. 1981. Study of the
pathogenic properties of substances used as wood protectives: Penta-
chlorophenol. Arch. Mai. Prof. Med. Trav. Secur. Soc. 42(65:334-337.
(Fre).
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Pentachlorophenol »*<~H.J March 31, 1987
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Chapman, J.B., and P. Robson. 1965. Pentachlorophenol poisoning from bath
water. Lancet. 1:1266-1267.
Dobbs, R.A., and J.M. Cohen. 1980. Carbon adsorption isotherms for toxic
organics. Report No. EPA-600/8-80-023. U.S. EPA. Office of Research
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Available from the Dow Chemical Co., Midland, MI 48640.
ESE. 1984. Environmental Science and Engineering. Review of treatability
data for removal of twenty-five synthetic organic chemicals from drinking
water. Prepared for U.S. EPA. Office of Drinking Water, Washington, DC.
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides. Int.
Chemical Carcinogenesis Essays. IARC Scientific Publ. No. 10. WHO,
1ARC. pp. 161-181.
Fahrig, R. C.A. Nilsson and C. Rappe. 1978. Genetic activity of chlorophenols
and chlorophenol impurities. In; Pentachlorophenol: Chemistry, Pharma-
cology and Environmental Toxicology, K.R. Rao, Ed. Plenum Press, NY.
pp. 325-338.
Gilbert, F.E., R.C. Duncan, W.H. Lederer and J.E. Wilkinson. 1983. Effects
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43(2):485-488.
Greichus, Y.A., G.W. Libal and D.D. Johnson. 1979. Diagnosis and physio-
logical effects of pentachlorophenols on young pigs. Part I. Effects
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23(3):418-422.
Grimm, H.G., B. Schellmann, K.H. Schaller and K. Gossler. 1981. Pentachloro-
phenol concentrations in tissues and body fluids of normal persons.
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(Ger).
IARC. 1982. International Agency for Research on Cancer. IAR" Monographs
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Chemicals, industrial processes and industries associated with cancer in
humans. WHO, IARC, Lyon, France. Supplement 4. pp. 205-206.
Jakobson, I., and S. Yllner. 1971. Metabolism of ^C-pentachlorophenol in the
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Johnson, R.L., P.J. Gehring, R.J. Kociba and B.A. Schwetz. 1973. Chlorinated
dibenzodioxins and pentachlorophenol. Environ. Health Perspect., Exp.
Issue No. 5, September, 1973. p. 171.
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Pentachlorophenol 236 March 31, 1987
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Klemmer, H.W., L. Wong, M.M. Sato, E.L. Retchert, R.J. Korsak and M.N. Rashad.
1980. Clinical findings in workers exposed to pentachlorophenol. Arch.
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Meerman, J.H.N., H.M.J. Sterenborg and G.J. Mulder. 1983. Use of pentachlo-
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Organic Solutes. National Academy Press, Washington, D.C. pp. 750-753.
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phenol poisoning in a nursery for newborn infants. I. Clinical features
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and commercial grade pentachlorophenol on rat embryonal and fetal
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Schwetz, B.A., J.F. Quast, P.A. Keeler, C.G. Humiston and R.J. Kociba. 1978.
Results of 2-year toxicity and reproduction studies on pentachlorophenol
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Pentachlorophenol March 31, 1987
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U.S. EPA. 1985a. U.S. Environmental Protection Agency. Drinking water
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238
March 31, 1987
TOXAPHENE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Hater (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. The chemical 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-hjt, Weibvll, 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 nodel is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Toxaphene
239
March 31, 1967
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This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for Toxaphene
(U.S. EPA, 1985a). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological 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 counter-
part (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 t 86-118049. The toll-free
number is (800) 336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. RN 8001-35-2
Structural Formula
0 Since the technical grade product is produced by free radical reac-
tions initiated by chlorine and ultraviolet light, toxaphene is a
complex mixture of poly chlorinated camphenes and bornanes with an
average empirical formula of C1oHiocl8 and an average molecular
weight of 414. In fact, more than 177 incompletely characterized
components have been separated (Holmstead et al., 1974).
Synonyms
Agricide maggot killer, Alltex, Camphechlor, Chem-Phene,
Compound 3956, and Kamfochlor.
Uses
Pesticide used to combat certain worms and insects such as cotton
boll weevils and ectoparasites on cattle and sheep (Meister, 1987).
Properties (Windholz, 1983)
Physical State
~oftening Range
Density
Vapor Pressure
Water Solubility
Octanol/Water Partition
Coefficient
Amber, waxy solid, with a mild
terpene odor
70-95eC
1.66 g/mL at 27°C
0.17-0.40 mmHg at 25CC
Approx. 3 mg/L or 0.5 mg/L at 25C<
Approx. 3,300 or 825
Occurrence
Toxaphene is an insecticide which up to 1982 was widely used on
soybeans, cotton, corn, wheat and other crops. The estimated pro-
duction of toxaphene in 1982 was 3.7 million Ibs. In 1982, all
major uses of toxaphene were canceled by EPA. Current use levels
of- toxaphene are believed to be very low.
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240
Toxaphene March 31, 1987
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Because toxaphene is a complex mixture of highly chlorinated com-
pounds, it is difficult to assess its behavior in the environment.
In general, toxaphene binds to soil and resists migration. Toxaphene
is considered to be an extremely persistent pesticide. Soil half-
lives of 20 years have been reported. Toxaphene is reported to
biodegrade slowly under certain anaerobic conditions.
Toxaphene is not degraded in surface water; however, its tendency
to bind tightly to sediments usually results in its rapid removal
from the water column. Toxaphene has been reported to bioaccumulate
readily in aquatic species.
