UNITED STATES ENVIRONMENTAL PROTECTION AGENCY
REGION II
DATE: 1 DEC "37
SUBJECT: Draft Health Advisories for 50 Pesticides
FROM: Walter E. Andrews, Chief .
Drinking/Ground Water Protection Branch^
T0: Barbara Metzger, Director
(2ESD), Bldg. 10, Edison, NJ
Transmitted herewith please find draft Health Advisories for fifty (50)
pesticides. Notices of availability will scon be published in the
Federal Register. For further information, please contact Ms. Jennifer
Qrme, ODW Health Advisory Program Coordinator, (202) 382-7586 or Edward
V. Ohanian, Chief, Health Effects Branch (202) 382-7571.
Enclosures
REGION II FORM 132O-1 (9/85)
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August, 1987
ACIFLUORFEN
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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Acifluorfen
August, 1987
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II. GENERAL INFORMATION AMD PROPERTIES
CAS No. 5094-66-6 (acid)
62476-59-9 (sodium salt)
Structural Formula
Sodium 5-(2-chloro-4-(trifluoromethyl)-phenoxy)-2-nitrobenzoate
Synonyms
• Blazer*) Carbofluorfen; RH-6201) Tackle9; Sodium acifluorfen (Meister,
1983).
Ust
• Acifluorfen is used as a selective pre- and post-emergence herbicide to
control weds and grasses in large-seeded legumes including soybeans
and peanuts (Meister, 1983).
Properties (Windholz et al., 1983; Meister, 1983; CHEMLAB, 1985)
Chemical Formula
Molecular Height
Physical State (25°C)
C14H7C1F3N05 (acid)
C14HgClF3NNa05 (sodium salt)
361.66 (acid)
383.65 (sodium salt)
Off-white solid (acid), brown crystalline
powder/white powder (sodium salt)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Hater Solubility (25°C)
Log Octanol/Hater Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
124-125°C (sodium salt)
151.5-157°C (acid)
>25% (sodium salt) (dimensions not
specified)
-4.85 (acid) (calculated)
Occurrence
• No information was found in the available literature on the occurrence
of acifluorfen.
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Acifluorfen August, 1987
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Environmental Fate
0 Acifluorfen is stable to hydrolysis; no degradation was observed
in solutions at pH 3, 6 or 9 within a 28-day interval. Varying
temperatures (18 to 40°C) did not alter this stability. The half-life
of the parent compound is 92 hours under continuous exposure to light
approximating natural sunlight. The decarboxy derivative of acifluorfen
was the primary degradate found in solution. It is suspected that a
substantial percentage of the photodegradate parent is lost from
solution (through volatilization or other mechanisms) (Registrant
CBI data).
0 The half-life of acifluorfen in an aerobically incubated soil was
found to be about 170 days; anaerobic degradation was more rapid
(half-life about 1 month). The dominant residue compounds after
6-months aerobic incubation were the parent compound and bound
materials. After 2 months under anaerobic conditions, the acetamide
of amino acifluorfen was the major degradate extracted from soil; the
amino analog itself was also significant, and denitro acifluorfen was
also formed (Registrant CBI data).
0 Acifluorfen applied at 0.75 Ib ai/A to a silt loam in Mississippi
dissipated with a tentative half-life of 59 days. Leaching of the
parent compound below 3 inches in the soil was negligible during the
179-day study. The dissipation of acifluorfen in two silt loam soils
in Illinois receiving multi-residue treatments was somewhat slower;
half-lives were 101 to 235 days (Registrant CBI data).
0 Acifluorfen applied to soil columns at highly excessive rates indica-
tive of spills (682 Ib ai/A) is very mobile. Acifluorfen leached
from the columns with 10 inches of water accounted for 79 to 93% of
the acifluorfen applied. Aerobic aging of the residues in the column
substantially reduced the mobility and pesticide movement was inversely
proportional to the soil CEC. Results from soil TLC (un-aged residues
only) predict mobility to be intermediate to mobile. Supplementary
data from a batch adsorption study indicate that un-aged acifluorfen
is weakly and reversibly adsorbed (Registrant CBI data).
0 Greenhouse studies have demonstrated that the uptake of acifluorfen
by rotational crops decreases with aging of residues in soil (Registrant
CBI data).
III. PHARMACOKINETICS
Absorption
0 No information was found in the available literature on the absorption
of acifluorfen.
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Distribution
• No information was found in the available literature on the distribution
of acifluorfen.
Metabolism
• No information was found in the available literature on the metabolism
of acifluorfen.
Excretion
0 No information was found in the available literature on the excretion
of acifluorfen.
IVo HEALTH EFFECTS
Humans
Short-term Exposure
0 No information was found in the available literature on the short-term
health effects of acifluorfen in humans.
Long-term Exposure
0 No information was found in the available literature on the long-term
health effects of acifluorfen in humans.
Animals
Short-term Exposure
0 The Whittaker Corporation (no date, a) reported that the oral LDso of
Tackle 2S (a formulation containing 20.2% sodium acifluorfen) in the
rat (strain not specified) was 2,025 mg/kg for males and 1,370 mg/kg
for females.
0 Meister (1983) reported that the acute dermal LD50 of Blazer* (tech-
nical grade, purity unspecified) in the rabbit is 450 mg/kg. The
acute dermal LD5o of Tackle* (purity unspecified) in the rabbit is
2,000 mg/kg.
0 Goldenthal et al. (1978a) presented the results of a two-week range-
finding study in which RH 6201 (a formulation containing 39.4% sodium
acifluorfen) was administered to Charles River CD-1 mice (10/sex/dose)
at dietary concentrations of 0, 625, 1,250, 2,500, 5,000 or 10,000
ppm. Assuming that 1 ppm in the diet of mice is equivalent to 0.15
mg/kg/day (Lehman, 1959), these doses correspond to about 0, 93.8,
187.5, 375.0, 750.0 or 1,500 mg/kg/day. No changes in general behavior
or appearance were reported at any dose level. During the second
week of the study, there was a decrease in body weight and food
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Acifluorfen f 1987
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consumption in animals receiving 10,000 ppm (1,500 mg/kg/day). Gross
pathological findings included pale kidneys, yellowish livers and
reddish foci of hyperemia in the stomachs of several mice at the
5,000- and 10,000-ppm (750 and 1,500 mg/kg/day) dose levels. Absolute
liver weight was increased in all test groups dosed at levels of
2,500 ppm (375 mg/kg/day) or greater. The increases were statistically
significant (p <0.01). A statistically significant (p <0.01) increase
in relative liver weight was reported at all dose levels. Based on
the results of this study, a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 625 ppm (93.8 mg/kg/day) was identified.
• Piccirillo and Robbins (1976) administered RH 6201 (a formulation
containing 39.8% sodium acifluorfen) to Wistar rats (5/sex/dose) for
4 weeks at dietary concentrations of 0, 5, 50, 500 or 5,000 ppm
(reported to be equivalent to 0, 0.7, 7.6, 55.4 or 506.4 mg/kg/day
for males and 0, 0.8, 8.3, 60.6 or 528.2 mg/kg/day for females).
Assuming that these dietary levels reflect the concentration of the
test compound and not the active ingredient, corresponding levels
of sodium acifluorfen are 0, 0.3, 3.0, 22.1 and 201.6 mg/kg/day for
males and 0, 0.3, 3.3, 24.0 and 210.2 mg/kg/day for females (Lehman,
1959). Results of the study indicated that body weight was decreased
in males at 22.1 and 201.6 mg/kg/day, and food consumption was decreased
in both males at 201.6 mg/kg/day and females at 210.2 mg/kg/day.
Biochemical analyses revealed that serum glutamic pyruvic transaminase
(SGPT) levels were increased in males at 22.1 and 201.6 mg/kg/day;
in males that received 201.6 mg/kg/day, blood urea nitrogen (BUN)
was increased and glucose levels were decreased. Changes in organ
weights included increased absolute liver and kidney weights in males
at 201.6 mg/kg/day, increased relative liver and kidney weights in
males at 201.6 mg/kg/day and females at 210.2 mg/kg/day and increased
relative liver weight in males only at 22.1 mg/kg/day. Based on the
results of this study, a No-Observed-Adverse-Effect-Level (NOAEL) of
3.0 mg/kg/day was identified.
Dermal/Ocular Effects
0 In a dermal irritation study (Whittaker Corp., no date, b). Tackle 2S
(a formulation containing 20.2% sodium acifluorfen) was applied
occlusively (dose not specified) to the intact and abraded skin of
rabbits. Effects observed included slight erythema, slight edema,
blanching of the skin, and eschar formation. Sign? of dermal
irritation at intact and abraded sites were absent by 8 days post-
application. The test substance was considered to be a moderate
dermal irritant at 72 hours.
0 In a dermal irritation study, Weatherholtz et al. (1979b) applied
RH 6201 (sodium acifluorfen) to the skin of New Zealand White rabbits
(five/sex/dose; ten/sex/control). Three different formulations of
RH 6201 were used in the study and each formulation was tested at
1.0 or 4.0 mLAg/day. The authors indicated that for all RH 6201
formulations tested, the dose levels correspond to 50 or 200 mg/kg/day
of the active ingredient. The test material was applied once daily
for 5 days, followed by 2 days with no applications, over a 4-week
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Acifluorfen August, 1987
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period (total of 20 applications). At both dose levels, two of
the formulations produced slight to well-defined irritation. At
200 mg/kg/day, central nervous system depression and a statistically
significant decrease in body weight gain and food consumption were
noted. The third formulation produced essentially the same effects,
with the addition of "thinness," ataxia, slight tremors and mortality
(2/5 males). Microscopic evaluations revealed chronic dermatitis,
acanthosis and hyperlceratosis at both dose levels for all formulations.
0 Madison et al. (1981) presented the results of the Buhler test for
dermal sensitization in Hartley-derived albino guinea pigs. In this
study, Tackle* (sodium acifluorfen; purity not specified) was not found
to be a sensitizer when applied topically at a dose of 0.25 mL under
occlusive binding.
• In an ocular irritation study (Whittaker Corp., no date, c). Tackle 2S
(a formulation containing 20.3% sodium acifluorfen) was instilled
into the eyes of rabbits. Signs of ocular irritation and lesions
included opacities of the cornea, iritis, redness and chemosis of the
conjunctiva and discharges from both washed and unwashed eyes. Four
of six unwashed eyes and one of three washed eyes exhibited blistering
of the conjunctiva. Three of six unwashed and one of three washed
eyes exhibited pannus where corneal opacity had been.
• In an ocular irritation study (Weatherholtz et al., 1979a), 0.1 mL of
Blazer 2S (purity not specified) was applied to the corneal surface
'of the eyes of rhesus monkeys. Corneal opacity and conjunctival
redness, swelling and discharge were observed in both washed and
unwashed eyes. All treated eyes were free of signs of irritation by
14 days posttreatment.
Long-term Exposure
0 Harris et al. (1978) administered RH 6201 (a formulation containing
39.4% sodium acifluorfen) in the diet to Sprague-Dawley rats (15/sex/
dose) for 3 months at dose levels of 0, 75, 150 or 300 mg/kg/day.
Assuming that these doses reflect levels of the test compound and not
the active ingredient, corresponding levels of sodium acifluorfen
would be 0, 29.6, 59.1 or 118.2 mg/kg/day. At the highest dose level
(118.2 mgAg/day), a number of effects were observed in male rats.
These effects included decreased body weight (13%) and decreased food
consumption (8%). Biochemical analyses of blood revealed increased
alkaline phosphatase levels (32%), decreased total protein (8%) and
decreased albumin (14%). No such effects were reported for female
rats. (These biochemical analyses were performed on control and high-
dose animals only.) Increased liver weight and microscopic liver
changes (enlarged hepatocytes) were observed in male rats that received
59.1 or 118.2 mgAg/day. m terms of the active ingredient, a NOAEL
of 29.6 mg/kg/day was identified.
• Barnett (1982) administered Tackle 25 (a formulation containing
20.4 to 23.6% sodium acifluorfen) to Fischer 344 rats (30/sex/dose)
for 90 days at dietary concentrations of 0, 20, 80, 320, 1,250,
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Acifluorfen August, 1987
•
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2,500 or 5,000 ppm. The author indicated that these dietary levels
correspond to average compound intake levels of 0, 1.5, 601, 23.7,
92.5, 191.8 or 401.7 mg/kg/day for males and 0, 1.8, 7.4, 29.7,
116.0, 237.1 or 441.8 mg/kg/day for females. Assuming that these
levels reflect test compound and not active ingredient intake,
corresponding levels of sodium acifluorfen intake are approximately
0, 0.4, 1.4, 5.6, 21.8, 45.3 or 94.8 mg/kg/day for males and 0, 0.4,
1.8, 7.0, 27.4, 56.0 or 104.3 mg/kg/day for females (based on 23.6%
active ingredient in test compound). At 5,000 ppm the following
effects were observed: decreased body weight and food consumption
in both sexes; decreased red blood cell (RBC) count, hemoglobin and
hematocrit in both sexes; increased serum cholesterol and serum calcium,
and decreased serum phosphorous in both sexes; increased alkaline
phosphatase, SGPT and BUN levels in males; elevated urobilinogen in
both sexes; increased liver size and discolored liver and kidneys
in both sexes; and liver cell hypertrophy and increases in mitotic
figures and individual cell deaths in both sexes. At 2,500 ppm the
following effects were observed: decreased body weight in males;
decreased RBC count, hemoglobin and hematocrit in both sexes; increased
BUN levels in males; elevated urobilinogen in both sexes; increased
liver size in both sexes; and liver cell hypertrophy and increases
in mitotic figures and individual cell deaths in both sexes. At
1,250 ppm, the following effects were observed: increased liver
size in males and liver cell hypertrophy in both sexes. The author
identified 320 ppm as the NOAEL in this study. In terms of active
ingredient concentration, this corresponds to a NOAEL of 5.6 mg/kg/day
for males and 7.0 mg/kg/day for females.
Mobil (1981) presented the 6-month interim results of a longer-term
study in which Tackle 25 (a formulation containing approximately 75%
sodium acifluorfen) was administered to beagle dogs (eight/sex/dose)
at dietary concentrations of 0, 20, 320 or 4,500 ppm. These dietary
levels were reported to be equivalent to 0, 0.7, 9.0 or 160 mg/kg/day.
Assuming that these levels reflect test compound and not active
ingredient intake, corresponding levels of sodium acifluorfen intake
are 0, 0.5, 6.8 or 120.0 mg/kg/day (based on 75% active ingredient in
the test compound). Following six months of compound administration,
two animals/sex/dose were sacrificed. The study reported a number of
effects at the highest dose tested. These effects included decreased
body weight and food consumption and increased liver weight in both
sexes. Additionally, RBC count and hemoglobin concentration were
decreased in both sexes. Clinical chemistry analyses revealed
depressed serum cholesterol, increased alkaline phosphatase, and
transient elevation of BUN in both sexes. Males only showed increased
levels of lactic dehydrogenase. No histopathological examinations
were conducted. The NOAEL reported in this study was 320 ppm. In
terms of the active ingredient, this corresponds to a NOAEL of
6.8 mg/kg/day.
Barnett et al. (1982b) administered Tackle 2S (a formulation con-
taining 19.1 to 25.6% sodium acifluorfen) to Fischer 344 rats
(73/sex/dose) for 1 year at dietary levels of 0, 25, 150, 500, 2,500
or 5,000 ppm. Assuming that these dietary levels reflect the
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Acifluorfen August, 1987
•
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concentrations of the test compound and not the active ingredient,
corresponding levels of sodium acifluorfen are 0, 6.4, 38.4, 128.0,
640.0 or 1,280 ppm (based on 25.6% active ingredient in the test
compound). Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day, these levels correspond approximately to 0, 0.3, 1.9,
6.4, 32.0 or 64.0 mg/kg/day (Lehman, 1959). No excess moribundity or
mortality was associated with the ingestion of the test substance.
At 5,000 ppm, the following effects were observed: decreased mean
body weight in both sexes; increased absolute and relative liver
weight in both sexes; decreased protein production, decreased serum
glucose, decreased triglyceride levels, increased alkaline phosphatase
and creatine phosphokinase levels, and sporadic increases in SCOT and
SGPT in both sexes; a slight increase in the excretion of urobilinogen
in both sexes; and the presence of acidophilic cells that were
considered to be evidence of cytotoxic changes in the livers of both
sexes. At 2,500 ppm, male rats showed increased absolute and relative
liver weights. Based on the information presented in this study, a
NOAEL of 500 ppm was identified for the test compound. In terms of
the active ingredient, this corresponds to a NOAEL of 6.4 mg/kg/day.
Spicer et al. (1983) administered Tackle 2S (a formulation containing
74.5 to 82.8% sodium acifluorfen) to beagle dogs (eight/sex/dose)
for 2 years at dietary concentrations of 0, 20, 300 or 4,500 ppm,
reported to be equivalent to 0, 0.5, 7.3 or 121 mg/kg/day for males
and 0, 0.5, 8.3 or 154 mg/kg/day for females. Assuming that these
dietary levels reflect the concentration of the test compound and not
the active ingredient, corresponding levels of sodium acifluorfen are
0, 0.4, 6.0 or 100.2 mg/kg/day for males and 0, 0.4, 6.9 or 127.5
mg/kg/day for females (based on 82.8% active ingredient in the test
compound). At the highest dose, body weight was decreased (not
statistically significant), and a corresponding (statistically sig-
nificant) decrease in food consumption was also reported. Physical
examination revealed heart anomalies in the high- and mid-dose groups.
At the high dose, irregular heart rhythms and rapid or slow heart
rates were reported in one male and four females. Also at this dose
level, one male was found to have a systolic murmur. At the mid-dose
level, one animal of each sex had an irregular heart rhythm (accompanied
by rapid heart rate in the male). At the highest dose tested, a
number of changes were reported, including a statistically significant
decrease in erythrocyte count, hemoglobin and hematocrit in both
nexes; reductions in albumin and cholesterol; increased absolute ar.1
telative liver and kidney weights; and histopathological liver changes
including centrilobular hepatocellular fatty -vacuolation, bilirubin
pigmentation and minimal foci of alteration. Renal tubules showed
bilirubin pigmentation at all dose levels (most pronounced at the
high dose). The authors concluded that this study showed clear
evidence of target organ toxicity affecting the liver and possibly
the kidney at the highest dose level. The authors identified 300 ppm
(of test compound) as the NOAEL. In terms of the active ingredient,
this corresponds to a NOAEL of 6.0 mg/kg/day.
Goldenthal (1979) administered RH 6201 (a formulation containing 39.4
to 40.5% sodium acifluorfen) to Charles River CD-1 mice (80/sex/dose)
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Acifluorfen August, 1987
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for two years in the diet at concentrations that provided dosage
levels of 0, 1.25, 7.5 or 45.0 ppm of the active ingredient. After
16 weeks of administration, the 1.25 ppm dose was increased to 270 ppm.
Assuming that 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day,
these levels correspond to about 0, 0.19 (increased to 40.5), 1.13 and
6.8 mg/kg/day (Lehman, 1959). Two control groups were used in this
study. One group received acetone in the diet (control 1), and the
other received water in the diet (control 2). At the 40.5 mg/kg/day
dose level, the following effects were observed: slight to marked
elevations in alkaline phosphatase and SGPT levels, in both sexes,
beginning after one year of exposure; increased absolute and relative
liver weight in males; increased absolute liver weight in females;
increased relative kidney weight in males; decreased absolute heart
weight in males; cellular alterations in the livers of males consisting
of focal pigmentation, focal hepatocytic necrosis, focal cellular
alteration, nodular hepatocellular proliferation and hepatocellular
carcinoma (the only statistically significant change was the focal
cellular alteration); and focal pigmentation in the livers of females.
At the 6.8 mg/kg/day dose level, the following effects were observed:
occasional increases in alkaline phosphatase and SGPT levels in both
sexes; decreased absolute heart weight in males; and focal pigmentation
in the livers of females. The author indicated that changes with an
apparent dose-related distribution included focal pigmentation,
hepatocellular vacuolation, focal hepatocytic necrosis and nodular
hepatocellular proliferation. The incidence of hepatocellular
carcinoma in males of all treatment groups was approximately the same.
A NOAEL of 7.5 ppm (1.13 mg/kg/day) was identified by the author.
Reproductive Effects
0 In a three-generation reproduction study, Goldenthai et al. (1978b)
administered RH 6201 (a formulation containing sodium acifluorfen) in
the diet to Charles River CD rats. During the course of the study,
the test compound was administered at various levels depending on the
age of the animals. The FI generation received dose levels of 2.9,
17.3 or 104 ppm during the first 2 weeks of the study, and 5, 30 and
180 ppm for the remaining weeks of the generation (study weeks 3 to
17) (Time-Weighted Average (TWA) dosage levels 4.8, 28.5 or 171.1 pm).
The ?2 and F$ generations received dosage levels of 180,
10 or 60 ppm during the first and second weeks of the generation;
312, 17.3 or 104 ppm during the third, fourth and fifth weeks of the
generation; and 540, 30 or 180 ppm for the remaining weeks of the
generation (TWA for ?2 generation 486.0, 27.0 or 162.0 ppm; TWA for
F3 generation 483.8, 26.7 or 161.3 ppm). The highest dietary TWA dose
tested in this study was 486 ppm of active ingredient. Assuming that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day, this
corresponds to a dose of 24.3 mg/kg/day (Lehman, 1959). No effects
related to compound administration were observed in parents or pups
in terms of general behavior, appearance or survival. Parental and
pup body weights and food consumption were similar to controls.
Fertility, gestation and viability indices were comparable for controls
and treated groups. There were no biologically meaningful teratogenic
effects in the second or third generation, based on mean number of
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Acifluorfen August, 1987
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viable fetuses, post-implantation losses, total implantations and
corpora lutea per dam, mean fetal body weight, number of fetal
anomalies and sex-ratio variations. No compound-related gross lesions
were noted in third-generation pups necropsied. Based on the infor-
mation presented, a NOAEL of 486 ppm (24.3 mg/kg/day) was identified.
This NOAEL represents the highest dose tested.
• In a two-generation reproduction study, Lochry et al. (1986) admini-
stered technical grade Tackle (sodium acifluorfen) of unspecified
purity to rats at levels of 0, 25, 500 and 2,500 ppm. The compound
was administered in the diet ad libitum to groups of 35 rats/sex/dose
beginning at 47 days of age and continuing until sacrifice. In addi-
tion, the compound was also administered to groups of 40 rats/sex/dose
from weaning until sacrifice. Reproductive paramaters, mortality,
body weight and a number of other end points were measured; in addition,
both gross and histopathological examinations were conducted. The
NOAEL for toxicity to both the parents and offspring was 25 ppm,
based on mortality and kidney lesions at higher doses. Assuming that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day, the NOAEL
of 25 ppm in this study corresponds to 1.25 mg/kg/day (Lehman, 1959).
Developmental Effects
• Lightkep et al. (1980) administered Tackle 2S (a formulation containing
22.4% sodium acifluorfen) by oral intubation at doses of 0, 3, 12 or
36 rag/kg/day to New Zealand White rabbits (16/dose) on days 6 to 29
of gestation. The authors indicated that the administered doses were
in terms of the active ingredient. At 36 mg/kg/day, there was a
slight (nonsignificant) inhibition of maternal body weight gain and a
marked (significant) inhibition of maternal food consumption. At this
dose level, there was also possible interference with implantation
and a slight decrease in average fetal body weight; neither of these
changes was statistically significant. No gross, soft-tissue or
skeletal malformations were observed in pups, fetuses or late resorp-
tions at any dose level. Based on the information presented in this
study, a NOAEL of 36 mg/kg/day was identified for maternal toxicity,
fetal toxicity and teratogenicity. This NOAEL represents the highest
dose tested.
0 Florek et al. (1981) administered Tackle 25 (a formulation containing
22.4% sodium acifluorfen) by gavpge at doses of 0, 20, 90 or 180
mg/kg/day to Sprague-Dawley rats (25/dose) on days 6 to 19 of gesta-
tion. The authors indicated that the administered doses were in
terms of active ingredient. At 180 mg/kg/day, dams gained signifi-
cantly less weight than controls. At 90 and 180 mg/kg/day, lower
average fetal body weight and significantly delayed ossification of
metacarpals and forepaw and hindpaw phalanges were noted. At 180
»9A9/day, there was delayed ossification of caudal vertebrae,
sternebrae and metatarsals. Additionally, at the highest dose level
there was a significantly increased incidence of slight dilation of
the lateral ventricle of the brain. The authors stated that the
fetal effects were indicative of delayed fetal development. Based
on the results of this study, a NOAEL of 90 mg/kg/day for maternal
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Acifluorfen August, 1987
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toxicity, a NOAEL of 20 mg/kg/day for fetotoxicity and a NOAEL of
180 mg/kg/day (the highest dose tested) for teratogenic effects were
identified.
0 Weatherholtz and Piccirillo (1979) administered RH 6201 (a formulation
containing 39.8% sodium acifluorfen) by gavage at doses of 0, 20, 60
or 180 mg/kg/day to New Zealand White rabbits on days 7 to 19 of
gestation. Maternal toxicity at 180 mg/kg/day included statistically
significant weight loss and mortality. At 180 mg/kg/day, there was
also evidence of fetal toxicity (mortality). Due to embryotoxicity
and maternal toxicity at 180 mg/kg/day, teratogenic evaluations could
not be performed at this dose level. At lower doses, no teratogenic
effects were observed. Based on the results of this study, NOAELs
of 60 mg/kg/day were identified for teratogenic effects, maternal
toxicity and fetal toxicity.
Mutagenieity
• Schreiner et al. (1980) tested Tackle 2S (purity unspecified) in an
Ames assay using Salmonella typhimurium strains TA 98, 100, 1535, 1537
and 1538. The test compound was not found to be mutagenic, with or
without metabolic activation, at concentrations up to 1.8 rag/plate.
0 Brusick (1976) tested RH 6201 (purity not specified) in a mutagenicity
assay using Saccharomyces cersvisiae strain 04 and £. typhimurium
strains TA 15?5, 1537, 1538, 98 and 100. The compound was not found
to be mutagenic, with or without metabolic activation, at concentrations
up to 500 ug/plate.
0 Putnam et al. (1981) tested Tackle 2S (purity not specified) in a
dominant lethal assay using Sprague-Dawley rats. The compound was
administered by gavage at doses of 0, 80, 360 or 800 mg/kg/day for
5 consecutive days. No detectable mutagenic activity, as defined by
induction of fetal death, was reported.
0 Myhr and NcKeon (1981) conducted a primary rat (Fischer 344) hepato-
cyte unscheduled DNA synthesis (UDS) assay using Tackle 2S (purity
not specified). The test compound did not induce a detectable level
of UDS over a concentration range of 0.10 to 25 ug/mL. Treatment of
hepatocytes with 50 ug/mL was almost completely lethal to the cells.
0 Schreiner et al. (1981) tested Tackle 2S (purity not specified) in a
bone marrow metaphase analysis using Sprague-Dawley rats. The animals
were given the test compound by intubation at doses of 0, 0.37, 1.11
or 1.87 g/kg/day for 5 days. The test compound did not significantly
increase clas*:ogenic events in the bone marrow cells.
0 Schreiner et al. (1980) tested Tackle 2S (purity not specified) in
a murine lymphoma assay. The compound was tested without metabolic
activation at 0.11 to 1.7 ug/mL, and with metabolic activation at
0.08 to 0.56 ug/mL. No detectable mutagenic activity was detected
either with or without activation.
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Acifluorfen August, 1987
-12-
0 Jagannath (1981) tested Tackle 2S (29.7% purity) in a mitotic recombi-
nation assay using Saccharomyces cerevisiae strain D5. The compound
was tested at 0, 2.5, 5.0 or 7.5 uL/plate without metabolic activation,
and at 7.5, 10.0 and 25.0 uL/plate with metabolic activation. In the
absence of metabolic activation, the compound induced a dose-related
increase in recombination events (significant at 5.0 uL/plate). With
metabolic activation, a dose of 10.0 uL/plate induced an increase in
recombination events. The authors reported that very few survivors
were observed at 25.0 uL/plate.
• Bowman et al. (1981) tested Tackle 2S (purity not specified) in
mutagenicity assays using Drosophila melanogaster. Assays included
the Biothorax test of Lewis, a dominant lethal assay, an assay for
Y-chromosome loss, and a White Ivory reversion assay. In all cases,
the test compound was tested at concentrations of 15 mg/mL. Results
of these assays were negative for somatic reversions of White Ivory
and the Biothorax test of Lewis and positive for Y-chromosome loss
and dominant lethal mutations.
Carcinogenicity
0 Barnett et al. (1982b) administered Tackle 2S (a formulation
containing 19.1 to 25.6% sodium acifluorfen) to Fischer 344 rats
(73/sex/dose) for one year at dietary levels of 0, 25, 150, 500,
2,500 or 5,000 ppm. Assuming that these dietary levels reflect the
concentrations of the test compound and not the active ingredient,
corresponding levels of sodium acifluorfen are 0, 6.4, 38.4, 128.0,
640.0 or 1,280 ppm (based on 25.6% active ingredient in the test
compound). Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day, these doses correspond to approximately to 0, 0.3,
1.9, 6.4, 32.0 or 64.0 mg/kg/day (Lehman, 1959). Histopathological
examinations revealed no evidence of carcinogenicity at any dose level.
0 Barnett et al. (1982a) administered Tackle* (a formulation containing
24% sodium acifluorfen) to B6C3Fi mice (60/sex/dose) for 18 months at
dietary concentrations of 0, 625, 1,250 or 2,500 ppm. (The high dose
was reported to be the maximum tolerated dose.) The authors reported
that the dietary levels corresponded to average compound intake values
of 0, 118.96, 258.73 or 655.15 mg/kg/day for males, and 0, 142.50,
312.65 or 710.54 mg/kg/day for females. Assuming that these levels
reflect test compound and not active ingredient intake, corresponding
levels of sodium acifluorfen intake are 0, 28.55, 62.10 or 157.24
mg/kg/day for males and 0, 34.20, 75.04 or 170.53 mg/kg/day for
females. An obvious dose-related depression of body weight was
reported for all doses. Beginning in week 52 of the study and
continuing with increasing frequency was the appearance of palpable
abdominal masses. Gross necropsy revealed a dose-related increase in
liver masses in both sexes. Histopathological examinations conducted
at the 52-week interval revealed that the livers of six animals per
sex of high-dose animals (157.24 mg/kg/day for males; 170.53 mg/kg/day
for females) showed evidence of acidophilic cells. Males receiving
this dose displayed a statistically significant increase in the
frequency of hepatocellular adenomas. After 18 months of treatment.
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Acifluorfen August, 1987
-13-
all 40 high-dose males and 27/47 high-dose females sacrificed were
found to have a single benign hepatoma, multiple benign hepatomas or
hepatocellular carcinomas. In the males, the incidence of single
benign hepatoma and hepatocellular carcinomas was statistically
significant. In the females, the incidence of single hepatomas was
statistically significant.
0 Goldenthal (1979) administered RH 6201 (a formulation containing 39.4
to 40.5% sodium acifluorfen) to Charles River CD-I mice (80/sex/dose)
for two years in the diet at concentrations that provided dose
levels of 0, 1.25, 7.5 or 45.0 ppm of the active ingredient. After
16 weeks of administration, the 1.25 ppm dose was increased to 270 ppm.
Assuming that 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day,
these doses correspond to approximately 0, 0.19 (increased to 40.5),
1.13 or 6.8 mg/kg/day (Lehman, 1959). Two control groups were used
in this study. One group received acetone in the diet (control 1)
and the other received water in the diet (control 2). In males
receiving the highest dose there was a nonstatistically significant
increase in the incidence of nodular hepatocellular proliferation and
hepatocellular carcinoma, which indicated to the authors that these
changes were dose-related.
0 Coleman et al. (1978) administered RH 6201 (a formulation containing
39.8% sodium acifluorfen) to Charles River Outbred albino CD COBS
rats (approximately 75/sex/dose) for 2 years at changing dietary
concentrations. Mean sodium acifluorfen intake values over the
course of the study were 0, 1.25, 7.54 and 17.56 mg/kg/day for males
and 0, 1.64, 9.84 and 25.03 mg/kg/day for females.
0 Acifluorfen is structurally similar to nitrofen [2,4-dichloro-1-(4-
nitrophenoxy) benzene; CAS No. 1836-75-7]. Nitrofen has been shown
to be carcinogenic in Osborne-Mendel rats and B6C3F1 mice (NCI, 1978,
1979; both as cited in NAS, 1985).
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 th«s following formula:
HA - (NOAEL or LOAEL) x (BW) = /L ( /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet-Level
in mg/kg bw/day.
BW s assumed body weight of a child (10 kg) or
an adult (70 kg).
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Acifluorfen August, 1987
-14-
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
No data were found in the available literature that were suitable for
determination of a One-day HA value for acifluorfen. It is therefore recom-
mended that the Ten-day HA value for a 10-kg child (2 mg/L, calculated below)
be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The study by Florek et al. (1981) has been selected to serve as the
basis for determination of the Ten-day HA for a 10-kg child. In this study,
Tackle 2S (a formulation containing 22.4% sodium acifluorfen) was administered
by gavage at doses of 0, 20, 90 or 180 mg/kg/day to Sprague-Dawley rats
(25/dose) on days 6 to 19 of gestation. The authors indicated that the
administered doses were in terms of active ingredient. At 180 mg/kg/day,
dams reportedly gained significantly less weight than controls. At 90 and
180 mg/kg/day, lower average fetal body weight and significantly delayed
ossification of metacarpals and forepaw and hindpaw phalanges were noted. At
180 mg/kg/day, there was* delayed ossification of caudal vertebrae, sternebrae
and metatarsals. Additionally, at the highest dose level there was a signifi-
cantly increased incidence of slight dilation of the lateral ventricle of the
brain. The authors stated that the fetal effects were indicative of delayed
fetal development. No effects on implantations, litter size, fetal viability,
resorption or fetal sex ratio were reported. Based on the results of this
study, a NOAEL of 20 mg/kg/day for fetotoxicity was identified.
The Ten-day HA for the 10-kg child is calculated as follows:
Ten-day HA = (20 mg/kg/day) (10 kg) . 2 mg/L (2,000 ug/L)
(100) (1 L/day)
where:
20 mg/kg/day = NOAEL, based on absence of fetal toxicity in rats
exposed to aciflucrfen via gavage during days 6 to 19
of gestation.
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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Acifluorfen August, 1987
-15-
Longer-term Health Advisory
The study by Barnett (1982) had been selected to serve as the basis
for determination of the Longer-term HA. In this study, the NOAEL was
5.6 mg/kg/day based on an increase in the size of the liver in male rats.
However, a lower NOAEL, 1.25 mg/kg/day, was recently identified in a two-
generation rat reproduction study by Lochry et al. (1986). Since the NOAEL
in the Lochry et al. (1986) study is numerically identical to the value on
which the Lifetime HA is based and since a two-generation reproduction study
is suitable for calculating a Longer-term HA, it was determined that it is
appropriate to base the Longer-term HA on the Lifetime HA.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (1.25 mg/kg/day) (10 kg) = 0.13 mg/L (130 ug/L)
* (100) (1 L/day)
where:
1.25 mg/kg/day « NOAEL (see Lifetime Health Advisory below).
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 = (1.25 mg/kg/day) (70 kg) = 0.44 mg/L (440 ug/L)
(100) (2 L/day}
where:
1.25 mg/kg/day = NOAEL (see Lifetime Health Advisory below).
70 kg = assumed body weight of an adult.
100 a uncertainty factor, chosen in accordance with NAS/OCW
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
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Acifluorfen August. 1987
-16-
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
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.
A 2-year Charles River CD-I mouse dietary study by Goldenthal (1979) was
originally selected to serve as the basis for determination of the Lifetime
HA for acifluorfen. In this study, a NOAEL of 1.13 mg/kg/day was identified.
More recently, however, a two-generation rat reproduction study by Lochry et
al. (1986) was identified that strongly supports the results of the Goldenthal
(1979) study and identifies a NOAEL of 1.25 mg/kg/day.
Using the NOAEL of 1.25 mg/kg/day, the Lifetime HA for acifluorfen is
calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (1.25 mg/kg/day) . 0.0125 mg/kg/day
(100)
where:
1.25 mg/kg/day = NOAEL, based on the absence of mortality and kidney
lesions 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 (DWEL)
DWEL = (0.0125 mq/kq/day) (70 kg) -, Q.437 mg/L (437 ug/L)
(2 L/day)
where:
0.0125 mg/kg/day - RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Acifluorfen August, 1987
-17-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0*437 mg/L) (20%) a 0.009 mg/L (9 ug/L)
(10)
where:
0.437 mg/L = DHEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 Four studies that evaluated the carcinogenic potential of sodium
acifluorfen were identified. The results of one of these studies
(Barnett et al.f 1982a) indicated that sodium acifluorfen was car-
cinogenic in B6C3Fj mice. The results of the other three studies
(Goldenthai, 1979; Coleman et al., 1978; Barnett et al., 1982b)
provided no evidence of carcinogenicity in two strains of rats and
one strain of mice. However, due to deficiencies in the three negative
studies, the results of these studies are not sufficient to contradict
the results of the positive study. Each of these studies is discussed
briefly below.
- In the positive study (Barnett et al., 1982a), B6C3F1 mice received
sodium acifluorfen in the diet for 18 months. At the end of the
study, the high-dose (157.24 mg/kg/day) male mice displayed a
statistically significant increase in the incidence of single
benign hepatomas and hepatocellular carcinomas. A statistically
significant increase in the incidence of single hepatomas was
observed in high-dose (170.53 mg/kg/day) females.
- In one of the studies with negative results (Goldenthal, 1979)
Charles River CD-1 mice received sodium acifluorfen in the diet for
two years at doses of 0, 0.19 (increased to 40.5 after 16 weeks),
1.13 or 6.8 mg/kg/day. Although no evidence of carcinogenicity was
observed in this study, the dose levels tested were considerably
lower than the level that produced positive results in the 18-month
mouse feeding study (157.24 mg/kg/day) (Barnett et al., 1982a).
- In the second study with negative results (Coleman et al., 1978),
Charles River outbred albino CD COBS rats received sodium acifluorfen
for two years at dietary levels up to 25.03 mg/kg/day (females) or
17.56 mg/kg/day (males). Although it is difficult to make cross-
species comparisons, these levels are considerably lower than the
level that produced positive results in the 18-month mouse feeding
study (157.24 mg/kg/day) (Barnett et al., 1982a). In addition,
no adverse effects were observed at any dose level used in this
study, indicating that the maximum tolerated dose was not used.
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Acifluorfen August, 1987
-18-
- In the third study with negative results (Barnett et al., 1982b),
Fischer 344 rats received sodium acifluorfen for 1 year at dietary
concentrations of 0, 0.3, 1.9, 6.4, 32.0 or 64.0 mg/kg/day.
Although the results of this study were negative, a study duration
of 1 year is not sufficient for assessing carcinogenic potential.
0 Acifluorfen is structurally similar to nitrofen [2,4-dichloro-1-(4-
nitrophenoxy) benzene; CAS No. 1836-75-7]. Nitrofen has been shown
to be carcinogenic in Osborne-Mendel rats and B6C3F<| mice [NCI, 1978,
1979; both as cited in MAS (1985)]. Although data on nitrofen cannot
be used to conclude that sodium acifluorfen is carcinogenic, these data
do, to some extent, support the positive results of Barnett et al.
(1982a).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of acifluorfen.
* Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), acifluorfen is classified in
Group C: possible human carcinogen. Category C is for substances
with limited evidence of carcinogenicity in animals in the absence
of human data.
VI. OTHER CRITERIA, GUIDANCE AMD STANDARDS
0 The U.S. EPA has established residue tolerances for sodium acifluorfen
in or on raw agricultural commodities that range from 0.01 to 0.1 ppm
(CFR, 1985). .
0 The EPA RfD Workgroup has concluded that an RfD of 0.013 mg/kg/day
is appropriate for acifluorfen.
VII. ANALYTICAL METHODS
0 Analysis of acifluorfen is by a gas chromatographic (GC) method
applicable to the determination of certain chlorinated acid pesticides
in water samples (U.S. EPA, 1986b). In this method, approximately
1 liter of sample is acidified. The compounds are extracted with
ethyl ether using a separator/ funnel. The derivatives are hydrolyzed
with potassium hydroxide, and extraneous organic material is removed
by a solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane as the derivatizing
agent. Excess reagent is removed, and the esters are determined by
electron capture GC. The method detection limit has not been deter-
mined for this compound, but it is estimated that the detection
limits for analytes included in this method are in the range of 0.5
to 2 ug/L.
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Acifluorfen August, 1987
-19-
VIII. TREATMENT TECHNOLOGIES
0 Reverse osmosis (RO) is a promising treatment method for pesticide-
contaminated water. As a general rule, organic compounds with
molecular weights greater than 100 are candidates for removal by RO.
Larson et al. (1980) report 99% removal efficiency of chlorinated
pesticides by a thin-film composite polyamide membrane operating at a
maximum pressure of 1,000 psi and at a maximum temperature of 113°F.
More operational data are required, however, to specifically determine
the effectiveness and feasibility of applying RO for the removal of
acifluorfen from water. Also, membrane adsorption must be considered
when evaluating RO performance in the treatment of acifluorfen-
contaminated drinking water supplies.
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Acifluorfen August, 1987
-20-
IX. REFERENCES
Barnett, J.* 1982. Evaluation of ninety-day subchronic toxicity of Tackle*
in Fischer 344 rats. GSRI Project No. 413-971-40. Rhone-Poulenc
Agrochemie No. 372-80. Unpublished study. MRID 0122730.
Barnett, J., L. Jenkins and R. Parent.* 1982a. Evaluation of the potential
oncogenic and toxicological effects of long-term dietary administration
of Tackle* to B6C3FI mice. GSRI Project No. 413-984-41. Final Report.
Unpublished Study. MRID 00122732.
Barnett, J., L. Jenkins and R. Parent.* 1982b. Evaluation of the potential
oncogenic and toxicological effects of long-term dietary administration
of Tackle* to Fischer 344 rats: GSRI Project No. 413-985-41. Interim
report. Unpublished study. MRID 00122735.
Bowman, J., C. Mackerer, S. Bowman, D.C. Jessup, R.C. Geil and B.W. Benson.*
1981. Drosophila mutagenicity assays of Mobil Chemical Company compound
MC 10109 (MRI 533). Study No. 009-275-533-9. Unpublished study.
MRID 00122737.
Brusick, D.* 1976. Mutagenicity evaluation of RH-6201. LBI Project No. 2547.
Unpublished study. MRID 00083057.
CFR. 1985. Code of Federal Regulations. July 1, 1985. 40 CFR 180.383.
p. 336.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD.
Coleman, M.E., T.E. Murchison, P.S. Sahota et al.* 1978. Three and twenty-four
month oral safety evaluation study of RH-6201 in rats. DRC 5800. Final
Report. Unpublished study. MRID 00087478.
Florek, M., M. Christian, G. Christian and E.M. Johnson.* 1981. Terato-
genicity study of TACU 06238001 in pregnant rats. Argus Project 113-004.
Unpublished study. MRID 00122743.
Goldenthal, E.I., D.C. Jessup, R.G. Geil and B.W. Benson.* 1978a. Two week
range finding study in mice: 285-016. Unpublished study. MRID 00080568.
Goldenthal, E.I., D.C. Jessup and D. Rodwell.* 1978b. Three generation
reproduction study in rats: RH-6201, 285-014a. Unpublished study.
MRID 00107486.
Goldenthal, E.I., D.C. Jessup, R.G. Geil and B.W. Benson.* 1979. Lifetime
dietary feeding study in mice: 285-013a. Unpublished study.
MRID 00082897.
Harris, J.C., G. Cruzan and w.R. Brown.* 1978. Three month subchronic rat
study. RH-6201. TRD-76P-30. Unpublished study. MRID 00080569.
Jagannath, D.* 1981. Mutagenicity of 06238001 lot LCM 266830-7 in the mitotic
recombination assay with the yeast strain D5. Genetics Assay No. 5374.
Final Report. Unpublished study. MRID 00122740.
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Acifluorfen August, 1987
-21-
Larson, R.E., P.S. Cartwright, P.K. Eriksson and R.J. Petersen. 1982.
Applications of the FT-30 reverse osmosis membrane in metal finishing
operations. Paper presented at Tokohama, Japan.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Lightkep, G., G. Christian et al.* 1980. Teratogenic potential of TACU
06238001 in New Zealand white rabbits (Segment II Evaluation). Argus
Project 113-003. unpublished study. MRID 00122744.
Lochry, E.A., Hoberman, A.M. and Christian, M.S.* 1986. Two-generation Rat
Reproduction Study, Argus Research Laboratories, Inc. Study Mo.218-002.
Madison, P., R. Becci and R. Parent.* 1981. Guinea pig sensitization study.
Buhler test for Mobil Corporation. Tackle 2S. FDRL Study No. 6738.
Unpublished study. MRID 00122729.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Mobil Environmental and Health Science Laboratory.* 1981. A study of the oral
toxicity of Tackle 2S in the dog. Mobil Study No. 1091-80. Six-month
status report. Unpublished study. MRID 00122733.
Myhr, B., and M. McKeon.* 1981. Evaluation of 06238001 in the primary rat
hepatocyte unscheduled DNA synthesis assay. MEHSL Study 1022-80. Final
Report. Unpublished study. MRID 00122742.
NAS. 1985. National Academy of Sciences. Drinking Water and Health. Vol. 6.
Chapter 9: Toxicity of Selected Contaminants. Washington, DC. National
Academy Press.
NCI. 1978. National Cancer Institute. Biloassay of nitrofen for possible
carcinogenicity. Technical Report Series No. 26. DHEW Publication No.
(NIH) 78-826. U.S. Department of Health, Education and Welfare.
Washington, DC. 101 pp. Cited in: NAS. 1985. National Academy of
Sciences. Drinking Water and Health. Vol. 6. Chapter 9: Toxicity of
Selected Contaminants. Washington, DC. National Academy Press.
NCI. 1979. National Cancer Institute. Bioassay of nitrofen for possible
carcinogenicity. Technical Report Series No. 184. DHEW Publication No.
(NIH) 79-1740. U.S. Department of Health, Education and Welfare.
Washington, DC. 57 pp. Cited in: NAS. 1985. National Academy of
Sciences. Drinking Water and Health. Vol. 6. Chapter 9: Toxicity of
Selected Contaminants. Washington, DC. National Academy Press.
Piccirillo, V.J., and T.L. Robbins.* 1976. Four week oral range finding
study in rats. RII-6201. Unpublished study. MRID 00071892.
Putnam, D., L. Schechtman and W. Moore.* 1981. Activity of T1689 in the
dominant lethal assay in rodents. MA Project No. T1689.116. Final Report.
Unpublished study. MRID 00122738.
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Acifluorfen August, 1987
-22-
Schreiner, C.A., M.A. McKenzie and M.A. Mehloan.* 1980. An Ames Salmonella/
mammalian nicrosome mutagenesis assay for determination of potential
mutagenicity of Tackle 25 MCZ0978. Study No. 511-80. Unpublished
study. HRID 00061622.
Schreiner, C., M. Skinner and M. Mehlraan.* 1981. Metaphase analysis of rat
bone marrow cells treated in vivo with Tackle 2S. Study No. 1041-80.
Unpublished study. MRID 00122741.
Spicer, E., L. Griggs, F. Marroquin, N.D. Jefferson and M. Blair.* 1983. Two
year dietary toxicity study in dogs. (Tackle*) 450-0395. Unpublished
study. MRID 00131162.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #3 -
Determination of Chlorinated Acids in Ground Hater by GC/ECD, January
1986 draft. Available from U.S. EPA's Environmental Monitoring and
Support Laboratory. Cincinnati, OH.
Weatherholtz, W., S. Moore and G. Wolfe.* 1979a. Eye irritation study in
monkeys. Blazer 2S. Project No. 417-396. Final Report. Unpublished
study. MRID 00140887.
Weatherholtz, W., K. Peterson, M. Koka and R.W. Kapp.* 1979b. Four-week
repeated dermal study in rabbits. RH-6201 formulations. Project No.
417-386. Final Report. Unpublished study. MRID 00140889.
Weatherholtz, W., and V. Piccirillo.* 1979. Teratology study in rabbits
(RH-6201 LC). Final Report. Project No. 417-374. Unpublished study.
MRID 00107485.
Whit taker Corporation.* No date, a. Acute oral 1*050 rats. Study No. 410-0249.
Unpublished study. MRID 00061625.
Whittaker Corporation.* No date, b. Primary dermal irritation — rabbit:
Study No. 410-0286. Unpublished study. MRID 00061629.
Whittaker Corporation.* No date, c. Primary eye irritation — rabbits.
Study No. 410-C252. Unpublished study. MRID 00061628.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck Index, loth ed. Rahway, NJ: Merck and Co., Inc.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
AMETRYN
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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 accuratel} 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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II- GENERAL INFORMATION AND PROPERTIES
CAS No. 834-12-8
Structural Formula
SCH,
H
2-(Ethylamino)-4-(isopropylamino)-6-(methylthio)-s-triazin9
Synonyms
N-ethyl-N'-(1-methylet*yl)-6-(methylthio)-1,3,5-triazine-2,4-diamine;
Ametrex; Ametryne; Cemerin; Crisatine; Evik SOW; Gesapax {WSSA, 1983-
Meister, 1983). '
Uses
0 A selective herbicide for control of broadleaf and grass weeds in
pineapple, sugarcane, bananas and plantains. Also used as a post-
directed spray in corn, as a potato vine dessicant and for total
vegetation control (WSSA, 1983).
Properties (WSSA, 1983)
Chemical Formula C9H17N5S
Molecular Weight 227.35
Physical State Colorless crystals
Boiling Point
Melting Point 84 to 85°C
Density
Vapor Pressure 8.4 x 10-? mm Hg
Specific Gravity
Water Solubility 135 mg/L
Log Octanol/Water Partition -1.72 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Ametryn has been found in 3 of 1,246 surface water samples analyzed
and in 27 of 653 ground water samples (STORET, 1987). Samples were
collected at 211 surface water locations and 544 ground water
locations, and ametryn was found in 6 states. The 85th percentile of
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all nonzero samples was 0.1 ug/L in surface water and 210 ug/L in
ground water sources. The maximum concentration found was 0.1 ug/L
in surface water and 450 ug/L in ground water.
Environmental Fate
0 In aqueous solutions, ametryn is stable to natural sunlight, with a
half-life of greater than 1 week. When exposed to artificial light
for 6 hours, 75% of applied ametryn remained. One photolysis product
was identified as 2-ethylamino-4-hydroxy-6-isopropylaminos-triazine
(Registrant CBI data).
0 Ametryn is stable to photolysis on soil (Registrant CBI data).
0 Soil metabolism of ametryn, under aerobic conditions, proceeds with
a half-life of greater than 2 to 3 weeks. Metabolic products include
2-amino-4-isopropylaraino-6-methylthio-s-triazine, 2-amino-4-ethylamino-
6-methylthio-s-triazine and 2,4-diamino-6-methylthio-triazine. Under
anaerobic conditions the rate of metabolism decreases (t1/2 = 122 days)
(Registrant CBI data).
0 Under sterile conditions ametryn does not degrade appreciably. There-
fore, microbial degradation is a major degradation pathway (Registrant
CBI data).
0 Neither ametryn nor its hydroxy metabolite leach past 0 to 6 in. depth
with normal rainfall. However, since both compounds are persistent
they may leach under exaggerated rainfall or flood and furrow irrigation.
This behavior is seen with other triazines (Registrant CBI data).
0 Ametryn's Freundlich soil-water partition coeficient values, Kd, range
from 0.6 in sands to 5.0 in silty clay soils. Specifically, the Kd
for a sandy loam is 4.8, and for 2 silty loams, 3.8 and 2.8,
respectively.
0 In the laboratory, Ametryn has a half-life of 36 days. In the field,
Ametryn degraded with a half-life of 125 to 250 days (Registrant CBI
data).
III. PHARMACOKINETICS
Absorption
0 Oliver et al. (1969) administered He-labeled ametryn orally to
Sprague-Dawley rats. Investigators stated that ametryn was admini-
stered by stomach tube to animals at dosage levels from 1 to 4 mg
per animal. When the label was in the ring, 32.1% was excreted in
the feces, indicating that over 70% had been absorbed. When the
label was in the ethyl or isopropyl side chains, only 2 to 5% was
excreted in the feces.
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Distribution
0 Oliver et al. (1969) administered ring-labeled ametryn orally to
male and female Sprague-Dawley rats and measured distribution of
label in tissues at 6, 48 and 72 hours after dosing. Tissue distri-
bution at 6 hours was greatest in kidney, followed by liver, spleen,
blood, lung, fat, carcass, brain, and muscle. Blood levels remained
relatively constant for 72 hours after dosing, while all other tissue
levels dropped rapidly to <0.1% of dose per gram of tissue.
Metabolism
0 Oliver et al. (1969) administered 14C-labeled ametryn orally to
groups of six male and six female Sprague-Dawley rats. When the
label was in the isopropyl side chain, 41.9% of the label appeared as
C02. When the label was in the ethyl side chain, 18.1% of the label
appeared as CO2. This indicated that the side chains were extensively
metabolized. When the ring was uniformly labeled with carbon-14 and
the compound fed orally to rats, 58% was excreted in the urine but it
was not determined whether excretion of the original compound or
metabolites had occurred.
Excretion
0 Oliver et al. (1969) studied the excretion of ametryn utilizing
uniformly labeled compound with 1*C-ametryn in the ring or in the
ethyl or isopropyl side chains. Forty-eight hours after oral dosing
of six male and six female Sprague-Dawley rats, 57.6% of the ring
labeled activity had been excreted in the urine with 32.1% excreted
in the feces (total 89.7% of dose). When the fed compound was labeled
in the side chains, however, much of the 14C was excreted in expired
air as carbon dioxide. When fed compound labeled in the isopropyl
side chain, rats excreted 41.9% of the label in expired air 20% in
the urine, 2% in the feces and 7% remained in the carcass (total
70.9%) at 48 hours. When the ethyl side chain contained the label,
18.1% of the label was excreted as carbon dioxide, 45% in the urine,
5% in the feces and 9% remained in the carcass (total 77.1% of dose).
After 72 hours, total recovery was approximately 88% for both of the
side-chain labeled compounds.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of ametryn in humans.
Animals
Short-term Exposure
0 The following acute oral 1.050 values for ametryn in rats were
reported: Charles River CD rats, 1,207 rag/kg (males), 1,453 mg/kg
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{females) (Grunfeld, 1981); mixed male and female rats (strain not
specified), 1,750 rag/kg (Stenger and Planta, 1961a); male and female
Wistar rats, 1,750mg/kg (Consultox Laboratories Limited, 1974).
0 Piccirillo (1977) reported the results of a 28-day feeding study in
male and female mice. Animals were 5 weeks of age and weighed
21 to 28 g at the beginning of the study. Animals (five/sex/dose)
were fed diets containing 0, 100, 300, 600, 1,000, 3,000, 10,000 or
30,000 ppm of ametryn (technical). Based on the assumption that 1 ppm
in the diet of mice is equivalent to 0.15 mg/kg/day (Lehman, 1959),
these doses correspond to 0, 15, 45, 90, 150, 450, 1,500 or 4,500
mg/kg/day. At 30,000 ppm in the diet, all animals died within 2
weeks. At 10,000 ppm, 3 of the 10 died within 2 weeks. No other
deaths occurred at'any other dose level. Clinical signs in the two
highest dose groups included hunched appearance, stained fur and
labored respiration. At the 3,000-ppra dose level, only 1 of the
10 animals showed clinical signs of toxicity. Body weight gain was
comparable in all survivors by the end of week 4. Gross pathology in
animals that died showed a dark-red mucosal lining of the gastro-
intestinal tract and ulcerated areas of the gastric mucosa. There
was no histopathological examination of tissues in this study.
0 Stenger and Planta (1961b) reported a 28-day study of the toxicity
of ametryn in rats. Dose levels of 100, 250 or 500 mg/kg/day were
administered 6 days/week by gavage to groups of five male and five
female rats. The study indicated that there was a control group but
no data were given. At the 500-mg/kg/day dose level, animals became
emaciated, weight gain was limited and 7 of 10 rats died. Histo-
pathological examination of the animals that died indicated severe
vascular congestion, centrilobular liver necrosis and fatty degeneration
of individual liver cells. At 250 mg/kg/day, 1 of 10 rats died
during the study and there was depressed growth rate in the survivors.
Histological examination of liver, kidney, spleen, pancreas, heart,
lung, intestine and gonads showed no major degenerative changes. No
effects were reported in animals administered 100 mg/kg/day, which
was identified as the No-Observed-Adverse-Effect-Level (NOAEL) in
this study.
0 Ceglowski et al. (1979) administered single oral doses of 88 or 880
mg/k? of ametryn to mice 5 days before, on the day of or 2 days after
immunization with sheep erythrocytes fpurity not specified). All
mice receiving the highest dose (880 mg/kg) of ametryn had significant
depression of splenic plaque-forming cell numbers when assayed 4 days
later. Animals receiving the low dose showed no effect. Similarly,
animals receiving 88 mg/kg for 8 or 28 consecutive days prior to
immunization exhibited no significant reduction in antibody plaque
formation.
Dermal/Ocular Effects
0 Two of six rabbits showed mild skin irritation when ametryn was left
in contact with intact or abraded skin (500 mg/2.5 cm2) for 24 hours
(Sachsse and Ullmann, 1977).
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0 In a sensitization study with Perbright White guinea pigs (Sachsse
and Ullmann, 1977), 10 male and 10 female guinea pigs weighing 400
to 450 g received 10 daily intracutaneous 0.1-mL injections of 0.1%
aoetryn in polyethylene glycol:saline (70:30). Fourteen days after
the last dose, animals were challenged by an occlusive dermal applica-
tion of ametryn or by an intradermal challenge. Animals showed no
sensitization reaction following the dermal application of the challenge
dose but there was a positive response after the intradermal challenge.
0 Kopp (1975) found that ametryn (technical grade) placed in the eyes
of rabbits produced slight conjunctival redness at 24 hours. This
cleared completely within 72 hours.
0 Sachsse and Bathe (1976) applied 2,150 mg/kg or 3,170 mg/kg ametryn
in suspension to the shaved backs of five male and five female rats
weighing 180 to 200 g. The occlusive covering was removed at
24 hours, the skin was washed and animals were observed for 14 days.
There was no local irritation or adverse reaction, and at necropsy
there were no gross changes in the skin. The acute dermal LD50 i°
male and female rats was reported to be >3,170 rag/kg.
0 Ametryn (2,000 mg/kg) was applied daily to the skin of five male and
five female rats weighing approximately 200 g (Consultox Laboratories
Limited, 1974). After 14 days of treatment, no deaths had occurred
and no other effects were reported. The 14-day dermal LDso was re-
ported to be >2,000 mg/kg/day.
Long-term Exposure
0 Domenjoz (1961) administered ametryn in water via stomach tube
6 days/week for 90 days to Neyer-Arendt rats (12/sex/dose). The
initial material was 50% ametryn in a powder vehicle. Two dose
levels of the material (20 or 200 mg/kg/day) provided dose levels of
ametryn of 10 or 100 mg/kg/day. Two control groups were included;
one group received water only and the other received the powder
vehicle only suspended in water. Over the 90-day period, all animals
gained weight at comparable rates and there was no visible effect on
appearance or behavior. One control rat and one rat in the 100-mg/kg
dosage group died. This death was not considered compound-related.
At the 90-day necropsy, organ-to-body weight ratios were comparable
to controls. Liver, Vidney, spleen, heart, gonads, small intestine,
colon, stomach, thyroid and lung were microscopically examined. The
Lowest-Observed-Adverse-Effect-Level (LOAEL) was associated with fatty
degeneration of the liver. Based on this study, a LOAEL of 100 mg/kg/day
(the highest dose tested) was identified. All tissues were comparable
to controls at the lowest dose (10 mg/kg/day), which was identified
as the NOAEL.
Reproductive Effects
0 No information was found in the available literature on the reproduc-
tive effects of ametryn.
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Developmental Effects
0 No information was found in the available literature on the developmental
effects of ametryn.
Mutagenicity
0 Anderson et al. (1972) reported that ametryn was not mutagenic in
eight strains of Salmonella typhimurium. No metabolic activating
system was utilized.
8 Simmons and Poole (1977) also reported that ametryn was not mutagenic
in five strains of Salmonella typhimurium (TA 98, 100, 1535, 1537 and
1538), with or without metabolic activation provided by an S9 fraction
from rats pretreated with Aroclor 1254.
0 Shirasu et al. (1976) reported ametryn was not mutagenic in the
rec-assay system utilizing two strains of Bacillus subtilis, in
reversion assays utilizing auxotrophic strains of Escherichia coli
(WP2) and in £. typhimurium strains TA 1535, 1536, 1537 and
1538 (without metabolic activation).
Carcinogenicity
0 No information was found in the available literature on the carcinogenic
effects of ametryn.
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 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 child
(1 L/day) or an adult (2 L/day).
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One-day Health Advisory
No data were found in the available literature that were suitable for
determination a One-day HA value for ametryn. It is, therefore, recommended
that the Ten-day HA value for the 10-kg child (8.6 mg/L, calculated below) be
used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The study by Stenger and Planta (1961b) has been selected to serve as
the basis for determination of the Ten-day HA value for the 10-kg child.
This study identified a NOAEL of 100 mg/kg/day, based on normal weight gain
and absence of histological evidence of injury in rats following 28 days of
exposure by gavage. The study also identified a LOAEL of 250 mg/kg/day,
based on reduced body weight gain, although no major histological changes
were noted. One death occurred in the 250-mg/kg/day group, but it could not
be determined if this was compound-related. The NOAEL identified in this
study (100 mg/kg/day) is supported by the 28-day feeding study in rats by
Piccirillo (1977), which identifed a NOAEL of 150 mg/kg/day and a LOAEL of
450 mg/kg/day, and by the study of Ceglowski et al. (1979), which identified
a NOAEL of 88 mg/kg/day and a LOAEL of 880 mg/kg/day.
Using the NOAEL of 100 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA - (100 mg/kg/day) (10 kg) (6/7) = 8.6 mg/L (8,600 ug/L)
(100) (1 L/day)
where:
100 mg/kg/day = NOAEL, based on absence of effects on weight gain
or histology in rats dosed by gavage for 28 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from a study in animals.
6/7 » conversion from 6 to 7 days.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The 90-day oral dosing study in rats by Domenjoz (1961) has been selected
to serve as the basis for determination of the Longer-term HA. At two dose
levels (10 or 100 mg/kg/day), no deaths were reported and no other effects
were noted during the 90-day period. Terminal necropsy findings and histo-
logical examination of tissues from treated animals were comparable to
controls. At the highest dose tested, there was fatty degeneration in the
livers examined. Based on these data, a NOAEL of 10 mg/kg/day (the lowest
dose tested) was identified.
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The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (1° "^Aa/day) (10 kg) (6/7) = 0.86 /L (860 ug/L)
y (100) (1 L/day)
where:
10 mg/kg/day = NOAEL, based on the absence of histological evidence
of toxicity in rats exposed to ametryn via gavage for
90 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from a study in animals.
6/7 = conversion from 6 to 7 days of exposure.
1 L/day = assumed daily water consumption of a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (10 mg/kg/day) (70 kg) (6/7) = 3 mg/L (3f0oO ug/L)
(100) (2 L/day)
where:
10 mgA9/day = NOAEL, based on the absence of histological evidence
of toxicity in rats exposed to ametryn via gavage for
90 days.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from a study in animals.
6/7 = conversion from 6 to 7 days of exposure.
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
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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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
Compound-specific, chronic ingestion data for ametryn are not available
at this time. In the absence of appropriate ingestion studies', the Lifetime
HA for ametryn is derived from the subchronic study in rats reported by
Domenjoz (1961). At two dose levels (10 or 100 mg/kg/day), no deaths were
reported during the 90-day period. Terminal necropsy findings and histological
examination of tissues from treated animals were comparable to controls at
the lowest dose level of 10 mg/kg/day. This study identified a NOAEL of 10
mg/kg/day (the lowest dose tested).
Using the NOAEL of 10 mg/kg/day, the Lifetime HA for ametryn is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (10 mg/kg/day) (6/7) = Oo0086 mg/kg/day
(1,000)
where:
10 mg/kg/day = NOAEL, based on absence of histological evidence of
toxicity in rats exposed to ametryn via gavage for
90 days.
6/7 = conversion from 6 to 7 days exposure.
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 - (0.0086 mq/kg/day) (70 kg) =0.3 mg/L (300 ug/L)
(2 L/day)
where:
0.0086 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.3 mg/L) (20%) = 0.06 mg/L (60 ug/L)
where:
0.3 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No carcinogenicity studies were found in the literature searched.
0 The International Agency for Research on Cancer (IARC) has not
evaluated the carcinogenic potential of ametryn.
8 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), ametryn may be classifed in
Group D: not classified. This category is for agents with indadequate
animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA has established residue tolerences for ametryn in or on
raw agricultural commodities that range from 0.1 to 0.5 ppm (CFR, 1985).
VII. ANALYTICAL METHODS
0 Analysis of ametryn is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing pesti-
cides in water samples. In this method, approximately 1 liter of
sample is extracted with methylene chloride. The extract is concen-
trated and the compounds are separated using capillary column GC.
Measurement is made using a nitrogen phosphorus detector. The method
detection limit has not been determined for ametryn, but it is estimated
that the detection limits for analytes included in this method are in
the range of 0.1 to 2 ug/L (U.S. EPA, 1986b).
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) adsorption
will remove ametryn from water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
ametryn on GAC.
0 Whittaker (1980) reported the results of GAC columns operating under
bench-scale conditions. At a flow rate of 0.8 gpm/ft2 and an empty
bed contact time of 6 minutes, ametryn breakthrough (when effluent
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concentration equals 10% of influent concentration) occurred after
896 bed volumes (BV). When a bi-solute ametryn-propham solution was
passed over the same column, ametryn breakthrough occurred after 240 BV.
In a laboratory study (Nye, 1984) GAC was employed as a possible
means of removing ametryn from contaminated wastewater. The results
show that the column exhaustion capacity was 111.2 mg ametryn adsorbed
on 1 g of activated carbon.
Treatment technologies for the removal of ametryn from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt ametryn
removal from water must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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IX. REFERENCES
Anderson, K.J., E.G. Leighty and M.T. Takahasi. 1972. Evaluation of herbicides
for possible mutagenic activity. J. Agr. Food Chem. 20:649-656.
Ceglowski, W.S., D.D. Ercegrovich and N.S. Pearson. 1979. Effects of pesticides
on the reticuloendothelial system. Adv. Exp. Med. Biol. 121:569-576.
CFR. 1985. Code of Federal Regulations. July 1, 1985. 40 CFR 180.258.
pp. 300-301.
Consultox Laboratories Limited.* 1974. Ametryn: Acute oral and dermal toxicity
evaluation. Unpublished study. MRID 00060310.
Domenjoz, R. 1961.* Ametryn: Toxicity in long-term administration. Unpub-
lished study. MRID 00034838.
Grunfeld, Y. 1981.* Ametryn 80 w.p.: Acute oral toxicity in the rat.
Unpublished study. MRID 00100573.
Kopp, R.W.* 1975. Acute eye irritation potential study in rabbits. Final
Report. Project No. 915-104. Unpublished study. MRID 00060311.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Nye, J.C. 1984. Treating pesticide-contaminated wastewater. Development
and evaluation of a system. American Chemical Society.
Oliver, W.H., G.S. Born and P.L. Zeimer. 1969. Retention, distribution, and
excretion of ametryn. J. Agr. Food Chem. 17:1207-1209.
Piccirillo, V.J.* 1977. 28-day pilot feeding study in mice. Final Report.
Project No. 483-126. Unpublished study. MRID 00068169.
Sachsse, K. and R. Bathe.* 1976. Acute dermal LD5Q in the rat of technical
G34162. Project No. Siss. 5665. Unpublished study. MRID 00068172.
Sachsse, K. and L. Ullmann.* 1977. Skin irritation in the rabbit after
single application of technical grade G34162. Unpublished study.
MRID 00068174.
Shirasu, Y.f M. Moriya, K. Kato, A. Furuhashi and T. Kada. 1976. Mutagenic
screening of pesticides in the microbial system. Mutat. Res. 40:19-30.
Simmons, V.F. and D. Poole.* 1977. In-vitro and in-vivo microbiological
assays of six Ciba-Geigy chemicals. SRI project LSC-5686. Final Report.
Unpublished study. MRID 00060642.
Stenger, P. and V. Plants.* 1961 a. Oral toxicity in rats. Unpublished
study. MRID 00048226.
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Ametryn August, 1987
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Stenger, P. and V. Planta.* 1961b. Subchronic toxicity test no. 257.
Unpublished study. MRID 00048228.
STORET. 1987.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
Carcinogen Risk Assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. EPA Method #1 - Determination of nitrogen and phosphorus
containing pesticides in ground water by GC/NPD, January 1986 draft.
Available from U.S. EPA's Environmental Monitoring and Support Laboratory,
Cincinnati, OH.
WSSA. 1983. Weed Science Society of America. Herbicide handbook. 5th
ed. Champaign, IL: Heed Society of America, pp. 16-19.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
•Confidential Business Information submitted to the Office of Pesticide
-------
August, 1987
DRAFT
I. INTRODUCTION
AMMONIUM SULFAMATE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
wa^ronw M"iso? e (HA) *rogr«, 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 .
st**. Ad^is"ief 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. «»ject to
UimrJUfi^ ?dvisories 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.
Ssk^r^^h168 d° "Ot
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Ammonium Sulfaoate August, 1987
-4-
0 Read and Hueber (1938) orally administered 1 mL of a 50% aqueous
solution of ammonium sulfamate (1.7 g/kg/day) to 10 rats on alternate
days. Five rats were killed on the 27th day of the study after nine
treatments, and the remaining five were killed on the 42nd day of the
study after 15 treatments. Investigators reported that there were no
gross pathological changes of importance in any of the animals.
Microscopic pathology indicated the following: in one animal, super-
ficial capillaries of the stomach mucosa occasionally contained
yellow-brown granules; in three animals, there was slight vacuolation
of the cytoplasm of liver cells about the central veins, but these
changes were very mild; and in the spleen, three of the sections had
moderate numbers of macrophages filled with hemosiderin. A fourth
spleen section showed marked erythrophagia.
Long-term Exposure
0 Gupta et al. (1979) reported the results of a 90-day study involving
oral administration of 0, 100, 250 or 500 mg/kg of ammonium sulfamate
to rats 6 days a week. No adverse effects were observed with respect
to appearance, behavior or survival of animals. No significant
difference in the body weights of rats was observed except in the
case of rats receiving 500 mg/kg, where body weight was signifi-
cantly less than controls after the end of 60 days. No significant
changes in relative organ weights were noticed in any group of rats.
Hematological examination conducted at 30, 60 and 90 days revealed
nonsignificant increases in the numbers of neutrophils in the female
adult and male weanling rats (500 mg/kg dose level) after 90 days.
In the histological examination, organs in all the groups of animals
appeared normal except that the liver of one adult rat (500 mg/kg)
showed slight fatty degenerative changes after 90 days.
0 Rosen et al. (1965) reported the findings of a study in female rats
following administration of ammonium sulfamate at dietary levels of 1.1%
(10 g/kg/day) or 2.1% (20 g/kg/day) for 105 days. No effect was detected
at the 1% (10 gAg/day) level of feeding, but growth retardation
and a slight cathartic effect were observed at the 2% (20 g/kg/day)
dietary level. No othef information was provided by the authors.
0 Sherman and Stula (1966) reported the results of a 19-month feeding
study in 29-day-old CHR-CD male and female rats. Ammonium sulfamate
was fed at dietary concentrations of 0, 350 (350 mg/kg) or 500
(500 mg/kg) ppm without any clinical or nutritional evidence of toxicity.
There were no histopathological changes that could be attributed to
the feeding of the test chemical. The observed pathologic lesions
were interpreted as a result of spontaneous diseases.
Reproductive Effects
0 Sherman and Stula (1966) reported the results of a three-generation
reproduction study in rats. Rats receiving 0, 350 (350 mg/kg) or
500 (500 mg/kg) ppm ammonium sulfamate in the diet showed no evidence
of toxicity as measured by histopathological evaluation and reproduction
and lactation indices.
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Ammonium Sulfamate August, 1987
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Developmental Effects
0 No information was found in the available literature on the develop-
mental effects of ammonium sulfamate.
Mutagenicity
0 No information was found in the available literature on the mutagenic
effects of ammonium sulfamate.
Carcinogenic!ty
0 No information was found in the available literature on the carcinogenic
effects of ammonium sulfamate.
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 0 (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 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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No data were located in the available literature that were suitable for
deriving a One-day HA value for ammonium sulfamate. It is recommended that
the Longer-term HA value for the 10-kg child (21.4 mg/L, calculated below)
be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
No data on ammonium sulfamate toxicity were located in the available
literature that were suitable for calculation of a Ten-day HA value. It is
recommended that the Longer-term HA value for the 10-kg child (21.4 mg/L,
calculated below) be used at this time as a conservative estimate of the
Ten-day HA value.
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Ammonium Sulfamate August, 1987
-6-
Lonqer-term Health Advisory
The subchronic oral toxicity study in rats by Gupta et al. (1979) may be
considered for the Longer-term HA. In this study, rats (female adults and
male and female weanlings) received ammonium sulfamate orally at dose levels
of 0, 100, 250 or 500 mg/kg/day for 90 days. Hematological and histological
examinations at 30, 60 and 90 days revealed nonsignificant changes in hemato-
logical and histological measures. However, adult rats fed 500 mg/kg ammonium
sulfamate showed lesser weight gain compared to other groups.
Using 250 mg/kg/day as a No-Observed-Adverse-Effect-Level (NOAEL), a
Longer-term HA for the 10-kg child is calculated as follows:
Longer-term HA = (250 mg/kg/day) (10 kg) (6/7) 0 21.4 mg/L (21,400 ug/L)
* (100) (1 L/day)
where:
250 mg/kg/day = NOAEL, based on the absence of hematological and
histopathological changes in rats.
10 kg = assumed body weight of a child.
6/7 3 conversion from 6 days to 7 days.
100 B uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
For the 70-kg adult:
Longer-term HA = (250 mg/kg/day) (70 kg) (6/7) . 75 mg/L (75,000 ug/L)
(100) (2 L/day)
where:
250 mg/kg/day = NOAEL, based on the absence of hematological and
histopathological changes in rats.
70 kg = assumed body weight of an adult.
6/7 = conversion from 6 days to 7 days.
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
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Ammonium Sulfamate August, 1987
-7-
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.
exposure to this chemical.
The study by Gupta et al. (1979) has been selected to serve as the basis
for determination of the Lifetime HA even though the results of this subchronic
study were based on 90 days' exposure. In this study, rats (female adults
and weanling males and females) received ammonium sulfamate orally in drinking
water at dose levels of 0, 100, 250 or 500 mg/kg/day for 90 days. The NOAEL
was identified as 250 mg/kg/day, since the highest dose level of 500 mg/kg/day
was associated with decreased body weight gain in rats over a 90-day exposure
period). In a chronic feeding study reported by Sherman and Stula (1966)
in rats, ammonium sulfamate was fed to rats at dietary levels of 0, 350 or
500 ppm over a 19-month period. The authors stated that these dose levels
did not produce any significant clinical or histological changes in rats
receiving the test compound, and any changes recorded were interpreted as
being lesions of spontaneous diseases.
Using a NOAEL of 250 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (250 mg/kg/day) (6/7) = 0>214 mg/kg/day
(1,000)
where:
250 mg/kg/day = NOAEL.
6/7 = conversion from 6 days to 7 days.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study
of less than a lifetime exposure.
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Ammonium Sulfamate August, 1987
-8-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.214 mg/kq/day) (70 kg) = 7.5 /L (7 500 /L)
(2 L/day)
wheres
0.250 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 = (7.5 mg/L) (20%) - 1.5 mg/L (1,500 ug/L)
where:
7.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No studies were found in the available literature investigating
the carcinogenic potential of ammonium sulfamate. Applying the
criteria described in EPA's final guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), ammonium sulfamate may be
classified in Group D: not classified. This category is used for
substances with inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The American Conference of Government Industrial Hygienists (ACGIH)
has adopted a Threshold Limit Value-Time-Weighted Average (TLV-TWA)
of 10 mg/m3 and a TLV short-term exposure limit (STEL) of 20 mg/m3
for inhalation exposure (ACGIH, 1984).
VII. ANALYTICAL METHODS
0 There is no standardized method for determination of ammonium sulfamate
in water samples. A procedure has been reported for the estimation of
ammonium sulfamate in certain foods, however (U.S. FDA, 1969). This
procedure involves a colorimetric determination of ammonium sulfamate
based on the liberation of S04 and reduction it to H2S, which is
measured after treating with zinc, p-aminodimethylaniline and ferric
chloride to form methylene blue.
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Ammonium Sulfamate August, 1987
-9-
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies capable of effectively removing ammonium sulfamate from
contaminated water.
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Ammonium Sulfamate August, 1987
-10-
IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH.
Bergen, D.S. and F.M. Wiley.* 1938. The metabolism of sulfamic acid and
ammonium sulfamate. Unpublished report. Submitted to U.S. EPA, Office
of Pesticide Programs, Washington, DC.
Gupta, B.N., R.N. Khanna and K.K. Datta. 1979. Toxicological studies of
ammonium sulfamate in rats after repeated oral administration. Toxicology.
13:45-49.
Konnai, M., Y. Takeuchi and T. Takematsu. 1974. Basic studies on the residues
and movements of forestry herbicides in soil. Bull. Coll. Agric.
Utsunomiya Univ. 9(1 ):995-1012.
Heister, R., ed. 1983. Farm chemicals handbook, willoughby, OH: Meister
Publishing Co.
Read, W.T. and K.C. Hueber.* 1938. The pathology produced in rats following
the administration of sulfamic acid and ammonium sulfamate. Unpublished
report. NRID GS0016-0040.
Rosen, D.E., C.J. Krisher, H. Sherman and E.E. Stula. 1965. Toxicity studies
on ammonium sulfamate. The Toxicologist. Fourth Annual Meeting, Williams-
burg, VA. March 8-1 0.
Sherman, H. and E. Stula.* 1966. Toxicity studies on ammonium sulfamate.
Unpublished report. MRID GS0016-0038.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (1 85):33992-34003. September 24.
U.S. FDA. 1969. U.S. Food and Drug Administration. Pesticide analytical
manual. Vol. II. Washington, DC.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
September, 1987
ATRAZINE
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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Atrazine
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1912-24-9
Structural Formula .
H H
2XThloro-4-ej:hylamir.o-6-isopropylan,ino-1,3, 5-triazine
Synonms
:, 1987).
Uses
0 Atrazine is used for nonselective weed e - i
noncropped land and selective weed control f °° lndustrial or
cane, pineapple and certain other Bi,«». 7M • ""' Sor9num' su?ar
^^x«i w w*et DJ.an gg ( MA le^A^ IQO^\
Properties (Meister, 1987; Windholz, 1976)
Chemical Formula
Molecular Weight
Physical State w
Boiling Point (25 mm Hg) . "' ordorless' crystalline solid
Melting Point 175
Density (20B) JJf J° 177°?
Vapor Pressure (20°O i'r, .« •»
°
70 mg/L at 22 °C
Log Octanol/wacer Partition --2°
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
or Juy (Newby and ed, 976"? 6St conc«tt^on. found in June
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Atrazine September, 1987
-3-
Samples were collected at 1,468 surface water locations and 2,123
ground water locations, and atrazine was found in 36 states. The
85th percentile of all non-zero samples was 2.3 ug/L in surface water
and 1.9 ug/L in ground water sources. The maximum concentration
found in surface water was 400 ug/L and in ground water it was
1,400 ug/L.
• Atrazine has been found also in ground water in Pennsylvania, Iowa,
Nebraska, Wisconsin and Maryland; typical positives were 0.3 to 3 ppb
(Cohen et al., 1986).
Environmental Fate
0 An aerobic soil metabolism study in Lakeland sandy loam, Hagerstown
silty clay loam, and Wehadkee silt loam soils showed conversion of
atrazine to hydroxyatrazine, after 8 weeks, to be 38, 40 and 47% of
the amount applied, respectively, (Harris, 1967). Two additional
degradates, deisopropylated atrazine and deethylated atrazine, were
identified in a sandy loam study (Beynon et al., 1972).
0 Hurle and Kibler (1976) studied the effect of water-holding capacity
on the rate of degradation and found a half-life for atrazine of more
than 125 days, 37 days and 36 days in sandy soil held at 4%, 35% and
70% water-holding capacity, respectively.
0 In Oakley sandy loam and Nicollet clay loam, atrazine had a half-life
of 101 and 167 days (Warnock and Leary, 1978).
0 Carbon dioxide production was generally slow in several anaerobic
soils: sandy loam, clay loam, loamy sand and silt loam (Wolf and
Martin, 1975; Goswami and Green, 1971; Lavy et al., 1973).
0 14C-Atrazine was stable in aerobic ground water samples incubated for
15 months at 10 or 25°C in the dark (Weidner, 1974).
0 Atrazine is moderately to highly mobile in soils ranging in texture
from clay to gravelly sand as determined by soil thin layer chroma-
tography (TLC), column leaching, and adsorption/desorption batch
equilibrium studies. Atrazine on soil TLC plates was intermediately
mobile in loam, sandy clay loam, clay loam, silt loam, silty clay
loam, and silty clay soils, and was mobile in sandy loam soils.
Hydroxyatrazine showed a low mobility in sandy loam and silty clay
loam soils (Helling, 1971).
0 Soil adsorption coefficients for atrazine in a variety of soils were:
sandy loam (0.6), gravelly sand (1.8), silty clay (5.6), clay loam
(7.9), sandy loam (8.7), silty clay loam (11.6), and peat (more than
21) (Weidner, 1974; Lavy 1974; Talbert and Fletchall, 1965).
0 Soil column studies indicated atrazine was mobile in sand, fine sandy
loam, silt loan and loam; intermediately mobile in sand, silty clay
loam and sandy loam; low to intermediately mobile in clay loam (Weidner,
1974; Lavy, 1974; Ivey and Andrews, 1964; Ivey and Andrews, 1965).
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Atrazine September, 1 987
-4-
0 In a Mississippi field study, atrazine in silt loam soil had a half-
life of less than 30 days (Portnoy, 1978). In a loam to silt loam
soil in Minnesota, atrazine phytotoxic residues persisted for more
than 1 year and were detected in the maximum-depth samples (30 to
42 inches) (Darwent and Behrens, 1968). In Nebraska, phytotoxic
residues persisted in silty clay loam and loam soils 16 months after
application of atrazine; they were found at depths of 12 to 24 inches.
But atrazine phytotoxic residues had a half-life of about 20 days in
Alabama fine sandy loam soil, although leaching may partially account
for this value (Buchanan and Hiltbold, 1973).
0 Under aquatic field conditions, dissipation of atrazine was due to
leaching and to dilution by irrigation water, with residues persisting
for 3 years in soil on the sides and bottoms of irrigation ditches,
to the maximum depth sampled, 67.5 to 90 cm (Smith et al., 1975).
III. PHARMACOKINETICS
Absorption
0 Atrazine appears to be readily absorbed from the gastrointestinal
tract of animals. Bakke et al. (1972) administered single 0.53-mg
doses of 14c-ring-labeled atrazine to rats by gavage. Total fecal
excretion after 72 hours was 20.3% of the administered dose; the
remainder was excreted in urine (65.6%) or retained in tissues (15.8%).
This indicates that at least 80% of the dose was absorbed.
Distribution
0 Bakke et al. (1972) administered single 0.53-mg doses of 14c-ring-
labeled atrazine to rats by gavage. Liver, kidney and lung contained
the largest amounts of radioactivity, while fat and muscle had lower
residues than the other tissues examined.
0 In a metabolism study by Ciba-Geigy (1983a), the radioactivity of
14C-atrazine dermally applied to Harlan Sprague-Dawley rats at
0.25 mg/kg was distributed to a minor extent to body tissues. The
highest levels were measured in liver and muscle at all time points
examined; 2.1% of the applied dose was in muscle and 0.5% in liver
at 8 hours.
0 Khan and Foster (1976) observed that in chickens the hydroxy metabo-
lites of atrazine accumulate in the liver, kidney, heart and lung.
Residues of both 2-chloro and 2-hydroxy moieties were found in chicken
gizzard, intestine, leg muscle, breast muscle and abdominal fat.
Metabolism
0 The principal reactions involved in the metabolism of atrazine are
dealkylation at the C-4 and C-6 positions of the molecule. There is
also some evidence of dechlorination at the C-2 position. These data
were reported by several researchers as demonstrated below.
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Atrazine September, 1987
-5-
Bakke et al. (1972) administered single 0.53-mg doses of 1^-ring-
labeled atrazine to rats by gavage. Less than 0.1% of the label
appeared in carbon dioxide in expired air. Most of the radioactivity
was recovered in the urine (65.5% in 72 hours), including at least 19
radioactive compounds. Approximately 47% of the urinary radioactivity
was identified as 2-hydroxyatrazine and its two mono-N-dealkylated
metabolites. None of the metabolites identified contained the 2-chloro
moiety (which may have been removed via hydrolysis during the isolation
technique or by a dechlorinating enzyme as suggested by the in vitro
studies of Foster et al. (1979), who found evidence for a dechlorinase
in chicken liver homogenates incubated with atrazine.
Bohme and Bar (1967) identified five urinary metabolites of atrazine
in rats: the two monodealkylated metabolites of atrazine, their
carboxy acid derivatives and the fully dealkylated derivative. All
of these metabolites contained the 2-chloro group. The in vitro
studies of Dauterman and Muecke (1974) also found no evidence for
dechlorination of atrazine in the presence of rat liver homogenates.
Similarly, Bradway and Moseman (1982) administered atrazine (50,
5, 0.5 or 0.005 mg/day) for 3 days to male Charles River rats and
observed that the fully dealkylated derivative (2-chloro-4,6-diamino-
s-triazine) was the major urinary metabolite, with lesser amounts of
the two mono-N-dealkylated derivatives.
Erickson et al. (1979) dosed Pittman-Moore mini-pigs by gavage with
0.1 g of atrazine (SOW). The major compounds identified in the urine
were the parent compound (atrazine) and deethylated atrazine (which
contains the 2-chloro substituent).
Excretion
Urine appears to be the principal route of atrazine excretion in
animals. Following the administration of 0.5 mg doses of Boring-
labeled atrazine by gavage to rats, Bakke et al. (1972) reported that
in 72 hours most of the radioactivity (65.5%) was excreted in the
urine, 20.3% was excreted in the feces, and less than 0.1% appeared
as carbon dioxide in expired air. About 85 to 95% of the urinary
radioactivity appeared within the first 24 hours after dosing,
indicating rapid clearance.
Dauterman and Muecke (1974) have reported that atrazine metabolites
are conjugated with glutathione to yield a mercapturic acid in the
urine. The studies of Foster et al. (1979) in chicken liver homo-
genates also indicate that atrazine metabolism involves glutathione.
Ciba-Geigy (1983b) studied the excretion rate of 14c-atrazine from
Harlan Sprague-Dawley rats dermally dosed with atrazine dissolved in
tetrahydrofuran at levels of 0.025, 0.25, 2.5 or 5 mg/kg. Urine and
feces were collected from all animals at 24-hour intervals for 144
hours. Results indicated that atrazine was readily absorbed, and
within 48 hours most of the absorbed dose was excreted, mainly in the
urine and to a lesser extent in the feces. Cumulative excretion in
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Atrazine September, 1987
-6-
urine and feces appeared to be directly proportional to the administered
dose, ranging from 52% at the lowest dose to 80% at the highest dose.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 A case of severe contact dermatitis was reported by Schlichter and
Beat (1972) in a 40-year-old farm worker exposed to atrazine formu-
lation. The clinical signs were red, swollen and blistered hands
with hemorrhagic bullae between the fingers.
Long-term Exposure
0 Yoder et al. (1973) examined chromosomes in lymphocyte cultures
taken from agricultural workers exposed to herbicides including
atrazine. There were more chromosomal aberrations in the workers
during mid-season exposure to herbicides than during the off-season
(no spraying). These aberrations included a four-fold increase in
chromatid gaps and a 25-fold increase in chromatid breaks. During
the off-season, the mean number of gaps and breaks was lower in this
group than in controls who were in occupations unlikely to involve
herbicide exposure. This observation led the authors to speculate
that there is enhanced chromosomal repair during this period of time
resulting in compensatory protection.
Animals
Short-term Exposure
0 Acute oral LDSO values of 3,000 mg/kg in rats and 1,750 rag/kg in
mice have been reported for technical atrazine by Bashmurin (1974);
the purity of the test compound was not specified.
• Molnar (1971) reported that when atrazine was administered by gavage
to rats at 3,000 mg/kg, 6% of the rats died within 6 hours, and 25%
of those remaining died within 24 hours. The rats that died during
the first day exhibited pulmonary edema with extensive hemorrhagic
foci, cardiac dilation and microscopic hemorrhages in the liver and
spleen. Rats that died during the second day had hemorrhagic broncho-
pneumonia and dystrophic changes of the renal tubular mucosa. Rats
sacrificed after 24 hours had cerebral edema and histochemical
alterations in the lungs, liver and brain.
• CSE Laboratories (1980) studied the acute oral lethality of atrazine
in Sprague-Dawley rats dosed at 1,500, 1,700, 1,900, 2,000 or
5,000 mg/kg. Deaths occurred within 48 hours in all groups except
for that given the 1,500-mg/kg dose. Toxic signs in other groups
included ataxia, diarrhea, oral discharge and chromorhinorrhea (bloody
nasal discharge). After 14 days, examination of surviving rats
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Atrazine September, 1987
-7-
revealed that body weights were generally normal, and gross necropsy
revealed no abnormalities.
0 Gaines and Linder (1986) determined the oral LDso for adult male and
female rats to be 737 and 672 rag/kg respectively and 2,310 mg/kg for
pups. This study also reflected that the dermal LDso f°r adult rats
was higher than 2,500 mg/kg.
0 An acute dermal LD50 value of 7.55 g/kg for technical atrazine applied
to rabbits has been reported (Frear, 1969).
0 Palmer and Radeleff (1964) administered atrazine as a fluid dilution
or in gelatin capsules to Delaine sheep and dairy cattle. Two doses
of 250 mg/kg atrazine caused death in both sheep and cattle. Sixteen
doses of 100 mg/kg were lethal to one sheep. At necropsy, degeneration
and discoloration of the adrenal glands and congestion in lungs,
liver and kidneys were observed.
0 Palmer and Radeleff (1969) orally administered 10 doses of atrazine SOW
(analysis of test material not provided) by capsule or by drench to sheep
at 5, 10, 25, 50 or 100 mg/kg/day and to cows at 10 or 25 mg/kg/day.
The number of animals in each dosage group was not stated, and the use
of controls was not indicated. Observed effects included muscular
spasms, stilted gait and stance and anorexia at all dose levels in
sheep and at 25 mg/kg in cattle. Necropsy revealed epicardial petechiae
(small hemorrhagic spots on the lining of the heart) and congestion
of the kidneys, liver and lungs. Effects appeared to be dose related.
A Lowest-Observed-Adverse-Effect-Level (LOAEL) of 5 mg/kg/day in
sheep and a No-Observed-Adverse-Effect-Level (NOAEL) of 10 mg/kg/day
in cows can be identified from this study.
• Bashmurin (1974) reported that oral administration of 100 mg/kg of
atrazine to cats had a hypotensive effect, and that a similar dose in
dogs was antidiuretic and decreased serum cholinesterase activity.
No other details of this study were reported.
Dermal/Ocular Effects
0 In a primary dermal irritation test in rats, atrazine at 2,800 mg/kg
produced erythema but no systemic effects (Hayes, 1982).
0 In primary eye irritation studies, atrazine was described as irritating
when applied at an unspecified concentration in rats (Hayes, 1982).
Long-term Exposure
0 Hazelton Laboratories (1961) fed atrazine to male and female rats for
2 years at dietary levels of 0, 1, 10 or 100 ppm. Based on the
dietary assumptions of Lehman (1959), these levels correspond to
doses of approximately 0, 0.05, 0.50 or 5.0 mg/kg/day. After 65
weeks, the 1.0-ppm dose was increased to 1,000 ppm (50 mg/kg/day) for
the remainder of the study. No treatment-related pathology was found
at 26 weeks, at 52 weeks, at 2 years, or in animals that died and
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Atrazine September, 1987
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were neeropsled during the study. Results of blood and urine analyses
were unremarkable. Atrazine had no effects on the general appearance
or behavior of the rats. A transient roughness of the coat and
piloerection were observed in some animals after 20 weeks of treatment
at the 10- and 100-ppm levels but not at 52 weeks. Body weight gains,
food consumption and survival were similar in all groups for 18
months, but from 18 to 24 months there was high mortality due to
infections (not attributed to atrazine) in all groups, including
controls, which limits the usefulness of this study in determining a
NOAEL for the chronic toxicity of atrazine.
• In a 2-year study by Woodard Research Corporation (1964), atrazine
(SOW formulation) was fed to male and female beagle dogs for 105
weeks at dietary levels of 0, 15, 150 or 1,500 ppm. Based on the
dietary assumptions of Lehman (1959), these levels correspond to
doses of 0, 0.35, 3.5 or 35 mg/kg/day. Survival rates, body weight
gain, food intake, behavior, appearance, hematologic findings,
urinalyses, organ weights and histologic changes were noted. The
15-ppm dosage (0.35 mg/kg/day) produced no toxicity, but the 150-ppm
dosage (3.5 mg/kg/day) caused a decrease in food intake as well as
increased heart and liver weight in females. In the group receiving
1,500 ppm (35 mg/kg/day) atrazine, there were decreases in food
intake and body weight gain, an increase in adrenal weight, a
decrease in hematocrit and occasional tremors or stiffness in the
rear limbs. There were no differences among the different groups in
the histology of the organs studied. Based on these results, a NOAEL
of 0.35 mg/kg/day can be identified for atrazine.
Reproductive Effects
• A three-generation study on the effects of atrazine on reproduction
in rats was conducted by Woodard Research Corporation (1966). Groups
of 10 males and 20 females received atrazine at dietary levels of 0,
50 or 100 ppm. Based on the dietary assumptions that 1 ppm in the
diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these
levels correspond to doses of approximately 0, 2.5 or 5 mg/kg/day.
After receipt, animals were fed only half of the dietary levels for
the first 3 weeks and were then changed to the stated levels for 74
days. After 74 days of dosing, rats within each group were paired
for mating. Approximately 13 days after the first weaning, the
females in each group were remated with different males in the same
group. The protocol employed following the first mating was repeated
with the pups from the second mating. After the second weaning, the
parents (F0 generation) were sacrificed and the weanlings (Fib genera-
tion) were used to form another three groups. The entire series of
tests was repeated following the dosing of the FIO generation with
50 or 100 ppm (2.5 or 5 mg/kg/day) atrazine for 105 days. The F2b
generation was fed atrazine for 75 days and the entire protocol
repeated again. After weaning of the F^b generation, the study was
terminated. There were no adverse effects of atrazine on reproduction
observed during the course of the three-generation study. Atrazine
had no effect on any of the following parameters: mean parental body
weight, survival, appearance, behavior, number of litters/group.
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Atrazine September, 1987
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number of live births, mean body weights at birth and weaning, and
percent of pups alive at weaning. A NOAEL of 100 ppn (5 mg/Jq/day)
was identified for this study. However, the usefulness of this study
is limited due to an alteration of the atrazine content of the diet
during important maturation periods of the neonates.
Developmental Effects
0 In the three-generation reproduction study in rats conducted by
Woodard Research Corporation (1966) (described above), atrazine at
dietary levels of 50 or 100 ppm (2.5 or 5 mg/Jq/day) resulted in no
observed histologic changes in the weanlings and no effects on fetal
resorption. No malformations were observed, and weanling organ
weights were similar in controls and atrazine-treated animals.
Therefore, a NOAEL of 100 ppn (5 mg/Jq/day) was also identified for
developmental effects in this study. However, the usefulness of this
study is limited due to an alteration of the atrazine content of the
diet during important maturation periods of the neonates.
0 Atrazine was administered orally to pregnant rats on gestation days
6 to 15 at 0, 100, 500 or 1,000 mg/Jq (Ciba-Geigy, 1971). The two higher
doses increased the number of embryonic and fetal deaths, decreased
the mean weights of the fetuses and retarded the sJeletal development.
No teratogenic effects were observed. The highest dose (1,000 mg/Jq)
resulted in 23% maternal mortality and various toxic symptoms. The
100 mg/Jq dose had no effect on either dams or embryos and is therefore
the maternal and fetotoxic NOAEL in this study.
0 In a study by Ciba-Geigy (1984a), Charles River rats received atrazine
(97%) by gavage on gestation days 6 to 15 at dose levels or 0, 10, 70,
or 700 mg/Jq/day. Excessive maternal mortality (21/27) was noted at
700 mg/Jq/day, but no mortality was noted at the lower doses; also
reduced weight gains and food consumption were noted at both 70 and
700 mg/Jq/day. Developmental toxicity was also present at these dose
levels. Fetal weights were severely reduced at 700 mg/Jq/day; delays
in sJeletal development occurred at 70 mg/Jq/ day, and a dose-related
runting was noted at 10 mg/Jq/day and above. The NOAEL for maternal
toxicity appears to be 10 mg/Jq/day, hower, this is also the LOAEL
for developmental effects.
0 New Zealand white rabbits received atrazine (96%) by gavage on gestation
days 7 through 19 at dose levels of 0, 1, 5 or 75 mg/Jq/day (Ciba-Geigy,
(1984b). Maternal toxicity, evidenced by decreased body weight gains
and food consumption, was present in the mid- and high-dose groups.
Developmental toxicity was demonstrated only at 75 mg/Jq/day by an
increased resorption rate, reduced fetal weights, and delays in
ossification. No teratogenic effects were indicated. The NOAEL
appears to be 1 mg/Jq/day.
0 Peters and Cook (1973) fed atrazine to pregnant rats (four/group)
at levels of 0, 50, 100, 200, 300, 400, 500 or 1,000 ppn in the diet
throughout gestation. The authors assumed a body weight of 300 g and
a daily food consumption of 12 g (based on Arrington, 1972); thus,
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Atrazine September, 1987
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these levels correspond to approximately 0, 2, 4, 8, 12, 16, 20 or
40 mg/fcg/day. The number of pups per litter was similar in all
groups, and there were no differences in weanling weights. This
study identified a NOAEL of 40 mg/fcg/day for developmental effects.
In another phase of this study, the authors demonstrated that sub-
cutaneous (sc) injections of 50, 100 or 200 mg/Jq atrazine on gestation
days 3, 6 and 9 had no effect on the litter size, while doses of 800
or 2,000 mg/Jg were embryotoxic. Therefore, a NOAEL of 200 mg/Jq by
the sc route was identified for embryotoxicity.
Mutagenicity
0 Loprieno et al. (1980) reported that single doses of atrazine
(1,000 mg/kg or 2,000 mg/kg, route not specified) produced bone marrow
chromosomal aberrations in the mouse. No other details of this study
were provided.
fi Murnik and Nash (1977) reported that feeding 0.01% atrazine to male
Drosophila melanogaster larvae significantly increased the rate of
both dominant and sex-linted recessive lethal mutations. They stated,
however, that dominant lethal induction and genetic damage may not be
directly related.
0 Adler (1980) reviewed unpublished work on atrazine mutagenicity
carried out by the European Economic Community. Mutagenic activity
was not induced when mammalian liver enzymes (S-9) were used; however,
the use of plant microsomes produced positive results. Also, in
in vivo studies in mice, atrazine induced dominant lethal mutations
and increased the frequency of chrooatid breaks in bone marrow.
Hence, the author suggested that activation of atrazine in mammals
occurs independently of the liver, possibly in the acidic part of the
s tomach.
0 As described previously, Yoder et al. (1973) studied chromosomal
aberrations in the lymphocyte cultures of farm workers exposed to
various pesticides including atrazine. During mid-season a 4-fold
increase in chromatid gaps and a 25-fold increase in chromatid breaks
was observed. During the off-season (no spraying), the number of
gaps and breaks was lower than in controls, suggesting to the authors
that there is enhanced chromosomal repair during the un ex posed period.
Carcinogenicity
0 Innes et al. (1969) investigated the tumorigenicity of 120 test com-
pounds including atrazine in mice. Two F1 hybrid stocks (C57BL/6 x AnF
and C57BL/6 x AKR) were used. A dose of 21.5 mg/kg/day was administered
by gavage to mice of both sexes from age 7 to 28 days. After weaning
at 4 weeks, this dose level was maintained by feeding 82 ppn atrazine
ad libitum in the diet for 18 months. At necropsy, thoracic and
abdominal cavities were examined, and histologic studies were performed
on all major organs and grossly visible lesions. Blood smears were
examined if the mice showed signs of splenomegaly or lynphadenopathy.
The incidence of hepatomas, pulmonary tumors, lymphomas and total
tumors in atrazine-treated mice was not significantly different from
that in the negative controls.
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Atrazine September, 1987
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0 In data supplied to EPA (U.S. EPA, 1986a) in support of pesticide
registration for atrazine, Ciba-Geigy Corporation (1985) submitted
preliminary summary incidence information (1-year interim report)
on the histopathological findings of their 2-year oncogenicity study
of atrazine in Spraque-Dawley rats. The summary tables contained
indications of increased numbers of tumors in the mammary glands of
the female rats. The statistical evaluation of this preliminary
data raised concerns of a dose-related response for increases in
mammary tumors. Unfortunately, the data are of a preliminary nature
and cannot be used for any further conclusions in this document before
the 2-year study is completed and evaluated. However, a subsequent
briefing paper by Ciba-Geigy (1987) indicated that this study is
positive. The evaluation of the recently submitted final report of
this 2-year rat study will be performed at a later date.
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 tj'.e following formula:
,_ ug/L,
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/Jgg bw/day.
BW = assumed body weight of a child (10 Jq) or
an adult (70 Jq).
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
No suitable information was found in the available literature for the
determination of the One-day HA value for atrazine. It is, therefore, recom-
mended that the Ten-day HA value calculated below for a 10- Jg child of
0.1 mg/L (100 ug/L), be used at this time as a conservative estimate of the
One-day HA value.
Ten-day Health Advisory
Two teratology studies by Ciba-Geigy one inthe rat (1984a) and one in the
rabbit (1984b) were considered for the calculation of the Ten-day HA value.
The rat study reflected a NOAEL of 10 mg/Jg/day for maternal toxicity but this
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Atrazine September, 1987
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value was also the LOAEL for developmental toxicity while the rabbit study
reflected NOAELs of 5 mg/kg/day for developmental toxicity and 1 mg/)q/day for
maternal toxicity. Thus, the rabbit appears to be a more sensitive species than
the rat for internal toxicity, hence, the rabbit study with a NOAEL of 1 mg/Jq/day
is used in the calculations below.
The Ten-day HA for a 10 kg child is calculated below as follows:
(1 ng/kg/d) x (10kg) = n.1 n.g/r. (inn »g/r.)
(100 )x (1 L/day)
where:
1 ing/kg/day = NOAEL, based on maternal toxicity evidenced by decreased
body weight gain and food consumption.
10 kg a assumed body weight of a child
100 = uncertainty factor, chosen in accordance with ODW/NAS guidelines
for use with a NOAEL from an animal study.
1 L/d = assumed daily consumption for a child
Longer-term Health Advisory
No suitable information was found in the available literature for the
determination of the longer-term HA value for atrazine. It is, therefore,
recommended that the adjusted DWEL for a 10-kg child of 0.035 mg/L
(35 ug/L) and the DWEL for a 70-Jq adult of 0.123 mg/L (123 ug/L) be used at
this time as a conservative estimate of the Longer-term HA values.
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 Intafe (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is litely 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%
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Atrazine September, 1987
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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, 1986b), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study in dogs by Woodard Research Corporation
(1964) has been selected to serve as the basis for the calculation of the
Longer-term HA, as well as the DWEL, and Lifetime HA. Atrazine (SOW formulation)
was fed to male and female beagle dogs for 105 weeks at nominal doses of 15,
150 or 1,500 ppm; based on measured analytical concentrations of 14.1, 141
and 1,410 ppm, however, these values correspond to approximately 0.35, 3.5
and 35 mg/kg/day (Lehman, 1959). Survival rate, body weight gain, food
intake, behavior, appearance, hematology, urinalysis, organ weights and
histology were determined. Ihe 15-ppm dosage (0.35 mg/kg/day) produced no
toxicity, but the 150-ppm dosage (3.5 mg/kg/day) caused a decrease in food
intake as well as increased heart and liver weight in females. In the group
receiving 1,500 ppm (35 mg/kg/day), there were decreases in food intake and
body weight gain, an increase in adrenal weight, a decrease in hematocrit and
occasional tremors or stiffness in the rear limbs. There were no differences
among the different groups in the histology of the organs studied. Based on
these results, a NOAEL of 0.35 mg/kg/day was identified for atrazine. This
NOAEL is supported by the available preliminary data by Ciba-Geigy (1985) on
a new two-year study in the Sprague-Dawley rats that will be completed for
the Agency review in the near future. This preliminary data reflected adverse
effects (mammary gland tumors) at 70 ppm (3.5 mg/kg/day) but no effects were
were noted at the lower dose level, 10 ppm (0.5 mg/kg/day). Other studies
(Woodard Research Corporation, 1966; Hazelton Laboratories, 1961) identified
long-term NOAEL values of 5 to 50 mg/kg/day and were not considered to be as
protective as the Woodard Research Corporation (1964) study in the dog for
use in calculating the HA values for atrazine.
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.35 mg/kg/day) = Q.0035 mg/kg/day
(100)
where:
0.35 mg/kg/day = NOAEL, based on the absence of adverse clinical,
hematological, biochemical and histopathological
effects in dogs.
100 = uncertainty factor, chosen in accordance with ODW/NAS
guidelines for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0035 mg/kg/day) (70 kg) = 0<123 /L (123 /L)
(2 L/day)
where:
0.0035 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
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Atrazine September, 1987
-14-
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.123 mg/L) (20%) = 0.0025 mg/L (3 ug/L)
10
where:
0.123 mg/L = OWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor, according to ODW policy,
to account for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 Preliminary data submitted by Ciba-Geigy Corporation (1985) in support
of the pesticide registration of atrazine indicate that atrazine
induced an increased incidence of mammary tumors in female Sprague-
Dawley rats. These findings have been further confirmed in a briefing
by Ciba-Geigy (1987) on the recently completed study. An evaluation
of this study will be performed in the near future.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of atrazine.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986b), atrazine may be classified in
Group C: possible human carcinogen. This category is used for
substances with limited evidence of carcinogenicity in animals in the
absence of human data. This classification is considered preliminary
until the Office of Pesticide Program completes a peer review of the
weight of the evidence for atrazine and its analogs. At present, ODW
has determined that at least one closely related analog, propazine,
is a group C oncogen based on an increased incidence of tumors in the
same target tissue (mammary gland) and animal species (rat) as was
noted for atrazine.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Toxicity data on atrazine were reviewed by the National Academy of
Sciences (NAS, 1977), and the study by Innes et al. (1969) was used
to identify a chronic NOAEL of 21.5 mg/kg/day. Although at that time
it was concluded that atrazine has low chronic toxicity, an uncertainty
factor of 1,000 was employed in calculation of the ADI from that
study, since only limited data were available. The resulting value
(0.021 mg/kg/day) corresponds to an ADI of 0.73 mg/L in a 70-kg adult
consuming 2 L of water per day.
0 Tolerances for atrazine alone and the combined residues of atrazine
and its metabolites in or on various raw agricultural commodities
have been established (U.S. EPA, 1986c). These tolerances range from
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Atrazine September, 1987
-15-
0.02 ppm (negligible) in animal products (meat and meat by-products)
to 15 ppm in various animal fodders.
VII. ANALYTICAL METHODS
0 Analysis of atrazine is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing pesti-
cides in water samples (U.S. EPA, 1986d). In this method, approximately
1 L of sample is extracted with methylene chloride. The extract is
concentrated, and the compounds are separated using capillary column
GC. Measurement is made using a nitrogen phosphorus detector. The
method detection limit has not been determined for this compound, but
it is estimated that the detection limits for the method analytes are
in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies which will remove atrazine from water include
activated carbon adsorption, ion exchange, reverse osmosis, ozone
oxidation and ultraviolet irradiation. Conventional treatment methods
have been found to be ineffective for the removal of atrazine from
drinking water (ESE, 1984; Miltner and Fronk, 1985a). Limited data
suggest that aeration would not be effective in atrazine removal
(ESE, 1984; Miltner and Fronk, 1985a).
0 Baker (1983) reported that a 16.5-inch GAC filter cap using F-300,
which was placed upon the rapid sand filters at the Fremont, Ohio
water treatment plant, reduced atrazine levels by 30 to 64% in the
water from the Sandusky River. At Jefferson Parish, Louisiana,
Lykins et al. (1984) reported that an adsorber containing 30 inches
of Westvaco WV-G* 12 x 40 GAC removed atrazine to levels below
detectable limits for over 190 days.
0 At the Bowling Green, Ohio water treatment plant, PAC in combination
with conventional treatment achieved an average reduction of 41% of
the atrazine in the water from the Maumee River (Baker, 1983).
Miltner and Fronk (1985a) reported that in jar tests using spiked
Ohio River water with the addition of 16.7 and 33.3 mg/L of PAC and
15-20 mg/L of alum, PAC removed 64 and 84%, respectively, of the
atrazine. Higher percent removals reflected higher PAC dosages.
Miltner and Fronk (1985b) monitored atrazine levels at water treat-
ment plants, which utilized PAC, in Bowling Green and Tiffin, Ohio.
Applied at dosages ranging from 3.6 to 33 mg/L, the PAC achieved 31
to 91% removal of atrazine, with higher percent removals again
reflecting higher PAC dosages.
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Atrazine September, 1987
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Harris and Warren (1964) reported that Amber lite IR-1 20 cation exchange
resin removed atrazine from aqueous solution to less than detectable
levels. Turner and Adams (1968) studied the effect of varying pH on
different cation and an ion exchange resins. At a pH of 7.2, 45%
removal of atrazine was achieved with Dowex® 2 anion exchange resin
and with H2?O4~ as the exchangeable ion species.
Chi an et al. (1975) reported that reverse osmosis, utilizing cellulose
acetate membrane and a cross-linked polyethelenimine (NS-100) membrane,
successfully processed 40% of the test solution, removing 84 and 98%,
respectively, of the atrazine in the solution.
Miltner and Fronk (1985a) studied the oxidation of atrazine with
ozone in both spiked distilled and ground water. Varying doses of
ozone achieved a 70% removal of atrazine in distilled water and 49 to
76% removal of atrazine in ground water.
Kahn et al. (1978) studied the effect of fulvic acid upon the photo-
chemical stability of atrazine to ultraviolet irradiation. A 50%
removal of atrazine was achieved much faster at higher pH conditions
than at lower pH conditions. In the presence of fulvic acids, the
time needed for ultraviolet irradiation to achieve 50% removal was
almost triple the time required to achieve similar removals without
the presence of fulvic acids. Since fulvic acids will be present in
surface waters, ultraviolet irradiation may not be a cost-effective
treatment alternative.
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Atrazine September, 1987
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IX. REFERENCES
Adler, I.D. 1980. A review of the coordinated research effort on the
comparison of test systems for the detection of mutagenic effects,
sponsored by the E.E.C. Mutat. Res. 74:77-93.
Arrington, L.R. 1972. The laboratory animals. In: Introductory laboratory
animal science. The breeding, care and management of experimental animals.
Danville, IL: Interstate Printers and Publishers, Inc., pp. 9-11.
Baker, D. 1983. Herbicide contamination in municipal water supplies in
northwestern Ohio. Final Draft Report 1983. Prepared for Great Lakes
National Program Office, U.S. Environmental Protection Agency. Tiffin, OH.
Bakke, J.E., J.D. Larson and C.E. Price. 1972. Metabolism of atrazine and
2-hydroxyatrazine by the rat. J. Agric. Food Chem. 20:602-607.
Bashmurin, A.F. 1974. Toxicity of atrazine for animals. Sb. Rab. Leningrad
Vet. Institute. 36:5-7. (English abstract only)
Beynon, K.I., G. Stoydin and A.N. Wright. 1972. A comparison of the
breakdown of the triazine herbicides cyanazine, atrazine and simazine
in soils and in maize. Pestic. Biochem. Physiol. 2:153-161.
Bohme, E., and F. Bar. 1967. Uber den Abbau von Triazin-Herbiciden in
tierischen Organismus. Food Cosmet. Toxicol. 5:23-28. (English abstract
only)
Bradway, D.E., and R.F. Moseman. 1982. Determination of urinary residue
levels of the n-dealkyl metabolites of triazine herbicides. J. Agric.
Food Chem. 30:244-247.
Buchanan, G.A., and A.E. Hiltbold. 1973. Performance and persistence of
atrazine. Weed Sci. 21:413-416.
Chian, E.S.K., W.N. Bruce and H.H.P. Fang. 1975. Removal of pesticides by
reverse osmosis. Environmental Science and Technology. 9(1):52-59.
Ciba-Geigy. 1971. Rat reproduction study-test for teratogenic or embryotoxic
effects. 10/1971; Teratology study of atrazine technical in Charles
River rats 9/1984, SCDFA, Sacramento.
Ciba-Geigy. 1983a. Dermal absorption of 14c-atrazine by rats. Ciba-Geigy
Corporation, Greensboro, NC. Report No. ABR-83005, May, 1983. Accession
No. 255815.
Ciba-Geigy. 1983b. Excretion rate of 14c-atrazine from dermally dosed rats.
Ciba-Geigy Corporation, Greensboro, NC. Report No. ABR-83081, October,
1983. Accession No. 255815.
Ciba-Geigy Ltd. 1984a. A teratology study of atrazine technical in Charles
River Rats: Toxicology/pathology report No. 60-84. MRID 00143008.
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Ciba-Geigy Ltd. 1984b. Segment II. Teratology study in rabbits: Toxicology/
pathology report No. 68-84. MRID 00143006.
Ciba-Geigy. 1985. Atrazine chronic feeding/oncogenicity study. One-year
interim report. May 17, 1985.
Ciba-Geigy. 1987. Briefing paper on atrazine. December, 1986. Analysis of
chronic rat feeding study results. Ciba-Geigy Corp., Greensboro, NC.
Cohen, S.2., C. Eiden and M.N. Lorber. 1986. Monitoring Ground Water for
Pesticides in the U.S.A. _In Evaluation of pesticides in ground water.
American Chemical Society Symposium Series, (in press).
Cosmopolitan Laboratories.* 1979. CBI, Document No. 00541, EPA Accession No.
2-41725.
CSE Laboratories.* 1980. CBI, Document No. 000850, EPA Accession No. 2-43485.
Darwent, A.L., and R. Behrens. 1968. Dissipation and leaching of atrazine
in a Minnesota soil after repeated applications. In Proc. North Cent.
Weed Control Conf., December 3-5, 1968, Indiana, pp. 66-68.
Dauterman, W.C., and W. Muecke. 1974. In vitro metabolism of atrazine by
rat liver. Pestic. Biochem. Physiol. 4:212-219.
ESE. 1984. Environmental Science and Engineering. Review of treatability
data for removal of 25 synthetic organic chemicals from drinking water.
U.S. Environmental Protection Agency, Office of Drinking Water, Washington,
DC.
Erickson, M.D., C.W. Frank and D.P. Morgan. 1979. Determination of s-triazine
herbicide residues in urine: Studies of excretion and metabolism in swine
as a model to human metabolism. J. Agric. Food Chem. 27:743-745.
Foster, T.S., S.U. Khan and M.H. Akhtar. 1979. Metabolism of atrazine by
the soluble fraction (105,000 g) from chicken liver homogenates.
J. Agric. Food Chem. 17:300-302.
Frear, E.H., ed. 1969. Pesticide index. State College, PA: College Science
Publications.
Gaines T.B., Linder, R.E. 1986. Acute toxicity of pesticides in adult and
weanling rats. Fundam. Appl. Toxicol. 7:299-308
Goswami, K.P., and R.E. Green. 1971. Microbial degradation of the herbicide
atrazine and its 2-hydroxy analog.
Harris, C.I., and G.F. Warren. 1964. Adsorption and desorption of herbicides
by soil. Weeds. 12:120-126.
Harris, C.I. 1967. Fate of 2-chloro-£-triazine herbicides in soil. J. Agric.
Food Chem. 15:157-162.
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Atrazine September, 1987
-19-
Hayes, W.J.,Jr. 1982. Pesticides studied in man. Baltimore, HD: Williams
and Wilkins.
Hazelton Laboratories.* 1961. Two-year chronic feeding study in rats.
CBI, Document No. 000525, MRID 00059211.
Helling, C.S. 1971. Pesticide mobility in soils. II. Applications of soil
thin-layer chromatography. Proc. Soil Sci. Soc. Am. 35:737-748.
Hurle, K., and E. Kibler. 1976. The effect of changing moisture conditions
on the degradation of atrazine in soil. Proceedings of the British Crop
Protection Conference—Weeds. 2:627-633.
Innes, J.R.M., B.M. Ulland, and M.G. Valeric. 1969. Bioassay of pesticides
and industrial chemicals for tumorigenicity in mice: A preliminary note.
J. Natl. Cancer Inst. 42:1101-1114.
Ivey, M.J., and H. Andrews. 1964. Leaching of simazine, atrazine, diuron,
and DCPA in soil columns. Unpublished study submitted by Ciba-Geigy,
Greensboro, N.C.
Ivey, M.J., and H. Andrews. 1965. Leaching of simazine, atrazine, diruon,
and DCPA in soil columns. Unpublished study prepared by University of
Tennessee, submitted by American Carbonyl, Inc., Tenafly, NJ.
Khan, S.U., and T.S. Foster. 1976. Residues of atrazine (2-chloro-4-ethyl-
air,ino-6-isopropylamino-s-triazine) and its metabolites in chicken tissues.
J. Agric. Food Chem. 24:768-771.
Khan, S.U., and M. Schnitzer. 1978. "UV irradiation of atrazine in aqueous
fulvic acid solution. Environmental Science and Health. 813:299-310.
Lavy, T.L. 1974. Mobility and deactivation of herbicides in soil-water
systems: Project A-024-NEB. Available from National Technical Information
Service, Springfield, VA; PB-238-632.
Lavy. T.L., F.W. Roeth and C.R. Fenster. 1973. Degradation of 2,4-D and atra-
zine at three soil depths in the field. J. Environ. Qual. 2:132-137.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food and Drug Off.
Loprieno, N., R. Barale, L. Mariani, S. Presciuttini, A.M. Rossi, I. Shrana,
L. Zaccaro, A. Abbondandolo and S. Bonatti. 1980. Results of mutagenicity
tests on the herbicide atrazine. Mutat. Res. 74:250.
Lykins, Jr., B.W., E.E. Geldreich, J.Q. Adams, J.C. Ireland and R.M. Clark.
1984. Granular activated carbon for removing nontrihaloraethane organics
from drinking water. U.S. Environmental Protection Agency, Office of
Research and Development, Municipal Environmental Research Laboratory,
Cincinnati, OH.
Meister, R.G., ed. 1987. Farm chemicals handbook. 3rd ed. Willoughby, OH:
Meister Publishing Co.
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Atrazine September, 1 987
-20-
Miltner, R.J., and C.A. Fronk. 1985a. Treatment of synthetic organic contami-
nants for Phase II regulations. Progress report. U.S. Environmental
Protection Agency, Drinking Hater Research Division. July 1985.
Miltner, R.J., and C.A. Fronk. 1985b. Treatment of synthetic organic contami-
nants for Phase II regulations. Internal report. U.S. Environmental
Protection Agency, Drinking Water Research Division. December 1985.
Molnar, V. 1971. Symptomatology and pathomorphology of experimental poisoning
with atrazine. Rev. Med. 17:271-274. (English abstract only)
Nurnik, M.R., and C.L. Nash. 1977. Mutagenicity of the triazine herbicides
atrazine, cyanazine, and simazine in Drosophila melanogaster. J. Toxicol.
Environ. Health. 3:691-697.
NAS. 1977. National Academy of Sciences. Drinking Water and Health.
Washington, DC: National Academy Press, pp. 533-539.
Newby, Lc, and B.C. Tweedy. 1976. Atrazine residues in major rivers and
tributaries. Unpublished study submitted by Ciba-Geigy Corporation,
Greensboro, N.C.
Palmer, J.S., and R.D. Radeleff. 1964. The toxicological effects of certain
fungicides and herbicides on sheep and cattle. Ann. N.Y. Acad. Sci.
111:729-736.
Palmer, J.S., and R.D. Radeleff. 1969. The toxicity of some organic herbicides
to cattle, sheep and chickens. Production Research Report No. 1066.
U.S. Department of Agriculture, Agricultural Research Service: 1-26.
Peters, J.W., and R.M. Cook. 1973. Effects of atrazine on reproduction in
rats. Bull. Environ. Contain. Toxicol. 9:301-304.
Portnoy, C.E. 1978. Disappearance of bentazon and atrazine in silt loam soil.
Unpublished study submitted by BASF Wyandotte Corporation, Parsippany, NJ.
Schlichter, J.E., and V.B. Beat. 1972. Dermatitis resulting from herbicide
use — A case study. J. Iowa Med. Soc. 62:419-420.
Smith, A.E., R. Grover, G.S. Emrnond and B.C. Korven. 1975. Persistence and
movement of atrazine, bromacil, monuron, and simazine in intermittently-
tilled irrigation ditches. Can. J. Plant Sci. 55:809-816.
STORET. 1987.
Talbert, R.E., and O.H. Fletchall. 1965. The adsorption of some S-triazines
in soils. Weeds. 13:46-52.
Turner, M.A., and R.S. Adams, Jr. 1968. The adsorption of atrazine and
atratone by anion- and cation-exchange resins. Soil Sci. Amer. Proc.
32:62-63.
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Atrazine September, 1987
-21-
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Atrazine chronic
feeding/oncogenicity study preliminary incidence table of tumors regarding
possible section 6(a)(2) effect. Washington, DC: U.S. EPA Office of
Pesticide Programs.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guideline for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986c. U.S. Environmental Protection Agency. Code of Federal
Regulations. Protection of the environment. Tolerances and exemptions
from tolerances for pesticide chemicals in or on raw agricultural commodi-
ties. 40 CFR 180.220. p. 216.
U.S. EPA. 1986d. U.S. Environmental Protection Agency. Method #1 - Determina-
tion of nitrogen and phosphorus containing pesticides in ground water by
GC/NPD. January, 1986 draft.
Warnock , R.E., and J.B. Leary. 1978. Paraquat, atrazine and Bladex—dissipa-
tion in soils. Unpublished study prepared by Chevron Chemical Company,
submitted by Shell Chemical Company, Washington, DC.
Weidner, C.H. 1974. Degradation in groundwater and mobility of herbicides.
Master's thesis. University of Nebraska, Department of Agronomy.
Wolf, D.C., and J.P. Martin. 1975. Microbial decomposition of ring-14C-
atrazine, cyanuric acid, and 2-chloro-4,6-diamino-S-triazine. J. Environ.
Qual. 4:134-139.
Woodard Research Corporation.* 1964. Two-year feeding study in dogs. CBI,
Document No. 000525, MRID 00059213.
Woodard Research Corporation.* 1966. Three-generation reproduction study in
in rats. CBI, Document No. 000525, MRID 00024471.
Yoder, J., M. Watson and W.W. Benson. 1973. Lymphocyte chromosome analysis
of agricultural workers during extensive occupational exposure to
pesticides. Mutat. Res. 21:335-340.
Windholz, M., ed. 1976. The Merck index. 9th ed. Rahway, NJ: Merck and
Co., Inc.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
DRAFT
BAYGON (Propoxur)
August, 1987
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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Baygon August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 114-26-1
Structural Formula
2-(1-Methylethoxy)-phenol methylcarbamate
Synonyms
0 Propoxur (proposed common name); Aprocarb; Blattenex; BAY 39007;
Bayer 39007; Pillargon; Propyon; Suncide; Tugon; QMS 33; Unden
(Meister, 1984).
Uses
e
A nonfood insecticide used on humans, animals and turf grass
(Meister, 1984).
Properties (ACGIH, 1984; Meister, 1984; and CHEMLAB, 1985)
Chemical Formula C
Molecular Weight 209.24
Physical State (at 2S°C) White to tan crystalline solid
Boiling Point
Melting Point 91 °C
Density (°C)
Vapor Pressure (120°C) 0.1 mmHg
Water Solubility (20°C) 2000 mg/L
Log Octanol/Water Partition 0.14
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Baygon has been found in none of the 58 ground water samples analyzed
from 55 locations. No surface water samples were analyzed (STORET,
1987).
Environmental Fate
(Forthcoming from OPP)
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Baygon August, 1987
-3-
III. PHARMACOKINETICS
Absorption
0 Vandekar et al. (1971) administered a single oral dose of 1.5 mg/kg
of propoxur, 95% active ingredient (a.i.), to a 42-year-old male
volunteer. About 45% of the dose was recovered in urine within
24 hours as o-isopropoxyphenol. Since vomiting occurred 23 minutes
after ingestion, the authors assumed that much of the dose was expelled
by this route, so the percent actually absorbed could not be calculated.
0 Chemagro Corp. (no date) investigated the dermal absorption of 14C-
labeled Baygon in human subjects. Baygon (4 ug/cm2, total dose less
than 1 mg) was applied to the forearm of the subjects(s) in four tests:
(1) application to the skin without preparation, (2) application
after stripping of the skin with an adhesive tape, (3) application
followed by occlusion and (4) application followed by induction of
sweating. The amounts excreted (route not specified, but presumably
in urine) after these treatments were 20, 51, 64 and 18%, respectively,
indicating that Baygon is well absorbed through the skin.
0 Krishna and Casida (1965) administered single oral doses of 14C-labeled
Baygon (50 mg/kg) to Sprague-Dawley rats. After 48 hours, about 4%
of the dose had been excreted in feces, and the remainder was detected
in urine (64 to 72%), expired air (26%) or the body (4.2 to 7.9%).
This indicated that Baygon had been well absorbed (at least 96%) from
the gastrointestinal tract. Similar findings were reported by Foss
and Krechniak (1980).
Distribution
0 Foss and Krechniak (1980) investigated the fate of Baygon after oral
administration of 50 mg/kg to male albino rats. Analysis of tissues
indicated that Baygon levels were greatest in the kidneys, with
somewhat lower levels in the liver, blood and brain.
Metabolism
0 Dawson et al. (1964) administered single oral doses of 92.2 mg of
Baygon (purity not specified) to six male volunteers, and single oral
doses of 50 mg to three subjects. Urine samples were collected and
analyzed for metabolites. A material identified as 2-isopropoxyphenol
was observed in the urine of both groups. Similar results were
reported by Vandekar et al. (1971).
0 Foss and Krechniak (1980) investigated the metabolism of Baygon after
both oral and intravenous administration of 50 mg/kg to male albino
rats. Isopropoxyphenol was detected in tissues 10 minutes following
administration, and the highest concentrations were attained between
30 and 60 minutes after dosing. This metabolite prevailed in the
blood and liver, but in the kidney only unchanged Baygon could be
detected. Eight hours postdosing, only traces of Baygon and its
metabolites were detected in these tissues.
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Saygon August, 1987
-4-
0 Everett and Gronberg (1971) studied the metabolism of Baygon in
Holtzman rats. Animals were dosed by gavage with Baygon (5 to 10
mg/kg) labeled with ^C or 3H in the carbonyl or the isopropyl groups.
Pooled urine from eight rats (four/sex) dosed with 20 mg/kg/day of
unlabeled Baygon for 4 days was used to isolate sufficient quantities
of metabolites for identification of structure. Results indicated
that the major pathway of Baygon metabolism involved depropylation to
2-hydroxyphenol-N-methyl carbamate and hydrolysis of the carbamate to
isopropoxyphenol. Minor pathways involved ring hydroxylation at the
five- or six-position, secondary hydroxylation of the 2'-carbon of
the isopropoxy group and N-methyl hydroxylation. Metabolites that
contained the 6-hydroxy group formed N-conjugates, while those that
contained the 5-hydroxy group formed O-glucuronides.
Excretion
Oawson et al. (1964) reported that in humans given a single oral
doses of 92.2 mg Baygon (purity not specified), 38% of the dose was
excreted as phenols in urine over the next 24 hours; most was excreted
in the first 8 to 10 hours.
Krishna and Casida (1965) administered single oral doses of 50 mg/kg
of 14c-carbonyl-labeled Baygon to Sprague-Dawley rats. After 48
hours, recovery of label in excretory products was as follows: 64%
(males) and 72% (females) in urine; 4% in feces (males and females);
and 26% in expired carbon dioxide (males and females). Residual
label in the body was 4.2% (males) and 7.9% (females). One-third of
the excreted dose was hydrolyzed, with most of the remainder being
intact.
Everett and Gronberg (1971) reported that 85% of orally administered
14c-carbonyl-labeled Baygon (5 to 8 mg/kg) was recovered from Holtzman
rats within 16 hours of dosing; 20 to 25% of tne radioactivity appeared
in the expired air, and 60% of the radioactivity appeared in the
urine as conjugates. Also, Foss and Krechniak (1980) indicated that
85 to 95% of an oral dose (50 mg/kg) administered to male albino rats
was excreted in urine with a half-life of 0.18 to 0.26 hour.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 Vandekar et al. (1971) studied the acute oral toxicity of Baygon in
human volunteers. A 42-year-old man ingested a single oral dose of
1.5 mg/kg of propoxur (Baygon) (95% a.i., recrystallized). Cholinergic
symptoms, including blurred vision, nausea, sweating, tachycardia and
vomiting, began about 15 to 20 minutes after exposure. Effects were
transient and disappeared within 2 hours. Cholinesterase (ChE)
activity (measured spectrophotometrically) in red blood cells decreased
to 27% of control values by 15 minutes after exposure, and returned
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Baygon ' 1987
-5-
to control levels by 2 hours. No effect was detected in plasma ChE
activity. In a second test, a single dose of 0.36 mg/kg caused
short-lasting stomach discomfort, blurred vision and moderate facial
redness and sweating. Red blood cell ChE activity fell to 57% of
control values within 10 minutes, then returned to control levels
within 3 hours.
0 Vandekar et al. (1971) administered five oral doses of 0.15 or 0.20
mg/kg to male volunteers at half-hour intervals (total dose of 0.75
or 1.0 mg/kg). In each subject, a symptomless depression of red
blood cell ChE was observed; the lowest level, about 60% of control
values, was reached between 1 and 2 hours following doses 3, 4 and 5.
After the final dose, red blood cell ChE activity rose to control
levels within about 2 hours. The authors noted that a dose of Baygon
was tolerated better if it was divided into portions and given over
time than if it was given as a single dose.
Long-term Exposure
0 Davies et al. (1967) described the effects of a large-scale spraying
operation in El Salvador in which Baygon (OMS-33, 100% a.i.) was
used. The trial was planned so that medical assistance would be
available, and appropriate clinical support could be provided to
those affected by the spraying. The total amount of OMS-33 sprayed
was 345 kg. Among the spraymen, exposure (expressed in person-days)
was 70.5; 19 experienced symptoms (26% incidence). In the general
population, the exposure was 3,340 person-days, and 35 experienced
symptoms (1% incidence). The primary symptoms were headache, vomiting
and nausea. In the spraymen, the symptoms occurred mostly in the
first days, with no visible symptoms after this time. In severe
cases, atropine was administered as antidote. It was concluded that
the acute toxicity symptoms were observed in a low incidence, and
they were, in general, mild, evanescent, reversible, responsive to
small doses of atropine, and tended to occur at the beginning of the
spray program.
0 Montazemi (1969) reported on the toxic effects of Baygon on the
population of 26 villages in Iran that were sprayed with Baygon at
the rate of 2 g/m2 daily for 18 days. Selected inhabitants from six
villages and sprayers were examined on days 2, 8 and 18 and after
the completion of the spraying. Depression of ChE activity was found
in the inhabitants and in the sprayers, but the sprayers generally
had more severe symptoms. Atropine or belladonna was adequate to
treat those exhibiting symptoms.
Animals
Short-term Exposure
0 The acute oral LD5Q value for technical Baygon (purity not spedified)
in male and female Sherman rats was reported to be 83 and 86 mg/kg,
respectively (Games, 1969). The oral LD50 was reported to be 32 mg/kg
in mice and 40 mg/kg in guinea pigs (NIOSH, 1983).
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Baygon August, 1987
-6-
0 Farbenfa'briken Bayer (1961) determined an oral LDso °f 100 to 150 rag/kg
(purity not specified) in male albino rats. Severe muscle spasms
were observed, but no dose-response information was provided.
0 Eben and Kimmerle (1973) studied the acute toxicity of Baygon in
SPF-Wistar rats. Single oral doses of propoxur (98.7% a.i.), diluted
with propylene glycol, were given by gavage to groups of three male
rats at levels of 15, 20, 40 or 60 mg/kg; female rats were given
doses of 10, 20, 40 or 60 mg/kg. Cholinesterase levels were measured
in plasma, erythrocytes and brain at 10, 20 and 180 minutes after
dosing. Maximum ChE depression was observed at 10 and 20 minutes in
the plasma and erythrocytes, and at 180 minutes in the brain. The
inhibition was dose-dependent and a no-effect level was not observed.
In plasma, ChE was inhibited from 19% (low dose) to 63% (high dose)
in males and from 0 to 32% in females. In erythrocytes, ChE was
inhibited from 27 (low dose) to 63% (high dose) in males and from 15
to 45% in females. Based on ChE inhibition, this study identified a
Lowest-Observed-Adverse-Effect-Level (LOAEL) of 10 mg/kg/day.
0 Farbenfabriken Bayer (1966) conducted a 9-week feeding study with
Bay 39007 (purity not specified) in male and female rats (Elberfeld FB).
Baygon was included in the diets of the male animals at dose levels
of 0, 1,000, 2,000, 4,000 or 8,000 ppm. Based on the assumption that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), this corresponds to doses of 0, 50, 100, 200 or 400 mg/kg/day,
respectively. Females were given only one dose (4,000 ppm). The study
was begun when the animals (15/dose level) were 4 weeks of age and
weighed about 48 g. In males, food consumption and body weight were
depressed in a dose-dependent manner. At the 4,000 and 8,000-ppm
levels, the males were less lively and exhibited slightly shaggy
coats. Gross pathologic examinations of all animals were conducted.
Two males exposed to 4,000 ppm died during the study, one at 11 days
(evidence of myocarditis) and one at 23 days. Two males also died
at the 8,000-ppm level (at 23 and 25 days); one showed necrotic
inflammation of the mucosa of the small intestine. Females (exposed
to 4,000 ppm only) displayed decreased food consumption and reduced
weight gain similar to that seen in exposed males. One of 15 female
controls died at day 12 (death attributed to pneumonia), and two of
15 exposed females died, one at 7 days and one at 45 days (in this
rat there was suppuration of the cerebellar bottom). There were
apparently no measurements of ChE activity or other clinical tests
performed during this study. It was concluded by the authors that
the observed pathology could not be directly attributed to the presence
of Baygon in the diet. Based on gross observations, the No-Observed-
Adverse-Effect-Level (NOAEL) for male animals was identified as
2,000 ppm (100 mg/kg/day) and the LOAEL as 4,000 ppm (200 mg/kg/day).
In females, 4,000 ppm (200 mg/kg/day, tho only dose tested) was a
toxic level.
0 Eben and Kimmerle (1973) exposed SPF-Wistar rats (four/sex/dose) by
gavage to doses of 3, 10 or 30 mg/kg/day of Baygon for 4 weeks. The
high-dose animals (30 mg/kg/day) displayed cholinergic symptoms.
Cholinesterase activity in plasma and red blood cells, measured 15
-------
Baygon ' 1987
-7-
minutes after dosing on days 3, 8, 14, 21 and 28, was generally
depressed in a dose-related manner at 10 and 30 mg/kg, but not at the
3-mg/kg dose. For example, on day 28, ChE activity in plasma was
reduced by 0, 21 or 27% in males and by 14, 27 or 41% in females. In
erythrocytes, ChE was inhibited by 9, 24 or 32% in males and by 11,
32 or 43% in females. No cumulative toxic effects were observed.
Based on ChE inhibition, the NOAEL for this study was 3 mg/kg/day,
and the LOAEL was 10 mg/kg/day.
Dermal/Ocular Effects
0 The acute dermal LD50 of technical Baygon (purity not specified) was
reported to be greater than 2,400 mg/kg for both male and female
Sherman rats (Gaines, 1969).
0 Crawford and Anderson (1971) indicated that 500 mg of technical
Baygon (purity not specified, dissolved in acetone) did not cause
any skin irritation within 72 hours of its application to the abraded
or unabraded skin of mature New Zealand White rabbits (six/group).
• Heimann (1982) demonstrated that Baygon (98.8% pure) is not a skin
sensitizer when tested in guinea pigs.
• Crawford and Anderson (1971) instilled 100 mg of technical Baygon
(purity not specified) in the left eye of six rabbits. Examination
at 48 and 72 hours revealed no evidence of ocular irritation or
corneal damage.
Long-term Exposure
e Eben and Kimmerle (1973) fed propoxur (98.7% a.i.) to male rats in
the diet for 15 weeks. Doses were 0, 250, 750 or 2,000 ppm. Assuming
that 1 ppm in the diet is equivalent to 0.05 mg/kg/day (Lehman, 1959),
this corresponds to doses of about 0, 12.5, 37.5 or 100 mg/kg/day.
Assays for ChE activity in plasma, erythrocytes and brain showed no .
constancy of inhibition and no dependence on the administered dose.
No other details were given.
8 Root et al. (1963) studied the effect of Bayer 39007 added to the
diet of Sprague-Dawley rats for 16 weeks. The rats (12/sex/dose,
weighing 72 to 145 g at the start of the feeding trial) were fed Baygon
(technical, 95.1% pure) at dose levels of 0, 100, 200, 400 or 800 pptr.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 0, 5, 10, 20 or 40 mg/kg/day,
Biweekly measurements revealed no changes in growth or food consump-
tion. Cholinesterase was assayed in blood, brain and submaxillary
glands of five animals of each sex at each dose level, and no inhi-
bition was detected. Necropies were performed on five animals of
each sex at the termination of the study, and no significant pathology
was found. It was concluded that the NOAEL for the rats was greater
than 800 ppm (40 mg/kg/day, the highest dose tested).
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Baygon *»*»**' 1987
-8-
0 Suberg and Loeser (1984) conducted a chronic (106-week) feeding study
of Baygon (99.4% a.i.) in rats (Elberfeld strain) at dose levels of
0, 200, 1,000 or 5,000 ppm. Based on the assumption that 1 ppm in
the diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), this
corresponds to doses of about 0, 10, 50 or 250 mg/kg/day. There were
50 rats of each sex per dose level, plus an additional 10 of each sex
for interim autopsies at the end of the first year. At the 200-ppm
dose, there was no effect on food consumption or body weight, there
were no cholinergic signs, and clinical chemistry, gross pathology,
histopathology and organ weights showed no changes from control
values. At 1,000 ppm, retarded weight gain was observed in males
during the first 20 weeks. At 1,000 and 5,000 ppm, there were
significant hyperplasia of urinary bladder epithelium (described in
more detail in the Carcinogenicity section) and increased incidence
of neuropathy. At the 5,000-ppm dose, both weight gain and food
consumption were significantly retarded throughout the study; males
showed increased thromboplastin time, and females had consistently
lower mean plasma ChE activity than did controls or other test groups.
Both sexes showed some degree of splenic atrophy, but there were no
other significant changes in other organs. Based on body weight
gain, the NOAEL for this study was identified as 200 ppm (10 mg/kg/day),
and the LOAEL as 1,000 ppm (50 mg/kg/day).
0 Loser (1968a) conducted a.2-year feeding study of Baygon in male and
female SPF-Wistar rats. Starting at 1 month of age, the test material,
BAY 39007 (99.8% a.i., technical), was included in the diet at levels
of 0, 250, 750, 2,000 or 6,000 ppm. Based on the assumption that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), this corresponds to doses of 0, 12.5, 37.5, 100 or 300 mg/kg/day.
The control group consisted of 50 animals of each sex, while test
groups contained 25 animals of each sex. Growth and behavior were
observed, liver function and ChE activity were tested, and blood and
urine were analyzed periodically. Necropsies on five animals of each
sex were conducted at the termination of the experiment. The major
adverse effects noted were low food consumption and low body weight
in all animals at the 6,000-ppm dose level, and low body weight in
the female (but not male) animals at the 2,000-ppm dose level.
Cholinesterase determinations on blood (measured at the high dose
only) revealed no changes; ChE activity was 9.8 and 9.9 units in
control males and females, respectively, compared with 9.9 and 10.0
in exposed males and females. The author indicated that the methodology
may have been too insensitive to detect small changes that may have
occurred. No spasms or other symptoms of ChE inhibition were observed.
No impairment of liver or kidney function was detected by clinical
tests, but necropsy revealed increased liver weight at all doses
greater than 250 ppm. Results of blood analysis were normal at all
dose levels except at 6,000 ppm. Apart from increased liver weights,
necropsy findings were unremarkable. Based on increased liver weights,
this study identified a NOAEL of 250 ppm (12.5 mg/kg/day) and a LOAEL
of 750 ppm (37.5 mg/kg/day).
0 Loser (1968b) conducted a 2-year study of Baygon toxicity in beagle
dogs. The product, BAY 39007 (technical, 99.8% pure), was included in
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Baygon August. 1987
-9-
the diet at levels of 0, 100, 250, 750 or 2,000 ppm. Assuming that
1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day (Lehman,
1959), this corresponds to doses of about 0, 2.5, 6.25, 18.7 or
50 mg/kg/day. The study was begun when the dogs (four/sex/dose) were
4 to 5 months old. Observations on the animals included weight and
food consumption at periodic intervals, ChE determinations in blood
at 16 weeks, clinical evaluations of blood and urine, and tests for
liver and kidney function. Necropsies were performed on animals that
died during the study and at termination of the study. The appearance,
behavior, and food consumption of dogs at the 100, 250 or 750 ppm
levels were comparable to those of the controls. At the 2,000-ppm
level, dogs of both sexes appeared to be weak and sick. One of the
males and all four females at this dose died before completion of the
study. During the first 6 months, dogs at this dose level exhibited
quivering and spasms, particularly in the abdominal region, and food
consumption was less than for the controls (especially in females);
as expected, the dogs showed statistically significant depression
in weight gain compared with the controls. Males, but not females,
showed lower weights than did controls at the 750-ppm dose level, but
the decrease was not statistically significant. Clinical analyses
did not reveal any aberrations in the blood or any changes in liver
or kidney function. However, increased liver weights were observed
at necropsy, and serum electrophoresis performed at the time of
sacrifice revealed decreased levels of some serum proteins, inter-
preted by the author as reflecting impaired protein synthesis'.
Cholinesterase determinations in whole blood at 16 weeks did not
reveal any significant inhibition of activity. In males, ChE inhibi-
tion at 100, 250, 750 and 2,000 ppm was 0, 11, 1 and 13%, respectively,
and in females ChE inhibition was 0, 10, 7 and 0%, respectively. The
author indicated that the assay method may have been too insensitive
to detect small changes that may have occurred. Emaciation was the
principal finding in dogs that died during the study; one female had
abnormal liver parenchyma. The NOAEL for this study was 250 ppm
(6.25 mg/kg/day), and the LOAEL (based on increased liver weight,
decreased body weight and altered blood proteins) was 750 ppm (18.7
mg/kg/day).
0 Bomhard and Loeser (1981) conducted a 2-year feeding study of propoxur
(99.5% a.i.) in SPF CFI/W71 mice at dose levels of 0, 700, 2,000 or
6,000 ppm. Assuming that 1 ppm in the diet of mice is equivalent to
0.15 mg/kg/day (Lehman, 1959), this corresponds to doses of about 0,
105, 300 or 900 mg/kg/day. Mice were 5 to 6 weeks of age, weighing
22 to 25 g at the beginning of the study; each group consisted of 50
animals of each sex, plus an additional 10/sex/group included for
interim autopsy at 1 year. Body weight gain was slightly depressed
in male mice at the 6,000-ppm level. Apart from this observation,
all aspects of behavior, appearance, food intake, weight and mortality
were comparable to control values. Clinical chemistry and blood
studies, including glucose and cholesterol levels, were within the
normal range for all groups, and there were no significant gross
pathological or histopathological findings that could be attributed
to the ingestion of Baygon. It was concluded that the male mice
tolerated the pesticide at levels up to and including 2,000 ppm.
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Baygon August, 1987
-10-
while the female mice tolerated doses up to and including 6,000 ppm
without adverse effects. Based on these conclusions, the NOAEL for
this study was 2,000 ppm (300 ing/kg/day), and the LOAEL (based on
depressed weight gain in males) was 6,000 ppm (900 mg/kg/day).
Reproductive Effects
0 No multigeneration studies of the effects of Baygon on reproductive
function of animals were found in the available literature.
0 In a developmental toxicity study in rabbits, Schlueter and Lorke
(1981) observed no adverse effects on several reproductive end points.
This study is described below.
Developmental Effects
0 Schlueter and Lorke (1981) studied the effect of propoxur (99.6% a.i.)
on Himalayan CHBBrHM rabbits during gestation. Propoxur was admini-
stered by gavage (in 0.5% cremophor) to 15 animals/dose at 0, 1, 3
or 10 rag/kg. No adverse effects were observed in the dams, and no
changes were detected in implantation index, mean placental weight,
resorption index or litter size. Embryos were examined for visceral
and skeletal defects grossly, then were stained with Alizarin, and
transverse sections were prepared using the Wilson technique. No
adverse fetal effects were found at any dose level with respect to
mean fetal weight, the percent of stunting, the percent of slight
skeletal deviations, or the malformation index. These results indicate
that the NOAEL for maternal toxicity, teratogenicity and fetotoxicity
is greater than 10 mg/kg/day (the highest dose tested).
0 Lorke (1971) fed Baygon (technical, 98.4% a.i., 0.82% isopropoxyphenol)
in the diet to female FB-30 rats on days 1 to 20 of gestation, at
levels of 0, 1,000, 3,000 or 10,000 ppn (10/dose). Assuming that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day, (Lehman,
1959), this corresponds to doses of about 50, 150 or 500 mg/kg/day.
The rats were 2.5 to 3.5 months of age, weighing 200 to 250 g at the
time of the experiment. Cesarean sections were performed on day 20.
External and internal examinations on fetuses were performed, and
fetuses were subjected to skeletal staining. At the 3,000- and
10,000-ppm dose levels, average fetal weights were significantly
lower than control values, but ether fetal measurements were in the
control range. No f.erata were observed at a higher incidence than in
the control group. Data on fetal ossification were not adequately
described for an adequate evaluation. Although this study appears to
reflect a NOAEL of 1,000 ppn (50 mg/kg/day) based on fetotoxic effects,
information obtained from this study is limited due to the small
number of animals tested and an apparent dose-rel-ated decrease in
maternal weight gain and fetal weight at the lowest dose tested
(although these effects were not statistically significant).
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Baygon August, 1987
-11-
Mutagenicity
0 DeLorenzo et al. (1978) evaluated the mutagenic properties of Baygon
and other carbamate pesticides by use of the Salmonella mutagenicity
test of Ames. In assays using five strains of Salmonella typhimurium,
no mutagenic activity was obtained with Baygon (with microsomal
activation).
0 Moriya et al. (1983) tested Baygon in five strains of jS. typhimurium
and one strain of Escherichia coli using the Ames technique (without
metabolic activation) and observed no evidence of mutagenic activity.
0 Blevins et al. (1977) used five mutants of £. typhimurium LT2 to
examine the mutagenic properties of Baygon and other methyl carbamates
and their nitroso derivatives. No mutagenic activity was observed
with Baygon in this experiment using the Ames technique.
Carcinogenicity
0 Suberg and Loeser (1984) conducted a chronic (106-week) feeding study
of Baygon (99.4% a.i.) in rats (Elberfeld strain) at dose levels of
0, 200, 1,000 or 5,000 ppm. Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 10, 50 or 250 mg/kg/day. The study utilized 50
rats/sex/dose, plus an additional 10 of each sex included for interim
necropsies at the end of the first year. At 5,000 ppm there was
significant hyperplasia of the urinary bladder epithelium was noted.
The incidence at this dose level after 2 years was 44/49 in males and
48/48 in females, as compared with 1/49 and 0/49 in control males and
females, respectively. At 1,000 ppm, there was a smaller increased
incidence (10/50 and 5/49 in males and females), respectively. No
significant effect occurred at 200 ppm (1/50 and 0/49, males and
females, respectively). Bladder papillomas were observed in both
males (26/57) and females (28/48) at the highest dose after 2 years.
In addition, at the 5,000-ppm level, carcinoma of the bladder was
found in 8/57 males and 5/48 females, and carcinoma of the uterus was
seen in 8/49 females, as compared with 3/49 for the control group.
At the mid-dose level (1,000 ppm) only papillomas were noted in one
male. The tumors of significance in this study are the uncommon
bladder tumors (carcinoma and papillomas) with high incidences at the
high dose level. The combined tumor incidences were 34/57 males and
33/48 females at 5,000 ppm; 1/59 males and 0/48 females at 1,000 ppm.
and none in the 200-ppm or control groups.
0 Bomhard and Loeser (1981) conducted a 2-year feeding study of propoxur
(99.5% a.i.) in SPF CFI/W71 mice at dose levels of 0, 700, 2,000 or
6,000 ppm. Assuming that 1 ppm in the diet is equivalent to 0.15
mg/kg/day (Lehman, 1959), this corresponds to doses of about 0, 105,
300 or 900 mg/kg/day. Mice were 5 to 6 weeks of age, weighing 22 to
25 g at the beginning of the study; each group consisted of 50 animals
of each sex, plus an additional 10/sex/group included for interim
necropsy at 1 year. Gross and histological examination of tissues
revealed no evidence of increased tumor frequency.
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Baygon August, 1987
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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 ing/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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
The study by Vandekar et al. (1971 ) has been selected to serve as the
basis for determination of the One-day HA for Baygon. In this study, human
volunteers who ingested single oral doses of 0.36 or 1.5 mg/kg displayed
transient cholinergic signs accompanied by marked (43 and 75%, respectively)
inhibition of red blood cell ChE (measured 10 to 15 minutes after exposure).
Total doses of 0.75 or 1.0 mg/kg administered in five equal portions over 2
hours did not cause clinical signs, but inhibited red blood cell ChE by about
40%. A NOAEL was not identified; 0.36 mg/kg is taken as the LOAEL for bolus
exposure, and 0.45 mg/kg (three-fifths of a 0.75-mg/kg/day total dose,
administered in the first 3/5 doses) is the LOAEL when exposure to this
dose is spread over several hours. It should be noted that both values are
considerably lower than the NOAEL values for Baygon identified in subchronic
and chronic feeding studies in animals, especially rodents. Possible reasons
for this disparity are that humans may be more sensitive to this chemical
than animals are; furthermore, single oral doses probably produce higher peak
inhibitions than if the same total dose is ingested over a longer period of
time. It is also likely that measurement of ChE activity 10 to 15 minutes
after exposure (as in the case of human studies) detects peak inhibition,
while sampling later reveals smaller effects (due to the reversible nature of
ChE inhibition with carbamates). Since a child's exposure is more likely to
occur in a manner similar to Vandekar's test, where doses were administered
in five equal portions over time, the LOAEL of 0.45 mg/kg (three-fifths of a
0.75 mg/kg total dose) is used for the calculation below:
The One-day HA for a 10-kg child is calculated as follows:
One-day HA = (0.45 mg/kg/day) (10 kg) = 0>045 /L (40 /L)
(100) (1 L/day)
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Baygon
August, 1987
-13-
where:
0.45 mg/kg/day
10 kg
100
LOAEL, based on an inhibition of 40% in red blood cell
ChE activity in humans as determined 10 minutes after
oral exposure to three-fifths of a 0.75-mg/kg dose,
each fifth given at half-hour intervals, and based on
the fact that complete recovery of the ChE activity
occurred within 2 hours after administration of the
last fifth of the total dose.
assumed body weight of a child.
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.
Ten-day Health Advisory
In addition to human studies by Vandekar et al. (1971) discussed above,
two studies were considered for determination of the Ten-day HA. In a tera-
tology study in rabbits by Schlueter and Lorke (1981), the NOAEL appeared to
be higher than 10 mg/kg/day, the highest dose tested. In a teratology study
in rats by Lorke (1971), the dietary administration of Baygon to animals
during gestation was designed to assess both maternal and fetal effects.
While sufficient data were obtained to derive a NOAEL of 50 mg/kg/day and
a LOAEL of 150 mg/kg/day in rats, it is important to note that a dosage of
50 mg/kg/day was sufficient to kill all female animals in a chronic study in
dogs by Loser (1968b); all deaths occurred before the end of the 2-year study
period. Because humans appear to be more sensitive to Baygon than animals,
the human study by Vandekar et al. (1971), used in the determination of the
One-day HA value, is also the most suitable study for calculation of the Ten-day
HA. The two LOAELs identified in this study, 0.36 mg/kg (bolus exposure) and
0.45 mg/kg/day (exposure to three-fifths of a 0.75-mg/kg total dose spread
out over the day) can be approximated to 0.40 mg/kg; this value is used below
for calculation of the Ten-day HA.
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-dav HA = (0.40 mg/kg/day) (10 kg) = Q.040 mg/L (40 ug/L)
y (100) (1 L/day)
where:
0.40 mg/kg/day = LOAEL, based on mild cholinergic signs and 40%
inhibition of red blood cell ChE in humans 10 minutes
after a single oral dose.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance NAS/ODW
guidelines for use with a LOAEL from a human study.
1 L/day = assumed daily water consumption of a child.
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Baygon August, 1987
-14-
Longer-term Health Advisory
No suitable information was found in the available literature for the
determination of the Longer-term HA value for Baygon. It is, therefore,
recommended that the modified Drinking Water Equivalent Level (DWEL) of
40 ug/L for a 10-kg child be used as a conservative estimate for a Longer-term
exposure. The DWEL of 100 ug/L, calculated below, should be used for the
Longer-term value for a 70-kg 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. 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study in dogs by Loser (1968b) and the human study
by Vandekar et al. (1971) have been considered for determination of the
Lifetime HA. In the 2-year dog study by Loser (1968b), the chronic NOAEL was
identified as 6.25 mg/kg/day and the LOAEL as 18.7 mg/kg/day. The dog NOAEL
value is supported by the data of Loser (1968a) and of Suberg and Loeser
(1984), which identified NOAEL values of 12.5 and 10 mg/kg/day, respectively,
in chronic studies in rats. However, t^e dog appears to be far more sensitive
at the higher doses than are rodents; all female dogs and some of the males
in the high-dose group, 50 mg/kg/day, died before the end of the study
period, while mild systemic toxicity was noted at this dose level in rats.
Cholinesterase determinations were not performed in the dog study for use in
comparison with human data. Due to the reversible nature of ChE inhibition
by carbamates, a large difference is noted between the dosages that can cause
biologically significant levels of ChE inhibition and the dosages that can
produce cholinergic symptoms of toxicity (including death). Hence, in the
absence of ChE data in the dog study, and because of the sensitivity of this
end point in the determination of the toxicity of this chemical, the study by
Vandekar et al. (1971) in humans has been selected to serve as the basis for
the Lifetime HA for Baygon. This study was discussed in the previous sections
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Baygon ' 1987
-15-
on the One-day and Ten-day HAs. The 2-year mouse study by Bomhard and Loeser
(1981) was not considered, since the data suggest that the mouse is even less
sensitive than the rat.
Using a human ChE LOAEL of 0.36 mg/kg/day, the Lifetime HA is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.36 mg/kg/day) = Q.004 mg/kg/day
(100)
where:
0.36 mg/kg/day = LOAEL, based on mild cholinergic signs and 43%
inhibition of red blood cell ChE in a human 10 minutes
after a single oral dose.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from a human study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
nwEL = (0*004 mg/kg/day) (70 kg) = Q.140 mg/L (140 ug/L)
(2 L/day)
where:
0.004 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 - (0.140 mg/L) (20%) = 0.003 ng/L (3 ug/L)
(10)
where:
0.140 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor in accordance with ODW
policy* to account for possible carcinogenicity.
*This policy is used only for group C oncogen. However, since there is a
potential that tnis chemical may be a more potent oncogen, its oncogenic
potency (q-j*) was calculated using the multistage model (U.S. EPA, 1987a).
The q, * was estimated to be 7.9 x 10~3 (mg/kg/day)~1; if the oncogenic risk
level associated with the above determined Lifetime HA value is computed
using this q,*, the risk level would be 7 x 10~7 (7 in 10,000,000).
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Baygon Au*ust' 1987
-16-
Evaluation of Carcinogenic Potential
0 Soberg and Loeser (1984) detected an increased frequency of urinary
bladder epithelium hyperplasia, bladder papillomas and carcinomas, and
carcinoma of the uterus in rats fed Baygon (250 mg/kg/day) for 2
years.
0 Bomhard and Loeser (1981) did not detect an increased incidence of
tumors in mice fed Baygon at doses up to 90 mg/kg/day for 2 years.
0 The International Agency for Research on Cancer (IARC) has not evalu-
ated the carcinogenic potential of Baygon.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), Baygon may be classified in
Group C: possible human carcinogen. This category is for substances
with limited evidence of carcinogenicity in animals in the absence of
human data. However, this classification group may be considered
preliminary at the present (U.S. EPA, 1987b) since the U.S. EPA
Office of Pesticide Programs (OPP) has classified this chemical in
Group B2: probable human carcinogen (U.S. EPA, 1987a). A resolution
will be reached between OPP and the Cancer Assessment Group (CAG) in
the near future.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
• Residue tolerances have not been established for Baygon by the OPP.
0 The American Conference of Governmental Industrial Hygienists (ACGIH,
1984) has proposed a threshold limit value of 0.5 mg/m3.
0 The World Health Organization (WHO) calculated an ADI of 0.02 mg/kg/day
for Baygon (Vettorazzi and Van den Hurk, 1985).
VII. ANALYTICAL METHODS
0 Analysis of Baygon is by a high-performance liquid chromatographic
(HPLC) procedure used for the determination of N-methyl carbamoyloximes
and N-methylcarbamates in water samples (U.S. EPA, 1986b). 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 that is detected with a
fluorescence detector. The method detection limit has not been
determined for Baygon, but it is estimated that the detection limits
for analytes included in this method are in the range of 0.5 to 3 ug/L.
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Baygon ' 1987
-17-
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate granular activated carbon (GAC) adsorption
to be a possible Baygon removal technique.
0 Adsorption of Baygon on GAC proceeds in accordance with both Freundlich
and Langmuir isotherms (El-Dib et al., 1974; Whittaker et al., 1982).
0 One full-scale laboratory test was carried out on a commercially
available system (Dennis et al., 1983; Kobylinski et al., 1984).
Different levels of Baygon (20 mg/L, 60 mg/L and 100 mg/L) were added
to tap water. At a flow rate of 67.4 gpm, the column removed 99% of
the Baygon in 3.5, 8.5, and 21 hours, respectively, using only 45 Ib
of granular carbon.
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Baygon ' 1987
-18-
IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH.
Blevins, R.D., M. Lee and J.D Regan. 1977. Mutagenicity screening of five
methyl carbamate insecticides and their nitroso derivatives using mutants
of Salmonella typhimurium LT2. Mutat. Res. 56:1-6.
Bomhard, E., and E. Loeser.* 1981. Propoxur, the active ingredient of Baygon:
Chronic toxicity study on mice (two-year feeding experiment). Bayer
Report No. 9954;69686. Bayer A.G, Institut fur Toxicologie. Unpublished
study. MRID 00100546.
Chemagro Corporation.* (no date). Toxicity study on humans. Report No. 28374.
Unpublished study. MRID 00045091.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Crawford, C.R., and R.H. Anderson.* 1971. The skin and eye irritating
properties of (R) Baygon technical and Baygon 70% WP to rabbits. Report
No. 29706. Unpublished study. MRID 00045097.
Davies, J.E., J.J. Freal and R.W. Babione. 1967. Toxicity studies: Field
trial of OMS-33 insecticide in El Salvador. Report No. 23933. World
Health Organization. CDL:091768-F. Unpublished Study. MRID 00052281.
Dawson, J.A., D.F. Heath, J.A. Rose, E.M. Thain and J.B. Word. 1964. The
excretion by humans of the phenol derived from 2-isopropoxyphenyl
N-methylcarbamate. Bull. WHO. 30:127-134.
DeLorenzo, F., N. Staiano, L. Silengo and R. Cortese. 1978. Mutagenicity of
Diallate, Sulfallate and Triallate and relationship between structure
and mutagenic effects of carbamates used widely in agriculture. Cancer
Res. 38:13-15.
Dennis, W.H., A.B. Rosencrance, T.M. Trybus, C.W.R. Wade and E.A. Kobylinski.
1983. Treatment of pesticide-laden wastewaters from Army pest control
facilities by activated carbon filtration using the carbolator treatment
system. U.S. Amy Bioengineering Research and Development Laboratory,
Ft. Detrick, Frederick, MD.
Eben, A., and G. Kimmerle.* 1973. Propoxur: Effect of acute and subacute
oral doses on acetylcholinesterase activity in plasma, erythrocytes, and
brain of rats. Report No. 4262. Report No. 39114. Unpublished study.
MRID 00055148.
El-Dib, M.A., F.M. Ramadan and M. Ismail. 1974. Adsorption of sevin and
baygon on granular activated carbon. Water Res. 9:795-798.
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Baygon August, 1987
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Everett, L.J., and R.R. Gronberg.* 1971. The metabolic fate of Baygon
(o-isopropoxyphenylmethyl carbamate) in the rat. Chemagro Corp. Research
and Development Department Report No. 28797. Unpublished study. MRID
00057737.
Farbenfabriken Bayer.* 1961. Toxicity of Bayer 39007 (Dr. Bocker 5812315):
Report No. 6686. Farbenfabriken Bayer Aktiengesellschaft. Unpublished
study. MRID 00040433.
Farbenfabriken Bayer.* 1966. Two-month feeding test with Bayer 39007. Report
No. 17466. Znstitut fur Toxicologie. Unpublished study. MRID 00035412.
Foss, W., and J. Krechniak. 1980. The fate of propoxur in rat. Arch. Toxicol.
4:346-349.
Games, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol.
14:515-534.
Heimann, K. 1982. Propoxur (the active ingredient of Baygon and Unden):
study of sensitization effects on guinea pigs: Bayer Report No. 11218.
(Mobay Report 82567, prepared by Bayer AG, Institute fuer Toxikologie).
Unpublished study. MRID 00141139.
Kobylinski, E.A., W.H. Dennis and A.B. Rosencrance. 1984. Treatment of
pesticide-laden wastewater by recirculation through activated carbon.
American Chemical Society.
Krishna, J.G., and J.E. Casida.* 1965. Fate of the variously labeled methyl-
and dimethyl-carbamate-14c insecticide chemicals in rats. Report No.
16440. Unpublished study. MRID 00049234.
Lehman, A. J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U. 5.
Lorke, D.* 1971. BAY 39007: Examination for embryotoxic effects among rats.
Report No. 2388. Report No. 29035. MRID 00045094.
Loser, E.* 1968a. BAY 39007: Chronic toxicological studies on rats. Report
No. 726. Report No. 22991. Unpublished study. MRID 00035425.
Loser, E.* 1968b. BAY 39007: chronic toxicological studies on dogs. Report
No. 669. Report No. 22814. Unpublished study. MRID 00035423.
Meister, R., ed. 1984. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Montazemi, K. 1969. Toxicological studies of Baygon insecticide in
Shabankareh area, Iran. Trop. Geogr. Med. 21:186-190.
Moriya, M., T. Ohta, K. Wantanabe, T. Miyazawa, K. Kato and Y. Shirasu. 1983.
Further mutagenicity studies on pesticides in bacterial reversion assay
systems. Mutat. Res. 116:185-216.
-------
Baygon August, 1987
-20-
NIOSH. 1983. National Institute for Occupational Safety and Health. Registry
of toxic effects of chemical substances. Tatken, R.L., and R.J. Lewis,
eds. Cincinnati, OH: National Institute for Occupational Safety and
Health. DHHS (NIOSH) Publication No. 83-107.
Root, M., J. Cowan and J. Doull.* 1963. Subacute oral toxicity of Bayer 39007
to male and female female (sic) rats: Report No. 10685. Unpublished
study. MRID 00040447.
Schlueter, G., and D. Lorke.* 1981. Propoxur, the active ingredient of
Baygon: Study of embryotoxic and teratogenic effects on rabbits after
oral administration. Bayer Report No. 10183; MOBAY ACD Report No. 80034.
Bayer AG Institut fur Toxicologie. Unpublished study. MRID 00100547.
STORET. 1987.
Suberg, H., and H. Loeser.* 1984. Chronic toxicological study with rats
(feeding study over 106 weeks): Report 12870. Unpublished MOBAY study
No. 88501 prepared by Bayer Institute of Toxicology. Unpublished study.
MRID 00142725.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (1 85):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Method #5. Measure-
ment of N-methyl carbamoyloxines and N-methylcarbamates in ground water
by direct aqueous injection HPLC with post column derivatization.
January 1986 draft. Cincinnati, OH: U.S. EPA Environmental Monitoring
and Support Laboratory.
U.S. EPA. 1987a. U.S. Environmental Protection Agency. Qualitative and
quantitative risk assessment for Baygon. Office of Pesticide Programs.
A memo from Bernice Fisher to Dennis Edwards, 4/3/87.
U.S. EPA. 1987b. U.S. Environmental Protection Agency. Supplemental
discussion of Baygon classification. Cancer Assessment Group. A memo
from Arthur Chiu to William H. Farland, 4/6/87.
Vandekar, M., R. Plestina and K. Wilhelm. 1971. Toxicity of carbamates for
mammals. Bull. WHO. 44:241-249.
Vettorazzi, G. and G.W. Van den Hurk. 1985. Pesticides Reference Index,
Joint Meeting on Pesticide Residues (JMPR) 1961-1984.
Whittaker, K.F., J.C. Nye, R.F. Wukash, R.J. Squires, A.C. York and H.A.
Razimier. 1982. Collection and treatment of wastewater generated by
pesticide application. U.S. Environmental Protection Agency, Cincinnati,
OH. EPA-600/2-82-028.
•Confidential Business Information submitted to the Office of Pesticide
Programs
-------
August, 1987
DRAFT
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.
lacause each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 25057-89-0
Structural Formula
3-( 1 -Methylethyl)-1H=2,1,3-benzothiadiazin-4(3H)-one,2,2-dioxide
Synonyms
0 Basagran; Bendioxide; Bentazone (Worthing, 1983).
Uses
0 Selective postemergent herbicide used to control broadleaf weeds in
soybeans, rice, corn, peanuts, dry beans, dry peas, snap beans for
seed, green lima beans and mint (Meister, 1986).
Properties (Worthing, 1983)
Chemical Formula C*0Hi 2N2O3S
Molecular Weight 240.4
Physical State Colorless crystalline powder
Boiling Point —
Melting Point 137 to 139°C
Density —
Vapor Pressure —
Water Solubility 500 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Bentazon was not found in sampling performed at two water supply
stations in the STORET database (STORET, 1987). No information
on the occurrence of bentazon was found in the available literature.
Environmental Fate
Bentazon, at 1 ppm, was stable to hydrolysis for up to 122 days in
unbuffered water (initial pH 5, 7, and 9) at 22°C (Drescher, 1972c).
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Bentazon August, 1987
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The bentazon degradate, 2-amino-N-isopropyl benzamide (AIBA) at
1 ppm, was stable to hydrolysis in unbuffered, distilled water at pH 5,
7, and 9, during 28 days1 incubation in the dark at 22°C (Drescher,
1973b).
0 14c-Bentazon at 2 to 10 ppm, degraded with a half-life of less than
2 to 14 weeks in a sandy clay loam, loam, and three loamy sand soils
(Drescher and Otto, 1973a; Drescher and Otto, 1973b). The soils were
incubated at 14 to 72% of field capacity and 23°C. The bentazon
degradation rate was not affected by soil moisture content but was
decreased by lowering the temperatures to 8 to 10°C. At pH 6.4, the
degradation rate in a loamy sand soil was 2.5 times longer than at
pH 4.6 and 5.5. The bentazon degradate, AIBA, was identified at less
than 0.1 ppm. AIBA degraded in loamy sand soil with a half-life of
1 to 10 days (43% of field capacity). 14c-Bentazon at 1.7 ppm did
not degrade appreciably in a loamy sand soil during 8 weeks of incubation;
AIBA was detected at a maximum concentration of 0.008 ppm.
0 Bentazon did not adsorb to Drummer silty clay loam, adjusted to pH 5
and 7, and 11 other soils tested at pH 5 (Abernathy and Wax,
1973). In the same study, using soil TLC, (!4C)bentazon was very
mobile in 12 soils, ranging in texture from sand to silty clay loam,
with an Rf value of 1.0.
0 Bentazon was very mobile in a variety of soils, ranging in texture
from loamy sand to silty clay loam and muck, based on soil column
tests (Drescher and Otto, 1972; Abernathy and Wax, 1973; Drescher,
1973a; Drescher, 1972a). Approximately 73 to 103% of the bentazon
applied to the columns was recovered in the leachate.
0 AIBA (100 ug applied to loamy sand soil) was very mobile (Drescher,
1972b). After leaching a 12-inch soil column with 500 ml (10 inches)
of distilled water, 86.3% of the applied material was found in the
leachate.
Bentazon has the potential to contaminate surface waters as a result
of its mobility in runoff water and application to rice fields
(Devine, 1972; Wuerzer, 1972).
0 In the field, bentazon at 0.75 to 10 Ib ai/A dissipated with a half-
life of less than or equal to 1 month in the upper 6 inches of soil,
ranging in texture from sand to clay (Daniels et al., 1976; Devine
and Hanes, 1973; Stoner and Hanes, 1974b; Stoner and Hanes, 1974a;
BASF Wyandotte Corporation, 1974; Devine and Tietjens, 1973; Devine
et al., 1973). In the majority of soils (6 of 9), bentazon had a
half-life of less than 7 days. AIBA was detected at less than or
equal to 0.09 ppm. Collectively, the available data indicated that
geographic region (NC, TX, MS, AL, MN, or ID) and crops treated
(peanuts, soybeans, corn or fallow soil) had little or no effect on
the dissipation rate of bentazon in soil.
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Bentazon August, 1987
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III. PHARMACOKINETICS
Absorption
0 Male and female rats (200 to 250 g) given 0.8 og ^C-bentazon in 1 mL
of 50% ethanol by stomach tube excreted 91% of the administered dose
in the urine within 24 hours, This suggests that bentazon is almost
completely absorbed when ingested (Chasseaud et al., 1972).
Distribution
0 Whole-body autoradiography of rats indicated high levels of radio-
activity in the stomach, liver, heart and kidneys after 1 hour of
dosing with 14c-bentazon. Radioactivity was not observed in the brain
or spinal cord (Chasseaud et al., 1972).
Metabolism
0 Bentazon is poorly metabolized. Two unidentified metabolites
were detected (Chasseaud et al., 1972).
Excretion
Rats given radiolabeled bentazon excreted 91% of the administered dose
in the urine as parent compound. Feces contained 0.9% of the administers
dose (Chasseaud et al., 1972).
IV. HEALTH EFFECTS
Humans
No information on the health effects of bentazon in humans was found
in the available literature.
Animals
Short-term Exposure
0 The oral LD50 of bentazon in the rat was reported to be 2,063 mg/kg
(Meister, 1986).
0 LD50 values for bentazon in the rat, dog, cat and rabbit are reported
to be 1,100, 900, 500 and 750 mg/kg, respectively (RTECS, 1985).
0 Acute, subchronic and chronic oncogenicity studies on bentazon have
been invalidated because of data gaps and deficiencies. However, the
RfD Workgroup (U.S. EPA 1986a) calculated a Reference Dose (RfD) for
bentazon from a 13-week study in dogs. This study is described in
detail in Section V. Quantification of Toxicological Effects. Note
that the calculated RfD value has a low confidence level.
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Bentazon August, 1987
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DermaI/Ocular Effects
0 No valid information on the dermal/ocular effects of bentazon was
found in the available literature.
Long-term Exposure
0 As indicated under Short-term Exposure, long-term studies, including
reproductive effects and carcinogenicity studies, have been invalidated,
Reproductive Effects
No valid information on the reproductive effects of bentazon was
found in the available literature.
Developmental Effects
0 No valid information on the developmental effects of bentazon was
found in the available literature.
Mutagenicity
0 No valid information on the mutagenic effects of bentazon was found
in the available literature.
Carcinogenic!ty
0 No valid information on the carcinogenic effects of bentazon was
found in the available literature.
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:
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 child
(1 L/day) or an adult (2 L/day).
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Bentazon August, 1987
-6-
One-day Health Advisory
No data were found in the available literature that were suitable for
determination of One-day HA values. It is, therefore, recommended that the
Longer-term HA value for a 10-kg child (0.25 mg/L, calculated below) be used
at this time as a conservative estimate of the One-day HA.
Ten-day Health Advisory
No data were found in the available literature that were suitable for
determination of Ten-day HA values. It is, therefore, recommended that the
Longer-term HA value for a 10-kg child (0.25 mg/L, calculated below) be used
at this time as a conservative estimate of the Ten-day HA.
Longer-term Health Advisory
A 13-week study in beagle dogs has been selected for the calculation of
a Longer-term HA (Leuschner et al., 1970). Beagle dogs (three dogs/sex/dose}
were gven 0 (control), 100, 300, 1,000 and 3,000 ppm (0, 2.5, 7.5, 25 and
75 mg/kg/day; Lehman, 1959) of bentazon for 13 weeks. At a dose level of
3,000 ppm, overt signs of toxicity, including weight loss and ill health, were
observed; 1/3 males and 2/3 females died. At 3,000 ppm, all males showed signs
of prostatitis. Similar signs were observed in one male each at the 300- and
1,000-ppm levels. This study suggests a NOAEL of 100 ppm (2.5 mg/kg/day).
Utilizing this NOAEL, a Longer-term HA for a 10-kg child is calculated
as follows:
Longer-term HA = (2.5 mg/kg/day) (10 kg) = 0>25 mg/L (250 uq/L)
(100) (1 L/day)
where:
2.5 mgAg/day - NOAEL, based on absence of prostatic effects 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.
The Longer-term HA for the 70-kg adult is calculated as follows:
Longer-term HA = (2.5 mq/kq/day) (70 kg) = 0>875 /L (8?5 /L)
(100) (2 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of prostatic effects in dogs.
70 kg = assumed body weight of an adult.
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Bentazon August, 1987
-7-
100 - uncertainty factor, chosen in accordance with NAS/OOH
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.
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, 1986b), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
Lifetime studies were not available to calculate a Lifetime HA. However,
with the addition of another safety factor of 10 for studies of less-than-
lifetime duration, the Lifetime HA may be calculated from the 13-week feeding
study in dogs (Leuschner et al., 1970).
Using the NOAEL of 2.5 mg/kg/day, the Lifetime HA for bentazon is
calculated as follows:
Step 1: Determination of a the Reference Dose (RfD)
RfD = (2.5 mg/kg/day) = 0.0025 mg/kg/day
(1,00 0)
where:
2.5 mg/kg/day = NOAEL, based on the absence of prostatic effects in
dogs.
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.
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Bentazon August, 1987
-8-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - (0.0025 "qWday) (70 kg) „ 0.0875 mg/L (87.5
1 2 L/day )
where:
0.0025 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 = (0.0875 mg/L) (20%) = 0.0175 mg/L (17.5 ug/L)
where:
0.0875 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
No valid data are available to make a determination of the carcino-
genic potential of bentazon.
e Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986b), bentazon may be classified
in Group D: not Classified. This category is for agents with inadequate
animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 In response to a bentazon-tolerance review petition, EPA's Office of
Pesticide Programs has concluded that "a tolerance cannot be supported
at this time."
VII. ANALYTICAL METHODS
Analysis of bentazon is by a gas chromatographic (GC) method applicable
to the determination of bentazon in water samples (U.S. EPA, 1985).
In this method, an aliquot of sample is acidified and extracted with
ethyl acetate. The extract is dried, concentrated to 1 to 2 mL, and
methylated with diazomethane. The methylated extracts are analyzed
by gas chroma tography with flame photometric detection. The method
detection limit for bentazon has not been determined.
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Bentazon August, 1987
-9-
VIII. TREATMENT TECHNOLOGIES
There is no information available regarding treatment technologies
used to remove bentazon from contaminated water.
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Bentazon August, 1987
-10-
IX. REFERENCES
Abernathy, J. R. and L. M. Wax. 1973. Bentazon mobility and adsorption in
twelve Illinois soils. Weed Science. 21(3):224-226.
BASF Wyandotte Corporation. 1974. Analytical residue reports (soil and
water): bentazon. Unpublished study.
Chasseaud, L.F., D.R. Hawkins, B.D. Cameron, B.J. Fry and V.H. Saggers. 1972.
The metabolic fate of bentazon. Xenobiotica. 2(3):269-276.
Daniels, J., J. Gricher and T. Boswell. 1976. Determination of bentazon
(BAS 351-H) residues in sand soil samples from Yoakum, Texas: Report
No. IRDC-3; BWC Project No. I-2-G-73. Unpublished study prepared by
International Research and Development Corporation, submitted by BASF
Wyandotte Corporation, Wyandotte, Ml.
Devine, J. M. 1972. Determination of BAS 351-H (3-isopropyl-lH-2,l,3-benzo-
thiadiazin-4(3H)-one-2,2-dioxide) residues in soil and runoff water.
Report No. 133.
Devine, J. M. and R. E. Hanes. 1973. Determination of residues of BAS
351-H(3-isopropyl-lH-2,l,3-benzothiadiazin-4(3H)-one-2,2-dioxide) and
its benzamide metabolite, AIBA (2-amino-N-isopropyl benzamide), in
Sharkey silty clay soil from Greenville, Mississippi: Field Experiment
No. 72-99. Unpublished study prepared by State University of New
York—Oswego, Lake Ontario Environmental Laboratory and United States
Testing Company, inc., submitted by BASF Wyandotte Corporation,
Parsippany, NJ.
Devine, J. M. and F. Tietjens. 1973. Determination of BAS 351-H (3-isopropyl-
lH-2,l,3-benzothiadiazin-4(3H)-one-2,2-dioxide) residues in Commerce
silt loam soil from Greenville, Mississippi: Field Experiment No. 72-76.
Unpublished study prepared by State Univeristy of New York—Oswego, Lake
Ontario Environmental Laboratory and United States Testing Company,
Inc., submitted by BASF Wyandotte Corporation, Parsippany, NJ.
Devine, J. M., C. Carter, L. W. Hendrick et al. 1973. Determination of
residues of BAS 351-H (3-isopropyl-lH-2,l,3-benzothiadiazin-4(3H)-one-
2,2-dioxide) and its benzamide metabolite, AIBA (2-amino-N-isoopropyl
benzamide), in Webster Glencoe silty clay loam soil from Prior Lake,
Minnesota: Field Experiment No. III-B-6-72. Unpublished study prepared
by State University of New York—Oswego, Lake Ontario Environmental
Laboratory and others, submitted by BASF Wyandotte Corporation,
Parsippany, NJ.
Drescher, N. 1972a. A comparison between the leaching of bentazon and 2,4-D
through a soil in a model experiment: Laboratory Report No. 679.
Drescher, N. 1972b. Leaching of 2-amino-N-isopropyl benzamide (AIBA) from
the soil. Laboratory Report No. 682.
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Bentazon August, 1987
-11-
Orescher, N. 1972c. The effect of pH on the rate of hydrolysis of bentazon
(BAS 351-H) in water: Laboratory Report No. 1107. Translation;
unpublished study prepared by Badische Anilin- and Soda-Fabrik, AG,
submitted by BASF Wyandotte Corporation, Parsippany, NJ.
Drescher, N. 1973a. Leaching of bentazon in a muck soil. Laboratory Report
No. 1138.
Drescher, N. 1973b. The influence of pH on the hydrolysis of the bentazon
metabolite A IB A (2-amino-N-isopropyl benzamide) in water. Laboratory
Report No. 1136.
Drescher, N. and S. Otto. 1972. Penetration and leaching of bentazon in
soil: Laboratory Report No. 1099. Translation; unpublished study
prepared by BASF, AG, submitted by BASF Wyandotte Corporation,
Parsippany, NJ.
Drescher, N. and S. Otto. 1973a. Degradation of bentazon (BAS 351-H) in
soil. Report No. 1140.
Drescher, N., and S. Otto. 1973b. Degradation of bentazon (BAS 351-H) in
soil. Report No. 1149.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S.
Leuschner, F., A. Leuschner, W. Schwerdtfeger and H. Otto. 1970. 13-Week
toxicity study of 3-isopropyl-lH-2,1,3-benzothiadiazin-4(3H)-one-2,2-
dioxide to beagles when administered with the food. Unpublished report
prepared by Laboratory of Pharmacology and Toxicology, W. Germany.
September 28.
Meister, R., ed. 1986. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. National
Institute for Occupational Safety and Health. National Library of
Medicine Online File.
Stoner, J.H., and R.E. Hanes. 1974a. Determination of residues of bentazon
and AIBA (2-amino-N-isopropyl benzamide) in Commerce silt loam soil from
Greenville, MS: Field Experiment No. 73-41. Unpublished study prepared
in cooperation with Stoner Laboratories, Inc., and United States Testing
Company, Inc., submitted by BASF Wyandotte Corporation, Parsippany, NJ.
Stoner, J. H., and R. E. Hanes. 1974b. Determination of residues of bentazon
(BAS 351-H) and AIBA i'n Commerce silt loam soil from Greenville, MS:
Field Experiment No. 73-43. Unpublished study prepared in cooperation
with Stoner Laboratories, Inc. and United States Testing Company, Inc.,
submitted by BASF Wyandotte Corporation, Parsippany, NJ.
STORET. 1987.
-------
August. 1987
-12-
U.S. EPA. 1985. U.S. Environmental Protection Agency. U.S. EPA Method 107
- Revision A, Bentazon. Fed. Reg. 50:40701. October 4, 1985.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. RfD Work Group.
Worksheet dated April 7. «oup.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185):33992-34003.
September 24.
Worthing, C.R., ed. 1983. The pesticide manual. Great Britain: The Lavenham
Press, Ltd., p. 39.
Wuerzer, B. 1972. Bentazon model box runoff study: Runoff Report 73-6.
Unpublished study prepared in cooperation with United States Testing
Company, submitted by BASF Wyandotte Corporation, Parsippany, NJ.
-------
August, 1987
BROMACIL
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 m-xiels 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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Bromacil August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No; 314-40-9
Structural Formula
Bf^sXJ-CHCH,CHl
0 CH8
5-Bromo-6-methyl-3-(1 -methylpropyl) -2,4(1H, 3H)-pyrimidinedione
Synonyms
0 Borea; Borocil IV: Bromazil? Cynogan; Herbicide 976; Hyvar X-WS;
Hyvar X; Hyvar X Weed Killer; Hyvar X-L; Hyvar ex; Krovar II; Nalkil;
Uragan; Urox HX; Urox B; Weed-Broom (Meister, 1983).
Uses
0 Herbicide for general weed or brush control in noncrop areas;
particularly useful against perennial grasses (Neister, 1983).
Properties (Windholz et al., 1983)
Chemical Formula CgH^C^^Br
Molecular Weight 261.11
Physical State (at 25'C) White crystalline solid
Boiling Point
Melting Point 158-160°C
Density
Vapor Pressure (100°C) 8 x 1 0~4 mm Hg
Specific Gravity
Water Solubility (20°C) 815 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold —
Conversion Factor
Occurrence
0 Bromacil has been found in Florida ground water; a typical positive
was 300 ppb (Cohen et al., 1986).
Environmental Fate
0 Bromacil in aqueous solution was stable when exposed to simulated
sunlight for 6 days (Moilanen and Crosby, 1974). Only one minor
«4%) photolysis product (5-bromo-6-methyluracil) was identified. An
aqueous solution of bromacil at 1 ppm lost all herbicidal activity
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Bromacil August, 1987
-3-
after exposure to UV light for 10 minutes, but at 10 ppm and 15
minutes' irradiation herbicidal activity was still present (Kearney
et al., 1964). However, bromacil in an aqueous solution (pH 9.4)
containing the photosensitizer methylene blue was rapidly degraded
under direct sunlight with a halflife of <1 hour (Acher and Dunkelblum,
1979).
0 More than 26 soil fungi representative of several taxonomic groups,
including Fungi Imperfecti, Ascomycetes and Zygomycetes, were capable
of metabolizing bromacil as their sole carbon source (Wolf et al.,
1975; Torgeson, 1969; Torgeson and Mee, 1967; Boyce Thompson Institute
for Plant Research, 1971).
0 Data from soil metabolism studies indicate that bromacil at 8 ppm had
a half-life of about 6 months in aerobic loam soil incubated at 31 °C
(Zimdahl et al., 1970). However, 10% of applied bromacil at approximately
3 ppm was slowly degraded to CO2 in an aerobic sandy loam soil after
330 days at 22°C (Wolf, 1974; Wolf and Martin, 1974). In anaerobic
sandy loam soil, bromacil at approximately 3 ppm had a calculated
half-life of approximately 144 days. No C02 evolved from the sterilized
soil treated with bromacil within 145 days, indicating that degradation
was microbial.
0 Bromacil is mobile in soil. Phytotoxic residues of bromacil leached
19 cm in clay and silty clay loam soils eluted with the equivalent of
4.3 acre-inches of water (Signori et al., 1978). In mucky peat, loam
and loamy sand soils eluted with the equivalent of 13 to 15 cm of water,
bromacil leached to 10-, 25-, and to >30-cm depths, respectively (Day,
1976). Utilizing soil thin-layer chromatographic techniques 1*0
bromacil was evaluated to be mobile (Rf 0.7) in a silty clay loam
soil (Helling, 1971). Bromacil is not adsorbed by montmorillonite,
illite, or humic acid to any great extent [Freundlich K (adsorption
coefficient) i10 at 25°C]; however, at 0°C bromacil was adsorbed
(Freundlich K 126) to humic acid (Haque and Coshow, 1971; Volk,
1972). Adsorption appeared to increase with decreasing temperatures.
0 Data from field dissipation studies showed that bromacil phytotoxic
residues persisted in soils ranging in texture from sand to clay for
>2 years following a single application of bromacil at i2.6 Ib ai/A
(active ingredient/acre) (Bunker et al., 1971; Stecko, 1971).
III. PHARMACOKINETICS
Absorption
0 Workers who were exposed to bromacil during production, formulation
and packaging excreted unchanged bromacil and the 5-bromo-3-sec-butyl-
6-hydroxymethyluracil metabolite in the urine (DuPont, 1966b).
Unchanged bromacil and the metabolite were also detected in the urine
of rats fed bromacil in the diet (DuPont, 1966a). Although these
data indicate that bromacil is absorbed, sufficient information was
not available to quantify the extent of absorption.
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Bromacil August. 1987
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Distribution
0 No information was found in the available literature on the distribu-
tion of bromacil.
Metabolism
0 Workers at a bromacil production plant excreted unchanged bromacil
and the S-bromo-3-sec-butyl-6-hydroxymethyluracil metabolite, present
as the glucuronide and/or sulfonate conjugate, in urine (DuPont, 1966b).
0 Gardiner et al. (1969) fed rats (age and strain not specified) food
containing 1,250 ppm bromacil for 4 weeks. Assuming 1 ppm equals
0.05 mg/kg/day in the older rat (Lehman, 1959), this dietary level
corresponds to about 62.5 mg/kg/day. Analysis of the urine of these
rats revealed the presence of unchanged bromacil and the 5-bromo-3-
sec-butyl-6-hydroxymethyluracil metabolite (primarily as the
glucuronide and/or sulfonate conjugate). Five other minor metabolites
were also detected: 5-bromo-3-(2-hydroxy-1-methylpropyl)-6-methyluracil;
5-bromo-3-(2-hydroxy-1-methylpropyl)-6-hydroxymethyluracil; 3-sec-butyl-
6-hydroxymethyluracil; 5-bromo-3-(3-hydroxyl-1-methylpropyl)6-methyluracil;
and 3-sec-butyl-6-methyluracil. An unidentified bromine-containing
compound with a molecular weight of 339 was also detected.
Excretion
In humans exposed to bromacil during its formulation and packaging,
urinary excretion products included 0.1 ppm parent compound and
6.3 ppm 5-bromo-3-sec-butyl-6-hydroxymethyluracil, present mostly as
a conjugate (DuPont, 1966b).
Rats were fed bromacil (1,250 ppm in the diet) for 4 weeks; urine
was collected daily during weeks 3 and 4 of the study. Analysis of
the urine revealed the presence of 20 ppm unchanged bromacil and
146 ppm of the 5-bromo-3-sec-butyl-6-hydroxymethyluracil metabolite
(conjugated and unconjugated form) (DuPont, 1966a; Gardiner et al.,
1969).
IV. HEALTH EFFECTS
Humans
No information was located in the available literature on the health
effects of bromacil in humans.
Animals
Most of the animal data available are from unpublished studies identified
prior to the published report by Sherman and Kaplan (1975). These
authors stated that an 80% wettable bromacil powder was used in all
tests discussed in their report except for eye irritation studies in
which a 50% wettable bromacil powder was used. All dosages and
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Bromacil August, 1987
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feeding levels, unless otherwise stated, were based on the active
ingredient, bromacil.
Short-term Exposure
0 The oral LD50 value for male ChR-CD rats was calculated to be 5,200
mgAg (Sherman and Kaplan, 1975). Clinical signs of toxicity included
rapid respiration, prostration and initial weight loss.
0 In male mongrel dogs, a single oral dose of 5,000 rag/kg caused nausea,
vomiting, fatigue, incoordination and diarrhea (Sherman and Kaplan,
1975). It was not possible to estimate a lethal oral dose for bromacil
in dogs because vomiting occurred almost immediately, even at doses
of 100 mg/kg.
0 Sherman and Kaplan (1975) administered bromacil to groups of six male
ChR-CD rats by gastric intubation at dose levels of 650, 1,035 or
1,500 mg/kg/day, 5 days/week for 2 weeks (10 doses). Four of six
animals died at the high dose. Five of six survived exposure to
1,035 mg/kg/day, but showed gastrointestinal and nervous system
disturbances, and there was focal liver cell hypertrophy and hyper-
plasia. All animals survived the low dose with similar, but less
severe, pathological changes. The 650-mg/kg/day dose is identified as
the Lowest-Observed-Adverse-Effect-Levels (LOAEL) in this study.
0 Palmer (1964) reported that sheep that received bromacil at oral
doses of 250 mg/kg for five days developed weakness in the legs and
incoordination. Recovery from these symptoms usually took several
weeks. Administration of 100 mg/kg/day for 11 days induced an 11%
weight loss but no observable clinical symptoms.
Dermal/Ocular Effects
0 Bromacil (applied as a 50% aqueous solution of the 80% wettable
powder) was only mildly irritating to the intact and abraded skin of
young guinea pigs exposed for periods of up to 3 weeks. It was more
irritating to the skin of older animals. Bromacil did not produce
skin sensitization (OuPont, 1962).
0 Sherman and Kaplan (1975) reported that when bromacil was applied
dermally to rabbits the lethal dose was greater than 5,000 mg/kg,
the maximum feasible dose. No clinical signs of toxicity and no
gross pathological changes were observed.
0 Bromacil, as a 50% aqueous suspension, was mildly irritating to the
skin of young guinea pigs, but only slightly more irritating to the
skin of older animals. It was not a skin sensitizer (Sherman and
Kaplan, 1975).
0 Sherman and Kaplan (1975) reported that bromacil (0.1 mL of a 10%
suspension in mineral oil) resulted in only mild temporary conjuncti-
vitis in both the washed and unwashed eyes of rabbits. No corneal
injury was observed when a dose of 10 mg dry powder was applied
directly to the eye.
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Bromacil August, 1987
-6=
Long-term Exposure
0 Zapp (1965) discussed a study, also reported by Sherman and Kaplan
(1975), in which 10 male and 10 female ChR-CD rats were fed dietary
levels of 0, 50, 500 or 2,500 ppm bromacil for 90 days. This
corresponds to doses of about 0, 2.5, 25 or 125 mg/kg/day, assuming
1 ppm equals 0.05 mg/kg/day in an older rat (Lehman, 1959). Because
no signs of toxicity were observed at any dose, the high dose was
increased to 5,000 ppm (about 250 mg/kg/day) after 6 weeks; to
6,000 ppm (about 300 mg/kg/day) after 10 weeks; and to 7,500 ppm
(about 375 mg/kg/day) after 11 weeks. This dosing pattern resulted
in reduced food intake and mild histological changes in thyroid and
liver. No compound-related effects on weight gain, hematology,
urinalysis or histology were detected at the two lowest doses; 25
mg/kg/day was identified as the No-Observed-Adverse-Effect-Level
(NOAEL) in this study.
0 Sherman et al. (1966, also reported by Sherman and Kaplan, 1975) fed
groups of 36 male and 36 female ChR-CD rats food containing 0, 50,
250 or 1,250 ppm bromacil for 2 years. This corresponds to doses
of about 0, 2.5, 12.5 or 62.5 mg/kg/day, assuming 1 ppm equals 0.05
mg/kg/day in older rats (Lehman, 1959). Females at the highest
dose showed decreased weight gain (p <0.001). No other toxic effects
were observed in a variety of parameters measured, including mortality,
hematology, urinalysis, serum biochemistry, gross pathology, organ
weight or histopathology, except for a slight thyroid hyperplasia at
the high dose. This study identified a NOAEL of 12.5 mg/kg/day.
9 Beagle dogs (three/sex/dose level) were fed a nutritionally complete
diet containing 0, 50, 250 or 1,250 ppm bromacil for 2 years (Sherman
et al., 1966; also reported by Sherman and Kaplan, 1975). This
corresponds to doses of about 0, 1.25, 6.25 or 31.2 mg/kg/day, assuming
1 ppm equals 0.025 mg/kg/day in the dog (Lehman, 1959). No nutritional,
clinical, hematological, urinary, blood chemistry or histopathologic
evidence of toxicity was detected in any group. This study identified
a NOAEL of 31.2 mg/kg/day.
0 Kaplan et al. (1980) administered bromacil (approximately 95% pure)
to CD-I mice (80/sex/dose) for 78 weeks at dietary levels of 0, 250,
1,250 or 5,000 ppm. Based on information presented by the authors,
these dietary levels correspond to doses of 0, 39.6, 195 or 871
mg/kg/day for males and 0, 66.5, 329 or 1,310 mg/kg/day for females.
During the first year of the study, a compound-related decrease in
body weight gain was observed in male mice receiving 5,000 ppm and in
female mice receiving 1,250 ppm. The treatment and control groups
exhibited no significant (p <0.05) differences in food consumption.
Mortality in the 5,000-ppm females was significantly (p <0.05) greater
than in the controls. Liver changes noted in treated mice consisted
of increased mean and relative weights in the 1,250-ppm females and
the 5,000-ppm males; an increased incidence of diffuse hepatocellular
hypertrophy in the 1,250- and 5,000-ppm males and in the 5,000-ppm
females; an increased incidence of centrilobular vacuolation in 250-pprn
males; an increased incidence of scattered hepatocellular necrosis in
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Bromacil Au*ust' 1987
-7-
5,000-ppm males; and the presence of extravasated erythrocytes in
the hypertrophied hepatocytes of the 1,250- and 5,000-ppn males. The
authors felt that centrilobular vacuolation and hypertrophy were
probably related to enzyme induction. The toxicological significance
of extravasated erythrocytes in the hypertrophied hepatocytes was
unclear. Compound-related changes in the testes of mice consisted of
an increased incidence of spermatocyte necrosis, sperm calculi and
mild interstitial-cell hypertrophy/hyperplasia in the 1,250- and
5,000-ppm males and a dose-related increase in the incidence of testi-
cular tubule atrophy in all male treatment groups. Based on changes
in testes, a LOAEL of 250 ppm (39.6 mg/kg/day) is identified for male
mice. A NOAEL of 250 ppm (66.5 mg/kg/day) was identified for female
mice.
Reproductive Effects
0 Sherman et al. (1966; also reported by Sherman and Kaplan, 1975)
reported the effects of bromacil on reproduction in a three-generation
study in rats. Twelve male and twelve female weanling ChR-CD rats were
fed bromacil in the diet at 0 or 250 ppm. This corresponds to doses
of about 0 or 12.5 mg/kg/day, assuming 1 ppm in the diet equals
0.05 mg/kg/day for older rats (Lehman, 1959). Animals were bred
after 12 weeks, and the F1b and the F2b generations were maintained on
the same diets as their parents. No evidence of adverse effects on
reproduction or lactation performance was observed. Examination of
the F2b generation revealed no evidence of gross or histopathological
effects. This study identified a minimum NOAEL of 12.5 mg/kg/day.
Developmental Effects
0 Paynter (1966; also reported by Sherman and Kaplan, 1975) administered
bromacil to New Zealand White rabbits (8 or 9 per dosage) at dietary
levels of 0, 50 or 250 ppm on days 8 through 16 of gestation. Assuming
1 ppm equals 0.03 mg/kg/day in the rabbit (Lehman, 1959), these
dietary levels correspond to about 0, 1.5 or 7.5 mg/kg/day. No
significant differences between the conception rates of the control
and test groups were observed. Control and test group litters were
comparable in terms of litter size, mean pup length, mean litter
weight, number of stillbirths and number of resorption sites. No
gross malformations were observed in any animals. Skeletal clearing
revealed no abnormalities in bone structure in any animals. Based
on reproductive and teratogenic end points, a NOAEL of 250 ppm
(7.5 mg/kg/day) was identified.
• Pregnant rats (strain not specified) were exposed to aerosols of
bromacil (165 mg/m3) on days 7 to 14 of gestation. No prenatal
changes or teratogenic effects were observed (no further details were
provided) (Dilley et al., 1977).
Mutagenicity
0 In a sex-linked recessive lethal test (Valencia, 1981), Drosophila
melanogaster (Canton-S wild-type stock) were exposed to bromacil in
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Bromacil August, 1987
-8-
food at levels of 2, 3, 5 or 2,000 ppm. Bromacil was found to be
weakly mutagenic at the 2,000-ppm dose level.
0 Riccio et al. (1981) reported that bromacil (tested concentrations
not specified) was not mutagenic with or without metabolic activation
in assays conducted using Saccharomyces cerevisiae strains D3 and D7.
0 Siebert and Lemperle (1974) reported that bromacil was not mutagenic
when tested at a concentration of 1,000 ppm using £. cerevisiae
strain D4.
0 Simmon et al. (1977) reported that bromacil was not mutagenic in an
in vivo mouse dominant-lethal assay and the following in vitro assays:
unscheduled DNA synthesis in human fibroblasts (WI-38 cells); reverse
mutation in Salmonella typhimurium strains TA1535, 1537, 1538 and
100, and in Escherichia coli WP2; mitotic recombination in £. cerevisiae;
and preferential toxicity assays in E. coli (strains W3110 and p3478)
and Bacillus subtilis (strains H17 and M45).
0 In a modified Ames assay (Rashid, 1974), bromacil was not mutagenic
in £. typhimurium strains TA1535 and 1538 when tested at
concentrations up to 325 ug/plate.
0 In an assay designed to test for thymine replacement in mouse DNA
(McGahen and Hoffman, 1963), Swiss-Webster white mice received bromacil
by oral intubation at 100 rag/kg twice daily for 2 days, followed by
50 mg/kg twice daily for 8 days. Under the conditions of the assay,
bromacil was not recognized as a thymine analog by the mouse.
0 Bromacil did not show any signs of mutagenicity in a variety of
microbial test systems (Jorgenson et al., 1976; Woodruff et al., 1984).
0 In the Ames test, bromacil (5% concentration) induced revertants in
three of six Salmonella strains tested (Njage and Gopalan, 1980).
0 Bromacil did not induce sex-linked recessive lethals in £. melanogaster
(Gopalan and Njage, 1981).
Carcinogenicity
0 Sherman et al. (1966) fed Croups of 36 male and 36 female weanling
ChR-CD rats bromacil in the diet for 2 years. Dietary levels were
0, 50, 250 or 1,250 ppm (about 0, 2.5, 12.5 or 62.5 mg/kg/day, based
on Lehman, 1959). There was no effect on mortality, and the only
treatment-related lesion detected by histological examination was a
slight increase in the incidence of light-cell and follicular-cell
hyperplasia in the thyroid at the high dose. One high-dose female
was found to have follicular-cell adenoma. The authors stated that
these observations suggest a compound-related effect.
0 Kaplan et al. (1980) administered bromacil (approximately 95% pure)
to CD-1 mice (80/sex/dose) for 78 weeks at dietary levels of 0, 250,
1,250 or 5,000 ppm. Based on information presented by the authors,
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Bromacil August, 1987
-9-
these dietary levels correspond to compound intake levels of 0, 39.6,
195 or 871 mg/kg/day for males and 0, 66.5, 329 or 1,310 mg/kg/day
for females. In males, the combined incidences of hepatocellular
adenomas plus carcinomas/number of animals examined were 10/74,
11/71, 8/71 and 19/70 (p <0.05) at 0, 250, 1,250 and 5,000 ppm,
respectively. Hepatocellular carcinoma incidences were 5/74, 4/71,
4/71 and 9/70 (p >0.05) at 0, 250, 1,250 and 5,000 ppm, respectively.
These tumors were found predominantly in mice that survived to terminal
sacrifice. No effect on liver tumor incidence was observed in females.
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 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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No studies were located which are suitable for derivation of a One-day HA
for bromacil. The Ten-day HA, derived below, of 4.6 mg/L for a 10-kg child
is proposed as a conservative One-day HA.
Ten-day Health Advisory
The 2-week oral study in rats by Sherman and Kaplan (1975) has been
selected as the basis for the Ten-day HA for bromacil. Animals were
dosed by gavage for 10 days over a period of 2 weeks. The lowest dose tested
(650 mg/kg/day) produced mild pathological changes in the liver, and this
value was identified as a LOAEL.
Using a LOAEL of 650 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (650 mg/kg/day) (5/7) (10 kg) = 4.6 mg/L (4,60o ug/L)
(1,000) (1 L/day)
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Brcxnacil August, 1987
-10-
where:
650 mg/kg/day = LOAEL, based on mild liver pathology in rats
exposed by gavage to bromacil for 2 weeks.
5/7 = correction for dosing 5 days per week.
10 kg = assumed body weight of a child.
1,000 = uncertainty factor chosen in accordance with NAS/OCW
guidelines for use with a LOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The 90-day study by Zapp (1965) has been selected to serve as the basis
for the Longer-term HA for bromacil. Rats were fed diets containing up to
500 ppm without any adverse effects. This study identified a NOAEL of
500 ppm (about 25 mg/kg/day).
Using a NOAEL of 25 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (25 mg/kg/day) (10 kg) = 2.5 mg/L (2,500 ug/L)
(100) (1 L/day)
where:
25 mg/kg/day = NOAEL, based on the absence of any pathological evidence
of toxicity in rats exposed to bromacil via oral feeding
for 90 days.
10 kg = assumed body weight of child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Using a NOAEL of 25 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = (25 mg/kg/day) (70 kg) = 8.7 mg/L (8,700 ug/L)
(100) (2 L/day)
where:
25 mg/kg/day - NOAEL, based on absence of any toxic effects in rats
exposed to bromacil via oral feeding for 90 days.
70 kg = assumed body weight of an adult.
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Bromacil August, 1987
-11-
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.
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 chronic feeding study in rats by Sherman et al. (1966) has been
selected to serve as the basis for the Lifetime HA. This study identified a
dietary LOAEL of 1,250 ppm and a NOAEL of 250 ppm, based on weight gain and
mild thyroid hyperplasia. This NOAEL corresponds to about 12 mg/kg/day. The
same NOAEL is evident in a three-generation reproduction study in rats by
Sherman et al. (1966). The long-term feeding studies in dogs by Sherman
et al. (1966) and mice by Kaplan et al. (1980) were not selected, since the
demonstrated NOAEL was the lowest in the rat study.
Using a NOAEL of 12 mg/kg/day, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (12 mg/kg/day) = 0.1 2 mg/kg/day
(100)
where:
12 mg/kg/day = NOAEL, based on absence of hepatic effects in rats
exposed to bromacil via the diet for 2 years.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
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Bromacil August, 1987
-12-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = <0-12 mg/kg/day) (70 kg) „ ^2 mg/L (4/2oO ug/L)
(2 L/day)
where:
0.12 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 = (4.2 mg/L) (20%) = 0.08 mg/L (80 ug/L)
10
where:
4.2 mg/L = Lifetime HA at 100% contribution from drinking water.
20% = assumed relative source contribution from water.
10 » additional uncertainty factor per ODW policy for use with
a Group C carcinogen.
Evaluation of Carcinogenic Potential
0 Bromacil has not been determined to be carcinogenic, although an
.increased incidence of hepatocellular adenomas plus carcinomas was
observed in male CD-1 mice fed bromacil in the diet at a dose level
of 871 mg/kg/day for 78 weeks (Kaplan et al., 1980).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of bromacil.
0 Applying the criteria described in EPA'3 guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), bromacil is classified in Group C:
possible human carcinogen. This category is for substances with
limited evidence of carcinogenicity in animals in the absence of
human data.
0 The U.S. EPA has not published excess lifetime cancer risks for this
material.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The NAS (1977) has calculated an acceptable daily intake (ADI) of
0.0125 mg/kg/day, based on a chronic NOAEL of 12.5 mg/kg/day in rats and
an uncertainty factor of 1,000. A suggested-no-adverse-response level
(SNARL) of 0.086 mg/L was calculated based on an assumed water consumption
of 2 L/day by a 70-kg adult, with 20% contribution from water.
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Brooacil August, 1987
-13-
0 The U.S. EPA Office of Pesticide Programs (EPA/OPP) previously
calculated an AOI of 62.5 ug/kg/day, based on a NOAEL of 6.25 mg/kg/day
in a 2-year feeding study in dogs (Sherman et al., 1966) and an
uncertainty factor of 100. This was updated to 130 ug/kg/day based
on a 2-year rat feeding study using a NOAEL of 12.5 mg/kg/day and a
100-fold uncertainty factor.
0 A tolerance of 0.1 ppm bromacil in or on citrus fruits and pineapples
has been set by the EPA/OPP (CFR, 1985). A tolerance is a derived
value based on residue levels, toxicity data, food consumption levels,
hazard evaluation and scientific judgment, and it is the legal maximum
concentration of a pesticide in or on a raw agricultural commodity or
other human or animal food (Paynter et al., undated).
0 The American Conference of Governmental Industrial Hygienists (ACGIH,
1984) has recommended a threshold limit value (TLV) of 1 ppm, and a
short-term exposure limit (STEL) of 2 ppm.
VII. ANALYTICAL METHODS
0 Analysis of bromacil is by a gas chromatographic (GC) method applicable
to the determination of certain organonitrogen pesticides in water
samples (U.S. EPA, 1985). This method requires a solvent extraction
of approximately 1 L of sample with methylene chloride using a
separatory funnel. The methylene chloride extract is dried and
exchanged to acetone during concentration to a volume of 10 mL or
less. The compounds in the extract are separated by GC, and measure-
ment is made with a thermionic bead detector. The method detection
limit for bromacil is 2.38 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies used to remove bromacil from contaminated water.
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Bromacil August, 1987
-14-
IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH, p. 11.
Acher, A.J., and E. Dunkelblum. 1979. Identification of sensitized
photooxidation products of bromacil in water. J. Agric. Food
Chem. 27(6):1184-1187.
Boyce Thompson Institute for Plant Research. 1971. Interaction of herbicides
and soil microorganisms. U.S. EPA, Office of Research and Monitoring,
Washington, D.C.
Bunker, R.C., W.C. LeCroy, D. Katchur and T.C. Ellwanger, Jr. 1971.
Preliminary evaluation of herbicides on native grassland in Florida.
Department of the Army, Fort Detrick, Frederick, MD. Department of the
Army Technical Memorandum No. 232. Available from: NTIS, Springfield, VA.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.210, p. 287, July 1.
Cohen, S.Z., C. Eiden and M. N. Lorber. 1986. Monitoring ground water for
pesticides in the U.S.A. _In_ Evaluation of pesticides in ground water.
American Chemical Society Symposium Series. In press.
Day, E.W.* 1976. Laboratory soil leaching studies with tebuthiuron. (Unpub-
lished studies received Feb. 18, 1977, under 1471-109; submitted by
Elanco Products Co., Div. of Eli Lilly and Co., Indianapolis, IN. CDL:
095854-1). MRID 00020782.
Dilley, J.V., N. Chernoff, D. Kay, N. Winslow and G.W. Newell. 1977.
Inhalation teratology studies of five chemicals in rats. Toxicol. Appl.
Pharmacol. 41:196.
DuPont.* 1962. E.I. duPont de Nemours & Co. Toxicological information:
5-Bromo-3-sec-butyl-6-methyl-uracil. Unpublished report. MRID 00013246.
DuPont.* 1966a. E.I. duPont de Nemours & Co. Effect of enzymatic hydrolysis
on the concentration of bromacil and the principal bromacil metabolite
in rat urine. Unpublished report by E.I. duPont de Nemours & Co.
MRID 00013274.
DuPont.* 1966b. E.I. duPont deNemours Company. Analysis of urine from
bromacil production workers. Unpublished report by E.I. duPont de Nemours
S Co. MRID 00013273.
Gardiner, J.A., R.W. Reiser, and H. Sherman. 1969. Identification of the
metabolites of bromacil in rat urine. J. Agri. Food Chem. 17:967-973.
Gopalan, H.N.B., and G.D.E. Njage. 1981. Mutagenicity testing of pesticides.
Genetics. 97:544.
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Bromacil August, 1987
-15-
Haque, R., and W.R. Coshow. 1971. Adsorption of isocil and bromacil from
aqueous solution onto some mineral surfaces. Environ. Sci. Tech.
5:139-141.
Helling, C.S. 1971. Pesticide mobility in soils. I. Parameters of thin-layer
chromatography. Proc. Soil Sci. Soc. Am. 35:732-737.
Jorgenson, T.A., C.J. Rushbrook and G.W. Newell. 1976. In vivo mutagenesis
investigation of ten commercial pesticides. Toxicol. Appl. Pharmacol.
37:109.
Kaplan, A.M., H. Sherman, J.C. Summers, P.W. Schneider, Jr. and C.K. Wood.*
1980. Long-term feeding study in mice with 5-bromo-3-sec-butyl-6-methyl-
uracil (INN-976; Bromacil). Haskell Laboratory Report No. 893-80.
Final Report. Unpublished study. MRID 00072782.
Kearney, P.C., E.A. Woolson, J.R. Plimmer and A.R. Zsensee. 1964. Decontami-
nation of pesticides in soils. Residue Rev. 29:137-149.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
McGahen, J.W., and C.E. Hoffman. 1963. Action of 5-bromo-3-sec-butyl-6-
methyluracil as regards replacement of thymine on mouse DNA. Nature 199:
810-811.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Moilanen, K.W., and D.G. Crosby. 1974. The photodecomposition of bromacil.
Arch. Environ. Contam. Toxicol. 2(1):3-8.
NAS. 1977. National Academy of Sciences. Drinking water and health. Vol. 1.
Washington, DC: National Academy Press.
Njage, G.D.E., and H.N.B. Gopalan. 1980. Mutagenicity testing of some
selected food preservatives, herbicides and insecticides: II Ames Test.
Bangladesh J. Bot. 9(21:141-146.
Palmer, J.S. 1964. Toxicity of methyluracil and substituted urea and phenol
compounds to sheep. J. Am. Vet. Med. Assoc. 145:787-789.
Paynter, O.E.* 1966. Reproduction study — rabbits. Project No. 201-163.
(Unpublished study including letter dated May 27, 1966 from O.E. Paynter
to Wesley Clayton, Jr.). MRID 00013275.
Paynter, O.E., J.G. Cummings and M.H. Rogoff. Undated. United States
pesticide tolerance system. U.S. EPA Office of Pesticide Programs,
Washington, DC. Unpublished.
Rashid, K.A.* 1974. Mutagenesis induced in two mutant strains of Salmonella
typhimurium by pesticides and pesticide degradation products. Master's
Thesis, Pennsylvania State Univ., Dept. of Entomology. Unpublished
study. MRID 00079923.
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Bromacil August, 1987
-16-
Riccio, E., G. Shepherd, A. Pomeroy, K. Mortelmans and N.D. Haters.* 1981.
Comparative studies between the £. cerevisiae D3 and D7 assays of eleven
pesticides. Environ. Mutagen. 3:327 (Abstract P63).
Sherman, H., and A.M. Kaplan. 1975. Toxicity studies with 5-bromo-3-secbutyl-
6-methyluracil. Toxicol. Appl. Pharmacol. 34:189-196.
Sherman, H., J.R. Barnes and E.F. Stula.* 1966. Long-term feeding tests with
5-bromo-3-secondary butyl-6-methyluracil (INN-976; Hyvar(R)X; Bromacil):
Report No. 21-66. Unpublished study. MRID 00076371.
Siebert, D., and E. Lemperle. 1974. Genetic effects of herbicides: Induction
of mitotic gene conversion in Saccharomyces cerevisiae. Mutat. Res. 22:111-
120.
Signori, L.H., R. Deuber and R. Forster. 1978. Leaching of trifluralin,
atrazine, and bromacil in three different soils. Noxious Plants.
Z(l):39-43.
Simmon, V.F., A.D. Mitchell and T.A. Jorgenson.* 1977. Evaluation of selected
pesticides as chemical mutagens: in vitro and in vivo studies. Unpub-
lished study. MRID 05009139.
Stecko, V. 1971. Comparison of the persistence and the vertical movement of
the soil-applied herbicides simazine and bromacil. In Proceedings of
the 10th British weed control conference, Vol. 1. Droitwich, England:
British Weed Control Conference, pp. 303-306.
Torgeson, D.C. 1969. Microbial degradation of pesticides in soil. _In
Current topics in plant science. J.E. Gunckel, ed. New York: Academic
Press, pp. 58-59.
Torgeson, D.C., and H. Mee. 1967. Microbial degradation of bromacil.
jn Proceedings of the Northeastern Weed Control Conference, Vol. 21.
Farmingdale, NY: Northeastern Weed Control Conference, p. 584.
U.S. EPA. 1985. U.S. Environmental Protection Agency. U.S. EPA Method 633-
Organonitrogen Pesticides. Fed. Reg. 50:40701, October 4.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment.- Fed. Reg. 51(185):33992-34003. September 24.
Valencia, R.* 1981. Mutagenesis screening of pesticides "Drosophilia."
Prepared by Warf Institutes, Inc., for the Environmental Protection
Agency; Available from the National Technical Information Service.
EPA 600/1/-81/017. Unpublished study. MRID 00143567.
Volk, V.V. 1972. Physico-chemical relationships of soil-pesticide interactions.
In Progress Report, Oregon State University Environmental
Health Science Centre. Corvallis, OR. pp. 186-199.
Windholz, J., S. Budaveri, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck index, 10th ed. Rahway, NJ: Merck and Company, Inc.
-------
Bromacil August, 1987
-17-
Wolf, D.C. 1974. Degradation of bromacil, terbacil, 2,4-D and atrazine in
soil and pure culture and their effect on microbial activity. Diss.
Abstr. Int. B. 34(10):4783-4784.
Wolf, D.C., and J.P. Martin. 1974. Microbial degradation of 2-carbon-14
bromacil and terbacil. Proc. Soil Sci. Soc. Am. 38:921-925.
Wolf, D.C., D.I. Bakalivanov and J.P. Martin. 1975. Reactions of bromacil
in soil and fungus cultures. Soil Sci. Ann. XXVI(2):35-48.
Woodruff, R.C., J.P. Phillips and D. Irwin. 1984. Pesticide-induced complete
and partial chromosome loss in screens with repair-defective females of
Drosophilia melanogaster. Environ. Mutagen. 5:835-846.
Zapp, J.A., Jr.* 1965. Toxicological information: bromacil: 5-bromo-3-sec-
butyl-6-methyluracil. Unpublished study. MRID 00013243.
Zimdahl, R.L., V.H. Freed, M.L. Montgomery and W.R. Furtick. 1970. The
degradation of triazine and uracil herbicides in soil. Weed Res.
10:18-26.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
BUTYLATE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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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Butylate
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 2008-41-5
Structural Formula
H5C2-S-C-N;
^^
'CH2CH(CH3)2
Carbamothioic acid, bis(2-methylpropyl)-, S-ethyl ester
Synonyms
Uses
0 S-ethyl di-isobutylthiocarbamate; S-ethyl bis(2-methylpropyl)
carbamothioate; ethyl N,N-di-isobutyl thiocarbamate; S-ethyl-di-isobutyl
thiocarbamate; ethyl-N,N-di-isobutyl thiolcarbamate; R-1910; Sutan*.
0 Selective preplant herbicide (Meister, 1983).
CnH23NOS
217.41
Clear liquid, aromatic odor
138°C
0.9417
1.3 x 10-3 mm Hg
45 mg/L
Properties (BCPC, 1977)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density (25°C)
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (20°C)
Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Butylate has been found in 298 of 431 surface water samples
analyzed and in none of 18 ground water samples (STORET, 1987).
Samples were collected at 52 surface water locations and 18 ground
water locations, and butylate was found in 5 states. The 85th
percentile of all nonzero samples was 0.17 ug/L in surface water and
0 ug/L in ground water sources. The maximum concentration found was
6 ug/L in surface water and in 0 ug/L in ground water.
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Butylate August, 1987
-3-
Environmental Fate
0 Butylate degrades fairly rapidly in moist soils under aerobic condi-
tions; half-lives were 3 to 10 weeks (Thomas and Holt, 1979; Shell
Development Company, 1975; Stauffer Chemical Company, 1975a). Under
anaerobic conditions, butylate degrades with a half-life of 13 weeks
(Thomas et al., 1978). Butylate sulfoxide is the major degradate,
but s-ethyl-2,2-dimethyl-2-hydroxyethylisobutyl thiocarbamate,
diisobutylformamide, diisobutylamine, diisobutylthiocarbamate, and
isobutylamine were also identified as degradates (Thomas and Holt,
1979; Thomas et al., 1978; Shell Development Company, 1975; Stauffer
Chemical Company, 1975a).
0
e
Butylate is slightly mobile to highly mobile in soils ranging in
texture from silty clay loam to gravelly sand (Gray and Weierich,
1966; Lavy, 1974; Thomas and Holt, 1979; Weidner, 1974).
Butylate is fairly volatile; 45 to 50% of 14C- butylate applied to
moist (20% moisture) Sorrento clay loam was recovered as volatile
radioactivity over 3 weeks following treatment. Volatile radioactivity
was characterized as butylate (Thomas and Holt, 1979).
In the field, butylate dissipated more readily in a soil in
Florida than in a' silty clay loam in California, probably leaching
beyond the 6-inch sampling depth. The estimated half-lives in the
upper 6 inches of the sand were 28 and 18 days when a 4 Ib/gal Mcap
and a 6.7 Ib/gal EC formulation, respectively, were applied at 8 Ib
ai/A. For the silty clay loam, estimated half-lives were more than
64 days for both the Mcap and a 7 Ib/gal EC formulation applied at
8 Ib ai/A (active ingredient/acre) (Stauffer Chemical Company, 1975b;
Stauffer Chemical Company, 1975c).
Butylate has a low bioaccumulation potential in bluegill sunfish. A
bioconcentration factor of 33 was found in the edible portion of fish
dosed with 14c-butylate at 0.01 or 1 ppm for 28 days. The nonedible
portion of fish dosed at 0.01 and 1 ppm exhibited bioconcentration
factors of 174 and 122, respectively. After 10 days of depuration,
50 to 67% of the day-28 residues was lost (Sleight, 1973).
III. PHARMACOKINETICS
Absorption
Data relating specifically to the absorption of butylate were not
located in the available literature; however, some information was
obtained from a metabolism study by Hubbell and Casida (1977). Doses
of 12.3 or 156.0 mg/kg 14co-labeled butylate were administered by
gavage to male albino Sprague-Dawley rats weighing 190 to 210 g.
Within 48 hours, 27.3 and 31.5% of the administered radioactivity
were recovered in the urine, and 60.9 and 64.0% were expired as 14CO2
in the low- and high-dose groups, respectively. These results indicate
that butylate is appreciably absorbed from the gastrointestinal tract
of rats.
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Butylate August, 1987
-4-
Distribution
0 Hubbell and Casida (1977) measured the tissue radioactivity 48 hours
after the administration by gavage of 12.3 or 156.0 mg 14co-labeled
butylate/kg to male Sprague-Dawley rats. At the low dose, 2.4% of
the administered radioactivity was retained in the body, with levels
of radioactivity equivalent to 276 ppb in the blood, 524 ppb in the
kidney, 710 ppb in the liver and a range of 182 to 545 ppb in other
tissues (brain, fat, heart, lung, muscle, spleen and testes). At the
high dose, 2.2% of the radioactivity was retained in the body with
2,076 ppb in the blood, 5,320 ppb in the kidney, 7,720 ppb in the
liver and 1,720 to 5,560 ppb in other tissues.
Metabolism
0 Hubbell and Casida (1977) followed the metabolism of butylate in male
Sprague-Dawley rats based upon identification of the 48-hour urinary
metabolites of 14co~labeled preparations of butylate (12.3 or 156
mg/Jcg). Degradation of administered butylate metabolites was also
assessed. Approximately 40% of the administered 14co-butylate was
metabolized by ester cleavage and 14CO2 liberation without going
through the sulfoxide (the major metabolite) as an intermediate. The
metabolites from all compounds were essentially the same qualitatively
and quantitatively. The metabolites for 14CO-butylate included, as
percent of urinary radioactivity, 4.3% as the N,N-di-isobutyl mercapturic
acid, 17.1% as the N-isobutylmercapturic acid, 0.8% as the mercaptoacetic
acid derivative, 11.7% as the glycine conjugate of the mercaptoacetic
acid derivative and about 66% as at least 15 other metabolites.
e s-(1-14c)ethyl-Sutan*f orally administered at about 110 mg Sutan*A9f
was readily degraded and excreted by male and female Sprague-Dawley
rats (Thomas et al., 1980). Cleavage of the S-ethyl moiety and the
incorporation of the two-carbon fragment into intermediary metabolic
pathways accounted for >70% of the total administered radiocarbon.
Urinary excretion of 14c-hippuric acid, ethyl methyl sulfoxide and
ethyl methyl sulfone was evident.
Excretion
Hubbell and Casida (1977) administered 12.3 or 156 mg/kg of 1 Re-
labeled butylate by gavage to adult male Sprague-Dawley rats,
24 hours, 60.9 and 64.0% of the administered radioactivity were
expired as C(>2, 27.3 and 31.5% were excreted in the urine and 3.3 and
4.7% were excreted in the feces in the low- and high-dose groups,
respectively.
A study by Bova et al. (1978) indicates that biotransformation of
S-(1-14c) ethyl-Sutan® in male and female Sprague-Dawley rats given
oral doses of 83.5 to 133.5 mg Sutan*/kg involves rapid cleavage of
the S-ethyl moiety. Degradation of this fragment of the molecule
results in the release of 14CO2 as the major product of metabolism,
accounting for 69% of the total administered dose. This rapid pro-
duction of 14C02 may account for the relatively high levels (7.8%) of
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Butylate August, 1987
-5-
14c found in the tissues after 8 days. Urine and feces accounted for
13.9 and 3.2% of the 14c dose, respectively.
0 Data obtained from a 3-day balance and tissue residue study by Thomas
et al. (1979) show that (1-14c-isobutyl)Sutan® is readily eliminated
by male and female Sprague-Dawley rats after a single oral dose (about
100 mg Sutan®/kg). More than 99% of the administered radiocarbon was
recovered from the animals within 72 hours after dosing. Most of the
dose (94%) was recovered within 24 hours after treatment. Less than
0.5% of the radiocarbon remained in the tissues after 72 hours, and
the Sutan® equivalents in organ and tissue samples were all less than
2 ppm. Urine, feces and expired 14CO2 accounted for 93.7, 4.0 and
2.0% of the dose, respectively.
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of butylate in humans.
Animals
Short-term Exposure
0 The acute oral LD5Q value in male and female rats given butylate
technical (85.71% pure) was 3.34 and 3.0 g/kg, respectively (Raltech,
1979).
Dermal/Ocular Effects
0 Skin irritation was observed in rabbits topically exposed to 2 g
butylate technical (85.71% pure) for 24 hours (Raltech, 1979).
0 Topical application of R-1910 6E technical (97.5% pure) at doses of
20 and 40 mg active ingredient (a.i.)Ag, 5 days per week for a total
of 21 applications, was without observed effect except for local skin
irritation (Woodard Research Corp., 1967a).
0 Application of butylate technical (85.71% pure) to the eyes of rabbits
resulted in irritation and corneal opacity. No corueal opacity was in
eyes washed after treatment (Raltech, 1979).
Long-term Exposure
Q Dietary feeding of R-1910 technical (97.5% pure) to male and female
Charles River rats at dose levels of 32, 16 and 8 mg/kg/day for 13 weeks
was without observable adverse effect. The high dose (32 mg/kg/day)
was identified as the No-Observed-Adverse-Effect-Level (NOAEL) for this
study (Woodard Research Corp., 1967b).
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Butylate August, 1987
-6-
0 Dietary feeding of Sutan* Technical and Sutan* Analytical (purities
not specified) to male Sprague-Dawley rats at dose levels as high as
ISO mg/kg/day for 15 weeks was without observable adverse effect
(NOAEL) (Scholler, 1976).
0 Results of a toxicity study in which male and female beagle dogs were
fed R-1910 Technical (97.6% pure) at dietary levels of 450, 900 and
1,800 ppm (corresponding to doses of 11, 23 and 45 mg/kg/day, assuming
1 ppm equals 0.025 mg/kg/day from Lehman, 1959) for 16 weeks were
unremarkable (Woodard Research Corp., 1967c). Hence, 45 mg/kg/day is
identified as a NOAEL.
0 Sutan* Technical (98% pure) was fed in the diet to male and female
Spcague-Dawley rats at dose levels of 10, 30 and 90 mg/kg/day for 56
weeks. One group of rats was given 90 mg/kg/day for 15 weeks followed
by 180 mg/kg/day for 41 weeks. No systemic effects were found at
10 mg/kg/day (NOAEL). Testes/body weight ratios were significantly
(p <0.05) lower in terminally sacrificed males given 30 and 90 mg/kg/day.
Slight (8 to 15%) nonsignificant (p >0.05) mean body weight decreases
were found in 30 and 90 mg/kg males and 90 mg/kg females. Liver to
body weight increases and testicular lesions were found with the
highest doses. Blood clotting parameters were affected at all doses,
with the effects at 10 mg/kg/day being significant (p <0.05) decreases
in factor II times in males and activated partial thromboplastin
times in females (Hazelton Laboratories, Inc., 1978).
0 R-1910 Technical (purity not specified) was fed in the diet to male
and female Sprague-Dawley CD rats at dose levels of 50, 100, 200 and
400 mg/kg/day for 2 years. Although significantly (p <0.05) elevated
liver-to-body weight ratios occurred in terminally sacrificed males
given 50 mg/kg/day, this effect was not observed in animals from this
dose group sacrificed at 12 and 18 months. Hence, 50 mg/kg/day was
identified as a NOAEL. In males and females, body weights were
significantly (p <0.05) reduced, and liver to body weight ratios were
significantly (p <0.05) increased with doses above 50 mg/kg/day.
Neoplastic nodules and periportal hypertrophy in the liver were
significantly (p <0.05) increased in males given 400 mg/kg/day
(Biodynamics, 1982).
0 Male and female Charles River CD-1 mice were given Sutan* Technical
(98% pure) in the diet at dose levels of 20, 80 and 120 mg/kg/day for
2 years. No effects were found at 20 mg/kg/day (NOAEL). Kidney and
liver lesions were noted with higher doses (International Research
and Development Corporation [IRDC], 1979).
Reproductive Effects
0 No information was found in the available literature on the effects
of butylate on reproduction.
Developmental Effects
0 Sutan9 Technical (98.2% pure) was administered by gavage to pregnant
rats at doses of 40, 400 and 1,000 mg/kg/day on days 6 through 20 of
-------
Butylate August, 1987
-7-
gestation. The 40 mg/kg/day dose was without observable effect (NOAEL).
Higher doses decreased body weight gain in dams, increased liver-to-
body weights in dams, decreased fetal body weights, increased incidences
of misaligned sternebrae and delayed ossification, and increased
early resorptions. Sutan* was not teratogenic in this study (Stauffer
Chemical Co., 1983).
0 Administration of R-1910 Technical (97.6% pure) in the diet to pregnant
Charles River mice at dose levels of 4, 8 and 24 mg/kg/day either on
days 6 through 18 or from day 6 until natural delivery was without
observable effect (NOAEL) on dams and fetuses (Woodard Research
Corp., 1967d).
Mutagenicity
0 Butylate was not mutagenic in Salmonella typhimurium strains TA1535,
TA1537, TA1538 and TA100 with or without the S-9 activating fraction
(Eisenbeis et al., 1981).
0 In Drosophila melanogaster, butylate treatment increased the frequency
of sex-linked recessive lethals but had no effect on the frequency of
dominant lethals (Murnik, 1976).
Carcinogenicity
0 R-1910 Technical was not determined to be carcinogenic in the 2-year
rat study by Biodynamics (1982), but a significant (p <0.05) increase
in neoplastic nodules in liver in high-dose males was evident.
Neoplastic nodules were found in 2/69, 6/69, 1/69, 1/70 and 9/70
males given 0 ppm (control), 50 ppm, 100 ppm, 200 ppm and 400 ppm,
respectively. Hepatocellular carcinomas were found in 2/69, 3/69,
4/69, 3/70 and 2/70 males given 0 ppm (control), 50 ppm, 100 ppm,
200 ppm and 400 ppm, respectively.
0 Sutan® Technical was not carcinogenic in the 2-year mouse study by
IRDC (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)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
-------
Butylate August, 1987
-8-
BW t> 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/OCW guidelines.
_____ L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Realth Advisory
No information was found in the available literature that was suitable
for determination of the One-day HA value for butylate. It is, therefore,
recommended that the Ten-day HA value (2.4 mg/L, calculated below) be used
at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The teratology study in mice by Woodard Research Corporation (1967d)
has been selected to serve as the basis for determination of the Ten-day HA
value for butylate because it provides a short-term NOAEL (24 mg/kg/day for
13 days) for both maternal and fetal toxicity. The teratology study in rats
by Stauffer (1983), which identified a NOAEL of 40 mg/kg/day (for 15 days)
for maternal and fetal effects, could also be considered; however, because
doses higher than the 24 mg/kg/day NOAEL were not included in the Woodard
study (1967d), the effect levels in this study are uncertain. Furthermore,
the agent was given in the diet in the Woodard study (1967d) and by gavage in
the Stauffer (1983) study. Therefore, dose-response comparisons in terms of
both effect and no-effect levels between the Woodard (1967d) and Stauffer
(1983) studies cannot be made.
Using a NOAEL of 24 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-Day HA = (24 mg/kg/day) (10 kg) = 2.4 mg/L (2,400 ug/L)
(1 L/day) (100)
where:
24 mg/kg/day = NOAEL based on the absence of fetal and maternal
effects in mice exposed to Sutan® Technical orally
for 13 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 National
Academy of Sciences/Office of Drinking Water (NAS/ODW)
guidelines for use with a NOAEL from an animal study.
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Butylate August, 1987
-9-
Longer-term Health Advisory
The DWEL (2.45 mg/L) is recommended for use as a conservative estimate
of the Longer-term HA by the rationale given below.
The 56-week feeding study with Sutan* Technical in rats by Hazelton
Laboratories (1978) is a possible basis for a Longer-term HA. However,
effects observed in this study were not evident with higher doses in the
2-year feeding study with R-1910 Technical in rats by Biodynamics, Inc.
(1982).
The 20 mg/kg/day NOAEL in the 2-year mouse study by the International
Research and Development Corporation (IRDC) (1979) used to calculate the
Lifetime HA is concluded to be consistent with the data in the 56-week study
by Hazelton (1978) in that it is between the 30 mg/kg/day dose, where the
observed effect was decreased testes/body weight ratios, and the 10 mg/kg/day
NOAEL in the latter study. Effects on blood clotting parameters (decrease in
factor II times in males and activated partial thromboplastin times in females)
at the 10 mg/kg/day dose and higher in the Hazelton (1978) study are considered
to be of questionable toxicological significance because it is not certain
whether they actually represent adverse effects, and these effects were not
found in the 2-year rat study by Biodynamics (1982).
The 16-week and 13-week feeding studies with R-1910 Technical in dogs
and rats, respectively, by Woodard Research Corp. (1967b,c) can also be
proposed for calculation of the Longer-term HA. However, the highest
estimated dose of 45 mg/kg/day was the NOAEL in the dog study, and the
highest dose of 32 mg/kg/day was the NOAEL in the rat study. These NOAELs
are also higher than the 30 mg/kg/day dose where testicular effects were
evident in the Hazelton (1978) study in rats, though these effects are
overshadowed by the failure to repeat them in the 2-year rat study by
Biodynamics (1982), and use of doses between the 20 mg/kg/day NOAEL and the
80 mg/kg/day LOAEL in the IRDC (1979) mouse study could have provided a closer
comparison of dose-response across species. Consequently, the 20 mg/kg/day
NOAEL in the mouse study by IRDC (1979) is concluded to be an effective NOAEL
across species used in presently available butylate toxicity studies.
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
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Butylate August, 1987
-10-
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. EPAa, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study on Sutan® Technical in mice by IRDC (1979) has
been selected to serve as the basis for the Lifetime HA value for butylate.
Although the NOAEL of 20 mg/kg/day is lower than the NOAEL of 50 mg/kg/day in
the 2-year feeding study with Sutan® Technical in rats by Biodynamics (1982),
the mouse study is used, following the reasons given under the Longer-term HA.
The Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD . (20 mg/kg/day) = 0.07 mg/kg/day (70 ugAg/day)
(100) (3)
where:
20 mg/kg/day • NOAEL, based on the absence of toxic signs in mice
exposed to butylate in the diet for 2 years.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = additional uncertainty factor used by EPA/Office of
Pesticide Programs to account for the absence of major
studies (chronic feeding in dogs, reproduction in
rats, teratology in rabbits) which does not make it
possible to establish the most sensitive end point for
butylate.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0*07 mg/kg/day) (70 kg) = 2.45 mg/L (2,450 ug/L)
(2 L/day)
where:
0.07 mg/kg/day = RfD.
70 kg = assumed body weight of adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (2.45 mq/L)(20%) = 0.05 mg/L (50 ug/L)
(10)
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Butylate August, 1987
-11-
where:
2.45 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = uncertainty factor, chosen in accordance with Office of
Drinking Water (ODW) policy for use with Group C carcinogens.
Evaluation of Carcinogenic Potential
0 Available toxicity data do not determine butylate to be carcinogenic,
although a significant (p <0.05) increase in neoplastic nodules in
the liver of male rats fed the highest dose in the 2-year study by
Biodynamics (1982) was found.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), butylate may be placed in
Group C: a possible human carcinogen. This category is for substances
that show limited evidence of carcinogenic!ty in animals and inadequate
evidence in humans.
0 The U.S. EPA has not calculated excess lifetime cancer risks for this
material.
VI. OTHER CRITERIA,. GUIDANCE AND STANDARDS
0 Residue tolerances for butylate have been established by the U.S. EPA
(1985) and include 0.1 ppm in or on corn grain, fresh corn, corn
forage and fodder, sweet corn and popcorn. A tolerance is a derived
value based on residue levels, toxicity data, food consumption levels,
hazard evaluation and scientific judgment, and it is the legal maximum
concentration of a pesticide in or on a raw agricultural commodity or
other human or animal food (Paynter et al., undated).
0 The U.S. EPA Office of Pesticide Programs has calculated a provisional
ADI of 70 ug/kg/day, based on the 20-mg/kg/day NOAEL in the 2-year
mouse study by IRDC (1979) and a 300-fold uncertainty factor (used
because of data gaps, including a chronic feeding study in dogs, a
reproduction study in rats and a teratology study in rabbits, in the
total data package).
VII. ANALYTICAL METHODS
9 Analysis of butylate is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen- and phosphorus-containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 L of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector.
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Butylate August, 1987
-12-
The method detection limit has not been determined for butylate, but
it is estimated that the detection limits for analytes included in
this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies capable of effectively removing butylate from contaminated
water.
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Butylate August, 1987
-13-
IX. REFERENCES
BCPC. 1977. British Crop Protection Council. Pesticide Manual, 5th ed.
Nottingham, England: Boots Company, Ltd., p. 5S3.
Biodynamics, Inc.* 1982. A two-year oral toxicity/carcinogenicity study of
R-1910 in rats. Project no. 78-2169. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 00125678.
Bova, D.L., J.R. DeBaun, J.C. Petersen and J.J. Menn. 1978.* Metabolism of
[ethyl-14c] Sutan in the rat: Balance and tissue residue. Stauffer
Chemical Co., Richmond, CA. Unpublished final report. MRID 00043681.
Casida, J.E., R.A. Gray and H. Tilles. 1974. Thiocarbamate sulfoxides.
Potent, selective and biodegradable herbicides. Science. 184:573-574.
Eisenbeis, S.J., D.L. Lynch and A.E. Hampel. 1981. The Ames mutagen assay
tested against herbicides and herbicide combinations. Soil Sci.
131 (1):44-47.
Gray, R.A., and A.J. Weierich.* 1966. Behavior and persistence of S-ethyl-
diisobutylthiocarbamate (Sutan) in soils. Unpublished study. Stauffer
Chemical Company, Richmond, CA.
Hazelton Laboratories America, Inc.* 1978. Fifty-six-week feeding study in
rats. Sutan Technical. Project no. 132-135. Submitted to Stauffer
Chemical Co., Richmond, CA. Unpublished final report. MRID 00035843.
Hubbell, J.P., and J.E. Casida. 1977. Metabolic fate of the N,N'-dialkyl-
carbamoyl moiety of thiocarbamate herbicides in rats and corn. J. Agric.
Food Chem. 25(25:404-413.
IRDC.* 1979. International Research and Development Corporation. Sutan
Technical. Lifetime oral study in mice. Submitted to Stauffer Chemical
Co., Richmond, CA. Unpublished final report. MRID 00035844.
Lavy, T.L. 1974. Mobility and deactivation of herbicides in soil-water
systems: Project A-024-NEB, University of Nebraska, Water Resources
Research Institute. Submitted by Shell Chemical Company, Washington
DC. Available from National Technical Information Service (NTIS),
Springfield, VA*-PB-238-632.
Lehman, A. J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Murnik, M.R. 1976. Mutagenicity of widely used herbicides. Genetics. 83:554.
Paynter, O.E., J.G. Cummings and M.H. Rogoff. Undated. United States
Pesticide Tolerance System. U.S. EPA, Office of Pesticide Programs.
Unpublished draft report.
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Butylate August, 1987
-14-
Raltech.* 1979. Project nos. 74489 and 733422. Submitted to Stauffer Chemical
Co., Richmond, CA. Unpublished final report.
Scholler, J.* 1976. Fifteen-week oral (diet) toxicity study with Sutan
Technical and Analytical in male rats: Experiment 7. Unpublished final
report. MRID 00021844.
Shell Development Company.* 1975. Dissipation of Bladex herbicide and Sutan
in soil following application of Bladex, Sutan, or a tank mix of Bladen
and Sutan: TIR-24-134-74. Unpublished study.
Sleight, B.H., III.* 1973. Exposure of fish to He-labeled Sutans Accumulation,
distribution, and elimination of He residues. Unpublished study prepared
by Bionomics, Inc., submitted by Stauffer Chemical Company, Richmond, CA.
Stauffer Chemical Company.* 1975a. Dissipation of Bladex herbicide and Sutan
in soil following application of Bladex, Sutan, or a tank mix of Bladex
and Sutan: TIR-24-134-74. Unpublished study submitted by Stauffer
Chemical Company, Richmond, CA.
Stauffer Chemical Company.* 1975b. Residues from Sutan on soil: FSDS Nos.
A-9229, A-9229-1, A-9229-2, A-10366. Unpublished study by Stauffer
Chemical Company, Richmond, CA.
Stauffer Chemical Company.* 1975c. Soil residue data of Sutan combinations
and R-25788: FSDS Nos. A-9229, A-9229-1, A-9229-2, A-10366. Unpublished
study by Stauffer Chemical Company, Richmond, CA.
Stauffer Chemical Company.* 1983. A teratology study in CD rats with Sutan
Technical. Project no. T-11713. Unpublished final report by Stauffer
Chemical Company, Richmond, CA. MRID 000131032.
STORET. 1987.
Thomas, D.B., J.B. Miaullis, A.R. Vispetto and J. Osuna.* 1979. Metabolism
of [isobutyl-14C] Sutan in the rat: Balance and tissue residue study.
Stauffer Chemical Co., Richmond, CA. Unpublished final report. MRID
00043680.
Thomas, D.L.B., J.C. Petersen and J.R. DeB=»un.* 1980. Metabolism of
[1_14c-ethyl] Sutan in the rat: Urinary metabolite identification.
Stauffer Chemical Co., Richmond, CA. Unpublished final report. MRID
00043682.
Thomas, V.M., and C.L. Holt.* 1979. Behavior of Sutan in the environment:
MRC-B-76; MRC-78-02. Unpublished study submitted by Stauffer Chemical
Company, Richmond, CA.
Thomas, V.M., C.L. Holt and P.A. Bussi.* 1978. Anaerobic soil metabolism
of Sutan selective herbicide: MRC-B-98; MRC-79-13. Unpublished study
submitted by Stauffer Chemical Comapny, Richmond, CA.
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Butylate August, 1987
-15-
U.S. EPA. 1985. U.S. Environmental Protection Agency. Residue tolerances
for S-ethyl-diisobutyl thiocarhamate. CFR 180.232. July 1. p. 294.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Weidner, C.W.* 1974. Degradation in groundwater and mobility of herbicides.
Master's thesis, University of Nebraska, Department of Agronomy.
Unpublished study submitted by Shell Chemical Company, Washington, DC.
Woodard Research Corporation.* 1967a. R-1910 6-E. Subacute dermal toxicity.
21-Day experiment with rabbits. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 00026312.
Woodard Research Corporation.* 1967b. R-1910. Safety evaluation by dietary
feeding to rats for 13 weeks. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 00026313.
Woodard Research Corporation.* 1967c. R-1910. Safety evaluation by dietary
feeding to dogs for 16 weeks. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 00026314.
Woodard Research Corporation.* 1967d. R-1910. Safety evaluation by
teratological study in the mouse. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 000129544.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
CARBARYL
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
Z. 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 accurate!.' 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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Carbaryl
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 63-25-2
Chemical Structure
0 H
ii I
0-C-N-CH,
1-Naphthalene! methylcarbamate
Synonyms
0 Arilate; Bercena NMC50; Caprolin; Sevin; Vioxan (Meister, 1983).
Uses
0 Contact insecticide used for the control of pests on more than 100
different crops, forests, lawns, ornamentals, shade trees and rangeland
(Meister, 1983).
Properties (Windholz et al., 1983; CHEMLAB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Water Solubility (30°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C12Hn02N
201.22
White crystals
142°C
<4 x 10-5 mm Hg
120 mg/L
0.14
Occurrence
Carbaryl has been found in 61 of 522 surface water samples analyzed
and in 28 of 1,125 ground water samples (STORET, 1987). Samples were
collected at 138 surface water locations and 1,100 ground water
locations, and Carbaryl was found in 8 states. The 85th percentile
of all nonznro samples was 260 ug/L in surface water and 10 ug/L in
ground water sources. The maximum concentration found was 180,000
ug/L in surface water and 10 ug/L in ground water.
Environmental Fate
14c-Carbaryl (purity unspecified) at 10 ppm was relatively stable
to hydrolysis in buffered solutions at pH 3 and 6. It hydrolized at
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Carbaryl August, 1987
-3-
pH 9 with a half-life of 3 to 5 hours when incubated at 2S*C (Khasawinah
and Rolsing, 1977a). At 35*C, 14C-carbaryl was stable at pH 3, and
hydrolyzed with a half-life of >28 days and 30 to 60 minutes at pH 6 and
9, respectively. 1-Naphthol was the major degradate formed.
• Hocarbaryl (purity unspecified) at 5 ppm photodegraded slowly in
0.1 M phosphate buffer solutions, with 4.39 to 4.49 ppm remaining as
parent compound after 18 days of irradiation (Khasawinah and Holsing,
1977b). In a 2% acetone solution, 14C-carbaryl accounted for 3.63 to
3.65 ppm after 18 days. 1-Naphthol and several unidentified compounds
were found at <0.07 ppm.
0 Under aerobic conditions, 14C-carbaryl (>99% pure) at 1 ppm degraded
with a half-life of 7 to 14 days in a sandy loam soil maintained at 15
or 23 to 25°C, and 14 to 28 days in a clay loam soil maintained at
23 to 25°C (Khasawinah and Holsing, 1978). Degradation was slightly
slower in sterile soils (half-lives of 14 to 56 days). The majority
of the applied radioactivity was bound to the soil or had been evolved
as 14CO2 by the end of the test period (112 days). No degradates were
found.
0 Under aerobic conditions, 14C-carbaryl (>99% pure) at 1 ppm degraded
with a half-life of 84 to 112 days in a flooded sandy loam soil (Khasa-
winah and Holsing, 1978). At 168 days after treatment, 14C-carbaryl
accounted for 42% of the applied radioactivity in the soil and water
layer. 4-Hydroxy carbaryl was found at <0.3% of the applied radio-
activity in soil samples taken after 112 days. Approximately 20% of
the total radioactivity was soil-bound at 112 days.
III. PHARMACOKINETICS
Absorption
0 Comer et al. (1975) reported the results of tests conducted in factory
workers exposed to carbaryl during the formulation of 4 and 5% carbaryl
dust. Carbaryl exposure via the skin was measured by attachment of a
special gauze pad to various parts of the body, and inhaled carbaryl
was measured by the use of special filter pads in face masks. Calcu-
lated exposures were 73.90 and 1.10 mg/hour for the dermal and respiratory
routes, respectively. The total exposure was 75 mg/hour, or 600 mg/day.
Absorption levels were determined by estimation of the carbaryl
metabolite 1-naphthol in urine. It was determined that during an
8-hour workday the total absorption of carbaryl would be 5.6 mg.
This is about 0.9% of the total exposure, and the authors interpreted
this to mean that dermal absorption was not complete.
0 Feldman and Maibach (1974) applied 4 ug/cm2 of 14C-iabeled carbaryl
(position of label not specified) dissolved in acetone to one or both
forearms of apparently healthy male volunteers. The area of application
was left unwashed and unprotected for 24 hours. Based on the excretion
rate, it was determined that 73.9% of the applied carbaryl was absorbed
through the skin.
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Carbaryl August, 1987
-4-
• Houston et al. (1974) reported that 1^C-carbamyl-labeled carbaryl
administered by gavage to Bale rats at doses of 0.5 mg/kg (given as
0.5 mL of 0.5% propylene glycol in water) rapidly appeared in the
systemic circulation. Within a few minutes, the plasma level was
50 ng/mL. A maximum level of 150 ng/mL was reached in less than
10 minutes and steadily declined to 20 ng/mL at 120 minutes. Only
4.6% of the dose was excreted in the feces, indicating that at least
95.3% had been absorbed. .
• Falzon et al. (1983) administered single doses of 20 mg/kg of 14C-
carbaryl (in olive oil) to six female rats by gavage. After 24
hours, 5.8% of the label was recovered in the feces, indicating that
about 94.2% had been absorbed.
Distribution
• The distribution of 14c-carbonyl-labeled carbaryl in male and female
rats after administration of 1.5 mg/kg by stomach tube was examined
in eight body tissues (Krishna and Casida, 1965). The amounts
detected (umol/kg) in males and females, respectively, were: cecum,
0.17 and 0.60} esophagus, 0.05 and 0.05; large intestine, 0.02 and
0.03; small intestine, 0.06 and 0.08; kidney, 0.06 and 0.07; liver,
0.11 and 0.112; spleen, 0.05 and 0.08; and stomach, 0.07 and 0.14.
0 Falzon et al. (1983) administered single oral doses of 20 mg/kg of
14C-carbaryl to female Wistar rats by gavage. The amounts detected
24 hours after administration were 0.11% in the brain, 3.87% in the
digestive tract and 13.31% in the carcass.
Metabolism
• Human tissues obtained by either biopsy or autopsy were incubated
using an in vitro organ-maintenance technique with 14c-(N-methyl)-
labeled carbaryl (Chin et al., 1974). The following tissues were
examined: for males — lung, liver and kidney; for females — liver,
placenta, vaginal mucosa, uterus and uterine tumor (leiomyoma).
Hepatic tissues metabolized carbaryl by hydrolysis and/or demethylation,
hydroxylation and oxidation followed by conjugation. The primary
hydrolytic product was 1-naphthol (42% by 24 hours at pH 7.8). The
kidney produced naphthyl glucuronide; the uterus, lung and placenta
produced naphthyl sulfate from carbaryl. The vaginal mucosa produced
glucuronide and sulfate conjugates, but only a slight amount of
conjugating activity (napthol sulfate) was found in the uterine
leiomyoma.
0 Houston et al. (1974) administered 1 ^c-carbamyl-labeled carbaryl
(0.5 mg/kg) to male rats by gavage. Within 48 hours, 54.5% of the
label had been excreted in the urine as metabolites (not identified).
In addition, 32.9% was excreted as C02. This indicated that carbaryl
was extensively metabolized in rats.
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Carbaryl August, 1987
-5-
Excretion
0 Comer et al. (1975) studied the excretion of 1-naphthol in the urine
of workers who were exposed to carbaryl in a pesticide formulation
plant. The workers were exposed to carbaryl both dermally (73.9
mg/hour) and by inhalation (1.1 mg/hour). Analyses of urine samples
indicated that the excretion rate of 1-napthol varied from 0.004 to
3.4 mg/hour, with a mean value of 0.5 mg/hour. This corresponds to
an excretion rate of 0.7 mg carbaryI/hour. Following exposure to
carbaryl at the start of the workday, the urinary level of 1-naphthol
increased, reached its maximum level during the late afternoon and
evening hours, and then dropped to a lower level before the start of
the next day's workday.
0 Urinary excretion of topically applied radiolabeled carbaryl in
healthy male volunteers was measured by Feldman and Maibach (1974).
A total of 26.1% of the dose was recovered in the urine over a
5-day period.
0 Krishna and Casida (1965) administered single doses of 1.5 mg/kg of
14c-carbonyl-labeled carbaryl orally to rats. Excretion of the label
for male and female animals, respectively, was as follows: expired
carbon dioxide, 26% and 26%; urine, 64.0% and 72.0%; and feces, 4.0%
and 4.0%.
0 Houston et al. (1974) administered 14c-carbamyl-labeled carbaryl
(0.5 mg/kg) by gavage to male rats. The label was almost completely
excreted within 48 hours, with the following distribution: expired
carbon dioxide, 32.9%; urine, 54.5%; and feces, 4.6%. Less than 1%
of the label in urine was unchanged carbaryl. About 6.0% of the
label remained in the body. Biliary excretion was examined by bile-
duct cannulation. Within 6 hours, 30 to 33% of the administered dose
was present in the bile; after 6 hours, the amount in the bile leveled
off.
IV. HEALTH EFFECTS
Humans
0 Vanderkar (1965) investigated the effects of large-scale carbaryl
spraying in a village in Nigeria. Mo quantitative estimates of
exposure were obtained, but plasma cholinesterase (ChE) activity was
decreased by about 15% in eight applicators (spraymen) and by an
average of 8% in 63 villagers.
• Wills et al. (1968) studied the subchronic toxicity of carbaryl in
human volunteers. Groups of five or six men were given daily oral
doses of 0, 0.06 or 0.13 mg/kg/day for 6 weeks. At the lower dose,
no significant effects were detected on kidney function, electroen-
cephalogram, hematology, blood chemistry, urinalysis or plasma and
red blood cell ChE activity. At the higher dose, the only detectable
effect was a slight increase in the urinary ratio of amino acid
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Carbaryl August, 1987
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nitrogen to creatinine. This was interpreted to suggest a slight
decrease in resorption of amino acids in the kidney, This effect
was fully reversible. Based on these observations, a No-Observed-
Adverse-Effect-Level (NOAEL) of 0.06 mg/kg/day was identified.
Animals
Short-tern Exposure
• Carpenter et al. (1961) investigated the acute oral toxicity of
carbaryl in several species. Cats were found to be most sensitive
(2/2 deaths at 250 mg/kg). Guinea pigs, rats and rabbits were less
sensitive, with calculated U>so values of 280, 510 and 710 mg/kg,
respectively. No deaths were reported in dogs administered doses up
to 795 mg/kg/day.
0 The acute oral toxicity of carbaryl in male Sprague-Dawley rats was
studied by Rittenhouse et al. (1972). Carbaryl (99.9% active)
dissolved in acetone and propylene glycol (10% v/v) was administered
in a single dose at four dose levels to six animals per level.
Animals were observed for 14 days following treatment. Dose levels
were 439, 658, 986 or 1,481 mg/kg. Mortalities observed at these
levels were 0/6, 0/6, 4/6 and 5/6 rats, respectively. Most deaths
occurred in the first 24 hours. The LD50 was calculated to be
988 mg/kgo Animals at all dose levels exhibited symptoms of ChE
inhibition, but ChE activity was not measured. No other parameters
were reported.
0 Carpenter et al. (1961) fed single oral doses of carbaryl in capsules
to female mongrel dogs as follows: 250 mg/kg (one animal), 375 mg/kg
(four animals) or 500 mg/kg (one animal). Signs of overstimulation
of the parasympathetic nervous system were observed at the two higher
doses, but not at 250 mg/kg. These signs included: increased
respiration, lacrimation, salivation, urination, defecation, muscular
twitching, constriction of pupils, poor coordination and vomiting.
Plasma ChE was not affected at 375 mg/kg, but a transient decrease
(24 to 33%) was observed in erythrocyte ChE at this dose. After 1
day, the appearance of the animals was normal and no adverse CNS
effects were noted. Based on the absence of visible external effects
or inhibition of ChE, this study identified a NOAEL of 250 mg/kg.
0 Carpenter et al. (1961) also administered single oral doses of carbaryl
(560 mg/kg, by gavage in corn oil) to three groups of rats (seven to
nine per group). Groups were sacrificed after 0.5, 4 or 24 hours,
and ChE activity was measured in plasma, erythrocytes and brain.
Plasma ChE was slightly lower (7 to 14%) than control, but this was
not statistically significant. In erythrocytes, ChE was inhibited
42% after 0.5 hours, but this returned to near normal (86% of control)
within 24 hours. Brain ChE activity was inhibited 30% after 0.5 hours,
and this returned toward normal (91% of control) by 24 hours.
0 Weil et al. (1968) fed carbaryl in the diet for 1 week to Harlan-
Histar albino rats (42-days old) at concentrations yielding ingested
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Carbaryl August, 1987
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dosea of 0, 10, 50, 250 or 500 mgAg/day. Body weight gain was
decreased in animals exposed to 50 mgAg/day or higher. At 10 ing/kg/day,
ChE activity was not significantly affected in plasma, red blood
cells or brain. At 50 mgAg/day, plasma ChE was decreased 15 to 17%
and red blood cell ChE was decreased 26 to 47% (males and females,
respectively). At higher doses, larger decreases in plasma and red
blood cell ChE were seen, and brain ChE was also decreased (23 to 25%
at 250 mg/kg/day and 33 to 58% at 500 mg/kg/day). After 1 day on
control diet, these effects on ChE were entirely reversed. Based on
these data, a NOAEL of 10 mg/kg/day and a Lowest-Observed-Adverse-
Effect-Level (LOAEL) of 50 mg/kg/day were identified in rats.
Dermal Exposure
0 Carpenter et al. (1961) applied 0.01 mL of 10% carbaryl in acetone
(a dose of 1 mg) to the clipped skin of the belly of five rabbits.
No irritation was detected.
0 Gaines (1960) applied a series of doses of carbaryl dissolved in
xylene to the skin of Sherman rats. The dermal LD5Q value was greater
than 4,000 mg/kg for both males and females.
0 Carpenter et al. (1961) detected a weak skin sensitization reaction
in 4 of 16 male albino guinea pigs given eight intracutaneous injec-
tions of 0.1 mL of 0.1% carbaryl (0.1 mg/dose). The challenge dose
(not specified) was given 3 weeks later, and examinations for sensiti-
zation reaction were performed 24 and 48 hours thereafter.
0 Carpenter et al. (1961) applied carbaryl to the eyes of rabbits and
evaluated corneal injury. Technical carbaryl (98% pure) applied as
a 10% suspension in propylene glycol caused mild injury in 1/5 eyes.
A 25% aqueous suspension caused no injury, and 50 mg of powder caused
only traces of corneal necrosis.
Long-term Exposure
0 Wistar rats (five/sex, 45-days old) were fed carbaryl (as Compound
7744; purity not specified) in the diet for 90 days at levels of
0.0037, 0.011, 0.033 or 0.10% (Weil, 1956). Assuming that 1 ppm in
the diet of young rats is equivalent to approximately 0.10 mg/kg/day
(Lehman, 1959), this corresponds to doses of about 3.7, 11, 33 or 100
mgAg/day. The author stated that there were no significant changes
in appetite or weight gain when compared to the control; micropathology
revealed no changes in lung, liver or kidney tissue at any dose level.
It was concluded that for these end points the effect level for
toxicity is higher than 0.10%, which is equivalent to a NOAEL of
about 100 mgAg/day (the highest dose tested).
0 Carbaryl was administered to male rats by gavage at a level of
200 mgAg. 3 days a week for 90 days (Dikshith et al., 1976). This
corresponds to an average dose of 86 mg/kg/day. The control animals
received vehicle (peanut oil) on a similar schedule. There were no
overt toxicological signs in these rats, and no marked biological
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Carbaryl August, 1987
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changes were seen in teatea, liver and brain (enzymatic determinations)
except for ChE activity, which was inhibited 34% in blood (p <0.001)
and 11% in brain (p <0.05)e No significant histological changes were
noted in testes, epididymis, liver or kidney. Based on ChE inhibition,
the LOAEL for this study was identified as 86 mg/kg/day.
0 Carpenter et al. (1961) fed carbaryl to male and female Basenji-Cocker
dogs (four or five per dose) for 1 year. Dietary levels were about
0, 24, 95 or 414 ppm, which were adjusted to supply ingested doses of
0, 0.45, 1.8 or 7.2 mg/kg/day. No compound-related effects were
detected on mortality, body weight, hematocrit, hemoglobin, leukocyte
count, blood chemistry, plasma or erythrocyte ChE activity, or liver
and kidney weights. Microscopic examination of tissues revealed dif-
fuse cloudy swelling of renal nephrons and focal debris in glomeruli
of dogs fed the higher dose. These conditions were also observed in
controls, but less frequently, and the authors judged they were not
early stages of toxic degeneration. One dog at the low dose displayed
a transient hind leg weakness after 189 days. This disappeared within
3 weeks, although dosing was continued throughout. Subsequent micro-
scopic examination revealed no differences between this dog and
others. A NOAEL of 7.2 mg/kg/day (the highest dose tested) was
identified.
4 Shering (1963) administered carbaryl (5.0 mg/kg/day) by gavage to 25
male and 25 female rats, 5 days per week for 18 months. No effects
were observed on weight gain, organ weights, urinalysis, heraatology
or histologic appearance of tissues. The authors concluded that 5.0
mg/kg/day was a NOAEL in rats.
0 Carpenter et al. (1961) studied the toxicity of carbaryl in a 2-year
feeding study in rats. Groups of 20 male and 20 female CF-N rats
(60-days old) were maintained on a diet containing 0, 50, 100, 200
or 400 ppm dry Sevin. Based on measured food consumption and body
weights, the authors reported the doses to be equivalent to 0, 2.0,
4.0, 7.9 or 15.6 mg/kg/day in males, and 0, 2.4, 4.6, 9.6 or 19.8
nig/kg/day in females. No adverse effects were detected on life span,
food consumption, body weight gain, liver and kidney weights, cataract
formation or hematocrit. Histological examination after 1 year
revealed mild changes in the kidney, characterized by cloudy swelling
of the nephrons. This was statistically significant (p <0.004) at
the high dose. Cloudy swelling of hepatic chords was also observed
at the high dose, and this was significant after 2 years (p <0.002).
No histological changes were detectable at the lower doses. Based on
these observations, a NOAEL of 7.9 mg/kg/day for males and 9.6 mg/kg/day
for females was identified.
Reproductive Effects
0 Weil et al. (1972) investigated the reproductive effects of carbaryl
in female rats exposed either by gavage or by feeding. Doses of 0,
2.5 and 10 mg/kg/day ingested from the diet for three generations
resulted in no statistically significant, dose-related effects on fer-
tility, gestation, lactation or pup viability. Doses of 100 mg/kg/day
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Carbaryl August, 1987
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given by gavage (5 days/week, beginning at 5 weeks of age) resulted
in maternal mortality, reduced fertility and signs of ChE inhibition.
These signs were not seen in animals ingesting doses of up to 200
from the diet.
Murray et al. (1979) assessed the reproductive effects of carbaryl
(99%) in rabbits (New Zealand White). Pregnant females were given
either 150 or 200 mg/kg/day by gavage from days 6 through 18 of
gestation. The incidence of pregnancy was not significantly affected
at either dose level. On days 6 through 11, carbaryl-treated rabbits
gained significantly less weight than the controls (p
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Carbaryl August, 1987
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veight gain and number of implantations, the NOAEL in this study was
identified as 10 mg/kg/day.
• Golbs et al. (1975) orally administered carbaryl to Wistar rats at
doses of 200 or 350 mg/kg on days 5, 7 and 9, or on days 11, 13 and
15 of the gestation period. In one group of rats, 200 mg/kg was admin-
istered on days 5, 7, 9, 11, 13 and 15. Doses of 350 mg/kg given during
late gestation (days 11 to 15) delayed fetal development, whereas the
same dose given at the earlier interval (days 5 to 9) resulted in
loss of fertilized ova and more pronounced retardation in development
of individual fetuses. Similar results were produced by the 200-mg/k?
dose given on alternate days from day 5 through day 15. It was
concluded that carbaryl produces dose-dependent effects on intrauterine
development in rats. Based on this study, a LOAEL of 200 mg/kg (100
mg/kg/day) was identified.
0 Collins et al. (1970) reported (abstract) the effects of carbaryl in
the diet on various reproductive parameters over three generations of
rats. Osborne-Mendel rats were fed 0, 2,000, 5,000 or 10,000 ppm
carbaryl in the diet. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to doses of about 0, 100, 250 or 500 mg/kg/day. At 10,000 ppm, no
litters were produced after the first litter of the second generation;
decreases were observed in the fertility, viability, survival and
lactation indices in all litters at this dose. The survival index
also showed a decrease at the 5,000-ppn level. Dose-related decreases
were observed in the ratio of average number of animals weaned per
number of litters at both 5,000 and 10,000 ppm. At all three dose
levels there was a decrease in weanling weights. In rats, the LOAEL
was identified as 2,000'ppm (100 mg/kg/day).
0 Collins et al. (1970) reported (abstract) the effects of carbaryl in
a three-generation study in gerbils. Carbaryl was fed at dose levels
of 0, 2,000, 4,000, 6,000 or 10,000 ppm. Assuming that 1 ppm in the
diet of gerbils is equivalent to 0.05 mg/kg/day (Lehman, 1959), this
corresponds to doses of about 0, 100, 200, 300 or 500 mg/kg/day. No
second litters were produced in the third generation at 10,000 ppm.
Decreases in the viability index were observed at 6,000 and 10,000 ppm.
Dose-related decreases in the survival index were also observed.
The average number of animals weaned per litter was also decreased.
Based on these findings, a LOAEL of 6,000 ppm (300 mg/kg/day) and a
NOAEL of 4,000 ppm (200 mg/kg/day) were identified.
Developmental Effects
0 Weil et al. (1972) exposed pregnant Harlan-Wistar rats to carbaryl
in the diet on days 5 to 15 of gestation. Ingested doses were 0, 20,
100 or 500 mg/kg/day. Animals were sacrificed on days 19 to 21, and
fetuses were examined for soft-tissue and skeletal abnormalities. No
increased incidence of teratogenic anomalies was detected at any dose
level. Based on this information, a NOAEL of 500 mg/kg/day (the
highest dose tested) was identified.
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Carbaryl August, 1987
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* Murray et al. (1979) administered 200 mg/kg/day carbaryl to female
rabbits by gavage on days 6 to 18 of gestation. Fetuses were removed
and examined for developmental defects. There was a significantly
(p <0.05) higher incidence of omphalocele in fetuses from exposed
animals than in the controls. The anomalies occurred in litters from
does that showed the greatest weight losses during the experimental
period. Mo other anomalies were seen at this dose level. At
150 mg/kg/day, there were single cases of omphalocele, hemivertebrae
and conjoined nostrils with missing nasal septum, but no fetal alterations
occurred at an incidence significantly different from that of the
control group. Based on fetal defects, the LOAEL for the rabbit was
identified as 150 mg/kg/day.
0 Murray et al. (1979) studied the teratogenic effects of carbaryl
in CF-1 mice. Carbaryl was administered by gavage at 100 or
150 mg/kg/day, or by feeding in the diet at 5,660 ppm (calculated by
the authors to be equivalent to 1,166 ing/kg/day). No major malformations
were detected among the offspring of dams given carbaryl by either
route at incidences significantly different than concurrent or histo-
rical controls. Delayed ossification of skull bones and of sternebrae
occurred significantly more often among litters from dams given
carbaryl in the diet, but not in litters from gavage-administered
dams. Based on developmental observations in fetuses, the NOAEL in
this study was identified as 150 mg/kg/day.
0 Lechner and Abdel-Rahman (1984) administered carbaryl to Sprague-Dawley
rats by gavage for 3 months prior to and throughout gestation at doses
of 0, 1, 10 or 100 mg/kg/day. Dams were sacrificed on day 20, and
fetuses were examined for external, skeletal and visceral malforma-
tions. There were no statistically significant increases of serious
anomalies at any dose level. The authors concluded that in the rats
tested, carbaryl displayed no evidence of teratogenicity. On this
basis, a NOAEL of 100 mg/kg/day (the highest dose tested) was identified.
0 Benson et al. (1967) fed mice carbaryl in their diet (intake levels
of 10 or 30 mg/kg/day) during gestation. Some dams were allowed to
deliver naturally, and others were delivered by Cesarean section.
There were no differences between the offspring of the two treated
groups and the controls in sex ratio, incidence of anomalies or in
ossification. Based on this information, a NOAEL of 30 mg/kg/day
(the highest dose tested) was identified.
Mutagenicity
0 The effects of pesticides on scheduled and unscheduled DNA synthesis
of rat thymocytes and human lymphocytes were studied by Rocchi et al.
(1980). Carbaryl (99.2% pure) in the rat thymocyte culture inhibited
thymidine uptake 15, 22 and 99% at levels of 1, 10 and 100 ug/mL,
respectively. In the human lymphocytes, a dose of 50 ug/mL produced
62% inhibition on scheduled DNA synthesis, but had no effect on
unscheduled DNA synthesis.
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Carbaryl August, 1987
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Carcinogenicity
• Carpenter et al. (1961) fed carbaryl to groups of CF-N rats
(20/sex/dose) for 2 years. Concentrations in the diet were 0, 50,
100, 200 or 400 ppm, reported by the authors to be equal to doses of
0, 2.0, 4.0, 7.9 or 15.6 mg/kg/day in males and 0, 2.4, 4.6, 9.6 or
19.8 mg/kg/day in females. Based on gross and histological examina-
tion of tissues, no increased frequency of any tumor type was detected.
The total number of tumors seen at each of the five concentrations
tested was 10, 12, 8, 9, 12 and 11, respectively.
0 Shering (1963) dosed 25 male and 25 female rats by gavage with
5.0 mg/kg/day carbaryl for 18 months. Based on histological examination
of tissues, no effects of carbaryl on tumor frequency were detected.
0 Carbaryl (30 mg/kg/day) was administered by gavage to mongrel rats
daily for 22 months (Andrianova and Alekseev, 1969). At the termi-
nation of the study, 46 of the original 48 controls survived and one
animal had a malignant tumor. In the treated rats, 12 of the original
60 survived to 22 months, and 4 of these had malignancies (25%). It
was concluded that carbaryl was carcinogenic in this investigation.
0 Zabezhinski (1970) studied the carcinogenicity of beta-Sevin (the
2-napthol analog of carbaryl, often an impurity in technical Sevin).
Mice and rats (CC57H) were fed beta-Sevin in the diet five times per
week for their lifetime. Mice were fed 10 mg for 24 months and rats
25 mg for 33 months. On the assumption that this refers to mgAg/day
(translation does not use that designation), the average daily
consumption would be 7 mg/kg/day for mice, and 17 mg/kg/day for rats.
At the end of the experiment, 31% (8/26) of the surviving mice had
malignancies. The author noted that some of the tumor types were
occasionally observed in control mice, but at a much lower frequency.
Of the original 50 rats, several died due to nephrosis and other ail-
ments that were attributed to the carbaryl. Of the 16 rats surviving
to the end of the study, 4 had malignancies. No malignancies were
observed in the controls. It was concluded that beta-Sevin had a
weak carcinogenic effect in mice and 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 ( /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL - No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
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Carbaryl August, 1987
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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
No data were found in the available literature that were suitable for
determination of the One-day HA value. It is recommended that the Ten-day HA
value for a 10-kg child (1.0 mg/L, calculated below) be used at this time as
a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The study by Weil et al. (1968) has been selected to serve as the basis
for determination of the Ten-day HA for the 10-kg child. This study identified
a NOAEL of 10 mg/kg/day in rats fed carbaryl in the diet for 7 days, based on
inhibition of ChE in plasma and red blood cells.
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA - (10 mg/kg/day) (10 kg) „ K0 mg/L (1f000 ug/L)
(100)0 L/day)
where:
10 mg/kg/day = NOAEL, based on absence of effects on ChE in rats
exposed to carbaryl in the diet for 7 days.
10 kg = assumed body weight of a child.
100 « uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
1 L/day « assumed daily water consumption of a child.
Longer-term Health Advisory
No data were found in the available literature that were suitable for
the determination of a Longer-term HA value. It is, therefore, recommended
that the DWEL, adjusted for a 10-kg child (1.0 mg/L) be used as a conservative
estimate of the Longer-term HA value.
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
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Carbaryl August, 1987
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(RfD), formerly called the Acceptable Daily Intake (ADZ). 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
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 2-year feeding study in rats by Carpenter et al. (1961) has been
selected to serve as the basis for determination of the Lifetime HA for
carbaryl. This study identified a NOAEL of 9.6 mg/kg/day, based on absence
of effects on mortality, body weight, organ weight, hematology, cataract
frequency or histopathology. This value is supported by a 1-year feeding
study in dogs, which identified a NQAEL of 7.2 mg/kg/day (Carpenter et al.,
1961), and an 18-month oral study in rats, which identified a NOAEL of 5.0
mg/kg/day (Shering, 1963); however, these latter studies were not selected
because exposure was less-than-lifetime.
Using the NOAEL of 9.6 mg/kg/day, the Lifetime HA for carbaryl is calcu-
lated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (9.6 mg/kg/day) « 0.,
(100)
where:
9.6 mg/kg/day = NOAEL, based on absence of adverse effects in rats
fed carbaryl in the diet for 2 years.
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 = (0.1 mg/kg/day) (70 kg) = 3.5 ng/L {3|500 ug/L)
(2 L/day)
where:
0.1 mgAg/day - RfD.
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Carbaryl August, 1987
-15-
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 - (3.5 mg/L) (20%) - 0.70 mg/L (700 ug/L)
where:
3.5 mg/L - OREL.
20% « assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer (IARC) (1976) has
classified carbaryl in Group 3; i.e., this chemical cannot be
classified as to its carcinogenicity for humans.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), carbaryl may be classified
in Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA/Office of Research and Development determined an Acceptable
Daily Intake (ADI) of 0.096 mg/kg/day based on a rat chronic oral NOAEL
of 9.6 mg/kg/day (Carpenter, 1961) with an uncertainty factor of 100.
0 The National Academy of Sciences (MAS) determined an ADI of 0.082
mg/kg/day based on a rat chronic oral NOAEL of 8.2 mg/kg/day (Union
Carbide, 1958) and an uncertainty factor of 100.
0 The NAS has also determined a Suggested-No-Adverse-Response-Level
(SNARL) of 0.574 mg/L, based on an ADI of 0.082 mg/kg/day (70-kg adult
consuming 2 L/day and a 20% source contribution factor) (NAS, 1977).
0 The U.S. EPA has established residue tolerances for carbaryl in or
on raw agricultural commodities that range from 0.1 to 100 ppm (CFR,
1985).
VII. ANALYTICAL METHODS
0 Analysis of carbaryl is by a high-performance liquid chromatographic
(HPLC) procedure used for the determination of N-methylcarbamoyloximes
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
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Carbaryl August, 1987
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column, the compounds are hydrolyzed with sodium hydroxide. The
methyl aaine formed during hydrolysis is reacted with o-phthalaldehyde
(OPA) to form a fluorescent derivative that is detected using a
fluorescence detector. The method detection limit has been estimated
to be approximately 0.7 ug/L for carbaryl.
VIII. TREATMENT TECHNOLOGIES
• Available data indicate that granular-activated carbon (GAG) adsorption,
ozonation and conventional treatment trill remove carbaryl from water.
The percentage removal efficiency ranged from 43 to 99%.
0 Whittaker (1980) determined adsorption isotherms using GAC on laboratory-
prepared carbaryl in water solutions.
0 Pilot studies proved that GAC is 99% effective for carbaryl removal
(Whittaker et al., 1980 and 1982). Two columns, each packed with 37 kg
(80 Ibs) of two different GAC, were studied at an empty bed contact
time of 8 minutes and an optimum flow rate of 1 gpm.
0 Laboratory studies for both batch and flow-through columns were used
to examine carbaryl adsorption on two different GAC particle sizes
(Whittaker et al., 1982). Data were fitted to both Langmuir and
Freundlich isotherms; the monolayer capacity was calculated to be
800 moles carbaryl/gm and 1,250 moles carbaryl/gm for the 1.2 mm and
0.6 mm GAC, respectively.
0 Ozonation has been 99% effective in removing carbaryl and its
hydrolysis product, napthol, from aqueous solution (Shevchenko et al.,
1982). Carbaryl and napthol were not detected in the treated effluent
after the addition of 24.8 mg/L and 4.8 mg/L, of ozone respectively.
Before ozonation can be used to treat carbaryl contaminated drinking
water, however, the identity and toxicity of the resulting degradates
must be established.
0 Conventional water treatment by alum coagulation, 30-minute settling
period and filtration removed 56% of the carbaryl present (Whittaker
et al., 1982). Alum dosage of 100 mg/L plus the addition of 1 mg/L
of anionic polymer achieved this degree of removal of carbaryl from
wast ewater.
0 A 3-^3ay settling period without any chemical treatment yield a 50%
carbaryl concentration reduction (Holiday and Hardin, 1981).
0 Treatment technologies for the removal of carbaryl from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt carbaryl
removal from water must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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Carbaryl August, 1987
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IX. REFERENCES
Andrianova, M.M. and Z.V. Alekseev.* 1969. Carcinogenic properties of
Sevin, Maneb, Ciram and Cineb. Vopr. Pitan. 29:71-74. unpublished
report. MRID 00080671.
Benson, B., W. Scott and R. Beliles.* 1967. Sevin: safety evaluation by
teratological study in the mouse. Unpublished report. MRID 00118363.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.169. July 1, 1985.
pp. 274-276.
Carpenter, C.P., C.S. Weil, P.E. Palm, M.W. Woodside, J.H. Nair and H.F. Smyth.
1961. Mammalian toxicity of 1-napthyl-N-methylcarbamate (Sevin insecticide),
J. Agr. Food Chen. 9:30-39.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. In; U.S. EPA.
1985. U.S. Environmental Protection Agency. Pesticide survey chemical
profile. Final Report. Contract No. 68-01-6750. Office of Drinking Water.
Chin, B.H., J.M. Eldridge and L.J. Sullivan. 1974. Metabolism of carbaryl
by selected human tissues using an organ-maintenance technique. Clin.
Tbxicol. 7(1):37-56.
Collins, T.F.X., W.H. Hansen and H.V. Keeler. 1970. The effects of carbaryl
on reproduction of the rat and the gerbil. Toxicol. Appl. Pharmacol.
17(1)i273.
Comer, S.W., D.C. Staiff, J.F. Armstrong and H.R. Wolfe. 1975. Exposure of
workers to carbaryl. Bull. Environ. Contain. Toxicol. 1 3(4):385-391.
Dikshith, T.S.S., P.K. Gupta, J.S. Gaur, K.K. Datta and A.K. Mathur. 1976.
Ninety day toxicity of carbaryl in male rats. Environ. Res. 12:161-170.
Feldman, R.J. and H.I. Maibach.* 1974. Percutaneous penetration of some
herbicides in man. Toxicol. Appl. Pharmacol. 28:126-132. Unpublished
report. MRID 00031050.
Falzon, M., Y. Fernandez, C. Cambon-Gros and S. Mitjavila. 1983. Influence
of experimental hepatic impairment on the toxicokinetics and the
anticholinesterase activity of carbaryl in the rat. J. Appl. Toxicol.
3(2):87-89.
Gaines, V.B.* 1960. The acute toxicity of pesticides to rats. Toxicol. Appl.
Pharm. 2:88-99. MRID 00005467.
Golbs, S., M. Kuehnert and F. Leue. 1975. Prenatal toxicity of Sevin
(carbaryl) for Wistar rats. Arch. Exp. Veterinaermed. 29(4):607-614.
Holiday, A.D. and D.P. Hardin. 1981. Activated carbon removes pesticides
from wastewater. Chem. Eng. 88(6):88-89.
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Carbaryl August, 1987
-18-
Houston, J.B., D.G. Upshall and J.W. Bridges. 1974. Pharmacokinetlcs and
metabolism of two carbamate insecticides, carbaryl and landrin, in the
rat. Xenobiotica. 5(10)i637-648.
1ARC. 1976. 'International Agency for Research on Cancer. IARC monographs
on the evaluation of carcinogenic risk of chemicals to man. Lyon, France:
IARC. 12x37-48.
Xhasawinah, A.M. and G.C. Holsing.* 1977a. Hydrolysis of carbaryl in aqueous
buffer solutions. In: Metabolism and environmental fate, Carbaryl
Registration Standard. Unpublished study received Nov. 30, 1984 under
264-327} submitted by Union Carbide Corporation, Research Triangle Park,
N.C. Accession No. 255799.
Khasavinah, A.M. and G.C. Holsing.* 1977b. Photodegradation of carbaryl in
aqueous buffer solutions. In: Metabolism and environmental fate,
Carbaryl Registration Standard. Unpublished study received Nov. 30,
1984 under 264-327; submitted by Union Carbide Corporation, Research
Triangle Park, N.C. Accession No. 255799.
Xhasawinah, A.M. and G.C. Holsing.* 1978. Fate of carbaryl in soil. In:
Metabolism and environmental fate, Carbaryl Registration Standard.
Unpublished study received Nov. 30, 1984 under 264-327; submitted by
Union Carbide Corporation, Research Triangle Park, N.C. Accession No.
255799.
Krishna, J.G. and J.E. Casida.* 1965. Fate of ten variously labeled methyl -
and dimethyl-carbamate-CI 4 insecticide chemicals in rats. Unpublished
report. MRID 00049134.
Lechner, D.M.W. and M.S. Abdel-Rahman. 1984. A teratology study of carbaryl
and malathion mixtures in rat. J. Toxicol. Environ. Health. 14:267-278.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., P.O. Box 1494, Topeka, Kansas.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Murray, F.J., R.E. Staples and B.A. Schwetz. 1979. Teratogenic potential of
carbaryl given to rabbits and mice by g-vage or by dietary inclusion.
Toxicol. Appl. Pharmacol. 51(1):81-89.
MAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
Rittenhouse, J.R., J.K. Narcisso and R.D. Cavalli.* 1972. Acute oral toxicity
to rats of Orthene in combination with five other cholinesterase-inhibiting
materials. Unpublished report. MRID C0014933.
Rocchi, P., P. Perocco, W. Alberghini, A. Fini and G. Prodi. 1980. Effect
of pesticides on scheduled and unscheduled DNA synthesis of rat thymocytes
and human lymphocytes. Arch. Toxicol. 45:101-108.
-------
Carbaryl August, 1987
-19-
Shering, A.G.* 1963. Promecarb (SN 34615): long-term feeding study in rats:
ZK No. 3858. Unpublished report. HRID 00081723.
Shevchenko, M.A., P.N. Taran and P.V. Marchenko. 1982. Modern methods for
purifying water from pesticides. Soviet Journal of Water Chemistry and
Technology. 4(4):53-71.
STORET. 1987.
Union Carbide. 1958. Chronic oral feeding of Sevin to rats. Internal Report
No. 21-88. Cited in: MAS. 1977. National Academy of Sciences. Drinking
water and health. Washington, DC: National Academy Press.
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.
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185):33992-34003. September 24.
Vanderkar, M. 1965. Observations of the toxicity of carbaryl, folithion and
3-isopropylphenyl-N-methy 1 car-hamate in a village-scale trial in southern
Nigeria. Bull. W.H.O. 33:107-115. MRID 000365173.
Weil, C.J.* 1956. Special report on subacute oral toxicity studies on
Compound 7744. Unpublished report. MRID 00076124.
Weil, C., M.W. Woodside, J. Bernard, D. Crawfod and P. Baker.* 1968. Sevin:
results of feeding in the diet of rats for one week and for one week plus
one day on control diets. Unpublished report. MRID 00118393.
Weil, C.S., M.W. Woodside, C.P. Carpenter and H.F. Smyth. 1972. Current
status of tests of carbaryl for reproductive and teratogenic effects.
Toxicol. Appl. Pharmacol. 21:390-404.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. PhD. Thesis, Purdue
University.
Whittaker, K.F., J.C. Nye, *•F. Wukasch and H.A. Kazimier. 1980. Cleanup
and collection of wastewater generated during the cleanup of pesticide
application equipment. Control of Hazardous Material Spills, Proceedings
of a National Conference, pp. 141-144.
Whittaker, K.F., J.C. Nye, R.F. Wukasch, R.J. Squires, A.C. York and H.A.
Kazimier. 1982. Collection and treatment of wastewater generated by
pesticide application. EPA Report No. 600/2-82-028.
Wills, J.H., E. Jameson and F. Coulston. 1968. Effects of oral doses of
carbaryl on man. Clin. Toxicol. 1:265-271.
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Carbaryl August, 1987
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Hindholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, 10th ed. Rahway, NJ: Merck and Co., Inc. pp. 246-247.
Zabezhinski, M.A.* 1970. Possible carcinogenic effect of (beta)-Sevin.
Voprosy Onkoologii. 16:106-107. In Russian: translation. Unpublished
report. MRID 00086672.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
CARBOXIN
August, 1987
DRAFT
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 th?se models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimate? that are
derived can differ by several orders of magnitude.
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Carboxin August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 5234-68-4
Structural Formula
5,6-Dihydro-2-methyl-N-phenyl-1,4-Oxathin-3-carboxamide
Synonyms
0 Carbathiin; Carboxine; D-735; DCMO; DMOC; F735; Vitavax (Meister,
1983).
Uses
0 Systemic fungicide; seed protectant; wood preservative (Meister,
19831.
Properties (Meister, 1983; Windholz et al., 1983; Vo and Shapiro, 1983;
Worthing, 1983; TOB, 1985)
Chemical Formula C12H1302NS
Molecular Weight 235.31
Physical State (25°C) Crystals
Boiling Point
Melting Point 93 to 95°C
Density
Vapor Pressure (20°C) <1 mm Hg
Specific Gravity
Water Solubility (25°C) 170 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 No information was found in the available literature on the occurrence
of carboxin.
Environmental Fate
0 Carboxin is rapidly metabolized (oxidized by flavin enzymes found in
fungi mitochondria) in aerobic soil. When applied to soil (aerobic
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Carboxin August, 1987
-3-
conditions), more than 95% of the carboxin was degraded within 7
days. The major degradation product was carboxin sulfoxide, which
represented 31 to 54% of the applied radioactivity at 7 days after
treatment. Several minor degradation products were also formed
(carboxin sulfone, £-hydroxy carboxin and 14CO2). Carboxin was
degraded in sterile soil but at a much slower rate than in nonsterile
soil (46 to 72% degraded in 7 days). This would indicate that soil
metabolism of carboxin under aerobic conditions is primarily by
microbial processes. Carboxin sulfoxide is stable in anaerobic soil
(Chin et al., 1972, 19&9, 1970a,b; Dzialo and Lacadie, 1978; Dzialo
et al., 1978; Spare, 1979).
0 Carboxin sulfoxide, a major metabolite of carboxin, photodegrades to
unknown compounds. After 7 days of incubation, 49% of the applied
radioactivity was present as unknown compounds (Smilo et al., 1977).
0 Carboxin does not readily adsorb to soil [K value (adsorption coeffi-
cient) <1] and both carboxin and carboxin sulfoxide are very mobile
in soil with about half of the applied radioactivity leaching through
12-inch columns of clay loam soils (Lacadie et al., 1978; Dannals
et al., 1976).
0 In aqueous solution, carboxin was oxidized to carboxin sulfoxide and
carboxin sulfone within 7 days (Chin et al., 1970a).
III. PHARMACOKINETICS
Absorption
0 Waring (1973) administered carboxin (Vitavax) by gavage to groups
of four to six female New Zealand White rabbits (age not specified;
2.5 to 3 kg) and Wistar rats (age not specified; 200 to 250 g) at
1 mmol/kg (235 mg/kg). In the rats, an average of 40% of the dose
was excreted in the feces, mostly as unchanged carboxin. In the
rabbits, an average of 10% was recovered in the feces. These data
suggest that carboxin is not completely absorbed from the gut,
especially in rats.
Distribution
0 waring (1973) administered single oral doses of carboxin (Vitavax,
6.3 uCi/rat) to female Wistar rats (age not specified; 200 to 250 g).
Carboxin was labeled either in the heterocyclic or aromatic ring and
distribution of label was assessed by autoradiography of whole-body
sections. After 2 hours, label was localized in the liver, intestinal
tract and salivary gland. After 6 hours, label was also present in
the kidney. Only trace levels remained in any tissue after 48 hours.
There were no differences in the distribution of the two labeled
compounds.
0 Nandan and Wagle (1980) fed carboxin to male albino rats (age not
specified) for 28 days at dietary levels of 0, 100, 1,000 or 10,000
-------
Carboxin
August, 1987
-4-
pptn. Based on the dietary assumptions of Lehman (1959), 1 ppm in the
diet of rats equals approximately 0.05 mg/kg/day. Therefore, these
levels correspond to 0, 5, 50 and 500 mg/kg/day. In animals fed the
highest dose, maximum levels were detected in the liver (140 ug/g),
with lower levels in the kidney (123 ug/g), heart (58 ug/g) and
muscle (22 ug/g).
Metabolism
0 In the study by Waring (1973), as described previously, female New
Zealand White rabbits (age not specified; 2.5 to 3 kg) and Wistar
rats (age not specified; 200 to 250 g) were given single oral doses of
carboxin by gavage at 1 mmol/kg (235 mg/kg). The principal metabolic
pathway was found to be ortho- or parahydroxylation, followed by
glucuronidation. In the rats, 32% of the dose was excreted in urine
as glucuronides and 7% as unconjugated phenols. In the rabbits, 85%
of the dose was excreted in urine as glucuronides and 3% as free
phenols. The pattern of phenolic metabolites was the same for carboxin
labeled in either the heterocyclic or the aromatic rings, indicating
that cleavage of the compound did not occur.
Excretion
In the study by Waring (1973), as described previously, female New
Zealand White rabbits (age not specified; 2.5 to 3 kg) and Wistar
rats (age not specified; 200 to 250 g) were given single oral doses
of carboxin by gavage at 1 mmol/kg (235 mg/kg). In the rats, 41% was
excreted in the feces (largely unchanged carboxin) and 54% was excreted
in the urine (15% parent compound, 32% glucuronides, 7% free phenols).
In the rabbits, 10% was excreted in the feces and 90% was excreted in
the urine (2% parent compound, 85% glucuronides, 3% free phenols).
IV. HEALTH EFFECTS
Humans
A seven-year-old boy developed headaches and vomiting within 1 hour
after ingesting several handfuls of wheat seed treated with carboxin.
He was administered ipecac (an emetic) and was asymptomatic 2 hours
later. No estimate o:T the ingested dose was provided (PIMS, 1980).
Animals
Short-term Exposure
Reagan and Becci (1983) reported that the acute oral LDso for tech-
nical carboxin (purity not specified) in young CD-I mice (age not
specified) was 4,150 mg/kg for males and 2,800 mg/kg for females.
The average LD$Q was reported to be 3,550 mg/kg.
RTECS (1985) reported that the acute oral LD$Q for carboxin (purity
not specified) in the rat (age not specified) was 430 mg/kg.
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Carboxin August, 1987
-5-
0 Nandan and Wagle (1980) fed carboxin to male albino rats (age not
specified) for 28 days at dietary levels of 0, 100, 1,000 or 10,000
ppm. Based on the authors' measurements of food consumption and
assuming average body weights of 0.1 kg, these levels corresponded to
doses of about 0, 5.5, 59.0 or 311 mg/kg/day. A Lowest-Observed-
Adverse-Effect-Level (LOAEL) of 100 ppm (5.5 mg/kg/day) was tentatively
identified in this study based on fluid accumulation in the liver.
However, due to a number of deficiencies in this study, it is not
possible to accurately evaluate its validity. These deficiencies
include a lack of information on the test animals (e.g., condition at
study initiation, numbers used) and the absence of statistical analyses.
Dermal/Ocular Effects
0 Holsing (1968a) applied carboxin (D-735; purity and vehicle not
specified) to the intact or abraded abdominal skin of rabbits
(10/sex/dose; age not specified) at concentrations of 1,500 or
3,000 mg/kg. Five animals of each sex served as controls. Test
animals were exposed occlusively for 6 to 8 hours, 5 days per week,
for 3 weeks (15 applications). No signs of dermal irritation were
observed. The test material stained the skin and precluded readings
for erythema.
Long-term Exposure
0 Ozer (1966) administered carboxin (D-735; purity not specified) to
weanling FDRL (Wistar-derived) rats (10/sex/dose; controls: 15/sex)
for 90 days at dietary concentrations of 0, 200, 600, 2,000, 6,000
or 20,000 ppm, intended by the author to correspond to approximate
dosage levels of 0, 10, 30, 100, 300 or 1,000 mg/kg/day. All animals
survived the 90-day treatment period. Growth, food efficiency,
hematology, blood chemistry and urinalysis were reported to be similar
in all groups with the exception of increased blood urea nitrogen and
decreased hemoglobin at the 12-week interval in females that received
20,000 ppm (1,000 mg/kg/day). No significant dose-related gross
pathological changes were observed. Microscopically, a significant
number of inflammatory degenerative renal changes were found in
animals that received doses of 600 ppm (30 mg/kg/day) or higher.
These changes included focal chronic inflammation, protein casts and
cortical tubular degeneration. In two animals that received 2,000 ppm
(100 mg/kg/day), some fibrosis in the medulla was observed. Based on
renal changes, a LOAEL of 600 ppm (30 mg/kg/day) and a No-Observed-
Adverse-Effect-Level (NOAEL) of 200 ppm (10 mg/kg/day) can be identified.
0 Jessup et al. (1982) administered carboxin (technical Vitavax; purity
not specified) to six-week old Charles River CD-I mice (50/sex/dose;
controls: 75/sex) for approximately 84 weeks at dietary concentra-
tions of 0, 50, 2,500 or 5,000 ppm. The authors indicated that these
dietary levels corresponded to doses of about 0, 8, 385 or 751 mg/kg/day
for males and 0, 9, 451 or 912 mg/kg/day for females. No compound-
related effects on general behavior or appearance were reported.
Survival rates of females receiving 5,000 ppm (912 mg/kg/day) were
significantly (p <0.01 ) lower than controls. No compound-related
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Carboxin August, 1987
-6-
effects on body weight gain, food consumption, or various hematological
parameters were reported. No gross pathologic lesions that were
considered to be related to compound administration were observed
at necropsy in any mice in any treatment group. Microscopically,
compound-related effects on the liver, consisting of hypertrophy of
the centrilobular parenchymal cells, were observed in mice in the
2,500- or 5,000-ppm dose groups (385 and 751 mg/kg/day for males; 451
and 912 mg/kg/day for females). No other nonneoplastic lesions that
could be attributed to compound administration were observed. The
NOAEL in this study is 50 ppm (8 mg/kg/day for males; 9 mg/kg/day for
females) based on hepatic effects.
Holsing (1969a) administered carboxin (technical D-735; considered
to be 100% active ingredient) to Charles River rats (30/sex/dose;
controls: 60/sex) for 2 years at dietary concentrations of 0, 100,
200 or 600 ppm. Based on the dietary assumptions of Lehman (1959),
1 ppm in the diet of rats equals approximately 0.05 mg/kg/day.
Therefore, these dietary levels correspond to dose levels of approxi-
mately 0, 5, 10 or 30 mg/kg/day. While the age of the animals was
not specified, the weights of the male rats at initiation ranged from
65 to 88 g and the weight of the female rats ranged from 59 to 85 g.
No compound-related effects in terms of physical appearance, behavior,
hematology, blood chemistry or urinalysis were reported at any dose
level. Observations at terminal necropsy did not reveal any compound-
related gross or microscopic changes in the organs of animals at any
dose level. At the 600-ppm level (30 mg/kg/day), body weight gain
was significantly depressed in both sexes, and food consumption by
males was lower than that of controls throughout most of the study
(significantly lower during the first 26 weeks). Food consumption by
females at all dose levels was generally comparable to controls.
Compound-related effects included an increase in mortality at 18
months in males that received 600 ppm (30 mg/kg/day), and changes in
absolute and relative organ weights at all dose levels, including
increases in thyroid weight and decreases in kidney, heart and spleen
weight and histopathological changes in the kidneys at the 12-month
interval in both sexes at 200 and 600 ppm. Most of these effects
were inconsistent and were not observed at the end of the study
period. At the end of the 2-year study, decreased kidney weights
were observed in males at 600 ppm (30 mg/kg/day). Therefore, based
on the information presented in this study, a NOAEL of 200 ppm
(10 mg/kg/day) was identified.
Holsing (1969b) administered carboxin (technical D-735; considered
to be 100% active ingredient) to young adult beagle dogs (4/sex/dose;
controls: 6/sex) for 2 years at dietary concentrations of 0, 100,
200 or 600 ppm. Based on the dietary assumptions of Lehman (1959),
1 ppm in the diet of rats equals approximately 0.05 mg/kg/day.
Therefore, these dietary levels have been calculated to correspond
approximately to 0, 2.5, 5.0 or 15.0 mg/kg/day. No treatment-related
effects were reported on survival, body weight gain, food consumption,
organ weights, organ-to-body weight ratios, hematological, blood
chemistry or urinary parameters, liver and kidney function tests or
gross and histopathological observations. Based on this information,
-------
Carboxin August, 1987
-7-
a NOAEL of 600 ppm (15 mg/kg/day; the highest dose tested) was
identified.
Reproductive Effects
0 In a three-generation reproduction study, Holsing (1968b) administered
carboxin (technical D-735; 97% active ingredient) to Charles River
rats (10 males/dose, 20 females/dose; controls: 15 males, 30 females)
(age not specified) at dietary concentrations of 0, 100, 200 or 600 ppm.
Based on the dietary assumptions of Lehman (1959), these dietary levels
have been calculated to correspond to dose levels of approximately
0, 5, 10 or 30 mg/kg/day. Criteria evaluated included fertility,
gestation, live birth and lactation indices, litter size and the
physical appearance and growth of the pups. No compound-related*
effects on reproductive performance were reported at any dose level.
A compound-related effect on the progeny (moderate growth suppression
in the nursing male and female pups of all three generations) was
observed at the 600-ppm (30 mg/kg/day) dose level. Based on the
information presented in this study, a NOAEL of 200 ppm (10 mg/kg/day)
was identified.
Developmental Effects
0 Schardein and Laughlin (1981) administered technical Vitavax
(carboxin; 99% active ingredient) by gavage at doses of 0, 75, 375
or 750 mg/kg/day to seven- to eight-month-old Dutch Belted rabbits
(10/dose) on days 6 through 27 of gestation. The compound was
administered in a 0.5% carboxymethyl cellulose vehicle. No treatment-
related effects on maternal mortality, appearance, behavior or body
weight were reported. Four females aborted on days 27 and 28 of
gestation (one at 375 mg/kg/day, three at 750 mg/kg/day). Examination
for fetal malformations revealed no compound-related differences
between the control and treatment groups. Based on the frequency of
abortion, a NOAEL of 75 mg/kg/day and a LOAEL of 375 mg/kg/day were
identified.
0 Knickerbocker (1977) administered carboxin (technical Vitavax; purity
not specified) in corn oil by gavage at doses of 0, 4, 20 or 40 mg/kg/day
to sexually mature (age not specified) Sprague-Oawley rats (20/dose)
on days 6 through 15 of gestation. No compound-related effects were
observed on reproduction, gestation ot in skeletal or soft tissue
development. Based on the information presented, a NOAEL of 40
mg/kg/day (the highest dose tested) was identified.
Mutagenicity
0 Brusick and Weir (1977) conducted a mutagenicity assay using Salmonella
typhimurium strains TA 1535, 1537, 1538, 98 and 100, and Saccharomyces
cerevisiae strain D4. Carboxin (purity not specified) was tested
without activation at concentrations up to 500 ug/plate and with
activation at concentrations up to 100 ug/plate. No mutagenic activity
was detected in this assay.
-------
Carboxin August, 1987
-8-
0 Byeon et al. (1978) reported that carboxin (Vitavax; purity not
specified) tested at concentrations up to 1 mg/plate was not found to
be mutagenic in an Ames assay using j>. typhimurium strains TA 1535,
1538, 98 and 100. ~~
• Brusick and Rabenold (1982) conducted an Ames assay using technical
carboxin (Vitavax, 98% active ingredient) at concentrations up to
5,000 ug/plate. No mutagenic activity was detected, with or without
activation, in £. typhimurium strains TA 1535, 1537, 1538, 98 and 100.
0 Myhr and McKeon (1982) reported the results of a primary rat hepatocyte
unscheduled DNA synthesis assay using carboxin (technical Vitavax;
98% active ingredient). The test compound produced significant
increases in the nuclear labeling of primary rat hepatocytes over a
concentration range of 5.13 to 103 ug/mL.
Carcinogenicity
0 Holsing (1969a) administered carboxin (technical D-735; considered to
be 100% active ingredient) to Charles River rats (30/sex/dose; controls:
60/sex) for 2 years at dietary concentrations of 0, 100, 200 or 600 ppm.
Based on the dietary assumptions of Lehman (1959), 1 ppm in the diet
of rats equals approximately 0.05 mg/kg/day. While the age of the
animals was not specified, the weights of the male rats at initiation
ranged from 65 to 88 g and the weights of the female rats ranged from
59 to 85 g. Therefore, dietary levels correspond to approximately 0,
5, 10 or 30 mg/kg/day. No evidence of increased tumor frequency*was
detected by either gross or histological examination of tissues.
0 Jessup et al. (1982) administered carboxin (technical Vitavax; purity
not specified) to six-week-old Charles River CD-1 mice (50/sex/dose;
controls: 75/sex) for approximately 84 weeks at dietary concentra-
tions of 0, 50, 2,500 or 5,000 ppm. The authors indicated that these
dietary levels corresponded to dosage levels of approximately 0, 8,
385 or 751 mg/kg/day for males and 0, 9, 451 or 912 mg/kg/day for
females. Survival rates of females receiving 5,000 ppm (912 mg/kg/day)
were significantly (p <0.01) lower than those of controls. No compound-
related gross pathologic lesions were observed at necropsy in any
treatment group. Microscopically, compound-related effects on the liver,
. consisting of hypertrophy of the centrilobular parenchymal cells, were
observed in mice in the 2,500 or 5,000 ppm dose groups (385 and 751
mg/kg/day for males; 451 and 912 mg/kg/day for females). In males, the
incidence of pulmonary adenoma/alveolar-bronchiolar adenoma was 13/75,
7/49, 7/50, and 17/50 at 0, 50, 2,500, and 5,000 ppm, respectively.
The incidence at the high dose (34%) may have been compound-related
based on comparison with the incidence in controls (17%). The difference
was statistically significant (p <0.01) using Cox's test for adjusted
trend and the Kruskail Wallis tests for life-table data and adjusted
incidence. However, based on the opinions of pathologists who reviewed
the data and on historical data on tumor incidence in control Charles
River CD-1 mice, the authors concluded that the increased incidence
was not compound-related. Historical data indicate that in six
18-month studies, the incidence of lung adenomas ranged from 6.3 to
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Carboxin August, 1987
-9-
16.7%; in seven 20- to 22-month studies, the incidence of lung adenomas
ranged from 4.0 to 31.1%.
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 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/OEW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
Appropriate data for calculating a One-day HA value are not available.
It is recommended that the Longer-term HA value for the 10-kg child (1.0 mg/L,
calculated below) be used as the One-day HA value.
Ten-day Health Advisory
Appropriate data for calculating a Ten-day HA value are not available.
The 22-day rabbit teratogenicity study by Schardein and Laughlin (1981)
was considered for the development of the Ten-day HA. However, the NOAEL
(75 mg/kg/day) identified in this study is far in excess of the NOAEL
(10 mg/kg/day) identified in the 90-day rat feeding study reported by Ozer
(1966) suggesting that the rat is the more sensitive species. It is, therefore,
recommended that the Longer-Term HA value for the 10-kg child (1.0 mg/L,
calculated below) be used as the Ten-day value.
Longer-term Health Advisory
The study by Ozer (1966) has been selected to serve as the basis for
calculating the Longer-term HA for carboxin. In this study, weanling rats
were exposed to carboxin in the diet for 90 days. At 30 mg/kg/day there was
histological evidence of renal injury. At 10 mg/kg/day, no effects were
detected on any parameter measured, including growth, hematology, blood
chemistry, urinalysis, gross pathology and histopathology. Based on these
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Carboxin August, 1987
-10-
data, a NOAEL of 10 mg/kg/day was identified. This value is supported by the
subchronic (84 week) feeding study in mice by Jessup et al. (1982) which
identified a NOAEL of 8 to 9 mg/kg/day, based on the absence of effects on
appearance, behavior, mortality, weight gain, hematology, gross pathology and
histopathology.
The Longer-term HA for the 10-kg child is calculated as follows:
Longer-term HA = (10 mg/kg/day) (10 kg) = K0 ng/L (1,000 ug/L)
(100) (1 L/day)
where:
10 mg/kg/day = NOAEL, based on absence of effects on growth, hematology,
blood chemistry, urinalysis, gross pathology and
histopathology in rats exposed to carboxin in the diet
for 90 dayse
10 kg a 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 - (1° ng/*g/day) (70 kg) = 3.5 ng/L (3,500 ug/L)
(100) (2 L/day)
where:
10 mg/kg/day = NOAEL, based on absence of effects on growth, hematology,
blood chemistry, urinalysis, gross pathology and
histopathology in rats exposed to carboxin in the diet
for 90 days.
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
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Carboxin August, 1987
-11-
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Holsing (1969a) has been selected to serve as the basis for
calculation of the Lifetime HA for carboxin. In this study, rats were exposed
to carboxin in the diet for 2 years. At 10 mg/kg/day, no significant effects
were detected on appearance, behavior, body weight, mortality, hematology,
blood chemistry, urinalysis, gross pathology or histopathology. Based on
these data, a NOAEL of 10 mg/kg/day was identified. This value is supported
by a 90-day rat study (Ozer, 1966) which also identified a NOAEL of 10 mg/kg/day,
a 2-year feeding study in dogs by Holsing (1969b) which identified a NOAEL of
15 mg/kg/day, and an 84-week mouse study (Jessup et al., 1982) which identified
a NOAEL of 8 mg/kg/day for males and 9 mg/kg/day for females.
Using the NOAEL of 10 mg/kg/day, the Lifetime HA for carboxin is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD „ (10 mg/kg/day) = Q., mgAg/day
(100)
where:
10 mg/kg/day = NOAEL, based on absence of effects on appearance,
behavior, body weight, mortality, hematology, blood
chemistry, urinalysis, gross pathology or histopathology
in rats exposed to carboxin in the diet for 2 years.
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 = (0-1 mg/kg/day) (70 kg) = 3.5 mg/L (3,500 ug/L)
(2 L/day)
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Carboxin August, 1987
-12-
where:
0.1 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 = (3.5 mg/L) (20%) =0.7 mg/L (700 ug/L)
where:
3.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Jessup et al. (1982) reported a possible compound-related increase
in pulmonary adenoma/alveolar-bronchiolar adenoma frequency in male
CD-1 mice that received carboxin in the diet at 751 mg/kg/day.
0 Holsing (1969a) fed Charles River rats carboxin at dietary levels up
to 30 mg/kg/day for 2 years, and detected no compound-related histo-
pathologic changes. This study is limited, however, by the following
factors: inadequate numbers of animals were used; survival was
generally poor and, therefore, late-developing lesions may not have
been detected; all tissues from all animals were not examined micro-
scopically; and there was no adjustment in dietary levels of carboxin
to account for growth of the test animals.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of carboxin.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), carboxin is classified in
Group D: not classified. This category is for substances with
inadequate human and animal evidence of carcinogenicity or for which
no data are available.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No existing criteria or standards for oral exposure to carboxin were
located.
0 The U.S. EPA (OPP) has proposed an Acceptable Daily Intake (ADI) of
0.4 mg/kg/day, based on a NOAEL of 200 ppm established in a 2-year
rat feeding study and an uncertainty factor of 100 (U.S. EPA, 1981).
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Carboxin August, 1987
-13-
• The U.S. EPA has established residue tolerances for carboxin in or
on raw agricultural commodities that range from 0.01 to 0.5 ppm
(CFR, 1979).
VII. ANALYTICAL METHODS
0 Analysis of carboxin is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing pesti-
cides in water samples (U.S. EPA, 1986b). In this method, approximately
1 liter of sample is extracted with methylene chloride. The extract
is concentrated and the compounds are separated using capillary
column GC. Measurement is made using a nitrogen phosphorus detector.
The method detection limit has not been determined for carboxin but
it is estimated that the detection limits for analytes included in
this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information regarding treatment techniques to remove carboxin from
contaminated waters is currently available.
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Carboxin August, 1987
-14-
IX. REFERENCES
Brusick, D.J., and R.J. Weir.* 1977. Mutagenicity evaluation of D-735.
CBZ Project No. 1683. Final Report 53727. Unpublished study.
MRID 00053727.
Brusick, D., and C. Rabenold.* 1982. Mutagenicity evaluation of technical
grade Vitavax in the Ames Salmonella microsome plate test. CBI Project
No. 20988. Final Report. Unpublished study. MRID 00132453.
Byeon, W., H.H. Hyun and S.Y. Lee.* 1978. Mutagenicity of pesticides in the
Salmonella/microsomal enzyme activation system. Korean J. of Microbiol.
14:128-134. MRID 00061590.
Chin, W.T., L.E. Dannals and N. Kucharczyk.* 1972. Environmental Fate studies
on Vitavax. (Unpublished study submitted by Uniroyal Chemical, Bethany,
Conn. CDL:093515-A.) MRID 00002935.
Chin, W.T., G.M. Stone and A.E. Smith.* 1969. Fate of D735 in soil.
(Unpublished study submitted by Uniroyal Chemical, Bethany, Conn.
CDL:091420.) MRID 00003041.
Chin, W.T., G.M. Stone and A.E. Smith.* 1970a. Degradation of carboxin
(Vitavax) in water and soil. J. Agric. Food Chem. 18(4):731-732.
MRID 05002176.
Chin, W.T., G.M. Stone, A.E. Smith and B. von Schmeling.* 1970b. Fate of
carboxin in soil, plants, and animals. In; Proc. Fifth British
Insecticide and Fungicide Conf., Nov. 17-20, 1969, Brighton, England.
Vol. 2. pp. 322-327. MRID 05004996.
CFR. 1979. Code of Federal Regulations. 40 CFR 180.301. July 1, 1979.
p. 527.
Dannals, L.E., C.R. Campbell and R.A. Cardona.* 1976. Environmental fate
studies on Vitavax. Status report II on PR 70-15. Includes three
updated methods. (Unpublished study submitted by Uniroyal Chemical,
Bethany, Conn. CDL:223866-A.) MRID 00003114.
Dzialo, D.G., and J.A. Lacadie.* 1978. Aerobic soil study of 14C-Vitavax in
sandy soil: Project no. 7746-1. (Unpublished study submitted by Uniroyal
Chemical, Bethany, Conn. CDL.-236662-F.) MRID 00003225.
Dzialo, D.G., J.A. Lacadie, and R.A. Cardona.* 1978. Anaerobic soil metabolism
of 14c-Vitavax in sandy soil. (Unpublished study submitted by Uniroyal
Chemical, Bethany, Conn. CDL:236662-G.) MRID 00003226.
Holsing, G.C.* 1968a. Summary: Repeated dermal (Leary design) - rabbits.
Project No. 798-148. Unpublished study. MRID 00021626.
Holsing, G.C.* 1968b. Three-generation reproduction study - rats. Final
Report. Project No. 798-104. Unpublished study. MRID 00003032.
-------
Carboxin August. 1987
-15-
Holsing, G.C.* 1969a. 24-Month dietary administration - albino rats. Final
Report. Project No. 798-102. Unpublished study. MRID 00003031.
Holsing, G.C.* 1969b. Two-year dietary administration - dogs. Final Report.
Project No. 798-103. Unpublished study. MRID 00003030.
Jessup, D., G. Gunderson and R. Gail.* 1982. Lifetime carcinogenicity study
in mice (Vitavax): 399-002a. Unpublished study. MRID 00114139.
Knickerbocker, M.* 1977. Teratologic evaluation of Vitavax technical in
Sprague-Dawley rats. Unpublished study. MRID 00003102.
Lacadie, J.A., D.R. Gerecke and R.A. Cardona.* 1978. Vitavax 1*C laboratory
column leaching study in clay loam: Project no. 7758. (Unpublished
study submitted by Uniroyal Chemical, Bethany, Conn. CDL:236662-H.)
MRID 00003227.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S. Q. Bull.
Matthews, R.J.* 1973. Acute LD50 rats, oral. Final Report. Unpublished
study. MRID 00003012.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Myhr, B., and M. McKeon.* 1982. Evaluation of Vitavax technical grade in the
primary rat hepatocyte unscheduled DNA synthesis assay. CBI Project No.
20991. Unpublished study. MRID 00132454.
Nandan, D., and D.S. Wagle. 1980. Metabolic effects of carboxin in rats.
Symp. Environ. Pollut. Toxicol. pp. 305-312.
Ozer, B.L.* 1966. Report: Subacute (90 day) feeding studies with D-735 in
rats. Unpublished study. MRID 00003063.
PIMS. 1980. Pesticide Incident Monitoring System. Summary of reported
incidents involving carboxin. Report No. 383. Health Effects Branch,
Hazard Evaluation Division, Office of Pesticide Programs, U.S. Environ-
mental Protection Agency, Washington, D.C. October 1980.
Reagan, E., and P. Becci.* 1983. Acute oral LD50 assay in mice: (Vitavax
Technical): FDRL Study No. 7581A. Unpublished study. MRID 00128469.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. National
Institute for Occupational Safety and Health. National Library of
Medicine Online File.
Schardein, J.L., and K.A. Laughlin.* 1981. Teratology study in rabbits:
399—042. Unpublished study. MRID 00086054.
-------
Carboxin August, 1987
-16-
Smilo, A.R., J.A. Lacadie and B. Cardona.* 1977. Photochemical fate of
Vitavax in solution. (Unpublished study submitted by Uniroyal Chemical,
Bethany, Conn. CDL:231932-C.) MRID 00003088.
Spare, W.* 1979. Report: Vitavax microbial metabolism in soil and its effect
on microbes, (unpublished study prepared by Biospherics, Inc., in
cooperation with United States Testing Co., Inc., submitted by Uniroyal
Chemical, Bethany, Conn. CDL:098029-A.) MRID 00005540.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA. 1981. U.S. Environmental Protection Agency. Carboxin. Pesticide
Registration Standard. Office of Pesticides and Toxic Substances,
Washington, DC.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of Nitrogen and Phosphorus Containing Pesticides in
Ground Water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Waring, R.N. 1973. The metabolism of Vitavax by rats and rabbits.
Xenobiotica. 3:65-71.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck Index, 10th ed. Rahway, NJ: Merck and Co., Inc.
Wo, C», and R. Shapiro.* 1983. EPA acute oral toxicity. Report No. T-3449.
Unpublished study. MRID 00143944.
Worthing, C. R. 1983. The Pesticide Manual. British Crop Protection Council.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
CHLORAMBEN
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 mod.il is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Chloraoiben
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 133-90-4
Structural Formula:
COOH
NH2 'CI
3-Amino-2-5-dichlorobenzoic acid
Synonyms
Uses
0 Acp-m-728; Ambiben; Abiben; Amibin; Amoben; Chlorambed; Chloranbene;
NCI-C00055 ornamental weeder; Ornamental weeder* Vegaben; Vegiven
(U.S. EPA, 1985).
0 Pre-emergent herbicide for weed control (Meister, 1983).
Properties (U.S. EPA, 1985; CHEMLAB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
C7H502NC12
206.02
Crystals
200-201 °C
7 x 10-3 mm Hg (100°C)
700 mg/L
2.32
Samples were collected at 5 surface water locations and 188 ground
water locations, and chloramben was found in only 1 state. The 85th
percentile of all nonzero samples was 2.1 ug/L in surface water and
1.7 ug/L in ground water sources. The maximum concentration found
was 2.3 ug/L in surface water and 1.7 ug/L in ground water (STORET,
1987).
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Chloramben August, 1987
-3-
Environmental Fate
0 Sodium chloramben appears to be resistant to hydrolysis. Limited
studies indicate that there is no loss of phytotoxicity when aqueous
solutions of chloramben are kept in the dark (Registrant CBI data).
0 Photodegradation of aqueous solutions of sodium chloramben appears
to occur readily in sunlight. Total loss of phytotoxicity occurs in
2 days. Loss of phytotoxicity on dry soil is somewhat slower, about
30% in 48 hours (Registrant CBI data).
0 Soil bacteria bring about a loss of phytotoxicity in sodium chloramben
after several weeks. It appears that this is due to a decarboxylation.
The rate of reaction appears to be independent of soil pH within the
range of 4.3 to 7.5 (Registrant CBI data).
0 The mobility of sodium chloramben is governed principally by its high
solubility in water and its apparent limited strength of adsorption
to soil particles. It appears to easily leach down in most soil
types by rainfall (Registrant CBI data).
0 Probably all plants grown in contact with sodium chloramben take up
the compound. In some plants the subsequent movement of compound
away from the roots is very slow, whereas in others it readily spreads
throughout the plant. The fate of chloramben in plants includes
decomposition, a detoxifying conjugation which proceeds fairly rapidly,
and a detoxifying conjugation which goes slowly, if at all (Registrant
CBI data).
0 The methyl ester of chloramben acid appears to have the expected
properties of a carboxylic acid ester. It is apparently not hydrolysed
after a short period in contact with water at slightly acid pH values
(5 to 6). Bacteria-mediated hydrolysis appears to be quick: approxi-
mately 50% of the ester is converted to the free acid in about 1 week
when in contact with wet soil. A subsequent and slower bacterial
reaction, shown by a loss of phytotoxicity, is probably a decarboxy-
lation, as with sodium chloramben (Registrant CBI data).
0 The leaching behavior of the methyl ester is governed by its aqueous
solubility, which is much lower than that of the sodium salt (120 ppm
and 250,000 ppm, respectively). For a given rainfall the ester seems
to leach down about 15% of the distance travelled by the sodium salt
(Registrant CBI data).
III. PHARMACOKINETICS
Absorption
0 Chloramben is rapidly absorbed from the gastrointestinal tract of
Sprague-Dawley female rats (Andrawes, 1984). Based on radioactivity
recovered in urine (96.7%) and expired air (0.2%), about 97% of an
oral dose (5 uCi/rat) of chloramben is absorbed.
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Chloramben August, 1987
-4-
Distribution
0 Andrawes (1984) reported low levels (up to 0.5% of the administered
dose) of chloramben in liver, kidney, lung, muscle, plasma and red
blood cells of rats 96 hours after a single oral dose (by gavage).
Metabolism
0 In rats dosed by gavage. Andrawes (1984) reported that the parent
compound accounted for 70% of the applied dose in 24-hour urine.
0 Andrawes (1984) identified 5 of 24 urinary metabolites: 3-amino-5-
chlorobenzoic acid; 3-aminobenzoic acid; 2,5-dihydroxybenzoic acid;
3,5-dihydroxybenzoic acid; and 2,5-dichloroaniline. Together, these
constituted 1.4% of the administered dose.
0 Metabolism of chloramben in rats proceeded through dechlorination,
deamination, decarboxylation and hydroxylation. Metabolism through
oxidative ring cleavage was negligible (Andrawes, 1984).
Excretion
Rats administered chloramben (5 uCi/rat) by gastric intubation excreted
over 99% of the dose within 3 to 4 days, mostly within the first
24 hours (Andrawes, 1984). Approximately 96.7% was eliminated in the
urine, with lesser amounts in the feces (4.1%) and respiratory gases
(0.2%). Only 0.6% remained in the carcass after 3 to 4 days.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the human
health effects of chloramben.
Animals
Short-term Exposure
0 Acute oral 1.050 values for chloramben range from 2,101 mg/kg (Field,
1980a) to 5,000 mg/kg (Field, 1978a) in rats; the acute dermal LDso
in rabbits has been reported to be >2,000 (Field, 1980b) or
>5,000 mg/kg (Field, 1978b).
0 Rees and Re (1978) reported an acute (1 hr) LC50 of >200 mg/L in rat
inhalation studies.
0 Keller (1959) fed male Holtzman Sprague-Dawley rats (10/dose) chloramben
(100% a.i.) for 28 days in the diet at dose levels of 0, 1,000, 3,000
or 10,000 ppm. Assuming that 1 ppm in the diet of rats is equivalent
to 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of 0, 50,
150 or 500 mg/kg/day. Body weights, food consumption, general appearance
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Chloramben August, 1987
-5-
and behavior and histopathology were evaluated. There were no statis-
tically significant differences between the treated rats and untreated
controls in any parameter measured. Based on this information, a No-
Observed-Adverse-Effect-Level (NOAEL) of 10,000 ppm (500 ing/kg/day),
the highest dose tested, was identified.
Dermal/Ocular Effects
0 Gabriel (1969) applied chloramben (4 or 8 g/kg) to intact and
abraded skin of 16 male albino rabbits (8/dose). Test animals were
observed for 14 days. No evidence of skin irritation was observed
under conditions of the study.
0 In a study by Myers et al. (1982), a 1.0% (w/w) chloramben sodium
salt suspension produced little or no sensitization reactions in male
albino Hartley guinea pigs.
Long-term Exposure
0 In studies by Beliles (1976), weanling Golden Syrian hamsters
(12/sex/dose) were administered technical chloramben (purity not
specified) at dose levels of 0, 100, 1,000 or 10,000 ppm (reported to
be equivalent to 0, 11, 115 or 1,070 mg/kg/day) in the diet for
90 days. Food consumption, body and organ weights and histopathology
were evaluated. No treatment-related adverse effects were reported
for any parameter evaluated. Based on this information, a NOAEL of
10,000 ppm (1,070 mg/kg/day), the highest dose tested, was identified.
0 In an 18-month feeding study (Huntingdon Research Center, 1978; cited
in U.S. EPA, 1981), Crl:COBS CD-I mice (50/sex/dose) were administered
technical chloramben (purity not specified) at dietary levels of 0,
100, 1,000 or 10,000 ppm. Assuming that 1 ppm in the diet of mice is
equivalent to 0.15 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 15, 150 and 1,500 mg/kg/day. No compound-related
effects were observed in terms of survival, general appearance,
behavior or changes in body weight. Statistically significant
(p <0.05) changes in organ weights included decreased liver weight in
males at 100 ppm, decreased kidney weight in males at 10,000 ppm, and
decreased kidney weight in females at 10,000 ppm. Since the values
for these observations were within normal ranges for this species and
no trends were established, the organ-weight changes were not attributed
to compound administration. Histopathological examinations revealed
alterations in the livers of all treated mice. The primary hepatocellular
reaction was a histomorphological hepatocellular alteration compatible
with that observed in enzyme induction. The typical cellular changes
included hepatocyte hypertrophy, increased nuclear size and chromatin
content, and dense granular eosinophilic cytoplasm. Other changes
included scattered foci of individual or small groups of degenerating
hepatocytes, hepatocyte vacuolation, cytoplasmic eosinophilic inclusions,
and multiple focal small granulomas. Based on the reported hepatic
effects, this study identifies a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 100 ppm (15 mg/kg/day).
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Chloramben August, 1987
-6-
0 NCI (1977) administered technical-grade chloramben (90 to 95% active
ingredient) to Osborne-Mendel rats (50/sex/dose) and B6C3Fi mice
(50/sex/dose) for 80 weeks at dietary levels of 10,000 or 20,000 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
and 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day (Lehman,
1959), this corresponds to doses of 500 or 1,000 mg/kg/day for rats
and 1,500 or 3,000 mg/kg/day for mice. Matched controls consisted of
10 animals per sex for each species. Pooled controls consisted of
the matched controls plus 75 rats/sex and 70 mice/sex from similarly
performed bioassays. Body weights and mortality did not differ
between control and treatment groups for both species, and the various
(unspecified) clinical signs observed were similar in the control and
treatment groups for both species. Based on this information, a
NOAEL of 20,000 ppm (1,000 mg/kg/day for rats and 3,000 mg/kg/day for
mice), the highest dose tested, was identified for each species.
0 In studies conducted by Paynter et al. (1963), albino rats
(35/sex/dose) were administered chloramben (97% pure) in the diet for
2 years at dose levels of 0, 100, 1,000 or 10,000 ppm. Assuming that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), this corresponds to doses of 0, 5, 50 or 500 mg/kg/day.
Untreated rats (70/sex/dose) were observed concurrently. The general
appearance and behavior, growth, food consumption, clinical chemistry,
hematology and histopathology in the treated rats did not differ
significantly from the untreated controls. Based on this information,
a NOAEL of 10,000 ppm (500 mg/kg/day), the highest dose tested, was
identified.
0 Hazleton and Farmer (1963) administered technical chloramben (97%
pure) in the feed to 16 young adult beagle dogs (4/sex/dose) for
2 years at dietary levels of 0, 100, 1,000 or 10,000 ppm. Assuming
that 1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day
(Lehman, 1959), this corresponds to doses of 0, 2.5, 25 or 250 mg/kg/day.
General appearance and behavior, food consumption, body weight,
hematology, biochemistry, urinalysis and histopathology of the treated
dogs did not differ significantly from the untreated controls. Based
on this information, a NOAEL of 10,000 ppm (250 mg/kg/day), the highest
dose tested, was identified.
0 Johnston and Seibold (1979) administered technical chloramben to
Sprague-Dawley rats for 2 years at dietary concentrations of 0,
100, 1,000 or 10,000 ppm. Assuming that 1 ppm in the diet: of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959) this corresponds to doses
of 0, 5, 50 and 500 mg/kg/day. No compound-related effects were
observed on any parameters measured including body weight, food
consumption, hematology, clinical chemistry, urinalysis, gross
pathology and histopathology. Based on this information, a NOAEL of
10,000 ppm (500 mg/kg/day), the highest dose tested, was identified.
Reproductive Effects
0 In a three-generation study (Gabriel, 1966), three groups of albino
rats (8 females and 16 males/dose) were administered 0, 500, 1,500 or
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Chloramben August, 1987
-7-
4,500 ppm chloramben (purity not specified) in the diet for 9 weeks
prior to breeding, during breeding and during weaning periods.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), these dietary levels correspond to doses of about 0,
25, 75 or 225 mg/kg/day. Untreated animals served as controls.
Following treatment, various parameters were measured, including
indices of fertility, gestation, viability and lactation. No adverse
effects were reported in any parameter measured. Based on this
information, a NOAEL of 4,500 ppm (225 mg/kg/day), the highest dose
tested, was identified for reproductive effects.
Developmental Effects
0 Beliles and Mueller (1976) administered technical chloramben (purity
not specified) to pregnant CFE rats (20/dose) by incorporation into
the diets on days 6 through 15 of gestation. No compound-related
changes were seen among dams treated at levels of 0, 500, 1,500 and
4,500 ppm. Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of about 0,
25, 75 or 225 mg/kg/day. Fetal mortality was increased, and data
suggestive of decreased fetal skeletal development were observed in
fetuses from dams treated at 4,500 ppm (225 mg/kg/day). At 1,500 ppm
(75 mg/kg/day), there was no significant increase in embryo mortality;
however, there was a generalized reduction in skeletal development.
Fetuses of dams treated with 500 ppm (25 mg/kg/day) were similar in
all respects to those of untreated control dams. Based on this
information, a NOAEL of 4,500 ppm (225 mg/kg/day), the highest dose
tested, was identified for maternal toxicity and teratogenicity. The
NOAEL for fetotoxicity was identified as 500 ppm (25 mg/kg/day).
0 Holson (1984) conducted studies in which New Zealand White rabbits
(24/dose) were administered chloramben (sodium salt, 83% a.i. by weight)
by gavage at dose levels of 0, 250, 500 or 1,000 rag/kg during days
6 through 18 of gestation. A NOAEL of 1,000 mg/kg/day, the highest
dose tested, was identified, since the test compound did not produce
maternal or fetal toxicity or teratogenic effects at any dose level
tested. Other end points were not monitored.
Mutagenicity
0 Chloramben was found to be negative in several indicator systems for
potential mutagenic activity, including several microbial assays
(Anderson et al., 1967; Eisenbeis et al., 1981; Jagannath, 1982), an
in vivo bone marrow cytogenetic assay (Ivett, 1985) and primary rat
hepatocytes unscheduled DNA synthesis test (Myhr and McKeon, 1982).
0 Results were positive for the in vitro cytogenic test using Chinese
hamster ovary cells (Galloway and Lebowitz, 1982).
Carcinogenicity
0 In an 18-month feeding study (Huntingdon Research Center, 1978; cited
in U.S. EPA, 1981), Crl:COBS CD-1 mice (50/sex/dose) were administered
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Chloranben August, 1987
-8-
technical chloramben (purity not specified) at dietary levels of 0,
100, 1,000 or 10,000 ppm. Assuming that 1 ppm in the diet of mice is
equivalent to 0.15 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 15, 150 and 1,500 mg/kg/day (Lehman, 1959).
Hepatocellular carcinomas (trabecular type) were present in 1/50 low-
dose and 1/50 high-dose males. In no case was vascular invasion or
secondary spread of the nodular carcinoma masses observed. Hepatocellular
adenomas were present only in males as follows: 5/50 control, 2/50
low-dose, 2/48 intermediate-dose and 5/50 high-dose. However, due to
a number of deficiencies in this study (e.g., missing data, significant
tissue autolysis), no conclusion can be made regarding the oncogenic
potential of the test material.
0 NCI (1977) administered 10,000 or 20,000 ppm technical chloramben
(90 to 95% active ingredient) in the feed to Osborne-Mendel rats
(50/sex/dose) and B6C3F-) mice (50/sex/dose) for 80 weeks followed by
up to 33 weeks of postexposure observation. Assuming that 1 ppm in
the diet of rats is equivalent to 0.05 mg/kg/day and 1 ppm in the
diet of mice is equivalent to 0.15 mg/kg/day (Lehman, 1959), this
corresponds to doses of 500 or 1,000 mg/kg/day for rats and 1,500 or
3,000 mg/kg/day for mice. Under conditions of the study, no compound-
related tumors were reported in male or female rats or male mice.
Hepatocellular carcinomas were reported in female mice, but in a
. retrospective audit of this bioassay by Drill et al. (1982), it was
reported that the incidence of hepatocellular carcinomas in both the
low-dose and high-dose female mice was lower than the maximal
incidence of corresponding tumors in historical groups. It was
concluded that there was no association between chloramben and the
occurrence of hepatocellular carcinomas under conditions of the assay.
However, since exposure was for only 80 weeks, this study may not
have been adequate to detect late-occurring tumors.
0 Paynter et al. (1963) reported no evidence of carcinogenic activity
in albino rats (35/sex/dose) that received chloramben (97% pure) in
the diet for 2 years at dose levels of 0, 100, 1,000 or 10,000 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959) this corresponds to doses of 0, 5, 50 or 500 mg/kg/day.
0 Johnston and Seibold (1979) reported no evidence of carcinogenic
activity in Sprague-Oawley rats administered 0, 100, 1,000 or
10,000 ppm technical chloramben in the diet for 2 years. Assuming
that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), this corresponds to doses of 0, 5, 50 or 500 mg/kg/day. No
compound-related effects were observed on any other parameters measured,
including body weight, food consumption, hematology, clinical chemistry,
urinalysis, gross pathology and histopathology.
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 toxicityc
The HAs for noncarcinogenic toxicants are derived using the following formula:
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Chloramben August, 1987
-9-
where:
HA 0 (NOAEL or LOAEL) X (BW) = mg/L ( ug/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).
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
No data were found in the available literature that were suitable for
determination of the One-day HA value. It is, therefore, recommended that
the Ten-day HA value for a 10-kg child (2.5 mg/L, calculated below) be used
at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The rat teratology study by Beliles and Mueller (1976) has been selected
to serve as the basis for determination of the Ten-day HA value for a 10-kg
child for Chloramben. In this study, a NOAEL of 225 mg/kg/day, the highest
dose tested, was identified for maternal toxicity and teratogenicity while a
NOAEL of 25 mg/kg/day was identified for fetotoxicity (skeletal development)
in rats exposed on days 6 to 15 of gestation. There is some question as to
whether it is appropriate to base a Ten-day HA for the 10-kg child on
fetotoxicity observed in a teratology study. However, this study is of
appropriate duration and the fetus may be more sensitive than the 10-kg
child.
The studies by Keller (1959) and*Holson (1984) have not been selected,
since the NOAEL values identified in these studies (500 and 1,000 mg/kg/day,
respectively) are much higher than the NOAEL identified by Beliles and Mueller
(1976).
Using the NOAEL of 25 mg/kg/day, the Ten-day HA for the 10-kg child is
calculated as follows:
Ten-day HA = (25 mg/kg/day) (10 kg) = 2.5 /L (2 50Q /L)
(100) (1 L/day)
where:
25 mg/kg/day = NOAEL, based on the absence of systemic toxic effects
in rats fed Chloramben for 10 days.
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Chloramben August, 1987
-10-
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 Advisories
No data were found in the available literature that were suitable for
the determination of the Longer-term HA. It is, therefore, recommended that
an adjusted DWEL for a 10-kg child (0.15 mg/L - 150 ug/L) and the DWEL for
a 70-kg adult (0.525 mg/L - 525 ug/L) be used at this time for the Longer-
term HA values.
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
(RfO), 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 18-month feeding study by the Huntingdon Research Center (1978;
cited in U.S. EPA, 1981) has been selected to serve as the basis for determina-
tion of the Lifetime HA for chloramben. In this study, Crl:COBS CD-1 mice
were administered technical chloramben at dietary levels of 0, 100, 1,000 or
10,000 ppm (0, 15, 150 or 1,500 mg/kg/day). Hepatocellular alterations were
observed in mice in all treatment groups, and a LOAEL of 100 ppm (15 mg/kg/day)
was identified. Other studies of appropriate duration identify NOAELs that
are higher than the LOAEL of 15 mg/kg/day. For example, Hazleton and Farmer
(1963) identified a NOAEL of 250 mg/kg/day in a 2-year study in dogs, and
both Paynter et al. (1963) and Johnston and Siebold (1979) identified a
NOAEL of 500 mg/kg/day in 2-year rat studies.
Using the LOAEL of 15 mg/kg/day, the Lifetime HA for chloramben is
calculated as follows:
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Chloramben August, 1987
-11-
Step 1: Determination of the Reference Dose (RfD)
KfD = (IS mg/kg/day) _ 0.015 nig/kg/day
(1,000)
where:
15 mg/kg/day = LOAEL, based on hepatic effects in mice exposed to
chloramben via the diet for 18 months.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.015 mg/kg/day) (70 kg) = 0.525 mg/L (525 ug/L)
(2 L/day)
where:
0.015 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 = (0.525 mg/L) (20%) = 0.105 mg/L (105 ug/L)
where:
0.525 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 NCI (1977) evaluated the carcinogenic potential of orally admini-
stered chloramben (10,000 or 20,000 ppm, equivalent to 500 or 1,000
mg/kg/day) to Osborne-Mendel rats (50/sex/dose) and B6C3Fi mice
(20/sex/dose) for 80 weeks. It was concluded in a retrospective
audit of this assay (Drill et al., 1982) that under conditions of
this study, chloramben is not carcinogenic. Since exposure was for
only 80 weeks, this experiment may not have been adequate to detect
late-occurring tumors. Johnston and Seibold (1979) reported no evidence
of carcinogenic activity in Sprague-Dawley rats that received chloramben
in the diet for 2 years at concentrations up to 500 mg/kg/day. The
Huntingdon Research Center (1978; cited in U.S. EPA, 1981) reported
no evidence of carcinogenicity in Crl:COBS CD-I mice that received
chloramben in the diet for 18 months at concentrations up to
1,500 mg/kg/day. However, due to a number of deficiencies in this
study, no conclusion can be made regarding the oncogenic potential
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Chloramben August, 1987
-12-
of the test material. Paynter et al. (1963) reported no evidence of
carcinogenicity in albino rats that received chloramben in the diet
for 2 years at concentrations up to 500 mg/kg/day.
0 The International Agency for Research on Cancer has not evaluated
the carcinogenicity of chloramben.
• Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), chloramben may be classified in
Group D: not classified. This category is for agents with inadequate
human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 NAS has determined an Acceptable Daily Intake of 0.25 mg/kg/day with
a Suggested-No-Adverse-Effeet-Level of 1.75 mg/L (U.S. EPA, 1985).
0 The U.S. EPA has established a residue tolerance for chloramben in or
on raw agricultural commodities of 0.1 ppm (CFR, 1985).
VII. ANALYTICAL METHODS
0 Chloramben may be analyzed using a gas chromatographic (GC) method
applicable to the determination of chlorinated acids, ethers and
esters in water samples (U.S. EPA, 1986b). In this method, approx-
imately 1 liter of sample is acidified. The compounds are extracted
with ethyl ether using a separatory funnel. The derivatives are
hydrolyzed with potassium hydroxide, and extraneous organic material
is removed by a solvent wash. After acidification, the acids are
extracted and converted to their methyl esters using diazomethane as
the derivatizing agent. Excess reagent is removed, and the esters
are determined by electron-capture (EC) gas chromatography. The
method detection limit has not been determined for this compound.
VTII. TREATMENT TECHNOLOGIES
0 No data were found for the removal of chloramben from drinking water
by conventional treatment.
0 No data were found for the removal of chloramben from drinking water
by activated carbon treatment. However, due to its low solubility
and its high molecular weight, chloramben probably would be amenable
to activated carbon adsorption.
0 No data were found for the removal of chloramben from drinking water
by ion exchange. However, chloramben is an acidic pesticide and
these compounds have been readily adsorbed in large amounts by ion
exchange resins. Therefore, chloramben probably would be amenable
to an ion exchange.
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Chloramben August, 1987
-13-
0 No data were found for the removal of chloramben from drinking water
by aeration. However, the Henry's Coefficient can be estimated from
available data on solubility (700 mg/L at 25°C) and vapor pressure
(7 x ID'3 mm Hg at 100°C). Due to its estimated Henry's Coeeficient
of 0.15 atm, chloramben probably would not be amenable to aeration or
air stripping.
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Chloramben August, 1987
-14-
IX. REFERENCES
Anderson, K.J., E.G. Leighty and M.T. Takahashi.* 1967. Evaluation of herbi-
cides for possible mutagenic properties. Unpublished study. MRID 00025376.
And r awes, N.* 1984. Amiben: Metabolism of 1 4c-chloramben in the rat. Project
No. 852R10. Union Carbide. Unpublished study. MRID 00141157.
Beliles, R.P.* 1976. Ninety-day toxicity study in hamsters; technical
chloramben. LBI Project No. 2595. Final Report. Unpublished study.
MRID 00131187.
Beliles, R.P. and S. Mueller.* 1976. Teratology study in rats: technical
chloramben. LBI Project No. 2577. Final Report. Unpublished study.
MRID 0096618.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.226. July 1, 1985.
p. 298.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., cited in U.S. EPA.
1984. U.S. Environmental Protection Agency. Pesticide survey chemical
profile. Final Report. Contract No. 68-01-6750. Office of Drinking Water,
Washington, DC.
Drill, Vo, S. Friess, H. Hayes et al. (names not specified).* 1982. Retro-
spective audit of the bioassay of chloramben for possible carcinogen icity.
Unpublished study. MRID 00126379.
Eisenbeis, S.J., D.L. Lynch and A.E. Hampel. 1981. The Ames mutagen assay
tested against herbicides and herbicide combination. Soil Sci.
131 (1):44-47.
Field, W.E. and W. Carter.* 1978a. Oral LD50 in rats. Study No. CDC -AM-01 5-78.
MRID 00100318.
Field, W.E.* 1978b. Acute dermal application (LDso) — rabbit. Study No.
CDC-AM-012-78. Unpublished study. MRID 00100319.
Field, W.* 1980a. Oral LD50 in rats: chloramben 10G. Study No. CDC-UC-1 58.
MRID 00128640.
Field, W. and G. Field.* 1980b. Acute dermal toxicity in rabbits: (AXF-1107).
Study No. CDC UC-16-180. Unpublished study. MRID 00128644.
Gabriel, K.L. * 1966. Reproduction study in albino rats with AmChem Products,
Inc. — AmiJben (3-amino-2,5-dichlorobenzoic acid). Project No. 20-064.
Unpublished study. MRID 00100202.
Gabriel, K.L.* 1969. Acute dermal toxicity-rabbits. Unpublished study.
MRID 00023483.
-------
Chloramben August, 1987
-15-
Galloway, S. and H. Lebowitz.* 1982. Mutagenicity evaluation of chloramben
(sodium salt), in an in vitro cytogenetic assay measuring chromosome
aberration frequencies in Chinese Hamster Ovary (CHO) cells. Project
No. 20990. Final Report. Unpublished study. MRID 00112855.
Hazleton, L.W. and K. Farmer.* 1963. Two year dietary feeding—dog. Final
Report. Unpublished study. MRID 00100201.
Holson, J.* 1984. Teratology study of chloramben sodium salt in New Zealand
White rabbits. Science Applications (1282018). MRID 00144930.
Huntingdon Research Center.* 1978. 18-Month oncogenic study in CD-1 mice.
Study No. HRC #1-362; October 20, 1978. Cited in U.S. EPA, 1981.
EPA Reg. #204-138, Chloramben; 18-month oncogenic study in mice; Accession
#242821-2. U.S. EPA, Office of Pesticide Programs. Washington, DC.
Memorandum from William Dykstra to Robert Taylor dated January 15, 1981.
Ivett, J. 1985.* Clastogenic evaluation of chloramben in the mouse bone
marrow cytogenetic assay. Final Report. LSI Project No. 22202. Unpub-
lished study. MRID 00144363.
Jagannath, D.* 1982. Mutagenicity evaluation of chloramben sodium salt in
Ames Salmonella/microsome plate test. Project No. 20988. Final Report.
Unpublished study. MRID 00112853.
Johnston, C.D. and H.R. Seibold.* 1979. Two-year carcinogenesis study in rats:
technical chloramben: LBI Project No. 20576. Final Report. Unpublished
study. MRID 00029806.
Keller, J.G.* 1959. Twenty-eight day dietary feeding — rats. Unpublished
study. MRID 00100199.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Myers, R., S. Christopher, H. Zimmer-Weaver et al.* 1982. Chloramben sodium
salt: Dermal sensitization study in the guinea pig. Project No. 45-162.
Unpublished study. MRID 00130275.
Myhr, B. and M. McKeon.* 1982. Evaluation of chloramben sodium salt in the
primary rat hepatocyte unscheduled DNA synthesis assay. Project No.
20991. Final report. Unpublished study. MRID 00112854.
NCI. 1977. National Cancer Institute. Bioassay of chloramben for possible
carcinogenicity. Technical Report Series No. 25.
Paynter, O.E., M. Kundzin and T. Kundzin.* 1963. Two-year dietary feeding
— rats. Final Report. Unpublished study. MRID 00100200.
-------
Chloramben August, 1987
-16-
Rees, D.C. and Re Ta.* 1978. Inhalation toxicity of amiben sodium salt 3599
in adult Sprague-Dawley rats. Laboratory No. 5764b. Unpublished study.
MRID 00100322.
STORET. 1987.
U.S. EPA.* 1981. U.S. Environmental Protection Agency. EPA Reg. #264-138,
chloramben; 18-month oncogenic study in mice; Accession #242821-2.
U.S. EPA, Office of Pesticide Programs. Washington, DC. Memorandum
from William Dykstra to Robert Taylor dated January 15, 1981.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Pesticide survey
chemical profile. Final Report. Contract No. 68-01-6750. Office of
Drinking Water. Washington, DC.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Method #3—Determi-
nation of chlorinated acids in ground water by GC/ECD, January, 1986 draft.
Available from U.S. EPA's Environmental Monitoring and Support Laboratory,
Cincinnati, OH 45263.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
CHLOROTHALONIL
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 it* based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Chlorothalonil August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1897-45-6
Structural Formula
Cl
ci
2, 4, 5, 6-Tetrachloro-1 , 3-benzenedicarbonitrile
Synonyms
0 Tetrachloroisophthalonitrile; Bravo; Chloroalonil; Chlorthalonil;
Daconil; Exothern; For turf; Nopcocide N96; Sweep; Termil; TPN; DAC-2787.
Uses (Meister, 1986)
0 Broad-spectrum fungicide.
Properties (Meister, 1986; CHEMLAB, 1985; Meister, 1983; Windholz et al. , 1983)
Chemical Formula
Molecular Weight 265.89
Physical State (25°C) White, crystalline solid
Boiling Point 350°C
Melting Point 250 to 251 eC
Density
Vapor Pressure (40°C) <0.01 mm Hg
Specific Gravity
Water Solubility (25°C) 0.6 mg/L
Octanol/Water Partition 1.32 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Chlorothalonil has been found in the 1 surface water sample analyzed
and in none of the 560 ground water samples (STORET, 1987). Samples
were collected at 1 surface water location and 556 ground water
locations; and the 1 location where it was found in Michigan, the
concentration was 6,500 ug/L.
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Chlorothalonil August, 1987
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Environmental Fate
0 Ring-labeled Hc-chlorothalonil, at °*5 to 1>5 PPm' was stable to
hydrolysis for up to 72 days in aqueous solutions buffered at pH 5
and 7 (Szalkowski, 1976b). At pH 9, Chlorothalonil hydrolyzed with
half-lives of 33 to 43 days and 28 to 72 days in solutions to which
ring-labeled Hc-chlorothalonil was added at 0.52 and 1.5 ppm,
respectively. After 72 days of incubation, pH 9-buffered solutions
treated with Chlorothalonil at 1.5 ppm contained 36.4% Chlorothalonil,
48.9% 3-cyano-2,4,5,6-tetrachlorobenzamide (DS-19211) and 11.3% 4-
hydroxy-2,5,6-trichloroisophthalonitrile (DAC-3701).
0 The degradate Hc-DAC-3701, at 1000 ppm, was not hydrolysed in aqueous
solutions buffered at pH 5, 7, and 9 after 72 days of incubation
(Szalkowski, 1976b).
0 Ring-labeled 14c-chlorothalonil and its major degradate, ring-labeled
14C-DAC-3701, were stable to photolysis on two silt loam and three
silty clay loam soils, after UV irradiation for the equivalent of 168
12-hour days of sunlight (Szalkowski, 19??).
0 14c-Chlorothalonil is degraded with half-lives of 1 to 16, 8 to 31,
and 7 to 16 days in nonsterile aerobic sandy loam, silt loam and peat
loam soils, respectively, at 77 to 95flF and 80% of field moisture
capacity (Szalkowski, 1976a). When Chlorothalonil (WP) was applied
to nonsterile soils ranging in texture from sand to silty clay loam,
at 76 to 100°F and 6% soil moisture, it was degraded with half-lives
of 4 to more than 40 days; increasing either soil moisture content
(0.6 to 8.9%) or incubation temperature (76 to 100°F) enhanced
Chlorothalonil degradation (Stallard and Wolfe, 1967). Soil pH
(6.5 to 8) does not appear to influence or only negligibly influences
the degradation rate of Chlorothalonil; however, soil sterilization
greatly reduced the degradation rate. The major degradate identified
in nonsterile aerobic soil was DAC-3701, representing up to 69% of the
applied radioactivity. Other identified degradates included DS-19221,
trichloro-3-carboxybenzamide, 3-cyanotrichlorohydroxybenzamide, and
3-cyanotrichlorobenzamide (Stallard and Wolfe, 1967; Szalkowski, 1976a;
Szalkowski et al., 1979).
0 14C-Chlorothalonil was immobile (Rf 0.0) and the degradate 14C-DAC-3701
was found to have low to intermediate mobility (Rf 0.25 to 0.43) in
two silt loam and three silty clay loam soils, as evaluated using soil
thin-layer chromotography (TLC) (Szalkowski, 19??). Based on batch
equilibrium tests, Chlorothalonil has a relatively low mobility (high
adsorption) in silty clay loam (K = 26), silt (K = 29), and sandy
loam (K = 20) soils but is intermediately mobile (low adsorption) in
a sand (K = 3) (Capps et al., 1982). Soil organic matter content did
not appear to influence the mobility of Chlorothalonil in soil.
0 The Chlorothalonil degradate DAC-3701 is mobile in sand, loam, silty
clay loam and clay soils (Wolfe and Stallard, 1968a). After eluting
a 6-in soil column with the equivalent of 5 inches of water, approxi-
mately 57, 84, 10 and 84% of the applied DAC-3701 was recovered in
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Chlorothalonil August, 1987
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the leachate of the sand, loan, silty clay loam and clay soil columns,
respectively.
0 Chlorothalonil (4.17 Ib/gal F1C) was degraded with a half-life of
1 to 3 months in sandy loam and silt loam soils when applied alone at
8.34 Ib ai/A or in combination with benomyl (50% wettable powder) at
1.35 Ib ai/A (Johnston, 1981). The treated soils were maintained at
80% of moisture capacity in a greenhouse.
0 Under field conditions, the half-life of Chlorothalonil (75% wettable
powder) in a sandy loam soil was between 1 and 2 months following the
last of five consecutive weekly applications totaling 15 Ib ai/A
(Stallard et al., 1972). Little movement of Chlorothalonil (0.01 to
0.17 ppm) below the 0- to 3-inch depth occurred throughout the 8-month
study. Small amounts (0.01 to 0.21 ppm) of the degradate DAC-3701
were found in soil samples collected up to 5 months post-treatment.
No Chlorothalonil or DAC-3701 was detected (less than 1 ppb) in a
nearby stream up to 7 months post-treatment, or in ground water
samples (10-foot depth) up to 8 months post-treatment. Cumulative
rainfall over the study period was 26.22 inches.
III. PHARMACOKINETICS
Absorption
0 Ryer (1966) administered '4C-Chlorothalonil (dose not specified)
orally to albino rats (3/sex; strain not specified). In 48 hours
post-treatment, 60.21% of the radioactivity was detected in the
feces, indicating that at least 40% of the oral dose was absorbed.
0 Skinner and Stallard (1967) reported that rats receiving 1.54 mg of
14c-chlorothalonil in a 500 mg/kg dose (route not specified)
eliminated 88% of the administered dose unchanged in the feces over
264 hours, indicating that 12% was absorbed.
0 Skinner and Stallard (1967) reported that mongrel dogs receiving
a single oral dose (by capsule) of 500 mg/kg of Chlorothalonil,
eliminated 85% of the administered dose as the parent compound
within 24 hours post-treatment, indicating that 15% was absorbed.
Distribution
0 Ryer (1966) administered 14c-chlorothalonil (dose not specified) to
albino rats (3/sex; strain not specified) by oral intubation. After
11 days, the carcasses retained 0.44% of the dose while 0.05% of the
dose remained in the gastrointestinal tract. The highest residues
occurred in the kidneys, which averaged 0.01% of the dose for the six
rats. Lesser amounts were detected in the eyes, brain, heart, lungs,
liver, thyroid and spleen.
0 Ribovich et al. (1983) administered single doses of 1 ^c-chlorothalonil
by oral intubation to CD-I mice at levels of 0, 1.5, 15 or 105 mg/kg.
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Chlorothalonil August, 1987
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Twenty-four hours post-treatment, the stomach, liver, kidneys, fat,
small intestine, large intestine, lungs and heart accounted for less
than 3% of the administered dose. The stomach and kidneys had the
highest concentration at all doses tested. The compound was
eliminated from the stomach and kidneys by 168 hours post-treatment.
0 Wolfe and Stallard (1968b) reported a study in which dogs and rats
received Chlorothalonil in the diet for 2 years at 1,500 to 30,000 ppm.
The amount of the 4-hydroxy-2,5,6-trichloroisophthalonitrile metabolite
that was detected in the kidney tissue of dogs was less than 1.5 ppm;
less than 3.0 ppm was detected in liver tissue from dogs and rats.
The authors concluded that the metabolite was not stored in animal
tissue.
Metabolism
0 In the Wolfe and Stallard (1968b) study, only a small amount of the
4-hydroxy-2,5,6-trichloroisophthalonitrile metabolite was detected in
the kidney tissue of dogs (<1.5 ppm) and in liver tissue from dogs
and rats (<3 ppm).
0 Marciniszyn et al. (1983) reported that when Osborne-Mendel rats were
administered single oral doses of 14C-chlorothalonil by intubation at
levels of 0, 5, 50, 200 or 500 mg/kg, no metabolites of Chlorothalonil
were unequivocally identified in urine.
Excretion
The Ryer study (1966) revealed that, at the end of 11 days, an average
of 88.45% of the administered dose was excreted in the feces, 5.14% in
the urine and 0.32% in expired gases as C02.
The Skinner and Stallard study (1967) presented results that demon-
strated that 88% of a dose (route unspecified) of Chlorothalonil was
eliminated unchanged in the feces. Only 5.2% was eliminated via the
urine and negligible amounts were detected in expired air.
Ribovich et al. (1983) administered single doses of 14c-chlorothalonil
by oral intubation to CD-I mice at levels of 0, 1.5, 15 or 105 mg/kg.
The total recoveries of radioactivity 24 hours post-treatment were
93% for the low dose, 81% for the mid dose and 62% for the high dose.
The major route of elimination was the feces and was complete at 24
hours post-treatment for the low- and mid-dose animals, and by 96
hours for the high dose animals.
Marciniszyn et al. (1981) reported a study in which single doses of
14c-chlorothalonil were administered intraduodenally to male Sprague-
Oawley rats at 0.5, 5, 10, 50, 100 or 200 mg/kg. Biliary excretion
of radioactivity was monitored for 24 hours. Percent recovery of
radioactivity was 27.8, 20.7, 16.8, 6.4, 7.8 and 6% for each dose
level, respectively.
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Chlorothalonil August, 1987
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Marciniszyn et al. (1983a) administered 14c-chlorothalonil intra-
duodenally to male Sprague-Dawley rats (donor animals) at a dose of
5 mg/kg« Bile was collected for 24 hours following administration.
Some of the collected bile was administered intraduodenally to recipient
rats; bile was also collected from these animals for 24 hours. Data
from the donor rats indicated that 1 to 6% of the administered radio-
activity was excreted in the bile within 24 hours after dosing.
Approximately 19% of the radioactivity in bile administered to recipient
rats was excreted within 24 hours after dosing. These data suggest
that enterohepatic recirculation plays a role in the metabolism of
Chlorothalonil in rats.
Pollock et al. (1983) administered 14C-chlorothalonil by gavage to
male Sprague-Dawley rats at dose levels of 5, 50 or 200 mg/kg. They
subsequently determined blood concentrations of radioactivity. The
authors hypothesized that, at 200 mg/kg, an elimination mechanism
(urinary, biliary and/or metabolism) was saturated, since the kinetics
were nonlinear at this dose.
IV. HEALTH EFFECTS
0 The purity of the administered Chlorothalonil is assumed to be
>90% for all studies described below, unless otherwise noted.
Humans
Johnsson et al. (1983) reported that Chlorothalonil exposure resulted
in contact dermatitis in 14 of 20 workers involved in woodenware
preservation. The wood preservative used by the workers consisted
mainly of "white spirit," with 0.5% Chlorothalonil as a fungicide.
Workers exhibited erythema and edema of the eyelids, especially the
upper eyelids, and eruptions on the wrist and forearms. Results of
a patch test conducted with 0.1% Chlorothalonil in acetone were posi-
tive in 7 of 14 subjects. Reactions ranged from a few erythematous
papules to marked papular erythema with a brownish hue without
infiltration.
Animals
Short-term Exposure
0 Powers (1965) reported that the acute oral LD^Q of Chlorothalonil
(75% wettable powder) in Sprague-Dawley rats was >10 g/kg.
0 Doyle and Elsea (1963) reported that the acute oral LD50 of Chloro-
thalonil in Sprague-Dawley rats was >10 gAg«
0 Rittenhouse and Narcisse (1974) reported that the acute oral LD50 of
Chlorothalonil in Sprague-Dawley rats was >17.4 g/kg.
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Chlorothalonil August, 1987
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Dermal/Ocular Effects
0 Doyle and Elsea (1963) reported that the dermal LD50 of DAC-2787
(technical Chlorothalonil) in albino rabbits was >10 gAg« At dermal
concentrations of 1, 2.15, 4.64 or 10 g/kg (24-hour exposure), the
compound produced mild to moderate skin irritation characterized by
erythema, edema, atonia and desquamation.
0 Doyle and Elsea (1963) reported that when 3 mg of DAC-2787 (technical
Chlorothalonil) was applied to the eyes of albino rabbits, eye
irritation was limited to mild conjunctivitis that subsided largely
or completely within 7 days.
0 Auletta and Rubin (1981) reported the results of eye irritation
studies in cynomologus monkeys and New Zealand White rabbits using a
formulation containing 96% Chlorothalonil. In both species, 0.1 mL
of the test substance was instilled into the conjunct!val sac of one
eye. Each species displayed mild and transient ocular irritation as
evidenced by corneal opacities that were reversed by 4 days post-
instillation. The animals also showed slight to moderate iridial
and conjunctival effects which were also reversible. Rinsing reduced
conjunctival and iridial effects and prevented formation of corneal
opacities.
Long-term Exposure
0 Blackmore and Shott (1968) administered technical grade DAC 2787
(Chlorothalonil) to Charles River rats for 90 days at dietary levels
of 0, 4, 10, 20, 30, 40 or 60 ppm (approximately 0, 0.2, 0.5, 1.0,
1.5, 2.0 or 3.0 mg/kg/day; Lehman, 1959). No compound-related effects
were reported regarding physical appearance, growth, survival, terminal
clinical values, organ weights or organ-to-body weight ratios.
Microscopically, the kidneys exhibited occasional vacuolation and
swelling of the epithelial cells lining the deeper proximal convoluted
tubules. These changes were more numerous and more severe in the two
highest dose groups. The authors stated that the difference between
the two highest dose groups (2.0 and 3.0 mg/kg/day) and the controls
was distinct, but the difference between the lower dose groups and
controls was not clear. Based on this information, a NOAEL of 30 ppm
(1.5 mg/kg/day) is identified.
0 Wilson et al. (1981) administered Chlorothalonil in the diet to
Charles River CD rats (20/sex/dose) for 90 days at doses of 0, 40,
80, 175, 375, 750 or 1,500 mg/kg/day. At doses of 375 mg/kg/day or
higher, significant decreases in body weight were reported. Decreases
in glucose levels, blood urea nitrogen and serum thyroxine were
attributed by the investigators to body weight effects. A dose-related
decrease in serum glutamic-pyruvic transaminase (SGPT) was noted in all
test groups. Significant increases in kidney weights were also noted
in males at 40, 80, 175 and 375 mg/kg, while in females increased
kidney weights were noted at 80, 175 and 750 mg/kg. These were
dose-related increases in kidney-to-body weight ratios in both sexes
at all doses. Focal acute gastritis occurred in some rats of both
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Chlorothalonil August, 1987
-8-
sexes at all doses and this effect was inversely related to dose.
A LOAEL of 40 mg/kg/day (the lowest dose tested) is identified in
this study.
Colley et al. (1983) administered technical-grade Chlorothalonil in
the diet to Charles River rats (27 males and 28 females per dose) for
13 weeks at concentrations of 0, 1.5, 3.0, 10 or 40 ing/kg/day.
Histopathological examination revealed that at a dose of 3.0 mg/kg/day
or greater, all males displayed an increased number of irregular
intracytoplasmic inclusion bodies in the renal proximal convoluted
tubules. A NOAEL of 1.5 mg/kg/day is identified in this study.
Shults et al. (1983) administered technical-grade Chlorothalonil to
CD-I mice for 90 days at dietary concentrations of 0, 7.5, 15, 50, 275
or 750 ppm (approximately 0, 1.1, 2.3, 7.5, 33.8 or 112.5 mg/kg/day;
Lehman, 1959). No treatment-related effects were noted on survival,
physical condition, body weight, food consumption or gross pathology.
At 750 ppm (112.5 mg/kg/day), an increase in alkaline phosphatase
levels was observed in females only. Increased kidney weight was
reported in males dosed at 750 ppm (112.5 mgAg/day) and in females
dosed at 275 and 750 ppm (33.8 and 112.5 mg/kg/day). Histopatho-
logically, dose-related changes in the forestomach of mice were
characterized by hyperplasia and hyperkeratosis of squamous epithelial
cells. These changes were observed in the 50-, 275- and 750-ppm dose
groups. No other treatment-related histopathological changes were
reported. A NOAEL of 15 ppm (2.3 mg/kg/day) is identified in this
study.
Paynter and Murphy (1967) administered DAC 2787 (Chlorothalonil) to
beagle dogs (4/sex/dose) for 16 weeks at dietary concentrations of 0,
250, 500 or 750 ppm (approximately to 0, 6.3. 12.5 or 18.8 mg/kg/day;
Lehman, 1959). No effects attributable to Chlorothalonil were noted
in terms of appearance, behavior, appetite, elimination, body weight
changes, gross pathology or organ weights. Hematological, biochemical
and urinalysis values were generally within accepted limits in treated
and control animals, except for slightly elevated protein-bound
iodine values in treated dogs (especially high-dose females). No
compound-related histopathology was noted. Based on this, a minimum
NOAEL of 750 ppm (18.8 mg/kg/day) is identified.
Hastings et al. (1975) admin.stered Chlorothalonil to Wistar albino
rats (15/sex/dose for treatment groups, 30/sex for controls) for four
months at dietary concentrations of 0, 1, 2, 4, 15, 30, 60 or 120 ppm
(approximately 0, 0.05, 0.1, 0.2, 0.8, 1.5, 3 or 6 mg/kg/day; Lehman,
1959). No significant differences between treated and control groups
were seen in body weight, food consumption, mortality or gross patho-
logical changes. Histopathological examination of the kidneys
revealed no demonstrable effects at any dose level. A minimum NOAEL
of 120 ppm (6 mg/kg/day) is identified.
Blackmore et al. (1968) administered DAC 2787 (Chlorothalonil) to
Charles River rats (35/sex/dose) for 22 weeks at dietary concentrations
of 0, 250, 500, 750 or 1,500 ppm (approximately 0, 12.5, 25, 37.5 or
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Chlorothalonil August, 1987
-9-
75 mgAg/day; Lehman, 1959). At all dose levels, male rats gained
less weight from weeks 11 to 22. Females gained less weight from
weeks 9 to 22 at 750 and 1,500 ppm (37.5 or 75 mg/kg/day). Food
consumption values were similar for all groups. No differences
between control and test animals were reported for various hematological
parameters, urinalysis and plasma and urine electrolytes. Results of
gross necropsy revealed that livers and kidneys of males treated at
750 or 1,500 ppm (37.5 or 75 mg/kg/day) were larger than controls.
Microscopic examinations demonstrated dose-related compound-induced
alterations in the kidneys of both sexes at all doses. These changes
were characterized by irregular swelling of the tubular epithelium,
epithelial degeneration and tubular dilatation. There was a signifi-
cant increase in renal tubular diameter in males at all dose levels.
Accordingly, a LOAEL of 250 ppm (12.5 mg/kg/day) is identified.
0 Blackmore and Kundzin (1969) administered technical-grade DAC 2787
(chlorothalonil) to rats (strain not specified) (35/sex/dose) for 1
year at dietary concentrations of 0, 4, 10, 20, 30, 40 or 60 ppm.
The authors indicated that these dietary levels correspond to 0, 0.2,
0.5, 1.0, 1.5, 2.0 or 3.0 mg/kg/day. No compound-related effects on
physical appearance, behavior, growth, food consumption, survival,
clinical laboratory values, organ weights or gross pathology were
noted. Microscopically, there were kidney alterations in both sexes
at 40 and 60 ppm (2.0 and 3.0 mg/kg/day). These alterations occurred
primarily in the deeper cortical tubules and consisted of increased
vacuolation of epithelial cells accompanied by swelling or hypertrophy
of the affected cells, often with the deposition of an eosinophilic
droplet material in the cytoplasm of the vacuole. Statistical
significance was not addressed. A NOAEL of 30 ppm (1.5 mg/kg/day)
is identified.
0 Holsing and Voelker (1970) administered technical-grade chlorothalonil
to beagle dogs (eight/sex/dose) for 104 weeks at dietary concentrations
of 0, 60 or 120 ppm (approximately 0, 1.5 or 3 mg/kg/day; Lehman, 1959).
After 2 years of administration, compound-related histopathological
changes were observed in the kidneys of males fed 120 ppm (3 mg/kg/day).
Males fed 60 ppm (1.5 mg/kg/day) and females fed both dose levels
were comparable to controls. The observed changes included increased
vacuolation of the epithelium in both the convoluted and collecting
tubules and increased pigment in the convoluted tubular epithelium.
Clinical findings, terminal body weight, organ-to-body weight ratios
and gross pathology revealed no conclusive compound-related trends.
A NOAEL of 60 ppm (1.5 mg/kg/day) is identified.
0 Tierney et al. (1983) administered technical grade chlorothalonil
to Charles River CD-1 mice (60/sex/dose) for 2 years at dietary
concentrations of 0, 750, 1,500 or 3,000 ppm. The authors indicated
tnat these dietary levels were approximately 0, 119.4, 251.1 or
517.4 mg/kg/day for males and 0, 133.6, 278.5 or 585.0 mg/kg/day
for females. No treatment-related effects on body weight, food
consumption, physical condition or hematological parameters were noted.
A slightly increased mortality rate was noted in males receiving
3,000 ppm (517.4 mg/kg/day). Also, kidney-to-body weight ratios and
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Chlorothalonil August, 1987
-10-
kidney-to-brain weight ratios were increased significantly in all
test groups. Gross necropsy revealed a number of renal effects
including kidney enlargement, discoloration, surface irregularities,
pelvic dilation, cysts, nodules and masses. Effects on the stomach
included an increased incidence in masses or nodules. In the stomach
and esophagus, nonneoplastic histopathological effects were noted at
all dose levels, and included hyperplasia and hyperkeratosis of the
squamous mucosa. This was considered to be indicative of mucosal
irritation. Other changes in the stomach included mucosal and
submucosal inflammation, focal necrosis or ulcers of mucosa and
hyperplasia of glandular mucosa. Reported histopathological effects
on the kidney included an increase in the incidence and severity of
glomerulonephritis, cortical tubular degeneration and cortical cysts.
These changes were not dose-related, but they did occur at higher
incidences in treated animals. Based on the information presented in
this study, a LOAEL of 750 ppm (119.4 mg/kg/day-males; 133.6 mg/kg/day-
females) is identified.
Reproductive Effects
0 In a three-generation reproduction study, Paynter and Kundzin (1967)
administered a mixture containing 93.6% Chlorothalonil to Charles River
rats (10 males and 20 females per dose) at dietary concentrations of
0 or 5,000 ppm (approximately 0 or 250 mg/kg/day; Lehman, 1959). At
the dose tested, the test material produced significant growth
suppression in the nursing litters of each generation. Reproductive
performance was not affected and pups showed no malformations attrib-
utable to the test substance. Body weight gains for exposed male and
female rats of each generation were lower than controls.
Developmental Effects
0 Rodwell et al. (1983) administered technical grade Chlorothalonil by
gavage at doses of 0, 25, 100 or 400 mg/kg/day to Sprague-Dawley rats
(25/dose level) on days 6 to 15 of gestation. No compound-related
external, internal or skeletal malformations were observed in fetuses.
At 400 mg/kg/day, maternal toxicity was noted (as evidenced by changes
in appearance, three deaths, decreased body weight gain and food con-
sumption). A slight increase in the number of early embryonic deaths
was associated with this maternal toxicity. This study identifies
a NOAEL of 400 mg/kg/day for teratogenic effects and a NOAEL of
100 mg/kg/day for maternal toxicity.
0 Wazeter et al. (1976) administered DTX-75-0016 (Chlorothalonil;
purity not specified) by oral intubation at doses of 0, 1, 2.5 or
5 mgAg to Dutch Belted rabbits (10/dose) on days 6 to 18 of gestation.
No compound-related changes in general behavior or appearance were
reported at the 1 or 2.5 mg/kg dose level. Occasional hypothermia
and hyperactivity were noted at a dose of 5 mg/kg. Maternal body
weight was not affected at any dose. No signs of toxicity were
reported regarding the number of implantation sites, numbers of live
or dead fetuses, live fetal weight, sex ratio or structural development.
However, an increase in the number of females with dead or resorbed
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Chlorothalonil August, 1987
-1 1-
fetuses (nine) and in the number of females aborting (four, two died
during the study) were seen at 5 mg/kg. Based on this information,
this study identifies a NOAEL of 2.5 mg/kg/day for maternal/fetal
toxicity and a NOAEL of 5 mg/kg/day for teratogenic effects.
0 Shirasu and Teramoto (1975) administered chlorothalonil by gavage to
Japanese white rabbits (eight controls, nine per dose) at doses of
0, 5 or 50 mg/kg/day on days 6 to 18 of gestation. At 50 mg/kg/day,
four of the nine does aborted. No compound-related growth retardation
or malformations were noted in offspring in any test group. This
study identifies a NOAEL of 50 mg/kg/day for teratogenic effects and
a NOAEL of 5 mg/kg/day for maternal toxicity.
Mutagenicity
0 Quinto et al. (1981) reported that chlorothalonil (concentrations not
specified) was not mutagenic, with or without metabolic activation,
in five tester strains of Salmonella typhimurium.
0 Wei (1982) reported that chlorothalonil, at concentrations up to
764 ug/plate, was not mutagenic in JJ. typhimurium strains TA 1535,
1537, 1538, 100 or 98, with or without liver or kidney activation
systems.
0 Kouri et al. (1977c) reported that DTX-77-0035 (chlorothalonil) at
concentrations up to 6.6 ug/plate did not induce point mutations in
S. typhimurium strains TA 1535, 100, 1537, 1538 or 98, with or without
S-9 activation.
0 Shirasu et al. (1975) reported the results of a reverse mutation test
using S. typhimurium strains TA 1535, 1537, 1538, 98 and 100 and
Escherichia coli WP2 hcr+ and WP2 her-. Chlorothalonil failed to pro-
duce an effect without activation at concentrations up to 500 pg/plate;
negative results also were obtained with activation at chlorothalonil
concentrations up to 100 pg/plate.
0 Kouri et al. (1977b) reported the results of a DNA repair assay using
_S. typhimurium strains TA 1978 and 1538. Chlorothalonil, dissolved
Tn dime thyIsulfoxide at 1 mg/mL and tested at 2, 10 and 20 uL of the
stock solution per plate, was found to be active in both strains with
or without metabolic activation.
0 DeBertoldi et al. (1978) reported that chlorothalonil (2,500 ppm) did
not induce mitotic gene conversions in Saccharomyces cerevisiae in the
presence or absence of metabolic activation systems. In tests on
Aspergillus nidulans using both-resting and germinating conidia,
chlorothalonil (up to 200 ppm) did not induce mitotic gene conversions.
0 Shirasu et al. (1975) reported that, at concentrations up to 200
ug/disk, chlorothalonil was negative in a rec-assay using Bacillus
subtilis strains H17 and M45.
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Chlorothalonil August, 1987
-12-
0 Kouri et al. (1977a) exposed Chinese hamster cells (V-79) and mouse
fibroblast cells (BALB/3T3) in vitro to Chlorothalonil at concentra-
tions of 0*3 ug/mL (for V-79 cells) or 0.03 ug/mL (for mouse fibroblast
cells). The V-79 cells were tested without metabolic activation; the
BALB/3T3 cells were tested with and without metabolic activation.
Chlorothalonil was not mutagenic in either cell type.
• Nizens et al. (1983a) reported the results of a micronucleus test in
Histar rats, Swiss CFLP mice and Chinese hamsters. Rats were dosed at
0, 8, 40, 200, 1,000 or 5,000 ing/kg; mice and hamsters received 0, 4,
20, 100, 500 or 2,500 rag/kg. All animals were dosed by gavage and all
received two doses, 24 hours apart. Chlorothalonil did not induce
bone marrow erythrocyte micronuclei in any of the species tested.
0 Legator (1974) reported the results of an in vivo cytogenetic test on
Chlorothalonil in mice (strain not specified) using the micronuclei
procedure. The test compound was administered by gavage for 5 days
at a concentration of 6.5 mg/kg/day. At this concentration,
Chlorothalonil did not increase the number of cells with micronuclei.
0 Legator (1974) presented the results of a host-mediated assay using
male Swiss albino mice and S, typhimurium strains G-46, TA1530, C-207,
TA1531, C-3076, TA1700, D-3056 and TA1724. Mice (10/dose) received
Chlorothalonil by gavage for 5 days at 6.5 mg/kg/day. The compound
did not produce any measurable mutagenic response when initially
evaluated in vitro against the eight tester strains of £. typhimurium.
When the tester strains were inoculated into treated mice, no increase
in mutation frequency was observed.
0 Legator (1974) presented the results of a dominant lethal assay in
which male mice (strain not specified) were dosed with Chlorothalonil
for five days at 6.5 mg/kg/day. These mice were mated with untreated
females, and the number of early fetal deaths and preimplantation
losses were measured. There was no significant difference in the
fertility rates between test and control animals during weeks 1 to 7.
At week 8, there was a significant decrease in fertility in the test
group.
0 Hizens et al. (1983b) presented the results of a chromosomal aberration
test in Chinese hamsters. The test animals received two doses of
Chlorothalonil, 24 hours apart, by gavage at concentrations of 0, 8,
40, 200, 1,000 or 5,000 mg/Kg. At 5,000 mgAg. a statistically
significant increase in bone marrow chromosomal abnormalities was
observed. However, the authors concluded that this effect could not
be attributed to Chlorothalonil because the animals exhibited toxic
responses to dosing.
Carcinogenicity
0 NCI (1980) reported the results of a study in which technical-grade
Chlorothalonil was administered to Osborne-Mendel rats (50/sex/dose)
for 80 weeks at Time-Weighted Average (TWA) dietary doses for both
males and females of 5,063 or 10,126 ppm, respectively. These dietary
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Chlorothalonil August, 1987
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doses have been calculated to correspond to approximately 253 and
506 mg/kg/day (Lehman, 1959). Matched controls consisted of groups
of 10 untreated rats of each sex; pooled controls consisted of the
matched controls combined with 55 untreated male or female rats from
other bioassays. An observation period of 30 to 31 weeks followed
dosing. Clinical signs that appeared with increased frequency in
dosed rats included hematuria and, from week 72 on, bright yellow
urine. Adenomas and carcinomas of renal tubular epithelium occurred
with a significant (p = 0.03, males; p = 0.007, females) dose-related
trend. The frequency of renal tumors was statistically greater in
the high-dose males (p = 0.035) and high-dose females (p » 0.016)
than in corresponding controls (males: pooled controls, 0/62; low
dose, 3/46; high dose, 4/49; females: pooled controls, 0/62; low
dose, 1/48; high dose, 5/50). The observed adenomas and carcinomas
were considered to be histogenically related. Results of this study
were interpreted as sufficient evidence of carcinogenicity in
Osborne-Mendel rats.
0 NCI (1980) also reported a study in which technical-grade chlorothalonil
was administered to B6C3F1 mice (50/sex/dose) for 80 weeks at TWA
dietary doses of 2,688 or 5,375 ppm for males and 3,000 or 6,000 ppm
for females. These dietary doses have been calculated to correspond
to approximately 403.2 or 806.3 rag/kg for males and 450 or 900 mg/kg
for females (Lehman, 1959). Matched controls consisted of 10 untreated
mice of each sex; pooled controls consisted of the matched controls
combined with 50 untreated male or female mice from other bioassays.
An observation period of 11 to 12 weeks followed dosing. Since the
dosed female mice did not show depression in mean body weights or
decreased survival compared with the controls, they may have been
able to tolerate a higher dose. No tumors were found to occur at a
greater incidence among dosed animals than among controls. It was
concluded that, under the conditions of this bioassay, chlorothalonil
was not carcinogenic in B6C3F<| mice.
0 Tierney et al. (1983) administered technical-grade chlorothalonil
(97.7% pure) to Charles River CD-I mice (60/sex/control and dose groups)
for 2 years at dietary concentrations of 0, 750, 1,500 or 3,000 ppm.
The authors indicated that these dietary levels were equivalent to
0, 119, 251 or 517 mg/kg/day for males and 0, 133, 278 or 585 mg/kg/day
for females. Increased incidences of squamous cell tumors of the
forestomacb were noted in all treatment groups. These tumors consisted
principally of carcinomas, although papillomas were also seen. This
increased incidence was statistically significant in females dosed
at 1,500 ppm (279 mg/kg/day). No clear dose-related trend in the
incidence of these tumors was observed. A slight increase in the
incidence of tumors of the glandular epithelium of the fundic stomach
was observed in dosed animals; this increase was neither statistically
significant nor dose-related. When the numbers of animals with
epithelial tumors of the fundic or forestomach were combined, the
incidence of these tumors showed a statistically significant increase
in the 1,500- and 3,000-ppm female dose groups (279 and 585 mg/kg/day).