Toxaphene has not been found in drinking water supplies in measurable
quantities. Studies of surface and ground waters generally have
not found detectable levels. A few positive samples of surface
water have reported levels of approximately 1 ppb (U.S. EPA, 1983).
Toxaphene has been reported to occur as a contaminant in food,
especially fish and other seafood. Toxaphene also has been detected
in air at low levels (U.S. FDA, 1984). Food appears to be the
major source of toxaphene exposure. Based upon the recent cancel-
lations of uses for toxaphene, occurrences of toxaphene in food and
water are expected to decline.
III. PHARMACOK.INETICS
Absorption
0 Toxaphene is absorbed through the skin (especially if mixed with
xylene), the lung and the gut (IUPAC, 1979; IARC, 1979). The rate
of absorption depends upon the vehicle for its administration.
Distribution
0 In general, toxaphene appears to be metabolized rapidly and its
metabolites quickly excreted in most species, with fat as the
preferred tissue of storage {IUPAC, 1979).
0 Twenty days after a single 20 ng/k
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241
Toxaphene March 31. 1987
-4-
Matsumura, 1979). The toxicity of toxaphene is potentiated in mice
by piperonylbutoxide, a known P-450 inhibitor (Saleh et al., 1977).
Excretion
The half-life of toxaphene in rats after administration of single
oral doses of 1*C- or 36ci-labeled toxaphene was less than a week
(Crowder and Dindal, 1974; Ohsawa et al., 1975). Most of the
36C1-labeled toxaphene was excreted within 6 to 7 days and little
remained in the tissues.
Approximately 49% of 36-Cl-toxaphene was eliminated in the urine of
rats 14 days after oral administration of 14.2 mg/kg bw. The feces
contained 27% of the ^^Cl-labeled metabolite after 14 days (Ohsawa
et al., 1975).
Toxaphene was excreted in the milk of cows fed 0 to 20 ppm toxaphene
in the diet for 77 days. Toxaphene levels in milk ranged from 0.043
to 0.179 ng/L and were dependent on the administered concentration
(Zweig et al., 1963). Following cessation of exposure, residues in
milk decreased to undetectable levels after two weeks in cows given
levels lower than 10 ppm. At the 20 ppm level, residues were still
detected 30 days after administration of the test diet was terminated.
IV. HEALTH EFFECTS
Humans
Toxaphene poisoning in humans is characterized by diffuse stimulation
of the central nervous system (CNS) resulting in salivation, rest-
lessness, hyperexcitability. muscle tremors or spasms, generalized
convulsions and sometimes loss of consciousness. Nausea and vomiting
may follow ingestion. Clonic convulsions also nay occur and can be
prevented by barbiturates (McGee et al., 1952).
The IUPAC (1979) has estimated an acute oral 1.050 of 60 m9Ag for
toxaphene.
At lease thirteen leaths from toxaphene poisoning have been recorded
(Hayes, 1975). Most of the fatal cases involved ingestion of toxaphene
by small childern.
Studies of human exposure to toxaphene in the workplace are confounded
because exposure to many chemicals occurred in all of the reported
studies. Two cases of acu?~e aplastic anemia after dermal exposure
to toxaphene/lindane have been reported; one death was due to myelo-
monocytic leukemia (IARC, 1979).
Animals
Short-term Exposure
Greater than ten-fold differences in toxicity have been documented
for various toxaphene fractions or components that differed from
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Toxaphene March 31, 1987
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each other in chemical composition, polarity and solubility (Pollack
and Kilgore, 1978a.b).
0 Toxaphene is a CKS stimulant in mammals. Effects of toxic exposures
in humans (hyper sensitivity, tremors and convulsions) are similar
to those observed in both rats and dogs (Lehman, 1951).
0 Rats fed a protein-deficient diet were more susceptiLle to toxaphene
poisoning than were rats fed regular laboratory chow, with LD^Q
values of 80 and 220 mg/kg bw, respectively (Boyd and Taylor, 1971).
Clinical signs of depression and stimulation of the CNS were the
same in both groups; however, signs appeared earlier and at lower
toxaphene concentrations in protein-deficient rats. This suggests
that humans who ingest a protein-deficient diet may represent *
sensitive subpopulation.
Long-term Exposure
0 In a study by Lackey (1949) dogs (breed not indicated, 3 to 8 per
dosage group) were administered toxaphene in corn oil by gavage at
5, 10, 15, 20, 25, 30, 40 or 50 mg/kg/day for an unstated number of
days. The author reported that at 5 mg/kg/day convulsions were
seen after a few days (number not stated). Deaths occurred in all
dosage groups except the two lowest (5 and 10 ag/kg/day). Deaths
were attributed to respiratory failure. Therefore, 5 mg/kg/day is
the LOAEL for convulsions after short-term exposure. No convulsions
were induced after a single dose of 5
In a second part of the Lackey (1949) study, dogs were administered
toxaphene at large cumulative doses (176 to 424 mg/kg) at 4 mgAg/day
for 44 to 106 days. At that level there was widespread degeneration
of the renal tubular epithelium, occasionally accompanied by pyelitis
(inflammation of the renal pelvis). Therefore, 4 mgAg/day is
identified as the LOAEL for renal effects in this study.
In a lifetime feeding study, Fitzhugh and Nelson (1951) observed
increased liver weights with minimal liver cell enlargement in rats
fed a diet containing toxaphene at 25 ppm (approximately 1.25
mg/kg/day based on the dietary assumptions of Lehman (1959)). In a
lifetime study in rats by L»hman (1952), this level resulted in no
effects, whereas 100 ppm (approximately 5 mgAg/day based on Lehman
(1959)) resulted in fatty degeneration of the liver. Boots Hercules
Agrochemicals, Inc. (not dated) reported liver necrosis in rats fed
toxaphene at 200 ppm (author's conversion: approximately 5 mgAg/day)
for 3.7 years. Clapp et al. (1971), however, observed no adverse
effects on liver histology even at doses up to 189 "ppm (approximately
9.45 mg/kg/day, based on Lehman (1959)). Based on these combined
observations, the LOAEL for liver effects is determined to be 1.25
mg/k9/<3ay. as reported in the Fitzhugh and Nelson (1951) study.