No treatment-related renal neoplasms were seen in any female dose
group. Increased incidences of adenomas and carcinomas in renal
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Chlorothalonil August, 1987
-14-
cortical tubules were noted in all treated groups of male mice.
These changes did not show a dose-response relationship; the increased
incidence was statistically significant only in the 750 ppm (251
nig/kg/day) group. The authors concluded that the administration of
Chlorothalonil caused an increase in the incidence of primary gastric
tumors and an increase in the incidence of renal tubular neoplasms.
0 Wilson et al. (1985) gave Chlorothalonil (98.1% pure with less than 0.03%
hexachlorobenzene) to Fischer 344 rats (60/sex/dose) in their diet at
dose levels of 0, 40, 80 or 175 mg/kg/day. Males were treated for
116 weeks, while females received the chemical for 129 weeks. Survival
among the various groups was comparable. In both sexes, at the high-
dose level, there were significant decreases in body weights. In
addition, there were also significant increases in blood urea nitrogen
and creatinine, while there were decreases in serum glucose and
albumin levels. In both sexes, there were dose-dependent increases
in kidney carcinomas and adenomas at doses above 40 mg/kg/day. In
the high-dose females, there was also a significant increase in
stomach papillomas. The data show that, in the Fischer 344 rat,
Chlorothalonil is a carcinogen.
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, 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).
One-day Health Advisory
No information was found in the available literature that was suitable
for determination of a One-day HA for Chlorothalonil. Accordingly, it is
recommended that the Ten-day HA value (250 ug/L, calculated below) for a
10-kg child be used at this time as a conservative estimate of the One-day
HA value.
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Chlorothalonil August, 1987
-15-
Ten-day Health Advisory
The rabbit teratology study by Wazeter et al. (1976) has been chosen to
serve as the basis for the calculation of the Ten-day HA. Animals received
0, 1, 2.5 or 5 mg/kg Chlorothalonil by gavage on days 6 through 18 of gestation.
No adverse effects were observed at either of the two lower treatment doses.
At 5 mg/kg, an increase in the number of females with dead or resorbed fetuses
and in the number of females aborting was observed. The NOAEL for maternal/
fetal toxicity is 2.5 mg/kg/day.
The Ten-day HA for the 10-kg child is calculated as follows:
where:
Ten-day HA = (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (250 ug/L)
(TOO) (1 L/day)
2.5 mg/kg/day = NOAEL, based on absence of maternal or fetal toxicity
in rabbits exposed to Chlorothalonil via gavage on
days 6 to 18 of gestation.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The studies by Colley et al. (1983), Blackmore and Kundzin (1969) and
Blackmore and Shott (1968) have been selected to serve as the basis for the
Longer-term HA for Chlorothalonil. In the study by Colley et al., technical-
grade Chlorothalonil was administered in the diet to Charles River rats for
13 weeks at concentrations of 0, 1.5, 3.0, 10 or 40 mg/kg/day. Histopatho-
logical examinations revealed that at doses of 3.0 mg/kg/day or greater, male
rats displayed an increased number of intracytoplasmic inclusion bodies in
the proximal convoluted renal tubules. Blackmore and Shott (1968), gave
technical-grade Chlorothalonil in the diet to Charles River rats for 90 days
at doses of 0, 0.2, 0.5, 1.0, 1.5, 2.0 or 3.0 mg/kg/day. At the two highest
dose levels, the kidneys exhibited occasional vacuolation and swelling of
the epithelial cells lining the deeper proximal convoluted tubules. In the
Blackmore and Kundzin (1969) study, technical-grade Chlorothalonil was admin-
istered in the diet to rats for 1 year at doses of 0, 0.2, 0.5, 1.0, 1.5, 2.0
or 3.0 mg/kg/day. At the 2 higher doses, there were alterations in the deeper
convoluted renal tubules in both sexes. Each of the studies identified a
NOAEL of 1.5 mg/kg/day.
The Longer-term HA for a 10 kg child is calculated as follows:
Longer-term HA = (1.5 mg/kg/day) (10 kg) = o.15 mg/L (150 ug/L)
(100) (1 L/day)
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Chlorothalonil August, 1987
-16-
where:
1.5 mg/kg/day = NOAEL, based on absence of kidney effects in rats
exposed to Chlorothalonil in the diet for 1 3 weeks .
10 kg = assumed body weight of a child.
1 00 B uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day a assumed daily water consumption of a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = <1 *5 »9/*9/day> <70 *g> =0.525 mg/L (525 ug/L)
(100) (2 L/day)
where:
1.5 mg/kg/day = NOAEL, based on absence of kidney effects in rats
exposed to Chlorothalonil in the diet for 13 weeks.
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, ass'iming 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, 1986a), then caution should be exercised
in assessing the risks associated with lifetime exposure to this chemical.
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Chlorothalonil August, 1987
-17-
The study by Holsing and Voelker (1970) has been selected to serve as
the basis for the Lifetime HA for chlorothalonil. In this study, technical-
grade chlorothalonil was administered to beagle dogs (eight/sex/dose) for 104
weeks at dietary concentrations of 0, 60 or 120 ppm (0, 1.5 or 3.0 mg/kg/day).
The results following 2 years of administration revealed compound-related
histopathological changes in the kidneys of males fed 120 ppm (3 mg/kg/day).
Males fed 60 ppm (1.5 mg/kg/day) and females fed both dose levels were
comparable to controls. The observed changes included increased vacuolation
of the epithelium in both the convoluted and collecting tubules and increased
pigment in the convoluted tubule epithelium. From these results, a NOAEL of
1.5 mg/kg was identified.
Using this NOAEL, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = H»5 mg/kg/day) = Q.015 mg/kg/day
(100)
where:
1.5 mg/kg/day = NOAEL, based on absence of histopathological changes
in dogs fed chlorothalonil 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)
DWEL = (0-015 mg/kg/day) (70 kg) = 0.525 /L (525 ug/L)
2 L/day
where:
0.015 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determine tion of the Lifetime Health Advisory
The estimated excess cancer risk associated with lifetime exposure to
drinking water containing chlorothalonil at 525 ug/L (the DWEL) is 3.5 x 10~4.
This estimate represents the upper 95% confidence limit from extrapolations
prepared by OFF and ODW using the linearized, multistage model. The actual
risk is unlikely to exceed this value, but there is considerable uncertainty
as to the accuracy of risks calculated by this methodology.
Evaluation of Carcinogenic Potential
0 In an NCI bioassay (1980), technical grade chlorothalonil was
administered in the diet at 253 or 506 mg/kg/day to Osborne-Mendel
-------
Chlorothalonil August, 1987
-18-
rats for 80 weeks. A statistically significant increase in the
frequency of renal tumors was observed in high-dose males and females.
0 NCI (1980) reported that chorothalonil was not carcinogenic in B6C3Fj
mice when administered in the diet, at 403 or 806 rag/kg and 450 or
900 mg/kg for males and females, respectively, for 80 weeks. However,
Tierney et al. (1983) concluded that Chlorothalonil was carcinogenic
in Charles River CO-1 which received the compound (0, 119, 251 or
517 mg/kg/day for males and 0, 134, 279 or 585 mg/kg/day for females)
in the diet for 2 years. Increased incidences of squamous cell
papilloma and carcinoma of the forestomach were noted in all treatment
groupse This increase was statistically significant only in the mid-
dose females. Increased incidences of adenoma and carcinoma of the
renal cortical tubules were observed in all treatment groups. Again,
no dose-response was noted, since these increases were statistically
significant only in the mid-dose males.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of Chlorothalonil.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), Chlorothalonil is classified in
Group B2: probable human carcinogen. This category is for chemicals
for which there is inadequate evidence from human studies and sufficient
evidence from animal studies.
0 From the Wilson et al. (1985) data, OPP calculated a q^ of 2.4 x
10-2 (mg/kg/day)-1. The 95% upper limit lifetime dose in drinking water
associated with a 10-6 excess risk level is 1.5 ug/L. Corresponding
levels for 10-5 and 10~4 are 15 and 150 ug/L, respectively. While
recognized as statistically alternative approaches, the range of
risks described by using any of these modelling 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
Agency has recommended use of the linearized multistage approach.
However, for completeness, the 10~6 risk numbers for other models
will be given. These values, at the 10~*> level, are: multihit -
9 ug/L; one hit - 2 ug/L; probit - 51 ug/L; logit - 0.8 ug/L; and
Weibel - 0.6 ug/L.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 WHO Temporary Acceptable Daily Intake = 0.005 mg/kg/day (Vettorazzi
and Van den Hurk, 1985).
0 EPA/OPP has calculated a PADI of 0.015 mg/kg/day based on the NOAEL
of 1.5 mg/kg/day identified in the 2-year dog study (Holsing and and
Voelker, 1970) and an uncertainty factor of 100 (U.S. EPA, 1984a).
0 U.S. EPA established tolerances in or on raw agricultural commodities
residue levels of 0.1 to 5 ppm (40 CFR 180.275, 1985).
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Chlorothalonil August, 1987
-19-
VII. ANALYTICAL METHODS
0 Analysis of Chlorothalonil is by a gas chromatographic (GC) method
applicable to the determination of certain chlorinated pesticides in
water samples (U.S. EPA, 1986b). In this method, approximately
1 liter of sample is extracted with methylene chloride. The extract
is concentrated and the compounds are separated using capillary
column GC. Measurement is made using an electron capture detector.
The method detection limit has not been determined for Chlorothalonil,
but it is estimated that the detection limits for analytes included
in this method are in the range of 0.01 to 0.1 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Reverse osmosis (RO) is a promising treatment method for pesticide-
contaminated water. As a general rule, organic compounds with
molecular weights greater than 100 are candidates for removal by RO.
Larson et al. (1982) reported 99% removal efficiency of chlorinated
pesticides by a thin-film composite polyamide membrane operating at
a maximum pressure of 1,000 psi and a maximum temperature of 113°F.
More operational data are required, however, to specifically determine
the effectiveness and feasibility of applying RO for the removal of
Chlorothalonil from water. Also, membrane adsorption must be consid-
ered when evaluating RO performance in the treatment of Chlorothalonil-
contaminated drinking water supplies.
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Chlorothalonil August, 1987
-20-
IX. REFERENCES
Auletta, C.S., and L.F. Rubin.* 1981. Eye irritation studies in monkeys and
rabbits with Bravo 500: Report DS-2787. Unpublished study. MRZD 00077176.
Blackmore, R.H., and L.D. Shott.* 1968. Final report: three-month feeding
study—rats. Project No. 200-205. Unpublished study. MRID 00087316.
Blackmore, R.H., L.D. Shott, M. Kundzin et al.* 1968. Final report:
four-month feeding study—rats (22 weeks). Project No. 200-198.
Unpublished study. MRID 00057701.
Blackmore, R.H., and M. Kundzin.* 1969. Final report: 12-month feeding
study—rats. Project No. 200-205. Unpublished study. MRID 00087358.
Capps, T.M., J.P. Marciniszyn, A.F. Markes, and J.A. Ignatoski.* 1982.
Document No. 555-4EF-81-0261-001, Section J, Vol. VI. Submitted by
Diamond Shamrock Corporation.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.275. July 1, 1985.
CHEMIAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD.
Colley, J., L. Syred, R. Heywood et al.* 1983. A 13-week subchronic toxicity
study of T-117-11 in rats (followed by a 13-week withdrawal period).
Unpublished study. MRID 00127852.
DeBertoldi, M., R. Barale, and M. Giovannetti. 1978. Mutagenicity of
pesticides evaluated by means of gene conversion in £. cerevisiae and £.
nidulans. Environ. Mut. 2:359-370.
Doyle, R.L., and J.R. Elsea.* 1963. Acute oral, dermal and eye toxicity and
irritation studies on DAC-2787: N-107. Unpublished study. MRID 00038909.
Hastings, T.F., M. Dickson, W.M. Busey et al.* 1975. Four-month dietary
toxicity study—rats Chlorothalonil. Project No. 24-201. Unpublished study.
MRID 00040463.
Holsing, G., and R. Voelker.* 1970. 104-week dietary administration—dogs:
Daconil 2787 (Technical). Project No. 200-206. Unpublished study.
MRID 00114304.
Johnsson, M., M. Buhagen, H.L. Leira and S. Solvang. 1983. Fungicide-induced
contact dermatitis. Contact Dermat. 9:285-288.
Johnston, E.F.* 1981. Soil disappearance studies with Benlate fungicide and
Bravo 500 fungicide, alone and in combination: Document No. AMR-06-81.
Unpublished study submitted by E.I. du Pont de Nemours and Co.,
Wilmington, DE.
Kouri, R.E., R. Joglekar and D.P.A. Fabrizio.* 1977a. Activity of
DTX-77-0034 in an ir± vitro mammalian cell point mutation assay.
Unpublished study. MRID 00030289.
-------
Chlorothalonil August, 1987
-21-
Kouri, R.E., A.S. Parmar, J.M. Kuzava et al.* 1977b. Activity of DTX-77-0033
in a test for differential inhibition of repair deficient and repair
competent strains of Salmonella typhimurium. Unpublished study.
MRID 00030288.
Kouri, R.E., A.S. Parmar, J.M. Kuzava et al.* 1977c. Activity of DTX-77-0035
in the SaImone1la/microsoma1 assay for bacterial mutagenicity. Unpublished
study. MRID 00030290.
Larson, R.E., P.S. Cartwright, P.K. Eriksson and R.J. Petersen. 1982. Appli-
cations of the FT-30 reverse osmosis membrane in the metal finishing
operations. Paper presented at Tokohama, Japan.
Legator, M.S.* 1974. Report on mutagenic testing with DAC 2787. Unpublished
study. MRID 00040464.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Published by the Association of Food and Drug Officials of
the United States.
Marciniszyn, J., J. Killeen and J. Ignatoski.* 1981. Dose-response
determination of the excretion of radioactivity in rat bile following
intraduodenal administration of Hc-chlorothalonil (14C-DS-2787).
Unpublished study. MRID 00137132.
Marciniszyn, J., J. Killeen and J. Ignatoski.* 1983a. Recirculation of
radioactivity in rat bile following intraduodenal administration of bile
containing 14C-chlorothalonil label. Unpublished study. MRID 00137130.
Marciniszyn, J., J. Killeen and J. Ignatoski.* 1983b. Identification of major
Chlorothalonil metabolites in rat urine. Unpublished study. MRID 00137129,
Meister, R., ed. 1986. Farm Chemicals Handbook. Hilloughby, OH: Meister
Publishing Company.
Mizens, M., J. Killeen and J. Ignatoski.* 1983a. The micronucleus test in
the rat, mouse and hamster using Chlorothalonil. Unpublished study.
MRID 00127853.
Mizens, M., J. Killeen and J. Ignatoski.* 1983b. The chromosomal aberration
test in the rat, mouse and hamster using Chlorothalonil. Unpublished
study. MRID 00r127854.
NCI. 1980. National Cancer Institute. Bioassay of Chlorothalonil for possible
carcinogenicity (NTP #TR-041). U.S. Public Health Service. U.S. Depart-
ment of Health, Education and Welfare.
Paynter, O.E., and M. Kundzin.* 1967. Final report: three-generation
reproduction study—rats. Project No. 200-155. Unpublished study.
MRID 00091289.
Paynter, O.E., and J.C. Murphy.* 1967. Final report: 16-week dietary
feeding—dogs. Project No. 200-200. Unpublished study. MRID 00057698.
-------
Chlorothalonil August, 1987
-22-
Pollock, G., j. Marciniszyn, J. Kllleen et al.* 1983. Levels of radioactivity
in blood following oral administration of 1*C-Chlorothalonil (14c-DS-2787)
to male rats. Unpublished study. MRID 00137127.
Powers, M.B.* 1965. Acute oral administration—rats. Project No. 200-167.
Unpublished study. MRID 00038910.
Cuinto, X., G. Nartire, G. Vricella, F. Riccardi, A. Perfumo, R. Giulivo and
F. DeLorenzo. 1981. Screening of 24 pesticides by Salmonella/microsome
assay: mutagenicity of benazolin, metoxuron and paraoxon. Mutat. Res.
85:265.
Ribovich, M., Pollock G., J. Marciniszyn et al.* 1983. Balance study of the
distribution of radioactivity following oral administration of 1*C-
chlorothalonil (14C-DS-2787) to male mice. Unpublished study. MRID
00137125.
Rittenhouse, J.R., and J.K. Narcisse.* 1974. The acute oral toxicity of
Daconil 2.88 F to adult male and female rats. Unpublished study.
MRID 00070358.
Rodwell, D., M. Mizens, N. Wilson et al.* 1983. A teratology study in rats
with technical Chlorothalonil. Unpublished study. MRID 00130733.
Ryer, F.Ho* 1966. Radiotracer metabolism study. Unpublished study.
MRID 00038918.
Shirasu, Y., M. Moriya and K. Watanabe.* 1975. Mutagenicity testing on
Daconil in microbial systems. Unpublished study. MRID 00052947.
Shirasu, Y., and S. Teramoto. 1975.* Teratogenicity study of Daconil in
rabbits. Unpublished study. MRID 00127855.
Shults, S., J. Laveglia, J. Killeen et al.* 1983. A 90-day feeding study in
mice with technical Chlorothalonil. Unpublishsd study. MRID 00138148.
Skinner, W.A., and D.E. Stallard.* 1967. Daconil 2787 animal metabolism
studies. Unpublished study. MRID 00038917.
Stallard, D.E., and A.L. Wolfe.* 1967. The fate of 2,4,5,6-tetrachloro-
isophthalonitrile (Daconil 2787) in soil. Unpublished study submitted
by Diamond Alkali Company, Cleveland, OH.
Stallard, D.E., A.L. Wolfe and W.C. Duane.* 1972. Evaluation of the leaching
of Chlorothalonil under field conditions and its potential to contaminate
underground water supplies. Unpublished study submitted by Diamond
Shamrock Company, Cleveland, OH.
STORETo 1987.
Szalkowski, M.S.* 197? Photodegradation and mobility of Daconil and
its major metabolite on soil thin films. Unpublished study submitted
by Diamond Shamrock Agricultural Chemicals, Cleveland, OH.
-------
Chlorothalonil August, 1987
-23-
Szalkowski, M.B.* 1976a. Effect of microorganisms upon the soil
metabolism of Daconil and 4-hydroxy-2,5,6-trichloroisophthalonitrile.
Unpublished study submitted by Diamond Shamrock Agricultural Chemi-
cals, Cleveland, OH.
Szalkowski, M.B.* 1976b. Hydrolysis of Daconil and its metabolite,
4-hydroxy-2,5,6-trichloroisophthalontrile, in the absence of light
at pH levels of 5,7, and 9. Updated method. Unpublished study sub-
mitted by Diamond Shamrock Agricultural Chemicals, Cleveland, OH.
Szalkowski, M.B., J.J. Mannion, D.E. Stallard et al.* 1979. Quant-
ification and characterization of the biotransformation products of
2,4,5,6-tetrachloroisophthalonitrile (chlorothalonil, DS-2787) in soil.
Unpublished study submitted by Diamond Shamrock Agricultural Chemi-
cals, Cleveland, OH.
Tierney, W., N. Wilson, J. Killeen et al.* 1983. A chronic dietary study in
mice with technical chlorothalonil. Unpublished study. MRID 00127858.
U.S. EPA. 1984a. U.S. Environmental Protection Agency. Proposed guidelines
for carcinogenic risk assessment; Request for comments. Fed. Reg.
49(227):46294-46301. November 23.
U.S. EPA. 1984b. U.S. Environmental Protection Agency. Chlorothalonil
(case GS0097) Pesticide Registration Standard. Office of Pesticide
Programs, Washington, DC.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
assessment of carcinogen risk. Fed. Reg. 51 (185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #2
- Determination of chlorinated pesticides in ground water by GC/EDC,
January 1986 draft. Available from U.S. EPA's Environmental Monitoring
and Support Laboratory, Cincinnati, OH.
Vettorazzi, G., and G.W. Van den Hurk. 1985. The pesticide reference index,
JMPR 1961-1984. World Health Organization, Geneva.
Wazeter, F.X., E.I. Goldenthal and S.B. Harris.* 1976. Teratology study in
rabbits. Unpublished study. MRID 00047944.
Wei, C. 1982. Lack of mutagenicity of the fungicide 2,4,5,6-tetrachloro-
isophthalonitrile in the Ames Salmonella/microsome test. Appl. Environ.
Microbiol. 43:252-4.
Wilson, N., J. Killeen, J. Ignatoski et al.* 1981. A 90-day toxicity study
of technical chlorothalonil in rats. Unpublished study. MRID 00127850.
Wilson, N. J. Killeen, J. Ignatoski.* 1985. A tumorigenicity study of
technical chlorothalonil in rats: Document No. 099-5TX-80-0234-008.
Unpublished study prepared by ADS Biotech Corp. 2269 p. MRID 00146945.
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Chlorothalonil August, 1987
-24-
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
Hie Merck index—An encyclopedia of chemicals and drugs. 10th ed.
Rahway, NJ: Merck and Company, Inc.
Wolfe, A.L., and D.E. Stallard.* 1968a. The fate of DAC-3701 (4-hydroxy-
2,5,6-trichloroisophthalonitrile) in soil. Unpublished study submitted
by Diamond Shamrock Chemical Company, Cleveland, OH.
Wolfe, A.Le, and O.E. Stallard.* 1968b. Analysis of tissues and organs for
storage of the Daconil metabolite 4-hydroxy-2,5,6-trichloroisophthalo-
nitrile. Unpublished study. MRID 00087254.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
DRAFT
CYANAZINE
August, 1987
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.
-------
Cyanazine
-2-
GENERAL INFORMATION AND PROPERTIES
CAS No. 21725-46-2
Structural Formula
H CSN
N-C(CH,)2
N-CH2CH,
H
2-[[4-Chloro-6-(ethylamino)-1,3,5-triazin-2-
Synonyms
Uses
August, 1987
-2-B. thylprop.ne,,i wile
Properties (U.S. EPA, 1984a; Meister, 1983; CHEMLAB, 1985)
Chemical Formula
Molecular Weight
C«H, ,C1NC
240 7
Melting Potnt
Log Octanol/water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
,67.5 to ,6,.c
2.24
Occurrence
*
9°° "9/L °£
«« -t.r fro. the
-------
Cyanazine August, 1987
-3-
0 Cyanazine was identified in drinking water in New Orleans, Louisiana,
in concentrations ranging from 0.01 to 0.35 ug/L.
0 Cyanazine was monitored in a newly-built reservoir on the Des Moines
River in Iowa during September 1977 through November 1978. Agri-
cultural runoff (from corn and soybeans) was a major source of
pollution in the river: levels of 71 to 457 ng/L were detected
during the active months of May through August; levels of 2 to 151
ng/L wre detected during September through December; and zero levels
were found from January through April (U.S. EPA, 1984a; MAS, 1977).
0 Cyanazine has been found in surface water in Ohio river basins
(Oatta, 1984).
8 Cyanazine has also been found in ground water in Iowa and Pennsylvania;
typical positives found were 0.1 to 1.0 ppb (Cohen et al., 1986).
0 Cyanazine has been found in 4,312 of 4,285 surface water samples
analyzed and in 21 of 1,564 ground water samples (STORET, 1987).
Samples were collected at 337 surface water locations and 1,066 ground
water locations, and cyanazine was found in 11 states. The 85th
percentile of all non-zero samples was 4.11 ug/L in surface water and
.20 ug/L in ground water sources. The maximum concentration found in
surface water was 900 ug/L and in ground water it was 3,500 ug/L.
Environmental Fate
0 14c-Cyanazine, at 5 to 10 ppm, degraded with a half-life of 2 to
4 weeks in an air-dried sandy clay loam soil, 7 to 10 weeks in a
sandy loam soil, 10 to 14 weeks in a clay soil, and 9 weeks in a
fresh sandy clay soil incubated in the dark at 22°C and field capacity
(Osgerby et al., 1968). Three-degradation products, the amide and
two acids, were identified in all four soils; a fourth degradate,
the anine, was found only in the air-dried sandy clay loam soil.
0 Freundlich K values were 0.72 for a sandy loan soil (2.0% organic
matter), 2.0 for a sandy clay soil (5.4% organic matter), 1.25 for
a sandy clay loam soil (6.8% organic matter) and 6.8 for a clay soil
(16% organic matter) treated with imaged 14C-cyanazine (Osgerby
et al., 1968). No linear correlation was found between organic
matter content and adsorption.
0 14c-Cyanazine readily moved tnrough columns of sandy clay loam (52%
of applied compound) and loamy sand (18% of applied) soil leached with
78 cm of water over a 13-day period; imaged !4C-cyanazine was inter-
mediately mobile on sandy clay loam and of low mobility on loamy sand
soil thin-layer chromatography (TLC) plates (Rf 0.36 and 0.20,
respectively; (McMinn and Standen, 1981). Aerobically and anaerobically
aged 14c-cyanazine residues, primarily the amide degradate (SO 20258),
were intermediately mobile to mobile on sandy clay loam soil TLC plates.
0 Aged 1^-cyanazine residues readily leached through columns containing
sand (47.8% of applied), loamy sand (69.7% of applied) and sandy
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Cyanazine August, 1987
-4-
loam (26;9% of applied) soils eluted with 20 cm of water (Eadsforth,
1984). The amide degradation product (SO 20258) was predominant in
the leachate from the sandy soil (45% of radioactivity in leachate);
the acid degradate (SD 20196) was predominant in leachate from the
loamy sand (84%) and sandy loam (47%) soils. Unaltered cyanazine
and SD 31222 were also identified in leachate from all three soils
(£6% of recovered).
III. PHARMACOKINETICS
Absorption
0 Studies by Shell Chemical Company (Shell Chemical Company, 1969) and
Hutson et al. (1970) indicated that cyanazine is rapidly absorbed
from the gastrointestinal tract when administered orally at low
dosage levels to three different animal species: rat, dog and cow.
Measurements of urinary, fecal and biliary excretion indicated that
80 to 88% of 2,4,6-14C-labeled cyanazine was eliminated within 4 days
from the rat and dog, and within 21 days from the cow. The initial
dosages were 1 to 4 mg/kg for the rat, 0.8 mg/animal for the dog and
5 ppm in the total ration of the cow. The dosages were administered
by gavage in the rat studies and in gelatin capsules in the dog study.
Distribution
0 In rats treated with a single oral dose of 4 mg/kg cyanazine,
samples of the carcass, skin and gut reflected 2.02, 0.62 and 2.73%
residual radioactivity, respectively, 4 days after exposure (Shell,
1969).
0 In cows, samples of brain, liver, kidney, muscle and fat reflected
concentrations of 0.55, 0.27, 0.24, 0.14 to 0.06 and less than 0.06
ppm cyanazine, respectively, after 21 days of continuous exposure
to feed that contained 5 ppm cyanazine; however, when a lower dosage
(0.2 ppm) was used in the feed, the detectable residues in each of
these tissues were less than 0.05 ppm (Shell, 1969).
Metabolism
0 Based on the analyses of metabolites in urine, the ma]or metabolic
pathways of cyanazine in the rat and cow involveo: (1) conversion of
the cyano group to an amide to form 2-chloro-4-ethylamino-6-(1-amido-
1-methylethylaniino)-s-thiazine; (2) N-deethylation to form 2-chloro-4-
amino-6-(l-cyanol-methyl-ethylamino)-s-triazine; (3) conversion of the
cyano group of deethylate cyanazine to form the amide of deethylated
cyanazine, 2-chloro-4-amino-6{ 1-aminn-1-methylethylamino)-s-triazine;
(4) dechlorination via glutathione, partial hydrolysis of glutathione
conjugate and N-acetylation to form mercapturic acid, N-acetyl-S-
[4-amino-6-(1-cyano-1-methylethylamino) L-cysteine; and (5)
dechlorination via hydrolysis (occurs only in the cow) to form
2-hydroxy-4-ethylamino-6-(1-carboxy-1-methylethylamino)-s-triazine
and 2-hydroxy-4-amino-6-( 1,carboxy-1 -methylanino)-s-triazine,
respectively (Shell, 1969).
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Cyanazine August, 1987
-5-
0 Studies by Shell Chemical Company (1969) and Hutson et al. (1970)
with ring-labeled and side-chain-labeled cyanazine (cyano-14c,
isopropyl-14C and ethylamino-14c) indicated that only the ethylaraino-14C
side chain underwent extensive degradation, since 47% of the initial
radioactivity was detected in the exhaled carbon dioxide. Thus,
N-deethylation was found to be a major route of degradation of
cyanazine.
0 Crayford and Hutson (1972) identified 5 metabolites in urine, an
additional 2 (total 7) in feces and 4 metabolites in bile.
0 Crayford et al. (1970) studied the metabolism of two major plant
metabolites, DW4385 and DW4394, in rats. These two compounds were
identified in the rat metabolism studies by Crayford and Hutson (1972)
as 2-hydroxy-4-ethylamino-6-(1-carboxy-1-methylamino)-s-triazine)
(DW4385) and as 2-hydroxy-4-amino-6-(1-carboxy-1 -methylethylamino)-
s-triazine) (DW4394). Approximately 91% of compound DW4385 and 84%
of compound DW4394 were recovered unchanged from urine and feces.
Excretion
Orally administered low doses of cyanazine were rapidly excreted
in the urine and feces of rats and dogs (Shell, 1969; Hutson et al.,
1970; Crayford and Hutson, 1972). See discussion of these studies
in the above sections.
In rats treated with 1 to 4 rag/kg cyanazine by gavage, a total of
88% of cyanazine was eliminated in 4 days. Elimination via urine was
almost equal to elimination via feces; about 5.37% of the administered
cyanazine remained in the body; and approximately 21% of the 1 mg/kg
dose appeared in the bile within the first 20 hours (Shell, 1969).
Hutson et al. (1970) reported that 33% of an oral dose of cyanazine
was excreted in the urine of rats within 24 hours.
A study in rats with 14c-labeled 4-ethyl-amino cyanazine indicated
that 47% of the radioactivity was eliminated in carbon dioxide
(Shell Chemical Company, 1969).
In dogs administered 0.8 mg of cyanazine in gelatin capsules, 51.67
and 36.29% of the dose were eliminated in the urine and feces,
respectively, over a 4-day period (Shell Chemical Company, 1969)•
In cows exposed to treated feed (5 ppm cyanazine) for 21 consecutive
days, the amount of daily excretion of radioactivity in urine and
feces was constant throughout the study period. The total cyanazine
equivalents in urine and feces were 53.7 and 26.8% of the dose,
respectively. The concentration in milk was reported as 0.022 ppm
(Shell Chemical Company, 1969).
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Cyanazine August. 1987
-6-
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of cyanazine in humans.
Animals
Short-term Exposure
0 The acute oral LD50 in rats ranged from 149 to 369 mg/kg (SRI, 1967b;
NIOSH, 1977; Young and Adamik, 1979b; Meister, 1983). In these
studies, the percentage of active ingredient (a.i.) in the tested
product(s) was not clearly identified. However, studies by Walker
et al. (1974) with technical cyanazine (97% a.i.) in three different
animal species reflected LD50s of 182, 380 and 141 mg/kg for the rat,
mouse and rabbit, respectively.
0 The acute dermal LD50 in rabbits treated with technical cyanazine
(purity unspecified) was >2,000 mg/kg (SRI, 1967a; Young and Adamik,
1979c); in rats, the LD50 was >1,200 mg/kg (97% a.i.) (Walker et al.,
1974).
0 The acute inhalation LCso f°r cyanazine dust (% a.i. not specified)
in rats was >2.28 mg/L/hr (Bishop, 1976) (toxicity category III).
0 In a study by Walker et al. (1968), groups of 10 female CFE rats,
5 months old, were treated by gavage with single oral doses of 1,
5 or 25 mg/kg of a wettable powder formulation (75% a.i.); the control
group received water. No diuretic effects were produced in the rats
receiving the formulation; however, serum protein and potassium
concentrations increased at the high dose, and serum osmolality
increased at 5 mg/kg, the Lowest-Observed-Adverse-Effect-Level (LOAEL).
The No-Observed-Adverse-Effect-Level (NOAEL) in this study appeared
to be 1 mg/kg; however, this study did not provide enough information
to determine the presence or absence of more significant effects at
this dosage level.
0 A 4-week oral toxicity study by Walker et al. (1968) was performed
using groups of 10 male and 10 female CFE rats, 5 weeks of age,
receiving diets containing 1, 10 or 100 ppm cyanazine (75% or 97% a.i.)
for 4 weeks; These doses are equivalent to 0.05, 0.5 or 5 mg/kg/day
(Lehman, 1959). A control group of 20 animals/sex was used. After
4 weeks, urine samples were collected for 16 hours (overnight), and
blood samples were used to determine the kidney function. Reductions
in body weight and food intake were noted at the high-dose level.
Osmolal clearance decreased in males, and tnis change was associated
with a decrease in free water clearance in both the low- and mid-dose
groups. In females, decreased urine and increased serum osmolality
were observed in the mid-dose group, and both creatinine clearance
and urine potassium concentrations increased in the low-dose group.
The LOAEL in this study appeared to be 0.05 mg/kg/day (lowest dose
-------
Cyanazine August, 1987
-7-
tested) based on kidney function tests, although additional
information was not available to determine if any other significant
adverse effects were noted at this level.
Dermal/Ocular Effects
0 Cyanazine caused mild eye irritation (100 mg) and slight skin irrita-
tion (2,000 mg) in rabbits. A skin sensitization test in guinea pigs
was negative (Walker et al., 1974; Young and Adamik, 1979d).
Long-term Exposure
0 In a 13-week oral study in dogs (Walker and Stevenson, 1968a, 1974),
groups of 5- to 7-month old beagle dogs, four animals/sex/treatment
group, were given daily doses of 1.5, 5 or 15 mg/kg/day cyanazine
in gelatin capsules. A control group of five animals/sex was given
empty capsules. The test material caused emesis within the first
hour of dosing in all of the high-dose males. Reduced body weight
gain was also noted in the high-dose group during the second half of
the study period as well as increased kidney and liver weights in the
females of this group. Thus, the LOAEL was 15 mg/kg/day and the
NOAEL was 5 mg/kg/day.
0 In a 13-week mouse feeding study (Fish et al., 1979), groups of 12
animals/sex/dose were fed diets containing 10, 50, 500, 1,000 or
1,500 ppm, equivalent to 1.5, 7.5, 75, 150 or 225 mg/kg/day (Lehman,
1959). The control group consisted of 24 animals/sex. Body weight
gain reduction was observed in both sexes at 75 mg/kg/day and above.
Statistically significant increases in liver weights were observed in
both sexes at 75 mg/kg/day and above. Thus, the LOAEL was 75 mg/kg/day
and the NOAEL was 7.5 mg/kg/day.
0 An initial 13-week rat feeding study by Walker and Stevenson (1968a)
was performed using 0.1, 1.0 or 100 ppm (equivalent to 0.005, 0.05
or 0.5 mg/kg/day; Lehman, 1959) of technical cyanazine (purity not
specified: 97% or 75% a.i. WP) in feed. Each dosage group had 20
animals/sex; the control group had 40 animals/sex. Body weight gain
decreased in all dosage groups in males and in the high-dose female
group. A NOAEL was not reflected in this study for males, although
it appeared to be 0.05 mg/kg/day for females.
0 Walker and Stevenson (1968b) repeated the above study in rats at dose
levels of 1.5, 3, 6, 12, 25, 50 or 100 ppm; these levels are equivalent
to 0.075, 0.15, 0.30, 1.25, 2.5 or 5 mg/kg/day (Lehman, 1959). Similar
effects were noted; however, a NOAEL of 25 ppm (1.25 mg/kg/day) was
identified.
0 In a 2-year study in dogs (Walker et al., 1970a), groups of 4- to
6-month-old beagle dogs were treated with technical cyanazine (97%
a.i., in gelatin capsules) at dose levels of 0.625, 1.25 or 5 mg/kg/
day. Each group consisted of four animals/sex. The control group
consisted of six animals/sex and received empty gelatin capsules.
Frequent emesis within 1 hour of dosing was observed throughout the
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Cyanazine August, 1987
-8-
study period in the high-dose group; this effect was associated with
reduction of growth rate and serum protein. The NOAEL appeared to be
1.25 mg/kg/day; however, this NOAEL should be considered with reser-
vations because the study did not provide adequate explanation relative
to missing histological data on one of four female dogs in the 1.25-
mg/kg/day dosage group. In addition, the reported data were limited
to a summary report.
e In a 2-year study in mice (Shell, 1981), cyanazine technical (purity
not specified) was given in feed to CD mice at 10, 25, 50, 250 or
1,000 ppm, equivalent to 1.5, 3.75, 7.5, 37.5 or 150 mg/kg/day (Lehman,
1959); 50 animals/sex were used in the treatment groups, and 100
animals/sex were used as controls. Toxic effects reported at the two
high-dose levels, 37.5 and 150 mg/kg/day, included poor appearance
and skin sores, increased mortality in the female animals in both
groups, increased relative brain weight in both sexes, increased
relative liver weight in the two female groups, and decreased absolute
and differential leukocyte values in both sexes. Anemia was noted at
150 ng/kg/day in the females, as well as increased blood protein and
increased relative kidney weight. Cyanazine did not demonstrate an
oncogenic potential in this study. The NOAEL for systemic toxicity
in mice appeared to be 50 ppm (7.5 mg/kg/day).
0 Two chronic feeding studies in rats were available for review. In
one study (Walker et al., 1970b; also cited in Walker et al., 1974),
groups of 24 CFE rats/sex/dose received diets containing 6, 12, 25
or 50 ppm, equivalent to 0.3, 0.6, 1.25 or 2.5 mg/kg/day (Lehman,
1959) cyanazine (97% a.i.); 45 rats/sex were used as controls. The
authors indicated that no effects due to cyanazine were noted in this
study, although reduction in growth rate was noted in both sexes at
2.5 mg/kg/day and in females at 1.25 mg/kg/day. A review of this
study (U.S. EPA, 1984b) indicated that cyanazine appeared to be
tumorigenic in both male and female rats based on the increased
incidences of thyroid tumors in all treatment groups as compared to
the study's control group; increased incidences of adrenal tumors
also were noted in all male treatment groups. However, this study
was considered unacceptable because of several deficiencies: a
limited number of tissues per animal were examined microscopically;
the tumor incidences were calculated based on the number of animals
tested rather than on the number of specific tissues histologically
examined; gross examination and histologic findings for nonneoplastic
lesions were not adequately reported; and only limited hematology,
clinical blood chemistry and urinalyses data were presented.
0 Simpson and Dix (1973) repeated the above 2-year study using 1, 3 or
25 ppm, equivalent to 0.05, 0.15 or 1.25 mg/kg/day (Lehman, 1959);
however, convulsions were noted in the rats 3 months after the study
initiation and throughout the remainder of tne study period.
Approximately 42% of the animals were affected, and the incidence was
not considered to be dose-related. The incidence of animals with
convulsions was similar in both the control and high-dose male groups
(21/48 and 11/24, respectively).
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Cyanazine August. 1987
-9-
Reproductive Effects
0 A three-generation reproduction study in Long-Evans rats (Eisenglord
et al., 1969) using technical cyanazine (unknown percentage a.i.) at
dietary levels of 3, 9, 27 or 81 ppm (0.15, 0.45, 1.35 or 4.05 mg/kg/day)
did not reflect a significant effect on reproduction parameters. The
NOAEL in this study appeared to be 1.35 mg/kg/day; the LOAEL was
4.05 mg/kg/day (highest dose tested) based on findings related to
reduced body weight gain in parental animals, and increased relative
brain weight and decreased relative kidney weight in F3b female
weanlings.
Developmental Effects
0 Cyanazine appeared to cause teratogenic effects and developmental
toxicity in two animal species, the rabbit and the rat (Bui, 1985b).
0 In the rabbit study (Shell Toxicology Laboratory, 1982), 7- to 11-
month-old New Zealand White rabbits were orally dosed with cyanazine
(98% a.i.) in gelatin capsules at levels of 0, 1, 2 or 4 mg/kg/day on
gestation days 6 through 18 (22 dams/dose/group). At 2 and 4 mg/kg/day,
maternal toxic effects included anorexia, weight loss, death and
abortion. Alterations in skeletal ossification sites, decreased
litter size, and increased postimplantation loss were observed at
2 and 4 mg/kg/day. Malformations were also noted at 4 mg/kg/day as
demonstrated by anophthaImia/microphthalmia, dilated brain ventricles,
domed cranium and thoracoschisis; however, these responses were
observed at levels in excess of maternal toxicity. The maternal and
developmental toxicity NOAELs were 1 mg/kg/day.
0 In a rat study by Lu et al. (1981, 1982), 122-day-old Fischer 344
rats (30 dams/group) were administered cyanazine (97% a.i.) by gavage
at dose levels of 0, 1.0, 2.5, 10.0 or 25.0 mg/kg/day on gestation
days 6 through 15; the dosages were suspended in a 0.2% Methocal
emulsion as vehicle. Maternal body weight reductions during dosing
were noted at the 10- and 25-mg/kg/day levels. Diaphragmatic hernia .
associated with liver microphthalmia was observed at the 25 mg/kg/day
dose level. A teratogenic NOAEL could not be determined from this
study.
0 The above study was repeated in the same strain of rats, Fischer 344,
by Lochry et al. (1985) in order to further examine the malformations
reported in the study by Lu et al. (1981). In this study, the dams
(70/dosage group) were 86 days old. Cyanazine (98% a.i.) was admini-
stered by gavage in an aqueous suspension of 0.25% (w/v) methyl
cellulose at dose levels of 0, 5, 25 or 75 mg/kg/day on days 6 through
15 of gestation. One-half of the dams in each group were selected
for Cesarean delivery on day 20 of gestation. The remaining half of
the dams in each group were allowed to deliver, and both they and
their pups were observed for 21 days before sacrifice. Maternal body
weight reductions during dosing were noted in all dosage groups and
appeared to be partly associated with lower food intake during the
dosing period. Alteration in skeletal ossification sites were also
observed in the fetuses at all dose levels. Teratogenic effects were
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Cyanazine August, 1987
-10-
demonstrated at 25 and 75 mg/kg/day as anophthalmia/microphthalmia,
dilated brain ventricles and cleft palate in the fetuses, and abnor-
malities of the diaphragm (associated with liver protrusion) in pups
sacrificed at time of weaning. The maternal and developmental toxicity
NOAELs were lower than 5 mg/kg/day (lowest dose tested), and the
teratogenic NOAEL was 5 mg/)cg/day (Bui, 1985a).
0 An additional study in Sprague-Dawley rats (Shell Development Company,
1983) did not reflect any maternal or developmental toxicity at the
highest dose tested, 30 mg/kg/day.
Mutagenicity
• The mutagenic potential of cyanazine has not been investigated
adequately, and only limited information was available for evaluation.
0 A study by Dean et al. (1975) using technical cyanazine (80% a.i.)
in mice of both sexes did not reflect any increase in chromosomal
aberrations in the bone marrow cells. The animals were examined at
8- and 24-hour intervals after oral dosing with 50 or 100 mg/kg
cyanazine. However, the sensitivity of this test was potentially
compromised because the positive control data did not reflect a
significant number of aberrations: the percent of cells showing
chromatid gaps in the positive control (cyclophosphamide) was not
statistically significant at the p <0.05 level (U.S. EPA, 1985b).
0 Dean et al. (1974) used technical cyanazine (purity not specified)
to induce dominant lethal effects in male CF1 mice. The test
was negative at the dose levels tested (80, 160 and 320 mg/kg).
However, this study appeared to be invalid because there was no
positive control for comparison of data, and a range-finding test was
not performed to select the appropriate dosages used in this study
(U.S. EPA, 1984b).
0 Cyanazine is a member of the triazine family of herbicides. It is known
that the triazines follow similar metabolic pathways (i.e., N-dealkyla-
tion, S-dealkylation or 0-dealkylation and conjugation with glutathion)
that result in common or closely related metabolites. Waters, et al.
(1980) noted that a triazine herbicide (atrazine) gave a positive
mutagenic response in the Drosophila sex-linked recessive lethal test
(DRL), although this chemical gave a negative response in an in vitro
test battery with microorganisms. Hence, the potential for cyanazine
to give a positive response in a similar test exists (U.S. EPA, 1984b).
Carcinogenic!ty
0 Cyanazine was not determined to have a carcinogenic potential in a
2-year mouse study (Shell, 1981).
0 Cyanazine was not oncogenic in 2-year rat studies by Walker et al.
(1970b) or by Simpson and Dix (1973); however, these studies were
deficient (see description of these studies under the section entitled
Long-term Exposure) and are considered to be inadequate by design to
determine the oncogenic potential of cyanazine.
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Cyanazine August, 1987
-1 1-
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 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/OOW 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 for determination
of the One-day HA for cyanazine. It is, therefore, recommended that the
Ten-day HA value for a 10-kg child, calculated below as 0.10 mg/L (100 ug/L),
be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The teratology study in rabbits by Shell Toxicology Laboaratory (1982) has
been selected as the basis for determination for the Ten-day HA for cyanazine
because it provides a short-term NOAEL (1 mg/kg/day for 13 days) for both
maternal and fetal toxicity. This study also reflects the lowest NOAEL when
compared with the teratology studies in rats described earlier, two in
Fischer 344 rats (Lu et al., 1981; Lochry et al., 1985) and one in Sprague-
Oawley rats (Shell Development Company, 1983).
Uting a NOAEL of 1 mg/kg/day, the Ten-day HA for a 10 kg child is
calculated as follows:
Ten-day HA = (1 mg/kg/day) (10 kg) = 0.10 mg/L (100 ug/L)
(100) (1 L/day)
where:
1 mg/kg/day = NOAEL based on maternal and fetal effects in rabbits
exposed to technical cyanazine orally for 13 days.
10 kg = assumed body weight of a child.
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Cyanazine August, 1987
-12-
TOO = 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 by a child.
Longer-term Health Advisory
No information was suitable for the determination of the Longer-term
HA for cyanazine. It is, therefore, recommended that the adjusted Drinking
Water Equivalent Level (DWEL) of 0.013 mg/L (13 ug/L) be used for a 10-kg
child as a conservative estimate for the Longer-term HA value and the DWEL
of 0.046 mg/L (46 ug/L), calculated below, be used for a 70-kg 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.
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.
Four chronic studies were available for evaluation: (1) a 2-year
oncogenic study in mice (Shell, 1981) with a potential NOAEL of 50 ppm
(approximately 7.5 mg/kg/day when using a conversion factor for food consumption
of 15% of the body weight); (2) a 2-year feeding study in dogs (Walker et al.,
1970a) with a NOAEL of 1.25 mg/kg/day; (3) a 2-year feeding/oncogenic study
in rats (Walker et al., 1970b, also cited in Walker et al., 1974) with a
NOAEL of 12 ppm (approximately 0.6 mg/kg/day when using a conversion factor
for food consumption of 5% of the body weight); however, this study was
considered unacceptable (U.S. EPA, 1984b) due to several deficiencies in the
study report (see Longer-term Exposure); and (4) a second 2-year feeding
study in rats (Simpson and Dix, 1973), which was also considered inadequate
because the control group reflected an effect, i.e., convulsions, that was
suggestive of cross-dosing.
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Cyanazine • 1987
-13-
The NOAEL in the mouse study (7.5 mg/kg/day) can be considered for this
calculation; however, this NOAEL is higher than the NOAEL in the Walker et al.
(1970a) dog study (1.25 mg/kg/day) or in the Walker et al. (1970b) rat study
(0.6 mg/kg/day). Since this rat study is considered unacceptable and since
the second rat study (Simpson and Dix, 1973) appeared to be flawed by the
invalidity of the control group, it is concluded that the 2-year dog study
(Walker et al., 1970a) will be used for the Lifetime HA.
The NOAEL of 1.25 mg/kg/day is used; however, because the data in this
study were of marginal acceptability, an uncertainty factor of 1,000 fold
will be applied to the HA calculations. This study NOAEL is also similar
to the NOAEL reflected in the suchronic study in rats by Walker and Stevenson
(1968b); thus the RfD value calculated below can be equally based on either
one of these studies (or both) using the same uncertainty factor.
Using a NOAEL of 1.25 mg/kg/day, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (1.25 mg/kg/day) = 0.0013 mg/kg/day
(1,000)
where:
1.25 mg/kg/day = NOAEL based on absence of toxicity in both the
2-year dog study and the 13-week rat study.
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 (as in the 13-week
rat study) or for a study with limited acceptability
(as in the 2-year dog study).
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-0013 mg/kg/day) (70 kg) = Q.0455 mg/L (46 ug/L)
(2 L/day)
where:
0.0013 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption by an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.046 mg/L) (20%) = 0.009 mg/L (9 ug/L)
-------
Cyanazine August, 1987
-14-
where:
0.046 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Available toxicity data indicate that cyanazine was not carcinogenic
in mice (Shell, 1981) or rats (Walker et al.f 1970b, 1974; Simpson
and Dix, 1973); however, in the rat, some increases were noted in the
incidences of both thyroid tumors (male and female rats) and adrenal
tumors (male rats); however, these increases were not statistically
significant. .
• Cyanazine is a chloro-s-triazine derivative that has a chemical
structure analagous to atrazine, propazine and simazine, the first
two of which were found to significantly (p <0.05) increase the
incidence of mammary tumors in rats. A new oncogenic study in rats
using simazine is not yet completed. Based on structure-activity
relationship, cyanazine may reflect a similar patteinof toxicity in
the rat. A new 2-year oncogenic study is required from the manufacturer
of this chemical to fill this data gap in the toxicity profile of
this chemical.
« Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), cyanazine may be classified in
Group D: not classified. This category is used for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 U.S. EPA Office of Pesticide Programs (OPP) has established residue
tolerances for cyanazine ranging from 0.05 to 0.10 ppm in or on raw
agricultural commodities (U.S. EPA, 1985a) based on a Provisional ADI
(PADI) of 0.0013 mg/kg/day.
VII. ANALYTICAL METHODS
0 Analysis of cyanazine is by a high-performance liquid chromatographic
(HPLC) method applicable to the determination of cyanazine in water
samples (U.S. EPA, 1985b). In this method, 1 L of sample is solvent
extracted with methylene chloride using a separatory funnel. The
methylene chloride extract is dried and exchanged to methanol during
concentration to a volume of 10 mL or less. Separation and measure-
ment of cyanazine is by HPLC with an ultraviolet (UV) detector. The
estimated method detection limit for cyanazine is 6 ug/L.
-------
Cyanazine Au*ust' 1987
-15-
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) adsorption
will remove cyanazine from water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
cyanazine on GAC.
0 GAC adsorption appears to be an effective method of cyanazine removal
from water. However, selection of individual or combinations of
technologies to attempt cyanazine removal from water must be based
on a case-by-case technical evaluation, and an assessment of the
economics involved.
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Cyanazine August, 1987
-16-
IX. REFERENCES
Bishop, A.L.* 1976. Report to Shell Chemical Company: Acute dust inhalation
toxicity study in rats. (Unpublished study received July 18, 1979 under
201-279; prepared by Industrial Bio-Test Laboratories, Inc., submitted
by Shell Chemical Co., Washington, D.C.; CDL:098395-A). MRID #00022789.
(Cited in U.S. EPA, 1984b)
Bui, Q.Q.* 1985a. Review of a developmental toxicity study (teratology and
post-natal study). U.S. EPA, internal memo from author to Robert Taylor
(reviewing study cited in Shell Development Company (1985), report no.
619-002, accession no. 257867).
Bui, Q.Q.* 1985b. Overview of the teratogenic potential of Bladex (cyanazine)
U.S. EPA, internal memo from author to Herb Harrison, dated June 5, 1985.
CHEMLAB. 1985. The Chemical Information System. CIS Inc., Bethesda, MD.
Cohen, S.Z., C. Eiden and M.N. Lorber. 1986. Monitoring ground water for
pesticides in the U.S.A. In; Evaluation of pesticides in ground water.
American Chemical Society Symposium Series. (in press)
Crayford, J.V., E.G. Hoadley, B.A. Pikering et al.* 1970. The metabolism of
the major plant metabolites of Bladex (DW 4385 and DW 4394) in the rat:
Group research report TLGR.0081.70. (Unpublished study prepared by'
Shell Research, Ltd). MRID #000223871. (Cited in U.S. EPA, 1984b)
Crayford, J.V., and D.H. Hutson.* 1972. Metabolism of the herbicide 2-chloro-
4-(ethylamino)-6-(1-cyano-1-methylethylamino)-S-triazine in the rat.
Pesticide Biochem. Physiol. 2:295-307. MRID #00022856. (Cited in
U.S. EPA, 1985a; U.S. EPA, 1984b)
Datta, P.R. 1984. Internal memorandum: Review of six documents regarding
monitoring of pesticides in northwestern Ohio rivers. U.S. Environmental
Protection Agency, Washington, DC.
Dean, B.J., E. Thorpe and D.E. Stevenson.* 1974. Toxicity studies on Bladex:
Dominant-lethal assay in male mice after single dose of Bladex. (Unpub-
lished study received August 13, 1976 prepared by Shell Research, Ltd.
for Shell Chemical Co., Washington, D.C.; CDL:095245-C). MRID #00023837.
(Cited in U.S. EPA, 1984b)
Dean, B.J., K.R. Senner, B.D. Perquin and S.M.A. Doak.* 1975. Toxicity
studies with Bladex chromosome studies on bone marrow cells of mice
after two daily oral doses of Bladex. (Unpublished study report no.
TLGR.0032074 received August 13, 1976 under 6F1729 prepared by Shell
Research, Ltd., submitted by Shell Chemical Co., Washington, D.C.;
CDL:095245-B). MRID #00023836. (Cited in U.S. EPA, 1984b)
Eadsforth, C.V. 1984. The leaching behavior of Bladex and its degradation
products in German soils under laboratory conditions. Expt. No. 2994.
Unpublished study submitted by Shell Chemical Company, Washington, DC.
-------
_ . August, 1987
Cyanazine ' '
-17-
Eisenglord, G., G.S. Loqunam and S. Leung.* 1969. Results of reproduction
study of rats fed diets containing SD 15418 over three generations:
Report No. 47. (Unpublished study received on unknown date under 9G0844;
prepared by Mine Laboratories. Inc., submitted by Shell Chemical Co.,
Washington, DC.; CDL:095023-D). MRID #00032346. (Cited in U.S. EPA,
1985b)
Fish, A., R.W. Hend and C.E. Clay.* 1979. Toxicity on the herbicide Bladex:
A three-month feeding study in mice: TLGR.0021.79. (Unpublished study
received July 19, 1979 under 201-279; submitted by Shell Chemical Co.,
Washington, DC.; CDL:09835-C). (Cited in U.S. EPA, 1984b)
Hutson, D.H., E.G. Hoadley, M.H. Griffiths and C. Donninger. 1970. Mercap-
turic acid formation in the metabolism of 2-chloro-4-ethylamino-6-
(l-methyl-l-cyanoethylamino)-s-triazine in the rat. J. Agric. Food. Chem.
18:507-512. (Data also available in U.S. EPA, 1984b, MRID # 00032348,
Shell Chemical Co., 1969.)
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S.
Lochry, E.A., A.M. Hoberman and M.S. Christian.* 1985. Study of the develop-
mental toxicity of technical Bladex herbicide (SO-15418) in Fischer-344
rats. (Unpublished report, submitted by Shell Oil Company; prepared by
Argus Research Laboratory, Inc., Horsham, PA, Report No. 619-002, dated
4/18/85)
Lu, C.C., B.S. Tang, E.Y. Chai et al.* 1981. Technical Bladex (R) (SD 15418)
teratology study in rats: Project no. 61230. (Unpublished study received
January 4, 1982 under 201-179; submitted by Shell Chemical Co., Washington,
DC.; CDL:098395-C). MRID #00091020. (Cited in Lu et al., 1982, and in
U.S. EPA, 1984b)
Lu, C.C., B.S. Tang and E.Y. Chai. 1982. Teratogenicity evaluations of
technical Bladex in Fischer-344 rats. Teratology. 25(2):59A-60A.
McMinn, A.L., and M.E. Standen. 1981. The mobility of Bladex and its
degradation products in soil under laboratory conditions. Unpublished
study submitted by Shell Chemical Company, Washington, DC.
Meister.R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Mirvish, S.S. 1975. Formation of N-nitroso compounds: Chemistry, kinetics,
and in vivo occurrence. Submitted by Shell Oil Co., Washington, DC.;
CDL:"070584-A). Fiche/Master ID 00000000.
HAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC.: National Academy Press.
NIOSH. 1977. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances. U.S. DHEW, PHS, CDC, Rockville,
MD. (Cited in U.S. EPA, 1984a)
-------
Cyanazine August, 1987
-18-
Osgerby, J.M., D.F. Clarke and A.T. Woodburn. 1968. The decomposition and
adsorption of DW 3418 (WL 19,805) in soils. Unpublished study submitted
by Shell Chemical Company, Washington, DC.
Plewa, M.J., and J.M. Gentile. 1976. Mutagenicity of atrizine: A maize-
microbe bioassay. Mutat. Res. 38:287-292.
Shell.* 1981. Two-year oncogenicity study in the mouse. (Unpublished report
submitted under pesticide petition number 9F2232, EPA accession number
247,295-298).
Shell Development Company.* 1983. Teratogenic evaluation of Bladex in SD CD
rats. (Unpublished report submitted by Shell Development Company,
prepared by Research Triangle Institute, Project No. 31T-2564, Report
dated 5/16/83, submitted to the EPA on 7/6/83; EPA Accession No. 071738).
(Cited in U.S. EPA, 1984b)
Shell Chemical Company.* 1969. Metabolism. Unpublished study. MRID #00032348.
(Cited in U.S. EPA, 1984b)
Shell Toxicology Laboratory (Tunstall).* 1982. A teratology study in New
Zealand White rabbits given Bladex orally. A report prepared by Sitting-
bourne Research Center, England; project no. 221/81, experiment no.
AHB-2321, November, 1982. Submitted on February 1, 1983 as document
SBGR.82.357 by Shell Oil Co., Washington, DC. under accession no.
071382. (Cited in U.S. EPA, 1984b)
Simpson, B.J., and K.M. Dix.* 1973. Toxicity studies on the s-triazine
herbicide Bladex: Second 2-year oral experiment in Research Limited,
London. Dated July 1973. EPA Accession No. 251954-955-956.
SRI.* 1967a. Stanford Research Institute Project 868-1, Report No. 39,
January 4, 1967. Acute dermal toxicity of SD-15418 (technical cyanazine).
Submitted by Shell Chemical Co., Washington, DC., Pesticide Petition
#9G0844, Accession #91460. (Cited in U.S. EPA, 1984b)
SRI.* 1967b. Stanford Research Institute Project 55 868, Report No. 43, May 26,
1967. Acute oral toxicity of SS-15418 (technical cyanazine). Submitted
by Shell Chemical Co., Washington, DC., Pesticide Petition #9G0844,
Accession #91460. (Cited in U.S. EPA, 1984b)
U.S. EPA. 1984a. U.S. Environmental Protection Agency. Draft health and
environmental effects profile for cyanazine. Cincinnati, OH: Environmental
Criteria and Assessment Office.
U.S. EPA.* 1984b. U.S. Environmental Protection Agency. Cyanazine toxicology
data review for registration standard. Washington, DC: Office of Pesticide
Programs.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. 40 CFR. 180.307.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. U.S. EPA Method 629
- cyanazine. Fed. Reg. 50:40701. October 4.
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Cyanazine August, 1987
-19-
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Walker, A.I.T., R. Kampjes and G.G. Hunter.* 1968. Toxicity studies in rats
on the s-triazine herbicide (DW 3418): (a) 13-Week oral experiments;
(b) The effect on kidney function: Group research report TLGR.0007.69.
(Unpublished study received Oct. 17, 1969 under 9G0844; prepared by
Shell Research, Ltd., England, submitted by Shell Chemical Co., Washington,
DC.; CDL:091460-H. ) MRID #00093200. (Cited in U.S. EPA, 1984b; Walker,
et al., 1974)
Walker, A.I.T., and D.E. Stevenson.* 1968a. The toxicity of the s-triazine
herbicide (DW 3418): 13-Week oral toxicity experiment in dogs: Group
research report TLGR.0016.68. (Unpublished study received Oct. 17, 1969
under 9G0844; prepared by Shell Research, Ltd., England, submitted by
Shell Chemical Co., Washington, DC.; CDL:091460-G.) MRID #00093199.
(Cited in U.S. EPA, 1984b; Walker et al., 1974)
Walker, A.I.T., and D.E. Stevenson.* 1968b. The toxicity of the s-triazine
herbicide (DW 3418): 13-Week oral experiment in rats: Group research
report TLGR.0017.68. (Unpublished study received Oct. 17, 1969 under
9G0844; prepared by Shell Research, Ltd., England, submitted by Shell
Chemical Co., Washington, DC.) MRID #00093198. (Cited in U.S. EPA,
1984b; Walker et al., 1974)
Walker, A.I.T., E. Thorpe and C.G. Hunter.* 1970a. Toxicity studies on the
s-triazine herbicide Bladex (DW 3418): Two-year oral experiment with
dogs: Group research report TLGR.0065.70. (Unpublished study received
December 4, 1970 under OF0998; prepared by Shell Research, Ltd., England,
submitted by Shell Chemical Co., Washington, DC.; CDL:091724-R.)
MRID #00065483.
Walker, A.I.T., E. Thorpe and C.G. Hunter.* 1970b. Toxicity studies on the
s-triazine herbicide Bladex (DW 3418): Two-year oral experiment with
rats: An unpublished report prepared by Tunstall Laboratory, submitted
by Shell Research, Ltd., London. (TLGR.0063.70). EPA Accession Nos.
251, 949-251, 953; PP# OF0998 (CDL:091724-Q). MRID #00064482.
Walker, A.I.T., V.K. Brown, J.R. Kodama, E. Thorpe and A.B. Wilson. 1974.
Toxicological studies with the 1,3,5-triazine herbicide cyanazine.
Pestic. Sci. 5(2):153-159. (Cited in U.S. EPA, 1984a)
Waters, M.D., V.F. Simmon, A.D. Mitchell, T.A. Jorgenson and R. Valencia.
1980. An overview of short-term tests for the mutagenic and carcinogenic
potential of pesticides. J. Environ. Sci. Health. 6:867-906.
Whittaker, K.F., 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis.
Lafayette, IN: Purdue University.
Wolfe, N.L., R.G. Zapp, J.A. Gordon and R.C. Fincher. 1975. N-Nitrosoatra-
zine: Formation and degradation. 170th Amer. Chem. Soc. Meeting.
Abstracts. American Chemical Society, p. 23.
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Cyanazine August, 1987
-20-
Young, S.M., and E.R. Adamik.* 1979a. Acute eye irritation study in rabbits
with SO 15418 (technical Bladex (R) herbicide): Code 16-8-0-0: Project
no. WIL-1223-78. (Unpublished study received Jan. 10, 1980 under 201-
281: submitted by Shell Chemical Co., Washington, DC.; CDL:099198-E.)
NRID #00026427. (Cited in U.S. EPA, 19845)
Young, S.M., and E.R. Adamik.* 1979b. Acute oral toxicity study in rats
with SO 15418 (technical Bladex (R) herbicide): Code 16-8-0-0: Project
no. WIL-1223-78. (Unpublished study received Jan. 10, 1980 under 201-
281: submitted by Shell Chemical Co., Washington, DC.; COL:099198-C.)
MRID #00026424. (Cited in U.S. EPA, 1984b)
Young, S.M., and E.R. Adamik.* 1979c. Acute oral toxicity study in rabbits
with SO 15418 (technical Bladex (R) herbicide): Code 16-8-0-0: Project
no. WIL-1223-78. (Unpublished study received Jan. 10, 1980 under 201-
281: submitted by Shell Chemical Co., Washington, DC.; CDL:099198-C.)
MRIO #00026425. (Cited in U.S. EPA, 1984b)
Young, S.M., and E.R. Adamik.* 1979d. Delayed contact in hypersensitivity
study in guinea pigs with SO 15418 (technical Bladex (R) herbicide):
Code 16-8-0-0: Project no. WIL-1223-78. (Unpublished study received
Jan. 10, 1980 under 201-281: submitted by Shell Chemical Co., Washington,
DC.; CDL:099198-F.) MRID #00026428. (Cited in U.S. EPA, 1984b)
Zendzian, R.P. 1985. Review of a study on Bladex dermal absorption. U.S. EPA,
internal memo to G. Werdig dated 2/20/85, reviewing study by Jeffcoat,
A.R. (Research Triangle Institute, RTI/3134/01F, Dec. 1984), Accession
no. 256324.
•Confidential Business Information submitted to the w*fice of Pesticide
Programs
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August, 1987
DACTHAL
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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Daethai
August, 1987
-2-
IX. GENERAL INFORMATION AND PROPERTIES
CAS No. 1861-32-1
Structural Formula
Dimethyl tetrachloroterephthalate
Synonyms
0 2,3,5,6-Tetrachlorodimethyl-1,4-benzenedicarboxylic acid; DCPA;
Chlorothal; Dacthalor; DAC; DAC-4; DAC-893; DCP (Meister, 1983).
Uses
Selective pre-emergence herbicide used to control various annual
grasses in turf, ornamentals, strawberries, certain vegetable
transplants, seeded vegetables, cotton, soybeans and field beans
(Meister, 1983).
Properties (Meister, 1983; Windholz et al., 1983; CHEMLAB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density (°C)
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
C10H604C14
331.99
Crystals
156°C
5,000 mg/L
4.15 (calculated)
Daethai has been found in 462 of 1,818 surface water samples analyzed
and in 33 of 615 ground water samples (STORET, 1987). Samples were
collected at 551 surface water locations and 576 ground water locations,
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Dacthal August, 1987
-3-
and dacthal was found in eight states. The 85th percentile of all
nonzero samples was 0.39 ug/L in surface water and 0.05 ug/L in
ground water sources. The maximum concentration found was 8.74 ug/L
in surface water and 0.05 ug/L in ground water.
Environmental Fate
0 In aqueous solutions, dacthal is stable to photolysis with a half-
life of greater than one week. Dacthal is stable to soil photolysis
(Registrant CBI data).
0 Soil metabolism of dacthal proceeds with a half-life of greater than
2-3 weeks. Degradation rate is affected by temperature. No degra-
dation of dacthal has been observed in sterile soils (half-life of
1,590 days) (Registrant CBI data).
0 Degradation products of dacthal include monomethyltetrachlorotere-
phthalate (MTP) and tetrachloroterephthalic acid (TTA) (Registrant
CBI data).
0 TTA has been shown to be very mobile in soils whereas dacthal is not
(Registrant CBI data).
III. PHARMACOKINETICS
Absorption
0 Hazleton Laboratories (no date) reported that humans receiving single
oral doses of dacthal (25 or 50 mg) excreted up to 6% of the 25 mg
dose in urine as metabolites over a 3-day period. Approximately
12% of the 50 mg dose was metabolized and excreted over a similar
time period. The data indicated that up to 12% of a 50 mg dose
could be absorbed in humans.
0 Skinner and Stallard (1963) reported that following administration of
single oral doses of dacthal (100 or 1,000 mg/kg) by capsule to dogs,
90 and 97% of the administered doses were eliminated as the parent
compound in the feces by 24 and 96 hours, respectively. Approximately
3% of dacthal was converted to the monoethylester of tetrachloro-
t»rephthalic acid (DAC 1449). Two percent was eliminated in the
urine and 1% in the feces. Less than 1% (0.07%) of DAC 1449 was
converted to tetrachloroterephthalic acid (DAC 954), which was also
excreted in the urine. The results indicated that dacthal was
absorbed poorly (about 3%) from the gastrointestinal tract of dogs.
Distribution
0 Skinner and Stallard (1963) reported that following a single oral
dose of dacthal (100 or 1,000 mg/kg) to dogs, there was no storage of
dacthal in the kidneys, liver or fat. However, DAC 954 was found in
the kidneys. The authors also reported that no dacthal was found in
the kidneys or liver of dogs that had been administered dacthal-T at
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Dacthal August, 1987
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10,000 ppm (250 mg/kg/day) in the diet for two years. The kidneys,
liver and fat contained DAC 1449, while the kidneys contained DAC 954
only. Both dacthal and DAC 1449 were found in the fat of dogs treated
with 10,000 ppm.
Metabolism
0 Hazleton Laboratories (no date) reported that humans who took single
oral doses of dacthal (25 or 50 mg) converted 3 to 4% of the dose to
DAC 1449 within 24 hours. After 3 days, approximately 6% of the
25 mg dose and 11% of the 50 mg dose were converted to DAC 1449. At
either dose, less than 1% was converted to DAC 954 in the 1- or
3-day time period.
0 Skinner and Stallard (1963) reported that in dogs administered single
oral doses of dacthal, small amounts were converted to DAC 1449 (3%)
or DAC 954 (0.07%).
• Hazleton and Dieterich (1963) reported similar results when dogs were
administered dacthal (10,000 ppm; 250 mg/kg bw) in the diet for
2 years.
Excretion
In human studies (Hazleton Laboratories, no date), 6% of a single
25 mg oral dose was excreted in urine as DAC 1449 and 0.5% as DAC 954
over a three-day period. Approximately 11% of the 50 mg dose was
converted to DAC 1449 and 0.6% was converted to DAC 954. The parent
compound was not found in the urine at either dose.
Skinner and Stallard (1963) reported that following the administra-
tion of a single oral dose (100 or 1,000 mg/kg) to dogs, 90 and 97%
was eliminated unchanged in the feces at 24 hours and 96 hours,
respectively. Approximately 3% was converted to DAC 1449; of this
3%, 2% was eliminated in the urine and 1% in the feces.
IV. HEALTH EFFECTS
Humans
Hazleton Laboratories (no date) reported that dacthal, administered
as single 25 mg or 50 mg oral doses to each of six volunteer subjects,
did not cause any observable effects. Assuming 70 kg body weight,
these amounts correspond to doses of 0.36 or 0.71 mg/kg. Kemograms,
liver, kidney and urine analyses from the six human volunteers were
normal.
Animals
Short-term Exposure
0 The acute oral LDso for male and female rats was reported to be
greater than 12,500 mg/kg (Wazeter et al., 1974a).
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Dacthal August, 1987
-5-
0 The acute oral LD5Q for male and female beagle dogs was reported to
be greater than 10,000 mg/kg (Wazeter et al., 1974b).
0 Keller and Kundzin (1960) administered pure dacthal to weanling
male Sprague-Dawley rats (10/dose) in the diet for 28 days at dose
levels of 0, 0.0082, 0.0824 or 0.824%. Based upon body weight and
compound consumption data provided by the investigators, these dietary
levels correspond to approximately 0, 7.6, 78.6 or 758 mg/kg/day.
Following treatment, no effects on growth, food consumption, survival,
body weights, organ weights, gross pathology and histopathology were
observed. This study identifies a NOAEL of 758 mg/kg/day (the highest
dose tested).
8 Keller (1961) reported that oral administration (by capsule) of
800 mgAg/day of DCPA (% a.i. unknown) to beagle dogs (two/sex) for
28 days resulted in loss of body weight, reduced appetite, increased
liver weight and liver to body weight ratio, centrilobular liver
congestion and degeneration.
Dermal/Ocular Effects
0 The acute dermal LD50 value for albino rabbits was reported to be
greater than 10,000 mgAg (Elsea, 1958). He also reported that
dacthal, when applied to rabbit skin, did not cause irritation or
sensitization.
0 Johnson et al. (1981) applied dacthal (2,000 mg/kg) for 24 hrs to
shaved intact or abraded back or flank skin of New Zealand rabbits
(five/sex) in a paste form. Desquamation (which ranged from very
slight to slight) and very slight erythema were observed. There was
no macroscopic or microscopic pathology noted, and dacthal caused no
signs of irritation or sensitization.
0 A single application of 3.0 rag of dacthal to the eyes of albino
rabbits produced a mild degree of irritation that subsided completely
within 24 hours following treatment (Elsea, 1958).
Long-term Exposure
0 Wazeter et al. (1977) fed CD rats (15/sex/dose) disodium dacthal in
the diet for 90 days at dose levels of 0, 50, 500, 1,000 or 10,000 ppm.
Based upon compound consumption and body weight data provided by the
authors, these dietary levels are approximately 0, 3.6, 36.4, 74 or
732 mg/kg/day for males and 0, 4.2, 43.2, 82.3 or 856 mg/kg/day
for females. General behavior, appearance, body weight, food con-
sumption, ophthalmoscopic evaluation, hematology, clinical chemistry,
urinalysis, gross pathology and histopathology were comparable for
treated and control groups. A NOAEL of 10,000 ppm (732 mg/kg/day
for males and 856 mg/kg/day for females, the highest dose tested)
was identified for this study.
0 Hazleton and Die tench (1963) fed beagle dogs (four/sex/dose) dacthal
in the diet at 0, 100, 1,000 or 10,000 ppm for two years. Based upon
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Dacthal August, 1987
-6-
body weight and food consumption data provided in the report, these
dietary levels are approximately 0, 2.6, 17.7 or 199 mg/kg/day for
males and 0, 3, 20.7 or 238 mg/kg/day for females. Physical
appearance, behavior, food consumption, hematology, biochemistry,
urinalysis, organ weight, organ-to-body weight ratio, gross pathology
and histopathology were comparable in treated and control groups at
all dose levels. A NOAEL of 10,000 ppm (199 mg/kg/day for males and
238 mgAg/day for females; the highest dose tested) was identified
for this study.
0 Paynter and Kundzin (1963) fed albino rats (35/sex/dose; 70/sex for
controls) dacthal in the diet for 2 years at 0, 100, 1,000 or
10,000 ppm. Based on food consumption and body weight data
provided in the report, these dietary levels correspond approximately
to 0, 5, 50 or 500 mg/kg/day. Physical appearance, behavior, hematology,
biochemistry, organ weights, body weights, gross pathology and histo-
pathology of treated and control animals were monitored. After 3
months at 10,000 ppm, slight hyperplasia of the thyroid was reported
in both sexes. After 1 year, increased hemosiderosis of the spleen
of females occurred at 10,000 ppm and there were slight alterations
in the centrilobular cells of the liver of both sexes. Kidney weights
were increased significantly in males fed 10,000 ppm at the end of
the 2-year study. Based on these data, a NOAEL of 1,000 ppm
(50 mgAg/day) was identified.
Reproductive Effects
0 Paynter and Kundzin (1964) conducted a two-generation study using
albino rats. Animals (8 males/16 females) were fed dacthal in the
diet at dose levels of 0, 0.1 or 1.0% for 24 weeks, prior to mating.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 6, 50 or 500 mg/kg/day.
This study reported an evaluation of data collected on the second
parental generation (P2) and through weaning of the first litter
(F2a)«> The authors reported that a second litter (Fyb^ was not
obtained. Following treatment, the following indices were evaluated;
fertility, gestation, live births and lactation. Since the fertility
index was 37% (6/16) at the 1% dose, 75% (12/16) at the 0.1% dose,
and only 19% (3/16) in controls, no conclusions could be reached.
The lactation index for the 0.1% group was significantly lower than
controls. No oth-»r adverse reproductive effects were observed.
0 Hazleton (1963) performed a one-generation reproduction study in
albino rats. Animals were given dacthal in the diet at 0, 1,000 or
10,000 ppm in the diet. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 50 or 500 mg/kg/day. No effects were detected on
fertility, gestation, number of live births or lactation. Based on
this information a NOAEL of 10,000 ppm (500 mg/kg/day; the highest
dose tested) was identified.
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Daethai August, 1987
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Developmental Effects
0 Powers (1964) fed pregnant New Zealand rabbits (six/dose) dietary
levels of dacthal-T (0, 1,000 or 10,000 ppm) on days 8 to 16 of
gestation. Assuming that 1 ppm in the diet of rabbits is equivalent
to 0.03 mg/kg/day (Lehman, 1959), this corresponds to about 0, 30 or
300 mgAg/day. Following treatment, fetal toxicity (number of live/dead
or resorptions), maternal effects (appearance, behavior, body weight)
and visceral and skeletal anomalies were evaluated. No adverse
effects were observed at any dose level tested. This study identified
a developmental NOAEL of 300 mg/kg/day (the highest dose tested).
Mutaqenicity
0 No significant increase in mutation frequency was observed in Droso-
phila melanogaster larvae that had been fed media containing 0.1 to
10 mM dacthal (Paradi and Lovenyak, 1981).
0 Dacthal had no mutagenic activity in Salmonella assays (Microbiological
Associates, 1977a), in in vivo cytogenetic tests (Microbiological
Associates, 1977b), in DNA repair tests (Microbiological Associates,
1977c) or in dominant lethal tests (Microbiological Associates, 1977d).
Carcinogenicity
0 Paynter and Kundzin (1963) fed albino rats (35/sex/dose; 70/sex for
controls) dacthal-T for 2 years at dose levels of 0, 100, 1,000 or
10,000 ppm. Based upon compound consumption and body weight provided
in the report, these dietary levels correspond approximately to 0, 5,
50 or 500 mg/kg/day. Based on gross and histologic examination, neo-
plasms of various tissues and organs were similar in type, localization,
time of occurrence, and incidence in control and treated animals.
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 mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
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Dacthal August, 1987
-8-
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
No information was found in the available literature that was suitable
for deriving a One-day HA. The study in humans by Hazleton Laboratories (no
date) was not selected since only low doses (0.36 or 0.71 mg/kg) were tested,
and longer-term studies in animals suggest the no-effect level may be much
higher. It is, therefore, recommended that the Ten-day HA value for the
10-kg child (75 mg/L; calculated below) be used at this time as a conservative
estimate of the One-day HA.
Ten-day Health Advisory
The 28-day feeding study in rats by Keller and Kundzin (1960) has been
selected to serve as the basis for determination of the Ten-day HA. In this
study, no adverse effects on growth, organ weight, food consumption, gross
pathology or histopathology were detected at 758 mg/kg/day.
The Ten-day HA for the 10-kg child is calculated as follows:
Ten^day HA = (758 mg/kg/day) (10 kg) . 75 ng/L (75fooO ug/L)
(100) (1 L/day)
where:
758 mg/kg/day = NOAEL, based on absence of effects on growth, organ
weight, food consumption, gross pathology or
histopathology in rats fed dacthai for 28 days.
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
No appropriate data were available for the calculation of a Longer-term
HA. Therefore, it is recommended that the modified DWEL, adjusted for a 10-kg
child (5 mg/L), be used at this time as a conservative estimate for a 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
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Dacthal *»«»•*• 1987
-9-
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year study in rats by Paynter and Kundzin (1963) has been selected
to serve as the basis for determination of the Lifetime HA value for dacthal.
This study identified a NOAEL of 50 mg/kg/day, based on absence of effects on
appearance, behavior, hematology, blood chemistry, organ weight, body weight,
gross pathology and histopathology in male rats. The LOAEL was 500 mg/kg/day,
based on thyroid hyperplasia, histological changes in the liver and increased
kidney weights.
Using this study, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (50 mg/kg/day) = 0.5 mg/]ig/day
(100)
where:
50 mg/kg/day = NOAEL, based on absence of toxic effects in rats
exposed to dacthal in the diet for two years.
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*5 mg/kq/day) (70 kg) =17.5 mg/L (17,500 ug/L)
(2 L/day)
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Dacthal August, 1987
-10-
where:
0.5 mgA9/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 = (17.5 mg/L) (20%) = 3.5 mg/L (3,500 ug/L)
where:
17.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Paynter and Kundzin (1963) fed dacthal to rats for 2 years and
reported no evidence of carcinogenic effects at dose levels up to
10,000 ppm (450 mg/kg/day for males and 555 mg/kg/day in females).
This study is limited in that the relatively small numbers of animals
used (35/sex/dose; 70/sex for controls) and the removal of animals
(10/sex/dose; 20/sex for controls) for interim sacrifice may have
resulted in there being too few animals available for observation of
late-developing tumors.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of dacthal.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), dacthal may be classified in
Group D: not classified. This category is for substances with inade-
quate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA has established residue tolerances for dacthal in or on
raw agricultural commodities that range from 0.5 ppm to 15.0 ppm
(U.S. EPA, 1985).
VII. ANALYTICAL METHODS
0 Analysis of dacthal is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated pesticides in water
samples (U.S. EPA, 1986b). In this method, approximately 1 liter of
sample is extracted with methylene chloride. The extract is concen-
trated and the compounds are separated using capillary column GC.
Measurement is made using an electron-capture detector. The method
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Dacthal August, 1987
-11-
detection limit has not been determined for dacthal, but it is estimated
that the detection limits for analytes included in this method are in
the range of 0.01 to 0.1 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Reverse osmosis (RO) is a promising treatment method for pesticide-
contaminated waters. As a general rule, organic compounds with
molecular weights greater than 100 are candidates for removal by RO.
Larson et al. (1982) report 99% removal efficiency of chlorinated
pesticides by a thin-film composite polyamide membrane operating at a
maximum pressure of 1,000 psi and a maximum temperature of 113°F.
• More operational data are required, however, to specifically determine
the effectiveness and feasibility of applying RO for the removal of
dacthal from water. Also, membrane adsorption must be considered when
evaluating RO performance in the treatment of dacthal-contaminated
drinking water supplies.
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Dacthal August, 1987
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IX. REFERENCES
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Elsea, J.R.* 1958. Acute oral administration; acute dermal application; acute
eye application. Unpublished study. MRID 00045823.
Hazleton Laboratories, Inc.* undated. Oral administration - humans. ODW
Document No. 0036.
Hazleton, L.N., and W.H. Dieterich.* 1963. Two-year dietary feeding - dogs.
Final Report. Unpublished study. MRID 00083584.
Hazleton Laboratories, Inc.* 1963. Reproduction study - albino rats. ODW
Document No. 0032.
Johnson, D., J. Myer and A. Olafsson.* 1981. Acute dermal toxicity (1,050)
study in albino rats. Unpublished study. MRID 00110553.
•
Keller, J.G. 1961.* 28-day oral administration - dogs. Unpublished study.
MRID 00083573.
Keller, J.G., and M. Kundzin.* 1960. Twenty-eight day dietary feeding study
in rats. Unpublished study. MRID 00083571.
Larson, R.E., P.S. Cartwright, P.K. Eriksson and R.J. Petersen. 1982. Appli-
cations of the FT-30 reverse osmosis membrane in metal finishing operations.
Paper presented in Tokohama, Japan.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Published by the Association of Food and Drug Officials of
the United States.
Meister, R., ed. 1983. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Company.
Microbiological Associates.* 1977a. Activity of DTX-0003 in the Salmonella/
microsomal assay for bacterial mutagenicity. ODW Document No. 0029.
Microbiological Associates.* 1977b. The activity of DTX-77-0006 in the in
vivo cytogenetic assay in rodents for mutagenicity. ODW Document No. 0029.
Microbiological Associates.* 1977c. Activity of DTX-77-0005 in a test for
differential inhibition of repair deficient and repair competent strains
of Salmonella typhimurium. ODW Document No. 0029.
Microbiological Associates.* 1977d. Activity of DTX-77-0004 in the dominant
lethal assay in rodents for mutagenicity. ODW Document No. 0029.
Paradi, E., and M. Lovenyak. 1981. Studies on genetical effect of pesticides
in Drosophila melanogaster. Acta Biol. Sci. Hung. 32:119-122.
-------
Dacthal August, 1987
-13-
Paynter, O.E., and M. Kundzin.* 1963. Two year dietary administration - rats.
Final Report. MRID 00083577.
Paynter, O.E., and M. Kundzin.* 1964. Reproductive study - rats. Unpublished
study. MRID 00053082.
Powers, M.B. 1964.* Reproductive study - rabbits. Unpublished study.
MRID 00053088.
Skinner, W.A., and D.E. Stallard.* 1963. Dacthal animal metabolism studies.
ODW Document No. 0033.
STORET. 1987.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.185. July 1, 1985. pp. 280-281.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. Septem-
ber 24.
U.S. EPA. 1986b. U.S. Environmental Agency. U.S. EPA Method #2 - Determina-
tion of chlorinated pesticides in ground water by GC/ECD, January 1986
draft. Available from U.S. EPA's Environmental Monitoring and Support
Laboratory, Cincinnati, OH.
Wazeter, F.X., E.I. Goldenthal and W.P. Dean.* 1974a. Acute oral toxicity
(LD50) male and female rats. Unpublished study. MRID 00031873.
Hazeter, F.x., E.I. Goldenthal and W.P. Dean.* 1974b. Acute oral toxicity
(LD50) in beagle dogs. Unpublished study. MRID 00031873.
Wazeter, F.X., E.I. Goldenthal et al.* 1977. Ninety-day toxicity study in
rats. ODW Document No. 0029.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index—an Encyclopedia of Chemicals and Drugs, lOth ed. Rahway, NJ:
Merck and Company, Inc.
•Confidential Business Information submitted to the Office of Pesticides
Programs.
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August, 1987
DALAPON
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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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Dalapon August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 75-99-0
Structural Formula
CH,CCIiCOOH
(2,2-Dichloropropionic acid)
Synonyms
• Dalapon (ANSI, BSI, WSSA), DPA, Basfapon and Basfapon B (discontinued
by BASF Wyandotte); Basfapon/Basfapon N, BH Dalapon and Crisapon
(Crystal Chemical Inter-America); Dalapon 85, Dalapon-Na, Ded-Weed
and Devipon (Devidayal); Dowpon, Dowpon M, Gramevin and Radapon (discon-
tinued by Dow); Revenge (Hopkins); Unipon (Meister, 1984).
Uses
0 Dalapon (2,2-dichloropropionic acid) is used as a herbicide in the
form of its sodium and/or magnesium salts to control grasses in crops,
drainage ditches, along railroads and in industrial areas (U.S. EPA,
1984).
Properties (U.S. EPA, 1984)
Chemical Formula C3H4C1202
Molecular Weight 143 (acid form)
Physical State (room temp.) liquid
Boiling Point 185 to 190°C
Melting Point 20°C
Density (°C)
Vapor Pressure —
Specific Gravity —
Hater Solubility (25°C) >800 mg/L
Log Octanol/Water Partition 3.87
Coefficient
Taste Threshold —
Odor Threshold ~
Conversion Factor --
Occurrence
0 Dalapon has been found in none of the surface water or ground water
samples analyzed from 14 samples taken at 14 locations (STORET, 1987K
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Dalapon August, 1987
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Environmental Fate
0 The sodium salt of dalapon has been shown to hydrolize slowly in
water to produce pyruvic acid, and the rate of hydrolysis increases
with increasing temperature. After 175 hours, the extent of hydrolysis
at 25°C for 1%, 5% and 18% dalapon solutions was 0.41%, 0.61% and 0.8%,
respectively (Brust, 1953).
0 Hydrolysis of solutions of either dalapon or dalapon sodium salt is
accelerated at alkaline pH values. For example, hydrolysis of dalapon
sodium salt at 60°C was 20% complete in 30 hours at which time the
equilibrium pH was 2.3. In contrast, hydrolysis was 50% complete
in 30 hours when the pH was maintained at 12 during the experiment
(Tracey and Bellinger, 1958).
0 Based on reaction rate studies, Kenaga (1974) concluded that both
dalapon sa31 and dalapon would have chemical hydrolysis half-lives of
several months at temperatures less than 25°C and at initial solution
concentrations of less than 1%. Considering the more rapid rate of
microbial degradation, those authors concluded that it does not appear
that chemical hydrolysis of dalapon is a particularly significant
degradative pathway in soils.
0 Because of its high water solubility and lack of affinity for soil
particles, appreciable adsorption of dalapon on suspended or bottom
sediments is not expected in natural waters. Chemical degradation
and volatilization probably occur too slowly to account for substantial
loss of dalapon from water. Aquarium studies conducted by Smith et al.
(1972) provide evidence that volatility is not a route for significant
loss of dalapon from water.
0 Microbial degradation is by far the most important process affecting
the fate of dalapon in soil. Other processes which are of lesser
importance are adsorption, leaching and runoff, chemical degradation
and volatilization. Based on the light absorption characteristics of
aqueous solutions of sodium salts of dalapon, it has been concluded
that photodecomposition of dalapon in field applications is improbable
(Kearney et al., 1965).
0 Although dalapon is subject to hydrolysis under field conditions,
chemical degradation is considered to be very slow and is unlikely
to be an important factor in the dissipation of dalapon from soil.
Smith et al. (1957) and Brust (1953) demonstrated that dalapon and
its sodium salt can undergo hydrolysis to pyruvate and HC1.
0 Although the laboratory studies indicate that dalapon is a hignly
mobile compound (Warren, 1954; Helling, 1971; Kenaga, 1974) and should
be readily leachable from soils, field data snow that under many
practical conditions dalapon does not move beyond the first six-inch
depth of soil. This is probably because microbial action proceeds at
a faster rate than leaching under favorable conditions (Kenaga, 1974).
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Dalapon August, 1987
-4-
0 The microbial degradation of dalapon in soil has been well established.
Thiegs (1955) compared the rates of degradation of dalapon in autoclaved
and non-autoclaved soils. The concentration of dalapon (59 ppm) in
the autoclaved soil did not change after incubation at 100°F for 1 week
while in the unsterilized soil, dalapon disappeared in 4 to 5 weeks
after one application and in 1 week after the second application of
50 ppm. Based on the observations that dalapon decomposition is
adversely affected by low soil moisture, low pH, temperatures below
20° to 25°C, and large additions of organic matter, Holstun and
Loomis (1956) concluded that dalapon degradation was a function of
microbiological activity.
III. PHARMACOKINETICS
Absorption
0 In both dogs and humans, orally administered dalapon is quickly excreted
in the urine. Dogs administered a single oral dose of 500 mg/kg
dalapon sodium salt excreted 65 to 70% of the administered dose in
48 hours (Hoerger, 1969). In a 60-day feeding study, dogs receiving
50 and 100 mg/kg of dalapon sodium salt excreted 25 to 53% of the
administered dose in the urine (Hoerger, 1969). Human subjects
consuming five successive daily oral doses of 0.5 mg of dalapon
sodium salt excreted approximately 50% of the administered dose over
an 18-day period (Hoerger, 1969). These data suggest that dalapon
is well absorbed from the gastrointestinal tract.
Distribution
Chronic oral administration of dalapon did not result in significant
bioaccumulation in either rats or dogs (Paynter et al., 1960). In both
rats and dogs, the highest levels of dalapon were found in the kidneys,
followed by the muscle and the fat (Paynter et al., 1960).
Metabolism
0 Although inadequate data are available to characterize dalapon
metabolism in hu-nans, data in cattle (Redemann and Hanaker, 1959)
suggest that dechlorination may be involved in the metabolism of
daldpon.
Excretion
Available information sug-j-^ts that at least 50% of orally admini-
stered dalapon is eliminated via the kidneys in dogs and hutaans
(Hoerger, 1963).
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Dalapon August, 1987
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IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 No information on the short-term health effects of dalapon in humans
was found in the available literature.
Long-term Exposure
0 No information on the long-term health effects of-dalapon in humans
was found in the available literature.
Animals
Short-term Exposure
0 The sodium salt of 'lalapon is relatively nontoxic, with an oral LO50
ranging from 3,860 mg/kg in the female rabbit to 7,570 mg/kg in the
female rat (Paynter et al., 1960).
Dermal/Ocular Effects
0 Concentrate^ sodium dalapon solutions have been found to be irritating
to the skin and eyes of rabbits (Paynter et al.f 1960).
Long-term Exposure
0 In a 90-day dietary study by Paynter et al. (1960), male and female
rats were exposed to sodium dalapon (65% pure) at levels of 0, 11.5,
34.6', 115, 346 or 1,150 mg/kg/day. Increases in kidney and liver
weight were observed in both sexes at 346 and 1,150 mg/tg/day. The
No-Observed-Adverse-Effect-Level (NOAEL) in this study was identified
as 11.5 mg/kg/day base! on increases in kidney weight at higher
doses. (See discussion under Longer-term Health Advisory below.)
0 In a 1-year study, sodium dalapon (65% pure) was administered to
dogs by capsule at levels of 0, 15, 50 or 100 mg/kg/day. Based on
increases in kidney weight at 100 mg/kg/day, the NOAEL was identified
as 50 mg/kg/day (Paynter et a:., 1960).
0 with the exception of an increase in kidney weight in male rats,
sodium dalapon (65% pure) was without eff«>.7t in a 2-year dietary study
(Paynter et al., I960); the NOAEL in this study «MS 15 mg/kg/day.
(See discussion under Longer-tern Health Advisory below.)
Reproductive Effects
0 Administers^ in the diet, sodium dalapon (65% pure) had no effects on
reproduction in the rat at dose levels of approximately 30, 100 or
300 mg/kg/day (Paynter et al., 1960).
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Dalapon August, 1987
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Developmental Effects
0 Sodium dalapon (purity not specified) was not teratogenic in the rat
at doses as high as 2,000 mg/kg/day (Emerson et al., 1971; Thompson
et al., 1971). (See Ten-day Health Advisory below.)
Mutagenicity
0 Dalapon was not mutagenic in a variety of organisms including Salmonella
typhimurium, Eseherichia coli, T4 bacteriophage, Streptomyces coelicolor
and Aspergillus nidulans (U.S. EPA, 1984). Although Kurinnyi et al.
(1982) reported that dalapon increased chromosome aberrations in mice,
the inadequate technical detail presented precluded an evaluation of
this study.
Carcinogenicity
0 No evidence of a carcinogenic response was observed in a 2-year
chronic feeding study in which sodium dalapon (65% pure) was
administered to rats at levels as high as 50 mg/kg/day for a period
of 2 years (Paynter et al., 1960).
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 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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No data were found in the available literature that were suitable for
determination of the One-day HA value for dalapon. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child (4.3 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
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Dalapon August, 1987
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Ten-day Health Advisory
The rat teratology study by Emerson et al. (1971) has been selected to
serve as the basis for determination of the Ten-day HA for a 10-kg child.
In this study, sodium dalapon (purity not specified; assumed to be 100%) was
orally administered to pregnant rats over a 10-day period (days 6 through
15 of gestation) at doses of 0, 500, 1,000 or 1,500 mg/kg/day. Although no
compound-related teratogenic response was seen, there was a decreased in
weight gain in the dams at the lowest level tested, 500 mg/kg/day. Decreased
weight gain was also observed in the pups, but only at higher levels (1,000
and 1,500 mg/kg/day). Standards for dalapon are commonly expressed in terms
of the acid rather than the salt. Thus, it is necessary to convert the LOAEL
for the sodium salt, 500 mg/kg/day, to the equivalent value for the acid.
The LOAEL for dalapon as acid = (500 mg/kg/day) (143) = 430 mgAg/day
1 65
where:
500 mg/kg/day = LOAEL for sodium dalapon.
143 = molecular weight of dalapon as acid in g/MWt.
165 = molecular weight sodium dalapon in g/MWt.
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA » (430 mg/kg/day) (10 kg) = 4.3 mg/1 (4300 ug/L)
r (1,000) (1 L/day)
where:
430 mg/kg/day = LOAEL for dalapon as acid based on body weight
decreases in dams.
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
The results of Paynter et al. (1960) suggest that the subchronic and
chronic toxicity of dalapon are much the same. Specifically, in a 97-day rat
subchronic dietary study, sodium dalapon (65% sodium dalapon; 16% sodium salts
of related chloropropionic acids; 2% sodium pyruvate; 5% sodium chloride; 5%
water; 7% undetermined) produced an increase in kidney weight in female rats
at 34.6 mg/kg/day and higher exposure levels but not at 11.5 mg/kg/day (NOAEL),
Similarly, in a two-year rat chronic dietary study, sodium dalapon exposure
(65% pure) resulted in an increase in male kidney weight at 50 mg/kg/day but
not at 15 mg/kg/day (NOAEL). Considering both Paynter et al. (1960) rat
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Dalapon August, 1987
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dietary studies together, the IS mg/kg/day NOAEL for sodium dalapon is
appropriate to calculate both a Longer-term HA and a Lifetime HA.
It is customary to express dalapon standards in terms of the acid rather
than the salt. The NOAEL used to derive the Longer-term HA is based on
studies (Paynter et al. , 1960) in which rats were exposed to sodium dalapon
that was 65% pure. Thus, a NOAEL for dalapon as the pure acid must be
calculated:
The NOAEL for dalapon as pure acid = (15 mg/kg/day) (0.65) (143) - Q nig/kg/day
1 65
where:
1 5 mg/kg/day » NOAEL for 65% pure sodium dalapon.
0.65 = purity of sodium dalapon used in determining NOAEL.
1 43 3 molecular weight of dalapon as acid.
165 = molecular weight of sodium dalapon.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (8 mg/kg/day) (10 kg) = 0.8 mg/L (800 ug/L)
(100) (1 L/day)
where:
8 mg/kg/day = NOAEL based on kidney weight increases in male rats.
1 0 kg » assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from animal study.
1 L/day = assumed daily water consumption of a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Lor jer-tenn HA - (8 "9Aq/day) (70 kg) „ 2>8 mg/L (2,800 ug/L)
(100) (2 L/day)
where all factors are the same except:
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Dalapon August, 1987
-9-
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 NQAEL (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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The data used to determine the Lifetime HA are identical to those used
to determine the Longer-term HA. Using the NOAEL of 8 mg/kg/day from the
2-year rat study by Paynter et al. (1960), the Lifetime HA for the 70-kg
adult is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (8 mgykg/day) = 0>08 mg/kg/day
(100)
where:
8 mg/kg/day = NOAEL for 100% dalapon as acid.
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 = (0.08 mg/kg/day) (70 kg) = 2>8 mg/L (2>aoo ug/L)
(2 L/day)
where:
0.08 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Dalapon August, 1987
-10-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (2.8 mg/L) (20%) = 0.56 mg/L (560 ug/L)
where:
2.8 mg/L » DWEL.
20% o assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenicity was found in a 2-year dietary study in
which sodium dalapon was administered to rats at levels as high as
50 mg/kg/day (Paynter et al., 1960).
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), dalapon may be classified
in Group D: not classified. This group is for substances with
inadequate human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The American Conference of Governmental Industrial Hygienists suggests
a Threshold Limit Value (TLV) of 1 ppm (6 mg/m3) as a time-weighted
average for an 8-hour work day.
0 Tolerances have been established for dalapon in a wide variety of
agricultural commodities (CFR, 1985) ranging from 0.1 ppm in milk to
75 ppm in flaxseed.
VII. ANALYTICAL METHODS
0 Analysis of dalapon is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated acid pesticides in water
samples (U.S. EPA, 1986b). In this method, approximately 1 liter of
sample is acidified. The compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolyzed with
potassium hydroxide, and extraneous organic material is removed by a
solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane as the derivatizing
agent. Excess reagent is removed, and the esters are determined by
electron-capture GC. The method detection limit has not been determined
for this compound.
VIII. TREATMENT TECHNOLOGIES
0 No information on treatment technologies capable of effectively
removing dalapon from contaminated water was found in the available
literature.
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Dalapon August, 1987
-11-
IX. REFERENCES
Brust, H. 1953. Hydrolysis of dalapon sodium salt solutions. E.G. Britton
Research Laboratory, The Dow Chemical Co., Midland, MI. November 4,
1953. Cited in Kenaga, 1974.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.150.
Emerson, J.L., D.J. Thompson and C.G. Gerbig. 1971. Results of teratological
studies in rats treated orally with 2,2-dichloropropionic acid (dalapon)
during organogenesis. Report HH-417, Human Health Research and Develop-
ment Laboratories, The Dow Chemical Co., Zionsville, IN (cited in
Kenaga, 1974).
Helling, C.S. 1971. Pesticide mobility in soils, I, II, III. Proc. Soil
Sci. Soc. Amer. 35:732-748.
Hoerger, F. 1969. The metabolism of dalapon. Blood absorption and urinary
excretion patterns in dogs and human subjects. Unpublished report.
Dow Chemical Company (cited in Kenaga, 1974).
Ho Iston, J.T., and W.E. Loomis. 1956. Leaching and decomposition of
2,2-dichloropropionic acid in several Iowa soils. Weeds. 4:205-217.
Kearney, P.C., et al. 1965. Behavior and fate of chlorinated aliphatic
acids in soils. Adv. Pest. Control Res. 6:1-30.
Kenaga, E.E. 1974. Toxicological and residue data useful in the environ-
mental safety evaluation of dalapon. Residue Rev. 53:109-151.
Kurinnyi, A.I., M.A. Pilinskaya, I.V. German and T.S. L'vova. 1982. Imple-
mentation of a program of cytogenic study of pesticides: Preliminary
evaluation of cytogenic activity and potential mutagenic hazard of 24
pesticides. Tsitologiya i Genetika. 16:45-49.
Meister, R., ed. 1984. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Paynter, O.E., T.W. Tusing, D.D.' McCollister and V.K. Rowe. 1960. Toxicology
of dalapon sodium (2,2-dichloropropionic acid, sodium salt). Agr. Food
Chem. 8:47-51.
Redemann, C.T., and J.W. Hanaker. 1959. The lactic secretion of metabolic
products of ingested sodium 2,2-dichloropropionate by the dairy cow.
Agricultural Research, the Dow Chemical Company. Seal Beach, CA (cited
in Kenaga, 1974).
Smith, G.N., M.E. Getzendaner and A.H. Kutschinski. 1957. Determination of
2,2-dichloropropionic acid (dalapon) in sugar cane. J. Agr. Food Chem.
5:675. Cited in Kenaga, 1974.
Smith, G.N., Y.S. Taylor and B.S. Watson. 1972. Ecological studies on dalapon
(2,2-dichloropropionic acid). Unpublished report NBE-16. Chemical
Biology Res., The Dow Chemical Co., Midland, MI (cited in Kenaga, 1974).
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Daiapon August, 1987
-12-
STORET. 1987.
Thiegs, B.J. 1955. The stability of dalapon in soils. Down to Earth, Fall
issue. Cited in Kenaga, 1974.
Thompson, D.J., C.G. Gerbig and J.L. Emerson. 1971. Results of tolerance
study of 2,2-dichloropropionic acid (dalapon) in pregnant rats.
Unpublished report HH-393. Human Research and Development Center, Dow
Chemical Company (cited in Kenaga, 1974).
Tracey, W.J., and R.R. Bellinger, Jr. 1958. Hydrolysis of sodium 2,2-dichloro-
ropionate in water solution. Midland Division, The Dow Chemical Co.,
Midland, MI (cited in Kenaga, 1974).
U.S. EPA. 1984. U.S. Environmental Protection Agency. Draft health and
environmental effects profile for dalapon. Environmental Criteria and
Assessment Office, Cincinnati, OH.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003, September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #3 -
Determination of chlorinated acids in ground water by GC/ECD, January
1986 draft. Available from U.S. EPA's Environmental Monitoring and
Support Laboratory, Cincinnati, OH.
Warren, G.F. 1954. Rate of leaching and breakdown of several herbicides
in different soils. NC Weed Control Conf. Proc., 11th Ann. Meeting,
Fargo, ND (cited in Kenaga, 1974).
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August, 1987
DIAZINON
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.
*N
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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Diazinon August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 333-41-5
Structural Formula
0,0-Diethyl-0-(6-methyl-2-{1-methylethyl)-4-pyrimidinyl)ester
Synonyms
0 Antigal; AG-500; Basudin; Bazudin; Ciazinon; Ducutox; Dassitox;
Dazzel; Dianon; Diater; Diaterr-Fos; Diazajet; Diazide; Oiazitol;
Diazol; Dicid; Dimpylat; Dizinon; Dyzol; Exodin; Flytrol; Galesan;
Kayazinon; Necidol/Nucidol; R-Fos; Spectacide; Spectracide (Meister,
1985).
Uses
Soil insecticide; insect control in fruit, vegetables, tobacco, forage,
field crops, range, pasture, grasslands and ornamentals; nematocide
in turf; seed treatment and fly control (Meister, 1985).
Properties (Meister, 1983; Windholz et al., 1983)
Chemical Formula C12H21°3N2SP
Molecular Height 304.36
Physical State (25°C) Colorless oil
Boiling Point 83 to 84°C (0.002 mm Hg)
Melting Point
Density
Vapor Pressure (20°C) 1.4 x 10~4
Water Solubility (20°C) 40 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold —
Odor Threshold —
Conversion Factor ~
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Diazinon August. 1987
-3-
Occurrence
0 Diazinon has been found in 7,230 of 23,227 surface water samples
analyzed and in 115 of 3,339 ground water samples (STORET, 1987).
Samples were collected at 3,527 surface water locations and 2,552
ground water locations, and diazinon was found in 46 states. The
85th percentile of all nonzero samples was 0.20 ug/L in surface
water and 0.25 ug/L in ground water sources. The maximum concen-
tration found was 33,400 ug/L in surface water and 84 ug/L in ground
water.
Environmental Fate
0 14c-Diazinon (99% pure), at 7 or 51 ppm on sandy loam soil, degraded
with a half-life of 37.4 hours after exposure to natural light (Blair,
1985). The degradate, oxypyrimidine, was detected at a maximum
concentration of 19.60% (13.5 hours) of applied material when exposed
to natural sunlight. After 35.5 hours (37.4 hours is the half-life)
of sunlight exposure, 20.7% of the radiolabeled material was in
soil-bound residues (some of which contained oxypyrimidine), 24.4%
was oxypyrimidine and 39.7% diazinon. Losses of 7% were attributed to
volatilization of diazinon and degradates (of which 0.5% was carbon
dioxide). The total He-radioactive material balance was 87-89% at
the 0 hour and 84% at all other experimental points.
• 14c-Diazinon (99% pure) degraded in sandy loam soil with a half-life
of 17.3 hours when exposed to natural sunlight (Martinson, 1985).
The degradate, oxypyrimidine, was detected at maximum concentrations
of 23.72% (32.6 hours) of applied after exposure to natural sunlight.
The degradate 2-(1'-hydroxy-1l-methyl)ethyl-4-methyl-6-hydroxypyrimidine
was present after 8 hours of natural sunlight exposure at 3.6% of the
applied material but was not present in the non-exposed samples. An
unidentified degradate resulting from non-photolytic degradation
(since it was also present in non-exposed samples), accounted for
about 7% of the applied material under sunlight.
0 In a Swiss sandy loam soil at 75% of field capacity and 25°C, ring-
labeled 14C-diazinon (97% pure) applied at 10 ppm rapidly degraded to
2-isopropyl-4-methyl-6-hydroxypyrimidine (IMHP) with a half-life of
less than one month. Within 14 days only 12.3% of the activity was
found as the parent; 72.9% was identified as ZMHP. Breakdown of IMHP
was slower than that of diazinon and 49% of the applied radioactive
material was in the form of IMHP after 84 days. After 166 days the
amount of IMHP decreased to 4.7% of the applied material. Increased
recoveries of 14C02 (55.6% after 166 days) and unextracted 4C residues
(15.1% after 166 days) corresponded to IMHP breakdown. No other major
metabolites were found. Radioactivity in the H2SO4 and ethylene
glycol traps was <1% of the applied 14C throughout the study and
material balance was generally above 80% of the applied material
(Keller, 1981).
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Diazinon August, 1987
-4-
III. PHARMACOKINETICS
Absorption
0 Mucke et al. (1970) reported that in both male and female rats, 69 to
80% of orally administered diazinon is excreted in the urine within
12 hours. This indicates that diazinon is well absorbed from the
gastrointestinal tract.
Distribution
0 The retention of diazinon labeled with 14C in the pyrimidine ring and
in the ethoxy groups was investigated in Wistar rats (Mucke et al.,
1970). Doses of 0.1 mg/rat were administered by stomach tube daily
for 10 days. Tissue levels 8 hours after the final dose were as
follows: stomach and esophagus, 0.25%; small intestine, 0.65%;
cecum/colon, 0.76%; liver, 0.16%; spleen, 0.01%; pancreas, 0.01%;
kidney, 0.04%; lung, 0.02%; testes, 0.02%; muscle, 0.77% and fat,
0.23%.
• Chickens were fed diazinon at levels of 2, 20 or 200 ppm in their food
for a period of 7 weeks (Mattson and Solga, 1965). Assuming that
1 ppm in the diet of chickens is equivalent to 0.125 mg/kg/day, this
corresponds to doses of about 0.25, 2.5 or 25 mg/kg/day (Lehman, 1959).
At the end of the feeding period, tissues from the animals fed 200
ppm (25 mg/kg/day) in the diet were analyzed for diazinon. There was
no diazinon detected in fat, white or dark muscle, heart, kidney,
liver, gizzard or eggs. The limit of sensitivity of the method was
0.05 ppm. There appeared to be no accumulation of diazinon in the
body at 200 ppm (25 mg/kg/day) in the diet.
Metabolism
0 The metabolism of diazinon 14C-labeled in the pyrimidine ring was
investigated in Wistar rats (200 g) after administration by stomach
tube (Mucke et al., 1970). In addition to some unchanged diazinon,
three major metabolites, all with the pyrimidine ring intact, were
identified in the urine, and to a lesser degree in the feces. A
fourth fraction containing polar materials was also found. The three
main metabolites were the result of a split at the oxygen-phosphorus
bond, with subsequent hydroxylation of the isopropyl side chain.
There was no significant expiration of labeled carbon dioxide, further
indicating that the pyrimidine nucleus remained intact.
0 The metabolism of diazinon was investigated ^n vitro in rat liver
microsomes obtained from adult male rats (Nakatsugawa et al., 1969).
It was found that diazinon underwent a dual oxidative metabolism
consisting of activation to diazoxon and degradation to diethyl
phosphorothioic acid. The authors noted that they had observed
similar pathways in studies with parathion and malathion, and these
results emphasized the importance of microsomal oxidation in the
degradation of organophosphate esters, indicating that many of the
so-called phosphatase products or hydrolysis products may actually be
oxidative metabolites.
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Diazinon August, 1987
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Excretion
The excretion of diazinon labeled with 14C in the pyrimidine ring and
in the ethoxy groups was investigated after administration by stomach
tube to wistar rats (Mucke et al., 1970). The diazinon was excreted
rapidly by both male and female animals, and 50% of the administered
dose was recovered within 12 hours. Of this, 69 to 80% was excreted
in the urine, and 18 to 25% in the feces. There was negligible
expiration of labeled carbon dioxide. There was no evidence of
accumulation of diazinon in any tissue.
IV. HEALTH EFFECTS
Diazinon is a reactive organophosphorus compound, and many of its
toxic effects are similar to those produced by other substances of
this class. Characteristic effects include inhibition of acetyl
cholinesterase (ChE) and central nervous system (CNS) depression.
Humans
Short-term Exposure
0 Weden et al. (1984) described a case report of diazinon poisoning
in a 26-year-old man who deliberately ingested a preparation
containing an unknown concentration of diazinon in an apparent suicide
attempt. Upon admission to the hospital, the patient exhibited most
of the usual symptoms of organophosphate poisoning, including muscarinic,
nicotinic and CNS manifestations. During treatment and monitoring, it
was noted that the urine output was very low and was dark and cloudy
in appearance. By the second day, the urine was found to contain
moderate amorphous crystals that could not be identified. With
increased intravenous fluids, the urine output increased, but the
crystaluria persisted and increased up to the 4th day, with a gradual
decrease for the next 5 days, at which time the patient was discharged.
Serum creatinine and urea nitrogen levels remained normal throughout
this period. It was noted that this phenomenon may have been related
to the specific pesticide formulation that had been ingested, but the
authors suggested that renal function should be monitored more closely
in persons with organophosphate poisoning.
0 Two men reportedly developed "marked" inhibition of plasma cholin-
esterase following the administration (route not specified) of five
doses of 0.025 mg/kg/day. A dose of 0.05 mg/kg/day for 28 days
reduced plasma cholinesterase in three men by 35 to 40%. In other
tests, each involving three to four men, doses ranging from 0.02 to
0.03 mg/kg/day produced reductions in plasma cholinesterase activity
of 0, 15 to 20 and 14%. In no case was there any effect on red
blood cell cholinesterase activity or on hematology, serum chemistry
or urinalysis. Thus, 0.02 mg/kg/day was identified as a No-Observed-
Adverse-Effect-Level (NOAEL) in humans (FAO/WHO, 1967, cited in
Hayes, 1982).
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Diazinon August, 1987
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Long-term Exposure
• No information was found in the available literature on the long-term
health effects of diazinon in humans.
Animals
Short-term Exposure
0 The acute oral toxicity of diazinon MG8 (a yellow oily liquid, 1,200
mg/mL) was studied in male albino rats (238 to 321 g) by DeProspero
(1972). Four groups of six rats each were given a single dose of
diazinon by gavage and then observed for 7 days. Dose levels
administered were 157, 313, 625 or 1,250 mg/kg. Within 4 hours of
administration, animals at the three higher levels displayed symptoms
of lethargy, tremors, convulsions and runny noses. Mortality in the
four groups was 0/6, 2/6, 5/6 and 6/6, respectively, with death
occurring between 8 and 24 hours after exposure. At 2 days, the
remaining animals at the two intermediate levels had recovered. There
was no mention of adverse symptoms at the lowest dose level. Gross
necropsy (performed only on animals that died) did not reveal abnormal
findings. The acute oral LDso value was calculated to be 395.6 mg/kg.
0 Hazelette (1984) investigated the effects of dietary hypercholesteremia
(HCOL) on sensitivity to diazinon in inbred male C56BL/6J mice. The
LD50 of diazinon in HCOL mice was nearly half that of diazinon admin-
istered to normal mice (45 versus 84 mg/kg). Cholesterol feeding
increased ChE activity in both blood and liver, and these increases
were negated by diazinon. Hepatic diazinon levels were also higher in
the HCOL animals. It was concluded that HCOL resulted in an increase
in susceptibility to, and toxicity of, diazinon.
0 Adult mongrel dogs (one/sex/dose) were fed diazinon (0 or 1.0 ppm in
the diet) for a period of 6 weeks (Doull and Anido, 1957). Assuming
that 1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day, this
corresponds to doses of about 0 or 0.025 mg/kg/day (Lehman, 1959).
Serum and erythrocyte ChE determinations were made on a weekly basis
before and during exposure. Neither plasma nor red blood cell ChE
varied by more than ±15% from control in exposed animals of either
sex, and there were no observed changes in body weight for the test
period. The apparent NOAEL for this study, based on blood chemistry
parameters, is 0.025 mg/kg.
0 The effect of diazinon on blood cell ChE activity was investigated
in sheep after the administration of single oral doses by gavage
of 50, 65, 100, 200 or 250 mg/kg (Anderson et al., 1969). Twenty-six
sheep were used in the study groups. Prior to dosing, 245 untreated
sheep were used to determine the normal range of erythrocyte ChE
values. A typical severe clinical response consisted of profuse
salivation, ataxia, dyspnea, dullness, anorexia and muscle twitching.
In mild cases, only dullness and anorexia were seen, but were suffi-
ciently pronounced to enable differentiation between normal and
affected animals. Sheep that were clinically affected by diazinon
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Diazinon August, 1987
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suffered a depression of ChE of more than 75%. However, there were
five animals (at the 50-mg/kg dose level) that tolerated depressions
of 80 to 90% without clinical effect. The ChE values fell to minimum
values within 1 to 4 hours, and remained close to this level until
about 8 hours after dosing, during which time symptoms were observed.
In those showing maximum depressions of 80% or more, the ChE activity
returned to about half its normal value by the 5th day, and thereafter
recovered only very slowly during a period of several weeks.
0 Davies and Holub (1980) compared the subacute toxicity of diazinon in
male and female Wistar rats. The diazinon was incorporated into a
semipurified diet at levels of 2 or 25 ppm. Assuming that 1 ppm in
the diet of rats is equivalent to 0.05 mg/kg/day, this corresponds to
doses of about 0.1 or 1.2 mg/kg/day (Lehman, 1959). Effects on ChE
activity were periodically assessed during a 28- to 30-day feeding
period. Levels of 25 ppm (1.2 mg/kg/day) diazinon in the diet for
30 days produced more significant reduction of ChE activity in plasma
(22 to 30%) and brain (5 to 9%) among treated females compared to
treated males. Erythrocyte ChE activity was significantly more
depressed (13 to 17%) in treated females relative to males at days
21 to 28 of the feeding period. At no time was ChE activity in any
tissue more reduced among treated males than females. At the 2-ppm
(0.1 mg/kg/day) dose level, diazinon failed to affect erythrocyte ChE
activity in either sex relative to controls. Plasma ChE activities
of treated males were not significantly different from control values,
but treated females showed significant depression (29%) of plasma ChE
activity. This investigation indicated that the female rat is more
sensitive to the toxicity of dietary diazinon than the male. Based
on the inhibition of ChE in the female animals observed at 2 ppm, the
Lowest-Observed-Adverse-Effect-Level (LOAEL) for this study was
identified as 0.1 mg/kg/day.
Dermal/Ocular Effects
0 Nitka and Palanker (1980) investigated the primary dermal irritation
and primary ocular irritation characteristics of a commercial formu-
lation of diazinon in New Zealand White rabbits. The percentage of
diazinon in the formulation was not given. After administration of a
single application of 0.5 mL to abraded and intact skin of six rabbits,
the formulation was judged not to be a primary dermal irritant. Nine
rabbits were used to examine the effect of administration of a single
dose of 0.1 mL of the formulation in one eye, and the results indicated
that it was not an ocular irritant.
Long-term Exposure
Female Wistar rats were fed a semipurified diet containing 0 or 0.1
to 15 ppm diazinon for up to 92 days with no visible toxic effects
(Davies and Holub, 1980). Weight gain and food consumption were
comparable to controls. Feeding studies up to 90 days revealed that
rats were highly sensitive to diazinon after 31 to 35 days of exposure,
as judged by reduction in plasma and erythrocyte cholinesterase (ChE)
activities. ChE was judged most sensitive. A NOAEL of 0.1 ppm,
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Diazinon August, 1987
-8-
which the authors translated to an equivalent daily intake of 9
ug/kg/day, is based on plasma ChE inhibition noted for up to 35 days
of feeding. Other data in this reference indicate that the depression
of plasma ChE is not further inhibited by continued dosing (up to 90
days).
0 Barnett and Kung (1980) fed Charles River CD-1 mice diazinon in the
diet at levels of 0, 4, 20 or 100 ppm for 18 months (males) or
19 months (females). Assuming that 1 ppm in the diet of mice is
equivalent to 0.15 mg/kg/day, this corresponds to doses of about 0,
0.6, 3 or 15 mg/kg/day (Lehman, 1959). Groups of 60 animals of each
sex were used at each treatment level, and a similar group served as
controls. In males, there was a significant reduction in weight gain
at the highest dose. Weight reduction was significant in all female
groups, although it did not appear to be dose- or treatment related.
There were no significant trends in mortality. Animals showed skin
irritation, loss of hair, skin lesions and piloerection. Gross and
microscopic examinations showed no inflammatory, degenerative, pro-
liferative or neoplastic lesions due to the administration of diazinon.
A LOAEL of 4 ppm (0.6 mg/kg/day) was identified for the mouse in this
study.
0 Horn (1955) fed diazinon to groups of 20 male and 20 female rats at
0, 10, 100 or 1,000 ppm in the diet for 104 weeks. Assuming that
1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day, this
corresponds to dose levels of about 0, 0.5, 5 or 50 mg/kg/day (Lehman,
1959). The rats were started on the diet as weanlings weighing 62 to
63 g. In preliminary studies, the highest dose caused significant
growth retardation. The animals for this group were initially given
100 ppm diazinon, which was increased gradually over a period of 11
weeks to the 1,000-ppm level. Mortality occurred in all groups,
including the controls, and pneumonia was common. In all groups,
body weight and food consumption were comparable to the controls.
Hematocrit values for males at 1,000 ppm were significantly depressed
when compared to controls. At 10 ppm, plasma ChE was inhibited by 60
to 73%, red blood cell ChE was inhibited 24 to 42% and brain ChE was
inhibited 8 to 10%. At 100 or 1,000 ppm, there was 95 to 100% inhibition
of ChE in plasma and blood cells. At 100 ppm, brain ChE was inhibited
19 (males) to 53% (females), and this increased to 41 (males) to 59%
(females) at 1,000 ppm. There were no significant gross pathological
findings. Based on inhibition of blood and plasma ChE, the LOAEL for
this study was identified as 10 ppm (0.5 mg/kg/day).
0 Woodard et al. (1965) exposed monkeys (three/sex/dose) to diazinon
orally for 52 weeks. The animals were started at doses of 0.1, 1.0
or 10 mg/kg/day for the first 35 days, but these doses were lowered
to 0.05, 0.5 or 5.0 mg/kg/day for the remainder of the study, apparent-
ly because of poor food consumption and decreased weight gain.
During the 52 weeks, body weight gain was slightly depressed in all
treated groups, and soft stools were observed in all animals, with
diarrhea in three animals (dose not specified). One female at the
0.5-mg/kg dose level had significant weight loss and signs of dehydra-
tion, emaciation, pale skin coloration and an unthrifty hair coat.
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Diazinon August, 1987
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One female at this level (it is not clear whether it is the same
animal just mentioned) exhibited decreased hemoglobin and a rapid
sedimentation rate at 39 and 53 weeks. Plasma ChE was inhibited 93%
at the high dose and 23% at the mid-dose, but no inhibition was noted
at 0.05 mg/kg (the low dose). Red blood cell ChE was inhibited 90%,
0% and 0% at the high, mid and low doses, respectively. Other bio-
chemical parameters were normal. Based on inhibition of ChE, a NOAEL
of 0.05 mg/kg/day and a LOAEL of 0.5 mg/kg/day were identified in
this study.
Reproductive Effects
0 Johnson and Cronin (1965) conducted a three-generation reproduction
study in Charles River rats. Beginning 70 days before mating, groups
of 20 females were fed diazinon (as 50% wettable powder) in the
diet at 4 or 8 ppm. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day, this corresponds to doses of about 0.2
or 0.4 mg/kg/day (Lehman, 1959). The end points monitored included:
general maternal condition, number of live and dead fetuses, number
of pups per litter, mean pup and litter weights, gross pathology of
F^a, F2a and F3a animals, and histopathology of F3b animals. All
findings were reported to be normal, but there were no detailed data
provided. A NOAEL of 8 ppm (0.4 mg/kg/day), the highest dose tested,
was identified in this study.
0 Diazinon was administered orally at dose levels of 0, 1, 25 or 100
mg/kg to groups of 18 to 22 New Zealand White rabbits on days 6 to
18 of gestation (Harris et al., 1981). At the 100-mg/kg level,
9/22 animals died. This was not quite significant (p <0.07) using
the Fisher Exact Test, although it was thought to be biologically
significant by the authors. Of these nine animals, seven showed
lesions indicative of gastrointestinal toxicity. At this dose,
animals also were observed to have tremors and convulsions and were
anorexic and hypoactive. These symptoms were not observed in animals
at the 7- and 25-mg/kg levels. One rabbit at the 25-mg/kg level
aborted on day 27, and all fetuses were dead. At this dose there
were no significant changes in weight gain compared to the control,
and no changes in the corpora lutea. There were also no statistically
significant changes in implantation sites, proportion of live, dead
or resorbed fetuses per litter, fetal weights or sex ratios. Based on
these data, the NOAEL for reproductive effects for the rabbit was
identified as 7 mg/kg/day.
Developmental Effects
0 Diazinon at dose levels of 7, 25 or 100 mg/kg was administered orally
to New Zealand White rabbits on days 6 to 18 of gestation (Harris
et al., 1981). Groups of 18 to 22 rabbits, 4 to 5 months of age and
weighing 3.0 to 4.1 kg, were given diazinon in 0.2% sodium carbo-
xymethyl cellulose (CMC) and a group of controls was given 0.2% CMC
only. At the 100-mg/kg level, 9/22 animals died, and although this
mortality was not quite significant (p <0.07) using the Fisher Exact
Test, it was thought to be biologically significant by the authors.
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Diazinon August, 1987
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There were no significant differences in abnormalities between the
control and treated groups, and it was concluded that diazinon was
neither fetotoxic nor teratogenic in the rabbit at these dose levels.
With respect to fetal effects, a NOAEL of 100 mg/kg/day, the highest
dose tested was identified. Based on maternal toxicity, a NOAEL of
25 ing/kg/day is identified.
Tauchi et al. (1979) administered diazinon by gavage to groups of
30 pregnant rats for 11 days (days 7 to 17 of gestation), at dose
levels of 0, 0.53, 1.45 or 4.0 mg/kg/day. In each group, 20 animals
were delivered by Cesarean section on day 17, while the remaining
10 were allowed to deliver normally. There were no effects on behavior
or learning ability, and no pathological lesions were detected at 10
weeks. It was concluded that diazinon was not teratogenic at the
doses tested. The NOAEL for fetal effects in this study was 4.0
the highest dose tested.
Mutagenicity
0 Fritz (1975) conducted a dominant lethal study in NMRI-derived albino
mice. Single doses of diazinon were administered orally to males at
levels of 15 or 45 rag/kg. After exposure, the males were mated to
untreated females several times over a period of 6 weeks. There were
no significant differences in mating ratios, the number of implantations
or embryonic deaths (resorptions) , and no adverse effects were observed
in the progeny at either dose level. It was concluded that diazinon
did not produce dominant lethal mutations in this test at the doses
used.
0 The mutagenicity of diazinon was tested in bacterial reversion-assay
systems with five strains of Salmonella typhimurium and one strain of
Escherichia coli (Moriya et al., 1983). No evidence of mutagenic
activity was noted in any of the test systems.
Four strains of Salmonella typhimurium were used to assay the muta-
genic potential of diazinon (Marshall et al., 1976). Negative
results were found by these investigators as well.
Carcinogenic! ty
0 A chronic bioassay for possible carcinogenicity of diazinon was
conducted in F-344 rats and B6C3F! mice (NCI, 1979). Groups of 50
animals were fed diazinon in the diet at the following levels: rats,
400 or 800 ppm; mice, 100 or 200 ppm. Assuming that 1 ppm in the
diet of rats and mice is equivalent to 0.05 and 0.15 mg/kg/day,
respectively, this corresponds to doses of about 20 or 40 mg/kg/day
in rats and about 15 or 30 mg/kg/day in mice (Lehman, 1959). There
was some hyperactivity notud in animals of both species, but there
was no significant effect on either weight gain or mortality. There
was no incidence of tumors that could be clearly related to diazinon,
and it was concluded that diazinon was not carcinogenic in either
species.
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Diazinon August, 1987
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0 Charles River CD-I mice were fed diazinon in the diet at levels of 4,
20 or 100 ppm for 18 months (males) or 19 months (females) (Barnett
and Kung, 1980). Assuming that 1 ppm in the diet of mice is equiva-
lent to 0.15 mg/kg/day, this corresponds to doses of about 0.6, 3 or
15 mg/kg/day (Lehman, 1959). Groups of 60 animals of each sex were
used at each treatment level, and a similar group served as controls.
In males at the highest dose level there was a significant difference
in weight gain from the controls. Weight reduction was significant
in all female treatment groups, but it did not appear to be dose-
or treatment-related. There were no significant trends in mortality.
Gross and microscopic examinations showed no inflammatory, degenerative,
proliferative or neoplastic lesions due to the administration of
diazinon, and the study was judged to be negative with respect to
carcinogenicity.
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 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 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. It is, therefore, recommended
that the Ten-day HA value for a 10-kg child (0.02 mg/L, calculated below) be
used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The most sensitive indicator of the effects of diazinon is inhibition of
ChE. However, this effect is reversible, and significant inhibition of this
enzyme often occurs without production of clinically significant effects.
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Diazinon August, 1987
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Consequently, selection of a NOAEL or LOAEL value based only on inhibition
of ChE, in the absence of any other toxic signs, is a highly conservative
approach.
The study in humans described by Hayes (1982) has been selected to serve
as the basis for determination of the Ten-day HA value for diazinon. Although
this study is a secondary source, it establishes a NOAEL in humans based on
the most sensitive end point, i.e., ChE. Hayes reported that in human volun-
teers, short-tern exposure to doses of 0.02 mg/kg/day did not result in
decreased ChE levels, while doses of 0.025 to 0.05 mg/kg/day caused ChE
reductions of 15 to 40%. This NOAEL (0.02 mg/kg/day) is supported by studies
in animals; e.g., based on blood and serum ChE, Doull and Anido (1957)
reported a NOAEL of 0.05 mg/kg/day in a 6-week study in dogs.
Using a NOAEL of 0.02 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (0.02 mg/kg/day) (10 kg) = Oo02 /L (20 /L)
(10) (1 L/day)
where:
0.02 mg/kg/day = NOAEL, based on absence of ChE inhibition in humans.
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
The study by Hoodard et al. (1965) has been selected to serve as the
basis for the Longer-term HA. Based on inhibition of plasma ChE in monkeys
exposed for 52 weeks, this study identified a NOAEL of 0.05 and a LOAEL of
0.5 mg/kg/day. These values are supported by the NOAEL for ChE inhibition of
0.025 mg/kg/day identified in a 6-week feeding study in dogs (Ooull and Anido,
1957) and by the LOAEL of 0.5 mg/kg/day identified by Horn (1955), based on
ChE inhibition in rats exposed for 2 years.
Using a NOAEL of 0.05 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-tera HA = (0'os "g/kg/day) (10 kg) = 0.005 mg/L (5.0 ug/L)
(100) (1 L/day) * *'
where:
0.05 mg/kg/day = NOAEL, based on absence of ChE inhibition in monkeys
given diazinon orally for 52 weeks.
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Diazinon August, 1987
-13-
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.
Using a NOAEL of 0.05 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = JO.05 mg/kg/day) (70 kg) „ Q.0175 mg/L (17.5 ug/L)
(100) (2 L/day)
where:
0.05 mg/kg/day = NOAEL, based on absence of ChE inhibition in monkeys
given diazinon orally for 52 weeks.
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.
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.
Available lifetime studies were not judged adequate for use in the deter-
mination of the Lifetime HAs since toxicological end points and numbers of
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Diazinon August, 1987
-14-
animals tested were limited. Therefore, the 13-week study of Davies and
Holub (1980) has been selected to serve as the basis for determination of
the Lifetime HA, with an additional safety factor of 10 for studies of less
than a lifetime. This study identified a NOAEL of 0.009 mg/kg/day.
Using a NOAEL of 0.009 mg/kg/day, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (0.009 mg/kg/day) = 0.00009 mg/kg/day
(100)
where:
0.009 mg/kg/day = NOAEL, based on plasma cholinesterase inhibition
in rats exposed to diazinon in the diet for up to
92 days.
100 = uncertainty factor of 10 for the end point of
toxicity-cholinesterase inhibition and an additional
factor of 10 for a study of less-than-lifetime
duration.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - <0-00009 mg/kg/day) (70 kg) = 0.00315 mg/L (3.15 ug/L)
(2 L/day)
where:
0.00009 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day =* assumed water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.00315 mg/L) (20%) = 0.00063 mg/L (0.63 ug/L)
where:
0.00315 mg/L *> DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Two studies on the carcinogenicity of diazinon in mice have been
reported (NCI, 1979; Barnett and Kung, 1980). Neither study revealed
any evidence of carcinogenicity.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of diazinon.
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Diazinon August, 1987
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0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), diazinon may be classified in
Group E: evidence of non-carcinogenicity for humans. This category
is for substances that show no 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 NAS (1977) has calculated an ADI of 0.002 mg/kg/day, based on a
NOAEL in humans of 0.02 mg/kg/day and an uncertainty factor of 10.
Assuming average body weight of human adult of 70 kg, daily consumption
of 2 liters of water and a 20% contribution from water, NAS (1977)
calculated a Suggested-No-Adverse-Effeet-Level of 0.014 mg/L.
VII. ANALYTICAL METHODS
0 Analysis of diazinon is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing pesti-
cides in water samples (U.S. EPA, 1986b). In this method, approximately
1 liter of sample is extracted with methylene chloride. The extract
is concentrated and the compounds are separated using a capillary
column GC. Measurement is made using a nitrogen-phosphorus detector.
The method detection limit has not been determined for diazinon but
it is estimated that the detection limits for analytes included in
this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that reverse osmosis (RO), granular-activated
carbon (GAC) adsorption and ozonation will remove diazinon from
water. The percent removal efficiency ranged from 75 to 100%.
0 Laboratory studies indicate that RO is a promising treatment method
for diazinion-contaminated waters. Chian (1975) reported 100% removal
efficiency using a cross-linked polyethylenimine (NS-100) membrane
and 99.88% removal efficiency with a cellulose acetate (CA) membrane.
Both membranes operated separately at 600 psi and a flux rate of
8-12 gal/ft2/day. Membrane adsorption, however, is a major concern
and must be considered as breakthrough of diazinon would probably
occur once the adsorption potential of the membrane was exhausted.
0 GAC is effective for diazinon removal. Dennis and Kobylinski (1983)
and Dennis et al. (1983) reported 94.5%, 90.5% and 76% diazinon
removal efficiency from wastewater in 6 hr. treatment periods with
45 Ibs of GAC. Also, 95% diazinon removal efficiency was achieved
in an 8-hr, treatment period with 40 Ibs of GAC.
0 Whittaker (1980) experimentally determined GAC adsorption isotherms
for diazinon and diazinon-methyl parathion solutions in distilled
water indicate that treatment with GAC can be used to remove diazinon.
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Diazinon August, 1987
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UV/03 oxidation treatment appears to be an effective diazinon removal
method. UV/03 oxidized 75% of diazinon at 3.4 gm/L ozone dosage and
a retention time of 204 minutes. When lime pretreatment was used,
UV/03 oxidized 99+% of diazinon at 4.1 gm/L ozone dosage and 240
minutes retention time (Zeff et alo, 1984).
Some treatment technologies for the removal of diazinon from water
are available and have been reported to be effective. However,
selection of individual or combinations of technologies to attempt
diazinon removal from water must be based on a case-by-case technical
evaluation, and an assessment of the economics involved.
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Diazinon August, 1987
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IX. REFERENCES
Anderson, P.M., A.F. Machin and C.N. Hebert. 1969. Blood cholinesterase
activity as an index of acute toxicity of organophosphorus pesticides in
sheep and cattle. Res. Vet. Sci. 10:29-33.
Barnett, J.W. and A.B.C. Kung. 1980. Carcinogenicity evaluation with
diazinon technical in albino mice. Industrial Bio-Test Laboratories, Inc.,
Chicago, IL.
Blair, J.* 1985. Photodegradation of diazinon on soil: Study No. 6015-208.
Unpublished study prepared by Hazleton Laboratories America, Inc. 130 pp.
(00153230)
Chian, E.S.K., W.N. Bruce and H.H.P. Fang. 1975. Removal of pesticides by
reverse osmosis. Environ. Sci. Technol. 9(1):52-59.
Davies, D.B. and B.J. Holub. 1980. Toxicological evaluation of dietary
diazinon in the rat. Arch. Environ. Contain. Toxicol. 9:637-650.
Dennis, W.H. and E.A. Kobylinski. 1983. Pesticide-laden wastewater treatment
for small waste generators. J. Environ. Sci. Health. B18(13):317-331.
Dennis, W.H., A.B. Rosencrance, T.M. Trybus, C.W.R. Wade and E.A. Kobylinski.
1983. Treatment of pesticide-laden wastewaters from Army pest control
facilities by activated carbon filtration using the carbolator treatment
system. U.S. Army Medical Bioengineering Research and Development
Laboratory, Frederick, MD. 21701. Technical Report 8203.
DeProspero, J.R.* 1972. Acute oral toxicity in rats: diazinon MG8.
Affiliated Medical Research, Princeton, New Jersey for Geigy Agricultural
Chemicals. MRID 00034096.
Doull, J. and P. Anido.* 1957. Effects of diets containing guthion and/or
diazinon on dogs. Department of Pharmacology, University of Chicago,
Chicago, XL. MRID 00046789.
FAO/WHO. 1967. Food and Agricultural Organization of the United Nations/World
Health Organization. Evaluation of some pesticide residues in food.
Geneva, Switzerland: FAO PL:CP/15, WHO/Food Add/67.32.
Fritz, H. 1975.* Mouse: dominant lethal study of diazinon technical.
Ciba-Geigy Ltd., Basle, Switzerland. MRID 00109037.
Harris, S.B., J.F. Holson and K.R. Fite.* 1981. A teratology study of diazinon
in New Zealand White rabbits. Science Applications, Inc., La Jolla, CA,
for Ciba-Geigy Corporation, Greensboro, NC. MRID 00079017.
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
Wilkins.
Hazelette, J.R. 1984. Dietary hypercholesteremia and susceptibility to the
pesticide diazinon. Diss. Abstr. Int. B. 44:2116.
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Diazinon August, 1987
-18-
Horn, H.J.* 1955. Diazinon 25W: chronic feeding-104 weeks. Hazleton Labora-
tories, Falls Church, VA for Geigy Agricultural Chemicals Division of
Ciba-Geigy Corp. MRID 00075932.
Johnson, C.D. and M.T.I. Cronin.* 1965. Diazinon: three generation repro-
duction study in the rat. Woodard Research Institute for Giegy Research
Laboratory. MRID 00055407.
Keller, A.* 1981. Degradation of Basudin in aerobic soil: Project Report 37/81.
Accession No. 251777. Report 7. Unpublished study received Nov. 5,
1982 under 4581-351; prepared by Ciba-Geigy, Ltd., Switz., submitted by
Agchem Div., Pennwait Corp., Philadelphia, PA; CDL:248818-L. (00118031)
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., P.O. Box 1494, Topeka, Kansas.
Marshall, T.C., H.W. Dorough and H.E. Swim. 1976. Screening of pesticides
for mutagenic potential using Salmonella typhimurium mutants. J. Agr.
Food Chem. 24(3):560-563.
Martinson, J.* 1985. Photolysis of diazinon on soil: Final Report: Biospherics
Project No. 85-E-044 SP. Unpublished study prepared by Biospherics Inc.
135 pp. (00153229)
Mattson, A.J. and J. Solga.* 1965. Analysis of chicken tissues for diazinon
after feeding diazinon for seven weeks. Geigy Research Laboratories.
MRID 00135229.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Meister, R., ed. 1985. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
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.
Mucke, W., K.O. Alt and H.O. Esser. 1970. Degradation of He-labeled
diazinon in the rat. J. Agr. Food Chem. 18(2):208-212.
Nakatsugawa, T., N.M. Tolman and P.A. Dahm. 1969. Oxidative degradation of
diazinon ty rat liver nicrosomes. Biochem. Pharmacol. 18:685-688.
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
NCI. 1979.* National Cancer Institute. Bioassay of diazinon for possible
carcinogenicity. Carcinogenic!ty Testing Program. NCI-NIH, Bethesda, MD.
DHEW Publication No. NIH 79-1392. MRID 00073372.
Nitka, S. and A.L. Palanker.* 1980. Primary dermal irritation in rabbits;
primary ocular irritation in rabbits. Final report: Study No. 80147
for Boyle-Midway, Cranford, NJ. MRID 00050966.
-------
Diazinon August, 1987
-19-
Tauchi, K., N. Igarashi, H. Kawanishi and K. Suzuki.* 1979. Teratological
study of diazinon in the rat. Institute for Animal Reproduction, Japan.
MRIO 00131150.
STORET. 1987.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1 -
Determination of nitrogen- and phosphorus-containing pesticides in ground
water by GC/NPO, January 1986 draft. Available from U.S. EPA's Environ-
mental Monitoring and Support Laboratory, Cincinnati, OH.
Weden, G.P., _._. Pennente and S.S. Sachdev. 1984. Renal involvement in organo-
phosphate poisoning. J. Am. Med. Assoc. 252:1408.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. thesis, Purdue
University.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, loth ed. Rahway, NJ: Merck and Co., Inc.
Woodard, M.W., K.O. Cockrell and B.J. Lobdell.* 1965. Diazinon SOW: Safety
evaluation by oral administration for 104 weeks; 52-week report. Woodard
Research Corporation. MRID 00064320.
Zeff, J.D., E. Leitis and J.A. Harris. 1984. Chemistry and application of
ozone and ultraviolet light for water reuse — Pilot plant demonstration.
Proceedings of Industrial Waste Conference. Vol. 38, pp. 105-116.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
DRAFT
August, 1987
1,3-DICHLOROPROPENE
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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1,3-Dichloropropene August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 542-75-6
Structural Formula
C1CH2 H C1CH2 Cl
\ / \ /
c = c c = c
/ \ / \
H Cl H H
(trans) (cis)
1,3-Dichloropropene
(approximately 46% trans/42% cis)
Synonyms
0 Dichloro-1,3-propene; 1,3-dichloro-1-propene; Telone; Telone II;
Dow Telone; cis/trans-1,3-dichloropropene; 1,3-D; DCP; D-D
(approximately 28% cis/27% trans).
Uses
• The pesticide 1,3-dichloropropene (DCP) is a broad spectrum soil
fumigant to control plant pests. Its major use is for nematode
control on crops grown in sandy soils of the Eastern, Southern and
Western U.S.
0 The usage of DCP has increased due to cancellation of the once widely
used product containing ethylene dibromide (EDB) and dibromochloro-
propane (DBCP) (U.S. EPA, 1986a).
0 Estimated usage of DCP containing products in 1984 to 1985 ranged from
about 34 to 40 million pounds (U.S. EPA, 1986a).
Properties (Dow Chemical USA, 1977, 1982; Patty, 1981)
Chemical Formula C3H4C12
Molecula- Weight 110.98 (pure isomers)
Physical State (25°C) Pale yellow to yellow liquid
Boiling Point about 104°C (104.3°C, cis; 112°C, trans)
Density (25°C) 1.21 g/mL
Vapor Pressure (25'C) 27.3 mm Hg
Specific Gravity about 1.2 (20/20°C)
Water Solubility (25°C) 0.1 to about 0.25% (1 to 2.5 g/L)
reported; miscible with most organic
solvents
Log Octanol/Water Partition 25
Coefficient
Flash Point about 28°C
Conversion Factor (25°C) 1 mg/L = 220 ppm; 1 ppm =» 4.54 mg/m3
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1,3-Dichloropropene August, 1987
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Occurrence
0 in California (Maddy et al., 1982), 54 veils were examined in areas
where Telone or D-D were used for several years. The well water did
not have measurable amounts of DCP (<0.1 ppb).
• Monitoring data from New York have shown positive results for DCP in
ground water (U.S. EPA, 1986b).
Environmental Fate
0 Available data indicate that DCP does leach to ground water. However,
the relative hydrolytic instability of the parent compound would
mitigate the potential for extensive contamination (U.S. EPA, 1986b;
U.S. EPA, 1986c).
0 The half-life of 1,3-DCP in soil was reported by Laskowski et al.
(1982) to be approximately 10 days while Van Dijk (1974) reported
3 to 37 days depending on soil conditions and analytical methods.
III. PHARMACOKINETICS
Absorption
0 Toxicity studies indicate that DCP is absorbed from skin, respiratory
and gastrointestinal systems (Patty, 1981).
0 Oral administration of DCP in rats resulted in approximately 90%
absorption of the administered dose (Hutson et al., 1971).
Distribution
0 Radiolabeled [C14] D-D (55% DCP) was administered orally in arachis
oil in rats. After 4 days, most of the administered dose was recovered
for the most part in urine and there were insignificant amounts (less
than 5%) remaining in the gut, feces, skin and carcass (Hutson et al.,
1971).
Metabolism
0 cis-Dichloropropene in corn oil was given as a single oral dose
(20 mg/kg bw) to two female Wistar rats. Urine and feces were
collected separately. The main urinary metabolite (92%) was N-acetyl-
S-[(cis)-3-chloroprop-2-enyl] cysteine. The cis-DCP has also been
shown to react with glutathione in the presence of rat liver cystol
to produce S[(cis)-3-chloroprop-2-enyl]glutathione. The cis-DCP is
probably biotransformed to an intermediate glutathione conjugate and
then follows the mercapturic acid pathway and is excreted in the
urine as a cysteine (Climie and Morrison, 1978).
0 In a study conducted by Dietz et al. (1984) rats and mice administered
(via gavage) up to 50 and 100 mg DCP/kg bw, respectively, demonstrated
no evidence of metabolic saturation.
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1,3-Dichloropropene August, 1987
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Excretion
In two studies (Hutson et al., 1971; Climie and Morrison, 1978)
[14c]cis- and/or trans-DCP, administered orally in rats, were excreted
primarily in the urine in 24 to 48 hours. When pulmonary excretion
was evaluated (Hutson et al., 1971), the cis and trans isomers were
3.9% and 23.6% of the administered dose, respectively. Most of the
cis-DCP was excreted in the urine.
IV. HEALTH EFFECTS
Humans
The only known human fatality occurred a few hours after accidental
ingestion of D-D mixture. The dosage was unknown. Symptoms were
abdominal pain, vomiting, muscle twitching and pulmonary edema.
Treatment by gastric lavage failed (Gosselin et al., 1976).
Inhalation of high vapor concentrations result in gasping, refusal to
breathe, coughing, substernal pain and extreme respiratory distress
at vapor concentrations over 1,500 ppm (Gosselin et al., 1976).
Venable et al. (1980) studied 64 male workers exposed to three carbon
compounds including DCP to determine if fertility was adversely
affected. The exposed study population was divided into 5 years exposure. Sperm counts and percent normal
sperm forms were the major variables evaluated. Although the study
participation rate for the exposed group was only 64%, no adverse
effects on fertility were observed.
Animals
Short-term Exposure
0 DCP is moderately toxic via single-dose oral administration. A
technical product containing 92% cis-/trans-DCP was fed as a 10%
solution in corn oil to rats. The oral LDSQS in male and female rats
were 713 and 740 mg/kg, respectively (Torkelson and Oyen, 1977). In
another study, the oral LDso in the mouse for both males and females
was 640 mg/kg (Toyoshima et al., 1978).
Dermal/Ocular Effects
• The percutaneous LD50s for male and female mice were greater than
1,211 mg/kg (Toyoshima et al., 1978).
0 The percutaneous administration of DCP in rabbits (3 g/kg) resulted
in mucous nasal discharge, depressed respiration and decreased body
movements. The LD50 by this route was 2.1 g/kg (Torkelson and Oyen,
1977).
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1,3-Dichloropropene August, 1987
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0 Primary eye irritation and primary dermal irritation studies in
rabbits indicated that DCP causes severe conjunctival irritation,
moderate transient corneal injury and slight skin erythema/edema.
Eye irritation was reversible 8 days post-instillation. The dermal
LD50 in rabbits was 504 rag/kg (Dow, 1978; Hanley et al., 1987).
Long-term Exposure
0 Rats, guinea pigs, rabbits and dogs were exposed to 4.5 or 13.6 mg/m3
DCP in air for 7 hours per day, and 5 days per week for 6 months.
The only effect noted was a slight apparently reversible microscopic
renal lesion in male rats exposed to the high dose (Torkelson and
Oyen, 1977).
0 Fischer 344 rats and CD-I albino mice were exposed to Telone II
(Production Grade) by inhalation exposure, 6 hours per day for 13
weeks at concentrations of 11.98, 32.14, and 93.02 ppm. Gross pathology
revealed an increased incidence of kidney discoloration in the treated
male rats relative to the control group. The significance of this
lesion is unknown (Coate et al., 1979).
0 Solutions of Telone (78.5% DCP) in propylene glycol were administered
by gavage to 10 rats/sex/dose for six days per week for a period of 13
weeks. The dose levels were 1, 3, 10 and 30 mg/kg/day. The control
groups were given propylene glycol. The daily administration of DCP
to rats by stomach intubation up to a dosage of 30 mg/kg/day did not
result in any major adverse effects. No significant effects on body
weight, food consumption, hematology and histopathology were noted.
However, at the 10 mg/kg/day dose, the relative weight of the kidney
of males was still higher than controls. The authors conclude that
the no-toxic-effect level for DCP was between 3 and 10 mg/kg/day.
The actual observed No-Observed-Adverse-Effect-Level (NOAEL) was
3 mg/kg/day (Til et al, 1973).
0 The National Toxicology Program (NTP, 1985) evaluated the chronic
toxicity and carcinogenicity of Telone II in rats and mice. These
studies utilized Telone II fumigant containing approximately 89%
cis- and trans-DCP. Groups of 52 male and female F344/N rats (doses
0, 25 or 50 mg/kg) and 50 male and female B6C3F1 mice (doses 0, 50
or 100 mg/kg) were gavaged with Telone II in corn oil, 3 days per
week up to 104 weeks. Arcillary studies were conducted in which
dose groups containing five male and female rats were killed after
receiving Telone II for 9, 16, 21, 24 or 27 months. Toxic effects
(noncarcinogenic) included basal cell or epithelial hyperplasia of
the forestomach of rats and mice at all treatment levels of DCP.
Epithelial hyperplasia of the urinary bladder of mice occurred at
both treatment levels in males and females. Kidney hydronephrosis
also occurred in mice. The study in male mice was considered inade-
quate due to the deaths of vehicle control animals. Many chronic
toxicity parameters (hematology/ clinical chemistry) were not deter-
mined. The DCP used in the NTP study had a different stabilizer from
the current Telone II.
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1,3-Dichloropropene August. 1987
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Reproductive Effects
0 Groups of male and female wistar rats were exposed to technical D-D
at 0, 64, 145 and 443 mg/m3 (0, 14, 12 and 94 ppm) for 5 days per
week over 10 weeks. Male mating indices, fertility indices and
reproductive indices were not affected by D-D exposure. No gross
morphological changes were seen in sperm. Female mating, fertility
and other reproductive indices were normal. Litter sizes and weights
were normal and pup survival over 4 days was not influenced by exposure
(Clark et al.f 1980).
Developmental Effects
0 Hanley et al. (1987) investigated the effects of inhalation exposure
to DCP on fetal development in rats. Pregnant Fischer 344 rats were
exposed to 0, 20, 60 and 120 ppm DCP for 6 hr/day during gestation
days 6 to 15. Maternal body weight gain was depressed in all of the
DCP-exposed rats in a dose-related manner. Therefore, the Lowest-
Observed-Adverse-Effect-Level (LOAEL) for this effect was 20 ppm DCP.
There was also significant depression of feed consumption in all
exposed rats, along with decreases in water consumption in rats
exposed to 120 ppm DCP. At 120 ppm there were significant increases
in relative kidney weights and decreases in absolute liver weights in
all exposed rats. There was a statistical increase in the incidence
of delayed ossification of the vertebral centra of rats exposed to
120 ppm DCP. This effect is of little toxicological significance due
to maternal toxicity observed at 120 ppm DCP.
0 Hanley et al. (1987) also studied the effects of inhalation exposure
to DCP on fetal development in rabbits. Pregnant New Zealand White
rabbits were exposed to 0, 20, 60 or 120 ppm DCP for 6 hr/day during
gestation days 6 through 18. In rabbits, evaluation of maternal
weight gain over the entire exposure period indicated significant
exposure-related decreases in both the 60- and 120-ppm groups.
Therefore, the NOAEL was 20 ppm DCP. Statistically significant
decreases in the incidence of delayed ossification of the hyoid and
presence of cervical spurs among the exposed group were considered
within normal variability in rabbits.
Mutagenicity
0 Tests of commercial formulations containing DCP (DeLorenzo et al.,
1975; Flessel, 1977; Neudecker et al., 1977; Brooks et al., 1978;
Sudo et al., 1978; Stolzenberg and Hine, 1980), a mixture of pure
cis-DCP and trans-DCP (DeLorenzo et al., 1975), and pure cis-DCP
(Brooks et al, 1978) were positive in the Salmonella typhimurium
strains TA1535 and TA100 with and without metabolic activation.
These results indicate that DCP acts by base-pair substitution and
is a direct acting mutagen.
0 DCP may be a mutagen that acts via frame shift mutation indicated
by studies (DeLorenzo et al, 1975) in which positive results were
obtained for TA1978 (with and without metabolic activation) for a
commercial mixture of DCP and a mixture of pure cis- and trans-DCP.
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1 , 3-Dichloropropene August, 1987
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0 A commercial mixture of DCP and pure cis-DCP were also positive with
and without metabolic activation in Salmonella typhimurium strain TA98
(Flessel, 1977; Sudo et al.( 1978; Brooks et al.f 1978).
0 Sudo et al. (1978) tested DCP in a reverse mutation assay with
Ji* coli B/r WP2 "it" negative results.
0 DCP was negative for reverse mutation in the mouse host-mediated test
with £. typhimurium G46 in studies by Shirasu et al. (1976) and Sudo
et al. (1978).
Carcinogenici ty
0 F344 rats of each sex were gavaged with Telone II in corn oil at
doses of 0, 25 and 50 mg/kg/day for 3 days per week. A total of
77 rats/sex were used for each dose group (52 animals/sex/group were
dosed for 104 weeks in the main oncogenicity study, and an ancillary
study where 5 animals/sex/ group were sacrificed after 9, 16, 21, 24
and 27 months' exposure to DCP). No increased mortality occurred in
treated animals. Neoplastic lesions associated with Telone II included
squamous cell papillomas of the forestomach (male rats: 1/52; 1/52;
9/52; female rats: 0/52; 2/52; 3/52), squamous cell carcinomas of
the forestomach (male rats: 0/52; 0/52; 4/52) and neoplastic nodules
of the liver (male rats: 1/52; 6/52; 7/52). The increased incidence
of forestomach tumors was accompanied by a positive trend for fore-
stomach basal cell hyperplasia in male and female rats of both treated
groups (25 and 50 mg/kg/day). The highest dose level tested in rats
(50 mg/kg/day) approximated a maximum tolerated dose level (NTP, 1985).
0 B6C3Fi mice of each sex were gavaged with Telone II in corn oil at
doses of 0, 50 and 100 mg/kg/day for 104 weeks. A total of 50 mice/sex
were used for each dose group. Due to excessive mortality in control
male mice from myocardial inflammation approximately 1 year after the
initiation of the study, conclusions pertaining to oncogenicity were
based on concurrent control data and NTP historical control data.
Neoplastic lesions associated with the administration of Telone II
included squamous cell papillomas of the forestomach (female mice:
0/50; 1/50; 2/50), squamous cell carcinomas of the forestomach (female
mice: 0/50; 0/50; 2/50), transitional cell carcinomas of the urinary
bladder (female mice: 0/50; 8/50; 21/48), and alveolar/bronchiolar
adenomas (female mice: 0/50; 3/50; 8/50). The increased incidence
of forestomach tumors was accompanied by an increased incidence of
stomach epithelial cell hyperplasia in males and females at the
highest dose level tested (100 mg/kg/day), and the increased incidence
of urinary bladder transitional cell carcinoma was accompanied by a
positive trend for bladder hyperplasia in male and female mice of
both treated groups (50 and 100 mg/kg/day) (NTP, 1985).
0 Thirty female Ha:ICR Swiss mice received weekly subcutaneous injections
of cis-DCP. The dose was 3 mg DCP/mouse in 0.05 mL trioctanoin
delivered to the left flank. After 77 weeks, there was an increased
incidence of fibrosarcomas at the site of injection. Six of the
30 exposed mice developed the tumors. There were no similar lesions
in the controls (Van Duuren, 1979).
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1,3-Dichloropropene August, 1987
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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 rag/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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
There are not sufficient data to derive a One-day Health Advisory value
for DCP. It is recommended that the Longer-term HA value for a 10-kg child
(30 ug/L, calculated below) be used at this time as a conservative estimate
of the One-day HA value.
Ten-day Health Advisory
There are not sufficient data to derive a Ten-day HA value for DCP. It
is recommended that the Longer-term HA value for a 10-kg child (30 ug/L) be
used as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
The Til et al. (1973) 90-day subchronic feeding study in rats has been
selected to serve as the basis for calculating th»- Longer-term HA for DCP.
This study resulted in a LOAEL of 10.0 mg/kg/day based on increased relative
kidney weight in males. No adverse biological effects were noted at the
next lowest dose (3.0 mg/kg/day). Therefore, the NOAEL is 3.0 mg/kg/day.
Based on the NOAEL of 3.0 mg/kg/day determined in this study, the Longer-
term HAs are calculated as follows:
For a 10-kg child:
Longer-term HA = (3.0 mg/kg/day) (10 kg) = 0>03 /L {30 /L)
(100) (10) (1 L/day)
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1,3-Dichloropropene August, 1987
-9-
where:
3.0 mg/kg/day = NOAEL based on the absence of increased relative kidney
weights 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.
10 = modifying factor, selected since this was the only
useful feeding study available and the study design was
not ideal for assessing exposure via drinking water.
1 L/day = assumed daily water consumption of a child.
For a 70-kg adult:
Longer-term HA - (3.0 mg/kg/day) (70 kg) = .105 ng/L (105 ug/L)
y (100) (10) (2 L/day)
where:
3.0 mg/kg/day = NOAEL based on the absence of increased relative kidney
weights 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.
10 = modifying factor, selected since this was the only
useful feeding study available and the study design was
not ideal for assessing exposure via drinking water.
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 wter 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
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1, 3-Dichloropropene August, 1987
-10-
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, then caution should be exercised in assessing the
risks associated with lifetime exposure to this chemical. For Group C
carcinogens, an additional safety factor of 10 is added to the CWEL.
The Lifetime HA for a 70-kg adult has been determined on the basis of
the study in rats by Til et al. (1973), as described above.
Using the NOAEL of 3.0 mg/kg/day, as determined in that study, the
DWEL is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (3.0 mg/kg/day) = Q.0003 mg/kg/day
(1,000) (10)
where:
3.0 mg/kg/day = NOAEL based on the absence of increased relative kidney
weights 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.
10 = modifying factor selected since this was the only useful
feeding study available and the study design was not
ideal for assessing exposure via drinking water.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0003 mg/kg/day) (70 kg) = .011 mg/L (11 ug/L)
(2 L/day)
where:
0.0003 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 HAs are not recommended for Group A or B carcinogens. DCP is
a Group B, probable human carcinogen. The estimated cancer risk associated
with lifetime exposure to drinking water containing DCP at 11 ug/L is
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1,3-Dichloropropene August, 1987
-11-
approximately 5.5 x 10-5. This estimate represents the upper 95% confidence
limit using the linearized multistage model. The actual risk is unlikely to
exceed this value.
Evaluation of Carcinogenic Potential
0 DCP may be classified as a B2, probable human carcinogen based on
sufficient evidence of tumor production in two rodent species and two
routes of administration.
0 Data on an increased incidence of squamous cell papilloma or carcinoma
of the forestomach in rats exposed to DCP (NCI, 1985) were used for a
quantitative assessment of cancer risk due to DCP. Based on the data
from this study and using the linearized multistage model, a carcinogenic
potency factor (q^) for humans of 1.75 x 10~1 (mg/kg/day)~1 was
calculated.
0 The drinking water concentrations corresponding to increased lifetime
cancer risks of 10~4, 10~5 and 10-6 (One excess cancer per one million
population) for a 70-kg adult consuming 2 L/day are 20 ug/L, 2 ug/L
and 0.2 ug/L, respectively.
0 The forestomach tumor data in male rats used to calculate the qi*
value (NCI, 1985) consisted of the 2-year study data excluding the
ancillary studies data. The ancilliary studies involved serial
sacrifice of animals (at 9, 16, 21, 24 and 27 months). It is not
appropriate to include these data in the lifetime predictive model
used (multistage).
0 For comparison purposes, drinking water concentrations associated
with an excess risk of 10~6 were 0.2 ug/L, 3.6 mg/L, 0.03 ug/L and
0.004 ug/L for the one-hit, Weibull, probit and logit models,
respectively.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The ACGIH recommended 1 ppm (5 mg/m3) as a Threshold Limit Value for
DCP (Patty, 1981).
VII. ANALYTICAL METHODS
0 No specific methods have been published by U.S. EPA for analysis of
DCP in water. However, EPA Method 524.2 (U.S. EPA, 1986d) and EPA
Method 502.2 (USEPA, 1986e) both for volatile organic compounds in
water should be suitable for analysis of DCP. Both are standard
purge and trap capillary column gas chromatographic techniques.
VIII. TREATMENT TECHNOLOGIES
0 There are no specific publications on treatment of 1,3-DCP. However,
adequate treatment by granular activated carbon (GAC) should be
possible. Freundlich carbon absorption isotherms for DCP indicate
reasonably high adsorption capacity (U.S. EPA, 1980).
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1,3-Dichloropropene August, 1987
-12-
IX. REFERENCES
Brooks, T.M., B.J. Dean, A.S. Wright et al.* 1978. Toxicity studies with
dichloropropenes: mutation studies with 1,3-D and cis-1,3-dichloropropene
and the influence of glutathione on the mutagenicity of cis-1,3-dichloro-
propene in Salmonella typhimurium; Group research report (Shell Research,
Ltd.) TLGR.0081 78. Unpublished study by Shell Chemical Co., Washington,
DC. NRID 61059.
Clark, D., D. Blair and S. Cassidy.* 1980. A 10 week inhalation study of
mating behavior, fertility and toxicity in male and female rats; Group
research report (Shell Research, Ltd.) TLGR.80.023. Unpublished study
Dow Chemical U.S.A., Midland, MI. MRIDs 117055, 103280, 39691.
Climie, I.J.G., and B.J. Morrison.* 1978. Metabolism studies on (Z)1,3-dichloro-
propene in the rat: Group research report (Shell Research, Ltd.) TLGR.0101.
78. Unpublished study by Dow Chemical U.S.A., Midland, MI. MRID 32984.
Coate, W.B., D.L. Keenan, R.J. Hardy and R.W. Voelker.* 1979. Inhalation-
toxicity study in rats and mice: Telone II: Project No. 174-127.
Final report. Unpublished study by Hazleton Laboratories America, Inc.,
for Dow Chemical U.S.A., Midland, MI. MRID 119191.
DeLorenzo, F., S. Degl Innocenti and A. Ruocco." 1975. Mutagenicity of
pesticides containing 1,3-dichloropropene: University of Naples, Italy.
Submitted by Dow Chemical U.S.A., Midland, MI. MRID 119179.
Dietz, F.K., E.A. Hermann and J.C. Ramsey. 1984. The pharmacokinetics of
14C-1,3-dichloropropene in rats and mice following oral administration.
Toxicologist. 4:585 (Abstract no.).
Dow Chemical U.S.A.* 1977. Telone II soil fumigant: Product chemistry.
MRID 00119178.
Dow Chemical U.S.A.* 1978. Summary of human safety data. Summary of studies
099515-1 and 099515-J. Unpublished study Dow Chemical U.S.A., Midland, MI.
MRID 39676.
Dow Chemical U.S.A. 1982. A data sheet giving the chemical and physical
properties of the chemical. A complete statement of the names and
percentages by weight of each active inert ingredient in the formulation
to be shipped. Dow Chemical U.S.A., Midlano, MI. MRID 115213.
Flessel, P.* 1977. Letter dated Apr. 8, 1977: Subject: Mutagen testing
program, mutagenic activity of Telone II in the Ames Salmonella assay.
Prepared by Calif. Dept. Health, submitted by Dow Chemical U.S.A.,
Midland, MI. MRIDs 120906, 67534.
Gosselin, R.E., H.C. Hodge, R.P. Smith and M.N. Gleason. 1976. Clinical
toxicology of commercial products. 4th ed. Baltimore, MD: The Williams
and Wilkins Co., p. 120.
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1,3-Dichloropropene August, 1987
-13-
Hanley, T.R., J.A. John-Greene, J.T. Young, L.L. Calhoun and K.S. Rao. 1987.
Evaluation of the effects of inhalation exposure to 1,3-dichloropropene
on fetal development in rats and rabbits. Fundamental and Applied
Toxicology. 8:562-570.
Hutson, D.H., J.A. Moss and B.A. Pickering.* 1971. The excretion and retention
of components of the soil fumigant D-D and their metabolites in the rat.
Food Cosmet. Toxicol. 9:677-680. Dow Chemical U.S.A., Midland, MI.
MRID 39690.
Laskowski, D., C. Goring, P. McCall and R. Swan. 1982. Terrestrial environment.
Environ. Risk Anal. Chem. 25:198-240.
Maddy, K., H. Fong, J. Lowe, D. Conrad and A. Fredrickson. 1982. A study
of well water in selected California communities for residues of
1,3 dichloropropene, chloroallyl alchohol, and 49 organophosphate or
chlorinated hydrocarbon pesticides. Bull. Environ. Contain. Toxicol.
29:354-359.
Neudecker, T., A. Stefani and D. Heschler. 1977. In vivo mutagenicity of
soil nematocide 1,3-dichloropropene. Experientia. 33:1084-1085.
NTP. 1985. National Toxicology Program. NTP Technical report on the toxi-
cology and carcinogenesis studies of Telone II in F344/N rats and B6C3F^
mice (gavage studies). NTP TR 269, NIH Pub. No. 85-2525, May, 1985.
Patty. 1981. Patty's Industrial hygiene and toxicology. 3rd ed., New York,
NY: Wiley-Interscience Co. Vol. 2B, pp. 3573-3577.
Shirasu, Y., M. Moriga and K. Kato.* 1976. Mutagenicity testing on D-D in
microbial systems. Prepared by Institute of Environmental Toxicology,
submitted by Shell Chemical Co., Washington, DC. MRID 61050.
Stolzenberg, S. and C. Mine. 1980. Mutagenicity of 2- and 3-carbon halo-
genated compounds in Salmonella/mammalian microsome test. Environmental
Mutagenesis. 2:59-66.
Sudo, S., M. Nakazawa and M. Nakazono.* 1978. The mutagenicity test on
1,3-dichloropropene in bacteria test systems. Prepared by Nomura Sogo
Research Institute, submitted by Dow Chemical U.S.A., Midland, MI.
MRID 39688.
Til, H.P., M.T. Spankers, V.J. Feron and P.J. Reuzel. 1973.* Subchronic
(90-day) toxicity study with Telone in albino rats: Report No. R4002.
Final report. Unpublished study (Central Institute for Nutrition
and Food Research) submitted by Dow Chemical U.S.A., Midland, MI.
MRIDs 39684, 67977.
Torkelson, T.R., and F. Oyen. 1977.* The toxicity of 1,3-dichloropropene is
determined by repeated exposure of laboratory animals. American Industrial
Hygiene Association Journal. 38:217-223. Dow Chemical U.S.A., Midland, MI.
MRID 39686.
-------
1,3-Dichloropropene August, 1987
-14-
Toyoshima, S., R. Sato and S. Sato. 1978. The acute toxicity test on
Telone II in mice. Unpublished study by Dow Chemical U.S.A., Midland, MI.
MRID 39683.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Carbon adsorption
isotherms for toxic organics. EPA-60018-80-023. Apr. 1980.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. 1,3-Dichloropropene,
a digest of biological and economic benefits and regulatory implications.
Benefits and Use Division, Office of Pesticide Programs.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. 1,3-Dichloropropene;
initiation of special review; availability of registration standard;
notice. Fed. Reg. 51(195):36161. October 8, 1986.
U.S. EPA. 1986c<> U.S. Environmental Protection Agency. Guidance for the
reregistration of pesticide products containing 1,3-dichloropropene as
the active ingredient. Office of Pesticides and Toxic Substances,
Washington, DC. September 1986, 111 pp.
U.S. EPA. 1986d. U.S. Environmental Protection Agency. Volatile organic
compounds in water by purge and trap capillary gas chromatography/mass
spectrometry. Office of Drinking Water, Washington, DC. Aug. 1986.
U.S. EPA. 1986e. U.S. Environmental Protection Agency. Volatile organic
compounds in water by purge and trap capillary column gas chromatography
with photoionization and electrolytic conductivity detectors in series.
Office of Drinking Water, Washington, DC.
Van Dijk, H. 1974. Degradation of 1,3-dichloropropenes in soil. Agro-
Ecosystems. 1:193-204.
Van Duuren, B.L., B.M. Goldschmidt and G. Loewengart.* 1979. Carcinogenicity
of halogenated olefinic and aliphatic hydrocarbons in mice. Journal of
the National Cancer Institute. 63(6):1433-1439. MRID 94723.
Venable, J.R., C.D McClimans, R.E. Flake and D.B. Demick.* 1980. A fertility
study of male employees engaged in the manufacture of glycerine. Journal
of Occupational Medicine. 22(2):87-91. Dow Chemical U.S.A., Midland,
MI: MRID 117052.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
DICAMBA
Health Advisory
Office of Drinking Water
U.S. Bwironmental Protection Agency
DRAFT
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 mechani ms 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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Dicamba
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-00-9
Structural Formula
3,6-Dichloro-2-methoxy-benzoic acid
Synonyms
0 Banes, Banex, Banlen, Banuel D, Banvel, Brush buster, Dianat, Dianate,
Dicambe, Mediben, Mondak, MDBA, Velsicol Compound R
Uses
0 Herbicide used to control broadleaf weeds in field and silage corn,
grain sorghum, small grains, asparagus, grass seed crops, turf,
pasture, rangeland, and non-cropland areas such as fence rows,
roadways and wastelands. For control of brush and vines in non-
cropland, pasture and rangeland areas (Meister, 1983).
Properties (Berg, 1986; CHEMLAB, 1985; Meister, 1983; Windholz et al., 1983;
Worthing, 1983)
Chemical Formula
Molecular weight
Physical State (at 25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (20°C)
Specific Gravity
Water SolubilJ ty (20°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
CQH6C1203
221.04
Crystals
114 to 116°C
3.75 x 10-3 mm Hg
6,500 mg/L at 25°C
3.67 (calculated)
Occurrence
Dicamba has been found in 249 of 624 surface water samples analyzed
and in 39 of 275 ground water samples (STORET, 1987). Samples were
collected at 148 surface water locations and 229 ground water locations;
dicamba was found in 12 states. The 85th percentile of all non-zero
samples was 0.15 ug/L in surface water and 0.07 ug/L in ground water.
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Dicamba August, 1987
-3-
The maximum concentration found in surface water was 3.3 ug/L, while
in ground water it was 0.8 ug/L.
Environmental Fate
0 In several aerobic soil metabolism studies, dicamba (acid or salt form
not specified) had half-lives of 1 to 6 weeks in sandy loam, heavy
clay, silty clay, clay loam, sand and silt loam soils at 18 to 38°C
and 40 to 100% of field capacity. Degradation rates decreased with
decreasing temperature and soil moisture (Smith, 1973a,b; Smith, 1974;
Smith and Cullimore, 1975; Suzuki, 1978;1979).
0 For the dimethylamine salt, half-lives in sandy loam and loam soils
ranged from 17 to 32 days (Altom and Stritzke, 1973). Phytotoxic
residues, detected by a non-specific bioassay method, have persisted
in aerobic soil for almost 2 years (Sheets, 1964; Sheets et al.,
1968).
0 Based on soil thin-layer chromatography (TLC), dicamba (acid or salt
form not specified) is highly mobile in sandy loam, silt loam, sandy
clay loam, clay loam, loam, silty clay loam and silty clay soils
(Helling, 1971; Helling and Turner, 1968).
0 The free acid of dicamba and the dimethylamine salt were not appre-
ciably adsorbed to any of five soils ranging from heavy clay to loamy
sand (Grover and Smith, 1974). The dicamba degradation product,
3,6-dichlorosalicylic acid, adsorbed to sandy loam (30%), clay and
silty clay (55%) (Smith, 1973a,b; Smith and Cullimore, 1975).
0 Losses of 12 to 19% of the applied radioactivity from nonsterile soils
indicated that metabolism contributes substantially more to 14C-dicamba
losses than does volatilization (Burnside and Levy, 1965; 1966).
0 Under field conditions, dicamba (acid or salt form not specified) had
half-lives of 1 to 2 weeks in a clay and a sandy loam soil when applied
at 0.27 and 0.53 Ib/A. At either application rate, less than 30 ppb
of dicamba remained after 4 weeks (Scifres and Allen, 1973). In
another study, using a nonspecific bioassay method of analysis,
dicamba phytotoxic residues dissipated within 2 years in loam and
silty clay loam (Burnside et al., 1971).
0 Ditchbank field studies indicated vertical movement of dicamba in
soil; the soil layers at 6 to 12 inches contained a maximum of 0.07 ppm
and 0.28 ppm in canals treated at 0.66 and 1.25 Ib/A, respectively
(Salman et al., 1972).
III. PHARMACOKINETICS
Absorption
0 Atallah and Yu (1980) reported that mice, rats, rabbits and dogs
administered single oral doses of 14c-dicamba (99% purity,
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Dicamba August, 1987
-4-
approximately). 100 ing/kg) excreted an average of 85% of the admini-
stered dose in urine in the 48 hours after dosing.
0 Similar findings were reported for rats by Tye and Engel (1967) (96%
excreted in 24 hours) and by Whitacre and Diaz (1976) (83% excreted
in 24 hours). The data indicate that dicamba is rapidly absorbed
from the gastrointestinal tract.
Distribution
0 The retention of dicamba (99% purity, approximately 100 mg/kg) was
investigated in rats, mice, rabbits and dogs following single doses
by oral intubation (Atjallah and Yu, 1980). Tissue levels 16 hours
after treatment were low. Tye and Engel (1967) also found low residue
levels of dicamba in kidneys, liver and blood. The data indicate
that dicamba does not accumulate in mammalian tissues.
Metabolism
0 The metabolism of 14c-dicamba (99% purity) was investigated in mice,
rats, rabbits and dogs after administration of single oral doses at
approximately 100 mg/kg (Atallah and Yu, 1980). Between 97 to 99%
of the dicamba was recovered unchanged in the urine of all four
species. 3,6-Dichloro-2-hydroxybenzoic acid (DCHBA, a metabolite)
was not detected in any urine sample at a level greater than 1% of
the dose. There wa= also a small amount of unknown metabolites
totaling about 1%.
Excretion
Atallah and Yu (1980) investigated the excretion of 14C-dicamba (99%
purity) after a single oral dose (approximately 100 mg/kg) in mice,
rats, dogs and rabbits, and reported that 67 to 93% of the adaiinistered
dose was excreted in urine of the four species within 16 hours. The
compound was found to a lesser degree in feces (0.5 to 5.7%) and
various tissues (0.17 to 0.5%) 16 hours postdosing.
IV. HEALTH EFFECTS
Humans
The Pesticide Incident Monitoring System data base revealed 10
incident reports involving humans from 1966 to March 1981 for
dicamba alone (U.S. EPA, 1981). Six of the ten reported incidents
involved spraying operations. No concentrations were specified.
Exposed workers developed muscle cramps, dyspnea, nausea, vomiting,
skin rashes, loss of voice or swelling of cervical glands. Four
additional incidences resulted in coughing and dizziness in one child
involved in an undescribed agricultural incident. Three children who
sucked mint leaves from a ditch bank previously sprayed with dicamba
were asymptomatic.
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Dicamba Au*ust' 198?
-5-
Animals
Short-term Exposure
0 Reported acute oral LD^Q values for technical dicamba [85.8% active
ingredient (a.i.)J range from 757 to 1,414 mg/kg (Witherup et al.,
1962) in rats. The acute oral LD50 in mice has been reported to be
>4,640 mg/Jcg (Kettering Laboratory, 1962) and 316 mg/Jcg in hens
(Roberts et al., 1983).
0 An acute inhalation LC50 of >200 mg/L was reported in rats (IRDC, 1973).
0 The neurotoxic effects of dicamba in hens were studied by Roberts
et al. (1983). Technical dicamba (86.2% a.i.) was administered per os
(10 hens/dose) in doses of 0, 79, 158 or 316 mg/kg. Two groups of
ten hens each were dosed at 316 mg/kg. The various groups were
observed for 21 days following treatment. No signs of ataxia were
observed at any dose level tested. Histopathological evaluation of
nervous tissue from 13 hens treated at 316 mg/kg demonstrated
sciatic nerve damage in 6 hens (46%). The authors attributed this
alteration to prolonged recumbency rather than a direct effect of
dicamba. Based on the absence of delayed neurotoxicity and sciatic
nerve damage, a NOAEL of 158 mg/kg is identified for this study.
0 Rats (two/sex/dose) of the CD strain were fed diets containing 658
or 23,500 ppm of technical dicamba (85.8% a.i.) for up to three weeks
(Witherup et al., 1962). Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to about 32.9 or 1,175 mg/kg/day. No adverse effects on physical
appearance, behavior, food consumption, body or organ weights, gross
pathology or histopathology were reported. Based on this information,
a NOAEL of 1,175 mg/kg/day (the highest dose tested) is identified.
Dermal/Ocular Effects
0 IRDC (1974) reported an acute LD50 of >2000 mg/kg in rabbit dermal
studies.
0 Heenehan et al. (1978) studied the sensitization potential of technical
dicamba (86.8% a.i.) in albino guinea pigs. The compound was applied
as a 10% suspension to the shaved backs of guinea pigs (five/sex) for
6 hours three times per week for 3 weeks. Following nine sensitizing
doses, two challenge doses were applied. Dicamba was ]udged to cause
moderate dermal sensitization.
0 Technical dicamba (86.8% a.i.) was applied to the shaved backs of
New Zealand White rabbits (four/sex/dose) in doses of 0, 100, 500 or
2,500 mg/kg/day, 5 days per week for 3 weeks (IRDC, 1979). Slight
skin irritation was observed at 100 mg/kg, and moderate irritation at
500 mg/kg/day and above. No changes were observed in general appearance,
behavior, body weight, organ weight, biochemistry, hematology or
urinalysis.
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Dicamba August, 1987
-6-
0 Thompson (1984) instilled single doses (0.1 g) of technical dicamba
(purity not specified) into the conjunctival sacs of nine New Zealand
rabbits; three eyes were washed and six were not washed. Dicamba was
severely irritating and corrosive to both washed and unwashed eyes.
Long-term Exposure
0 Laveglia et al. (1981) fed CD rats (20/sex/dose) technical dicamba
(86.8% a.i.) in the diet for 13 weeks in doses of 0, 1,000, 5,000 or
10,000 ppm. Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of about
0, 50, 250 or 500 mg/kg/day. No compound-related effects were observed
in general appearance, hematology, biochemistry or in urinalysis values,
survival and gross pathology at any dose levels tested. There was an
absence or reduction of cytoplasmic vacuolation of hepatocytes and a
decrease in mean body weight for both sexes (6.3% in females and 7.5%
in males) at 10,000 ppm (500 mg/kg/day). The body weight decrease
was lower (p <0.05) at week 13 when compared to controls. A NOAEL of
5,000 ppm (250 mg/kg/day) can be identified for this study.
0 Male Wistar rats (20/dose) were fed diets containing technical dicamba
at 0, 31.6, 100, 316, 1,000 or 3,162 ppm for 15 weeks (Edson and Sand-
erson, 1965). Assuming that 1 ppm in the diet of rats is equivalent
to 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of about
0, 1.6, 5, 15.8, 50 or 158 mg/kg/day. Following treatment, general
behavior, physical appearance, food consumption, organ weights, gross
pathology and histopathology were evaluated. However, the authors
presented data only for the evaluation of body and organ weights.
Hematological, urinalysis or clinical chemistry studies were not
reported. No adverse effects were observed in the parameters measured
at 316 ppm (15.8 mg/kg/day) or less. Relative liver-to-body weight
ratios increased (p value not specified) in at 1,000 and 3,162 ppm
(50 and 158 mg/kg/day). Based on these data, the authors identified
a NOAEL of 316 ppm (15.8 mg/kg/day).
0 Davis et al. (1962) fed beagle dogs (three/sex/dose) technical dicamba
(90% a.i.) in the diet in doses of 0, 5, 25 or 50 ppm for 2 years.
Assuming that 1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day,
(Lehman, 1959), this corresponds to doses of about 0, 0.125, 0.625 or
1.25 mg/kg/day. No compound-related effects were observed on survival,
food consumption, hematology, urinalysis and organ weights. A decrease
in body weight was observed in males at 25 and 50 ppm and in females
at 50 ppm. No individual data except for body weight were reported,
and no statistical evaluations were made. The authors did not present
data on gross pathology. Histopathology was done only on the heart,
lung, liver and kidney. Based on marginal information, a NOAEL of
5 ppm (0.125 mg/kg/day) can be identified.
0 Sprague-Dawley rats (32/sex/dose) were fed technical dicamba (90% a.i.)
in the diet for 2 years in doses of 0, 5, 50, 100, 250 or 500 ppm
(Davis et al., 1962). Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 0.25, 2.5, 5, 12.5 or 25 mg/kg/day. The authors
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Dicamba August, 1987
-7-
reported no adverse effects upon survival, body weight, food consump-
tion, organ weight, hematologic values or histology at the dose
levels tested. No data were presented for evaluation of pharmacologic
effects, gross pathology, urinalysis or clinical chemistry. Incomplete
histological data were presented. A NOAEL could not be determined for
this study due to insufficient data.
Reproductive Effects
0 Charles River CD rats (20 females or 10 males/dose) were fed diets
containing technical dicamba (87.2% a.i.) in doses of 0, 5, 50, 100,
250 or 500 ppm through three generations (Kettering Laboratory,
1966). Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of about 0,
0.25, 2.5, 5, 12.5 or 25 mg/kg/day. Fertility index, gestation
index, viability index, lactation index and pup development were
comparable in treated and control rats. A NOAEL of 500 ppm
(25 mg/kg/day) was identified.
Developmental Effects
0 Technical dicamba (87.7% a.i.) was administered per os to pregnant New
Zealand White rabbits (23-27/dose) at doses of 0, 1, 3 or 10 mg/kg/day
from days 6 through 18 of gestation (IRDC, 1978). No maternal toxicity,
fetotoxicity or teratogenic effects were observed at 1 and 3 mg/kg/day.
There were slightly reduced fetal and maternal body weights and
increased postimplantation losses in the 10 mg/kg/day dose group when
compared to untreated controls. The author did not consider these
differences to be statistically significant. The author identified
a developmental toxicity NOAEL of 10 mg/kg/day (the highest dose
tested). Based on a reduction in body weights and increased post-
implantation losses at the highest dose, a maternal and fetotoxic
NOAEL of 3 mg/kg/day was identified by EPA/OPP.
0 Pregnant albino rats (20-24/dose) were administered technical-grade
dicamba by gavage at dose levels of 0, 64, 160 or 400 mg/kg/day on
days 6 throug.it 19 of gestation (Toxi Genetics, 1981). No maternal
toxicity was observed up to 160 mg/kg/day. Dicamba-treated dams in
the 400-mg/kg/day dosage group exhibited ataxia and reduced body
weight gain; they consumed less food during the dosing period when
compared with controls given vehicle alone (p <0.05). No fetotcucity
or developmental effects were observed at the dose levels tested.
Based on these findings, a NOAEL for maternal toxicity of 160 mg/kg/day
is identified. The NOAEL for fetotoxic and developmental effects is
400 mg/kg/day (the highest dose tested).
Mutagenicity
0 Moriya et al. (1983) reported that dicamba (up to 5,000 ug/plate)
exhibited no mutagenic activity against Salmonella typhimunum
(TA 98, TA 100, TA 1535, TA 1537 and TA 1538) or Escherichia coll
(WF2 her) either with or without metabolic activation.
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Dicair.ba August, 1987
-8-
0 An increased number of chromosomal aberrations (p <0.01) were reported
in mouse bone marrow cells exposed to 500 mg/kg dicamba (Kurinnyi
et al., 1982). No other details were presented.
Carcinogenicity
0 Sprague-Dawley rats (32/sex/dose) were administered dicamba (90% a.i.)
in the diet for two years at doses of 0, 5, 50, 100, 250 or 500 ppm
(Davis et al., 1962). Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of about 0, 0.25, 2.5, 5, 12.5 or 25 mg/kg/day. The treated
rats did not differ from the untreated control animals with respect
to the incidence, types and time of appearance 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) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/)cg 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
No information was found in the available literature that was suitable
for determination of the One-day HA value for dicamba. Accordingly, it is
recommended that the Ten-day HA value of 0.3 mg/L (calculated below) for a
10 kg child be used at this time as a conservative estimate of the One-day HA.
Ten-day Health Advisory
The developmental toxicity study by IRDC (1978) has been selected to
serve as the basis for the Ten-day HA value for dicamba. In this study,
pregnant rabbits administered technical dicamba (87.7 % a.i.) by gastric
intubation at dosage levels of 1, 3 or 10 mg/kg/day from days 6 through 18
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Dicamba August, 1987
-9-
of gestation showed slightly reduced maternal body weights at 10 mg/kg/day.
Similarly, fetal body weights were slightly reduced, and postimplantation
losses were increased in the 10-mg/Jcg/day dose group.
Based on these data, a maternal and fetal toxicity NOAEL of 3 mg/kg/day
is identified. A rat study (Toxi Genetics, 1981) of comparable duration deter-
mined higher maternal and fetal NOAELs (160 and 400 mg/kg/day, respectively).
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA = (3 mg/kg/day) (10 kg) = 0.3 mg/L (300 ug/L)
(1 L/day) (100)
where:
3 mg/kg/day = NOAEL, based on absence of body weight loss and post-
implantation losses.
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
No studies found in the available literature were suitable for
determining a Longer-term HA value for dicamba. One 13-week rat study (Laveglia
et al., 1981) and one 15-week rat study (Edson and Sanderson, 1965) reported
NOAELs (250 mg/kg/day and 15.8 mg/kg/day, respectively) that were higher than
the NOAEL (3 mg/kg/day) of the rabbit study (IRDC, 1978) selected to derive
the Ten-day HA value. It is therefore recommended that the Reference Dose
(RfD) derived below in the calculation of the Lifetime HA (0.0013 mg/kg/day)
be used at this time as the basis for the Longer-term HA values. As a result,
the Longer-term HA is 13 ug/L for the 10-kg child and is 50 ug/L for the
70-kg 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.
The DWEL is derived from the multiplication of the RfD by the assumed body
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Dicamba August, 1987
-10-
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 2-year dog study by Davis et al. (1962) has been selected to serve
as the basis for deriving the Lifetime HA for dicamba. In this study, beagle
dogs were administered technical dicamba at dietary levels of 0, 5, 25 or
50 ppm (Or 0.125, 0.625 or 1.25 mg/kg/day). A decrease in body weight was
observed in males at 25 and 50 ppm and in females at 50 ppm. A NOAEL of
25 ppm (0.125 mg/kg/day) was identified.
The Lifetime HA is derived from this NOAEL as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = °'125 "q/kq/day _ Q.0013 mg/kg/day
(100)
where:
0.125 mg/kg/day = NOAEL based on the absence of body weight loss.
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 = (0.0013 mg/kg/day) (70 kg) = 0.046 mg/L (46 ug/L)
(2 L/day)
where:
0.0013 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 = (0.046 mg/L) (20%) = 0.009 mg/L (9 ug/L)
where:
0.046 mg/L = DWEL.
20% = assumed relative source contribution from water.
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Dicamba August, 1987
-1 1-
Evaluation of Carcinogenic Potential
8 One study on the carcinogenic!ty of dicamba in rats has been reported;
it revealed no evidence of carcinogenicity (Davis et al., 1962).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenicity of dicamba.
0 Applying the criteria described in EPA'3 guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), dicamba is classified in Group D:
not classified. This category is used for substances with inadequate
evidence of carcinogenicity in animal studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The MAS (1977) has calculated an ADI of 0.00125 mg/kg/day based on a
NOAEL of 1.25 mg/kg/day from a 2-year feeding study in dogs and an
uncertainty factor of 1,000. Assuming a body weight of 70 kg and a
20% source contribution factor,' they calculated a Suggested-No-Adverse-
Reaction-Level (SNARL) of 0.009 mg/L.
0 Residue tolerances from 0.05 to 40 ppm have been established for a
variety of agricultural products (U.S. EPA, 1985a).
VII. ANALYTICAL METHODS
0 Analysis of dicamba is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated acid pesticides in water
samples (U.S. EPA, 1985b). In this method, approximately 1 L of
sample is acidified. The compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolized with
potassium hydroxide, and extraneous organic material is removed by
a solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane as the derivatizing
agent. Excess reagent is removed, and the esters are determined by
electron capture (EC) GC. The method detection limit for dicamba has
been estimated to be 0.27 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate granular-activated carbon (GAC) adsorption
to be a possible removal technique for dicamba.
0 Whittaker et al. (1982) report that a reduction of pH from 7 to 3
increased the extent of dicamba GAC adsorption. No system performance
was reported.
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Dicamba August, 1987
-12-
IX. REFERENCES
Atallah, Y.H., and C.C. Yu.* 1980. Comparative pharmacokinetics and
metabolism of dicamba in mice, rats, rabbits and dogs. MRID 00128088.
Altom, J.D., and J.R. Stritzke. 1973. Degradation of dicamba, picloram, and
four phenoxy herbicides in soils. Weed Sci. 21:556-560.
Berg, G.L. 1986. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
Burnside, O.C., and T.L. Levy. 1965. Dissipation of dicamba. Unpublished
study prepared by the University of Nebraska, Department of Agronomy,
submitted by Velsicol Chemical Corporation, Chicago, 111.
Burnside, O.C., and T.L. Levy. 1966. Dissipation of dicamba. Weeds
14:211-214.
Burnside, O.C., G.A. Wicks and C.R. Fenster. 1971. Dissipation of dicamba,
picloram, and 2,3,6-TBA across Nebraska. Weed Sci. 19:323-325.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc.
Davis, R.K., W.P. Jolly, K.L. Stemmer et al.* 1962. The feeding for two
years of the herbicide 2-methoxy-3,6-dichlorobenzoic acid to rats and
dogs. MRID 00028248.
Edson, E.F., and D.M. Sanderson. 1965. Toxicity of the herbicides 2-
methoxy-3,6-dichlorobenzoic acid (Dicamba) and 2-methoxy-3,5,6-tri-
chlorobenzoic acid (tricamba). Food Cosmet. Toxicol. 3:299-304.
Grover, R., and A.E. Smith. 1974. Adsorption studies with the acid and
dimethylamine forms of 2,4-D and dicamba. Can. J. Soil Sci. 54:179-186.
Heenehan, P.R., W.E. Rinehart and W.G. Brun.* 1978. A dermal sensitization
study in guinea pigs. Compounds: Banvel 45, Banvel technical:
Project No. 5055-78. MRID 00023691.
Helling, C.S. 1971. Pesticide mobility in soils: II. Applications of soil
thin-layer chromatography. Soil Sci. Soc. Amer. Proc. 35:737-748.
Helling, C.S., and B.C. Turner. 1968. Pesticide mobility: Determination of
soil thin-layer chromatography. Science. 162:562-563.
IRDC.* 1973. International Research and Development Corporation. Acute
inhalation exposure in the male albino rats. Report No. 163-191.
MRID 00028234.
IRDC.* 1974. International Research and Development Corporation. I. Acute
toxicity studies in rats and rabbits. Report No. 163-295. MRID 00025372.
IRDC.* 1978. International Research and Development Corporation. Teratology
study in rabbits. Report No. 163-436. MRID 00025373.
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Dicamba ' 198?
-13-
IRDC.* 1979. International Research and Development Corporation. Three-
week dermal toxicity study in rabbits. Report No. 163-620. MRID 00128090.
Kettering Laboratory.* 1966. The effects exerted upon the fertility of rats
and upon the viability of their offspring by the introduction of Banvel
D into their diets. MRID 00028249.
Kettering Laboratory.* 1962. The cumulative toxicity of 2-methoxy-3,6-
dichlorobenzoic acid (Banvel D) and 2-methoxy-3,5,6-trichlorobenzoic
acid (Banvel T) when fed to rats. MRID 00022503.
Kurinnyi, A.I.. M.A. Pilinskaya, I.V. German and T.S. L'voya. 1982. Imple-
mentation of a program of cytogenetic study of pesticides: Preliminary
evaluation of cytogenetic activity and potential mutagenic hazard of 24
pesticides. Tsitol. Genet. 16:45-49.
Laveglia, J., D. Rajasekaran, L. Brewar.* 1981. Thirteen week dieting
toxicity study in rats with dicamba. IRDC No. 163-671. MRID 00128093.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Assoc. Food Drug Off. U.S.
Meister, R., ed. 1983. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
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. 1977. National Academy of Sciences. Drinking Water and Health.
Washington, DC: National Academy of Science Press.
Roberts, N., C. Fairley, C. Fish et al.* 1983. The acute oral toxicity
(LD5o) and neurotoxicity effects of dicamba in the domestic hen. HRC
Report No. 24/8355. MRID 00131290.
Salman, H.A., T.R. Hartley and A.R. Hattrup. 1972. Progress report of
residue studies on dicamba for ditchbank weed control. U.S. Department
of the Interior, Bureau of Reclamation, Applied Sciences Branch, Division
of General Research, Engineering and Research Center. USDI, Br. Report
No. REC-ERC-72-6; available from National Technical Information Center,
Springfield, VA. 22151.
Scifres, C.F., and T.J. Allen. 1973. Dissipation of dicamba from grassland
soils of Texas. Weed Sci. 21:393-396.
Sheets T.J. 1964. Letter sent to Warren H. Zick dated Jan.3, 1964. Greenhouse
persistence study with dicamba and tricamba. U.S. Agricultural Research
Service, Crops Research Division, Crops Protection Research Branch,
Pesticide Investigations—Behavior in soils; unpublished study.
Sheets, T.J., J.W. Smith and D.D. Kaufman. 1968. Persistence of benzoic
and phenylacetic acids in soils. Weed Sci. 16:217-222.
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Dicamba August, 1987
-14-
Smith, A.E. 1973a. Degradation of dicamba in prairie soils. Weed Res.
13:373-378.
Smith, AoE. 1973b. Transformation of dicamba in Regina heavy clay.
J. Agric. Food Chem. 21:708-710.
Smith, A.E. 1974. Breakdown of the herbicide dicamba and its degradation
products 3,6-dichlorosalicylic acid in prairie soils. J. Agric. Food Chem.
22:601-605.
Smith, A.E., and D.R. Cullimore. 1975. Microbiological degradation of the
herbicide dicamba in moist soils at different temperatures. Weed Res.
15:59-62.
STORET. 1987.
Suzuki, H.K. 1978. Dissipation of Banvel and in combination with other
herbicides in two soil types: Report NO. 196. Unpublished study prepared
in cooperation with International Research and Development Corporation,
submitted by Velsicol Chemical Corporation, Chicago, 111.
Suzuki, H.K. 1979. Dissipation of Banvel or Banvel in combination with
other herbicides: Two soil types: Report No. 197. Unpublished study
prepared in cooperation with Craven Laboratories, Inc.; submitted by
Velsicol Chemical Corporation, Chicago, 111.
Thompson, G.* 1984. Primary eye irritation study in albino rabbits with
technical dicamba. Study No. Will 15134. Will Research Laboratories,
inc. MRID 00144232.
Toxi Genetics.* 1981. Teratology study in albino rats with technical dicamba.
Study No. 450-0460. MRID 00084024.
Tye, R., and D. Engel. 1967. Distribution and excretion of dicamba by rats
as determined by radiotracer technique. J. Agric. Food Chem. 15:837-840.
U.S. EPA. 1981. U.S. Environmental Protection Agency. Summary of reported
incidents involving dicamba. Pesticide incident monitoring system.
Report No. 432. Office of Pesticide Programs, Washington, DC.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.227. July 1, 1985.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. U.S. EPA Method 615
- Chlorinated phenoxy acids. Fed. Reg. 50:40701, October 4, 1985.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment Fed. Reg. 51 (185):33992-34003. September 24.
Whitacre, D.M. and L.I. Diaz.* 1976. Metabolism of Hc-dicamba in female
rats. MRID 00025363.
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Dicamba August, 1987
-15-
Whittaker, K.F., J.C. Nye, R.F. Weekash, R.J. Squires, A.C. York and H.A.
Razemier. 1982. Collection and treatment of wastewater generated by
pesticide application. U.S. Environmental Protection Agency.
EPA-600/2-82-028, Office of Environmental Criteria and Assessment,
Cincinnati, Ohio.
Windholz, M., S. Budavari, R.F. Blumetti, E.S. Otterbein, eds. 1983. The
Merck Index — An Encyclopedia of Chemicals and Drugs, 10th ed.
Rahway, NJ: Merck and Company, Inc.
Witherup, S., K.L. Stemmer and H. Schlect.* 1962. The cumulative toxicity
of 2-methoxy-3,6-dichlorobenzoil acid (Banvel D) and.2-methoxy-3, 5,6-
trichlorobenzoil acid (Banvel T) when fed to rats. MRID 00022503.
Worthing, C.R, ed. 1983. The Pesticide Manual: A World Compendium, 7th Ed,
London: BCPC Publishers.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
DIELDRIN
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 car differ by several orders of magnitude.
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Dieldrin August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 60-57-1
Structural Formula
Cl
Cl
Dieldrin; 3, 4, 5, 6, 9, 9-hexachloro-1af 2, 2a, 3, 6,6a, 7, 7a-octahydro-
2,7:3,6-dimethanonaphth[2,3-b]oxirene (Windholz, 1983).
Synonyms
e HEOD; Alvit; Quintox; Octalox (IPCS, 1987).
Uses
0 Formerly used for control of soil insects, public health insects,
termites and many other pests. These uses have been cancelled and
manufacture discontinued in the United States (Meister, 1983).
Properties (NAS, 1977; Weast and As tie, 1982; Windholz, 1983)
Chemical Formula C^HgCigO
Molecular Weight 380.93
Physical State Crystals
Boiling Point
Melting Point 175 to 176°C
Density — .
Vapor pressure (20°C) 3.1 x 1 0~6 mci Hg
Water Solubility (25°C) 0.25 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold (water) 0.04 mg/L
Conversion Factor ~
Occurrence
Dieldrin has been found in 9,809 of 52,453 surface water samples
analyzed and in 217 of 6,042 ground water samples (STORET, 1987).
Samples were collected at 8,831 surface water locations and 4,522
ground water locations, and Dieldrin was found in 48 states, Canada
and Puerto Rico. The 85th percentile of all nonzero samples was
0.01 ug/L in surface water and 0.10 ug/L in ground water sources.
The maximum concentration found was 301 ug/L in surface water and in
10.08 ug/L in ground water.
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Dieldrin August, 1987
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Environmental Fate
0 Dieldrin is stable and highly persistent in the environment.
0 Dieldrin has the longest half-life of the chlorinated hydrocarbons in
water 1-m deep (half-life = 723 days) (MacKay and Wolkoff, 1973).
III. PHARMACOKINETICS
Absorption
0 A single oral dose of dieldrin at 10 mg/kg body weight (bw) administered
in corn oil to male Sprague-Dawley rats produced consistent concentrations
of dieldrin in plasma, muscle, brain, kidney and liver for periods up
to 48 hours suggesting slow absorption of the substance (Hayes, 1974).
Distribution
0 Rats given a single oral dose of dieldrin at 10 mg/kg showed concen-
trations of dieldrin in fat, muscle, liver, blood, brain and kidney.
Hie highest concentration of dieldrin was in fat. The lowest con-
centration was in the kidney (Hayes, 1974).
Metabolism
0 Both the CFE rat and CF1 mouse, following a single oral dose of
dieldrin (not less than 85% HEOD) at 3 and 10 mg/kg in olive oil,
respectively, metabolized dieldrin to 9-hydroxydieldrin, 6,7-trans-
dihydroaldrindiol and some unidentified metabolites. The rat, but
not the mouse, also metabolized dieldrin to pentachloroketone (Baldwin
and Robinson, 1972).
Excretion
Female rats infused with total doses of 8 to 16 mg 36ci-dieldrin/kg bw
excreted approximately 70% of the infused dose in the feces over a
period of 42 days, while only about 10% of the dose was recovered in
the urine. Excretion was markedly increased by restriction of the
diet indicating that the1 concentration of dieldrin in the blood
increased as fat was mobilized (Heath and Vandekar, 1964).
IV. HEALTH EFFECTS
Humans
Dieldrin has been reported to cause hypersensitivity and muscular
fasciculations that may be followed by convulsive seizures and
respective changes in the EEC pattern. Acute symptoms of intoxication
include hyperirritability, convulsions and/or coma sometimes accompanied
by nausea, vomiting and headache, while chronic intoxication may result
in fainting, muscle spasms, tremors and loss of weight. The lethal
dose for humans is estimated to be about 5 g (ACGIH, 1984).
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Dieldrin August, 1987
-4-
Animals
Short-term Exposure
0 RTECS (1985) reported the acute oral LD50 values of dieldrin in the
rat, mouse, dog, monkey, rabbit, pig, guinea pig and hamster as 38.3,
38, 65, 3, 45, 38, 49 and 60 mg/fcg, respectively.
Dermal/Ocular Effects
0 Aldrin or Dieldrin (dry powder) applied to rabbit skin for 2 h/day,
5 days/week had no discernible effects (IPCS, 1987).
Long-term Exposure
0 Groups of Osborne-Mendel rats, 12/sex/level, were fed 0, 0.5, 2, 10,
50, 100 or 150 ppm dieldrin (recrystallized, 100% active ingredient)
in their diet for 2 years. These doses correspond to approximately
0, 0.025, 0.1, 0.5, 2.5, 5.0 or 7.5 mg/kg/day, respectively (Lehman,
1959). Survival was markedly decreased at levels of 50 ppm and
above. Liver-to-body weight ratios were significantly increased at
all treatment levels, with females showing the effect at 0.5 ppm and
males at 10 ppm and greater. Microscopic lesions were described as
being characteristic of chlorinated hydrocarbon exposure. These
changes were minimal at the 0.5 ppm level. Hale rats, at the two
highest dose levels (100 and 150 ppm), developed hemorrhagic and/or
distended urinary bladders usually associated with considerable
nephritis (Fitzhugh et al., 1964). A Lowest-Observed-Adverse-Effect-
Level (LOAEL) of 0.025 mg/kg/day, the lowest dose tested, was identified
in this study.
0 Dogs, one/sex/dose level (two/sex at 0.5 mg/kg/day), fed dieldrin
(recrystallized, 100% active ingredient) at 0.2 to 10 mg/kg/day,
6 days/week for up to 25 months, showed toxic effects including weight
loss and convulsions at dosages of 0.5 mg/kg/day or more. Survival was
inversely proportional to dose level. No toxic effects, gross or
microscopic, were seen at a dose level of 0.2 mg/kg/day (Fitzhugh et
al., 1964).
0 Groups of Carworth Farm "E" strain rats, 25/sex/dose level, were fed
dieldrin (>99% purity) in the diet at 0.0, 0.1, 1.0 or 10.0 ppm for
2 years. These doses correspond to approximately 0, 0.005, 0.05 or
0.5 mg/kg/day, respectively (Lehman, 1959). At 7 months, the 1-ppm
intake level was equivalent to approximately 0.05 and 0.06 mg/kg/day
for males and females, respectively. No effects on mortality, body
weight, food intake, hematology and blood or urine chemistries were
seen. At the 10-ppm level, all animals became irritable after 8 to
13 weeks of treatment and developed tremors and occasional convulsionsc
Liver weight and liver-to-body weight ratios were significantly
increased in females receiving both 1.0 and 10 ppm. Pathological
findings described as organochlorine-insecticide changes of the liver
were found in one male and six females at the 10-ppm level. No
evidence of tumorigenesis was found (Walker et al., 1969).
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Dieldrin August, 1987
-5-
0 Groups of beagle dogs (five/sex/dose) were treated daily by capsule
with dieldrin (>99% purity) at 0.0, O.OOS or 0.05 rag/kg in olive oil
for 2 years. No treatment-related effects were seen in general
health, behavior, body weight or urine chemistry. A significant
increase in plasma alkaline phosphate in both sexes and a significant
decrease in serum protein concentration in males receiving the high
dose were not associated with any clinical or pathological change.
Liver weight and liver-to-body weight ratios were significantly
increased in females receiving the high dose, 0.05 mg/kg/day, but no
gross or microscopic lesions were found. There was no evidence of
tumorigenic activity (Walker et al., 1969).
0 Dieldrin (>99% pure) was administered to CF1 mice of both sexes in
the diet for 128 weeks. Dosages were 1.25, 2.5, 5, 10 or 20 ppm
dieldrin. These doses are equivalent to 0.19, 0.38, 0.75, 1.5 or 3
mg/kg body weight (Lehman, 1959). At the 20-ppm dose level,-approximately
25% of the males and nearly 50% of the females died during the first
3 months of the experiment. Palpable intra-abdominal masses were
detected after 40, 75 or 100 weeks in the 10, 5 and 2.5-ppm-treated
groups, respectively. At 1.25 ppm, liver enlargement was not palpable
and morbidity was similar to that of controls. A No-Observed-Adverse-
Effect-Level (NOAEL) cannot be established because clinical chemistry
parameters were not determined (Walker et al., 1972).
Reproductive Effects
0 Coulston et al. (1980) studied the reproductive effects of dieldrin
in Long Evans rats. Pregnant rats were administered 0 or 4 mg/kg bw
dieldrin by gavage daily from day 15 of gestation through 21 days
postpartum. The treated group did not differ from the control group
when examined for fecundity, number of stillbirths, perinatal mortality
and total litter weights.
Developmental Effects
0 Pregnant Syrian golden hamsters given 30 mg/kg bw dieldrin (^99% pure)
in corn oil on days 7, 8 or 9 of gestation manifested an embryo-
cidal and teratogenic response as evidenced by a statistically
significant increase in fetal deaths, a decrease in live fetal weight
and an increased incidence of webbed foot, cleft palate and open eye
(Ottolenghi et al., 1974). Similar anomalies were observed in
CD] mice administered 15 mg/kg bw dieldrin on day 9 of gestation, but
nc effect was seen on fetal survival or weight.
0 Dieldrin (87% pure) was not found to be teratogenic in the CD rats and
CD-I mice administered doses of 1.5, 3.0 or 6.0 mg/kg/day by gastric
intubation on days 7 through 16 of gestation. Fetal toxicity, as
indicated by a significant decrease in numbers of caudal ossification
centers at the 6.0-mg/kg/day dose level and a significant increase
in the number of supernumerary ribs in one study group at both the
3.0- and 6.0-mg/kg/day dose level, was reported in the experiments in
mice. Maternal toxicity in the high-dose rats was indicated by a 41%
mortality and a significant decrease in weight gain; similarly, mice
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Dieldrin August, 1987
-6-
receiving 6.0 mgAg/day showed a significant decrease in maternal
weight gain. A significant increase in liver-to-body weight
ratio in one group of maternal mice was reported at both 3.0 and 6.0
mgAg/day (Chernoff et al., 1975).
Mutagenicity
0 Dieldrin was not mutagenic in the Salmonella/microsome test with and
without S-9 mix (McCann et al., 1975).
0 Dieldrin significantly decreased the mitotic index and increased
chromosome abnormalities in STS mice bone marrow cells in an ^n vivo
study. Similar observations were made in human WI-38 embryonic lung
cells in an in vitro test that also gave evidence of cytotoxicity, as
indicated by degree of cell degeneration (Majumdar et al., 1976).
Carcinogenicity
A dose-related increase in the incidence of hepatocellular carcinomas
was observed in B6C3F1 mice, with the incidence in the high-dose
males being significantly higher when compared to pooled controls
(NCI, 1978). Mice were given dieldrin (technical grade, >85% purity)
in the diet at concentrations of 2.5 or 5 ppm for 80 weeks. These
doses correspond to approximately 0.375 or 0.75 mgAg/day, respectively
(Lehman. 1959).
0 Osborne-Mendel rats treated with dieldrin at Time-Weighted Average (TWA)
doses of 29 or 65 ppm in the diet (approximately 1.45 or 3.25 mgAg/day,
respectively, based on Lehman, 1959) for 80 weeks, did not elicit
treatment-related tumors (NCI, 1978).
0 Diets containing 0.1, 1.0 or 10 ppm dieldrin (>99% purity), when
given to mice of both sexes for 132 weeks, were associated with an
increased incidence of liver tumors at all dose levels tested (Walker
et al., 1972). These doses are equivalent to approximately 0.015,
0.15 or 1.5 mg/kg/day, respectively (Lehman, 1959).
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) 9
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
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Dieldrin August, 1987
-7-
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
No data were found in the available literature that was suitable for
determination of a One-day HA value for dieldrin. It is, therefore, recommended
that the modified DUEL for a 10-kg child (0.0005 mg/L, calculated below) be
used as a conservative estimate for the One-day HA value.
Ten-day Health Advisory
No data were found in the available literature that was suitable for
determination of a Ten-day HA value for dieldrin. It is, therefore, recommended
that the modified DWEL for a 10-kg child (0.0005 mg/L, calculated below) be
used as a conservative estimate for the Ten-day HA value.
Longer-term Health Advisory
No data were found in the available literature that was suitable for
determination of a Longer-term HA value for dieldrin. It is, therefore,
recommended that the modified DWEL for a 10-kg child (0.0005 mg/L,
calculated below) be used as a conservative estimate for the Longer-term HA
value.
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
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Dieldrin August, 1987
-8-
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study of Walker et al. (1969), in which rats were fed dieldrin in
the diet at 0.0, 0.1, 1 or 10 ppm for 2 years (approximately 0, 0.005, 0.05
or 0.5 mg/kg/day based on Lehman, 1959), has been selected as the basis for
calculating the DWEL. In this study, liver weight and liver-to-body weight
ratios were significantly increased in females receiving 1 and 10 ppm, while
pathological changes consistent with exposure to organochlorides were evident
at the 10-ppm level. This study established a NOAEL of 0.1 ppm (equivalent
to 0.005 mg/kg/day).
Using a NOAEL of 0.005 mg/kg/day, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RFD = 0.005 mg/kg/day = 0.00oo5 mg/kg/day
where:
0.005 mg/kg/day - NOAEL, based on the absence of hepatic effects in
rats fed dieldrin in the diet.
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 (DWEL)
DWEL - (0.00005 m^/kg/day)(70 kg) , 0.00175 fflg/L (1>?5 ,
2 L/day
where:
0.00005 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
Dieldrin may be classified in Group B2: probable human carcinogen. A
Lifetime HA is not recommended for dieldrin.
The estimated excess cancer risk associated with lifetime exposure to
drinking water containing dieldrin at 1.75 ug/L is approximately 8.05 x 10-4.
This estimate represents the upper 95% confidence limit from extrapolations
prepared by EPA's Carcinogen Assessment Group (U.S. EPA, 1987) 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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Dieldrin August, 1987
-9-
Evaluation of Carcinogenic Potential
0 Applying the criteria described in EPA's proposed guidelines for
assessment of carcinogenic risk (U.S. EPA, 1986), dieldrin may be
classified in Group B2: probable human carcinogen.
0 Evidence has been presented in several carcinogenicity studies showing
that dieldrin is carcinogenic to mice. Thirteen data sets from these
studies are adequate for quantitative risk estimation. Utilizing the
linearized multistage model, the U.S. EPA performed potency estimates
for each of these data sets. The geometric mean of the potency
estimates, Q1 • = 16 (mg/kg/day)"1 , was estimated as the potency for the
general population (U.S. EPA, 1987).
0 Using this Qj* value and assuming that a 70-kg human adult consumes
2 liters of water a day over a 70-year lifespan, the linearized
multistage model estimates that concentrations of 0.219, 0.0219 and
0.00219 ug dieldrin per liter may result in excess cancer risk of
10-4, 10~5 and 10-6, respectively.
0 The linearized multistage model is only one method of estimating
carcinogenic risk. From the data contained in U.S. EPA (1987), it
was determined that five of the thirteen data sets were suitable for
determining slope estimates for the probit, logit; Weibull and gamma-
multihit models. Using the geometric mean of these slope estimates
(13 for multistage, 5 for other models) at their upper 95% confidence
limits, the following comparisons of unit risk (i.e., a 70-kg man
consuming 2 liters of water per day containing 1 ug/L of dieldrin over
a lifetime) can be made: multistage, 4.78 x 10-4; probit, 7.7 x 10-12-
logit, 5.09 x 10~6; Weibull, 1.13 x 10-4; multihit, 5.68 x 10-4. Each
model is based on different assumptions. No current understanding of
the biological mechanisms of carcinogenesis is able to predict which
of these models is more accurate than another.
0 While recognized as statistically alternative approaches, the range
of risks described by using any of these modelling 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 Agency has recommended use of the linearized multistage
approach.
0 IARC (1982) concluded that there is limited evidence that dieldrin is
carcinogenic in laboratory animals.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 ACGIH (1984) has established a short-term exposure limit (STEL) of
0.75 mg/m3 and an 8-hour Threshold Limit Value (TLV)-TWA exposure
0.25 mg/m3 for dieldrin.
0 U.S. EPA (1980) has recommended ambient water quality criteria of
0.71 ng/L for dieldrin. It is based on a carcinogenic potency factor
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Dieldrin August, 1987
-10-
(q^*) of 30.37 (mg/kg/day)"1 derived from the incidence of hepato-
cellular carcinoma in a mouse feeding study conducted by Walker et
al. (1972).
0 Residue tolerances ranging from 0.02 to 0.1 ppcn have been established
for dieldrin in or on agricultural commodities (U.S. EPA, 1985).
0 WHO (1982) established guidance of 0.03 ug dieldrin/L in drinking water.
VII. ANALYTICAL METHODS
0 Determination of dieldrin is by a liquid-liquid extraction gas
chromatographic (GC) procedure (U.S. EPA, 1984a). In this procedure,
a 1-liter sample is extracted with methylene chloride using a separatory
funnel. The methylene chloride extract is dried and exchanged to
hexane during concentration to a volume of 10 mL or less. The extract
is separated by GC, and the components are then measured with an
electron-capture detector. Identification may be corroborated through
the use of two unlike columns or by gas chromatography-mass
spectroscopy (GC-MS). A GC-MS procedure is available (U.S. EPA,
1984b) that allows for the qualitative and quantitative confirmation
of results obtained by the GC procedure.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that reverse osmosis (RO), granular-activated
carbon (GAC) adsorption, ozonation and conventional treatment will
remove dieldrin from water. The percent removal efficiency ranges
from 50 to 99+%.
0 Laboratory studies indicate that RO is a promising treatment method
for dieldrin-contaminated waters. Chian et al. (1975) reported 99+%
removal efficiency for two types of membranes operating at 600 psig and
a flux rate of 8 to 12 gal/ft2/day. Membrane adsorption, however, is
a major concern and must be considered, since breakthrough of dieldrin
would probably occur once the adsorption potential of the membrane
was exhausted.
0 GAC is effective for dieldrin removal. Pirbazari and Weber (1983)
reported 99+% dieldrin removal efficiency of a GAC column operating
at an empty bed contact time (EBCT) of 15 minutes and a hydraulic
loading of 1.4 gal/ft2/min, for the entire test period (approximately
7.5 months).
0 Pirbazari and Weber (1983) determined adsorption isotherms using GAC
on dieldrin in water solutions. Resin adsorption was also found to
remove dieldrin from water. The Freundlich values determined by
The authors indicate that the tested resins are not quite as effective
as GAC in the removal of dieldrin from water.
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Dieldrin August, 1987
-11-
Ozonation treatment appears to be an effective dieldrin removal
method. Treatment with 36 mg/L ozone (03) removed 50% of dieldrin
while 11 mg/L 03 removed only 15% of dieldrin (Robeck et al., 1965).
Conventional water-treatment techniques using alum coagulation,
sedimentation and filtration proved to be 55% effective in removing
dieldrin from contaminated potable water supplies (Robeck et al.,
1965). Lime- and soda-ash softening with ferric chloride as a coagulant
did not improve upon the removal efficiency achieved with alum alone.
Oxidation with chlorine and potassium permanganate is ineffective in
degrading dieldrin (Robeck et al., 1965).
Treatment technologies for the removal of dieldrin from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt dieldrin
removal from water must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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Dieldrin August, 1987
-12-
IX. REFERENCES
ACGXH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air. 3rd ed. Cincinnati, OH: ACGIH. p. 139.
Baldwin, M.K. and J. Robinson. 1972. A comparison of the metabolism of
HEOO (Dieldrin) in CF1 mouse with that in the CFE rat. Food Cosmet.
Toxicol. 10:333-351.
Chernoff, N., R.J. Kavlock, J.R. Kathrein, J.M. Dunn and J.K. Haseman.
1975. Prenatal effects of dieldrin and photodieldrin in mice and rats.
Toxicol. Appl. Pharmacol. 31:302-308.
Chian, E.S., W.N. Bruce and H.H.P. Fang. 1975. Removal of pesticides by
reverse osmosis. Environ. Sci. Technol. 9(1):52-59.
Coulston, F., R. Abraham and R. Mankes. 1980. Reproductive study in female
rats given dieldrin, alcohol or aspirin orally. Albany, NY: Albany
Medical College of Union University. Institute of Comparative and
Human Toxicology. Cited in IPCS, 1987.
Fitzhugh, O.G., A.A. Nelson and M.L. Quaife. 1964. Chronic oral toxicity of
aldrin and dieldrin in rats and dogs. Food Cosmet. Toxicol. 2:551-562.
Hayes, W.J., Jr. 1974. Distribution of dieldrin following a single oral
dose. Toxicol. Appl. Pharmacol. 28:485-492.
Heath, D.F. and M. Vandekar. 1964. Toxicity and metabolism of dieldrin in
rats. Br. J. Ind. Med. 21:269-279.
IARC. 1982. International Agency for Research on Cancer. IARC monographs
on the evaluation of the carcinogenic risk of chemicals to humans.
Chemicals, industry process and industries associated with cancer in
humans. IARC Monographs Vols. 1-29, Supplement 4. Geneva: World Health
Organi zation.
IPCS. 1987. International Programme on Chemical Safety. Environmental Health
Criteria for Aldrin and Dieldrin. United Nations Environment Programme.
International Labour Organization. Geneva: World Health Organization.
Lehman, A. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
MacKay, D. and A.W. Wolkoff. 1973. Rate of evaporation of low-solubility
contaminants from water bodies to atmosphere. Environ. Sci. Technol.
7:611.
Majumdar, S.K., H.A. Kopelman and M.J. Schnitman. 1976. Dieldrin-induced
chromosome damage in mouse bone marrow and WI-38 human lung cells.
J. Hered. 67:303-307.
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Dieldrin August, 1987
-13-
McCann, J., E. Choi, E. Yamasaki and B.N. Ames. 1975. Detection of carcinogens
as mutagens in the Salmonella/microsome test: Assay of 300 chemicals.
Proc. Natl. Acad. Sci. 72(12):5135-5139.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
HAS. 1977. National Academy of Sciences. Drinking water and health.
Vol. 1. Washington, DC: National Academy Press, pp. 556-571.
NCI. 1978. National Cancer Institute. Bioassay of aldrin and dieldrin for
possible carcinogenicity. Technical Report Series No. 21.
Ottolenghi, A.D., J.K. Haseman and F. Suggs. 1974. Teratogenic effects of
aldrin, dieldrin, and endrin in hamsters and mice. Teratology.- 9:11-16.
Pirbazari, M. and W.J. Weber. 1983. Removal of dieldrin from water by
activated carbon. J. Environ. Eng. 110(3):656-669.
Robeck, G.G., K.A. Dostal, J.M. Cohen and J.F. Kreessl. 1965. Effectiveness
of water treatment processes in pesticide removal. J. AWWA. (Feb):181-199.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. National
Institute for Occupational Safety and Health. National Library of
Medicine Online File.
STORET. 1987.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality
criteria for aldrin/dieldrin. EPA 440/5-80-019. Washington, DC: U.S.
EPA. NTIS Ace. No. PB 81-117301.
U.S. EPA. 1984a. U.S. Environmental Protection Agency. Method 608, organo-
chlorine pesticides and PCBs. Fed. Reg. 49(209):43234-43443. October 26.
U.S. EPA. 1984b. U.S. Environmental Protection Agency. Method 625, base/
neutrals and acids. Fed. Reg. 49(209):43234-43443. October 26.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal Regu-
lations. 40 CFR 180.137. July 1.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(1 85):33992-34003.
September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Carcinogenicity
assessment of aldrin and dieldrin. Carcinogen Assessment Group, Office
of Research and Development, U.S. EPA, Washington, DC 20460.
Walker, A.I.T., D.E. Stevenson, J. Robinson, E. Thorpe and M. Roberts. 1969.
The toxicology and pharmacodynamics of dieldrin (HEOD)). Two-year oral
exposures of rats and dogs. Toxicol. Appl. Pharmacol. 15:345-373.
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Dieldrin August, 1987
-14-
Walker, A.I.T., E. Thorpe and D.E. Stevenson. 1972. The toxicology of
dieldrin (HEOD). I. Long-term oral toxicity studies in mice.
Food Cosmet. Toxicol. 11:415-432.
Weast, R.c. and M. Astle, eds. 1982. CRC handbook of chemistry and physics
— A ready reference book of chemical and physical data, 63rd ed.
Cleveland, OH: CRC Press.
WHO. 1982. World Health Organization. Guidelines for drinking water quality.
Unedited final draft.
Windholz, M. 1983. The Merck index. 10th ed. Rahway, NJ: Merck and Co., Inc.
pp. 450-451.
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August, 1987
DIMETHRIN
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 pre.ict 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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Dimethrin
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 67239-16-1
Structural Formula
2,4-Dimethylbenzyl-2,2-dimethyl-3(2-methylpropenyl)-cyclopropane carboxylate
Synonyms
0 ENT 21,170; Chrysanthemumic acid; 2,4-Dimethylbenzylester.
Uses
0 Insecticide for use in ponds and swamps as a mosquito larvicide
(Meister, 1986).
Properties
C18H24°2
286.39 (Ambrose, 1964)
Amber liquid
175°C
0.98
Insoluble (further details not provided)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (2S°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
No information is available on the occurrence of dimethrin in water.
Environmental Fate
0 No information is available on the environmental fate of dimethrin.
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Dimethrin August, 1987
-3-
III. PHARMACOKINETICS
Absorption
0 In a preliminary metabolic study by Ambrose (1964), four rabbits were
given 5 mL/kg (5 mg/Jcg) of undiluted dimethrin by intubation. Urine
was collected every 24 hours over a 72-hour period. Identification
of two possible metabolites in the urine indicated that dimethrin was
absorbed. Sufficient data were not available to quantify the extent
of absorption.
Distribution
0 No information on the distribution of dimethrin was found in,the
available literature.
Metabolism/Excretion
0 Information presented by Ambrose (1964) indicates that dimethrin
(5 mg/kg), administered by intubation to rabbits, is metabolized (by
reduction) and excreted in the urine as chrysantheinumic acid and the
glucuronic ester of 2,4-dimethyl benzoic acid. Sufficient information
was not presented to determine if these are the only metabolites of
dimethrin or if any unchanged dimethrin is excreted.
IV. HEALTH EFFECTS
Humans
0 No information on the health effects of dimethrin in humans was found
in the available literature.
Animals
Short-term Exposure
0 The acute oral 1*050 value of dimethrin for male and female Sherman
rats was reported to be >15,000 mg/kg (Gaines, 1969).
0 Ambrose (1964) conducted an acute oral study in which male and
female albino rabbits (two/sex/dose) and male albino Wistar-CWL rats
(five/dose) were given a single dose of 10 or 15 mL/kg (9.8 or 14.7
mg/kg) of technical-grade dimethrin (98% pure) by gavage. Albino
guinea pigs (four/sex) received a single dose of 10 mL/kg (9.8 mg/kg)
by gavage. No effects were observed in rats or rabbits during a
2-week observation period. (Specific parameters observei were not
identified). In guinea pigs, the only effect reported during a
similar observation period was a refusal to eat or drink for 24 hours
following dosing.
0 Ambrose (1964) administered 10 mL/kg (9.8 mg/kg) of technical-grade
dimethrin (98% pure) to 15 male albino Wistar-CWL rats by gavage,
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Dimethrin August, 1987
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5 days per week for 3 weeks. This corresponds to an average daily
dose of 7 ing/kg. No adverse effects, as judged by general appearance,
behavior and growth, were observed. At necropsy, no gross abnormalities
were observed. No histopathological examinations were performed.
Dermal/Ocular Effects
0 Ambrose (1964) conducted a dermal irritation study in which dimethrin
(98% pure) was applied at a dose level of 10 mL/kg (9.8 mg/kg) to the
intact or abraded skin of four albino rabbits (two/sex) for a 24-hour
exposure period. No skin irritation was observed immediately after
the removal of the dimethrin or during a 2-week observation period.
0 Ambrose (1964) reported that single or multiple (3 consecutive days)
instillations of 0.1 mL of undiluted dimethrin (98% pure) into the
conjunctival sac of eight albino rabbits caused no visible irritation
or chemosis and no injury to the cornea as detectable by means of
fluorescein staining. When 0.2 mL of dimethrin was applied to the
penile mucosa of five albino rabbits on two occasions 6 days apart,
no irritation or sloughing of the mucosa was observed during a 1-week
observation period.
0 Masri et al. (1964) applied 3 mL of undiluted dimethrin-to the shaved
back and sides of three albino rabbits 10 times over a 2-week period
(frequency of application not specified). The only reported reaction
was the development of a slight scaliness which disappeared after
cessation of application.
0 Ambrose (1964) applied dimethrin (98% pure) to the skin of albino
rabbits (five/dose) 5 days per week for 13 weeks (65 applications).
Doses administered were 0.5 mL/kg undiluted dimethrin or 0.5 mL/kg
of a 50% solution of dimethrin in cottonseed oil (equivalent to
0.25 mL/kg of dimethrin); controls received 0.5 mL/kg of cottonseed
oil only. No evidence of any cutaneous reaction was observed.
Occasionally, a slight, nonpersistent erythema was observed in all
groups of rabbits. At necropsy, all organs from treated animals were
indistinguishable from the controls. No histopathological differences
between control and treated animals were observed.
Long-term Exposure
0 Masri et al. (1964) administered dimethrin to male (five/dose) and
female (six/dose) weanling albino rats for 16 weeks at dietary levels
of 0, 0.2, 0.6, 1.5 or 3.0%. Based on food consumption and body
weight data presented in the study, these dietary levels of dimethrin
were calculated to correspond to about 0, 120, 320, 1,000 or 2,300
mg/kg/day for males, and 0, 130, 400, 1,100 or 2,500 mg/kg/day for
females. Results indicated a significant reduction in body weight
in males receiving 0.6 or 3.0% and females receiving 1.5 or 3.0%.
Absolute liver weight and liver-to-body weight ratios were signifi-
cantly higher in both the male and female 1.5- and 3.0%-dose groups.
Kidney-to-body weight ratios were also significantly higher for these
groups. Scattered gross pathologic changes did not appear to bear a
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Dimethrin August, 1987
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relationship to dose. Histopathological examination revealed dose-
related morphological changes in the liver that consisted of a round
eosinophilic ring in the cytoplasm, approximately the size of the
nucleus. Amorphous material within the ring stained less densely
than the rest of the cytoplasm. Also, many hepatic cells of rats
receiving 1.5 or 3.0% dimethrin appeared larger than those of controls
and had less distinct basophilic cytoplasmic particles. Hepatic
changes were less pronounced in the 0.6% group. No cell inclusions
were seen in rats receiving 0.2% dimethrin. The effects of increased
liver and kidney-to-body weight ratios as well as histopathological
changes in the liver were shown to be reversible after withdrawal of
dimethrin. The No-Observed-Adverse-Effect-Level (NOAEL) identified
in this study was 0.2% dimethrin (120 mg/Jcg/day for males; 130 mg/kg/day
for females).
0 Ambrose (1964) administered dimethrin to male and female albino
Wistar-CWL rats (10/sex/dose} for 52 weeks at dietary levels of 0,
0.05, 0.1, 0.5, 1.0 or 2.0%. These dietary levels correspond to 0,
30, 60, 300, 600 or 1200 mg/kg/day. The only statistically significant
effect reported in this study was an increase in the liver-to-body
weight ratios in both male and female animals receiving 1.0 or 2.0%
dimethrin. Withdrawal of dimethrin from the diet for 6 weeks resulted
in return of liver weights to levels indistinguishable from the
controls. No differences in hemoglobin parameters were noted between
the treated and control animals at any time during the 52-week period.
Histologically, no significant changes or lesions that could be attrib-
uted to dimethrin in the diet were observed in any of the test groups
of animals. A NOAEL of 300 mg/kg was identified from this study.
0 Dimethrin has been implicated as a hypolipidemic agent and causes an
increase in hepatic peroxisome proliferation (Cohen and Grasso,
1981). Dietary administration of hypolipidemic agents to rodents has
resulted in induced liver carcinomas.
Reproductive Effects
0 No information on the reproductive effects of dimethrin was found in
the available literature,
Developmental Effects
0 No information on che developmental effects of dimethrin was found in
the available literature.
Mutagenicity
0 No information on the mutagenicity of dimethrin was found in the
available literature.
Carcinogenicity
0 No information on the carcinogenicity of dimethrin was found in the
available literature. However, the report by Cohen and Grasso (1981)
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Dimethrin August, 1987
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implicating dimethrin as a hypolipidemic agent may indicate that
dimethrin has carcinogenic potential in rodents. (It should be noted
that the relationship between hypolipodemic agents and liver carcinomas
in rodents has not been observed in humans.)
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 ( u /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, 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 information was found in the available literature that was suitable
for determination of the One-day HA values for dimethrin. It is therefore
recommended that the Longer-term HA for a 10-kg child (12 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
No information was found in the available literature that was suitable
for determination of tne Ten-day HA values for dimethrin. It is therefore
recommended that the Longer-term HA for a 10-kg child (12 mg/L, calculated
below) be used at this time as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
The 16-week rat study by Masri et al. (1964) has been selected to serve
as the basis for determination of the Longer-term HA. In this study, male
and female rats were administered dimethrin at dietary levels of 0, 0.2, 0.6,
1.5 or 3.0% for 16 weeks. Results of this study indicated a statistically
significant reduction in body weights of males receiving 0.6 or 3.0%, and
in females receiving 1.5 or 3.0%. Absolute liver weight and liver-to-body
weight ratios were significantly higher in the 1.5- and 3.0%-dose groups.
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Dimethrin August, 1987
-7-
Kidney-to-body weight ratios were also significantly higher in those groups.
Histopathological examinations revealed dose-related morphological changes in
the liver occurring at dose levels as low as 0.6%. A NOAEL of 0.2% dimethrin
(120 mg/kg/day for males; 130 mg/kg/day for females) was identified in this
study.
Using a NOAEL of 120 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Lonqer-term HA = (12° mg/kg/day) (10 kg) = 12 ng/L (12,000 ug/L)
9 (100) (1 L/day)
where:
120 mg/kg/day = NOAEL, based on absence of hepatic effects in male
rats exposed to dimethrin via the diet for 16 weeks.
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.
Using a NOAEL of 120 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = ***
where:
120 mg/kg/day = NOAEL, based on absence of hepatic effects in rats
exposed to dimethrin via the diet for 16 weeks.
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 f actor (s). From the RfD, a Drinking Water Equivalent Level
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Dimethrin August, 1987
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(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 52-week study in rats by Ambrose (1964) has been selected to serve
as the basis for determination of the Lifetime HA for dimethrin. In this
study, dimethrin was administered to albino Wistar-CWL rats for 52 weeks at
dietary levels of 0, 0.05, 0.1, 0.5, 1.0 or 2.0%. A statistically significant
increase in the liver-to-body weight ratio was observed in both male and
female rats receiving 1.0 or 2.0% dimethrin (600 and 1,200 mg/kg/day).
Histologically, no changes that could be attributed to dimethrin were observed
in any of the test groups. No adverse effects were reported in rats receiving
dimethrin at 0.5% (300 mg/kg/day for males) or lower.
Using a NOAEL of 300 mg/kg/day, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (300 mg/kg/day) = o.3 mg/kg/day
(1,000)
where:
300 mg/kg/day - NOAEL, based on absence of increased liver-to-body
weight ratio in rats exposed to dimethrin in the diet
for 52 weeks.
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 = (0.3 mg/kg/day) (70 kg) = 10.5 mg/L (10,500 ug/L)
(2 L/day)
where:
0.3 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Dimethrin August, 1987
-9-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA - (10.5 mg/L) (20%) - 2.1 mg/L (2,100 ug/L)
where:
10.5 mg/L = DWEL
20% » assumed percentage of daily exposure contributed by
ingestion of drinking water.
It should be noted that the Lifetime HA of 2.1 mg/L apparently exceeds the
water solubility of dimethrin (insoluble).
Evaluation of Carcinogenic Potential
0 No information on the carcinogenic!ty of dimethrin was found in the
available literature. However, the report by Cohen and Grasso (1981)
implicating dimethrin as a hypolipidemic agent may indicate that
dimethrin has carcinogenic potential in rodents. (It should be noted
that the relationship between hypolipidemic agents and liver carcinomas
in rodents has not been observed in humans.)
0 The International Agency for Research on Cancer has not evaluated the
carcinogenicity of dimethrin.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), dimethrin may be classified in
Group O: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No information on existing criteria, guidance, or standards pertaining
to dimethrin was found in the available literature. However, tolerances
for pyrethroids, of which dimethrin is a member, range from 0.05 ppm
in potatoes (post-harvest) to 3 ppm in wheat, barley, rice and oats
(CFR, 1985).
VII. ANALYTICAL METHODS
0 No information on the analytical methods for measuring dimethrin in
water was found in the available literature.
VIII. TREATMENT TECHNOLOGIES
0 The manufacture of this compound was discontinued (Meister, 1986). No
information was found in the available literature on treatment tech-
nologies capable of effectively removing dimethrin from contaminated
water.
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Dimethrin August, 1987
-10-
IX. REFERENCES
Ambrose, A.M. 1964. Tbxicologic studies on pyrethrin-type esters of chrysan-
themumic acid II. Chrysanthemumic acid, 2,4-dimethylbenzyl ester.
Toxicol. Appl. Pharmacol. 6:112-120.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.128.
Cohen, A.J., and P. Grasso. 1981. Review of hepatic response to hypolipidemic
drugs in rodents and assessment of its toxicological significance to
man. Food Cosmet. Toxicol. 4:585-605.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol.
14:515-534.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs,
cosmetics. Assoc. Food Drug Off. U.S. Q. Bull.
Masri, M.S., A.P. Henderson, A.J. Cox and F. De, eds. 1964. Subacute toxicity
of two Chrysanthemumic acid esters: barthrin and dimethrin. Toxicol.
Appl. Pharmacol. 6:716-725.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company, p. C81.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg;51(1851:33992-34003. September 24.
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August, 1987
DINOSEB
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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 differim assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Dinoseb
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 88-85-7
Structural Formula
2-sec-butyl=4,6-dinitropher,ol
Synonyms
Uses
DNBP, dinitro, dinoseb (BSI, ISO, WSSA); dinosebe (France); Basanite
(BASF Wyandotte); Caldon, Chemox General, Chemox PE, Cherasect DNBP,
ON-289 (product discontinued), Dinitro, Dinitro-3, Dinitro General,
Dynamite (Drexel Chemical); Elgetol 318, Gebutox, Hel-Fire (Helena);
Kiloseb, Nitropone C, Premerge 3(Agway), Sinox General (FMC Corp.);
Subitex, Unicrop DNBP, Vertac Dinitro Weed Killer 5, Vertac General
Weed Killer, Vertac Selective Weed Killer (Neister, 1984).
Dinoseb is used as a herbicide, desiccant and dormant fruit spray
(Heister, 1984).
Properties (WSSA, 1983)
Chemical Formula
Molecular Weight
Physical State (room temp.)
Boiling Point
Melting Point
Density (°C)
Vapor Pressure
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Dinoseb has been found in 1 of 79 surface water samples analyzed and
in 21 of 819 ground water samples (STORET, 1987). Samples were
collected at 70 surface water locations and 814 ground water locations.
ClOH12N2°5
240
Dark amber crystals
32°C
1.2647 (45°C)
(262°C) 100 mmHg
0.05 g/100 mL
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Dinoseb August, 1987
-3-
and dinoseb was found in California, Georgia and Ohio. The 85th
percentile of all non-zero samples was 1 ug/L in surface water and
10 ug/L in ground water sources. The maximum concentration found in
surface water was 1 ug/L and in ground water it was 100 ug/L.
0 Dinoseb has been found in New York ground water; typical positives
were 1 to 5 ppb (Cohen et al., 1986).
Environmental Fate
0 Dinoseb was stable to hydrolysis at pH 5, 7, and 9 at 25°C over a
period of 30 days (Dzialo, 1984).
0 With natural sunlight on a California sandy loam soil, dinoseb had a
half-life of 14 hours; with artificial light, it had a half-life of
30 hours, indicating that dinoseb is subject to photolytic degradation
(Dinoseb Task Force, 1985a).
0 In water with natural sunlight, dinoseb had a half-life of 14-18
days; with artificial light, it had a half-life of 42-58 days (Dinoseb
Task Force, 1985b).
0 With soil TLC plates, dinoseb was intermediate to very mobile in a
silt loam, sand, sandy loam and silty clay loam (Dinoseb Task Force,
1985c).
0 Soil adsorption studies gave a K
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Dinoseb August, 1987
-4-
acid)-4,6-diaminophenol, 2-(2-butyric acid)-4,6-dinitrophenol, 2-sec-
butyl-4-nitro-6-aminophenol, 2-sec-butyl-4-acetamido-6-nitrophenol and
2-(3-butyric acid)-4,6-dinitrophenol (Ernst and Bar, 1964; Froslie and
Karlog, 1970; Bandal and Casida, 1972).
Excretion
0 In mice, dinoseb is excreted in both urine (20%) and feces (30%)
following oral administration (specific means of administration not
specified) (Gibson and Rao, 1973).
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 While minimal data are available concerning human toxicity, at least
one death has been attributed to an accidental exposure of a farm worker
to sprayed dinoseb and dinitro-ortho-cresol (Heyndrickx et al., 1964).
Long-term Exposure
0 No information was found in the available literature on the long-term
health effects of dinoseb in humans.
Animals
Short-term Exposure
0 In rats and mice, the acute oral 1050 of dinoseb ranges from 20 to
40 mg/kg (Bough et al., 1965).
Dermal/Ocular Effects
0 In rats, the acute dermal toxicity of dinoseb ranges from 67 to
134 mg/kg (Noakes and Sanderson, 1969).
0 No information was found in the available literature on the dermal
or ocular effects of dinoseb in animals.
Long-term Exposure
0 Hall et al. (1978) reported the results (abstract only) of a feeding
study in male and female rats. Eight groups of rats, each group
composed of 14 male's and 14 females, were exposed to levels of 0, 50,
100, 150, 200, 300, 400 or 500 ppm of dinoseb (80% pure) in the diet
for 153 days, respectively. Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to 0, 2.5, 5.0, 7.5, 10.0, 15.0, 20.0 and 25.0 mg/kg/day. Mortality
was observed at 300 ppm (15 mg/kg/day) and above, and growth was
depressed at all dose levels. The LOAEL for this study was identified
as 50 ppm (2.5 mg/kg/day), the lowest dose tested.
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Dinoseb August, 1987
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0 In a 6-month dietary study by Spencer et al. (1948), groups of male
rats were exposed to dinoseb (99% pure) at levels of 0 (30 animals),
1.35, 2.7, 5.4 (20 animals) and 13.5 mg/kg/day (10 animals). Based
on increased mortality at the highest dose and an increase in liver
weight at intermediate doses, the NOAEL for dinoseb was identified as
2.7 mg/kg/day.
0 In a study submitted to EPA in support of the registration of dinoseb
(Hazleton, 1977), four groups of rats (60/sex/dose) were exposed to
dinoseb (purity not specified) in their diets for periods up to two
years at dose levels of 0, 1, 3 and 10 mg/kg/day, respectively.
Although no evidence of dose-related changes in histopathology,
hematology, blood chemistry or certain other parameters were observed,
a dose-related decrease in mean thyroid weight was observed in all
treated males. The LOAEL in this study was identified as 1 mg/kg/day.
Reproductive Effects
0 In a reproduction study by Linder et al. (1982), four groups of ten
male rats each were exposed to dinoseb (97% pure) in the diet at
levels of 0, 3.8, 9.1 or 15.6 mg/kg/day over an 11-week period,
respectively. In addition, a group of five animals was exposed to
22.2 mg/kg/day. The fertility index was reduced to 0 at 22.2 mg/kg
and to 10% at 15.6 mg/kg/day; in neither case did the fertility index
improve in 104 to 112 days following treatment. A variety of other
effects were seen at levels of 9.1 mg/kg/day and higher, including
decreased weight of the seminal vesicles, decreased sperm count and
an increased incidence of abnormal sperm. The NOAEL for dinoseb in
this study was 3.8 mg/kg/day based on a decrease in sperm count and
other effects at higher levels.
0 In a 2-generation rat reproduction study (Irvine, 1981), four groups
of rats (25/sex/dose) were exposed to 0, 1, 3, and 10 mg/kg/day of
dinoseb in the diet for 29 weeks. Although no reproductive effects
were observed in this study per se, a decrease in pup body weight was
observed at day 21 post-parturition for all dose levels. Thus, based
on a compound-related depression in pup body weight at all dose
levels, the LOAEL in this study was 1 mg/kg/day.
Developmental Effects
0 Although dinoseb has been reported to be teratogenic (e.g., oligodactyly,
imperforate anus, hydrocephalus, etc.) when administered to mice
intraperitoneally (Gibson, 1973), it was not teratogenic when admini-
stered orally to mice (Gibson, 1973; Gibson and Rao, 1973) or rats
(Spencer and Sing, 1982).
0 Dinoseb (95% pure), administered to pregnant rats in the diet on
days 6 through 15 of gestation, produced a marked reduction in fetal
survival at doses of 9.2 mg/kg/day and above but not at doses of
6.9 mg/kg/day (NOAEL) and below (Spencer and Sing, 1982).
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Dinoseb August, 1987
-6-
0 Dinoseb (purity not specified) was without effect in a study in which
pregnant mice were orally exposed to a single dose of 15 mg/kg/day
(Chernoff and Kavlock, 1983).
0 In a developmental toxicity study by Research and Consulting Company
(1986), four groups of 16 Chinchilla rabbits were exposed to dinoseb
(98% pure) by oral gavage at levels of 0, 1, 3 or 10 mg/kg/day from
day 6 to 18 of gestation. At the highest dose level dinoseb produced
a statistically significant increase in malformations and/or anomalies
when compared to the controls, with external, internal (body cavities
and cephalic viscera) and skeletal defects being observed in 11/16
litters examined. Neural tube defects, the major developmental toxic
effect, included dyscrania associated with hydrocephaly, scoliosis,
kyphosis, malformed or fused caudal and sacral vertebrae and
encephalocele. The NOAEL for dinoseb in this study was identified as
3.0 mg/kg/day, based on the occurrence of neural tube defects at the
highest dose level.
0 In a study by the Dinoseb Task Force (1986), developmental toxicity
was observed in Wistar/Han rats. Groups of 25 rats received dinoseb
(purity 96.1%) by gavage at levels of 0, 1, 3 or 10 mg/kg/day from
day 6 to 15 of gestation. Developmental toxicity was observed at the
high dose as evidenced by a slight depression in fetal body weight,
increased incidence of absence of skeletal ossification for a number
of sites and an increase in the number of supernumerary ribs. Slight
to moderate decreases in body weight gain and food consumption was
observed in dams at the intermediate- and high-dose levels. Based on
the occurrence of developmental effects at the highest dose level, a
NOAEL of 3.0 mg/kg/day was identified.
Mutaqenicity
0 With the exception of an increase in DNA damage in bacteria (Waters,
et al., 1982), dinoseb was not mutagenic in a number of organisms
including Salmonella typhimurium, Escherichia coli, Saccharomyces
cerevisiae, Drosophila melanogaster or Bacillus subtilis (Simmon
et al., 1977; Waters et al., 1982; Moriyta et al., 1983).
Carcinogenicity
0 No evidence of a carcinogenic response was observed in a 2-year
chronic feeding study in which dinoseb was administered to rats at
levels as high as 10 mg/kg/day (Hazleton, 1977).
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)
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-6a-
ATTENTION
I. BACKGROUND
Over approximately the last 18 months, HEB/ODW has been
developing a Health Advisory (HA) for the herbicide Dinoseb.
Among other toxic endpoints, the Dinoseb HA notes that there is
a positive rabbit oral teratology study with a NOAEL of 3 mg/kg/day-
the basis of the proposed Ten-day HA value.
Subsequent to the latest HEB revision of the Dinoseb HA, a
rabbit dermal teratology study and certain other studies became
available. Both the rabbit dermal teratology study and the -
other studies are currently under Agency review. However, the
rabbit dermal teratology is positive with a NOAEL of 1 mg/kg/day.
In addition, the same toxic effect, neural tube defects, was
observed in both the oral and dermal teratology studies.
II. ISSUE
While no final decision concerning Dinoseb can be made until
all available data have undergone Agency review, the dermal
teratology raises certain issues of concern to ODW. Specifically:
0 Exposure to both the embryo and fetus is determined by the
mother's exposure. Thus, in the case of a teratogen, woman of
child bearing age are the group of principal interest.
0 In the case of an adult - i.e. woman of child bearing age - the
HA values are based on the consumption of 2 liters of water per
day by a 70-kg adult.
0 Considerably more water is used to bathe (roughly 100 L/day)
than is ingested (2 L/day).
0 Toxic amounts of Dinoseb can be readily absorbed dermally - i.e.,
the dermal NOAEL of 1 mg/kg/day is less than the oral NOAEL of
3 mg/kg/day.
0 Since bathing and other practices involve dermal exposure to
drinking water contaminants, it is at least possible that the
dermal absorption of Dinoseb may result in significant exposure.
Until the issue of the dermal absorption of Dinoseb is
resolved, ODW believes the following procedure should be used to
allow for the positive dermal teratology study.
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-6b-
III. RESOLUTION OF ISSUE
A. Interim
Until such time as detailed data concerning the dermal
absorption of Dinoseb are available, it is suggested that, on an
interim basis, an HA value of 3.5 ug/L be used to evaluate all
exposure situations (e.g. One-day, Ten-day etc.) where significant
dermal exposure may be involved. This conclusion is based on
the following analysis which suggests that a level of 3.5 ug
Dinoseb/L will offer adequate protection against both the oral
and dermal teratogenic potential of Dinoseb:
Oral and dermal HA =
Where:
1 mg/kg/day =
70 kg =
100 =
102 L/day =
(1 mg/kg/day)(70 kg)
^»^»^o^B^ ^•^^•^^•^^M ^m^^m^mm^m^m^**
(100)(102 L/day)
0.007 mg/L (7 ug/L)
tentative NOAEL in rabbit dermal teratogenic
study.
assumed body weight of a woman of child
bearing age.
uncertainty factor, chosen in accordance with
NAS/ODW guidelines for use with a NOAEL from
an animal study.
possible volume of water from which all
Dinoseb is either absorbed dermally (100 L)
or ingested (2 L). While this value is
possibly overly conservative, it provides
an interim worst case until such time as
Dinoseb dermal absorption studies (in
progress) are available.
Normally, ODW uses a Relative Source Contribution (RSC) factor
of 20% when the actual RSC is unknown. However, since it is at
least possible that the RSC may be of some magnitude (due to
dermal absorption), ODW has determined that it is appropriate to
use an RSC of 50% in this case. Using an RSC of 50%, ODW recommends
that an HA value of 3.5 ug/L (7.0 ug/L x 50%) not be exceeded.
B. Final
Any final conclusion must await the results of ongoing
Dinoseb dermal absorption studies.
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Oinoseb August, 1987
-7-
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 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. It is therefore recommended that
the Ten-day HA value for a 10-kg child (0.3 mg/L, calculated below) be used
as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The rabbit developmental toxicity study (Research and Consulting Co.,
1986) in which dinoseb produced neural tube defects at doses greater than 3
mg/kg/day (NOAEL) was selected as the basis for determination of the Ten-day
HA. While it is reasonable to base a Ten-day HA for the adult on a positive
developmental toxicity study, there is some question as to whether it is
appropriate to base the Ten-day HA for a 10-kg child on a such a study.
However, since this study is of appropriate duration and since the fetus may
be more sensitive than a 10-kg child, it was judged that, while it may be
overly conservative, it is reasonable to base the Ten-day HA for a 10-kg
child on such a study.
Using a NOAEL of 3.0 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (3.0 mg/kq/day) (10 kg) = 0.3 mg/L (300 ug/L)
(100) (1 L/day)
where:
3.0 mg/kg/day = NOAEL, based on the absence of teratogenic effects
in rabbits.
10 kg a 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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Dinoseb August, 1987
-8-
Lonqer-term Health Advisory
The Hall et al. (1978) 153-day dietary dinoseb study in rats was
originally selected to serve as the basis for determination of the Longer-
term HA (decreased growth was observed at all exposure levels with a LOAEL of
2.5 mg/kg/day). Subsequently, however, a 2-generation reproduction study in
rats (Irvine, 1981) was identified with a LOAEL of 1 mg/kg/day (based on a
decrease in pup body weight at all dose levels). Since a reproduction study
is of appropriate duration, the Irvine (1981) study has been selected to serve
as the basis for determination of the Longer-term HA.
Using a LOAEL of 1 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA - (1>0 mg/fcg/day) (10 kg) = 0.010 mg/L (10 ug/L)
(1,000) (1 L/day)
where:
1.0 mg/kg/day = LOAEL, based on decreased pup body weight.
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.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (1'° **/*<*/***) (70 *«*) = 0.035 mg/L (35 ug/L)
(1,000) (2 L/day)
where:
1.0 mg/kg/day = LOAEL, based on decreased pup body weight.
70 kg = assumed body weight of an adult.
1,000 = uncertainty factor; chosen in accordance with NAS/ODW
guidelines for use with a LOAEI 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
-------
Dinoseb August, 1987
-9-
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 2-year dietary rat study by Hazelton (1977) was selected to
serve as the basis for determination of the Lifetime HA. In this study, a
compound-related decrease in mean thyroid weights was observed in all males
(LOAEL = 1 mg/kg/day) treated with dinoseb (purity not specified).
Using a LOAEL of 1 mg/kg/day, the Lifetime HA for a 70 kg adult is
calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (1 mg/kg/day) = 0.001
(1,000)
where:
1 mg/kg/day = LOAEL, based on decreased thyroid weight in male rats
exposed to dinoseb via the diet for up to two years.
1,000 = uncertainty factor; chosen in accordance with NAS/ODH
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0*001 mg/kg/day) (70 kg) = 0.035 mg/L
(2 L/day)
where:
0.001 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 = (0.035 mg/L) (20%) = 0.007 mg/L (7 ug/L)
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Dinoseb August, 1987
-10-
where:
0.035 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenicity was found in a 2-year dietary study
in which dinoseb was administered to rats at levels as high as 10
ng/kg/day (Hazleton Labs, 1977).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of dinoseb.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), dinoseb is classified in
Group D: not classified. This group is for agents with indadequate
human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Tolerances have been established for dinoseb (40 CFR 180.281) at
0.1 ppm on a wide variety of agricultural commodities.
0 The~EPA RfD Workgroup approved a 0.001 mg/kg/day RfD for dinoseb.
The EPA RfD Workgroup is an EPA wide group whose function is to
ensure that consistent RfD values are used throughout the EPA.
VII. ANALYTICAL METHODS
0 Analysis of dinoseb is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated acid pesticides in water
samples (U.S. EPA, 1985). In this method, approximately 1 liter of
sample is acidified. The compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolyzed with
potassium hydroxide, and extraneous organic material is removed by
a solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane as the derivatizing
agent. Excess reagent is removed, and the esters are determined by
electron capture GC. The method detection limit has been estimated
at 0.07 ug/L for dinoseb.
VIII. TREATMENT TECHNOLOGIES
0 The treatment technologies which will remove dinoseb from water include
activated carbon and ion exchange. No data were found for the removal
of dinoseb from drinking water by conventional treatment or by aeratione
However, limited data suggest that aeration would not be effective in
the removal of dinoseb from drinking water (ESE, 1984).
-------
Dinoseb August, 1987
-11-
• Becker and Wilson (1978) reported on the treatment of a contaminated
lake water with three activated carbon columns operated in series.
The columns processed about 2 million gallons of lake water and
achieved a 99.98 percent removal of dinoseb. Weber and Gould (1966)
performed successful isotherm tests using Columbia LC carbon, which
is coconut based, and reported the following Langmuirian equilibrium
constants:
Q = 444 mg dinoseb per g of carbon
1/b = 1.39 mg/L
Though the Langmuir equation provides a good fit over a broad
concentration range, greater adsorption would probably be achieved at
lower concentrations (less than 100 ug/L) than predicted by using
these constants.
0 Weber (1972) has classified dinoseb as an acidic pesticide; and such
compounds have been readily adsorbed in large amounts by ion exchange
resins. Harris and Warren (1964) studied the adsorption of dinoseb
from aqueous solution by anion exchanger (Amberlite® IRA-400) and a
cation exchanger (Amberlite® IR-200). The anion exchanger adsorbed
dinoseb to less than detectable limits in solution.
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Dinoseb August, 1987
-12-
IX. REFERENCES
Bandal, S.K., and J.E. Casida. 1972. Metabolism and photoalteration of
2-sec-butyl-4,6-dinitrophenol (DNBP herbicide) and its isopropyl carbonate
derivative (dinobuton acaricide). J. Agr. Food Chem. 20:1235-1245.
Becker, D.L. and Wilson, S.C. 1978. The use of activated carbon for the
treatment of pesticides and pestididal wastes. In Carbon Adsorption
Handbook (D.H. Cheremisinoff and F. Ellerbusch, Eds.). Ann Arbor Science
Publishers, Ann Arbor, MI.
Bough, R.G., E.E. Cliffe and B. Lessel. 1965. Comparative toxicity and blood
level studies on binapacryl and DNBP. Toxicol. Appl. Pharmacol. 7:353-360.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.281. July 1, 1985.
Chernoff, N., and R.J. Kavlock. 1983. A teratology test system which
utilizes postnatal growth and viability in the mouse. Environ, Sci. Res.
27:417-427.
Cohen, S.Z., C. Eiden and M.N. Lorber. 1986. Monitoring ground water for
pesticides in the USA. ^n American Chemical Society Symposium Series
titled Evaluation of Pesticides in Ground water (in press).
Dinoseb Task Force. 1985a. Photodegradation of dinoseb on soil. Prepared
by Hazleton Laboratories America, Inc. Report No. 6015-191 (Tab 3),
July 19, 1985.
Dinoseb Task Force. 1985b. Photodegradation of dinoseb in water. Prepared
by Hazleton Laboratories America, Inc. Report No. 6015-190 (Tab 4),
July 19, 1985.
Dinoseb Task Force. 1985c. Determination of the mobility of dinoseb in
selected soils by soil TLC. Prepared by Hazleton Laboratories America,
Inc. Report No. 6015-192 (Tab 1). July 19, 1985.
Dinoseb Task Force. 1985d. The adsorption/desorption of dinoseb on repre-
sentative agricultural soils. Prepared by Hazleton Laboratories America,
Inc. Report No. 6015-193 (Tab 2), July 19, 1985.
Dinoseb Task Force. 1986. Probe embryotoxicity study with dinoseb technical
grade in wistar rats. Prepared by Research and Consulting Company.
Project No. 045281. April 22, 1986.
Dzialo, D. 1984. Hydrolysis of dinoseb: Project No. 84239. Unpublished
study prepared by Uniroyal Inc.
Environmental Science and Engineering (ESE). 1984. Review of treatability
data for removal of twenty-five synthetic organic chemicals from drinking
water. U.S. Environmental Protection Agency, Office of Drinking Water,
Washington, DC.
Ernst, W., and F. Bar. 1964. Die umwandlung des 2,4-dinitro-6-sec-butylphenols
and seiner ester im tierischen organismus. Arzenimittel Forschung 14:81-84.
-------
Dinoseb August, 1987
-13-
Froslie, A., and 0. Karlog. 1970. Ruminal metabolism of DNOC and DNBP. Acta
Vet. Scand. 11:31-43.
Gibson, J.E. 1973. Teratology studies in mice with 2-sec-butyl-4,6-dinitro-
phenol (dinoseb). Fd. Cosmet. Toxicol. 11:31-43.
Gibson, J.E, and K.S. Rao. 1973. Disposition of 2-sec-butyl-4,6-dinitrophenol
(dinoseb) in pregnant mice. Food Cosmet. Toxicol. 11:45-52.
Hall, L., R. Linder, T. Scotti, R. Bruce, R. Moseman, T. Heidersheit, D. Hinkle,
T. Edgerton, S. Chaney, J. Goldstein, M. Gage, J. Farmer, L. Bennett,
J. Stevens, W. Durham and A. Curley. 1978. Subchronic and reproductive
toxicity of dinoseb. Toxicol. Appl. Pharmacol. 45:235-236. (abstract
only)
Harris, C.I. and G.F. Warren. 1964. Adsorption and desorption of herbicides
by soil. Weeds, 12:120.
Hazleton.* 1977. Hazleton Labs. 104-Week dietary study in rats. Dinoseb DNBP.
Final Report. Unpublished study. MRID 00211
Heyndrickx, A., R. Maes and F. Tyberghein. 1964. Fatal intoxication by man
due to dinitro-ortho-cresol (DNOC) and dinitro butylphenol (DNB°). Mededel
Lanbovwhoge School Opzoekingstaa Staa Gent 29:1189-1197.
Irvine, L.F.H.* 1981. 3-Generation reproduction study; Hazelton Laboratories
Europe, Ltd.
Lehman, A. J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
Linder, R.E., T.M. Scotti, D.J. Svendsgaard, W.K. McElroy and A. Curley.
Testicular effects of dinoseb in rats. Arch. Environ. Toxicol.
11:475-485.
Meister, R., ed. 1984. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
Moriya, M., T. Ohta, T. Watanabe1, K. Kato and Y. Shirasu. 1983. Further
mutagenicity studies on pesticides in bacterial reversion assay systems.
Mut. Res. 116:185-216.
Noakes, D.N., and D.M. Sanderson. 1969. A method for determining the dermal
toxicity of pesticides. Brit. J. Ind. Med. 26:59-64.
Research and Consulting Company. 1986. Embryotoxicity study with dinoseb
technical grade in the rabbit (oral administration). Unpublished study.
Simmon, V.F., A.D. Mitchell and T.A. Jorgenson. 1977. Evaluation of selected
pesticides as chemical mutagens in vitro and in vivo studies. Research
Triangle Park, NC: U.S. Environmental Protection Agency, EPA 600/1-77-028.
-------
Dinoseb August, 1987
-14-
Spencer, F. and L.T. Sing. 1982. Reproductive toxicity in pseudopregnant
and pregnant rats following postimplantational exposure: Effects of the
herbicide dinoseb. Pestic. Biochem. Physiol. 18:150-157.
Spencer, H.C., V.K. Rowe, E.M. Adams and D.D. Irish. 1948. Toxicological
studies on laboratory animals of certain alkyldinitrophenols used in
agriculture. J. Ind. Hyg. Toxicol. 30:10-25.
STORET. 1987.
U.S. EPA. 1985. U.S. EPA Method 615 - Chlorinated Phenoxy Acids. 50 FR
50701, October 4, 1985.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Waters, M.D., S. Shahbeg, S. Sandhu et al. 1982. Study of pesticide
genotoxicity. Basic Life Sci. 21:275-326.
Weber, J.B. 1972. Interaction of organic pesticides with particulate matter
in aquatic and soil systems. In^ Advances in Chemistry Series 111 (R.F.
Gould, Ed.). American Chemical Society, Washington, DC.
Weber, W.J.,Jr. and J.P. Gould. 1966. Sorption of organic pesticides from
aqueous solution. In Advances in Chemistry Series 60 (R.F. Gould,
Ed.). American Chemical Society, Washington, DC.
WSSA. 1983. Weed Science Society of America. Herbicide handbook, 5th edition.
Champaign, IL.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
DIPHENAMID
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
August, 1987
DRAFT
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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Diphenamid August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 957-51-7
Structural Formula
0
HC-C-N(CH,)2
N,N-dimethyl-alpha-phenyl-benzeneacetamide
Synonyms
0 Dymid; Enide (Meister, 1983).
Uses
0 Pre-emergent and selective herbicide for tomatoes, peanuts, alfalfa,
soybean, cotton and other crops (Meister, 1986).
Properties (Windholz et al., 1983)
Chemical Formula C16H17ON
Molecular Weight 239.30
Physical State (at 25°C) White crystalline solid
Boiling Point
Melting Point 135°C
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (27°C) 260 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Diphenamid has not been found in any of the water samples collected
and analyzed from 567 ground water locations (STORET, 1987).
Environmental Fate
0 Diphenamid is stable to hydrolysis at pH 5, 7 and 9 for 7, 12 and
10 days, respectively, at elevated temperature (49°C or 120°F)
(NOR-AM, 1986).
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Diphenamid August, 1987
-3-
0 Diphenaraid is intermediately mobile (class 3) on silt loam and silty
clay loam soil TLC plates; on sandy loam, it is in class 5, indicating
that it would leach readily in this soil (Helling and Turner, 1968).
III. PHARMACOKINETICS
Absorption
0 No information was found in the available literature on the absorption
of diphenamid.
Distribution
0 No information was found in the available literature on the distri-
bution of diphenamid.
Metabolism
0 No information was found in the available literature on the metabolism
of diphenamid.
Excretion
0 No information was found in the available literature on the excretion
of diphenamid.
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of diphenamid in humans.
Animals
Short-term Exposure
0 RTECS (1985) reported the acute oral LD50 values in the rat, mouse,
dog, monkey and rabbit to be 60C, 700, 1,000, 1,000 and 1,500 mg/kg,
respectively.
Dermal/Ocular Effects
0 Weddon and Brown (1976) applied a 90% wettable powder formulation of
diphenamid to intact or abraded skin of New Zealand rabbits (two/sex/
dose) for 24 hours at 0, 200, 1,000 or 2,000 mg/kg. No adverse
responses were observed in any of the exposed animals.
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Diphenamid August, 1987
-4-
Long-term Exposure
0 Woodard et al. (1966b) administered technical diphenamid (purity
not specified) in the feed to beagle dogs (three/sex/dose) at dose
levels of 0, 3, 10 or 30 mg/kg/day for 103 weeks. No pathological
effects were reported at 3 mg/kg/day for clinical chemistry, hematology,
urinalysis, gross pathology and histopathology. Liver weights were
slightly increased in the 10- and 30-mg/kg/day dosage groups of both
sexes, and there were slight increases in numbers of portal macrophages
and/or fibroblasts when compared to untreated controls. Liver enzyme
levels were normal in all treated groups, except for elevation of
serum glutamic-oxaloacetic transaminase (SCOT) after 8 weeks in one
female dosed with 3 mg/kg/day. A No-Observed-Adverse-Effect-Level
(NOAEL) of 3 mg/kg/day and a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 10 mg/kg/day were identified by this study.
0 Hollingsworth et al. (1966) fed technical diphenamid (>98% pure) to
rats (30/sex/dose) at dose levels of 0, 3, 10 or 30 mg/kg/day for 101
weeks. A slight increase in the mean absolute liver weights of males
and the relative liver and thyroid weights of females in the high-
dose groups was observed. No other adverse effects were reported
at 10 mg/kg/day or less in general behavior, feed consumption, body
and organ weights, hematology, gross pathology and histopathology.
A NOAEL of 10 mg/kg/day was identified by this study.
Reproductive Effects
0 In a three-generation reproduction study, Woodard et al. (1966a)
supplied diphenamid to albino rats (10 males and 20 females/dose) at
dose levels of 0, 10 or 30 mg/kg/day. No reproductive or pathological
effects were observed for the parental generations (F0, F1b, F-jt,)
at any dose tested. Weanlings of the F3D generation dosed with
30 mg/kg/day showed reversible liver changes, including slight
congestion, glycogen depletion and irregular size of the hepatocytes.
Based on reproductive end points, this study identifies a NOAEL of
30 mg/kg/day. Based on fetal toxicity, a NOAEL of 10 mg/kg/day and
a LOAEL of 30 mg/kg/day are identified.
Developmental Effects
0 Woodard et al. M966a) reported no developmental effects in rat pups
at any dose level. Reversible liver changes were observed in weanling
pups of the F3b generation dosed with 30 mg/kg/day. A NOAEL based on
fetotoxicity of 10 mg/kg/day can be identified.
Mutaqenicity
0 Moriya et al. (1983) reported that diphenamid (up to 5,000 ug/plate)
did not increase reversion frequency in £. typhimurium or £. coli
test systems, either with or without metabolic activation.
0 Shirasu et al. (1976) reported that diphenamid (1%) was not mutagenic
in a recombination assay utilizing B. subtills or in reversion assays
with E. coli or S. typhimurium.
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Diphenamid August, 1987
-5-
Carcinogenicity
0 In a 2-year feeding study in rats by Hollingsworth et al. (1966),
diphenamid was administered to albino rats (30/sex/dose) at dose
levels of 0, 3, 10 or 30 mg/kg/day for 101 weeks. Based on
histopathological examination of a variety of tissues and organs,
the authors reported that the type and incidence of neoplasms were
comparable in treated and control rats.
0 In a 2-year feeding study in dogs by Woodard et al. (1966b>, diphenamid
was administered in the feed to beagle dogs (three/sex/dose) at dosage
levels of 0, 3, 10 or 30 mg/kg/day for 103 weeks. Histopathological
examinations were performed on a variety of tissues and organs, and
no evidence of increased tumor frequency was reported.
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 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 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 diphenamid. It is therefore
recommended that the Drinking Water Equivalent Level (DWEL), adjusted for a
10-kg child (0.3 mg/L, calculated below), be used at this time as a conservative
estimate of the One-day HA value.
For a 10-kg child, the adjusted DWEL is calculated as follows:
(0.03 mg/kg/day) (10 kg) _ 0.3 mg/L
(1 L/day)
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Diphenamid A"9ust' 1987
-6-
where:
0.03 mg/kg/day = RfD (see Lifetime Health Advisory Section).
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of a child.
Ten-day Health Advisory
No information was found in the available literature that was suitable
for determination of the Ten-day HA value for diphenamid. It is therefore
recommended that the DWEL, adjusted for a 10-kg child (0.3 mg/L) be used at
this time as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
No information was found in the available literature that was suitable
for determination of the Longer-term HA value for diphenamid. It is therefore
recommended that the DWEL value, adjusted for a 10-kg child (0.3 mg/L) be
used at this time as a conservative estimate of the Longer-term HA value.
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to rhis chemical.
The feeding study in dogs by Woodard et al. (1966b) has been selected to
serve as the basis for determination of the Lifetime HA value for diphenamid.
In this study, dogs were administered technical diphenamid (0, 3, 10 or 30
mg/kg/day) in the diet for 103 weeks. Based on clinical chemistry, hematology,
urinalysis, gross pathology and histopathology, this study identified a NOAEL
of 3 mg/kg/day and a LOAEL of 10 mg/kg/day. The study by Hollingsworth et al.
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Diphenamid August, 1987
-7-
(1966), which identified a NOAEL of 10 mg/kg/day in a 101-week experiment in
rats, was not selected, since the rat appears to be somewhat less sensitive
than the dog (the NOAEL in the rat is the same as the LOAEL in the dog).
Using a NOAEL of 3 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (3 mg/kg/day) . 0.03 mg/kg/day
(100)
where:
3 mg/kg/day = NOAEL, based on absence of organ weight loss, clinical
chemistry, hematology, urinalysis, gross pathology and
histopathology in dogs exposed to diphenamid via the
diet for 103 weeks.
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*03 mg/kg/day) (70 kg) =1.0 mg/L (1,000 ug/L)
(2 L/day)
where:
0.03 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 = (1.0 mg/L) (20%) =0.2 mg/L (200 ug/L)
where:
1.0 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic potential was detected in rats (30/sex/dose)
fed diphenamid in the diet for 2 years at a dose level of 30 mg/kg/day
(Hollingsworth et al., 1966), or in dogs (three/sex/dose) fed diphenamid
in the diet for 2 years, also at a dose of 30 mg/kg/day (Woodward
et al., 1966b). These studies are limited by the low doses and the
small number of animals employed.
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Diphenamid August, 1987
-8-
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of diphenamid.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), diphenamid is classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Tolerances in or on raw agricultural commodities of 0.01 ppm for milk
to 2 ppm for peanut hay and forage have been set for diphenamid (U.S.
EPA, 1985).
VII. ANALYTICAL METHODS
0 Analysis of diphenamid is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen phosphorus
detector. The method detection limit has not been determined for
diphenamid but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAC) adsorp-
tion will remove diphenamid from water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
diphenamid on GAC.
0 Whittaker (1980) reported the results of GAC columns operating under
bench-scale conditions. At a flow rate of 0.8 gpm/sq ft and an empty
bed contact time of 6 minutes, diphenamid breakthrough (when effluent
concentration equals 10% of influent concentration) occurred after
500 bed volumes (BV). When two bi-solute diphenamid solutions were
passed over the same column, diphenamid breakthrough occurred after
235 BV for diphenamid-propham solution and after 290 BV for diphenamid-
fluometuron solution.
0 GAC adsorption appears to be the most effective treatment technique
for the removal of diphenamid from contaminated water. However,
selection of individual or combinations of technologies to attempt
diphenamid removal from water must be based on a case-by-case technical
evaluation, and an assessment of the economics involved.
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Diphenamid August, 1987
-9-
IX. REFERENCES
Helling, C.S., and B.C. Turner. 1968. Pesticide mobility: Determination
by soil TLC. Science. 16:562-563.
Hollingsworth R.L., M.W. Woodard and G. Woodard.* 1966. Diphenamid safety
evaluation by dietary feeding to rats for 101 weeks. Final Report.
Unpublished study. MRID 00076381.
Meister, R.T., ed. 1986. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
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.
NOR-AM. 1986. NOR-AM Chemical Company. Diphenamid: Hydrolysis study (ground
water data call-in). Wilmington, DE. Unpublished study submitted to the
Office of Pesticide Programs.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. National
Institute for Occupational Safety and Health. Washington, DC. National
Library of Medicine On-Line File.
Shirasu, Y., M. Moriya, K. Kato, A. Furuhashi and T. Kada. 1976. Mutagenicity
screening of pesticides in the microbial system. Mutat. Res. 40:19-30.
STORET. 1987.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.230.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24,
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Weddon T.E., and P.K. Brown.* 1976. Enide 90 W—Dermal LD50 and skin
irritation evaluation in New Zealand rabbits. Technical Report No.
124-961O-MWG-76-6. Unpublished study. MRID 00054611.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, 10th ed. Rahway, NJ: Merck and Co., Inc.
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Diphenamid August, 1987
-10-
Woodard M.W., G. Woodard and M.T. Cronin.* 1966a. Diphenamid: three-genera-
tion reproduction study in rats. Unpublished study. MRID 00076383.
Woodard M.W., G. Woodard and M.T. Cronin.* 1966b. Diphenamid safety evaluation
by dietary feeding to dogs for 103 weeks. Final Report. Unpublished
study. MRID 00076382.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
DRAFT
DISULFOTON
August, 1987
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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Di8Ulfoton August. 1987
-2-
II- GENERAL INFORMATION AND PROPERTIES
CAS No. 298-04-4
Structural Formula
0,0-Diethyl-S-[2-(ethylthio)-ethyl], phosphorodithioate
Synonyms
Si!11?*0?"11 Disyston; Oisystox; Dithiodemeton; Bayer 19639; Di-syston;
Ethyl thiometon; Frumin AL; M-74 (Meister, 1983).
Uses
Systemic insecticide-ac-ricide (Meister, 1983).
Properties (Meister, 1983; Windholz et al., 1983)
Chemical Formula CsHig02PS3
Molecular Height 274.38
Physical State (at 25»C) Pale yellow liquid
SiSS JXE l°8°c (0-01 -H9); 132 to 133°c (1-5 ran B^
Density (20°C) 1.144
Vapor Pressure (at 20°C) l.a x 10-4 mm Hg
Water Solubility (at 23"C) 25 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Disulfoton has been found in only 1 of the surface water samples
and none of the ground water samples analyzed from 835 samples
taken at 764 locations. (STORET, 1987).
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Disulfoton August, 1987
-3-
Environmental Fate
0 Disulfoton has a low mobility in Hugo sandy loam soil; 28% of the
pesticide applied to a 6-inch-high soil column was eluted with a
total of 110 feet of dilute buffer (McCarty and King, 1966). In
another study, disulfoton sulfoxide and disulfoton sulfone were more
mobile in sandy loam, clay loam and silty clay loam soils than the
parent compound. Aging 32p-disulfoton prior to elution increased the
adsorption to 10 to 20 times that of unaged 32p_
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Disulfoton August, 1987
-4-
III. PHARMACOKINETICS
Absorption
0 Puhl and Fredrickson (1975) administered by gavage single oral
doses of disulfoton-o-ethyl-l-Hc (99% purity) to Sprague-Dawley
rats (12/sex/dose). Males received 1.2 mg/kg and females received
0.2 mg/kg. In the 10 days following dosing, an average of 81.6,
7.0 and 9.2% of the dose was recovered in the urine, feces and
expired air, respectively. Males excreted 50% of the administered
dose in the urine in the first 4 to 6 hours; females required
30 to 32 hours. These data indicate that disulfoton is absorbed
readily from the gastrointestinal tract.
Distribution
0 In the study by Puhl and Fredrickson (1975), described above,
4.1 and 16.1% of the administered dose was detected in the livers of
males and females, respectively, and 0.4 and 1.2% of the dose was
detected in the kidneys of males and females, respectively, 48 hours
postdosing.
Metabolism
0 March et al. (1957) studied the metabolism of disulfoton in vivo and
.in vitro in mice (strain not specified). In the in vivo portion of
the study, mice received radiolabeled disulfoton intrapentoneally
(dose not specified). Results indicated that unspecified urinary
metabolites consisted mainly of hydrolysis products. In vitro
metabolism data indicated the presence of dithio-systox sulfoxide
and sulfone, and the thiol analog sulfoxide and sulfone. The dithio-
systox sulfoxide was present in the greatest quantity followed by
thiol analog sulfoxide, dithio-systox sulfone and thiol analog
sulfone. Based on a review of these data (U.S. EPA, 1984a), it was
concluded that the metabolism of disulfoton in mice involves at least
two reactions: (1) the sequential oxidation of the thioether sulfur
and/or oxidative desulfuration; and (2) hydrolytic cleavage of the
ester, producing phosphoric acid, thiophosphoric acid and dithio-
phosphoric acid.
0 In the above study by Puhl and Fredrickson (1975), the major urinary
metabolites detected in both sexes were diethylphosphate (DEP) and
diethylphosphorothioate (DEPT). These products were formed from
hydrolysis of disulfoton and/or its oxidation products. Minor urinary
metabolites included the oxygen analog sulfoxide, oxygen analog
sulfone and disulfoton sulfoxide.
Excretion
In the above study by Puhl and Fredrickson (1975), 96 to 99% of the
administered dose was recovered (81.6% in urine, 7.0% in feces and
9.2% as expired carbon dioxide during a 10-day postdosing period.
Excretory pathways were similar for males and females, but the rate
of excretion was slower for females.
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Disulfoton August, 1987
-5-
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 No significant anticholinesterase effects were observed in human
subjects (five test subjects, two controls) who received disulfoton
in doses of 0.75 mg/day (orally) for 30 days (Rider et al., 1972).
0 Quinby (1977) reported that three carpenters were sprayed accidentally
with disulfoton while the compound was being applied by airplane to
a wheat field adjacent to their work site. The individuals were
reexposed as they handled contaminated building materials in the
days following spraying. Exposure levels were not identified. The
older two carpenters experienced coronary attacks and one had at
least two severe cerebral vascular effects subsequent to exposure.
The author postulated that the effects may have been due to disturbances
of clotting mechanisms.
Long-term Exposure
0 No Long-term human studies were identified for disulfoton.
Animals
Short-term Exposure
0 Reported acute oral LD50 values for adult rats administered disulfoton
(approximately 94 to 96% purity when identified) ranged from 1.9 to
2.6 mg/kg for females and 6.2 to 12.5 mg/kg for males (Crawford and
Anderson,1973b; Bombinski and DuBois, 1957); a value of 5.4 mg/kg was
reported for weanling male rats (Brodeur and Dubois, 1963).
0 In guinea pigs, acute oral LD50 values ranged from 8.9 to 12.7 mg/kg
(Bombinski and Dubois, 1957; Crawford and Anderson, 1973a).
0 Mihail (1978) reported acute oral LD50 values of 7.0 mg/kg and 8.2
mg/kg in male and female NMRI mice, respectively.
0 Hixson (1982) reported that the acute oral LDsp of disulfoton (98%
pure) in white Leghorn hens was 27.5 mg/kg. Hixson (1983) reported
the results of an acute delayed neurotoxicity study in which 20 white
Leghorn hens were administered technical disulfoton (97.8% pure) by
gavage at a dose level of 30 mg/kg on two occasions, 21 days apart.
The study also employed live animals for each of the negative controls,
antidote controls and positive controls. Disulfoton did not produce
acute delayed neurotoxicity under the conditions of this study.
Based on this information, a No-Observed-Adverse-Effect-Level (NOAEL)
of 30 mg/kg (the only dose tested) was identified in this study.
8 Taylor (1965) reported the results of a demyelination study in which
white Leghorn hens (six/dose) were administered disulfoton in the diet
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Disulfoton August, 1987
-6-
for 30 days at concentrations of 0, 2, 10 or 25 ppm. Assuming that
1 ppm in the diet of hens is equivalent to 0.06 mg/kg/day (Lehman,
1959), these dietary levels correspond to doses of about 0, 0*1, 0.6
and 1.5 mg/kg/day. The author indicated that no evidence of demyelina-
tion was observed in any of the tissues examined. Based on this
information, a NOAEL of 1.5 mg/kg/day (the highest dose tested) was
identified.
Dermal/Ocular Effects
0 DuBois (1957) reported that the acute dermal LDso of technical
disulfoton in male Sprague-Dawley rats was 20 mg/kg. Mihail (1976)
reported acute dermal LDso values of 15.9 mg/kg and 3.6 mg/kg in male
and female wistar rats, respectively.
0 No information was found in the available literature on the effects
of ocular exposure to disulfoton.
Long-term Exposure
0 Hayes (1983) presented the results of a 23-month feeding study in
which CD-I mice (50/sex/dose) were administered disulfoton (98.2%
pure) at dietary concentrations of 0, 1, 4 or 16 ppm. Assuming that
1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day (Lehman,
1959), these dietary levels correspond to doses of about 0, 0.15, 0.6
and 2.4 mg/kg/day. No treatment-related effects were observed in*
terms of body weight, food consumption or hematology. A statistically
significant increase in mean kidney weight and kidney-to-body weight
ratio was noted in high-dose females; this increase may have been
associated with a nonsignificant increase in the incidence of malignant
lymphomas of kidneys in this group. Plasma, red blood cell and brain
cholinesterase (ChE) activity was decreased significantly in both
sexes at the highest dose tested (16 ppm). However, since ChE activity
was measured only in the control and high-dose groups, a NOAEL for
this effect could not be determined.
0 In a study by Hoffman et al. (1975), beagle dogs (four/sex/dose) were
administered disulfoton (95.7% pure) at dietary concentrations of 0,
0.5 or 1.0 ppm for 2 years. Assuming that 1 ppm in the diet of dogs
is equivalent to 0.025 mg/kg/day (Lehman, 1959), these dietary levels
correspond to doses of about 0, 0.0125 and 0.025 mg/kg/day. A fourth
group of animals received disulfcton in the diet at 2 ppm for 69
weeks, then 5 ppm for weeks 70 to 72, and finally 8 ppm from week 73
until termination (104 weeks); these doses correpond to 0.05, 0.125 and
0.2 mg/kg/day, respectively. No treatment-related effects were observed
in terms of general appearance, behavior, ophthalmoscopic examinations,
food consumption, body weight, organ weight, hematology, clinical
chemistry or histopathology. Additionally, no effects on ChE activity
were observed in animals that received 0.5 or 1.0 ppm (0.0125 or
0.025 mg/kg/day). However, exposure at 2.0 ppm (0.05 mg/kg/day)
for 69 weeks caused ChE inhibition in plasma and red blood cells in
both sexes. Maximum inhibition occurred at week 40, when males
exhibited 50% and 33% inhibition of Che in red blood cells and plasma;
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Disulfoton August, 1987
-7-
respectively; females exhibited 22 and 36% inhibition of ChE in red
blood cells and plasma, respectively. At a dose level of 8 ppm
(0.2 mg/kg/day), males exhibited 56 to 66* and 63 to 70% inhibition
of red blood cell and plasma ChE, respectively; females exhibited 46
to 53% and 54 to 64% inhibition of red blood cell and plasma ChE,
respectively. Based on these data, a NOAEL of 1.0 ppm (0.025 mg/kg/day)
was identified.
0 Carpy et al. (1975) presented the results of a 2-year feeding study
in which Sprague-Dawley rats (60/sex/dose) were administered disulfoton
(95.7% pure) at dietary concentrations of 0, 0.5, 1.0 or 2.0 ppm.
Based on data presented by the authors, these dietary levels correspond
to doses of about 0, 0.02, 0.05 and 0.1 mg/kg/day for males and 0,
0.03, 0.04 and 0.1 mg/kg/day for females. At week 81 of the study,
the 0.5-ppm dose was increased to 5.0 ppm (0.2 and 0.3 mg/kg/day for
males and females, respectively) since no adverse effects were observed
in the 1.0-ppm dose group. No treatment-related effects were observed
in terms of food consumption, body weight, hematology, clinical
chemistry, urinalysis and histopathology. A trend was observed at
all dose levels toward increased absolute and relative spleen, liver,
kidney and pituitary weights in males and toward decreased weights of
these organs in females. In males receiving 5 ppm, the increases
were statistically significant (p <0.05) for absolute spleen and
liver weights. In females receiving 5 ppm, the decrease in absolute
and relative kidney weights was statistically significant (p <0.05).
At all levels tested, the brain showed a trend toward decreased
absolute and relative weights in males and increased weights in
females. Additionally, statistically significant inhibition of
plasma, red blood cell and brain ChE was observed in both sexes at
2.0 and 5.0 ppm. At 1.0 ppm brain ChE in females was inhibited 11%
(p <0.01). Based on this information, a Lowest-Observed-Adverse-
Effect-Level (LOAEL) of 1.0 ppm (0.04 ng/kg/day for females) was
identified for ChE inhibition. It was concluded (U.S. EPA, 1984a)
that a NOAEL for systemic toxicity could not be identified due to the
inadequacy of histopathology and necropsy data.
0 Hayes (1985) presented the results of a 2-year feeding study in
which Fischer 344 rats (60/sex/dose) were administered disulfoton
(98.1% pure) at dietary concentrations of 0, 0.8, 3.3 or 13 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), these dietary levels correspond to doses of about
0, 0.04, 0.17 and 0.65 mg/kg/day. Mortality was generally low for
all groups with the exception of increased mortality in high-dose
females during the last week of the study. No compound-related
effects were observed in terms of clinical chemistry, hematology or
urinalysis. A dose-related trend in ChE inhibition was observed in
both sexes at all dose levels. Statistically significant inhibition
of plasma, red blood cell and brain ChE occurred in all dose groups
throughout the study. Histopathologically, a statistically significant
increase (p <0.05) in corneal neovascularization was observed in both
sexes at 13 ppm (0.65 mg/kg/day). A dose-related increase in the
incidence of optic nerve degeneration was also observed. This effect
was statistically significant (p <0.05) in mid-dose males and in mid-
and high-dose females. Additionally, a significantly (p <0.05)
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Disulfoton August, 1987
-8-
higher incidence of cystic degeneration of the Harderian gland was
observed in females at all doses and in mid-dose males. A significantly
(p <0.05) increased incidence of atrophy of the pancreas also was
observed in high-dose males. On the basis of ChE inhibition, this
study identified a LOAEL of 0.8 ppm (0.04 mg/Jcg/day) (lowest dose
tested).
Reproductive Effects
0 Taylor (1966) conducted a three-generation reproduction study in
which albino Holtzman rats (20 females and 10 males) were administered
disulfoton (98.5% pure) at dietary concentrations of 0, 2, 5 or 10 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), these dietary levels correspond to doses of about
0, 0.1, 0.25 and 0.5 mg/kg/day. At 10 ppm (0.5 mg/kg/day), litter
size was reduced by 21% in the Fa and 33% in the Fb in both the first
and third generations. Also in these generations, a 10 to 25% lower
pregnancy rate was noted for Fa ma tings. Histopathologically, F^
litters at 10 ppm (0.5 mg/kg/day) exhibited cloudy swelling and fatty
infiltration of the liver (both sexes), mild nephropathy in kidneys
(females) and juvenile hypoplasia of the testes. No histopathological
examinations were conducted on the 2- and 5-ppm dose groups.
Cholinesterase determinations revealed a 60 to 70% inhibition of red
blood cell ChE in F3b litters and their parents at 5 and 10 ppm (0.25
and 0.5 mg/kg/day). At 2 ppm (0.1 mg/kg/day), the inhibition was
insignificant in males and moderate (30 to 40%) in females. Based on
these data, a LOAEL of 2 ppm (0.1 mg/kg/day) was identified for ChE
inhibition. It was concluded (U.S. EPA, 1984a) that a reproductive
NOAEL could not be determined due to deficiencies in data reporting
(e.g., insufficient data on reproductive parameters, no statistical
analys.es, incomplete necropsy report and insufficient histopathology
data).
Developmental Effects
0 Lamb and Hixson (1983) conducted a study in which CO rats (25/dose)
were administered disulfoton (98.2% pure) by gavage at levels of 0,
0.1, 0.3 or 1 mg/kg/day on days 6 through 15 of gestation. Mean
plasma and red blood cell ChE activities were decreased significantly
in dams receiving 0.3 and 1 mg/kg/day. Examination of the fetuse^
after Cesarean section reflected no increases in the incidence of
soft tissue, external or skeletal abnormalities. However, at the
1,0-mg/kg/day dose level, increased incidences of incompletely ossified
parietal bones and sternebrae were observed. This is considered a
fetotoxic effect, since it is indicative of retarded development.
Based on the information presented in this study, a developmental
NOAEL of 0.3 mg/kg/day was identified based on fetotoxic effects. A
NOAEL of 0.1 mg/kg/day was identified for ChE inhibition in treated dams,
0 Tesh et al. (1982) conducted a teratogenicity study in which New
Zealand White rabbits were administered disulfoton (97.3% pure) at
initial doses of 0, 0.3, 1.0 or 3.0 mg/kg on days 6 through 18 of
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Disulfoton August, 1987
-9-
gestation. Due to mortality and signs of toxicity, the high dose was
reduced to 2.0 mg/kg/day and finally to 1.5 mg/kg/day. The control
group consisted of 15 animals, the low- and mid-dose groups consisted
of 14 does each and the high-dose group contained 22 animals. No
signs of maternal toxicity were observed in the low- or mid-dose
groups. In the high-dose group, signs of maternal toxicity included
muscular tremors, unsteadiness and incoordination, increased respiratory
rate and increased mortality. No compound-related effects on maternal
body weight or fetal survival, growth and development were observed.
Based on this information, a NOAEL of 1.0 mg/kg/day was identified for
maternal toxicity. The NOAEL for teratogenic and fetotoxic effects was
1.5 mg/kg/day (the highest dose tested).
Mutagenicity
0 Hanna and Dyer (1975) reported that disulfoton (99.3% pure) was
mutagenic in Salmonella typhimurium strains C 117, G 46, TA 1530 and
TA 1535, and in Escherichia coll strains WP 2, WP 2uvrA, CM 571,
CM 611, WP 67 and WP 12. These tests were performed without metabolic
activation; however, demeton, the major metabolite of disulfoton,
was also mutagenic in these microbial tests (U.S. EPA, 1984a).
0 Simmon (1979) presented the results of an unscheduled DNA synthesis
assay using human fibroblasts (W 138). Disulfoton (purity not specified)
was positive in this assay only in the absence of metabolic activation.
Carcinogenicity
0 Carpy et al. (1975) presented the results of a 2-year feeding study
in which Sprague-Dawley rats (60/sex/dose) were administered disulfoton
(95.7% pure) at dietary concentrations of 0, 0.5, 1.0 or 2.0 ppm.
Based on data presented by the authors, these dietary levels correspond
to doses of about 0, 0.02, 0.05 and 0.1 mg/kg/day for males and 0,
0.03, 0.04 and 0.1 mg/kg/day for females. At week 81 of the study,
the 0.5-ppm dose was increased to 5.0 ppm (reported to be equivalent
to 0.2 and 0.3 mg/kg/day for males and females, respectively) since
no adverse effects were observed in the 1.0-ppm dose group. The
number of tumor-bearing animals at all dose levels was comparable to
that of controls suggesting that, under the conditions of this study,
disulfoton is not oncogenic. However, a review of this study (U.S.
EPA, 1984a) concluded that due to numerous deficiencies (e.g., invalid
high dose, insufficient necropsy data, insufficient histology data),
the data presented were inadequate for an oncogenic evaluation.
0 Hayes (1983) presented the results of a 23-month feeding study in
which CD-I mice (50/sex/dose) were administered disulfoton (98.2%
pure) at dietary concentrations of 0. 1, 4 or 16 ppm. Assuming that
1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day (Lehman,
1959), these dietary levels correspond to doses of about 0, 0.15, 0.6
and 2.4 mg/kg/day. The incidence of specific neoplasms was similar
among treated and control animals. There was an increased incidence
of malignant lymphoma (the most frequently observed neoplastic lesion)
in both males and females at 16 ppm (2.4 mg/kg/day) when compared with
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Disulfoton August, 1987
-10-
controls, but this change was not statistically significant. Therefore,
under the conditions of this study, disulfoton was not oncogenic in
mice at dietary concentrations up to 16 ppm (2.4 mg/kg/day).
0 Hayes (1985) presented the results of a 2-year feeding study in
which Fischer 344 rats (60/sex/dose) were administered disulfoton
(98.1% pure) at dietary concentrations of 0, 0.8, 3.3 or 13 ppm,
corresponding doses of about 0, 0.04, 0.17 and 0.65 mg/kg/day (Lehman,
1959). The most commonly occurring neoplastic lesions included
leukemia, adenoma of the adrenal cortex, pheochromocytoma, fibroadenoma
of the mammary glands, adenoma and carcinoma of the pituitary glands,
interstitial cell adenoma of the testes, and uterine stromal polyps.
The incidences of these lesions showed no dose-related trend and were
not significantly different in treated versus control animals.
Therefore, under the conditions of this assay, disulfoton was not
oncogenic in male or female Fischer 344 rats at dietary concentrations
up to 13 ppm (0.65 mg/kg/day).
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 = or LOAEL) x (BW) = _ mg/L ( _ ug/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).
One-day Health Advisory
No suitable information was found in the available literature for the
determination of a One-day HA value for disulfoton. It is, therefore,
recommended that the Ten-day HA value for a 1 0-kg child of 0.01 mg/L (10 ug/L),
calulated below, be used at this time as a conservative estimate of the One-day
HA value.
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Disulfoton August, 1987
-11-
Ten-day Health Advisory
The developmental toxicity study by Lamb and Hixson (1983) has been
selected to serve as the basis for the Ten-day HA value for disulfoton.
In this study, CD rats were administered disulfoton (98.2% pure) by gavage
at doses of 0, 0.1, 0.3 or 1 mg/kg/day on days 6 through 15 of gestation.
Mean plasma and red blood cell ChE activities were decreased significantly
in dams receiving 0.3 and 1 mg/kg/day. Based on this information, a NOAEL of
0.1 mg/kg/day was identified. The only other study of comparable duration
was a rabbit teratology study (Tesh et al., 1982). This study identified
NOAELs of 1.0 mg/kg/day for maternal toxicity and 1.5 mg/kg/day (the highest
dose tested) for developmental toxicity. The rabbit appeared to be less
sensitive to disulfoton than the rat, therefore the rat study was selected
for this calculation.
Using a NOAEL of 0.1 mg/kg/day, the Ten-day HA for a 1 0-kg child is
calculated as follows:
Ten-dav HA = (0<1 *9/*9/&*y) <10 k9> = 0.01 mg/L (10 ug/L)
(100) (1 L/day)
where:
0.1 mg/kg/day = NOAEL, based on the absence of ChE effects in female
rats administered disulfoton by gavage on days 6
through 15 of gestation.
10 kg = assumed body weight of a child.
1 00 = 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 by a child.
Longer-term Health Advisory
The 2-year dog feeding study by Hoffman et al. (1975) has been selected
to serve as the basis for the Longer-term HA values for disulfoton. In this
study, beagle dogs were administered disulfoton (95.7% pure)
at dietary concentrations of 0, 0.5 or 1.0 ppm (0, 0.0125 and 0.025 mg/kg/day).
A fourth group of dogs received disulfoton at 2.0
ppm (0.05 mq/kg/day) for 69 weeks, then 5.0 ppm (0.125 mg/kg/day) for weeks
70 to 72, and finally 8.0 ppm (0.2 mg/kg/day) from weeks 73 to 104. Exposure
to 2.0 ppm (0.05 mg/kg/day) for 69 weeks caused plasma and red blood cell ChE
inhibition in both sexes. Brain ChE inhibition was also noted at termination
in this group. Based on this information, a NOAEL of 1.0 ppm (0.025 mg/kg/day)
was identified. No other suitable studies were available for consideration
for the Longer-term HA. Since the effects in the study by Hoffman et al. (1975)
were observed following 69 weeks of exposure, the study is considered to be
of appropriate duration for derivation of a Longer-term HA.
Using a NOAEL of 0.025 mg/kg/day, the Longer-term HA for a 10-kg child
is calculated as follows:
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Disulfoton August, 1987
-12-
Longer-term HA = (0.025 mg/kg/day) (10 kg) = Q.0025 mg/L (3 ug/L)
where:
0.025 mg/kg/day = NOAEL, based on the absence of ChE effects in dogs
administered disulfoton in the diet; ChE effects
were noted at the higher dose during the first 40
to 69 weeks of exposure and thereafter.
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.
Using a NOAEL of 0.025 mg/kg/days, the Longer-term HA for a 70-kg
adult is calculated as follows:
Longer-term HA = (0*025 mg/kg/day) (70 kg) = 0.0088 mg/L (9 ug/L)
(100) (2 L/day)
where:
0.025 mg/kg/day = NOAEL, based on the absence of ChE effects in dogs
administered disulfoton in the diet; ChE effects
were noted at the higher dose during the first 40
to 69 weeks of exposure and thereafter.
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. 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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Disulfoton Au9ust' 1987
-13-
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The studies by Hayes (1985) and Carpy et al. (1975) have been selected
to serve as the bases for the Lifetime HA values for disulfoton. Each of
these studies identifies a LOAEL of 0.04 mg/kg/day. In the Hayes (1985)
study, Fischer 344 rats were administered disulfoton (98.1% pure) at dietary
concentrations of 0, 0.8, 3.3 or 13 ppm (0, 0.04, 0.17 and 0.65 mg/kg/day)
for 2 years. Dose-related, statistically significant inhibition of ChE in
plasma, red blood cell and brain was observed in both sexes at all doses;
also, a dose-related optic nerve degeneration was observed in females. Based
on this information, a LOAEL of 0.04 mg/kg/day was identified. In the Carpy
et al. (1975) 2-year study, Sprague-Dawley rats were administered disulfoton
(95.7% pure) at dietary concentrations of 0, 0.5, 1.0 or 2.0 ppm (0, 0.02,
0.05 and 0.1 mg/kg/day for males and 0, 0.03, 0.04 and 0.1 mg/kg/day for
females). At week 81 of the study, the 0.5 ppm dose was increased to 5.0 ppm
(equivalent to 0.2 and 0.3 mg/kg/day for males and females, respectively).
Statistically significant inhibition of plasma and red blood cell ChE was
observed in both sexes at 2.0 and 5.0 ppm. Additionally, at 1 ppm (0.04
mg/kg/day), brain ChE was inhibited significantly (p <0.01) in females.
Since the initial low dose used in the study (0.5 ppm) was raised to 5.0 ppm,
the 1.0-ppm dose is the lowest dose tested and represents the study LOAEL.
Using a LOAEL of 0.04 mg/kg/day, the Lifetime HA is calculated as follows
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.04 mg/kg/day) _ Q.00004 mg/kg/day
(1,000)
where:
0.04 mg/kg/day = LOAEL, based on ChE inhibition an^ optic nerve
degeneration in rats exposed to disulfoton in the
diet for 2 years.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.00004 mg/kg/day) (70 kg) = Q.0014 mg/L (1 ug/L)
(2 L/day)
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Disulfoton August, 1987
-14-
where:
0.00004 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 (HA)
Lifetime HA = (0.0014 mg/L)(20%) = 0.0003 mg/L (0.3 ug/L)
where:
0.0014 mg/L - DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Three studies were available on the carcinogenicity of disulfoton.
The chronic study in rats by Carpy et al. (1975) was inadequate for
an oncogenic evaluation. The remaining two studies presented results
indicating that disulfoton was not carcinogenic in mice (Hayes, 1983)
or in rats (Hayes, 1985).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenicity of disulfoton.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), disulfoton may be classified in
Group E: no evidence of carcinogenicity in humans. This category is
used for substances that show no evidence of carcinogenicity in at least
two adequate animal tests or in both epidemiologic and animal studies.
However, disulfoton and its metabolites are mutagenic compounds (see
section on Mutagenicity).
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Academy of Sciences {NAS, 1977) has calculated an ADI of
0.0001 mg/kg/day, based on a NOAEL of 0.01 mg/kg/day from a subcnronic
dog feeding study on phorate (a closely related organophosphorus
insecticide) and an uncertainty factor of 100, with a Suggested-No-
Adverse-Response-Level (SNARL) of 0.0007 mg/L.
• The World Health Organization (WHO, 1976) has identified an ADI of
0.002 mg/kg/day based on chronic data from a 2-year chronic feeding
study in dogs (Hoffman et al., 1975) with a NOAEL of 0.025 mg/kg/day.
0 U.S. EPA Office of Pesticide Programs (OPP) has established residue
tolerances for disulfoton at 0.1 to 0.75 ppm in or on a variety of
raw agricultural commodities (U.S. EPA, 1985). At the present time,
these tolerances are based on a Provisional ADI (PADI) of 0.00004
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Disulfoton August, 1987
-15-
mg/kg/day. As for the RfD calculation, this PADI is calculated based
on a LOAEL of 0.8 ppm (0.04 mg/kg/day) for both ChE inhibition and
optic nerve degeneration that were identified in the 2-year rat
feeding study by Hayes (1985) and using a safety factor of 1,000.
VII. ANALYTICAL METHODS
0 Analysis of disulfoton is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus-containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 L of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
disulfoton, but it is estimated that the detection limits for
analytes included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature regarding treat-
ment technologies used to remove disulfoton from contaminated water.
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Disulfoton August, 1987
-16-
IX. REFERENCES
Bombinski, T.J., and K.P. Dubois.* 1957. The acute mammalian toxicity and
pharmacological actions of Di-Syston. Report No. 1732. Unpublished
study received Nov. 20, 1957 under 3125-58; prepared by Univ. of Chicago,
Dept. of Pharmacology, submitted by Mobay Chemical Corp., Kansas City,
MO. CDL:100153-B. NRID 00069347.
Brodeur, J., and K.P. Dubois.* 1963. Comparison of acute toxicity of
anticholinesterase insecticides to weanling and adult male rats. In
Proceedings of the Society for Experimental Biology and Medicine.
Vol. 114. New York: Academic Press, pp. 509-511. MRID 05004291.
Carpy, S., C. Klotzsche and A. Cerioli.* 1975. Disulfoton: 2-year feeding
study in rats: AGRO DOK CBK 1854/74. Report No. 47069. Unpublished
study received December 15, 1976 under 3125-58; prepared by Sandoz,
Ltd., Switzerland, submitted by Mobay Chemical Corp., Kansas City, MO.
CDL:095641-C. MRID 00069966.
Chemagro Corporation. 1969. Di-Syston soil persistence studies. Unpublished
study.
Crawford, C.R., and R.H. Anderson.* 1973a. The acute oral toxicity of Di-Syston
technical to guinea pigs. Report No. 39113. Unpublished study received
December 15, 1976 under 3125-58; submitted by Mobay Chemical Corp.,
Kansas City, MO. CDL:095640-F. MRID 00071872.
Crawford, C.R., and R.H. Anderson.* 1973b. The acute oral toxicity of several
Di-Syston metabolites to female and male rats. Report No. 39687.
Unpublished study received December 15, 1976 under 3125-58; submitted by
Mobay Chemical Corp., Kansas City, MO. CDL:095640-Gc MRID 00071873.
Doull, J.* 1957. The acute inhalation toxicity of Di-Syston to rats and
mice. Report No. 1802. Unpublished study received November 20, 1957
under 3125-58; prepared by Univ. of Chicago, Dept. of Pharmacology,
submitted by Mobay Chemical Corp., Kansas City, MO. CDL:100153-D.
Fiche Master ID 00069349.
DuBois, K.P.* 1957. The dermal toxicity of Di-Syston to rats. Report No.
2063. Unpublished study received January 23, 1958 under unknown admin.
no.; prepared by Univ. of Chicago, Dept. of Pharmacology, submitted by
Mobay Chemical Corp., Kansas City, MO. CDL:109216-8. MRID 00043213.
DuBois, K.P., and F.K. Kinoshita.* 1971. Effect of repeated inhalation
exposure of female rats to Di-Syston. Submitted 30571. Unpublished
study received November 30, 1971 under 3125-119; prepared by Univ. of
Chicago, Toxicity Laboratory, submitted by Mobay Chemical Corp., Kansas
City, MO. CDL:10059-A. MRID 00087935.
Flint, D.R., D.D. Church, H.R. Shaw and J. Armour II. 1970. Soil runoff,
leaching and adsorption, and water stability studies with Di-Syston:
Report No. 2899. Unpublished study submitted by Mobay Chemical Corporation
Kansas City, MO.
-------
Disulfoton August, 1987
-17-
Hanna, P.J., and K.F. Dyer. 1975. Mutagenicity of organophosphorus compounds
in bacteria and Drosophila. Mutat. Res. 28:405-420.
Hayes, R.H.* 1983. Oncogenicity study of disulfoton technical on mice. An
unpublished report of study No. 80-271-04 prepared by the Corporate
Toxicology Department, Mobay Chemical Corp., Stilwell, KS. Dated Aug. 10,
1983. MRID 0000000.
Hayes, R.H.* 1985. Chronic feeding/oncogenicity study of technical disulfoton
(Di-Syston) with rats. Unpublished study no. 82-271-01. Prepared by
Mobay Chemical Corp. Accession No. 258557.
Hixson, E.J." 1982. Acute oral toxicity of Di-Syston technical in hens. An
unpublished report (No. 341 ) prepared by the Environmental Health Research
Institute of Mobay Chemical Corp., Stilwell, KS. Study No. 82-018-01,
dated Oct. 25, 1982. MRID 00000000.
Hixson, E.J.* 1983. Acute delayed neurotoxicity study on disulfoton.
Toxicology Report No. 365 (Study No. 82-418-01) prepared by Agricultural
Chemicals Divison, Mobay Chemical Corp., Kansas City, MO., dated Mar. 7,
1983. MRID 00000000.
Hoffman, K. , C.H. Weischer, G. Luckhaus et al.* 1975. S 276 (Disulfoton)
chronic toxicity study on dogs (two-year feeding experiment). Report
No. 5618; Report No. 45287. Unpublished study received Dec. 15, 1976
under 3125-58; prepared by A.G. Bayer, w. Germany, submitted by Mobay
Chemical Corp., Kansas City, MO. CDL: 095640-N. MRID 00073348.
Kadoum, A.M., and D.E. Mock. 1978. Herbicide and insecticide residues in
tailwater pits: water and pit bottom soil from irrigated corn and
sorghum fields. J. Agric. Food Chem. 26:45-50.
Kawamori, I., T. Saito and K. lyatomi, 1971a. Fate of organophosphorus
insecticides in soils. Part I. Botyu-Kagaku. 36:7-12.
Kawamori, I., T. Saito and K. lyatomi. 1971b. Fate of organophosphorus
insecticides in soils. Part II. Botyu-Kagaku. 36:12-17.
Lamb, D.W., and E.J. Hixson.* 1983. Embryotoxic and teratogenic effects of
disulfoton. Unpublished study no. 81-611-02. Prepared by Mobay Chemical
Company .
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Assoc. Food Drug Off. U.S.
Lichtenstein, E. , K. Schulz, R. Skrentny and Y. Tsukano. 1966. Toxicity and
fate of insecticide residues in water: insecticide residues in water
after direct application or by leaching of agricultural soil. Arch.
Environ. Health. 12:199-212.
Loeffler, W.W. 1969. A summary of Dasanit and Di-Syston soil persistence
data: Report No. 25122. Unpublished study submitted by Mobay Chemical
Corporation, Kansas City, MO.
-------
Disulfoton August, 1987
-18-
March, R.B., T.R. Fukuto and R.L. Metcalf.* 1957. Metabolism of P-32-dithio-
systox in the white mouse and American cockroach: Submitter 1830.
Unpublished study. NRIO 00083215.
McCarty, P.L., and P.M. King. 1966. The movement of pesticides in soils.
Pages 156-171, _ln_ Proceedings of the 21st Industrial Haste Conference:
May 3-5, 1966, Lafayette, IN. Purdue University Engineering Extension
Series No. 121. pp. 156-176.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Mihail, P.* 1978. S 276 (Disyston active ingredient) acute toxicity studies.
Report No. 7602a prepared by A.G. Bayer, Institut Fur Toxikologie, for
Mobay Chemical Corp. June 12, 1978. MRID 00000000.
Mobay Chemical Corporation. 1964. Synopsis of analytical and residue infor-
mation on Di-syston (clover). Includes method dated Mar. 5, 1964.
Mobay Chemical Corporation. 1972. Dasanit - Di-syston: analytical and
residue information on tobacco. Includes methods dated Mar. 5, 1964;
Mar. 28, 1966; Oct. 27, 1967; and others. Unpublished study, including
published data.
NAS. 1977. National Academy of Sciences. Volume I. Drinking water and
health. Washington, DC: National Academy Press.
Puhl, R.J., and D.R. Fredrickson.* 1975. The metabolism and excretion of
Di-Syston by rats. Unpublished report submitted by Mobay Chemical Corp.,
Report No. 44261, prepared by Chemagro Agricultural Division, Mobay
Chemical Corp. Dated May 6, 1975. MRID 00000000.
Quinby, G.E. 1977. Poisoning of construction workers with disulfoton.,
Clin. Toxicol. 10:479.
Rider, J.A., J. I. Swader and E.J. Pulette. 1972. Anticholinesterase toxicity
studies with Guthion, Phosdrin, Di-syston and Trithion in human subjects.
Proc. Fed. Am. Soc. Exp. Biol. 31:520.
Simmon, V.F.* 1979. In vitro microbiological mutagenicity and unscheduled
DNA synthesis studies of eighteen pesticides. Report No. EPA-600/1-79-042.
Unpublished study including submitter summary, received April 3, 1980
under 279-2712/; prepared by SRI International, submitted by FMC Corp.,
Philadelphia, PA. CDL.-099350-A. MRID 00028625.
STORET. 1987.
Suett, D.L. 1975. Persistence and degradation of chlorfenvinphos, chlormephos,
disulfoton, phorate and. primiphos-ethyl following spring and late-summer
soil application. Unpublished study submitted by ICI Americas, Inc.,
Wilmington, DE.
-------
Disulfoton August, 1987
-19-
Taylor, R.E.* 1965. Letter sent to Chemagro Corporation dated Jan. 5, 1965:
Report on demyelination studies on hens. Report No. 15107. Unpublished
study received March 24, 1965 under 6F0478; prepared by Harris Labora-
tories, Inc., submitted by Mobay Chemical Corp., Kansas City, MO.
CDL:090534-C. MRID 00057265.
Taylor, R.E.* 1966. Letter sent to D. MacDougall dated May 5, 1966: Di-Syston,
three generation rat breeding studies: Submitter 18154. Unpublished
study received March 7, 1977 under 3125-252; prepared by Harris Labora-
tories, Inc., submitted by Mobay Chemical Corp., Kansas City, MO.
CDL:096021-L. MRID 00091104.
Tesh, J.M. et al.* 1982. S 276: Effects of oral administration upon
pregnancy in the rabbit. An unpublished report (Bayer No. R 2351)
prepared by Life Science Research, Essex, England and submitted to
A.G. Bayer, Wuppertal, Germany. Dated Dec. 22, 1982. MRID 00000000.
Thornton, J.S., J.B. Hurley, and J.J. Obrist. 1976. Soil thin-layer mobility
of twenty-four pesticide chemicals: Report No. 51016. Unpublished
study submitted by Mobay Chemical Corporation, Kansas City, MO.
U.S. EPA.* 1984a. U.S. Environmental Protection Agency. Disulfoton
(Di-Syston) Registration Standard. Washington, DC: Office of Pesticide
Programs.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.183. July 1, 1985.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1 -
Determination of nitrogen and phosphorus containing pesticides in ground
water by GC/NPD, January 1986 draft. Cincinnati, OH: U.S. EPA's Environ-
mental Monitoring and Support Laboratory.
WHO. 1976. World Health Organization. Pesticide Residues Series No. 5,
City, Country or State: World Health Organization, p. 204.
Windholz, M., S. Budavari, R.F. Blumetti, E.S. Otterbein, eds. 1983. The
Merck index — an encyclopedia of chemicals and drugs, 10th ed.
Rahway, NJ: Merck and Company, Inc.
•Confidential Business Information submitted to the Office of Pesticide
Programs
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August, 1987
DIURON
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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Diuron August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 330-54-1
Structural Formula
N'-(3,4-Dichlorophe.T/l)-N,N-di.Tiethylurea
Synonyms
0 Crisuron; Dailor.; Di-on; Dichlorfendism; Diurex, Drexel; Duran;
Dynex; DCMU; Herbatox; HW 920; Karmex; Sup'r flo; Telvar, Urox D;
Vonduron (Meister, 1983).
Uses
0 Pre-emergence herbicide (Meister, 1984).
Properties (Meister, 1984; Windholz et al., 1983)
Chemical Formula CgHjQ^OC^
Molecular Weight 233.10
Physical State (at 25°C) White crystalline solid
Boiling Point
Melting Point 158-159°C
Vapor Pressure (20°C) 3.1 x 10-6 nun Hg
Specific Gravity --
Water Solubility (25°C) 42 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor ~
Occurrence
0 Diuron has been found in none of the 8 surface water samples analyzed
and in 25 of 939 ground water samples (STORET, 1987). Samples were
collected at 6 surface water locations and 930 ground water locations,
and diuron was found only in California and Georgia. The 85th percentile
of all non-zero samples was 1 ug/L in ground water sources only. The
maximum concentration found in ground water was 5 ug/L.
0 Diuron residues as a result of agricultural practice have been detected
in ground waters in California in wells at low (e.g., 2 to 3 ppb)
levels (California Department of Food and Agriculture, 1986).
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Diuron August, 1987
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Environmental Fate
0 Radiolabeled diuron and its degradation products 3-(3,4-dichlorophenyl)-
1-methylurea (DCPMU) and 3-(3,4-dichlorophenyl)urea (DCPU) had half-lives
of 4 to 8, 5, and 1 month, respectively, in aerobic soils maintained
at 18 to 29°C and moisture levels at approximately field capacity
(Walker and Roberts, 1978; Elder, 1978). 3,4-Oichloroaniline (DCA)
was identified as a minor degradation product of diuron (Belasco,
1967; Belasco and Pease, 1969; Elder, 1978). Increasing soil organic
matter content appears to increase the rate of decline of diuron
phytotoxic residues (McCormick, 1965; Corbin and Upchurch, 1967;
McCormick and Hiltbold, 1966; Liu et al., 1970).
0 Degradation of diuron phytotoxic residues is much (28 to 50%) slower
in flooded soil than in aerobic soil (Imamliev and Bersonova, 1969;
Wang et al., 1977).
0 Diuron has a low-to-intermediate mobility in fine to coarse-textured
soils and freshwater sediments (Hance 1965a; Hance, 1965b; Harris and
Sheets, 1965; Harris, 1967; Helling and Turner, 1968; Grover and
Hance, 1969; Gerber et al., 1971; Green and Corey, 1971; Helling,
1971; Guth, 1972; Grover, 1975; Helling, 1975). Mobility is correlated
with organic matter content and (CEC). Soil texture apparently is not,
by itself, a major factor governing the mobility of diuron in soil.
0 In a study using radiolabeled material, the diuron degradation products
(96% pure) had Kd values of 66 and 115 in silty clay loam soils,
indicating that they are relatively immobile or less mobile than diuron
(Elder, 1978).
0 In the field, diuron residues (nonspecific method used) generally
persisted for up to 12 months in soils that ranged in texture from sand
to silt loam treated with diuron at 0.8 to 4 Ib/A (Cowart, 1954; Hill
et al., 1955; Weed et al., 1953; Weed et al., 1954; Miller et al.,
1978). These residues may leach in soil to a depth of 120 cm (4 feet).
Diuron was detectable (3 to 74 ppb) in runoff-water sediment and soil
samples for up to 3 years after the last application to pineapple-
sugarcane fields in Hawaii (Mukhtar, 1976; Green et al., 1977).
0 Phytotoxic residues persisted for up to 12 months in soils ranging in
texture from sand to silty clay loam tc boggy meadow soil following
the last of one to six annual applications of diuron at 1 to 18 Ib/A
(Weldon and Timmons, 1961; Dalton et al., 1965; Bowmer, 1972; Dawson
et al., 1978; Arle et al., 1965; Wang and Tsay, 1974; Spiridonov et al.,
1972; Addison and Bardsiey, 1968; Cowart, 1954; Hill et al., 1955;
Weed et al, 1953; Weed et al., 1954). Diuron persistence in soil
appears to be a function of application rate and amount of rainfall
and/or irrigation water. Three degradation products (DCPMU, DCPU,
and DCA) were identified in soil (planted to cotton) that had received
multiple applications of diuron (80% wettable powder totaling 5 to 5.7
Ib/A (Dalton et al., 1965).
0 Diuron persists in irrigation-canal soils for 6 or more months following
application at 33 to 46 kg/ha (Evans and Duseja, 1973a; Evans and
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Diuron August, 1987
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Duseja, 1973b; Bowmer and Adeney, 1978a; Bowmer and Adeney,
1978b). The relative percentages of diuron and its degradates DCPMU
and DCPU were 60-90:10-25:1-30 in clay and sandy clay soils, 4.5 to 17
weeks after treatment. Oiuron levels in water samples were highest
(0.5 to 8 ppm) in the initial flush of irrigation water. These levels
declined rapidly, probably as a function of dilution and not degradation.
III. PHARMACOKINETICS
Absorption
0 Diuron is absorbed through the gastrointestinal tract of rats and dogs.
Hodge et al. (1967) fed diuron to rats and dogs at dietary levels
from 25 to 2,500 ppm and from 25 to 1,250 ppm active ingredient (a.i.),
respectively, for periods up to two years. These doses are equivalent
to 1.25 to 125 mg/kg/day for the rat and 0.635 to 31.25 mg/kg/day for
the dog. Urinary and fecal excretion products after one week to 2
years accounted for about 10% of the daily dose ingested. The
excretion data provided evidence that gastrointestinal absorption
ocurred in rats and dogs.
Distribution
0 Hodge et al. (1967) fed diuron (80% wettable powder) for 2 years
to rats at dietary levels of 25 to 2,500 ppm a.i. and to dogs at
dietary levels of 25 to 1,250 ppm a.i. Assuming that 1 ppm in the
diet is equivalent to 0.05 mg/kg/day in rats and 0.025 mg/kg/day in
dogs, this corresponds to doses of 1.25 to 125 mg/kg/day in rats and
0.625 to 31.25 mg/kg/day in dogs (Lehman, 1959). Analysis of tissue
samples for diuron residues revealed levels ranging from 0.2 to 56 ppm,
depending on dose. This constituted only a minute fraction of
the total dose ingested. The authors concluded that there was little
diuron storage in tissues.
Metabolism
0 Geldmacher von Mallinckrodt and Schlussier (1971) analyzed the urine
of a woman who had ingested a dose cf 38 mg/kg of diuron along with
20 mg/kg of aminotriazole. The urine was found to contain
1-(3,4-dichlorophenyl)-3-methylurea and 1-(3,4-dichloropheny1 )-urea,
and may also have contained some 3,4-dichloroaniline. No unaltered
diuron was detected.
0 Hodge et al. (1967) fed diuron (80% wettable powder) to male beagle
dogs at a dietary level of 125 ppm active ingredient for 2 years.
Assuming that 1 ppm in the diet is equivalent to 0.025 mg/kg/day
(Lehman, 1959), this corresponds to a dose of 3.1 mg/kg/day. Analysis
of urine at weeks one to four or after two years revealed the major
metabolite was N-(3,4-dichlorophenyl)-urea. Small amounts of
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Diuron August, 1987
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N-(3,4-dichlorophenyl)-N'-methylurea, 3,4-dichloroanaline,
3,4-dichlorophenol and unmetabolized diuron also were detected.
Excretion
0 Hodge et al. (1967) fed diuron (80% wettable powder) for 2 years
to rats at dietary levels of 25 to 2,500 ppm and to dogs at dietary
levels of 25 to 1,250 ppm. Assuming that 1 ppm in the diet is equivalent
to 0.05 mg/kg/day in rats and 0.025 mg/kg/day in dogs, this corresponds
to doses of 1.25 to 125 mg/kg/day in rats and 0.625 to 31.25 mg/kg/day
in dogs (Lehman, 1959). In rats, urinary excretion (6.3 to 492 ppm,
depending on dose) was consistently greater than fecal excretion
(1.0 to 204 ppm). In dogs, urinary excretion (6.3 to 307 ppm) was
similar to fecal excretion (7.9 to 308 ppm). For both rats and dogs,
combined urinary and fecal excretion accounted for only about 10% of
the daily diuron ingestion.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of diuron in humans.
Animals
Short-term Exposure
0 Acute oral LD50 values of 1,017 mg/kg and 3,750 mg/kg have been
reported in albino rats by Boyd and Krupa (1970), NIOSH (1985) and
Taylor (1976a), respectively. Signs of central nervous system
depression were noted after treatment.
0 Hodge et al. (1967) administered single oral doses of recrystallized
diuron in peanut oil to male CR rats. The approximate lethal dose was
5,000 mg/kg, and the LD5Q was 3,400 mg/kg. Survivors sacrificed after
14 days showed large and dark-colored spleens with numerous foci of
blood formation.
0 Hodge et al. (1967) administered oral doses of 1,000 mg/kg of
recrystallized diuron five times a week for 2 weeks (10 doses) to
six male CR rats. At necropsy, 3 or 11 days after the final dose,
the spleens were large, dark and congested, and foci of blood formations
were noted in both the spleen and bone marrow.
0 Hodge et al. (1967) fed Wistar rats (five/sex/dose) diuron (purity
not specified) in the diet for 42 days at dose levels of 0, 200, 400,
2,000, 4,000 or 8,000 ppm a.i. Assuming that 1 ppm in the diet is
equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of 0, 10, 20, 100, 200 or 400 mg/kg/day. Following treatment
body weight, clinical chemistry, food consumption, hematology,
urinalysis and histology were evaluated. No effects were observed at
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Diuron August, 1987
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400 ppm (20 ing/kg/day) or less. At 2,000 ppm (100 mg/kg/day) or
greater, red blood cell counts and hemoglobin values were decreased.
A marked inhibition of growth occurred in the 4,000 ppm (200 mg/kg/day)
or greater dosage groups, and there was increased mortality at 8,000
ppm. Based on these data, a No-Observed-Adverse-Effect-Level (NOAEL)
of 400 ppm (20 mg/kg/day) and a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 2,000 ppm (100 mg/kg/day) were identified.
Dermal/Ocular Effects
0 Taylor (1976b) applied diuron (98% pure) to the intact or abraded skin
of eight albino rabbits at dose levels of 1,000 to 2,500 mg/kg for 24
hours. After treatment, a slight erythema was observed, but no other
symptoms of toxicity were noted during a 14-day observation period.
The dermal LD5Q was reported as >2,500 mg/kg.
0 Larson (1976) applied diuron (98% pure) at doses of 1, 2.5, 5 or 10
mg/kg to intact abraded skin of rabbits for 24 hours. Adverse effects
were not detected in exposed animals.
0 In studies conducted by DuPont (no date), diuron (50% water paste)
was not irritating to intact skin and was moderately irritating to
abraded skin of guinea pigs. No data were available on skin
sensitization. See also DuPont (1961).
0 In studies conducted by Larson and Schaefer (1976), 10 mg of a fine
dry powder of diuron (98% a.i.) was instilled into the conjunctival
sac of one eye of each of six New Zealand White rabbits. Ocular
irritation was not detected within 72 hours.
Long-term Exposure
0 Hodge et al. (1967) fed albino Charles River rats (five/sex/dose)
diuron (98%'pure) for 90 days at dietary levels of 0, 50, 500 or
5,000 ppm. Assuming that 1 ppm in the diet is equivalent to 0.05
nig/kg/day (Lehman, 1959), this corresponds to doses of 0, 2.5, 25 or
250 mg/kg/day. Following treatment, body weight, food consumption,
clinical chemistry and histopathology were evaluated. No adverse
effects were observed in any parameter at 50 ppm. At 500 ppm there
were no effects on males, but females gained less weight than controls
an.1 appeared cyanotic. At the 5,000-ppm dose level, body weights
were reduced in both sexes, spleens were enlarged and exhibited
hemosiderosis, and there was clinical and pathological evidence of
chronic methemoglobinemia. Based on these data, a NOAEL of 50 ppm
(2.5 mg/kg/day) and a LOAEL of 500 ppm (25 mg/kg/day) were identified.
0 Hodge et al. (1967) fed diuron (80% wettable powder) to groups of
Cnarles River rats (20/sex/dose) for 90 days at dietary levels of 0,
250 or 2,500 ppm active ingredient. Assuming that 1 ppm in the diet
is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of 0, 12.5 or 125 mg/kg/day. At 2,500 ppm, both males and
females ate less and gained less weight did than controls. There was
a slight decrease in red blood cell count, greater in females than in
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Diuron August, 1987
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males. No effect on food consumption or weight gain- was noted at
250 ppm, but hematological changes were evident in females. This
study identified a LOAEL of 250 ppm (12.5 mg/kg/day), the lowest
dose tested.
0 In a 2-year feeding study conducted by Hodge et al. (1964a, 1967),
beagle dogs (two males/dose and three females/dose) were administered
technical diuron (80% a.i.) in the diet at dose levels of 0, 25, 125,
250 or 1,250 ppm active ingredient. Assuming that 1 ppm in the diet
of dogs is equivalent to 0.025 mg/kg/day (Lehman, 1959), this corresponds
to doses of diuron of 0, 0.625, 3.12, 6.25 or 31.25 mg/kg/day.
Following treatment, body weight, clinical chemistry, hematology,
organ weight, gross pathology and histopathology were evaluated. No
adverse effects were reported at 25 ppm in any parameter measured.
Abnormal blood pigment was observed at 125 ppm or greater. .Hemato-
logical alterations (depressed red blood cells (RBC), hematocrit and
hemoglobin) were observed at 250 ppm or greater. In the 1,250 ppm
group, a slight weight loss occurred as well as increased erythrogenic
activity in bone marrow and hemosiderosis of the spleen. Based on
these data, a NOAEL of 25 ppm (0.625 mg/kg/day) and a LOAEL of 125 ppm
(3.12 mg/kg/day) were identified.
0 Hodge et al. (1964b, 1967) administered technical diuron (80% a.i.)
in the diet of rats (35/sex/dose) for 2 years at dose levels of 0,
25, 125, 250 or 2,500 ppm active ingredient. Assuming that 1 ppm in
the diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), this
corresponds to doses of diuron of 0, 1.25, 6.25, 12.5 or 125 mg/kg/day.
Following treatment, body weight, clinical chemistry, hematology,
food consumption, urinalysis, organ weights and histopathology were
evaluated. No adverse effects were reported at 25 ppm (1.25 mg/kg/day)
for any parameters measured. Abnormal blood pigments (sulfhemoglobin)
were observed at 125 ppm (6.25 mg/kg/day) or greater. Hematological
changes (decreased RBC, reduced hemoglobin), growth depression,
hemosiderosis of the spleen and increased mortality were observed at
250 ppm (12.5 mg/kg/day) or greater. Based on these data, a NOAEL of
25 ppm (1.25 mg/kg/day) and a LOAEL of 125 ppm (6.25 mg/kg/day) were
identified.
Reproductive Effects
0 Hodge et al. (1964b, 1967) studied the effects of diuron (80% wet-
table powder) in a three-generation reproduction study in rats.
Animals were supplied food containing 125 ppm active ingredient.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/
day (Lehman, 1959), this corresponds to a dose of 6.25 mg/kg/day.
Fertility rate, body weight, hematology and histopathology were
monitored. No effect was seen on any parameter except body weight,
which significantly decreased in the ?2b anc* ?3a litters. A LOAEL
of 125 ppm (6.25 mg/kg/day) was identified.
Developmental Effects
0 Khera et al. (1979) administered by gavage a formulation containing
80% diuron at dose levels of 125, 250 or 500 mg/kg of formulation to
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Diuron August, 1987
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pregnant Wistar rats (14 to 18/dose) on days 6 through 15 of gestation.
Vehicle (corn oil) controls (19 dams) were run concurrently. No
maternal or teratogenic effects were observed at 125 mg/kg/day.
Developmental effects appeared to increase in all treatment groups,
ioe. wavy ribs, extra ribs and delayed ossification. The incidence
of wavy ribs was statistically significant at 250 mg/kg and greater.
Maternal and fetal body weights decreased significantly at 500 mg/kg
(p <0.05). A NOAEL was not determined from this study for fetotoxic
effects; hence, a LOAEL of 125 mg/kg of formulation per day was
identified.
Mutagenicity
0 Andersen et al. (1972) reported that diuron did not exhibit mutagenic
activity in T4 bacteriophage test systems (100 ug/plate) or in tests
with eight histidine-requiring mutants of Salmonella typhimurium
(small crystals applied directly to surface of plate).
0 Fahrig (1974) reported that diuron (purity not specified) was not
mutagenic in a liquid holding test for mitotic gene conversion in
Saccharomyces cerevisiae, in a liquid holding test for forward mutation
to streptomycin resistance in Escherichia coli, in a spot test for
back mutation in £. marcescens or in a spot test for forward mutation
in _E. coli.
0 Recent studies by DuPont (1985) did not detect evidence of mutagenic
activity for diuron in reversion tests in several strains of _S_.
typhimurium (with or without metabolic activation), in a Chinese
hamster ovary/hypoxanthine guanine phosphoribosyl-transferase (CHO/HGPRT)
forward gene mutation test or in unscheduled DNA synthesis tests in
primary rat hepatocytes. However, in an in vivo cytogenetic test in
rats, diuron was observed to cause clastogenic effects.
Carcinogenicity
e Hodge et al. (1964b, 1967) fed Wistar rats (35/sex/dose) diuron (80%
wettable powder) in the diet at levels of 0, 25, 125, 250 or 2,500 ppm
a.i. for 2 years. Assuming that 1 ppm in the diet of rats corresponds
to 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of 0,
1.25, 12.5 or 125 mg/kg/day. There was some early mortality in males
at 250 and 2,500 ppm, but the authors ascribed this to viral infection.
Histological examination of tissues showed no evidence of changes
related to diuron; however, only 10 animals or fewer were examined
per group. Tumors and neoplastic changes observed were similar in
exposed and control groups, and the authors concluded there was no
evidence that diuron was carcinogenic in 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
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Diuron August, 1987
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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 j
(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 leg) or
an adult (70 kg).
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).
One-day Health Advisory
No suitable information was found in the available literature for use in the
determination of the One-day HA value for diuron. It is, therefore, recommended
that the Ten-day HA value for a 10-kg child, calculated below as 1.0 mg/L
(1,000 ug/L) be used at this time as a conservative estimate of the One-day
HA value.
Ten-day Health Advisory
The study by Khera et al. (1979) has been selected to serve as the
basis for the Ten-day HA for diuron. In this study, pregnant rats were
administered diuron (80%) on days 6 through 15 of gestation at dose levels
of 125, 250 or 500 mg/kg/day. Developmental effects appeared to increase in
the diuron-treated groups as compared to the control group, i.e. wavy ribs,
extra ribs and delayed ossification. The incidence of wavy ribs was
statistically significant at 250 mg/kg/day (p <0.05). Fetal and maternal
body weights were decreased at 500 mg/kg (p <0.05). A NOAEL was not determined
from this study at the lowest dose tested (LOT) based on developmental toxicity;
hence, the LOAEL for this study was 125 mg/kg/day (LOT).
Using a LOAEL of 125 m'7/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-Day HA = (125 mq/kg/day) (10 kg) (0.80) = uo mg/L (1,000 ug/L)
(1,000) (1 L/day)
where:
125 mg/kg/day = LOAEL, based on fetotoxicity in rats exposed to
diuron via the diet for days 6 through 15 of gestation.
10 kg = assumed body weight of a child.
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Diuron August, 1987
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0.80 = correction factor to account for 80% active ingredient.
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
The 90-day feeding study in rats by Hodge et al. (1967) has been chosen
to serve as the basis for determination of the Longer-term HA values for diuron.
In this study, five animals per sex were fed diuron (98% pure) at dose levels
of 0, 2.5, 25 or 250 mg/kg/day. Based on decreased weight gain and
methemoglobinemia, this study identified a NOAEL of 2.5 mg/kg/day and a LOAEL
of 25 mg/kg/day. These values are supported by the 42-day feeding study of
Hodge et al. (1964b), in which a NOAEL of 20 mg/kg/day and a LOAEL of 100
mg/kg/day were identified. This study was not selected, however, since the
duration of exposure was only 42 days.
Using a NOAEL of 2.5 mg/kg/day, the Longer-term HA for a 10-kg child
is calculated as follows:
Lonqer-term HA - (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (250 ug/L)
(100) (1 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of effects on weight gain or
blood chemistry in rats exposed to diuron via the
diet for 90 days.
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 a 70-kg adult is calculated as follows:
Lon.-or-term HA - (2.5 mg/kg/day) (70 kg) = 0.875 mg/L (875 ug/L)
(100) (2 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of effects on weight gain or
blood chemistry in rats exposed to diuron via the
diet for 90 days.
70 kg = assumed body weight of an adult.
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Diuron August, 1987
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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.
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study in dogs by Hodge et al. (1964a, 1967) has been
selected to serve as the basis for the Lifetime HA for diuron. In this
study, dogs (three/sex/dose) were fed diuron at doses of 0.625, 3.12, 6.25 or
31.15 mg/kg/day of active ingredient. Hematological alterations were observed
at 3.12 mg/kg/day or greater, and this was identified as the LOAEL. No effects
were reported at 0.625 mg/kg/day in any parameter measured, and this was
identified as the NOAEL. This value is supported by a lifetime study in rats
by the same authors (Hodge et al., 1964b). In this study, rats were fed
diuron at dose levels of 0, 1.25, 6.25, 12.5 or 125 mg/kg/day for 2 years.
Hematological changes were observed at 6.25 mg/kg/day or greater, and a NOAEL
of 1.25 mg/kg/day was identified.
Using a NOAEL of 0.625 mg/kg/day, the Lifetime KA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0-625 mg/kg/day) = 0.002 mg/kg/day
(100) (3)
where:
0.625 mg/kg/day = NOAEL, based on absence of hematological effects in
dogs exposed to diuron via the diet for 2 years.
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Diuron August, 1987
-12-
100 o uncertainty factor chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = additional uncertainty factor used in the Office of
Pesticide Programs (U.S. EPA, 1987). This factor
is used to account for a lack of adequate chronic
toxicity studies in the data base preventing estab-
lishment of the most sensitive toxicological end
point.
Step 2: Determination of the Drinking Water Equivalent Level (OWED
DWEL = (0*002 mg/kg/day) (70 kg) „ 0.07 mg/L (70 ug/L)
(2 L/day)
where:
0.002 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 = (0.07 mg/L) (20%) = 0.014 mg/L (14 ug/L)
where:
0.07 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Hodge et al. (1964b, 1967) fed rats (35/sex/dose) diuron in the diet
at ingested doses of up to 125 mg/kg/day for 2 years. Histological
examinations did not reveal increased frequency of tumors; however,
fewer than half of the survivors were examined.
0 The International Agenry for Research on Cancer has not evaluated the
carcinogenic potential of diuron.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), diuron may be classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
0 Structurally related analogue(s) (e.g., linuron) of diuron appears to
reflect some oncogenic activity. In addition, a Russian study by
Rubenchik et. al. (1973) reported gastric carcinomas and pancreatic
adenomas in rats (strain not designated) given 450 mg/kg/ day for
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Diuron August, 1987
-13-
22 months. However, the actual data for the study is unavailable
for Agency review.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 An Acceptable Daily Intake (ADI) of 0.002 mg/kg/day, based on a
NOAEL of 0.625 rag/kg from a dog study and an uncertainty factor of
300 has been calculated (U.S. EPA, 1986b).
0 Residue tolerances have been established for diuron in or on agricul-
tural commodities that range from 0.1 to 7 ppm (U.S. EPA, 1985).
VII. ANALYTICAL METHODS
0 Analysis of diuron is by a high-performance liquid chromatographic
(HPLC) method applicable to the determination of certain carbamate
and urea pesticides in water samples (U.S. EPA, 1986c). This method
requires a solvent extraction of approximately 1 L of sample with
methylene chloride using a separatory funnel. The methylene chloride
extract is dried and concentrated to a volume of 10 mL or less. HPLC
is used to permit the separation of compounds, and measurement is
conducted with an ultraviolet (UV) detector. The method detection
limit has not been determined for diuron, but it is estimated that the
detection limits for analytes included in this method are in the
range of 1 to 5 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) and
powdered activated carbon (PAC) adsorption and chlorination effectively
remove diuron from water.
0 El-Dib and Aly (1977b) determined experimentally the Freundlich
constants for diuron on GAC. Although the values do not suggest a
strong adsorption affinity for activated carbon, diuron is better
adsorbed than other phenylurea pesticides.
0 El-Dib and Aly (1977b) calculated, based on laboratory tests, that
66 mg/L of PAC would be required to reduce diuron concentration by
98%, and 12 mg/L of PAC to reduce diuron concentration by 90%.
0 Conventional water treatment techniques of coagulation with ferric
svlfate, sedimentation and filtration proved to be only 20% effective
in removing diuron from contaminated water (El-Dib and Aly, 1977a).
Aluminum sulfate was reportedly less effective than ferric sulfate.
0 Oxidation with chlorine for 30 minutes removed 70% of diuron at a pH 7.
Under the same conditions, 80% of diuron was oxidized by chlorine
dioxide (EL-Dib and Aly, 1977a). Chlorination, however, will produce
several degradation products whose environmental toxic impact should
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Diuron August, 1987
-14-
be evaluated prior to selection of oxidative chlorination for treatment
of diuron-contaminated water.
The treatment technologies cited above for the removal of diuron from
water are available and have been reported to be effective. However,
selection of individual or combinations of technologies to attempt
diuron removal from water must be based on a case-by-case technical
evaluation and an assessment of the economics involved.
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Diuron August, 1987
-15-
IX. REFERENCES
Addison, D.A., and C.E. Bardsley. 1968. Chlorella vulgaris assay of the
activity of soil herbicides. Weed Sci. 16:427-429.
Andersen, K.J., E.G. Leighty and M.T. Takahasi. 1972. Evaluation of herbi-
cides for possible mutagenic properties. J. Agr. Food Chem. 20:649-656.
Arle, H.F., J.H. Miller and T.J. Sheets. 1965. Disappearance of herbicides
from irrigated soils. Weeds. 13(1):56-60.
Belasco, I.J. 1967. Absence of tetrachloroazobenzene in soils treated with
diuron and linuron. Unpublished study submitted by E.I. du Pont de
Nemours & Company, Inc., Wilmington, DE.
•
Belasco, I.J., and H.L. Pease. 1969. Investigation of diuron- and linuron-
treated soils for 3,3',4,4'-tetrachloroazobenzene. J. Agric. Food
Chem. 17:1414-1417.
Bowmer, K.H. 1972. Measurement of residues of diuron and simazine in an
orchard soil. Aust. J. Exp. Agric. Anim. Husb. 12(58):535-539.
Bowmer, K.H., and J.A. Adeney. 1978a. Residues of diuron and phytotoxic
degradation products in aquatic situations. I. Analytical methods for
soil and water. Pestic. Sci. 9(4):342-353.
Bowmer, K.H., and J.A. Adeney. I978b. Residues of diuron and phytotoxic
degradation products in aquatic situations. II. Diuron in irrigation
water. Pestic. Sci. 9(4): 354-364.
Boyd, E.M. and V. Krupa. 1970. Protein deficient diet and diuron toxicity.
J. Agric. Food Chem., 18:1104-1107.
Corbin, F.T., and R.P. Upchurch. 1967. Influence of pH on detoxication of
herbicides in soil. Weeds. 15(4 ):370-377.
Cowart, L.E. 1954. Soil-herbicidal relationships of 3-(£-chlorophenyD-
1,1-dimethylurea and 3-(3,4-dichlorophenyl)-l,l-dimethylurea. In
Proceedings of the Western Weed Control Conference. Vol. 14.
Salt Lake City, UT: Western Weed Control Conference, pp. 37-45.
Dalton, R.L., A.H. Evans and R.C. Rhodes. 1965. Disappearance of diuron in
cotton field soils, ^n Proceedings of the Southern Weed Conference.
Vol. 18. Athens, GA: Southern Weed Science Society, pp. 72-78.
Dawson, J.H., V.G. Bruns and W.J. Clore. 1968. Residual monuron, diuron,
and simazine in a vineyard soil. Weed Sci. 16(1):63-65.
DuPont.* 1961. E. I. du Pont de Nemours & Co., Inc. Condensed technical
information (Diuron).
DuPont.' No date. E. I. du Pont de Nemours & Co., Inc. Toxicity of
3-(3,4-dichlorophenyl)-1,1-dimethylurea. Medical Research Project Nos.
MR-48 and MR-263. Unpublished study. MRID 00022036.
-------
Diuron August, 1987
-16-
DuPonto* 1985. E. I. du Pont de Nemours & Co., Inc. Mutagenicity studies
with diuron. Salmonella test, No. HLR 471-84 (7185); CHO/HGPRT forward
gene mutation assay, HR No. 282-85 (06/28/85); Unscheduled DNA synthesis
test in primary rat hepatocytes, HLR No. 349-85 (07/10/85); and in vivo
cytogenetic test. No. 36685 (06/20/85).
Elder, V.A. 1978. Degradation of specifically labeled diuron in soil and
availability of its residues to oats. Doctoral dissertation. Honolulu,
HZ: University of Hawaii. Available from: University Microfilms,
Ann Arbor, MI. Report No. 79-13776.
El-Dib, M.A., and O.A. Aly 1977a. Removal of phenylamide pesticides from
drinking waters. I. Effect of chemical coagulation and oxidants.
Water Res. 11:611-616.
El-Dib, M.A., and O.A. Aly. 1977b. Removal of phenylamide pesticides from
drinking waters. II. Adsorption on powdered carbon. Water Res.
11:617-620.
Evans, J.O., and D.R. Duseja. 1973a. Herbicide contamination of surface
runoff waters. Washington, DC: U.S. Environmental Protection Agency,
Office of Research and Monitoring. EPA-R2-73-266; available from National
Technical Information Service, Springfield, VA. PB-222283.
Evans, J.O., and D.R. Duseja. 1973b. Results and discussion: Field experi-
ments. In Herbicide contamination of surface runoff waters. Utah State
University, pp. 33-35, 38-43. EPA-R2-73-266; project no. 13030 FDJ;
available from Superintendent of Documents, U.S. Government Printing
Office, Washington, DC.
Fahrig, R. 1974. Comparative mutagenicity studies with pesticides.
International Agency for Research on Cancer (IARC), Lyon, France.
Sci. Pub. 10. pp. 161-181.
Geldmacher von Mallinckrodt, M., and F. Schlussier.* 1971. Metabolism and
toxicity of 1-{3,4-dichlorophenyl)-3,3-dimethylurea (diuron) in man.
Arch. Toxicol. 27(3):31 1-314. Cited in Weed Abst. 21:331.
MRID 00028010.
Gerber, H.R., P. Ziegler and P. Dubah. 1971. Leaching as a tool in the
evaluation of herbicides. Iri Proceedings of the 10th British Weed
Control Conference (1970).. Vol. 1. Droitwich, England: British Weed
Control Conference, pp. 118-125.
Green, R.E., and J.C. Corey. 1971. Pesticide adsorption measurement by flow
equilibration and subsequent displacement. Proc. Soil Sci. Soc. Am.
35:561-565.
Green, R.E., K.P. Goswami, M. Mukhtar and H.Y. Young. 1977. Herbicides
from cropped watersheds in stream and estuarine sediments in Hawaii.
J. Environ. Qual. 6(2):145-154.
Grover, R. 1975. Adsorption and desorption of urea herbicides on soils.
Can. J. Soil Sci. 55:127-135.
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Diuron August, 1987
-17-
Grover, R., and R.J. Hance. 1969. Adsorption of some herbicides by soil and
roots. Can. J. Plant Sci. 40:378-380.
Guth, J.A. 1972. Adsorption and leaching characteristics of pesticides in
soil. Unpublished study including German test, prepared by Ciba-Geigy,
AG, submitted by Shell Chemical Company, Washington, DC.
Hance, R.J. I965a. Observations on the relationsip between the adsorption
of diuron and the nature of the adsorbent. Weed Res. 5:108-114.
Hance, R.J. 1965b. The adsorption of urea and some of its derivatives by a
variety of soils. Weed Res. 5:98-107.
Harris, C.I. 1967. Movement of herbicides in soil. Weeds. 15(3):214-216.
Harris, C.I., and T.J. Sheets. 1965. Influence of soil properties on
adsorption and phytotoxicity of CIPC, diuron, and simazine. Weeds.
13(3):215-219.
Helling, C.S. 1971. Pesticide mobility in soils: II. Applications of soil
thin-layer chromatography. Proc. Soil Sci. Soc. Am. 35:737-748.
Helling, C.S. 1975. Soil mobility of three Thompson-Hayward pesticides.
Interim Report. U.S. Agricultural Research Service, Pesticide Degradation
Laboratory; unpublished study.
Helling, C.S., and B.C. Turner. 1968. Pesticide mobility: Determination by
soil thin-layer chromatography. Method dated Nov. 1, 1968. Science.
162:562-563.
Hill, G.D., J.W. McGahen, H.M. Baker, D.W. Finnerty and C.W. Bingeman. 1955.
The fate of substituted urea herbicides in agricultural soils. Agron. J.
47(2):93-104.
Hodge, H.C., W.L. Downs, E.A. Maynard et al.* 1964a. Chronic feeding studies
of diuron in dogs. Unpublished study. MRID 00017763.
Hodge, H.C., W.L. Downs, E.A. Maynard et al.* 1964b. Chronic feeding studies
of diuron in rats. Unpublished study. MRID 00017764.
Hodge, H.C., W.L. Downs, B S. Panner, D.W. Smith and E.A. Maynard. 1967.
Oral toxicity and metabolism of diuron (N-(3,4)-dichlorophenyl)-N',N'-
dimethylurea) in rats and dogs. Food Cosmet. Toxicol. 5:513-531.
Imamliev, A.I., and K.A. Bersonova. 1969. Movement of detoxication of dalapon
and diuron in soil. _£n_ Problems of physiology and biochemistry of the
cotton plant. A.I. Imamliev and E.A. Popova, eds. Tashkent, USSR:
Akademii Nauk Uzbekskoi, Institut Eksperimental'noi Biologii Rastenii.
pp. 266-274.
Khera, K.S., C. Whalen, G. Trivett and G. Angers. 1979. Teratogenicity
studies on pesticidal formulations of dimethoate, diuron and lindane in
rats. Bull. Environ. Contarn. Toxicol. 22:522-529.
-------
Diuron August, 1987
-18-
Larson, K.A.* 1976. Acute dermal toxicity—Diuron. Unpublished study.
MRID 00017795.
Larson, K.A., and J.H. Schaefer.* 1976. Eye irritation study using the
pesticide diuron. For Colorado International Corporation. Unpublished
study. MRID 00017797.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S.
Liu, L.C., H.R. Cibes-Viade and J. Gonzalez-Ibanez. 1970. The persistence
of atrazine, ametryne, prometryne, and diuron in soils under greenhouse
conditions. J. Agric. Univ. Puerto Rico. 54(4): 631-639.
McCormick, L.L. 1965. Microbiological decomposition of atrazine and diuron
in soil. Doctoral dissertation. Auburn, AL: Auburn University.
Available from: University Microfilms, Ann Arbor, MI. Report No. 65-6892.
McCormick, L.L., and A.E. Hiltbold. 1966. Microbiological decomposition of
atrazine and diuron in soil. Weeds. 14(1):77-82.
Meister, R., ed. 1984. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company, p. C85.
Miller, J.H., P.E. Keeley, R.J. Thullen and C.H. Carter. 1978. Persistence
and movement of ten herbicides in soil. Weed Sci. 26(1):20-27.
Mukhtar, M. 1976. Desorption of adsorbed ametryn and diuron from soils and
soil components in relation to rates, mechanisms, and energy of adsorption
reactions. Doctoral dissertation. Honolulu, HI: University of Hawaii.
Available from University Microfilms, Ann Arbor, MI. Report No. 77-14,601.
NIOSH. 1985. National Institute for Occupational Safety and Health.
Registry of Toxic Effects of Chemical Substances (RTECS). National
Library of Medicine Online File.
Rubenik, B.L., N.E. Botsman, G.P. Gorman and L.I. Loevskaya. 1973.
Relation between the chemical structure and carcinogenic activity
of urea derivatives. Oukalogiya (Kiev) 4:10-16.
Spiridoncv, Y.Y., V.S. Skhiladze and G.S. Spiridonova. 1972. The effects --f
diuron and monuron in a meadow-bog soil of the moist subtropics of
Adzhariia. Subtrop. Crops. (1):150-155.
STORET. 1987.
Taylor, R.E.* 1976a. Acute oral toxicity (LD50). Project T1001. Unpublished
study. MRID 00028006.
Taylor, R.E.* 1976b. Primary skin irritation study. Project T1002.
Unpublished study. MRID 00028007.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.106, p. 252. July 1, 1985.
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Oiuron August, 1987
-19-
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(1 85):33992-34003. Septem-
ber 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Acceptable Daily
Intake Data; Tolerances Printout, February 21. Office of Pesticide
Programs. Office of Pesticides and Toxic Substances.
U.S. EPA. 1986c. U.S. Environmental Protection Agency. U.S. EPA Method #4
- Determination of Pesticides in Ground Water by HPLC/UV, January 1986
draft. Available from U.S. EPA's Environmental Monitoring and Support
Laboratory, Cincinnati, OH.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Interim guidance for
establishing Rfd dated May 1, 1987 as an addendum to TOX SOP #1002.
Office of Pesticide Programs.
Walker, A., and M.G. Roberts. 1978. The degradation of methazole in soil.
II. Studies with methazole, methazole degradation products, and diuron.
Pestic. Sci. 9(4):333-341.
Wang, C.C., and J.S. Tsay. 1974. Accumulative residual effect and toxicity
of some persistence herbicides in multiple cropping areas. Med. Coll.
Med. Natl. Taiwan Univ. 14(1):1-13.
Wang, Y.S., T.C. Wang and Y.L. Chen. 1977. A study on the degradation of
herbicide diuron in soils and under the light. J. Chinese Agric. Chem.
Soc. 15(1/2):23-31.
Weed, M.B., R. Sutton, G.D. Hill and L.E. Cowart. 1953. Substituted ureas
for pre-emergence weed control in cotton. Unpublished study submitted
by E.I. du Pont de Nemours & Co. Inc., Wilmington, DE.
Weed, M.B., A.W. Welch, R. Sutton and G.D. Hill. 1954. Substituted ureas
for pre-emergence weed control in cotton. In Proceedings of the Southern
Weed Conference. Vol. 7. Athens, GA: Southern Weed Science Society.
pp. 68-87.
Weldon, L.W., and F.L. Timmons. 1961. Penetration and persistence of diuron
in soil, weeds. 9(2):195-203.
Windholz, M., S. Budavari, R.F. Blumetti and E.s. Otterbein, eds. 1983. The
Merck Index—an encyclopedia of chemicals and drugs, 10th ed. Rahway, NJ:
Merck and Company, Inc.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
ENDOTHALL
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, Logi± or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is ab'e 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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Endothall
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 145-73-3 Q
/
Structural Formula
COOH
COOH
H20
7-Oxabicyclo-(2,2,1)-heptane-2,3-dicarboxylic acid
Synonyms
0 1,2-dicarboxy3,6-endoxocyclohexane; Aquathol; Hydrothol; Des-i-cate;
Accelerate
Uses
0 Endothall is used as a defoliant and an herbicide on both terrestrial
and aquatic weeds.
Properties (Carlson et al., 1978; Simsiman et al., 1976)
C8H1005NSP
186.06
White crystalline solid
144°C to the anhydride
Negligible
100 g/L (acid monohydrate)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 No information was found in the available literature on the occurrence
of endothall.
Environmental Fate
0 No information was found in the available literature on the environ-
mental fate of endothall.
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Endothall August, 1987
-3-
III. PHARMACOKINETICS
Absorption
0 Few data exist regarding endothall pharraacolcinetics in mammals.
Soo et al. (1967) performed pharmacokinetic experiments with male
and female Wistar rats. Approximately 82% of a 5-mg/kg oral dose
of 14c-labeled endothall (dissolved in 20% ethanol to a concentration
of 1 mg/mL) was absorbed by the rats within 72 hours. The rats had
received 5 mg/kg of unlabeled endothall in the diet for 2 weeks prior
to treatment with 14c-endothall.
0 Deaths in rabbits directly exposed to endothall in the eye or on the
skin (Pharmacology Research, Inc, 1975a, 1975b) indicate the potential
for absorption by these routes.
Distribution
0 In the Soo et al. (1967) study, the absorbed endothall was distrib-
uted in low levels through most body tissues. Peak levels in all
tissues were observed 1 hour after dosing, with most of the dose
(about 95%) found in the stomach and intestine. Otherwise, the
tissues with the highest concentrations after 1 hour were the liver
and kidney (1.1 and 0.9% respectively), with lower concentrations
(0.02 to 0.1%) in heart, lung, spleen and brain. Very low concentra-
tions were observed in muscle, and endothall was not detected in fat.
No marked preferential accumulation was apparent.
Metabolism
0 The metabolism of endothall is not known to be characterized.
Excretion
Soo et al. (1967) described excretion as follows:
0 Clearance of 14C-endothall was biphasic in the stomach (t1 >2 ° 2«2 and
14.2 hours) and kidney (tj = 1.6 and 34.6 hours) and monopnasic
in the intestine and liver (t^ = 14.4 and 21.6 hours, respectively).
Total excretion of the 14C label was over 95% complete by 48 hours and
over 99% complete by 72 hours, suggesting that no significant
bioaccumulation occurred.
0 Approximately 90% of the administered dose was excreted in the feces.
Urinary excretion accounted for approximately 7% of the dose, and
approximately 3% of the radioactive label was recovered in expired
carbon dioxide.
0 Approximately 20% of the dose excreted in the feces was unchanged
endothall. The remaining radioactivity was presumed to be metabolites
or conjugates.
0 Soo et al. (1967) also found no radioactivity in pups from lactating
dams given oral doses of 14c-endothall.
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Endothall August, 1987
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IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of endothall in humans except for one case history of a young
male suicide victim who ingested an estimated 7 to 8 g of disodium
endothall in solution (approximately 100 mg endothall ion/kg).
Repeated vomiting was evident. Autopsy revealed focal hemorrhages
and edema in the lungs and gross hemorrhage of the gastrointestinal
(GI) tract (Allender, 1983).
Animals
Short-term Exposure
0 Early acute studies report cardiac arrest (Goldstein, 1952) or
respiratory failure (Srensek and Woodard, 1951) as causes of death
in dogs and rabbits. Endothall was injected intravenously in both
studies with these effects observed at doses of 5 mg/kg (lowest)
and higher.
0 The available acute oral dose studies are essentially restricted to
mortality data without biochemical or histopathological observations.
The acute toxicity of endothall aoid appeared to be greater than that
of the salt forms normally used in herbicide formulations. In rats,
the oral LD50 of endothall was reported as 35 to 51 mg/kg for the
acid form and 182 to 197 mg/kg for the sodium salt (Simsiman et al.f
1976; Tweedy and Houseworth, 1976).
0 Rats were given 1,000 or 10,000 ppm disodium endothall in the diet
(Brieger, 1953a) and doses were calculated by assuming a body weight
of 0.4 kg and daily food consumption of 20 g. Slight liver degeneration
and focal hemorrhagic areas in the kidney were reported for male and
female rats dosed orally with approximately 40 mg endothall ion/kg/day
for 4 weeks; most of the rats receiving approximately 400 mg endothall
ion/kg/day died within 1 week. The liver and kidney effects from
endothall ingestion are consistent with the pharmacokinetic tissue
distribution results reported by Soo et al. (1967).
0 Nine male dogs (one dog/dose) were do'ed orally with capsules containing
1 to 50 mg disodium endothall/kg/day vO.8 to 40 mg endothall ion/kg/day)
for 6 weeks (Brieger, 1953b). All dogs that were administered 20 to
50 mg disodium endothall/kg/day died within 11 days. Vomiting and diarrhea
were observed in the group given 20 mg disodium endothall/kg/day.
Pathological changes in the GI tract, described as congested and
edematous stomach walls and edematous upper intestines, were indicated
as common in all dogs. Erosion and hemorrhages in the stomach were
observed with doses of 20 mg/kg/day or more.
Dermal/Ocular Effects
0 Goldstein (1952) reported that a 1% solution of endothall applied to
the unbroken skin of rabbits produced no effects. The same solution
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Endothall August, 1987
-5-
applied to scarified skin resulted in mild skin lesions. Ten to
twenty percent solutions or applications of the pure, powdered
material to intact or scarified skin resulted in more severe damage,
including necrosis, and the deaths of some treated animals.
0 Topical exposure of six rabbits to 200 mg endothall technical/kg
resulted in the death of all rabbits within 24 hours (Pharmacology
Research, Inc., 1975a).
0 Technical endothall (0.1 g equivalent to 80 mg endothall ion) produced
severe eye irritation in three rabbits when directly applied to the
conjunctiva. Effects included corneal opacity, con^unctival irritation
and iridic congestion. Furthermore, technical endothall apparently
produced systemic effects, by this route of absorption, since several
animals died within 24 hours as a result of this exposure. Eyes were
rinsed with water 20 to 30 seconds after treatment in three rabbits;
conjunctival irritation and iridic congestion reversed in 4 days in
two rabbits but persisted along with corneal opacity in one rabbit
for 7 days (Pharmacology Research, Inc., 1975b).
Long-term Exposure
0 Beagle dogs (four/sex/group) fed diets containing 0, 100, 300 or
800 ppm disodium endothall (equivalent to 0, 2, 6 or 16 mg endothall
ion/kg/day for 24 months showed no gross signs of toxicity (Keller,
1965). Values for hematology, urinalysis, weight gain and food
consumption were within normal limits and comparable to those for
control animals. Increased stomach and small intestine weights were
observed in the intermediate and high-dose groups. However, microscopic
examination of essentially all tissues in the high-dose group revealed
no pathological changes that could be attributed to endothall ingestion.
A No-Observed-Adverse-Effect-Level (NOAEL) of 2 mg endothall ion/kg/day
is identified from this study.
0 Brieger (1953b) reported no toxic effects in female rats given dietary
levels as high as 2,500 ppm disodium endothall (about 100 mg endothall
ion/kg/day, assuming food intake of 20 g/day and mean body weight of
0.4 kg) for 2 years.
Reproductive Effects
0 A three-generation study in rats was reported by Scientific Associates
(1965). Groups of male and female rats were fed diets containing 0,
100, 300 or 2,500 ppm disodium endothall (equivalent to 0, 4, 12 or
100 mg endothall ion/kg/day) until they were 100 days old and were
then mated. Three successive generations of offspring were maintained
on the test diet for 100 days and then bred to produce the next tast
generation. Pups in the 4-rng/kg/day dose group were normal, pups in
the 12-mg/kg/day group had decreased body weights at 21 days of age
and pups in the 100 mg/kg/day group did not survive more than 1 week.
A NOAEL for reproductive effects of 4 mg endothall ion/kg/day was
identified from this study.
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Endothall August, 1987
-6-
Developmental Effects
0 A short-term teratology study in rats by Science Applications, Inc.
(1982) indicated no observable signs of developmental toxicity at
dose levels that were fatal to the dams. This study suggests that
the dams are more susceptible to endothall than are the embryos or
fetuses. Groups of 25 or 26 female rats were mated and then orally
dosed with 0, 10, 20 or 30 mg/kg/day of aqueous endothall technical
(0, 8, 16 or 24 mg endothall lon/kg/day) on days 6 to 19 of gestation.
Two dams died from the 20-mg/kg/day dose, and 10 dams died from the
30-mg/kg/day dose. No clinical signs were noted prior to death, and
no lesions were observed at necropsy. The researchers concluded that
endothall technical was not embryotoxic or teratogenic at maternal
doses of 30 mg/kg/day or below. A NOAEL of 10 mg endothall
technical/kg/day based on maternal effects was identified.
Mutagenicity
0 Mutagenicity results from short-term in vitro tests are mixed, with
various forms of endothall reported as test agents. Mutagenicity
studies utilizing Salmonella with and without metabolic activation
resulted in negative findings for endothall technical (Andersen
et al., 1972; Microbiological Associates, 1980a). Mutagenic activity
was not found in BALB/3T3 Clone A31 mouse cells exposed to endothall
technical (Microbiological Associates, 1982b).
e For the following studies, Wilson et al. (1956) used "commercial
Endothall" with no further description, whereas the remaining investi-
gators used Aquathol K, a commercial formulation containing dipotassium
endothall at a level of 28.6% acid equivalent. In Drosophila melano-
gaster, mutagenic results were mixed, with Wilson et al. (1956) and
Sandier and Hamilton-Byrd (1981) reporting positive and negative
results, respectively. Sandier and Hamilton-Byrd (1981) reported
negative results in a mutagenicity assay with the mold Neurospora
crassa. A sister chromatid exchange study in human lymphocytes was
negative (Vigfusson, 1981). Transformation was induced in a BALB/c
3T3 test for malignant transformation (Litton Bionetics, Inc., 1981).
Carcinogenicity
8 No statistically significant numbers or types of tumors were observed
in rats fed as much as 100 mg endothall ion/kg/day for 2 years
(Brieger, 1953b).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs)* are generally determined for one-day, ten-day,
longer-tern (approximately 7 years) and lifetime exposures if adequate data
•Because the test material in the various toxicity studies was salt or acid
forms of endothall, the HAs described below are expressed in terms of
endothall ion.
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Endothall August, 1987
-7-
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 ing/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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No studies were located in available literature that were suitable for
calculation of the One-day HA. The single-dose studies measured mortality as
the toxicological end point and are not suitable for use in calculating an HA.
The value of 0.8 mg/L calculated as the Ten-day HA can be used as a conservative
estimate of the One-day HA.
Ten-day Health Advisory
The teratology study by Science Applications, Inc. (1982) has been
selected as the basis for the Ten-day HA. It is the only study that defined
a short-term NOAEL (8 tng endothall ion/kg/day, based on maternal toxicity).
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-dav HA = (8 mg/kg/day) (10 kg) = 0.8 mg/L (800 ug/L)
(100) (1 L/day)
where:
8 mg/kg/day = NOAEL based on the absence of fetal and maternal
effects in rats exposed to endothall acid orally for
13 days.
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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Endothall August, 1987
-8-
Longer-term Health Advisory
There is concluded to be insufficient data for calculation of a Longer-
term HA. Therefore, the DWEL adjusted for a 10-)g child (0.2 mg/L) is proposed
as a conservative estimate for a 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 IntaJe (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.
The 2-year feeding study in dogs by Keller (1965), which identified a
NOAEL of 2 mg endothall ion/kg/day, has been selected to serve as the basis
for the Lifetime HA for endothall. The study by Scientific Associates (1965)
was of shorter duration (100 days/generation) and did not as completely
define a NOAEL (except for 4 mg endothall ion/Jg/day for reproductive effects);
however, the NOAEL in this study approximates that in the Keller (1965)
study. The 2-year study in rats by Brieger (1953b) showed no adverse effects
from doses up to 100 mg endothall ion/kg/day, but no information was provided
on the parameters tested and the levels at which effects did occur.
Using the NOAEL of 2 mg/Jg/day, the Lifetime HA for endothall is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (2 mg/)g/day) _ Q.02 mg/Jg/day
(100)
where:
2 mg/Jg/day = NOAEL, based on absence of increased organ weight and
organ-body weight ratios in the stomach and small
intestine in dogs exposed to endothall in the diet
for 2 years.
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Endothall August, 1987
-9-
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*02 mg/kg/day) (70 kg) = 0.7 mg/L (70o ug/L)
(2 L/day)
where:
0.02 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 = (0.7 mg/L) (20%) =0.14 mg/L (140 ug/L)
where:
0.7 mg/L = DWEL.
20% = assumed percentage of daily exposure contributed by
ingestion of drinking water.
Evaluation of Carcinogenic Potential
0 Available toxicity data do not show endothall as carcinogenic.
0 Endothall can be placed in Group D (inadequate evidence in humans
and animals) by the EPA's guidelines for carcinogenic risk assessment
(U.S. EPA, 1986).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of endothall (WHO, 1982).
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 An interim tolerance of 200 ug/L has been published for residues of
endothall, used to control aquatic plants, in potable water (CFR,
1979).
0 Residue tolerances for endothall published by the U.S. EPA (CFR,
1977) include 0.1 ppm in or on cottonseed, 0.1 ppm in or on potatoes,
0.05 ppm in or on rice grain and 0.05 ppm in or on rice straw.
0 A tolerance is a derived value based on residue levels, toxicity
data, food consumption levels, hazard evaluation and scientific
judgment; it is the legal maximum concentration of a pesticide
in or on a raw agricultural commodity or other human or animal food
(Paynter et al., undated).
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Endothall August, 1987
-10-
0 The ADI set by the U.S. EPA Office of Pesticide Programs is 0.02
mg/kg/day based on the 2 mg/kg/day NOAEL in the 2-year dog study by
Keller (1965) and a 100-fold uncertainty factor.
VII. ANALYTICAL METHODS
0 No information was found in the available literature on the analytical
methods used to detect endothall in drinking water.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies capable of effectively removing endothall from contaminated
water.
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Endothall August, 1987
-11-
IX. REFERENCES
Allender, W.J. 1983. Suicidal poisoning by endothall. J. Anal. Toxicol.
7:79-82.
Andersen, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of herbi-
cides for possible mutagenic properties. J. Agr. Food Chem. 20:649-654.
Brieger, H.* 1953a. Preliminary studies on the toxicity of endothall
(disodium). EPA Pesticide Petition No. 6G0503, redesignated No. 7F0570,
1966. Accession No. 246012.
Brieger, H.* 1953b. Endothall, long term oral toxicity test—rats. EPA
Pesticide Petition No. 6G0503, redesignated No. 7F0570, 1966. Accession
No. 246012.
Carlson, R., R. Whitaker and A. Landskov. 1978. Endothall. Chapter 31.
Jin G. Zweig and J. Sherma, eds. Analytical methods for pesticides and
plant growth. New York: Academic Press, pp. 327-340.
CFR. 1977. Code of Federal Regulations. 40 CFR 180.293.
CFR. 1979. Code of Federal Regulations. 21 CFR 193.180. April 1, 1979.
Goldstein, F. 1952. Cutaneous and intravenous toxicity of endothall
(disodium-3-endohexahydrophthalic acid). Pharmacol. Exp. Ther. 11:349.
Keller, J.* 1965. Two year chronic feeding study of disodium endothall to
beagle dogs. Scientific Associates report. EPA Pesticide Petition
6G0503, redesignated No. 7F0570, June 1966. Accession No. 24601.
Litton Bionetics, Inc. 1981. Evaluation of Aquathol K in the in vitro
transformation of BALB/3T3 cells with and without metabolic activation
assay. Project No. 20992. Report to Municipality of Metropolitan
Seattle, Seattle, WA, by Litton Bionetics, Inc., Rockville, MD.
Microbiological Associates.* 1980a. Activity of T1604 in the Salmonella/
microsomal assay for bacterial mutagenicity. Unpublished final report
for Pennwalt Corp. by Microbiological Associates, Bethesda, MD.
Microbiological Associates.* 1980b. Activity of T1604 in the in vitro
mammalian cell point mutation assay in the absence of exogenous metabolic
activation. Unpublished final report for Pennwalt Corp. by Microbiological
Associates, Bethesda, MD.
Paynter, O.E., J.G. Cummings and M.H. Rogoff. Undated. United States
Pesticide Tolerance System. U.S. EPA, Office of Pesticide Programs,
Washington, DC. Unpublished draft report.
Pharmacology Research, Inc.* 1975a. U.S. EPA Pesticide Resubmission File
4531-EIE. Summary data on acute oral toxicity and dermal irritation in
rabbits (Endothall). Accession No. 244125.
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Endothall August, 1987
-12-
Pharmacology Research, Inc.* 1975b. U.S. EPA Pesticide Resubmission File.
Summary data, primary eye irritation in the rabbit and inhalation toxicity
in several species (Endothall). Accession No. 246012.
Sandier, L., and E.L. Hamilton-Byrd. 1981. The induction of sex-linked
recessive ethal mutations in Drosophila melanogaster by Aquathol K, as
measured by the Muller-5 test. Report to Municipality of Metropolitan
Seattle, Seattle, WA.
Scientific Associates.* 1965. Three generation rat reproductive study,
disodium endothall. EPA Pesticide Petition No. 6G0503, redesignated
7F0570, 1966. EPA Accession No. 114667.
Science Applications, Inc.* 1982. A dose range-finding teratology study of
endothall technical and disodium endothall in albino rats. Resubmission
of Pesticide Application for Aquathol K Aquatic Herbicide (EPA Registra-
tion No. 4581-204) and Hydrothal 191 Aquatic Algicide and Herbicide
(EPA Registration No. 4581=174). EPA Accession No. 071249.
Simsiman, G.V., T.C. Daniel and G. Chesters. 1976. Diquat and endothall:
Their fates in the environment. Res. Rev. 62:131-174.
Soo, A., I. Tinsley and S.C. Fang. 1967. Metabolisir. of Hc-endothall in
rats. J. Agric. Food Chem. 15:1018-1021.
Srensek, S.E., and G. Woodard. 1951. Pharmacological actions of "endothall"
(disodium-3,6-endoxo-hexahydrophthalic acid). Fed. Proc. 10:337.
(Abstract)
Tweedy, B.C., and L.D. Houseworth. 1976. Miscellaneous herbicides. In
Herbicides-chemistry, degradation and mode of action. P.C. Kearney and
D.D. Kaufman, eds. Chapter 17. New Yorks Marcel Dekker, Inc., pp.
815-833.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(1851:33992-34003.
September 24.
Vigfusson, N.V. 1981. Evaluation of the mutagenic potential of Aquathol K
by induction of sister chromatid exchanges in hjr.an lymphocytes in vitro.
Report to Municipality of Metropolitan Seattle, Seattle, WA.
WHO. 1982. World Health Organization. IARC monographs on the evaluation of
the carcinogenic risk of chemicals to humans. Chemicals, industry
processes and industries associated with cancer to humans. International
Agency for Research on Cancer Monographs Vol. 1 to 29. Supplement 4.
Geneva: World Health Organization.
Wilson, S.M., A. Daniel and G.B. Wilson. 1956. Cytological and genetical
effects of the defoliant endothall. J. Hered. 47:151-154.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
ETHYLENE THIOUREA
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 the--e 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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Ethylene Thiourea August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
Ethylene thiourea (ETU) is a common degradation product of the ethylene
bisdithiocarbamate (EBDC) pesticides and is, itself, toxic.
Although the toxicity of ETU may be similar to the toxic effects observed
with the EBDCs, the One-day, Ten-day, Longer-term and Lifetime HAs for ETU
should not necessarily be considered protective of exposure to individual
EBDCs at this time. The mechanisms of toxicity for these substances are
still under evaluation.
CAS No. 96-45-7
Structural Formula
¥
-NH
2-Imidazolidinethione
Synonyms
• ETU
Uses
0 Degradation product of several EBDC pesticides.
Properties
Chemical Formula C3H5N2S
Molecular Weight 102.2
Physical State (25°C) White crystals
Boiling Point
Melting Point 203°
Density
Vapor Pressure
Specific Gravity
Water Solubility (30°C) 20 g/L
Log Octanol/Water Partition —
Coefficient
Taste Threshold
Odor Threshold
Occurrence
0 ETU was not found in sampling performed at 250 ground water stations,
according to the STORET database (STORET, 1987).
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Ethylene Thiourea August, 1987
-3-
Environmental Fate
0 Ethylene thiourea can be degraded by photolysis (U.S. EPA, 1982).
0 14C-Ethylene thiourea was intermediately mobile (Rf 0.61) to very
mobile (Rf 100) in muck and sandy loam soils, respectively, as determined
by soil TLC (U.S. EPA, 1986a). Adsorption was correlated to organic
matter. Following 6 days of incubation in dry silty clay loam soil,
ETU residues were immobile; however, ETU residues subjected to a
wet-dry cycle were slightly mobile (Rf 0.2).
0 Levels of ETU (purity unspecified) declined at an unspecified rate in
sand, with a half-life of 1-6 days (U.S. EPA, 1986a). Concentrations
of ETU declined from 220 ppm at day 0 to 116 ppm by day 1 and 86 ppm
by day 6.
0 The ethylene bisdithiocarbamates (EBDCs) are generally unstable in
the presence of moisture and oxygen, as well as in biological systems
(U.S. EPA, 1982).
0 The EBDCs decompose rapidly in water. Mancozeb has been shown to have
a half-life of less than 1 day in sterile water before degrading to
ETU (U.S. EPA, 1982).
0 Photolysis is a major degrading pathway for ET'J (U.S. EPA, 1982).
III. PHARMACOKINETICS
Absorption
0 Allen et al. (1978) reported a very high rate of absorption of !*C-ETU
gastrically administered at 40 mg/kg to female rhesus monkeys and
female Sprague-Dawley rats. In both species, feces accounted for less
than 1.5% of the excreted radioactivity at 48 hours after administration.
0 Absorption was also high in male Sprague-Dawley rats orally administered
14c-ETU at 4 mg/kg, with 82.7% of the total administered dose detected
in the urine at 24 hours (Iverson et al., 1980).
Distribution
0 Allen et al. (1978) reported that in rhesus monkeys administered
14c-ETU at 40 mg/kg by gastric intubation, total tissue distribution
at 48 hours was approximately 25% of the aditinistered dose; approximately
half of that was concentrated in muscle, with measurable amounts
noted in blood, skin and liver. In Sprague-Dawley rats, however,
total tissue distribution was less than 1% of the administered dose.
0 Except in the thyroid, ETU was not found to accumulate in rats given
an oral dose (amount not specified) (U.S. EPA, 1982). Up to 80% of the
absorbed dose was eliminated in the urine 24 hours after administration.
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Ethylene Thiourea August, 1987
-4-
Metabolism
0 Iverson et al. (1980) identified the 24-hour urinary metabolites of
14c-ETU orally administered to male Sprague-Dawley rats at 4 mg/kg.
Imidazoline was present at 1.9% of the total recovered dose, imidazolone
at 4.9%, ethylene urea at 18.3% and unchanged ETU at 62.6%. In female
cats, intravenous (iv) administration of this dose resulted in unchanged
ETU present in the urine at only 28% of the total recovered dose, with
S-methyl ETU at 64.3% and ethylene urea at 3.5%.
0 One hundred percent of the ETU (dose not specified) fed to mice was
recovered rapidly (time not specified) with 50% recovered in the
urine (U.S. EPA, 1982). Of the urinary products, 52% was unchanged
ETU, 12% was ethylene urea, and 37% were polar products.
0 All animals that have been tested appear to metabolize EBDCs rapidly.
ETU and ethylene bisdiisothiocyanato sulfide (EBIS) are the major
metabolites formed (U.S. EPA, 1982). Approximately 18% of an EBDC
dose is converted to ETU in vivo.
Excretion
Allen et al. (1978) reported that 48 hours after gastric admini-
stration of 14C-ETU at 40 mg/kg to rhesus monkeys, approximately 55%
of the administered dose was detected in the urine and 0.5% in the
feces. In Sprague-Dawley rats dosed identically, 82% was recovered
in the urine and 1.3% in the feces.
Iverson et al. (1980) reported that 82.7 and 80.6% of the total
radioactivity of a single 4-mg/kg dose of 14C-ETU was eliminated in the
24-hour urine of orally treated male Sprague-Dawley rats and iv-treated
female cats, respectively.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of ETU in humans.
Animals
Short-term Exposure
0 The acute oral LD50 for 5TU is 1,832 mg/kg in rats (U.S. EPA, 1982).
0 Graham and Hansen (1972) measured 131I uptake in male Osborne-Mendel
rats administered ETU (purity not stated) in the diet at 50, 100, 500
or 750 ppm for various time periods (e.g., 30, 60, 90 or 120 days).
Assuming that 1 ppm in the diet of younger rats is equivalent to
approximately 0.1 mg/kg/day (Lehman, 1959), these levels correspond
to doses of about 5, 10, 50 or 75 mg/kg/day. Four hours after the
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Ethylene Thiourea August, 1987
-5-
injection of 131I, uptake was decreased significantly in rats that had
ingested ETU at 500 or 750 ppm for all time periods. At 24 hours after
1311 injection, uptake was significantly decreased in rats that had
ingested 100, 500 or 750 ppm for all time periods. Histologically,
the thyroid glands of rats ingesting ETU at approximately 5.0 mg/kg,
the No-Observed-Adverse-Effect-Level (NOAEL) for this study, were not
different from those of control rats. There was slight hyperplasia
of the thyroid in rats given 100 ppm (10 mg/kg/day). At doses of 500
or 750 ppm (50 or 75 mg/kg/day), the thyroid had moderate to marked
hyperplasia.
0 In an 8-day maximum tolerated dose (MTD) study by Plasterer et al.
(1985), dose levels of 0, 75, 150, 300, 600 and 1,200 mg/kg ETU were
given by gavage to mice (10/group, sex not specified). Body weight
and mortality were evaluated. No significant effects were noted on
body weight at the end of the eighth day. Based on mortality, ETU was
considered moderately toxic by the authors. An MTD of 600 mg/kg was
determined.
0 In a study by Freudenthal (1977), ETU (>95% pure) was fed to rats
(20/sex/group) in the diet at levels of 0, 1, 5, 25, 125 or 625 ppm
for 30 days. Assuming that 1 ppm in the diet of a young rat is
equivalent to 0.1 mg/kg (Lehman, 1959), these levels correspond to
doses of about 0, 0.1, 0.5, 2.5, 12.5 or 62.5 mg/kg. Thyroid function,
food consumption, body weight gain and histopathology were assessed
in the animals. Rats in the 625-ppm groups showed signs of toxicity
after 8 days of exposure. Hair loss, dry skin, increased salivation
and decreased food consumption and body weight gain were observed.
Other effects noted in the 625-ppm dose group were decreased iodine
uptake and percent triiodothyronine (T3) bound to thyroglobulin.