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••> 11
.•v'iO
Toxaphene March 31, 1987
-6-
• Allen et al. (1983) reported that IgG antibody titers were
depressed significantly and liver weights were increased in female
Swiss-Webster mice (an average of 14 animals per dosage group)
administered toxaphene for 8 weeks at dietary concentrations of
100 ppm and 200 ppm but not at 10 ppm (approximately 15, 30 and
1.5 mgAg/day, respectively, based on the dietary assumptions of
Lehman (1959)). Macrophage phagocytosis was significantly suppressed
in offspring exposed to toxaphene, both transplacentally and while
nursing, at all three dosage levels. Therefore, the LOAEL for this
effect is 1.5 ing/kg/day .
Reproductive Effects
0 Chernoff and Carver (1976) studied the potential fetal toxicity of
toxaphene in CD rats administered toxaphene at doses of 15, 25 or
35 mg/kg/day in corn oil by gastric intubation on days 7 through
16 of gestation. The numbers of rats per group were as follows:
33 control rats, 39 rats at 15 and 25 mgAg/day, and 16 rats at
35 ing/kg/day . At 35 mg/kg, toxicity was evidenced by 31% maternal
mortality. There was also a dose-related reduction in the weight
gain of dams (p<0.001) at 15 and 25 mg/kg/day. Even though there
was significant maternal toxicity in all treated groups, there were
no dose-related changes in fetal mortality or in the occurrence of
fetal anomalies. Therefore, the LOAEL for this study is 15 mg/kg/day.
Mutagenicity
• NTP (1983) reported toxaphene to be mutagenic when tested in the
Salmonella/microsomal reverse mutation assay. Hill (1977), summa-
rizing tests done by Litton Bionetics, Inc. (1978) for Hercules, Inc.,
indicated that toxaphene was directly mutagenic only for Salmonella
typhimurium strains TA98 (which detects frameshift mutagens) and
TA100 (nonspecific). By contrast, a "high tenperature" toxaphene
(high boiling component) was mutagenic only with activation by rat
liver microsomal preparation (S-9).
• Toxaphene has been tested in sister chromatid exchange systems,
dominant lethal assays and ONA breakage assays and produced negative
results (U.S. EPA, 1985a).
Carcinogenicity
• The most definitive study of toxaphene carcinogenic!ty was conducted
by the Tracer Jitco Co. under contract to the National Cancer Insti-
tute (NCI, 1979), despite the fact that the study was not conducted in
strict accordance with NCI guidelines (control groups contained only
10 animals each and paired-feeding was not done). Osborne-Mendel
rats and B6C3F1 mice (50/sex/treatment group; 10/sex/control group)
were used. Diets fed to male rats initially contained toxaphene at
2,560 and 1,280 ppm and the females received 1,280 and 640 ppm.
For mice of both sexes, the doses were 320 and 160 ppm. Because of
overt toxicity, these concentrations were lowered later. For male
rats, the high dose was was lowered to 1,280 ppm at two weeks, and
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o.V-1
Toxaphene **'*'* March 31/ 1987
-7-
to 640 ppm at 53 weeks after initiation of the study for an average
dose of 1,112 ppm. The low dose was similarly lowered to 640 ppm
after two weeks and 320 ppm 53 weeks after feeding had begun for an
average dose of 556 ppm. For female rats, both dose levels were
halved after 55 weeks, and average doses were calculated to be 540
and 1,080 ppm. For both sexes, toxaphene treatment was discontinued
after 80 weeks, and the animals were fed control diets without corn
oil for twenty weeks and then with corn oil for an additional 8
weeks. In male and female mice, both doses were halved 19 weeks
after treatment was initiated and average doses were 99 and 198
ppm. Toxaphene treatment was discontinued after 80 weeks, and
animals were fed control diets without corn oil for 7 weeks then
diets with corn oil for an additional 3 to 4 weeks.
Although none of the tumors observed in the animals was uncommon for
the animal strain used, certain,tumors and hyperplastic lesions were
present with higher incidence in the treated animals. In rats
these included thyroid follicular cell adenomas and carcinomas
(7/41 (17%) at the low dose; 9/35 (26%) at the high dose; and 1/7
(14%) in control males). Taking thyroid follicular cell adenomas
and carcinomas together, a statistically significant increase was
found for the high-dose group compared with the matched controls
for both male and female rats. Increased incidence of these lesions
was also significant in comparison with historical controls from
the same laboratory. In the female rats, there was also an elevated
cumulative incidence of tumors of the pituitary (chromophobe adenomas,
chromophobe carcinomas) in the high dose compared with the control
group.
In the mice, toxaphene was reported to be more toxic. Hepatocellular
carcinomas were observed with incidences of 691 and 98% in males at
the low and high doses, respectively, and at 10% and 69% in females
at the low and high doses, respectively. These neoplasms were not
observed in control animals of either sex, but hepatic nodules were
observed in 20% of matched-control males, but not in females. On
the basis of these findings, toxaphene was carcinogenic in B6C3F-|
mice and caused an increase incidence of thyroid tumors in Osborne-
Mendel rats.
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) . „ ( /L,
CUF) x { L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
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Toxaphene March 31, 1987
-8-
BW « assumed body weight of a child (10 kg) or
an adult (70 kg).