Thyroid-stimulating hormone (TSH) was increased, and T3 and thyroxine
(14) decreased in the 625-ppm dose group. Thyroid hyperplasia was also
noted in this group. Animals exposed to 125 ppm exhibited increased
TSH; decreased 14, and thyroid hyperplasia. Other thyroid parameters
were not affected. Based on the absence of adverse effects in rats
exposed to 25 ppm or less after 30 days, a NOAEL of 25 ppm (2.5 mg/kg)
was identified.
0 Arnold et al. (1983) showed that the thyroid effects of ETU (purity
not stated) administered in the diet for 7 weeks to male and female
Sprague-Dawley rats were reversible when ETU was removed from the
diet. Dose-related significant decreases in body weight and increases
in thyroid weight were observed in all treated animals, starting at
dose levels of 75 ppm (approximately 7.5 mg/kg/day based on Lehman,
1959). This dose was identified as the Lowest-Observed-Adverse-Effect-
Level (LOAEL) for this study.
0 In a 60-day study, which was a continuation of the above study by
Freudenthal (1977), 14/40 rats in the 625-ppm group died. Thyroid
hyperplasia and altered thyroid function were observed in the two
high-dose groups. Thyroid hyperplasia was also observed in the
25-ppm group. This effect, however, was not observed in this dose
group when esposure was continued to 90 days. Thus, the NOAEL for
this study is presumed to be 25 ppm, or 2.5 mg/kg.
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Ethylene Thiourea August, 1987
-6-
Dermal/Ocular Effects
0 No information was found in the above literature on the dermal/ocular
effects of ETUc
Long-term Exposure
0 Freudenthai (1977) described alterations in thyroid function and
changes in thyroid morphology when Sprague-Dawley rats were admini-
stered ETU (96.8% pure) in the diet at levels of 1 to 625 ppm (approxi-
mately 0.1 to 62.5 mg/kg/day based on Lehman, 1959) for up to 90 days.
The NOAEL was reported to be 19.5 mg/kg/day at week 1 and 12.5 mg/kg/day
at week 12.
0 Graham and Hansen (1972) measured 1 31I uptake in male Osborne-Mendel
rats administered ETU (purity not specified) in the diet at 50, 100,
500 or 750 ppm for up to 120 days. Assuming that 1 ppm in the diet
of older rats is equivalent to approximately 0.05 mg/kg/day (Lehman,
1959), these dosages are equivalent to approximately 2.5, 5, 25 and
37.5 mg/kg/day. Four hours after the injection of radioactive iodine,
uptake was decreased significantly in rats ingesting ETU at 500 or
750 ppm (25 or 37.5 mg/kg/day) for all feeding periods. At 24 hours
after 1311 injection, uptake was significantly decreased in rats
ingesting the 100-, 500- and 750-ppm doses for all feeding periods.
Histologically, the thyroid glands of rats ingesting ETU at approximately
2.5 mg/kg, the NOAEL for this study, were not different from those of
control rats. There was slight hyperplasia of the thyroid in rats
given 100 ppm (5 mg/kg/day). At doses of 500 or 750 ppm (25 or 37.5
mg/kg/day), the thyroid had moderate to marked hyperplasia.
0 The thyroid appears to be the primary target organ for ETU toxicity
in longer-term exposure studies. Graham et al. (1973) measured
131i uptake in male and female Charles River rats fed ETU (purity
not specified) in the diet at 5, 25, 125, 250 or 500 ppm for up to
12 months. Assuming that 1 ppm in the diet of older rats is equivalent
to approximately 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to doses of about 0.25, 1.25, 6.25, 12.5 or 25 mg/kg/day. Adverse
effects were noted at 2, 6 and 12 months. At 12 months, significant
decreases in body weight and increases in thyroid weight were seen at
the 125-, 250- and 500-ppm levels. Uptake of 131I was significantly
decreased in male rats after 12 months at 500 ppm, but was increased
in females. Microscopic examination of the thyroid revealed the
development of nodular hyperplasia at dose levels of 125 ppm and
higher. The NOAEL for thyroid effects in this study was 25 ppm
(approximately 1.25 mg/kg/day).
0 Ulland et al. (1972) reported a dose-related increased incidence of
hyperplastic goiter in male and female rats fed ETU at 175 and 350 ppm
in their diet for 18 months (approximately 8.75 and 17.5 mg/kg/day,
based on Lehman, 1959). An increased incidence (significance not
specified) of simple goiter was also reported in all treatment groupst
0 In a 2-year study by Graham et al. (1975), Charles River rats were fed
ETU (purity not specified) in the diet at 5, 25, 125, 250 or 500 ppm
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Ethylene Thiourea August, 1987
-7-
{approximately 0.25, 1.25, 6.25, 12.5 or 25 mg/kg/day, based
on Lehman, 1959). Statistically significant (p <0.01) decreases in
body weight were observed in both sexes fed at 500 ppm. Increases
in thyroid-to-body weight ratios were apparent at 250 and 500 ppm
(p <0.01). There was an increased iodine (131j) uptake at 5 ppm and
a decreased uptake at 500 ppm, as well as slight thyroid hyperplasia
at the 5- and 25-ppm dose levels (significance not stated). Based on
these results, a LOAEL for lifetime exposure of 5 ppm (0.25 mg/kg/day)
was identified.
Reproductive Effects
0 Plasterer et al. (1985) administered ETU (purity not specified) by
gavage as a water slurry to CD-I mice at 600 mg/kg/day on days 7 to
14 of gestation. At this dose level, maternal toxicity was not
observed but the reproductive index was significantly decreased
(p <0.05), indicating severe prenatal lethality.
0 New Zealand White rabbits were dosed with ETU at 10, 20, 40 or 80
mg/kg/day on days 7 to 20 of pregnancy (Khera, 1973). Observed
effects included an increase (p <0.05) in resorption sites at 80 rag/kg.
No adverse effects on fetal weight or on the number of viable fetuses
per pregnancy were noted at any dose level, and no signs of maternal
toxicity were observed. Based on the results of this study, a NOAEL
of 80 mg/kg/day for maternal toxicity and a NOAEL of 40 mg/kg/day for
fetotoxicity were identified.
Developmental Effects
0 The ability of ETU to induce various adverse effects, including
teratogenicity and maternal toxicity, has been demonstrated by several
investigators using various animal models. Available data indicate
that rats are probably the most sensitive species.
0 Khera (1973) orally administered ETU (100% pure) to Wistar rats at
daily doses of 5, 10, 20, 40 or 80 mg/kg from 21 or 42 days before
conception to pregnancy day 15 and on days 6 to 15 or 7 to 20 of
pregnancy. Dose-dependent lesions of the fetal central nervous and
skeletal systems were produced, irrespective of 'the time at which ETU
was administered. Teratogenic effects seen at the two highest dose
levels included meningoencephalocele, neningorrhagia, meningorrhea,
hydrocephalus, obliterated neural canal, abnormal pelvic limb posture
with equinovarus, micrognathia, oligodactyly, and absent, short or
kinky tail. Less serious defects were seen at 20 mg/kg, and at 10 mg/kg
there was only a retardation of parietal ossification and of cerebellar
Purkinje-cell migration. Retarded parietal ossification was the only
abnormality seen at 5 mg/kg (significance not stated), its incidence
being limited to small areas and to a few large litters. No signs of
maternal toxicity were observed in rats administered ETU at 40 mg/kg/day
for 57 days (42 days preconception to day 15 of gestation). Based on
the results of this phase of the study, the NOAEL for maternal toxicity
was 40 mg/kg/day, and the Lowest-Observed-Adverse-Effect-Level (LOAEL)
for developmental effects was 5 mg/kg/day.
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Ethylene Thiourea August, 1987
-8-
0 In the same study (Khera, 1973) New Zealand White rabbits were dosed
with ETU at 10, 20, 40 or 80 mg/kg/day on days 7 to 20 of pregnancy.
Observed effects included a reduction in fetal brain:body weight
ratio at 10 and 80 mg/kg (p <0.01). Renal lesions, characterized by
degeneration of the proximal convoluted tubules, were noted micro-
scopically (dose level not specified), but there were no skeletal
abnormalities that were attributed by the authors to ETU. A LOAEL
of 10 mg/kg/day was identified.
0 Dose-related central nervous system (CNS) lesions in Wistar rat
fetuses were reported by Khera and Tryphonas (1985). Ethylene thiourea
(>98% pure) was administered by gastric intubation at 0, 15 or 30 mg/kg
to dams on day 13 of pregnancy. Observed lesions at 30 mg/kg included
histopathological changes of the CNS such as karyorrhexis in the
germinal layer of basal lamina extending from the thoracic spinal
cord to the telencephalon, and obliteration and duplication of the
central canal and disorganization of the germinal and mantle layers.
In the brain, the ventricular lining was fully denuded, neuroepithelial
cells were arranged in the form of rosettes and nerve cell proliferation
was disorganized. In the 15-mg/kg/day group, cellular necrosis was
less severe and consisted of small groups of cells dispersed in the
germinal layers of the neuraxis. None of the dams treated with ETU
at any level in this study showed any overt signs of toxicity. Based
on the results of this study, the NOAEL for maternal toxicity was 30
mg/kg and the LOAEL for developmental toxicity was 15 mg/kg.
0 Sato et al. (1985) investigated the teratogenic effects of ETU (purity
not specified) on Long-Evans rats exposed by gastric intubation to a
single dose of 80, 120 or 160 mg/kg on one day between days 11 and 19
of gestation. Fetal malformations were related to both the day of
administration and the dosage level. A short or absent tail was
noted, for example, in 100% of fetuses exposed to ETU on gestational
day 11 to 14. On day 11, a dose-dependent incidence of spina bifida
and myeloschisis with hind-brain crowding were observed. A high
incidence (48 to 87.5%, not dose-related) of macrocephaly with occipital
bossing was noted, with administration of ETU on day 12, and an almost
total incidence (96 to 100%) with administration on day 13. Other
abnormalities seen in this study were exencephaly, microcephaly and
hypognathia, and extremely high incidences (100% in many groups) of
hydroencephaly and hydrocephalus, especially associated with administrate:
days 14 through 19. Maternal toxicity was not addressed by the
authors. The results of this study are not useful in determining
LOAELs or NOAELs for teratogenicity or maternal toxicity, but serve
instead as evidence of the kinds of developmental effects that a single
dose of ETU at 80 mg/kg can induce in rats.
9 Khera and Iverson (1978) reported that there was no clear evidence of
teratogenicity in kittens whose mothers had been administered ETU
(purity not specified) at 5, 10, 30, 60 or 120 mg/kg by gelatin capsule
for days 16 to 35 of gestation. However, fetuses from cats in a
moribund state subsequent to ETU toxicosis (30 to 120 mg/kg dosage
groups) did show a high incidence (11/35) of malformations including
coloboma, umbilical hernia, spina bifida and cleft palate. Maternal
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Ethylene Thiourea August, 1987
-9-
toxicity and death were observed at dose levels of 10 mg/kg and
above, manifesting signs of toxicity that were delayed in onset and
characterized by progressive loss of body weight, ataxia, tremors and
hind-limb paralysis. In this study, the NOAEL for maternal toxicity
was identified as 5 mg/kg/day and the NOAEL for developmental effects
was 10 mg/kg/day.
0 Chernoff et al. (1979) demonstrated the teratogenic effects of ETU
in Sprague-Dawley rats, CD-I mice and golden hamsters. The rats
were administered ETU (purity not specified) by gastric intubation
at 80 mg/kg/day on days 7 to 21 of gestation. Gross defects of the
skeletal system (micrognathia, micromelia, oligodactyly, kyphosis)
and the CNS (hydrocephalus, encephalocele), as well as cleft palate
were noted in a majority of fetuses at this dose level. No clear
evidence of teratogenicity was seen in groups of rats administered
dose levels of 5 to 40 mg/kg/day. No similar pattern of defects was
observed in CD-I mice dosed at 100 or 200 mg/kg/day on days 7 to 16
of gestation or in golden hamsters dosed at 75, 150 or 300 mg/kg/day
on days 5 to 10 of gestation. Observations of maternal toxicity
included a marked decrease in the average weight gain of pregnant
rats dosed at 80 mg/kg/day (p <0.001). No significant effects were
observed in mice or hamsters. Based on the results of this study,
the NOAELs for maternal and developmental toxicity were 40 mg/kg/day
in the rat, 200 mg/kg/day in the mouse and 300 mg/kg/day in the
hamster.
0 Adverse developmental effects of orally administered ETU, including
teratogenicity and/or maternal toxicity, have been reported at 60,
100 and 240 mg/kg in rats (Khera, 1982; Teramoto et al., 1975; Ruddick
and Khera, 1975) and at 400 and 1,600 to 2,400 mg/kg in mice (Teramoto
et al., 1980; Khera, 1984).
Mutagenicity
0 Seiler (1973) described ETU as exhibiting weak but significant
mutagenic activity in Salmonella typhimurium HIS G-46. A 2.5-fold
increase in mutation frequencies (p <0.001) was seen at intermediate
concentrations (100 or 1,000 ppm/plate), but at higher concentrations
(10,000 and 25,000 ppm) ETU was somewhat lethal to the test colonies
resulting in lower relative mutagenic indices (1.60 and 1.16,
respectively).
0 Schubach and Hummler (1977) reported that ETU induced mutations of
the base-pair substitution type in £. typhimurium TA 1530 in vitro as
well as in a host-mediated assay. In the host-mediated assay, a
dose of 6,000 mg/kg (LDgo = 5,400 mg/kg) resulted in a slight but
significant increase of the reversion frequency by a factor of 2.37.
Results of a micronucleus test were negative after twofold oral
applications of 700, 1,850 or 6,000 mg/kg to Swiss albino mice;
it was concluded that ETU induces hardly any chromosomal anomaly
in the bone marrow. No dominant-lethal effect was observed after
single oral doses of 500, 1,000 and 3,500 mg/kg were given to male
mice.
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Ethylene Thiourea August, 1987
-10-
Carcinogenicity
0 Graham et al. (1975) reported that ETU was a follicular thyroid
carcinogen in male and female Charles Raver rats that were fed the
compound (purity not specified) for 2 years at dietary levels of
250 and 500 ppm (approximately 12.5 and 25 mg/kg/day based on
Lehman, 1959).
9 In a survey of several compounds for tumongenicity, Innes et al.
(1969) reported that ETU (purity not stated) administered by diet to
two strains of specific pathogen-free hybrid mice at a daily dosage
of 215 mg/kg/day for 18 months resulted in statistically significant
(p <0.01) increases in hepatomas (14/16 or 18/18 for males and 18/18
or 9/16 for females) and in total tumor incidence. Pulmonary tumors
and lymphomas were also investigated, but were not found to occur in
the ETU group. The thyroid was not evaluated in this study. No
other dose level was tested.
0 Dose-related incidences of follicular and papillary thyroid cancers
with pulmonary metastases and related lesions such as thyroid solid-
cell adenomas were reported in Charles River CD rats by Ulland et al.
(1972). Ethylene thiourea (97% pure) was administered by diet for
18 months at 175 or 350 ppm followed by administration of a control
diet for 6 months. Assuming that 1 ppm in the diet of older rats is
equivalent to approximately 0.05 nig/kg/day (Lehman, 1959) these
levels correspond to doses of about 8.75 and 17.5 mg/kg/day. The
first tumor was found after 68 weeks, and most were detected after
18 to 24 months when the study was terminated. The statistical
significance of the reported findings was not addressed.
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 mg/kg bw/day.
BW = assumed body weight of a child {1 0 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODri guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
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Ethylene Thiourea August, 1987
-11-
One-day Health Advisory
No data located in the available literature were suitable for determination
of the One-day HA value. It is therefore recommended that the Ten-day HA
value for the 10-kg child (0.25 mg/L, calculated below) be used at this time
as a conservative estimate of the One-day HA value.
Ten-Day Health Advisory
The study by Freudenthal (1977) has been selected to serve as the basis
for determination of the Ten-day HA for a 10-kg child. ETU was fed to a
group of rats (20/sex/group) for up to 90 days at levels of 0, 1, 5, 25, 125
or 625 ppm (0, 0.1, 0.5, 2.5, 12.5 or 62.5 mg/kg/day assuming that 1 ppm in
the diet of a young rat equals 0.1 mg/kg/day, based on Lehman, 1959),. Toxic
effects on thyroid function and morphology were observed after 30 days'
exposure to 125 ppm or greater. No adverse effects were noted in the 25-ppm
group (2.5 mg/kg). Developmental effects reported in other studies have been
reported in rats exposed _in utero at 5 mg/kg (delayed parietal ossification)
(Khera, 1973). The adversity of this effect is unclear. Khera and Iverson
(1978) have reported maternal toxicity and death in cats exposed to 10 mg/kg.
Therefore, 2.5 mg/kg was selected as a conservative NOAEL for deriving the
Ten-day HA.
Using the NOAEL of 2.5 mg/kg/day, the Ten-day HA for a 10-kg child is
' calculated as follows:
Ten-day HA = (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (250 ug/L)
(100) (1 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of fetal or maternal toxicity
in rats exposed to ETU for 30 days.
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
The study by Graham et al. (1973) has been selected to serve as the
basis for determination of the Longer-term HA. In a 12-month study, 1311
uptake was measured in male and female Charles River rats fed ETU (purity not
specified) in the diet at 5, 25, 125, 250 or 500 ppm for 2, 6 or 12 months.
Assuming that 1 ppm in the diet of older rats is equivalent to approximately
0.05 mg/kg/day (Lehman, 1959), these levels correspond to doses of about
0.25, 1.25, 6.25, 12.5 or 25 mg/kg/day.
Adverse effects were noted at all three test intervals. At 12 months,
significant decreases in body weight and increases in thyroid weight were
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Etnylene Thiourea August, 1987
-12-
seen at the 125-, 250- and 500-ppm levels. Uptake of 131I was significantly
decreased in male rats after 12 months at 500 ppm but was increased in females.
Microscopic examination of the thyroids revealed the development of nodular
hyperplasia at dose levels of 125 ppm and higher. The NOAEL for thyroid
effects in this study was 25 ppm (approximately 1.25 mg/kg/day).
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (1 «25 ""?/*g/d«y> MO *g> = 0.125 mg/L (125 ug/L)
y (100) (1 L/day)
where:
1.25 mg/kg/day = NOAEL, based on absence of thyroid effects in male
rats exposed to ETU in the diet for up to 12 months.
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 water consumption by a 10-kg child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = 1'25 mg/kg/day) (70 kg) = 0.44 mg/L (440 ug/L)
(100) (2 L/day)
where:
1.25 mg/kg/day = NOAEL, based on absence of thyroid effects in male
rats exposed to ETU in the diet for up to 12 months.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption by 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., dnnkimj
water) lifetime exposure level, assuming 100% exposure from that medium, at
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Ethylene Thiourea August, 1987
-13-
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, 1986b), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Graham et al. (1975) was selected as the most appropriate
basis for the calculation of a DWEL. In this 2-year study (presumably a
continuation of the Graham et al. (1973) study, Charles River rats were fed
ETU (purity not stated) in the diet at 5, 25, 125, 250 or 500 ppm (approxi-
mately 0.25, 1.25, 6.25, 12.5 or 25 mg/kg/day based on Lehman, 1959).
Statistically significant (p <0.01) decreases in body weight were observed
in both sexes fed at 500 ppm. Increases (p <0.01) in thyroid-to-body weight
ratios were apparent at 250 and 500 ppm. There was an increased iodine (131i)
uptake at 5 and 125 ppm and a decreased uptake at 500 ppm as well as slight
thyroid hyperplasia at the 5- and 25-ppm dose levels (statistical
significance not stated). This effect is considered to be biologically
significant. Tumors were evident in animals in the 125-ppm group. Based on
these results, the LOAEL for lifetime exposure was identified as 5 ppm
(approximately 0.25 mg/kg/day).
Using the LOAEL of 0.25 mg/kg/day, the DWEL is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (Q-25 mg/kg/day) = 0.000025 mg/kg/day (0.03 ug/kg/day)
(1,000) (10)
where:
0.25 mg/kg/day = LOAEL, based on increased iodine intake as well as
thyroid hyperplasia in rats exposed to ETU in the
diet for 2 years.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
10 = additional uncertainty factor to account for the
severity of effect and response at this dose level.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.00003 mg/kg/day) (70 kg) = Q. 00105 mg/L (1.05 ug/L)
(2 L/day)
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Ethylene Thiourea August, 1987
-14-
where:
0.00003 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
According to EPA's guidelines for assessment of carcinogenic risk (U.S.
EPA, 1986b),
ETU is classified in Group B: Probable human carcinogen.
Therefore, a Lifetime Health Advisory is not recommended for ETU. The
estimated cancer risk level associated with lifetime exposure to ETU at
1.05 ug/L is approximately 2.5 x 10-5.
Evaluation of Carcinogenic Potential
0 Three studies that evaluated the carcinogenic potential of ETU were
identified. The results of these studies indicate that ETU is a
thyroid carcinogen in rats (Graham et al., 1975; Ulland et al., 1972)
and increases the incidence of hepatomas as well as total tumor
incidence in mice (Innes et al., 1969).
0 Graham et al. (1975) reported ETU to be a thyroid carcinogen in male
and female Charles River rats that were fed the compound (purity not
specified) for 2 years at dietary levels of 250 and 500 ppm (approxi-
mately 12.5 and 25 mg/kg/day in the diet of older rats based on
Lehman, 1959). At 125 ppm (approximately 6.3 mg/kg/day), ETU was a
thyroid oncogen.
0 Dose-related incidences of follicular and papillary thyroid cancers
with pulmonary raetastases and related lesions such as thyroid solid-
cell adenomas were reported in Charles River CD rats by Ulland et al.
(1972). Ethylene thiourea (97% pure) was administered in the diet
for 18 months at 175 and 350 ppm followed by administration of a
control diet for 6 months. Assuming that 1 ppm in the diet of older
rats is equivalent to approximately 0.05 mg/kg/day (Lehman, 1959),
these levels correspond to doses of about 8.75 and 17.5 mg/kg/day.
The first tumor was found after 68 weeks, and most were detected
after 18 to 24 months when the study was terminated. The statistical
significance of the reported findings was not addressed.
0 Innes et al.
(1969) reported that ETU (purity not stated) administered by diet to
specific pathogen-free hybrid mice at a daily dosage of 215 mg/kg/day
for 18 months resulted in statistically significant (p <0.01) increases
in hepatomas and in total tumor incidence. No other dose level was
tested. (Pulmonary tumors and lymphomas were also investigated in
this study.)
0 Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986b), ETU may be classified in
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Ethylene Thiourea August, 1987
-15-
Group B2: probable human carcinogen based on sufficient evidence
from animal studies.
0 The EPA Carcinogen Assessment Group estimated a one-hit slope of
0.1428/mg/kg/day based on the Innes et al. (1969) study identifying
male mouse liver tumors as the sensitive sex/species end point (U.S.
EPA, 1979). An assumed consumption of 2 liters of water per day by a
70-kg adult over a lifetime results in drinking water concen-
trations of 25, 2.4 and 0.24 ug/L for 10-4, 1 Q-5 and 10-6 cancer risk
levels, respectively.
0 Data are not available to estimate excess cancer risks using other
mathematical models.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No other data have been located for ETU.
VII. ANALYTICAL METHODS
0 Analysis of ethylene thiourea is by a gas chromatographic (GC) method
applicable to water samples (Method #6 - Determination of Ethylene
Thiourea in Ground Water by Gas Chromatography with a Nitrogen-
Phosphorus Detector, 1987). In this method, the ionic strength and
pH of 50-ml of the sample is adjusted. The sample is extracted in a
column and then eluted with raethylene chloride. The extract is
solvent exchanged to ethyl acetate and concentrated to 5-ml. Compounds
are separated using bonded fused silica capillary column GC. Measure-
ment is made using a nitrogen-phosphorus detector. The estimated
detection limit for ethylene thiourea using this method is 5 ug/1.
VIII. TREATMENT TECHNOLOGIES
0 No data were found on the removal of ethylene thiourea from drinking
water by conventional treatment.
0 No data were found on the removal of ethylene thiourea from drinking
water by activated carbon adsorption. However, since ethylene thiourea
has a high solubility and is hydrophylic, treatment with activated
carbon probably would not be effective.
0 No data were found on the removal of ethylene thiourea from drinking
water by ion exchange. However, the structure of ethylene thiourea
indicates it is not ionic and thus probably would not be amenable to
ion exchange.
0 No data were found on the removal of ethylene thiourea from drinking
water by aeration. Since vapor pressure data are unavailable, Henry's
Coefficient, and thus the effectiveness of aeration, cannot be
estimated. However, the high melting point and the high solubility
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Ethylene Thiourea August, 1987
-16-
indicate that Henry's Coefficient would be low and that aeration or
air stripping probably would not be an effective form of treatment*
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Ethylene Thiourea August, 1987
-17-
IX. REFERENCES
Allen, J.R., J.P. Von Miller and J.L. Seymour. 1978. Absorption, tissue
distribution and excretion of 14C ethylenethiourea by the Rhesus monkey
and rat. Res. Comm. Chem. Path. Pharmacol. 20:109-115.
Arnold, D.L., D.R. Krewski, D.B. Jenkins , P.F. McGuire, C.A. Moodie and
I.C. Munro. 1983. Reversibility of ethylenethiourea-induced lesions.
Toxicol. Appl. Pharmacol. 67:264-273.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD.
Chernoff, N., R.J. Kavlock, E.H. Rogers, B.D. Carver and S. Murray. 1979.
Perinatal toxicity of Maneb, ethylene thiourea, and ethylenebisthio-
cyanate sulfide in rodents. J. Toxicol. Environ. Health. 5:821-334.
Freudenthal, R.I. 1977. Dietary subacute toxicity of ethylene thiourea in
the laboratory rat. EPA-600/1-77-023. Health Effects Research Lab.
U.S. EPA, Office of Research and Development, Research Triangle Park,
North Carolina 27711.
Graham, S.L. and W.H. Hansen. 1972. Effects of short-term administration
of ethylene thiourea upon thyroid function of the rat. Bull. Environ.
Contam. Toxicol. 7(1):19-25.
Graham, S.L., W.H. Hansen, K.J. Davis and C.H. Perry. 1973. One-year
administration of ethylenethiourea upon the thyroid of the rat. J. Agr.
Food Chem. 21:324-329.
Graham, S.L., K.J. Davis, W.H. Hansen and C.H. Graham. 1975. Effects of
prolonged ethylene thiourea ingestion on the thyroid of the rat. Food
Cosmet. Toxicol. 13:493-499.
Innes, J.R., B.M. Ulland, M.G. Valeric, L. Petrucelli, L. Fishbein, E.R. Hart
and A.J. Pallotta. 1969. Bioassay of pesticides and industrial chemicals
for tumorigeniciy in mice: A preliminary note. J. Natl. Cancer Inst.
42:1101-1114.
Iverson, F., K.S. Khera and S.L. Hierlihy. 1980. In vivo and in vitro
metabolism of ethylene thiourea in the rat and the cat. Toxicol. Appl.
Pharmacol. 52:16-21.
Khera, K.S. 1973. Ethylene thiourea: teratogenicity study in rats and
rabbits. Teratology. 7:243-252.
Khera, K.S. 1982. Reduction of teratogenic effects of ethylenethiourea in
rats by interaction with sodium nitrite iri vivo. Food Cosmet. Toxicol.
20:273-278.
Khera, K.S. 1984. Etnylenethiourea-induced hindpaw deformities in mice and
effects of metabolic modifiers on their occurrence. J. Toxicol. Environ.
Health. 13:747-756.
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Ethylene Thiourea August, 1987
-18-
Khera, K.S.. and F. Iverson. 1978. Toxicity of ethylenethiourea in pregnant
cats. Teratology. 18:311-314.
Khera, K.S., and L. Tryphonas. 1985. Nerve cell degeneration and progeny
survival following ethylenethiourea treatnent during pregnancy in rats.
Neurol. Toxicol. 6:97-102.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
pesticides. Published in the Assoc. of Food and Drug Officals of the U.S.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Plasterer, M.R., W.S. Bradshaw, G.M. Booth, M.w. Carter, R.L. Schuler and
B.D. Hardin. 1985. Developmental toxicity of nine selected compounds
following prenatal exposure in the mouse: Naphthalene, p-nitrophenol,
sodium selenite, dimethyl phthalate, ethylenethiourea, and four glycol
ether derivatives. J. Toxicol. Environ. Health. 15:25-38.
Ruddick, J.A. and K.S. Khera. 1975. Pattern of anomalies following single
oral doses of ethylenethiourea to pregnant rats. Teratology. 12:277-282.
Sato, K., N. Nakagata, C.F. Hung, M. Wada, T. Shimoji and S. Ishii. 1985.
Transplacental induction of myeloschisis associated with hindbrain
crowding and other malformations in the central nervous system in Long-
Evans rats. Child Nerv. Syst. 1:137-144.
Schubach, M. and H. Hummler. 1977. A comparative study on the mutagenicity
of ethylenethiourea in bacterial and mammalian test systems. Mut. Res.
56:111-120.
Seiler, J.P. 1973. Ethylenethiourea (ETU), a carcinogenic and mutagenic
metabolite of ethylenebis-diothiocarbamate. Mut. Res. 26:189-191.
STORET. 1987.
Teramoto, S., R. Saito and Y. Shirasu. 1980. Teratogenic effects of combined
administration of ethylenethiourea and nitrite in mice. Teratology.
21:71-78.
Teramoto, S., A. Shingu, M. Kaneda, R. Saito, T. Harada, Y. Karo and Y. Shirasu.
1975. Teratogenicity of ethylenethiourea in rats. II. Mode of terato-
genic action. Teratology. 12:216.
Ulland, 3.M., J.H. Weisburger, E.K. Weisburger, J.M. Rice and R. Cypher. 1972.
Brief communication: Thyroid cancer in rats from ethylene thiourea intake.
J. Natl. Cancer Inst. 49:583-584.
U.S. EPA. 1979. U.S. Environmental Protection Agency. The Carcinogen
Assessment Group's Risk Assessment on Ethylene Bisdithiocarbamate.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Ethylene Bisdithio-
carbamate Pesticides. Decision Document. Final Resolution of Rebuttable
Presumption Against Registration. Office of Pesticide Programs.
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Ethylene Thiourea August, 1987
-19-
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Final report.
Task 2: Environmental Fate and Exposure Assessment. June 10.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (1 85):33992-34003. September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Method #6 - Determi-
nation of Ethylene Thiourea (ETU) in Ground Water by Gas Chromatography
with a Nitrogen-Phosphorus Detector, 1987 Draft. Available from U.S.
EPA's Environmental Monitoring and Support Laboratory, Cincinnati, OH
45268.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, 10th ed. Rahway, N.J.: Merck and Co., Inc.
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August, 1987
FENAMIPHOS
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 ible 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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Fenamiphos August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 22224-92-6
Structural Formula
0 H
0-P-N-CH(CH3)2
OC,H5
(1-Methylethyl)-ethyl-3-methyl-4-(methylthio)phenyl-phosphoramidate
Synonyms
0 Nemacur; B 68138; Bay 68138; Bayer 63138; ENT 27572; Phenamiphos
(Meister, 1983).
Uses
o
0 Systemic nematicide (Meister, 1983).
Properties (Meister, 1983)
Chemical Formula C1 3^263^?
Molecular Weight 303 (calculated)
Physical State (at 25°C) Brown, waxy solid
Boiling Point —
Melting Point 49.2°C
Density —
Vapor Pressure (30°C) 7.5 x 1C'7 mr. Hg
Water Solubility (25°C) 400 =ig/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Cor version Factor --
Occurrence
0 Fenamiphos has been found in only 2 ground water samples out of
452 analyzed (STORET, 1987). Botn locations were in California with
the highest concentration found being 5 ug/L. No surface water
locations were tested.
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Fenamiphos August, 1987
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Environmental Fate
0 Ring-labeled 1*C-fenamiphos (radiochemical purity 94%), at 1 and 10 ppm,
degraded with half-lives of 7 to 14 days in a buffered aqueous solution
at pH 3 and >30 days at pH 9, and appeared to be stable at pH 7 when
incubated in the dark at 30°C (McNamara and Wilson, 1981). In the
pH 3 buffer solution, the primary degradation product was deaminated
fenamiphos accounting for 74 to 78% of the applied material. Degradates
identified in methylene chloride extracts from the pH 3, 7 and 9
solutions included fenamiphos sulfoxide, fenamiphos sulfone, fenamiphos
phenol, fenamiphos sulfoxide phenol and fenamiphos sulfone phenol.
0 Ring-labeled ! ^-fenamiphos (radiochemical purity >99%), at 12 ppm,
degraded with a half-life of 2 to 4 hours in pH 7 buffered water
irradiated with artificial light (approximately 5200 uW/cm2, 300 to
600 nm) (Dime et al., 1983). After 24 hours of irradiation, fenamiphos
accounted for approximately 4% of the applied radioactivity, fenamiphos
sulfonic acid phenol for approximately 19%, fenamiphos sulfoxide for
approximately 17%, fenamiphos sulfonic acid for approximately 6% (tenta-
tive identification), and fenamiphos sulfoxide phenol for approximately
1%. In the dark control, fenamiphos accounted for approximately 94% of
the applied at 24 hours post-treatment.
0 Ring-labeled 14c-fenamiphos (radiochemical purity >99%), at approxi-
mately 20 ppm, degraded with a half-life of <1 hour on sandy loam soil
irradiated with artificial light (approximately 6200 uW/cm2, 300 to
600 nm) (Dime et al., 1983). After 48 hours of irradiation, fenamiphos
and the degradates fenamiphos sulfoxide and fenamiphos sulfone accounted
for approximately 12, 55 and 6% of the extractable radioactivity,
respectively. In the dark control, fenamiphos accounted for approxi-
mately 93% of the extractable compound at 48 hours post-treatment.
o 14c-Fenamiphos (purity 86%), at 3 ppm, degraded with a half-life of
<4 days in silty clay loam soil previously treated with fenamiphos
(Green et al., 1982). Fenamiphos sulfoxide comprised up to approxi-
mately 74% of the applied radioactivity (maximum at 11 days post-
treatment); fenamiphos sulfone comprised approximately 10% and volatile
14c-residues comprised 17% of the applied material at 55 days post-
treatment. At 55 days post-treatment, 1.13% of the applied fenamiphos
remained undegraded in the soil previously treated with fenamiphos,
5.41% remained undegraded in soil with no prior history of fenamiphos
treatment, and 40.58% remained undegraded in sterile soil. Fenamiphos
sulfoxide was the major degradate in all three treatments.
o 14c-Fenamiphos (test substance uncharacterized), at 0.29 to 2.30 ug/ml
of water, was adsorbed to sandy loam and clay loam soils with 26.3 to
30.0% and 42.2 to 52.3% of the applied radioactivity, respectively,
adsorbed after 16 hours (Church, 1970).
0 Fenamiphos (3 Ib/gallon SC and 15% G), at approximately 20 Ib ai/A,
was mobile in columns (16-cm length) of sandy soil eluted with 10
inches of water. Fenamiphos was detected throughout the columns, and
0.9 to 2.2% of the applied material was recovered in the leachate
(Gronberg and Atwell, 1980).
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. August, 1987
Fenamiphos ^
-4-
0 Aged (30 days) 1*C-fenamiphos residues, at approximately 4 Ib ai/A,
were slightly mobile in a column (12-inch length) of sandy loam soil
leached with 22.5 inches of water; approximately 2.3% of the applied
radioactivity leached from the column and approximately 91% of the
applied radioactivcity remained in the top 5 inches of the soil
column (Tweedy and Houseworth, 1980).
III. PHARMACOKINETICS
Absorption
0 Gronb.erg (1969) administered ! 4C-labeled fenamiphos (99% purity)
by oral intubation to rats. Only 5 to 7% was recovered in feces,
indicating that 93 to 95% was absorbed from the gastrointestinal
tract.
Distribution
0 Gronberg (1969) administered single oral doses of 2 mg/kg of ethyl-
14c-fenamiphos (99% purity) by oral intubation to rats. Forty-eight
hours after treatment, residues measured in tissues were: brain
<0.1 ppm; heart 0.1 ppm; liver 0.8 to 1.7 ppm; kidney 0.4 to 0.5 ppm;
fat 0.2 to 0.4 ppm; muscle <0.1 ppm; and gastrointestinal tract 0.2 ppm.
Metabolism
0 In studies conducted by Gronberg (1969), rats were administered 2 mg/kg
oral doses of fenamiphos (99% purity) using ethyl-14c, methylthio-3H or
isopropyl-14C label. The authors proposed a pathway of fenamiphos
metabolism involving oxidation to the sulfoxide and sulfone analogs.
Subsequent hydrolysis, conjugation and excretion via urine gave high
molecular-weight compounds (600 to 800). No other details were
provided.
Excretion
Gronberg (1969) administered ethyl-14c, methylthio-3H or isopropyl-
14c-labeled fenamiphos (2 mg/kg, 99% purity) to rats by gavage.
Thirty-nine to forty-two percent or 50% of the administered radio-
activity was expired as C02, respectively. Thirty-eight to 40% of
the ethyl-1 4c labels were in urine and 5% in feces, respectively.
Eighty percent of the methylthio-3H label was found in urine. The
majority of the administered dose was excreted 12 to 15 hours after
treatment.
IV. HEALTH EFFECTS
Humans
No information on the health effects of fenamiphos in humans was
found in the available literature.
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Fenamiphos August, 1987
-5-
Animals
Short-term Exposure
0 NIOSH (1985) reported the acute oral LDso of fenamiphos in the rat,
mouse, dog, cat, rabbit and guinea pig as 8, 22.7, 10, 10, 10 and
75 mg/kg, respectively.
0 Kimmerle and Lorke (1970) fed chickens (eight/dose) diets containing
technical fenamiphos at levels of 0, 1, 3, 10 or 30 ppm active
ingredient (a.i.) for 30 days. The authors stated that this corre-
sponded to doses of 0, 2, 5, 16 or 26 mg/kg/day. Following treatment,
feed consumption, neurotoxicity and cholinesterase (ChE) activity
were determined. Histopathological sections of the brain, spinal
cord and peripheral nerves were also evaluated. No neuropathy was
observed at any dose level tested. No ChE symptoms were reported,
but ChE activity in whole blood was inhibited in a dose-dependent
manner from 21% at 3 ppm to 65% at 30 ppm. Based on ChE inhibition,
a No-Observed-Adverse-Effect-Level (NOAEL) of 1 ppm (2 mg/kg/day) was
identified.
Dermal/Ocular Effects
0 DuBois et al. (1967) .reported acute dermal LD50 values of 78 mg/kg
for rats and 55 mg/kg for guinea pigs.
0 Crawford and Anderson (1973) applied 120 mg of a spray concentrate of
fenamiphos (37.47% a.i.) to shaved intact and abraded skin of six New
Zealand White rabbits and reported slight erythema 24 and 72 hours
post-treatment.
0 In ocular studies conducted by Crawford and Anderson (1973), the
instillation of 0.1 mL of a spray concentrate of fenamiphos (37.47%
a.i. ) into the eyes of New Zealand White rabbits resulted in corneal
and conjunctival damage at 24 and 72 hours post-treatment. These
effects had not subsided by 21 days post-treatment.
Long-term Exposure
0 In feeding studies conducted by Mobay Chemical Corporation (1983),
Fischer 344 rats (50/sex/dose) were administered technical fenamiphos
(89% purity) at dose levels of 0, 0.36, 0.60 or 1.0 ppm a.i. for
90 days. Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), this corresponds to dose levels of 0,
0.018, 0.03 or 0.05 mg/kg/day. Following treatment, brain, plasma
and erythrocyte ChE levels were measured. Cholinesterase levels were
not significantly reduced at any dose tested. Other parameters were
not evaluated. The author reported a NOAEL of 1 ppm (0.05 mg/kg/day,
the highest dose tested).
0 Loser and Kimmerle (1968) fed Wistar rats (15/sex/dose) fenamiphos
for 90 days in the diet at dose levels of 0, 4, 8, 16 or 32 ppm
active ingredient. Assuming that 1 ppm in the diet is equivalent to
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Fenamiphos ugust, 1987
-6-
0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of 0, 0.2,
0.4, 0.8 or 1.6 mg/kg/day. Following treatment, body weight, food
consumption, hematology, ChE activity, urinalysis and gross pathology
were evaluated. No histologic examination was performed. No effects
on any end point were reported except for ChE inhibition. No effect
was seen at 4 ppm (0.2 mg/kg/day). At 8 ppm (0.4 mg/kg/day), ChE in
whole blood and plasma was decreased by 11% and 19%, respectively.
Higher doses produced larger decreases in ChE. Based on these data,
a NOAEL of 4 ppm (0.2 mg/kg/day) was identified.
0 Loser (1970) administered technical fenamiphos (99.4% purity) in the
feed of beagle dogs (two/sex/dose) for 3 months at dietary levels of
0, 1, 2 or 5 ppm. Assuming that 1 ppm in the diet of dogs is equivalent
to 0.025 mg/kg/day (Lehman, 1959), this corresponds to doses of 0,
0.025, 0.05 or 0.125 mg/kg/day. Untreated controls (three/sex) were
run concurrently. Following treatment, body weight, feed consumption,
clinical chemistry, urinalysis, ChE activity and gross pathology were
evaluated. At 5 ppm, there was some slight decrease in weight gain,
although the author did not consider this to be important. No compound-
related effects were reported in any other parameters measured except
ChE activity. At 1 ppm, plasma ChE was inhibited 13% and 18%, and
red blood cell ChE was inhibited 6% and 19% in males and females,
respectively. At 2 ppm, plasma and red blood cell ChE was comparable
to control levels in males, and was inhibited 13% in pla-sma and 15%
in red blood cells in females. At 5 ppm, ChE in plasma was inhibited
44% and 41%,and red blood cell ChE was inhibited 22% and 26% (females
and males, respectively). No brain ChE measurements were reported.
Based on the absence of significant (>20%) ChE inhibition at 1 or
2 ppm, a NOAEL of 2 ppm (0.05 mg/kg/day) is identified.
0 Hayes et al. (1982) administered fenamiphos (90% purity) in the diet
to CD albino mice (50/sex/dose) at dose levels of 0, 2, 10 or 50 ppm
for 18 months. Assuming that 1 ppm in the diet of mice is equivalent
to 0.15 mg/kg/day (Lehman, 1959), this corresponds to doses of 0, 0.3,
1.5 or 7.5 mg/kg/day. Following treatment, body weight, food con-
sumption, hematology and mortality were evaluated. Absolute brain
weights were decreased at 2 ppm (0.3 mg/kg/day) or greater. At 50 ppm
(7.5 mg/kg/day), there was a decrease in body weight. Based on these
data, a Lowest-Observed-Adverse-Effect-Level (LOAEL) of 2 ppm (0.3
mg/kg/day), lowest dose tested, was identified, but not a NOAEL.
0 Loser (1972a) administered technical fenamiphos (78.8% purity) in the
diet of Wistar rats (40/sex/dose) for 2 years at dose levels of 0, 3,
10 or 30 ppm a.i. Assuming that 1 ppm in the diet of rats is equiva-
lent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of
0, 0.15, 0.5 or 1.5 mg/kg/day. Untreated controls (50 males, 60
females) were run concurrently. Following treatment, body weight,
food consumption, hematology, urinalysis, plasma and erythrocyte ChE
activity, gross pathology and histopathology were evaluated. At the
highest dose (30 ppm), a slight increase in female mortality (38%
versus 29% in controls) was noted, but the author did not consider
this significant. There were statistically significant (p <0.05)
increases in thyroid gland and lung weights in females and in heart
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Fenamiphos Au*ust' 1987
-7-
weight in males. No compound-related effects were observed in any of
the other parameters measured except an inactivation of plasma and
erythrocyte ChE. At 10 ppm, ChE was decreased by 18 to 41%, and at
30 ppm, ChE was decreased by 28 to 60%. No brain ChE measurements
were reported. Based on ChE inhibition, the author identified a NOAEL
of 3 ppm (0.15 mg/kg/day). Based on organ weight changes, the NOAEL
was 10 ppm (0.5 mg/kg/day).
e In chronic feeding studies by Loser (1972b), beagle dogs (four/sex/dose)
were administered technical fenamiphos (78.8% purity) in the feed for
2 years at 0, 0.5, 1, 2, 5 or 10 ppm active ingredient. Assuming
that 1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day (Lehman,
1959), this corresponds to doses of 0, 0.013, 0.025, 0.050, 0.125 or
0.250 mg/kg/day. Following treatment, no compound-related effects
were observed on appearance, general behavior, food consumption,
clinical chemistry, hematology, gross pathology or histopathology at
any dose tested. Plasma and erythrocyte ChE levels were inhibited
about 26% at 2 ppm, about 21 to 57% at 5 ppo and about 32 to 51% at
10 ppm. Cholinesterase was not inhibited at 1 ppm (0.025 mg/kg/day)
or below. Based on ChE inhibition, this study identified a NOAEL of
1 ppm (0.025 mg/kg/day) and a LOAEL of 2 ppm (0.05 mg/kg/day).
Reproductive Effects
0 In a three-generation study conducted by Loser (1972c), FB30 rats
(10 males or 20 females/dose) were fed technical fenamiphos (78.8%)
in the diet at dose levels of 0, 3, 10 or 30 ppm active ingredient.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 0, 0.15, 0.5 or 1.5
mg/kg/day. Fertility, lactation performance, pup development and
parental and litter body weights were evaluated. No compound-related
effects were observed in any parameter in animals exposed to 10 ppm
(0.5 mg/kg/day) or less. At 30 ppm (1.5 mg/kg/day), one male of the
F2b generation showed a lower body weight gain than the untreated
controls, but there were no differences in body weight gain in any
other generation. Based on these data, a reproductive NOAEL of 30
ppm (1.5 mg/kg/day) was identified.
Developmental Effects
0 MacKenzie et al. (1982) administered fenamiphos (88% a.i. by gavage
to pregnant New Zealand White rabbits (20/dose) at dose levels of 0,
0.1, 0.3 or 1.0 mg/kg/day on days 6 to 18 of gestation. Following
treatment, there was a decrease in maternal body weight at 0.3 mg/kg/day
or above. At the 1.0-mg/kg/day level, eight dead pups and seven late
resorptions were reported, and fetal weight was depressed. A signifi-
cant (p <0.05) increase in the incidence of chain-fused sternebrae
was also observed at 1.0 mg/kg. Based on maternal body weight, a
NOAEL of 0.1 mg/kg was identified. Based on fetotoxicity and terato-
genicity, a NOAEL of 0.3 mg/kg/day was identified.
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Fenamiphos Au*ust' 1987
-8-
Mutaqenicity
0 Herbold (1979) reported that fenamiphos was not mutagenic in Salmonella
typhimurium (TA 1535, 1537, 98 or 100) up to 2,500 ug/plate, either
with or without activation.
0 In a dominant lethal test with male mice {Herbold and Lorke, 1980),
acute oral doses of 5 mg/kg did not produce mutagenic effects.
Carcinoqenicity
0 Hayes et al. (1982) administered fenamiphos (90% purity) for 18 months
in the diet to CD albino mice (50/sex/dose) at dose levels of 0, 2, 10
or 50 ppm (0, 0.3, 1.5 or 7.5 mg/kg/day). Based on gross and histo-
pathologic examination, neoplasms in various tissues and organs were
similar in type, organization, time of occurrence and incidence in
control and treated animals.
0 Loser (1972a) administered technical fenamiphos (78.8% purity) in the
diet of Wistar rats (40/sex/dose) for 2 years at dose levels of 3, 10
or 30 ppm active ingredient. Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of 0.15, 0.5 or 1.5 mg/kg/day. Untreated controls (50 males,
60 females) were run concurrently. No evidence of carcinogenicity
was detected either by gross or histological examination.
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/u ( ug/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, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption by a child
(1 L/day) or an adult (2 L/day).
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Fenamiphos August, 1987
-9-
One-day Health Advisory
No information was found in the available literature that was suitable
for determination of the One-day HA value for fenamiphos. It is therefore
recommended that the Ten-day HA value for the 10-kg child of 0.009 mg/L (9 ug/L,
calculated below) be used at this time as a conservative estimate of the
One-day HA value.
Ten-day Health Advisory
The study by MacKenzie et al. (1982) has been selected to serve as the
basis for determination the Ten-day HA value for fenamiphos. In this study,
pregnant rabbits (20/dose) were administered technical fenamiphos (88% purity)
by gavage at dose levels of 0, 0.1, 0.3 or 1.0 rag/kg on days 6 through 18 of
gestation. A decrease in maternal body weight was observed in animals dosed
with 0.3 mg/kg/day or above. No maternal toxicity was reported at 0.1 mg/kg/day.
No fetotoxicity or teratogenic effects were observed at 1.0 mg/kg or less or
0.3 mg/kg or less, respectively. Chain fusion of sternebrae were observed in
the 1.0 mg/kg group. Based on maternal effects, a NOAEL of 0.1 mg/kg/day was
identified.
Using a NOAEL of 0.1 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ton-dav HA - (0'1 "9Aq/day) (10 kg) (0.88) = Q.009 mg/L (9 ug/L)
y (100) (1 L/day)
where:
0.1 mg/kg/day = NOAEL, based on absence of maternal or fetal toxicity
in rabbits exposed to fenamiphos via gavage on days
6 through 18 of gestation.
10 kg = assumed body weight of a child.
0.88 = correction factor to account for 88% active ingredient
in administered doses.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = asoumed water consumption of a child.
Longer-term Health Advisory
The study by Loser (1970) has been selected to serve as the basis for
determination of the Longer-term HA value for fenamiphos. In this study,
beagle dogs (two/sex/dose) were fed technical fenamiphos (99.4% purity) in
the diet at dose levels of 0, 1, 2 or 5 ppm (0, 0.025, 0.05 or 0.125 mg/kg/day)
for 3 months. No effects were detected on body weight, food consumption,
clinical chemistry, urinalysis and gross pathology. The only effect observed
was inhibition of plasma and erythrocyte ChE activity at the 5-ppm dose
level (0.125 mg/kg/day). No significant effect was seen at 2 ppm or less
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Fenamiphos Au*ust' 1987
-10-
(0.05 mg/kg/day), which was identified as the NOAEL. The 90-day study in
F344 rats by Mobay Chemical Corporation (1983) identified a NOAEL of 1 ppm
(0.05 mg/kg/day), but this was not considered, since it was the highest dose
tested and a LOAEL was not identified. The study by Loser and Kimtnerle
(1968) identified a NOAEL of 0.2 mg/kg/day in rats, but this was not chosen,
since available data (Loser et al., 1972a,b) suggest that the rat is less
sensitive than the beagle dog.
Using a NOAEL of 0.05 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (0.05 mg/kg/day) (10 kg) = Oo005 mg/L (5 ug/L)
* (100) (1 L/day)
where:
0.05 mg/kg/day = NOAEL, based on absence of significant cholinesterase
inhibition in dogs exposed to fenamiphos via the diet
for 3 months.
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 by a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (0-05 mg/kg/day) (70 kg) = 0.018 mg/L (18 ug/L)
(100) (2 L/day)
where:
0.05 mg/kg/day = NOAEL, based on absence of significant cholinesterase
inhibition in dogs exposed to fenanuphos via the diet
for 3 months.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/CW
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
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Fenamiphos August, 1987
-11-
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. 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 Loser (1972b) has been selected to serve as the basis for
determination of the Lifetime HA value for fenamiphos. In this study, dogs
(four/sex/dose) were fed technical fenamiphos (78.8% purity) in the diet for
2 years at dose levels of 0, 0.5, 1, 2, 5 or 10 ppm active ingredient (0,
0.013, 0.025, 0.05, 0.125 or 0.25 mg/kg/day). The only effect detected was
inhibition of plasma and erythrocyte cholinesterase at dose levels of 2, 5 or
10 ppm (0.05, 0.125 or 0.25 mg/kg/day). The NOAEL identified in this study
was 1 ppm (0.025 mg/kg/day). The chronic studies in rats by Loser (1972a)
and by Hayes et al. (1982) were not chosen, since the data indicate the rat
is less sensitive than the dog.
Using a NOAEL of 0.025 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0-025 mg/kg/day) = Q.00025 mg/kg/day
(100)
where:
0.025 mg/kg/day = NOAEL, based on absence of cholinesterase inhibition
in dogs exposed to technical fenamiphos via the diet
for 2 years.
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»00025 mg/kg/day) (70 kg) = Q.009 mg/day (9 ug/L)
(2 L/day)
where:
0.00025 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Fenamiphos August, 1987
-12-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.009 mg/L) (20%) = 0.0018 mg/L (1.8 ug/L)
where:
0.009 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic potential was detected in chronic feeding
studies in rats (Loser, 1972a) or mice (Hayes et al., 1982).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of fenamiphos.
8 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), fenamiphos may be classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenic!ty.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Residue tolerances have been established for fenamiphos and its
cholinesterase-inhibiting metabolites in or on various agricultural
commodities at 0.02 to 0.60 ppm based on an ADI for fenamiphos
of 0.0025 mg/kg/day (U.S. EPA, 1985).
G The World Health Organization (WHO) calculated a TADI of 0.0003
mg/kg/day for fenamiphos (Vettorazzi and Van den Hurk, 1985).
VII. ANALYTICAL METHODS
0 There is no standarized method for the determination of fenamiphos
in water samples. A procedure has been reported for the estimation
of fenamiphos and other pesticides in foods and feeds (FDA, 1979).
This procedure involves extraction and isolation in an organic phase;
the extract is then dried and concentrated, and an aliquot of the
concentrated organic phase is injected into a gas chromatograph
equipped with a phosphorus-selective detector.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies used to remove fenamiphos from contaminated water.
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Fenamiphos August, 1987
-13-
IX. REFERENCES
Church, D.D.* 1970. Bay 68138 — leaching, runoff, and water stability.
Report No. 26849. Unpublished study received May 27, 1970 under OF0982;
submitted by Chemagro Corp., Kansas City, MO; CDL:091690-H. MRID 00067117.
Crawford, C., and R. Anderson.* 1973. The eye and sJcin irritancy of Nemacur
3 Ibs/gal spray concentrate to rabbits. Report No. 37549. Unpublished
study. MRID 00119227.
Dime, R.A., C.A. Leslie and R.J. Puhl.* 1983. Photodecomposition of Nemacur
in aqueous solution and on soil. Report No. 86171. Mobay Chemical Corp.
1983. Supplement No. 4 to brochure entitled: Nemacur: The effects on
the environment — environmental chemistry (dated Feb. 1, 1973). Document
No. AS83-2611. Compilation? unpublished study received Dec. 9, 1$83
under 3125-236; CDL:251891-A. MRID 00133402.
DuBois, K.P., M. Flynn and F. Kinoshita.* 1967. The acute toxicity and anti-
cholinesterase action of Bayer 68138. Unpublished study. MRID 00082807.
FDA. 1979. Food and Drug Administration. Pesticide analytical manual.
Revised June 1979.
Green, R., C. Lee and W. Apt.* 1982. Processes affecting pesticides and
other organics in soil and water systems: Assessment of soil and
pesticide properties important to the effective application of nematicides
via irrigation. Hawaii contributing project to Western Regional Research
Project W-82. Unpublished study. MRID 00154533.
Gronberg, R.R.* 1969. The metabolic fate of (Bay 68138), (Bay 68138 sulfoxide),
and (Bay 68138 sulfone) by white rats. Report No. 26759. Unpublished
study. MRID 00052527.
Gronberg, R.R., and S.H. Atwell.* 1980. Leaching of Nemacur residues in
Florida sand. Report No. 66409. Rev. Unpublished study received Aug. 28,
1980 under 3125-236; submitted by Mobay Chemical Corp., Kansas City, MO;
CDL:243126-Y. MRID 00045607.
Hayes, R.H., D.W. Lamb and D.R. Ma'llicoat. * 1982. Technical fenamiphos
oncogenicity study in mice. Report No. 3037. Unpublished study.
MRID 00098614.
Herbold, B.* 1979. Nemacur: Salmonella/microsome test for detection of
Point-mutagenic effects: Report No. 8730; 82210. Unpublished study.
MRID 00121287.
Herbold, B., and D. Lorke.* 1980. SRA 3386: Dominant lethal study on male
mouse to test for mutagenic effects. Report No. 8838; 69377. Unpublished
Study. MRID 00086981.
Kimmerle, G., and D. Lorke.* 1970. Bay 68138: Subchronic neurotoxicity
studies on chickens. Report No. 1831; 27489. Unpublished study.
MRID 00082105.
-------
Fenamiphos August, 1987
-14-
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off.
Loser, E.* 1970. Bay 68138: Subchronic toxicological studies on dogs (three
months feeding test). Report No. 1655; Report No. 26906. Unpublished
study. MRID 00064616.
Loser, E.* 1972a. Bay 68138: Chronic toxicological studies on rats (two-year
feeding experiment). Report No. 3539; Report No. 34344. Unpublished
study. MRID 00038490.
Loser, E.* 1972b. Bay 68138: Chronic toxicological studies on dogs (two-year
feeding experiment). Report No. 3561; Report No. 34345. Unpublished
study. MRID 00037965.
Loser, E.* 1972c. Bay 68138: Generation studies on rats. Report No. 3424;
Report No. 34029. Unpublished study. MRID 00037979.
Loser, E., and G. Kimmerle.* 1968. Bay 68138: Subchronic toxicological study
on rats. Report No. 74523307. Unpublished study. MRID 00082810.
MacKenzie, K., S. Dickie, B. Mitchell et al.* 1982. Teratology study with
Nemacur in rabbits. Unpublished study. MRID 00121286.
McNamara, F.T., and C.M. Wilson.* 1981. Behavior of Nemacur in buffered
aqueous solutions. Report No. 68582. Unpublished study received July 23,
1981 under 3125-236; submitted by Mobay Chemical Corp., Kansas City, MO;
CDL:245613-A. (00079270).
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Mobay Chemical Corporation.* 1983. Combined chronic toxicity/oncogenicity
study of Technical Fenamiphos with rats. Unpublished study. MRID
00130774.
NIOSH. 1985. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances (RTECS). National Library of
Medicine Online File.
Tweedy, B.C., and L.D. Houseworth.* 1980. Leaching of aged Nemacur residues
in sandy loam soil. Report No. 40506. Unpublished study received Aug. 28,
1980 under 3125-236; submitted by Mobay Chemical Corp., Kansas City, MO;
CDL:243126-N. MRID 00045598.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Summary of reported
incidents involving fenamiphos. Pesticide Incident Monitoring System.
Report No. 208. Washington, DC: U.S. Environmental Protection Agency.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.349, p. 324. July 1, 1985.
-------
Fenamiphos August, 1987
-15-
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
Vettorazzi, G., and G.W. Van den Hurk. 1985. Pesticides reference index,
JMPR 1961-1984. p. 10.
•Confidential Business Information.
-------
FLUOMETURON
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
August, 1987
DRAFT
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 -nodel is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Fluometuron August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 2164-17-2
Structural Formula 0
H-N-C-N(CH,)2
N,N-Dimethyl-N-( 3-(trifluoromethyl )phenyl )-urea
Synonyms
0 C 2059; Cotoron; Cottonex; Lanex (Meister, 1983).
Uses
0 Herbicide (Windholz et al., 1983).
Properties (Windholz et al., 1983; CHEMIAB, 1985; TDB, 1985)
Chemical Formula C10H11ON2F3
Molecular Weight 232.21
Physical State (25°C) White crystals
Boiling Point —
Melting Point 163-164.5°C
Density —
Vapor Pressure (20°C) 5 x 1(T7 nri Hg
Specific Gravity
Water Solubility (25»C) 80 mg/L
Octanol/Water Partition 1.88 (calculated)
Coefficient
Taste Threshold —
Odor Threshold
Conversion Factor —
Occurrence
0 Fluometuron was not found in any of 31 ground water samples analyzed
from 29 locations (STORET, 1937). No sjrface water samples were
tested.
Enviroronental Fate
0 l4C-Fluometuron (test substance not characterized) was intermediately
mobile (Rf = 0.50) in a silty clay loam soil (2.5% organic matter)
based on thin-layer chromatograohy (TLC) tests of soil (Helling, 1971;
Helling et al., 1971).
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Fluometuron August, 1987
-3-
0 14c-Fluometuron (test substance not characterized), at various concen-
trations, was very mobile in a Norge loam soil (1.7% organic matter)
with a Freundlich-K of 0.31 (Davidson and McDougal, 1973). Freundlich-K
values, determined in soil:water slurries (5-10 g/100 mL) treated with
14c-fluometuron (test substance not characterized) at 0.05 to 10.0 ppm,
were 0.37 for Uvrier sand (1% organic matter), 1.07 for Collombey sand
(2.2% organic matter), 1.66 for Les Evouettes loam (3.6% organic matter),
3.16 for Vetroz sandy clay loam (5.6% organic matter), and 1.36 for
Illarsatz high organic soil (22.9% organic matter) (Guth, 1972).
0 Fruendlich-K values were positively correlated with the organic matter
content of the soil. Fluometuron (test substance not characterized),
at 10 to 80 uM/kg, was adsorbed at 10 to 51% of the applied amount to a
loamy sand soil (1.15% organic matter) and 16 to 67% of the applied to a
sandy loam soil (1.9% organic matter) in water slurries during a test
period of 1 minute to 7 days, with adsorption increasing with time
(LaFleur, 1979). Approximately 22% of the applied fluometuron desorbed
in water from the loamy sand soil and 15% desorbed from the sandy loam
soil during a 7-day test period.
0 Fluometuron (50% wettable powder, WP) dissipated from the 0- to 5-cm
depth of a sandy clay loam soil (3.2% organic matter) in central
Europe with a half-life of less than 30 days (Guth et al., 1969).
Fluometuron residues (not characterized) dissipated with a half-life
of 30 to 90 days.
III. PHARMACOKINETICS
Absorption
0 Boyd and Foglemann (1967) reported that fluometuron is slowly absorbed
from the gastrointestinal (GI) tract of female CFE rats (200 to 250 g).
Based on the radioactivity recovered in the urine and feces of four
rats given 50 mg 14C-labeled fluometuron after a 2-week pretreatment
with 1,000 ppm unlabeled fluometuron [estimated as 100 mg/kg/day,
assuming 1 ppm equals 0.1 mg/kg/day in the young rat (Lehman, 1959)],
the test compound appears not to have been fully absorbed within 72
hours. Of an orally administered dose (50 mg/kg), up to 15% was
excreted in the urine and 49% in the feces.
Distribution
0 Boyd and Foglemann (1967) detected radioactivity in the liver, kidneys,
adrenals, pituitary, red blood cells, blood plasma and spleen 72 hours
after oral administration of 14C-labeled fluometuron at dose levels of
50 or 500 mg/5cg in rats. The highest concentration was detected in
red blood cells.
Metabolism
0 Boyd and Foglemann (1967) concluded that, by thin-layer chromatographic
analysis, the urine of rats in their study contained m-trifluoromethyl-
aniline, desmethyl-fluometuron, demethylated fluometuron, hydroxylated
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Fluometuron Au*ust« 1987
-4-
desmethyl-fluometuron, hydroxylated demethylated fluometuron, and
hydroxylated aniline.
0 Lin et al. (1976) reported that after incubation of 14CF3-labeled
fluometuron with cultured human embryonic lung cells for up to 72
hours, 95% of the compound remained unchanged. Human embryonic lung
cell homogenate metabolized small amounts of fluometuron through
oxidative pathways to N-(3-trifluoromethylphenyl)-N-formyl-N-methylurea,
N-(3-trifluoromethylphenyl)-N-methylurea, and N-(3-trifluoromethylphenyl)
urea.
Excretion
0 Boyd and Foglemann (1967) reported that urinary excretion of radio-
active label peaked at 24 hours after administration of 14C-fluometuron
(50 rag/kg) and decreased during the remaining 48 hours. Seventy-two
hours after oral administration of the radioactive label, up to 15%
of the administered dose was eliminated in the urine.
• In the study by Boyd and Foglemann (1967), fecal excretion of fluometuron
peaked by 48 hours postdosing and decreased over the remaining 24 hours.
Forty-nine percent of the administered dose (50 mg/kg) was eliminated
in the feces.
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of fluometuron in humans.
Animal^
Short-term Exposure
0 NIOSH (1985) reported the acute oral LD50 values of fluometuron as
6,416, 2,500, 900 and 810 mg/kg in the rat, rabbit, mouse and guinea
pig, respectively.
0 Sachsse and Bathe (1975) reported an acute oral LD50 value of
4,636 mg/kg for both male and female Tif RA1 rats.
• Foglemann (1964a) reported the acute oral LD5(p values for CFW albino
mice as 2,300 mg/kg in females and 900 mg/kg in males.
Dermal/Ocular Exposure
0 Siglin et al. (1981) conducted a primary dermal irritation study in
which undiluted fluometuron powder (80%) was applied to intact and
abraded skin of six young adult New Zealand White rabbits for 24
hours. The test substance was severely irritating, with eschar
formation observed at 24 and 72 hours.
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Fluometuron August, 1987
-5-
0 Foglemann (1964b) exposed the skin of eight albino rabbits (four/sex)
to a 10% aqueous suspension of fluometuron (applied under rubber
dental damming) for 6 hours/day for 10 days. No contact sensitization
developed during the exposure period. Weight depression at day 130
was evident in the treated group.
0 Galloway (1984) reported no sensitizing reactions in Hartley albino
guinea pigs exposed to undiluted fluometuron on alternate days for
22 days and on day 36.
0 Technical fluometuron was not found to be an eye irritant in rabbits
(Foglemann, 1964c).
Long-term Exposure
0 Foglemann (1965a) conducted a 90-day feeding study in which CFE rats
(15/sex/dose) were administered technical fluometuron (purity not
specified) in the diet at dose levels of 100, 1,000 or 10,000 ppm
(reported as 7.5, 75 or 750 mg/kg/day). Following exposure, various
parameters including hematology, clinical chemistry and histopathology
were evaluated. Enlarged, darkened spleens were observed grossly in
male rats given 75 mg/kg/day. At the highest dose level, a depression
in body weight and congestion in the parenchyma of the spleen, adrenals,
liver and kidneys were evident. A mild deposition of hemosiderin in
the spleen was also evident. Spleens were large and dark; livers
were brownish and muddy colored; and kidneys were small with discolored
pelvises in high-dose males. Histopathological findings were confined
to mild congestion in various organs and mild hemosiderin deposits
in the spleens of high-dose rats. No effects were evident in rats
given the 7.5 mg/kg/day dose level for any parameter measured. This
dose level was identified as the No-Observed-Adverse-Effect-Level
(NOAEL) for this study.
0 Foglemann (1965b) administered technical fluometuron (purity not
stated) in feed to three groups of beagle pups (three/sex/dose) at
dose levels of 40, 400 or 4,000 ppm (reported as 1.5, 15 or 150
mg/kg/day) for 90 days. At 150 mg/kg/day, mild inflammatory-type
reactions and congestion in the liver and kidneys and mild congestion
and hemosiderin deposits in the spleen were observed. Also at this
high dose, the spleen to body weight ratio was slightly increased.
No adverse systemic effects were observed in dogs administered 1.5 or
15 mg/kg/day (NOAEL).
0 In the NCI (1980) study, B6C3Fi mice and F344 rats (10 of each sex)
were given fluometuron (>99% pure) in the diet for 90 days to estimate
1,000, 2,000, 4,000, 8,000, and 16,000 ppm. Decreased body weight gain
(>10%) was apparent with doses above 2,000 ppm. Treatment-related
splenomegaly was found in rats with doses above 1,000 ppm. Microscopic
examination was done on spleens only from rats given more than 2,000
ppm, and this assessment indicated dose-related changes including
hyperemia of red pulp with atrophy of Malpighian corpuscles and
depletion of lymphocytic elements. Body weight gain was reduced
(>10%) in male and female mice given more than 2,000 ppm. Assuming
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Fluometuron August, 1987
-6-
that 1 ppm in the diet equals 0.10 ng/kg/day in the young rat and
0.15 mg/kg/day in the mouse (Lehman, 1959), 1,000 ppm (NOAEL)
corresponds to 100 mg/kg/day in rats and 2,000 ppm (NOAEL) corresponds
to 300 mg/kg/day in mice.
0 Hofmann (1966) administered 0, 3, 10, 30 or 100 mg/kg technical
fluometuron (Cotoron = C-2059, purity not specified) as a suspension
in 1% Mulgafarin six times per week for 1 year by pharnyx probe to
four groups of Wistar rats (25/sex/dose). Following treatment,
general behavior, mortality, growth, food consumption, clinical
chemistry, blood, urine, and histopathology were evaluated. Males
dosed with 30 or 100 mg/kg/day and females dosed with 100 mg/kg/day
showed significant (p <0.05) reductions in body weight at the end of
the study compared to controls. No toxicological effects were observed
in rats administered 3 or 10 mg/kg/day (NOAEL).
0 In the NCI (1980) study, F344 rats (10 of each sex) were given
fluometuron (>99% pure) at dietary levels of 250, 500, 1,000, 2,000
and 4,000 ppm in a repeat of the 90-day study to examine splenic
effects more closely. Splenomegaly in all treated groups was noted.
A dose-related increase in spleen weights and a dose-related decrease
in circulating red blood cells was observed in females fed 250 ppm
and higher. Increased spleen weights were evident in males given
doses above 500 ppm. However, statistical analysis of the data was
not done. Stated in the report without presentation of data is the
observation of a dose-related increase in red blood cells with
polychromasia and anisocytosis in male and female rats and congestion
of red pulp with corresponding decrease of white pulp in spleen.
Assuming that 1 ppm equals 0.10 mg/kg/day in the young rat (Lehman,
1959), a Lowest-Observed-Adverse-Effect-Level (LQAEL) of 250 ppm (25
mg/kg/day) is suggested in this study.
0 No noncarcinogenic effects (survival, body weight and pathological
changes) in B6C3Fj mice and F344 rats were found in the NCI (1980)
bioassay discussed under Carcinogenicity.
Reproductive Effects
0 No information was found in the available literature on the effects
of fluometuron on reproduction.
0 A reproduction study with technical fluometuron in rats is in progress
to satisfy U.S. EPA Office of Pesticide Programs (OPP) data requirements
Developmental Effects
0 Fritz (1971) reported a teratology study in rats in which dams were
given C-2059 suspension in carboxymethylcellulose during days 6
through 15 of gestation. Offspring were removed on day 20 of ges-
tation for examination. The NOAEL was indicated as 100 mg/kg/day,
and higher doses reduced fetal body weight. However, this study was
invalidated by the U.S. EPA OPP because of inadequate reporting.
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Fluometuron August, 1987
-7-
0 A teratology study in which pregnant Spf New Zealand rabbits were
given technical fluometuron (purity not specified) by gavage at dose
levels of 50, 500, and 1,000 mg/kg/day during gestation days 6 through
19 was reported by Arhur and Triana (1984). Does were examined for
body weight, food consumption and pathological and developmental
effects, and laparohysterectomy was done on gestation day 29 for
pathological evaluation of fetuses. Increased liver weights and
increased mean number of resorptions were found with all doses
(p <0.05 at the low and mid doses; insufficient number of fetuses for
statistical analysis at the high dose). A LOAEL of 50 mg/kg/day was
identified. Reductions in body weights and food consumption occurred
in does given 500 and 1,000 mg/kg/day. Deaths, abortions and perforated
stomachs were observed in does given 1,000 mg/kg/day.
Mutagenicity
0 In bacterial assays (Dunkel and Simmon, 1980), fluometuron (6.6 mg/plate)
was not mutagenic in Salmonella strains TA 1535, TA 1537, TA 1538,
TA 98 and TA 100, either with or without metabolic activation.
0 Seiler (1978) reported that fluometuron (2,000 rag/kg bw) given as a
single oral dose of an aqueous suspension by gavage resulted in a
strong inhibition of mouse testicular DNA synthesis in mice killed
3.5 hours after treatment. Results were inconclusive in a subsequent
micronucleus test.
0 In yeast assays (Seibert and Lemperle, 1974), a commercial formulation
of fluometuron was ineffective in inducing mitotic gene conversion
in Saccharomyces cerevisiae strain D4 without exogenous metabolic
activation.
Carcinogenicity
0 In a long-term bioassay (NCI, 1980), fluometuron was administered in
feed to F344 rats and B6C3F! mice. Groups of rats (50/sex/dose) were
fed diets containing 125 or 250 ppm fluometuron for 103 weeks. Mice
(50/sex/dose) were fed 500 or 1,000 ppm for an equivalent period
of time. Assuming that 1 ppm equals 0.05 mg/kg/day in the older rat
and 0.15 mg/kg/day in the mouse (Lehman, 1959), 125 and 250 ppm
equaled 6.25 and 12.5 mg/kg/day in rats and 500 and 1,000 ppm equaled
75 and 150 mg/kg/day in mice. Results based on survival, body weights,
and nonneoplastic pathology (including spleen) were negative in rats.
Following treatment, there were no significant increases in tumor
incidences in male or female F344 rats or in female B6C3F]. mice com-
pared to controls. In male B6C3F! mice, an increased incidence
of hepatocellular carcinomas and adenomas was noted. The incidences
were dose-related and were marginally higher than those in the corre-
sponding matched controls or pooled controls from concurrent studies
[matched control, 4/21 or 19%; low dose, 13/47 or 28%; high dose,
21/49 or 43% (p = 0.049); pooled controls, 44/167 or 26%]. NCI (1980)
concluded that additional testing was needed because of equivocal
findings for male mice and because both rats and mice may have been
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Fluometuron August, 1987
-8-
able to tolerate higher doses, the NOAELs identified for rats and
nice are 12.5 and 75 mg/kg/day, respectively.
0 Chronic feeding studies with technical fluometuron in rats and mice
are ongoing to satisfy OPP data requirements.
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 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/OCW 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 fluometuron. The teratology
study by Arhur and Triana (1984) was not selected because a NOAEL was not
identified. It is therefore recommended that the Longer-term HA value for a
10-kg child (1.5 mg/L, calculated below) be used at this time as a conservative
estimate of the One-day HA value.
Ten-day Health Advisory
No information was found in the available literature that was suitable
for determination of the Ten-day HA value for fluometuron. The teratology
study by Arhur and Triana (1984) was not selected because a NOAEL was not
identified. It is therefore recommended that the Longer-term HA value for a
10-kg child (1.5 mg/L, calculated below) be used at this time as a conservative
estimate of the Ten-day HA value.
Longer-term Health Advisory
The 90-day feeding study in dogs by Foglemann (1965b) has been selected
to serve as the basis for the Longer-term HA value for fluometuron. In this
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Fluometuron August, 1987
-9-
study, dogs given technical fluometuron at dose levels of 0, 1.5, 15 or 150
nig/kg/day in the diet for 90 days showed pathological effects in spleen,
liver and kidney at the highest dose and no observable effects at the lower
doses. The 90-day feeding studies with rats by Foglemann (1965a) and NCI
(1980) were not selected because the 15 mg/kg/day NOAEL in the Foglemann
(1965b) study was below the lowest doses of 75 mg/kg/day in the Foglemann
(1965a) and 25 mg/kg/day (estimated) in the NCI (1980) repeat 90-day study
where effects were noted. Additionally, pathological changes in spleen found
with the lowest dose (250 ppm) in the repeat NCI (1980) study in rats were
not found with this dose in the initial 90-day study and in the 2-year bioassay
in rats by the NCI (1980). Because 7.5 mg/kg/day in the Foglemann (1965a)
study and 12.5 mg/kg/day (estimated) in the NCI (1980) carcinogenicity bioassay
were NOAELs, it is concluded that 15 mg/kg/day would be consistent with a
NQAEL in these 90-day studies in rats. The study by Hofmann (1966) in which .
rats were given technical fluometuron as a suspension by gavage at dose
levels of 0, 3, 10, 30 and 100 mg/kg, six times per week for 1 year, was not
selected because feeding the substance in the diet is preferred over giving
it as a suspension by gavage for estimating exposure from drinking water,
although the 10 mg/kg NOAEL in this study approximates the 15 mg/kg/day NOAEL
in the Foglemann (1965b) study. The 90-day feeding study in mice by NCI
(1980) was not selected because the NOAEL of 300 mg/kg/day (estimated) is
above the effect levels in the other studies considered. The 15 mg/kg/day
dose level in dogs was, therefore, identified as the NOAEL.
Using a NOAEL of 15 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (15 mg/kg/day) (10 kg) = U5 mg/L (1,500 ug/L)
(100) (1 L/day)
where:
15 mg/kg/day = NOAEL, based on absence of pathological changes in the
spleen, liver and kidneys of dogs exposed to the test
substance in the diet for 90 days.
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 a 70-kg adult is calculated as follows:
where:
Longer-term HA = (TS "SAg/day) (7° *9 > = 5.3 mg/L (5,300 ug/L)
(100) (2 L/day)
15 mg/kg/day = NOAEL, based on absence of pathological changes in the
spleen, liver and kidneys of dogs exposed to the test
substance in the diet for 90 days.
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Fluometuron August, 1987
-10-
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.
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 NCI (1980) carcinogenicity bioassay in F344 rats has been selected
to serve as the basis for determination of the Lifetime HA value for fluo-
meturon. Rats were exposed to dose levels of 0, 125 and 250 ppm fluometuron
in the diet (estimated as 6.25 and 12.5 mg/kg/day) for 103 weeks. No observable
effects were evident in this study. Although pathological changes in spleens
of rats given 250 ppm fluometuron in the diet (estimated as 25 mg/kg/day)
were noted in the repeat 90-day study in rats by NCI (1980), it appears that
splenic lesions were either not evident or were able to reverse in the rats
given the 250-ppm dietary level for 2 years (only one rat died by 1 year into
the bioassay). Furthermore, pathological changes in the spleen were not
evident with doses below 2,000 ppm in the initial 90-day study in F344 rats
by NCI (1980). The 90-day and 1-year studies discussed under Longer-term
Health Advisory have not been selected for calculation of a Lifetime HA
because of their short duration compared to the 103-week NCI (1980) bioassay
and because, although not as many end points were assessed in the NCI (1980)
bioassay compared to these studies, major effects observed in these studies
(pathology, body weight) were evaluated in the NCI (1980) bioassay. The NCI
(1980) bioassay in B6C3FT mice was not considered because higher dose levels
(500 and 1,000 ppm, estimated as 75 and 150 mg/kg/day) were used.
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Fluometuron August, 1987
-1 1-
Using the NCI (1980) bioassay in rats with a NOAEL of 12.5 mg/kg/day,
the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD m (12.5 mg/kg/day) = Q.0125 mg/kg/day
(100) (10)
where:
12.5 mg/kg/day = NOAEL, based on absence of observable effects in rats
exposed to fluometuron in the diet for 103 weeks.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
10 = additional uncertainty factor used by U.S. EPA OPP
to account for data gaps (chronic feeding studies in
rats and dogs, reproduction study in rats, teratology
studies in rats and rabbits).
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
nWEL - tO-0125 mq/kg/day) (70 kg) = Q.438 mg/L (438 ug/L)
(2 L/day)
where:
0.0125 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 = (0.438 mg/L) (20%) = 0.09 mg/L (90 ug/L)
where:
4.38 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 NCI (1980) determined that fluometuron was not carcinogenic in male
and female F344 rats and female mice (B6C3F!). The marginal increase
in the incidence of hepatocellular carcinomas and adenomas in male
B6C3F! mice was concluded to be equivocal evidence in the NCI (1980)
report on its bioassay.
0 IARC (1983) has classified fluometuron in Group 3: This chemical
cannot be classified as to its carcinogenic!ty for humans.
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Fluometuron August, 1987
-12-
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), fluometuron may be classified in
Group D: not classified. This category is used for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA/OPP previously calculated an ADI of 0.008 mg/kg/day
based on a NOAEL of 7.5 mg/kg/day in a 90-day feeding study in rats
(Foglemann, 1965a) and an uncertainty factor of 1,000 (used because
of data gaps). This has been updated to 0.013 mg/kg/day, based on a
2-year feeding study in rats using a NOAEL of 12.5 mg/kg/day and an
uncertainty factor of 1,000.
0 Tolerances have been established for negligible residues of fluometuron
in or on cottonseed and sugar cane at 0.1 ppm (U.S. EPA, *1985a). A
tolerance is a derived value based on residue levels, toxicity data,
food consumption levels, hazard evaluation and scientific judgment,
and it is the legal maximum concentration of a pesticide in or on a
raw agricultural commodity or other human or animal food (Paynter
et al., undated).
VII. ANALYTICAL METHODS
0 Analysis of fluometuron is by a high-performance liquid chromatographic
(HPLC) method applicable to the determination of certain carbamate
and urea pesticides in water samples (U.S. EPA, 1985b). This method
requires a solvent extraction of approximately 1 liter of sample with
methylene chloride using a separatory funnel. The methylene chloride
extract is dried and concentrated to a volume of 10 mL or less. HPLC
is used to permit the separation of compounds, and measurement is
conducted with a UV detector. The method detection limit for
fluometuron is 11.1 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAC) adsorption
will remove fluometuron from water.
8 Whittaker (1980) experimentally determined adsorption isotherms for
fluometuron on GAC.
0 Whittaker (1980) reported the results of GAC columns operating under
bench-scale conditions. At a flow rate of 0.8 gpm/sq ft and an empty
bed contact time of 6 minutes, fluometuron breakthrough (when effluent
concentration equals 10% of influent concentration) occurred after
1,640 bed volumes (BV). When a bi-solute solution of fluometuron
diphenamide was passed over the same column, fluometuron breakthrough
occurred after 320 BV.
-------
Fluometuron August, 1987
-13-
0 GAC adsorption appears to be the most promising treatment technique
for the removal of fluometuron from contaminated water. However,
selection of individual or combinations of technologies to attempt
fluometuron removal from water must be based on a case-by-case
technical evaluation, and an assessment of the economics involved.
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Fluometuron August. 1987
-14-
IX. REFERENCES
Arhur, A., and V. Triana.* 1984. Teratology study (with fluometuron) in
rabbits. Ciba-Geigy Corporation. Report No. 217-84. Unpublished
study. MRZO 842096.
Boyd, V.F., and R.W. Foglemann.* 1967. Metabolism of fluometuron (1,1-dimethyl-
3-(alpha, alpha, alpha-trifluoro-m-tolyl) urea) in the rat. Ciba
Agrochemical Company. Research Report CF-1575. Unpublished study.
MRIO 00022938.
CHEMLAB. 1985. The chemical information system. CIS, Inc., Bethesda, MD.
Davidson, J., and J. McDougal. 1973. Experimental and predicted movement
of three herbicides in a water-saturated soil. J. Environ. Qual.
2(4):428-433.
Dunkel, V.C., and V.F. Simmon. 1980. Mutagenic activity of chemicals
previously tested for carcinogenic!ty in the National Cancer Institute
bioassay. PROGRAM. IARC. Sci. Publ. 27:283-302.
Foglemann, R.W.* 1964a. Compound C-2059 technical — acute oral toxicity —
male and female mice. AME Associates for CIBA Corporation. Project No.
20-042. Research Report CF-735. Unpublished study. MRID 00019032.
Foglemann, R.N.* 1964b. Compound C-2059 80 WP-repeated rabbit dermal toxicity.
AME Associates for CIBA Corporation. Project No. 20-0242. Research
Project CF-740. Unpublished study. MRID 00018593.
Foglemann, R.W.* 1964c. Compound C-2059 Technical — Acute eye irritation —
Rabbits. AME Associates for CIBA Corporation. Project No. 20-042.
Unpublished study. MRID 0019032. MRID 00018593.
Foglemann, R.W.* 1965a. Cotoran — 90-day feeding rats. AME Associates for
CIBA Corporation. Project No. 20-042. Unpublished study. MRID 00019034.
Foglemann, R.W.* 1965b. Subacute toxicity — 90 day administration — dogs.
AME Associates for CIBA Corporation. Project No. 20-042. Unpublished
study. MRID 00019035.
Fritz, H.* 1971. Reproduction study: Segment II. Preparation C-2059:
Experiment No. 22710100. CIBA-GEIGY, Ltd. Unpublished study. MRID
000019211.
Galloway, Do* 1984. Guinea pig skin sensitization. Project No. 3397-84.
Unpublished study. Stillmeadow, Inc. for CIBA-GEIGY Corporation. MRID
00143601.
Guth, J.A. 1972. Adsorption and elution behavior of plant protective agents
in soils. A translation of: Adsorptions- und einwasch ver halten von
pflanzenschutzmitteln in boeden. Schriftenreihe des vereins fuer wasser,
boeden, and lufthygiene, Berlin-Dahlem (37):143-154.
-------
Fluometuron August, 1987
-15-
Guth, J. A., H. Geissbuehler and L. Ebner. 1969. Dissipation of urea
herbicides in soil. Meded. Rijksfac. Landbouwwet. XXXIV(3):1027-1037.
Helling, C.S. 1971. Pesticide mobility in soils: II. Applications of soil
thin-layer chromatography. Soil Sci. Soc. Am. Proc. 35:737-738.
Helling, C.S., D.D. Kaufman and C.T. Dieter. 1971. Algae bioassay detection
of pesticides mobility in soils. Weed Sci. 19(6):685-690.
Hofmann, A.* 1966. Examinations on rats of the chronic toxicity of preparation
Bo-27 690 (Cotoran = C-2059). Hofmann-Battelle-Geneva. (Translation;
Unpublished study). MRID 00019088.
IARC. 1983. International Agency for Research on Cancer. Vol. 30. . IARC
monographs on the evaluation of carcinogenic risk of chemicals to man.
Lyon: IARC.
LaFleur, K. 1979. Sorption of pesticides by model soils and agronomic
soils: Rates and equilibria. Soil Sci. 127(2):94-101.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs,
and cosmetics. Association of Food and Drug Officials of the
United States.
Lin, T.H., R.E. Menzer and H.H. North. 1976. Metabolism in human embryonic
lung cell cultures of three phenylurea herbicides; chlorotoluron,
fluometuron and metobromuron. J. Agric. Food Chem. 24:759-763.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
NCI. 1980. National Cancer Institute. Bioassay of fluometuron for possible
carcinogenicity. NCI-CG-TR-195. Bethesda, MD.
NIOSH. 1985. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances. National Library of Medicine
Online File.
Paynter, O.E., J.G. Cummings and M.H. Rogoff. Undated. United States
Pesticide Tolerance System. U.S. EPA Office of Pesticide Programs.
Washington, DC. Unpublished draft report.
Sachsse, K., and R. Bathe.* 1975. Acute oral LD50 of technical fluometuron
(C-2059) in the rat. Project No. Siss. 4574. Unpublished study. MRID
00019213.
Seiler, J.P. 1978. Herbicidal phenylalkylurea as possible mutagens. I.
Mutagenicity tests with some urea herbicides. Mutat. Res. 58:353-359.
Siebert, D., and E. Lemperle. 1974. Genetic effects of herbicides: Induction
of mi totic gene conversion in Saccharomyces cerevisiae. Mutat. Res.
22:111-120.
-------
Fluometuron August, 1987
-16-
Siglin, J.C., P.J. Becci and R.A. Parent.* 1981. Primary skin irritation in
rabbits (EPA - FIFRA): FDRL: Study No. 681 7A. Food and Drug Research
Laboratories for Ciba-Geigy. Unpublished study. MRZD 00068040.
STORET. 1987.
TOB. 1985. Toxicology Data Bank. Medlars II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.229. July 1, p. 293.
U.S. EPA. 198Sb. U.S. Environmental Protection Agency. U.S. EPA Method 632
- Carbamate and urea pesticides, Fed. Reg. 50:40701. October 4.
U.S. EPA. 1986. U.S* Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck Index — an encyclopedia of chemicals and drugs, 10th ed.
Rahway, NJ: Merck and Company, Inc.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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FONOFOS
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
August, 1987
DRAFT
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 thes^ 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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Fonofos August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 944-22-9
Structural Formula
0-Ethyl-S-phenylethylphosphonodithioate
Synonyms
0 Difonate; Difonatal; Dyfonate; Dyfonate*; Dyphonate®; ENT 25, 796;
Fonophos; Stauffer N2790 (Meister, 1983).
Uses
• Soil insecticide (Meister, 1983).
Properties (Windholz et al., 1983; TDB, 1985)
Chemical Formula C10H15OS2P
Molecular Weight 246.32
Physical State (25°C) Light yellow liquid
Boiling Point 1 30°C
Melting Point
Vapor Pressure (25°C) 2.1 x 10~4 mm Hg
Specific Gravity (20eC) 1.154
Water Solubility (25°C) Practically insoluble
Octanol/Water Partition
Coefficient
Taste Threshold --
Odor Threshold
Conversion Factor --
Occurrence
0 Fonofos has been detected in ground waters in California at 0.01 to
0.03 ppb (U.S.G.S. Regional Assessment Project, C. Eiden, 1985).
0 Fonofos has been found in tailwater pit sediment and water samples.
Monitoring studies conducted in 1973 and 1974 in Haskell County,
Kansas, showed that the highest concentrations found were 770 ppb
for sediment and 5.9 ppb for water during 1974. Mean peak concen-
trations were highest in June and July (Xadoum and Mock, 1978).
0 Fonofos (Dyfonate) has been found in Iowa ground water; a typical
positive sample found was 0.1 ppb (Cohen et al., 1986).
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Fonofos August, 1987
-3-
Environmental Fate
0 Under aerobic conditions, fonofos at 10 ppm was degraded at a moderate
rate with a half-life ranging from 3 to more than 16 weeks in soils
varying in texture from loamy sand to clay loam to peat (MeBain and
Menn, 1966; Hoffman et al., 1973; Hoffman and Ross, 1971; Miles
et al., 1979). The major degradate identified was 0-ethylethane
phosphonothioic acid; other degradates identified included fonofos
oxon, O-ethylethane phosphonic acid, O-ethyl O-methylethyl phosphonate,
diphenyl disulfide, methylphenyl sulfoxide, and methylphenyl sulfone
(Hoffman et al., 1973; Hoffman and Ross, 1971). The soil fungus,
Rhizopus japonicus rapidly degraded Hc-fonofos to 7ield dyfoxon,
thiophenol, ethylethoxy phosphonic acid and methylphenyl sulfoxide
(Lichtenstein et al., 1977).
9 Fonofos is relatively immobile in a silt loam and sandy loam soil but
relatively mobile in quartz sand. After 7 to 12 inches of water were
added to 7-inch soil columns, 2 to 9% of the applied Hc-fonofos
leached from the treated soil layer in Piano silt loam and Fox fine
sandy loam columns. When a quartz sand was leached with 7 inches of
water, 50% of the applied radioactivity was detected in the leachate.
Dyfoxon, a fonofos degradate, and two unidentified compounds were
found in the leachate of the silt loam soil (Lichtenstein et al.,
1972).
0 Fonofos is relatively mobile in runoff water from loam sand. After
30 days, only 0.54 to 1.2% of the applied !4C-fonofos was recovered
in runoff water from drenching a Sorrento loam soil on an inclined
plane at a 15-degree slope. Fonofos accounted for most of the
recovered radioactivity, which was primarily adsorbed to the silt
fraction (Hoffman et al., 1973).
0 Fonofos is not volatile from soil but is fairly volatile from water.
Within 24 hours after application, 15 to 16% of the 14C-fOnofos applied
volatilized from soil water (a suspension of fine sand in tapwater or
tapwater alone; 1% volatilized from a silt loam soil alone).
14C-Fonofos volatilized from soil water with a half-life of 5 days;
80% of the applied radioactivity was volatilized at the end of 10
days (Lichtenstein and Schulz, 1970).
0 In the field, fonofos dissipated with a half-life of 28 to 40 days
when either a 10% G or a 4 ]b/gal EC formulation was applied at 4.8
to 10 Ib ai/A to a sandy loam and two silt loam soils (Kiigemagi and
Terriere, 1971; Schulz and Lichtenstein, 1971; Talekar et al., 1977).
Using a root maggot bioassay, toxic fonofos residues in a sandy loam
field soil were detected up to 17 weeks after the 10% G formulation
was applied at 2 to 5 Ib ai/A. Residues were detected up to 28 weeks
after treatment when the same soil was maintained in a greenhouse
(Ahmed and Morrison, 1972).
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Fonofos August, 1987
-4-
III. PHARMACOKINETICS
Absorption
0 McBain et al. (1971) administered 14c-phenyl-labeled fonofos (99%
purity, dissolved in corn oil) orally to albino rats (two/dose) at
doses of 2, 4 or 8 mgAg. Only 7% of the label was recovered in
feces, indicating that absorption was nearly complete (about 93%).
Hoffman, et al. (1971) reported essentially identical results in rats
dosed with 0.8 mg/kg fonofos. Measurements of urinary, fecal and
biliary excretion indicated that about 80 to 90% of the dose was
absorbed from the gastrointestinal tract.
0 Hoffman et al. (1971) administered single oral doses of 35s-labeled
fonofos (2.0 mg/kg; 99% purity) to rats. About 32% of the label was
excreted in feces. Measurement of biliary excretion indicated that
15% of the label in the feces came from the bile. The authors
concluded that about 17% had not been absorbed.
Distribution
0 Hoffman et al. (1971) administered 35S-labeled fonofos (2.0 mg/kg,
13.4 mCi/mmol; 99% purity) to rats by gavage (in safflower oil);
the levels of label in blood and tissues were measured for 16 days.
Higher levels of radioactivity were found in the kidneys, blood,
liver and intestines, and lower levels were found in bone, brain,
fat, gonads and muscle. Concentration values at 2 days ranged from
about 400 ppb in the kidneys to about 70 ppb in other tissues. All
values were 10 ppb or lower by day 8. Tissue levels declined in
first-order fashion, with near total (99.3%) elimination during 2 to
16 days after dosing.
Metabolism
0 McBain et al. (1971) administered single oral doses of 2, 4 or 8 mg/kg
of ethyl or phenyl-1 ^-labeled fonofos (97.5% or 99% purity) to male
albino rats (two/dose). Only 2.6 to 7.1% was recovered as unchanged
fonofos in the urine. The remainder was converted to a variety of
terminal metabolites, including: 0-ethylethane phosphonothioic acid,
O-ethylethane phosphonic acid, and 0-con]ugates of 3- and 4-(hydrox-
phenyDmethyl sulfone. McBain et al. (1971) reported that fonofos
was converted by rat liver microsomes in vitro to the more toxic
fonofos oxon, but only traces of this compound were excreted by the
intact animal.
Excretion
McBain et al. (1971) administered single oral doses of 2, 4 or 8 mg/kg
of 14C-labeled fonofos (97.5% or 99% purity) orally to male albino rats
(two rats/dose). When the label was on the phenyl ring, recovery of lab*
was 90.7% in urine and 7.4% in feces. When the label was on the ethyl
group, recovery of label was 62.8% in urine and 31.8% in feces. Of
this fecal label, 15.3% was found to be excreted in the bile.
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Fonofos August, 1987
-5-
0 Hoffman et al. (1971) dosed rats orally with 1^C-ethyl-labeled fonofos
(0.8 mg/kg; 98% purity). After 15 days, average recovery of label
was 91% in urine, 7.4% in feces and 0.35% in expired air. Essentially
all of the excretion occurred within 4 days. In rats dosed with
35s-labeled fonofos (2 mg/kg; 99% purity), average recovery of label
after 4 days was 62.5% in the urine, 31.8% in feces and 0.1% in
expired air. Bile duct cannulation studies indicated that about 15%
of the label in feces arose from biliary excretion.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 The Pesticide Incident Monitoring System (PIMS) database reported 21
cases between 1966 and 1979 of human toxicity resulting from exposure
to fonofos. Fourteen of the cases involved fonofos only, and seven
involved mixtures. Two fatalities occurred, and four individuals
required medical treatment. No quantitative exposure data and no
description of adverse effects were provided (U.S. EPA, 1979).
0 One reported case of accidental ingestion involved a woman who ate
pancakes prepared with a formulation containing fonofos. No quanti-
tative estimate of the dose level was provided. The individual
developed nausea, vomiting, salivation, sweating and suffered
cardio-respiratory arrest. She was treated at a hospital and was
found to have muscle fasciculation, blood pressure of 64/0 mm Hg, a
pulse rate of 46, pinpoint pupils, and profuse salivary and bronchial
secretions. The patient also developed a pancreatic pseudocyst. The
woman was discharged after 2 months of treatment. A second individual
who also ate the contaminated pancakes died (Hayes, 1982).
Long-term Exposure
0 No information on the long-term exposure effects of fonofos on humans
was found in the available literature.
Animals
Short-term Exposure
0 Fonofos is an organophosphorus compound. Acute toxic effects of
such compounds are due largely, if not entirely, to inhibition of
cholinesterase (ChE) and acetylcholine accumulation in the body
(Derache, 1977).
0 Reported values for the oral LD50 of fonofos for female rats range
from 3.2 to 7.9 mg/kg, and values for male rats range from 6.8 to
18.5 mg/kg (Horton, 1966a,b; Dean, 1977).
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Fonofos August, 1987
-6-
0 Morton (1966a) administered single oral doses of fonofos (purity not
specified) to rats (strain not specified). Doses of 1.0 or 2.15 mg/kg
did not produce visible symptoms. Doses of 4.6 to 46 mg/kg elicited
rapid appearance of fasciculations and tremors, salivation, exophthalmia
and labored respiration, with females being somewhat more sensitive
than males. Gross autopsy of animals that died revealed congested
liver, kidneys and adrenals and lung erythema. Autopsy of survivors
showed no effects. Based on gross changes, a No-Observed-Adverse-Effect-
Level (NOAEL) of 2.15 mg/kg was identified by this study.
0 Cockrell et al. (1966) fed fonofos in the diet to dogs at levels of
0 or 8 ppm for 5 weeks. Based on the assumption that 1 ppm in the
diet of dogs is equivalent to 0.025 mg/kg/day (Lehman, 1959), these
doses correspond to 0 or 0.2 mg/kg/day. Plasma and red blood cell
cholinesterase were measured at 2 and 4 weeks; organ weights, brain
cholinesterase and changes in gross pathology were measured at termination
(5 weeks). Following treatment, no systemic toxicity was observed;
brain and plasma or red blood cell cholinesterase levels were
unaffected. No other details were provided. This study identified
a NOAEL of 8 ppm (0.2 mg/kg/day).
0 In a demyelination study, groups of 10 adult hens each received
fonofos in the diet for 46 days (Woodard and Woodard, 1966). Levels
fed were equivalent to 0, 2, 6.32 or 20 mg/kg/day. Only hens at
20 mg/kg showed impairment of locomotion and equilibrium, and one
showed histological evidence of possible demyelination of the
peripheral nerves. A NOAEL for demyelination of 6.32 mg/kg/day was
indicated by the study.
Dermal/Ocular Effects
0 Reported dermal LDso values of fonofos for the rabbit (both sexes)
ranged from 121 to 147 mg/kg (Morton, 1966a,b). However, Dean (1977)
determined a different LD50 in rabbits: 25 mg/kg for females and
TOO mg/kg for males.
0 Instillation of 0.1 mL undiluted fonofos (about 23 mg/kg/day)
in one eye of each of three rabbits caused negligible local irritation,
but was lethal to all within 24 hours (Horton, 1966a,b; Dean, 1977).
0 Dean (1977) applied 0.5 mL undiluted fonofos to closely clipped
intact skin of rabbits; no dermal irritation was reported but all
animals died within 24 hours.
0 Horn et al. (1966) applied fonofos (10% granular) to intact or abraded
skin of New Zealand rabbits (five/sex/dose; the five animals included
both normal and abraded skin animals) 5 days per week for 3 weeks at
doses of 0, 35 or 70 mg/kg. Following treatment, dermal effects,
general appearance and behavior, hematology, organ weights, cholinesterase-
levels, gross pathology and histopathology were evaluated. No
difference was observed in any of the responses between the intact or
abraded skin animals. One normal and one abraded skin males and one
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Fonofos August, 1987
-7-
normal skin female died in the 70 mg/kg group; and one intact skin
male died in the 35 mg/kg group. No irritation of the skin was
observed at any dose tested for either intact or abraded skin. In
males, adrenal weights were increased by about 50% at 35 mg/kg, and
by 70% at 70 mg/kg (p value not given). Similar but smaller (15 to
20%) increases in adrenal weights were seen in females. No hematological
effects were observed at any dose tested. No histopathological
changes occurred except slight to moderate liver glycogen depletion
at 70 mg/kg. Reductions were observed in red blood cell, plasma and
brain cholinesterase activity for both sexes of the treated groups.
At 35 mg/kg, ChE in red blood cells was inhibited 70% (for both
sexes), while plasma ChE levels were inhibited 74% (males) and 91%
(females), and brain ChE was inhibited 66% (males) and 89% (females).
At 70 mg/kg, ChE in red blood cells was inhibited 36% (males) and 45%
(females). ChE in plasma was inhibited 67% inhibited for both sexes.
ChE in brain was inhibited 59% (males) and 57% (females).
Long-term Exposure
0 Daily oral doses of fonofos in corn oil (at 0, 2, 4 or 8 mg/kg/day)
for 90 days failed to elicit delayed neurotoxicity in adult hens
(Miller et al., 1979, abstract only). A minimum NQAEL of 8 mg/kg/day
for delayed neurotoxicity was indicated by these reported results.
0 In a similar experiment (Cockrell et al., 1966), rats were fed diets
containing 0, 10, 31.6 or 100 ppm for 13 weeks. Based on the
assumption that 1 ppm in the diet is equivalent to 0.05 mg/kg/day,
these doses correspond to 0, 0.5, 1.58 or 5 mg/kg/day (Lehman, 1959).
Cholinesterase was measured in serum and red blood cells before and
after exposure, and brain ChE was measured at termination. At 100 ppm,
there was significant inhibition of ChE in serum (70%, females only),
red blood cells (85%, females only) and brain (51% to 60%, both
sexes). Decreases of over 50% in red blood cell ChE in both males
and females were reported at the 31.6-ppm level. At 10 ppm, the
largest difference detected was a 23% decrease in red blood cell ChE
in females; the authors did not consider this to be significant.
All other ChE measurements at this dose were comparable between
exposed and control animals. Other observations were negative for
compound effect, and there were no histopathological findings. Based
on ChE inhibition, the NOAEL in rats was identified as 10 ppm
(0.5 mg/kg/day).
0 Cockrell et al. (1966) fed fonofos in the diet to dogs at levels
of 0, 16, 60 or 240 ppm for 14 weeks. Based on the assumption that
1 ppm in the diet is equivalent to 0.025 mg/kg/day, these doses
correspond to 0, 0.4, 1.5 or 6 mg/kg/day (Lehman, 1959). Dogs showed
increased lacrimation and salivation plus convulsions (at 16 ppm),
bloody diarrhea (at 60 ppm) or tremors and anxiety and increased
liver weight (at 240 ppm). At 16 ppm, there was about 60% ChE inhibi-
tion in erythrocytes and slight ChE inhibition in brain (female only).
At 60 ppm, ChE in red blood cells was inhibited 60% or more, and
plasma ChE was decreased about 20% (in males only) at week 13. At
the high dose (240 ppm), ChE was nearly totally inhibited in red
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Fonofos August, 1987
-8-
blood cells; about 50% inhibited in plasma; and moderately inhibited
in brain. Based on cholinesterase inhibition and systemic toxicity,
a Lowest-Observed-Adverse-Effect-Level (LOAEL) of 16 ppm (0.4 mg/lcg/day),
the lowest dose tested, was identified.
0 Pure-bred beagle dogs were fed fonofos in the diet for 2 years
(Hoodard et al., 1969). Groups of four males and four females each
received 0, 16, 60 or 240 ppm fonofos. Based on the assumption that
1 ppm in the diet is equivalent to 0.025 mg/kg/day, these doses
correspond to 0, 0.4, 1.5 or 6 mg/kg/day (Lehman, 1959). After 14
weeks, the low dose (16 ppm) was reduced to 8 ppm (0.2 mg/kg/day),
and this dose level was maintained for the duration of the study.
Cholinesterase levels in plasma were inhibited about 50% at 240 ppm,
about 25% to 50% at 60 ppm, and were not different from controls at
the low dose (16 or 8 ppm). In red blood cells, ChE levels were
inhibited almost completely at the 240-ppm level and about 65% at
60 ppm. In animals receiving 16 ppm for 14 weeks, ChE in red blood
cells was inhibited about 30%. After reduction of the dose to 8 ppm,
ChE levels returned to values comparable to controls. At sacrifice,
no inhibition of ChE in brain was detected at any dose level. At
240 ppm, nervous, apprehensive behavior and tremors were seen, and
three dogs died, each with marked acute congestion of tissues and
hemorrhage of the small intestinal mucosa. At this dose level, also,
serum alkaline phosphatase was increased, as were liver weights.
Histopathological examination of animals receiving 240 ppm revealed a
marked increase in basophilic granulation of the myofibril of the
inner layer of the muscularis of the small intestine, and there were
slight changes in the liver. At 60 ppm, increased liver weight was
observed. At the low dose (16/8 ppm), the only effect was a single
brief episode of fasciculation in one male dog at 5 months. The
author judged that this could not be ascribed with certainty to
fonofos exposure. For this study, the NOAEL for ChE inhibition and
for systemic toxicity was 8 ppm (0.2 mg/kg/day).
0 Albino rats received fonofos in the diet for 2 years at 0, 10, 31.6
or 100 ppm (0, 0.5, 1.58 or 5 mg/kg/day, Lehman, 1959) (Bannerjee
et al., 1968). Fonofos was judged not to have affected survival,
food intake, body weight gain, organ weights or gross and histopatho-
logical findings. At 100 ppm, females showed tremors and nervous
behavior, and males had reduced hemoglobin and packed-cell volume.
At 100 ppm, ChE was markedly decreased in plasma (50 to 75%), red
blood cells (close to 100%) and brain (about 40%, in females only).
At 31.6 ppm, there was moderate (about 50%) inhibition of ChE in red
blood cells and plasma (at weeks 26 and 52 only). At 10 ppm, no
decrease in ChE was seen in brain or red blood cells, and no effect
was seen in plasma, except for a moderate decrease (40 to 56%) in
males at weeks 19 and 26 only. Based on cholinesterase inhibition,
a NOAEL of 10 ppm (0.5 mg/kg/day) is identified.
Reproductive Effects
0 Woodard et al. (1968) exposed three generations of rats to dietary
fonofos at 0, 10 or 31.6 ppm. Based on the assumption that 1 ppm in
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Fonofos August, 1987
-9-
the diet is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corre-
sponds to doses of 0, 0.5 or 1.58 mg/kg/day. No differences were
detected in exposed dams with respect to mortality, body weight or
uterine implantation sites. No effects were seen in offspring on
conception ratio, litter size, number of live-born and still-born,
litter weight and weanling survival. Skeletal and visceral examina-
tion of offspring revealed no evidence of developmental defects.
A minimum NOAEL of 31.6 ppm (1.58 mg/kg/day, the highest dose tested)
was identified.
Developmental Effects
0 'Groups of pregnant mice each received 10 daily doses of fonofos by
gavage (0, 2, 4, 6 or 8 mg/kg/day) on gestational days 6 through 15
(Minor et al., 1982). At 8 mg/kg/day, maternal food intake and body
weight gain were decreased. At 6 mg/kg/day, two dams experienced
tremors and died. Increased incidences of variant ossifications of
the sternebrae (8 mg/kg/day) and a slight dilatation of the fourth
ventricle of the brain (4 and 8 mg/kg/day) were observed, but the
authors did not interpret this as evidence of teratogenicity. The
NOAEL for fetotoxicity identified in this study was 2 mg/kg/day.
Mutagenicity
0 Fonofos, with or without metabolic activation, was not mutagenic in
each of five microbial assay systems (the Ames (Salmonella) test;
reverse mutation in an Escherichia coli strain; mitotic recombination
in the yeast, Saccharomyces cerevisiae D3; and differential toxicity
assays in strains of _E. coli and Bacillus subtilis) and in a test for
unscheduled DNA synthesis in human fibroblast cells (Simmon, 1979).
Carcinogenicity
0 Groups of 30 male and 30 female CD albino rats (Charles River) each
received 0, 10, 31.6 or 100 ppm fonofos in the diet (0, 0.5, 1.58 or
5 mg/kg/day) for 2 years (Bannerjee et al., 1968). Based on gross
and histological examination, the authors detected no carcinogenic
effects.
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 m (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 mg/kg bw/day.
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Fonofos August, 1987
-10-
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/OCW guidelines.
L/day = assumed daily water consumption by 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 fonofos. It is therefore
recommended that the Longer-term HA value for a 10-kg child of 0.02 mg/L
(20 ug/L, calculated below) be used at this time as a conservative estimate
of the One-day HA value.
Ten-day Health Advisory
No information was found in the available literature that was suitable
for determination of the Ten-day HA value for fonofos. It is therefore
recommended that the Longer-term HA value for a 10-kg child (0.02 mg/L,
calculated below) be used at this time as a conservative estimate of the
Ten-day HA value.
Longer-term Health Advisory
The 2-year feeding study in dogs by Woodard et al. (1969) has been
selected to serve as the basis for the Longer-term HA for fonofos. In this
study, dogs received dietary fonofos at 0, 16, 60 or 240 ppm (0, 0.4, 1.5 or
6 mg/kg/day). After 14 weeks, marginal (about 30%) inhibition of ChE was
noted in red blood cells at the 16-ppm level; this dose was reduced to 8 ppm
(0.2 mg/kg/day) for the remainder of the study. Following dose reduction,
ChE levels returned to those of controls. At 60 ppm, dogs showed increased
liver weights and significant inhibition (25 to 65%) of ChE activity in
plasma and erythrocytes. At 240 ppm, there was increased ChE inhibition and
increased mortality. There were no toxic effects in dogs at 8 ppm (0.2 mg/kg/day),
with the possible exception of one brief episode of fasciculation in one dog
at 5 months. This was not judged to be significant, and a NOAEL of 8 ppm
(0.2 mg/kg/day) was identified. The 13-week feeding study in rats by Cockrell
et al. (1966) has not been selected, since the rat ?ppears to be less sensitive
than the dog. The 14-week feeding study in dogs by Cockrell et al. (1966)
has not been selected since frank toxic responses were noted at the lowest
dose tested in this study (0.4 mg/kg/day).
Using a NOAEL of 0.2 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (0-2 mg/kg/day) (10 kg) = 0.02 mg/L (20 ug/L)
(100) (1 L/day)
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Fonofos August, 1987
-11-
where:
0.2 mg/kg/day = NOAEL, based on absence of systemic toxicity or ChE
inhibition in dogs exposed to fonofos in the diet for
2 years.
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 by a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (0-2 mg/kg/day) (70 kg) = 0.070 mg/L (70 Ug/L)
(100) (2 L/day)
where:
0.2 mg/kg/day = NOAEL, based on absence of systemic toxicity or ChE
inhibition in dogs exposed to fonofos in the diet for
1 month.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/OEW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption by 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. 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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Fonofos August, 1987
-12-
The 2-year feeding study in dogs by Woodard et al. (1969) has been
selected to serve as the basis for the Lifetime HA for fonofos. Dogs received
dietary fonofos at 0, 16, 60 or 240 ppm (0, 0.4, 1.5 or 6 mg/kg/day) for 14
weeks. Marginal (about 30%) inhibition of ChE was noted in red blood cells
at the 16-ppm level; this dose was reduced to 8 ppm (0.2 mg/kg/day). Following
dose reduction, ChE levels returned to control. At 60 ppm, dogs showed
increased liver weights and significant inhibition (25 to 65%) of ChE
activity in plasma and erythrocytes. At 240 ppm, there was increased ChE
inhibition and increased mortality. There were no toxic effects in dogs at
8 ppm (0.2 mg/kg/day), with the possible exception of one brief episode of
fasciculation in one dog at 5 months. This was not judged to be significant,
and a NOAEL of 8 ppm (0.2 mg/kg/day) was identified. The 2-year feeding
study in rats by Bannerjee et al. (1968) has not been selected, since rats
appear to be less sensitive than dogs when doses are calculated on a body
weight (mg/kg) basis.
Step 1: Determination of the Reference Dose (RfD)
RfD = (0«2 mg/kg/day) = 0.002 mg/kg/day
(100)
where:
0.2 mg/kg/day = NOAEL, based on absence of systemic toxicity or ChE
inhibition in dogs exposed to fonofos in the diet
for 2 years.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guideline for use with a NOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - (0-002 mg/kg/day) (70 kg) = 0>07 mg/L (70 ug/L)
(2 L/day)
where:
0.002 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption by an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.07 mg/L) (20%) = 0.014 mg/L (14 ug/L)
where:
0.07 mg/L = DWEL.
20% = assumed relative source contribution from water.
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Fonofos August, 1987
-13-
Evaluation of Carcinogenic Potential
0 Groups of 30 male and 30 female albino rats (Charles River, Cesarean-
derived) each received 0, 10, 31.6 or 100 ppm fonofos in the diet
(0, 0.5, 1.58 or 5 nig/kg/day} for 2 years (Bannerjee et al., 1968).
Based on gross and histological examination, the authors detected no
carcinogenic effect.
0 IARC (1982) has not evaluated the carcinogenic potential of fonofos.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), fonofos may be classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No existing criteria, guidelines or standards for oral exposure to
fonofos were located.
0 The U.S. EPA Office of Pesticide Programs (OPP) has calculated an
ADI of 0.002 mg/kg/day for fonofos. This was based on a NOAEL of
0.2 mg/kg/day (8 ppm) for both ChE inhibition and systemic effects,
in a 2-year feeding study in dogs (Woodard et al., 1969), and an
uncertainty factor of 100.
0 The Threshold Limit Value (TLV) for fonofos is 100 ug/m3 (ACGIH,
1984).
0 The U.S. EPA (1985) has established tolerances for fonofos in or on
raw agricultural commodities that range from 0.1 to 0.5 ppm.
VII. ANALYTICAL METHODS
0 Analysis of fonofos is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus-containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen phosphorus
detector. The method detection limit has not been determined for
fonofos, but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information on treatment technologies used to remove fonofos from
contaminated water was found in the available literature.
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Fonofos August, 1987
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IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH.
Ahmed, J. and F.O. Morrison. 1972. Longevity of residues of four organo-
phosphate insecticides in soil. Phytoprotection. 53(2-3): 71-74.
Bannerjee, B.M., D. Howard and M.W. Woodard.* 1968. Dyfonate (N-2790) safety
evaluation by dietary administration to rats for 105 weeks. Woodard
Research Corporation. Unpublished study. MRID 00082232.
Cockrell, K.O., M.W. Woodard and G. Woodard.* 1966. N-2790 Safety evaluation
by repeated oral administration to dogs for 14 weeks and to rats for 13
weeks. Woodard Research Corporation. Unpublished study. MRID 0090818.
Cohen, S.Z., C. Eiden and M.N. Lorber. 1986. Monitoring ground water for
pesticides in the U.S.A. American Chemical Society Symposium Series
titled: Evaluation of Pesticides in Ground Water (in press).
Dean, W.P.* 1977. Acute oral and dermal toxicity (LD5Q) i° male and female
albino rats. Study No. 153-047. International Research and Development
Corporation. Unpublished study. MRIDS 00059860, 00059856 and 00059857.
Derache, R. 1977. Organophosphorus pesticides. Criteria (dose/response
effect relationships) for organophosphorus pesticides). Published for
the Commission of the European Communities. Oxford, England: Pergamon
Press.
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
wilkins.
Hoffman, L.J., J.M. Ford and J.J. Menn. 1971. Dyfonate metabolism studies.
I. Absorption, distribution, and excretion of O-ethyl S-phenyl ethyl-
phosphonodithioate in rats. Pesticide Biochemistry and Physiology.
1:349-355.
Hoffman, L.J., J.B. McBain and J.J. Menn. 1973. Environmental behavior
of O-ethyl S-phenyl ethylphosphonodithioate (Dyfonate): ARC-B-35.
Unpublished study submitted by Stauffer Chemical Company, P.ichmond, CA.
Hoffman, L.J. and J.H. Ross. 1971. Dyfonate soil metabolism: Project
038022. Unpublished study submitted by Stauffer Chemical Company,
Richmond, CA.
Horn, H.J., G. Woodard and M.T. Cronin.* 1966. N-2790 10% granular:
Subacute dermal toxicity: 21-day experiment in rabbits. Unpublished
study. MRID 00092438.
Horton, R.J.* 1966a. N-2790: Acute oral LD50 - rats; acute dermal toxicity -
rabbits; acute eye irritation - rabbits. Technical Report T-986. Stauffer
Chemical Company. Unpublished study. MRID 00090806.
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Fonofos August, 1987
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Horton, R.J.* 1966b. N-2790: Acute oral LD50 - rats; acute dermal toxicity -
rabbits; acute eye irritation - rabbits. Technical Report T-985. Stauffer
Chemical Company. Unpublished study. MRID 00090807.
IARC. 1982. International Agency for Research on Cancer, World Health
Organization. IARC monographs on the evaluation of the carcinogenic risk
of chemicals to humans. Chemicals, industrial processes and industries
associated with cancer in humans. International Agency for Research on
Cancer Monographs. Vols. 1 to 29, Supplement 4. Geneva: World Health
Organization.
Kadoum, A.M. and D.E. Mock. 1978. Herbicide and insecticide residues in
tailwater pits: water and pit bottom soil from irrigated corn and
sorghum fields. J. Agric. Food Chem. 26(1):45-50.
Kiigemagi, U. and L.C. Terriere. 1971. The persistence of Zinophos and
Dyfonate in soil. Bull. Environ. Contain. Toxicol. 6(4): 355-361.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S.
Lichtenstein, E.P., H. Parlar, F. Korte and A. Suss. 1977. Identification
of fonofos metabolites isolated from insecticide-treated culture media of
the soil fungus Rhizopus japonicus. J. Agric. Food Chem. 25(4):845-848.
Lichtenstein, E.P., and K.R. Schulz. 1970. Volatilization of insecticides
from various substrates. J. Agric. Food Chem. 18(5):814-818.
Lichtenstein, E.P., K.R. Schulz and T.W. Fuhremann. 1972. Movement and
fate of Dyfonate in soils under leaching and nonleaching conditions.
J. Agric. Food Chem. 20(4):831-838.
McBain, J.B. and J.J. Menn. 1966. Persistence of £-Ethyl-S-phenyl
ethylphosphonodithioate (Dyfonate) in soils: ARC-B-10. Unpublished
study submitted by Stauffer Chemical Company, Richmond, CA.
McBain, J.B., L.J. Hoffman and J.J. Menn. 1971. Dyfonate metabolism studies
II. Metabolic pathway of 0-ethyl S-phenyl ethylphosphonodithioate in
rats. Pesticide Biochem. Physiol. 1:356-365.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Miles, J.R.W., C.M. Tu and C.R. Harris. 1979. Persistence of eight
organophosphorus insecticides in sterile and non-sterile mineral and
organic soils. Bull. Environ. Contarn. Toxicol. 22:312-318.
Miller, J.L., L. Sandvik, G.L. Sprague, A.A. Bickford and T.R. Castles. 1979.
Evaluation of delayed neurotoxic potential of chronically administered
Dyfonate in adult hens. Toxic. Appl. Pharmacol. 48:A199.
Minor, J., J. Downs, G. Zwicker et al.* 1982. A teratology study in CD-I
mice with Dyfonate technical T-10192. Final report. Stauffer Chemical
Company. Unpublished study. MRID 00118423.
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Fonofos August, 1987
-16-
Schulz, K.R. and E.P. Lichtenstein. 1971. Field studies on the persistence
and movement of Dyfonate in soil. J. Econ. Entonol. 64(1) : 283-287.
Simmon, V.F. 1979. _In vitro microbiological mutagenicity and unscheduled
DNA synthesis studies of eighteen pesticides. National Technical Infor-
mation Service, Springfield, Virginia, publication EPA-600/1 -79-041 ,
Research Triangle Park, North Carolina, p. 164.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
Talekar, N.S., L.T. Sun, E.M. Lee and J.S. Chen. 1977. Persistence of some
insecticides in subtropical soil. J. Agric. Food Chem. 25(2) : 348-352.
U.S. EPA. 1979. U.S. Environmental Protection Agency, Office of Pesticide
Programs. Summary of reported incidents involving fonofoso Pesticide
Incident Monitoring Systems. Report No. 220. Washington, DC: U.S.
Environmental Protection Agency.
U.S. EPA. 1985. United States Environmental Protection Agency. Code of
Federal Regulations. 40 CFR 180.221, p. 290.
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. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, Ohio.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck index — an encyclopedia of chemicals and drugs, 10th ed. Rahway,
NJ: Merck and Company, Inc.
Woodard, M.W., J. Donoso, J.P. Gray et al.* 1969. Dyfonate (N-2790) safety
evaluation by dietary administration to dogs for 106 weeks. Woodard
Research Corporation. Unpublished study. MRID 00082223.
Woodard, M.W., C.L. Leigh and G. Woodard.* 1968. Dyfonate (N-2790) three-
generation reproduction study in rats. Woodard Research Corporation.
Unpublished study. MRID 00082234.
Woodard, M.W. and G. Woodard.* 1966. N-2790 (Dyfonate): Demyelination
study in chickens. Woodard Research Corporation. Unpublished study.
MRID 00090819.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
GLYPHOSATE
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 tl an 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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Glyphosate August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1071-83-6
Structural Formula
O O
i1
HO-C-CH2-N-CH2-P-OH
H OH
Glycine, N-(Phosphonomethyl)
Synonyms
Rodeo9; Roundup®.
Uses
0 Herbicide for control of grasses, broad leaved weeds and woody brush
(U.S. EPA, 1986b).
Properties (Meister, 1983)
Chemical Formula C3H8NO5P
Molecular Weight 169.07
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 200°C
Density 1.74
Vapor Pressure —
Water Solubility 10 g/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Glyphosate has been found in none of the surface water samples and
in only 1 of the ground water samples (in the state of California)
analyzed from 64 samples taken at 61 locations (STORET, 1987).
Environmental Fate
0 14c-Glyphosate (94% glyphosate, 5.9% aminomethylphosphonic acid) and
aminomethylphosphonic acid were stable in sterile buffered water at
pH 3, 6, and 9 during 35 days of incubation in the dark at 5 and 35°C
(Brightwell and Malik, 1978).
0 14c-Glyphosate (94% glyphosate, 5.9% aninomethylphosphonic acid) was
adsorbed to Drummer silty clay loam, Ray silt, Spinks sandy loam,
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Glyphosate August, 1987
-3-
Lintonia sandy loam, and Cattail Swamp sediment with Freundlich-K
values of 62, 90, 70, 22, and 175, respectively (Brightwell and
Malik, 1978). For each soil preparation, the maximum percentages
of applied glyphosate desorbed were 5.3, 3.7, 3.6, 11.5, and 0.9%,
respectively. At concentrations ranging from 0.21 to 50.1 ppm,
14c-Glyphosate was highly adsorbed to five soils, with organic matter
contents ranging from 2.40 to 15.50% (Monsanto Company, 1975).
Adsorption of glyphosate ranged from 71 (Soil E, 2.4% organic matter,
pH 7.29) to 99% (Soil C, 15.5% organic matter, pH 5.35).
0 14c-Glyphosate (94% glyphosate, 5.9% aminomethylphosphonic acid)
was slightly mobile to relatively immobile, with less than 7% of the
applied 14C detected in the leachate from 30-cm silt, sand, clay,
sandy clay loam, silty clay loam, and sandy loam soil columns eluted
with 20 inches of water (Brightwell and Malik, 1978). Aged (30 days)
!4C-glyphosate residues were relatively immobile in silt, clay and
sandy clay loam soils with less than 2% of the radioactivity detected
in the leachate following elution with 20 inches of water. Both
glyphosate and aminomethylphosphonic acid were detected in the leachate
of aged and un-aged soil columns.
III. PHARMACOKINETICS
Absorption
0 Feeding studies with chickens, cows and swine showed that ingestion
of up to 75 ppm glyphosate resulted in nondetectable glyphosate
residue levels (<0.05 ppm) in muscle tissue and fat (Monsanto Company,
1983). The duration of exposure was not reported in this report.
Glyphosate residue levels were not detectable «0.025 ppm) in milk
and eggs from cows and chickens on diets containing glyphosate.
Distribution
0 No information on the distribution of glyphosate was found in the
available literature.
Metabolism
0 No information on the metabolism of glyphosate was found in the
available literature.
Excretion
After a single oral or intraperitoneal dose, less than 1% of the
administered dose was retained after 120 hours of treatment (U.S. EPA,
1986b). In rats fed 1, 10 or 100 ppm of 14C-glyphosate for 14 days,
a steady-state equilibrium between intake and excretion of label was
reached within about 8 days. The amount of radioactivity excreted
in the urine decreased rapidly after withdrawal of treatment. Ten
days after withdrawal, radioactivity was detectable in the urine and
feces of rats fed 10 or 100 ppm of the test diet. Minimal residues
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Glyphosate August, 1987
-4-
of 0.1 ppm or less remained in the tissues of high-dose rats after
10 days of withdrawal. No single tissue showed a significant
difference in the amount of label retained.
IV. HEALTH EFFECTS
Humans
0 No information on the health effects of glyphosate in humans was
found in the available literature.
Animals
Short-term Exposure
0 An oral LD50 of 5,600 rag/kg in the rat is reported for glyphosate
(Monsanto Company, 1982a).
0 Bababunmi et al. (1978) reported that daily intraperitoneal admini-
stration of 15, 30, 45 or 60 rag/kg to rats for 28 days resulted in
reduced daily body weight gain, decreased blood hemoglobin, decreased
red blood cell count and hematocrit values and elevated levels of
serum glutamic-pyruvic transaminase and leucine-amino peptidase during
the experimental period. The investigators did not specify the dose
levels at which these effects were observed.
Dermal/Ocular Effects
0 A dermal LD^Q for glyphosate in the rabbit was reported to be
>5,000 mg/kg (Monsanto Company, 1982a).
Long-term Exposure
0 In subchronic studies reported by the Weed Science Society of America
(1983), technical-grade glyphosate was fed to rats at dietary levels
of 20, 60 or 200 mg/kg/day and to dogs at 50, 150 or 500 mg/kg/day
for 90 days. Mean body weights, food consumption, behavioral reactions,
mortality, hematology, blood chemistry and urinalysis did not differ
significantly from controls. There were no relevant gross or histo-
patholocical changes. No other details or data were provided.
0 Bio/dynamics, Inc. (1981 a) conducted a study in which glyphosate
was administered in the diet to four groups of Sprague-Dawley rats
(50/sex/dose) at dose levels of 0, 3.1, 10.3 or 31.5 mg/kg/day to
males or 0, 3.4, 11.3 or 34.0 mg/kg/day to females. After 26 weeks,
body weight, organ weight, organ-to-body weight ratios and hematological
and clinical chemistry parameters were evaluated. No significant
differences between control and exposed animals were observed at any
dose level.
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Glyphosate August, 1987
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Reproductive Effects
0 Bio/dynamics, Inc. (1981b) investigated the reproductive toxicity of
glyphosate in rats. The glyphosate (98.7% purity) was administered
in the diet at dose levels of 0, 3, 10 or 30 mg/kg/day to Charles
River Sprague-Dawley rats for three successive generations. Twelve
males and 24 females (the F0 generation) were administered test diets
for 60 days prior to breeding. Administration was continued through
mating, gestation and lactation for two successive litters (F1a,
Fib>- Twelve males and 24 females from the F1b generation were
retained at weaning for each dose level to serve as parental animals
for the succeeding generation. The following indices of reproductive
function were measured: fetal, pup and adult survival; parental and
pup body weight; food consumption; and mating, fertility or gestation.
Necropsy and histopathologic evaluation were performed as well.
No compound-related changes in these parameters were observed when
compared to controls, although an addendum to the pathological report
for this study reported an increase in unilateral focal tubular
dilation of the kidney in the male F3b pups when compared to concurrent
controls. Based on data from this study, the authors concluded that
the highest dose tested (30 mg/kg/day) did not affect reproduction
in rats under the conditions of the study.
Developmental Effects
0 Glyphosate was also administered to pregnant rabbits (route not
specified) at dose levels of 75, 175 or 350 mg/kg/day on days 6
through 27 of gestation (Monsanto Company, 1982a). No evidence of
fetal toxicity or birth defects in the offspring was observed.
However, at dose levels of 350 mg/kg/day, death, soft stools, diarrhea
and nasal discharge were observed in the animals.
Mutagenicity
0 The Monsanto Company (1982a) reported that glyphosate did not cause
mutation in microbial test systems. A total of eight strains (seven
bacterial and one yeast), including five Salmonella typhimurium strains
and one strain of Bacillus subtilis, Escherichia coli and Saccharomyces
cerevisiae, were tested. No mutagenic effects were observed in any
strain.
0 N]agi and Gopalan (1980) found that glyphosata did not induce reversion
mutations in Salmonella typhimurium histidine auxotrophs.
Carcinogenicity
0 Bio/dynamics, Inc. (1981b) conducted a study to assess the oncogenicity
of glyphosate (98.7% purity). The chemical was given in the diet to
four groups of Sprague-Dawley rats at dose levels of 0, 3.1, 10.3 or
31.5 mg/kg/day to males or 0, 3.4, 11.3 or 34.0 mg/kg/day to females.
After 26 weeks, animals were sacrificed and tissues were examined for
histological lesions. A variety of benign and malignant tumors were
observed in both the treated and control groups, the most common tumor
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Glyphosate August, 1987
-6-
occurring in the pituitary of both sexes and in the mammary glands of
females. The total number of rats of both sexes that developed
tumors (benign and malignant) was 72/100 (low dose), 79/100 (mid
dose), 85/100 (high dose) and 87/100 (control). An increased rate of
interstitial cell tumors of the testes was reported in the high-dose
males when compared to concurrent controls (6/50 versus 0/50), but
this was not considered to be related to compound administration.
Based on the data from this study, the authors concluded that the
highest dose level tested (31.5 and 34.0 mg/kg/day for males and
females, respectively) was not carcinogenic in 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) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in ing/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/OCW 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 glyphosate. It is, therefore,
recommended that the Ten-day HA value be used at -Jiis time as a conservative
estimate of the One-day HA value.
Ten-day Health Advisory
The teratology study in pregnant rabbits has been selected to se*-ve as
the basis for determination of the Ten-day HA for the 10-kg child. In this
study, pregnant rabbits that received glyphosate at dose levels of 0, 75,
175 or 350 mg/kg/day on days 6 through 27 of gestation showed effects at
350 mg/kg/day; however, no treatnent-related effects were reported at lower
dose levels. The No-Observed-Adverse-Effect-Level (NOAEL) identified in
this study is, therefore, 175 mg/kg/day. While a developmental end point may
not be the most appropriate basis for derivation of an HA for a 10-kg child,
use of this study provides an extra margin of safety. .
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Glyphosate August, 1987
-7-
Using a NOAEL of 175 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-dav HA = (175 mg/kg/day) (10 kg) = 17.50 mg/L (17,500 ug/L)
y (100) (1 L/day)
where:
175 mg/kg/day = NOAEL, based on absence of altered physical changes
and mortality in rabbits.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
No information was found in the available literature that was suitable
for determination of the Longer-term HA value for glyphosate. It is, therefore,
recommended that the adjusted DWEL for a 10-kg child be used at this time as
a conservative estimate of the Longer-term HA value.
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 ky 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, 1986a), then caution should be exercised
in assessing the risks associated with lifetime exposure to this chemical.
The study by Bio/dynamics (1981b) has been selected to serve as the
basis for determination of the Lifetime HA value for glyphosate. In this
study, the reproductive toxicity of glyphosate in rats was investigated over
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Glyphosate August, 1987
-8-
three generations. Even though no compound-related changes in the reproductive
indices were observed when compared to controls at a dose level of 30 rag/kg/day,
there were pathological changes of renal focal tubular dilation in male F3b
weanling rats at this level. Therefore, the lower dose level of 10 mg/kg/day
was identified as the MOAEL.
Using a NOAEL of 10 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD 0 (10 mg/kg/day) = Oe, mg/kg/day
where:
10 mg/kg/day = NOAEL, based on absence of renal focal tubular
dilation in rats.
100 3 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 = (0.1 mg/kg/day) (70 kg) = 3o5 mg/L (3,500 ug/L)
(2 L/day)
where:
0.1 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 = (3.5 mg/L) (20%) = 0.70 mg/L (700 ug/L)
where:
3.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), glyphosate may be classified
in Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
0 The evidence of carcinogenicity in animals is considered equivocal by
the Science Advisory Board (Pesticides), and has been classified in
Category D [Office of Pesticide Programs has requested the manufacturer
to conduct another study in animals (U.S. EPA, 1986)].
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Glyphosate ^gust, 1987
-9-
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No other criteria, guidelines or standards were found in the available
literature pertaining to glyphosate.
0 Tolerance of 0.1 ppm has been established for the combined residues
of glyphosate and its metabolite in or on raw agricultural commodities
(U.S. EPA, 1985a).
VII. ANALYTICAL METHODS
0 Analysis of glyphosate is by a high-performance liquid chromatographic
(HPLC) method applicable to the determination of glyphosate in water
samples (U.S. EPA, 1985B). In this method, a known volume of sample
is applied to a Bio-Rad prefilled AG 50W-X8 column. The column
effluent is injected via an auto injector onto a primary column
packed with a cation exchange resin, but used in an anion-exclusion
mode to eliminate interferences. The effluent from this column flows
onto a strong anion-exchange column where the analytical separation
is accomplished. Detection and quantitation are made with a spectro-
photometer at 570 nm. The method detection limit for glyphosate is
5 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies capable of effectively removing glyphosate from contami-
nated water.
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Glyphosate August, 1987
-10-
IX. REFERENCES
Bio/dynamics, Inc.* 1981a. Lifetime feeding study of glyphosate (Roundup
Technical). Project No. 77-2062 for Monsanto Co., St. Louis, MO. EPA
Accession Nos. 246617 and 246621. (Unpublished report)
Bio/dynamics, Inc.* 1981b. A three-generation reproduction study in rats
with glyphosate. Project No. 77-2063 for Monsanto Co., St. Louis, MO.
EPA Accession Nos. 245909 and 247793. (Unpublished report)
Brightwell, B., and J. Malik. 1978. Solubility, volatility, adsorption and
partition coefficients, leaching and aquatic metabolism of MON 0573 and
MON 0101: Report No. MSL-0207.
Meister, R.T., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company, p. C117.
Monsanto Company. 1975. Residue and metabolism studies in sugarcane and
soils. Montsanto Agricultural Products Company, 800 Lindbergh Blvd.,
St. Louis, MO.
Monsanto Company. 1982a. Material safety data sheet, glyphosate technical.
800 N. Lindbergh Blvd., St. Louis, MO. MSDS No. 107-83-6.
Monsanto Company. 1982b. Rodeo herbicide for aquatic vegetation management.
Technical manual. 800 N. Lindbergh Blvd., St. Louis, MO. 82-L01.
Monsanto Company. 1982c. The health and environmental safety aspects of
Roundup herbicide: An overview. 800 N. Lindbergh Blvd., St. Louis, MO.
Roundup Herbicide Bulletin No. 3.
Monsanto Company. 1983. Rodeo herbicide: Toxicological and environmental
properties. 800 N. Lindbergh Blvd., St. Louis, MO. Rodeo Herbicide
Bulletin No. 1.
NAS. 1977. National Academy of Sciences. Drinking water and health. Vol. I.
Washington, DC: National Academy of Sciences.
NAS. 1980. National Academy of Sciences, National Research Council. Drinking
water and health. Vol. 3. Washington, DC: National Academy Press.
pp. 77-80.
Njagi, G.D.E., and H.N.B. Gopalan. 1980. Mutagenicity testing of some
selected food preservatives, herbicides and insecticides. Bangladesh
J. Bot. 9:141-146. (abstract only)
Olorunsogo, 0.0. 1981. Inhibition of energy-dependent transhydrogenase
reaction by N-(phosphonomethyl)glycine in isolated rat liver mitochondria.
Toxicol. Lett. 10:91-95.
Olorunsogo, O.O., and E.A. Bababunmi. 1980. Inhibition of succinate-linked
reduction of pyridine nucleotide in rat liver mitochondria "in vivo" by
N-(phosphonomethyl)glycine. Toxicol. Lett. 7:149-152.
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Glyphosate August, 1987
-11-
Olorunsogo, O.O., E.A. Bababunmi and 0. Bassir. 1977. Toxicity of glyphosate.
Proceedings of the 1st International Congress on Toxicology. G.L. Plaa
and W.A.M. Duncan, eds. New York: Academic Press, p. 597. (abstract
only)
Olorunsogo, O.O., E.A. Bababunmi and O. Bassir. 1979a. Effect of glyphosate
on rat liver mitochondria in vivo. Bull. Environ. Contam. Toxicol.
22:357-364.
Olorunsogo, O.O., E.A. Bababunmi and 0. Bassir. 1979b. The inhibitory effect
of N-(phosphonomethyl)qlycine in vivo on energy-dependent, phosphate-
induced swelling of isolated rat liver mitochondria. Toxicol. Lett.
4:303-306.
Rueppel, M.L., B.B. Brightwell, J. Schaefer and J.T. Marvel. 1977. Metabolism
and degradation of glyphosate in soil and water. J. Agric. Food Chem.
25:517-528.
Seiler, J.P. 1977. Nitrosation in vitro and in vivo by sodium nitrite, and
mutagenicity of nitrogenous pesticides. Mutat. Res. 48:225-236.
Shoval, S., and S. Yariv. 1981. Infrared study of the fine structures of
glyphosate and Roundup. Agrochimica. 25:377-386.
STORET. 1987.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.364. July 1.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. U.S. EPA Method 140
- Revision A - Glyphosate. Fed Reg. 50:40701. October 4, 1985.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34002. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidance for the
registration of pesticide products containing glyphosate as the active
ingredient. Case No. 0178, June, 1986.
Weed Science Society of America. 1983. Herbicide handbook, 5th ed.
Champaign, IL: Weed Science Society of America, pp. 258-263.
'Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
August, 1987
HEXAZINONE
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 thar another.
Btcause each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Hexazinone August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No.: 51235-04-2
Structural Formula:
a
V V
N^
I
CHa
3-Cyclohexyl-6-(dimethylamino)-1 methyl-1,3,5-triazine-2,4(lH,3H)-dione;
Synonyms
0 Velpar; Hexazinone.
Use
0 Contact and residual herbicide (Meister, 1983).
0 Usage areas include plantations of coniferous trees, railroad right-
of-ways, utilities, pipelines, petroleum tanks, drainage ditches, and
sugar and alfalfa (Kennedy, 1984).
Properties (Kennedy, 1984; CHEMLAB, 1985)
Chemical Formula C11H20°2N3
Molecular Weight 226 (calculated)
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 115-117°C
Density —
Vapor Pressure (86°C) 6.4 x 10"5 mir Hg
Specific Gravity
Water Solubility (25°C) 33,000 mg/L
Log Octanol/Water Partition -4.40 (calculated)
Coefficient
Taste Threshold
Odor Threshold odorless
Conversion Factor —
Occurrence
0 Hexazinone has been found in none of the surface water samples
or ground water samples analyzed from 13 samples taken at 6
locations (STORET, 1987).
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Hexazinone August, 1987
-3-
Environmental Fate^
0 Hexazinone did not hydrolyze in water within the pH range of 5.7 to 9
during a period of 8 weeks (Rhodes, 1975a).
0 In a soil aerobic metabolism study, hexazinone was added to a Fallsington
sandy loam and a Flanagan silt loam at 4 ppm. 14c-Hexazinone residues
had a half-life of about 25 weeks. Of the extractable 14C residues,
half was present as parent compound and/or 3-cyclohexyl-l-methyl-6-
methylamino-l,3,5-triazine-2,4-(lH,3H)-dione. Also present were
3_(4-hydroxycyclohexyl)-6-(dimethylamino)-l-methyl-l-(lH,3H)-dione
and the triazine trione (Rhodes, 1975b).
0 A soil column leaching study used 14c-hexazinone, half of which was
aged for 30 days and applied to Flanagan silt loam and Fallsington
sandy loam. Leaching with a total of 20 inches of water showed that
unaged hexazinone leached in the soils; however, leaching rates were
slower for the aged samples, indicating that the degradation products
may have less potential for contaminating ground water (Rhodes, 1975b).
0 A field soil leaching study indicated that 14c-hexazinone residues
were leached into the lower sampling depths with increasing rainfall.
A Keyport silt loam (2.75% organic matter; pH 6.5} and a Flanagan
silt loam (4.02% organic matter; pH 5.0) were used. For the Keyport
silt loam, 14C residues were found at all depths measured, including
the 8- to 12-inch depth, when total rainfall equaled 8.43 inches,
1 month after application of hexazinone. For the Flanagan silt loam,
14c residues were found at all depths sampled, including the 12- to
15-inch depth, 1 month after application, when a total of 7.04 inches
of rain had fallen (Rhodes, 1975c).
0 A soil TLC test for Fallsington sandy loam and Flanagan silt loam
gave Rf values for hexazinone of 0.85 and 0.68, respectively. This
places hexazinone in Class 4, indicating it is very mobile in these
soils (Rhodes, 1975c).
0 In a terrestrial field dissipation study using a Keyport silt loam
in Delaware, hexazinone had a half-life of less than 1 month. In a
field study in Illinois ' (Flanagan silt loam), hexazinone had a half-
life of more than 1 month (62% of the parent compound remained at
1 month) (Rhodes, 1975b). Tn a separate study with Keyport silt
loam, some leaching of the parent compound to a depth of 12 to 18
inches was observed (Holt, 1979).
III. PHARMACOKINETICS
Absorption
0 Rapisarda (1982) reported that a dose of 14 mg/kg 14c-labeled
hexazinone (>99% pure) was about 80% absorbed in 3 to 6 days
(77% recovery in urine, 20% in feces) when administered by gastric
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Hexazinone November 11, 1986
-4-
intubation to male and female Charles River CD rats with or without
3 weeks of dietary preconditioning with unlabeled hexazinone.
0 Rhodes et al. (1978) administered 2,500 ppm (125 mg/kg) hexazinone in
the diet to male rats for 17 days. This was followed by a single dose
of 18.3 mg/300 g (61 mg/kg) 14c-labeled hexazinone. The hexazinone
was rapidly absorbed within 72 hours, with 61% detected in the urine
and 32% in the feces. Trace amounts were found in the gastro-
intestinal (GI) tract (0.6%, tissues not specified) and expired air
(0.08%).
Distribution
0 Orally administered hexazinone has not been demonstrated to accumulate
preferentially in any tissue (Rhodes et al., 1978; Holt et al., 1979;
Rapisarda, 1982).
0 Studies in rats by Rapisarda (1982) and Rhodes et al. (1978) showed
that no detectable levels of 14C-hexazinone were found in any body
tissues when the animals were administered >14 mg/kg hexazinone by
gastric intubation with or without dietary preconditioning.
0 In a study with dairy cows by Holt et al. (1979) hexazinone was given
in the diet at 0, 1, 5 or 25 ppm for 30 days. Assuming that 1 ppm in
the diet of a cow equals 0.015 mg/kg (Lehman, 1959), these levels
correspond to 0, 0.015, 0.075 or 0.37 mg/kg/day. The investigators
reported no detectable residues in milk, fat, liver, kidney or lean
muscle.
Metabolism
0 Major urinary metabolites of hexazinone in rats identified by Rhodes
et al. (1978) were 3-(4-hydrocyclohexyl)-6-(dimethylamino)1-methyl-
1,3,5-triazine-2,4-{lH,3H)-dione (metabolite A); 3-cyclohexyl-6-
(methylamino)-1-methyl-1,3,5-triazine-2,4-(1H, 3H)-dione (metabolite B);
and 3-(4-hydrocyclohexyl)-6-(methylamino)-1-methyl-1,3,5-triazine-2,4-
(lH,3H)-dione (metabolite C). The percentages of these metabolites
detected in the urine were 46.8, 11.5 and 39.3%, respectively.
The major fecal metabolites detected by Rhodes et al. (1978) were
A (26.3%) and C (55.2%). Less than 1% unchanged hexazinone was
detected ir the urine or the feces. Similar results were reported
by Rapisarda (1982).
Excretion
Rapisarda (1982) and Rhodes et al. (1978) reported that excretion of
14c-hexazinone and/or its metabolites occurs mostly in the urine
(61 to 77%) and in the feces (20 to 32%).
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Hexazinone August, 1987
-5-
IV. HEALTH EFFECTS
Humans
0 The Pesticide Incident Monitoring System data base (U.S. EPA, 1981)
indicated that 3 of 43,729 incident reports involved hexazinone.
Only one report cited exposure to hexazinone alone, without other
compounds involved. A 26-year-old woman inhaled hexazinone dust
(concentration not specified). Vomiting occurred within 24 hours.
No other effects were reported and no treatment was administered.
The other two reports did not involve human exposure.
Animals
Short-term Exposure
0 Reported oral LD^g values for technical-grade hexazinone in rats range
from 1,690 to >7,500 mg/kg (Matarese, 1977; Dashiell and Hinckle,
1980; Kennedy, 1984).
0 Henry (1975) and Kennedy (1984) reported the oral LD50 value of
technical-grade hexazinone in beagle dog<* to be >3,400 mg/kg.
0 Reported oral LD5Q values for hexazinone in guinea pigs range from
800 to 860 mgAg (Dale, 1973; Kennedy, 1984).
0 Kennedy (1984) studied the response of male rats to repeated oral
doses of hexazinone (89 or 98% active ingredient). Groups of six
rats were in tuba ted with hexazinone, 300 mg/kg, as a 5% suspension
in corn oil. Animals were dosed 5 days/week for 2 weeks (10 total
doses). Clinical signs and body weights were monitored daily. At
4 hours to 14 days after exposure to the last dose, microscopic
evaluation of lung, trachea, liver, kidney, heart, testes, thymus,
spleen, thyroid, GI tract, brain, and bone marrow was conducted. No
gross or histological changes were noted in animals exposed to either
active ingredient percentage of hexazinone.
0 In an 8-week range-finding study (Kennedy and Kaplan, 1984), Charles
River CD-I mice (10/sex/ddse) received hexazinone (>98% pure) in the
diet for 8 consecutive weeks at concentrations of 0, 250, 500, 1,250,
2,500 or 10,000 ppm. Assuming 1 ppm in the diet of mice equals
0.15 mg/kg (Lehman, 1959), these dietary concentrations correspond to
doses of about 0, 37.5, 75.0, 187.5, 375.0 or 1,500 mg/kg/day. No
differences were observed in general behavior and appearance, mortality,
body weights, food consumption or calculated food efficiency between
control and exposed groups. No gross pathologic lesions were detected
at necropsy. The only dose-related effects observed were increases
in both absolute and relative liver weights in mice fed 10,000 ppm. A
No-Observed-Adverse-Effect-Level (NOAEL) of 2,500 ppm (375.0 mg/kg/day)
was identified by the authors.
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Hexazinone August, 1987
-6-
Dermal/Ocular Effects
0 In an acute dermal toxicity test performed by McAlack (1976), up to
7,500 mg/kg of a 24% aqueous solution of hexazinone (reported to be
1,875 mgAg of active ingredient) was applied occlusively for 24
hours to the shaved backs and trunks of male albino rabbits. No
deaths were observed throughout a 14-day observation period.
0 Morrow (1973) reported an acute dermal toxicity test in which 60 mL
of a 24% aqueous solution of hexazinone (reported as 5,278 mg/kg) was
applied occlusively to the shaved trunks of male albino rabbits for 24
hours. No mortalities were observed through an unspecified observation
period. One animal exhibited a mild, transient skin irritation.
0 In a 10-day study conducted by Kennedy (1984), semiocclusive dermal
application of hexazinone for 6 hours/day for 10 days to male rabbits
at 70 or 680 mg/kg/day resulted in no signs of skin irritation or
toxicity. A trend toward elevated serum alkaline phosphatase (SAP)
and serum glutamic pyruvic-transaminase (SGPT) activities was observed,
but no hepatic damage was seen by microscopic evaluation. In a
second 10-day study using 35, 150 or 770 mg/kg/day, the highest dose
again resulted in elevated SAP and SGPT activities, but they returned
to normal after 53 days of recovery. Histopathological evaluations
were not performed in the second study.
0 Edwards (1977) applied 6,000 mg/kg hexazinone as a 63% solution occlu-
sively to the shaved backs and trunks of male albino rabbits. No
treatment-related mortalities were reported after a 14-day observation
period.
0 Morrow (1972) reported the results of dermal irritation tests in which
a single dose of 25 or 50% hexazinone was applied to the shaved, intact
shoulder skin of each of 10 male guinea pigs. To test for sensitization,
four sacral intradermal injections of 0.1 mL of a 15% solution were first
given over a 3-week period. After a 2-week rest period, the guinea
pigs were challenged with 25 or 50% hexazinone applied to the shaved,
intact shoulder skin. The test material was found to be nonirritating
and nonsensitizing at 48 hours post-application.
0 Using a 10% solution, Goodman (1976) repeated the Morrow study with
guinea pigs and observed no irritation or sensitization.
0 Dashiell and Henry (1980) reported that in albino rabbits, a single
dose of hexazinone applied as 27% (vehicle not specified) solution to
one eye per animal and unwashed was a severe ocular irritant. When
applied at 27% (vehicle not specified) and washed or at 4% (aqueous
solution) unwashed, mild to moderate corneal cloudiness, iritis
and/or conjunctivitis resulted. By 21 days post-treatment with the
higher dose, two of the three rabbit eyes had returned to normal; a
small area of mild corneal cloudiness persisted through the 35-day
observation period in one of the three eyes. Eyes treated with lower
doses were normal within 3 days.
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Hexazinone August, 1987
-7-
Long-term Exposure
0 In a 90-day feeding study, Sherman et al. (1973) fed beagle dogs
(four/sex/dose) hexazinone (97.5% active ingredient) in the diet
at levels of 0, 200, 1,000 or 5,000 ppm. Assuming 1 ppm in the diet
of a dog equals 0.025 mg/kg/day (Lehman, 1959), these levels correspond
to about 0, 5, 25 or 125 mg/kg/day. At the highest dose level tested,
decreased food consumption, weight loss, elevated alkaline phosphatase
activity, lowered albumin/globulin ratios and slightly elevated liver
weights were noted. No gross or microscopic lesions were observed
at necropsy. Based on the results of this study a NOAEL of 1,000 ppm
(25 mg/kg/day) and a Lowest-Observed-Adverse-Effect-Level (LOAEL) of
5,000 ppm (125 mg/kg/day) were identified.
0 In a 90-day feeding study (Kennedy and Kaplan, 1984), Crl-CD rats
(16/sex/dose) received hexazinone (>98% pure) at dietary levels of
0, 200, 1,000 or 5,000 ppm. Assuming 1 ppm in the diet of rats
equals 0.05 mg/kg/day (Lehman, 1959), these levels correspond to
about 0, 10, 50 or 250 mg/kg/day. Hematological and biochemical
tests and urinalyses were conducted on subgroups of animals after 1,
2 or 3 months of feeding. Following 94 to 96 days of feeding, the
rats were sacrificed and necropsied. The only statistically significant
effect reported was a decrease in body weight in both males and
females receiving 5,000 ppm. No differences in food consumption were
reported. Results of histopathological examinations from the control
and high-dose groups were unremarkable. The authors identified a
NOAEL of 1,000 ppm (50 mg/kg/day).
0 In a 1-year feeding study (Kaplan et al., 1975) weanling Charles River
CD rats (36/sex/dose) received hexazinone (94 to 96% pure) at dietary
levels of 0, 200, 1,000 or 2,500 ppm (which, according to the authors,
corresponds to 0, 11, 60 or 160 mg/kg/day for males and 0, 14, 74 or
191 mg/kg/day for females). Results of this study indicated a decrease
in weight gain by both sexes at 2,500 ppm and by females at 1,000 ppm.
The authors indicated that various unspecified clinical, hematological
and biochemical parameters revealed no evidence of adverse effects.
No significant gross or histopathological changes attributable to
hexazinone were noted. From the information presented in the study,
a NOAEL of 200 ppm (11 mg/kg/day for males and 14 mg/kg/day for
females) can be identified.
0 In a 2-year study, Goldenthal and Trumball (1981) fed hexazinone
(95 to 98% pure) to Charles River CD-I mice (80/sex/dose) at dietary
levels of 0, 200, 2,500 or 10,000 ppm. Assuming that 1 ppm in the
diet of a mouse equals 0.15 mg/kg/day (Lehman, 1959), these levels
correspond to 0, 30, 375 or 1,500 mg/kg/day. Corneal opacity sloughing
and discoloration of the distal tip of the tail were noted as early
as the fourth week of the study in mice receiving 2,500 or 10,000 ppm.
A statistically significant decrease in body weight was observed in
male mice receiving 10,000 ppm and in female mice receiving 2,500 or
10,000 ppm. Statistically significant increases in liver weight were
noted in male mice receiving 10,000 ppm; male and female mice in the
10,000-ppm dose group also displayed statistically significant increases
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Hexazinone August, 1987
-8-
in relative liver weight. Sporadic occurrence of statistically
significant changes in hematological effects were considered by
the authors to be unrelated to hexazinone treatment. Histologically,
a number of liver changes were observed among mice fed 2,500 or
10,000 ppm. The most characteristic finding was hypertrophy of
centrilobular parenchymal cells. Other histological changes included
an increased incidence of hyperplastic liver nodules and an increased
incidence and severity of liver cell necrosis. Mice fed 200 ppm
showed no compound-related histopathological changes. A NOAEL of
200 ppm (30 mg/kg/day) was identified by the authors.
0 Kennedy and Kaplan (1984) presented the results of a 2-year feeding
study in which Crl-CD rats (36/sex/dose) received hexazinone (94 to
96% pure) at dietary levels of 0 (two groups), 200, 1,000 or 2,500 ppm
(approximately 0, 10, 50 or 125 mg/kg/day assuming that 1 ppm in the
diet of a rat equals 0.05 mg/kg/day)(Lehman, 1959). After 2 years
of continuous feeding, all rats in all groups were sacrificed and
examined. Males fed 2,500 ppm and females fed either 1,000 or 2,500
ppm had significantly lower body weights than controls (p 98% pure) for
approximately 90 days at dietary levels of 0, 200, 1,000 or 5,000 ppm.
Assuming that 1 ppm in the diet of rats equals 0.05 mg/kg/day (Lehman,
1959), this corresponds to approximately 0, 10, 50 and 250 mg/kg/day.
Following the 90-day feeding period, six rats/sex/dose were selected
to serve as the parental generation. The authors concluded that the
rats had normal fertility. The young were delivered in normal numbers,
and survival during the lactation period was unaffected. In the
5,000 ppm group, weights of pups at weaning (21 days) were significantly
(p <0.01) lower than controls or other test groups. The results of
this study identify a NOAEL of 1,000 ppm (50 mg/kg/day) (no decrease
in weanling weight).
0 In a three-generation reproduction study (DuPont, 1979), Crl-CD rats
(36/sex/dose) received hexazinone (98% pure) at dietary levels of 0,
200, 1,000 or 2,500 ppm for 90 days (approximately 0, 10, 50 or 125
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Hexazinone August, 1987
-9-
mg/kg/day, assuming the above assumptions for a rat). Following
90 days of feeding, 20 rats/sex/dose were selected to serve as the
parental (F0) generation. Reproductive parameters tested included
the number of matings, number of pregnancies and number of pups per
litter. Pups were weighed at weaning, and one male and female were
selected from each litter to serve as parental rats for the second
generation. Similar procedures were used to produce a third generation;
the same reproductive parameters were collected for the second and
third generations. The authors stated that there were no significant
differences between the control and treated groups with respect to
the various calculated indices (fertility, gestation, viability and
lactation). However, body weights at weaning of pups in the 2,500 ppm
dose group were significantly (p <0.05) lower than those of controls
for the F2 and F3 litters. The results of this study identify a
NOAEL of 1,000 ppm (50 mg/kg/day).
Developmental Effects
0 Kennedy and Kaplan (1984) presented the results of a study in which
Charles River Crl-CD rats (25 to 27/dose) received hexazinone (97.5%
pure) at dietary concentrations of 0, 200, 1,000 or 5,000 ppm (approxi-
mately 0, 10, 50 or 250 mg/kg/day following the previously stated
assumptions for the rat) on days 6 through 15 of gestation. Rats
were observed daily for clinical signs and were weighed on gestation
days 6, 16 and 21. On day 21, all rats were sacrificed and ovaries
and uterine horns were weighed and examined. The number and location
of live fetuses, dead fetuses and resorption sites were noted.
Fetuses from the 0 and 5,000 ppm dose groups were evaluated for
developmental effects (gross, soft tissue or skeletal abnormalities).
At sacrifice, no adverse effects were observed for the dams. No
malformations were noted in the fetuses. However, pup weights in the
high-dose group were significantly lower than in the controls. This
study identified a NOAEL of 1,000 ppm (50 mg/kg/day).
0 Kennedy and Kaplan (1984) presented the results of a study in which
New Zealand white rabbits (17/dose) received hexazinone suspended in
a 0.5% aqueous methyl cellulose vehicle by oral intubation on days 6
through 19 of gestation at levels of 0, 20, 50 or 125 mg/kg/day.
Rabbits were observed daily and body weights were recorded throughout
gestation. On day 29 of gestation, all rabbits were sacrificed, uteri
were excised and weighed, and the number of live, dead and resorbed
fetuses was recorded. Each fetus was examined externally and internally
for gross, soft tissue and skeletal abnormalities. No clinical signs
of maternal or fetal toxicity were observed. Pregnancy rates among
all groups compared favorably. The numbers of corpora lutea and
implantations per group were not significantly different. Resorptions
and fetal viability, weight and length were also similar among all
groups. Based on the information presented in this study, a minimum
NOAEL of 125 mg/kg/day for maternal toxicity, fetal toxicity, and
teratogenicity can be identified.
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Hexazinone August, 1987
-10-
Mutagenicity
0 The ability of hexazinone to induce unscheduled ONA synthesis was
assayed by Ford (1983) in freshly isolated hepatocytes from the livers
of 8-week-old male Charles River/Sprague-Dawley rats. Hexazinone
was tested at half-log concentrations from 1 x 10-5 to 10.0 mM and at
30.0 mM. No unscheduled DNA synthesis was observed.
0 Valachos et al. (1982) conducted an in vitro assay for chromosomal
aberrations in Chinese hamster ovary cells. Hexazinone was found to
be clastogenic without S-9 activation at concentrations of 15.85 mM
(4.0 mg/mL) or 19.82 mM (5.0 mg/mL); no significant increases in
clastogenic activity were seen at 1.58, 3.94 and 7.93 mM (0.4, 1.0
and 2.0 mg/mL). With S-9 activation, significant increases in aber-
rations were noted only at a concentration of 15.85 mM (4.0 mg/mL).
0 In a study designed to evaluate the clastogenic potential of hexazinone
in rat bone marrow cells (Farrow et al., 1982), Sprague-Dawley CD rats
(12/sex/dose) were given a single dose of 0, 100, 300 or 1,000 mg/kg
of the hexazinone by gavage (vehicle not reported). No statistically
significant increases in the frequency of chromosomal aberrations were
observed at any of the dose levels tested. The authors concluded that,
under the conditions of this study, hexazinone was not clastogenic.
0 Hexazinone was tested for mutagenicity in Salmonella typhimurium
strains TA1535, TA1537, TA1538, TA98 and TA1OO at concentrations up
to 7,000 ug/plate. The compound was not found to be mutagenic, with
or without S-9 activation (DuPont, 1979).
Carcinogenic!ty
0 Goldenthal and Trumball (1981) fed hexazinone (98% pure) for 2 years
to mice (80/sex/dose) in the diet at 0, 200, 2,500, or 10,000 ppm
(0, 30, 375 or 1,500 mg/kg/day, based on Lehman [1959]). A number
of liver changes were observed histologically at the 2,500- and
10,000-ppm level. These included hypertrophy of the centrilobular
parenchymal cells, increased incidence of hyperplastic liver nodules
and liver cell necrosis. The authors concluded that hexazinone was
not carcinogenic to mice.
0 No carcinogenic effects were observed in C:1-CD rats (36/sex/dose)
given hexazinone (94 to 96% pure) in the diet at 0, 200, 1,000, or
2,500 ppm (0, 10, 50, or 125 mg/kg/day) for 2 years (Kennedy and
Kaplan, 1984). The authors concluded that none of the tumors were
attributable to hexazinone.
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:
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Hexazinone Au*ust' 1987
-11-
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 ing/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 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 for hexazinone. It is, therefore,
recommended that the Longer-term HA value of 2.5 mg/L (2,500 ug/L, calculated
below) for a 10-kg child be used at this time as a conservative estimate of
the One-day HA value.
Ten-day Health Advisory
The study reported by Kennedy and Kaplan (1984) in which pregnant rabbits
(17/dose) received hexazinone by oral intubation at levels of 0, 20, 50 or
125 mg/kg/day on days 6 through 19 of gestation was considered to serve as
the basis for deriving the Ten-day HA for a 10-kg child. Since no signs of
maternal or fetal toxicity were observed in this study, a NOAEL of 125 mg/kg/day
(the highest dose tested) was identified. The NOAEL from this study is
greater than that identified in a 90-day rat feeding study (50 mg/kg; Kennedy
and Kaplan, 1984). The LOAEL from the one-generation rat reproduction study
was 250 mg/kg based on decreased weanling weight. Effects at doses between
50 and 250 mg/kg have not been reported for the rat. However, in a 90-day
dog study, a LOAEL of 125 mg/kg was identified (Sherman et al., 1973).
Therefore, the rabbit study was hot selected to derive a Ten-day value.
It is, therefore, recommended that the Longer-term HA value of 2.5 mg/L
(2,500 ug/L) for the 10-kg child be used at this time as a conservative
estimate of the Ten-day HA value.
Longer-term Health Advisory
The 90-day feeding study in dogs (Sherman et al., 1973) has been selected
to serve as the basis for determination of the Longer-term HA for hexazinone.
In this study, dogs received hexazinone in the diet at levels of 0, 200,
1,000 or 5,000 ppm (0, 5, 25, or 125 mg/kg/day) for 90 days. Decreased food
consumption and body weight gain, elevated alkaline phosphatase activity,
lowered albumin/globulin ratios and elevated liver weights were observed at
the highest dose. A NOAEL of 1,000 ppm (25 mg/kg/day) and a LOAEL of 5,000 ppm
(125 mg/kg/day) were identified. This NOAEL is generally supported by a 90-day
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Hexazinone Au*ust' 1987
-12-
rat feeding study that reported a NOAEL of 50 mg/kg/day (Kennedy and Kaplan,
1984). Effects in dogs exposed to hexazinone at 50 mg/kg/day have not been
reported.
Using a NOAEL of 25 mg/kg/day, the Longer-term HA for a 10-kg child
is calculated as follows:
Longer-term HA = (25 mg/kg/day) (10 kg) = 2.5 mg/L (2.500 ug/L)
(100) (1 L/day)
where:
25 mgAg/day = NOAEL, based on absence of hepatic effects or weight loss
in dogs exposed to hexazinone via the diet for 90 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
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 a 70-kg adult is calculated as follows:
Longer-term HA = (25 mg/kg/day) (70 kg) = 8.75 mg/L (8f750 ug/L)
(100) (2 L/day)
where:
25 mgA9/day = NOAEL, based on absence of hepatic effects or weight
loss in dogs exposed to hexazinone via the diet for
90 days.
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 ar. 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
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Hexazinone Au*ust' 1987
-13-
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.
A 2-year rat feeding/oncogenicity study (Kennedy and Kaplan, 1984) was
selected as the basis for determination of the Lifetime HA for hexazinone.
Crl-CD rats (36/sex/dose) received 0, 200, 1,000, or 2,500 ppm hexazinone (0,
10, 50, or 125 mg/kg/day) for 2 years. Body weight gain in males and females
in the 2,500-ppm group, and females in the 1,000-ppm group, was significantly
lower than that in controls. No clinical, hematological or urinary evidence
of toxicity was reported. Based on decreased body weight gain, a NOAEL of
200 ppm (10 mg/kg/day) and LOAEL of 1,000 ppm (50 mg/kg/day) were identified.
Using a NOAEL of 10 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (10 mg/kg/day) = 0.03 mg/kg/day
(100) (3)
where:
10 mg/kg/day = NOAEL, based on absence of body weight effects in rats
exposed to hexazinone via the diet for 2 years.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = modifying factor; to account for data gaps (chronic
dog-feeding study) in the total data base for hexazinone,
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - (0.03 mg/kg/day) (70 kg) = 1.05 mg/day (1,050 ug/L)
(2 L/day)
where:
0.03 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Hexazinone August, 1987
-14-
Step 3: Determination of Lifetime Health Advisory
Lifetime HA = (1.05 mg/L) (20%) = 0.21 mg/L (210 ug/L)
where:
1.05 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic!ty in rats or mice has been observed.
0 The International Agency for Research on Cancer has not evaluated
the carcinogenic potential of hexazinone.
0 The criteria described in EPA's guidelines for assessment of car-
cinogenic risk (U.S. EPA, 1986), place hexazinone in Group O: not
classified. This category is for substances with inadequate animal
evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Residue tolerances range from 0.5 to 5.0 ppm for the combined residues
of hexazinone and its metabolites in or on the raw agricultural
commodities pineapple, pineapple fodder and forage (U.S. EPA, 1985a).
VII. ANALYTICAL METHODS
0 Analysis of hexazinone is by a gas chromatographic method applicable
to the determination of certain organonitrogen pesticides in water
samples (U.S. EPA, 1985b). This method requires a solvent extraction
of approximately 1 liter of sample with methylene chloride using a
separately funnel. The methylene chloride extract is dried and
exchanged to acetone during concentration to a volume of 10 mL or
less. The compounds in the extract are separated by gas chromatography,
and measurement is made with a thermionic bead detector. The method
detection limit for hexazinone is 0.72 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies used to remove hexazinone from contaminated water.
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Hexazinone Au*ust' 198?
-15-
IX. REFERENCES
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD.
Dale, N.* 1973. Oral LD50 test (guinea pigs). Haskell Laboratory Report
No. 400-73, unpublished study. MRID 00104973.
Dashiell, O.L., and J.E. Henry.* 1980. Eye irritation tests in rabbits—United
Kingdom Procedure. Haskell Laboratory Report No. 839-80, unpublished
study. MRID 00076958.
Dashiell, O.L., and L. Hinckle.* 1980. Oral LD50 test in rats—EPA proposed
guidelines. Haskell Laboratory Report No. 943-80, unpublished study.
MRID 00062980.
DuPont.* 1979. E.I. duPont de Nemours & Co. Supplement to Haskell Laboratory
Report. No. 352-77. Reproduction study in rats with sym-triazine-2,4(lH,
3H)-dione, 3-cyclohexyl-1-methyl-6-dimethylamino (INA 3674, hexazinone).
Accession No. 97323.
Edwards, D.F.* 1977. Acute skin absorption test on rabbits LD50. Haskell
Laboratory Report No. 841-77, unpublished study. MRID 00091140.
Farrow, M, T. Cartina, M. Zito et. al.* 1982. In vivo bone marrow cytogenetic
assay in rats. HLA Project No. 201-573. Final Report. (Unpublished
study received July 11, 1983 under 352-378.) Submitted by E.I. duPont
de Nemours & Co., Inc., Wilmington, DE. MRID 0013155.
Ford, L.* 1983. Unscheduled DNA synthesis/rat hepatocytes in vitro.
(INA-3674-112). Haskell Laboratory Report No. 766-82, unpublished
study. MRID 00130708.
Goldenthal, E.I. and R.R. Trumball.* 1981. E.I. duPont de Nemours & Co.,
Inc. Two-year feeding study in mice. IRDC No. 125-026, unpublished
study. Submitted to the Office of Pesticide Programs. MRID No. 0079203.
Goodman, N.* 1976. Primary skin irritation and sensitization tests on guinea
pigs. Report No. 434-76, unpublished study. Submitted to the Office of
Pesticide Programs. MRID 00104433.
Henry, J.E.* 1975. Acute oral test (dogs). Haskell Laboratory Report No.
617-75, unpublished study. MRID 00076957.
Holt, R.F., F.J. Baude and D.W. Moore.* 1979. Hexazinone livestock feeding
studies; milk and meat. Unpublished study. Submitted to the Office of
Pesticide Programs. MRID 00028657.
Holt, R.F. 1979. Residues resulting from application of DPX-3674 to soil.
E. I. duPont de Nemours & Co., Inc., Wilmington, DE.
Kaplan, A.M., Z.A. Zapp, Jr., C.F. Reinhardt et al.* 1975. Long-term
feeding study in rats with sym-triazine-2,4(1H,3H)dione, 3-cyclohexyl-1-
methyl(-6-dimethylamino (INA-3674). One-year Interim Report. Haskell
Laboratory Report No. 585-75. MRID 00078045.
-------
August, 1987
Hexazinone *
-16-
Kennedy, G.L. 1984. Acute environmental toxicity studies with hexazinone.
Fund. Appl. Tbxicol. 4:603-611.
Kennedy, G.L.. and A.M. Kaplan. 1984. Chronic toxicity, "P'0*"^^' and
teratogenic studies of hexazinone. Fund. Appl. Tbxicol. 4:960-971.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs, and
cosmetics. Assoc. Food Drug Off. of the U.S.
Matarese, C.* 1977. Oral LD50 test. Haskell Laboratory Report No. 1037-77,
unpublished study. MRID 0011477.
McAlack, J.W.* 1976. Skin absorption LD50. Haskell Laboratory Report No.
353-76, unpublished study. MRID 00063971.
Meister, R, ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Morrow, R.« 1972. Primary skin irritation and sensitization tests on
pigs. Haskell Laboratory Report No. 489-72, unpublished study. MRID
00104978.
Morrow, R.* 1973. Skin absorption toxicity ALD and skin i«itancy test.
Haskell Laboratory Report No. 503-73, unpublished study. MRID 00104974.
Rapisarda, C.* 1982. Metabolism of 1 4C-labeled hexazinone in the rat. E.I.
duPont de Nemours & Co. Document No. AMR-79-82, unpublished study.
Accession No. 247874.
Rhodes, Robert C. 1975a. Studies with "Velpar" weed killer in water.
Biochemicals Department Experimental Station, E. I. duPont de Nemours
& Co., Inc., Wilmington, DE.
Rhodes, Robert C. 1975b. Decomposition of "Velpar- weed killer in soil.
Biochemicals Department Experimental Station, E. I. duPont de Nemours
& Co., Inc., Wilmington, DE.
Rhodes, Robert C. 1975c. Mobility and adsorption studies with -Velpar"
weed killer on soils. Biochemicals Department Experimental Station,
E. I. duPont de Nemours & Co. , Inc. , Wilmington, DE.
Rhodes, R, R.A. Jewell and H. Sherman.* 1978. Metabolism of Velpar (R) weed
killer in the rat. Unpublished study. E. I. duPont de Nemours & Co., Inc.
MRID 00028864.
Sherman, H, N. Dale and L. Adams et al.« 1973. Three month feeding study in
dogs with sym-triazine-2,4UH,3H)-dione, 3-cyclohexyl-1 -methy (-6-dimethyl-
amino-(INA-3674). Haskell Laboratory Report No. 408-73. MRID
STORET. 1987.
U.S. EPA. 1981. U.S. Environmental Protection Agency. Pesticide Incident
Monitoring System. Office of Pesticide Programs, Washington, DC.
February.
-------
Hexazinone August, 1987
-17-
U.S. EPA. 1982. U.S. Environmental Protection Agency. Toxicology Chapter.
Registration Standard for Hexazinone. Office of Pesticide Programs,
Washington, DC.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.396.
U.S. EPA. 19855. U.S. Environmental Protection Agency. U.S. EPA Method 633
- Organonitrogen Pesticides. Fed. Reg. 50:40701. October 4, 1985.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
Valachos, D, J. Martenis and A. Horst.* 1982. In vitro assay for chromosome
aberrations in Chinese Hamster Ovary (CHO) cells. Haskell Laboratory
Report No. 768-82, unpublished study. MRIO 00130709.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
DRAFT
MALE1C HYDRAZIDE
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 ->ne of these models is able to predict risk more accurately than ai ather.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Maleic Hydrazide
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 123-33-1
Structural Formula
2-Dihydro-3,6-pyridazinedione
Synonyms
0 Antergon; Chemform; De-Sprout; Retard; Slo-Gro; Sucker-Stuff;
(Meister, 1983).
Uses
0 Plant growth retardant (Meister, 1983).
Properties (Meister, 1983; CHEMLAB, 1985; TDB, 1985)
C4H402N2
112.09
Crystalline solid
292°C
1.60
0 mm Hg
6,000 mg/L
-3.67 (calculated)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (50°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Hater Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 No information was found in the available literature on the occurrence
of maleic hydrazide.
Environmental Fate
0 Maleic hydrazide is very soluble in water (6,000 ppm) and in most
organic solvents (>1,000 ppm). The vapor pressure is essentially
zero (Registrant CBI data; WSSA, 1983).
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Maleic Hydrazide August, 1987
-3-
0 Salts of maleic hydrazide will dissociate in solutions above pH 4.5
and exist only as maleic hydrazide. Maleic hydrazide is stable to
hydrolysis at pHs of 3, 6 and 9. Photolysis potential has not been
addressed (Registrant CBI data; WSSA, 1983).
0 In field dissipation studies using various soils from the eastern,
southern and midwestern U.S., the half-lives were reported to be
between 14 and 100 days. There is no pattern, but the half-life may
be related to organic matter content. Degradation by soil micro-
organisms appears to be rapid (Registrant CBI data; WSSA, 1983).
0 There is some indication that maleic hydrazide is highly mobile
in imaged soils. Aerobic aging of maleic hydrazide results in a
lowering of leaching potential (Registrant CBI data; WSSA, 1983).
III. PHARMACOKINETICS
Absorption
0 Mays et al. (1968) administered single oral doses of 14C-labeled
maleic hydrazide to rats. After 6 days, only 12% had been excreted
in the feces, indicating that 88% had been absorbed.
Distribution
0 Kennedy and Keplinger (1971) administered ^c-labeled maleic hydrazide
to pregnant rats in daily doses of either 0.5 or 5.0 mg/kg. Fetuses
from dams sacrificed on day 20 were found to contain label equivalent
to 20 to 35 ppb of the parent compound at the 0.5-mg/kg dose level,
and 156 to 308 ppb at the 5.0-mg/kg dose level. Pups from females
that were allowed to litter were sacrificed at 8 and at 24 hours, and
stomach coagulum was analyzed to determine transfer through the milk.
At the 0.5 mg/kg dose, the coagulum contained 4 to 7 ppb at 8 hours
and 2 ppb at 24 hours; at the 5.0 mg/kg dose, the figures for 8 and
24 hours were 79 to 89 ppb and 7 to 8 ppb, respectively. These
results showed that maleic hydrazide crossed the placenta and was
also transmitted to the pups via the milk.
Metabolism
0 Barnes et al. (1957) reported that rabbits administered a single oral
dose of 100 mg/kg of maleic hydrazide excreted 43 to 62% of the dose,
unchanged, within 48 hours. The route of excretion (urinary or
fecal) was not stated. The results were similar following a dose of
2,000 mg/kg, and no glucuronide or ethereax sulfate conjugates were
found.
0 Oral administration of maleic hydrazide labeled with 14C to rats
resulted in excretion of 0.2% labeled carbon dioxide in the expired
air over a 6-day observation period (Mays et al., 1968). Urinary
products (77% of the total dose) were largely unchanged maleic
hydrazide (92 to 94% of the urinary total) and conjugates of maleic
hydrazide (6 to 8%).
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Maleic Hydrazide August, 1987
-4-
Excretion
Mays et al. (1968) administered single oral doses of 14C-maleic hydra-
zide to rats. Over a 6-day observation period, the animals excreted
0.2% of the label as carbon dioxide in the expired air, 12% in the
feces and 77% in the urine. Only trace amounts were detected in
tissues and blood after 3 days.
IV. HEALTH EFFECTS
Humans
0 No information on human exposure to maleic hydrazide was found in the
available literature.
Animals
Short-term Exposure
0 The acute oral toxicity of maleic hydrazide (purity not specified) in
rats was determined with administration of four dose levels to groups
of five animals, with a 15-day observation period (Reagan and Becci,
1982). At dose levels of 5,000, 6,300, 7,940 or 10,000 rag/kg, deaths
occurring in the male animals were 0/5, 0/5, 1/5 and 5/5, respectively,
while those for female animals were 1/5, 1/5, 4/5 and 5/5, respectively.
The LD50 values were calculated to be 6,300 mg/kg for males, 6,680
mg/Jcg for females and 7,500 mg/kg for both sexes combined. Adverse
effects noted included ataxia, diarrhea, salivation, decreased motor
activity and blood in the intestines and stomach.
0 Sprague-Dawley rats (five males and five females) were fasted for
16 hours and then given a single oral dose of technical maleic hydra-
zide (purity not specified) at a level of 5,000 mg/kg and observed
for 14 days (Shapiro, 1977a). No deaths occurred during this period.
Necropsies were not performed, and no details were given with respect
to adverse effects that may have been observed.
0 The acute oral toxicity of the diethanolamine salt of maleic hydrazide
(MH-DEA) (purity not specified) was determined in rats and rabbits
(Uniroyal Chmical, 1971). In both species, MH-DEA was lethal at a
level of 1,000 mg/kg, 'while doses between 300 and 500 mg/kg showed no
toxicity in either species. The LDso value for both species was cal-
culated to be 700 mg/kg.
0 Rats were used for a comparison of the acute oral toxicity of the
sodium and diethanolamine sains (purities not specified) of maleic
hydrazide (Tate, 1951). The diethanolamine salt showed an LDso
value of 2,350 mg/kg, while the LDso for the sodium salt (MH-Na)
was 6,950 mg/kg. No details of the study were given.
0 The acute oral LDso value of technical-grade maleic hydrazide (purity
not specified) for rabbits was greater than 4,000 mg/kg (Lehman,
1951). No details of the study were available.
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Maleic Hydrazide August, 1987
-5-
0 The acute oral toxicity of maleic hydrazide (purity not specified) in
four species (mouse, rat, rabbit and dog) was studied by Mukhorina
(1962). For all species, the LD50 was reported as 700 mg/kg, with an
LD1OO of 1»000 n>9A9 and a toxicity range from 300 to 500 rag/kg. For
rats and rabbits, adverse effects noted were cyanosis, tachypnea,
convulsions and paralysis; no other details were given.
Dermal/Ocular Effects
0 Technical-grade maleic hydrazide was tested on male and female New
Zealand rabbits for both skin and eye irritation (Shapiro, 1977b,c).
Applied at 0.5 mL, the maleic hydrazide was scored as a mild primary
skin irritant. In the eye test, 100 mg of the material was used, and
maleic hydrazide was judged not to be an eye irritant.
0 The acute dermal toxicity of maleic hydrazide (purity and form not
specified) was determined in five male and five female New Zealand
rabbits (Shapiro, 1977d). The skin of two males and three females
was abraded. A single dose of 20,000 mg/kg was applied, and the
animals were observed for 14 days. On the first day, two males
(one with abraded skin) and one female died. The animals that died
exhibited ataxia, shallow respiration and were comatose.
0 In an evaluation of the acute dermal toxicity of Royal MH-30 (30%
MH-DEA) and maleic hydrazide-technical, both formulations were stated
to be mild primary skin irritants and slight eye irritants (Uniroyal
Chemical, 1977). Individual details of the study were not given.
Long-term Exposure
0 Rats were fed MH-Na or MH-DEA (purity not specified) in the diet for
11 weeks (Tate, 1951). The MH-Na was given at dose levels of 0.5%
or 5.0% (5,000 or 50,000 ppm). Assuming that 1 ppm in the diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these doses
correspond to 250 or 2,500 mg/kg/day. No significant mortality or
other adverse effects were noted (no details given). The No-Observed-
Adverse-Effect-Level (NOAEL) for MH-Na in this study is 2,500 mg/kg
(the highest dose tested). The MH-DEA was fed at a level of 0.1%
(1,000 ppm) for 11 weeks. This is equivalent to a dose of 50 mg/kg/day
(Lehman, 1959). At the end of 11 weeks, 21/24 animals had died. The
author stated that after further investigation (details not gi^en),
it was concluded that the observed mortality was due to the DEA
component of the formulation.
0 The toxicity of maleic hydrazide in the diet for 1 year (320 to
360 days) was investigated in rats and dogs (Mukhorina, 1962). Rats
received oral doses of maleic hydrazide at 0.7, 1.5 or 3 mg/kg/day,
and a fourth group received 7 mg/kg MH-DEA. Dogs were administered
an oral dose of 0.7 mg/kg/day maleic hydrazide. Other details in
this translation on study design and conduct were not clear. Rats
exposed at the high dose had hyperemia and hemorrhage of the lungs,
myocardium, liver and brain, abnormal glucose-tolerance curves,
lowered liver glycogen, dystrophic changes in the liver, nephritis,
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Maleic Hydrazi.de August, 1987
-6-
interstitial pneumonia, loss of hair and significant reduction in
weight gain compared with the controls (at 4 months, controls had
gained 30%; those fed MH-DEA at 3 mg/kg/day had gained only 21%).
Dogs fed 0.7 mg/kg/day maleic hydrazide showed no significant adverse
changes, and it appears that for both the rat and the dog the level
of 0.7 mg/kg/day MH-DEA was a NOAEL.
0 Mukhorina (1962) also reported on a study done in mongrel mice given
0.7 mg/kg/day maleic hydrazide (purity not specified) in the diet for
320 to 360 days. No pathological changes were found. Based on these
data, the NOAEL for MH-DEA in the mouse is 0.7 mg/kg/day.
0 In a study by Food Research Labs (1954), MH-Na was fed in the diet
to rats (number not specified) from weaning for two years. Levels
of MH-Na (expressed as the free acid) were 0.0, 0.5, 1.0, 2.0 or 5.0%
(0, 5,000, 10,000, 20,000 or 50,000 ppm). Assuming that 1 ppm in the
diet of rats corresponds to 0.05 mg/kg/day (Lehman, 1959), this is
equivalent to doses of 0, 250, 500, 1,000 or 2,500 mg/kg/day. There
were no changes in blood or urine and no dose- or time-dependent
effects on longevity. Other study details were not presented.
Based on these observations, the NOAEL identified from this study
is 2,500 mg/kg/day (highest dose tested) for the rat.
0 In a similar study in dogs (Food Research Labs, 1954) animals were
fed doses of 0.0, 0.6, 1.2 or 2.4% maleic hydrazide (as MH-Na) in
the diet for 1 ye»r. Assuming 1% (10,000 ppm) in the diet of dogs
corresponds to 250 mg/kg/day (Lehman, 1959), this is equivalent to
a dose of 500 mg/kg/day. No effects attributable to exposure were
detected.
0 Van Der Heijden et al. (1981) fed technical maleic hydrazide, 99%
active ingredient (a.i.) and containing less than 1.5 mg hydrazine/kg
as an impurity to rats at dietary levels of 1.0 or 2.0% (10,000 or
20,000 ppm) for 28 months. Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to
doses of 500 or 1,000 mg/kg/day. These two levels of maleic hydrazide
in the diet caused proteinuria and increased the protein/creatinine
ratio in the urine of both sexes, although there were no detectable
histopathological changes in the kidney or the urinary tract. Based
on the effects on kidney function, the no-effect level was considered
by the authors to be lower than 1.0% maleic hjIrazide in the diet of
rats. On this basis, a Lowest-Obcerved-Adverse-Effeet-Level (LOAEL)
of 500 mg/kg is identified.
Reproductive Effects
0 In a two-generation reproduction study by Kehoe and MacKenzie (1983),
Charles River CD(SD)BR rats (15 males and 30 females/dose) were
administered the potassium salt of maleic hydrazide (K-MH) (purity
not specified) at dietary concentrations of 0, 1,000, 10,000 or
30,000 ppm. Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), these doses correspond to 0, 50, 500
and 1,500 mg/kg/day. No adverse effects on reproductive indices were
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Maleic Hydrazide August, 1987
-7-
observed at any dietary level. However, increased mortality was
observed in FI parents that received 30,000 ppm. Also at this dose
level, body weights were reduced in FQ parents during growth and
reproduction and in FT and F2 pups during lactation. Based on the
postnatal decrease in the body weight of pups, a reproductive NOAEL
of 10,000 ppm (500 mg/kg/day) is identified.
0 In a four-generation reproduction study in rats (Food Research Labs,
1954), animals were fed MH-Na (purity not specified) in the diet at
dose levels of 0.5, 1.0, 2.0 or 5.0% (5,000, 10,000, 20,000 or 50,000
ppm) (expressed in terms of free acid). Assuming 1 ppm in the diet
of rats corresponds to 0.05 mg/kg/day (Lehman, 1959), this is equivalent
to 250, 500, 1,000 or 2,500 mg/kg/day. The authors reported that
there were no effects on fertility, lactation or other reproductive
parameters, but no data from the study were presented for an adequate
assessment of these findings.
Developmental Effects
0 Khera et al. (1979) administered maleic hydrazide (97% purity) to
pregnant rats by gavage on days 6 to 15 of gestation at doses of 0,
400, 800, 1,200 or 1,600 mg/kg/day. Animals were sacrificed on day
22. No sign of toxicity or adverse effect on maternal weight gain
was observed at any dose level tested. Values for corpora lutea,
total implants, resorptions, dead fetuses, male/female ratio and
fetal weight were within the control range. The number of live fetuses
was decreased at the 1,200-mg/kg dose, but this was not statistically
significant and did not occur at the highest dose tested. Fetuses
examined for external, soft-tissue and skeletal abnormalities showed
no increase in frequency of abnormalities at any dose level tested.
Based on the results of this study, a NOAEL of 1,600 mg/kg/day (the
highest dose tested) is identified for maternal effects, fetotoxicity
and teratogenic effects.
0 Hansen et al. (1984) studied the teratogenic effects of MH-Na and
the monoethanolamine salt (MH-MEA) on fetuses from female rats exposed
by gavage to doses of 500, 1,500 or 3,000 mg/kg/day in the diet at
various stages of gestation. Replicate tests were run. No increased
frequency of gross, skeletal or visceral abnormalities was observed in
animals dosed by gavage on days 6 to 15 of gestation with 500 mg/kg/day
of either MH-Na or MH-MEA. to increased frequency of minor skeletal
variants (asymmetrical and bipartite sternebrae, wavy ribs, fused
ribs, rudiment of cervical rib, single bipartite or other variations
in thoracic vertebrae) was observed in animals receiving 1,500
(p <0.01) or 3,000 (p <0.1) mg/kg/day of MH-MEA on days 6 to 15, but
this was observed neither in animals exposed to 3,000 mg/kg/day for
days 1 to 21 of gestation nor in a replicate experiment. Similarly,
MH-Na produced marginal increases in minor skeletal variants in one
experiment at doses of 1,500 mg/kg/day for days 6 to 15 (p <0.1) or
3,000 mg/kg/day for days 1 to 21 (p <0.1), but this was not observed
in a replicate experiment. Rats dosed with 3,000 mg/kg/day MH-MEA in
the diet exhibited a significant decrease in maternal body weight and
in weight gain compared to the controls. This effect was not observed
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Maleic Hydrazide August, 1987
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when 3,000 mg/kg was given on days 1 to 21 by gavage, and there was
no significant effect on food intake. Exposure to 3,000 mg/kg in the
diet caused a significant increase in resorptions (p <0.001) and a
decrease in mean fetal weight (p <0.001). Similar but less pronounced
effects were observed when this dose was given by gavage. In addition,
postimplantation loss was increased significantly (p <0.01) in both
experiments. The authors theorized that the more severe effects
observed when the MH-MEA was fed in the diet (versus gavage) could be
due to an alteration in the palatability of the diet, resulting in
decreased food consumption. In contrast to the results with MH-MEA,
MH-Na had no adverse effects on the dams except for a reduction in
food consumption for days 1 to 6 in the group exposed from days 1 to
21 at 3,000 mg/kg. There were significant differences in body weight
of the pups (up to age 35 days) of dams administered MH-MEA by gavage
at 3,000 mg/kg/day from day 6 of gestation through day 21 of lactation;
a significant delay in the pups' startle response to an auditory
stimulus, significantly higher brain weight in both male and female
pups, and a delay in unfolding of the pinna were noted also. The
authors attributed the increase in relative brain weight to the lower
body weight. The delay in the startle response in MH-MEA dosed
offspring was considered the most significant effect, since it was
observed in both sexes, but the authors noted that it cannot be
explained. Based on these data, maternal, fetotoxic and teratogenic
NOAELs of 1,500, 1,500 and 500 mg/kg/day, respectively, were identified
for both MH-MEA and MH-Na.
0 Aldridge (1983, cited in U.S. EPA, 1985a) administered K-MH by gavage
at doses of 0, 100, 300 or 1,000 mg/kg/day to Dutch Belted rabbits
(16/dose) on days 7 through 27 of gestation. No signs of maternal
toxicity were reported, and the NOAEL for this effect is identified
as 1,000 mg/kg/day (the highest dose tested). Malformed scapulae
were observed in fetuses from the 300- and 1,000-mg/kg/day dose
groups. An evaluation of this study by the Office of Pesticide
Programs (U.S. EPA, 1985a) concluded that scapular malformations are
rare and considered to be a major skeletal defect. Historical data
for Dutch Belted rabbits from the testing laboratory (IRDC) indicated
that scapular anomalies were observed in only 1 of 1,586 fetuses
examined from 264 litters. Based on this information, a NOAEL of
1 00 mg/kg/day is identified for developmental effects.
Mutagenicity
0 The mutagenic activity of maleic hydrazide and its formulations has
been investigated in a number of laboratories. These studies are
complicated by the fact that hydrazine (a powerful mutagen) is a common
contaminant of these preparations, and N-nitrosoethanolamine (also a
mutagen) may be present in MH-DEA. Present data are inadequate to
determine with certainty whether any mutagenic activity of maleic
hydrazide is due to impurities and not the maleic hydrazide itself.
0 Tosk et al. (1979) reported that maleic hydrazide (purity not
specified), at levels of 5, 10 and 20 rag, was not mutagenic in
Salmonella typhimurium (TA 1530). However, two formulations (MH-30
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Maleic Hydrazide August, 1987
-9-
and Royal MH), at 50, 100 and 200 uL (undiluted), were highly mutagenic
in this system.
0 Moriya et al. (1983) reported that maleix hydrazide was not mutagenic
in five strains of j>. typhimurium.
0 Ercegovich and Rashid (1977) observed a weak mutagenic response with
maleic hydrazide (purity not specified) in five strains of £. typhimurium.
0 Shiau et al. (1980) reported that maleic hydrazide was mutagenic,
with and without activation, in several Bacillus subtilis strains.
0 Epstein et al. (1972) reported that maleic hydrazide (500 rag/kg) was
not mutagenic in a dominant-lethal assay in the mouse.
0 Nasrat (1965) reported a slight increase in the frequency of sex-
linked recessive lethals in the progeny of Drosophila melanogaster
males reared on food containing 0.4% maleic hydrazide.
0 Manna (1971) indicated that exposure to a 5% aqueous solution of
maleic hydrazide produced chromosomal aberrations in the bone marrow
of mice in a manner similar to that produced by x-rays and other
known mutagens.
0 Chaubey et al. (1978) reported that intraperitoneal injection of 100
or 200 mg/kg maleic hyerazide (purity not specified) did not affect
the incidence of bone marrow erythrocyte micronuclei or the ratio of
poly- to nonnochromatic erythrocytes in male Swiss mice.
0 Sabharwal and Lockhard (1980) reported that at concentrations above
100 ppm, maleic hydrazide induced dose-related increases in sister
chromated exchange (SCE) in human lymphocytes and V79 Chinese hamster
cells. Commercial formulations of maleic hydrazide (Royal MH and
MH-30) at the 250- and 500-mg/kg doses did not cause an increase in
micronucleated polychromatic erythrocytes in a mouse micronucleus test.
0 Stetka and Wolff (1976) reported that maleic hydrazide (11 and 112 mg/L;
purity not specified) caused no significant effect in an SCE assay.
0 Nishi et al. (1979) reported that maleic hydrazide (1,000 ug/L; purity
not specified), MH-DEA (20,000 ug/mL) and MH-K (20,000 ug/mL) produced
cytogenetic effects in Chinese hamster V79 cells in vitro.
0 Paschin (1981) reported that in the concentration range of 1,800 to
2,500 mg/L maleic hydrazide (purity not specified) was mutagenic for
the thymidine kinase locus of mouse lymphoma cells.
Carcinogenic!ty
0 The carcinogenicity of maleic hydrazide (purity not specified)
was evaluated in two hybrid strains of mice (C57BL/6 x AKR and
C57BL/6 x C3H/Anf) (Kotin et al., 1968; Innes et al., 1969). Beginning
at 7 days of age, mice were given maleic hydrazide at 1,000 mg/kg/day
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Maleic Hydrazide August. 1987
-10-
(suspended in 0.5% gelatin) by stomach tube. After 28 days of age,
they were given maleic hydrazide in the diet at 3,000 ppm for 18
months. Assuming that 1 ppm in the diet of mice corresponds to
0.15 mg/kg/day (Lehman, 1959), this is equivalent to a dose of
450 mg/kg/day. These were the maxiirum tolerated doses. No evidence
of increased tumor frequency was detected in gross or histologic
examination.
Barnes et al. (1957) fed maleic hydrazide at a level of 1% (10,000 ppm)
in the diet of rats and mice (10 to 15/sex/dose) for a total of 100
weeks. Assuming that 1 ppm in the diet corresponds to 0.05 mg/kg/day
in rats and 0.15 mg/kg/day in mice (Lehman, 1959), this is equivalent
to a dose of 500 mg/kg/day in rats and 1,500 mg/kg/day in mice.
A concurrent study was conducted in which the maleic hydrazide
(500 mg/kg/week, corresponding to 71 mg/kg/day) was injected subcu-
taneously (sc) for the same length of time. No increase in the
incidence of tumors was observed in animals exposed by either route
when compared with controls (data were pooled).
Cabral and Ponomarkov (1982) administered maleic hydrazide by gavage
in weekly doses of 510 mg/kg in 0.2 mL olive oil to male and female
C57BL/B6 mice for 120 weeks. Controls received 0.2 mL olive oil
alone, and a third group served as untreated controls. A simultaneous
study was conducted using sc injection as the route of administration.
There was no evidence of carcinogenicity in the study.
Van Der Heijden et al. (1981) fed maleic hydrazide (99% pure)
containing less than 1.5 mg hydrazine/kg as impurity to rats at
dietary levels of 1.0 or 2.0% (10,000 or 20,000 ppm) for 28 months.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 500 or 1,000 mg/kg/day.
Histological examination revealed no increase in the tumor incidence
in exposed animals compared with the control group.
In a study by Uniroyal Chemical (1971), mice were administered maleic
hydrazide (0.5% in water) by gavage twice weekly beginning at 2 months
of age (weight 15 to 18 g) for a total of 2 years. A parallel study
was conducted using sc administration. No carcinogenic effect was
reported, but specific details of the study were not presented.
Uniroyal Chemical (1971) reported a 2-year stjdy in Wistar-derived
rats in which MH-Na was included in the diet at levels of 0, 0.5, 1,0,
2.0 or 5.0% (0, 5,000, 10,000, 20,000 or 50,000 ppm). Assuming that
1 ppm in the diet of rats corresponds to 0.05 mg/kg/day (Lehman, 1959),
this is equivalent to doses of 0, 250, 500, 1,000 or 2,500 mg/kg/day.
Although no experimental details were presented, it was concluded
that the MH-Na resulted in no blood dyscrasias or tissue pathology,
and no indication of carcinogenic potential was detected.
Epstein and Mantel (1968) used random-bred infant Swiss mice (ICR/Ha)
to assess the carcinogenic effects of maleic hydrazide when admini-
stered during the neonatal period. The free acid form of maleic
hydrazide (containing 0.4% hydrazine impurity) was prepared as an
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Maleic Hydrazide August, 1987
-1 1-
aqueous solution of 5 mg/mL, or as a suspension in redistilled
tricaprylin at a concentration of 50 mg/mL. The mice were given
injections in the nape of the neck on days 1, 7, 14 and 21 following
birth. Six litters received the maleic hydrazide aqueous solution
(total dose: 3 mg), and 16 litters received the maleic hydrazide
suspension (total dose: 55 mg). One litter received one injection
of the suspension at a higher dose (100 mg/mL, total dose: 10 mg),
but this was lethal to all mice. A total of 16 litters served as
controls (treated with solvents alone). The experiment was terminated
between 49 and 51 weeks. The mice that received a total dose of
55 mg in the 3-week period had a high incidence of hepatomas: 65% of
26 male mice alive at 49 weeks, in contrast to solvent controls in
which hepatomas occurred in 8% of 48 male mice. The males that
received 3 mg total had an 18% incidence of hepatomas. In addition
to these lesions, hepatic "atypia" was observed in five males
(at 55 mg) and eight females, which the authors judged might be
preneoplastic. At the 3-mg level, one atypia was seen in each sex.
It was concluded that maleic hydrazide was highly carcinogenic in the
male mice. The authors also noted that since there was a complete
absence of multiple pulmonary adenomas and pulmonary carcinomas, it
was unlikely that the carcinogenicity of maleic hydrazide was due
to hydrazine (either present as trace contamination or formed by
metabolism), since hydrazine is a potent lung carcinogen in several
species of rats and mice (including CBA 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:
NQAEL 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
accordanca with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
Several studies (Tate, 1951; Mukhorina, 1962; Hansen et al., 1984)
indicate that the DEA ion is toxic and may contribute to the toxicity of the
MH-DEA salt. For this reason, studies involving MH-DEA have not been consid-
ered as candidates in calculating HA values for maleic hydrazide.
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Maleic Hydrazide August, 1987
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One-day Health Advisory
No information was found in the available literature that was suitable
for deriving a One-day HA value for maleic hydrazide. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child (10 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA
value.
Ten-day Health Advisory
The developmental toxicity study by Aldridge (1983, cited in U.S. EPA,
1985a) has been selected to serve as the basis for the Ten-day HA. In this
study, the potassium salt of maleic hydrazide (K-MH) was administered by
gavage at doses of 0, 100, 300 or 1,000 mg/kg/day to Dutch Belted rabbits
(16/dose) on days 7 through 27 of gestation. Malformed scapulae were observed
in fetuses from the 300- and 1,000-aig/kg/day dose groups. Although the
incidence of these malformations was not statistically significant and did
not occur in a dose-related fashion, malformed scapulae are a rare, major
skeletal defect. Additionally, historical data for this breed of rabbits
indicate that scapular anomalies were observed in only 1 of 1,586 fetuses
examined from 264 litters. For these reasons U.S. EPA (1985a) concluded that
the possibility of teratogenic activity at these dose levels cannot be ruled
out. The NOAEL for teratogenic effects is identified as 100 mg/kg/day.
Although a teratogenic response is clearly a reasonable basis upon which
to base an HA for an adult, there is some question about whether the Ten-day HA
for a 10-kg child can be based upon such a study. However, a teratogenic
study is of appropriate duration and does supply some information concerning
fetotoxicity. Since the fetus may be more sensitive to the chemical than
a 10-kg child and since a teratogenic study is of appropriate duration,
it is judged that, though possibly overly conservative, it is reasonable in
this case to base the Ten-day HA for a 10-kg child on a developmental toxicity
study.
Using a NOAHL of 100 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (100 mg/kg/day) (10 kg) = 10 mg/L (10,000 ug/L)
(100) (1 L/day)
where:
100 mg/kg/day = NOAEL, based on the absence of teratogenic effects
in rabbits exposed to K-MH by gavage on days 7 to 27
of gestation.
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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Maleic Hydrazide August, 1987
-13-
Longer-term Health Advisory
No studies were found that were adequate for calculation of Longer-
term HA values for maleic hydrazide. An 11-week feeding study in rats by
Tate (1951) identified a NOAEL of 2,500 mg/kg/day, and 2-year feeding
studies in rats and dogs by Food Research Laboratories (1954) identified
NQAEL values of 2,500 and 500 mg/kg/day, respectively. These studies have
not been selected because they provided too little experimental detail to be
suitable for calculation of an HA value. It is, therefore, recommended that
the Drining Water Equivalent Level (DWEL) of 17.5 mg/L, calculated below, be
used as a conservative estimate of the Longer-term HA for a 70-kg adult and
that the modified DWEL of 5 mg/L (adjusted for a 10-kg child) be used as a
conservative estimate of the Longer-term HA for a 10-kg child.
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 28-month feeding study in rats by Van Der Heijden et al. (1981) has
been selected to serve as the basis for the Lifetime HA value for mali-ic
hydrazide. Based on proteinuria (in the absence of visible histological
effects in kidney), a LOAEL of 500 mg/kg/day was identified. This is a
conservative selection, since 2-year feeding studies in dogs and rats by Food
Research Laboratories (1954) identified NOAEL values of 500 and 2,500 ng/kg/day,
respectively; those studies were not selected, however, because few data or
details were provided.
Using the LOAEL identified by Van Der Heijden et al. (1981), the Lifetime
HA is calculated as follows:
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Maleic Hydrazide August, 1987
-14-
Step 1: Determination of the Reference Dose (RfD)
RfD = (500 mg/kg/day) = Oo5 mgAg/day
(1,000)
where:
500 mg/kg/day = LOAEL, based on decreased ammo acid resorption in
kidney of rats exposed to maleic hydrazide in the
diet for 28 months.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a. LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.5 mg/kg/day) (70 kg) = 17.5 mg/L (17,500 ug/L)
(2 L/day)
where:
0.5 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 = (17.5 mg/L) (20%) =3.5 mg/L (3,500 ug/L)
where:
17.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic activity was detected in five studies in
which maleic hydrazide was administered oral.1/ to mice or rats for
periods from 18 to more than 2 years (Kotin et al., 1968; Innes et al.,
1969; Barnes et al., 1957; Cabral and Ponomarkov, 1982; Van Der Heijden
et al., 1981; Uniroyal Chemical, 1971). Increased incidence of
hepatomas has been reported in mice exposed by sc injection during
the first 3 weeks of life (Epstein and Mantel, 1968).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of maleic hydrazide.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), maleic hydrazide may be classified
in Group D: not classified. This group is used for substances with
inadequate human or animal evidence of carcinogenicity.
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Maleic Hydrazide August, 1987
-15-
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA (1985b) has established residue tolerances for maleic
hydrazide in or on raw agricultural commodities that range from 15.0
to 50.0 ppm.
VII. ANALYTICAL METHODS
0 There is no standardized method for the determination of maleic
hydrazide in water samples. A procedure has been reported for the
estimation of maleic hydrazide residues on various foods (U.S. FDA,
1975). In this procedure, the sample is boiled in an alkaline solution
to drive off volatile basic interferences. Distillation with zinc and
a nitrogen sweep expel hydrazine liberated from maleic hydrazide.
Hydrazine is reacted in acid solution with p-dimethylaminobenzaldehyde
to form a yellow compound that is measured spectrophotometrically.
VIII. TREATMENT TECHNOLOGIES
0 Currently available treatment technologies have not been tested for
their effectiveness in removing maleic hydrazide from drinking water.
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Maleic Hydrazide August, 1987
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IX. REFERENCES
Aldridge, 0.* 1983. Teratology study in rabbits with potassium salt of maleic
hydrazide. Unpublished report prepared by International Research and
Development Corporation for Uniroyal Chemical Company. Accession No.
250523. Cited in: U.S. EPA. 1985. U.S. Environmental Protection
Agency. Memorandum dated 3/7/85 from G. Ghali to R. Taylor concerning
EPA Reg. Numbers 400-84, 400-94 and 400-165; Maleic Hydrazide, K-Salt.
Barnes, J.M., P.N. Magee, E. Boyland, A. Haddow, R.D. Passey, W.S. Builough,
C.N.D. Cruickshank, M.H. Salaman and R.T. Williams. 1957. The non-
toxicity of maleic hydrazide for mammalian tissues. Nature. 180:62-64.
Cabral, J.R.P., and V. Ponomarkov. 1982. Carcinogenicity study of the
pesticide maleic hydrazide in mice. Toxicology. 24:169-173.
Chaubey, R.C., B.R. Kavi, P.S. Chauhan and K. Sundaram. 1978. The effect of
hycanthone and maleic hydrazide on the frequency of micronuclei in the
bone-marrow erythrocytes of mice. Mutat. Res. 57:187-191.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Epstein, S.S., E. Arnold, J. Andrea, W. Bass and Y. Bishop. 1972. Detection
of chemical mutagens by the dominant lethal assay in the mouse. Toxicol.
Appl. Pharmacol. 23:288-325.
Epstein, S.S., and N. Mantel. 1968. Hepatocarcinogenicity of the herbicide
maleic hydrazide following parenteral administration to infant Swiss mice.
Intl. J. Cancer. 3:325-335.
Ercegovich, C.D., and K. A. Rashid. 1977. Mutagenesis induced in mutant
strains of Salmonella typhimurium by pesticides. (Abstract of Paper)
Am. Chem. Soc. 174:Pest 43.
Food Research Labs, Inc.* 1954. Chronic toxicity studies with sodium maleic
hydrazide. Unpublished report. HRID 00112753.
Hansen, E., 0. Meyer and E. Kristiansen. 1984. Assessment of teratological
effect and developmental effect of maleic hydrazide salts in rats.
Bull. Environ. Contain. Toxicol. 33:184-192.
Innes, J.R.M, B.M. Ulland, M.G. Valerio, L. Petrucelli, L. Fishbein, E.R. Hart,
A.J. Pallotta, R.R. Bates, H.L. Falk, J.J. Gart, M. Klein, I. Mitchall
and J. Peters. 1969. Bioassay of pesticides and industrial chemicals
for tumorigenicity in mice: A preliminary note. J. Natl. Cancer Inst.
42:1101-111 4.
Kehoe, D.F., and K.M. MacKenzie.* 1983. Two-generation reproduction study
with KMH in rats. Study No. 81065 prepared by Hazleton Raltech, Inc.
for Uniroyal Chemical Company. Accession No. 250522. Cited in: U.S. EPA
1985. U.S. Environmental Protection Agency. Memorandum dated 3/7/85
from G. Ghali to R. Taylor concerning EPA Reg. Numbers 400-84, 400-94
and 400-165; Maleic Hydrazide, K-Salt.
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Maleic Hydrazide August, 1987
-17-
Kennedy G., and M.L. Keplinger.* 1971. Placental and milk transfer of maleic
hydrazide in albino rats. Unpublished report. MRID 00112778.
Khera, K.S., C. Whalen, C. Trivett and G. Angers. 1979. Teratologic assess-
ment of maleic hydrazide and daminozide, and formulations of ethoxyquin,
thiabendazole and naled in rats. J. Environ. Sci. Health (B). 14:563-577.
Kotin, P., H. Falk and A.J. Pallotta.* 1968. Evaluation of carcinogenetic,
teratogenic, and mutagenic activities of selected pesticides and industrial
chemicals. Unpublished report. MRID 0017801.
Lehman, A.J. 1951. Chemicals in food: A report to the Association of Food
and Drug Officials. Assoc. Food Drug Off. U.S.Q. Bull. T5:122. Cited
in Ponnampalam R., N.I. Mondy and J.G. Babish. 1983. A review of
environmental and health risks of maleic hydrazide. Regul. Toxicol.
Pharmacol. 3:38-47.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Manna, G.K. 1971. Bone marrow chromosome aberrations in mice induced by
physical, chemical and living mutagens. J. Cytol. Genet. (India) Congr.
Suppl. 144-150.
Mays, D.L., G.S. Born, J.E. Christian and B.J. Liska. 1968. Fate of
C14-maleic hydrazide in rats. J. Agric. Food Chem. 16:356-357. Cited in
Swietlinska Z and J. Zuk. 1978. Cytotoxic effects of maleic hydrazide.
Mutat. Res. 55:15-30.
Meister, R. , ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
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.
Mukhorina, K.V.* 1962. Action of maleic hydrazide on animal organisms.
Unpublished report (translation from Russian). MRID 00106969.
Nasrat, G.E. 1965. Maleic hydrazide, a chemical mutagen in Drosophila
melanogaster. Nature. 207:439.
Nishi, Y., M. Mori and N. Inui. 1979. Chromosomal aberrations induced by
maleic hydrazide and related compounds in Chinese hamster cells in vitro.
Mutat. Res. 67:249-257.
Paschin, Y.V. 1981. Mutagenicity of maleic acid hydrazide for the TK locus
of mouse lymphoma cells. Mutat. Res. 91:359-362.
Reagan E. and P. Becci.* 1982. Acute oral LDso in rats of Royal-DRI-60-DG.
Food and Drug Research Labs. Unpublished report. MRID 00110459.
Registrant Confidential Business Informathion data.* Complete citation not
available.
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Maleic Hydrazide August, 1987
-18-
Sabharval, P.S., and J.M. Lockard. 1980. Evaluation of the genetic toxicity
of maleic hydrazide and its commercial formulations by sister chromatid
exchange and micronucleus bioassays. In Vitro. 16(3):205.
Shapiro, R.* 1977a. Acute oral toxicity: Report no. T-235. Unpublished
report. MRID 00079657.
Shapiro, R.* 1977b. Primary skin irritation: Report no. T-212. Unpublished
report. MRID 00079660.
Shapiro, R.* 1977c. Eye irritation: Report no. T-220. Unpublished report.
MRID 00079661.
Shapiro, R.* 1977d. Acute dermal toxicity: Report no. T-242. Unpublished
report. MRID 00079658.
Shiau, S.Y., R.A. Huff, B.C. Wells and I.C. Felkner. 1980. Mutagenicity and
DNA-damaging activity for several pesticides with Bacillus subtilis
mutants. Mutat. Res. 71:169-179.
Stetka, D.G., and S. Wolff. 1976. Sister chromatid exchange as an assay
for genetic damage induced by mutagen-carcinogens. II. In vitro test
for compounds requiring metabolic activation. Mutat. Res. 41:343-350.
Tate, H.Do* 1951. Progress report on mammalian toxicity studies with maleic
hydrazide. Unpublished report. MRID 00106972.
TDB. 1985. Toxicity Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
Tosk, Jo, I. Schmeltz and D. Hoffmann. 1979. Hydrazines as mutagens in a
histidine-requiring auxotroph of Salmonella typhimurium. Mutat. Res.
66:247-252.
Uniroyal Chemical Co., Bethany, Connecticut.* 1971. Summary of toxicity
studies on maleic hydrazide: Acute oral toxicity in rats and rabbits.
Unpublished report. MRID 00087385.
Uniroyal Chemical Co., Bethany, Connecticut.* 1977. Results from acute
toxicology tests run with Royal MH-30(R) and MH Technical (R). Unpub-
lished report. MRID 00079651.
U.S. EPA. 1985a.* U.S. Environmental Protection Agency. Memorandum dated
3/7/85 from G. Ghali to R. Taylor concerning EPA Reg. Numbers 400-84,
400-94 and 400-165; Maleic Hydrazide, K-Salt.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.175. July 1, p. 277.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24,
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Maleic Hydrazide August, 1987
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U.S. FDA. 1975. U.S. Food and Drug Administration. Pesticide analytical
manual. Vol. II. Washington, DC.
Van Der Heijden, C.A., E.M. Den Tonkelaar, J.M. Garbis-Berkvens and G.J. Van
Esch. 1981. Maleic hydrazide, carcinogenicity study in rats. Toxicology,
19:139-150.
WSSA. 1983. Weed Science Society of America. Herbicide handbook, 5th ed.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
MCPA
(4-Chloro-2-Methylphenoxy)-Acetic Acid
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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 "hat
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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MCPA August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 94-74-6
Structural Formula CHj
•OCH2COOH
(4-Chloro-2-methylphenoxy)-acetic acid
Synonyms
0 MCPA; MCP; Agroxone; Hormotuho; Metaxon.
Uses
0 MCPA is a hormone-type herbicide used to control annual and perennial
weeds in cereals, grassland and turf (Hayes, 1982).
Properties (CHEMLAB, 1985; Meister, 1983)
Chemical Formula
Molecular Weight 200.63
Physical State (25°C) Light brown solid
Boiling Point
Melting Point 118 to 119°C
Vapor Pressure (25eC)
Density (25°C) 1.56
Water Solubility 825 mg/L (room temperature)
Log Octanol/Water Partition 2.07 (calculated)
Coefficient
Taste Threshold
Odor Threshold —
Conversion Factor —
Occurrence
MCPA has been found in 4 of 1 2 surface water samples analyzed and in
none of 99 ground water samples (STORET, 1987). Samples were collected
at 8 surface water locations and 97 ground water locations. MCPA was
found only in California. The 85th percentile of all nonzero samples
was 0.54 ug/L in surface water, and the maximum concentration found
was 0.54 ug/L.
Environmental Fate
MCPA is not hydrolyzed at pH 7 and 34 to 35°C (Soderquist and Crosby,
1974, 1975). MCPA in aqueous solution ( pH 8.3) has a photolytic
half-life of 20 to 24 days in sunlight. With fluorescent light, MCPA
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MCPA August, 1987
-3-
in aqueous solution (pH 9.8) produced three minor (less than 10%)
photolysis products: 4-chloro-2-methyl-phenol, 4-chloro-2-formylphenol
and ^-cresol in 71 hours (Soderquist and Crosby, 1974, 1975).
0 MCPA is degraded more rapidly (1 day) in soils containing less than
10* organic matter than in soil containing higher levels (3 to 9 days)
(Torstensson, 1975). This may be due to adsorption to the soil
organic matter. MCPA, when applied a second time to soil, is degraded
twice as fast (6 to 12 days) as it is after one application (15 to 28
days). Persistence does not depend greatly upon the soil type (Loos
et al., 1979).
0 Unlabeled MCPA in rice paddy water under dark conditions is totally
degraded by aquatic microorganisms in 13 days (Soderquist and Crosby,
1974, 1975).
0 MCPA would be expected to leach readily in most soils. Phytotoxic
levels of MCPA leached 30 cm in a sandy soil column eluted with 50 cm
of water (Herzel and Schmidt, 1979). Using soil thin-layer chromato-
graphic techniques, MCPA was mobile (Rf 0.6 to 1.0) in calcium
montmorillonite clay (Helling, 1971) and in sandy loam, silt loam,
and silty clay loam soils (Helling and Turner, 1968). Mobility
increases as organic matter content decreases, possibly due to
adsorption of MCPA to this soil component.
0 MCPA does not volatilize from aqueous solution (pH 7.0) heated for
13 days at 34 to 35°C (Soderquist and Crosby, 1974, 1975).
0 Using bioassays, MCPA appears to dissipate fairly rapidly (3 to 7
weeks) from soil treated with levels of 0.75 to 1.5 ppm for 6 to 19
previous years (DeRose, 1946; Fryer and Kirkland, 1970; Torstensson
et al., 1975). An initial application of MCPA may require up to 20
weeks for complete dissipation. In another study, MCPA dissipated
to nondetectable levels from sandy and silt loam soils in 30 to 60
days (Suzuki, 1977).
0 In the aquatic environment, MCPA disipates rapidly (14 to 32 days)
in water, but residue levels in the flooded soil remain unchanged
(Soderquist and Crosby, 1974, 1975; Sokolov et al., 1974, 1975).
A common metabolite, 5-chloro-£-cresol, is formed at low levels
(1.3% or less) within 1 day of treatment. Frank et al. (1979) detected
MCPA residues (1.1 to 1000 ppb) in 2 of 237 wells in Ontario, Canada,
between 1969 and 1978.
0 In the forest ecosystem, MCPA remains in soil (0 to 3 cm) and leaf
litter at 0.7 and 32 ppm, respectively, 10 months after application
at 2.5 kg active ingredient per hectare (ai/ha) (Eronen et al.,
1979). MCPA residues in moss decline to 7% of the initial level
within 40 days. Residues in soil (3 to 15 cm deep) are not detectable
after 40 days.
0 MCPA has not been found in U.S. ground water.
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MCPA August, 1987
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III. PHARMACOKINETICS
Absorption
0 No information on the absorption of MCPA was found in the available
literature.
Distribution
0 Elo and Yltalo (1979) treated rats with 8 mg of 1*C-MCPA [98% active
ingredient (a.i.)] intravenously and measured the distribution of
radioactivity in nine tissues 1.5 hours after treatment. Highest
levels were found in plasma, kidney, lung, liver and heart with
lesser amounts found in brain/cerebrospinal fluid (CSF), testis and
muscle. Prior treatment of rats with MCPA (intravenous injections
of 25 to 500 mg/kg 3 hours before administration of radiolabeled
compound or chronic exposure to 500 or 2,500 mg/L in drinking water)
lead to decreased levels of 14C-MCPA in the plasma and kidney and
increased levels in brain/CSF.
0 Elo and Yltalo (1977) treated rats with 8 mg of 14C-MCPA (purity not
specified) intravenously and measured the distribution of radioactivity
in brain, CSF, muscle, liver and kidney 1.5 to 120 hours after treat-
ment. Prior treatment of rats with MCPA (subcutaneous injections of
250 or 500 mg/kg) caused a decrease in the amount of radioactivity
found in the plasma. Increased levels were found in other tissues
with the largest increases found in the CSF (39- to 67-fold) and
brain (11- to 18-fold).
Metabolism
0 MCPA is metabolized by the liver. Stimulation of microsomal oxidation
by phenobarbital increases the rate of MCPA breakdown (Buslovich et al.,
1979). Gaunt and Evans (1961) found that 5-chloro-methyl-catechol is
one of the metabolites of MCPA (Hattula et al., 1979).
Excretion
In studies by Fjeldstad and Wannag (1977), four healthy human volun-
teers each ingested a dose of 5 mg of MCPA (purity not specified).
Approximately 50% (2.5 mg) of the dose was detected in the urine
within several days. Urinary levels were not detectable on the fifth
day following exposure.
Rats treated orally with MCPA (purity not specified) excreted nearly
all of the MCPA during the first 24 hours after intake (90% in urine
and 7% in feces) (Elo, 1976).
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MCPA August, 1987
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IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 Palva et al. (1975) reported- one case of MCPA (purity not specified)
exposure (dose and duration not specified) in a farmworker involved in
spraying operations. Exposure resulted in reversible aplastic anemia
as well as muscular weakness, hemorrhagic gastritis and slight signs
of liver damage that were later followed by pancytopenia of all of
the myeloid cell lines. In a followup study in the exposed farmer,
Timonen and Palva (1980) reported the occurrence of acute myelomono-
cytic leukemia.
Long-term Exposure
0 No information on the human health effects of chronic exposure to
MCPA was found in the available literature.
Animals
Short-term Exposure
0 Reported acute oral LD50 values for MCPA (purity not specified) in
mice and rats are 550 mg/kg and 700 mg/kg, respectively (RTECS, 1985).
0 Gurd et al. (1965) reported an acute oral LD50 value for MCPA (purity
not specified) of 560 mg/kg in mice.
0 Elo et al. (1982) showed that MCPA (sodium salt; 99% a.i.) causes a
selective damage of the blood-brain barrier. These authors observed
that the penetration of intravenous tracer molecules such as 14c-MCPA,
14C-PABA, 14C-sucrose, 14C-antipyrine and iodinated human albumin
(125i_HA) in the brain and CSF of MCPA-intoxicated rats (200 to
500 mg/kg, sc) was increased compared to controls. The tissue-plasma
ratios of 14C-sucrose, 14C-antipyrine and 125I-HA treated rats were
also increased in the brain and CSF of intoxicated animals, but the
increases were less pronounced than those of 14C-MCPA or 14C-PABA.
0 In oral studies by Vainio et al. (1983), Wistar rats administered an
ester of MCPA (purity not specified) (0, 100, 150 or 200 mg/kg/day),
5 days per week for 2 weeks, showed hypolipidemia and peroxisome
proliferation in the liver. A Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 100 mg/kg was identified.
Dermal/Ocular Effects
0 Raltech (1979) reported acute dermal LD50 values for MCPA (purity not
specified) in rabbits of 4.8 g/kg for males and 3.4 g/kg for females.
0 In acute dermal studies conducted by Verschuuren et al. (1975), an
aqueous paste of MCPA (80.6% a.i.) (0.5 g) was applied to the abraded
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MCPA August, 1987
-6-
skin of five chinchilla rabbits. Slight erythema resulted; the
skin became sclerotic after 5 to 6 days and healed by 12 days.
0 In subacute dermal studies, Verschuuren et al. (1975) applied an
aqueous paste of MCPA (80.6% active ingredient; 0, 0.5, 1.0 or 2.0 g)
five times weekly for 3 weeks to the shaved skin of rabbits. Slight
to moderate erythema occurred at all dose levels, and elasticity of
the skin was decreased. The effects subsided at 2 weeks post-treatment.
Weight loss was observed at all dose levels. High mortality and
histopathological alterations were observed in the liver, kidneys,
spleen and thymus at the 1.0- and 2.0-g dose levels.
Long-term Exposure
0 Verschuuren et al. (1975) administered MCPA (80.6% a.i.) in the diet
for 90 days to SPF weanling rats (10/sex/dose) at levels of 0, 50,
400 or 3,200 ppm. Assuming that 1 ppm in the diet of rats is equiva-
lent to 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of
about 0, 2.5, 20 or 160 mg/kg/day. Following treatment, growth, food
intake, mortality, hematology, blood and liver chemistry, organ
weights and histopathology were measured. No compound-related effects
were reported for any of these parameters except for growth retard-
ation and elevated relative kidney weights at 400 ppm (20 mg/kg/day)
or moree A No-Observed-Adverse-Effect-Level (NOAEL) of 50 ppm
(2.5 mg/kg/day) and a LOAEL of 400 ppm (20 mg/kg/day) were identified.
0 Holsing and Kundzin (1970) administered MCPA (considered to be 100%
a.i.) in the diet of rats (10/sex/dose) for 3 months. Doses were
reported as 0, 4, 8 or 16 mg/kg/day; the concentration in the diet
was not specified. Following treatment, no compound-related effects
were observed in the physical appearance, behavior, growth, food
consumption, survival, clinical chemistry, organ weights, organ-to-
body weight ratios, gross pathology or histopathology at any dose
tested, except for increases in kidney weight in males at 16 mg/kg/day.
A NOAEL of 8 mg/kg/day and a LOAEL of 16 mg/kg/day were identified
by this study.
0 Holsing and Kundzin (1968) administered oral doses of MCPA to rats at
dose levels of 0, 25, 50, and 100 mg/kg/day for 13 weeks. Cytopatho-
logical changes in the liver and kidneys were observed at all doses.
Kidney effects included focal hyperplasia of thr eptithelial lining,
interstitial nephritis, tubular dilation and/or hypertrophy. A LOAEL
of 25 mg/kg/day (the lowest dose tested) is identified by this study.
0 Reuzel and Hendriksen (1980) administered MCPA (94% a.i.) in feed to
dogs in two separate 13-week studies. Dosing regimens of 0, 3, 12 or
48 mg/kg/day, and 0, 0.3, 1 or 12 mg/kg/day, respectively, were
employed. Decreased kidney and liver function, characterized by
increases in blood urea, SGPT and creatinine were observed at doses
as low as 3 mg/kg/day. Low prostatic weight and mucopurulent conjunc-
tivitis were observed at higher doses. A NOAEL of 1 mg/kg/day and a
LOAEL of 3 mg/kg/day were identified by these studies.
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August, 1987
-7-
0 Hellwig (1986) administered oral doses of MCPA (95% a.i.) to dogs at
doses of 0, 6, 30, or 150 ppm for 1 year. Assuming that 1 ppm in the
diet of dogs is equivalent to 0.025 rag/kg (Lehman, 1959), this corre-
sponds to doses of 0, 0.15, 0.75 or 1.5 mg/Jcg/day. Renal toxicity
was observed at the two highest doses and was characterized by elevated
serum levels of creatinine, urea and potassium, coloration of the
Icidneys and increased storage of pigment in the renal tubules. A
NOAEL of 0.15 mg/kg/day and a LOAEL of 0.75 mg/kg/day were identified
by this study.
0 Holsing (1968) administered oral doses of MCPA (considered to be
100% a.i.) (0, 25, 50 or 75 mg/kg/day) to beagle dogs (three/sex/dose)
for 13 weeks. Histopathological changes and alterations in various
hematologic and biochemical parameters indicative of bone marrow,
liver and kidney damage were observed at all dose levels. The
hematological findings included decreased hematocrit, hemoglobin and
erythrocyte counts. Several dogs had elevated blood urea nitrogen,
serum glutamic-pyruvic transaminase, serum-oxaloacetic transaminase,
alkaline phosphatase and serum bilirubin. Histopathological alterations
were seen in the liver, kidney, lymph nodes, testes, prostate and
bone marrow. All dogs of all three groups had various degrees of
hepatic, renal and bone marrow injury. A LOAEL of 25 mg/kg/day (the
lowest dose tested) was identified.
0 Gurd et al. (1965) administered technical MCPA (purity not specified)
in the feed to rats (five/sex/dose) for 7 months at dose levels of 0,
100, 400, 1,000 or 2,500 ppm. Assuming that 1 ppm in the diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), this corresponds
to doses of 0, 5, 20, 50 or 125 mg/kg/day. Following treatment, there
was a marked decrease in body weight gain at 1,000 ppm (50 mg/kg/day)
or 2,500 ppm (125 mg/kg/day), and some deaths occurred at 2,500 ppm
(125 mg/kg/day). At 400 ppm (20 mg/kg/day) or greater, there was a
reduction in numbers of red blood cells, hemoglobin content and
hematocrit. Relative kidney weights were increased at 100 ppm
(5 mg/kg/day), but no effects on body weight were evident. No
histopathological changes were reported at any dose level tested.
A LOAEL of 5 mg/kg/day (the lowest dose tested) was identified.
Reproductive Effects
0 No effects on reproduction were found in rats exposed to doses of
0, 50, 150, or 450 ppm MCPA (95* a.i.) in the diet over a period of
two generations (MacKenzie, 1986). Assuming that 1 ppm in the diet
of rats corresponds to 0.05 mg/kg/day (Lehman, 1959), this corresponds
to doses of 0, 2.5, 7.5 or 15 mg/kg/day. Body weight depression was
observed in the F-\ and F2 generations at the two highest doses. A
NOAEL of 15 mg/kg/day was identified for reproductive function, and
a NOAEL of 2.5 mg/kg/day was identified for fetoxtoxicity (depressed
weight gain).
Developmental Effects
0 Irvine et al. (1980) administered MCPA (purity not specified) (0, 5,
12, 30 or 75 mg/kg/day) by gavage to rabbits (15 to 18/dose) on days
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MCPA August, 1987
-8-
6 to 18 of gestation. No fetotoxicity or teratogenicity was observed
at any dose level tested. Body weights of the does were markedly
reduced in the 75 mg/kg/day dosage group. A fetal NOAEL of 75 mg/Jcg/day
and a maternal NOAEL of 30 mg/kg/day were identified.
0 Irvine (1980) administered MCPA (purity not specified) (0, 20, 50 or
125 mg/kg/day) by gavage to pregnant CD rats (16 to 38/dose) on days
6 to 15 of gestation. No maternal or fetal toxicity or teratogenic
effects were observed. A NOAEL of 125 mg/kg/day (the highest dose
tested) was identified.
0 Palmer and Love11 (1971) administered oral doses of MCPA (75% a.i.;
0, 5, 25 or 100 mg/kg/day of the active ingredient) to mice (20/dose)
on days 6 to 15 of gestation. Dams were monitored for pregnancy rate,
body weight, and gross toxicity; no significant effects were observed.
At 100 mg/kg/day, fetal weights were significantly reduced and there
was delayed skeletal ossification. A NOAEL of 25 mg/kg/day and a
LOAEL of 100 mg/kg/day based on fetal weights were identified.
Mutagenicity
0 Moriya et al. (1983) reported that MCPA (purity not specified) (5,000
ug/plate) did not produce mutagenic activity in Salmonella typhimurium
(TA 100, TA 98, TA 1535, TA 1537, TA 1538) and in Escherichia coli
(WP2 her) either with or without metabolic activation.
0 In studies conducted by Magnusson et al. (1977), there were no
effects on chromosome disjunction, loss or exchange in Drosophila
fed MCPA (250 or 500 ppm).
0 In studies by Linnainmaa (1984), no increases were observed in the
frequency of sister chromatid exchange (SCE) in blood lymphocytes
from rats intragastrically administered MCPA (purity not specified)
at 100 mg/kg/day for 2 weeks. A slight increase in SCE was observed
in bone marrow cells from Chinese hamsters given daily oral doses of
100 mg/kg for 2 weeks. In Chinese hamster ovarian cell cultures,
SCE was slightly increased following treatment with MCPA (10-5, iQ-4,
10~3M, 1 hour) with and without activation.
Carcinogenicity
0 No information on the potential carcinogenicity of MCPA was found in
the available literature. However, MCPA stimulates liver peroxisomal
proliferation, which has been implicated in carcinogenicity (Vainio
et al., 1983).
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:
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MCPA August, 1987
-9-
(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 mg/Jg bw/day.
BW = assumed body weight of a child (10 Jg) or
an adult (70 Jg ) .
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
No information was found in the available literature that was suitable
for determination of the One-day HA value for MCPA. It is therefore recom-
mended that the Longer-term HA value for a 10- Jq child (0.1 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
No information was found in the available literature that was suitable
to serve as the basis for determining the Ten-day HA value for MCPA. Several
reproductive/teratology studies have been performed in which rats or rabbits
have been given oral doses of MCPA for acute duration (Irvine, 1980; Irvine
et al., 1980; Palmer and Lovell, 1971; MacKenzie, 1986). The only signs of
maternal toxici-ty observed in these studies was a reduction in body weight in
rats exposed to 75 mg/Jq (Irvine, 1980). Estimates of maternal NOAELs range
from 30 to 125 mg/Jq/day (Irvine, 1980; Irvine et al, 1980). Fetotoxicity
has been observed at dose levels as low as 7.5 mg/Jg/day (MacKenzie, 1986).
The toxicity of MCPA from acute exposure has not been well characterized. It
is therefore recommended that the Longer-term HA value for a 10- kj child of
(0.1 mg/L, calculated below) be used at this time as a conservative estimate
of the Ten-day HA value.
Longer-term Health Advisory
Evidence of renal dysfunction has been observed in both 13-week (Reuzel and
Hendriteen, 1980; Holsing, 1968) and 1-year (Hellwig, 1986) feeding studies in
beagle dogs and serves as the basis for the Longer-term HA. In subchronic studies
changes in blood urea and creatinine levels have been observed at doses of 25
mg/Jg/day iHolsing, 1968) and 3 mg/Jg/day (Reuzel and HendriJsen, 1986). Renal
toxicity is not unique to dogs and has been observed in rats after 90-day
exposure at dose levels of 20 mg/Jg/day (Verschuuren et al., 1975) and 25
mg/Jg/day (Holsing, 1968). The rat and dog may have similar sensitivities; a
conservative estimate of the NOAEL was obtained from the studies described by
Reuzel and Hendri Jsen (1980). In these studies, oral doses of 0, 3, 12 or 48
mg/Jg/day, and 0, 0.3, 1 or 12 mg/Jg/day, respectively, were administered to
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MCPA August, 1987
-10-
dogs for 13 wee Is. Increases in blood urea, SGPT and creatinine levels were
observed at dose levels as low as 3 mg/)q/day; low prostatic weight and
mucopurulent conjunctivitis were observed at higher dose levels. A NOAEL of
1 ng/)q/day was identified by these studies.
Using a NOA^L of 1 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (1.0 mg/kg/day) (10 lg) = 0.10 mg/L (100 ug/L)
(100) (1 L/day)
where:
1.0 mg/)q/day = NOAEL, based on the absence of renal effects in dogs
exposed to MCPA in the diet for 90 days.
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 a 70-kg adult is calculated as follows:
Longer-term HA = (1.0 ing/kg/day) (70 kg) = Oi35 mg/L (350 ug/L)
(100) (2 L/day)
where:
1.0 mg/kg/day = NOAEL, based on the absence of renal effects in dogs
exposed to MCPA in the diet for 90 days.
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
i
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., drink} ri
water) lifetime exposure level, assuming 100% exposure from that medium, at
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MCPA August, 1987
-11-
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 chronic toxicity study in dogs (Hellwig, 1986) has been selected to
serve as the basis for the determination of the Lifetime HA. Beagle,dogs were
exposed to 0, 6, 30 and 150 ppm (0.15, 0.75, or 3.75 mg/kg/day) for 1 year.
Renal toxicity was observed at the two highest doses and was characterized by
elevated serum levels of creatinine, urea and potassium, coloration of the
kidneys and increased storage of pigment in the renal tubules. A NOAEL of
0.15 mg/kg/day was identified, which is supported by the findings from
subchronic feeding studies. From 90-day feeding studies, NOAELs of 1 mg/kg/day
and 2.5 mg/kg/day have been identified for dogs (Reuzel and Hendriksen, 1980)
and rats (Verschuuren et al., 1975), based on the absence of effects on the
kidney seen at higher doses. In a 7-month feeding study, Gurd (1965) observed
increased kidney weight in rats exposed to doses as low as 5.0 mg/kg/day, the
lowest dose tested.
Using a NOAEL of 0.15 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.15 mg/kg/day) „ 0.0005 mg/kg/day
(100) (3)
where:
0.15 mg/kg/day = NOAEL, based on the absence of kidney effects in
dogs exposed to MCPA in the diet for 1 year.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = additional uncertainty factor, chosen in accordance
with U.S. EPA Office of Pesticide Programs (OPP)
policy to account for the incomplete database on
chronic toxicity.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0005 mg/kg/day) (70 kg) = 0>018 mg/L {18 ug/L)
(2 L/day)
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MCPA August, 1987
-12-
where:
0.0005 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 = (0.018 mg/L) (20%) = 0.0036 mg/L (3.6 ug/L)
where:
•
0.018 = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No studies on the carcinogenic potential of MCPA were found in the
available literature.
0 The International Agency for Research on Cancer (IARC, 1983) concluded
that the potential carcinogenicity of MCPA in both humans and laboratory
animals was indeterminate.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), MCPA may be classified in Group D:
not classified. This category is used for substances with inadequate
animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Academy of Sciences has recommended an ADI of 0.00125
mg/Jcg/day and a Suggested-No-Adverse-Response-Level (SNARL) of
0.009 mg/L, based on a LOAEL of 1.25 mg/kg/day in a 90-day study in
rats (NAS, 1977).
0 Residue tolerances have been established for MCPA at 0.1 ppm in milk
and meat. Feed and forage residue tolerances range from 0.1 to
300 ppm (U.S. EPA, 1985a).
VII. ANALYTICAL METHODS
0 Analysis of MCPA is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated acid pesticides in water
samples (U.S. EPA, 1985b). In this method, approximately 1 liter of
sample is acidified. The compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolized with
potassium hydroxide, and extraneous organic material is removed by
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August, 1987
-13-
a solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane as the derivatizing
agent. Excess reagent is removed, and the esters are determined by
electron-capture GC. The method detection limit has been estimated
at 249 ug/L for MCPA.
TREATMENT TECHNOLOGIES
Oxidation by ozone of 500 mg/L MCPA, after 50 to 80% disappearance
of initial compound, produced no identifiable degradation products
(Legube et al., 1981). This indicates that oxidation by ozone may
be a possible MCPA removal technique.
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MCPA August, 1987
-14-
IX. REFERENCES
Buslovich, S.Y.i Z.A. Aleksashina and V.M. Kolosovskaya. 1979. Effect of
phenobarbital on the embryotoxic action of 2-methyl-4-chlorophenoxyacetic
acid (a herbicide). Russ. Pharmacol. Toxicol. 24(2): 57-61.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
DeRose, H.R. 1946. Persistence of some plant growth-regulators when applied
to the soil in herbicidal treatments. Botanical Gazette. 107:583-589.
Elo, H.A. 1976. Distribution and elimination of 2-methyl-4-chlorophenoxy-
acetic acid (MCPA) in male rats. Scand. J. Work Environ. Health.
3:100-103.
Elo, H.A., and P. Ylitalo. 1977. Substantial increase in the levels of
chlorophenoxyacetic acids in the CNS of rats as a result of severe
intoxication. Acta Pharmacol. Toxicol. 41:280.
Elo, H.A., and P. Ylitalo. 1979. Distribution of 2-methyl-4-chlorophenoxyacetic
acid and 2,4-dichlorophenoxyacetic acid in male rats: Evidence for the
involvement of the central nervous system in their toxicity. Toxicol.
Appl. Pharm. 51:439-446.
Elo, H.A., P. Ylitalo, J. Kyottila and H. Herronen. 1982. Increase in the
penetration of tracer compounds into the rat brain during 2-methyl-4-
chlorophenoxyacetic acid (MCPA) intoxication. Acta Pharmacol. Toxicol.
50:104-107.
Eronen, L., R. Julkunen and A. Saarelainen. 1979. MCPA residues in developing
forest ecosystem after aerial spraying. Bull. Environ. Contain. Toxicol.
21:791-798.
Fjeldstad, P., and A. Wannag. 1977. Human urinary excretion of the herbicide
2-methyl-4-chlorophenoxyacetic acid. Scand. J. Work Environ. Health.
3:100-103.
Frank, R., G.J. Siron and B.D. Ripley. 1979. Herbicide contamination of well
waters in Ontario, Canada, 1969-78. Pestic. Monitor. J. 13:120-127.
Fryer, J.D., and K. Kirkland. 1970. Field experiments to investigate long
term effects of repeated applications of MCPA, tri-allate, simazine and
linuron: report after 6 years. Weed Res. 10(2): 1 33-1 58.
Gurd, M.R., G.L.M. Harmer and B. Lessel. 1965. Summary of toxicological
data: Acute toxicity and 7-month feeding studies with mecoprop and
MCPA. Food Cosmet. Toxicol. 3:883-885.
Hattula, M.L., H. Reunanen, R. Krees, A.V. Arstila and J. Knuutinen. 1979.
Toxicity of 5-chloro-3-methyl-catechol to rat: Chemical observations
and light microscopy of the tissue. Bull. Environ. Contarn. Toxicol.
22:457-461.
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MCPA August, 1987
-15-
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
Wilkins.
Helling, C.S. 1971. Pesticide mobility in soils. II. Application of soil
thin-layer chromatography. Proc. Soil Sci. Soc. Am. 35(5):737-743.
Helling, C.S., and B.C. Turner. 1968. Pesticide mobility: Determination by
soil thin-layer chromatography. Science. 162(3853):562-563.
Hellwig, J. 1986. Report on the study of the toxicity of MCPA in beagle
dogs after 12-month administration in the diet. Project No. 33D0046/8341.
Unpublished study. MRID 164352.
Herzel, F., and G. Schmidt. 1979. Testing the leaching behavior of herbicides
on lysimeters and small columns. WaBoLu-Berichte. (3):1-16.
Holsing, G.C.* 1968. Thirteen-week dietary/oral administration in dogs.
Final Report. Project No. 517-101. Unpublished study. MRID 00004756.
Holsing, G.C., and M. Kundzin.* 1968. Three-month dietary administration
study in rats. Project No. 517-100. Unpublished study. MRID 00004775.
Holsing, G.C., and M. Kundzin.* 1970. Three-month dietary administration
study in rats. Final Report. Project No. 517-106. Unpublished study.
MRID 00004776.
IARC. 1983. International Agency for Research on Cancer. IARC monograph on
the evaluation of carcinogenic risk to chemicals to man. Lyon, France:
IARC.
Irvine, L.F.H., D. Whittaker, J. Hunter et al.* 1980. MCPA/oral teratogenicity
study in the Dutch belted rabbit. Report No. 1737R-277/5. Unpublished
study. MRID 00041637.
Irvine, L.F.H.* 1980. MCPA oral teratogenicity study in the rat. Report No.
1996-277/76. Unpublished study. MRID 00066317.
Legube, B., B. Langlaia, B. Sohm and M. Dore. 1981. Identification of
ozonation products of aromatic hydrocarbon micropollutants: Effect on
chlorination and biological filtration. Ozone: Sci. Eng. 3(1):33-48.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
Linnainmaa, K. 1984. Induction of sister chromatid exchanges by the peroxisome
proliferators 2,4-D, MCPA and clofibrate in vivo and in vitro. Carcino-
genesis. 5(6):703-707.
Loos, M.A., I.F. Schlosser and W.R. Mapham. 1979. Phenoxy herbicide degrada-
tion in soils: quantitative studies of 2,4-D- and MCPA-degrading
microbial populations. Soil Biol. and Biochem. 11 (4): 377-385.
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MCPA August, 1987
-16-
MacKenzie, K.M. 1986. Two-generation study with MCPA in rats. Final report.
Study No. 6148-100. Unpublished study.
Magnusson, J. et al. 1977. Mutagenic effects of chlorinated phenoxyacetic
acids in Drosophila melanogaster. Hereditas. 87:121-123.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Moriya, M., T. Ohta, T. Watanabek, T. Miyazawa, K. Kato and Y. Shirasu.
1983. Further mutagenicity studies on pesticides in bacterial reversion
assay systems. Mutat. Res. 116:185-216.
HAS. 1977. National Academy of Sciences. Drinking water and health. Vol. 1.
Washington, DC: National Academy Press.
Palmer, A.K., and M.R. Lovell.* 1971, Effect of MCPA on pregnancy of the
mouse. Unpublished study. MRID 00004447.
Palva, H.L.A., 0. Koivisto and I.P. Palva. 1975. Aplastic anemia after
exposure to a weed killer, 2-methyl-4-chlorophenoxyacetic acid. Acta.
Haemat. 53:105-108.
Raltech.* 1979. Raltech Scientific Services, Inc. Dermal LD50. Unpublished
study. MRID 00021973.
Reuzel, P., and Hendriksen, C.* 1980. Subchronic (13-week) oral toxicity
study of MCPA in Beagle dogs: Final report: Project No. B77/1867:
Report Nos. R6478 and R6337. Unpublished study prepared by Central
Institute for Nutrition and Food Research.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. NIOSH,
National Library of Medicine On-Line File.
Soderquist, C.J., and D.G. Crosby. 1974. The dissipation of 4-chloro-2-
methylphenoxyacetic acid (MCPA) in a rice field. Unpublished study
prepared by Univ. of California, Davis, Department of Environmental
Toxicology, submitted by Dow Chemical Company, Midland, MI.
Soderquist, C.J., and D.G. Crosby. 1975. Dissipation of 4-chloro-2-methyl-
phenoxyacetic ac'd (MCPA) in a rice field. Pestic. Sci. 6(1):17-33.
Sokolov, M.S., L.L. Knyr, B.P. Strekozov, V.D. Agarkov, A.P. Chubenko, and
B.A. Kryzhko. 1974. The behavior of some herbicides under the conditions
of a rice irrigation system. Khimiya v Sel'skom Khozyaistve (Chemistry
in Agriculture). 13:224-234.
Sokolov, M.S., L.L. Knyr, B.P. Strekozov, and V.D. Agarkov. 1975. Behavior
of proanide, yalan, MCPA and 2,4-D in rice irrigation systems of the
Kuban River. Agrokhimiya (Agricultural Chemistry). 3:95-106.
Suzuki, H.K.* 1977. Dissipation of Banvel, bromozynil or MCPA or combination
thereof in two soil types: Report No. 181. Unpublished study submitted
by Velsicol Chemical Corporation, Chicago, IL.
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MCPA August, 1987
-17-
STORET. 1987.
Timonen, T.T., and I. P. Palva. 1980. Acute leukemia after exposure to a
weed killer, 2-methyl-4-chlorophenoxyacetic acid. Acta Haemat. 63:170-171
Torstensson, N.T.L. 1975. Degradation of 2,4-D and MCPA in soils of low pH.
In: Pesticides: IUPAC Third International Congress; July 3-9, 1974,
Helsinki, Finland. Coulston, P., and F. Korte, eds. Stuttgart, West
Germany: George Thieme. (Environmental Quality and Safety, Supplement,
Vol. 3). pp. 262-265.
Torstensson, N.T.L., J. Stark and B. Goransson. 1975. The effect of repeated
applications of 2,4-D and MCPA on their breakdown in soil. Weed Res.
15(3):159-164.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.339.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. U.S. EPA Method 615
- Chlorinated phenoxy acids. Fed. Reg. 50:40701. October 4.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for car-
cinogen risk assessment. Fed. Reg. 51(185):33992-34002. September 24.
Vainio, H., K. Linnainmaa, M. Kahonen, J. Nickels, E. Hietanen, J. Marniemi
and P. Peltoneu. 1983. Hypolipidemia and peroxisome proliferation
induced by phenoxyacetic acid herbicide in rats. Biochem. Pharmacol.
32(18):2775-2779.
Verschuuren, H.G., R. Kroes and E.N. denTonkecaar. 1975. Short-term oral
and dermal toxicity of MCPA and MCPP. Toxicology. 3:349-359.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
METHOMYL
DRAFT
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
understai 3ing 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.
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Methorny1 August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 16752-77-5
Structural Formula
0
CH,-CeN-0-C-N-CH3
S-CH, H
S-Methyl-N [ (methylcarbamoyl )oxy] -thioacetimidate
Synonyms
0 Dupont Insecticide 1179; Dupont 1179; Insecticide 1,179; Insecticide
1179; IN 1179, Lannate; Mesomile; Nudrin; SO 14999; WL 18236 (Meister,
1983).
Uses
0 Methorny1 is a carbamate insecticide used to control a broad spectrum
of insects in agricultural and ornamental crops (Meister, 1983).
Properties (Meister, 1983; Windholz et al., 1983; Cohen, 1984; CHEMLAB, 1985;
and TDB, 1985)
Chemical Formula C5H1002N2S
Molecular Weight 162.20
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 78 to 79°C
Density (24°C) 1.29
Vapor Pressure (25°C) 5 x 10~5 mm Hg
Specific Gravity
Water Solubility (25°C) 10,000 mg/L
Log Octanol/Water Partition -3.56
Coefficient
Taste Threshold —
Odor Threshold —
Conversion Factor --
Occurrence
0 Me thorny 1 has been found in 2 of 446 surface water samples analyzed
and in 25 of 1,023 ground water samples (STORET, 1987). Samples were
collected at 110 surface water locations and 1,000 ground water
locations, and methorny1 was found in California, Georgia and Texas.
The 85th percentile of all non-zero samples was 2 ug/L in surface
water and 10 ug/L in ground water sources. The maximum concentration
found in surface water was 2 ug/L and in ground water it was 10 ug/L.
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Methomyl August, 1987
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Environmental Fate
0 In laboratory and greenhouse studies, methomyl was more rapidly
degraded in a sandy loam and a California soil than in silt loam
soils, with 21, 31, and 44 to 48% of the applied methomyl remaining in
the respective soils 42-45 days after treatment. The major degra-
dation product was carbon dioxide, which accounted for 23 to 47% of
the applied methomyl after 42 to 45 days. A minor degradation product,
S-methyl-N-hydroxy-thioacetimidate (a possible hydrolysis product),
was also found. Methomyl half-lives were less than 30 days in sandy loam
soil, less than 42 days in California soil, and approximately 45 days
in muck and silt loam soils. In a sterilized Flanagan silt loam
soil, 89% of the methomyl remained 45 days after application, indicating
that methomyl degradation in soil is primarily a microbial process
(Harvey, 1977a,b).
0 The nitrogen-fixing ability of some bacteria was severely reduced
(by as much as 85%) when methomyl was applied at 20 to 160 ppm (Huang,
1978).
0 In another study, methomyl (18 ppm) had no effect on fungal and
bacterial population or on carbon dioxide production in .either silt
loam or fine sand soils (Peeples, 1977).
0 No methomyl residues were detected in a muck soil 7 to 32 days after
treatment (E.I. DuPont de Nemours and Co., 1971).
0 The environmental fate of methomyl has also been the subject of
several undated, unpublished reports (Harvey, undated a,b; Harvey
and Pease; Han).
III. PHARMACOKINETICS
Absorption
0 Single oral doses of 1-14c-methomyl (purity not specified) were ad-
ministered via gavage to female CD rats as a suspension in 1% aqueous
methylcellulose. Ninety-five percent of the dose could be accounted
for in excretory products or tissue residues, indicating virtual
complete absorption from the gastrointestinal tract (Andrawes et al,
1976).
0 Baron (1971) reported that in rats given a single oral dose of 5 mg/kg
of I-1 ^-labeled methomyl (purity not specified), approximately 2% of
the original label was excreted in the feces after 3 days, indicating
essentially complete gastrointestinal absorption.
Distribution
0 Baron (1971) fed a single oral dose of 1-14c-labeled methomyl (5 mg/kg,
purity not specified) to rats and analyzed 13 ma^or tissues for residues
at 1 and 3 days after dosing. Only 10% of the label was present in
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Methomyl August, 1987
-4-
tissues 24 hours after dosing, with no evidence of accumulation at
any site. By this time, over 40% of the label had been excreted via
the lung. At 3 days after dosing, tissue residues were essentially
unchanged from day 1, suggesting incorporation of label into tissue
components.
0 Baron (1971) reported that feeding methorny1 to a lactating cow at
levels of 0.2 or 20 ppm in the diet (duration not specified) resulted
in very low residues (less than 0.02 ppm) in the milk, meat, fat,
liver and kidney.
Metabolism
0 According to Baron (1971), in 72 hours approximately 15 to 23% of a
5-mg/kg oral dose of l-14C-labeled methomyl in rats could be accounted
for as carbon dioxide, 33% as another metabolite in expired air, and 25%
as metabolites in the urine.
0 Harvey (1974) reported that in the rat, 1-14C-labeled methomyl (dose
and purity not specified) was metabolized to carbon dioxide (25%) or
acetonitrile (50%) within 72 hours.
0 Andrawes et al. (1976) reported that single oral doses of 4 mg/kg
were rapidly metabolized in the rat. In exhaled air, carbon dioxide
and acetonitrile were the major metabolites. In 24-hour urine samples,
polar metabolites (80%) and acetonitrile (18%), both free and conjugated,
were found with free methomyl, methy(o), the oxime and the sulfoxide
oxime detected at low levels.
0 Dorough (1977), in a series of studies with 14C-labeled isomeric forms
of methomyl, confirmed the report by Harvey (1974) of the excretion of
labeled carbon dioxide and acetonitrile in the expired air of treated
rats. In addition, nearly complete (79 to 84%) hydrolysis of the
ester linkage was apparent within 6 hours, prior to the major
formation of carbon dioxide and acetonitrile from methomyl. The
author suggested the following pathway: partial isomerization of
methomyl is followed by hydrolysis of the two isomeric forms to yield
two isomeric oximes that then break down to carbon dioxide and
acetonitrile at different rates. No additional metabolites were
identified.
Excretion
0 Baron (1971) stated that within 72 hours after receiving a single
oral dose of I-14C-labeled methomyl, rats excreted 15 to 23% as
carbon dioxide, 33% as other metabolites in the expired air and
approximately 16 to 27% as methomyl and metabolites in the urine.
0 Harvey (1974) reported that 75% of an oral dose of 1-1 ^-labeled
methomyl (dose and purity not specified) was excreted by rats within
72 hours, 50% as acetonitrile and 25% as carbon dioxide in the expired
air. In contrast to other carbamates, sulfur-containing metabolites
were not found in the urine.
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Methomyl August, 1987
-5-
0 Andrawes et al. (1976) reported that single oral doses (4 mg/kg) of
1_14c-labeled methomyl were rapidly excreted, with 32% of the dose
recovered in urine, 19% in feces and 40% in exhaled air after 4 days.
IV. HEALTH EFFECTS
Humans
^^^••^^
Short-term Exposure
0 Liddle et al. (1979) reported a case of methomyl poisoning in Jamaica,
W.I., involving five men who had eaten a meal that included unleavened
bread. Methomyl was discovered in an unlabeled plastic bag in a tin
can, and had evidently been used as salt in preparation of the bread.
Approximately 3 hours after the meal, the men were found critically
ill, frothing at the mouth, twitching and trembling. Three were dead
on arrival at the hospital. One of the two survivors showed generalized
twitching and spasms, fasciculation, and respiratory impairment
thought to be due to severe bronchiospasms. The other patient walked
unaided and appeared generally normal. Both patients were given
atropine intravenously, and the symptomatic patient recovered within
2 hours after treatment. Methomyl was confirmed in the stomach
contents of each of the men who died, and analysis of the bread
indicated that it contained 1.1% methomyl. It was stated that two of
the victims had eaten about 75 to 100 g of bread each, or 0.82 to
1.1 g of methomyl. From these data it may be calculated that a dose
of 12 to 15 mg/kg body weight can be fatal in humans.
0 Araki et al. (1982) reported a case of a 31-year old woman who
committed suicide, giving methomyl in drinks to herself and her two
children. The 9-year-old elder son survived. In autopsies performed
on the mother and the 6-year-old son, the mucous membranes of the
stomach were blackish-brown, markedly edematous and congested. The
lungs were heavy and congested. On the basis of measured stomach
contents and tissue levels, it was estimated that the total doses
taken were 2.75 g (55 mg/kg) by the mother and 0.26 g (13 mg/kg) by
the child.
Long-term Exposure
0 Morse and Baker (1979) reported on a survey of the health of workers
in a plant that manufactured methomyl. The plant had also manufactured
propanil, an herbicide manufactured from 3,4-dichloroaniline. The
plant employed 111 workers in seven job categories. A complete work
history, symptoms or history of poisoning, personal habits, and
sources of other chemical exposure were obtained. Blood samples were
collected from 100 of the 111 workers (96% males). Blood chemistries,
blood counts, and cholinesterase (ChE) determinations were carried
out. A routine urinalysis was also performed. Average employment at
the plant was 2 years. Packaging workers had the highest rate of
"methorny 1" symptoms: small pupils (46%), nausea and vomiting (46%),
blurred vision (46%) and increased salivation (27%). Biomedical
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Methorayl August, 1987
-6-
examination did not demonstrate significant effects, and acetylcholin-
esterase findings were normal. Other effects, such as chloracne,
were reported but were considered related to propanil exposure.
Animals
Short-term Exposure
0 The acute oral LD^Q reported for methorny1 in the fasted male and female
rat ranged from 17 to 25 rag/kg (Bedo and Cieieszky, 1980; Dashiell
and Kennedy, 1984; Kaplan and Sherman, 1977). The oral LD50 in the
nonfasted rat was 40 mg/kg (Dashiell and Kennedy, 1984). Clinical signs
in rats included chewing motions, profuse salivation, lacrimation,
bulging eyes, fasciculatlons and tremors characteristic of ChE inhibition.
0 The acute oral 1*050 for methomyl in the mouse ranged from 27 to
55 mg/kg (Boulton et alo, 1971; El-Sebae et al., 1979; Natoff and
Reiff, 1973).
0 The oral 1.050 in hens was 28 mg/kg and in Japanese quail, 34 mg/kg.
(Kaplan and Sherman, 1977).
0 The 4-hour inhalation 1^50 of methomyl in rats was 300 rag/m^. Animals
showed the typical signs of ChE inhibition, including salivation,
lacrimation and tremors (ACGIH, 1984).
0 Bedo and Cieieszky (1980) administered single oral doses of methomyl
(purity not specified) by gavage to stock colony rats at dose levels
of 0, 2, 3 or 10 mg/kg. The high dose (10 mg/kg) produced typical
tremors in rats, and brain ChE levels were decreased. Mixed-function
oxidase, glucose-6-phosphatase activity, glycogen, and vitamin A
levels in the liver were unaffected. Apparently, dose levels of 2 or
3 mg/kg did not produce these effects.
0 Woodside et al. (1978) fed methomyl (purity not specified) in the diet
to male and female Wistar rats for 7 days at dose levels of 0, 5.0,
17 or 41 mg/kg/day in males and 0, 6.3, 15 or 39 mg/kg/day in females.
Body weight gain was depressed at doses of 17 and 41 mg/kg/day in the
males and at 15 and 39 mg/kg/day in the females. Liver and kidney
weight were also depressed at 41 mg/kg/day in the male rat and at
15 and 39 mg/kg/day in the female rat. No effects were noted at the
lowest doses. This study did not mention clinical signs of toxicity,
and no measurements of plasma or brain ChE activity were reported.
The No-Observed-Adverse-Effect-Level (NOAEL) identified in this
study is 5.0 mg/kg/day.
0 Bedo and Cieieszky (1980) fed methomyl (purity not specified) in the
diet at levels of 0, 100, 400 or 800 ppm to young adult male and female
stock colony rats for 10 days. Assuming that 1 ppm in the diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these doses
correspond to 0, 5, 20 or 40 mg/kg/day. Brain ChE inhibition could
not be detected at any dietary level. The only findings were increased
mixed-function oxidase activity in the livers of female rats at 400
and 800 ppm. This study identified a NOAEL of 800 ppm (40 mg/kg/day).
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Methomyl August, 1987
-7-
0 Kaplan and Sherman (1977) administered methorny1 (90% pure) to six
male Charles River-Cesarian Derived (ChR-CD) rats at 0 or 5.1 mg/kg/day,
five times a week for 2 weeks. Following treatment, survival, clinical
signs, ChE activity and histopathology were evaluated. All rats
survived the dosing period. Clinical signs in treated rats included
chewing motions, profuse salivation, lacrimation, bulging eyes,
fasciculations and tremors characteristic of ChE inhibition. The
authors reported that the signs became less pronounced after the
first week of dosing, indicating some degree of adaptation. Plasma
ChE was comparable to control levels, and no compound-related histo-
pathologic effects were reported. A Lowest-Observed-Adverse-Effect-
Level (LOAEL) of 5.1 mg/kg/day was identified from this study.
Dermal/Ocular Effects
0 Kaplan and Sherman (1977) applied a 52.8% aqueous suspension of
methomyl to the clipped, intact skin of six adult male albino rabbits
and covered the area with an occlusive patch for a 24-hour period.
The lethal dose was found to be greater than 5,000 mg/kg, the maximum
feasible dose.
0 McAlack (1973) reported a 10-day subacute exposure of rabbit skin to
methomyl. Male albino rabbits, six per dosage group, were treated
with 0, 50 or 100 mg/kg/day for 10 days. The compound was diluted in
water (29% solution), placed on the skin and covered with an occlusive
covering for 6 hours per day. No signs of ChE inhibition were noted
in any of the animals.
0 Ten rabbits survived 15 daily doses of 200 mg/kg/day of methomyl
applied to intact skin. When the same dose of methomyl was applied
to abraded skin, rabbits showed labored respiration, nasal discharge,
salivation, excessive mastication, tremors, poor coordination, hyper-
sensitivity and abdominal hypertonia. These effects occurred within
1 hour after dosing in most animals. One animal died after the first
dose, and another died after the eighth application. These deaths
appeared to be compound-related (Kaplan and Sherman, 1977).
Long-term Exposure
0 Kaplan and Sherman (1977) reported a 90-day feeding study in
ChR-CD rats (10/sex/group) given food containing methomyl (90% purity)
at dietary levels of 0, 10, 50, 125 or 250 ppm active ingredient (a.i.).
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of about 0, 0.5, 2.5, 6.2 or
12.5 mg/kg/day. After 6 weeks, the 125-ppm dose was increased to 500
ppm (25 mg/kg/day) for the remainder of the study. Clinical signs,
biochemical analyses (including plasma ChE) and urinalyses were not
abnormal. In a few cases, lower hemoglobin valves were observed at one
month in females receiving 50 ppm (2.5 ug/kg/day) and at two months
in males receiving 250 ppm. At three months, the red cell count of
female rats at 250 ppm was somewhat lower than controls, but still
within normal limits. These findings were consistent with moderate
increases of erythroid components observed histologically in the bone
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Me thorny1 August, 1987
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narrow. Microscopic examination of all other tissues showed no
consistent abnormalities. Based on these observations, this study
identified a NOAEL of SO ppm (2.5 ing/kg/day) and a LOAEL of 250 ppm
(12.5 mg/kg/day).
• In a 90-day study using dogs, Kaplan and Sherman (1977) fed me thorny 1
(90% pure) to four males and four females, 11 to 13 months of age,
at dietary levels of 0, 50, 100 or 400 ppm a.i. Assuming that 1 ppm
in the diet of dogs is equivalent to 0.025 mg/kg/day (Lehman, 1959),
this corresponds to doses of about 0, 1.25, 2.5 or 10 mg/kg/day.
Hematological, biochemical and urine analyses were conducted at least
three times on each dog prior to the study and then at 1, 2 and 3
months during the exposure period. Body weight was monitored weekly.
At necropsy, organ weights were recorded, and over 30 tissues were
prepared for histopathologic examination. No effects attributable to
methorny1 were found during or at the conclusion of the study. Based
on these data, a NOAEL of 10 mg/kg/day was identified.
0 Honan et al. (1978) reported a 13-week dietary study of methomyl
(purity not specified) in F-344 rats. Dose levels were reported
to be 0, 1, 3, 10.2, or 30.2 mg/kg/day for male rats, and 0, 1, 3, 9.9
or 29.8 mg/kg/day for female rats. There were no deaths or clinical
signs of toxicity. The body weight gain of females (but not males)
was significantly depressed at all dose levels from day 28 until
completion of the study. Kidney weight to body weight ratios were
significantly increased in female rats at the two highest dose levels.
Red blood cell ChE activity was elevated at the high dose levels, but
plasma and brain ChE levels were normal at all dose levels. Histo-
pathological examination of 31 tissues from representative high-dose
and control animals revealed no significant effects. Weights of
brain, liver, kidney, spleen, heart, adrenals and testes were not
altered. This study identified a NOAEL of 3 mg/kg/day and a LOAEL of
9.9 mg/kg/day.
0 Bedo and Cieleszky (1980) reported a 90-day feeding study of methomyl
in male and female rats receiving dietary levels of 100 or 200 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 5 or 10 mg/kg/day. At
200 ppm, the female rats showed decreased brain ChE activity, decreased
liver vitamin A content and elevated total serum lipids. This study
identified a NOAPL of 100 ppm (5 mg/kg/day).
0 Kaplan and Sherman (1977) reported a 22-month dietary feeding study
in which Charles Raver-CD male and female rats were fed methomyl
(90 or 100% pure) at dietary levels of 0, 50, 100, 200 or 400 ppm
a.i. Assuming that 1 ppm in the diet of rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of about 0,
2.5, 5, 10 or 20 mg/kg/day. Mortality data were not reported. At
autopsy, 9 of 13 males and 21 of 23 females at the 400-ppm level had
kidney tubular hypertrophy and vacuolization of epithelial cells of
the proximal convoluted tubules. Compound-related histological
alterations were also seen in the spleens of female rats at the
200-ppm dose level. No effects were seen on ChE levels in plasma or
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Methomyl August. 1987
-9-
red blood cells. This study identified a LOAEL of 200 ppm
(10 mg/kg/day) and a NOAEL of 100 ppm (5 mg/kg/day).
0 Kaplan and Sherman (1977) performed a 2-year feeding study in beagle
dogs (four/sex/dose). Methomyl (90 or 100% pure) was supplied at
dietary levels of 0, 50, 100, 400 or 1,000 ppm a.i. Assuming that
1 ppm in the diet of dogs is equivalent to 0.025 mg/kg/day (Lehman,
1959), this corresponds to doses of about 1.25, 2.5, 10 or 25 mg/kg/day.
Hematological, biochemical (including plasma- and red-blood-cell ChE
activity) and urinanalyses were conducted once on each dog prior to
the start of the study, at 3, 6, 12, 18 months during the exposure
period and at 24-month sacrifice. At 1 year, one male and one female
per dose group were sacrificed for histopathological examination.
One female dog at the 1,000-ppm dose level died after 8 weeks ,in the
study, and a replacement dog died after 18 days. Death was preceded
by convulsive seizures and coma. These deaths appear to be compound-
related. Two male dogs in the 1,000-ppm dose group showed clinical
signs during week 13, including tremors, salivation, incoordination
and circling movements. Hematological studies revealed slight-to-
moderate anemia in five dogs (1,000-ppm dose group) at 3 months,
which persisted in one dog to sacrifice. No compound-related signs
or effects were noted with respect to appetite, body weight changes,
biochemical studies (including ChE) and urinanalyses. Dose-related
histopathological changes were seen in kidney and spleen of animals
receiving 400 and 1,000 ppm. Changes were also seen in livers and
bone marrow of animals receiving 1,000 ppm. Pigment deposition was
noted in the epithelial cells of the proximal convoluted tubules of
the kidney in males at 400 and 1,000 ppm and in females at 1,000 ppm.
A minimal-to-slight increase in bile duct proliferation and a slight
increase in bone marrow activity was seen in animals receiving
1,000 ppm. The authors concluded that histological results indicated
a NOAEL of 100 ppm (2.5 mg/kg/day). Minimal histopathological changes
seen in the kidneys and spleen of animals receiving 400 ppm (10 mg/kg/day),
identified this level as the LOAEL.
0 Hazelton Laboratories (1981) reported a 2-year study of methorny1
(purity not specified) in mice. Male and female CD-1 mice (80/sex/dose)
were fed methomyl in the diet at dose levels of 0, 50, 100, or 800 ppm
for 104 weeks. Assuming 1 ppm in the diet to be equivalent
to 0.15 mg/kg/day (Lehman, 1959), this corresponds to doses of about
0, 7.5, 15 or 120 mg/kg/day. Survival was significantly reduced (-.o
details provided) in both males and females at the 800-ppm dose level by
week 26. The 800 ppm dose level was reduced to 400 ppm (1.0 mg/kg/day)
at week 28 and then further reduced to 200 ppm (30 mg/kg/day) at week
39. At week 39, the 100 ppm was decreased to 75 ppm (11.2 mg/kg/day).
Survival was depressed in all groups of treated males at 104 weeks.
No compound-related histopathological changes were noted in tissues
of animals necropsied at 104 weeks. A LOAEL of 50 ppm (7.5 mg/kg/day;
the lowest dose tested) may be identified based on decreased survival.
Reproductive Effects
0 Male and female weanling Charles River-CD rats were fed methomyl
(90% pure) at dietary levels of 0, 50, or 100 ppm a.i. for 3 months.
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Methomyl August, 1987
-10-
Assuming that 1 ppm in the diet of weanling rats is equivalent to
0.05 mg/kg/day (Lehman, 1959), these doses correspond to about 0, 2.5
or 5 mg/kg/day. Ten males and twenty females from each group were
bred and continued on the diet through three generations. No adverse
effects were reported on reproduction or lactation, and no pathologic
changes were found in the weanling pups of the F3b generation (Kaplan
and Sherman, 1977). A NOAEL of 5 mg/kg/day was identified from the
highest dose tested.
Developmental Effects
0 New Zealand White rabbits, five per group, were dosed with 0, 2, 6 or
16 mg/kg of methomyl (98.7% pure) on days 7 through 19 of gestation.
One animal died at the 16 mg/kg dose level, exhibiting characteristic
signs of ChE inhibition, including tremors, excitability, salivation
and convulsions. No adverse effects were observed at any dose level
on embryo viability or on the frequency of soft-tissue or skeletal
malformations (Feussner et al., 1983). This study identified a
maternal NOAEL of 6 mg/kg and a teratogenic NOAEL of 16 mg/kg/day,
the highest dose tested.
0 Kaplan and Sherman (1977) fed methomyl (90% pure) to pregnant New
Zealand White rabbits on days 8 to 16 of gestation at dietary levels
of 0, 50 or 100 ppm active ingredient. Assuming that 1 ppm in the
diet of rabbits is equivalant to 0.03 mg/kg/day (Lehman, 1959), this
corresponds to doses of about 0, 1.5 or 3 mg/kg/day. One-third of
the fetuses were stained with Alizarin Red S and cleared for skeletal
examination. Since no soft tissue or skeletal abnormalities were
observed at any dose level tested, a NOAEL of 3 mg/kg/day was identified.
Mutagenicity
0 Methomyl has been reported to be negative in the Ames test utilizing
Salmonella typhimurium strains TA 98, TA 1OO, TA 1535, TA 1537, and
TA 1538 without metabolic activation (Blevins et al., 1977; Moriya
et al., 1983). Waters et al. (1980) reported methomyl as negative
with and without metabolic activation in strains TA 1OO, TA 1535,
TA 1537 and TA 1538.
Carcinogenicity
0 Kaplan and Sherman (1977) fed ChR-CD rats (35/sex/dose) methomyl (90%
pure) in the diet at levels of 0, 50, 100, 200 or 400 ppm active
ingredient for 22 months. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), these doses correspond
to about 0, 2.5, 5, 10 or 20 mg/kg/day. Gross and histological
examination revealed no increased tumor incidence in either male or
female rats.
0 Hazelton Laboratories (1981) reported the results of a 2-year study
of methomyl (purity not specified) in CD-I mice (80/sex/dose). Initial
dose levels were 0, 50, 100, or 800 ppm. Assuming that 1 ppm in the
diet of mice is equivalent to 0.15 mg/kg/day (Lehman, 1959), these
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Methomyl August, 1987
-11-
doses correspond to 0, 7.5, 15 or 120 mg/kg/day. Because of early
mortality, the 800-ppm dose was reduced to 400 ppm (60 mg/kg/day)
at week 28, and then to 200 ppm (30 mg/kg/day) at week 39. At week
29, the 100-ppm dose was reduced to 75 ppm (11.2 mg/kg/day). Histo-
logical examination at necropsy did not reveal any treatment-related
effects on tumor incidence.
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 ing/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 child
(1 L/day) or an adult (2 L/day).
One-Day Health Advisory
No information found in the available literature was suitable for deter-
mination of the One-day HA value for methomyl. It is, therefore, recommended
that the Drinking Water Equivalent Level (DWEL), adjusted for a child,
(0.25 mg/L) be used at this time as a conservative estimate of the One-day HA
value.
Ten-day Health Advisory
The health effects associated with acute and subchronic exposure to
methomyl are primarily associated with cholinesterase (ChE) inhibition.
Symptoms of ChE inhibition have been shown in rats at doses (via gavage) as
low as 5.1 mg/kg/day for 2 weeks (Kaplan and Sherman, 1977). Methomyl
incorporated into the diet may have less dramatic effects; no ChE effects
were observed in rats exposed subchronically to methomyl at dietary levels of
100 ppm (5 mg/kg/day)(Kaplan and Sherman, 1977; Bedo and Cieleszky, 1980).
Animal studies may be misleading in assessment of human toxicity. No
controlled human studies have been performed, but human fatalities from
methomyl ingestion after a single exposure to an estimated dose of 12 mg/kg
in bread or 13 mg/kg in drinks have been reported (Liddle et al., 1979; Araki
et al., 1982).
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Methorny1 August, 1987
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Because the timing and nature of administration can profoundly
affect the expression of methomyl toxicity, and little margin of safety
can be expected between doses that are fatal and those that cause little or
no acute toxicity, the available studies were judged to be inadequate for the
basis of the Ten-day HA value. Therefore, it is recommeded that the DWEL,
adjusted for a 10-kg child (0.25 mg/L), be used at this time as a conservative
estimate of the Ten-day HA value.
Longer-term Health Advisory
The onset of subchronic or chronic methomyl toxicity appears to occur at
doses similar to those that cause acute toxicity. Kidney toxicity (increased
kidney weight and hypertrophy) in acute, subchronic and chronic conditions
has been reported at doses of 15, 9.9 and 10 mg/kg/day, respectively (Woodside
et al., 1978; Homan et al., 1978; Kaplan and Sherman, 1977). Acute ChE
inhibition in rats exposed to methomyl via gavage has been reported to occur
at doses as low as 5.1 mg/kg/day, and human fatalities from methomyl ingestion
of approximately 12 mg/kg in bread and 13 rag/kg in drinks have been reported
(Liddle et al, 1979; Araki et al., 1982).
Little margin of safety can be expected between doses of methomyl that
are fatal and those that cause little or no longer-term toxicity. Therefore,
it is recommended that the DWEL adjusted for the child (0.25 mg/L) be used at
this time as a conservative estimate of the Longer-term HA value.
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 population that is likely to be
without appreciable risk of deleterious effects over a lifetime, and is
derived from the NQAEL (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 -f 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 classifi-
cation scheme of carcinogenic potential (U.S. EPA, 1986), then caution should
be exercised in assessing the risks associated with lifetime exposure to this
chemical.
Chronic exposure to methomyl in the diet induces renal toxicity in rats
and dogs. Rats exposed to 900 ppm (20 mg/kg/day) for 22 months exhibited
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Methomyl August, 1987
-13-
kidney tubular hypertrophy and vacuolation of the eptithelial cells, and
dogs exposed to 400 ppm (10 mg/kg/day) for 2 years exhibited swelling and
increased pigmentation of the epithelial cells of the proximal tubules
(Kaplan and Sherman, 1977). Effects on the kidney (increased weight) have
also been observed in rats exposed to 9.9 mg/kg/day in the diet for 13 weeks
(Homan et al., 1978). The NOAEL of 2.5 mg/kg/day identified from the dog
study is a conservative estimate of the NOAEL and serves as the basis for the
Lifetime HA.
In the Kaplan and Sherman (1977) study, beagle dogs (4/sex/dose) were
exposed to 50, 100, 400 or 1,000 ppm methomyl in the diet for 2 years (1.25,
2.5, 10 and 25 mg/kg/day). Dogs receiving 1.25 or 2.5 mg/kg/day showed no
evidence of toxic effects. Those receiving 10 mg/kg/day exhibited histopatho-
logical changes in the kidney and spleen. In addition to these effects,
animals receiving the highest dose also exhibited symptoms of central nervous
system (CNS) toxicity, as well as liver and bone marrow effects.
Using a NOAEL of 2.5 mg/kg/day, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (2.5 mg/kg/day) = Q.025 mg/kg/day
(100)
where:
2.5 mg/kg/day = NOAEL, based on absence of effects on blood chemistry
(including ChE activity), hematology, urinalysis,
histopathology or body weight in dogs exposed in the
diet for 2 years.
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 mg/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 adult.
For the 10-kg child, the DWEL is calculated as follows:
DWEL = (0'025 mg/kg/day) (10kg) = 0.25 mg/L (250 ug/L)
child (1 L/day)
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Methomyl August, 1987
-14-
where:
0.025 ing/kg/day = RfD
10 kg = assumed body weight of a child
1 L/day =» assumed daily water consumption of child
Step 3: Determination of a Lifetime Health Advisory
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 Two-year carcinogenicity studies in rats and mice (Kaplan and Sherman,
1977; Hazelton Laboratories, 1981) have not revealed any evidence of
carcinogenicity.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of methomyl.
0 Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986), methomyl is classified
in Group D: not classifiable as to human carcinogenicity. This group
is used for agents with inadequate human and animal evidence of
carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Academy of Sciences (NAS, 1983) has a Suggested-No-Adverse-
Response-Level (SNARL) of 0.175 mg/L, which was calculated using an
uncertainty factor of 100 and a NOAEL of 2.5 mg/kg/day identified in
the 2-year dog study by Kaplan and Sherman (1977).
0 Residue tolerances have been established for methomyl in or on raw
agricultural commodities (U.S. EPA, 1985). These tolerances are
based on an ADI value of 0.025 mg/kg/day, based on a NOAEL of
2.5 mg/kg/day in dogs and an uncertainty factor of 100. Residues
range from 0.1 (negligible) to 40 ppm.
0 The World Health Organization identified a Temporary ADI of 0.01
mg/kg/day (Vettorazzi and Van den Hurk, 1985).
0 ACGIH (1984) has adopted a threshold limit value (TLV) of 0.2 mg/m3
as a time-weighted average exposure for an 8-hour day.
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Methomyl Au*ust' 1987
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VII. ANALYTICAL METHODS
0 Analysis of methomyl is by a high-performance liquid chromatographic
(HPLC) 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. Compounds are separated by 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 to form a
fluorescent derivative that is detected using a fluorescence detector.
The method detection limit for methomyl has been estimated to be
approximately 0.7 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) adsorption
will remove methomyl from water. Whittaker (1980) experimentally
determined adsorption isotherms for methomyl solutions on GAC.
0 Whittaker (1980) reported the results of GAC columns operating under
benchscale conditions. At a flow rate of 0.8 gpm/sq ft and empty
bed contact time of 6 minutes, methomyl breakthrough (when effluent
concentration equals 10% of influent concentration) occurred after
124 bed volumes (BV). When a bi-solute methomyl-metribuzin solution
was passed over the same column, methomyl breakthrough occurred after
55 BV.
0 Treatment technologies for the removal of methomyl from water are
available and have been reported to be effective (Whittaker, 1980).
However, the selection of individual or combinations of technologies
must be based on a case-by-case technical evaluation, and an assessment
of the economics involved.
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Methomyl August, 1987
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IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH.
Andrawes, W.R., R.H. Bailey and G.C. Holsing.* 1976. Metabolism of acetyl-
1-14C-methomyl in the rat. Report No. 26946. Unpublished study.
Araki, M., K. Yonemitsu, T. Kambe, D. Idaka, S. Tsunenari, M. Kanda and
T. Kambara. 1982. Forensic toxicological investigation on fatal cases
of carbamate pesticide methomyl (Lannate) poisoning. Nippon Hoigaku
Zasshi. 36:584-588.
Baron, R.L. 1971. Toxicological considerations of metabolism of carbamate
insecticides: methomyl and carbaryl. Pesticide Terminal Residues, Invited
Paper, Int. Symp. Washington, DC. pp. 185-197.
Bedo, M., and V. Cieleszky. 1980. Nutritional toxicology in the evaluation
of pesticides. Bibl. "Nutr. Dieta." 29:20-31.
Blevins, R.D., M. Lee and J.D. Regan. 1977. Mutagenicity screening of five
methyl carbamate insecticides and their nitroso derivatives using mutants
of Salmonella typhimurium LT2. Mutat. Res. 56:1-6.
/
Boulton, J.J., C.B. Boyce, P.J. Jewess and R.F. Jones. 1971. Comparative
properties of N-acetyl derivatives of oxime N-methylcarbamates and aryl
N-methylcarbamates as insecticides and acetylcholinesterase inhibitors.
Pestic. Sci. 2:10-15.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Cohen, S.Z. 1984. List of potential groundwater contaminants. Memorandum to
I. Pomerantz. Washington, DC: U.S. Environmental Protection Agency.
August 28.
Dashiell, O.L. and G.L. Kennedy. 1984. The effects of fasting on the acute
oral toxicity of nine chemicals in the rat. J. Appl. Toxicol. 5:320-325.
Dorough, H.W. 1977. Metabolism of carbamate insecticides. Available from
the National Techni-al Information Service, Springfield, VA. PB-266 233,
Springfield, VA.
E.I. du Pont de Nemours and Co. 1971.* Methomyl decomposition in muck soil—
a field study. Unpublished study.
El-Sebae, A.H., S.A. Soliman, A. Khalil and E. Sorya. 1979. Comparative
selective toxicity of some insecticides to insects and mammals. Proc. Br.
Crop Prot. Conf.-Pest. Dis. pp. 731-736.
Feussner, E., M. Christian, G. Lightkep et al.* 1983. Embryo-fetal toxicity
and teratogenicity study of methomyl in the rabbit. Study No. 104-005.
Unpublished study. MRID 00131257.
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-17-
Han, J.C. Undated.* Evaluation of possible effects of methomyl on nitrifying
bacteria in soil. E.I. duPont de Nemours and Company, Inc., Wilmington,
DE. Unpublished study.
Harvey, J. Undated(a).* Decomposition of 14c-methomyl in a high organic
matter soil in the laboratory. E.I. duPont de Nemours and Company, Inc.,
Wilmington, DE. Unpublished study.
Harvey, J. Undated(b).* Exposure of S-methyl N-{ (methylcarbamoyDoxy)thioaceti-
midate in sunlight, water and soil. E.I. duPont de Nemours and Company,
Inc., Wilmington, DE. Unpublished study.
Harvey, J., Jr. and H.L. Pease. Undated.* Decomposition of methomyl in soil.
E.I. duPont de Nemours and Company, Inc., Wilmington, DE. Unpublished
study.
Harvey, J., Jr. 1974.* Metabolism of aldicarb and methomyl. Environmental
quality and safety supplement, Vol. III. Pesticides. International
Union of Pure and Applied Chemistry Third International Congress.
Helsinki, Finland.
Harvey, J., Jr. 1977a.* Decomposition of Homethomyl in a sandy loam soil
in the greenhouse. Unpublished study prepared in cooperation with the
University of Delaware, Soil Testing Laboratory, submitted by E.I. du
Pont de Nemours and Co., Wilmington, DE.
Harvey, J., Jr. 1977b.* Degradation of 14C-methomyl in Flanagan silt loam
in biometer flasks. Unpublished study prepared in cooperation with the
University of Delaware, Soil Testing Laboratory, submitted by E.I. du
Pont de Nemours and Co., Wilmington, DE.
Hazelton Laboratories.* 1981. Final report: 104-week chronic toxicity and
carcinogenicity study in mice. Project No. 201-510. Unpublished study.
MRID 00048423.
Homan, E.R., R.R. Maronpot and J.B. Reid.* 1978. Methomyl: inclusion in the
diet of rats for 13 weeks. Project Report 41-64. Unpublished study.
MRID 00044881.
Huang, C.Y. 1978. Effects of nitrogen fixing activity of blue-green algae
or. the yield of rice plants. Botanical Bull. Academia Sinica. 19(1 ):41-
52.
Kaplan, M.A. and H. Sherman. 1977. Toxicity studies with methyl N-[[(methyl-
amino)carbonyl]oxy]-ethanimidothioate. Toxicol. Appl. Pharmacol. 40:1-17.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food and Drug Off. U.S. Q. Bull.
Liddle, J.A., R.D. Kimbrough, L.L. Needham, R.E. Cline, A.L. Smrak, L.W. Yert
and D.D. Bayse. 1979. A fatal episode of accidental methomyl poisoning.
Clin. Toxicol. 15:159-167.
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Methorny1 August, 1987
-18-
McAlack. J.W.* 1973. 10-day subacute exposure of rabbit skin to lannate (R)
insecticide: Haskell Lab report No. 24-73. Unpublished study.
MRIO 00007032.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Moriya, M.f 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.
Morse, D.L., and E.L. Baker. 1979. Propanil-chloracne and methomyl toxicity
in workers of a pesticide manufacturing plant. Clin. Toxicol. 15:13-21.
NAS. 1983. National Academy of Sciences. Drinking water and health.
Volume 5. Washington, DC: National Academy Press.
Natoff, I.e. and B. Reiff. 1973. Effects of oximes on acute toxicity of
anticholinesterase carbamates. Toxicol. Appl. Pharmacol. 25:569-575.
Peeples, J.L. 1977.* Effect of methomyl on soil microorganisms. Unpublished
study submitted by E.I. du Pont de Nemours and Co., Inc., Wilmington, DE.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
STORET. 1987.
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.
OH: Environmental Monitoring and Support Laboratory, ECAO, Cincinnati.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal Regu-
lations. 40 CFR 180.253, July 1. p. 278.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185):33992-34003. Septem-
ber 24.
Vettorazzi, G. and G.W. Van den Hurk. 1985. Pesticides Reference Index,
Joint Meeting of Pesticide Residues. 1961-1984, p. 10.
Waters, M.D., V.F. Summon, A.D. Mitchell, T.A. Jorgenson and R. Valencia.
1980. An overview of short-term tests for the mutagenic and carcinogenic
potential of pesticides. J. Environ. Sci. Health. B15(6):867-906.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
Windholz, M., S. Budavari, R.F. Blumetti, E.S. Otterbein, eds. 1983. The
Merck index—an encyclopedia of chemicals and drugs, 10th ed. Rahway, NJ:
Merck and Company, Inc.
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Methorny1
August, 1987
-19-
Woodside, M.D., L.R. DePasso and J.B. Reid.1
feeding in the diet of rats for 7 days;
MRIO 00044880.
1978. UC 45650: Results of
Pro] 41-102. Unpublished study.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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METHYL PARATHION
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
August, 1987
DRAFT
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 accurate!'' 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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Methyl Parathion
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 298-00-0
Structural Formula
f
-P(OCH,)2
0,0-Dimethyl-0-(4-nitrophenyl) phosphorothioic acid
Synonyms
Matron; Meptox; Metaphos; Dimethyl parathion; Nitrox; Azofos; Nitrox 80;
BAY 11405; Metacide; Folidol M; Azophos; Methyl-E 605; DaIf; Meticide;
Methylthiophos; Pencap M; Penncap M; Sinafid M-48; Wofotox; Vofatox;
Thiophenit; Wofatox (Meister, 1983).
Uses
0 A restricted-use pesticide for control of various insects of economic
importance; especially effective for boll weevil control (Meister, 1983).
Properties (Hawley, 1981; Meister, 1983; CHEMLAB, 1985; TDB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (20°C)
Specific Gravity
Water Solubility (258C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C8H1005NSP
263.23
White crystalline solid
35 to 36°C
0.97 x 10~5 mm Hg
55 to 60 mg/L
3.11 (calculated)
Occurrence
Methyl parathion has been found in 1,402 of 29,002 surface water
samples analyzed and in 25 of 2,878 ground water samples (STORET,
1987). Samples were collected at 3,676 surface water locations and
2,026 ground water locations, and methyl parathion was found in 22
states. The 85th percentile of all nonzero samples was 1.18 ug/L
in surface water and 1 ug/L in ground water sources. The maximum
concentration found was 13 ug/L in surface water and 1.6 ug/L in
ground water.
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Methyl Parathion August, 1987
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Environmental Fate
0 Methyl parathion (99% pure) at 10 ppm was added to sea water and
exposed to sunlight; some samples were also kept in the dark (controls).
After 6 days, 57% of the parent compound had degraded but the degradates
were not identified. Since only 27% of the parent compound had degraded
in the dark controls, this indicates that methyl parathion is subject
to photodegradation in sea water (U.S. EPA, 1981).
0 The degradation rate of two formulations (EC and MCAP) of methyl
parathion, applied at 0.04 ppm, was compared in a sediment/water
system. Degradates were not identified; however, the parent compound
had a half-life of 1 to 3 days in water. In the hydrosoil plus
sediment, methyl parathion applied as an emulsifiable concentrate
formulation had a half-life of 1 to 3 days, whereas for the micro-
encapsulated formulation, the half-life was 3 to 7 days (Agchem, 1983).
0 Methyl parathion was relatively immobile in 30-cm soil columns of sandy
loam, siity clay loam and silt loam soils leached with 15.7 inches of
water, with no parent compound found below 10 cm or in the column
leachate, which was the case for the column of sand (Pennwalt Corporation,
1977).
0 Methyl parathion (MCAP or EC formulation) at 5 Ib ai/A (active
ingredient/acre) was detected in runoff water from field plots irrigated
4 to 5 days posttreatment. Levels found in soil and turf plots ranged
from 0.13 to 21 ppm and 0.17 to 0.20 ppm, respectively (Pennwalt
Corporation, 1972).
0 A field dissipation study with methyl parathion (4 Ib/gal EC) at 3 Ib
ai/A, applied alone or in combinaton with Curacron, dissipated to
nondetectable levels (<0.05 ppm) within 30 days in silt loam and
loamy sand soils (Ciba-Geigy Corporation, 1978).
III. PHARMACOKINETICS
Absorption
0 Braeckman et al. (1983) administered a single oral dose of 35S-methyl
parathion (20 mg/kg) by stomach tube to four mongrel dogs. Peak
concentrations in plasma ranged from 0.13 to 0.96 ug/mL, with peak
levels occurring 2 to 9 hours after dosing. In two dogs given single
oral doses of 35s-methyl parathion (3 mg/kg) in this study, absorption
was estimated to be 77 and 79%, based on urinary excretion of label.
The authors concluded that methyl parathion was well absorbed from
the gastrointestinal tract.
0 Hollingworth et al. (1967) gave a single oral dose of 32p_iaDeied
methyl parathion by gavage (3 or 17 mg/kg, dissolved in olive oil) to
male Swiss mice. Recovery of label in the urine reached a maximum of
about 85%, most of this occurring within 18 hours of dosing. The
amount of label in the feces was low, never exceeding 10% of the
dose. This indicated that absorption was at least 90% complete.
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Methyl Parathion August, 1987
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Distribution
0 Ackermann and Engst (1970) administered methyl parathion to pregnant
albino rats and examined the dams and fetuses for the distribution
of the pesticide. The pregnant rats (weighing about 270 g each) were
given 3 mg (11.1 mg/kg) of methyl parathion orally on days 1 to 3 of
gestation and sacrificed 30 minutes after the last dose. Methyl
parathion was detected in the maternal liver (25 ng/g), placenta
(80 ng/g), and in fetal brain (35 ng/g), liver (40 ng/g) and back
musculature (60 ng/g).
Metabolism
0 Hollingworth et al. (1967) gave 32p-labeled methyl parathion by
gavage (3 or 17 mg/kg, dissolved in olive oil) to male Swiss mice.
About 85% of the label appeared in the urine within 72 hours. Urinary
metabolites identified -24 hours after the low dose were: dimethyl
phosphoric acid (53.1%); dimethyl phosphorothioic acid (14.9%);
desmethyl phosphate (14.1%); desmethyl phosphorothioate (11.7%);
phosphoric acid (2.0%); methyl phosphoric acid (1.7%); and phosphate
(0.6%). The radioactivity in the urine was fully accounted for by
hydrolysis products and P=0 activation products. No evidence was
found for reduction of the nitro group to an amine, oxidation of the
ring methyl group, or hydroxylation of the ring. A generally similar
pattern was observed at the high dose, except for a lower percentage
of dimethyl phosphoric acid (31.9%) and higher percentages of desraethyl
phosphate (23.1%) and desmethylphosphorothionate (18.8%). Based on
this, the authors proposed a metabolic scheme involving oxidative
desulfuration, oxidative cleavage of the phospho group from the ring
and hydrolysis of the phosphomethyl esters.
0 Heal and DuBois (1965) investigated the in vitro detoxification of
methyl parathion and other phosphorothioates using liver microsomes
prepared from adult male Sprague-Dawley rats. Metabolism was found
to involve oxidative desulfuration followed by hydrolysis to yield
p-nitrophenol. Extracts from livers of adult male rats exhibited
higher metabolic activity than that of adult females (3.2 versus
1.9 units, where one unit equals 1 ug p-nitrophenol/50 mg liver
extract) (p
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Methyl Parathion August, 1987
-5-
NADPH2. The amounts of phenol and oxygen analog formed were 3.8 and
3.7 uM in the rabbit liver extract and 2.5 and 5.4 uM in the rat
liver extract, respectively.
Excretion
Braeckman et al. (1983) administered individual doses of 3 mg/kg of
35s-methyl parathion to two mongrel dogs. In each dog, the agent was
given once intravenously and, 1 week later, once orally via stomach
tube. This dosing pattern was repeated once in one dog. Urine was
collected every 24 hours for 6 days after each'treatment. Urinary
excretion 6 days after oral dosing was 63% in the animal without
repeated dosing and 70% and 78% in the other. Urinary excretion
6 days after intravenous dosing was 80% in the animal without repeated
dosing and 95 to 96% in the other. Most of the label appeared in urine
within two days. Other excretory routes were not monitored.
Hollingworth et al. (1967) gave 32p_iabeled methyl parathion (3 or
17 mg/kg, dissolved in olive oil) by gavage to male Swiss mice.
Recovery of label in the urine reached a maximum of about 85%, most
of this occurring within 18 hours of dosing. The amount of label in
the feces was low, never exceeding 10% of the dose. This indicated
that absorption was at least 90% complete.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 Nemec et al. (1968) monitored cholinesterase (ChE) levels in two
workers (entomologists) who examined plants in a cotton field after
it had been sprayed with an ultra-low-volume (nonaqueous) preparation
of methyl parathion (1.5 to 2 Ib/acre). The men entered a cotton
field to examine the plants on 3 different days over a 2-week period;
two of these occasions were within 2 hours after the ultra-low-volume
spraying, and the third occasion was 24 hours after a spraying.
After each field trip their arms were washed with acetone and the
adhering methyl parathion determined. It was found that contact with
the plants 2 hours after spraying resulted in 2 to 1 0 r.g of methyl
parathion residue on the arms; exposure 24 hours after spraying
resulted in a residue on the arms of 0.16 to 0.35 mg. The amount of
pesticide absorbed was not estimated. No toxic symptoms were experienced
by either man, but measurement of red blood cell ChE activity immediately
after the third of these exposures showed a decrease in activity to
60 to 65% of preexposure levels. These values did not increase
significantly over the next 24 hours. It was concluded that workers
should not enter such a field until more than 24 hours, and preferably
48 hours, have elapsed after spraying with ultra-low-volume insecticide
sprays. Water emulsion sprays were not tested, but the authors
cautioned that it cannot be assumed that they are less hazardous than
the ultra-low-volume spray residues.
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Methyl Parathion August, 1987
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0 Rider et al. (1969, 1970, 1971) studied the toxicity of technical
methyl parathion (purity not specified) in human volunteers. Each
phase of the study was done with different groups of seven male
subjects, five of whom were test subjects and two were vehicle
controls (Rider et al., 1969). Each study phase was divided into a
30-day pre-test period for establishing cholinesterase baselines, a
30-day test period when a specific dose of methyl parathion was
given, and a post-test period.
0 Thirty-two different dosages were evaluated by Rider et al. (1969),
ranging from 1 to 19 rag/day. Early in the study, several of the
groups were given more than one dose level during a single phase.
The initial amount was 1.0 mg with an increase of 0.5 mg during each
succeeding test period up to 15.0 rag/day. At this point, the dose was
increased by 1.0 ing/day to a total dose of 19.0 mg/day. Pesticide in
corn oil was given orally in capsules, once per day for each test
period of 30 days. At no time during any of the studies were there
any significant changes in blood counts, urinalyses, or prothrombin
times, or was there any evidence of toxic side effects. Cholinesterase
activity of the plasma and red blood cells (RBCs) was measured twice
weekly prior to, during and after the dosing period. The authors
considered a mean depression of 20 to 25% or greater in ChE activity
below control levels to be indicative of the toxic threshold. At
11.0 mg/day, a depression of 15% in plasma ChE occurred, but doses up
to and including 19 mg/day did not produce any significant ChE
depression.
0 Rider et al. (1970) studied the effects of 22, 24 and 26 mg/day
technical methyl parathion. There were no effects observed at
22 mg/day. At 24 mg/day, plasma and RBC ChE depression was
produced in two subjects, the maximum decreases being 24 and 23% for
plasma, and 27 and 55% for RBC. The mean maximal decreases (in all
five subjects) were 17% for plasma and 22% for RBC. With 26 mg/day
RBC ChE depression was again produced in only two of the subjects,
with maximum decreases of 25 and 37%. The mean maximum decrease was
18%. Plasma cholinesterase was not significantly altered.
0 Rider et al. (1971) assessed the effects of 28 and 30 mg/day technical
methyl parathion. At 28 mg/day, a significant decrease in RBC ChE
was produced in three subjects (data not given), with a maximum mean
decrease of 19%. With a dose of 30 mVday, a mean maximum depression
of 37% occurred. Based on their criteria of 20 to 25% average
depression of ChE activity, the authors concluded that this was the
level of minimal incipient toxicity. Body weights of the test subjects
were not reported, but assuming an average body weight of 70 kg, a
dose of 22 mg/day corresponds to a No-observed-Adverse-Ef feet-Level1
(NOAEL) of 0.31 mg/kg/day, and the 30 mg/day dose corresponds to 0.43
mg/kg/day. The NOAEL is considered to be 22 mg/day herein because of
the apparent sensitivity of some individual subjects at higher doses
to have met the 20 to 25% criteria for ChE depression as an effect.
Long-term Exposure
0 No information was found in the available literature on the health
effects of methyl parathion in humans.
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Methyl Parathion August, 1987
-7-
Animals
Short-term Exposure
0 Reported oral LD5Q values for methyl parathion include 14 and 24 mg/kg
in male and female Sherman rats, respectively (Gaines, 1969); 14.5 and
19.5 mg/kg in male and female CD-I mice, respectively (Haley et al.,
1975); 30 mg/kg in male ddY mice (Isshiki et al., 1983); 18.0 and
8.9 mg/kg in male and female Sprague-Dawley rats, respectively (Sabol,
1985); and 9.2 mg/kg in rats of unreported strain (Galal et al., 1977).
0 Galal et al. (1977) determined the subchronic median lethal dose
(C-LD50) of methyl parathion (purity not specified) in adult albino
rats. Groups of 10 animals received an initial daily oral dose (by
gavage) of 0.37 mg/kg (4% of the acute oral LD50). Every 4th day the
dose was increased by a factor of 1.5 (dose based on the
body weight of the animals as recorded at 4-day intervals). Treatment
was continued until death or termination at 36 days. Hematological
and blood chemistry analyses were performed initially and on the 21st
and 36th days of the study. Histopathological studies of the liver,
kidneys and heart were also carried out on the 21st and 36th days of
treatment. The C-LDso obtained was 13 mg/kg. The authors concluded
that the most predominant hazards of subchronic exposure to methyl
parathion were weight loss, hyperglycemia and macrocytic anemia, all
probably secondary to hepatic toxicity. Since an increasing dose
protocol was used, this study does not identify a NOAEL or a Lowest-
Observed-Adverse-Effect-Level (LOAEL).
0 Daly et al. (1979) administered methyl parathion (technical, 93.65%
active ingredient) to Charles River CD-I mice for 4 weeks at levels
of 0, 25 or 50 ppm in the diet. Assuming that 1 ppm in the diet of
mice corresponds to 0.15 mg/kg/day (Lehman, 1959), this is equivalent
to doses of about 0, 3.75 or 7.5 mg/kg/day. Five animals of each sex
were used at each dose level. Mean body weights were lower (p <0.05)
than control for all treated animals throughout the test period. Mean
food consumption was lower (p <0.05) throughout for all test animals
except females at the 25-ppm level. Mortality, physical observations,
and gross postmortem examinations did not reveal any treatment-related
effects. Cholinesterase measurements were not performed. Based on
body weight gain, the LOAEL for this study was identified as 25 ppm
(3.75 mg/kg/day).
0 Tegeris and Underwood (1977) examined the effects of feeding methyl
parathion (94.32%,pure) to beagle dogs (4 to 6 months of age, weighing
5 to 10 kg) for 14 days. Two animals of each sex were given doses
of 0, 2.5, 5 or 10 mg/kg/day. All animals survived the 14-day test
period. Mean feed consumption and weight gain were significantly
(p <0.05) depressed for both sexes at the 5 and 10 mg/kg/day dose
levels. After the 3rd day, animals in the high-dose group began
vomiting after all meals. Vomiting was observed sporadically at the
lower dose levels, particularly during the 2nd week. The authors
attributed this to acetylcholinesterase inhibition, but no measure-
ments were reported. No other symptomatology was described. Based
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Methyl Parathion August, 1987
-8-
on weight loss and vomiting, this study identified a LOAEL of
2.5 mg/kg/day in the dog.
0 Fan et al. (1978) investigated the immunosuppressive effects of methyl
parathion administered orally to Swiss (ICR) mice. The pesticide
(purity not specified) was fed in the diet at dose levels corresponding
to 0, 0.08, 0.7 or 3.0 mg/kg/day for 4 weeks. Active immunity was
induced by weekly injection of vaccine (acetone-killed Salmonella
typhimurium) during the period of diet treatment. Defense against
microbial infection was tested by intraperitoneal injection of a
single LD5Q dose of active £. typhimurium cells. Protection by
immunization was stated to be decreased in methyl parathion-treated
animals, but no dose-response data were provided. The authors stated
that pesticide treatment extending beyond 2 weeks was required to
obtain significant increases in mortality. Increased mortality was
associated with an increased number of viable bacteria in blood,
decreased total gamma-globulins and specific immunoglobins in serum,
and reduced splenic blast transformation in response to mitogens.
0 Shtenberg and Dzhunusova (1968) studied the effect of oral exposure to
methyl parathion (purity not specified) on immunity in albino rats
vaccinated with MUSI polyvaccine. Three tests (six animals each)
were conducted in which: (a) the vaccination was done after the
animals had been on a diet supplying 1.25 mg/kg/day metaphos (methyl
parathion) for 2 weeks; (b) the diet and vaccinations were initiated
simultaneously; and (c) the diet was initiated 2 weeks after vaccina-
tion. The titer of agglutins in immunized control rats was 1:1,200.
This titer was decreased in all exposed groups as follows: 1:46 in
series (a), 1:75 in series (b) and 1:33.3 in series (c). The authors
judged this to be clear evidence of inhibition of immunobiological
reactivity in the exposed animals. Changes in blood protein fractions
and in serum concentration of albumins were not statistically significant.
Based on immune suppression, a LOAEL of 1.25 mg/kg/day was identified.
Dermal/Ocular Effects
0 Gaines (1969) reported a dermal LD5Q of 67 mg/kg for methyl parathion
in male and female Sherman rats.
0 Galloway (1984a,b) studied the skin and eye irritation properties of
methvl parathion (technical; purity not specified) using albino New
Zealand White rabbits. In the skin irritation test, 0.5 mL undiluted
pesticide was applied and the treated area occluded for 4 hours.
This treatment resulted in dermal edema that persisted for 24 hours,
and in erythema that lasted for 6 days. After a total observation
period of 9 days, a score of 2.0 was derived, and technical methyl'
parathion was rated as a weak irritant. In the eye irritation test,
0.1 mL of the undiluted pesticide was applied to nine eyes. Three
were washed after exposure, and six were left unwashed. Conjunctival
irritation was observed starting at 1 hour and lasting up to 48 hours
postexposure. Maximum average irritation scores of 11 and 10.7 were
assigned for nonwashed and washed eyes, respectively, and technical
methyl parathion was considered a weak irritant.
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Methyl Parathion August, 1987
-9-
0 Galloway (1985) used guinea pigs to examine the sensitizing potential
of methyl parathion (technical; purity not stated). Ten doses of
0.5 mL of a 10% solution (w/v in methanol) were applied to the clipped
intact skin of 10 male guinea pigs (albino Hartley strain) over a
36-day period. This corresponds to an average dose of 13.9 mg/kg/day.
Another group was treated with 2,4-dinitrochlorobenzene as a positive
control. No skin sensitization reaction was observed in methyl
parathion-treated animals.
0 Skinner and Kilgore (1982) studied the acute dermal toxicity of methyl
parathion in male Swiss-Webster mice, and simultaneously determined
ED50 values for cholinesterase and acetylcholinesterase inhibition.
Methyl parathion (analytical grade, 99% pure) was administered in
acetone solution to the hind feet of the mice; the animals were
muzzled to prevent oral ingestion through grooming. The dermal LD50
was 1,200 mg/kg. The ED50 was 950 mg/kg for cholinesterase inhibition
and 550 mg/kg for acetylcholinesterase inhibition.
Long-term Exposure
0 Daly and Ranchart (1980) conducted a 90-day feeding study of methyl
parathion (93.65% pure) using Charles River CD-1 mice. Groups of 15
mice of each sex were given diets containing the pesticide at levels
of 0, 10, 30 or 60 ppm. Assuming that 1 ppm in the diet of mice corre-
sponds to 0.15 mg/kg/day (Lehman, 1959), this is equivalent to doses
of about 0, 1.5, 4.5 or 9.0 mg/kg/day. All mice survived the test.
Mean body weights were significantly (p <0.05) depressed for both
sexes at 60 ppm throughout the study and for males during the first
5 weeks at 30 ppm. Animals of both sexes had a slight but not
significant (p >0.05) increase in the mean absolute and relative
brain weights at 60 ppm. There were dose-related decreases (p <0.05)
in the mean absolute and relative testes weights of all treated
males and in the ovary weights of the females at 30 and 60 ppm.
Gross and microscopic examination revealed no dose-related effects.
Histological examination revealed no findings in the brain, testes or
ovary to account for the observed changes in the weights of these
organs. Measurements on ChE were not performed. Based on decreased
testes weight, the LOAEL for this study was 10 ppm (1.5 mg/kg/day).
0 Tegeris and Underwood (1978) investigated the toxicity of methyl
parathion (94.32% active ingredient) in beagle dogs fed the pesticide
for 90 days at dose levels of 0, 0.3, 1.0 or 3.0 mg/kg/day. Four dogs
(4-months old, 4.5 to 8.0 kg) of both sexes were used at each dose
level. Soft stools were observed in all treatment groups throughout,
and there was also occasional spontaneous vomiting. There were no
persistent significant (p >0.05) effects on body weight gain, feed
intake, fasting blood sugar, BUN, SGPT, SCOT, hematological, or
urological indices. Organ weights were within normal limits, with
the exception of pituitary weights of females at 3.0 mg/kg, which
were significantly (p <0.05) higher than the control values. Gross
and microscopic examination revealed no compound-related abnormalities!
Plasma ChE was significantly (p <0.05) depressed in both sexes at 6
and 13 weeks at 3 mg/kg/day, and in the males only at 1.0 mg/kg/day
-------
Methyl Parathion August, 1987
-1 fl-
at 13 weeks; erythrocyte ChE was also significantly (p <0.05) depressed
in all animals at 6 and 13 weeks at 3 mg/kg/day, and in both sexes at
13 weeks at 1.0 mg/kg/day; brain ChE was significantly (p <0.05)
depressed in both sexes at 3.0 mg/kg/day. Based on ChE depression,
the NOAEL and LOAEL for this study were identified as 0.3 mg/kg/day
and 1.0 mg/kg/day, respectively.
0 Ahmed et al. (1981) conducted a 1-year feeding study in beagle dogs.
Methyl parathion (93.6% pure) was administered in the diet at ingested
dose levels of 0, 0.03, 0.1 or 0.3 mg/kg/day. Eight animals of each
sex were included at each dose level, with no overt signs of toxicity
noted at any dose. There were no treatment-related changes in food
consumption or body weight. Cholinesterase determinations in plasma,
red blood cells and brain revealed marginal variations, but the
changes were not consistent and were judged by the authors to be
unrelated to dosing. Organ weight determinations showed changes in
both males and females at 0.1 and 0.3 mg/kg/day, but the changes were
neither dose-related nor consistent. It was concluded that there was
no demonstrable toxicity of methyl parathion fed to the dogs at these
levels. The NOAEL for this study was 0.3 mg/kg/day.
0 NCI (1978) conducted a 2-year feeding study of methyl parathion
(purity not specified) in F344 rats (50/sex/dose) at dose levels of
0, 20 or 40 ppm in the diet. Assuming that 1 ppm in the diet of rats
corresponds to 0.05 mg/kg/day (Lehman, 1959), this is equivalent to
dose levels of about 0, 1 or 2 mg/kg/day. Cholinesterase levels were
not measured, but no remarkable clinical signs were noted, and no
significant (p <0.05) changes were observed in mortality, body weight,
gross pathology or histopathology. Based on this, a NOAEL of 40 ppm
(2 mg/kg/day) was identified in rats.
• NCI (1978) conducted a chronic (105-week) feeding study in B6C3F!
mice (50/sex/dose). Animals were initially fed methyl parathion
(94.6% pure) at dose levels of 62.5 or 125 ppm. Assuming that 1 ppm
in the diet of mice corresponds to 0.15 mg/kg/day (Lehman, 1959),
this is equivalent to doses of about 9.4 or 18.8 mg/kg/day. Because
of severely depressed body weight gain in males, their doses were
reduced at 37 weeks to 20 or 50 ppm, and the time-weighted averages
were calculated to be 35 or 77 ppm. This corresponds to doses of
about 5.2 or 11.5 mg/kg/day, respectively. Females were fed at the
original levels throughout. Mortalif was significantly (p <0.05)
increased only in female mice at 125 ppm. Body weights were lower
(p <0.05) for both sexes throughout the test period and decreases
were dose-related. No gross or histopathologic changes were noted,
and ChE activity was not measured. Based on body weight, this study
identified a LOAEL of 35 ppm (5.2 mg/kg/day) in male mice.
0 Daly et al. (1984) conducted a chronic feeding study of methyl
parathion (93.65% active ingredient) in Sprague-Dawley (CO) rats
(60/sex/dose) at dose levels of 0, 0.5, 5 or 50 ppm in the diet.
Using food intake/body weight data given in the study report, these
levels approximate doses of about 0, 0.025, 0.25 or 2.5 mg/kg/day.
At 24 months, five animals of each sex were sacrificed for qualitative
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Methyl Parathion August, 1987
-11-
and quantitative tests for neurotoxicity. Ophthalmoscopic examinations
were conducted on females at 3, 12 and 24 months and terminally.
Hematology, urinalysis and clinical chemistry analyses were performed
at 6, 12, 18 and 24 months. Mean body weights were reduced (p <0.05)
throughout the study for both sexes at 50 ppm. At this dose level,
food consumption was elevated (p <0.05) for males during weeks 2
to 13, but reduced for females for most of the study. Hemoglobin,
hematocrit and RBC count were significantly (p <0.05) reduced for
females at 50 ppm at 6, 12, 18 and 24 months. For males at 5 and
50 ppm at 24 months, hematocrit and RBC count were significantly
(p <0.05) reduced and hemoglobin was reduced, but not significantly
(p >0.05). At 50 ppm, plasma and erythrocyte ChE were significantly
(p <0.05) depressed for both sexes during the test, and brain ChE was
significantly (p <0.05) decreased at termination. Slight decreases
in ChE activity were also observed in animals at 5 ppm, but these
changes were not statistically significant (p >0.05). For males, the
absolute weight and the ratio to brain weight of the testes, kidneys
and the liver were reduced by 10 to 16% (not significant, p >0.05) in
both the 5- and 50-ppm groups, while for females absolute and organ/body
weights for the brain and heart (also heart/brain weight) were found
to be elevated significantly (p <0.05) at the same dose levels. Overt
signs of cholinergic toxicity (such as alopecia, abnormal gait and
tremors) were observed in the 50-ppm animals and in one female at
5 ppm. At 24 months, 15 females were observed to have retinal degen-
eration. There was also a dose-related occurrence of retinal posterior
subcapsular cataracts, possibly related or secondary to the retinal
degeneration, since 5 of the 10 cataracts occurred in rats with retinal
atrophy. The incidence of retinal atrophy was 20/55 at 50 ppm, 1/60 at
5 ppm, 3/60 at 0.5 ppm and 3/59 in the control group. Examination of
the sciatic nerve and other nervous tissue from five rats per sex
killed at week 106 gave evidence of peripheral neuropathy (abnormal
fibers, myelin corrugation, myelin ovoids) in both sexes at 50 ppm
(p <0.05). Too few fibers were examined at the lower doses to perform
statistical analyses, but the authors stated that nerves from both
sexes in low- and mid-dose groups could not be distinguished qualita-
tively from controls. Slightly greater severity of nerve changes
found in two males was not clearly related to treatment. No other
lesions were observed that appeared to be related to ingestion of
methyl parathion. Based on hematology, body weight, organ weights,
clinical chemistry, retinal degeneration and cholinergic signs, a
NOAEL of 0.5 ppm (0.025 mg/kg/day) was identified in this study.
Reproductive Effects
0 Lobdel and Johnston (1964) conducted a three-generation study in
Charles River rats. Each parental dose group included 10 males and
20 females. The investigators incorporated methyl parathion (99% pure)
in the diet of males and females at dose levels of 0, 10 or 30 ppm,
except for reduction of each dose by 50% during the initial 3 weeks
of treatment, to produce dose equivalents of 0, 1.0 and 3.0 mg/kg/day,
respectively. There was no pattern with respect to stillbirths,
although the 30-ppm groups had a higher total number of stillborn.
Survival was reduced in weanlings of the F]_a, F^ and F2a groups at
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Methyl Parathion August, 1987
-12-
30 ppm, and in weanlings of the F^a 9rouP at 10 PP"- At 30 ppm,
there was also a reduction in fertility of the F2b dams at the second
nating; the first mating resulted in 100% of the animals having
litters, while at the second mating, only 41% had litters. Animals
exposed to 1 0 ppm methyl par a th ion did not demonstrate significant
deviations from the controls. A NOAEL of 10 ppm (1.0 mg/kg/day) was
identified in this study.
0 Daly and Hogan (1982) conducted a two-generation study of methyl
parathion (93.65% pure) toxicity in Sprague-Dawley rats. Each parental
dose group consisted of 1 5 males and 30 females. The compound was
added to the diet at levels of 0, 0.5, 5.0 or 25 ppm. Using compound
intake data from the study report, equivalent dose levels are about
0, 0.05, 0.5 or 2.5 mg/kg/day. Feeding of the diet was initiated
14 weeks prior to the first mating and then continued for the remainder
of the study. Reduced body weight (p <0.05) was observed in FQ and
FI dams at the 25-ppm dose level. A slight decrease in body weight
was noted in Fla and F2a pups in the 25-ppm group, but this was not
significant (p >0.05). Overall, the authors concluded that there was
no significant (p >0.05) effect attributable to methyl parathion in
the diet. Based on maternal weight gain, the NOAEL for this study
was 5.0 ppm (0.5 mg/kg/day).
Developmental Effects
0 Gupta et al. (1985) dosed pregnant Wistar-Furth rats (10 to 12 weeks
of age) with methyl parathion (purity not specified) on days 6 to 2C
of gestation. Two doses were used: 1.0 mg/kg (fed in peanut butter)
or 1.5 mg/kg (administered by gavage in peanut oil). The low dose
produced no effects on maternal weight gain, caused no visible signs
of cholinergic toxicity and did not result in increased fetal resorp-
tions. The high dose caused a slight but significant (p <0.05)
reduction in maternal weight gain (11% in exposed versus 16% in
controls, by day 15) and an increase in late resorptions (25% versus
0%). The high dose also resulted in cholinergic signs (muscle fasicu-
lation and tremors) in some dams. Acetylcholesterase (AChE) activity,
choline ace tyltransf erase (CAT) activity, and quinuclidinyl benzilate
(QNB) binding to muscarinic receptors were determined in several
brain regions of fetuses at 1 , 7, 14, 21 and 28 days postnatal age,
and in maternal brain at day 19 of gestation. Exposure to 1 . 5 mg/kg
reduced (p <0.05) the AChE and increased CAT activity in all fetal
brain regions at each developmental period and in the maternal brain.
Exposure to 1.0 mg/kg caused a significant (p <0.05) but smaller and
less persistent reduction of AChE activity in offspring, but no change
in brain CAT activity. Both doses reduced QNB binding in maternal
frontal cortex (p <0.05), but did not alter the postnatal pattern of
binding in fetuses. In parallel studies, effects on behavior (cage
emergence, accommodated locomotor activity, operant behavior) were
observed to be impaired in rats exposed prena tally to 1.0 mg/kg, but
not to the 1.5-mg/kg dose. No morphological changes were observed in
hippocampus or cerebellum. It was concluded that subchronic prenatal
exposure to methyl parathion altered postnatal development of
cholinergic neurons and caused subtle alterations in selected
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Methyl Parathion August, 1987
-13-
behaviors of the offspring. The fetotoxic LOAEL for this study was
1.0 ing/kg.
0 Gupta et al. (1984) administered oral doses of 1.0 or 1.5 mg/kg/day
of methyl parathion (purity not specified) to female Wistar-Furth rats
on days 6 through 1 5 or on days 6 through 19 of gestation. Protein
synthesis in brain and other tissues was measured on day 15 or day 19
by subcutaneous injection of radioactive valine. The specific activity
of this valine in the free amino acid pool and protein-bound pool
(measured 0.5, 1.0 and 2.0 hours after injection) was significantly
(p <0.05) reduced in various regions of the maternal brain and in
maternal viscera, placenta and whole embryos (day 15), and in fetal
brain and viscera (day 19). The inhibitory effect of methyl parathion
on protein synthesis was dose dependent, greater on day 19 than on
day 15 of gestation and more pronounced in fetal than in maternal
tissues. With respect to protein synthesis in both maternal and
fetal tissues, the LOAEL of this study was 1.0 mg/kg.
Mutagenicity
0 Van Bao et al. (1974) examined the lymphocytes from 31 patients exposed
to various organophosphate pesticides for indications of chromosome
aberrations. Five of the examined patients had been exposed to methyl
parathion. Blood samples were taken 3 to 6 days after exposure and
again at 30 and 180 days. A temporary, but significant (p <0.05)
increase was found in the frequency of chromatid breaks and stable
chromosome-type aberrations in acutely intoxicated persons. Two of
the methyl parathion-exposed persons were in this category, having
taken large doses orally in suicide attempts. The authors concluded
that the results of this study strongly suggest that the organic
phosphoric acid esters exert direct mutagenic effects on chromosomes.
0 Shigaeva and Savitskaya (1981) reported that metophos (methyl para-
thion) induced visible morphological mutations and biochemical mutations
in Pseudomonas aeruginosa at concentrations between 100 and 1,000 ug/mL,
and significantly (p <0.05) increased the reversion rate in Salmonella
typhimurium at concentrations between 5 and 500 ug/mL.
0 Grover and Malhi (1985) examined the induction of micronuclei in bone
marrow cells of Wistar male rats that had been injected with methyl
parathion at doses between one-third and one-twelfth of the LD5Q*
The increase in micronuclei formation led the authors to conclude
that methyl parathion has high mutagenic potential.
0 Mohn (1973) concluded that methyl parathion was a probable mutagen,
based on the ability to induce 5-methyltryptophan resistance in
Escherichia coli. Similar results were obtained using the streptomycin-
resistant system of JS. coli and the trp-conversion system of Saccharo-
myces cerevisiae.
0 Rashid and Mumma (1984) found methyl parathion to be mutagenic to £.
typhimurium strain TA100 after activation with rat liver microsomal
and cytosolic enzymes.
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Methyl Parathion August, 1987
-1 4-
0 Chen et al. (1981) investigated sister-chromatid exchanges (SCE) and
cell-cycle delay in Chinese hamster cells (line V79) and two human
cell lines (Burkitt lymphoma B35N and normal human lymphoid cell
Jeff), and found methyl parathion to be the most active pesticide
of eight tested with respect to its induction potential.
0 Riccio et al. (1981) found methyl parathion to be negative in two
yeast assay systems (diploid strains 03 and D7 of Saccharomyces
cerevisiae), based on mitotic recombination (in D3), and mitotic
crossing over, mitotic gene conversion, and reverse mutation (in D7).
Carcinogenici ty
0 NCI (1978) conducted chronic (105-week) feeding studies of methyl
parathion in F344 rats and B6C3F^ mice (50/sex/dose). Rats were fed
methyl parathion (94.6% pure) at dose levels of 0, 20 or 40 ppm
(equivalent to doses of 0, 1 or 2 mg/kg/day). Mice were initially
fed dose levels of 62.5 or 125 ppm, but because of severely depressed
body weight gain in males, their doses were reduced at 37 weeks to
20 or 50 ppm, respectively. Time-weighted averages for males were
calculated to be 35 or 77 ppm (about 5.2 or 11.5 mg/kg/day). Females
received the original dose level throughout. Based on gross and
histological examinations, no tumors were observed to occur at an
incidence significantly higher than that of the control value in either
the mice or rats. The authors concluded that methyl parathion was
not carcinogenic in either species under the conditions of the test.
0 Daly et al. (1984) fed Sprague-Dawley rats (60/sex/dose) methyl
parathion (93.65%) in the diet for 2 years. Doses tested were 0,
0.5, 5 or 50 ppm, estimated as equivalent to doses of 0, 0.025, 0.25
or 2.5 mg/kg/day. There were no significant (p >0.05) increases in
neoplastic lesions between treated and control groups.
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/Ij ( ug/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).
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Methyl Parathion
August, 1987
-15-
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
No data were located in the available literature that were suitable for
deriving a One-day HA value. It is recommended that the Ten-day HA value for
the 10-kg child (0.31 mg/L calculated below) be used at this time as a
conservative estimate of the One-day HA value.
Ten-day Health Advisory
The studies by Rider (1969, 1970, 1971) have been selected to serve as
the basis for calculation of the Ten-day HA for methyl parathion. In these
studies, human volunteers ingested methyl parathion for 30 days at doses
ranging from 1 to 30 mg/day. The most sensitive indicator of effects was
inhibition of plasma ChE. No effects in any subject were observed at a dose
of 22 mg/day (about 0.31 mg/kg/day with assumed 70-kg body weight), and this
was identified as the NOAEL. Doses of 24 mg/day inhibited ChE activity in
plasma and red blood cells in two of five subjects, maximum decreases being
23 and 24% in plasma and 27 and 55% in red blood cells. Higher doses (26 to
30 mg/day) caused greater inhibition. On this basis, 24 mg/day (0.34 mg/kg/day)
was identified as the LOAEL. Short-term toxicity or teratogenicity studies
in animals identified LOAEL values of 1.0 to 2.5 mg/kg/day (Gupta et al.,
1984, 1985; Shtenberg and Dzhunusova, 1968; Tegeris and Underwood, 1977), but
did not identify a NOAEL value.
Using a NOAEL of 0.31 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA
(0.31 mq/kg/day) (10 kg)
(10) (1 L/day)
0.31 mg/L (310.0 ug/L)
where:
0.31 mg/kg/day - NOAEL, based on absence of toxic effects or inhibition
of ChE in humans exposed orally for 30 days.
10 kg = assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from a study in humans.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The 90-day feeding study in dogs by Tegeris and Underwood (1978) has
been selected to serve as the basis for calculation of the Longer-term HA
for methyl parathion. In this study, a NOAEL of 0.3 mg/kg/day was identified,
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Methyl Parathion August, 1987
-16-
based on absence of effects on body weight, food consumption, clinical chem-
istry, hematology, urinalysis, organ weights, gross pathology, histopathology
and ChE activity. The LOAEL, based on ChE inhibition, was 1.0 mg/kg/day.
These values are supported by the results of Ahmed et al. (1981), who
identified a NOAEL of 0.3 mg/kg/day in a 1-year feeding study in dogs, and
by the study of Daly and Rinehart (1980), which identified a LOAEL of
1.5 mg/kg/day (based on decreased testes weight) in a 90-day feeding study in
mice.
Using a NOAEL of 0.3 mg/kg/day, the Longer-term HA for a 1 0-kg child is
calculated as follows:
Longer-term HA = (0.3 mg/kg/day) (10 kg) = 0.03 mg/L (30 ug/L)
y (100) (1 L/day)
where:
0.3 mg/kg/day = NOAEL, based on absence of effects on body weight,
food consumption, clinical chemistry, hematology,
urinalysis, organ weights, gross pathology, histo-
pathology and ChE activity in dogs fed methyl parathion
for 90 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL. from an animal study.
1 L/day = assumed daily water consumption of a child.
Using a NOAEL of 0.3 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
where:
Longer-term HA = (0'3 "gAg/day) (70 kg) . 0. 1 3 mg/L ( T 00 ug/L)
(100) (2 L/day)
0.3 mg/kg/day = NOAEL, based on absence of effects on body weight,
food consumption, clinical cnemistry, hematology,
urinalysis, organ weights, gross pathology, histo-
pathology and ChE activity in dogs fed methyl parathion
for 90 days.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption by an adult.
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Methyl Parathion August, 1987
-17-
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study in rats by Daly et al. (1984) has been selected
to serve as the basis for calculation of the Lifetime HA for methyl parathion.
In this study, a NOAEL of 0.025 mg/kg/day was identified, based on the absence
of effects on body weight, organ weights, hematology, clinical chemistry, retinal
degeneration and cholinergic signs. A LOAEL of 0.25 mg/kg/day was identified,
based on decreased hemoglobin, red blood cell counts, and hematocrit (males),
changes in organ-to-body weight ratios (males and females) and one case of
visible cholinergic signs. There was increased retinal degeneration at
2.5 mg/kg/day, but this was not greater than control at 0.25 or 0.025 mg/kg/day.
This LOAEL value (0.25 mg/kg/day) is lower than most other NOAEL or LOAEL
values reported in other reports. For example, NOAEL values of 0.3 to 3.0
nig/kg/day have been reported in chronic studies by Ahmed et al. (1981), NCI
(1978), Lobdell and Johnston (1964) and Daly and Hogan (1982).
Using a NOAEL of 0.025 mg/kg/day, the Lifetime HA for a 70-kg adult is
calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0-025 mg/kg/day) = 0.00025 nig/kg/day
(100)
where:
0.025 mg/kg/day = NOAEL, based on absence of cholinergic signs or
other adverse effects in rats exposed to methyl
parathion in the diet for 2 years.
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Methyl Parathion August, 1987
-18-
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 = (°'OOQ25 mg/kg/day) (70 kg) = 0.009 mg/L (9 ug/L)
(2 L/day)
where:
0.00025 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 = (0.009 mg/L) (20%) = 0.002 mg/L (2 ug/L)
where:
0.009 mg/L = DWEL.
20% = relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcinogenic activity was detected in either rats or
mice in a 105-week feeding study (NCI, 1978).
0 Statistically significant (p <0.05) increases in neoplasm frequency
were not found in a 2-year feeding study in rats (Daly et al., 1984).
0 The International Agency for Research on Cancer (IARC) has not
evaluated the carcinogenicity of methyl parathion.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), methyl parathion may be classified
in Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 NAS (1977) concluded that data were inadequate for calculation of an
ADI for methyl parathion. However, using data on parathion, NAS
calculated an ADI for both parathion and methyl parathion of 0.0043
mg/kg/day, using a NOAEL of 0.043 mg/kg/day in humans (Rider et al.,
1969) and an uncertainty factor of 10 (NAS, 1977). Prom this ADI,
NAS calculated a chronic Suggested-No-Ad verse-Response Level (SNARL)
of 0.03 mg/L, based on water consumption of 2 L/day by a 70-kg adult,
and assuming a 20% RSC.
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Methyl Parathion August, 1987
-19-
0 The U.S. EPA Office of Pesticide Program (EPA/OPP) previously calcu-
lated a provisional ADI (PAOI) of 0.0015 mg/kg/day, based on a NOAEL
of 0.3 mg/kg/day. This is based on the 90-day dog study by Tegeris
and Underwood (1978) and a 200-fold uncertainty factor. This PADI
has been updated to use a value of 0.0025 mg/kg/day based on a NOAEL
of 0.0250 mg/kg/day in a 2-year rat chronic feeding study and a
100-fold uncertainty factor.
0 ACGIH (1984) has proposed a time-weighted average threshold limit
value of 0.2 mg/m3.
0 The National Institute for Occupational Safety and Health has recom-
mended a standard for methyl parathion in air of 0.2 mg/ra3 (TDB, 1985).
0 The U.S. EPA has established residue tolerances for parathion-and
methyl parathion in or on raw agricultural commodities that range
from 0.1 to 0.5 ppm (CFR, 1985). A tolerance is a derived value
based on residue levels, toxicity data, food consumption levels,
hazard evaluation and scientific judgment; it is the legal maximum
concentration of a pesticide in or on a raw agricultural commodity or
other human or animal food (Paynter et al., undated).
0 The World Health Organization established an ADI of 0.02 mg/kg/day
(Vettorazi and van den Hurk, 1985).
VII. ANALYTICAL METHODS
0 Analysis of methyl parathion is by a gas chromatographic (GC) method
applicable to the determination of certain nitrogen-phosphorus
containing pesticides in water samples (U.S. EPA, 1986b). In this
method, approximately 1 liter of sample is extracted with methylene
chloride. The extract is concentrated and the compounds are separate.3
using capillary column LGC. Measurement is made using a nitrogen-
phosphorus detector. The method detection limit has not been determined
for methyl parathioi, but it is estimated that the detection limits
for analytes included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) and
reverse osmosis (RO) will effectively remove methyl parathion from
water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
methyl parathion and methyl parathion diazinion bi-solute solutions.
As expected, the bi-solute solution showed a lesser overall carbon
capacity than that achieved by the application of pure solute solution.
8 Under laboratory conditions, GAC removed 99+% of methyl parathion
(Whittaker et al., 1982).
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Methyl Parathion August, 1987
20
IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists, Inc.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed. Cincinnati, OH: ACGIH.
Ackermann, H., and R. Engst. 1970. The presence of organophosphorus
insecticides in the fetus. Arch. Toxikol. 26(1):17-22. (In German)
Agchem.* 1983. Persistence and release rate of Penncap M insecticide in
water and hydrosoil: Project No. WT-5-82. Unpublished study.
Ahmed, F.E., J.W. Sagartz and A.S. Tegeris.* 1981. One-year feeding study in
dogs. Pharmacopathics Research Laboratories Inc., Laurel, Maryland for
Monsanto Company. Unpublished study. MRID 00093895.
Braeckman, R.A., F. Audenaert, J. L. Willems, F. M. Belpaire and M.G. Bogaert.
1983. Toxicokinetics of methyl parathion and parathion in the dog after
intravenous and oral administration. Arch. Toxicol. 54:71-82.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.121. July 1, 1985.
p. 484.
CHEMLAB. 1985. The chemical information system. CIS, Inc., Bethesda, MD.
Chen, H.H., J.L. Hsueh, S.R. Sirianni and C.C. Huang. 1981. Induction of
sister-chromatid exchanges and cell cycle delay in cultured mammalian cells
treated with eight organophosphorus pesticides. Mut. Res. 88:307-316.
Chian, E.S., W.N. Bruce and H.H.P. Fang. 1975. Removal of pesticides by
reverse osmosis. Environ. Sci. and Tech. 9(1)s52-59.
Ciba-Geigy Corporation.* 1978. Residue of CGA-15324 Curacron* (R) +4E and
methyl parathion 4E on soil. Compilation; unpublished study, including
AG-A Nos. 4635 I, II, II, and 5023.
Daly, I., and G. Hogan.* 1982. A two-generation reproduction study of methyl
parathion in rats. Bio/Dynamics, Inc. for Monsanto Company. Unpublished
study. MRID 00119087.
Daly, I., G. Hogan and J. Jackson.* 1984. A two-year chronic feeding study
of metnyl parathion in rats. Bio/Dynamics, Inc. for Monsanto Company.
Unpublished study. MRID 00139023.
Daly, I.W., and W.E. Rinehart.* 1980. A three month feeding study of methyl
parathion in mice. Bio/Dynamics, Inc., for Monsanto Company. Unpublished
study. MRID 00072513.
Daly, I.W., w.E. Rinehart and M. Cicco.* 1979. A four week pilot study in
mice with methyl parathion. Bio/Dynamics, Inc., for Monsanto Company.
Unpublished study. MRID 00072514.
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Methyl Parathion August, 1987
21
Fan, A., J.C. Street and R.M. Nelson. 1978. Immunosuppression in mice
administered methyl parathion and carbofuran by diet. Toxicol. Appl.
Pharmacol. 45(1):235.
Gabovich, R.D., and I.L. Kurennoy. 1974. Ozonation of water containing
humic compounds, phenols and pesticides. Army Medical Intelligence and
Information Agency. USAMIIA-K-4564.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol.
14:515-534.
Galal, E.E., H.A. Samaan, S. Nour El Dien, S. Kamel, M. El Saied, M. Sadek,
A. Madkour, K.H. El Saadany and A. El-Zawahry. 1977. Studies on the
acute and subchronic toxicities of some commonly used anticholinesterase
insecticides rats. J. Drug Res. Egypt. 9(1-2): 1-1 7.
Galloway, C.* 1984a. Rabbit skin irritation: methyl parathion technical
(Cheminova). STILLMEADOW, Inc., Houston, TX. for Gowan Company.
Unpublished study. MRID 00142804.
Galloway, C.* 1984b. Rabbit eye irritation: methyl parathion technical
(Cheminova). STILIMEADOW, Inc., Houston, TX. for Gowan Company.
Unpublished study. MRID 00142808.
Galloway, C.* 1985. Guinea pig skin sensitization: methyl parathion tech-
nical (Cheminova). STILI/1EADCW, Inc., Houston, TX. for Gowan Company.
Unpublished study. MRID 00142005.
Grover, I.S., and P.K. Malhi. 1985. Genotoxic effects of some organophos-
phorus pesticides. I. Induction of micronuclei in bone marrow cells in
rat. Mutat. Res. 155:131-134.
Gupta, R.C., R.H. Rech, K.L. Lovell, F. Welsch and J.E. Thornburg. 1985.
Brain cholinergic, behavior, and morphological development in rats exposed
in utero to methyl parathion. Toxicol. Appl. Pharmacol. 77:405-413.
Gupta, R.C., J.E. Thornburg, D.B. Stedman and D.B. Welsch. 1984. Effect of
subchronic administration of methyl parathion on in vivo protein synthesis
in pregnant rats and their conceptuses. Toxicol. Appl. Pharmacol.
72:457-468.
Haley, T.J., J.H. Farmer, J.R. Harmon and K.L. Dooley. 1975. Estimation of
the LDi and extrapolation of the LD0.i for five organothiophosphate
pesticides. J. Eur. Toxicol. 8(4):229-235.
Hawley, G.G. 1981. The Condensed Chemical Dictionary, 10th ed. NY: Van
Nostrand Reinhold Company.
Hollingworth, R.M., R.L. Metcalf and I.R. Fukuto. 1967. The selectivity of
sumithion compared with methyl parathion. Metabolism in the white mouse.
J. Agr. Food Chem. 15:242-249.
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Methyl Parathion August, 1987
22
Isshiki, K., K. Miyata, S. Matsui, M. Tsutsumi and T. Watanabe. 1983.
Effects of post-harvest fungicides and piperonyl butoxide on the acute
toxicity of pesticides in mice. Safety evaluation for intake of food
additives. III. Shokuhin Eiseigaku Zasshi. 24(3):268-274.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Lobdell, J.L., and C.D. Johnston.* 1964. Methyl parathion: three-generation
reproduction study in the rat. Virginia: Woodard Research Corporation
for Monsanto Company. Unpublished study. MRID 0081923.
Meister, R., ed. 1983. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Company.
Mohn, Go 1973. 5-Methyltryptophan resistance mutations in Escherichia coli
K-12: Mutagenic activity of monofunctional alkylating agents including
organophosphorus insecticides. Mut. Res. 20:7-15.
Nakatsugawa, T., N.M. Tolman and P.A. Dahm. 1968. Degradation and activa-
tion of parathion analogs by microsomal enzymes. Biochem. Pharmacol.
17:1517-1528.
NAS. 1977. National Academy of Sciences. Drinking water and health. Vol. 1.
Washington, DC: National Academy Press.
NCI. 1978. National Cancer Institute. Bioassay of methyl parathion for
possible carcinogenic!ty. Bethesda, MD: NCI, National Institutes of
Health. NCI-CG-TR-157.
Neal, R.A., and K.P. DuBois. 1965. Studies on the mechanism of detoxifi-
cation of cholinergic phosphorothioates. J. Pharmacol. Exp. Ther.
148(2):185-192.
Nemec, S.J., P.L. Adkisson and H.W. Dorough. 1968. Methyl parathion absorbed
oh the skin and blood cholinesterase levels of persons checking cotton
treated with ultra-low-volume sprays. J. Econ. Entomol. 61 (6): 1 740-1742.
Paynter, O.E., J.G. Cummings and M.H. Rogoff. Undated. United States
Pesticide Tolerance System. U.S. EPA, Office of Pesticide Programs,
Washington, DC. Unpublished draft report.
Pennwalt Corporation.* 1972. Soil and water run off test for Penncap M
versus methyl parathion E.G. Compilation. Unpublished study.
Pennwalt Corporation.* 1977. Penncap-M* (R)+ and Penncap-e* (TM)+ insecti-
cides—soil leaching. Unpublished study.
Rashid, K.A., and R.O. Mumma. 1984. Genotoxicity of methyl parathion in
short-term bacterial test systems. J. Environ. Sci. Health.
B19(6):565-577.
-------
Methyl Parathion August/ 1987
23
Riccio, E., G. Shepherd, A. Pomeroy, K. Mortelmans and M.D. Waters. 1981.
Comparative studies between the £. cerevisiae D3 and 07 assays of eleven
pesticides. Environ. Mutagen. 3(3):327.
Rider, J.A., H.C. Moeller, E.J. Puletti and J.I. Swader. 1969. Toxicity of
parathion, systox, octamethy1 pyrodophosphamide and methyl parathion in
man. Toxicol. Appl. Pharmacol. 14:603-611.
Rider, J.A., J.I. Swader and E.J. Puletti. 1970. Methyl parathion and
guthion anticholinesterase effects in human subjects. Federation Proc.
29(2):349. Abstracts.
Rider, J.A., J.I. Swader and E.J. Puletti. 1971. Anticholinesterase toxicity
studies with methyl parathion, guthion and phosdrin in human subjects.
Federation Proc. 30(2):443. Abstracts.
Sabol, E.* 1985. Rat: Acute oral toxicity of methyl parathion technical
(Cheminova). STILLMEADOW, Inc., Houston, TX. for Gowan Company.
Unpublished study. MRID 00142806.
Saunders, P.F. and J.N. Seiber. 1983. A chamber for measuring volatilization
of pesticides from model soil and water disposal systems. Chemosphere.
12(7/8):999-1012.
Shevchenko, M.A., P.N. Taran and P.V. Marchenko. 1982. Modern methods of
purifying water from pesticides. Soviet J. Water Chera. Techno1.
4(4):53-71.
Shigaeva, M.K. and I.S. Savitskaya. 1981. Comparative study of the mutagenic
effect of some organophosphorus insecticides on bacteria. Tsitol. Genet.
15(3):68-72.
Shtenberg, A.I. and R.M. Dzhunusova. 1968. Depression of immunological
reactivity in animals by some organophosphorus pesticides. Bull. Exp.
Biol. Med. 65(3):317-318.
Skinner, C.S. and W.W. Kilgore. 1982. Acute dermal toxicities of various
organophosphate insecticides in mice. J. Toxicol. Environ. Health.
9(3):491-497.
STORET. 1987.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
Tegeris, A.S. and P.C. Underwood.* 1977. Fourteen-day feeding study in the
dog. Pharmacopathics Research Laboratories, Laurel, MD., for Monsanto
Company. Unpublished study. MRID 00083109.
Tegeris, A.S. and P.C. Underwood.* 1978. Methyl parathion: Ninety-day
feeding to dogs. Pharmacopathics Research Laboratories, Inc., Laurel,
Maryland. Unpublished study. MRID 00072512.
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Methyl Parathion August, 1987
24
U.S. EPA. 1981. U.S. Environmental Protection Agency. Acephate, aldicarb,
carbophenothion, DEF, EPN, ethoprop, methyl parathion, and phorate:
their acute and chronic toxicity, bioconcentration potential, and
persistence as related to marine environment. Environmental Research
Laboratory. Unpublished study. Report No. EPA-600/4-81-023.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in ground
water by GC/NPD, January 1986 draft. Available from U.S. EPA's Environ-
mental Monitoring and Support -Laboratory, Cincinnati, Ohio.
Van Bao, T., I. Szabo, P. Ruzicska and A. Czeizel. 1974. Chromosome
aberrations in patients suffering acute organic phosphate insecticide
intoxication. Human Genetik 24(1): 33-57.
Vettorazzi, G. and G.W. van den Hurk, eds. 1985. Pesticides Reference Index.
J.M.P.R., p. 41.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. PhD. Thesis, Purdue
University.
Whittaker, K.F., J.C. Nye, R.F. Weekash, R.J. Squires, A.C. York and H.A.
Razemier. 1982. Collection and treatment of wastewater generated by
pesticide application. EPA-600/2-82-028, Cincinnati, Ohio.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
METOLACHLOR
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, Logifor 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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Metolachlor
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 51218-45-2
Structural Formula
2-Chloro-N-(2-ethyl-6-nethyiphenyl)-N-(2-methoxy-1-methylethyl) acetamide
Synonyms
• o-Acetanilide; 2-chloro-6'-ethyl-N-(2-methoxy-1-methylphenyl);
Dual8; Bleep9; Metetilachlor; Pimagram; Primextra; CGA-24705e
Uses (Meister, 1986)
0 Selective herbicide for pre-emergence and preplant incorporated weed
control in corn, soybeans, peanuts, grain sorghum, pod crops, cotton,
safflower, woody ornamentals, sunflowers and flax.
Properties (Meister, 1986; Ciba-Geigy, 1977; Hindholz et al., 1983; Worthing,
1983)
Chemical Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure (20°C)
Specific Gravity
Water Solubility (20°C)
Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C15H22N02C1
283.46
White to tan liquid
100°C (at 0.001 mm Hg)
1.3 x 10-5 mm Hg
530 mg/L
Occurrence
Metolachlor has been found-in 1,644 of 1,997 surface water samples
analyzed and in 45 of 239 ground water samples (STORET, 1987).
Samples were collected at 312 surface water locations and 297 ground
water locations, and Metolachlor was found in 14 states. The 85th
percentile of all nonzero samples was 11.5 ug/L in surface water and
0.25 ug/L in ground water sources. The maximum concentration found
was 138 ug/L in surface water and 0.25 ug/L in ground water.
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Metolachlor August, 1987
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0 Metolachlor residues resulting from agricultural use have also been
detected in ground water in Iowa and Pennsylvania with concentrations
ranging from 0.1 to 0.4 ppb.
Environmental Fate
(Forthcoming from OPP)
III. PHARMACOKINETICS
Absorption
0 In studies conducted by Hambock (1974a,b), rats were administered a
single oral dose (28.6 or 52.4 mgAg) of metolachlor (purity riot
specified, but 14C-labeled and unlabeled metolachlor were synthesized
for these experiments). The chemical was readily absorbed, since 70
to 90% of the metolachior was excreted as metabolites within 48 hours.
Distribution
0 Data from rats given radioactive metolachlor (approximately 3.2 to
3.5 mgAg) orally demonstrated that the chemical is rapidly metabolized.
Residues in meat tissues and blood were very low and only blood
contained residue levels in excess of 0.1 ppm (Hambock, 1974c).
Metabolism
0 Studies conducted to identify urinary and fecal metabolites in the
rat indicated that metolachlor is metabolized via dechlorination,
0-methylation, N-dealkylation and side-chain oxidation (Hambock, 1974
a,b). Urinary metabolites included 2-ethyl-6-methylhydroxyacetanilide
(MET-002) and N-(2-ethyl-6-methylphenyl)-N-(hydroxyacetyl)-DL-alanine)
(MET-004). Fecal metabolites included 2-chloro-N-(2-ethyl-6-methyl-
phenyl)-N-{2-hydroxy-1-methylethyl) (MET-003) and MET-004.
Excretion
When treated with 14c-metolachlor (approximately 31 mgAg orally),
male rats excreted 21.5% and 51.4% of the administered dose in the
urine and feces, respectively, in 48 hours (Hambock, 1974a,b). The
excreta contained 1, 15 and 22% of the administered dose as MET-002,
MET-003 and MET-004, respectively. No unchanged chemical was isolated,
and no glucuronide or sulfate conjugates were identified.
IV. HEALTH EFFECTS
Humans
Signs of human intoxication from metolachlor and/or its formulations
(presumably following acute deliberate or accidental exposures)
include abdominal cramps, anemia, ataxia, dark urine, methemoglobinemia,
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Metolachlor August, 1987
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cyanosis, hypothermia, collapse, convulsions, diarrhea, gastrointestinal
irritation, jaundice, weakness, nausea, shock, sweating, vomiting, CMS
depression, dizziness, dyspnea, liver damage, nephritis, cardiovascular
failure, skin irritation, dermatitis, sensitization dermatitis, eye
and mucous membrane irritation, corneal opacity and adverse reproductive
effects (HAZARDLINE, 1985).
Animals
Short-term Exposure
0 The acute oral LD50 of technical metolachlor [>90% active ingredient
(a.i.)] in the rat was reported to be 2,780 mg/kg (95% confidence
range of 2,180 to 3,545 mg/kg; Bathe, 1973).
0 Technical metolachlor in corn oil (>90% a.i.) was shown to be emetic
in beagle dogs, precluding the establishment of an LD50 (AMR, Inc.,
1974a). However, an "emetic dose" of 19 _+ 9.7 mg/kg was established.
0 Beagle dogs were fed technical metolachlor in the diet for 7 days in
a range-finding study (Goldenthal et al., 1979). Each test group
consisted of one male and one female. Doses were 1,000, 3,000 or
5,000 ppm with the controls receiving a basic diet plus the test
material solvent (ethanol). The mean doses were 0, 13.7, 22.7 or
40.2 mg/kg. Decreased food consumption and body weight indicated
that the two higher doses were unpalatable. No changes were observed
at the lowest dose, although the animals exhibited soft stools and/or
diarrhea over the study period. No other signs of overt toxicity,
morbidity or mortality were observed in any animal. Accordingly, the
lowest dose (13.7 mg/kg) is the NOAEL in this study.
Dermal/Ocular Effects
0 The LD50 of technical metolachlor (> 90% a.i.) in the rabbit when
tested by the unabraded dermal route is greater than 10,000 mg/kg
(AMR, Inc., 1974b).
0 Sachsse (1973b) evaluated the dermal irritation potential of technical
metolachlor (>90% a.i.) on the New Zealand rabbits. The chemical was
applied to abraded and unabraded skin for observation periods up to
72 hours. The results demonstrated that technical metolachlor is
non-irritating to rabbit skin.
0 Sachsse (1977) studied skin sensitization in the guinea pig by the
intradermal-injection method. Technical metolachlor (>90% a.i.)
dissolved in the vehicle (propylene glycol) and the vehicle alone
were intradermally injected into the skin of two groups of Pilbright
guinea pigs. A positive reaction was observed in the animals injected
with metolachlor in vehicle, but not in animals treated with the
vehicle alone. It was concluded that technical metolachlor is a skin
sensitizer.
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Metolachlor August, 1987
-5-
0 A study of eye irritation by technical metolachlor (>90% a.i.) in the
New Zealand White rabbit was conducted by Sachsse (1973a). The
chemical was applied at a dose level of 0.1 mL/eye. Evaluation of
both washed and unwashed eyes 24 hours and 7 days later revealed no
evidence of irritation.
Long-term Exposure
0 Beagle dogs (four/sex/dose) were administered technical metolachlor
(>90% a.i.) in their feed for up to 15 weeks (Coquet et al., 1974).
Initial doses were 0, 50, 150 or 500 ppm (equivalent to 0, 4 to 5,
or 14 to 19 mg/kg/day). However, after 8 weeks, the group receiving
50 ppm was switched to a diet that delivered 1,000 ppm (27 to
36 mg/kg/day) for the remaining 6 weeks. The dose was increased
because no signs of toxicity were observed in the 500-ppm group after
8 weeks. No animals died during the study and no significant changes
were observed in gross or histological pathology, blood or urine
analyses. Except for a decrease in food consumption and associated
slight weight loss at the 1,000-ppm dose, no compound-related effects
were observed. The NOAEL for this study is 500 ppm (14 to 19 mg/kg/day).
0 A 6-month feeding study in dogs was conducted at levels of 0, 100,
300 or 1,000 ppm (Jessup et al., 1979). The average compound consump-
tion was 0, 2.9, 9.7 or 29.6 mg/kg/day for the males and 0, 3, 8.8 or
29.4 mg/kg/day for the females, as determined by the investigators.
The control and high-dose groups consisted of eight animals/sex; the
low- and mid-dose groups consisted of six animals/sex. The extra
control and high-dose animals were used in a recovery period study
following sacrifice of the remaining animals at 6 months. The following
significant changes were observed at the end of the study. Mean body
weight gain was reduced in animals of both sexes fed 1,000 ppm; in
addition, food consumption was reduced in the females at this level.
Male dogs at the 300- and 1,000-ppm levels had significantly reduced
activated partial thromboplastin time (APTT) after 5 and 6 months of
observation. In females, significant changes in this parameter were
observed for dogs at month 4 fed 100 ppm, at month 6 at the 300 ppm
level, and at months 5 and 6 in the 1,000 ppm group. Additional
studies demonstrated that the changes were not attributable to the
pesticide. There were sporadic, but not treatment-related, changes
in platelet and red blood cell counts and hemoglobin over the course
of the study. Serum alkaline phosphate (SAP) levels decreased more
slowly in the test groups than in the controls. These changes were
significant in the groups fed 300 and 1,000 ppm. Therefore, the
NOAEL in this study was 100 ppm (3 mg/kg/day).
0 Tisdel et al. (1980) presented the results of a study in which
metolachlor (95% a.i.) was administered to Charles River CD-1 mice
(68/sex/dose) for 2 years at dietary concentrations of 0, 300, 1,000
or 3,000 ppm. Time-Weighted Average (TWA) concentrations, based upon
diet analyses, were equal to 0, 287, 981 and 3,087 ppm. The dietary
doses, from reported food intake and body weight data, were calculated
to be equal to 0, 50, 170 or 526 mg/kg/day for the males and 0, 64,
224 or 704 mg/kg/day for the females. No treatment-related effects
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Metolachlor August, 1987
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were observed in terms of physical appearance, food consumption,
heraatology, serum chemistry, urinalysis or gross or histopathology.
However, mortality was increased significantly in females fed
3,000 ppm (704 mgAg/day). Statistically significant reductions in
body weight gain were observed in both sexes at the highest dose.
Also, statistically significant changes in absolute and organ-to-body
weight ratios were noted occasionally (e.g. kidney- and liver-body
weight ratios and decreased seminal vesicle to body weight ratio in
high dose males). Based on this information, a NOAEL of 1,000 ppm
(170 mg/kg/day for males and 224 mg/kg/day for females) is identified.
0 Tisdel et al. (1983) presented the results of a study in which
metolachlor (purity not specified) was administered to CD-Crl:CD
(SD) BR rats for 2 years at dietary concentrations of 0, 30, 300
or 3,000 ppm. Assuming that 1 ppm in the diet of rats is equal to
0.05 mg/kg/day (Lehman, 1959), these dietary concentrations would be
equal to 0, 1.5, 15 or 150 mg/kg/day. The control and 3,000-ppm
groups consisted of 70 rats/sex. The 30- and 300-ppm groups consisted
of 60 rats/sex. No treatment-related effects were noted in terms of
mortality, organ weight and organ-to-body weight ratios. A variety
of differences in clinical pathology measurements was found between
control and treatment groups at various time intervals, but no
consistent dose-related effects were apparent with the exception of
a decrease in glutamic-oxaloacetic transaminase activity in high dose
males at 12 months. Mean body weights of high-dose females were
consistently less than controls from week 2 until termination of the
study. This difference was statistically significant (p <0.01) for
26 of the 59 intervals at which such measurements were made. Food
consumption in high-dose females also was generally less than controls.
Gross pathology findings were described by the investigators as being
unremarkable. Microscopically, atrophy of the testes with degenera-
tion of the tubular epithelium was noted to a greater extent in the
300- and 3,000-ppm groups than in the controls. Additionally, an
increased incidence of eosinophilic foci was observed in the livers
of both sexes exposed at 3,000 ppm. Based on this data, a NOAEL of
30 ppm (1.5 mg/kg/day) is identified.
Reproductive Effects
0 A three-generation rat reproduction study was reported by Smith and
Adler (1978). Targeted dietary exposures were 0, 30, 300 or 1,000
ppm. T.ie actual exposures were analyzed to be 0, 30, 250 or 850 ppm.
Assuming that 1 ppm equals 0.05 mg/kg/day (Lehman, 1959), the doses
were calculated to be 0, 1.5, 22.5 or 42.5 mg/kg bw/day. No adverse
effects were noted at any dose. A minimal NOAEL of 42.5 mg/kg is
identified for reproductive effects.
0 Smith et al. (1981) conducted a two-generation reproduction study
in which Charles River CD rats (15 males and 30 females/dose) were
administered technical-grade metolachlor (purity not specified) at
dietary doses of 0, 30, 300 or 3,000 ppm. The TWA concentrations of
metolachlor, based upon dietary analysis, were 0, 32, 294 or 959 ppm.
Assuming that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
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Metolachlor August, 1987
-7-
(Lehman, 1959), these dietary concentrations are approximately equal
to 0, 1.6, 14.7 or 48 mg/kg/day. Mating, gestation, lactation, and
female and male fertility indices were not affected in either generation
Additionally, pup survival was not affected. However, pup weights in
the 959-ppm dose group, but not the 32- and 294-ppm dose groups, were
significantly reduced in the F1a and F2a litters. Food consumption was
reduced significantly for FT females receiving 32 ppm (1.6 mg/kg/day)
and greater at various study intervals. Other effects that appeared
to be treatment-related included increased liver-to-body weight ratios
for both FT parental males and females at 1,000 ppm and increased
thyroid-to-body weight and thyroid-to-brain weight in Fj males at
1,000 ppm. Based on reduced pup weights, a reproductive NOAEL of
294 ppm (14.7 mg/kg/day) is identified.
0 Tisdel et al. (1980) gave metolachlor (95% a.i.) to CD-I mice
(68/sex/dose) in the food for 2 years at concentrations of 0, 300,
1,000 or 3,000 ppm (the TWAs based on diet analyses were 0, 287, 981
or 3,087 ppm and corresponded to 0, 50, 170 or 520 mg/kg/day in males
and to 0, 64, 224 or 704 mg/kg/day in the females). At the high dose,
males were found to have a reduced seminal vesical-to-body weight
ratio.
0 Tisdel et al. (1983) exposed CD-Crl:CD (SD) BR rats (70/sex/dose) to
metolachlor (purity not specified) in the diet for 2 years at 0, 30,
300 or 3,000 ppm (the doses correspond to 0, 1.5, 15 or 150 mg/kg/day).
They observed greater testicular atrophy and degeneration of the
tubular epithelium in the 300- and 3,000-ppm groups than in the
control group.
Developmental Effects
0 Fritz (1976) conducted a rat teratology study in which pregnant
females (25/dose level) were administered doses of technical metola-
chlor (purity not specified) orally at 0, 60, 180 or 360 mg/kg/day
during days 6 to 15 of gestation. No fetotoxic or developmental
effects were noted.
0 Lightkep et al. (1980) evaluated the teratogenic potential of metola-
chlor in New Zealand White rabbits (16/dose). The compound was
administered as a suspension in aqueous 0.75% nydroxymethylcellulose
at levels of 0, 36, 120 or 360 mg/kg/day. Single oral dcses were
given on days 6 to 18 of gestation. Abortions occurred in two rabbits:
one in the 120-mg/kg/day group on day 25 (one early resorption) and
one in the 360-mg/kg/day group on day 17 (one fetus) and day 20 (eight
additional implantations). They did not consider these abortions to
be treatment-related. Maternal toxicity (decreased food consumption
and pupillary constriction) was observed in animals receiving the two
highest doses. The highest dose group also exhibited blood in the
cage pan and body weight loss over the treatment period. No signifi-
cant developmental or fetotoxic effects were observed in the 319
fetuses, pups or late resorptions evaluated from all dose groups.
Thus, a minimal NOAEL of 360 mg/kg/day for fetotoxicity and a NOAEL
of 36 mg/kg/day for maternal toxicity were identified.
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Metolachlor August, 1987
-8-
Mutagenicity
o
Technical metolachlor (purity not specified) was tested in the Ames
Salmonella test system, using £5. typhimurium strains TA1535, TA1537,
TA98 and TA100 (Ami and Muller, 1976). No increase in mutagenic
response was observed, with or without microsomal activation, at
concentrations of 10, 100, 1,000 or 10,000 ug/plate. Toxicity was
observed at 1,000 and 10,000 ug/plate without activation and at
10,000 ug/plate with activation.
0 Ciba-Geigy (1976) reported the results of a dominant lethal study in
the mouse using technical metolachlor (purity not specified). The
compound was administered orally in single doses of 0, 100 or 300 mg/kg
to males that then were mated to untreated females over a period of
6 weeks. No evidence of adverse effects were observed, as expressed
by increased implantation loss or resorptions.
Carcinogenicity
0 Marias et al. (1977) presented the results of a study in which
technical-grade metolachlor (purity not specified) was administered
to Charles River CD-I mice (50/sex/dose) at dietary concentrations of
0, 30, 300 or 3,000 ppm. Assuming that 1 ppm in the diet of the mouse
is equal to 0«15 mg/kg/day (Lehman, 1959), these dietary levels are
approximately 0, 4.5, 150 or 450 mg/kg/day. Males received the test
material for 18 months; females received the test material for 20
months. Results of this study indicated no evidence of oncogenicity
in either sex.
0 Tisdel et al. (1980) presented the results of a study in which
metolachlor (95% a.i.) was administered to Charles River CD-1 mice
(68/sex/dose) for 2 years at dietary concentrations of 0, 300, 1,000
or 3,000 ppm. From food intake and body weight data, the doses were
calculated to be equal to 0, 50, 170 or 526 mg/kg/day for the males
and 0, 64, 224 or 704 mg/kg/day for the females. A statistically
significant increase in the incidence of alveolar tumors in high-dose
males was noted at the 18-month sacrifice; however, this effect was
not confirmed by the final sacrifice at 24 months or by total tumor
incidences for all animals.
0 In 1979, Ciba-Geigy reported the results of a study in which technical
metolachlor was administered to Charles River albino rats in their
diet for 2 years at doses equivalent to 0, 1.5, 15 or 50 mg/kg/day.
A statistically significant increase in the incidence of primary
liver tumors was observed in the high-dose females (15/60 compared
with 5/60 at mid doses and 3/60 at the low dose and control). These
tumors included hypertrophic-hyperplastic nodules, angiosarcoma,
cystic cholangioma and hepatocellular carcinoma. The variety of
tumor expression forms suggests that a variety of cell types and
locations may be affected in the liver.
8 Tisdel et al. (1983) presented the results of a study in which
metolachlor (purity not specified) was administered to CD-Crl:CD
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Metolachlor August, 1987
-9-
(SO) BR rats for 2 years at dietary concentrations of 0, 30, 300 or
3,000 ppm. These doses were assumed to be equal to 0, 1.5, 15 or
150 mg/kg/day. An increased incidence of proliferative hepatic
lesions (combined neoplastic nodules/carcinomas) was found in the
high-dose females at terminal sacrifice (p <0.018 by the Fisher exact
test). Six of the 60 had neoplastic nodules (p <0.05) and 7 of the
60 had liver tumors (one additional tumor was diagnosed as a carcinoma;
p <0.01).
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 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 child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
No suitable information was found in the available literature for
determination of a One-day HA for metolachlor. Accordingly, it is recommended
that the Ten-day HA value for the'lO kg child (1.4 mg/L, calculated below) be
used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The 7-day dietary study in dogs by Goldenthal et al. (1979) has been
selected to serve as the basis for the Ten-day HA. Doses were 1,000, 3,000
or 5,000 ppm with the controls receiving a basic diet plus the solvent (ethanol)
(one/sex/group). Actual mean doses were 0, 13.7, 22.7 or 40.2 mg/kg. The
results indicated that the two higher doses were unpalatable, resulting in
decreased food consumption and body weight. No changes were observed at the
lowest dose, although the animals exhibited soft stools and/or diarrhea over
the study period. No other signs of overt toxicity, morbidity or mortality
were observed in any animal. The lowest dose, 13.7 mg/kg/day, is identified
as the NOAEL.
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Metolachlor August, 1987
-10-
The Ten-day HA for a 10-Jcg child is calculated as follows:
Ten-day HA = (13.7 mg/kg/day) (10 kg) = K4 ng/L (1f40o ug/L)
(100) (1 L/day)
where:
13.7 mg/kg/day = NOAEL, based on absence of decreased food consumption
and body weight loss.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OOW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The study by Jessup et al. (1979) has been selected to serve as the
basis for the Longer-term HA. A 6-month feeding study in dogs was conducted
at average compound consumption levels of 0, 2.9, 9.7 and 29.6 mg/kg/day
(males) and 0, 3.0, 8.8 and 29.4 mg/kg/day (females). Significant changes
observed at the end of the study, included reduced mean body weight gain in
animals of both sexes fed 1,000 ppm and reduced food consumption in the
females at this level. Serum alkaline phosphate levels decreased more slowly
in the test groups than in the controls. These changes were statistically
significant in the groups fed 300 and 1,000 ppm. Therefore, the NOAEL in
this study is identified as 100 ppm (3 mg/kg/day).
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (3 »g/k9/day)(10 kg) . 0>3 /L (300 /L)
(100) (1 L/day)
where:
3 mg/kg/day = NOAEL.
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 a 70-kg adult is calculated as follows:
Longer-term HA = (3 mg/kg/day) (70 kg) = U05 mg/L (1 050 ug/L)
(100) (2 L/day)
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Metolachlor August. 1987
-11-
where:
3 mg/kg/day = NOAEL.
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.
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Tisdel et al. (1983) has been selected to serve as the
basis for the Lifetime HA. In this study, rats were given dietary doses of
metolachlor equivalent to 0, 1.5,'15 or 150 mg/kg/day. No treatment-related
effects were noted in terms of mortality, organ weight and organ-to-body
weight ratios. The investigators noted a statistically significant decrease
in glutamic-oxaloacetic transaminase activity in high-dose males at 12 months.
Mean body weights of high-dose females were consistently less than controls
from week 2 until termination of the study. This difference was significant
(p <0.01) for 26 of the 59 intervals at which such measurements were made.
Food consumption in high-dose females also was generally less than controls.
Gross pathology findings were described as unremarkable. Microscopically,
testicular atrophy with degeneration of the tubular epithelium was observed
to a greater extent in the 300- and 3,000-ppm groups than in controls.
Additionally, an increased incidence of eosinophilic foci was observed in the
livers of both sexes exposed at 3,000 ppm. Based on the data presented,
a NOAEL of 30 ppm (1.5 mg/kg/day) was identified.
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Metolachlor August. 1987
-12-
The Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD):
RfD m 1.5 mcr/kg/day = Q.015 mg/kg/day
100
where:
1.5 mg/kg/day = NOAH, based upon the absence of systemic effects in
rats exposed to metolachlor in the diet for two years.
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 = (0.015 mg/kg/day)(70 kg) = 0.525 mg/L (525 ug/L)
(2 L/day)
where:
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 = (0.525 mg/L) (20%) = 0.01 mg/L (10 ug/L)
(10)
where:
0.525 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenic!ty.
Evaluation of Carcinogenic Potential
0 Four studies evaluating the carcinogenic potential of metolachlor
have been identified. In two of these studies (Marias et al., 1977,
and Tisdel et al., 1980), no evidence of carcinogenicity in mice was
observed. The other studies, bet*! conducted using rats, showed an
increased tumor incidence related to treatment. Ciba-Geigy (1979)
reported a statistically significant increase in primary liver tumors
in female Charles River rats exposed to 150 mg/kg/day in the diet
for 2 years. Tisdel et al. (1983) also reported a statistically
significant increase in the incidence of proliferative hepatic lesions
(neoplastic nodules and carcinomas) in female rats at the same
dietary dose over the same time period. Additionally, there was a
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Metolachlor Au*ust' 1987
-13-
nonstatistically significant increase in the frequency of adenocarcinoma
of the nasal turbinates and fibrosarcoma of the nasal tissue in the
high-dose males (150 mg/kg/day).
0 The International Agency for Research on Cancer has not evaluated the
carcinogenicity of metolachlor.
0 Applying the criteria described in EPA's guidelines for the assessment
of carcinogenic risk (U.S. EPA, 1986a), metolachlor is classified in
Group C: possible human carcinogen. This category is for substances
with limited evidence of carcinogenicity in animals and absence of
human data.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 EPA/OPP has identified an ADI for metolachlor of 0.015 mg/kg/day based
on the NOAEL of 30 ppm (1.5 mg/kg/day) from the chronic rat feeding
study (Tisdel et al., 1983) and an uncertainty factor of 100 (U.S. EPA,
1986b). Using this ADI and an assumed body weight of 60 kg, the maximum
permissible intake has been calculated to be 0.9 mg/day. The total
maximum residue concentration is 0.07209 mg/day or about 8% of the ADI.
0 Residue tolerances ranging from 0.02 to 30 ppm have been established
for a variety of agricultural products (CFR, 1985).
VII. ANALYTICAL METHODS
0 Analysis of metolachlor is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986c). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen phosphorus
detector. The method detection limit has not been determined for
metolachlor but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Whittaker (1980) experimentally determined adsorption isotherms for
metolachlor on granular-activated carbon (GAC) Nuchar WV-G. Nuchar
WV-G, reportedly, exhibited the following adsorption capacities at
20°C: 0.173, 0.148 and 0.105 mg metolachlor/mg carbon at concentra-
tions of 79.84 mg/L, 10 mg/L and 1.74 mg/L, respectively.
0 Holiday and Hardin (1981) reported the results of GAC treatment of
wastewater contaminated with pesticides including metolachlor. The
column, 3.5 ft in diameter, was packed with 10 ft of granular acti-
vated carbon, or 3,150 Ib carbon/column. The column was operated at
1.04 gpm/ft2 hydraulic load and 72 minutes contact time. Under these
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Metolachlor August, 1987
-14-
conditions, 99.5% of the metolachlor was removed from wastewater at
an initial average concentration of 16.4 mg/L.
GAC adsorption appears to be the most promising treatment technique
for the removal of metolachlor from water. However, more actual data
are required to determine the effectiveness of GAC in removing
metolachlor from contaminated drinking water supplies.
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Metolachlor August, 1987
-15-
IX. REFERENCES
AMR, Inc.* 1974a. Affiliated Medical Research, Inc. Emetic dose 50 in
beagle dogs with CGA-24705-Technical: Contract No. 120-2255-34. Received
September 26, 1974, Greensboro, NC. MRID 15525.
AMR, Inc.* 1974b. Affiliated Medical Research, Inc. Acute dermal LD50 of
CGA-24705- Technical in rabbits: Contract No. 120-2255-34. Received
September 26, 1974 under 5G1553. Unpublished study prepared for Ciba-Geigy
Corp., Greensboro, NC. MRID 15526.
Ami, P., and D. Muller.* 1976. Salmonella/mammalian-microsome mutagenicity
test with CGA 24705. Test for mutagenic properties in bacteria. PH 2.632.
Received January 19, 1977 under 7F1913. MRID 15397.
Bathe, R. 1973.* Acute oral LD50 of technical CGA-24705 in the rat: Project
No. Siss 2979. Received September 26, 1974 under 5G1553. Unpublished
study prepared by Ciba-Geigy Corp., Ltd., Basle, Switzerland. MRID 15523.
Ciba-Geigy Corporation.* 1976. Dominant lethal study on CGA 24705 technical:
Mouse (test for cytotoxic or mutagenic effects on male germinal cells)
PH 2.632. Received January 18, 1978 under 7F1913. Unpublished study
including addendum. MRID 15630.
Ciba-Geigy Corporation.* 1977. Section A General Chemistry. Unpublished
study received January 19, 1977 under 7F1913. MRID 15392.
Ciba-Geigy Corporation.* 1979. Two-year chronic oral toxicity study with
CGA-24705 technical in albino rats: Study No. 8532-07926. Conducted by
Industrial Bio-Test Laboratories. Unpublished study received December 11,
1979 under 8F2098. MRID 130776.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.368. July 1, 1985.
Coquet, B., L. Gallard, D. Guyot, X. Pouillet and J.L Rounaud.* 1974. Three-
month oral toxicity study trial of CGA 24705 in the dog. IC-CREB-R740119.
Received September 26, 1974 under 5G1553. Unpublished study prepared by
the Oncins Research and Breeding Center for Ciba-Geigy Corp., Greensboro,
NC. MRID 52477.
Fritz, H.* 1976. Reproduction study on CGA-24705 Tech. Rat: Segment II test
for teratogenic or embryotoxic effects: PH 2.632. Unpublished study
received January 19, 1977 under 7F1913. Prepared by Ciba-Geigy Ltd.,
Basle, Switzerland. MRID 15396.
Goldenthal, E.I., D.C. Jessup and J.S. Mehring.* 1979. Range-finding study
with metolachlor technical in beagle dogs: IRDC No. 382-053. Unpublished
study received December 11, 1979 under 100-597. Prepared by International
Research and Development Corp. Submitted to Ciba-Geigy, Corp., Greensboro,
NC. MRID 16631.
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Metolachlor August, 1987
-16-
Hambock, H.* I974a. Project 7/74: Metabolism of CGA 24705 in the rat.
(Status of results gathered up until June 10, 1974): AC 2.52. Unpub-
lished study received September 26, 1974 under 5G1553. Prepared by
Ciba-Geigy Ltd., Basle, Switzerland. MRID 39193.
Hambock, H.* 1974b. Project 12/74: Addendum to Project 7/74: Metabolism of
CGA 24705 in the rat: AC 2.52. Unpublished study received September 26,
1974 under 6G1708. Prepared by Ciba-Geigy Ltd., Basle, Switzerland.
MRID 15425.
Hambock, H.* 1974c. Project 1/74: Distribution, degradation and excretion of
CGA 24705 in the rat: AC 2.52. Unpublished study received September 26,
1974 under 5G1553. Prepared by Ciba-Geigy Ltd., Basle, Switzerland.
MRID 39192.
HAZARDLINE. 1985. National Library of Medicine. National Institutes of
Health. Bethesda, MD.
Holiday, A.D., and D.P. Hardin. 1981. Activated carbon removes pesticides
from wastewater. Chem. Big. 88:88-89.
Jessup, D.C., F.L. Estes, N.D. Jefferson et al.* 1979. Six-month chronic
oral toxicity study in beagle dogs: IRDC No. 382-054. Unpublished study
including addendum and AG-A No. 5358 received December 11, 1979 under
100-597. Prepared by International Research and Development Corporation.
Submitted by Ciba-Geigy Corporation. MRID 16632.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Published by the Association of Foods and Drugs Officials of
the United States.
Lightkep, G.E., M.S. Christian, G.D. Christian et al.* 1980. Teratogenic
potential of CGA-24705 in New Zealand White rabbits; Segment II
evaluation—Project 203-001. Unpublished study received September 15,
1980 under 100-597. Prepared by Argus Research Laboratories, Inc.
Submitted by Ciba-Geigy Corporation, Greensboro, NC. MRID 41283.
Marias, A.J., J. Gesme, E. Albanese et al.* 1977. Revised final report to
Ciba-Geigy Corporation: Carcinogenicity study with CGA-24705 technical
in albino mice: IBT No. 622-07925 (8532-07925). Unpublished study
received September 30, 1977 under 100-597. Prepared by Whittaker Corp.
Submitted by Ciba-Geigy Corporation, Greensboro, NC. MRID 84003.
Meister, R., ed. 1986. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
Sachsse, K.* 1973a. Irritation of technical CGA-24705 in the rabbit eye:
Project No. Siss 2979. Received September, 1974 under 5G1553. Unpublished
study prepared by Ciba-Geigy Ltd., Basle, Switzerland. MRID 15528.
Sachsse, K.* 1973b. Skin irritation in the rabbit after single application
of Technical CGA-24705. Project No. Siss 2979. Received September,
1974 under 5G1553. Unpublished study prepared by Ciba-Geigy Ltd.,
Basle, Switzerland. MRID 15530.
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Metolachlor August, 1987
-17-
Sachsse, K.* 1977. Skin sensitizing (contact allergenic) effects in guinea
pigs of Technical CGA-24705. Project No. Siss 5726. Received October 17,
1977. Unpublished study prepared by Ciba-Geigy Ltd., Basle, Switzerland.
MRID 15631.
Smith, S.H., and G.L. Adler.* 1978. Final report to Ciba-Geigy Corp: Three-
generation reproduction study with CGA-24705 technical in albino rats:
IBT No. 8533-07928. Received January 18, 1978 under 7F1913. Unpublished
study prepared by Industrial Bio-Test Laboratories, Inc. for Ciba-Geigy
Corp., Greensboro, NC. MRID 15632.
Smith, S.H., C.K. O'Loughlin, C.M. Salamon et al.* 1981. Two-generation
reproduction study in albino rats with metolachlor technical. Study No.
450-0272. Final report. Unpublished, study received September 30, 1981
under 100-597. Prepared by Whittaker Corporation; submitted by Ciba-Geigy
Corp., Greensboro, NC. MRID 80897.
STORET. 1987.
Tisdel, M., M.W. Balk, T. Jackson et al.* 1980. Toxicity study with metola-
chlor on mice. Unpublished study No. 79020 received July 25, 1980 under
100-587. Prepared by Hazleton/Raltech Scientific Services and American
College of Laboratory Animal Medicine. Submitted by Ciba-Geigy Corp.,
Greensboro, NC. MRID 39194.
Tisdel, M., T. Jackson, P. MacWillianis et al.* 1983. Two-year chronic oral
toxicity and oncogenicity study with metolachlor technical in albino
rats: Raltech study No. 80030. Final report. Unpublished study received
May 24, 1983 under 100-587. Prepared by Hazleton-Raltech, Inc. Submitted
by Ciba-Geigy Corp., Greensboro, NC. MRID 129377.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185)33992-34003.
September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Draft guidance for
the reregistration of products containing as the active ingredient:
metolachlor. Office of Pesticide Programs, Washington, DC.
"J.S. EPA. 1986c. U.S. Environmental Protection Agency. U.S. EPA Metnod -«1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Whittaker, K.F. 1980. Absorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck Index - An Encyclopedia of Chemicals and Drugs. 10th ed.
Rahway, NJ: Merck and Co., Inc.
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Metolachlor August, 1987
-18-
Worthing, C.R., ed. 1983. The Pesticide Manual: A World Compendium, 7th ed.
London: BCPC Publishers.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
METRIBUZIN
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
DRAFT
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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Metribuzin
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 21087-64-9
Structural Formula
4-Amino-6-(1,1-dimethylethyl)-3-methylthio-1,2,4-triazin-5(4H)-one
Synonyms
0 Bayer 6159; Bayer 6443H; Bayer 94337; Lexone; Sencor; Sencoral;
Sencorer; Sencorex
Uses
Herbicide used for the control of a large number of grass and broadleaf
weeds infesting agricultural crops (Meister, 1983).
C8H14ON4S
214.28
white crystalline solid
125-126°C
10-5 mmHg (20°C)
1,200 mg/L
-5.00 (calculated)
properties (CHEMLAB, 1985)
Chemical Formula
Molecular Height
Physical State (at 25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor —
Occurrence
0 Metribuzin has been found in 1,517 of 3,580 surface water samples
analyzed and in 54 of 240 ground water samples (STORET, 1987). These
samples were collected at 407 surface water locations and 204 ground
water locations; metribuzin was found in 14 states. The 85th
percentile of all nonzero samples was 4.79 ug/L in surface water and
0.1 ug/L in ground water sources. The maximum concentration found in
surface water was 22.79 ug/L and in ground water, 1.25 ug/L.
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Metribuzin August, 1987
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0 Metribuzin has been found in Iowa ground water resulting from
agricultural uses; typical positives were 1 to 4.3 ppb (Cohen et al.,
1986).
Environmental Fate
0 The rate of hydrolysis of metribuzin is pH dependent. During a
28-day test, little or no degradation was observed at pH 6 or 9 at
25°C, or at pH 6 at 37«C or 52°C (Day et al., 1976).
o 14c-Metribuzin on silty clay soil degraded, with a half-life of 15
days, when exposed to natural sunlight (Khasawinah, 1972). The half-life
in control samples kept in the dark was 56 days. After 10 weeks,
20.6, 6.5 and 7.0% of the applied radioactivity was present in the
irradiated soil as 6-t-butyl-l,2,4-triazin-3,5-(2H,4H)-dione (DADK),
6-t-butyl-3-(methylthio)-l,2,4-trizin-5(4H)-one (DA) and parent compound,
respectively. A substantial portion of the applied radioactivity
(56%) was bound to the soil. In the dark control, 4.6, 16.9, 44.0 and
34% of the applied radioactivity was present as DADK, DA, parent or
bound compound, respectively.
0 Under aerobic conditions, metribuzin at 10 ppci degraded with a
half-life of 35-63 days in silt loam and sandy loam soils treated
with a 50% wettable powder (WP) formulation, and 63 days in soils
treated with a 4-lb/gal F1C formulation (Pither and Gronberg, 1976).
Degradates found were: 6-t-butyl-l,2,4-triazin-3,5-(2H,4H)-dione
(DADK); 4-amino-6-butyl-l-2,4,-triazin-3,5-(2H,4H)-dione (DK); and
6-t-butyl-3-(methylthio)-l,2,4-triazin-5-(4H)-one (DA).