UF « uncertainty factor (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
The study by Lackey (1949) has been chosen to serve as the basis for calcu-
lating the One-day HA for toxaphene. This study reported convulsions in
dogs exposed to toxaphene at 5 mgAg/day for a few days (number not specified),
while 4 mgAg/day induced only occasional convulsions even with much longer
exposures. Based on the author's statements, no convulsions were induced
in dogs after a single dose of 5 mg/k? bw» while 10 rag/kg induced convulsions
in 4/5 animals. Therefore, 5 mg/kg/day has been identified as the NOAEL
for a single oral exposure to toxaphene.
The One-day HA for the 10 kg child is calculated as follows:
One-day HA = (5 mgAg/day) (10 kg) = 0,5 mg/L (500 ug/L)
(1 L/day) (100)
where:
5 mg/kg/day = NOAEL for convulsions after a single oral dose of
toxaphene in dogs.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = Assumed daily water consumption of a child.
Ten-day Health Advisory
Results of the Lackey (1949) study suggest that the chresholf of
toxicity for toxaphene may be slightly lower in dogs than in rodents. In
that study, degeneration of the renal tubular epithelium and pyelitis
(inflammation of the renal pelvis) were reported in dogs exposed to toxaphene
at 4 mgAg bw/day from 44 to 106 days. These effects were not observed at
higher doses in other species. It is uncertain from the study whether
these effects occurred within 10 days of exposure. Occasional convulsions
were also noted at 4 mgAg/day; therefore, this dose level is viewed as a
LOAEL for subchronic exposure to toxaphene.
Using the LOAEL of 4 mgAg/day, the Ten-day HA is calculated as follows:
Ten-day HA = (4 mgAg/day) (10 kg) = 0.04 mg/L (40 ug/L)
(1000X1 L/day)
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Toxaphene March 31, 1967
-9-
where:
4 mgAg/day * LOAEL for kidney effects in dogs.
10 kg « assumed body weight of a child.
1000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from ar animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
A Longer-term HA has not been calculated due to the lack of appropriate
data.
Lifetime Health Advisory and DWEL
There are no acceptable studies in the available literature for
derivation of a lifetime DWEL.
Evaluation of Carcinogenic Potential
0 Toxaphene may be classified in Group B: Probable Human Carcinogen,
according to EPA's proposed weight-of-evidence scheme for the
classification of carcinogenic potential. Because of this, caution
must exercised in making a decision on how to deal with possible
lifetime exposure to this substance. The risk manager must balance
this assessment of carcinogenic potential against the likelihood of
occurrence of health effects related to non-carcinogenic end-points
of toxicity. In order to assist the risk manager in this process,
drinking water concentrations associated with estimated excess
lifetime cancer risks over the range of one in ten thousand to one
in a million for the 70 kg adult, drinking 2 liters of water per day,
are provided.
0 Since the results of two bioassays (NCI, 1979; Litton Bionetics,
Inc., 1978) were positive for cancer induction, estimated risk
levels for toxaphene in drinking water can be calculated using the
linearized mulitstage model as discussed in the appendices to the
October Federal Register notice regarding the availability of Water
Quality Criteria Documents (U.S. EPA, 1980).
0 Drinking water concentrations estimated to result in lifetime
excess cancer risks of 10~4, 10~5 and 10~6 for a 70 kg adult drinking
2 liters of water per day over a 70 year lifespan are 3.1 ug/L,
0.31 ug/L and 0.031 ug/L, respectively (upper 95% confidence limit).
e The International Agency for Research on Cancer (IARC, 1979) has
placed toxaphene in category 2B, meaning that toxaphene is probably
carcinogenic in humans.
-------
Toxaphene **'lt March 31, 1987
-10-
• Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), toxaphene nay be classified
Group B2: Probable Human Carcinogen. This category is for agents
for which there is inadequate evidence frorr. human studies and
sufficient evidence from animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
• A TWA of 500 ug/m3 and a tentative short-tenr exposure limit of 1.0
mg/m3 have been set for toxaphene by ACGIH (1977).
• The National Interim Primary Drinking Water Standard for toxaphene is
5 ug/L (U.S. EPA, 1976).
• The NAS (1977) estimated the ADI of toxaphene for humans at 1.25 ugAg<
VII. ANALYTICAL METHODS
0 Determination of toxaphene is by a liquid-liquid extraction gas
chromatographic procedure (U.S. EPA, 1978; Standard Methods, 1985).
This procedure involves the use of 15% methylene chloride in hexane
for sample extraction, followed by drying with anhydrous sodium
sulfate, concentration of the extract and identification by gas
chromatography. Detection and measurement is accomplished by
electron capture, microcoulometric or electrolytic conductivity
gas chromatography. Identification may be confirmed through the
use of gas chromatography-mass spectroscopy (GC-MS). The method
sensitivity is 0.001 to 0.010 ug/L for single component pesticides
and 0.050 to 1.0 ug/L for multiple component pesticides when
analyzing a 1-liter sample with the electron capture detector
(Holmstead et al., 1974).
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies with limited data for removal of toxaphene
from drinking water are adsorption by granular activated carbon
(GAC) and powdered activated carbon (PAC), air stripping and
:oagulation/filtration. Other technologies adaptable to drinking
water systems might be able to remove toxaphene; however, such data
are not readily available.
• GAC columns mounted in a mobile unit have been used for the treatment
of hazardous waste spills. This unit proved to be 97% effective in
removing toxaphene from pond water in Virginia from an initial
concentration of 36 ug/L. The GAC columns, with a contact time of
26 minutes, treated the water at a rate of 70,000 gpd.