0 14c-Metribuzin residues degraded slowly in silty clay soil under
anaerobic conditions with a half-life of more than 70 days (Khasawinah,
1972). After 10 weeks of incubation, 10, 10.9, 57, and 19% of the
applied radioactivity was present as DADK, DA, parent compound or
bound to the soil, respectively.
0 Metribuzin adsorption was significantly correlated to soil organic
matter, clay and bar soil water contents (Savage, 1976). Calculated
KJJ values ranged from 0.27 for a sandy loam soil (0.75% organic
matter, 11% clay and 12% Of 0.33 bar soil water content), to 3.41 for
a clay soil (42% organic matter, 71% clay and 42% of 0.33 bar soil
water content).
o 14c-Metribuzin residues were very mobile in Amarillo sandy loam and
Louisiana Commerce silt loam soils; after leaching 12-inch soil
columns with 20 inches of water, 96.6 and 91.6% of the applied radio-
activity, respectively, was found in the leachate (Houseworth and
Tweedy, 1973). 14c-Metribuzin residues were relatively immobile in
Indiana silt loam and New York muck soils; after leaching 12-inch
soil columns, 90.6 and 89.4% of the applied radioactivity was detected
in the top 3 cm of the Indiana silt loam and New York muck, soil
columns, respectively. No radioactivity was detected in column
leachates.
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Metnbuzin August, 1987
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0 14c-Metribuzin residues (test substance not characterized) aged 30
days were moderately mobile in an Amarillo sandy loam soil column;
after leaching a 12-inch column with 22.5 inches of water, 7.3% of
the applied radioactivity was found in the leachate (Tweedy and
Houseworth, 1974). In the soil column, 85.2% of the applied radio-
activity remained within the top 2 inches.
o 14c-Metribuzin residues (test substance not characterized) were
intermediately mobile in sandy clay loam and silt loam soils
(Rf 0.61 to 0.62) and mobile in sandy, sandy loam, and two silty
clay soils (Rf 0.68 to 0.77), based on soil thin-layer chromatography
(TLC) tests (Thornton et al., 1976). 14c-Metnbuzin residues (test
substance not characterized) were intermediately mobile in sand (Rf
0.61), sandy clay loam (Rf 0.64), two silty clay soils (Rf 0.62 and
0.71), silt loam (Rf 0.66) and sandy loam (Rf 0.82) soils, based on
soil TLC tests (Christ and Thorton, 1979). 14c-Metribuzin (purity not
specified) at 1.5 ug/spot had low mobility (Rf 0.13 to 0.26) in two
muck soils and intermediate mobility (Rf 0.42 to 0.53) in six mineral
soils ranging in texture from sand to clay, based on soil TLC plates
developed in water (Sharon and Stephenson, 1976).
0 In the field, netribuzin dissipates with half-lives of less than
1 month to 6 months. Three metribuzin degradates were detected:
6-t-butyl-l,2,4-triain-3,5-(2H,4H)-dione (DADK); 4-amino-6-t-butyl-
l,2,4-triazin-3,5-(2H,4H)-dione (DK); and 6-t-butyl-3-(methylthio)l,2,4-
triazin-5-(4H)-one (DA). Soil type and characteristics, chemical
formulation or application rates did not discernibly affect the
dissipation rate of metribuzin (Stanley and Schumann, 1969; Finlayson,
1972; Rockwell, 1972a; Rockwell, 1972b; Rockwell, 1972c; Rowehl,
1972a; Rowehl, 1972b; Schultz, 1972; Mobay Chemical, 1973; Fisher,
1974; Murphy, 1974; United States Borax and Cnemical Corp., 1974;
Potts et al., 1975; Analytical Biochemistry Laboratories, 1976;
Ballantine, 1976; and Ford, 1979).
III. PHARMACOKINETICS
Absorption
0 A study was conducted in four dogs using oral dosing of radiolabeled
metribuzin (Khasawinah, 1972) to evaluate absorption, distribution
and metabolites. Analysis of blood samples showed a peak level
at 4 hours.
Distribution
0 No information was found in the available literature on the distribution
of metribuzin.
Metabolism
0 No information was found in the available literature on the metabolism
of metribuzin.
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Metribuzin August, 1987
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Excretion
0 Khasawinah (1972) reported that 52 to 60% of the administered dose of
metribuzin in dogs was excreted in the urine and 30% in the feces.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of metribuzin in humans.
Animals
Short-term Exposure
0 Crawford and Anderson (1974) reported the acute oral LD50 values
following the administration of technical metribuzin to guinea pigs
and rats as 245 and 1,090 ing/kg, respectively, for male animals,
and 274 and 1,206 rag/kg, respectively, for females.
0 Mobay Chemical (1978) reported the ac.::te oral LD50 values for a
wettable granular formulation of metribuzin to be 2,379 and
2,794 mg/kg for male and female rats, respectively.
0 Mobay Chemical (1978) reported the acute dermal LD50 for a wettable
granular formulation of metribuzin to be >5,000 mg/kg for both male
and female rats.
0 Mobay Chemical (1978) reported the acute (1-hour) inhalation LCso in
rats for a wettable granular formulation to be >20 mg/L.
Dermal/Ocular Effects
0 In studies conducted by Mobay Chemical (1978), metribuzin (wettable
granular) was determined to be a very slight irritant to rabbit eyes
and skin.
Long-term Exposure
0 Loser et al. (1969) administered metribuzin to wistar rats ( 1 5/sex/dose)
for 3 months in their feed at levels of 0, 50, 150, 500 or 1,500 ppm
(about 2.5, 7.5, 25 or 75 mg/kg/day, based on calculations in Lehman
et al., 1959). Following treatment, food consumption, growth, body
weight, organ weight, clinical chemistry, hematology, urinalysis and
histopathology were measured. No significant efects on these parameters
were observed in either sex at 50 ppm (2.5 mg/kg/day). Among females,
enlarged livers were found in the 150, 500 or 1,500 ppm (7.5, 25 or
75 mg/kg/day) dosage groups (p <0.05), and thyroid glands were also
enlarged in the 500 or 1,500 ppm (25 or 75 mg/kg/day) groups (p <0.05
and p <0.01, respectively). In the males, enlarged thyroids were
reported among the 500 (25 mg/kg/day) (p <0.05) and 1,500 ppm
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Metribuzin August, 1987
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(75 mg/kg/day) (p <0.01) dosage groups, while an enlarged heart was
reported at 1,500 ppm (75 mg/kg/day) (p <0.05). At 1,500 ppm
(75 mg/kg/day), lower body weights (p <0.01) were reported in both
sexes when compared to untreated controls.
0 In studies conducted by Lindberg and Richter (1970), beagle dogs
(four/sex/dose) administered oral doses of 50, 150 or 500 ppm (about
1.25, 3.75 or 12.5 mg/kg/day, based on calculations in Lehman et al.,
1959) technical metribuzin for 90 days showed no significant differences
in body weights, food consumption, behavior, mortality, hematologic
findings, urinalysis, gross pathology or histopathology.
0 Loser and Mirea (1974) reported that dietary concentrations of 1.5,
2 or 20 mg/kg/day metribuzin did not significantly affect physical
appearance, behavior, mortality, hematologic clinical chemistry,
urinalysis or histopathology in rats (40/sex/dose) fed technical
metribuzin in the diet for 24 months. The body weights of females at
the 20 mg/kg/day dose level were usually lower (p <0.05) than those of
controls; at the end of the test period, however, no significant
differences were noted.
0 Hayes et al. (1981) administered technical metribuzin in the diet
to albino CD mice (50/sex/dose) at 200, 800 or 3,200 ppm (about 30,
120 or 480 mg/kg/day, based on calculations in Lehman et al., 1959)
for 24 months. Following treatment, feed consumption, general behavior,
body and organ weights, mortality, hematology and histopathology were
analyzed. No adverse effects were observed in these parameters in
either sex at 800 ppm (120 mg/kg/day). However, a significant
(p <0.05) increase in absolute and relative liver and kidney weights
was observed in female mice receiving 3,200 ppm (480 mg/kg/day).
0 In studies conducted by Loser and Mirea (1974), four groups of beagle
dogs (four/sex/dose) were administered metribuzin in the diet at dose
levels of 0, 25, 100 or 1,500 ppm (about 0, 0.625, 2.5 or 37.5 mg/kg/day,
based on calculations in Lehman, 1959) for 24 months. Following
treatment, food consumption, general behavior and appearance, clinical
chemistry, hematology, urinalysis, body and organ weights and histo-
pathology were evaluated. No toxicologic effects were reported in
animals administered 100 ppm metribuzin (2.5 mg/kg/day) or less for
any of the parameters measured. Necrosis of the renal tubular cells,
slight iron deposition as well as slight hyp^rglycemia and temporary
hypercholesterolemia were noted in animals administered 1,500 ppm
(37.5 mg/kg/day).
Reproductive Effects
0 In a 3-generation reproduction study, Loser and Siegmund (1974)
administered technical metribuzin in the feed at dose levels of 0,
35, 100 or 300 ppm (about 0, 1.75, 5 or 15 mg/kg/day, based on
calculations in Lehman et al., 1959) to FB30 (Elberfeld breed) rats
during mating, gestation and lactation. Following treatment,
fertility, lactation performance and pup development were evaluated.
No treatment-related effects were reported at any dose tested.
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Metribuzin August, 1987
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Developmental Effects
0 Unger and Shellenberger (1981) administered technical metribuzin by
gastric intubation to pregnant female rabbits (16 to 17/dose) on days
6 through 18 of gestation at daily doses of 15, 45 or 135 mg/kg/day.
Following treatment, there was a statistically significant (p <0.05)
decrease in body weight gain in the high-dose does (135 mg/kg). No
maternal toxicity was reported in animals administered metribuzin at
levels of 45 mg/Jcg/day or less. No treatment-related effects were
reported at any dose level in fetuses based on gross, soft tissue and
skeletal examinations.
0 Machemer (1972) reported no maternal toxicity, embryotoxicity or
teratogenic effects following oral administration (via stomach tube)
of technical metribuzin to FB30 rats (21 to 22/dose) on days 6 through
15 of gestation at dose levels of 5, 15, 50 or 100 mg/kg/day.
Mutagenicity
0 Metribuzin showed no mutagenic activity in several bacterial assays
(Inukai and lyatomi, 1977; Shirasu et al., 1978) or in dominant
lethal tests in mice (Machemer and Lorke, 1974, 1976). The results
of microbial point mutation assays (Machemer and Lorke, 1974) did not
indicate a mutagenic potential for metribuzin in the test systems
utilized. The results of dominant lethal mutations in mice or
chromosomal aberrations in hamster spermatogonia at dose levels of
300 mg/kg and 100 mg/kg, respectively, did not indicate any mutagenic
effects of metribuzin.
Carcinogenic!ty
0 Hayes et al. (1981) conducted studies in which technical metribuzin
was administered in the diet to albino CD-I mice (50/sex/dose) at 200,
800 or 3,200 ppm (30, 120 or 380 mg/kg/day) for 24 months. Minimal
toxic effects were observed at the high-dose level in the form of
increased liver weight and changes in the hematocrit and hemoglobin
measurements. Although some increase in the number of tumor-bearing
animals was observed in low- and mid-dose animals, significant
increases in the incidence of specific tumor types were not observed
at any dose level. It was concluded that, under the conditions of the
test, there was no increase in the incidence of tumors in 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) , mg/L ( ug/L)
(UF) x ( L/day)
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Metribuzin August. 1987
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where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW a 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
No information was found in the available literature that was suitable
for determination of a One-day HA for metribuzin. It is therefore recommended
that the Ten-day HA value for a 10-kg child (4.5 mg/L, calculated below) be
used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The study by Unger and Shellenberger (1981) has been selected to serve
as the basis for determination of the Ten-day HA for metribuzin. In this
study, pregnant rabbits (16 or 17/dose) that were administered technical
metrizubin by gastric intubation at dosage levels of 0, 15, 45 or 135 mg/kg/day
on days 6 through 18 of gestation showed a statistically significant (p <0.05)
decrease in body weight gain at the 135-mg/kg dose. No maternal toxicity was
reported at or below the 45-mg/kg dose. No treatment-related effects were
reported at any dose level in fetuses based on gross, soft tissue and skeletal
examinations. The NOAEL identified in this study was, therefore, 45 mg/kg/day.
While a reproductive end point is not the most appropriate basis for derivation
of an HA for a 10-kg child, this study is the only one available for the
appropriate duration.
Using a NOAEL of 45 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (45 mg/kg/day) (10 kg) = 4.5 mg/L (4,500 ug/L)
(100) (1 L/day)
where:
45 mg/kg/day = NOAEL, based on absence of body weight reduction in
rabbits exposed to metribuzin via gastric intubation
on days 6 through 18 of gestation.
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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Longer-term Health Advisory
The study by Loser et al. (1969) has been selected to serve as the basis
for the Longer-term HA for metribuzin. In this study, rats (15/sex/dose)
were fed diets containing metribuzin at doses of 50, 150, 500 or 1,500 ppm
(about 2.5, 7.5, 25 or 75 mg/kg/day based on calculations in Lehman et al.,
1959) for 90 days. Thyroid glands were enlarged in males in the 500 or
1,500 ppm (25 or 75 mg/kg/day) dosage groups, while the heart was enlarged at
the 1,500 ppm (75 mg/kg/day) dose level. In females, enlarged livers were
detected in the 150, 500 or 1,500 ppm (7.5, 25 or 75 mg/kg/day) dosage groups,
and the thyroid was enlarged in the 500 or 1,500 ppm (25 or 75 mg/kg/day)
dosage groups. Body weights were reduced in both sexes at 1,500 ppm
(75 mg/kg/day), compared to untreated controls. The NOAEL identified in this
study was, therefore, 50 ppm (2.5 mg/kg/day). Lindberg and Richter (1970)
determined a NOAEL of 12.5 mg/kg/day in dogs; however, this study was'not
chosen, since the NOAEL was higher than the LOAEL of 7.5 mg/kg/day identified
by Loser et al. (1969) in the rat.
Using a NOAEL of 2.5 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (250 ug/L)
(100) (1 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of increased absolute organ
weights in rats exposed to metribuzin via the diet
for 90 days.
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.
Using a NOAEL of 2.5 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = (2.5 mg/kg/day) (70 kg) = 0.975 mg/L (875 ug/L)
(100) (2 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of increased absolute organ
weights in rats exposed to metribuzin via the diet
for 90 days.
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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Metribuzin August, 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 study by Loser and Mirea (1974) has been selected to serve as the
basis for the Lifetime HA for metribuzin. In this study, dogs (four/sex/dose)
were administered metribuzin in the diet at dose levels of 0, 25, 100 or
1,500 ppm (0, 0.625, 2.5 or 37.5 mg/kg/day) for 24 months. Necrosis of the
renal tubular cells was reported as well as slight and temporary changes in
certain clinical chemistry parameters (e.g., blood glucose and cholesterol)
at the high-dose level. No other toxicologic effects were reported. Based
on this information, a NOAEL of 100 ppm (2.5 mg/kg/day) and a LOAEL of
1,500 ppm (37.5 mg/kg/day) were reported. Loser and Mirea (1974) reported a
NOAEL of 20 rag/kg/day in rats. This study was not selected because no dose-
related toxicologic responses were observed, and the rat may be less sensitive
than the dog. Hayes et al. (1981) determined a NOAEL of 120 mg/kg/day in
mice; however, this value exceeded the LOAEL (37.5 mg/kg/day) reported by Loser
and Mirea (1974).
Using this study, the 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 organ toxicity and clinical
chemistry effects in dogs exposed to raetribuzin via
the diet for 24 months.
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Metribuzin August, 1987
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100 = uncertainty factor, chosen in accordance with NAS/OCW
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 mg/day (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 Advisory
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 In a study by Hayes et al. (1981), metribuzin was administered in the
feed of mice (50/sex/dose) at dose levels of 200, 800 or 3,200 ppm
(30, 120 or 480 mg/kg/day) for 24 months. Following treatment, the
incidence of tumor formation was analyzed in a variety of tissues.
Neoplasms of various tissues and organs were similar in type,
localization, time of occurrence and incidence in control and treated
animals. It was concluded that under the conditions of the test,
there was no increase in the incidence of tumors in mice.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of metribuzin.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogen risk (U.S. EPA, 1986), metribuzin may be classified in
Group D: not classified. This category is used for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 A Threshold Limit Value-Time-Weighted Average (TLV-TWA) of 5 mg/m3
was determined, based on animal studies substantiated by repeated
inhalation tests, a safety factor of 5, and assuming a total pulmonary
absorption (ACGIH, 1984).
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Metnbuzin August, 1987
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VII. ANALYTICAL METHODS
0 Analysis of metribuzin is by a gas chromatography (GC) method appli-
cable to the determination of certain organonitrogen pesticides in
water samples (U.S. EPA, 1985). This method requires a solvent
extraction of approximately 1 L of sample with methylene chloride
using a separatory funnel. The methylene chloride extract is dried
and exchanged to acetone during concentration to a volume of 10 mL or
less. The compounds in the extract are separated by GC and measurement
is made with a thermionic bead detector. The method detection limit
for metribuzin is 0.46 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAG) adsorption
and a conventional treatment scheme will remove metribuzin from water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
metribuzin on GAC.
0 Whittaker (1980) reported the results of GAC columns operating under
bench-scale conditions. At a flow rate of 0.8 gpm/sq ft and an empty
• bed contact time of 6 minutes, metribuzin breakthrough (when effluent
concentration equals 10% of influent concentration) occurred after
112 bed volumes (Bv).
0 In the same study, Whittaker (1980) reported the results for four
metribuzin bi-solute solutions when passed over the same GAC continuous
flow column.
0 Another study investigated the effectiveness of two different GAC
columns in removing metribazin from contaminated wastewater (Whittaker,
et al., 1982). One type of GAC showed breakthrough for metribuzin
(6 mg/L) from an initial concentration of 140 mg/L after 50 gallons
of the wastewater had been treated. No pesticide was found in the
effluent from the second type of GAC.
0 Conventional water treatment, coagulation and sedimentation with alum
and an anionic polymer removed more than 50% of the metribuzin present
(Whittaker et al., 1980). The optimum alum dosage was 200 mg/L. Also
equivalent dosages of ferric chloride were found to be equally effective,
0 Treatment technologies for the removal of metribuzin from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt metribuzin
removal from water must be by a case-by-case technical evaluation,
and an assessment of the economics involved.
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IX. REFERENCES
ACGIH. 1984. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 3rd ed., Cincinnati, OH: ACGIH, p.
Analytical Biochemistry Laboratories. 1976. Chemagro agricultural division
— Mobay Chemical Corporation soil persistence study: MW-HR-409-75;
Report No. 50842. Unpublished study prepared in cooperation with Mobay
Chemical Corp., submitted by Ciba-Geigy Corp., Greensboro, NC.
Ballantine, L.G. 1976. Metolachlor plus metribuzin tank mix soil dissipation:
Report No. ABR-76092. Summary of studies 095763-B through 095763-F.
Unpublished study submitted by Ciba-Geigy Corp., Greensboro, NC.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD: p.
Cohen, S.Z., C. Eiden and M.N. Lober. 1986. Monitoring ground water for
pesticides in the U.S.A. _In Evaluation of Pesticides in Ground Water.
American Chemical Society Symposium Series. American Chemical Society,
City, State: p. . (in press).
Crawford, C.R. and R.H. Anderson.* 1974. The acute oral toxicity of Sencor
technical, several Sencor.metabolites and impurities to rats and guinea
pigs: Report no. 38927. Rev. unpublished study. MRID 00045270.
Day, E.W., W.L. Sullivan and O.D. Decker. 1976. A hydrolysis study of the
herbicides oryzalin and metribuzin. Unpublished study submitted by
Blanco Products Co., Div. of Eli Lilly Co., Indianapolis, IN.
Finlayson, D.G. 1972. Soil persistence study: Victoria, British Columbia,
Canada, _In Supplement No. 4 to brochure entitled: Sencor: The effects
on the environment: Document No. AS77-1968. Unpublished study submitted
by Mobay Chemical Corp.
Fisher, R.A. 1974. Mobay Chemical Corporation residue experiment, Mentha,
Michigan. Sencor residues in soil: Report No. 41395. Unpublished study
including report nos. 41625, 41626, 41627. Prepared in cooperation with
Missouri Analytical Laboratories, submitted by Mobay Chemical Corp.,
Kansas City, MO.
Fjrd, J.J. 1979. Herbicide combination—soil dissipation study involving
Antor herbicide with three commercial herbicides: RI 47-003-06. Submitted
by Hercules, Inc., Wilmington, DE.
Hayes, R.H., D.W. Lamb, D.R. Mallicout et al.» 1981. Metribuzin (R) (Sencor)
oncogenicity study in mice: 80050. Unpublished study. MRID 00087795.
Houseworth, L.D. and B.C. Tweedy. 1973. Report on parent leaching studies
for Sencor: Report No. 37180. Unpublished study prepared by Univ. of
Missouri, Dept. of Plant Pathology, submitted by Mobay Chemical Corp.,
Kansas City, MO.
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Metribuzin August, 1987
-14-
Inufcai, H. and A. lyatomi.* 1977. Bay 94337: Mutagenicity test on bacterial
systems: Report no. 67; 54127. Unpublished study. MRID 00086770.
Khasawinah, A.M. 1972. The metabolism of Sencor (Bay 94337) in soil:
Report No. 31043. Unpublished study submitted by Mobay Chemical Corp.,
Kansas City, MO.
Lehman, W.J., W.F. Reehl and D.H. Rosenblatt. 1959. Handbook of chemical
property estimation methods. New York: McGraw Hill.
Lindberg, D. and W. Richter.* 1970. Report to Chemugor Corporation: 90-Day
subacute oral toxicity of Bay 94337 in beagle dogs: IBT no. C776; 26488.
Unpublished study. MRID 00106162.
Loser, E., D. Lorke and L. Mandesley-Thomas.* 1969. Bay 94337. Subchronic
toxicological studies on rats (3-month feeding test): Report no. 1719;
26469. Unpublished study. MRID 00106161.
Loser, E. and D. Mirea.* 1974. Bay 94337: Chronic toxicity studies on dogs
(two-year feeding experiments): Report no. 4887; Report no. 41814.
Unpublished study. MRID 00061261.
Loser, E. and F. Siegmund.* 1974. Bay 94337. Multigeneration study on rats:
Report no. 4889; Report no. 41818. Unpublished study. MRID 00061262.
Machemer, L.* 1972. Sencor (Bay 94337): Studies for possible embryotoxic
and teratogenic effects on rats after oral administration: Report nos.
3678 and 35073. Unpublished study. MRID 00061257.
Machemer, L. and D. Lorke.* 1974. Evaluation of (R) Sencor for mutagenic
effects on the raouset Report no. 4942; 43068. Unpublished study.
MRID 00086766.
Machemer, L. and D. Lorke.* 1976. (R) Sencor: Additional dominant lethal
study on male mice to test for mutagenic effects by an improved method.
Report no. 6110; 49068. Unpublished study. MRID 00086768.
Meister, R., ed. 1983. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Company.
Mobay Chemical. 1973. Mobay Chemical Corporation. Sencor: Metabolic,
analytical, and residue information for sugarcane (Hawaii). Unpublished
study by Mobay Chemical Corp., Kansas City, MO.
Mobay Chemical.* 1978. Mobay Chemical Corporation. Supplement to synopsis
of human safety of Sencor: Supplement no. 3. Summary of studies 235396-B
through 235396-E. Unpublished study. MRID 00078084.
Murphy, H. 1974. Mobay Chemical Corporation residue experiment, Presque
Island, Maine. Sencor residues in soils: Report No. 41395. Unpublished
study including report nos. 41625, 41626, 41627, prepared in cooperation
with Missouri Analytical Laboratories, submitted by Mobay Chemical
Corp., Kansas City, MO.
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Metribuzin August, 1987
-15-
Obrist, J.J. and J.S. Thornton. 1979. Soil thin-layer mobility of Baycor
(TM), Baytan, Drydene and Peropal (TM). Unpublished study prepared in
cooperation with Agricultural Consultants, Inc., submitted by Mobay
Chemical Corp., Kansas City, MO.
Pither, K.M. and R.R. Gronberg. 1976. A comparison of the rate of metabolic
degradation of Sencor in soil using the 50% wettable powder and 4 flowable
formulations: Report No. 45990. Unpublished study submitted by Mobay
Chemical Corp., Kansas City, MO.
Potts, C.R., M.M. Laporta, J. Devine et al. 1975. Prowl (CL 92, 553):
Determination of CL 92,553 (N-(l-Ethylpropyl)-3,4-dimethyl-2,6-dinitro-
benzenamine and Sencor 4-Amino-6-t-butyl-3-(methylthio)-l,2,4-triazin-
5(4H)-one in soil: Report No. C-801. Unpublished study submitted by
American Cyanamid Company, Princeton, NJ.
Rockwell, L.F. 1972a. Soil persistence study of BAY 94337; Plot F-17,
Research Farm, Stanley, Kansas. In Sencor: The effects on the environ-
ment. Compilation; unpublished study submitted by Mobay Chemical Corp.,
Kansas City, MO.
Rockwell, L.F. 1972b. Soil persistence study; plot F-2, Research Farm,
Stanley, Kansas. In Supplement No. 4 to brochure entitled: Sencor:
The effects on the environment: Document No. AS77-1968. Unpublished
study submitted by Mobay Chemical Corp., Kansas City, MO.
Rockwell, L.F. 1972c. Soil persistence study of DADK; Plot F-17, Research
Farm, Stanley, Kansas. In Supplement No. 4 to brochure entitled:
Sencor: The effects on the environment: Document No. AS77-1968.
Compilation; unpublished study submitted by Mobay Chemical Corp.,
Kansas City, MO.
Rowehl, E.R. 1972a. Soil persistence study of BAY 94337; Vero Beach, Florida.
_In Sencor: The effects on the environment. Compilation; unpublished
study submitted by Mobay Chemical Corp., Kansas City, MO.
Rowehl, E.R. 1972b. Soil persistence study of DADK; Vero Beach, Florida.
In Supplement No. 4 to brochure entitled: Sencor: The effects of the
environment: Document No. AS77-1968. Unpublished study submitted by
Mobay Chemical Corp., Kansas City, MO.
Savage, K.E. 1976. Adsorption and mobility of metribuzin in soil. Weed
Sci. 24(5):525-528.
Schultz, T.H. 1972. Soil persistence study. Report No. 33131. Unpublished
study submitted by Chemagro, In Supplement No. 4 to brochure entitled:
Sencor: The effects on the environment: Document No. AS77-1968.
Compilation; unpublished study.
Sharon, M. and G.R. Stephenson. 1976. Behavior and fate of metribuzin.
Weed Sci. 24(2) .-153-160. Submitter report no. 49127. In unpublished
study submitted by Mobay Chemical Corp., Kansas City, MO.
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Metribuzin August, 1987
-16-
Shirasu, Y., M. Moriya and T. Ohta.* 1978. Metribuzin mutagenicity test on
bacterial systems. Submitter Report No. 66748. Unpublished study.
MRID 00109254.
Stanley, C.W. and S.A. Schumann. 1969. A gas chromatographic method for
the determination of BAY 94337 residues in potatoes, soybeans, and corn:
Report No. 25,838. Unpublished study submitted by Mobay Chemical Corp.,
Kansas City, MO.
STORET. 1987.
Thornton, J.S., J.B. Hurley and J.J. Obrist. 1976. Soil thin-layer mobility
of twenty-four pesticide chemicals. Report No. 51016. Unpublished
study submitted by Mobay Chemical Corp., Pittsburgh, PA.
Tweedy, B.C. and L.D. Houseworth. 1974. Leaching of aged residues of
Sencor-3-14C in sandy loam soil: Report No. 40567. Unpublished study
prepared by Univ. of Missouri, Dept. of Plant Pathology, submitted by
Mobay Chemical Corp., Kansas City, MO.
*Unger, T.M. and T.E. Shellenberger. 1981. A teratological evaluation of
Sencor (R) in mated female rabbits: 80051. Final report. Unpublished
study. MRID 00087796.
United States Borax and Chemical Corp. 1974. Cobex plus Sencor (or Lexone):
Degradation in soil. Compilation; unpublished study.
U.S. EPA. 1985. U.S. Environmental Protection Agency. U.S. EPA Method 633
- Organonitrogen Pesticides. Fed. Reg. 50:40701. October 4, 1985.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University, Lafayette, IN.
Whittaker, K.F., J.C. Nye, R.F. Wukasch and H.A. Kazimier. 1980. Cleanup
and collection of wastewater generated during cleanup of pesticide
application equipment. Paper presented at National Hazardous Waste
Symposium, Louisville, KY.
Whittaker, K.F., J.C. Nye, R.F. Wukasch, R.J. Squires, A.C. York and H.A.
Kazimier. 1982. Collection and treatment of wastewater generated by
pesticide application. EPA report no. 600/2-82-028.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
August, 1987
PARAQUAT
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, logifor 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 pre'ict 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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Paraquat
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
Paraquat, with a chemical name 1,1 '-dimethyl-A^'-dipyridinium
ion, is present mostly as the dichloride salt (CAS No. 1910-42-5) or
as the dimethyl sulfate salt (CAS No. 2074-50-2, molecular weight 408.48)
(Meister, 1987). Contents discussed below pertain to paraquat dichloride,
CAS No. 1910-42-5
Structural Formula
CH.-N,
2CI
1,1l-Dimethyl-4,4'-bipyridinium-dichloride
Synonyms
Uses
o-Paraquat dichloride, Gramixel, Gramonol, Gramoxone, Gramuron,
Pathelear, Totacol, Weedol (Meister, 1985).
0 Contact herbicide and desiccant used for desiccation of seed crops,
for noncrop and industrial weed control in bearing and nonbearing
fruit orchards, shade trees, and ornamentals, for defoliation and
desiccation of cotton, for harvest aid in soybeans, sugarcane, guar,
and sunflowers, for pasture renovation, for use in "no-till" or before
planting or crop emergence, dormant alfalfa and clover, directed
spray, and for killing potato vines. Paraquat is also effective for
eradication of weeds on rubber plantations and coffee plantations and
against paddy bund (Neister, 1985).
Properties (ACGIH, 1980; Meister, 1985; CHEMLAB, 1985; TDB, 1985)
Chemical Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Vapor Pressure
Specific Gravity
Hater Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C12H14N2.2C1
257.18
Colorless to yellow crystalline
solid
175 to 180°C
No measurable vapor pressure
1.24 at 20°C/20°C
Very soluble
2.44 (calculated)
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Paraquat August, 1987
-3-
Occurrence
0 Paraquat was found in only one sample, at a concentration level of
20 ug/L, from 721 ground water samples analyzed (STORET, 1987).
Samples were collected at 715 ground water locations, with paraquat
found in one location in California. No surface water samples were
collected for analysis.
Environmental Fate
8 14c-Paraquat dichloride (>96.5% pure) at 91 mg/L was stable to
hydrolysis at 25 and 40°C at pH 5, 7 and 9 for up to 30 days (Upton
et al.f 1985).
0 Uniformly ring-labeled 14c-paraquat (99.7% pure) at approximately
7.0 ppm in sand did not photodegrade when irradiated with natural
sunlight for 24 months (Pack, 1982). No degradation products were
detected at any sampling interval. After 24 months of irradiation,
>84% of the applied radioactivity was extractable and <4% was
unextractable.
0 Paraquat was essentially stable to photolysis in soil (Day and
Hemingway, 1981). Four degradation products, 1-methyl-4,4'-bipyridylium
ion, 4-(1,2-dihydro-1-methyl-2-oxo-4-pyridyl)-1-methyl pyridylium
ion, 4-carboxy-1-methyl pyridylium ion, and an unknown, individually
constituted <6.0% of the total radioactivity in either irradiated
(undisturbed) or dark control soils.
0 Paraquat (test substance uncharacterized) at 0.05 to 1.0 ppm in water
plus soil declined with a half-life of >2 weeks (Coats et al., 1964).
In water only, paraquat declined with a half-life of approximately
23 weeks.
8 14c-Paraquat (test substance uncharacterized) was immobile in silt
loan and silty clay loam (Rf 0.00), and slightly mobile in sandy loam
(Rf 0.13) soils, based on soil thin-layer chromatography (TLC) tests
(Helling and Turner, 1963).
8 Methyl-labeled 14c-paraquat (test substance uncharacterized) at 1.0
ppm was stable to volatilization at room temperature over a 64-day
period (Coats et al., 1964).
8 In a pond treated with paraquat (test substance uncharacterized) at
1.14 ppm (Frank and Comes, 1967), paraquat residues (uncharacterized)
declined from 0.55 ppm 1 day after treatment to nondetectable (<0.001
ppm) 18 days after treatment. The dissipation of paraquat residues
(uncharacterized) in water was accompanied by a concomitant increase
of paraquat residues (uncharacterized) in the soil. Paraquat (test
substance uncharacterized) at 0.04 ppm dissipated in pond water with
a half-life of approximately 2 days (Coats et al., 1964). For more
details, see Calderbank's chapter on paraquat in Herbicides
(Calderbank, 1976).
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Paraquat August, 1987
-4-
III. PHARMACOKINETICS
Absorption
0 In Wistar rats given single oral doses of 14C-paraquat dichloride or
dimethyl sulfate by gavage (0.5 to 50 mg/kg, purity not stated),
69 to 96% was excreted unchanged, mostly in feces, and no radioactivity
appeared in bile (Daniel and Gage, 1966). Some systemic absorption of
the degradation products that were produced in the gut was noted.
Approximately 30% of the administered dose appeared in feces in a
degraded form.
0 14C-Methyl-labeled paraquat (99.7% purity) was administered orally
to a cow in a single dose of approximately 8 mg cation/kg (Leahey
et al., 1972). A total of 95.6% of the dose was excreted in feces in
the first 3 days. A small amount, 0.7% of the dose, was excreted in
the urine, 0.56% during the first 2 days. Only 0.0032% of the dose
appeared in the milk.
0 A goat was administered 1*C-ring-labeled paraquat dichloride (>99%
purity) orally at 1.7 mg/kg for 7 consecutive days (Leahey et al.,
1976a). At sacrifice, 2.4% and 50.3% of the radioactive dose had been
excreted in the urine and feces, respectively, and 33.2% was recovered
in the contents of the stomach and intestines. The radioactivity was
associated with unchanged paraquat.
0 In studies with pigs, 1 ^-methyl-labeled (Leahey et al., 1976b) and
14c-ring-labeled (Spinks et al., 1976) paraquat (>99% purity) at
dose levels of 1.1 and 100 mg ion/kg/day, respectively, was given
for.up to 7 days. At sacrifice, 69 to 72.5% and 2.8 to 3.4% of the
total radioactive dose had been excreted in the feces and urine,
respectively.
Distribution
0 Pigs were given oral doses of 14c-methyl-labeled (Leahey et al.,
1976b) and 14c-ring-labeled (Spinks et al., 1976) paraquat dichloride
(>99% purity) for up to 7 consecutive days at dose levels of 1.1 and
100 mg ion/kg/day, respectively. At sacrifice, radioactivity associated
mostly with unchanged paraquat was identified in the lungs, heart,
liver and kidneys, with trace amounts in the brain, muscle and fat.
0 The distribution of radioactivity was studied in a goat fed 14c-ring-
labeled paraquat dichloride (1.7 mg/kg/day, 99.7% purity) in the
diet for 7 consecutive days (Hendley et al., 1976). Most of the
radioactivity was found in the lungs, kidneys and liver. The ma]or
residue was unchanged paraquat.
Metabolism
0 Paraquat dichloride or paraquat dimethyl sulfate (radiochemical
purity: 99.3 to 99.8%), labeled with 1*C in either methyl groups or
in the ring, was poorly absorbed from the gastrointestinal tract of a
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Paraquat August, 1987
-5-
cow (Leahey et al.y 1972), goats (Hendley et al., 1976), pigs (Leahey
et al., 1976b; Spinks et al., 1976) and rats (Daniel and Gage, 1966),
and was excreted in the feces mostly as unchanged paraquat. However,
after an oral dose, there was microbial degradation of paraquat in
the gut. In one study with rats (Daniel and Gage, 1966), 30% of a
dose of paraquat appeared in the feces in a degraded form. A portion
of these microbial degradation products can be absorbed and excreted
in the urine, whereas the remainder is excreted in the feces.
Excretion
In studies with a cow (Leahey et al., 1972) and rats (Daniel and
Gage, 1966), about 96% and 69 to 96%, respectively, of the administered
radioactivity (single oral doses, 1*Olabeled) from paraquat was
excreted in the feces within 2 to 3 days as unchanged paraquat.
Goats (Hendley et al., 1976) and pigs (Leahey et al., 1976b; Spinks
et al., 1976) that received single oral doses of 14c-labeled paraquat
(1.7 and 1.1 or 100 mg ion/kg/day, respectively) for up to 7 days
excreted 50 and 69%, respectively, of the total administered dose in
feces unchanged.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 The Pesticide Incident Monitoring System (U.S. EPA, 1979) indicated
numerous cases of poisoning from deliberate or accidental ingestion
of'paraquat or by dermal and inhalation exposure from spraying,
mixing and loading operations. Generally, the concentrations of the
ingested doses or of amounts inhaled or spilled on the skin were not
specified. Symptoms reported following these exposures included
burning of the mouth, throat, eyes and skin. Other effects noted
were nausea, pharyngitis, episcleritis and vomiting. No fatalities
were reported following dermal or inhalation exposure. Deliberate
and accidental ingestion of unspecified concentrations of paraquat
resulted in respiratory distress and subsequent death. See also
Cooke et al. (1973).
Long-term Exposure
0 No information was found in the available literature on long-term
human exposure to paraquat.
Animals
Short-term Exposure
0 Acute oral LD50 values for paraquat (99.9% purity) were reported as
112, 30, 35 and 262 mg paraquat ion/kg in the rat, guinea pig, cat
and hen (Clark, 1965). Signs of toxicity included respiratory distress
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Paraquat August, 1987
-6-
and cyanosis among rats and guinea pigs, blood-stained droppings
among the hens, and muscular weakness, incoordination and frequent
vomiting of frothy secretion among the cats.
• Acute (4-hour) inhalation LCso values for paraquat ranged from 0.6 to
1.4 mg ion/m3 paraquat (McLean Head et al., 1985).
Dermal/Ocular Effects
0 Acute dermal LD50 values for rabbits (Standard Oil, 1977) were
59.9 mg/kg and 80 to 90 mg paraquat ion/kg for rats (FDA, 1970).
• Paraquat concentrate 3 (34.4% paraquat ion) was applied (0.5 mL or
172 mg paraquat ion) to intact and abraded skin of six male New
Zealand White rabbits for 24 hours (Bullock, 1977). Very slight,
moderate or severe erythema and slight edema were noted during the
7-day observation period for both intact and abraded skin.
0 Paraquat concentrate 3 (0.1 mL, 34.4% paraquat ion) was instilled
into the conjunct!val sac of one eye in each of six male New Zealand
White rabbits (Bullock and MacGregor, 1977). Untreated eyes served
as controls. Unwashed eyes were examined for 14 days. Complete
opacity of the cornea was reported in three of six rabbits. Roughened
corneas, severe pannus, necrosis of the conjunctivae, purulent discharge,
severe chemosis of the conjunctivae and mild iritis were also reported.
Long-term Exposure
0 Beagle dogs (three/sex/dose) were fed technical o-paraquat (32.2%
cation) in the diet for 90 days at dose levels of 0, 7, 20, 60 or
120 ppm (Sheppard, 1981). Assuming that 1 ppm is equivalent to
0.025 mg/kg/day, these levels correspond to doses of 0, 0.18, 0.5,
1.5 or 3 mg paraquat ion/kg/day (Lehman, 1959), respectively.
Increased lung weight, alveolitis and alveolar collapse were observed
at 60 ppm. The No-Observed-Adverse-Effect-Level (NOAEL) identified
for this study was 20 ppm (0.5 mg paraquat ion/kg/day).
0 Alderley Park beagle dogs (six/sex/dose) were fed diets containing
technical paraquat (32.3%) cation daily for 52 weeks at dietary levels
of 0, 15, 30 or 50 ppm (Kalinowski et al., 1983). Based on actual
group mean body weights and food consumption, these values correspond
to doses of 0, 0.45, 0.93 and 1.51 mg/kg/day for male dogs and 0,
0.48, 1.00 or 1.58 for females. Clinical and behavioral abnormali-
ties, food consumption, body weight, hematology, clinical chemistry,
urinalysis, organ weights, gross pathology and histopathology were
comparable for treated animals and controls at 15 ppm (the lowest
dose tested). An increased severity and extent of chronic pneumonitis
occurred at 30 ppm in both sexes, but especially in the males. Based
on the results of this study, the NOAEL identified was 15 ppm (0.45 mg
paraquat cat-ion/kg/day).
0 Technical paraquat dichloride (32.7% paraquat ion) was fed to Alderley
Park mice (60/sex/dose) for 97-99 weeks at levels of 0, 12.5, 37.5
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Paraquat August, 1987
-7-
and 100/125 ppm (100 ppm for the initial 35 weeks and then 125 ppm
until termination of the study) (Litchfield et al., 1981). Based on
the assumption that 1 ppm in the diet of mice is equivalent to 0.15
mg/kg/day (Lehman, 1959), these levels correspond to doses of 0,
1.87, 5.6 and 15/18.75 mg/kg. The animals were observed for toxic
signs, and body weights, food consumption and utilization, urinalysis,
gross pathology and histopathology were evaluated. Renal tubular
degeneration in the males and weight loss and decreased food intake
in the females, were the only effects observed, and occurred in the
37.5-ppm dose group. Based on these findings, a NOAEL of 12.5 ppm
(1.87 mg/kg/day) was identified.
0 Fischer 344 rats (70/sex/dose) were fed diets containing 0, 25, 75
or 150 ppm of technical paraquat (32.69% cation) for 113 to 117 weeks
(males) and 122 to 124 weeks (females) (Woolsgrove et al., 1983). Based
on the assumption that 1 ppm in the diet is equivalent to 0.05 mg/kg/day
(Lehman, 1959), these levels correspond to doses of 0, 1.25, 3.75 or
7.5 mg/kg/day. Clinical signs, food and water consumption, clinical
chemistry, urinalysis, hematology, ophthalmoscopic effects, gross
pathology and histopathology were evaluated. Increased incidences of
slight hydrocephalus were noted in the female rats dying between week
53 and termination of the study; these incidences were 5/60, 8/30,
9/27 and 9/30 rats in the control, low, mid and high dose, respectively.
Also, increased incidences of spinal cord cysts and cystic spaces
were noted in the male rats dying between week 53 and termination of
the study. These incidences were 0/53, 6/36 and 4/35 rats at the
control, low and mid-level doses, respectively; no incidence was
reported at the high dose. Eye opacities, cataracts and nonneoplastic
lung lesions (alveolar macrophages and epithelialization, and slight
peribronchiolar lymphoid hyperplasia) were observed at 75 ppm and
above. Similar eye lesions occurred at 25 ppm (the lowest dose
tested). These effects did not appear to be biologically significant,
since they were either minimal or occurred after 104 weeks of treatment
and appeared, therefore, to be only an acceleration of the normal
aging process. Based on these results, an approximate NOAEL of
25 ppm (1.25 mg/kg/day) was identified.
Reproductive Effects
0 Lindsay et al. (1982) fed Alderley Park rats technical paraquat
dichloride (32.7% cation w/w) in unrestricted diet for three ge-era-
tions at dose levels of 0, 25, 75 or 150 ppm paraquat ion.
Based on the assumption that 1 ppm in the diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels
correspond to doses of 0, 1.25, 3.75 or 7.5 mg/kg/day. No adverse
reproductive effects were reported at 150 ppm (the highest dose
tested) or less. An increased incidence of alveolar histiocytosis in
the lungs of male and female parents (F0, ?-\ and F2) was observed in
the 75- and 150-ppm dose groups. Based on these results, a reproductive
NOAEL of >150 ppm (7.5 mg/kg/day, the highest dose tested) and a
systemic NOAEL of 25 ppm (1.25 mg/kg/day, the lowest dose tested)
were identified.
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Paraquat August, 1987
-8-
Developmental Effects
0 Young adult Alderley Park mice (number not stated) were administered
paraquat dichloride (100% purity) orally by gavage at dose levels of
0, 1, 5 or 10 mg paraquat ion/kg/day on days 6 through 15 of gestation
(Hodge et al., 1978a). No teratogenic responses were reported at
10 mg ion/kg/day (the highest dose tested) or lower. Partially
ossified sternebrae in 26.3% of the fetuses in the high-dose group
(10 mg ion/kg/day) and decreased maternal weight gain in the 5-mg
ion/kg/day dose group were observed. Based on these results, the
developmental NOAEL identified for this study was 5 mg/kg/day, while
the maternal NOAEL was 1 mg/kg/day.
0 Hodge et al. (1978b) dosed Alderley Park rats (29 or 30/dose) by
gavage with paraquat dichloride (100% purity) on days 6 through
15 of gestation at dose levels of 0, 1, 5 and 10 mg paraquat ion/kg/day.
No teratogenic effects were reported at 10 mg ion/kg/day (the highest
dose tested). Maternal body weight gain was significantly decreased
(p £0.001) at 5 mg ion/kg/day and above. Fetal body weight gain was
significantly (p = 0.05) decreased at the mid-dose (5 mg/kg/day) and
above. Based on these findings, the developmental and maternal NOAEL
of 1 mg paraquat ion/kg/day was identified.
Mutaqenicity
e Analytical-grade paraquat dichloride (99.6% purity) was weakly
mutagenic in human lymphocytes, with and without metabolic activation,
at cytotoxic concentrations (1,250 to 3,500 ug paraquat dichloride/mL)
(Sheldon et al., 1985).
0 Technical-grade, 45.7% active ingredient (a.i.) and analytical-grade
(99.6% a.i.) paraquat dichloride were weakly positive in the L5178Y
mouse lymphoma assay with and without metabolic activation in studies
by Clay and Thomas (1985) and Cross (1985), respectively. Statistically
significant increases in mutant colonies were observed only at doses
below 29% cell survival (Cross, 1985).
0 Analytical-grade paraquat dichloride (99.4% a.i.) increased sister-
chromatid exchanges (SCE) at nontoxic doses (_O24 ug/mL in non-
activated cultures and ^245 ug/mL in S9-supplemented cultures. The
induction of increased SCE was more marked in the absence of the S9
fraction (Howard etal., 1985).
0 Mutagenic activity was detected in various assays with Salmonella
typhimurium (Benigni et al., 1979), human embryo epithelial cells
(Benigni et al., 1979) and Saccharomyces cerevisiae (Parry, 1977).
Carcinogenicity
0 Technical paraquat dichloride (32.7% paraquat ion) fed to Alderley
Park mice (60/sex/dose) for 99 weeks did not induce statistically
significant dose-related oncogenic responses at dose levels of 0,
12.5, 37.5 or 100/125 ppm (100 ppm for the initial 35 weeks and then
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Paraquat August, 1987
-9-
125 ppm until termination of the study) (Litchfield et al., 1981).
Based on the assumption that 1 ppm in food in mice is equivalent to
0.15 mg/kg/day (Lehman, 1959), these levels correspond to doses of 0,
1.87, 5.6 and 15/18.75 mg/kg. The study appeared to have been conducted
properly, except that hematological and organ weight determinations
were not performed. The absence of these parameters do not compromise
the results, since the occurrence of certain toxicological end points
(e.g., leukemia) detected by these tests are rare in mice. The
results, therefore, provide evidence that paraquat is not oncogenic
at the dose levels tested.
0 Woolsgrove et al. (1983) fed Fischer 344 rats (70/sex/dose) diets
containing technical paraquat (32.69%) for 113 to 117 weeks (males)
and 122 to 124 weeks (females) at dietary levels of 0, 25, 75 and
150 ppm. Based on the assumption that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to doses of 0, 1.25, 3.75 and 7.5 mg paraquat cation/kg/day. The
predominant tumor types noted in this study were tumors of the lungs,
endocrine glands (pituitary, thyroid and adrenal) and of the skin and
subcutis. Both the lung and endocrine tumors occurred at a frequency
similar to the incidence of these kinds of tumors in the historical
control. Only the squamous cell neoplasia of the skin and subcutis
were determined to be treatment-related. The squamous cell carcinoma
was a predominant tumor in the head region of the male and female
rats. This uncommon tumor occurred in 51.6% of all rats with skin and
subcutis tumors in the head region. The incidence of these tumors in
this study was 2, 4, 0 and 8% in the control, low-, mid- and high-dose
male groups, respectively and 0, 0, 4 and 3% in the control, low-,
mid- and high-dose female groups, respectively. When these incidences
were compared with incidences in historical controls (0 to 2.0% in
males and 1.9 to 4.0% in females) the high-dose male group reflected a
significant increase (p = 0.01). Also when squamous cell carcinoma and
papilloma (including those of the head region) were combined, only
the tumor incidence in the high-dose male group exceeded the historical
and concurrent controls (U.S. EPA, 1985 and 1986a).
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) _ ng/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
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Paraquat August, 1987
-10-
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
No suitable information was found in the available literature for the
determination of the One-day HA value for paraquat. It is therefore recommended
that the Ten-day HA value for the 10-kg child of 0.1 mg/L (100 ug/L), calculated
below, be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The rat developmental study (Hodge et al., 1978b) has been selected to
serve as the basis for the determination of the Ten-day HA value for paraquat.
In this study, Alderley Park rats were administered paraquat (100% purity)
during gestation days 6 through 15 at dose levels of 0, 1, 5 or 10 mg paraquat
ion/kg/day. There was a statistically significant (p £0.001; p = 0.05)
decrease in maternal and fetal body weight gain at the 5-mg paraquat ion/kg/day
dose; also at 5 mg/kg/day, there was a slight retardation in ossification.
The fetotoxic and maternal NOAEL identified in this study was 1 mg paraquat
lon/kg/day. An adequate study of comparable duration reported a NOAEL that
was higher than that in the study selected for derivation of the Ten-day HA.
A NOAEL of 5 mg/kg/day was identified for developmental effects, while the
maternal NOAEL was similar (1 mg/kg/day) (Hodge et al., 1978a).
Using a NOAEL of 1 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = M mq/kg bw/day) (10 kg) = 0.1 mg/L (100 ug/L)
(100) (1 L/day)
where:
1 mg/kg/day = NOAEL, based on the absence of fetotoxic and maternal
effects in rats exposed to paraquat by gavage on days
6 through 15 of gestation.
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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Paraquat August, 1987
-11-
Longer-term Health Advisory
No studies were found in the available literature that were suitable for
deriving the Longer-term HA value for paraquat. The 90-day oral study of dogs
(Sheppard, 1981) reported a NOAEL (0.5 mg ion/kg/day) which is similar to the
NOAEL (0.45 mg ion/kg/day) of the 52-week oral dog study (Kalinowski et al.,
1983) used to derive the Lifetime HA. It is, therefore, recommended that the
Drinking Water Equivalent Level (DWEL) of 0.16 mg/L (160 ug/L), calculated below,
be used for the Longer-term HA value for an adult, and that the DWEL adjusted
for a 10-kg child, 0.045 mg/L (45 ug/L), be used for the Longer-term HA value
for a child.
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, 1986b), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Kalinowski et al. (1983) has been selected to serve as the
basis for the Lifetime HA value for paraquat. In this 52-week feeding study
in beagle dogs, a NOAEL of 15 ppm (0.45 mg paraquat lon/kg/day) was identified
based on the absence of hematological, biochemical, gross pathological and
histclogical effects as well as the absence of any significant changes in
food consumption, or in body and organ weights for treated and control groups.
Adequate studies of comparable duration reported NOAELs higher than those of
the critical study selected for derivation of the Lifetime HA. A lifetime
oral study in rats (Woolsgrove et al., 1983) reported a NOAEL of 25 ppm
(about 1.25 mg/kg/day); a NOAEL of 12.5 ppm (about 1.87 mg/kg/day) was
identified for mice (Litchfield et al., 1981).
Step 1: Determination of the Reference Dose (RfD)
RfD = tO-45 mg ion/kg/day) = 0.0045 mg/kg/day
(100)
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Paraquat August, 1987
-12-
where:
0.45 mg ionAg/day «* NOAEL, based on the absence of biochemical,
hematological, gross pathological and histo-
pathological effects in dogs fed paraquat in
the diet for 52 weeks.
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 _ (0.0045 mg/kg/day) (70 kg) _ Q.16 mg/L (160 ug/L)
(2 L/day)
where:
0.0045 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Calculation of the Lifetime Health Advisory
Lifetime HA = (0.16 mg/L) (20%) = 0.003 mg/L (3 ug/L)
10
where:
0.16 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 In studies with mice, technical paraquat dichloride (32.7% paraquat
ion) did not induce significant oncogenic responses at dose levels of
0, 12.5, 37.5 or 100/125 ppm (0, 1.87, 5.6 or 15/18.75 mgAg, respec-
tively) (Litchfield et al., 1981). The oncogenic potential of paraquat
has been determined in studies in which rats were fed technical
paraquat for 113 to 124 weeks at dose levels of 0, 25, 75 and 150 ppm
(0, 1.25, 3.75 and 7.5 mg/kg/day), respectively. The incidences of
pulmonary, thyroid, skin and adrenal tumors were not clearly associated
with treatment; however, the incidence of skin carcinomas was signifi-
cantly increased (p = 0.01) in the high-dose males (Woolsgrove et al.,
1983).
0 The International Agency for Research on Cancer has not evaluated the
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Paraquat August, 1987
-13-
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986b), paraquat may be classified in
Group C: possible human carcinogen. This group is used for substances
with limited evidence of carcinogenicity in animals in the absence of
human data.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The Office of Pesticide Programs (OPP) has established tolerances on
raw agricultural commodities for paraquat ion derived from either the
bis(methyl sulfate) or dichloride salt ranging from 0.01 to 5 ppm
(U.S. EPA, 1984). The tolerances are based on an ADI of 0.0045
mg/kg/day derived from a 1-year feeding study in dogs, with a .NOAEL
of 0.45 mg/kg/day and a safety factor of 100.
0 The National Academy of Sciences (NAS, 1977) has a Suggested-No-
Adverse-Response-Level (SNARL) of 0.06 mg/L. This was calculated
using an uncertainty factor of 1,000 and a NOAEL of 8.5 mg/kg/day
identified in the 2-year rat study by Chevron Chemical Company (1975),
with an assumed consumption of 2 L/day of water by a 70-kg adult, with
the assumption that 20% of total intake of paraquat was from water.
0 American Conference of Governmental Hygenists has presented a threshold
limit value of 0.1 mg/m3 for paraquat of respirable particle sizes
(ACGIH, 1980).
VII. ANALYTICAL METHODS
0 There is no standarized method for the determination of paraquat in
water samples. A method has been reported for the estimation of para-
quat residues on various crops (FDA, 1979). In this method, paraquat
is reduced by sodium dithionite to an unstable free radical that has
an intense blue color and also a strong absorption peak at 394 run.
VIII. TREATMENT TECHNOLOGIES
0 Weber et al. (1986) investigated the adsorption of paraquat and other
compounds by charcoal and cation and anion exchange resins and their
desorption with water. They developed Freundlich adsorption-desorption
isotherms for paraquat on charcoal. When 250 mg of charcoal was added
to paraquat solutions, it exhibited the following adsorptive capacities:
37.3 and 93.2 mg paraquat/g charcoal at concentrations of 0.373 mg/L
and 37.3 mg/L, respectively. Paraquat was also adsorbed by IR-120
exchange resins (H+ and Na+ forms). The IR-120-H resin showed more
affinity towards paraquat than the IR-120-Na resin. When 665 mg of
paraquat in solution was added to 15 mg of resin, IR-120-H adsorbed
70% of paraquat while the IR-120-Na adsorbed 66% of paraquat.
0 MacCarthy and Djebbar (1986) evaluated the use of chemically modified
peat for removing paraquat from aqueous solutions under a variety of
-------
Paraquat August, 1987
-14-
experimental conditions. Paraquat sorption isotherms on treated
Irish peat were determined by equilibrating 100-mL volumes of 3.66 mg/L
paraquat with 0.1 g of peat at ambient conditions. Tests indicated
that equilibrium for paraquat was achieved after 6 days. Peat exhib-
ited the following paraquat sorption capacities: 40, 55 and 60 mg
paraquat/g peat at concentrations of 2, 4 and 6 mg/L, respectively.
The effects of pH, ionic strength and flow rate on paraquat removal
efficiency were also investigated. When 45 mL of 16-mg/L paraquat
solution was gravity fed to a column with a diameter of 6 mm that had
been packed with 700 mg treated peat, 95 to 99% paraquat removal
efficiency was reported without a significant effect by variations in
pH, ionic strength or flow rate.
In summary, several techniques for the removal of paraquat from water
have been examined. While data are not unequivocal, it appears that
adsorption of paraquat by charcoal, ion exchange and modified peat are
effective treatment techniques. However, selection of individual or
combinations of technologies for paraquat removal from water must be
based on a case-by-case technical evaluation and an assessment of
the economics involved.
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Paraquat August, 1987
-15-
IX. REFERENCES
ACGIH. 1980. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 4th ed. Cincinnati, OH: ACGIH.
Benigni, R., M. Bignami, A. Carere, G. Conti, L. Conti, R. Crebelli, E. Dogliotti,
G. Gualandi, A. Novelletto and V. Ortali. 1979. Mutational studies
with diquat and paraquat in vitro. Mutat. Res. 68:183-193.
Bullock, C.H.* 1977. The skin irritation potential of ortho paraquat 3 Ibs/
gal concentrate. Standard Oil Company of California, Report No. SOCAL
1061/30:71, August 1. MRID 00054576.
Bullock, C.H. and J.A. MacGregor.* 1977. The eye irritation potential of
ortho paraquat 3 Ibs/gal concentrate. Standard Oil Company of California,
Report No. SOCAL 1060/30:70, August 1. MRID 00054575.
Calderbank, A. 1970. The fate of paraquat in water. Outlook Agric.
6(3):128-130.
Calderbank, A. 1976. In Herbicides: Chemistry, degradation and mode of
action. 2nd ed. G. Kearney, C. Phillips and D. Kaufman, eds. New York:
Marcel Dekker.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc, Bethesda, MD.
Chevron Chemical Company. 1975. Paraquat poisoning; a physician's guide for
emergency treatment and medical management. San Francisco, CA: Chevron
Environmental Health Center. (Cited in NAS, 1977)
Clark, D.G.* 1965. The acute toxicity of paraquat. Imperial Chemical
Industries Limited. Report No. IHR/170, January 1. MRID 00081825.
Clark, D.G., T.S. McElligott and E.W. Hurst. 1966. The toxicity of paraquat.
Brit. J. Ind. Med. 23(2):126-132.
Clay, P. and M. Thomas.* 1985. Paraquat dichloride (technical liquor):
Assessment of mutagenic potential using L5178Y mouse lymphoma cells.
Imperial Chemical Industries PLC, England. Report No. CTL/P/1398,
September 24. MRID GS 0262-009.
Coats, G.E., H.H. Funderburk, Jr. and J.H. Lawrence et al.* 1964. Persistence
of diquat and paraquat in pools and ponds. Proceedings, Southern Weed
Control Conference. 17:308-320. Also in Unpublished submission
received Apr. 7, 1971 under unknown admin, no.; submitted by Chevron
Chemical Co., Richmond, CA; CDL:180000-1. MRID 00055093.
Cooke, N.J., D.C. Flenley and H. Matthew. 1973. Paraquat poisoning. Serial
studies of lung function. Q. J. Med. New Ser. 42:683-692.
Cross, M. 1985.* Paraquat dichlorde: Assessment of mutagenic potential
using L5178Y mouse lymphoma cells. Imperial Chemical Industries PLC,
England. Report No. CTL/P/1374, September 17. MRID GS 0262-009.
-------
Paraquat August, 1987
-16-
Daniel, J.W. and J.C. Gage.* 1966. Absorption and excretion of diquat and
paraquat in rats. Imperial Chemical Industries Limited, England. Brit.
J. Ind. Med. 23:133-136. MRID 00055107.
Day, S.R. and R.J. Hemingway.* 1981. 14C-Paraquat: Degradation on a sandy
soil surface in sunlight. Report No. RJ 01688. Unpublished study
submitted by Chevron Chemical Co. under Accession No. 257105.
FDA. 1970. Food and Drug Administration. Acute LDso - rat. Project No.
stated. Chambers, GA. MRID GS 0262-003.
FDA. 1979. Food and Drug Administration. Pesticide analytical manual,
revised June 1979, Food and Drug Administration, Washington, DC.
Frank, P. A. and R.D. Comes.* 1967. Herbicidal residues in pond water and
hydrosoil. Weeds. 15:210-213.
Helling, C. and B. Turner.* 1968. Pesticide mobility: Determination by
soil thin-layer chromatography. Science. 167:562-563.
Hendley, P., J.P. Leahey, C.A. Spinks, D. Neal and P.K. Carpenter.* 1976.
Paraquat: Metabolism and residues in goats. Huntingdon Research Centre,
England. Project No. AR 2680A, July 16. MRID 00028597.
Hodge, M.C.E., S. Palmer, T.M. Wright and J. Wilson. 1978a. Paraquat
dichloride: teratogenicity study in the mouse. Imperial Chemical Indus-
tries Limited, England. Report No. CTL/P/364, June 12. MRID 00096338.
Hodge, M.C.E., S. Palmer, T.M. Wright and J. Wilson. 1978b. Paraquat
dichloride: teratogenicity study in the rat. Imperial Chemical Indus-
tries Limited, England. Report No. CTL/P/365, June 5. MRID 00113714.
Howard, C.A., J. Wildgoose, P. Clay and C.R. Richardson. 1985. Paraquat
dichloride: An ^n vitro sister chromatid exchange study in Chinese
hamster lung fibroblasts. Imperial Chemical Industries PLC, England.
Report No. CTL/P/1392, September 24. MRID GS 0262-009.
Kalinowski, A.E., J.E. Doe, I.S. Chart, C.W. Gore, M.J. Godley, K. Hollis,
M. Robinson and B.H. Woollen. 1983. Paraquat: One-year feeding study
in dogs. Imperial Chemical Industries, England. Report No. CTL/P/734,
April 20. MRID 00132474.
Leahey, J.P., R.J. Hemingway, J.A. Davis and R.E. Griggs. 1972. Paraquat
metabolism in a cow. Imperial Chemical Industries Ltd, England. Report
No. AR 2374A, November 17. MRID 00036297.
Leahey, J.P., C.A. Spinks, D. Neal and P.K. Carpenter. 1976a. Paraquat
metabolism and residues in goats. Huntingdon Research Centre, England.
Project No. AR 2680 A, July 16. MRID 00028597.
Leahey, J. P., P. Hendley and C.A. Spinks. 1976b. Paraquat metabolism and
residues in pigs. Huntingdon Research Centre, England. Project No.
AR 2694 A, October 4. MRID 00028598.
-------
August, 1987
Paraquat
-17-
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food and Drug Off. U.S., Q. Bull.
Lindsay, S., P.B. Banham, M.J. Godley, S. Moreland, G.A. Wickramaratue and
B.H. Woollen. 1982. Paraquat multigeneration reproduction study in
rats: Three generation. Imperial Chemical Industries PLC, England.
Report No. CTL/P/719, December 22 and Report No. CTL/P/719S, MRID
00126783. Chevron response to EPA comments on rat reproduction study.
No date. Received by EPA on 9/10/85.
Litchfield, M.H., M.F. Sotheran, P.B. Banham, M.J. Godley, S. Lindsay, I.
Pratt, K. Taylor and B.H. Woollen. 1981. Paraquat lifetime feeding
study in the mouse. Imperial Chemical Industries Limited, England.
Report No. CTL/P/556, June 22. MRID 00087924.
MacCarthy, P. and K.E. D3ebbar. 1986. Removal of paraquat, diquat and
amitrole from aqueous solution by chemically modified peat. J. Environ.
Qual. 15(2):103-107.
McLean Head, L., J.R. Marsh and S.W. Millward. 1985. Paraquat: 4-hour acute
inhalation toxicity study in the rat. Imperial Chemical Industries.
Report no. CTL/P/1325 and CTL/P/1325, September 24.
Meister, R., ed. 1985. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Meister, R., ed. 1987. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
Pack, D.E.* 1982. Long term exposure of 14c-paraquat on a sandy soil to
California sunlight. Unpublished submission by Chevron Chemical Co.
under Accession No. 257105.
Parry, J.M. 1977. The use of yeast cultures for the detection of environmental
mutagens using a fluctuation test. Mutat. Res. 46:165-176.
Sheldon, T., C.A. Howard, J. Wildgoose and C.R. Richardson. 1985. Paraquat
dichloride: A cytogenetic study in human lymphocytes .in vitro. Imperial
Chemical Industries PLC, England. Report No. CTL/P/1351, September 3.
MRID GS 0262-009.
Sheppard, D.B. 1981. Paraquat thirteen week (dietary administration) toxicity
study in beagles. Hazleton Laboratories Europe Ltd, England. Report No.
CTL/C/1027. HLE PronectNo. 2481-72/111A, February 17. MRID 00072416.
Spinks. C.A., P. Hendley, J.P. Leahey and P.K. Carpenter. 1976. Metabolism
and residues in pigs using 14C-ring-labelled paraquat. Huntingdon Research
Centre, England. Project No. AR 2692 A, October 1. MRID 00028599.
-------
Paraquat August, 1987
-18-
Standard Oil Company. 1977. Acute dermal LD$Q - rabbit. Project No. SOCAL
1059/29:40. MRID 00054574.
STORET. 1987.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Summary of reported
pesticide incidents involving paraquat. Pesticide Incident Monitoring
System. Report no. 200. July.
U.S. EPA. 1984. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.205. July 1.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Registration standard
for paraquat. Memo from K. Locke. December 20.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Peer reviews of Paraquat
oncogenic potential. Memo from E. Rinde. September 18.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Upton, B.P., P. Hendley and M.W. Skidmore.* 1985. Paraquat: Hydrolytic
stability in water pH 5, 7 and 9. ICI Plant Protection Division.
Report series RJ0436B. Submitted Sept. 3, 1985. Chevron Chemical Co.,
Richmond, CA.
Weber, J.B., T.M. Ward and S.B. Weed. 1986. Adsorption and desorption of
diquat, paraquat, prometone. Proc. Soil Sci. Soc. Amer. 32:197-200.
Windholz, M., S. Budvari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, 10th edition. Rahway, NJ: Merck and Co., Inc.
Woolsgrove, B., R. Ashby, P. Hepworth, A.K. Whimmey, P.M. Brown, J.C. Whitney
and J.P. Finn. 1983. Paraquat: Combined toxicity and carcinogenicity
study in rats. Life Sciences Research, England. Report No. 82/1LY217/328,
October 27. MRID 00138637.
Worthing, C.R. 1983. The pesticide manual. Published by the British Crop
Council.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
PICLORAM
Health Advisory
Office of Drinking Water
U.S. Qwironmental Protection Agency
o£
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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Picloran
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-02-01
Structural Formula
Cl
(4-amino-3,5,6-trichloropicolinic acid)
Synonyms
e Amdon; ACTP; Borolin; K-PIN; Tordon (Meister, 1987).
Uses
0 Broad-spectrum herbicide for the control of broadleaf and woody plants
in rangelands, pastures and rights-of-way for powerlines and highways
(Meister, 1987).
Properties (Meister, 1987)
Chemical Formula
Molecular Weight
Physical State (Room Temp.)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor ~
Occurrence
0 Picloram has been found in 359 of 653 surface water samples analyzed
and in 5 of 77 ground water samples (STORET, 1987). Samples were
collected at 124 surface water locations and 49 ground water locations,
and picloram was found in 7 states. The 85th percentile of all
nonzero samples was 0.13 ug/L in surface water and 1.00 ug/L in
ground water sources. The maximum concentration found was 4.6 ug/L
in surface water and 1.00 ug/L in ground water.
241.6
White powder
Decomposes
21 5°C (decomposes)
6.2 x 10~7 mm Hg
0.043 g/100 mL (free acid)
40 g/100 mL (salts)
(Chlorine-like)
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Picloram August, 1987
-3-
Environmental Fate
0 The main processes for dissipation of picloram in the environment are
photodegradation and aerobic soil degradation. Field tests conducted
in Texas with a liquid formulation of picloram have indicated that
approximately 74% of the picloram originally contained in the test
ecosystems, which included the soil, water and vegetation, was
dissipated within 28 days after application (Scifres et al., 1977).
0 Photodegradation of picloram occurs rapidly in water (Hamaker, 1964;
Redemann, 1966; Youngson, 1968; Youngson and Goring, 1967), but is
somewhat slower on a soil surface (Bovey et al., 1970; Merkle et al.,
1967; Youngson and Goring, 1967). Hydrolysis of picloram is very
slow (Hamaker, 1976).
0 Laboratory studies have shown that under aerobic soil conditions, the
half-life of picloram is dependent upon the applied concentration,
and the temperature and moisture of the soil. The major degradation
product is 002; other metabolites are present in insignificant amounts
(McCall and Jefferies, 1978; Merkle et al., 1967; Meikle et al., 1970,
1974; Meikle, 1973; Hamaker, 1975). In the absence of light under
anaerobic soil and aquatic conditions, picloram degradation is extremely
slow (McCall and Jefferies, 1978).
0 Following normal agricultural, forestry and industrial applications
of picloram, long-term accumulation of picloram in the soil generally
does not occur. In the field, the dissipation of picloram will occur
at a faster rate in hot, wet areas compared to cool, dry locations
(Hamaker et al., 1967). The half-life of picloram under most field
conditions is a few months (Youngson, 1966). There is little potential
for picloram to move off treated areas in runoff water (Fryer et al.,
1979). Although picloram is considered to have moderate mobility
(Helling, 1971a,b), leaching is generally limited to the upper portions
of most soil profiles (Grover, 1977). Instances of picloram entering
the ground water are largely limited to cases involving misapplications
or unusual soil conditions (Frank et al., 1979).
III. PHARMACOKINETICS
Absorption
0 Picloram is readily absorbed from the gastrointestinal (GI) tract of
rats (Nolan et al., 1980). Within 48 hours after dosing rats with
1400 mg/kg body weight (bw), 80 to 84% of the dose was found in
urine.
0 A 500-kg Holstein cow was administered 5 mg/kg picloram in the feed
for 4 days (approximately 0.23 mg/kg/day). Ninety-eight percent of
the total dose was excreted in the urine, demonstrating nearly
complete absorption (Fisher et al., 1965).
0 Similar results were observed in three male Fischer CDF rats receiving
14c-picloram (dose not specified), where 95% of the dose was absorbed
(Dow, 1983).
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Picloram August, 1987
-4-
Distribution
0 Picloram appears to be distributed throughout the body, with the
highest concentration in the kidneys (Redemann, 1964). In rats
(strain, age and sex not specified) administered a single 20 mg/kg
dose of He-labeled picloram in food, radioactivity was found in
abdominal fat, liver, muscle and kidneys with maximum levels occurring
2 to 3 hours after dosing.
0 Hereford-Holstein steers fed picloram at daily doses of 3.2 to 23 mg/kg
for 2 weeks had tissue concentrations of 0.05 to 0.32 mg/kg in
muscle, 0.06 to 0.45 mg/kg in fat, 0.12 to 1.6 mg/kg in liver, 0.18
to 2.0 mg/kg in blood and 2 to 18 mg/kg in kidney (Kutschinski and
Riley, 1969).
0 In a similar study, two steers (strain not specified) fed 100 or 200 mg
picloram (3 or 6 mg/kg bw/day) for 31 days had picloram concentrations
of 4 or 10 mg/kg, respectively, in the kidneys, while concentrations
in other tissues (muscle, omenturn fat, heart, liver, brain) were less
than 0.5 mg/kg (Leasure and Getzander, 1964).
Metabolism
0 Picloram administered to rats or cattle was excreted in the urine in
unaltered form (Fisher et al., 1965; Nolan et al., 1980; Dow, 1983),
and no ^ 4C02 was detected in expired air of rats given 1 ^-carbon-
labeled picloram (Redemann, 1964; Nolan et al., 1980; Dow, 1983).
These studies indicate that picloram is not metabolized significantly
by mammals.
Excretion
Picloram administered to rats is excreted primarily in the urine
(Redemann, 1964; Nolan et al., 1980; Fisher et al., 1965).
Male (F344) rats that were administered a single oral dose of picloram
at 1,400 mg/kg bw, within 48 hours excreted 80 to 84% of the dose in
the urine, 15% in the feces, less than 0.5% in the bile and virtually
no measurable amount as expired CC>2 (Nolan et al., 1980).
One Holstein cow administered 5 ppm picloram in fee! for 4 consecutive
days excreted more than 98% of the dose in the urine (Fisher et al.,
1965).
In male F344 rats administered picloram at 10 mg/kg bw orally, clearance
of picloram from the plasma was biphasic, showing half-lives of 29 and
228 minutes. When administered the same dose intravenously, biphasic
clearance occurred with half-lives of 6.3 and 128 minutes (Nolan
et al., 1980).
Cattle excrete picloram primarily in the urine (Fisher et al., 1965),
although small amounts may appear in the milk (Kutschinski and Riley,
1969). In Holstein cows fed picloram for 6 to 14 days at doses of
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Picloram Au*ust' 1987
-5-
2.7 mg/kg/day or less, no picloram could be found in the milk, while
cows fed picloram at doses of 5.4 to 18 mg/kg/day had milk levels up
to 0.28 mg/L. This corresponds to 0.02% of the ingested dose. When
picloram feeding was discontinued, picloram levels in milk became
undetectable within 48 hours.
0 Nolan et al. (1983) investigated the excretion of picloram in humans.
Six male volunteers (40- to 51-years old) ingested picloram at 0.5 or
5 mg/kg in approximately 100 mL of grape juice. Seventy-six percent
of the dose was excreted unchanged in the urine within 6 hours (half-
life of 2.9 hours). The remainder was eliminated with an average
half-life of 27 hours. The authors did not report observations, if
any, of adverse effects. Thus, excretion of picloram in humans was
biphasic as had been demonstrated in rats by Nolan et al. (1980).
IV. HEALTH EFFECTS
Humans
0 No information on the health effects of picloram in humans was found
in the available literature. In the excretion study by Nolan et al.
(1983), described above, the authors did not address the presence of
toxic effects in human volunteers ingesting picloram at 0.5 or 5 mg/kg.
Animals
Short-term Exposure
0 The acute oral toxicity of picloram is low. Lethal doses have been
estimated in a number of species, with LD50 values ranging from
2,000 to 4,000 mg/kg (NIOSH, 1980; Dow, 1983).
0 In a 7- to 14-day study by Dow (1981), beagle dogs (number per group
not specified) were administered picloram (79.4% Tordon) at dose
levels of 0, 250, 500 or 1000 mg/kg/day. Based on 79.4% active
ingredient, actual doses administered were 200, 400 or 800 mg/kg/day.
The No-Observed-Adverse-Effect-Level (NOAEL) was determined to be
200 mg/kg/day, the lowest dose tested, based on the absence of reduced
food intake.
0 In a 9-day feeding study by Dow (1980a), picloram was fed to dogs
(one/dose) at dose levels of 400, 800 or 1,600 mg/kg bw/day. Picloram
was acutely toxic to female dogs at the higher doses and not toxic
at 400 mg/kg/day (the lowest dose tested), which was identified as
the NOAEL.
0 In a 32-day feeding study by Dow (1980b), picloram was administered
to mice at dose levels of 0, 90, 270, 580, 900 or 2,700 mg/kg/day.
The NOAEL was 900 mg/kg/day, and the Lowest-Observed-Adverse-Effect-
Level (LOAEL) was 2700 mg/kg/day, based on increased liver weight.
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Picloram ' 987
-6-
Dermal/Ocular Effects
8 Most formulations of picloram have been evaluated for the potential
to produce skin sensitization reactions in humans. Dow (1981) reported
in summary data that Tordon 22K was not a sensitizer following an
application as a 5% solution. A formulation of Tordon 101 containing
6% picloram acid and 2,4-D acid was not a sensitizer as a 5% aqueous
solution in humans (Gabriel and Gross, 1964). However, when the
triisopropanolamine salts of picloram and 2,4-D (Tordon 101) were
applied as a 5% solution, sensitization occurred in several individuals;
however, when applied alone, the individual components were nonreactive.
Long-term Exposure
0 Subchronic studies with picloram have been conducted by Dow (1983)
using three species (dogs, rats, mice) over periods of 3 to 6 months.
A 6-month study was conducted with beagle dogs that received picloram
at daily doses of 0, 7, 35 or 175 mgAg/day (six/sex/dose group)
(Dow, 1983). Increased liver weights were observed at the highest
dose (175 mg/kg/day) for males and females, and at the intermediate
dose (35 mgAg/day) f°r males. Therefore, the 7-mgAg/day dose level
was considered to be a NOAEL.
• In a 13-week feeding study, CDF Fischer 344 rats ( 1 5/sex/dosage group)
were fed picloram in their diet at dose levels of 0, 15, 50, 150, 300
or 500 mg/kg/day (Dow, 1983). Liver swelling was observed in both
sexes at the 150- and 300-mgAg/day dose levels. The NOAEL in this
study was identified as 50
0 Os borne-Mendel rats receiving picloram at 370 or 740 mgAg/day in the
diet for 2 years had renal disease resembling that of the natural
aging process (NCI, 1978). Increased indices of parathyroid hyperplasia,
polyarteritis, testicular atrophy and thyroid hyperplasia and adenoma
were observed. Polyarteritis may be indicative of an autoimmune
effect.
0 Ten male and female B6C3F! mice were administered picloram in their
diet at dose levels of 0, 1,000, 1,400 or 2,000 mg/kg/day for 13 weeks
(Dow, 1983). Liver weights were increased significantly (p values not
reported) in fen-ales and males at all dose levels tested.
Reproductive Effects
0 As described above in the 2-year feeding study by NCI (1978), testicular
atrophy was observed in male Osborne-Mendel rats receiving picloram at
370 or 740 mg/kg/day.
0 Groups of 4 male and 1 2 female rats were maintained on diets containing
0, 7.5, 25 or 75 mg/kg/day of Tordon (95% picloram) through a three-
generation (two litters per generation) fertility, reproduction,
lactation and teratology study (McCollister et al. , 1967). The rats
were 11 -weeks old at the start of the study and were maintained on
the test diets for 1 month prior to breeding to produce the
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Picloram August, 1987
-7-
generation. Records were kept of numbers of pups born live, born
dead or killed by the dam; litter size was culled to eight pups after
5 days. Lactation continued until the pups were 21-days old, when
they were weaned and weighed. After a 7- to 10-day rest, the dam was
returned for breeding the F^j generation. The second generation (F2a
and F3b) was derived from F2b animals after 110 days of age. Two
weanlings per sex per level of both litters of each generation were
observed for gross pathology. Gross pathology was also performed on
all parent rats and all females not becoming pregnant. Five male and
five female weanlings from each group of the F3b litter were selected
randomly for gross and microscopic examination (lung, heart, liver,
kidney, adrenals, pancreas, spleen and gonads). Picloram reduced
fertility in the 75 mg/kg/day dose group. No other effects were
noted. Based on these results, a NQAEL of 25 mg/kg/day was identified.
Developmental Effects
0 In the McCollister et al. (1967) study described above, the Flc, F2c
and F3c litters were used to study the teratogenic potential of
picloram. The dams were sacrificed on day 19 or 20 of gestation, and
offspring were inspected for gross abnormalities, including skeletal
and internal structures, and placentas were examined for fetal death
or resorptions. None were observed at any dose level. Picloram
reduced fertility in the -75-mg/kg/day dose group. Based on these
results, a NQAEL of 25 mg/kg/day was identified.
0 Thompson et al. (1972) administered picloram in corn oil to pregnant
Sprague-Dawley rats on days 6 to 15 of gestation. Four groups of 35
rats (25 for the teratology portion and 10 for the postnatal portion
of the study) received picloram at 0, 500, 750 or 1,000 mg/kg/day by
gavage. Rats were observed daily for signs of toxicity. Prebreeding
and gestation day 20 body weights were obtained on teratology rats
and prebreeding and postpartum day 21 body weights were obtained for
signs of maternal toxicity, while rats given 750 or 1,000 mg/kg/day
developed hyperesthesia and mild diarrhea after 1 to 4 days of treatment;
and 14 maternal deaths occurred between days 8 and 17 of gestation in
these dose groups. Evidence of retarded fetal growth, as reflected
by an increase in unossified fifth sternebrae, was observed in all
treatment groups but not in a dose-related manner; i.e., the occurrence
of bilateral accessory ribs was increased significantly in fetuses of
dams given 1,000 mg/kg for 10 days during gestation. At this dose
level, there was maternal toxicity and, therefore, no NOAEL was
determined. The LOAEL was 500 mg/kg, the lowest dose tested.
Mutagenicity
0 The mutagenic activity of picloram has been studied in a number of
microbial systems. Ames tests in several Salmonella typhimurium
strains indicated that picloram was not mutagenic with or without
activation by liver microsomal fractions (Andersen et al., 1972;
Torracca et al., 1976; Carere et al., 1978).
0 One study using the same system as above found picloram to be weakly
mutagenic (Ercegovich and Rashid, 1977).
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Picloram August, 1987
-8-
0 Picloram was shown to be negative in the reversion of bacteriophage
AP72 to T4 phenotype (Andersen et al., 1972), but positive in the
forward mutation spot test utilizing Streptomyces coelicolor (Carere
et al., 1978).
0 Irrespective of a weak mutagenic response in the Salmonella typhimurium
test (Ercegovich and Rashid, 1977) and a positive forward mutation,
the authors take the position that picloram is not mutagenic. This
view is supported by studies in male and female Sprague-Dawley rats
fed picloram at dose levels of 20, 200 or 2,000 mg/kg/day in which no
cytological changes in bone marrow cells were observed (Mensik et
al., 1976).
Carcinogenicity
0 Picloram (at least 90% pure) was administered by diet to Osborne-
Mendel rats and B6C3F! mice (NCI, 1978; also reviewed by Reuber,
1981). Pooled controls from carcinogenicity studies run in the same
laboratory (and room, at the Gulf South Research Institute) and over-
lapping this study by at least 1 year were used. Fifty male rats
were dosed with picloram at 208 or 417 mg/kg/day and 50 female rats
were dosed at 361 or 723 mg/kg/day. During the second year, rough
hair coats, diarrhea, pale mucous membranes, alopecia and abdominal
distention were observed in treated rats. In addition, a relatively
high incidence of follicular hyperplasia, C-cell hyperplasia and
C-cell adenoma of the thyroid occurred in both sexes. However, the
statistical tests for adenoma did not show sufficient evidence for
association of the tumor with picloram administration. An increased
incidence of hepatic neoplastic nodules (considered to be benign tumors)
was observed in treated animals. In male rats, the lesion appeared
in only three animals of the low-dose treatment group and was not
significant when compared to controls. However, the trend was signifi-
cantly dose-related in females (p = 0.016). The incidence in the
high-dose group was significant (p = 0.014) when compared with that
of the pooled control group. The incidences of foci of cellular
alteration of the liver were: female rats - matched controls 0/10,
low-dose 8/50, high-dose 18/49; male rats - matched controls 0/10,
low-dose 12/49, high-dose 5/49. Thus, there is evidence that picloram
induced benign neoplastic nodules in the livers of rats of both
sexes, but especially those of the females. Subsequent laboratory
review by the National Toxicology Program (NTP) has questioned the
findings of this study because animals with exposure to known carcinogens
were placed in the same room with these animals and cross-contamination
might have occurred. In the sane study, NCI (1978), 50 male and
50 female mice received picloram at 208 or 417, and 361 or 723 mg/k.g/day,
respectively. Body weights of mice were unaffected, and no consistent
clinical signs attributable to treatment were reported during the
first 6 months of the study, except isolated incidences of tremors
and hyperactivity. Later, particularly in the second year, rough
hair coats, diarrhea, pale mucous membranes, alopecia and abdominal
distention occurred. No tumors were found in male or female mice or
male rats at incidences that could be significantly related to treatment.
It was concluded that picloram was not a carcinogen for B6C3Fi mice.
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Picloram August, 1987
-9-
0 Dow (1986) retested picloram (93% pure) in a 2-year chronic feeding/
oncogenicity study in Fisher 344 rats. Rats (50/sex/dose) were fed
20, 60 or 200 mg/kg/day. Oncogenic effects above those of controls
were absent 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-Effect-Level
in rag/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 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 picloram. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child (20 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The 7- to 14-day study in dogs by Dow (1981) has been selected to serve
as the basis for the Ten-day HA value for picloram because dogs appear to be
the most sensitive species. Doses of 200, 400 or 800 mg/kg/day were used and
the dose of 200 mg/kg/day was identified as the NOAEL for short-term exposures
based on reduced food intake. Other short-term studies include a 9-day study
in dogs by Dow (1980a) with a NOAEL of 400 mg/kg/day and a 32-day study in
mice by Dow (1980b) with a NOAEL of 900 mg/kg/day.
Using a NOAEL of 200 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA - (200 mg/kq/day) (10 kg) = 20 mg/L (20,000 ug/1)
(100) (1 L/day)
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Picloram August, 1987
-10-
where:
200 mg/kg/day » NOAEL based on the absence of reduced feed intake in
beagle dogs exposed to picloram for 7 to 14 days.
1 0 kg = assumed body weight of a child.
1 00 - 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
The study by Dow (1983) has been selected to serve as the basis for the
Longer-term HA value for picloram because dogs have been shown to be
the species most sensitive to picloram. In this study, picloram was fed for
6 months to beagle dogs (six/sex/group) in the diet at dose levels of 0, 7,
35 or 175 mg/kg/day. At 175 mg/kg/day, the following adverse effects were
observed in both male and female dogs: decreased body weight gain, food
consumption and alanine transaminase levels, increased alkaline phosphatase
levels, absolute liver weight and relative liver weight. At 35 mg/kg/day,
increased absolute and relative liver weights were noted in males. No
compound-related effects were detected in females at 35 mg/kg/day or in males
or females at 7 mg/kg/day. Based on these data, 7 mg/kg/day was identified
as the NOAEL for dogs for a 6-month exposure.
Using this study, the Longer-term HA for a 1 0-kg child is calculated as
follows:
Longer-term HA = (7, * ° * = °'7 m^/L <700 U9/L)
where:
7 mg/kg/day = NOAEL, based on the absence of relative and absolute
liver weight changes.
1 0 kg = assumed body weight of a child.
1 00 = uncfirtainty factor, chosen in accordance with NAS/OEW
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 a 70-kg adult is calculated as follows:
where:
Longer-term HA = (7 "gAg/day) (70) = 2.45 mg/L (2,450 ug/L)
(100) (2 L/day)
7 mg/kg/day = NOAEL, based on the absence of relative and absolute
liver weight changes.
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Picloram August, 1987
-11-
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/OEW
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 NQAEL (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 divfted 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Dow (1983), chosen for the Longer-term Health Advisory has
also been chosen to calculate the Lifetime HA value for picloram. In this
study, picloram was fed for 6 months to beagle dogs (six/sex/group) in the diet
at dose levels of 0, 7, 35 or 175 mg/kg/day. At 175 mg/kg/day, the following
adverse effects were observed in both male and female dogs: decreased body
weight gain, food consumption and alanine transaciinase levels, increased
alkaline phosphatase levels, absolute liver weight and relative liver weight.
At 35 mg/kg/day, increased absolute and relative liver weights were noted in
males. Ho compound-related effects were detected in females at 35 mg/kg/dey
or in males or females at 7 mg/kg/day. Based on these data, 7 mg/kg/day was
identified as the NOAEL for dogs for a 6-month exposure. Therefore, the
Lifetime HA for picloram is determined as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (7 mg/kg/day) = 0.07 mg/kg/day
(100)
where:
7 mg/kg/day = NOAEL, based on the absence of relative and absolute
liver weight changes.
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Picloram August, 1987
-12-
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 = (0.07 mg/kg/day) (70) = 2.45 mg/L (2450 ug/L)
(2 L/day)
where:
0.07 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 = (2.45 mg/L) (20%) = 0.49 mg/L (490 ug/L)
where:
2.45 mg/L = DWEL.
20% - assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The National Cancer Institute conducted studies on the carcinogenic
potential of picloram in rats and mice (NCI, 1978; this study
was also reviewed by Reuber, 1981). In the study with mice, there
was no indication of an oncogenic response from dietary exposure
which included levels of more than 5,000 ppm picloram (723 mg/kg/day)
for the greater part of their lifetime. The rat study, however, was
negative for oncogenic effects in males, while female rats exhibited
a statistically significant increase in neoplastic nodules in the
liver. On a time-weighted average, exposures ranged up to 14,875 ppm
(743 mg/kg/day) picloram in the diet.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of picloram.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986b), picloram may be classified
in Group D: not classified. This group is generally used for sub-
stances with inadequate human and animal evidence of carcinogenicity
or for which no data are available.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA Office of Pesticide Programs has set an RfD for picloram
at 0.07 mg/kg/day (U.S. EPA, 1986b).
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Picloram August, 1987
-13-
0 Tolerances have been established for picloram in or on raw agricultural
commodities (U.S. EPA, 1986c).
0 The National Academy of Sciences (NAS, 1983) has calculated a chronic
Suggested-No-Adverse-Response-Level (SNARL) of 1.05 mg/L for picloram.
An uncertainty factor of 1,000 was used because the issue of carcino-
genicity had not yet been resolved and also because the Johnson (1971)
study used by NAS does not provide enough information for a complete
judgment of its adequacy.
VII. ANALYTICAL METHODS
0 Analysis of picloram is by a gas chromatographic (GC) method applicable
to the determination of certain chlorinated acid pesticides in water
samples (U.S. EPA, 1986d). In this method, approximately 1 liter of
sample is acidified. The compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolyzed with
potassium hydroxide and extraneous organic material is removed by a
solvent wash. After acidification, the acids are extracted and
converted to their methyl esters using diazomethane. Excess reagent
is removed, and the esters are determined by electron-capture gas
chromatography. The method detection limit has not been determined
for picloram.
VIII. TREATMENT TECHNOLOGIES
0 The manufacture of this compound has been discontinued (Meister,
1987). No information was found on treatment technologies capable of
effectively removing picloram from contaminated water.
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Picloram August, 1987
-14-
IX. REFERENCES
Andersen, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of herbi-
cides for possible mutagenic properties. J. Agr. Food Chem. 20:649-658.
Bovey, R.W., M.I. Ketchersid and M.G. Merkle.* 1970. Comparison of salt and
ester formulations of picloram. Weed Science. 18(4):447-451. MRID
00111466.
Carere, A., V.A. Ortali, G. Cardamone, A.M. Torracca and R. Raschetti. 1978.
Microbiological mutagenicity studies of pesticides in vitro. Mutat. Res.
57:277-286.
Dow.* 1980a. Dow Chemical, Texas. Nine-day feeding study — dog.
TXT:K-38323(24). Received 3/16/80. EPA Accession No. 247156.
Dow.* 1980b. Dow Chemical Laboratories, Midland, Michigan. (No Dow number).
Received May 16, 1980. EPA Accession No. 247156.
Dow.* 1981. Dow Chemical U.S.A. Repeated insults patch tests of Tordon 22K
5% solution. Received Jan. 30, 1981. MRID CDL 250606g.
Dow.* 1983. Dow Chemical U.S.A. Agricultural Products Department, an
operating unit of the Dow Chemical Company. Toxicology profile of Tordon
herbicides. Form No. 137-1640-83.
Dow.* 1986. Dow Chemical U.S.A. Picloram: A two-year dietary chronic
toxicity-oncogenicity study in Fisher 344 rats. EPA Accession
Nos. 261129-261133.
Ercegovich, C.D. and K.A. Rashid. 1977. Mutagenesis induced in mutant
strains of Salmonella typhimurium by pesticides. Am. Chem. Soc. Abstr.
174:43.
Fisher, D.E., L.E. St. John, Jr., W.H. Gutenmann, D.G. Wagner and D.J. Lisk.
1965. Fate of Bonvel T, Toxynil, Tordon and Trifluorilin in the dairy
cow. J. Dairy Sci. 48:1711-1715.
Frank, R., G.J. Sirons and B.D. Ripley. 1979. Herbicide contamination and
decontamination of well waters in Ontario, Canada, 1969-78. Pest. Mon. J.
13(3):120-127.
Fryer, J.D., P.O. Smith and J.W. Ludwig. 1979. Long-term persistence of
picloram in a sandy loam soil. J. Env. Qual. 8{1):83-86.
Gabriel, K.L., and B.A. Gross. 1964. Repeated insult patch test study with
Dow Chemical Company TORDON 101. Received November 16, 1964. MRID
0004117.
Grover, R.* 1977. Mobility of dicamba, picloram, and 2,4-D in soil columns.
Weed Science. 25:159-162. MRID 00095247.
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Picloram August, 1987
-15-
Hamaker, J.W.* 1964. Decomposition of aqueous TORDON* solutions by sunlight.
The Dow Chemical Company. Bioproducts Research. Seal Beach, California.
MRID 00111477.
Hamaker, J.W., C.R. Youngson and C.A.I. Goring.* 1967. Prediction of the
persistence and activity of Tordon herbicide in soils under field
conditions. Down to Earth. 23(2):30-36. MRID Nos. 00109132-00111430.
Hamaker, J.W.* 1975. Distribution of picloram in a high organic sediment-
water system: Uptake phase. R&D Rep. Ag-Org. Res. The Dow Chemical
Company. Midland, MI. MRID 00069075.
Hamaker, J.W. 1976. The hydrolysis of picloram in buffered, distilled water-
GS-1460. Dow Chemical Co. Agr. Prods. Dept., Walnut Creek, CA. -
Helling, C.S.* 1971a. Pesticide Mobility in Soils. I. Parameters of thin-
layer chromatography. Soil Sci. Soc. Amer. Proc. 35:732-736. MRID
00111516.
Helling, C.S.* 1971b. Pesticide Mobility in Soils. II. Applications of
soil thin-layer chromatography. Soil Sci. Soc. Amer. Proc. 35:737-743.
MRID 00044017.
Johnson, J.E. 1971. The public health implication of widespread use of the
phenoxy herbicides and picloram. Bioscience. 21:899-905.
Kutschinski, A.H., and V. Riley. 1969. Residues in various tissues of steers
fed 4-amino-3,5,6-trichloropicolinic acid. J. Agric. Food Chem. 17:283-287,
Leasure, J.K. and M.E. Getzander. 1964. A residues study on tissues from
beef cattle fed diets containing Tordon herbicide. Unpublished Report.
Midland, MI. The Dow Chemical Company. GS-P 141. Reviewed in NRCC.
McCall, P.J. and T.K. Jeffries.* 1978. Aerobic and anaerobic soil degradation
of 14c-picloram. Agricultural Products R&D Report GH-C 1073, The Dow
Chemical Company, Midland, MI.
McCollister, D.D., J.R. Copeland and F. Oyen.* 1967. Dow Chemical Company,
Toxicology Research Laboratory, Midland, MI. Results of fertility and
reproduction studies in rats maintained on diets containing TORDON*
herbicide. Received January 24, 1967 under OF0863, CDL:094525-H.
MRID 00041098. EPA Accession No. 091152.
Meikle, R.W.* 1973. Comparison of the decomposition rates of picloram and
4-amino-2,3,5-trichloropyridine in soil. Unpublished report. MRID
00037883.
Meikle, R.W., C.R. Youngson, R.T. Hedlund, C.A.I. Goring and W.W. Addington.*
1974. Decomposition of picloram by soil microorganisms: A proposed
reaction sequence. Weed Science. 22:263-268. MRID 00111505.
Meikle, R.W., C.R. Youngson and R.T. Hedlund.* 1970. Decomposition of picloram
in soil: Effect of a pre-moistened soil. Report of The Dow Chemical
Company. GS-1097.
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Picloram August, 1987
-16-
Meister, R., ed. 1987. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Merkle, M.G., R.W. Bovey and F.S. Davis.* 1967. Factors affecting the
persistence of picloram in soil. Agronomy Journal. 39:413-415.
MRID 00111441.
Mensik, D.C., R.V. Johnston, M.N. Pinkerton and E.B. Whorten.* 1976. The
cytogenic effects of picloram on the bone marrow of rats. Unpublished
report. The Dow Chemical Company. Freeport, TX. 11 pp.
MAS. 1983. National Academy of Sciences. Drinking water and health. Vol. 5.
Washington, DC: National Academy Press, pp. 60-63.
NCI. 1978. National Cancer Institute. Bioassay of picloram for possible
carcinogenicity. Technical Report Series No. 23. Washington, DC:
Department of Health, Education and Welfare.
NIOSH. 1980. National Institute for Occupational Safety and Health. RTECS,
Registry of Toxic Effects of Chemical Substances. Vol. 2. U.S. Department
of Health and Human Services, p. 354. DHHS Publ. (NIOSH) 81-116.
Nolan, R.J., F.A. Smith, C.J. Mueller and T.C. Curl. 1980. Kinetics of
14c-labeled picloram in male Fischer 344 rats. Unpublished report.
Midland, MI. The Dow Chemical Co. 34 pp.
Nolan, R.J., N.L. Freshour, P.E. Kastl and J.H. Saunders. 1983. Pharmaco-
kinetics of picloram in human volunteers. Toxicologist. 4:10.
Redemann, C.T. 1964. The metabolism of 4-amino-3,5,6-trichloropicolinic
acid by the rat. Unpublished report. Seal Beach, CA: The Dow Chemical
Co. GS-623. Reviewed in NRCC. 1974. National Research Council.
Picloram: The effects of its use as a herbicide on environmental quality.
Ottawa, Canada. NRCC No. 13684.
Redemann, C.T.* 1966. Photodecomposition rate studies of 4-amino-3,5,6-
trichloropicolinic acid. The Dow Chemical Company. Bioproducts Research.
Walnut Creek, CA.
Reuber, M.D. 1981. Carcinogenicity of Picloram. J. Tox. Environ. Health.
7:207-222.
Scifres, C.J., H.G. McCall, R. Maxey and H. Tai. 1977. Residual properties
of 2,4,5-T and picloram in sandy rangeland soils. J. Env. Qual.
6(11:36-42.
STORET. 1987.
Thompson, D.J., J.L. Emerson, R.J. Strebing, C.C. Gerbig and V.B. Robinson.
1972. Teratology and postnatal studies on 4-amino-3,5,6-trichloro-
picolinic acid (picloram) in the rat. Food Cosmet. Toxicol. 10:797-803.
-------
Picloram August, 1987
-17-
Torracca, A.M., G. Cordamone, V. Ortali, A. Carere, R. Raschette and G. Ricciardi.
1976. Mutagenicity of pesticides as pure compounds and after metabolic
activation with rat liver microsomes. Atti. Assoc. Genet. Ital. 21:28-29.
(In Italian; abstract in English)
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for car-
cinogenic risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Registration
standard for picloram. Office of Pesticide Programs, Washington, DC.
U.S. EPA. 1986c. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.292.
U.S. EPA. 1986d. U.S. Environmental Protection Agency. U.S. EPA Method #3
- Determination of chlorinated acids in ground water by GC/ECD, January
1986 draft. Available from U.S. EPA's Environmental Monitoring and
Support Laboratory, Cincinnati, OH.
Youngson, C.R.* 1966. Residues of Tordon in soils from fields treated for
selective weed control with tordon herbicide. Report by the Dow Chemical
Company. Bioproducts Research, Walnut Creek, CA. MRID 00044023.
Youngson, C.R., and C.A.I. Goring.* 1967. Decomposition of Tordon herbicides
in water and soil. GS-850 Research Report, The Dow Chemical Company.
Bioproducts Research, Walnut Creek, CA. MRID 00111415.
Youngson, C.R.» 1968. Effect of source and depth of water and concentration
of 4-amino-3,5,6-trichloropicolinic acid on rate of photodecomposition
by sunlight. The Dow Chemical Company. Agricultural Products Research,
Walnut Creek, CA. MRID 00059425.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
PRON AMIDE
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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Pronamide
August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 23950-58-5
Structural Formula
0 H CH3
C-N-C-CSCH
i
CH9
3,5-Dichloro(N-1,1-dimethyl-2-propynyl)benzamide
Synonyms
Uses
Kerb*; Kerb* SOW; Propyzamide; RH315 (Meister, 1983),
Pronamide is used as an herbicide for pre- or postemergence weed and
grass control in small, seeded legumes grown for forage or seed,
southern turf, direct seeded or transplanted lettuce, endive, escarole,
woody ornamentals, nursery stock and Christmas trees (Meister, 1983).
C12H11C12ON
256.14
White crystals
154 to 156°C
8.5 x 10-5 mm Hg
0.48 gm/cc
0.015 mg/L
3.05 to 3.27
Properties (NIOSH, 1985; TDB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Vapor Pressure (25°C)
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Pronamide has been found in 18 of 258 ground water samples analyzed
(STORET, 1987). No surface water samples were collected, and samples
were collected from 252 ground water locations. Pronamide was found
only in California. The 85th percentile of all nonzero samples was
1 ug/L, and the maximum concentration found was 1 ug/L.
Environmental Fate
0 14c-Pronamide (100% radiopurity) at 1.5 ppm hydrolyzes very slowly
(10% of applied material) in sterile, deionized water buffered to
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Pronamide August, 1987
-3-
pH 5, 7, and 9 and incubated at 20°C for 28 days in the dark (Rohm
and Haas Bristol Research Laboratories, 1973). The following minor
hydrolysis products were identified: RH-24,644 (2-(3,5-dichlorophenyl)-
4,4-dimethyl-5methyleneoxazoline); RH-24,580 (3,5-dichloro-N-{1,1-
dimethylacetonyl) benzamide); and RH-25,891 (2-(3,5-dichlorophenyl)-
4,4-dimethyl-5-hydroxymethyl-oxazoline). Similar results were obtained
in other hydrolysis studies (Rohm and Haas Bristol Research Laboratories,
1970).
Pronamide has a half-life of 10 to 120 days in aerobic soils (Fisher,
1971; Walker, 1976; Walker and Thompson, 1977; Walker, 1978; Hance,
1979;). Complete experimental conditions and purity were not specified,
and/or a formulated product was applied. The degradation rate does
not appear to depend upon soil texture. However, increasing soil
temperature, and to a lesser extent, soil moisture and pH enhance
pronamide degradation. The major degradates are RH-24,580 and
RH-24,644. Soil sterilization greatly reduced the degradation rate
of pronamide. Pronamide (at a recommended application rate of 0.5 to
2 Ib/A) does not inhibit the growth or CO2 evolution of bacteria and
fungi (Lashen, 1970).
Pronamide is moderately mobile in soils ranging in texture from loamy
sand to clay based on preliminary soil column and adsorption/desorption
tests (Walker and Thompson, 1977; Rohm and Haas Company, 1971; Fisher
and Satterthwaitte, 1971). The two major degradates of pronamide
(RH-24,580 and RH-24,644) are mobile in sand and clay soils (Fisher,
1973). The mobility of pronamide and its two major degradates tends
to decrease as the organic matter content, clay content and cation
exchange capacity of the soil increases.
The dissipation rate of pronamide from terrestrial field sites is
quite variable, with half-lives ranging from 10 to 90 days (Benson,
1973; Walker, 1976; Hance et al., 1978a; Hance et al., 1978b; Kostowska
et al., 1978; Walker, 1978; Zandvoort et al., 1979). Data are insuf-
ficient to determine the effect, if any, of meteorological conditions
or the role leaching may play in pronamide dissipation.
The environmental fate of pronanide is the subject of several unpub-
lished, undated reports (Cummings and Yih; Fisher and Cummings; Rohm
and Haas; Satterthwaite and Fisher; Yih).
III. PHARMACOKINETICS
Absorption
0 No information on the absorption of pronamide was found in the
available literature.
Distribution
0 No information on the distribution of pronamide was found in the
available literature.
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Pronaoide August, 1987
-4-
Metabolism
0 About 54 and 0.6% of the radioactivity was recovered as unmetabolized
Kerb* in the feces and urine, respectively, of rats treated orally with
(14c-carbonyl)-pronamide (dose not specified) (Yih and Swithenbank,
undated). The major metabolite in the feces was 2-(3,5-dichlorophenyl)-
4,4-dimethyl-5-hydroxyraethyloxazoline (15%), and the major metabolites
in the urine were °-(3,5-dichlorobenzamido) isobutyric acid (22.4%),
B-(3,5-dichlorobenzamido)-a-hydroxy-B methyl-butyric acid (19.2%), and
two unknown metabolites (24.1 and 16.7%).
0 Unmetabolized Kerb* did not appear in the urine of cows treated orally
with (14C-carbonyl) Kerb*; the major metabolite was ^-(3,5-dichloro-
benzamido)-a-hydroxy-B-methyl-butyric acid (71.4%)(Yih and Swithenbank,
undated).
Excretion
0 After oral ingestion of radiolabeled Kerb* by rats, unmetabolized
Kerb* accounted for 54 and 0.6% of the radioactivity recovered in
feces and urine, respectively. In the cow, oral ingestion of Kerb*
produced no unmetabolized Kerb* in the urine (Yih and Swithenbank,
undated).
IV. HEALTH EFFECTS
Humans
No information on the health effects of pronamide in humans was found
in the available literature.
Animals
Short-term Exposure
0 The acute oral LD^Q in rats for pronamide (technical) is in the range
of 8,350 Dig/kg bw (Meister, 1984) to 16,000 mg/kg bw (Powers, 1970s).
Dermal/Ocular Effects
0 Pronamide is not a primary dermal irritant to albino rabbits. In two
separate studies, an aqueous paste of 500 mg pronamide [50% active
ingredient (a.i.)] was applied to the skin of six rabbits for 24 hours
(Powers, 1970c; Regel, 1972). No signs of irritation were observed
by Powers (1970c). Twenty-four hours after exposure, Regel (1972)
observed erythema, which subsided at 72 hours.
8 Powers (197Cb) administered 100 mg of Kerb* (50% a.i.) in the con-
junctival sac of 12 rabbits. Eye irritation and chemosis were noted
at 24 hours but disappeared by day 7, as confirmed by fluorescein
examination.
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Pronamide August. 1987
-5-
Lonq-term Exposure
0 Rats (10/sex/dose) were fed a diet containing 0, 50, 150, 450, 1,350
or 4,050 ppm pronamide (100% a.i.) for 3 months (Larson and Borzelleca,
1967a). This corresponds to 0, 2.5, 7.5, 22.5, 67.5 or 202.5
nig/kg/day, assuming 1 ppm in feed is equivalent to 0.05 mg/kg/day
(Lehman, 1959). Significant body weight depression was observed at
the 4,050 ppm dose level. Initial significant body weight depression
also occurred in the rats fed 1,350 ppm, but disappeared on continued
feeding. At the 150 ppm dose, absolute and relative liver weights in
females were-significantly higher than in controls; no histological
lesions were seen, and no dose-related trend was observed for this
increase in relative liver weight. Individual data were not presented
for organ weights and several other parameters, clinical observations
were not presented and analytical determination of the test compound
was not reported. The No-Observed-Adverse-Effect-Level (NOAEL)
identified in this study was 2.5 mg/kg/day.
0 Beagle dogs (10 months old; one/sex/dose) were fed a diet containing
0, 450, 1350 or 4050 ppm pronamide (100% a.i.) for 3 months (Larson
and Borzelleca, 1967b). This corresponds to approximate doses of
0, 10, 30 or 90 mg/kg/day, assuming 1 ppm in feed is equivalent to
0.025 mg/kg/day (Lehman, 1959). A decrease in weight gain and food
consumption and an increase in serum alkaline phosphatase, liver
weight and liver-to-body weight ratios, as compared to controls,
were seen in the animals dosed at 4,050 ppm. No histological changes
were seen in the livers. The hematological and urinalysis findings
were within normal ranges. The NOAEL identified in this study was
30 mg/kg/day.
0 In a 2-year feeding study in beagle dogs (four/sex/dose) the addition
of pronamide (97% a.i.) to the diet at dose levels of 0, 30, 100 or
300 ppm (0, 0.75, 2.5 or 7.5 mg/kg/day, assuming 1 ppm in feed is
equivalent to 0.025 mg/kg/day; Lehman, 1959) caused no adverse effects
at any of the doses tested (Larson and Borzelleca, 1970b). A NOAEL
of 7.5 mg/kg/day (the highest dose tested) was identified in this
study.
0 Smith (1974) administered Kerb* (97% a.i.) to 6-week-old (C57 BL16 x
C3H AnfjF! male and female mice (100/sex/dose), for 78 weeks at
dietary concentrations of 0, 1000 or 2000 ppm (0, 150 or 300 mg/kg/day,
assuming 1 ppm in feed is equivalent to 0.15 mg/kg/day; Lehman, 1959)
pronamide. Male and female mice that ingested 2000 ppm gained sig-
nificantly less weight (p <0.05); males also exhibited adenomatous
hyperplasia, degeneration, hyperplasia, intrahepatic cholestasis,
necrosis and/or fatty changes of the liver. Liver weights were
significantly increased over controls for males and females in both
treatment groups. Based on this information, a Lowest-Observed-Adverse-
Effect-Level (LOAEL) of 1,000 ppm (150 mg/kg/day) was identified.
0 Newberne et al. (1982) administered pronamide (94% a.i.) to male
B6C3F1 mice at dose levels of 0, 20, 100, 500 or 2,500 ppm (0, 3,
15, 75 or 375 mg/kg/day, assuming 1 ppm in feed is equivalent to
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Pronamide August, 1987
-6-
0.15 mg/kg/day: Lehman, 1959) for up to 24 months. Another group was
fed 2,500 ppm (375 mg/kg/day) pronamide for 6 months. The mean body
weight of the mice fed 2,500 ppm was significantly depressed at 14 days
and thereafter throughout the study. At the 24-month sacrifice, the
mean body weight of this group was approximately 70% of the control
group. Survival of the mice was unaffected. The highest dose level
(2,500 ppm) resulted in liver lesions including bile duct hyperplasia,
parenchymal cell hypertrophy, parenchymal cell necrosis, hyperplasia
and cholestasis at all time periods examined. Based on this infor-
mation, a NOAEL of 500 ppm (75 mg/kg/day) was identified.
Reproductive Effects
0 In a teratogenicity study in New Zealand White rabbits (18/dose),
pronamide was administered at levels of 0, 5, 20 or 80 mg/kg/day
(technical, 97% pure) during gestation days 7 to 1 9 (Costlow and
Kane, 1985). Five abortions were observed in the 80 mg/kg/day group.
There were no compound-related effects on the incidence of implantations,
resorptions, fetal deaths or on fetal body weight at any dose tested.
Maternal toxicity (anorexia, vacuolation of hepatocytes) was observed
in the 20-mg/kg/day group. A NOAEL of 20 mg/kg/day was identified
based upon the absence of developmental/reproductive effects and a
NOAEL of 5 mg/kg/day was identified based upon the absence of maternal
toxicity.
0 In a three-generation reproduction study, 20 to 25 albino CD rats were
fed a diet containing pronamide (RH-315; purity not stated) at dose
levels of 0, 30, 100 or 300 ppm (Larson and Borzelleca, 1970c).
Assuming 1 ppm in the diet is equivalent to 0.05 mg/kg/day, this
corresponds to doses of 0, 1.5, 5 or 15 mg/kg/day (Lehman, 1959).
The authors reported no adverse reproductive effects in parents or
pups, but individual animal data were not available to validate the
above conclusions. Based on this information a NOAEL of 300 ppm (15
the highest dose tested) was identified.
Developmental Effects
0 In a teratogenicity study in New Zealand White rabbits (18/dose),
pronamide was administered at levels of 0, 5, 20 or 80 mg/kg/day
(technical, 97% pure) during gestation days 7 to 1 9 (Costlow and
Kane, 1985). An inc-eased incidence of gross and microscopic liver
lesions, one materna^ death, five abortions and a significant
(p <0.05) decrease in the maternal body weight gain were observed at
the 80-mg/kg/day dose. At the 20-mg/kg/day dose, rabbits exhibited
anorexia, vacuolation of hepatocytes and a slight decrease in body
weight gain. There were no compound-related effects on the incidence
of implantations, resorptions, fetal deaths or on fetal body weight
at any dose tested. The NOAEL in this study was 5 mg/kg/day based
on maternal effects, and 80 mg/kg/day based on developmental effects.
0 In a study designed to evaluate fetal development, adult female rats
(FDRL) were administered 0, 7.5 or 15 mg/kg/day pronamide by gavage
in corn oil from days 6 through 16 of gestation (Vogin, 1972). No
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Pronamide August, 1987
-7-
adverse effects were reported for the mean number of implantation
sites, the number of live or dead fetuses or the mean fetal weight.
The authors concluded that pronamide administered orally to rats at
doses up to 15 mg/kg/day was not teratogenic, but individual animal
data were not available to validate these conclusions. Based on this
information a NOAEL of 15 mg/kg/day (the highest dose tested) was
identified.
Mutagenicity
0 In a cytogenetic study, pronamide (Kerb®, analytical) administered
by intragastric intubation at dose levels of 5, 50 or 500 mg/kg to
rats did not produce any aberrations of the bone marrow chromosomes
(Fabriaio, 1973).
Carcinogenicity
0 In a study evaluating the carcinogenic potential of Kerb®, 6-week-old
(C57 BL16 x C3H Anf)F, male and female mice (100/sex/dose) were fed
pronamide (97% a.i.) in the diet at doses of 0, 1,000 or 2,000 ppm
(0, 150 or 300 mg/kg/day, assuming 1 ppm in feed is equivalent to
0.15 mg/kg/day; Lehman, 1959) for 78 weeks (Smith, 1974). Male and
female mice that ingested 2,000 ppm gained significantly less weight
(p <0.05); the animals also gained slightly less weight at the 1,000-ppm
level, but the change was not significant. No increase in tumors was
observed for female mice treated with pronamide over controls. For
male mice, a total of 35 of the 99 animals in the high-dose group,
21 of the 100 animals in the low-dose group and 7 of the 100 animals
in the control group developed hepatic neoplasms, of which 24, 18
and 7 were carcinomas in the high-dose, low-dose and control groups,
respectively. A total of 28 of 99 male mice that ingested 2,000 ppm
exhibited intrahepatic cholestasis, but did not have carcinomas of
the liver.
0 In a 2-year study in male B6C3F! mice (Newberne et al., 1982),
pronamide was fed to the animals (63 animals/dose) at dose levels of 0,
20, 100, 500 or 2,500 ppm (0, 3, 15, 75 or 375 mg/kg/day, assuming 1
ppm in feed is equivalent to 0.15 mg/kg/day; Lehman, 1959). Another
group was fed 2,500 ppm ('375 mg/kg/day) pronaaide for 6 months. The
mean body weight of mice fed 2,500 ppm was significantly depressed at
14 days and thereafter throughout the study. At the 24-month sacrifice,
the mean body weight of this group was approximately 70% of the con-
trol group. Survival of the mice was unaffected. The highest dose
(2,500 ppm) resulted in liver lesions, including bile duct hyperplasia,
parenchymal cell hypertrophy, parenchymal cell necrosis, hyperplasia
and cholestasis at all time periods examined. At 18 months, the
2,500-ppm dose group had increased parenchymal cell neoplasms, but
this was not statistically different from the controls. At 24 months,
there was a statistically significant increased incidence of hepatic
adenomas and carcinomas in the 500- and 2,500-ppm dose groups. The
incidence of hepatic carcinomas was 5/63, 9/63, 12/63, 18/63 and
14/61 in the control, 20-ppm, 100-ppm, 500-ppm and 2,500-ppm groups,
respectively. Thus, the liver appears to be the target organ for
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Pronamide August, 1987
-8-
neoplasia. According to the authors, hypertrophy and hyperplasia
are not uncommon in untreated older mice of this strain. However,
pronamide tended to shift the onset of these lesions to an earlier age.
0 Pronamide in the diet at dose levels of 0, 30, 100 or 300 ppm (0,
1.5, 5 or 15 mg/kg/day, assuming 1 ppm in feed is equivalent to
0.05 mg/kg/day; Lehman, 1959) fed to rats (30/sex/group) for 2 years
did not produce any carcinogenic effects (Larson and Borzelleca,
1970a). However, doses used in this study were too low to assess the
carcinogenic potential of pronamide.
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 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/OOW 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 pronaaiide. It is therefore
recommended that the Lifetime HA value of 0.052 mg/L (52 ug/H be used at
this time as a conservative estimate of the One-day HA value for pronamide.
Ten-day Health Advisory
Little information is available on the acute toxicity of pronamide.
Toxicity from acute exposure to pronamide has been assessed in three
reproduction/teratology studies, but it is not possible to evaluate the
most sensitive end point for acute toxicity from these studies. No effects
were observed in rats exposed to pronamide via gavage (Vogin, 1972) or in
feed (Larson and Borzelleca, 1967b) at doses as high as 15 mg/kg/day. No
higher doses were tested in the rat, but higher doses have been tested in the
rabbit (Costlow and Kane, 1985). In this study, New Zealand White rabbits
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Pronanide August, 1967
-9-
were administered pronamide during gestation days 7 through 19 at dose levels
of 0, 5, 20 or 80 mg/kg/day. Toxic effects observed at the highest dose
include a statistically significant decrease in maternal body weight gain
and an increased incidence of gross and microscopic liver lesions. Less
significant effects on body weight and liver toxicity were observed at the
20-mg/kg/day dose, and a NOAEL of 5 mg/kg/day was identified. This value
is similar to the NOAEL identified from a 2-year feeding study in dogs
(7.5 mg/kg/day; Larson and Borzelleca, 19705), which is used as the basis
for the Lifetime HA. Considering the limitations of the database on pronamide,
it is therefore recommended that the Lifetime HA value of 0.052 mg/L (52 ug/L),
calculated below, be used at this time as a conservative estimate of the
Ten-day HA value for pronamide.
Longer-term Health Advisory
Liver toxicity has been observed after acute, subchronic and chronic
administration of pronamide to experimental animals. Adverse effects on the
liver have been observed after acute exposure of rabbits to 80 mg/kg/day via
gavage (Costlow and Kane, 1985), subchronic exposure of rats and dogs to
7.5 mg/kg/day and 90 mg/kg/day, respectively (Larson and Borzelleca, 1967a,b),
and chronic feeding of 300 and 375 mg/kg/day to mice (Smith, 1974; Newberne
et al, 1982). In contrast to the subchronic rat feeding study, a NOAEL of
15 mg/kg/day was identified in a 2-year rat feeding study (Larson and
Borzelleca, 1970a); however, this study was invalidated (U.S. EPA, 1985).
Both rat studies suffer similar deficiencies, which make them unsuitable to
serve as the basis for HA values (U.S. EPA, 1985a). Considering the limita-
tions of the database on pronamide and the potential for this compound to
cause liver damage, it is therefore recommended that the Lifetime HA value
of 0.052 mg/L (52 ug/L) be used at this time as a conservative estimate of
the Longer-term HA value for pronamide.
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
(RfO), 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 chroric (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
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Pronamide August, 1987
-10-
carcinogenic potential (U.S. EPA, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
Two-year chronic pronamide feeding studies have been performed in three
species: the rat (Larson and Borzelleca, 1970a), dog (Larson and Borzelleca,
1970b), and mouse (Newberne et al., 1982). For the rat and dog studies, only
low doses were used and no toxic effects were observed. The highest doses
tested, 15 mg/kg/day (rat) and 7.5 mg/kg/day (dog), were identified as NOAELs
for these studies. Because of various deficiencies in the rat study, this
study was not validated (U.S. EPA, 1985), and is therefore not acceptable as
the basis for the Lifetime HA value. The 2-year study performed on mice
(Newberne et al., 1982) was rejected as the basis for the Lifetime HA because
of the relative insensitivity of mice to pronamide compared to other species.
The NOAEL of 75 mg/kg/day identified in this study was higher than doses
causing liver toxicity in subchronic feeding studies in both the rat and dog
(Larson and Borzelleca, 1967a,b). Taking all of these studies into consid-
eration, the 2-year feeding study in dogs (Larson and Borzelleca, 1970b) was
selected as the basis for determination of the Lifetime HA for pronamide.
In this study, beagle dogs fed a diet containing pronamide at dose levels of
0, 30, 100 or 300 ppm (0, 0075, 2.5 or 7.5 mg/kg/day) for 2 years showed no
adverse effects at any of the doses tested. A NOAEL of 7.5 mg/kg/day (the
highest dose tested) was identified in this study.
Using a NOAEL of 7.5 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (7.5 mg/kgyday) = 0.075 mg/kg/day
(100) 7
where:
7.5 mg/kg/day = NOAEL, based on the absence of adverse effects in
dogs administered pronamide in the diet for 2 years.
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*075 mg/kg/day) (70 kg) = 2.6 mg/L (2,600 ug/L)
2 L/day
where:
0.075 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
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Pronamide August, 1987
-11-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA - (2.6 mg/L) (20%) = 0>052 mg/L (52 ug/L)
(10)
where:
2.6 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986a), pronamide has tentatively
been classified in Group C: possible human carcinogen. This category
is for substances with limited evidence of carcinogenicity in animals
in the absence of human data.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 A Provisional Acceptable Daily Intake (PADI) of 0.0750 mg/kg/day and
a calculated Theoretical Maximum Residue Concentration (TMRC) of
0.0409 ing/day that utilizes 0.91% of the PADI has been established
(U.S. EPA, 1985a).
0 Residue tolerances have been established for pronamide and its metabo-
lites in or on raw agricultural commodities that range from 0.02 ppm
to 10.0 ppm (U.S. EPA, 1985b).
VII. ANALYTICAL METHODS
0 Analysis of pronamide is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986o). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
pronamide, but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Reverse osmosis (RO) is a promising treatment method for pesticide-
contaminated water. As a general rule, organic compounds with
molecular weights greater than 100 are candidates for removal by RO.
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Pronamide August, 1987
-12-
Larson et al. (1982) report 99% removal efficiency of chlorinated
pesticides by a thin-film composite polyamide membrane operating at a
maximum pressure of 1,000 psi and at a maximum temperature of 113°F.
More operational data are required, however, to specifically determine
the effectiveness and feasibility of applying RO for the removal of
pronamide from water. Also, membrane adsorption must be considered
when evaluating RO performance in the treatment of pronamide-contami-
nated drinking water supplies.
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Pronamide August, 1987
-13-
IX. REFERENCES
Benson, N.R. 1973. Efficacy, leaching and persistence of herbicides in
apple orchards. Bulletin 863. Washington State University, College of
Agriculture Research Center.
Costlow, R.D., and W.W. Kane.* 1985. Teratology study with Kerb technical (no
clay) in rabbits. Unpublished study no. 83R-026 prepared and submitted
by Rohm and Haas Company, Spring House, PA. Accession no. 256590.
Cummings, T.L., and R.Y. Yih. Undated. Metabolism of Kerb (3,5-dichloro-N-
(l,l-dimethyl-2-propynyl)benzamide) in different types of soil.
Unpublished report prepared by Rohm and Haas Co., Philadelphia, PA.
Memorandum Report No. 52.
Fabrizio, P.O.A.* 1973. Final report: Cytogenetic study: Kerb analytical.
Unpublished report no. CDL:093756-D prepared by Litton Bionetics, Inc.,
Kensington, MD for Rohm and Haas Company, Philadelphia, PA. April 16.
MRID 00038031.
Fisher, J.D. 1971. Dissipation and metabolism study of Kerb in soil and its
effects on microbial activity. Unpublished report prepared by Rohm and
Haas Co., Philadelphia, PA. Lab. 11 Research Report No. 11-229.
Fisher J.D. 1973. Soil leaching study with Kerb degradation products RH-24,
580 and RH-24,644. Unpublished report prepared by Rohm and Haas Co.,
Philadelphia, PA. Tech. Report No. 3923-73-4.
Fisher, J.D., and T.L. Cummings. Undated. Biodegradation study of carbonyl-
14c-Kerb and ring-!4C-3,5-dichlorobenzoate in a semicontinuous activated
sludge test. Unpublished study prepared by Rohm and Haas Co, Philadelphia,
PA. Report No. 16.
Fisher, J.D., and S.T. Satterthwaite. 1971. Leaching and metabolism studies
of He-Kerb in soils. Unpublished report prepared by Rohm and Haas Co.,
Philadelphia, PA. Lab. 11 Research Report No. 11-228.
Hance, R.J. 1979. Effect of pH on the degradation of atrazine, dichlorprop,
linuron and propyzamide in soil. Pestic. Sci. 10(1):83-36.
Hance, R.J., P.D. Smith, T.H. Byast and E.G. Cotterill. 1978a. Effects of
cultivation on the persistence and phytutoxicity of atrazine and propy-
zamide. Proc. Br. Crop Prot. Conf. Weeds. 14(2):541-547.
Hance, R.J., P.D. Smith, E.G. Cotterill and D.C. Reid. 1978b. Herbicide
persistence: Effects of plant cover, previous history of the soil and
cultivation. Med. Fac. Landbouww. Rijksuniv. Gent. 43(2):1127-1134.
KostowsJca, B., J. Rola and H. Slawinska. 1978. Decomposition dynamics of
propyzamide in experiments with winter rape. Pamiet. Pulawski.
70:199-205.
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Pronamide August, 1987
-14-
Larson, P.S., and J.F. Borzelleca.* 1967a. Toxicologic study on the effect
of adding RH-315 to the diet of rats for a period of three months. Unpub-
lished study no. CDL:091422-D prepared by the Medical College of Virginia,
Dept. of Pharmacology, for Rohm and Haas Company, Philadelphia, PA.
November 27. MRID 00085506.
Larson, P.S., and J.F. Borzelleca.* 1967b. Toxicologic study on the effect of
adding RH-315 to the diet of beagle dogs for a period of three months.
Unpublished study no. CDL:091422-E prepared by the Medical College of
Virginia, Dept. of Pharmacology, for Rohm and Haas Company, Philadelphia,
PA. November 22. MRID 00085507.
Larson, P.S., and J.F. Borzelleca.* 1970a. Toxicologic study on the effect
of adding RH-315 to the diet of rats for a period of two years. Unpub-
lished study no. CDL:004357-A prepared by the Medical College of Virginia,
Dept. of Pharmacology, for Rohm and Haas Company, Philadelphia, PA.
June 11. MRID 00133111.
Larson, P.S., and J.F. Borzelleca.* 1970b. Toxicologic study on • the effect
of adding RH-315 to the diet of beagle dogs for a period of two years.
Unpublished study no. CDL:090918-A prepared by the Medical College of
Virginia, Dept. of Pharmacology, for Rohm and Haas Company, Philadelphia,
PA. June 12. MRID 00107949.
Larson, P.S., and J.F. Borzelleca.* 1970c. Three-generation reproduction study
on rats receiving RH-315 in their diets. Unpublished study prepared by
the Medical College of Virginia, Dept. of Pharmacology, for Rohm and Haas
Company, Philadelphia, PA. April 11. MRID 00107950.
Larson, R.E., P.S. Cartwright, P.K. Eriksson and R.J. Petersen. 1982.
Applications of the FT-30 reverse osmosis membrane in metal finishing
operations. Paper presented at Tokohama, Japan.
Lashen, E.S. 1970. Inhibitory effects of Kerb and Kerb transformation
products on typical soil microorganisms. Unpublished report prepared
by Rohm and Haas Co., Philadelphia, PA. Memorandum Report No. 22.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Newberne, P.M., R.G. McConnell and E.A. Essigmann.* 1982. Chronic study in
the mouse. Final report no. 81RC-157 prepared by the MIT Animal Pathology
Laboratory. Submitted by Rohm and Haas Company. August 10. EPA Accession
No. 248233.
NIOSH. 1985. National Institute for Occupational Safety and Health. Registry
of Toxic Effects Chemical Substances.
-------
Pronamide August, 1987
-15-
Powers, M.B.* 1970a. Final Report (Study 1) - Acute Oral - Rats. Unpublished
study, Project No. 417-337, prepared by TRW, Inc., Vienna, VA for Rohm
and Haas Co., Philadelpia, PA, dated October 6, 1970.
Powers, M.B.* 1970b. Final Report (Study 2) - Draize Eye - Rabbits. Unpub-
lished study. Project No. 417-337, prepared by TRW, Inc., Vienna, VA for
Rohm and Haas Co., Philadelpia, PA, dated October 6, 1970. MRID 00083663.
Powers, M.B.* 1970c. Final Report (Study 4) - Primary Skin - Rabbits.
Unpublished study, Project No. 417-337, prepared by TRW, Inc., Vienna,
VA for Rohm and Haas Co., Philadelpia, PA, dated October 6, 1970.
Regel, L.* 1972. Primary skin irritation study in albino rabbits. Unpublished
study no. 2060619, prepared by WARF Institute, Inc., Madison, WI for
O.M. Scott & Sons, Marysville, OH, dated June 28, 1972. MRID 0001265.
Rohm and Haas Bristol Research Laboratories. 1970. Fate and persistence of
Kerb (3,5-dichloro-N-(l,l-dimethyl-2-propynyl)-benzamide) in aqueous
systems. Unpublished report prepared by Rohm and Haas Co., Philadelphia,
PA. RAR Report No. 597.
Rohm and Haas Bristol Research Laboratories. 1973. A study of the hydrolysis
of the herbicide Kerb in water. Unpublished report prepared by Rohm and
Haas Co., Philadelphia, PA. Lab. 23. Technical Report No. 23-73-8.
Rohm and Haas Company. Undated. Research Report No. XXXXVI. Field dissipation
studies. Unpublished report prepared by Rohm and Haas Co., Philadelphia, PA.
Rohm and Haas Company. 1971. Soil adsorption studies with Kerb. Unpublished
report prepared by Rohm and Haas Co., Philadelphia, PA. Lab. 23 Tech.
Report No. 23-71-12.
Satterthwaite, S.T., and J.D. Fisher. Undated. Photodecomposition of Kerb in
water. Unpublished report prepared by Rohm and Haas Co., Philadelphia,
PA. Lab. 11 Memorandum Report No. 7.
Satterthwaite, S.T.* 1977. 14C-Kerb mouse feeding study. Unpublished study
no. 34H-77-3 prepared and submitted by Rohm and Haas Company, Philadelphia,
PA. February 19. MRID 0062604.
Smith, J.* 1974. Eighteen month study on the carcinogenic potential of Kerb
(RH-315: pronamide) in mice. Unpublished study received September 16
under 3F1317; prepared in cooperation with the Medical College of Virginia,
submitted by Rohm and Haas Company, Philadelphia, PA; CDL:094304-A.
MRID 008201601.
STORET. 1987.
TDB. 1985. Toxicology Data Book. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Sevice.
U.S. EPA. 1985a. U.S. Environmental Protection Agency, Office of Pesticide
Programs. Pronamide registration standard.
-------
Pronamide August, 1987
=16-
U.S. EPA. 198Sb. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.106. p. 252. July 1, 1985.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Vogin, E.E.* 1972. Effects of RH-315 on the development of fetal rats.
Unpublished study no. 0512 by Food and Drug Research Laboratories, Inc.,
Maspeth, NY for Rohm and Haas Company, Spring House, PA. October 22.
MRID 00125789.
Walker, A. 1976. Simulation of herbicide persistence in soil. III. Propy-
zamide in different soil types. Pestic. Sci. 7:59-64.
Walker, A. 1978., Simulation of the persistence of eight soil-applied herbi-
cides. Weed Res. 18:305-313.
Walker, A., and J»A. Thompson. 1977. The degradation of simazine, linuron
and propyzamide in different soils. Weed Res. 17(6):399-405.
Yin, R.Y., and C. Swithenbank.* Undated. Identification of metabolites of
N-(1,1-dimethylpropynyl)-3,5-dichlorobenzamide in rat and cow urine and
rat feces. Unpublished report prepared by Rohm and Haas Company, Spring
House, PA. MRID 00107954.
Yih, R.Y. Undated. Metabolism of N-(l,l-dimethylpropynyl)-3,5-dichlorobenzamide
(Rh-315) in soil, plants and mammals. Unpublished report prepared by
Rohm and Haas Co., Philadelphia, PA. Lab. 11 Research Report No. 11-210.
Zandvoort, R., D.C. van Dord, M. Leistra and J.G. Verlaat. 1979. The decline
of propyzamide in soil under field conditions in the Netherlands.
Weed Res. 19:157-164.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
PROMETON
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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Prometon August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1610-18-0
Structural Formula
OCR,
M<
H
^ — | •» |
H H
2,4-bis(isopropylamino)-6-methoxy-s-triazine
Synonyms
0 Gesafram 50; Ontracic 800; Primatol 25E; Pramitol; Methoxypropazine
(Meister, 1983).
Uses
A nonselective herbicide that controls most perennial broadleaf weeds
and grasses (Meister, 1983).
Properties (Meister, 1983; TDB, 1985; CHEMLAB, 1985)
Chemical Formula CIQH19^50
Molecular Weight 225.34
Physical State (258C) White crystals
Boiling Point --
Melting Point 91 to 92°C
Density 1.088 g/cm3
Vapor Pressure (20°C) 2.3 x 10~6 mm Hg
Specific Gravity
Water Solubility (20°C) 750 mg/L
Log Octanol/Water Partition -1.06 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Prometon has been found in 385 of 1,459 surface water samples analyzed
and in 40 of 757 ground water samples (STORET, 1987). Samples were
collected at 240 surface water locations and 650 ground water locations,
and proraeton was found in 12 states. The 85th percentile of all
nonzero samples was 0.6 ug/L in surface water and 50 ug/L in ground
water sources. The maximum concentration found was 8.5 ug/L in
surface water and 250 ug/L in ground water.
0 Prometon residues resulting from agricultural practice have been detected
in California ground waters at 0.21 - 80 ppb (Eiden, 1987).
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Prometon August, 1987
-3-
Environmental Fate
0 Prometon is stable to hydrolysis at pH 5, 7, and 9 at 25°C for 40
days (Ciba-Geigy Corporation, 1985a).
0 Prometon in aqueous solution was stable to natural sunlight for 2
weeks (Ciba-Geigy, 1985b).
0 Prometon has the potential to leach through soil, based on adsorption/
desorption tests and soil thin-layer chroma tog raphy (TLC). K^'s for
five soils were: sandy loam (2.61), silt loam (2.90), silty clay
loam (2.40), silt loam (1.20) and sand (0.398); organic matter content
ranged from 0.8 to 5% (Ciba-Geigy, 1985c).
0 Rf values for soil Thin Layer Chromatography (TLC) plates of five
soils put prometon in Class 4 (Very Mobile), Class 3 (Intermediate
Mobile), and Class 2 (Low Mobility). Prometon was very mobile in a
Mississippi silt loam and Plainfield sand, intermediately mobile in a
Hagerstown silty clay loam and Dubuque silt loam, and had low mobility
in a California sandy loam (Ciba-Geigy, 1985d).
0 In field dissipation studies, prometon was shown to have a half-life
>459 to 1,123 days at 3 different sites. Residues were found at all
depths sampled, down to 18 inches. There was no deeper sampling.
At 2 out of 3 sites, dealkylated prometon was found at the 0- to
18-inch depth (Ciba-Geigy, 1986)
III. PHARMACOKINETICS
Absor ption
0 Prometon is rapidly absorbed from the gastrointestinal tract. Based
on the radioactivity recovered in the urine and feces, prometon is
completely absorbed within 72 hours in the rat (BakJe et al., 1967).
Distribution
0 Seventy-two hours after intragastric intubation of 14C-prometon in
rats, no detectable levels of radioactivity were detected in any of
th<* tissues examined (BakJe et al., 1967).
Metabolism
0 Eleven metabolites of prometon have been identified in the urine of
rats treated with 14C-prometon. 2-Methoxy-4,6-diamino-S triazine and
ammeline represented 14% and 31%, respectively, of the radiolabel
excreted in the urine (Ciba-Geigy Corp., 1971).
0 Based on the metabolites formed, triazine ring cleavage apparently
does not occur during prometon metabolism (Ciba Geigy-Corp., 1971).
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Prometon August, 1987
-4-
Excretion
Excretion of prometon and/or its metabolites in rats was most rapid
during the first 24 hours after administration of 14c-prometon and
decreased to trace amounts at 72 hours. The radioactivity was quanti-
tatively excreted in the urine (91%) and feces (9%) within 72 hours
after dosing with 14C-prometon (Bakke et al., 1967).
IV. HEALTH EFFECTS
Humans
No information on the health effects of prometon in humans was found
in the available literature.
Animals
Short-term Exposure
0 The acute oral LDso value for prometon ranges from 1,750 to 2,980 mg/kg
in rats and is 2,160 mg/lcg in mice (Meister, 1983; NIOSH, 1985).
0 The acute inhalation LC50 value in rats is >3.6 mg/L for 4 hours
(Meister, 1983).
0 Long-Evans rats of both sexes (five/sex/dose) were fed a diet containing
0, 10, 30, 100, 300, 600, 1,000, 3,000, 6,000 or 10,000 ppm prometon
[technical, 97% active ingredient (a.i.)] for 4 weeks (Kileen et al.,
1976a). This corresponds to doses of 0, 0.5, 1.5, 5, 15, 30, 50,
150, 300, or 500 mg/kg/day, assuming 1 ppm in the diet corresponds to
0.05 mg/kg/day (Lehman, 1959). Rats fed 3,000 or more ppm prometon
showed a reduction in body weight during the treatment period; at
6,000 or 10,000 ppm (300 or 500 mg/kg/day) the reduction in body
weight was statistically significant (p <0.05 and 0.01, respectively).
At 1,000 ppm or less, mean body weight of both males and females were
comparable to controls. Gross pathology performed at the time of
sacrifice did not show any compound-related effects. The No-Observed-
Adverse-Effect-Level (NOAEL) and Lowest-Observed-Adverse-Effect-Level
(LOAEL) identified in this study are 3,000 and 6,000 ppm (150 and
300 mg/kg/day), respectively.
0 Beagle dogs (one/sex/dose) were administered 100, 300 or 3,000 ppm
prometon (technical) in the diet (2.5, 7.5 or 75 mg/kg/day, assuming
1 ppm in the diet is equivalent to 0.025 mg/kg/day; Lehman, 1959) for
2 weeks after which the 100- and 300-ppm doses were changed to 1,000
and 2,000 ppm (25 and 50 mg/kg/day) for the next 2 weeks (Killeen
et al., 1976b). Dogs that consumed 3,000 ppm showed a decrease in body
weight and food consumption. The body weight of the females receiving
1,000 or 2,000 ppm (25 or 50 mg/kg/day) was also decreased slightly;
food consumption was also slightly lower for the females receiving
2,000 ppm prometon (50 mg/kg/day). At 300 ppm and less, the body
weight and food consumption for both males and females were comparable
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Prometon August, 1987
-5-
to those of the controls. The NOAEL and LOAEL identified in this
study are 300 and 600 ppm (7.5 and 25 mg/kg/day), respectively.
Dermal/Ocular Effects
0 Prometon is a minimal dermal irritant (Meister, 1983). Barely
perceptible erythema was observed in rabbits exposed to 500 mg
prometon (97%) applied to one abraided and one intact site for 24 hours.
At 2,000 mg/kg, mild edema and slight desquamation was also observed
(Ciba-Giegy, 1977).
Long-term Exposure
0 Sprague-Dawley rats (30/sex/group) were fed a diet containing technical
prometon (98% active ingredient) at levels of 0, 10, 50, 100 or 300
ppm for 90 days (Johnson and Becci, 1982). Based on the assumption
that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), these doses correspond to approximately 0, 0.5, 2.5, 5 or 15
mg/kg/day. Although female rats exposed to 300 ppm showed an increase
in mean absolute weight of the kidneys, this was considered of no
toxicological significance, since the relative kidney to body weight
ratios were not changed (U.S. EPA, 1985). The NOAEL identified in
this study is, therefore, 300 ppm (15.0 mg/kg/day, the highest dose
tested).
Reproductive Effects
0 Prometon (technical, 98% a.i.) in corn oil was administered to Charles
River rats (25/dose) via gavage at levels of 0, 36, 120 or 360 mg/kg/day
from days 6 through 15 of gestation (Florek et al., 1981). Rats treated
with 120 or 360 mg/kg/day gained less body weight than the controls
during treatment; body weight gain in the 36-mg/kg/day group was
similar to that of the controls. Rats in all dosage groups exhibited
excessive salivation. Increased respiratory rate and lacrimation
were also seen in the 360-mg/kg/day group. No effects on implantation,
litter size, fetal viability, resorption, average fetal body weight
or gross external, soft tissue or skeletal variation in the fetuses
were observed at any dose level. This study identified a maternal
NOAEL of 36 mg/kg/day and a maternal LOAEL of 120 mg/kg/day.
0 New Zealand White rabbits (16/dose) administered prometon at dose levels
of 0, 0.5, 3.5 or 24.5 mg/kg/day (98* a.i.) from days 6 through 30 of
gestation showed reduced pregnancy rates at all dosage levels (Lightkep
et al., 1982). Pregnancy occurred in 16, 13, 13 and 11 rabbits given
0, 0.5, 3.5 and 24.5 mg/kg/day, respectively. Anorexia and excess
lacrimation were observed more frequently in the high-dose group.
Maternal body weight was significantly retarded during treatment in
the 24.5-mg/kg/day group. The maternal NOAEL identified in this study
is 3.5 mg/kg/day and the maternal LOAEL is 24.5 mg/kg/day.
Developmental Effects
0 In a teratogenicity study, prometon (technical) was administered to
albino rats at dose levels of 25 or 50 mg/kg/day on days 6 through
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Prometon August, 1987
-6-
15 of gestation (Haley, 1972). No significant differences between test
and control groups were seen in the maternal body weight, resorption
sites, viable fetuses, fetal external abnormalities, fetal skeletal
development or fetal internal development (details of the protocol
and individual data were not provided). Based on this information,
a NOAEL of 50 mg/kg/day (the highest dose tested) was identified.
0 Florek et al. (1981) reported no effects on fetal viability, resorp-
tion, average fetal body weight or gross external, soft tissue or
skeletal variations in the fetuses of Charles River rats (25/dose)
administered prometon via gavage at levels of 0, 36, 120 or 360
mg/kg/day (98% a.i.) in corn oil. A teratogenic NOAEL of 360 mg/kg/day
(the highest dose tested) and a maternal-toxicity NOAEL of 36 mg/kg/day
were identified.
• Lightkep et al. (1982) observed no gross, soft tissue or skeletal
variations in fetuses of New Zealand White rabbits (16/dose) administered
prometon at dose levels of 0, 0.5, 3.5 or 24.5 mg/kg/day (98% a.i.) on
days 6 through 30 of gestation. A teratogenic NOAEL of 24.5 mg/kg/day
(the highest dose tested) and a maternal-toxicity NOAEL of 3.5
were identified.
Mutagenicity
0 No information on the mutagenicity of prometon was found in the
available literature.
Carcinogenicity
0 No information on the carcinogenicity of prometon was found in the
available literature.
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 I 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/OOW guidelines.
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Proneton August, 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 information was found in the available literature that was suitable
for determination of the One-day HA value for prometon. It is therefore
recommended that the DWEL value adjusted for the 10-kg child (0.15 mg/L,
calculated below) be used at this time as a conservative estimate of the
One-day HA value.
Ten-day Health Advisory
Reduced body weight compared to controls has been observed in acute
and subchronic toxicity studies in the rat, dog and rabbit. Male and female
rats fed diets containing 3,000 ppm prometon for 4 weeks (300 mg/kg/day;
Killeen et al., 1976a) and pregnant rats administered 120 and 360 mg/kg/day
on days 6 through 15 of gestation (Florek et al., 1981) exhibited lower body
weights compared to controls. Dogs exhibited decreased body weight in a
4-week feeding study with dosing regimens as low as 2 weeks of initial dosing
at 100 ppm followed by 2 weeks at 1,000 ppm (25 mg/kg/day) (Killeen et al.,
1976b). Lightkep et al. (1982) treated rabbits via gavage with doses of 0.5,
3.5 and 24.5 mg/kg/day on days 6 through 15 of gestation and observed decreased
weights in animals exposed to the highest dose. From these studies, it can
be concluded that the rat is less sensitive to the effects of prometon on
weight gain than the dog. The rabbit appeared to exhibit a similar sensitivity
to the dog, but the method of oral dosing differed (gavage vs. feed). The
NOAEL identified from the rabbit study (3.5 mg/kg/day) is lower than that
identified in the dog study (7.5 mg/kg/day) and provides a more conservative
estimate of prometon toxicity.
Prometon toxicity is not well characterized, and fluctuations in weight
gain may not be an appropriately sensitive end point of toxicity. For this
reason, it is recommended that the DWEL, adjusted for a 10-kg child (0.15 mg/L,
calculated below) be used as a conservative estimate of the Ten-day HA value
for prometon.
Longer-term Health Advisory
The only species to be tested in subchronic studies of prometon toxicity
was the rat. In the study by Johnson and Becci (1932), rats were fed a diet
containing 0, 10, 50, 100 or 300 ppm prometon (0, 0.5, 2.5 or 15 mg/kg/day)
for 90 days. A NOAEL of 15 mg/kg/day (the highest dose tested) was identified.
A NOAEL of 100 mg/kg and a LOAEL of 300 mg/kg/day were identified from the
4-week rat feeding study by Killeen et al. (1976a). More conservative NOAEL
values can be identified from acute studies of other species (3.5 mg/kg/day,
rabbit, Lightkep et al., 1982; 7.5 mg/kg/day, dog, Killeen et al, 1976b).
The toxicity of prometon is not well characterized. It is therefore recommended
that the DWEL adjusted for a 10-kg child (0.15 mg/L, calculated below) be used
as a conservative estimate of the Longer-term HA value for prometon.
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Prometon August, 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 NQAEL (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, 1986a), then caution should be exercised
in assessing the risks associated with lifetime exposure to this chemical.
No suitable chronic or lifetime studies were available for the calculation
of a Lifetime HA for prometon. The available studies all reported on acute
health effects except that of Johnson and Becci (1982). In this study, rats
were fed diets containing 0, 10, 50, 100, or 300 ppm prometon for 90 days.
No toxic effects were observed at any of the dose levels tested, and a NOAEL
of 15 mg/kg/day was identified. This value may be a conservative estimate
of the NOAEL for rats; a NOAEL of 100 mg/kg was identified from the study
by Killeen et al. (1976a). In contrast, lower NOAELs were identified from
studies of acute exposure via gavage in other species (3.5 mg/kg/day, rabbit,
Lightkep et al., 1982; 7.5 mg/kg/day, dog, Killeen et al., 1976b). Taking
into consideration both the acute and subchronic test results, the study
of Johnson and Becci (1982) has been selected to serve as the basis for
determination of the RfD.
Step 1: Determination of the Reference Dose (RfD)
where:
RfD = mgay = 0.015 ngAg/day
(1,000) y y *
15 mg/kg/day = NOAEL, based on the absence of effects on the absolute
weight of the kidneys and on the mean kidney-to-brain
ratios in rats exposed to prometon in the diet for
90 days.
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.
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Prometon August, 1987
-9-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.015 mg/kg/day) (70 kg) „ 0.525 mg/L (525 ug/L)
(2 L/day)
where:
0.015 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 = 0.525 mg/L x 20% = 0.1 mg/L (100 ug/L)
where:
0.525 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No carcinogenicity studies were found in the literature searched.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of prometon.
0 Applying the criteria described in EPA's final guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), prometon may be classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No information was found'in the available literature on other existing
criteria, guidelines and standards pertaining to prometon.
VII. ANALYTICAL METHODS
0 Analysis of prometon is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
prometon, but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
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Prometon August, 1987
-10-
VIII. TREATMENT TECHNOLOGIES
0 Whittaker (1980) experimentally determined the adsorption isotherms
for prometon on granular activated carbon (GAC).
0 One study (Rees and Au, 1979) reported 95% removal efficiency when
prometon-contaminated water was passed over a 1 x 20 cm column packed
with resin.
0 Available data indicate that GAC adsorption and resin adsorption will
remove prometon from water (Whittaker, 1980; Rees and Au, 1979).
However, selection of individual or combinations of technologies to
attempt prometon removal from water must be based on a case-by-case
technical evaluation, and an assessment of the economics involved.
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Prometon August, 1987
-11-
IX. REFERENCES
Bakke, J.E., J.D. Robbins and V.J. Fell. 1967. Metabolism of 2-chloro-4,6-
bis(isopropylamino)-s-triazine(propazine) and 2-methoxy-4,6-bis(isopro-
pylamino)-s-triazine (prometon) in the rat. Balance study and urinary
metabolite separation. J. Agr. Food Chem. 15(4):628-631.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Cited in U.S. EPA.
1985. Pesticide survey chemical profile. Final report. Contract no.
68-01-6750. Office of Drinking Water, Washington, DC.
Ciba-Geigy Chemical Corporation.* 1971. Metabolisr. of s-triazine herbicides.
Unpublished study. EPA Accession No. 55672.
Ciba-Geigy Corporation.* 1977. Acute toxicity studies with prometon tech-
nical (97%). Industrial Bio-Text Laboratories, Inc. 1ST No. 8530-09308.
Unpublished study. EPA Accession No. 231815.
Ciba-Geigy Corporation. 1985a. Hydrolysis of prometon (Hazleton Study
6015-165). In: Environmental fate data required by special ground water
data call-in. May 30, 1985. Greensboro, NC.
Ciba-Geigy Corporation. 1985b. Photolysis of proraeton in aqueous solution
under natural sunlight and artifical sunlight conditions (1972), Ciba-
Geigy Report No. 72127. In: Environmental fate data required by special
ground water data call-in, May 30, 1985. Greensboro, NC.
Ciba-Geigy Corporation. 1985c. The adsorption/desorption of radiolabeled
prometon on representative agricultural soils (Hazleton Study 6015-164).
In: Environmental fate data required by special ground water data call-in,
May 30, 1985. Greensboro, NC.
Ciba-Geigy Corporation. 1985d. Mobility determination of prometon in soils
by TLC (Hazleton Study No. 6015-167). In: Environmental fate data
required by special ground water data call-in, May 30, 1985. Greensboro,
NC.
Ciba-Geigy Corporation. 1986*. Field disposition studies in California,
Nebraska and New York. Preprared by Daniel Sunuier. August 21, 1986.
Eiden, C. 1987. Assessing the leeching potential of pesticides: national
perspectives. Draft report prepared by the U.S. Environmental Protection
Agency, Office of Pesticide Programs, Washington, DC.
Florek, C., G. Christian et al.* 1981. Teratogenicity study of prometon
technical in pregnant rats. Argus Project 203-003. Unpublished study.
EPA Accession No. 129983.
Haley, s.* 1972. Report to Geigy Agricultural Chemicals, Division of Ciba-
Geigy Corporation. Teratogenic study with prometon technical in albino
rats. IBT No. B904. Unpublished study.
-------
Prometon August, 1987
-12-
Killeen, J.C., Jr., W.E. Rinehart, S. Munulkin et al.* 1976a. A four-week
range-finding study with technical prometon in rats. Project no. 76-1445.
Unpublished study. EPA Accession No. 54308.
Killeen, J.C., Jr., W.E. Rinehart, S. Nunulkin et al.* 1976b. A four-week
range-finding study with technical prometon in beagle dogs. Project no.
76-1446. Unpublished study. EPA Accession No. 54309.
Johnson, W., and P. Becci.* 1982. 90-Day subchronic feeding study with
prometon technical in Sprague-Dawley rats. FDRL Study No. 6805.
Unpublished study. EPA Accession No. 129985.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
Lightkep, G., M. Christian, G. Christian et al.* 1982. Teratogenic poten-
tial of prometon technical in New Zealand White rabbits. Segment II -
evaluation. Project No. 203-002. Final report. Unpublished study.
EPA Accession No. 129984.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
NIOSH. 1985. National Institute for Occupational Safety and Health. Registry
of toxic effects of chemical substances. National Library of Medicine
Online File.
Rees, G.A.V., and L. Au. 1979. Use of XAD-2 macroreticular resin for the
recovery of ambient trace levels of pesticides and industrial organic
pollutants from water. Bull. Environ. Contarn. Toxicol. 22(4/5):561-566.
STORET. 1987.
TDB. 1985. Toxicology Data Book. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA.* 1985. U.S. Environmental Protection Agency. Prometon, EPA I.D.
No. 100-544, Caswell No. 96. EPA Accession No. 259108.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Peg. 51(185):33992-34003. September 24,
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. Ph.D. Thesis, Purdue
University.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
PROPACHLOR
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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Propachlor August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-16-7
Structural Formula
CH(CH3)2
2-chloro-N-isopropylacetinilide
Synonyms
0 Bexton; Prolex; Ramrod (Meister, 1983).
Uses
0 Selective postemergence herbicide used for control of many grasses
and certain broadleaf weeds (Meister, 1983).
Properties (Rao and Davidson, 1982; HSDB, 1986)
Chemical Formula CnH14ClNO
Molecular Weight 211.69
Physical State (room temp.) White crystalline solid
Boiling Point 110°C at 0.03 mm HG
Melting Point 67 to 76°C
Density (25°C) 1.13 g/mL
Vapor Pressure 2.3 x 10~4 mm Hg
Specific Gravity
Water Solubility (20°C) 700 mg/L
Log Octanol/Water Partition 1.61
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Propachlor has been found in 132 of 1,144 surface water samples
analyzed and in 2 of 76 ground water samples (STORET, 1987). Samples
were collected at 314 surface water locations and 94 ground water
locations, and propachlor was found in eight states. The 85th
percentile of all nonzero samples was 2 ug/L in surface water and
0.12 ug/L in ground water sources. The maximum concentration found
was 10 ug/L in surface water and 0.12 ug/L in ground water.
Environmental Fate
0 Propachlor is degraded in aerobic soils in the laboratory and in the
field with half-lives of 2 to approximately 14 days, when the soils
are treated with propachlor at recommended application rates. However.
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Propachlor August, 1987
-3-
degradation was relatively slower in soil treated at 500 ppm, and 90%
of the applied material remained after 21 days (Registrant CBI data).
0 The major propachlor degradates produced under aerobic soil conditions
are [(1-methylethyl)phenylamino]oxoacetic acid and [(2-methylethyl)-
phenylamino]-2-oxoethane sulfonic acid. These degradates are recalci-
trant to further degradation in soil under anaerobic conditions. The
half-life of propachlor in anaerobic soil is <4 days (Registrant CBI
data).
0 Propachlor degrades very slowly (84.5% remaining after 30 days) in
lake water (Registrant CBI data).
0 Propachlor is moderately mobile to very mobile in soils ranging in
texture from sand to clay. Mobility appears to be correlated with
clay content and to a lesser degree with organic matter content and
CEC. Aged 14c-propachlor residues were mobile in a silt loam soil
(Registrant CBI data).
0 The rapid degradation of low levels of propachlor in soils is expected
to result in a low potential for groundwater contamination by propachlor
degradates. 14C-Propachlor residues are taken up by rotated corn
planted under confined conditions; <3% of the radioactivity remained
in soil at the time of planting (Registrant CBI data).
III. PHARMACOKINETICS
Absorption
0 No direct data on rate of gastrointestinal absorption of propachlor
were found in the available literature. Based on recovery studies,
propachlor appears to be rapidly absorbed by the oral route of admin-
istration. An estimated 68% of a single dose of 10 mg of ring-labeled
14-c propachlor administered to 12 rats was recovered in urine 56
hours after compound administration (Malik, 1986). These results are
supported by other studies in which 54 to 64% (Lamoureux and Davison,
1975) and 68.8% (Bakke et al., 1980) of the administered dose was
recovered in urine 24 hours and 48 hours after dose administration,
respectively.
Distribution
0 Fifty-six hours following oral administration of 10 mg of ring-
labeled 14c-propachlor (purity not specified) to rats, no detectable
levels of radioactivity were identified in any tissue samples (Malik,
1986).
Metabolism
0 Metabolism of propachlor occurs by initial glutathione conjugation
followed by conversion via the mercapturic acid pathway; oxidative
metabolism also occurs (Lamoureux and Davison, 1975; Malik, 1986).
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Propachlor August, 1987
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Eleven urinary metabolites have been identified as the result of
propachlor metabolism in rats. The primary metabolic end products
of propachlor are mercapturic acid and glucuronic acid conjugates
(approximately 20 to 25%), methyl sulfones (30 to 35%), and phenols
and alcohols (Lamoureux and Davison, 1975; Malik, 1986).
Excretion
0 Propachlor (purity not specified) was excreted in the form of metabo-
lites in the urine (68%) and feces (19%) of rats within 56 hours after
dosing with ring-labeled 14c-propachlor. Methyl sulfonyl metabolites
accounted for 30 to 35% of the administered dose (Malik, 1986).
0 In studies with germ-free rats, 98.6% of the administered dose (not
specified) for propachlor (purity not specified) was identified in
the urine (68.8%) and feces (32.1%) within 48 hours. The major
metabolite was mercapturic acid conjugate, which accounted for 66.8%
of the administered dose (Bakke et al., 1980).
IV. HEALTH EFFECTS
Humans
Schubert (1979) reported a case study in which occupational exposure
to propachlor for 8 days resulted in erythemato-papulous (red pimply)
contact eczema on the hands and forearms.
Animals
Short-term Exposure
0 The acute oral LDg0 values for technical-grade (approximately 96.5%)
and wettable powder (WP) (65%) propachlor range from 1,200 to 4,000
mg/kg in rats. Technical-grade and wettable powder propachlor both
produced a low LD50 value of 1,200 mg/kg (Keeler et al., 1976;
Heenehan et al., 1979; Auletta and Rinehart, 1979; Monsanto, (undated).
0 Beagle dogs (two/sex/dose) were administered propachlor (65% WP} in
the diet for 90 days at dose levels of 0, 1.3, 13.3 or 133.3 mg/kg/day
(Wazeter et al.. 1964). Body weight, survival rates, food consump-
tion, behavior, general appearance, hematology, biochemical indices,
urinalysis, histopathology and gross pathology were comparable in
treated and control animals. The No-Observed-Adverse-Effect-Level
(NOAEL) identified for this study is 133.3 mg/kg/day (the highest
dose tested).
0 Naylor and Ruecker (1985) fed propachlor [96.1% active ingredient
(a.i.)] to beagle dogs (six/sex/dose) in the diet for 90 days at dose
levels of 0, 100, 500 or 1,500 ppm. Based on the assumption that
1 ppm in food is equivalent to 0.025 mg/kg/day (Lehman, 1959), these
doses are equivalent to 0, 2.5, 12.5 or 37.5 mg/kg/day. Clinical
signs, ophthalmoscopic, clinicopathologic, gross pathology and
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Propachlor August, 1987
-5-
histopathologic effects were comparable for treated and control
groups. The reduction in food consumption and concomitant reductions
in body weight gain at all test levels were considered by the author
to be due to poor diet palatability. Based on these responses, a
NOAEL of 1,500 ppm (37.5 mg/kg/day) was identified.
Dermal/Ocular Effects
0 The acute dermal LD50 value of technical propachlor and WP (65% propa-
chlor) in the rabbit ranges from 380 mg/kg to 20 g/kg (Keeler et al.,
1976; Monsanto, undated; Braun and Rinehart, 1978). Wettable powder
produced the lowest LD50 in rabbits (380 mg/kg); the lowest LD5Q produced
by technical propachlor was between 1,000 and 1,260 mg/kg in rabbits.
0 Propachlor (94.5% a.i.) (1 g/mL) applied to abraded and intact skin
of New Zealand White rabbits (three/sex) for 24 hours produced erythema
and slight edema at treated sites 72 hours post-treatment (Heenehan
et al., 1979).
0 Heenehan et al. (1979) instilled single applications (0.1 cc) of
propachlor into one eye of tested New Zealand rabbits for 30 seconds.
Corneal opacity with stippling and ulceration, slight iris irritation,
con;junctival redness, chemosis, discharge and necrosis were reported
at 14 days. Similar responses were reported by Keeler et al. (1976)
for a corresponding observation period and by Auletta (1984) during
3 to 21 days post-treatment.
Long-term Exposure
0 Albino rats (25/sex/dose) administered 0, 1.3, 13.3 or 133.3 mg/kg/day
propachlor (65% WP = 65% a.i.) in the diet for 90 days showed decreased
weight gain (10 to 12% less than control levels) in and increased
liver weights in both sexes (10% greater than control levels) at
133.3 mg/kg/day (the highest dose tested) (Wazeter et al., 1964).
The body and liver weights of rats of both sexes that received the
low dose and mid dose were comparable to control levels. Survival,
biochemical indices, hematology, urinalysis, gross pathology and
histopathology did not differ significantly between treated and
control groups. The NOAEL identified in this study is 13.3 mg/kg/day.
The Lowest-Observed-Adverse-Effect-Level (LOAEL) is 133.3 mg/kg/day
(the highest dose tested).
0 Reyna et al. (1984a) administered propachlor (96.1% a.i.) to rats
(30/sex/dose) in the diet for 90 days at mean dose levels of 0, 240,
1,100 or 6,200 ppm. Assuming that 1 ppm is equivalent to 0.05 mg/kg/day,
these concentrations correspond to 0, 12, 55 or 310 mg/kg/day (Lehman,
1959). Body weights and food consumption were significantly decreased
(no p value specified) at 55 mg/kg/day and 310 mg/kg/day in both
sexes. Final body weights for females were 7 and 36% less than
controls at the mid- and high-dose levels, respectively. In males,
final body weights were 8 and 59% less than control levels for mid-
and high-dose levels, respectively. However, histopathological
examination showed no changes. Mid- and high-dose levels produced
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Propachlor August, 1987
-6-
increased platelet counts, decreased white blood cell counts and mild
liver cell dysfunction. Mild hypochromic, microcytic anemia was
reported at the high dose. A NOAEL of 12 mg/kg/day can be identified
for this study.
0 Albino mice (30/sex/dose) were fed propachlor (96.1% a.i.) in the
diet for 90 days at mean dose levels of 0, 385, 1,121 or 3,861 ppm
(Reyna et al., 1984b). Based on the assumption that 1 ppm in food
is equivalent to 0.15 mg/kg/day (Lehman, 1959), these doses correspond
to 0, 58, 168 or 579 mg/kg/day. Reduced body weight gain, decreased
white blood cell count, liver and kidney weight changes and increased
incidences of centrolobular hepatocellular enlargement were reported
at the mid (168 mg/kg/day) and high (579 mg/kg/day) doses when
compared to controls. Based on these responses, a NOAEL of 385 ppm
(58 mg/kg/day) can be identified.
Reproductive Effects
0 No information on the reproductive effects of propachlor was found in
the available literature.
Developmental Effects
0 Miller (1983) reported no signs of maternal toxicity in New Zealand
female rabbits (16/dose) that were administered propachlor (96.5%)
orally by gavage at doses of 0, 5, 15 or 50 mg/kg/day on days 7 to 19
of gestation. Statistically significant increases in mean implantation
loss with corresponding decreases in the mean number of viable fetuses
were reported at 15 and 50 mg/kg/day when compared to controls. Two
low-dose and one mid-dose rabbit aborted on gestation days 22 to 25.
These effects, however, do not appear to be treatment-related since
no abortions occurred in the high-dose animals. No treatment-related
effects were present in the 5-mg/kg/day group (the lowest dose tested).
The authors reported that the maternal and embryonic NOAELs were 50
and 5 mg/kg/day, respectively.
0 Schardein et al. (1982) administered technical propachlor orally by
gavage to rats (25/dose) at dose levels of 0, 20, 60 or 200 mg/kg/day
during days 6 to 19 of gestation. There were no adverse fetotoxic or
maternal effects reported at any dose level. Based on this information,
the NOAEL identified in this study was 200 ing/kg/day (the highest
dose tested).
Mutagenicity
0 Technical propachlor was not genotoxic in assays of Salmonella
typhimunum with or without plant and animal activation; however,
genotoxic activity was reported in yeast assays (Saccharomyces
cerevisiae) at 1.3 x 10"3 M and 3 mg per plate after plant activation
(Plewa et al., 1984).
0 In a cytogenic study, propachlor administered for 24 hours by intra-
peritoneal injection at dose levels of 0.05, 0.2 or 1.0 mg/kg to F344
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Propachlor August, 1987
-7-
rats did not induce chromosomal aberrations in bone marrow cells
(Ernst and Blazak, 1985).
0 Gene mutation was not detected in assays employing Chinese Hamster
Ovary (CHO) cells. Primary rat hepatocytes exposed to 1,000 and
5,000 ug/mL technical-grade propachlor showed no effect on unscheduled
DNA synthesis when compared to controls (Flowers, 1985; Steinmetz and
Mirsalis, 1984).
Carcinogenicity
0 No information was found in the available literature to evaluate the
carcinogenic potential of propachlor. However, several chemicals
analogous to this compound, i.e., alachlor and acetochlor, were found
to be oncogenic in two animal species'.
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 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 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 propachlor. It is therefore
recommended that the Ten-day HA value for the 10-kg child (0.5 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The developmental toxicity study in rabbits by Miller (1983) has been
selected as the basis for determination of the Ten-day HA value for propachlor.
Pregnant rabbits administered propachlor (96.5%) by gavage at a dose level of
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Propachlor August, 1987
-8-
5 mg/kg/day showed no clinical signs of toxicity in the adult animals and no
reproductive or developmental effects in the fetuses. The study identified a
NOAEL of 5 mg/kg/day. These results are supported by a reproduction study
reported by Schardein et al. (1982) in which rats were administered doses
ranging from 20 to 200 mg/kg/day during gestation, with no adverse fetotoxic
or maternal effects reported at any dose level. The NOAEL identified in that
study was 200 mg/kg/day (the highest dose tested). However, since the rabbit
appears to be the more sensitive species, the NOAEL identified in the rabbit
study will be used to derive the Ten-day HA.
Using a NOAEL of 5 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (5 mg/kg/day) (10 kg) = 0.5 mg/L (500 ug/L)
(100) (1 L/day)
where:
5 mg/kg/day = NOAEL, based on the absence of clinical signs of toxicity
and the lack of reproductive or teratogenic effects in
rabbits exposed to propachlor by gavage for 12 days
during gestation.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODH
guidelines for use with a NOAEL from an animal study.
1 L/day =» assumed daily water consumption of a child.
Longer-term Health Advisory
Because no suitable long-term studies were available to calculate a
Longer-term HA, it was decided that it would be more appropriate to use the
Reference Dose of 0.013 mg/kg/day and adjusting this number to protect a
10-kg child and a 70-kg adult. The resulting Longer-term HA thus becomes
0.13 mg/L and 0.46 mg/L for a 10-kg child and a 70-kg adult, respectively.
Lifetime Health Advisory
The Lifetime HA reoresents 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
(OWED 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
-------
Propachlor August, 1987
-9-
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 90-day study by Wazeter et al. (1964) has been selected to serve as
the basis for determination of the Lifetime HA value for propachlor. Based
on body and liver weight effects, a NOAEL of 13.3 mg/kg/day was identified.
These results were further verified by the results of a similar study with
rats conducted by Reyna et al. (1984a) in which a NOAEL of 12 mg/kg/day was
identified.
Step 1: Determination of the Reference Dose (RfO)
RfD = (13.3 mg/kg/day) = 0.013 mgAg/day
(1,000)
where:
13.3 mg/kg/day = NOAEL based on the absence of effects on body weight
and liver weight in rats exposed to propachlor for
90 days.
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 Level (DWEL)
DWEL = (0.013 mg/kg/day) (70 kg) = 0<46 /L (460 /L)
(2 L/day)
where:
0.013 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 Healtn Advisory
Lifetime HA = (0.46 mg/L) (20%) = 0.092 mg/L (92 ug/L)
where:
0.46 mg/L = DWEL.
20% = assumed relative source contribution from water.
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Propachlor August, 1987
-10-
Evaluation of Carcinogenic Potential
0 No studies on the carcinogenic potential of propachlor were found in
the available literature. However, other structurally similar compounds
such as alachlor and acetochlor have been found to be potent carcinogens.
0 Applying the criteria described in EPA's final guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), propachlor may be classified
in Group D: not classified. This category is for substances with
inadequate human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Residue tolerances ranging from 0.02 to 10.0 ppm have been established
for propachlor in or on agricultural commodities (U.S. EPA, 1985).
0 NAS (1977) has recommended an ADI of 0.1 mg/kg/day and a Suggested-
No-Adverse-Effect Level (SNARL) of 0.7 mg/L, based on a NOAEL of
100 mg/kg/day in a rat study (duration of study not available).
VII. ANALYTICAL METHODS
(to be provided by STB)
VIII. TREATMENT TECHNOLOGIES
0 No data were found for the removal of propachlor from drinking water
by conventional treatment or by activated carbon treatment.
0 No data were found for the removal of propachlor from drinking water
by aeration. However, the Henry's Coefficient can be estimated from
available data on solubility (700 mg/L at 20°C) and vapor pressure
(2.3 x 10~4 mm Hg at 25°C). Propachlor probably would not be amenable
to aeration or air stripping because its Henry's Coefficient is
approximately 0.0051 atm. Baker and Johnson (1984) reported the
results of water and pesticide volatilization from a waste disposal
pit. Over a 2-year period, approximately 66.4 mg of propachlor
evaporated for every liter of water which evaporated and only 8.3%
of the propachlor was removed. These results support the assumption
that aeration would not effectively remove propachlor from drinking
water.
0 Propachlor is similar in structure to alachlor and has similar physical
properties. The effectiveness of various processes for removing
propachlor would probably be similar to that of alachlor.
0 Alachlor is amenable to the following processes:
- GAC (Miltner and Fronk, 1985; DeFilippi et al., 1980).
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Propachlor August, 1987
-11-
- PAC (Miltner and Fronk, 1985; Baker, 1983).
- Ozonation (Miltner and Fronk, 1985).
- Reverse osmosis (Miltner and Fronk, 1985).
0 Chlorine and chlorine dioxide oxidation were partially effective in
removing alachlor from drinking water (Miltner and Fronk, 1985).
0 The following processes were not effective in removing alachlor from
drinking water:
- Diffused aeration (Miltner and Fronk, 1985).
- Potassium permanganate oxidation (Miltner and Fronk, 1985).
- Hydrogen peroxide oxidation (Miltner and Fronk, 1985).
- Conventional treatment (Miltner and Fronk, 1985; Baker, 1983).
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Propachlor August, 1987
-12-
IX. REFERENCES
Auletta, C., and W. Rinehart.* 1979. Acute oral toxicity in rats: Project No.
4891-77, BDN-77-431. Unpublished study. MRID 104342.
Auletta, C.* 1984. Eye irritation study in rabbits. Propachlor. Project No.
5050-84. Unpublished study. Biodynamics, Inc. MRID 151787.
Baker, D. 1983. Herbicide contamination in municipal water supplies in
northwestern Ohio. Final draft report. Prepared for Great Lakes National
Program Office, U.S. Environmental Protection Agency, Tiffin, OH.
Baker, J.L., and L.A. Johnson. 1984. Water and pesticide volatilization
from a waste disposal pit. Transactions of the American Society of
Agricultural Engineers. 27:809-816. May/June.
Bakke, J., J. Gustafsson and B. Gustafsson. 1980. Metabolism of propachlor
by the germ-free rat. Science. 210:433-435. October.
Braun, W., and W. Rinehart.* 1978. Acute dermal toxicity in rabbits [due to
propachlor (technical)]. Biodynamics, Inc. Project No. 4888-77, BDN-77-
430. Unpublished study. MRID 104351.
DeFilippi, R.P., V.J. Kyukonis, R.J. Robey and M. Modell. 1980. Super-
critical fluid regeneration of activated carbon for adsorption of
pesticides. Research Triangle Park, U.S. Environmental Protection
Agency. EPA-600/2-80-054.
Ernst, T., and W. Blazak.* 1985. An assessment of the mutagenic potential of
propachlor utilizing the acute ir± vivo rat bone marrow cytogenetics assay
(SR 84-180): Final Report: SRI Project LSC-7405. SRI International.
Unpublished study. MRID 00153940.
Flowers, L.* 1985. CHO/HGPRT gene mutation assay with propachlor: Final
Report: EWL 840083. Unpublished study. MRID 00153939.
Heenehan, P., W. Rinehart and W. Braun.* 1979. Acute oral toxicity study in
rats. Project No. 4887-77. BDN-77-430. Biodynamics, Inc. MRID 104350.
HSDB. 1986. Hazardous Substances Database. National Library of Medicine,
Bethesda, MD.
Keeler, P.A., D.J. Wroblewski and R.J. Kociba.* 1976. Acute toxicological
properties and industrial handling. Hazards of technical grade propachlor.
Unpublished study. MRID 54786.
Lamoureaux, G., and K. Davison.* 1975. Mercapturic acid formation in the
metabolism of propachlor, CDAA, Fluorodifen in the rats. Pesticide
Biochem. Physiol. 5:497-506.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs anc?
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
-------
Propachlor August, 1987
-13-
Malik, J.* 1986. Metabolism of propachlor in rats: Report No. MSL-5455;
Job/Project No. 7815 (Summary). Unpublished study. MRID 157495.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Miller, L.* 1983. Teratology study in rabbits (IR-82-224):401-190. Inter-
national Research and Development Corporation. Unpublished study.
MRID 00150936.
Miltner, R.J., and C.A. Fronk. 1985. Treatment of synthetic organic contami-
nants for Phase II regulations. Internal report. U.S. Environmental
Protection Agency, Drinking Hater Research Division. December.
Monsanto Company.* Undated. Toxicology. Summary of studies 241666-C through
241666-E. Unpublished study. MRID 25527.
MAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
Naylor, M., and F. Ruecker.* 1985. Subchronic study of propachlor admini-
stered in feed to dogs: DMEH Project No. ML-84-092. Unpublished study.
MRID 00157852.
Plewa, M.J., et al. 1984. An evaluation of the genotoxic properties of herbi-
cides following plant and animal activation. Mutat. Res. 136(3):233-246.
Rao, P.S.C., and J.M. Davidson. 1982. Retention and transformation of
selected pesticides and phosphorus in soil-water systems: A critical
review. U.S. EPA, Athens, GA. EPA-600/53-82-060.
Reyna, M., W. Ribelin, D. Thake et al.* 1984a. Three month feeding study of
propachlor to albino rats: Project No. ML-83-083. Unpublished study.
MRID 00152151.
Reyna, M., W. Ribelin, D. Thake et al.* 1984b. Three month feeding study of
propachlor to albino rats: Project No. ML-81-72. Unpublished study.
MRID 00152365.
Schardein, J., D. Wahlberg, S. Allen et al.* 1982. Teratology study in rats
(IR-81-264):401-171. Unpublished study. MRID 00115136.
Schubert, H. 1979. Allergic contact dermatitis due to propachlor. Dermatol.
Monatsschr. 165(71:495-498. (Ger.) (PESTAB 80:115)
Steinmetz, K., and J. Mirsalis.* 1984. Evaluation of the potential of
propachlor to induce unscheduled DNA synthesis in primary rat hepatocyte
culture. Final report: Study No. LSC-7538. Unpublished study.
MRID 00144512.
STORET. 1987.
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Propachlor August, 1987
-14-
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.211. July 1.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 (185): 33992-34003. September 24.
Wazeter, F.X., R.H. Buller and R.G. Geil.* 1964. Ninety-day feeding study in
the rat. Ninety-day feeding study in the dog: 138-001 and 138-002.
Unpublished study. MRID 00093270.
•Confidential Business Information submitted to the Office of Pesticide
Programs
-------
August, 1987
PROPAZINE
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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Propazine August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 139-40-2
Structural Formula i
H
'a~ i r» |
H H
6-Chloro-N,NI-bis(l-methylethyl)-l-3,5-triazine-2,4-diamine
Synonyms
0 Geigy 30,028; Gesomil; Milogard; Plantulin; Primatol P; Propasin;
Prozinex (Meister, 1983).
Uses
0 Selective preemergence and preplant herbicide used for the control of
most annual broadleaf weeds and annual grasses in milo and sweet
sorghum (Meister, 1983).
Properties (Meister, 1983; IPC, 1984; CHEMLAB, 1985; TDB, 1985)
Chemical Formula
Molecular Weight 230.09
Physical State (25*C) Colorless crystals
Boiling Point — ~
Melting Point 212 to 214°C
Density ~
Vapor Pressure (20°C) 2.9 x 1 0~8 mm Hg
Water Solubility (29°C) 8.6 mg/L
Octanol/Water Partition -1.21
Coefficient
Taste Threshold ~
Odor Threshold —
Conversion Factor
Occurrence
Propazine has been found in 132 of 1,231 surface water samples
analyzed and in 20 of 1,056 ground water samples (STORET, 1987).
Samples were collected at 253 surface water locations and 639 ground
water locations, and propazine was found in 8 states. The 85th
percentile of all nonzero samples was 2.3 ug/L in surface water and
0.2 ug/L in ground water sources. The maximum concentration found
was 20 ug/L in surface water and 300 ug/L in ground water
Propazine was detected in ground water in California at trace levels
(<0.1 ppb) (U.S.G.S., 1985).
-------
Propazine Au*ust' 1987
-3-
Environmental Fate
The following data were submitted by Ciba-Geigy and reviewed by the Agency
(U.S. EPA, 1987):
0 Hydrolysis studies show propazine to be resistant to hydrolysis.
After 28 days, at pH 5, 60% remains; at pH 7, 92% remains; and at pH 9,
100% remains. Hydroxypropazine (2-hydroxy-4,6-bis-isopropylamino)-s-
triazine) is the hydrolysis product.
0 Propazine at 2.5 ppm in aqueous solution was exposed to natural
sunlight for 17 days. In that time, 5% degraded to hydroxy-propazine.
0 Under aerobic conditions, 10 ppm propazine was applied to a loamy
sand (German) soil with 2.2% organic carbon. The soil was incubated
at 25°C in the dark and kept at 70% of field capacity. Propazine
degraded with a half-life of 15 weeks. Hydroxypropazine was the
major degradate from aerobic soil metabolism; its concentration
increased from 14% at 12 weeks to a maximum of 31% after 52 weeks of
incubation. Trapped volatiles identified as CO2 accounted for 1% of
the applied propazine after 52 weeks. Bound residues increased up to
35% after 12 weeks of incubation.
0 Under anaerobic conditions, further degradation of propazine was
slight.
0 Freundlich soil-water partition coefficient (Kd) values for propazine
and hydroxypropazine were determined for four soils: a sand loam
(0.7% OM), a sand loam (1.4% OM), a loam soil (2.9% OM) and a clay
loam {8.3% OM). The Kd values were: 0.34, 1.13, 2.69 and 3.19,
respectively, for propazine. On the same four soils the Kd values
for hydroxypropazine were: 1.13, 2.94, 31.8 and 10.6, respectively.
All Kd values have units of ml/gm.
0 Leaching studies for propazine performed on four soils under worst-case
conditions (30-cm columns leached with 20 inches of water) for
propazine indicate propazine's mobility in soil-water systems. In a
loamy sand (0.7% OM), a sandy loam (1.4% OM), a loam (1.7% OM), and a
silt loam (2.4% OM), 82.5%, 18%, 69.5%, and 23.6% leached, respectively.
0 In column studies using aged propazine, degradation products leached
from a loamy sand soil with 2.2% OM. About 25% of the aged propazine
added to the columns leached. In a loam soil with 3.6% OM, <0.05% of
the aged propazine added to the columns leached.
0 In field dissipation studies, propazine was found at 18 inches the
deepest depth in the soil sampled. Hydroxypropazine was found at
all depths and sites up to 3 years after application. Field half-
lives for propazine were 5 to 33 weeks in the 0- to 6-inch depth,
and 17 to 51 weeks at the 6 to 12 inch depth.
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Propazine August, 1987
-4-
III. PHARMACOKINETICS
Absorption
0 Bakke et al. (1967) administered single oral doses of ring-labeled
14c-propazine to Sprague-Dawley rats. After 72 hours, about 23% of
the label was recovered in the feces and about 66% was excreted in
the urine. This indicates that gastrointestinal absorption was at
least 77% complete.
Distribution
0 Bakke et al. (1967) administered ring-labeled 14C-propazine (41 to
56 mg/kg) to rats by gastric intubation. Radioactivity in a variety
of tissues was observed to decrease from an average of 46.7 ppm 2 days
posttreatment to 22.3 ppm after 8 days. Radioactivity was detected
in the lung (30 ppm), spleen (25 ppm) heart (27 ppm), kidney (17 ppm)
and brain (13 ppm) for up to 8 days. After 12 days, the only detectable
quantities were in hide and hair (3»35% of administered dose), viscera
(0.1%) and carcass (2.22%).
Metabolism
• Eighteen metabolites of propazine have been identified in the urine of
rats given single oral doses of 1*C-propazine (Bakke et al., 1967).
No other details were provided. Based on metabolites found in urine,
Bakke et al. (1967) reported that dealkylation is one reaction in the
metabolism of propazine. No other details were provided.
Excretion
Bakke et al. (1967) administered single oral doses of 1 ^C-ring-labeled
propazine to rats. Most of the radioactivity was excreted in the
urine (65.8%) and feces (23%) within 72 hours. Excretion of propazine
and/or metabolites was most rapid during the first 24 hours after
administration, decreasing to smaller amounts at 72 hours.
IV. HEALTH EFFECTS
Humans
Contact dermatitis was reported in workers involved in propazine
manufacturing (Hayes, 1982). No other information on the health
effects of propazine in humans was found in the available literature.
Animals
Short-term Exposure
The reported acute oral 1*05^ values for propazine (purity not soecifieci'
were >5,000 mg/kg in rats (Stenger and Kindler, 1963a), >5,000 mq/kq
in mice (Stenqer and Kindler. 1963b) and 1.200 ma/ka in ouinea oias
(HIOSH. 1985).
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Propazine August, 1987
-5-
0 Stenger and Kindler (1963a) reported that dietary administration of
propazine (purity not specified) to rats (five/sex/dose) at doses of
1,250 or 2,500 mg/kg for 4 weeks resulted in a decrease in body
weight, but there were no pathological alterations in organs or
tissues. No other details were provided.
Dermal/Ocular Effects
0 The acute dermal LD^g value in rabbits for propazine (90% water dis-
persible granules) was reported as >2,000 mg/kg (Cannelongo et al.,
1979).
0 Stenger and Huber (1961) reported that rats were unaffected when
a 5% gum arable suspension of propazine (0.4 mL/animal) was applied
to shaved and intact skin of five rats then washed away 3 hours
later.
0 Palazzolo (1964) reported that propazine (1 or 2 g/kg/day) applied to
intact or abraded skin of albino rabbits (five/sex/dose) for 7 hours
produced mild erythema, drying, desquamation and thickening of the
skin. Body weights, mortality, behavior, hematology, clinical chemistry
and pathology of the treated and untreated groups were similar.
jjong-term Exposure
0 In 90-day feeding studies by Wazeter et al. (1967a), beagle dogs
(12/sex/dose) were fed propazine (80 WP) in the diet at 0, 50, 200
or 1,000 ppm active ingredient. Based on the assumption that 1 ppm
in the diet of dogs is equivalent to 0.025 mg/kg/day (Lehman, 1959)
these doses correspond to 0, 1.25, 5.0 or 25 mg/kg/day. No compound-
related changes were observed in general appearance, behavior,
hematology, urinalysis, clinical chemistry, gross pathology or histo-
pathology at any dose tested. In the 1,000 ppm dose group, four
dogs lost 0.3 to 1.1 kg in body weight, which the author suggested
may have been compound-related (no p value reported). Based on these
results, a No-Observed-Adverse-Effect-Level (NOAEL) of 200 ppm
(5 mg/kg/day) and a LOAEL of 1,000 ppm (25 mg/kg/day) were identified.
0 Wazeter et al. (1967b) supplied CD rats (80/sex/dose) with propazine
(80 WP) in the diet for 90 days at dose levels of 0, 50, 200 or
1,000 ppm active ingredient. Based on th
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Propazine August, 1987
-6-
90-day study, a reduction in body weight (30%, no p value given) and
feed consumption were reported at 2,500 mgAg/day, but no effects
were seen at 250 mg/kg/day. No histopathological evaluations were
performed at the high-dose level. After 180 days, rats administered
propazine at 250 mg/kg/day were similar to untreated controls in
growth rates, daily food consumption, gross appearance and behavior,
mortality, gross pathology and histopathologyo This study identified
a NQAEL of 250 mg/kg/day and a LOAEL of 2,500 mg/kg/day.
0 Jessup et al. (1980a) fed CD mice (60/sex/dose) technical propazine
(purity not specified) for 2 years at dose levels of 0, 3, 1,000 or
3,000 ppm. Based on the assumption that 1 ppm in the diet of mice is
equivalent to 0.15 mg/kg/day (Lehman, 1959), these doses correspond
to 0, 0.45, 150 or 450 mg/kg/day. The general appearance, behavior,
survival rate, body weights, organ weights, food consumption and
incidence of inflammatory, degenerative or proliferative alterations
in various tissues and organs did not differ significantly from
untreated controls. The author identified a NOAEL of 3,000 ppm (450
mg/kg/day, the highest dose tested).
0 Jessup et al. (1980b) fed CD rats (60 to 70/sex/dose) technical
propazine (purity not specified) in the diet for 2 years at dose
levels of 0, 3, 100 or 1,000 ppm. Based on the assumption that 1 ppm
in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959),
this corresponds to doses of 0, 0.15, 5 or 50 mg/kg/day. No compound-
related effects were observed in behavior, appearance, survival, feed
consumption, hematology, urinalysis and in nonneoplastic alterations
in various tissues and organs. Mean body weight gains appeared to be
lower in the treatment groups than the control groups. Body weights
at 104 weeks were lower than controls at all dose levels. The percent
decreases in males and females were as follows: -6.3 and -3.9% (3
ppm); -4.6 and -5.6% (100 ppm); -13.1 and -11.4% (1,000 ppm). These
decreases were statistically significant in males at 3 and 1,000 ppm,
and in females at 100 and 1,000 ppm. The decreases at 3 or 100 ppm
appeared to be so small that they may not be considered biologically
significant; a NOAEL was identified at 100 ppm (5 mg/kg/day).
Reproductive Effects
0 Jessup et al. (1979) conducted a three-generation study in which CD
rats (20 females and 10 males/dose) were administered technical
propazine in the diet at 0, 3, 100 or 1,000 ppm. Based on the
assumption that 1 ppm in the diet is equivalent to 0.05 mg/kg/day
(Lehman, 1959), this corresponds to doses of 0, 0.15, 5 or 50 mg/kg/day.
No compound-related effects were observed in any dose group in
general behavior, appearance or survival of parental rats or pups.
The mean parental body weights were statistically lower at 1,000 ppm
(50 mg/kg/day). No differences were reported in feed consumption
for treated and control animals. No treatment-related effects were
observed in fertility, length of gestation or viability and surivival
of the pups through weaning. Mean pup weights at lactation were not
adversely affected at 3 or 100 ppm (0.15 or 5 mg/kg/day). However,
at 1,000 ppm (50 mg/kg/day), there was a statistically significant
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Propazine August, 1987
-7-
decrease in mean pup weights for all generations except Fja. Based
on these data, a NOAEL of 100 ppm (5 mg/kg/day) was identified.
Developmental Effects
0 Fritz (1976) administered technical propazine (0, 30, 100, 300 or
600 mg/kg/bw) orally by intubation to pregnant Sprague-Dawley rats
(25/dose) on days 6 through 15 of gestation. No maternal toxicity,
fetotoxicity or teratogenic effects were observed at 100 mg/kg/day
or lower. Maternal body weight and feed consumption were reduced at
300 mg/kg/day or higher. Fetal body weight was reduced, and there
was delayed skeletal ossification (of calcanei) at 300 mg/kg/day or
higher. Based on body weights, a maternal NOAEL of 100 mg/kg/day
and a fetal NOAEL of 100 mg/kg/day were identified.