• Another study examined the effectiveness of PAC for the removal of
several fish poisons, including toxaphene. PAC was added to water
containing 0.3 mg/L of toxaphene. Toxaphene removals of 95% were
achieved at a carbon dosage of 9 mg/L. The results of this experiment
indicate that PAC is effective for toxaphene removal (U.S. EPA, 1985b).
-------
Toxaphene March 31, 1987
-11-
A theoretical nodel of an air stripping column was developed and
applied to the removal of some synthetic organic chemicals (SOC)
including toxaphene. The mass transfer coefficients for each SOC
were developed according to Perry and Chilton (1973). The results
show that toxaphene was 99\ theoretically removed at an air-to-water
ratio of 30. Actual air stripping performance data for removal of
toxaphene are not readily available.
Air stripping is a simple and relatively inexpensive process for
removing organics from water. However, use of this process then
transfers the contaminant directly to the air stream. When consid-
ering use of air stripping as a treatment process, it is suggested
that careful consideration be given to the overall environmental
occurrence, fate, route of exposure, and various hazards associated
with the chemical.
A conventional water treatment plant consisting of coagulation, sedi-
mentation and filtration reportedly had little effect on reducing
toxaphene from water. The influent toxaphene concentrations did
not exceed 0.41 ug/L (U.S. EPA, 1985b).
Treatment technologies for the removal of toxaphene from drinking
water have not been extensively evaluated except on an experimental
level. Individual or combinations of technologies'selected for
toxaphene reduction must be based on a case-by-case technical
evaluation, and an assessment of the economics involved.
-------
Toxaphene .'* 19 March 31, 1967
-12-
IX. REFERENCES
ACGIH. 1977. American Conference of Governmental Industrial Hygienists.
TLVs: Threshold limit values for chemical substances and physical agents
in the workroom environment with intended changes for 1977. Cincinnati, OH.
Allen, A.L., L.D. Roller and G.A. Pollack. 1963. Effect of toxaphene exposure
on immune responses in mice. J. Toxicol. Environ. Health 11:61-69.
Boots Hercules Agrochemicals. Hot dated. Boots Hercules toxaphene insecticide
summary of toxicological investigations. Bulletin T-105D.
Boyd, E.M., and F.I. Taylor. 1971. Toxaphene toxicity in protein-deficient
rats. Toxicol. Appl. Pharmacol. 18:156-167.
Chandurkar, P.S., and F. Matsumura. 1979. Metabolise of toxaphene components
in rats. Arch. Environ. Contain. Toxicol. 8:1-24.
Chernoff, N., and B.D. Carver. 1976. Fetal toxicity of toxaphene in rats
and mice. Bull. Environ. Contain. Toxicol. 15:660-664.
Clapp, K.L., D.M. Nelson, J.T. Bell et al. 1971. A study of the effects
of toxaphene on the hepatic cells of rats. In; Proc. Ann. Meet.
Western Section, Am. Soc. Anim. Sci. 22:313-323.
Crowder, L.A., and E.F. Dindal. 1974. Fate of ^^Cl-toxaphene *n rats.
Bull. Environ. Contain. Toxicol. 12:320-327.
Fitzhugh, O.G., and A.A. Nelson. 1951. Comparison of chronic effects
produced in rats by several chlorinated hydrocarbon insecticides.
Fed. Proc. 10:295.
Hayes, W.J. 1975. Toxicology of pesticides. Willians and Wilkins Co.
Baltimore, MD.
Hill, R.M. 1977. Memorandum to Fred Hageman. December 15. (U.S. EPA).
Holmstead, R.L., S. Khalifa and J.E. Casida. 1974. Toxaphene composition
analyzed by combined gas chromatography-chemical ionization mass
spectrometry. J. Agric. Ford Chem. 22:939-944.
IARC. 1979. International Agency for Research on Cancer. Toxaphene (poly-
chlorinated camphenes). IARC monographs on the evaluation of the
carcinogenic risk of chemicals to humans. 20:327-348.
IUPAC. 1979. International Union of Pure and Applied Chemistry. IUPAC
Reports on pesticides. 7. Toxaphene (camphechlor). A special report.
Pure Appl. Chem. 51:1583-1601.
Lackey, R.W. 1949. Observations on the acute and chronic toxicity of
toxaphene in the dog. J. Ind. Hyg. Toxicol. 31:117-120.
-------
Toxaphenc March 31, 1967
-13-
Lehman, A.J. 1951. Chemicals in foods: A report to the Association of
Food and Drug Officials on current developments. Part II. Pesticides.
Q. Bull. Assoc. Food Drug Off. U.S. 15:122-133.
Lehman, A.J. 1952. Oral toxicity of toxaphene. Q. Bull. Assoc. Food Drug
Off. U.S. 16:47.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Q. Bull. Assoc. Food Drug Off. U.S.
Litton Bionetics, Inc. 1978. Carcinogenic evaluation in mice. Toxaphene.
Final rep. LBI Project No. 20602, Kensington, MD. Submitted to
Hercules, Inc., Wilmington, DE.
McGee, L.C., H.L. Reed and J.P. Fleming. 1952. Accidental poisoning by
toxaphene. J. Amer. Med. Assoc. 1.49:1124-1125.
Meister, R.T., ed. 1987. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
NCI. 1979. National Cancer Institute. Bioassay of toxaphene for possible
carcinogenesis. NCI Carcinogenesis Tech. Rep. Ser. No. 37, DHEW Publ.
No. (NIH) 79-837.
NAS. 1977. National Academy of Sciences. Safe Drinking Water Committee.
Drinking Water and Health. Volume 1. National Academy Press. Washington,
D.C.
NTP. 1963. National Toxicology Program. Annual plan for Fiscal Year 1983.
NTP-82-119. p. 66.
Ohsawa, T., J.R. Knox, S. Khalifa et al. 1975. Metabolic dechlorination
of toxaphene in rats. J. Agric. Food Chem. 23:98-103.
Perry, R.H., and C.H. Chilton. 1973. Chemical Engineers Handbook. 5th
Edition. McGraw Hill Book Company.
Pollock, G.A., and R.lHillstrand. 1982. The elimination, distribution and
metabolism-of 14C-toxaphene in the pregnant rat. J. Environ. Sci.
Health. B17:635-648.
Pollock, G.A., and W.W. Kilgore. 1978a. Toxaphene. Residue Rev. 69:87-140.
Pollock, G.A., and W.W. Kilgore. 1978b. The metabolism and excretion of
toxaphene and selected toxaphene fractions. Toxicol. Appl. Pharraacol.
45:235.
Saleh, M.A., W.V. Turner and J.E. Casida. 1977. Polychlorobornane components
of toxaphene: Structure-toxicity relations and metabolic reductive
dechlorination. Science 198:1256-1258.
-------
Toxaphene 251 March 31. 1987
-14-
Standard Methods. 1985. Method 509A. Organochlorine Pesticides, Standard
Methods for the Examination of Mater and Wastewater, 16th Edition,
APHA, AWWA, WPCF, 19B5.
U.S. EPA. 1976. U.S. Environmental Protection Agency. National interim
primary drinking water regulations. Office of Hater Supply.
EPA-570/9-76-003.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for organo-
chlorine pesticides in drinking water. Methods for organochlorine
pesticides and chlorophenoxy acid herbicides in drinking water and raw
source water. Interim. July 1978.
U.S. EPA. 1980. U.S Environmental Protection Agency. Water quality
criteria documents: Availability. Fed. Reg. 45:79340-79341.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Occurrence of
pesticides in drinking water, food, and air. Office of Drinking
Water.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Draft health
effects criteria document for toxaphene. Office of Drinking Water..
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Draft technologies
- and costs for the removal of synthetic organic chemicals from potable
water supplies. Science and Technology Branch, CSD, ODW, Washington,
D.C.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for car-
cinogen risk assessment. Fed. Reg. • 51(185):33992-34003. September 24.
U.S. FDA. 1984. U.S. Food and Drug Administration. Surveillance Index
for Pesticides. Bureau of Foods.
Windholz, M. 1983. The Merck Index. 10th Edition. Merck and Co., Inc.,
Rahway, N.J.
Zweig, G., E.L. Pye, R. Sitlani, et al. 1963. Residues in milk from dairy
cows fed low levels of toxaphene in their daily ration. J. Agric. Food
Chert. 11:70-72.
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March 31, 1987
2,4,5-TRICHLOROPHENOXYPROPIONIC ACID
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTIOH
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. The chemical 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 r.odels 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.
-------
2,4,5-Trichlorophenoxypropionic Acid March 31, 1987
-2-
This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for 2,4,5-Trichloro-
phenoxypropionic Acid (2,4,5-TP) (U.S. EPA, 1985). The HA and CD formats are
similar for easy reference. Individuals desiring further information on the
toxicological 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 # 86-117926/AS. The
toll-free number is (800) 336-4700; in the Washington, DC. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 93-72-1
Structural Formula
COOH
2, 4, 5-Trichlorophenoxypropionic Acid
Synonyms
Silvex, 2,4,5-TP, Kurosal (Meister, 1984).
Uses
Herbicide (Meister, 1984).
Properties (Anonymous, 1979; Windholz et al, 1983)
Chemical Formula
Molecular Weight 269.5
Physical State crystals
Boiling Point —
Melting Point 181. 6°C
Densit ' 1.64 at 30°C
Vapor Pressure —
Water Solubility 140 mg/L at 25°C,
Log Octanol/Water Partition
Coefficient
Taste Threshold —
Odor Threshold —
Conversion Facton —
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2,4,5-Trichlorophenoxypropionic Acid /<-"* March 31, 1987
-3-
Occurrence
2,4,5-TP had a production volume of approximately 3 million pounds
per year in the early 1980s. 2,4,5-TP was used both directly and in
the form of various salts and esters. The use of 2,4,5-TP for weed and
brush control in specific areas was suspended by the EPA in 1979, and
all registrations for herbicides containing 2,4,5-TP are now cancelled
(U.S. EPA, 1985b).
2,4,5-TP is degraded in the environment and is not considered to be a
persistent compound. It is metabolized by plants to trichlorophenol.
2,4,5-TP has a half-life of 1.6 weeks in grasses and is degraded by
soil bacteria with soil half lives of 3-4 months. It photodegrades
in surface waters. 2,4,5-TP and some of its salts and esters have
been demonstrated to migrate in sandy and clay soils. Although
2,4,5-TP does not tend to accumulate in soils, it has been reported
to bi©accumulate in the fat and liver of some animals.
2,4,5-TP has been included in a number of national and regional surveys.
2,4,5-TP has been detected in only a small number of drinking water
supplies. Levels of contamination in ground water systems have been
below 0.3 ug/L, with most levels below 0.1 ug/L. Contamination has
occurred more frequently in surface waters than ground waters.
Contamination of surface waters appears to be the result of surface
water runoff from agricultural usage. There has been-a single
report of a surface water supply exceeding the 10 ug/L maximum concen-
tration limig (MCL) over a 3-month period.
2,4,5-TP has not been reported to occur in foods at detectable levels.
The present data are insufficient to determine whether food or water
is the greater source of exposure for 2,4,5-TP.
III. PHARMACOKINETICS
Absorption
Ingestion of small doses of 2,4,5-TP and 2,4,5-trichlorophenoxyacetic
acid (2,4,5-T) by mammals was followed by rapid and efficient absorp-
tion (up to 98 to 100%) from the gastrointestinal tract (Erne, 1966;
NRCC, 1978).
In rats, more than 90% of single oral doses of 2,4,5-T (0.17 to
41 mgAg) given by Fang et al. (1973) or 100 to 200 mgAg by Piper et
al. (1973) were absorbed from the digestive tract.
Distribution
Once the phenoxy acids are absorbed, they are distributed rapidly
throughout the body (U.S. EPA, 1985). They are bound to plasma
proteins, resulting in the retention in the plasma or extracellular
compartments or both.
-------
2,4,5-Trichlorpphenoxypropionic Acid March 31, 1987
-4-
0 Senczuk and Pogorzelska (1981) reported in an abstract that 2,4,5-TP
and 2,4,5-T blood levels remained high for 24 to 72 hours in rats
treated with 50 mg of the compound.
Metabolism
0 No information was found in the available literature on the metabolism
2,4,5-TP.
Excretion
0 No information was found in the available literature on the excretion
of 2,4,5-TP.
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the effects
resulting from acute exposure of humans to 2,4,5-TP.
0 Early studies using 2,4,5-TP and 2,4,5-T must be interpreted cautiously
because the herbicide samples may have been contaminated with fairly
high concentrations of 2,3,7,8-TCDD (dioxin) (U.S. EPA, 1985).
• Occupational exposure to 2,4,5-TP and 2,4,5-T (along with other
chemicals such as 2,4-D, di- and triphenols and 2,3,7,8-TCDD) resulted
in reduced nerve conduction velocities (Singer et al., 1982).
0 Case-controlled epidemiological sturHes of populations in Scandinavian
countries exposed to the phenoxy herbicides (as well as other chemicals
and contaminants) indicate excess risk of the development of soft-
tissue sarcomas and malignant lymphomas (Hardell et al., 1981).
Animals
Short-term Exposure
0 The acute oral LD50 value for 2,4,5-TP for rats is 650 mgAg (Rowe .-nd
H>mas, 1954).
Long-term Exposure
0 Subchronic exposure of rats to 2,4,5-TP by ingestion (5 mgAg/dav for
90 days) resulted in histopathologic changes in the liver and kidney
(Dow Chemical Company, 1962).
0 Dogs fed 2 mgAg/day for 89 days were reported to be without adverse
health effects. Details of this study were not available (Dow Chemical
Company, 1963).
0 Adverse effects such as mild hepatic degeneration and necrosis of
hepatocytes accompanied by elevations of SCOT and SGPT were observed
-------
2,4,5-Trichlorophenoxypropionic Acid "^ ° March 31, 1987
-5-
in dogs fed 2.6, 8.2 or 9.9 mg (acid equivalent) a.e./kg/day for
2 years (Gehring and Betso, 1978).
In beagle dogs fed 0.9 mg a.e.Ag bw/day (actual dose 0.75 mgAg/day)
for 2 years, no effects -were noted on body weight, food intake,
hematology, blood urea nitrogen or histopathology (Mullison, 1966).
Reproductive Effects
0 No information was found in the available literature on the reproduc-
tive effects of 2,4,5-TP.
Developmental Effects
0 Two studies have been located on the potential teratogenic activity
of 2,4,5-TP (Courtney, 1977; U.S. EPA, 1979). Both studies reported
teratogenic effects at dose levels of approximately 398 mg/kg/day and
25 to 100 mgAg/day in mice and rats, respectively; however, the lack
of conclusive evidence from these studies precludes the evaluation of
the teratogenic potential of 2,4,5-TP (U.S. EPA, 1985).
Mutagenicity
0 No information was found in the available literature on the mutagenic
effects of 2,4,5-TP.
Carcinogenici ty
0 Chronic oral exposure to 2,4,5-TP did not elicit significantly
increased incidences of tumors at any site in two strains of mice
exposed to approximately 17.2 mgAg/day for 80 to 81 weeks (Bionetics
Research Laboratories, 1968; Innes, 1969).
9 Rats exposed to 2,4,5-TP at doses up to 7.9 mg a.e.Ag/day for 2
years (Gehring and Betso, 1978) or dogs exposed at doses up to 9.9 mg
a.e.Ag/day for 2 years (Gehring and Betso, 1978) did not show any
significantly increased incidence of tumors.
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) = „ ( , }
(UF) x ( L/day) y/
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2,4, 5-Trichlorophenoxypropionic Acid -C.O » March 31, 1987
-6-
where:
NOAEL or LOAEL - No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW « assumed body weight of a child (10 kg) or
an adult (70 kg).
UF m uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODH guidelines.
L/day - assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No information was found in the available literature that was suitable for
determination of the One-day HA value for 2,4,5-TP. It is,therefore, recommended
that the Ten-day HA value of 0.2 mg/L (calculated below) for a 10-kg child be
used at this time as a conservative estimate of the One-day value.
Ten-day Health Advisory
The Dow Chemical Company (1963) study that identified a NOAEL of 2 mgAg
for liver and kidney effects in dogs has been selected to serve as the basis
for determination of the Ten-day HA value for 2,4,5-TP.
Using a NOAEL of 2 mg/kg/day, the Ten-day HA is calculated as follows:
Ten-day HA = (2 mgAg/day) (10 kg) , 0.20 mg/L (200 ug/L)
(1 L/day) (100)
where:
2 mgAg/day = NOAEL, based on the absence of liver and kidney effects
in dogs exposed to 2,4,5-TP for 89 days.
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with NOAEL from animal study.
Longer-term Health Advisory
No information was found in the available literature that was suitable
for determination of the Longer-term HA value for 2,4,5-TP. It is, therefore,
recommended that the DWEL, adjusted for a 10-kg child (0.07 mg/L) be used as
a conservative estimate of a Longer-term HA.
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258
2,4,5-Trichlorophenoxypropionic Acid March 31, 1987
-7-
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 study by Mullison (1966) has been selected to serve as the basis for
the Lifetime HA value to 2,4,5-TP. Beagle dogs were fed 0.9 mg a.e 2,4,5-TP/da;
for 2 years, with no observed effects on body weights, food intake, hematology,
blood urea nitrogen or histopathology. Based on these results, a NOAEL of
0.75 mg/kg/day (diet adjusted from 0.9 mg/kg) was identified.
Using the NOAEL from this study, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = _(p_._75 mg/kg/day) = 0.0075 mgAg/day
(100)
where:
0.75 mg/kg/day = NOAEL, based on the absence of toxic effects in dogs
esposed via the diet for 2 years.
100 = uncertainty factor, chosed in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0075 mg/kg/day) (70 kg) = 0.260 mg/L (260 ug/L)
(2 L/day)
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2,4,5-Trichlorpphenoxypropionic Acid March 31, 1987
-8-
where:
0.0075 mgAg/day - RfD.
70 kg «= assumed body weight of an adult.
2 L/day * assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA « (0.260 mg/L) (20%) - 0.052 mg/L (52 ug/L)
where:
0.260 mg/L « DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer has not classified
2,4,5-TP into categories for carcinogenic potential.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), 2,4,5-TP may be classified in
Group D: not classified. This category is for substances with in-
adequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Interim Primary Drinking Water Regulations (U.S. EPA,
1979) allows up to 10 ug/L of 2,4,5-TP in drinking water.
0 The National Academy of Sciences (NAS, 1977) has recommended an ADI
for humans of 0.00075 mgAg/day for 2,4,5-TP, based on a NOAEL of
0.75 mgA9/day and a 1,000-fold uncertainty factor. From the ADI, they
derived a suggested-no-adverse-response-level (SNARL) of 0.00525 mg/L.
0 The EPA has proposed a Recommended Maximum Contaminant Level (RMCL)
of 0.052 mc,/L based on. the Adjusted ADI (AADI) of 0.26 mg/L with
20% drinking water contribution (U.S. EPA, 1985b).
VII. ANALYTICAL METHODS
0 Determination of 2,4,5-TP is by a liquid-liquid extraction gas
chromatographic procedure (U.S. EPA, 1978; Standard Methods, 1985).
Specifically, the procedure involves the extraction of chlorophenoxy
acids and their esters from an acidified water sample with ethyl
ether. The esters are hydrolyzed to acids, and extraneous organic
material is removed by a solvent wash. The acids are converted to
methyl esters which are extracted from the aqueous phase. Separation
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2, 4, 5-Trichlorophenoxypropionic Acid March 31, 1987
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and identification of the esters is made by gas chromatography.
Detection and measurement is accomplished by an electron-capture,
microcoulometric or electrolytic conductivity detector. Identifica-
tion may be corroborated through the use of two unlike columns. The
detection limit is dependent on the sample size and instrumention
used. Typically, using a 1-liter sample and a gas chromatograph with
an electron-capture detector results in an approximate detection
limit of 10 ng/L for 2,4,5-TP.
VIII. TREATMENT TECHNOLOGIES
0 Robeck et al. (1965) demonstrated removal of 2,4,5-T (BE) in spiked
(1 or 10 ug/L), distilled or river water mixed with powdered activated
carbon (PAC). The river water was treated with PAC in a pilot plant.
After PAC addition and mixing, the sample was flocculated and allowed
to settle, then filtered through sand and coal and final'ly through
granular-activated carbon (GAC) beds. Cumulative removals of 10 ug/L
(raw water concentration) of 2,4,5-T (BE) were reported as 80, 90 and
95* at carbon slurry concentrations of 5, 10 and 20 mg/L, respectively.
Greater than 99% removal was reported using a GAC bed.
0 Treatment technologies for the removal of 2,4,5-TP from water have not
been evaluated, although information on 2,4,5-T (BE) indicates that
adsorption merits further investigation. Any selection must be based
on a case-by-case evaluation, and an assessment of the economics
involved.
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IX. REFERENCES
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Bionetics Research Laboratories. 1968. Evaluation of carcinogenic, terato-
genic and mutagenic activities of selected pesticides and industrial
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Courtney, K.D. 1977. Prenatal effects of herbicides. Evaluation by the
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Dow Chemical Company. 1962. Results of the 90-day dietary feeding of the
propylene glycol isobutyl ether ester of Silvex (Dowco 171) to rats.
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Dow Chemical Company. 1963. Results of 89-day dietary studies — Kurosal,
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Erne, K. 1966. Distribution and elimination of chlorinated phenoxyacetic
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Fang, S.C., E. Fallin, H.L. Montgomery and V.H. Freed. 1973. The metabolism
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Toxicol. Appl. Pharmacol. 24:555-563.
Gehring, P.J., and J.E. Betso. 1978. Phenoxy acids: Effects and fate in
mammals. In; Chlorinated Phenoxy Acids and Their Dioxins, C. Ramel, Ed.
Ecol. Bull. (Stockholm). 27:122-133.
Hardell, L., M. Eriksson, P. Lenner and E. Lundgren. 1981. Malignant
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Innes, J.R.M., B.M. Ulland, M.G. Valerio and L. Petrucelli. 1969. Bioassay
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Meister, R. ed. 1984. Farm chemic .Is handbook, Willoughby, OH:. Meister
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NRCC. 1978. National Research Council of Canada. Phenoxy Herbicides —
Their effects on environmental quality with accompanying scientific
criteria for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). NRCC No. 16075.
Associate Committee on Scientific Criteria for Environmental Quality,
Ottawa, Canada.
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