0 Salamon (1985) dosed pregnant CD rats (21 to 23 animals per dose
group) with technical propazine (99.1% pure) by gavage at dose levels
of 0, 10, 100 or 500 mg/kg/day on days 6 through 15 of gestation.
Maternal body weight and feed consumption were statistically signifi-
cantly (p <0.05) decreased at doses of 100 mg/kg/day or higher.
Fetal body weight was reduced, and ossification of cranial structures
was delayed at 500 mg/kg/day. Based on maternal toxicity, a NOAEL of
100 mg/kg/day was identified.
Mutagenicity
0 Puri (1984a) reported that propazine (0, 0.4, 20, 100 or 500 ug/mL)
did not produce DNA damage in human fibroblasts in vitro.
0 Puri (1984b) reported that propazine (0, 0.50, 2.5, 12.5 or 62.5
ug/mL) did not cause DNA damage in rat hepatocytes in vitro.
0 Strasser (1984) reported that propazine administered to Chinese
hamsters by gavage (0, 1,250, 2,500 or 5,000 mg/kg) did not cause
anomalies in nuclei of somatic interphase cells.
Carcinogenicity
0 Innes et al. (1969) fed propazine in the diet to 72 mice (C57BL/6
x AKRF1, C57BL/6 x C3H/ANF1) for 18 months at a dose level of 46.4
ing/kg/day. Based on histopathological examination of tissues (no data
reported), the authors stated that propazine, at the one dose tested,
did not cause a statistically significant increase in the frequency
of any tumor type in any sex-strain subgroup or combination of groups.
0 Jessup et al. (1980b) fed CD rats (60 to 70/sex/dose) technical
propazine (purity not specified) in the diet for 2 years at dose
levels of 0, 3, 100 or 1,000 ppm. Based on the assumption that 1 ppm
in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959),
this corresponds to doses of 0, 0.15, 5 or 50 mg/kg/day. Tumor inci-
dence was evaluated for a variety of organs and tissues. The most
commonly occurring tumors were mammary gland tumors in female rats.
At the highest dose tested (1,000 ppm, 50 mg/kg/day), the authors
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Propazine August, 1987
-8-
reported an increase in adenomas (10/55, 18%), adenocarcinomas (9/55,
16%) and papillary carcinomas (8/55, 15%) compared to corresponding
tumor levels in untreated controls (3/55, 5%), (6/55, 11%) and
(4/55, 7%), respectively. Also, it was reported that the percentage
of tumor-bearing rats was 73% in the high-dose treated group compared
to 50% in corresponding untreated controls. The authors did not
consider these increases to be statistically significant. However,
in 1981, Somers reported historical control values of 122/1,248 (10%)
for adenomas and of 769/1,528 (50%) for percentage of tumor-bearing
animals. Further evaluations by Somers (1981) of the above data
(control and treated) and historical control data indicated that the
increase in mammary gland adenomas and the number of rats bearing one
or more tumor was statistically significant (p <0.02).
0 Jessup et al. (1980a) fed CO mice (60/sex/dose) technical propazine
(purity not stated) for 2 years at dose levels of 0, 3, 1,000 or
3,000 ppm. Assuming that 1 ppm in the diet of mice is equivalent
to Oci5 mg/kg/day (Lehman, 1959), this corresponds to doses of 0,
0.45, 150 or 450 mg/kg/day. The incidence of proliferative and
neoplastic alterations in the treated groups did not differ signifi-
cantly from the control group at any dose level.
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) = mq/L ( ug/L)
(UF) x ( L/day)
where;
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in rag/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/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 propazine. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child, 1.0 mg/L (1,000 ug/L,
calculated below), be used at this time as a conservative estimate of the
One-day HA value.
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Propazine August, 1987
-9-
Ten-day Health Advisory
The study by Salamon (1985) has been selected to serve as the basis for
the determination of Ten-day HA value for propazine. In this teratogenicity
study in rats, body weight was decreased in dams dosed on days 6 to 15 of
gestation with 100 mg/kg/day or greater. No adverse effects were observed
in either dams or fetuses at 100 mg/kg/day. The rat study by Fritz (1976)
reported maternal and fetal toxicity at 300 mg/kg/day, but not at 100 mg/kg/day.
This NOAEL was not selected, since maternal weight loss was noted at this dose
by Salamon (1985).
Using a NOAEL of 10 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (10 mg/kg/day) (10 kg) . u 0 mg/L (j 000 ug/L)
(100) (1 L/day)
where:
10 mg/kg/day = NOAEL, based on absence of maternal and developmental
toxicity in rats exposed to propazine by gavage on
days 6 through 15 of gestation.
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
The 90-day feeding study in dogs by Wazeter et al. (1967a) has been
selected to serve as the basis for the Longer-term HA for propazine. In this
study, body weight loss occurred at 1,000 ppm (25 mg/kg). A NOAEL of 200 ppm
(5 mg/kg/day) was identified. This is supported by the 90-day rat feeding
study by Wazeter et al. (1967b), which identified a NOAEL of 10 mg/kg/day and
a LOAEL of 50 mg/kg/day. The 90-day study in rats by Geigy (1960) has not
been selected, since the NOAEL (250 mg/kg/day) is higher than the LOAEL
values reported above.
Using a NOAEL of 5 mg/kg/day, the Longer-term HA for the 10-kg child is
calculated as follows:
Longer-term HA = (5 mg/kq/day) (10 kg) = Q.5 mg/L (500 ug/L)
(100) (1 L/day)
where:
5 mg/kg/day = NOAEL, based on absence of effects on appearance,
behavior, hematology, urinalysis, clinical chemistry,
gross pathology, histopathology and body weight gain
in dogs exposed to propazine via the diet for 90 days.
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Propazine August, 1987
-10-
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-tern HA for a 70-kg adult is calculated as follows:
Longer-term HA = (5 mg/kg/day) (70 kg) = 1075 mg/L (1750 Ug/L)
(100) (2 L/day)
where:
5 mg/kg/day = NOAEL, based on absence of effects on appearance,
behavior, hematology, urinalysis, clinical chemistry,
gross pathology, histopathology and body weight gain
in dogs exposed to propazine via the diet for 90 days.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/OCW
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.
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, 1986a), then caution should be exercised
in assessing the risks associated with lifetime exposure to this chemical.
The 2-year feeding study in rats by Jessup et al. (1980b) has been
selected to serve as the basis for determination of the Lifetime HA for
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Propazine August, 1987
-11-
propazine. No effects were detected on behavior, appearance, mortality, food
consumption, hematology, urinalysis or body weight gain at doses of 5 mg/kg/day.
At 50 mg/kg/day, decreased weight gain was noted, and there was evidence of
increased tumor frequency in the mammary gland. This NOAEL value (5 mg/kg/day)
is supported by the NOAEL of 5 mg/kg/day in the three-generation reproduction
study in rats by Jessup et al. (1979). The 2-year feeding study in mice by
Jessup et al. (1980a) has not been selected, since the data suggest that the
mouse is less sensitive than the rat.
The Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (5 mg/kg/day) = 0.02 mg/kg/day
(100) (3)
where:
5 mg/kg/day = NOAEL, based on absence of effects on behavior,
appearance, mortality, hematology, urinalysis or
body weight gain in rats exposed to propazine via
the diet for 2 years.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = additional uncertainity factor to account for data gaps
(chronic feeding dog study) in the propazine database.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
r/kg/day)
[2 L/day)
DWEL = (0*02 mg/kg/day) (70 kg) - Q.70 mg/L (700 ug/L)
where:
0.02 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 = (0'70 mg/L > (20%) = 0.014 mg/L (14 ug/L)
(10)
where:
0.70 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenicity.
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Propazine August, 1987
-12-
Evaluation of Carcinogenic Potential
0 No evidence of increased tumor frequency was detected in a 2-year
feeding study in mice at doses up to 450 mg/kg/day (Jessup et al.,
1980a) or in an 18-month feeding study in mice at a dose of 46.4
mg/kg/day (Innes et al., 1969).
0 Jessup et al. (1980b) reported that the occurrence of mammary gland
tumors in female rats administered technical propazine in the diet for
2 years at 1,000 ppm (50 mg/kg/day) was increased but did not differ
significantly from concurrent controls. However, a reevaluation of
the data by Somers (1981) that considered historical control data
indicated that the increase in mammary gland adenomas and the number of
rats bearing one or more tumors was statistically significant (p <0.02)c
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of propazine.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), propazine may be classified in
Group C: possible human carcinogen. This category is for substances
with limited evidence of carcinogenicity in animals in the absence of
human data.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA (1986c) has established residue tolerances for propazine
in or on various agricultural commodities of 0.25 ppm (negligible)
based on a Provisionary Acceptable Daily Intake (PADI) of 0.005 mg/kg/day.
0 NAS (1977) determined an Acceptable Daily Intake (ADI) of 0.464
m9Ag/day, based on a NOAEL of 46.4 mg/kg identified in an 80-week
feeding study in mice with an uncertainty factor of 1,000.
0 NAS (1977) calculated a chronic Suggested-No-Adverse-Effeet-Level
(SNARL) of 0.32 mg/L, based on an ADI of 0.0464 mg/kg/day and a
relative source contribution factor of 20%.
VII. ANALYTICAL METHODS
0 Analysis of propazine is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
propazine, but it is estimated that the detection limits for analytes
included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information regarding treatment technologies applicable to the
removal of propazine from contaminated water was found in the available
literature.
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Propazine August, 1987
-13-
IX. REFERENCES
Bakke, J.E. , J.D. Robbins and V.J. Fell. 1967. Metabolism of 2-chloro-4, 6-
bis(isopropylamine), S-triazine (propazine) and 2-methoxy-4, 6-bis (isopro-
pylamino)-s-triazine (prometone) in the rat. Balance study and urinary
metabolite separation. J. Agr. Food Chem. 1 5(4):628-631 .
Cannelongo, B. , E. Sabol, R. Sabol et al.* 1979. Rabbit acute dermal toxicity.
Project No. 1132-79. Unpublished study. MRID 00111700.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Fritz, H.* 1976. Reproduction study. G 30028 technical rat study. Segment II,
Test for teratogenic or embryotoxic effects. Experiment No. 227642.
Unpublished study. MRID 00087879.
Geigy, S.A.* 1960. Chronic toxicity of propazine 50 WP. Unpublished study.
MRID 00111671.
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
Wilkins.
IPC.* 1984. Industrie Prodotti Chimici. Atrazine product chemistry data.
Unpublished compilation. MRID 00141156.
Innes, J. , B. Ulland, M.G. Valeric, L. Petrucelli, L. Fishbein, E. Hart and
A. Pallotta. 1969. Bioassay of pesticides and industrial chemicals for
tumor igenicity in mice. A preliminary note. J. Natl. Can. Inst.
Jessup, D.C., R.J. Arceo and J.E. Lowry.* 1980a. Two-year carcinogenicity
study in mice. IRDC No. 382-004. Unpublished study. MRID 00044335.
Jessup, D.C., G. Gunderson, L.J. Ackerman et al.* 1980b. Two-year chronic
oral toxicity study in rats. IRDC No. 382-007. Unpublished
study. MRID 00041408.
Jessup, D.C., C. Schwartz, R.J. Arceo et al.* 1979. Three generation study
in rat. IRDC NO. 382-010. Unpublished study. MRID 00041409.
Lehman, A.J. 1959. Appraisal of the safety of rhemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off.
Meister, R. , ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
NIOSH. 1985. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances (RTECS). National Library of
Medicine Online File.
-------
Propazine August, 1987
-14-
Palazzolo, R.« 1964. Report to Geigy Research Laboratories. Repeated dermal
toxicity of propazine 80 W. Unpublished study. MRID 00111670.
Puri, E.* 1984a. Autoradiographic DNA repair test on human fibroblasts with
G30028 technical. Test No. 831373. Unpublished study. MRID 00150024.
Puri, E.* 1984b. Autoradiographic DNA repair test on rat hepatocytes with
G30028 technical. Test Report No. 831371. Unpublished study. MRID
00150623.
Salamon, C.* 1985. Teratology study in albino rats with technical propazine.
Report No. 450-1788. Unpublished study. American Biogenics Corporation.
MRID 00150242.
Somers, J.A.* 1981. Letter sent to Robert J. Taylor dated April 14, 1981.
Propazine herbicide chemical no. 080808, 6(a)(2): submission of treated
vs. control data involving mammary tumors in rats in IRDC study no.
382-007; response to November 18, 1980. MRID 00076955.
Stenger and Kindler.* 1963a. Subchronic oral toxicity in the rat. A trans-
lation of: subchronische toxizitat—ratte p.o. Unpublished study.
MRID 00111678.
Stenger and Kindler.* 1963b. Acute toxicity - mouse, oral. Translation of
akute toxizitat—maus per OS. Unpublished study. MRID 00111675.
Stenger and Huber.* 1961. Subchronic toxicity—rat skin. A translation of:
subchronic toxizitat—ratte, haut. Unpublished study, including German
text. MRID 00111677.
Strasser, P.* 1984.' Nucleus anomaly test in somatic interphase nuclei of
Chinese hamster. Test Report No. 831372. Unpublished study. MRID
00150622.
TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51M85):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
U.S. EPA. 1986c. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.243. July 1, 1985. p. 296.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Environmental fate
of propazine. Memo from C. Eiden to D. Tarkas, June 9.
U.S.G.S. 1985. U.S. Geological Survey. Regional assessment project. C. Eidcn,
-------
Propazine August, 1987
-15-
Wazeter, F., R. Buller, R. Geil et al.* 1967a. Ninety-day feeding study in
the beagle dog. Propazine SOW. Report No. 248-002. Unpublished study.
MRID 00111680.
Wazeter, F., R. Buller, R. Geil et al.* 1967b. Ninety-day feeding study in
albino rats. Propazine SOW. Report No. 248-001. Unpublished study.
MRID 00111681.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
Auiust. 19&7
PROPHAM
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 mode., is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
-------
Propham August, 1987
-3-
Environmental Fate
0 Ring-labeled 14c-propham (purity unspecified), at 4 ppm in unbuffered
distilled water declined to 2.4 ppm during 14 days of irradiation
with a Pyrex-filtered light (uncharacterized) at 25°C (Gusik, 1976).
Degradation products included isopropyl 4-hydroxycarbanilate (3.5% of
applied propham.7, isopropyl 4-aminobenzoate (approximately 0.1%),
1-hydroxy-2-propylcarbanilate (approximately 0.1%), and polymeric
materials (10 to 12%). No degradation occurred in the dark control
during the same period.
0 Under aerobic conditions, ring-labeled Hc-propham (test substance
uncharacterized), at 2 ppm, degraded with a half-life of 2 to 7 days in
silt loam soil, (Hardies, 1979; Hardies and Studer, 1979a), 4 to 7 days
in loam soil (Hardies and Studer, 1979b), and 7 to 14 days in'sandy '
loan soil (Hardies and Studer, 1979c) when incubated in the dark at '
approximately 25°C and 60% of water holding capacity.
0 Under anaerobic conditions, ring-labeled 14C-propham (test substance
uncharacterized) declined from 8.5 to <5% of the applied radioactivity
during 60 days of incubation in silt loam soil in the dark at approxi-
mately 25°C and 60% of water holding capacity (Hardies 1979; Hardies
and Studer, 1979a). Under anaerobic conditions, ring-labeled 14C-
propham (test substance uncharacterized) declined from approximately
0.08 to approximately 0.04 ppm during 61 days of incubation in loam
soil in the dark at approximately 25°C and 60% of water holding
capacity (Hardies and Studer, 1979b); in sandy loam soil, the decline
was from approximately 0.06 to 0.03 ppm during 63 days of incubation
(Hardies and Studer, 1979c).
0 14c-Propham (purity unspecified) at 0.2 to 20 ppm was adsorbed to two
silt loams, a silty clay loam, a sandy clay loam, and two sandy loam
soils with Freundlich K values of 0.74 and 2.72, 1.77, 0.65, and 0.27
and 1.58, respectively (Hardies and Studer, 1979d). Ring-labeled
14C-propham (purity unspecified) was very mobile (>98% of applied
propham in leachate) in 30.5-cm columns of sandy clay loam and sandy loam
soil leached with 20 inches of water (Hardies and Studer, 1979e). It
was less mobile in columns of Babcock silt loam (42.3% in leachate),
silty clay loam (approximately 62% at 11- to 27-cm depth), and Wooster
silt loam (approximately 54% at 7.6- to 15-cm depth) soils. Aged
(30-day) residues were relatively immobile in Wooster silt loam soil;
<1% of the applied radioactivity moved from the treated soil.
0 Propham residues dissipated from the upper 6 inches of sandy loam,
sandy clay loam, silty loam, and silty clay loam field plots with
half-lives of 42 to 94, 57 to 160, 42 to 147, and approximately
21 to 42 days, respectively, following application of propham (ChemHoe
135, 3 Ib/gal F1C) at 4 and 8 Ib active ingredient (a.i.) per acre
in September-November, 1977 (Pensyl and Wiedmann, 1979). Residues
were nondetectable «0.02 ppm) within 164 to 283 days after treatment
at all rates and sites. In general, propham residues in the 6- to
12-inch depth were <0.04 ppm. Propham (3 Ib/gal F1C) applied at
6 Ib a.i./A in mid-May dissipated with a half-life of 10 to 15 days in
-------
Propham August, 1987
-5-
Excretion
14c-Propham is rapidly excreted primarily in the urine of rats. Peak
urinary concentrations were reached 6 hours post-treatment. It was
found that 96% and 2% of the administered dose of Kc-propham (100
ing/kg 99% a.i.) was excreted in the urine and feces, respectively (Chen,
1979; Paulson e£al.f 1972).
Fang et al. (1972) reported that after oral administration of ring-
or chain-Hc-labeled Propham (99% a.i.) to rats, 80 to 85% of the
administered dose was excreted in the urine over a 3-day period. In
animals dosed with 14C-isopropyl-labeled propham, 5% was detected as
expired carbon dioxide.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of propham in humans.
Animals
Short-term Exposure
0 Terrell and Parke (1977) administered single oral doses of propham
(technical grade, purity not specified) to groups of 10 male and 10
female rats and monitored adverse effects for 14 days. Doses of
2,000 mg/kg produced loss of righting reflex, ptosis, piloerection,
decreased locomotor activity, chronic pulmonary disease and rugation
and irregular thickening of the stomach. The acute oral LDsg values
in male and female rats were reported to be 3,000 ± 232 mg/kg and
2,360 i 118 mg/kg, respectively. A No-Observed-Adverse-Effect-Level
(NOAEL) cannot be derived from the study because the doses used were
too high, and adverse effects were found at all doses tested.
Brown and Gross (1949) reported that when a single dose of 1,140
mg/kg propham (purity not specified) was administered orally to rats
(number not specified), no adverse effects were observed. Doses of
2,200 to 3,320 mg/kg resulted in periods of light anesthesia. Deep
anesthesia was produced when 4,420 mg/kg of propham was administered,
with subsequent death of 38% of the test animals.
0 The acute inhalation LC50 value in albino rats was reported to
be 10.71 mg/L (PPG Industries, 1978).
Dermal/Ocular Effects
The acute dermal LD50 value in albino rabbits was reported to be
greater than 3,000 mg/kg (PPG Industries, 1978).
0 Propham (3% aqueous solution) was slightly irritating when applied to
the skin and eyes of albino rabbits (PPG Industries, 1978).
-------
Propham Augunt, 1S87
-7-
Mutagenicity
0 Using the Ames Salmonella test. Mar gar d (1978) reported that propham
(purity not specified, 1,000 ug/plate) did not show any indications
of mutagenic activity either with or without activation.
when propham (1O0 ug/mL, purity not specified) was applied to cultures
containing BALB/c 3T3 cell lines, no clonal transformation was evident
(Margard, 1978).
0 Friedrick and Nass (1983) reported that propham (1.1 to 2.2 mM) did
not induce mutation in S49 mouse lymphoma cells.
Carcinogenicity
Innes et al. (1969) administered propham to C57BL/6XC3H/AMF or
C57BL/6XAKR mice (18/sex) in the diet at 560 ppm for 18 months.
Assuming that 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day
(Lehman, 1959), this corresponds to a dose -of about 84 mg/kg/day.
The incidence of tumors was not significantly increased (p >0.05)
for any tumor type in any sex-strain subgroup or in the combined
sexes of either strain. This duration of exposure and this dose
level may not be sufficient for detecting late-occurring tumors.
0 Hueper (1952) fed 15 Osborne Mendel rats (sex not specified) dietary
propham (20,000 ppm, purity not specified) for 18 months. The animals
were alternately placed from 1 to 2 months on the diet followed by
1 to 2 weeks on normal diet. Assuming that 1 ppm in the diet of rats
is equivalent to 0.05 mg/kg/day (Lehman, 1959), the dietary level was
equivalent to 1,000 mg/kg/day. The time-weighted average can not be
calculated due to a lack of detailed reporting of the study design.
No tumors were observed in 6 of 8 surviving rats that were evaluated
histologically. This study is limited by the low number of animals
used, the poor survival rate, short duration, limited histopathological
examination and method of treatment.
0 Van Esch and Kroes (1972) fed groups of 23 to 26 golden hamsters 0 or
0.2% propham (2,000 ppm, purity not specified) in the diet for
33 months. Assuming that 1 ppm in the diet of hamsters is equivalent
to 0.04 mg/kg/day (Lehman, 1959), these levels are equivalent to 0 or
80 mg/kg/day. Based on histological examination, the authors reported
no significant increase in tumor incidence.
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:
. _ Bg/l ,_ ug/L)
-------
Propham Augvst, Iy87
-9-
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from &n animal study.
1 L/day = assumed daily water consumption of a child.
The Longer-term HA ijpx a 70-kg adult is calculated as follows:
Longer-term HA ='(SO mq/kg/day) (70 kg) . 17.5 mg/L (17 500 ug/L)
(100) (2 L/day) *
where:
50 mg/kg/day = NOAEL, based on the absence of inhibition of cholin-
esterase or effects on organ weights in rats fed
propham in the diet for 91 days.
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). Bie 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 or. 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
No chronic study was found in the available literature that was suitable
for determination of the Lifetime HA value for propham. The chronic studies
by Innes et al. (1969), Hueper (1952) and Van Esch and Kroes (1972) did not
provide adequate data on noncarcinogenic end points. In the absence of
appropriate chronic data, the 90-day study by Tisdel et al. (1979), which
-------
Prophan August, 1987
-11-
Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), propham may be classified
in Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No information on other existing criteria, guidelines and standards
was found in the available literature.
VII. ANALYTICAL METHODS
0 Analysis of propham is by a high-performance liquid chromatographic
(HPLC) method applicable to the determination of certain carbamate
and urea pesticides in water samples (U.S. EPA, 1986b). This method
requires a solvent extraction of approximately 1 L of sample with
methylene chloride using a separatory funnel. The me thy 1 en e chloride
extract is dried and concentrated to a volume of 10 mL or less.
Compounds are separated by HPLC, and measurement is conducted with a
UV detector. The method detection limit has not been determined for
propham, but it is estimated that the detection limits for analytes
included in this method are in the range of 1 to 5 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAC) adsorption
will remove propham from water.
0 Whittaker (1980) experimentally determined adsorption isotherms for
propham on GAC.
0 Whittaker (1980) reported the results of studies with GAC columns
operating under bench scale conditions. At a flow rate of
0.8 gal/min/sq ft and an empty bed contact time of 6 minutes, propham
breakthrough (when effluent concentration equals 10% of influent
concentration) occurred after 720 bed volumes (BV).
0 In the same study, Whittaker (1980) reported the results for seven
propham bi -solute solutions when passed over the same GAC continuous-
flow column.
0 The studies cited above indicate that GAC adsorption ia the most
promising treatment technique for the removal of propham from water.
However, selection of individual or combinations of technologies for
propham removal from water must be based on a case-by-case technical
evaluation and an assessment of the economics involved.
-------
JProphan August, 1987
-13-
Hardies, D.E. and D.Y. Studer.* 1979e. A laboratory study of the leaching of
isopropyl carbanilate in soils. Unpublished study prepared and submitted
on Nov. 1, 1984, by PPG Industries, Inc., Chemical Division, Barberton,
OH: Accession No. 255364.
Hueper, W.C.* 1952. Carcinogenic studies on isopropyl-n-phenyl-carbamate.
Indus. Med. Surg. £l(2):71-74. Also unpublished submission. MRID
00091228.
IARC. 1976. International Agency for Research on Cancer. IARC monographs
on the evaluation of carcinogenic risk of chemicals to man. Lyon: IARC.
Vol. 12.
Innes, J., B. Ulland, M.G. Valeric, L. Petrucelli, L. Fishbein, E. Hart and
A. Pallotta. 1969. Bioassay of pesticides and industrial chemicals for-
tumorigenicity in mice. A preliminary note. J. Natl. Can. Inst.
42:1101-1114.
Lehman, A.j. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Margard, W.* 1978. Summary report on in vitro bioassay of selected compounds.
Unpublished study. MRID 00115428.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Paulson, G. and A. Jacobsen.* 1974. Isolation and identification of
propham metabolites from animal tissues and milk. Unpublished study.
MRID 00115440.
Paulson, G., A. Jacobsen and R. Zaylskie.* 1972. Propham metabolism in the
rat and goat: Isolation and identification of urinary metabolites.
Unpublished study. MRID 00115397.
Pensyl, J. and J.L. Wiedmann.* 1979. Field dissipation of IPC and PPG-124
from soil treated with ChemHoe 135 FL3: BR 21574. Unpublished study
received Sept. 17, 1979 under 748-224; submitted by PPG Industries, Inc.,
Barberton, OH; CDL:240987-E. MRID 00038947.
PPG Industries, Inc.* 1970. Primary rabbit eye irritation study. Inter-
national Bio-Test Laboratories. (#A-9252D). Unpublished study.
EPA Accession No. 097066.
PPG Industries, Inc.* 1978. Study: IPC toxicity to test subjects.
Unpublished study. MRID 00115420.
Ravert, J.* 1978. Three generation reproductive study of IPC in Sprague
Dawley rats. Unpublished study. MRID 00115425.
Ravert, J. and G. Parke.* ' 1977. Investigation of teratogenic and toxic
potential of IPC-50%-rats. Unpublished study. MRID 00115434.
-------
August, 1987
SIMAZINE
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)f 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 ris * i stimates may also be calculated
using the One-hit, Weibull, Logit or Problt 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.
-------
Simazine August, 1987
-2-
The information used in preparing this Health Advisory was collected
primarily from the open literature and the Simazine Registration Standard
(U.S. EPA, 1983).
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 122-34-9
Structural Formula
Cl
kNHC2H8
2-Chloro-4,6-bis(ethylamino)-1,3,5-triazine
Synonyms
0 Aquazine, Cekusan, Framed (discontinued by Farmoplant), G-27692,
Gesatop, Primatol, Princep, Simadex, Simanex, Tanzene (Meister, 1984)
Uses
0 Simazine is used as a selective preemergence herbicide for control of
most annual grasses and broadleaf weeds in corn, alfalfa, established
bermuda grass, cherries, peaches, citrus, different kinds of berries,
grapes, apples, pears, certain nuts, asparagus, certain ornamental
and tree nursery stock, and in turf grass soil production (Meister,
1984). It is also used to inhibit the growth of most common forms of
algae in aquariums, ornamental fish ponds and fountains. At higher
rates, it is used for nonselective weed control in industrial areas.
Properties (Berg, 1984; Freed, 1976; Windholz et al., 1983)
Chemical Formula
Molecular Weigh* 201.69
Physical State (room temperature) White, crystalline solid
Boiling Point
Melting Point 225 to 227°C
Density 1.302 g/cm3
Vapor Pressure (20°) 6.1 x 10-9 mm Hg
Water Solubility (20e) 3.5 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
-------
Simazine August, 1987
-3-
Occurrenee
0 Simazine has been found in 877 of 5,067 surface water samples analyzed
and in 229 of 2,282 ground water samples (STORET, 1987). Samples
were collected at 472 surface water locations and 1,730 ground water
locations, and simazine was found in 22 states. The 85th percentile
of all non-zero samples was 2.18 ug/L in surface water and 1.60 ug/L
in ground water sources. The maximum concentration found in surface
water was 1,300 ug/L, and in ground water it was 800 ug/T .
0 Simazine has been found in ground water in California, Pennsylvania
and Maryland; typical positives were 0.2 to 3.0 ppb (Cohen et al., 1986),
Environmental Fate
c Simazine did not hydrolyze in sterile aqueous solutions buffered at
pH 5, 7 or 9 (20°C) over a 28-day test period (Gold et al., 1973).
0 Under aerobic soil conditions, the degradation of simazine depends
largely on soil moisture and temperature (Walker, 1976). In a sandy
loam soil, half-lives ranged from 36 days to 234 days. Simazine
applied to loamy sand and silt loam soils and incubated (25 to 30°C)
for 48 weeks, dissipated with half-lives of 16.3 and 25.5 weeks,
respectively (Monsanto Company, date not available). Simazine degra-
dation products, 2-chloro-4-ethylamino-6-amino-s-triazine (G-28279),
2-chloro-4,6bis(amino)-s-triazine, and several unidentified polar
compounds were detected 32 and 70 days after a sandy loam soil had
been treated with 14c-simazine (Beynon et al., 1972). The degradates
2-hydroxy-4,6=bis(ethylamino)-s-triazine and 2-hydroxy-4-ethylamino-
6-amino-s-triazine were also detected in aerobic soil (Keller, 1978).
0 Under anaerobic conditions, 14c-simazine had a half-life of 8 to 12
weeks in a loamy sand soil (Keller, 1978). The treated soil (10 ppm)
was initially maintained for 1 month under aerobic conditions,
followed by 8 weeks under anaerobic conditions (flooded with water
and nitrogen). Degradates found i-.cluded G-28279, 2-caj.oro-4,6-
bis(amino)-s-triazine, 2-hydroxy-4,6-bis(ethylamino)-s-triazine, and
2-hydroxy-4-ethylamino-6-amino-s-triazine.
0 Simazine is expected to be slightly to very mobile in soils ranging
in texture from clay to sandy loam based on column leaching, soil
thin-layer chromatography (TLC), and adsorption/desorption (batch
equilibrium) studies. Using batch equilibrium tests, K$ values
determined for 25 Missouri soils ranged from 1.0 for a sandy loam
to 7.9 for a silty loam (Talbert and Fletchall, 1965). Simazine
adsorption was correlated with soil organic matter content and, to a
lesser extent, with cation exchange capacity (CEC) and clay content
{Talbert and Fletchall, 1965; Helling and Turner, 1968; Helling,
1971). Simazine exhibited low mobility in peat and peat moss (K.J
more than 21) and a higher mobility in clay fractions (Kd values
ranged from 0.0 for kaolinite to 12.2 for montmorillonite (Talbert
and Fletchall, 1965). Freundlich K and n values were determined to
be 7.25 and 0.88, respectively, for a silty clay loam soil.
-------
Simazine August, 1987
-4-
e Simazine, as determined by soil TLC, is mobile to very mobile in sandy
loam soil (Rf 0.80 to 0.96), and of low to intermediate mobility in
loam and silty clay loam (Rf 0.45), sandy clay loam (Rf 0.51), silt
loam (Rf 0.16 to 0.51), clay loam (Rf 0.32 to 0.45) and silty clay
(Rf 0.36) soils. Rf values were positively correlated with soil
organic matter and clay content (Helling, 1971; Helling and Turner,
1968).
Q Based on results of soil column leaching studies, simazine phytotoxic
residues were slightly mobile to mobile in soils ranging in texture
from clay loam to sand (Rodgers, 1968; Harris, 1967; Ivey and Andrews,
1965). Upon application of 18 inches of water to 30-inch soil columns
containing clay loam, loam, silt loam or fine sandy loam soils,
simazine phytotoxic residues leached to depths of 4 to 6, 10 to 12,
22 to 24, and 26 to 28 inches, respectively (Ivey and Andrews, 1965).
• In field studies, simazine had a half-life of about 30 to 139 days in
sandy loam and silt loam soils (Walker, 1976; Martin et a*., 1975;
Mattson et al., 1969). The degradate, 2-chloro-4-ethylariTno-6-
amino-s-triazine (G-28279) was detected at the 0- to 6-inch depth and
at the 6- to 12-inch depth (Martin et al., 1975; Mattson et al., 1969).
0 Simazine residues (uncharacterized) may persist up to 3 years in soil
under aquatic field conditions. Dissipation of simazine in pond and
lake water was variable, with half-lives ranging from 50 to 700 days.
The degradation compound G-28279 was identified in lake water samples,
but was no more persistent than the parent compound (Flanagan et al.,
1968; Kahrs, 1969; Larsen et al., 1966; LeBaron, 1970; Kahrs, 1977;
Smith et al., 1975).
III. PHARMACOKINETICS
Absorption
0 No quantitative information on the gastrointestinal absorption of
simazine in monogastric mammals was located. Bakke and Robbins (1968)
reported that in goats and sheep, from 67 to 77% of a dose of 14C-
simazine (given orally in gelatin capsules) was excreted in urine.
This suggests that absorption was approximately 70%.
Distribution
0 No studies providing data on the tissue distribution of absorbed
simazine in monogastric mammals were found in the available literature.
Metabolism
0 Bradway and Moseman (1982) administered simazine to male Charles
River rats by gavage. Two doses of 0.017, 1.7, 17 or 167 mg/kg
were given 24 hours apart. In 24-hour urine samples, the di-N-dealky-
lated metabolite (2-chloro-4,6-diamino-s-triazine) appeared to be the
major product, ranging from 1.6% at the 1.7 mg/kg-dose to 18.2% at
-------
Simazine August, 1987
-5-
the 167-mg/kg dose, while the mono-N-dealkylated metabolite ranged
from 0.35% at the 1.7-mg/kg dose to 2.8% at the 167-mg/kg dose.
Similar results were obtained by Bohme and Bar (196"), who fed simazine
(formulation and purity not stated) at levels of 200 or 800 mg/kg to
albino rats and at 240 to 400 mg/kg to rabbits. Of the several
metabolites identified, all retained the triazine ring intact. The
principal species were the mono- and di-N-dealkylated metabolites.
Bakke and Robbins (1968) administered 14c-simazine orally by gelatin
capsules to goats and sheep. The sheep were given simazine labeled
on the triazine ring or on the ethylamino side-chain, while goats
were given the ring-labeled compound only. Based on the metabolites
identified in the urine of animals receiving the ring-labeled compound,
there was no evidence to suggest that the triazine ring was metabolized.
In sheep that received chain-labeled triazines, at least 40% of the
ethylamino side-chains were removed. Using ion-exchange chromatography,
18 labeled metabolites were found in urine.
Bohme and Bar (1967) and Larsen and Bakke (1975) observed that rat
and rabbit urinary metabolites from the 2-chloro-s-triazines were all
2-chloro analogs of their respective parent molecules and none of the
metabolites contained the 2-hydroxy moiety. Total N-dealkyla-*on,
partial N-dealkylation, and N-dealkylation with N-alkyl oxidation
were suggested as the major routes of the metabolism of 2-chloro-s-
triazines in rats and rabbits.
Excretion
No quantitative study of simazine excretion routes in monogastric
animals was found in the available literature.
Bakke and Robbins (1968) studied the excretion of 14c-simazine in
goats and sheep using triazines labeled on the ring or on the ethylamino
side-chains. Approximately 67 to 77% of the administered ring-labeled
activity was found in the urine, and 13 to 25% was found in the feces.
Negligible residue was present in the milk immediately after treatment
and within 48 hours of treatment.
Hapke (1968) reported that simazine residues were present in the
urine of sheep for up to 12 days after administration of a single
oral dose. The maximum concentration in the urine occurred from 2
to 6 days after administration.
IV. HEALTH EFFECTS
Humans
Long-term Exposure
0 There were 124 cases of contact dermatitis noted by Yelizarov (1977)
in the Soviet Union among workers manufacturing simazine and propazine.
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Simazine August, 1987
-6-
Mild cases lasting 3 or 4 days involved pale pink erythema and slight
edema. Serious cases lasting 7 to 10 days involved greater erythema
and edema, and also a vesiculopapular reaction that sometimes progressed
to the formation of bullae.
Animals
Short-term Exposure
0 Oral LD5Q values for simazine have been reported to be greater than
5,000 mg/kg in the rat (Martin and Worthing, 1977), the mouse and the
rabbit (USDA, 1984).
0 Mazaev (1965) administered a single oral dose of simazine (formulation
and purity not stated) to rats at 4,200 mg/kg. Anorexia and weight
loss were observed, with some of the animals dying in 4 to 10 days.
When 500 mg/kg was administered daily, all the animals died in 11 to
20 days, with the time of death correlating with the loss of weight.
0 Sheep and cattle seem to be much more susceptible than laboratory
animals to simazine toxicity. Hapke (1968) reported that a single
oral dose of simazine, 50% active ingredient (a.i.), as low as
500 mg/kg was fatal to sheep within 6 to 25 days after administration.
The animals that survived the exposure were sick for 2 to 4 weeks
after treatment and showed loss of appetite, increased intake of
water, incoordination, tremor and weakness. Some of the animals
exhibited cyanosis and clonic convulsions.
0 Palmer and Radeleff (1969) orally exposed cattle by drench to 10 doses
of simazine SOW (purity not stated) at 10, 25 or 50 mg/kg/day and
sheep by drench or capsule to 10 doses at 25, 50 or 100 mg/kg. The
number of test animals in each group was not stated, and the use of
controls was not indicated. Anorexia, signs of depression, muscle
spasms, dyspnea, weakness and uncoordinated gait were commonly observed
in treated animals. Necropsy showed congestion of lungs and kidneys,
swollen, friable livers, and small, hemorrhagic spots on the surface
of the lining of the heart.
0 Palmer and Radeleff (1964) found that repeated oral administration of
simazine 80W (purity not stated) at either 31 daily doses of 50
mg/kg or 14 dai^y doses of 100 mg/kg was fatal to sheep. Simazine
was also lethal when administered at 100 mg/day for 14 days by drench
(Palmer and Radeleff, 1969).
0 The acute inhalation LC50 value of simazine is reported to be more
than 2.0 mg/L of air (4-hour exposure) (Weed Science Society of
America, 1983).
Dermal/Ocular Effects
0 The acute dermal toxicity in rabbits is greater than 8,000 mg/kg
(NAS, 1977).
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August, 1987
Simazine *
-7-
0 In a 21-day subacute dermal toxicity study in rabbits, Ciba-Geigy
(1980) reported that 15 dermal applications of technical simazine at
doses up to 1 g/kg produced no systemic toxicity or any dose-related
alterations of the skin.
0 In primary eye irritation studies in rabbits, simazine at 71 mg/kg
caused transient inflammation of conjunctivae (USDA, 1984).
Long-term Exposure
0 Tai et al. (1985a) conducted a 13-week subacute oral toxicity study
in Sprague-Dawley rats fed technical simazine at 0, 200, 2,000 or
4,000 ppm in their diets. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels correspond
to doses of about 0, 10, 100 or 200 mg/kg/day. Significant dose-
related reductions in food intake, mean body weight and weight gain
occurred in all treated groups. Significant weight loss occurred
in mid- and high-dose animals during the first week of dosing. At
13 weeks, various dose-related effects were noted in hematological
parameters (decreased mean erythrocyte and leukocyte counts and
increased neutrophil and platelet counts), clinical chemistry (lowered
mean blood glucose, sodium, calcium, blood urea nitrogen (BUN),
lactic dehydrogenase (LDH), serum glutamic-oxaloacetic transaminase
(SCOT) and creatinine and increased cholesterol and inorganic phosphate
levels), and urinalysis determinations (elevated ketone levels and
decreased protein levels). Relative and absolute adrenal, brain,
heart, kidney, liver, testes and spleen weights increased, and overy
and heart weights decreased. Necropsies revealed no gross lesions
attributable to simazine. A dose-related incidence of renal calculi
and renal epithelial hyperplasia were detected microscopically in
treated rats, primarily in the renal pelvic lumen and rarely in the
renal tubules. Microscopic examinations revealed no other lesions
that could be attributed to simazine. It appeared to the authors
that reduced mean food intake in treated rats was most likely due to
the unpalatability of simazine. Lower individual body weights and
reduced body weight gains paralleled mean food intake in treated
rats. The majority of the alterations in clinical chemistry values
may have been related to reduced food consumption. Since these
dietary levels of simazine seriously affected the 'nutritional status
of treated rats, the results of this study are of limited value.
0 Tai et al. (1985b) also conducted a 13-week dietary study with beagle
dogs fed technical simazine at 0, 200, 2,000 or 4,000 ppm. Based on
Lehman (1959), these levels correspond to doses of about 0, 5, 50 or
100 mg/kg/day. As in the previously described study in rats, reduced
daily food consumption was attributed to the palatability of simazine
in the diet and corresponded with weight loss, decreased weight gain
and various effects on hematology, clinical chemistry, and urinalysis
determinations. Changes in these parameters were generally similar
to those noted in the rat study (Tai et al., 1985a). Due to the
seriously affected nutritional status of the test animals, the results
of this study are of limited value.
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Simazine August, 1987
-8-
0 Cshurov (1979) studied the histological changes in the organs of
21 sheep following exposures to simazine (50% a.i.) by gavage at 0,
1.4, 3.0, 6.0, 25, 50, 100 or 250 mg/kg/day for various time durations
up to about 22 weeks. Fatty and granular liver degeneration, diffuse
granular kidney degeneration, neuronophagia, diffuse glial proliferation
and degeneration of ganglion cells in the cerebrum and medulla were
found. In sheep that died, spongy degeneration, hyperemia and edema
were observed in the cerebrum; the degree of severity was related to
the dose of simazine and the duration of exposure, "h e thyroid
showed hypofunction after daily doses of 1.4 to 6.0 mg/kg was admini-
stered for periods of 63 to 142 days. The most severe antithyroid
effect followed one or two doses of 250 mg/kg, which in one sheep
produced parenchymatous goiter and a papillary adenoma. This type of
goiter was also seen in sheep administered simazine at 50 or 100 mg/kg
once per week for approximately 22 weeks. Based on these data, a
Lowest-Observed-Adverse-Effect-Level (LOAEL) of 1.4 mg/kg can be
identified; however, it is not clear from the study details whether
the authors considered the 50% formulation when providing the dosage
levels.
Reproductive Effects
0 Woodard Research Corporation (1965) reported that no adverse effects
on reproductive capacity were observed in a three-generation study in
rats. In this study, two groups of 40 weanling rats (20/sex) were
used; one served as the control and the other was fed simazine SOW
at 100 ppm. This corresponds to a dose of about 5 mg/kg/day, based
on the assumptions that 1 ppm in the diet of rats corresponds to
0.05 mg/kg/day (Lehman, 1959). After 74 days of dosing, animals were
paired and mated for 10 days, resulting in F1a litters. After weaning
first litters, parents were remated to produce F1b litters. Weanlings
of parents in the 100 ppm group were divided into two groups and fed
simazine at 50 ppm (approximately 2.5 mg/kg/day) or at 100 ppm.
After 81 days they were mated to produce the F2a and F2b litters.
F2b weanlings were fed the same dietary levels of simazine (0, 50
or 100 ppm). F2b rats were mated to produce F3a and F3D litters.
Reproductive performance of rats fed simazine was basically similar
to that of controls, and no developmental changes were detected. The
No-Observed-Adverse-Effect-Level (NOAEL) for this study is approximately
5 mg/kg/day.
0 Dshurov (1979) reported that repeated administration of simazine (50%
a.i.) to sheep (6.0 mg/kg for 142 days or 25 mg/kg for 37 to 111 days)
caused changes in the germinal epithelium of the testes and disturbances
of spermatogenesis.
Developmental Effects
0 No treatment-related developmental effects were observed by Newell
and Dilley (1978) in the offspring of rats exposed to simazine at 0,
17, 77 and 317 mg/m3 via inhalation for 1 to 3 hours/day on days 7
through 14 of gestation.
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Simazine "»«""• 1987
-9-
• Woodard Research Corporation (1965), as described above in Reproductive
Effects, conducted a three-generation study in which rats were fed
simazine SOW in mixed dosage groups of 50 and 100 ppm (approximately
2.5 and 5 mg/kg/day). No developmental effects wjrr noted in the
offspring.
Mutagenicity
0 Simazine has shown negative results in a variety of microbial
mutagenicity assay systems including tests with the following
organisms: Salmonella typhimurium (Simmons et al., 1979; Commoner,
1976; Eisenbeis et al.( 1981; Anderson et al., 1972); Escherichia
coli (Simmons et al., 1978; Fahring, 1974); Bacillus subtilis
(Simmons et al., 1978); Serratia marcescens (Fahring, 1974); and
Saccharomyces cerevisiae (Simmons et al., 1978).
0 Simazine induced lethal mutations in the sex-linked recessive lethal
test using the fruitfly Drosophila melanogaster (Valencia, 1981).
In a study reported by Murnik and Nash (1977), simazine increased
X-linked lethals when injected into male D. melanogaster, but
failed to do so when fed to larvae.
0 There are contradictory data concerning the ability of simazine to
cause DNA damage. According to Simmons et al. (1979), simazine
induced unscheduled DNA synthesis in a human lung fibroblast assay.
However, in the same test conducted by Waters et al. (1982), simazine
showed a negative response.
0 Simazine does not produce chromosomal effects as indicated by the
sister-chromatid exchange test and mouse micronucleus assay (Waters
et al., 1982).
Carcinogenicity
0 Simazine was not tumorigenic in an 18-month feeding study in mice at
the highest tolerated dose of 215 mg/kg/day (Innes et al., 1969). In
this bioassay of 130 compounds, male and female mice of two hybrid
strains (C57BL/6 x C3H/Anf)F-| and (C57BL/6 x AKR)F-| were exposed to
simazine (purity not stated) at the maximum tolerated dose of 215 rag/kg
by gavage from ages 7 to 28 days. For the remainder of the study,
the animals were maintained on a diet with simazine at 215 mg/kg/day.
Based on information presented only in tabular form, gross necropsy
and histological examination revealed no significant increase in
tumors related to treatment with simazine. Other toxicological data
were not provided. This study is not considered to provide adequate
data to fully assess the carcinogenic potential of simazine.
0 Hazelton Laboratories (1960) conducted a 2-year dietary study in
Charles River rats administered simazine SOW (49.9% a.i.) in the feed
at 0, 1, 10 and 100 ppm (expressed on the basis of 100% a.i.). Based
on the dietary assumptions of Lehman (1959), these levels are equivalent
to approximately 0, 0.05, 0.5 and 5 mg/kg/day. These authors reported
an excess of thyroid and mammary tumors in high-dose females. However,
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Simazine August, 1987
-10-
complete histopathological details were not provided and statistical
significance was not evaluated. Furthermore, the high incidence of
respiratory and ear infections in all groups renders this study
unsuitable for evaluating the carcinogenic potential of simazine.
0 Simazine was found to produce sarcomas at the site of subcutaneous
injection in both rats and mice (Pliss and Zabezhinsky, 1977; abstract
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 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 mg/kg bw/day.
BW a 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
No suitable studies were found in the available literature for the deter-
mination of the One-day HA value for simazine. It is therefore recommended
that 0.05 mg/L (50 ug/L), the Drinking Water Equivalent Level (OWED calculated
below and adjusted for a 10-kg child, be used at this time as a conservative
estimate of the One-day HA value.
Ten-day Health Advisory
No suitable studies were found in the available literature for the deter-
mination of the Ten-day HA value for simazine. It is therefore recommended
that the adjusted DWEL for a 10-kg child of 0.05 mg/L (50 ug/L) be used at
this time as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
No suitable studies were found in the available literature for the deter-
mination of the Longer-term HA values for simazine. It is therefore recommended
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Simazine August, 1987
-1 1-
that the adjusted DWEL of 0.05 mg/L (50 ug/L) be used at this time as a
conservative estimate of the Longer-term HA value for a 1 0-kg child and that
the DWEL of 0.175 mg/L (175 ug/L) be used for a 70-kg 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 exposu-«. 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 mediuo-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 ol 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The three-generation reproduction study in rats by Woodard Research
Corporation (1965) has been selected to serve as the basis for calculation
of the DWEL and Lifetime HA for simazine. In this study, two groups of 40
weanling rats (20/sex) were used; one served as the control, and the other
was fed simazine SOW at 100 ppm (approximately 5 mg/kg/day). After 74 days
of dosing, animals were paired and mated for 10 days, resulting in F-|a litters.
After weaning first litters, parents were remated to produce F1b litters.
Weanlings of parents in the 100 ppm group were divided into two test groups:
one group was fed simazine at 50 ppm (about 2.5 mg/kg/day) and the other at
100 ppm. After 81 days of dosing, animals were mated to produce the F2a and
F2b litters. The F2b weanlings were then divided into 50- and 100-ppm dosage
groups. F2b rats were mated to produce F3a and F2b litters. Reproductive
performance of rats fed simazine was the same as that of controls, and no
teratological changes were detected. The NOAEL for this study is approximately
5 mg/kg/day.
It is important to note that, in this study, rats in the FQ generation were
exposed to simazine at the high dose (100 ppm) only. However, considering that
the F1 and F2 generations treated with 100 ppm did not reflect any adverse
reproductive effects, this feature of the study design did not seem to affect
the results. Therefore, the NOAEL of 5 mg/kg/day is used for calculation of
the RfD.
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Simazine August, 1987
-12-
Step 1: Determination of the Reference Dose (RfD)
RfD = 5 mg/kg/day = 0.005 mg/kg/day
(1,000)
where:
5 mg/kg/day = NOAEL for reproductive and developmental effects in a
three-generation rat study.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study
of less- than-life time duration.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (O'OOS mg/kg/day) (70 kg) = 0.175 mg/L (175 ug/L)
(2 L/day)
where:
0.005 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 = (0.175 mg/L) (20%) = 0.035 mg/L (35 ug/L)
where:
0. 1 75 mg/L = DWEL.
20% = Assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Based on the available data, there is no evidence to show that simazine
is carcinogenic, and no calculations of carcinogenic risk factors for
simazine have been performed. Neither the study in mice by Innes
et al. (1969) nor the study in rats by Haz»'ton Laboratories (1960)
is considered adequate for assessment of the carcinogenicity of this
substance.
0 Simazine is a chloro-s-triazine derivative, with a chemical structure
analogous to atrazine and propazine. Both these two structurally-
related compounds were found to significantly (p >0.05) increase the
incidence of mammary tumors in rats. The structure-activity relation-
ship of this group of chemicals indicates that simazine is likely to
reflect a similar pattern of oncogenic response in rats as atrazine
and propazine. However, a conclusion on this issue must await the
completion of a new 2-year oncogenic study in rats.
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Simazine August, 1987
-13-
0 Applying the criteria described in EPA's guidelines for the assessment
of carcinogenic risk (U.S. EPA, 1986a), simazine may be classified in
Group D: not classified. This category is used for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 A tolerance level of 10 ug/L has been established for simazine and
its metabolites in potable water when present as a result of application
to growing aquatic weeds (U.S. FDA, 1979).
0 Residue tolerances have been established for simazene alone and the
combined residues of simazine and its metan-lites in or on various
raw agricultural commodities (U.S. EPA, 1986b). These tolerances
range from 0.02 ppm (negligible) in animal products to 15 ppm in
various animal fodders.
VII. ANALYTICAL METHODS
0 Analysis of simazine is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus-containing
pesticides in water samples (U.S. EPA, 1986c). In this method,
approximately 1 L of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary column GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
this compound but it is estimated that the detection limits for the
method analytes are in the range of 0.1 to 2 ug/L.
7111. TREATMENT TECHNOLOGIES
0 Treatment technologies which will remove simazine from water include
activated carbon adsorption; ion exchange; and chlorine, chlorine
dioxide, ozone, hydrogen peroxide and potassium permanganate oxidation.
Conventional treatment processes were relatively ineffec:i^e in
removing simazine (Miltner and Fronk, 1935a). Limited data suggest
that aeration would not be effective in simazine removal (ESE, 1984;
Miltner and Fronk, 1985a).
0 Baker (1983) reported that a 16.5-inch GAC filter cap using F-300,
which was placed upon the rapid sand filters at the Fremont, Ohio
water treatment plant and had been in service for 30 months, reduced
the simazine levels by 35 to 89% in the water from the Sandusky
River. Miltner and Fronk (1985a) developed adsorption capacity data
using spiked, distilled water treated with Filtrasorb 400. The
following Freundlich isotherm values were reported for simazine:
K = 490 mg/g; 1/n = 0.56.
At the Bowling Green, Ohio water treatment plant, PAC in conjunction
with conventional treatment achieved an average reduction of 47% of
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Simazine August, 1987
-14-
the simazine levels in the water from the Maumee River (Baker, 1983).
Niltner and Fronk (1985b) monitored simazine levels at water treatment
plants, which utilized PAC, in Bowling Green and Tiffin, Ohio.
Applied at dosages ranging from 3.6 to 33 mg/L, the PAC achieved 43
to 100% removal of simazine with higher percent removals reflecting
higher PAC dosages. Andersen (1968) reported that activated charcoal
(wood charcoal, 300-mesh A.C. from Harrison Clark, Ltd.) was effective
in "inactivating" simazine when mixed into simazine-treated soils,
though no quantitative data on simazine concentrations were reported.
Rees and Au (1979) reported that an adsorption column containing XAD-2
resin removed 81 to 95% of the simazine in spiked tap water.
Turner and Adams (1968) reported that, in a study on the adsorption
of simazine by ion exchange resins (Sheets, 1959), duolite C-3 cation
exchange resin removed from solution up to 2,000 ug of simazine per
gram of resin. Little adsorption was observed with Duolite A-2 anion
exchange resin.
Miltner and Fronk (1985b) reported the bench scale testing results of
the addition of various oxidants to spiked, distilled water. Chlorine
oxidation achieved 62 to 74 percent removal of simazine. However,
when spiked Ohio River water was treated with smaller chlorine dosages
during shorter time intervals, less than 17% removal was achieved.
Chlorine dioxide oxidation of spiked, distilled water achieved only a
22% removal and achieved 8 to 27% removal of simazine in spiked Ohio
River water when applied at a smaller dosage over a shorter time
interval. Ozonation of spiked, distilled water resulted in a 92%
removal of simazine. Oxidation of spiked, distilled water with
hydrogen peroxide obtained a 19 to 42% removal of simazine, and in
spiked Ohio River water, a smaller dosage over a shorter time interval
obtained a simazine removal of 1 to 25%. Potassium permanganate
oxidized up to 26% of the simazine present in spiked distilled water.
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Simazine Au*ust' 1987
-15-
IX. REFERENCES
Andersen, A.H. 1968. The inactivetion of simazine and linuron in soil by
charcoal. Weed Res. 8:58-60.
Anderson, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of herbi-
cides for possible mutagenic properties. J. Agric. Food Chem. 20:649-65-6.
Baiter, D. 1983. Herbicide contamination in municipal water supplies in
northwestern Ohio. Final Draft Report 1983. Prepared for Great Lakes
National Program Office, U.S. Environmental Protection Agency. Tiffin, OH.
Bakke, J.E., and J.D. Robbins. 1968. Metabolism of atrazine and simazine by
the goat and sheep. Abstr. Pap. 155th National Meeting Am. Chem. Soc.
(A43).
Beynon, K.I., G. Stoydin and A.N. Wright. 1972. A comparison of the breakdown
of the triazine herbicides cyanazine, atrazine, and simazine in soils
and in maize. Pestic. Biochem. Physiol. 2:153-161.
Bohme, C., and F. Bar. 1967. The transformation of triazine herbicides in the
animal organism. Food Cosmet. Toxicol. 5:23-28.
Bradway, D.E., and R.F. Moseman. 1982. Determination of -urinary residue
levels of the N-dealkyl metabolites of triazine herbicides. J. Agric.
Food Chem. 30:244-247.
Ciba-Geigy Corporation. 1980. 21-Day dermal study in rabbits. Bio-Research;,
#12012. January 14.
Cohen, S.Z., C. fiiden and M.N. Lorber. 1986. Monitoring ground water for
pesticides in the U.S.A. In; Evaluation of pesticides in ground water,
American Chemical Society Symposium Series. (in press).
Commoner, B. 1976. Reliability of bacterial mutagenesis techniques to
distinguish carcinogenic and noncarcinogenic chemicals. Available from:
National Technical Information System (NTIS), Springfield, VA.
Dshurov, A. 1979. Histological changes in organs of sheep in chronic simazine
poisoning. Zentralbl. Veterinaermed. Reihe A. 26:44-54. [In German
with English abstract]
Eisenbeis, S.J., D.L. Lynch and A.E. Hampel. 1981. The Ames mutagen assay
tested against herbicides and herbicide combinations. Soil Sci. 131:44-47,
ESE. 1984. Environmental Science and Engineering. Review of treatability
data for removal of 25 synthetic organic chemicals from drinking water.
U.S. Environmental Protection Agency, Office of Drinking Water, Washington,
DC.
Fahring, R. 1974. Comparative mutagenicity studies with pesticides. IARC
Sci. Publ. 10:161-181.
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Simazine August, 1987
-16-
Flanagan, J.H., J.R. Foster, H. Larsen et al. 1968. Residue data for
simazine in water and fish. Unpublished study prepared in cooperation
with the University of Maryland and others; submitted by Geigy Chemical
Company, Ardsley, NY.
Gold, B., K. Balu and A. Hofberg. 1973. Hydrolysis of simazine in aqueous
solution: Report No. GAAC-73044. Unpublished study submitted by Ciba-
Geigy Corporation, Greensboro, NC.
Hapke, H. 1968. Research into the toxicology of weedkiller simazine. Berl.
Tieraerztl. Wochenschr. 81:301-303.
Harris, C.I. 1967. Fate of 2-chloro-£-triazine herbicides in soil. J. Agric.
Food Chem. 15:157-162.
Hazelton Laboratories. 1960. A two-year dietary feeding study in rats.
Unpublished study submitted by Ciba-Geigy Corporation. MRID 00037752,
00025441, 00025442, 00042793 and 00080626.
Helling, C.S. 1971. Pesticide mobility in soils: II. Applications of soil
thin-layer chromatography. Proc. Soil Sci. Soc. 35:737-748.
Helling, C.S., and B.C. Turner. 1968. Pesticide Mobility: Determination by
soil thin-layer chromatography. Method dated Nov. 1, 1968. Science.
162:562-563.
Innes, J.R.M., B.M. Ulland, M.G. Valeric et al. 1969. Bioassay of pesticides
and industrial chemicals for tumorigenicity in mice: A preliminary note.
j. Natl. Cancer Inst. 42:1101-1114.
Ivey, M.J., and H. Andrews. 1965. Leaching of simazine, atrazine, diuron,
and DCPA in soil columns. Unpublished study prepared by the University
of Tennessee and submitted by American Carbonyl, Inc., Tenafly, NJ.
Kahrs, R.A. 1969. Determination of simazine residues in fish and water by
microcoulometric gas chromatography. Method No. AG-111 dated Aug. 22,
1969. Unpublished study submitted by Geigy Chemical Company, Ardsley, NY.
Kahrs, R.A. 1977. Simazine lakes—1975 EUP Program: Status Report—1977:
Report No. ABR-77082. Unpublished study submitted by Ciba-Geigy
Corporation, Greensboro, NC.
Keller, A. 1978. Degradation of simazine (Gesatop) in soil under aerobic-
anaerobic and sterile-aerobic conditions: Project Report 05/78.
Unpublished study submitted by Ciba-Geigy Corporation, Greensboro, NC.
Larsen, G.L., and J.E. Bakke. 1975. Metabolism of 2-chloro-4-cyclo-propylamino-
6-isopropylamino-£-triazine (cyprazine) in the rat. J. Agric. Food Chem.
23:388-392.
Larsen, H., D.L. Button, A.R. Eaton et al. 1966. Summary of residue studies—
simazine SOW. Unpublished study prepared in cooperation with U.S. Fish
and Wildlife, Fish Control Laboratory an1 others, submitted by Ciba-Geigy
Corporation, Greensboro, NC.
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August, 1987
Simazine *
-17-
LeBaron, H.M. 1970. Fate of simazine in the aquatic environment: Report No.
GAAC-70013. Unpublished study submitted by Geigy Chemical Co., Ardsley, NY.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, rfrugs
and cosmetics. Assoc. Food Drug Off. U.S.
Martin, H., and C.R. Worthing, eds. 1977. Pesticide manual. Worcester,
England: British Crop Protection Council.
Martin, V., L. Motko, B. Gold et al. 1975. Simazine residue tests: AG-A
No. 1022. Unpublished study prepared in cooperation with University of
Missouri and submitted by Ciba-Geigy Corporation, Greensboro, NC.
Mattson, A.M., R.A. Kahrs and R.T. Murphy. 1969. Quantitative determination
of triazine herbicides in soils by chemical analysis: GAAC-69014. Method
dated Mar. 18, 1969. Unpublished study submitted by Ciba-Geigy Corpo-
ration, Greensboro, NC.
Mazaev, V.T. 1965. Experimental determinations of the maximum permissible
concentrations of cyanuric acid, simazine, and a 2-hydroxy derivative of
simazine in water reservoirs. Chem. Abstr. 62:15304.
Meister, R., ed. 1984. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Co.
Miltner, R.J., and C.A. Fronk. 1985a. Treatment of synthetic organic contami-
nants for Phase II regulations. Progress report. U.S. Environmental
Protection Agency, Drinking Water Research Division. July 1985.
Miltner, R.J., and C.A. Fronk. 1985b. Treatment of synthetic organic contami-
nants for Phase II regulations. Internal report. U.S. Environmental
Protection. Agency, Drinking Water Research Division. December 1985.
Monsanto Company. (date not available). The soil dissipation of glyphosate,
alachlor, atrazine and simazine herbicides. Unpublished study.
Murnik, M.R., and C.L. Nash. 1977. Mutagenici^y of the triazine herbicides
atrazine, cyanazine and simazine in Drosophila melanogaster. J. Toxicol.
Environ. Health. 3:691-697.
NAS. 1977. National Academy of Sciences. Safe Drinking Water Committee.
Drinking water and health. Part l, Chap. 1-5. Washington, D.C.:
National Academy Press, pp. V-184-V-348.
Newell, G.W., and J.C. Dilley. 1978. Teratology and acute toxicology of
selected chemical pesticides administered by inhalation. EPA-600/1-78-003,
U.S. Environmental Protection Agency, Washington, DC.
Palmer, J.S., and R.D. Radeleff. 1964. The toxicologic effects of certain
fungicides and herbicides on sheep and cattle. Ann. N.Y. Acad. Sci.
111:729-736.
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Simazine August, 1987
-18-
Palmer, J.S., and R.D. Radeleff. 1969. The toxicity of some organic herbicides
to cattle, sheep, and chickens. Production Research Report No. 1066,
U.S. Department of Agriculture, Agricultural Research Service. 1-26.
Pliss, G.B., and M.A. Zabezhinsky. 1977. Carcinogenicity of symmetric
triazine derivatives. Pest. Abstr. 5:72-1017.
Rees, G.A.V., and L. Au. 1979. Use of XAD-2 macroreticular resin for the
recovery of ambient trace levels of pesticides and industrial organic
pollutants from water. Bull. Environ. Con tarn. Toxicol. 22:561-566.
Rodgers, E.G. 1968. Leaching of seven s-triazines. Weed Sci. 16:117-120.
Sheets, T.J. 1959. The uptake, distribution, and phytotoxicity of 2-chloro-
4,6-bis(ethylamine)-s-triazine. Ph.D. Thesis. University of California.
Cited by Turner, M.A., and R.S. Adams, Jr. 1968. The adsorption of
atrazine and atratone by anion- and cation-exchange resins. Soil Sci.
Soc. Amer. Proc. 32:62-63.
Simmons, V.F., D.C. Poole, E.S. Riccio, D.E. Robinson, A.D. Mitchell and
M.D. Waters. 1979. In vitro mutagenicity and genotoxicity assays of
38 pesticides. Environ. Mutagen. 1:142-143.
Smith, A.E. , R. Grover, G.S. Emmond and H.C. Korven. 1975. Persistence and
movement of atrazine, bromacil, monuron, and simazine in intermittently-
filled irrigation ditches. Can. J. Plant Sci. 55:809-816.
Tai, C.N. , C. Breckenridge and J.D. Green.* 1985a. Simazine technical subacute
oral 13-week toxicity study in rats. Ciba-Geigy Pharmaceuticals Division.
Report No. 85018, Ace. No. 257693.
Tai, C.N. , C. Breckenridge and J.D. Green.* 1985b. Simazine technical subacute
oral 1 3 -week toxicity study in dogs. Ciba-Geigy Pharmaceuticals Division.
Report No. 85022, Ace. No. 257692.
Talbert, R.E., and O.H. Fletchall. 1965. The adsorption of some s-triazines
in soils. Weeds 13:46-52.
Turner, M.A. , and R.S. Adams, Jr. 1968. The adsorption of atrazine and
atratone by anion- and cation-exchange resins. Soil Sci. Amer. Proc.
32:62-63.
USDA. 1984. United States Department of Agriculture. Forest Service.
Pesticide background statements. Vol. 1. Herbicides.
U.S. EPA. 1983. U.S. Environmental Proter-ion Agency. Simazine registration
standard. Office of Pesticide Programs, Washington, DC. November 7.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185)33992-34003. September 24.
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Simazine August, 1987
-19-
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Code of Federal
Regulations. Protection of the environment. Tolerances and exemptions
from tolerances for pesticide chemicals in or on raw agricultural
commodities. 40 CFR 180.213.
U.S. EPA. 1986c. U.S. Environmental Protection Agency. Method #1 -
Determination of nitrogen- and phosphorus-containing pesticides in ground
water by GC/NPD, January, 1986. Draft. Cincinnati, OH: Environmental
Monitoring and Support Laboratory.
U.S. FDA. 1979. U.S. Food and Drug Administration. Code of Federal Regula-
tions. 21 CFR 193.400. April 1.
Valencia, R. 1981. Mutagenesis screening of pesticides using Drosophila.
Project summary. Research Triangle Park, NC: Health Effects Research
Laboratory, U.S. Environmental Protection Agency. EPA-600/S1-81-017.
Walker, A. 1976. Simulation of herbicide persistence in soil. Pestic. Sci.
7:41-49.
Waters, M.D., S.S. Saindhu, Z.S. Simmon et al. 1982. Study of pesticide
genotoxicity. Basic Life Sciences 21:275-326.
Weed Science Society of America. 1983. Herbicide handbook. 5th ed.
Champaign, IL: Weed Science Society of America, p. 433-437.
Windhloz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck index — an encyclopedia of chemicals and drugs, 10th ed.
Rahway, NJ: Merck and Company, Inc.
Woodard Research Corporation.* 1965. Three-generation reproduction study of
simazine in the diet of rats. MRID 00023365, 00080631
Yelizarov, G.P. 1977. Occupational skin diseases caused by simazine and
propazine. Pest. Abstr. 6:73-0352.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
TEBUTHIURON
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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Tebuthiuron August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 3401 4-1 8-1
Structural Formula
- 9
(CH.j.CHCH.SN-C-N
-«, SCH-
N-[ 5- (1,1 -Dimethyl ethyl )-1 , 3,4-thiadiazol-2-yl]-N,N'-dimethylurea
Synonyms
0 Brulan; Per flan; Prefunid; Spike; Trebulan; Turolan.
Uses
0 Herbicide for total vegetation woody plant control in noncropland
areas and for brush and weed control in rangeland (Meister, 1983).
Properties (Meister, 1983)
Chemical Formula C9H16ON4S
Molecular Weight 228 (calculated)
Physical State (25°C) White crystalline, odorless powder;
colorless solid
Boiling Point
Melting Point 159 to 161°C
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25°C, pH 7) 2,500 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold —
Odor Threshold —
Conversion Factor —
Occurrence
0 Tebuthiuron has been detected in groundwater in Texas over a 4 month
period at levies between 10 to 300 ppb (STORET, 1987).
Environmental Fate
0 Tebuthiuron is resistant to hydrolysis. ^C-Tebuthiuron, at 10
and 100 ppm, did not degrade during 64 days of incubation in sterile
aqueous solutions at pH 3, 6 and 9 in the dark at 25 °C (Nosier and
Saunders, 1976).
0 After 23 days of irradiation with artificial light (20-W black light),
tebuthiuron accounted for 87 to 89% of the applied radioactivity in
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Tebuthiuron August, 1987
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deionized (pH 7.1) and natural (pH 8.1) water treated with thiadiazole
ring-labeled 14c-tebuthiuron at 25 ppm (Blanco Products Company, 1972;
Rainey and Magnussen, 1976b). After 15 days of irradiation with a
black light or a sunlamp, tebuthiuron accounted for approximately
82 and 53%, respectively, of the applied compound in natural water
treated with 14C-tebuthiuron at 2.5 ppm.
Thiadiazole ring-labeled 14c-tebuthiuron in loam soil degraded from
8 ppm immediately post-treatment to 5.7 ppm at 273 days posttreatment
indicating a half-life greater than 273 days (Rainey and Magnussen,
1976a, 1978).
14c-Tebuthiuron, at 1.0 ppm, degraded with a half-life of greater
than 48 weeks in a loam soil maintained under anaerobic conditions in
the dark at 23°C (Berard, 1977). N-[5-(1,1-Dimethylethyl)-1,3,4-
thiadiazol-2-yl]-N-methylurea was the major degradate.
Ring-labeled 14C-tebuthiuron was very mobile (>94% of that applied
was found the leachate) in a 12-inch column of Lakeland fine sand
soil leached with 20 inches of water (Holzer et al., 1972). It was
mobile in columns of loamy sand (approximately 73% at 6 to 10 inches),
loam (approximately 84% at 1 to 8 inches) and muck (100% at 0 to 4
inches) soils leached with 4 to 8 inches of water.
Based on column leaching studies, tebuthiuron is mobile to very mobile
in loam, loamy sand, and Lakeland sand soils and has low mobility in
silty loam soil (Day, 1976a).
14C-Tebuthiuron residues aged 30 days were mobile in a column of
sandy loam soil; 39% of 14c-residues were found in the soil and 40%
of He-residues were in the leachate (Day, 1976b).
14c-Tebuthiuron degraded with half-lives of greater than 33 months
in field plots in California (loam soil), 12 to 15 months in Louisiana
(clay soil), and 12 to 15 months in Indiana (loam soil). The three
sites were treated with thiadiazole ring-labeled 14C-tebuthiuron at
8.96, 2.24 and 8.96 kg/ha, respectively (Rainey and Magnussen, 1976a,
1978). N-[5-(1,1-Dimethylethyl)-1,3,4-thiadiazol-2-yl]-N-methylurea
was the major degradate at all three sites. Radioactivity was detected
in the 15- to 30-cm depth of soil (10.2% of the applied compound at
18 months) at the California site, in the 30- to 45-cm depth of soil
(1.3% of the applied compound at 33 months) at the Louisiana site,
and in the 30- to 45-cm depth of soil (4.7% of the applied compound
at 15 months) at the Indiana site. 14c-Tebuthiuron residues did not
appear to accumulate in silt loam soil in Louisiana after three
applications of 14c-tebuthiuron (0.84 kg/ha at zero time; 1.4 kg/ha at
22 and 73 weeks).
III. PHARMACOKINETICS
Absorption
0 Morton and Hoffman (1976) reported that 94 to 96% of a single oral
dose of tebuthiuron (10 mg/kg) was excreted in the urine of rats,
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Tebuthiuron August, 1987
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rabbits and dogs. In mice, 66% was excreted in the urine, and 30% in
the feces. These data indicate that tebuthiuron was well absorbed
(about 70 to 96%) from the gastrointestinal tract.
Distribution
0 No quantitative data were found in the available literature on the
tissue distribution of tebuthiuron in exposed animals.
0 Adams et al. (1982) administered tebuthiuron in the diet to 20
pregnant Wistar rats at levels of 100 or 200 ppm for 6 days prior
to delivery. Forty-eight hours after delivery, radiolabeled tebu-
thiuron was reintroduced into the diet at the same levels as before.
Radioactive label was detected in the milk at levels of 2.7 and
6.2 ppm for the 100- and 200-ppn groups, respectively.
Metabolism
0 Morton and Hoffman (1976) reported that tebuthiuron was metabolized
extensively by mice, rats, rabbits and dogs. Tebuthiuron was
administered by gavage to male and female ICR mice, Harlan rats,
Dutch-Belted rabbits and beagle dogs at a dose of 10 mg/hg. Examin-
ation of urine extracts by thin-layer chromatography (TLC) showed the
presence of eight radioactively labeled metabolites in rat, rabbit
and dog urine and seven in mouse urine. Small amounts of unchanged
tebuthiuron also were detected in each case (except for the mouse).
The major metabolites were formed by N-demethylation of the substituted
urea side-chain in each species examined. Oxidation of the dimethylethyl
group also occurred in all species examined.
Excretion
Morton and Hoffman (1976) reported that tebuthiuron was excreted
rapidly in the urine of several species. Radiolabeled tebuthiuron
was administered to male and female ICR mice, Harlan rats, Dutch-
Belted rabbits and beagle dogs at a dose of 10 rag/kg by gavage.
Elimination of radioactivity was virtually complete within 72 hours
and recovery values at 96 hours were 96.3, 94.5, 94.3 and 95.7% in
the mouse, rat, rabbit and dcg, respectively. In the rats, rabbits
and dogs, the radioactivity was excreted almost exclusively in the
urine. In the mice, 30% of the radioactivity was excreted in the
feces.
IV. HEALTH EFFECTS
Humans
0 No information on the health effects of tebuthiuron in humans was
found in the available literature.
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Tebuthiuron August, 1987
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Animals
Short-term Exposure
0 Todd et al. (1974) reported the acute oral LDso values of tebuthiuron
in rats, mice and rabbits to be 644, 579 and 286 mg/kg, respectively.
In cats, oral doses of 200 mg/kg were not lethal, while 500 mg/kg
given orally was not lethal to dogs, quail, ducks or chickens.
0 An acute oral LD50 value of 1.9 g/kg for males and 2.1 g/kg for female
CD rats were reported by Choie and Katz (1984a).
0 Todd et al. (1972a) supplied Sprague-Dawley rats (age, sex and number
not specified) with food containing tebuthiuron (purity not stated)
at levels of 2,500 ppm for 15 days. Based on the dietary assumptions
of Lehman (1959), 1 ppm in the diet of a rat is equivalent to 0.05
mg/kg/day; therefore, this level corresponds to 125 mg/kg/day. The
animals were observed for an additional 15-day recovery period. All
the animals exhibited reduced body weight gain during the treatment
period. Light and electron microscopic evaluation revealed formation
of vacuoles containing electron-dense bodies and myeloid figures in
pancreatic acinar cells. This condition was rapidly reversed during
the recovery period.
Dermal/Ocular Effects
0 An acute dermal toxicity study using New Zealand White rabbits showed
a dermal LDcn value of greater than 20.0 g/kg for tebuthiuron (Choie
and Katz, 1984b).
0 Todd et al. (1974) administered 200 mg/kg tebuthiuron to the shaved,
abraded backs of male and female New Zealand White rabbits. During
the study, one rabbit died following development of diarrhea and
emaciation. All surviving rabbits gained weight over the 14-day
observation period and were without signs of dermal irritation.
0 Wolfe et al. (1982a,b) reported that tebuthiuron produced no erythema,
edema or other dermal effects when administered topically to the
intact dorsal or abdominal skin of male and female rabbits at a level
of 2,000 mg/kg.
0 Todd et al. (1'.-74) tested tebuthiuron for sensitization in 2- to
3-month-old female albino guinea pigs. Each animal received topical
applications of 0.1 mL of an ethanolic solution containing 2% tebu-
thiuron to the region of the flank three times per week for 3 weeks.
Ten days after the last of the nine treatments, a challenge application
was made, followed by a second challenge 15 days after the first.
Tebuthiuron induced no dermal or systemic responses indicative of
contact sensitization.
0 Todd et al. (1974) instilled 0.1 mL (71 mg) of tebuthiuron into one
eye and conjunctival sac of each of six New Zealand White rabbits (2-
to 3-months old). No irritation of the cornea or iris was observed.
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Tebuthiuron August, 1987
-6-
but there was slight transient hyperemia of the conjunctiva.
All eyes were normal by the end of the 7-day test period.
Long-term Exposure
0 Todd et al. (1972b) administered tebuthiuron (purity not stated) in
the diet to groups of male and female Harlan rats (10/sex/group, 28-
to 35-days old, 74 to 156 g) at levels of 0, 40, 100 or 250 mg/Jq/day
for 3 months. Body weights and food consumption were measured weekly.
Blood obtained prior to necropsy was evaluated for blood sugar, blood
urea nitrogen (BUN) and serum glutamic-pyruvic transaminase (SGPT).
Sections of organs and tissues were prepared for gross and microscopic
evaluation. There were no clinical signs of toxicity or mortality in
any of the groups. A moderate reduction in body weight gain and a
decrease in efficiency of food utilization in males and females in
the highest dose group (250 mg/kg/day) was evident from week 1 of the
study. Tebuthiuron had no clinically important effects on any of the
hematological or clinical-chemistry parameters measured. All rats
receiving 250 mg/Jq/day tebuthiuron showed diffuse vacuolation of
the pancreatic acinar cells. The degree of this change ranged from
moderate to severe, but the effect was not associated with necrosis
or with the presence of an inflammatory response. One rat receiving
100 rag/kg/day tebuthiuron showed similar but very slight pancreatic
changes. Based on these results, a No-Observed-Adverse-Effect-Level
(NOAEL) of 40 mg/hg/day and a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 100 mg/Jq/day were identified.
0 Todd et al. (1972c) administered tebuthiuron (purity not stated) in
gelatin capsules to groups of four beagle dogs (two/sex/group, 13- to
23-months old, 7 to 23 kg) at dose levels of 0, 12.5, 25 or 50 mg/Jg/day
for 3 months. The physical condition of the animals was assessed
daily, and body weights were recorded weekly. Gross and microscopic
histopathology examinations were performed. Anorexia was noted,
especially in the high-dose animals, leading to some weight loss.
There was no mortality. Behavior and appearance were unremarkable at
all test levels. No abnormalities were seen in the hematological or
urinalysis studies. Clinical chemistry findings indicated increased
BUN in the 50-mg/kg females. In addition, this group and the 50-mg/kg
males exhibited increasing levels of alkaline phosphatase, up to
four-fold over those of controls; however, these levels had returned to
normal at the terminal sampling. There were no urinary abnormalities.
The 25-mg/kg females and males demonstrated increased thyroid-to-body
weight ratios, and the 50-mg/Jg females also showed increased spleen-
to-body weight ratios. Histopathological findings were unremarkable.
The LOAEL was identified as 12.5 rag/kg, based on increased thyroid-to-
body weight ratios, increased alkaline phosphatase values and increased
BUN levels in test animals.
0 Todd et al. (1976a) administered tebuthiuron (purity not stated)
in the diet to groups of Harlan rats (40/sex/dose) for 2 years at
dietary levels of 0, 400, 800 or 1,600 ppm. Based on the dietary
assumptions of Lehman (1959), 1 ppm in the diet of a rat is equivalent
to 0.05 mg/kg/day; therefore, these doses correspond to 20, 40 or
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Tebuthiuron Au<3ust' 1987
-7-
80 mg/kg/day. Physical appearance, behavior, food intate, body
weight gain and mortality were recorded. Hematologic and blood
chemistry values were obtained throughout the study; urinalysis was
also performed. At necropsy, organ weights were determined and
organs and tissues were examined grossly and histologically. Mortality
in exposed animals was similar to, or less than, that observed in the
control group. Variations in hematology, blood chemistry and urinalysis
data from all groups were slight and unrelated to the test compound.
Reduced body weight gain (10% or greater) was observed in the highest
dose group animals. There was also a slight increase in the kidney
weights of the high-dose males. Microscopic examination revealed a
low incidence of slight vacuolation of the pancreatic acinar cells in
animals in the highest dose group. The NOAEL for this study, based
on acinar vacuolation, was 40 mg/kg.
0 Todd et al. (1976b) administered tebuthiuron (purity not stated) in
the diet for 2 years to groups of Harlan ICR mice (40/sex/dose) at
levels of 0, 400, 800 or 1,600 ppm. Based on the dietary assumptions
of Lehman (1959), 1 ppjn in the diet of a mouse is equivalent to 0.150
mg/hg/day; therefore, these dietary levels correspond to approximately
60, 120 or 240 ing/kg/day. Physical appearance, behavior, appetite,
body weight gain and mortality were recorded. Hematologic, blood
chemistry and organ weight values were obtained for animals surviving
the test period. Gross and microscopic evaluations were conducted on
organs and tissues obtained at necropsy. No important differences
were observed between treated and control groups for any of the
parameters evaluated. The vacuolation of pancreatic acinar cells
noted in the Todd (1976a) rat studies was not evident in this study
in mice. Based on this, the NOAEL for this study was identified as
240 mg/kg/day.
0 In a 2-year rat feeding study by Jessup et al. (1980), Charles River
CD rats received technical tebuthiuron at levels of 0, 2, 300 or
3,000 ppn (0, 0.08, 12 or 120 mg/kg, based on Lehman, 1959). Decreased
body weight gain for both males and females and an increased incidence
of focal cytomegaly in livers of female rats were seen at 120 mg/kg.
Due to incomplete hematological and clinical chemistry data, a NOAEL
for these parameters could not be determined (U.S. EPA, 1986b).
Reproductive Effects
0 Hoyt et al. (1981) studied the effects of tebuthiuron (98% active
ingredient) in a two-generation reproduction study in rats. Weanling
Wistar rats (25/sex/dose, FQ generation) were maintained on diets
containing tebuthiuron at 0, 100, 200, and 400 ppm based on the
active ingredient (0, 5, 10 or 20 mg/kg/day, based on Lehman, 1959)
for a period of 101 days preceding two breeding trials. First gene-
ration (FI ) offspring were maintained on the same diets for a period
of 124 days preceding two breeding trials. Spermatogenesis and sperm
morphology were examined in 10 F0 males per treatment group. In
addition, representative Fia and F2a weanlings and F-| adults were
necropsied and given histopathologic examinations after live-phase
observations were completed. No changes in the efficiency of food
utilization (EFU) were noted during the F0 growth period, but during
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Tebuthiuron August, 1987
-8-
the F1 growth period, a statistically significant (p £0.05) depression
in cumulative (124 days) EFU values occurred in both male and female
rats receiving 20 mg/kg/day. EFU was not affected at the other dose
levels. A dose-related depression in mean body weight occurred among
female rats of the FI generation receiving 10 or 20 mg/kg/day; mean
body weight was depressed significantly (p £0.05) only in the high-
dose females. In the 5-mg/kg/day group, body weights of either sex
were not affected. The reproductive capacity of the animals was not
affected at any level; no dose-related conditions or lesions were
found in any offspring. In adult males from the FQ generation, no
dose-related histologic lesions were found, and sperm morphology and
spermatogenesis were normal. A LOAEL of 10 mg/kg/day was determined
for a lower rate of body weight gain during the 101-day pre-mating
period in F^ females, and a NOAEL of 5 mg/kg/day, the lowest dose
tested, was identified.
Developmental Effects
0 Todd et al. (1972d) administered tebuthiuron (purity not stated) in
the diet to groups of 25 adult Wistar-derived female rats (245 to
454 g) at levels of 0, 600, 1,200 or 1,800 ppm based on the active
ingredient (0, 30, 60 or 90 mg/kg/day, based on Lehman, 1959) on days
6 to 15 of gestation. Fetal and uterine parameters were normal and
the fetal defects that occurred were not attributed to the test
compound. The NOAEL for developmental effects was greater than
1,800 mg/kg/day, the highest dose tested.
0 Todd et al. (1975) administered tebuthiuron (purity not stated) by
gavage to groups of 15 adult female Dutch-Belted rabbits at levels of
10 or 25 mg/kg/day on days 6 to 18 of gestation. No developmental or
toxic effects were observed.
0 Teratology studies were conducted in New Zealand White rabbits (Infurna
and Arthur, 1985) and in rats (Infurna et al., 1985). Tebuthiuron was
not found to be teratogenic to either species (U.S. EPA, 1986b). At
a dose level of 50 mg/kg/day in rabbits, there was decreased food
consumption, an increase in the food efficiency index, decreased body
weight gain and stool changes. In rats at 500 mg/kg/day, there was
increased mortality and salivation, urine staining, bloody discharge
and weight loss. Based on these results, the NOAEL for maternal
toxicity was 10 mg/kg/day in the rabbit and 50 mg/kg/day in the rat.
The NOAEL for fetotoxicity in ijoth the rabbit and rat was 50 mg/kg/day,
based on reduced ossification of sternebrae in rabbits at 75 rag/kg,
and on reduced ossification and misalignment of the sternebrae,
reduced ossification of the metacarpals and distal phalanges of the
forepaws and reduced ossification of distal phalanges of the hindpaws
in rats at 500 mg/kg.
Mutagenicity
0 Hill (1984) reported that primary cultures of adult rat hepatocytes
incubated with concentrations of tebuthiuron ranging from 0.5 to
1,000 ug/mL did not induce unscheduled DNA synthesis.
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Tebuthiuron August, 1987
-9-
0 Rexroat (1984) reported that tebuthiuron did not induce Salmonella
revertants (strains TA1535, 1537, 1538, 98 and 100) when tested at
concentrations ranging between 100 and 5,000 ug/plate, with or without
metabolic activation. It was concluded that tebuthiuron was not
mutagenic in the Ames Salmonella/mammalian microsome test for bacterial
mutation.
0 Neal (1984) reported that tebuthiuron did not induce sister chromatid
exchange in vivo in bone marrow cells of Chinese hamsters administered
oral doses~of 200, 300, 400 or 500 mg/kg tebuthiuron.
0 Cline et al. (1978) reported that histadine auxotrophs of Salmonella
typhimunum (strains G46, TA1535, 100, 1537, 1538, 98, C3076 and
D3052) and tryptophan auxotrophs of Escherichia coli were not
reverted to the prototype by tebuthiuron at levels of 0.1 to 1,000
ug/mL, with or without metabolic activation.
Carcinogenicity
0 Todd et al. (1976a) administered tebuthiuron (purity not stated) in
the diet to groups of Harlan rats (40/sex/dose) at levels of 0, 400, 800
or 1,600 ppm based on the active ingredient (0, 20, 40 or 80 mg/kg/day,
based on Lehman, 1959) for 2 years. The authors reported no influence
of the test compound on the incidence of neoplasms at any dose level.
0 Todd et al. (1976b) administered tebuthiuron in the diet to groups
of Harlan ICR mice (40/sex/dose) at levels of 0, 400, 800 or 1,600
ppm (0, 60, 120 or 240 mg/kg/day, based on Lehman, 1959) for 2 years.
The authors reported no statistical evidence of increased incidence
of tumors at any dose level.
0 In the 2-year feeding study reported by Jessup et al. (1980), Charles
River CD rats received technical tebuthiuron in the diet at levels
of 0, 2, 300 or 3000 ppm (0, 0.028, 12 or 120 mg/kg/day, based on
Lehman, 1959). This study demonstrated that at a dose level of
120 mg/kg/day, the highest dose tested, in female rats, there was a
statistically significant increase in combined mammary tumors (adenomas,
fibroadenomas and adenocarcinomas) and in combined hepatocellular
adenomas and carcinomas. At this level in male rats, there was a
statistically significant increase in combined thyroid follicular
adenomas and carcinomas and in testicular interstitial cell adenomas
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)
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Tebuthiuron August, 1987
-10-
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in ing/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).
One-day Health Advisory
No information was found in the available literature that was suitable
for the determination of the One-day HA value for tebuthiuron. It is therefore
recommended that the Ten-day value for a 10-kg child, 1 mg/L (1,000 ug/L,
calculated below), be used at this time as a conservative estimate of the
One-day HA value.
ten-day Health Advisory
The study by Infurna and Arthur (1985) has been selected to serve as the
basis for the Ten-day HA value for tebuthiuron because the maternal toxicity
in the New Zealand White rabbit was the most sensitive end point observed in
a short-term study. This study identified a NOAEL of 10 mg/kg/day for maternal
toxicity; however, at higher levels, tebuthiuron was shown to be fetotoxic.
Using a NOAEL of 10 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (1° mg/kg/day) (10 kg) = , mg/L ( , f 000 ug/L )
(100) (1 L/day)
where:
10 mg/kg/day = NOAEL, based on maternal toxicity in New Zealand White
rabbits exposed to tebuthiuron by diet.
1 0 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 Advisories
The subchronic (90-day) feeding study in beagle dogs reported by Todd
et al. (1972c) has been selected to serve as the basis for the Longer-term HA
values for tebuthiuron. The study identified a dose-response relationship
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Tebuthiuron August, 1987
-1 1-
and a LOAEL for female dogs administered tebuthiuron in gelatin capsules at
dose levels of 0, 12.5, 25 or 50 mg/kg/day for 3 months. There was an increased
BUN in the 50-mg/kg females and a four-fold increase in alkaline phosphatase.
The males also had a four-fold increase in alkaline phosphatase. Both the
males and females demonstrated increased thyroid-to-body weight ratios. Based
on these results, the LOAEL was 12.5 mg/kg/day, the lowest dose tested. The
two-generation reproduction study by Hoyt et al. (1981) was not selected,
even though an apparent LOAEL of 10 mg/kg/day was identified. This LOAEL was
based on a slight decrease in weight gain in exposed females, along with a
decrease in EFU values. This value was rejected because it is not clear that
the effects are biologically significant, and because no effects on weight
gain or EFU were seen at comparable dose levels in subchronic feeding studies
in rats and dogs (Todd et al., 1972b,c) or in chronic studies in rats and
mice (Todd et al., 1976a,b).
Using a LOAEL of 12.5 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (12.5 mg/kg/day) (10 kg) = 0.125 /L (125 /L,
(1,000) (1 L/day)
where:
12.5 mg/kg/day = LOAEL, based on a four-fold increase in alkaline
phosphatase levels, increased BUN levels and increased
thyroid-to-body weight ratios in dogs exposed to
tebuthiuron in the diet for 3 months.
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.
The Longer-term HA for the 70-kg adult is calculated as follows:
Longer-term HA = (12.5 mg/kg/day) (70 kg) = 0.438 mg/L (438 ug/L}
(1,000) (2 L/day)
where:
12.5 mg/kg/day = LOAEL, based on a four-fold increase in alkaline
phosphatase levels, increased BUN levels and increased
thyroid-to-body weight ratios in dogs exposed to
tebuthiuron in the diet for 3 months.
70 kg = assumed body weight of an adult.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
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Tebuthiuron August, 1987
-12-
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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The two-generation reproduction study in rats (Hoyt et al., 1981) has been
selected to serve as the basis for the Lifetime HA value for tebuthiuron. In
this study, four groups of Wistar rats (25/sex) were fed tebuthiuron at 0, 5,
10 or 20 mg/kg/day in the diet for 101 days (F0 rats) or 121 days (FI rats)
and then for a further period sufficient to mate, deliver and rear two
successive litters of young to 21 days of age (i.e., the test diet was fed
throughout mating, gestation and lactation). The Fla rats were parents of
the F2 offspring. No adverse effects were reported in this study except for
a lower rate of body weight gain during the premating period in F-\ females at
dietary levels of 10 and 20 mg/kg. The NOAEL was identified as 5 mg/kg/day.
The chronic study by Todd et al. (1976b) in mice was not selected because the
weight loss and vacuolation of pancreatic acinar cells noted in rats was not
observed in mice even at dose levels as high as 160 mg/kg/day, indicating
that the mouse is less sensitive than the rat.
Using the NOAEL of 5 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (5 mg/kq/day) = 0.05 mg/kg/day
(100)
where:
5 mg/kg/day = NOAEL, based on effects on the rate of weight gain in
rats exposed to tebuthiuron in the diet for 101 days.
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Tebuthiuron August, 1987
-13-
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 = (0.05 mg/kg/day) (70 kg) =1.75 mg/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*75 mg/L > (20%) = Q.035 mg/L (35 ug/L)
(10)
where:
1.75 mg/L = DWEL.
20% = assumed relative source contribution from water.
10 = additional uncertainty factor per ODW policy to account
for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of tebuthiuron.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), tebuthiuron may be classified in
Group C: possible human carcinogen. This category is for substances
that show limited evidence of carcinogenicity in animals in the
absence of human data.
0 In the 2-year chronic oral toxicity study in rats by Jessup (1980),
the 120-mg/kg/day dose level induced statistically significant increases
in combined hepatocellular adenomas and carcinomas, mammary adenomas
and carcinomas in female rats, and in thyroid adenomas and carcinomas
in males. Testicular adenomas were also increased.
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Tebuthiuron August, 1987
-14-
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 No other criteria, guidance or standards were found in the available
literature.
VII. ANALYTICAL METHODS
0 Analysis of tebuthiuron is by a gas chromatographic (GC) method
applicable to the determination of certain nitrogen-phosphorus-
containing pesticides in water samples (U.S. EPA, 1986c). In this
method, approximately 1 liter of sample is extracted with methylene
chloride. The extract is concentrated and the compounds are separated
using capillary column GC. Measurement is made using a nitrogen
phosphorus detector. The method detection limit has not been deter-
mined for tebuthiuron but it is estimated that the detection limits
for analytes included in this method are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information on treatment technologies capable of effectively
removing tebuthiuron from contaminated water was found in the available
literature.
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Tebuthiuron August, 1987
-15-
IX. REFERENCES
Adams, E., J. Magnussen, J. Emmerson et al.* 1982. Radiocarbon levels in the
milk of lactating rats given !4c-tebuthiuron (compound 75503) in the diet.
Eli Lilly and Company, Greenfield, IN. Unpublished study. MRID 00106081.
Berard, D.F.* 1977. 14C-Tebuthiuron degradation study in anaerobic soil.
Prepared and submitted by Eli Lilly and Co., Greenfield, IN.
MRID 00900098.
Choie, D., and R. Katz.* 1984a. Acute oral toxicity study in rats: Toxicology/
Pathology Report No. 137-84 on Igran 80 WDG. Unpublished study prepared
by Ciba-Giegy Corp., Greensboro, NC. MRID 00146725.
Choie, D., and R. Katz.* 1984b. Acute dermal toxicity study in New Zealand
White rabbits on Igran 80 WDG. Toxicology/Pathology report No. 137-84.
Unpublished study prepared by Ciba-Geigy Corp., Greensboro, NC. MRID 00146726.
Cline, J.C., G.Z. Thompson and R.I. McMahon.* 1978. The.effect of Lilly Com-
pound 75503 (tebuthiuron) upon bacterial systems known to detect mutagenic
events. Eli Lilly and Company, Greenfield, IN. Unpublished study.
MRID 000416090.
Day, G.W.* 1976a. Laboratory soil leaching studies with tebuthiuron. Unpublished
study received Feb. 18, 1977 under 1471-109; submitted by Blanco Products
Co., Div. of Eli Lilly and Co., Indianapolis, IN. CDL:095854-1.
MRID 00020782.
Day, G.W.* 1976b. Aged soil leaching study with herbicide tebuthiuron. Unpub-
lished study received Feb. 18, 1977 under 1471-109; submitted by Blanco
Products Co., Div. of Eli Lilly and Co., Indianapolis, IN. CDL:095854-J.
MRID 00020783.
Blanco Products Company.* 1972. Environmental safety studies with EL-103.
Unpublished study received Mar. 13, 1973 under 1471-97; prepared in
cooperation with United States Testing Co., Inc. CDL:120339-1.
MRID 00020730.
Hill, L.* 1984. The effect of tebuthiuron (Lilly Compound 75503) on the
induction of DNA repair synthesis in primary cultures of adult rat
hepatocytes. Eli Lilly and Company., Greenfield, IN. Unpublished study.
MRID 00141692.
Holzer, F.J., R.F. Siek, R.L. Large et al.* 1972. EL-103: Leaching study.
Unpublished study received Mar. 13, 1973 under 1471-97 and prepared in
cooperation with Purdue Univ., Agronomy Dept., and United States Testing
Co., Inc., and submitted by Elanco Products Co., Division of Eli Lilly
and Co., Indianapolis, IN. CDL:120339-K. MRID 00020732.
Hoyt, J.A., E.R. Adams and N.V. Owens.* 1981. A two-generation reproductive
study with tebuthiuron in the Wistar rat. Eli Lilly and Company, Green-
field, IN. Unpublished study. MRID 00090108.
-------
Tebuthiuron August, 1987
-16-
Xnfurna, R., and A. Arthur.* 1985. A teratology study in New Zealand White
rabbits: (MIN 842105): Report 85010. Unpublished study prepared by
Ciba-Geigy Corp., Greensboro, NC. MRID 00152763.
Infurna, R., K. Wimbert, J. Mainiero et al.* 1985. Terbutryn Technical:
A teratology study in rats: (MIN 842292): Report 85111. Unpublished
study prepared by Ciba-Geigy Corp..Greensboro, NC. MRID 00152764.
Jessup, D.C., G. Gunderson and J.F. Ferrell.* 1980. 2-Year chronic oral
toxicity study in rats: IRDC No. 382-011. Unpublished study by Interna-
tional Research and Development Corporation in cooperation with Experimental
Pathology Laboratories, Inc., submitted by Ciba-Geigy Corp., Greensboro,
NC. MRID 00035923.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs, and
cosmetics, Assoc. Food Drug Off. U.S., Q. Bull.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Morton, D.M., and D.G. Hoffman. 1976. Metabolism of a new herbicide, tebu-
thiuron (1-(5-(1,1-dimethylethyl)-1, 3,5-thiadiazol-2-yl)-1,3-dimethylurea),
in mouse, rat, rabbit, dog, duck and fish. J. Toxicol. Environ. Health.
1:757-768.
Nosier, J.W., and D.G. Saunders.* 1976. A hydrolysis study on the herbicide
tebuthiuron. Includes undated method. Unpublished study received
Feb. 18, 1977 under 1471-109; submitted by Blanco Products Co., Div. of
Eli Lilly and Co., Indianapolis, IN. CDL:095854-F. MRID 00020779.
Neal, S.B.* 1984. The effect of tebuthiuron (Lilly Compound 75503) on the
in vivo induction of sister chromatid exchange in bone marrow of Chinese
hamsters. Eli Lilly and Company, Greenfield, IN. Unpublished study.
MRID 00141693.
Rainey, D.P., and J.D. Magnussen.* 1976a. Behavior of 14c-tebuthiuron in
soil. Unpublished study received Feb. 18, 1977 under 1471-109; prepared
in cooperation with A & L Agricultural Laboratories and United States
Testing Co., Inc., and submitted by Elanco Products Co., Div. of Eli
Lilly and Co., Indianapolis, IN. CDL: 095854-C. MRID 00020777.
Rainey, D.P., and J.D. Magnussen.* 1376b. Photochemical degradation studies
with 14c-tebuthiuron. Unpublished study received Feb. 18, 1977 under
1471-109; submitted by Elanco Products Co., Div. of Eli Lilly and Co.,
Indianapolis, IN. CDL:095854-D. MRID 00020778.
Rainey, D.P., and J.D. Magnussen.* 1978. Behavior of 14c-tebuthiuron in
soil: Addendum report. Unpublished study received June 1, 1978 under
1471-109; submitted by Elanco Products Co., Div. of Eli Lilly and Co.,
Indianapolis, IN. CDL:097100-C. MRID 00020693.
Rexroat, M.* 1984. The effect of tebuthiuron (Lilly Compound 75503) on the
induction of reverse mutations in Salmonella typhimurium using the Ames
test. Eli Lilly and Company, Greenfield, IN. Unpublished study. MRID 00140691
-------
Tebuthiuron August, 1987
-17-
STORET. 1987.
Todd, G.E., W.J. Griffing, W.R. Gibson et al.* 1972a. Special subacute rat
toxicity study. Eli Lilly and Company, Greenfield, IN. Unpublished study.
MRID 00020798.
Todd, G.C., W.R. Gibson and G.F. Kiplinger.* 1972b. The toxicological
evaluation of EL-103 in rats for 3 months. Unpublished study.
MRID 00020662.
Todd, G.C., W.R. Gibson and G.F. Kiplinger.* 1972c. The toxicological
evaluation of EL-103 in dogs for 3 months. Unpublished study.
MRID 00020663.
Todd, G.C., J.K. Markham, E.R. Adams et al.* 1972d. Rat teratology study
with EL-103. Unpublished study. MRID 00020803.
Todd, G.C., W.R. Gibson and C.C. Kehr. 1974. Oral toxicity of tebuthiuron
(1-(5-tert-butyl-1,3,4-thiadiazol-2-yl)-1,3-dimethylurea) in experimental
animals. Food Cosmet. Toxicol. 12:461-470.
Todd, G.C., J.K. Markham, E.R. Adams, N.V. Owens, F.O. Gossett and D.M. Morton.*
1975. A teratology study with EL-103 in the rabbit. Eli Lilly and
Company, Greenfield, IN. Unpublished study. MRID 00020644.
Todd, G.C., W.R. Gibson, D.G. Hoffman, S.S. Young and D.M. Morton.* 1976a.
The toxicological evaluation of tebuthiuron (EL-103) in rats for two
years. Eli Lilly and Company, Greenfield, IN. Unpublished study.
MRID 00020714.
Todd, G.C., W.R. Gibson, D.G. Hoffman, S.S. Young and D.M. Morton.* 1976b.
The toxicological evaluation of tebuthiuron (EL-103) in mice for two
years. Eli Lilly and Company, Greenfield, IN. Unpublished study.
MRID 00020717.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Registration Standard
for Tubutryn. Office of Pesticide Programs, Washington, DC.
l.S. EPA. 1986c. U.S. Environmental Protection Agency. U.S. EPA Method #1
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, January 1986 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
Wolfe, G., M. Johnson and J. Gargus.* 1982a. Acute dermal toxicity study in
rabbits: Sprakil SK-26. Hazelton Laboratories America, Inc., Vienna, VA.
Unpublished study. MRID 00120278.
Wolfe, G., M. Johnson and J. Gargus.* 1982b. Acute dermal toxicity study in
rabbits: Sprakil SK-13. Hazelton Laboratories America, Inc., Vienna, VA.
Unpublished study. MRID 00120287.
Confidential Business Information submitted to the Office of Pesticide
Programs
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August, 1987
OB AFT
TERBACIL
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
anv one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates chat are
derived can differ by several orders of magnitude.
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Terbacil August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 5902-51-2
Structural Formula
H
CHa^xN^O
;H3^xN^C
C.V*
5-Chloro-3-(1,1-dimethylethyl)-6-methyl-2,4(1H, 3H)-pyrimidinedione
Synonyms
0 Sinbar; Turbacil (Meister, 1983).
Uses
0 Herbicide used for the selective control of annual and perennial weeds
in crops such as sugarcane, alfalfa, apples, peaches, blueberries,
strawberries, citrus, pecans and mint (Meister, 1983).
Properties (Meister, 1983)
Chemical Formula • CgHi3O2N2Cl
Molecular Weight 216.65
Physical State (at 25°C) White crystals
Boiling Point (at 25 mm Hg)
Melting Point 175-177°C
Vapor Pressure (54°CJ 5.4 x 10-6 mm Hg
Specific Gravity —
Water Solubility (25°C) 710 mg/L
Log Octanol/Water Partition -1.41
Coefficient
Taste Threshold —
Odor Threshold --
Conversion Factor —
Occurrence
0 Terbacil was not sampled at any water supply stations listed in the
STORET database (STORET, 1987). No information was found in available
literature on the occurrence of terbacil.
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Terbacil Au*U8t' 1987
-3-
Environmental Fate
« 14c-Terbacil at 5 ppm was stable (less than 2% degraded) in buffered
aqueous solutions at pH 5, 7, and 9 for 6 weeks at,15°C in the dark
(Davidson et al. 1978).
0 After 4 weeks of irradiation with UV light (300 to 400 nm), about 16*
of the applied 14c-terbacil (5 ppm) was photodegraded in distilled
water (pH 6.2) (Davidson et al., 1978).
0 Soil metabolism studies indicate that terbacil is persistent in soil.
At 100 ppm, terbacil was slowly degraded in an aerobic sandy loam
soil (80% remained after 8 months) (Marsh and Davies, 1978). Terbacil
at 8 ppm had a half-life of about 5 months in aerobic loam soil
(Zimdahl et al.f 1970). 14c-Terbacil at 2 ppm had a half-life of
2 to 3 months in aerobic silt loam and sandy loam soils (Rhodes, 1975;
Gardiner, 1964; Gardiner et al., 1969). The formation of carbon
dioxide is slow; for example, 28% of the applied 14c-terbacil at 2.88
ppm on sandy loam soil degraded to carbon dioxide in 600 days (Wolf,
1973; Wolf, 1974; Wolf and Martin, 1974).
0 Degradation of terbacil in an anaerobic soil environment is also slow.
In anaerobic silt loam and sandy soils, 14c-terbacil at 2.1 ppm was
slightly degraded (less than 5% after 60 days) in the dark (Rhodes,
1975). Only trace amounts of 14c_terbacil, applied at 2.88 ppm, were
degraded to 14c-carbon dioxide after 145 days in an anaerobic environment
when metabolized by microbes in the dark (Rhodes, 1975). At least
90% of the lable remained as terbacil after 90 days of incubation in
both sterile and nonsterile soils. Small amounts (0.8 to 1.5% of the
label of carbon dioxide were evolved from nonsterile soil, whereas
0.01% was evolved from sterile soil (Rhodes, 1975).
0 Terbacil was mobile in soil columns of sandy loam and fine sandy soil
(Rhodes, 1975; Mansell et al., 1972). However, in a silt loam soil
column, only 0.4% of the applied 14c-terbacil leached with 20 inches
of water (Rhodes, 1975). In an aged soil column leaching study of
the leaching characteristics of degradates, about 52% and 4% of the
applied radioactivity in aged sandy loam and silt loam soils leached,
respectively (Rhodes, 1975). Terbacil phytotoxic residues were
mobile to depths of 27.5 to 30 cm in a sandy soil column treated with
terbacil at 5.6 kg/ha and eluted with 10 or 20 cm water (Marriage,
1977). Terbacil was negligibly adsorbed to soils ranging in texture
from sand to clay (Davidson et al., 1978; Liu et al., 1971; Rao and
Davison, 1979). Terbacil was adsorbed (54%) to a muck soil (36%
organic matter) (Liu et al., 1971).
0 Data from field dissipation studies showed that terbacil persistence
in soil varied with application rate, soil type and rainfall. In
the field, terbacil phytotoxic residues persisted in soil for up to
16 months following a single application of terbacil. Residues were
found at the maximum depths sampled (3 to 43 inches) (Gardiner, undated
a,b; Gardiner et al., 1969; Isom et al., 1969; Isom et al., 1970; Liu
et al., undated; Mansell et al., 1977; Mansell et al., 1979; Morrow
and McCarty, 1976; Rahman, 1977; Rhodes, 1975).
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Terbac'il August, 1987
-4-
0 Phy to toxic residues resulting from multiple applications of terbacil
persisted for 1 to more than 2 years following the final application
(Skroch et al., 1971; Tucker and Phillips, 1970; Benson, 1973;
Doughty, 1978).
0 Terbacil has not been found in ground water; however, its soil
persistence and mobility indicate that it has the potential to get
into ground water.
III. PHARMACOKINETICS
Absorption
0 No information was found in the available literature on the absorption
of terbacil.
Distribution
0 No information was found in the available literature on the distribution
of terbacil.
Metabolism
0 No information was found in the available literature on the metabolism
of terbacil.
Excretion
0 No information was found in the available literature on the excretion
of terbacil.
IV. HEALTH EFFECTS
Humans
0 No information was found in the available literature on the health
effects of terbacil in humans.
Animals
Short-term Exposure
0 It was not possible to perform an acute oral toxicity study in dogs
because repeated emesis prevented dosing with terbacil in amounts in
excess of 5,000 mg/kg (Paynter, 1966). However, in a dog receiving
one oral dose of terbacil at 250 mg/kg followed 5 days later by a
dose of 100 mg/kg, emesis, diarrhea and mydriasis were noted.
In rats (details not available), the LDso was between 5,000 and 7,500
mg/kg (Sherman, 1965). At 2,250 mg/kg, inactivity, weight loss and
incoordination were noted.
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Terbacil August, 1987
-5-
DermaI/Ocular Effects
0 Hood (1966) reported that no compound-related clinical or pathological
changes were observed when terbacil was applied to the clipped dorsal
skin of rabbits (five males, five females) at a dose level of 5,000
ng/kg (as a 55% aqueous paste), for 5 hours/day, 5 days/week for
3 weeks (15 applications). The parameters observed included body
weight, dermal reaction, organ weights and histopathology.
0 Reinke (1965) reported that no dermal reactions were observed when
terbacil was administered to the intact dorsal skin of 10 guinea pigs
as a 15% solution in 1:1 acetonetdioxane containing 13% guinea pig fat.
0 Reinke (1965) reported no observed sensitization in ten albino guinea
pigs when terbacil was administered nine times during a 3-week period,
with half of the animals in each group receiving dermal applications
on aoraded dorsal s ta.n and the others receiving intraderma1 injections.
After 2 wee Is, the animals were challenged by application of terbacil
to intact and abraded skin. The challenge application was repeated
2 weeks later.
Long-term Exposure
0 Wazeter et al. (1964) administered terbacil, 82.7% (a.i.), in the
diet to Charles River pathogen-free albino rats (20/sex/level) at
levels of 0, 100, 500 or 5,000 ppm of a.i. for 90 days. This corresponds
to doses of about 0, 5, 25 or 250 mg/kg/day based on the dietary
assumptions of Lehman (1959). The parameters observed included body
weight, food consumption, hematology, liver function tests, urinalyses,
organ weights and gross and histologic pathology. No adverse effects
with respect to behavior and appearance were noted. All rats survived
to the end of the study. No effect on body weight gain was observed
in either sex when terbacil was administered at 5 or 25 mg/kg/day.
Females administered 250 mg/kg/day gained slightly less weight (15%)
than controls. Males at this level showed no effect. No compound-
related hematological or biochemical changes were found, and urinalyses
were normal at all times. No gross or microscopic pathological
changes were noted in animals administered terbacil at 5 or 25 mg/kg/day.
Morphological changes in animals receiving the highest dose level
were limited to the liver and consisted of statistically significant
increases in liver weights. This change was accompanied by a moderate-
to-marked hypertrophy of hepatic parenchymal cells associated with
vacuolation of scattered hepatocytes. Similar microscopic changes,
but with reduced severity, were found in one rat at the 25 mg/kg/day
level. This study identified a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 25 mg/kg/day and a No-Observed-Adverse-Effect-Level (NOAEL)
of 5 mg/kg/day.
0 Goldenthal et al. (1981) administered terbacil (97.8% a.i.) in the
diet to CD-I mice (80/sex/level) at levels of 0, 50, 1,250 or 5,000 to
7,500 ppm for 2 years. Based on the dietary assumptions of Lehman
(1959), 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day;
therefore, these levels correspond to doses of about 0, 7.5, 187 or
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Terbacil August, 1987
-6-
750 to 1,125 mg/kg/day. The 5,000-ppm dose level was increased
slowly to 7,500 ppm by week 54 of the study. Mortality was signifi-
cantly higher (p <0.05) in mice at the high dosage levels throughout
the study. No changes considered biologically important or compound-
related occurred in the hematological parameters. An increased
incidence of hepatocellular hypertrophy was seen microscopically in
male and female mice administered 750 to 1,125 mg/kg/day and in male
mice administered 187 mg/kg/day. An increased incidence of hyperplastic
liver nodules also occurred in male mice administered 750 to 1,125
mg/kg/day. Female mice from the 187-mg/kg/day group and both male
and female mice from the 7.5-mg/kg/day group were free of compound-
related microscopic lesions. This study identified a LOAEL of
187 mg/kg/day and a NOAEL of 7.5 mg/kg/day.
0 Wazeter et al. (1967b) administered terbacil (80% a.i.) in the diet
to CO albino rats (36/sex/level) at levels of 0, 50, 250 or 2,500 ppm
to 10,000 ppm of a.i. for 2 years. Based on the dietary assumptions
of Lehman (1959), 1 ppm in the diet of a rat corresponds to
0.05 mg/kg/day; therefore, these dietary levels correspond to doses
of about 0, 2.5, 12.5 or 125 to 500 mg/kg/day. The 2,500 ppm level
was increased slowly to 10,000 ppm by week 46 of the study. No
adverse compound-related alterations in behavior or appearance occurred
in any test group. No significant differences in body weight gain in
males and females administered 2.5 or 12.5 mg/kg/day were observed.
Rats administered 125 to 500 mg/kg/day exhibited a significantly
lower rate of body weight gain. This difference occurred early and
became more pronounced with time in the female rats than in the male
rats. Maximum differences were 14 to 17% in the male rats and 24 to
27% in the females when compared to the controls. No compound-related
gross pathological lesions were seen at necropsy in rats from any
groups. The only compound-related variation in organ weights was a
slight increase in liver weights among rats from the 125- to
500-mg/kg/day dose level at final sacrifice. Histological changes
were observed in the livers of rats fed terbacil at 12.5 mg/kg/day
for 1 year and in the high-dose group fed 125 to 500 mg/kg/day for
1 and 2 years. These changes consisted of enlargement and occasional
vacuolation of centrilobular hepatocytes. No compound-related microscopic
changes were observed in livers or in any tissues examined in rats from
the 12.5-mg/kg/day group sacrificed after 2 years. Due to an outbreak
of respiratory congestion observed in all study groups at week 27,
all animals were placed on antibiotic treat-lent (tetracycline hydrochloride)
at a dose level of 25 mg/kg/day in drinking water for 1 week. In the
29th week, all rats were administered 50,000 units of penicillin G
intra-muscularly and 1/16 g of streptomycin. Some rats still exhibiting
respiratory congestion were administered a second dose on the following
day. This study identified a LOAEL of 125 to 500 mg/kg/day, based on
irreversible histological changes in the liver, and a NOAEL of
12.5 mg/kg/day.
Wazeter et al. (1966) administered terbacil (80% a.i.) in the diet to
young purebred beagle dogs (4 to 6 months old, four/ sex/dose) at
dose levels of 0, 50, 250 or 2,500 to 10,000 ppm of a.i. for 2 years.
Based on the dietary assumptions of Lehman (1959), 1 ppm in the diet
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Terbacil August, 1987
-7-
of a dog corresponds to 0.025 mg/kg/day; therefore, these dietary
levels correspond to approximately 0, 1.25, 6.25 or 62.5 to
250 mg/kg/day. The 2,500-ppm level was gradually increased to
10,000 ppm from week 26 to week 46 of the study. All animals underwent
periodic physical examinations, hematologic tests, and determinations
of 24-hour alkaline phosphatase, prothrombin time, serum glutamate
oxaloacetate transaminase (SCOT), serum glutamate pyrurate transaminase
(SGPT) and cholesterol. No adverse compoundrelated alterations in
behavior or appearance occurred among any of the control or treated
dogs. No mortalities occurred during the 2-year course of treatment.
Although there were some fluctuations in body weight throughout the
study, these were not considered to be compound-related. No alterations
in hematology, plasma biochemistry or urinalysis were observed. No
compound-related gross or microscopic pathological changes were seen
in any of the dogs sacrificed after 1 or 2 years of feeding. A
slight increase in relative liver weights and elevated alkaline
phosphatase occurred in dogs from the 62.5- to 250-mg/kg/day group
and the 6.25-mg/kg/day group, which were sacrificed after 1 or
2 years. Also at 6.25 mg/kg/day, there was an increase in thyroid-
to-body weight ratio. This study identified a NOAEL of 1.25 mg/kg/day
(50 ppm) and a LOAEL of 6.25 mg/kg (250 ppm).
Reproductive Effects
0 Wazeter et al. (1967a) administered terbacil (80% a.i.) in the diet
to male and female rats of three generations (10 males and 10 females
per level per generation) at dietary levels of 0, 50 or 250 ppm of
a.i. Based on the dietary assumptions of Lehman (1959), 1 ppm in the
diet of a rat is equivalent to 0.05 mg/kg/day; therefore, these
dietary levels correspond to doses of about 2.5 or 12.5 mg/kg/day.
Each parental generation was administered terbacil in the diet for
100 days prior to mating. No abnormalities in behavior, appearance
or food consumption of the parental rats were observed in any of the
three generations. Males at the 12.5 mg/kg/day level in all three
generations exhibited reduced body weight gains. Females in all
three generations were similar to controls in body weight gain. No
abnormalities were observed in the breeding cycle of any of the three
generations relative to the fertility of the parental male and female
rats, development of the embryos and fetuses, abortions, deliveries,
live births, sizes of the litters, viability of the newborn, survival
of the pups until weaning o- growth of the pups during the nursing
period. Gross examination of pups surviving at weaning from both
litters of all three generations did not reveal any evidence of
abnormalities. No compound-related histopathological lesions were
observed in any of the tissues examined from weanlings of the F3b
litter. This study identified a NOAEL of 2.5 mg/kg/day and a LOAEL
of 12.5 mg/kg/day.
Developmental Effects
0 E.I. DuPont (1984a) administered terbacil by gavage as a 0.5% suspen-
sion in methyl cellulose to groups of 18 female New Zealand White
rabbits (5 months old) from days 7 to 19 of gestation at dose levels
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Terbacil August, 1987
-8-
of 0, 30, 200 or 600 mg/kg/day. Maternal mortality was significantly
increased (p £0.05) at the 600-mg/kg/day level. Additional indicators
of maternal toxicity at 600-mg/kg/day were a significant increase
(p <0.05) in adverse clinical signs (anorexia and liquid or semi-solid
yellow, orange or red discharges found below the cages) and a significant
decrease (p <0.05) in body weight gain. Mean body weight gains and
the incidence of adverse effects were similar in controls and in the
30- and 200-mg/kg/day groups. Fetal toxicity at doses of 600 mg/kg/day
included a significant decrease (p £0.05) in fetal body weight and a
significant increase (p £0.05) in the frequency of extra ribs and
partially ossified and unossified phalanges and pubes. This increase
was not due to a statistically significant increase in any specific
malformation, and occurred only at a dosage level that was overtly
toxic to the dams, suggesting to the authors that it may be the result
of maternal toxicity. No increase in the incidence of adverse effects
was noted among fetuses produced by animals administered 30 or
200 mg/kg/day terbacil. Based on maternal and fetal toxicity, this
study identified a NOAEL of 200 mg/kg/day and a LOAEL of 600 mg/kg/day.
0 Culik et al. (1980) administered terbacil (96.6% a.i.) in the feed to
female rats from days 6 to 15 of gestation at levels of 0, 250, 1,250
or 5,000 ppm. Based on the measured food consumption, these dietary
levels correspond to doses of about 0, 23, 103 or 391 mg/kg/day.
Maternal parameters observed included clinical signs of toxicity and
changes in behavior, body weight and food consumption. Statistically
significant (p£0.05), compound-related reductions in mean body
weight, weight gain and food consumption were seen in animals administered
103 or 391 mg/kg/day. No other clinical signs or gross pathological
changes were observed in any animals. The mean number of live fetuses
per litter and mean final maternal body weight were significantly
lower (p £0.05) in the groups administered 103 or 391 mg/kg/day than
in the control group; the mean number of implantations per litter was
also significantly lower (p £0.05) than in control animals. Anomalies
occurred in the renal pelvis, and ureter dilation was found in all
the treatment groups. This study identified a LOAEL of 23 mg/kg/day,
based on anomalies of the renal pelvis and ureter dilation.
Mutagenicity
0 E.I. DuPont (1984b) reported that terbacil did not induce unscheduled
DNA synthesis in primary cultures of rat hepatocytes (0.01 and 1.0 uM),
did not exnibit mutagenic activity in the CHO/HGPRT assay (0 to 5.0 uM)
with or without metabolic activation, and did not produce statistically
significant differences between mean chromosome numbers, mean mitotic
indices or significant increases in the frequency of chromosomal
aberrations when tested by in^ vivo bone marrow chromosome studies in
Sprague-Dawley CD rats (15/sex/level) administered a single dose of
terbacil by gavage at 0, 20, 100 or 500 mg/kg.
0 Murnik (1976) reported that terbacil significantly elevated the rates
. of apparent dominant lethals when tested in Drosophila melanogaster,
but the authors concluded that the significant reductions in egg
hatch were probably due to physiological toxicity of the treatment.
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Terbacil August, 1987
-9-
since genetic assays did not indicate the induction of chromosomal
breakage or loss.
Carcinoqenicity
0 Goldenthal et al. (1981) administered terbacil (97.8% a.i.) in the
diet to CD-1 mice (80/sex/level) at levels of 0, 50, 1,250 or
5,000 to 7,500 ppm for 2 years. These levels correspond to doses of
about 0, 7.5, 187 or 750 to 1,125 mg/kg/day (Lehman, 1959). The
5,000-ppm dose level was increased slowly to 7,500 ppm by week 54 of
the study. The authors reported no increased incidence of cancer in
the treated animals.
0 Wazeter et al. (1967b) administered terbacil (80% a.i.) in the diet
to CD albino rats (36/sex/level) at levels of 0, 50, 250 or 2,500 to
10,000 ppm of active ingredient for 2 years. These levels
correspond to doses of about 0, 2.5, 12.5 or 125 to 500 mg/kg/day
(Lehman, 1959). The authors reported no evidence of compound-related
carcinogenic effects.
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 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 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 terbacil. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child, 0.24 mg/L (240 ug/L),
calculated below, be used at this time as a conservative estimate of the
One-day HA value.
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Terbacil August, 1987
-10-
Ten-day Health Advisory
The dietary reproductive study in rats by Wazeter et al. (1967a) has
been selected to serve as the basis for the Ten-day HA value for terbacilo
It identifies a LOAEL of 12.5 mg/L, based on a reduced body weight gain in
the males in all three generations, and a NOAEL of 2.5 mg/kg/day, yielding
a Ten-day HA of 0.25 mg/L (see calculation below). The teratology study in
rats by Culik et al. (1980) provides support for this conclusion. This teratology
study identifies a LOAEL of 23 mg/L (no doses lower than 23 mg/kg/day were
tested) and essentially the same Ten-day HA value (0.23 mg/L) can be derived
from this LOAEL by using an uncertainty factor of 1,000.
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA - (2.S mg/kg/day) (10 kg) a 0.25 mg/L (250 ug/L)
(100) (1 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of reduced body weight gain in
male rats.
10 kg = assumed body weight of a child.
100 a uncertainty factor, chosen in accordance with NAS/OOW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The dietary reproductive study in rats by Wazeter et al. (1967a) has been
selected to serve as the basis for the Longer-term HA values for terbacil. A
NOAEL of 2.5 mg/kg/day is identified in this study. A 90-day subchronic study
in rats (Wazeter et al., 1964) identifying a NOAEL of 5 mg/kg/day supports
this conclusion.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (250 ug/L)
(100) (1 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of reduced body weight gain
in male 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.
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Terbacil August, 1987
-11-
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (2.5 mg/kg/day) (70 kg) = 0.875 mg/L (875 ug/L)
(100) (2 L/day)
where:
2.5 mg/kg/day = NOA EL, based on absence of reduced body weight gain
in male rats.
70 kg = assumed body weight of an adult.
100 a 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.
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 wit> lifetime exposure to this chemical.
The 2-year dog feeding study by Wazeter et al. (1966), selected to serve
as the basis for the Lifetime HA value for terbacil, identifies a NOAEL of
1.25 mg/kg/day, based on relative liver weight increases and an increase in
alkaline phosphatase. A number of other studies provide information that
supports the conclusion that the overall NOAEL for lifetime exposure of rats,
mice and dogs to terbacil is less than 25 mg/kg/day. These include a 2-year
feeding study in mice that identifies a NOAEL of 7.5 mg/kg/day for liver
changes (Goldenthal, 1981) and a 2-year feeding study in rats that identifies
a NOAEL of 12.5 mg/kg/day for lower body weight gain and liver effects (Wazeter
et al., 1967b).
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Terbacil August, 1987
-12-
Using a NOAEL of 1.25 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (1.25 mg/kg/day) . Q.0125 mg/kg/day
(100)
where:
1.25 mg/kg/day = NOAEL, based on slight increase in relative liver
weight and elevated alkaline phosphatase.
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 = <°'0125 mg/kg/day) (70 kg) = 0
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Terbacil August, 1987
-13-
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
• Tolerances have been established for residues of terbacil in or on
many agricultural commodities by the U.S. EPA
Office of Pesticide Programs
(U.S. EPA, 1985a).
VII. ANALYTICAL METHODS
0 Analysis of terbacil is by a gas chromatographic method applicable
to the determination of certain organonitrogen pesticides in water
samples (U.S. EPA, 1985b). This method requires a solvent extraction
of approximately 1 L of sample with methylene chloride using a
separatory funnel. The methylene chloride extract is dried and exchanged
to acetone during concentration to a volume of 10 mL or less. The
compounds in the extract are separated by gas chromatography, and
measurement is made with a thermionic bead detector. The method
detection limit for terbacil has not been determined.
VIII. TREATMENT TECHNOLOGIES
0 Treatment technologies currently available have not been tested for
their effectiveness in removing terbacil from drinking water.
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August, 1987
Terbacil
-14-
IX. REFERENCES
Benson, N.R. 1973. Efficacy, leaching and persistence of herbicides i«
apple orchards. Bulletin No. 863. Washington State University, College
of Agriculture Research Center.
Culik, R., C.K. wood, A.M. Kaplan et al.* 1980. Teratogenicity study in
rats with 3-tert-butyl-5-chloro-6-methyluracil. Haskell Laboratory
Report No. 481-79. Haskell Laboratory for Toxicology and Industrial
Medicine. Newark, DE. Unpublished study. MRID 00050467.
Davidson, J.M., L.T. Ou and P.S.C. Rao. 1978. Adsorption, -«v-J"J.J»J
biological degradation of high concentrations of selected herbicides in
soils? in: Land disposal of hazardous wastes. U.S. Environmental
Protection" Agency, Office of Research and Development, pp. 233-244.
EPA-600/9-78-016.
Doughty, C.C. 1978. Terbacil phytotoxicity and quackgrass ^^«^
light, and nonsterile soil. Unpublished study submitted by E.I. du Pont
de Nemours and Company, Inc., Wilmington, DE.
h^c;^r;^^^
Food Chem. 17:980-986.
Corporation. Unpublished study. MRID 00126770.
Hood D.* 1966. Fifteen exposure skin absorption studies with 3-tert-butyl-
S^hloro-elmethyluracil. Report No. 33-66. Unpublished study.
MRID 00125785.
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Terbacil August, 1987
-15-
Isom, W.H., H.P. Ford, M.P. Lavalleye and L.S. Jordan.* 1969. Persistence
(sic) of herbicides in irrigated soils. Unpublished study prepared by
Sandoz-Wander, Inc., submitted by American Carbonyl, Inc., Tenafly, NJ.
Isom, W.H., H.P. Ford, M.P. Lavalleye and L.S. Jordan. 1970. Persistence
(sic) of herbicides in irrigated soils. Proc. Ann. Calif. Weed Conf.
22:58-63.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food and Drug Officials of the United States.
Liu, L.C., H.R. Cibes-Viade and J. Gonzalez-Ibanez. Undated. Persistence of
several herbicides in a soil cropped to sugarcane. J. Agric. Univ.
Puerto Rico (volume no. not available):147-152.
Liu, L.C., H. Cibes-Viade and F.K.S. Koo. 1971. Adsorption of atrazine and
terbacil by soils. J. Agric. Univ. Puerto Rico 55(4): 451-460.
Mansell, R.S., D.V. Calvert, E.E. Stewart, W.B. Wheeler, J.S. Rogers, D.A.
Graetz, L.E. Allen, A.F. Overman and E.B. Knipling. 1977. Fertilizer
and pesticide movement from citrus groves in Florida flatwood soils.
Athens, GA: U.S. Environmental Protection Agency, Environmental Research
Laboratory. EPA report number EPA-600/2-77-177. Also available from
NTIS, Springfield, VA, PB-272 889.
Mansell, R.S., W.B. Wheeler, D.V. Calvert and E.E. Stewart. 1979. Terbacil
movament in drainage waters from a citrus grove in a Florida flatwood
soil. Proc. Soil Crop Sci. Soc. Fl. 37:176-179.
Mansell, R.S., W.B. Wheeler, L. Elliott and M. Shaurette. 1972. Movement of
acarol and terbacil pesticides during displacement through columns of
Wabasso fine sand. Proc. Soil Crop Sci. Soc. Fl. 31:239-243.
Marriage, P.B., S.U. Kahn and W.J. Saidak. 1977. Persistence and movement
of terbacil in peach orchard soil after repeated annual applications.
Weed Res. 17:219-225.
Marsh, J.A.P., and H.A. Davies. 1978. The effect of herbicides on respiration
and transformation of nitrogen in two soils. III. Lenacil, terbacil,
chlorthiamid and 2,4,5-T. Weed Res. 18:57-62.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Morrow, L.A., and M.K. McCarty. 1976. Selectivity and soil persistence of
certain herbicides used on perennial forage grasses. J. Environ. Qual.
5:462-465.
Murnik, M.R.* 1976. Mutagenicity of widely used herbicides. Genetics. 83:554.
Paynter, O.F.* 1966. Final report. Acute oral toxicity study in dogs.
Haskell Laboratory for Toxicology and Industrial Medicines, Newark, DE.
Unpublished Study. MRID 00012206.
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Terbacil Au9ust' 1987
-16-
Rahman, A. 1977. Persistence of terbacil and trifluralin under different
soil and climatic conditions. Weed Res. 17:145-152.
Rao, P.S.C., and J.M. Davidson. 1979. Adsorption and movement of selected
pesticides at high concentrations in soils. Water Res. 13:375-380.
Reinke, R.E.* 1965. Primary irritation and sensitization skin tests. Haskell
Laboratory Report No. 79-65. E.I. duPont deNemours and Company, Inc.
Haskell Laboratory for Toxicology and Industrial Medicine, Newark, DE.
Unpublished study. MRID 0006803.
Rhodes, R.C.* 1975. Biodegradation studies with 2-14c-terbacil in water and
soil. Unpublished study prepared in cooperation with University of
Delaware, College of Agricultural Sciences, submitted by E.I. duPont
deNemours and Company, Inc., Wilmington, DE.
Sherman, H.* 1965. Oral LD50 test. Haskell Laboratory Report No. 160-65.
E.I. duPont deNemours and Company, Inc. Haskell Laboratory for Toxi-
cology and Industrial Medicine. Newark, DE. Unpublished study.
MRID 00012235.
Skroch, W.A., T.J. Sheets and J.W. Smith. 1971. Herbicides effectiveness,
soil residues, and phytotoxicity to peach trees. Weed Sci. 19:257-260.
STORET. 1987.
Tucker, D.P., and R.L. Phillips. 1970. Movement and degradation of herbicides
in Florida citrus soil. Citrus Ind. 51(3):11-13.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Terbacil; tolerances
for residues. 40 CFR 180.209.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. U.S. EPA Method 633-
Organonitrogen Pesticides, 50 FR 40701, October 4,
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
Wazeter, F.X., R.H. Buller and R.G. Geil.* 1964. Ninety-day feeding study in
rats. IRDC No. 125-004. International Research and Development Corp.
Unpublished study. MRID 00068035.
Wazeter, F.X., R.H. Buller and R.G. Geil.* 1966. Two-year feeding study in
the dog. IRDC No. 125-011. International Research and Development
Corp. Unpublished study. MRID 00060851.
Wazeter, F.X., R.H. Buller and R.G. Geil.* 1967a. Three-generation reproduc-
tion study in the rat. IRDC No. 125-012. International Research and
Development Corp. Unpublished study. MRID 00060852.
Wazeter, F.X., R.H. Buller and R.G. Geil.* 1967b. Two year feeding study in
the albino rat. IRDC No. 125-100. International Research and Development
Corp. Unpublished study. MRID 00060850.
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Terbacil August, 1987
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Wolf, D.C. 1973. Degradation of bromacil, terbacil, 2,4-D and atrazine in
soil and pure culture and their effect on microbial activity.
Ph.D. Dissertation, University of California, Riverside.
Wolf, D.C. 1974. Degradation of bromacil, terbacil, 2,4-D and atrazine in
soil and pure culture and their effects on microbial activity. Disser-
tation Abstracts International B. 34(10):4783-4784.
Wolf, D.C., and J.P. Martin. 1974. Microbial degradation of 2-carbon-14-
bromacil and terbacil. Proc. Soil Sci. Soc. Am. 38:921-925.
Zimdahl, R.L., V.H. Freed, M.L. Montgomery and W.R. Furtick. 1970. The
degradation of triazine and uracil herbicides in soil. Weed Res.
10:18-26.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
TERBUFOS
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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Terbufos August, 1987
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 13071-79-9
Structural Formula
CH3CH20
P - S - CH2 - S - C - CH3
S-[ [ (1,1-Dimethylethyl)thio]methyl]0,0-diethyl phosphorodithioate
Syropyms
0 Counter; Cor tr aver (Meister, 1986).
Uses
0 Control of corn rootworm and other soil insects and nematodes infesting
corn. Control of sugarbeet maggots in sugarbeets; green bug on
grain sorghum (Meister, 1986).
Properties (Windholz et al., 1983; Meister, 1986)
Chemical Formula CgH^iO^PS?
Molecular Weight 288.41
Physical State (room temp.) Clear, slightly brown liquid
Boiling Point 69°C/0.01 mm Hg
Melting Point -29»2°C
Density (24°C) 1.105
Vapor Pressure (25°C) 34.6 mPa
Specific Gravity
Water Solubility (25°C) 15 mg/L
Log Octanol/Water Partition 595
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor —
Technical 87 to 97% pure
Occurrence
0 Terbufos has been found in 444 of 2,016 surface water samples
analyzed and jn 9 of 283 ground water samples (STORET, 1987).
Samples were collected at 55 surface water locations and 261 ground
water locations, and terbufos was found in 5 states. The 85th
percentile of all nonzero samples was .10 ug/L in surface water and
3 ug/L in ground water sources. The maximum concentration found
was 2.25 ug/L in surface water and 3 ug/L in ground water.
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Terbufos August, 1987
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Ervirormertal Fate
Forthcoming from OPP, EPA
III. PHARMACOKINETICS
Absorptiop
0 North (1973) reported that 83% of a single oral dose of technical
14c-terbufos (0.8 mg/kg) was excreted in the urine of rats 168 hours
after dosing. (The carbon atom of the thiomethyl portion of
terbufos was radiolabeled.) An additional 3.5% was recovered in
feces. This study indicates that terbufos was well absorbed (about
80 to 85%) from the gastrointestinal tract.
Distribution
0 North (1973) reported that maximum residues of cholinesterase-inhib-
iting compounds (phosphory la ted metabolites), resulting from a single
oral dose of technical ^C-terbufos (0>8 m9/k9) given to rats, were
found in rat liver (0.08 ppm) 6 hours after dosing. In the same
study, residues of hydrolysis (nonphosphorylated metabolites) products
reached a maximum in rat kidney 12 hours after dosing (0.9 ppm).
After 168 hours, each body tissue in the rat contained less than
0.1 ppm radiolabeled) terbufos.
Metabolism
0 North (1973) reported that terbufos was extensively metabolized in
the rat. 14c-Radiolabeled terbufos was administered in a single dose
to 16 male Wistar rats at a dose level of 0.8 mg/kg via gavage.
Examination of urine extracts by thin-layer chromatography (TLC)
showed the presence of 10 radiometabolites in the rat urine. Approxi-
mately 96% of the radioactivity present in the urine was composed of
an S-methylated series of metabolites, which result from the cleavage
of the sulfur-phosphorus bond, methylation of the liberated thiol group
and oxidation of the resulting sulfide to sulfoxides and sulfones.
Of the remaining radioactivity, about 2% was composed of various
oxidation products of the intact parent organophosphorus compound and
2% was an unknown metabolite.
Excretion
North (1973) reported that technical terbufos and its metabolites
were rapidly excreted in the urine of the rat. Radiolabeled terbufos
was administered in a single dose to male 'Jistar rats at a dose level
of 0.8 mg/kg by gavage. Of all the radioactivity recovered ir the
urine, 50% was excreted after 15 hours. After 168 hours, the termina-
tion of the test, 83% of the terbufos was excreted via the urine and
3.5% was recovered in the feces.
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Terbufos August, 1987
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IV. HEALTH EFFECTS
Humans
0 Peterson et al. (1984) reported the results of farm worker exposure
to Counter 15-G (a 15% granular formulation of terbufos). Five
farmers (one loader, one flagger and three scouts) were exposed for
varying time periods (loader, 5 minutes; flagger, 15 minutes;
scouts, twice for 30 minutes) during a typical workday while Counter
15-G was applied aerially to a young corn crop. The mean exposure
via inhalation was <0.25 ug/hour, the sensitivity of the monitoring
method, for all samples collected. The exposure values for the five
farm workers were: 331 ug/hour for the loader, 0 ug/hour for the
flagg'er, 381 ug/hour for scouts (after 3 days) and 250 ug/hour for
scouts (after 7 days). All of the farm workers were men and weighed
between 65.9 and 90.9 kg. Analysis of urinary metabolites showed no
indication of any adverse effects to any of the exposed workers. All
urinary alkyl phosphate analyses were negative (detection level,
0.1 ppm), indicating no significant absorption of terbufos. Plasma
and red blood cell cholinesterase values of the exposed workers
showed no significant (95% confidence level) decrease in activity
when compared to pre-exposed samples, indicating no adverse physiological
effects from exposures.
0 Devine et al. (1985) reported results similar to Peterson et al. (1984)
for 11 farmers who were exposed to terbufos during a typical workday
while planting corn and applying Counter 15-G. The average estimated
dermal exposure was 72 ug/hour, and the estimated respiratory exposure
was 11 ug/hour. The results of urinary alkyl phosphate analyses were
all negative, showing no detectable absorption of terbufos. Plasma
and red blood cell cholinesterase (ChE) values of the exposed farmers
showed no significant difference in activity when compared to pre-
exposure or control values, indicating no adverse physiological
effects from the exposure. The report concluded that, based on the
study results, the use of Counter 15-G does not present a significant
hazard, in terms of acute toxicity, to farmers using this product for
the control of corn insects.
Animals
Short-term Exposure
0 Parke and Terrell (1976) reported that the acute oral 1*050 value
of technical-grade (86%) terbufos in Wistar rats was 1.73 mg/kg.
Terbufos was administered in doses of 1.0 to 3.0 mg/kg via gavage in
corn oil to a total of 50 rats (10/sex/dose). Average weight of the
rats ranged from 200 to 300 g. The lowest dose (1.0 mg/kg) did not
result in any mortality. Observed effects to the rats were: respir-
atory depression, piloerection, clonic convulsions, exophthalmus,
ptosis, lacrimation, hemorrhage and decreased motor activity.
0 Consultox Laboratories (1975) reported that the acute oral LD50 value of
technical-grade (86%) terbufos in male Wistar rats was 1.5 mg/kg.
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Terbufos August, 1987
-5-
Terbufos was administered by gavage in doses of 0.50 to 2.5 mg/kg to a
total of 50 rats (1 0/s ex/dose) at an average weight of 200 ± 20 g.
No mortality was reported at the low dose (0.50 mg/kg). Ten percent
mortality was reported at the 0.75-mg/kg dose. Other effects reported
were: salivation, diuresis, diarrhea, disoriertatior, chromodocryorrhea ,
piloerectior and body tremors.
American Cyanamid (1972a) reported acute oral LDso values (for 96.7%
technical-grade terbufos) in dogs, mice and rats of 4.5 mg/kg (male)/
6.3 mg/kg (female), 3.5 mg/kg (male)/9.2 mg/kg (female), and
4.5 mg/kg (male)/ 9.0 mg/kg (female), respectively. No details were
given as to age, weight or route of exposure.
0 American Cyanamid (1972b) reported additional acute oral LDso values
in male Wistar rats and female CF1 mice of 1.6 mg/kg and 5.0 mg/kg,
respectively. Other effects reported included cholinesterase inhibition
in both sexes.
0 Berger (1977) reported that plasma ChE was inhibited by as much as
79% in eight beagle dogs that were dosed via corn oil with
0.05 mg/kg/day technical terbufos for 28 days. Red blood cell ChE
was not inhibited at the dose tested.
Dermal/Ocular Effects
0 Kruger et al. (1973) conducted a subacute dermal toxicity test in
New Zealand White rabbits. Technical-grade terbufos was administered
at doses varying from 0.004 to 0.1 mg/kg to the shaved, abraded backs
of male and female rabbits (2.5 to 3.5 kg). All animals survived the
30-day test and showed no adverse effects with regard to food and
water intake, elimination, behavior, pharmacological effects and
weight gain differences. There were no observed changes in hemato-
logical determinations (hematocrit, total erythrocyte and total
leukocyte levels). Minor changes reported were increased numbers of
eosirophils and basophils in all groups, occasional minimal edema
that abated by day 21, and occasional mild erythema. All observed
changes occurred on intact and abraded skin sites.
0 American Cyanamid (1972a,b) conducted a series of tests with 96.7
and 85.8% terbufos using New Zealand White rabbits. Twenty male
rabbits (2.56 to 2.73 kg) were administered doses of 0.4 to 3.5 ng/kg
terbufos to their shaved backs. Dermal contact with terbufos was
maintained for 24 hours. The dermal LDso value was 1.0 mg/kg. An
acute dermal test with 96.7% terbufos resulted in an LDso of 1.1 mg/kg
in male rabbits (no other details were given). In another test with
96.7% terbufos, 0.5 mL (500 mg) of terbufos was applied to the backs
of rabbits; all of these animals died within 24 hours after dosing.
0 American Cyanamid (1972a) reported the results of an application of
0.1 mg of technical-grade (96.7%) terbufos to the eyes of New Zealand
albino rabbits. All animals died within 2 to 24 hours after dosing.
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Terbufos August, 1987
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Lonq-term Exposure
0 Daly et al. (1979) administered terbufos (90% active ingredient
(a.i.)) in the diet to groups of male and female Sprague-Dawley rats
(10/sex/group, 24 to 39 days old, 95 to 150 g) at levels of 0, 0.125,
0.25, 0.5 or 1.0 ppm (estimated doses of 0, 0.01, 0.02, 0.04 or 0.08
mg/kg/day based on feed conversions given by the authors) for 90
days. Body weights and food consumption were measured weekly. Blood
samples were obtained weekly and analyzed for plasma, erythrocyte and
brain ChE. Body organs were weighed and analyzed for histopathology.
The No-Observed-Adverse-Effect-Level (NOAEL) was determined to be
0.02 mg/kg/day, based on the absence of effects on ChE. The statistically
significant Lowest-Observed-Adverse-Effect-Level (LOAEL) was determined
to be 0.046 mg/kg based on the observed 17% decrease in plasma ChE in
females. There were no depressions of erythrocyte or brain ChE at
the highest dose tested (0.09 mg/kg/day).. In addition, gross postmortem
observations and histopathologic evaluation of selected tissues
revealed no findings related to the test substance. Systemically, the
LOAEL for increased liver weight in females and for a dose-related
increase in liver extra-medullary hematopoiesis was 0.046 mg/kg/day.
The systemic NOAEL based on absence of liver effects was determined to be
0.02 mgAg in this study.
0 Morgareidge et al. (1973) administered technical-grade terbufos in
the diet to groups of male and female beagle dogs (four/sex/group,
10 to 13 months old, 9.0 to 13.8 kg) at levels of 0.0025, 0.01 and 0.04
mg/kg/day, 6 days a week for 26 weeks. Plasma, red blood cell and
brain ChE levels, body weight and food, urinalysis, gross necropsy
examination and histopathology were evaluated. Observed effects
included a decrease in ChE activity in plasma at all dose levels;
however, decreased ChE activity was statistically significant only
for doses of 0.01 mg/kg/day and above. At 0.01 mg/kg/day, plasma ChE
was inhibited by 26% and red blood cell ChE was inhibited by 14%.
The systemic NOAEL was determined to be greater than the highest dose
tested (0.04 ing/kg/day). No statistical analyses were performed on
body weight changes, food consumption, hematology, clinical chemistry,
urinalyses and organ weight data. The LOAEL (based on ChE effects)
determined by the study was 0.01 mg/kg/day and the NOAEL was determined
to be 0.0025 mg/kg/day.
0 Rapp et al. (1974) administered technical-grade terbufos in the diet
to groups of Long-Evans rats (six/sex/dose, weanlings, 122 to 138.8 g)
at levels of 0.25, 1.0, 2.0, 4.0, and 8.0 ppm for 2 years. These doses
correspond to 0.0125, 0.05, 0.1, 0.2 and 0.4 mg/kg/day (Lehman, 1959.
The original high doses (2.0 ppm) were increased to 4.0 and then to
8.0 ppm for males, and were increased from 2.0 to 4.0 to 8.0 and then
reduced to 4.0 ppm for females. Body weight and food consumption
were measured weekly. Hematology, clinical chemistry and urinalyses
were also performed. Red blood cell ChE and brain ChE were significantly
inhibited at 0.05 mg/kg/day (20% inhibition for brain ChE and 43% for
red blood cell ChE in females) and above. Red blood cell ChE was
also inhibited at 0.0125 mg/kg/day (12% in males and 15% in females).
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Tterbufos August, 1987
-7-
At the high dose (0.1 to 0.4 mg/Jq/day), there was a noticeable inhibition
in mean body weight and mean food consumption. Mortality rates were
24 and 27% (males and females, respectively) at the high dose,
19% (males) at the mid-dose and 10% (males) at the low dose. The
incidence of exophthalmia was in high-dose females (exophthalmia was
also noted in low- and mid-dose control females). This study did not
establish a NOAEL. The LOAEL was equivalent to the lowest dose
tested (0.0125 mg/Jq/day).
0 McConnell (1983) administered technical-grade terbufos in the diet to
groups of Long-Evans rats (60/dose/sex) at levels of 0.25, 1.0, 2.0,
4.0 and 8.0 ppn for 2 years. These doses are equivalent to 0.0125,
0.05, 0.1, 0.2 and 0.4 rag/kg/day (Lehman, 1959). The original high
dose (2.0 ppn) was increased to 4.0 and then 8.0 ppn for males after the
first 3 months, and increased from 2.0 to 4.0 to 8.0 and then reduced
to 4.0 ppm for females after the first 3 months. At the end of the
2-year study, tissues were prepared for microscopic examination.
Mortality occurred in all groups (control and test) due to broncho-
pneumonia, with mortality rates ranging from 17 to 35% in controls
and low-dose groups, respectively. Mortality rates at the high dose
(0.4 mg/Jq/day) were 58% and 43% in male and female rats, respectively.
Other effects reported were gastric ulceration and/or erosion of
glandular and nonglandular stomach mucosa in high-dose rats. No
similar effect was seen in lowand mid-dose rats. Acute bronchopneumonia
and granuloma of lungs occurred in high-dose rats more frequently
than in low-dose, mid-dose or control rats. The authors reported
that lung inflammation did not appear directly associated with the
compound. No LOAEL or NOAEL was established in this study.
0 Shellenberger (1986) administered technical-grade terbufos (89.6%a.i.)
in capsule form to groups of beagle dogs (six/sex/dose, 6.8 to 7.5 Jq,
5 to 6 months old) at doses of 0, 0.015, 0.060, 0.090, 0.120, 0.240 and
0.480 mg/Jq/day for 1 year. High doses were eventually reduced to
0.090 and 0.060 mg/Jq/day after the 8th week of the study. Body
weight and food consumption were measured together with assessment of
urinalyses, organ weights and cholinesterase levels. One male and one
female at the high dose and one female at 0.240 mg/Jq/day were found
dead. At the two highest doses (0.240 and 0.480 mg/Jq/day), decreased
body weights and food consumption were observed. Mean erythrocytic
parameters of high-dose males and females were significantly reduced at
3 months but not at 6 months ar at termination of the study. Plasma
ChE activity was significantly inhibited to 55% of controls at 0.015
mg/Jq/day. Slight inhibition of erythrocyte ChE activity occurred at
0.120 mg/Jq/day in females but not in males. No inhibition of erythrocyte
ChE in males or females was observed at the lower doses. Brain ChE
activities were similar for both sexes at all dose levels. Urinalyses and
organ weight data revealed no significant differences. The report
suggests that the NOAEL was 0.120 mg/Jq/day in males and 0.090 mg/Jq/day
in females.
Reproductive Effects
0 Smith and Kasner, (1972a) administered technical terbufos via the diet
to Long-Evans and Blue Spruce rats (10 males/dose, weighing 276.3 g; 20
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Terbufos August, 1987
-8-
females/dose, weighing 185.6 g) for a period of 6 months at levels
of 0, 0.25 and 1 ppn. These levels correspond to doses of 0, 0.0125
and 0.05 mg/kg/day, based on a conversion factor of 0.05 for rats
(Lehman, 1959). The first parental generation (F0) was dosed for 60
days. No reproductive effects were observed in males or females at
any dose tested. The authors concluded that the reproductive NOAEL
was greater than 0.05 mg/kg/day, the highest dose tested.
Developmental Effects
0 HacKenzie (1984) administered terbufos (87.8% a.i.) by gavage to
groups of 18 female New Zealand White rabbits (3.5 kg) at levels of
0, 0.1, 0.2 and 0.4 mg/kj/day on days 7 to 19 of gestation. Reproductive
indices monitored were female mortality, corpora lutea or implants,
sex ratio, implantation efficiency, fetal body weight, fetal mortality
and skeletal development. Cesarean sections were performed on day 29
of gestation. Survival of adult female rabbits was 100% in controls
and in the 0.2-mg/kg/day dose group; 89% in the 0.1-mg/kg/day dose
group; and 67% in the high-dose (0.4 mg/kg/day) group. There were
no statistically significant dose-related differences in mean body
weight, weight changes or gravid uterine weights, mean number of
corpora lutea, implantation efficiency, sex ratio, fetal body weight
or number of live or resorbing fetuses. The incidence of fetuses
with accessory left subclavian artery was significantly greater in
the high-dose (0.4 rag/kg/day) group. The incidence of an extra
unilateral rib and of chain fusion of sternebrae was significantly
lower in the high-dose group than in the controls. According to the
author, terbufos appears to be maternally toxic at 0.4 mg/kg/day, the
highest dose tested.
0 Rodwell (1985) administered terbufos (87.8% a.i.) via gavage to
groups of 25 Charles River female rats (226 to 282 g, 71-days old) at
doses of 0.05, 0.10 and 0.20 mg/kg/day on days 6 to 15 of gestation.
Cesareans sections were performed on day 20; half of the fetuses were
stained for skeletal evaluation. Parent survivability, body weight
and embryonic and fetal development were all assessed. All parents
survived the test. No changes in general appearance or behavior were
observed. Slightly decreased mean body weights were observed during
days 12 to 16 and following treatment in the 0.10- and 0.20-mg/kg/day
dose groups. The study demonstrates that terbufos is slightly
maternally f-xic at dose levels of 0.10 and 0.20 mg/kg/day. A NOAEL
of 0.05 mg/kg/day, the lowest dose tested, was identified.
Mutagenicity
0 Thilager et al. (1983) reported that Chinese hamster ovary cells
tested with and without S-9 rat liver activation at concentrations of
100, 50, 25, 10, 5 and 2.5 nL/mL (ppm) terbufos did not cause any
significant increase in the frequencies of chromosomal aberrations.
Only a concentration of 100 nL/mL proved to be cytotoxic.
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Terbufos August, 1987
-9-
0 Allen et al. (1983) conducted mutagenicity tests with terbufos
(87.8% a.i.) in the presence of metabolic activation and Chinese
hamster ovary cells and in the absence of S-9 activation. Initial
tests were conducted with doses of 100 to 10 ug/L, and then followed up
with activation at doses of 50, 42, 33, 25, 10 and 5 mg/ml. Terbufos
proved to be cytotoxic at 75 to 100 ug/mL with activation and at 50
to 70 mg/mL without activation. There were no increases in the
frequency of chromosomal aberrations. The authors concluded that
terbufos reflected a negative mutagenic potential.
e Godeket al. (1983) conducted a rat hepatocyte primary culture/DNA
repair test with terbufos (87.8% a.i.) at doses ranging from 100 to
33 ug/well (a well contains 2 mL of media). Unscheduled DNA repair
synthesis was quantified by a net nuclear increase of black silver
grains for 50 cells/slide. This value was determined by taking a
nuclear count and three adjacent cytoplasmic counts (100 ug/well was
cytotoxic). The results for terbufos were negative in the rat hepato-
cyte primary culture/DNA repair test. These findings are based on
the inability of terbufos to produce a mean grain count of 5 or
greater than the vehicle-control mean grain count at any level. The
authors concluded that terbufos reflected a negative mutagenic
potential.
Carcinogenicity
0 Smith and Kasner (1972b) administered technical terbufos in the diet
to groups of mice (15/sex/dose) at levels of 0, 0.5, 2.0 and 8.0 ppn
for 18 months. These doses correspond to.0.075, 0.3 and 1.2 mg/Jq/day
(Lehman, 1959). The authors reported no signs of tumors or neoplasia.
Effects noted include alopecia and signs of ataxia; exophthalmia in
males, corneal cloudiness and opacity and eye rupture. Organ tissues
examined were liver, kidney, heart and lung. No pathological changes
in these four organs were observed.
0 Rapp et al. (1974) administered technical terbufos in the diet to
groups of Long-Evans rats (six/sex/dose) at levels of 0, 0.25, 1.0,
2.0, 4.0 and 8.0 ppn for 2 years. These doses correspond to 0.0125,
0.05, 0.1, 0.2 and 0.4 mg/kg/day (Lehman, 1959). There were no
indications of tumorigenic effects at any dose tested.
McConnell (1983) administered technical terbufos in the diet tt-
groups of Long-Evans rats (60/sex/dose) at levels of 0, 0.25, 1.0,
2.0, 4.0 and 8.0 ppn for 2 years. These doses correspond to 0,
0.125, 0.05, 0.1, 0.2 and 0.4 mg/kg/day (Lehman, 1959). The author
concluded that the compound had no effect on tumorigenesis.
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:
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Terbufos August, 1987
-10-
HA = (NOAEL or LOAEL) x (BW) = fflg/L
(UF) x ( L/day)
where:
NOAEL or LOAEL « No- or Lowest-Observed-Adverse-Effeet-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 __
No information was found in the available literature that was suitable
for the determination of the One-day HA value for terbufos. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child (0.005 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The teratogenicity study in rats by Rodwell (1985) has been selected to
serve as the basis for the Ten-day HA value for terbufos. Pregnant rats
administered terbufos via gavage at a level of 0.05 mg/kg/day showed no
clinical signs of toxicity in the adult animals and no reproductive or terato-
genic effects in the fetuses. The study identified a NOAEL of 0.05 mg/kg/day.
These results are supported by the results of studies by MacKenzie (1984)
with rabbits and by Smith and Kasner (1972a) with rats.
Using a NOAEL of 0.05 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (0.05 mg/kg/day) (10 kg) = 0.005 mg/L (5 ug/L)
(100) (1 L/day)
where:
0.05 mg/kg/day = NOAEL, based on the absence of clinical signs of
toxicity and the lack of reproductive or teratogenic
effects in rats exposed to terbufos via gavage for
10 days during gestation.
10 kg =» assumed body weight of a child.
100 a 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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Terbufos August, 1987
-11-
Longer-term Health Advisories
No suitable studies were available to serve as the basis for the Longer-
term HA value for terbufos. It is recommended, however, that the modified
Drinking Water Equivalent Level (DWEL) (adjusted for a 10-kg child) be used as
a conservative estimate for a longer-term exposure. Accordingly, the Longer-term
HA for a 10-kg child is 0.00025 mg/L and the Longer-term HA for an adult is
0.00088 mg/L.
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
noncarcinogenic 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, 1986a), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The 6-month feeding study in beagle dogs by Morgareidge et al. (1973)
has been selected to serve as the basis for the Lifetime HA value for terbufos.
In this study, beagle dogs were administered terbufos in the diet at doses of
0.0025, 0.01 and 0.04 mg/kg/day. At 0.01 mg/kg/day and above, plasma and red
blood cell ChE activity were significantly inhibited. At 0.01 mg/kg/day,
plasma ChE was inhibited by 26% and red blood cell ChE was inhibited by 14%.
These effects were not observed at O.OO25 mg/kg/day, which was identified as
the NOAEL. Other studies were not selected because a clear NOAEL was not
identifed or the respective NOAELs/LOAELs were an order of magnitude
higher than the NOAEL derived from the Morgareidge et al. (1973) study.
Using this study, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.0025 mg/kg/day) = 0.000025 mg/kg/day
(100)
-------
Terbufos ' 198?
-12-
where:
0.0025 mg/kg/day = NOAEL, based on absence of inhibition of cholin-
esterase in beagles exposed to terbufos in the
diet for 6 months (180 days).
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)
DMEL ., (0.000025 mg/kg/day) (70 kg) -- o.00088 mg/L/day (0.88 ug/L)
(2 L/day)
where:
0.000025 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 - (0.00088 mg/L) (20%) = 0.00018 mg/L (0.18 ug/L)
where:
0.00088 mg/L » DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of terbufos.
0 The U. S. EPA's Cancer Assessment Group (CAG) has assessed the carcino-
genic potential of terbufos and has concluded that there are not
enough data to determine whether terbufos is carcinogenic.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), terbufos may be classified in
Group E: no evidence of carcinogenicity for humans. This group is for
substances that show no evidence of carcinogenicity in at least two
adequate animal tests in different species or in both epidemiologic
and animal studies. The studies by Smith and Kasner (1972b) on mice
and by Rapp et al. (1974) and McConnell (1983) on rats reported no
statistically significant influence on the incidence of neoplasms or
tumors at any dose level tested.
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Terbufos August, 1987
-13-
VI. OTHER CRITERIA, GUIDANCE AMD STANDARDS
0 No other criteria, guidance or standards were found in the available
literature.
VII. ANALYTICAL METHODS
8 Analysis of terbufos is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1986b). In this method,
approximately 1 liter of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
capillary collumn GC. Measurement is made using a nitrogen-phosphorus
detector. The method detection limit has not been determined for
this compound but it is estimated that the detection limits for the
method analytes are in the range of 0.1 to 2 ug/L.
VIII. TREATMENT TECHNOLOGIES
* No data were found- for the removal of terbufos from drinking water by
conventional treatment.
0 No data were found on the removal of terbufos from drinking water by
activated- carbon adsorption. However, due to its low solubility and
high molecular weight, terbufos probably would be amenable to activated
carbon adsorption.
0 No data were found on the removal of terbufos from drinking water by
ion exchange. However, the structure of this ester indicates that it
is not ionic and thus probably would not be amenable to ion exchange.
0 No data were found for the removal of terbufos from drinking water by
aeration. However, the Henry's Coefficient can be estimated from
available data on solubility (10 to 15 mg/L) and vapor pressure
(0.01 mm Hg at 69°C). Terbufos probably would not be amenable to
aeration or air stripping because its Henry's Coefficient is
approximately 12 atm.
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Terbufos August, 1987
-14-
IX. REFERENCES
Allen, J., E. Johnson and B. Fine. 1983. Mutagenicity testing of AC 92,100
in the in vitro CHO/HGPRT mutation assay. Project No. 0402. Final
report. Unpublished study. MRID 133297.
American Cyanamid Company. 1972a. Summary of data: Investigations made
with respect to the safety of AC 92, 100. Summary of studies 093580-A
through 093580-D. Unpublished study. MRID 35960.
American Cyanamid Company. 1972b. Toxicity data: 0,0-Diethyl-S(tert,butyl
thiomethyl) phosphorodiothiolate technical 85.8% AC 2162-42. Report
A-72-95. Unpublished study. MRID 37467.
Berger, H. 1977. Toxicology report on experiment L-1680 and L-1680-A:
Cholinesterase activity of dogs receiving Counter soil insecticide for
28 days. Toxicology Report No. A A77-158. Unpublished study. MRID 63189.
Consultox Laboratories. 1975. Acute oral and percutaneous toxicity evaluation.
Unpublished study. MRID 29863.
Daly, I., W. Rinehart and A. Martin. 1979. A three-month feeding study of
Counter terbufos insecticide in rats. Project No. 78-2343. Unpublished
Study. MRID 109446.
.- Devine, J.M., G.B. Kinoshita, R.P. Peterson and G.L. Picard. 1985. Farm
worJer exposure to terbufos during planting operations of corn. Arch.
Environ. Contarn. Toxicol. 15(1):113-120.
Godek, E.f R. Naismith and R. Mathews. 1983. Rat hepatocyte primary culture/
DNA repair test: (AC 92,100). PH 311-AC-001-83. Unpublished study.
MRID 133298.
Kruger, R., and H. Feinman. 1973. 30-Day subacute dermal toxicity in rabbits
of AC-92,100. Food and Drug Research Labs, Inc. July 17. Submitted to
American Cyanamid Co. Princeton, NJ. Unpublished study.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., C« Bull.
MacKenzie, K. 1984. Teratology study with AC 92,100 in rabbits. Study No.
6123-116. Unpublished study prepared by Hazelton Laboratories America, Inc.
MRID 147532.
McConnell, R. 1983. Twenty-four month oral toxicity and carcinogenicity
study in rats: AC 92,100: Pathology report. Unpublished study.
Biodynamics. April 22. MRID 130845.
Meister, R.T., ed. 1986. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Morgareidge, K., S. Sistner, M. Daniels et al. 1973. Final report: Six-month
feeding study in dogs on AC-92,100. Laboratory No. 1193. Unpublished
study. Food and Drug Laboratories, Inc. February 14. MRID 4"! 139.
-------
Terbufos August, 1987
-15-
North, N.H. 1973. Counter® insecticide: Rat metabolism of CL 92,100:
PD-M10:1008-1080. Progress report, March 1, 1973 through Sept. 28, 1973.
Unpublished study submitted by American Cyanamid Co., Princeton, NJ.
MRID 87695.
Parks, G.S.E., and Y. Terrell. 1976. Acute oral toxicity in rats: Compound:
Enlist technical insecticide (terbufos). EPA file symbol 2749-VEL.
Laboratory No. 6E-3164. Unpublished study. MRID 35121.
Peterson, R., G. Picard, J» Higham et al. 1984. Farm worter study with
aerial application of"counter 15-G. Report No. C-2370. Unpublished study.
MRID 137760.
Rapp, W., N. Wilson, M. Mannion et al. 1974. A three- and 24-month oral
toxicity and carcinogenicity study of AC-92,100 in rats. Project No.
71R-725. Unpublished study. Biodynamics, Inc. July 31. MRID 49236.
Rodwell, D. 1985. A teratology study with AC 92,100 in rats. Project No.
WIL-35014. Final report. Unpublished study prepared by WIL Research
Laboratories, Inc. MRID 147533.
Shellenberger, T. 1986. One-year oral toxicity study in purebred beagle
dogs with AC 92,100. Final report. Report No. 8414. Unpublished study.
Report No. 981-84-118. Prepared by Tegeris Laboratories, Inc. for
American Cyanamid Co., Princeton, NJ. MRID 161572.
Smith, J.M., and J. Kasner. 1972a. Status report for American Cyanamid Co.,
Nov. 28, 1972: A three-generation reproduction study of AC-92,100 in
rats. Project No. 71R-727. Unpublished study. MRID 37473.
Smith, J.M., and J. Kasner. 1972b. Status report for American Cyanamid Co.,
Nov. 24, 1972: An 18-month carcinogenicity study of AC-92,100 in
mice. Project No. 71R-728. Unpublished study.
STORET. 1987.
Thilager, A., P. Kumaroo and S. Knott. 1983. Chromosome aberration in Chinese
hamster ovary cells (test article AC-92,100). Microbiological Associate
Study No. T1906 337006. Sponsor Study No. 981-83-106. Unpublished study.
MRID 133296.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24,
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.352.
Windholz, M., S. Budvari, R.F. Blumetti and E.S. Otterbein. 1983. The Merck
Index, 10th ed. Rahway, NJ: Merck and Company.
-------
August, 1987
TRIFLURALIN
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 sub3ect 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«sk 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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Trifluralin
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1582-09-8
Structural Formula
N(CH2CH2CH,)2
alpha, alpha, alpha-Trifluro-2,6-dinitro-N,N-dipropyl-p-toluidine
Synonyms
• 2,6-Dinitro-N, N-dipropyl-4-trifluoromethylaniline; Agreflan; Crisalin;
Treflan; L-36352 Trifluralin (U.S. EPA, 1985a,b).
Uses
0 A selective herbicide (preemergent) for control of annual grasses and
broad-leafed weeds. Applied to soybean, cotton and vegetable crops;
fruit and nut trees, shrubs; and roses and other flowers. Also used
on golf courses, rights-of-way, and domestic outdoor and industrial
sites (U.S. EPA, 1985b).
Properties
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25"C)
C13"16F3N3°4
335.2
Orange, crystalline solid
139 to 140°C
46 to 49°C
1."\ x 10-* am Hg
0.3 mg/L
Log Octanol/Water Partition 4.69
Coefficient
Taste Threshold —
Odor Threshold --
Conversion Factor ~
Occurrence
0 Trifluralin is not a potential ground water contaminant due to its
strong adsorption to soil and negligible leaching (U.S. EPA, 1985b).
0 Trifluralin has been detected in finished drinking water supplies
(NAS, 1977).
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Trifluralin August, 1987
-3-
0 Trifluralin has been found in 318 of 377 surface water samples
analyzed and in 13 of 283 ground water samples (STORET, 1987).
Samples were collected at 194 surface water locations and 251 ground
water locations, and trifluralin was found in 9 states. The 85th
percentile of all nonzero samples was 0.10 ug/L in surface water and
.54 ug/L in ground water sources. The maximum concentration found was
16 ug/L in surface water and 0.54 ug/L in ground water.
Environmental Fate
0 Trifluralin at 5 ppm degraded with 15% of the applied trifluralin
lost after 20 days in a silt loam soil (aerobic metabolism) study
(Parr and Smith, 1973). The samples were incubated in the dark at
25°C and 0.33 bar moisture.
0 Trifluralin, applied alone or in combination with chlorpropham or
chlorpropham plus PPG-124, dissipated with a half-life of 42 to 84
days in sandy loam or silt loam soil incubated at 72 to 75°F and 18%
moisture content under laboratory conditions (Haliani, 1976).
0 In an anaerobic soil metabolism study, trifluralin at 5 ppm degraded
in nonsterile silt loam soil, with less than 1% of applied trifluralin
detected after 20 days of incubation (0.33 bar moisture in the dark
at 25°C; anaerobicity was maintained with nitroyen gas). Autoclaving
and flooding the soil decreased the degradation rate of the compound
(Parr and Smith, 1973).
• 14c-Trifluralin at 1.1 kg/ha was relatively immobile in sand, sandy
loam, silt, loam and clay loam soil columns (30-cm height) eluted
with 60 cm of water, with more than 90% of the applied radioactivity
remaining in the top 0- to 10-cm segment (Gray et al., 1982).
0 Trifluralin concentrations in runoff (water/sediment suspensions)
were less than 0.04% of the applied amount for 3 consecutive years
following treatment at 1.4 kg/ha and 13 to 27 cm of rainfall (Willis
et al., 1975). The field plots (silty clay loam soil, 0.2% slope)
were planted with cotton or soybeans.
0 In the field, Hc-trifluralin (99% pure) at 0.84 to 6.72 kg/ha dissipated
in the top 0- to 0.5 cm layer of a silt loam soil, with 14, 4, and 1.5%
of the applied amount remaining 1, 2 and 3 years, respectively, after
application (Golab et al., 1978). Approximately 30 minor degradates
were identified and quantified,- none represented more than 2.8% of
the applied amount. Trifluralin (4 Ib/gal EC) at 0.75 and 1.5 Ib/A
dissipated in a medium loam soil, with 20 and 32%, respectively, of
the applied remaining 120 days after treatment (Helmer et al., 1969;
Johnson, 1977).
e Trifluralin (4 Ib/gal EC) dissipated from a sandy loam soil treated
at 1.0 Ib ai/A, with a half-life of 2 to 4 months (Miller, 1973).
0 Trifluralin was detected in 107 soil samples taken nationwide at less
than 0.01 to 0.98 pj>m in fields treated with trifluralin at various
rates for 1, 2, 3 or 4 consecutive years (Parka and Tepe, 1969).
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Trifluralin August, 1987
-4-
0 Trifluralin was detected in 12% of the soil samples taken from 80
sites in 15 states in areas considered to be regular pesticide-use
areas based on available pesticide-use records (Stevens et al.,
1970). Concentrations detected in soils ranged from less than 0.01
to 0.48 ppm. Trifluralin residues were detected in only 3.5% of the
1,729 agricultural soils sampled in 1969 (Wiersma et al., 1972).
0 Trifluralin was detected at a maximum concentration of 0.25 ppm.
Residues of volatile nitrosamines (dimethylnitrosoamine, N-nitro-sodi-
propylamine, or N-butyl-N-ethyl-N-nitrosoamine) were not detected in
water samples taken from ponds and wells located in or near fields that
had been treated with trifluralin at various rates (Day et al., 1977).
III. PHARMACOKINETICS
Absorption
0 Emmerson and Anderson (1966) indicated that trifluralin is not readily
absorbed from the gastrointestinal (GZ) tract and that the fraction
that is absorbed is completely metabolized. Of an orally administered
dose (100 mg/kg), only 11 to 14% was excreted in the bile after 24
hours, indicating low GI absorption.
Distribution
0 No information was found in the available literature on the distri-
bution of trifluralin.
Metabolism
0 Four metabolites of trifluralin were identified in rats. Twelve rats
were given 100 mg/kg 14CF3-trifluralin in corn oil by gavage for 2
weeks. The metabolites, identified by thin-layer chromatography,
were produced by removal of both propyl groups or dealkylation and
reduction of a nitro group to an amine (Emmerson and Anderson, 1966).
0 An in vitro study using rat hepatic microsomes indicated that trifluralin
undergoes aliphatic hydroxylation of the N-alkyl substituents,
N-dealkylation and reduction of a nitro group (Nelson et al., 1976).
0 There are insufficient data to characterize the general metabolism of
trifluralin in animals (U.S. EPA, 1986a).
Excretion
Rats given an oral dose (100 mg/kg) of 14CF3-trifluralin excreted
virtually all of the dose within 3 days. The radioactivity was
excreted during the first 24 hours. Approximately 78% of the dose
was eliminated in the feces and 22% in the urine (Emmerson and
Anderson, 1966)*
" l\f
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Trifluralin August, 1987
-5-
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 The Pesticide Incident Monitoring System database revealed 105
incident reports involving trifluralin from 1966 to April of 1981.
Of the 105 reports, 49 cases involved humans exposed to trifluralin
alone. Twenty-seven cases involved human exposure to mixtures con-
taining trifluralin. The remaining incidents involved nonhuman
exposures (U.S. EPA, 1981a).
0 Among reports of human exposure to trifluralin alone, one fatality
was reported. A 9-year-old girl suffered cardiac arrest following
the ingestion of an unknown amount of trifluralin (U.S. EPA, 1981a).
0 Verhalst (1974) reported that the symptoms observed in trifluralin
poisonings appeared to be related to the Solvent used (e.g., acetone
or xylene) rather than trifluralin itself.
Long-term Exposure
0 The majority of reported trifluralin exposure cases were occupational
in nature. Trifluralin exposure has resulted in dermal and ocular
irritation in humans. Other reported symptoms include respiratory
involvement, abdominal cramps, nausea, diarrhea, headache, lethargy
and parasthesia following dermal or inhalation exposure. Specific
exposure levels or durations were not reported (U.S. EPA, 1981a).
Animals
Short-term Exposure
0 The acute oral toxicity of trifluralin is low. The following oral
LD50 values have been reported: mice >5 g/kg; rats >10 g/kg; dogs,
rabbits and chickens >2 g/kg (Meister, 1983; RTECS, 1985).
8 An inhalation LC5o value (41% trifluralin; species not specified) of
>2.44 mg/L/hour was reported (U.S. EPA, 1985c). No other information
was available.
Dermal/Ocular Effects
0 The results of a primary dermal-irritation study in the rabbit were
negative. No dermal irritation was observed at 72 hours following
application of a 41.2% trifluralin solution (U.S. EPA, 1985c).
0 Treflan, containing 10% trifluralin, was tested for sensitization in
female guinea pigs. A dose of 50 mg was applied to the skin of
12 animals, three times a week for 2 weeks. No dermal irritation or
contact sensitization developed during this time (ELANCO, 1984a).
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Trifluralin August, 1987
-6-
0 In a similar study, a 95% technical trifluralin solution was shown to be
a potential skin sensitizer in guinea pigs using the Buehler topical-
patch method (U.S. EPA, 1985c).
0 A 14-day study in which rabbits were exposed to 2 mL/kg trifluralin
topically produced diarrhea and slight dermal erythema in exposed
animals. No other effects were reported (ELANCO, 1979).
0 Technical-grade trifluralin applied as a powder to rabbit eyes was
reported as nonirritating. Slight conjunctivitis developed but
cleared within a week (U.S. EPA, 1985c).
0 When applied as a liquid to rabit eyes, technical trifluralin produced
corneal opacity that cleared in 7 days (U.S. EPA, 1985c).
Long-term Exposure
0 In a modified subacute study, female Harlan-Wistar rats were given 0,
0.05, 0.1 or 0.2% (0, 500, 1,000 or 2,000 ppm) trifluralin in their
diet for 3 months. Assuming that 1 ppm in the diet of rats equals
0.05 mg/kg/day (Lehman, 1959), these levels correspond to doses of
0, 25, 50 and 100 mg/kg/day. Physical appearance, behavior, body and
organ weights, mortality and clinical chemistries were monitored in
progeny from 10 females. No significant effects were observed in
survival or appearance. Liver weights in progeny continuously fed
diets of 0.1% and 0.2% trifluralin were increased over those of control
animals. The study identified a No-Observed-Adverse-Effect-Level
(NOAEL) in progeny of 0.05% (25 mg/kg) trifluralin (ELANCO, 1977a).
0 In a 90-day study, male F344 rats were fed dietary levels of 0 (n = 60),
0.005% (n = 60), 0.02% (n = 45), 0.08% (n = 45), 0.32% (n = 45) and
0.64% (n = 45). These concentrations are equivalent to dose levels of
0, 50, 200, 800, 3,200 and 6,400 ppm trifluralin, respectively (ELANCO,
1985). Assuming that 1 ppm in the diet of a rat equals 0.05 mg/kg/day
(Lehman, 1959), these levels correspond to doses of 0, 2.5, 10, 40,
160 and 320 mg/kg/day. After 90 days, alpha-1, alpha-2 and beta-
globulin levels were significantly increased in all treatment groups.
Other effects included increased aspartate transaminase, urinary
calcium, inorganic phosphorus and magnesium at levels y\60 mg/kg/day.
A Lowest-Observed-Adverse-Effect-Level (LOAEL) of 2.5 mg/kg/day (the
lowest dose tested) can be identified from this study.
0 Sixty weanling Harlan rats were fed 0, 20, 200, 2,000 or 20,000 ppm
trifluralin in the diet for 729 days (24 months). Assuming that
1 ppm in the diet of a rat equals 0.05 mg/kg (Lehman, 1959), these
concentrations correspond to doses of 0, 1, 10, 100 or 1,000 mg/kg/day.
No significant effects were observed in growth rate, mortality or
food consumption of treated animals at the three lower dose levels.
Animals in the highest dose group (1,000 mg/kg/day) were significantly
smaller than controls and ranked lower in food consumption. No effects
on hematology were noted. Animals in the high-dose group displayed a
slight proliferation of the bile ducts. No other histopathological
effects were observed. A NOAEL of 2,000 ppm (100 mg/kg/day) was
reported (ELANCO, 1966a).
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Trifluralin August, 1987
-7-
0 In a 2-year chronic carcinogenicity study with F344 rats, doses
greater than 128 mg/kg/day in males and 154 mg/kg/day in females were
reported to produce overt toxicity. Groups of 60 animals/sex/dose
were fed dietary levels of 0.08, 0.3 or 0.65% (30, 128 or 272 mg/kg/day
for males, and 37, 154 or 336 mg/kg/day for females) trifluralin. Body
weights of the high-dose groups were significantly decreased in both
sexes. This may be related to the decreased food consumption observed
in those groups. Increased blood urea nitrogen (BUN) levels and
increased liver and testes weights were note in the two high-dose
groups. Kidney and heart weights were significantly decreased in
females in the 0.3- and 0.65%-trifluralin groups. Other noncarcino-
genic effects included decreased hemoglobin values and erythrocyte
counts in both sexes of the high-dose group (BLANCO, 1980a). This
study appears to identify a NOAEL of 0.08% trifluralin (30 to 37 mg/kg^flay)
0 B6C3F1 mice (40/sex/group) were exposed to dose levels of 40, 180 or
420 mg/kg/day trifluralin in the diet for 2 years. Animals exposed
to the two higher levels exhibited decreased body weight and renal
toxicity. Other noncarcinogenic effects included decreased erythrocytic
and leukocytic values in the high-dose group, increased BUN and
alkaline phosphatase levels in the 180- and 420-mg/kg/day group,
decreased kidney weights in the two high-dose groups and decreased
spleen and uterine weights with increased liver weights in the high-
dose group (BLANCO, 1980b). No effects were noted at the low-dose
level (40 mg/kg/day).
• Occasional emesis and increased liver-to-body weight ratios were
observed in dogs (three/sex/dose) fed 25 mg/kg/day trifluralin for 3
years. No adverse effects were observed in animals fed 10 mg/kg/day
(Worth, 1970). An intermediate dose was not tested.
Reproductive Effects
e In a four-generation reproduction study (BLANCO, 1977b), rats were
given 0, 200 or 2,000 ppm trifluralin in the diet (0, 10 or 100
ing/kg/day). A reproductive NOAEL of 200 ppm (10 mg/kg/day) was
identified. The number of animals used in the study was not reported.
However, a review of this study (U.S. EPA, 1985c) indicated that an
insufficient number of animals were used and that several other
deficiencies in the study may have compromised the integrity of the
results.
0 In a 3-year feeding study in dogs a NOAEL of 10 mg/kg/day was
identified in adults (BLANCO, 1967). Dogs (three/sex/dose) were
given 10 or 25 mg/kg/day trifluralin in the diet. When bred after 2
years of exposure, no differences in litter size, survival or growth
of the pups were reported. An occasional emesis and increased liver
weights were reported in adults in the 25-mg/kg/day group.
Developmental Effects
0 Female rabbits (number not specified) were fed 0, 100, 225, 500, or
800 mg/kg/day by gavage during pregnancy (BLANCO, 1984b). No adverse
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Trifluralin August, 1987
-8-
reproductive effects were observed at the two lower dose levels.
The 500 and 800 mg/kg/day levels resulted in anorexia, aborted litters
and decreased live births. The NOAEL for maternal effects was identi-
fied as 225 mg/kg/day.
0 Rabbits (number not specified) exposed to 100, 225 or 500 mg/kg/day
trifluralin during pregnancy exhibited anorexia and cachexia at all
dose levels (U.S. EPA, 1985c). Aborted litters were observed at the
two high-dose levels. Fetotoxicity as evidenced by decreased fetal
weight and size was observed at the high-dose level.
0 In a rabbit teratology study, a total of 32 mated females were given
up to 1,000 mg/kg/day trifluralin by gavage (ELANCO, 1966b). Specific
dose increments were not reported. Animals were dosed until the 25th
day of gestation and then sacrificed. Does in the 1,000 mg/kg/day
group weighed slightly less than controls. Two fetuses were found to
be underdeveloped in the high-dose group; however, this was not
considered by the investigators to be treatment related. Average
litter size and weight were not significantly affected. The authors
reported that their results identified a safe level of 1,000 mg/kg/day.
0 Rabbit does (number per group not specified) were given 100, 225, 500
or 800 mg/kg/day trifluralin by gavage during pregnancy (ELANCO, 1964b)
The 500 and 800 mg/kg/day levels resulted in decreased live births,
cardiomegaly and wavy ribs in the progeny. No effects en progeny were
observed at 225 mg/kg/day or less (ELANCO, 1984b).
Mutagenicity
0 Anderson et al. (1972) reported that trifluralin did not induce point
mutations in any of the three microbial systems tested. No further
details were provided in the review.
0 Trifluralin was tested for genotoxicity in several in vivo and
in vitro systems (ELANCO, 1983). No reverse mutations were observed
in Salmonella typhimurium or Escherichia coli when incubated with 25
to 400 mg trifluralin/plate without activation; trifluralin was also
negative when tested at levels of 50 to 800 mg/plate with activation.
Negative results were obtained in mouse lymphoma L5178Y TK+ cells
incubated with 0.5 to 20 ug/mL trifluralin with and without activation.
An in vivo sister-chromatid exchange study in Chinese Hamster Ovary
(CHO) cells following exposure to 500 mg/kg trifluralin was also
negative.
Carcinogenicity
0 NCI (1978) conducted bioassays on B6C3F1 nice and Osborne-Mendel rats
using technical-grade trifluralin (which contained 84 to 88 ppm of the
contaminant dipropylnitrosamine). Two dietary levels were used in
each bioassay. Mice (50/sex/group) were exposed to trifluralin at
dose levels of 2,000 or 3,444 ppm (males) or 3,740 or 5,192 ppm
(females) for 78 weeks and observed for an additional 13 weeks after
exposure. A significant close-related increase in hepatocellular
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Trifluralin Au9ust' 1987
-9-
carcinoma was observed in female mice (0/20 control, 12/47 low dose,
21/44 high dose). An increased incidence of alveolar/ bronchiolar
adenomas was also observed (0/19 control, 6/43 low dose, 3/30 high
dose) in female mice. Squamous cell carcinomas in the forestomach of
a few treated female mice were also observed. Although the incidence
of squamous cell carcinoma in the forestomach was not statistically
significant when compared to pooled and matched controls, NCI deemed
this finding to be treatment related, since it was an unusual type of
lesion. Male mice were not significantly affected by trifluralin
exposure.
0 Rats (50/sex/group) were exposed to two levels of trifluralin in the
feed (4,125 or 8,000 ppm for males; 4,125 or 7,917 ppm for females)
for 78 weeks followed by a 33-week observation period (NCI, 1978).
Assuming 1 ppm in the diet of rats equals 0.05 mg/kg/day (Lehman,
1959), these doses correspond to 206 or 400 mg/kg/day. Several
neoplasms were observed and compared to pooled and matched controls.
These neoplasm types were reported to occur spontaneously in the
Osborne-Mendel strain and were not considered treatment related by
NCI.
0 In a 2-year feeding study, B6C3F1 mice were given 563, 2,250 or 4,500
ppm trifluralin (assuming 1 ppm in the diet of a mouse equals 0.15
mg/kg/day, these doses correspond to 40, 180 or 420 mg/kg/day (Lehman,
1959) in the diet (ELANCO, 1980b). Levels of a nitrosamine contaminant
of trifluralin, NDPA, were below the 0.01-ppm analytical detection
limit. A total of 40 animals/sex/treatment group was used. At the
lowest dose level, 40 mg/kg/day, no adverse effects were observed in
either sex. Decreased body weight and renal effects were noted in
mice in the mid- and high-dose groups. Pathology revealed progressive
glomerulonephritis in females of the high-dose group. Hepatocellular
hyperplasia and hypertrophy were also observed in the treated mice.
The specific dose level was not reported. No evidence of increased
incidence or decreased latency for any type of neoplasm was found in
any of the mice.
0 Trifluralin was administered to F344 rats (60/sex/group) at dose
levels of 813, 3,250 or 6,500 ppm [assuming 1 ppm in the diet of a
rat equals 0.05 mg/kg/day (Lehman, 1959), these doses correspond to
30, 128 or 272 mg/kg/day for males and 37, 154 or 336 mg/kg/day for
females] in the diet for 2 years (ELANCO, 1980a). A significant
increase in malignant renal neoplasms and thyroid tumors in male rats
and in neoplasms of the bladder in both sexes was reported. A high
incidence (20/30) of renal calculi was also observed in animals in
the high-dose groups.
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:
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Trifluralin August, 1987
-10-
HA = (NOAEL or LOAEL) X (BW) a mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW a 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
No information was found in the available literature that was suitable
for determination of the One-day HA value for trifluralin. Therefore, it is
recommended that a modified OWEL (0.025 mg/L, calculated below) for a 10-kg
child be used as a conservative estimate for the One-day HA value.
For a 10-kg child, the adjusted OWEL is calculated as follows:
DWEL - (0-0025 mq/kq/day) (IP kg) = Q.025 mg/L
1 L/day
where:
0.0025 mg/kg/day = Rfd (see Lifetime Health Advisory Section).
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of a child.
Ten-day Health Advisory
No information was found in the available literature that was suitable
for determination of the Ten-day HA value for trifluralin. It is, therefore,
recommended that a modified JWEL (0.025 mg/L) for a 10-kg child be used as a
conservative estimate for the Ten-day HA value.
Longer-term Health Advisory
No information was found in the available literature that was suitable
for determination of the Longer-term HA value for trifluralin. It is, therefore,
recommended that a modified DWEL (0.025 mg/L) for a 10-kg child be used as a
conservative estimate for a Longer-term exposure.
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Trifluralin August, 1987
-11-
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
(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 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 BLANCO (1985) study has been selected to serve as the basis for the
Lifetime HA value for trifluralin. F344 rats were fed diets containing
0.005, 0.02, 0.08, 0.32 or 0.64% trifluralin (2.5, 10, 40, 160 or 320
mg/kg/day) for 90 days. Significant increases in urinary alpha-1, alpha-2,
and beta-globulins were observed in all treated animals. A NOAEL was not
identified. Other longer-term studies report NOAELs at higher doses.
Using a LOAEL of 2.5 mg/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (2.5 mg/kg/day) _ 0.0025 mg/kg/day
(1,000)
where: n
2.5 Lig/kg/day = LOAEL, based on increased urinary globulins in rats
consuming a trifluralin diet for 3 months.
1,000 B uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a LOAEL from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0025 mg/kg/day) (70 kg) „ Q.088 mg/L (87 ug/L)
(2 L/day)
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Trifluralin August, 1987
-12-
where:
0.0025 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»088 mg/L) (20%) = Q.0017 mg/L (2 ug/L)
10
where:
0.088 mg/L *• OWEL.
20% » assumed relative source contribution from water.
10 » additional uncertainty factor per ODW policy to
account for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986b), trifluralin may be classified
in Group C: possible human carcinogen. This category is used for
substances that show limited evidence of carcinogenicity in animals
and inadequate evidence in humans.
0 In an NCI (1978) study of female B6C3F1 mice, a significant dose-
related increase in hepatocellular carcinomas and alveolar adenomas
was observed when the animals were exposed to 33 or 62 mg/kg/day
trifluralin in the diet for 78 weeks. The trifluralin used in this
study contained 84 to 88 ppm dipropylnitrosamine. Male rats, when
exposed to 30, 128 or 272 mg/kg/day trifluralin in the diet for 2
years, exhibited significant increases in the incidences in kidney,
urinary bladder and thyroid tumors (ELANCO, 1980a).
0 The evidence from the ELANCO (1980a) and NCI (1978) studies indicates
that trifluralin has carcinogenic potential. Based of 7.66 x
10-3 mg/kg/day based on the combined incidence of tumors in male rats.
Assuming that a 70-kg human adult consumes 2 liters of water a day
over a 70-year lifespan, the estimated cancer risk would be 10-4,
10-5 and 10-6 at concentrations of 500, 50 and 5 ug/L, respectively.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Residue tolerances from 0.05 to 2.0 ppm trifluralin have been established
for a variety of agricultural commodities (U.S. EPA, 1985).
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Trifluralin August, 1987
-13-
0 NAS (1977) has calculated an ADI of 0.1 mg/kg bw/day with a Suggested-
No-Adverse-Response-Level (SNARL) of 700 ug/L.
VII. ANALYTICAL METHODS
e Determination of trifluralin is by a liquid-liquid extraction gas
chromatographic procedure applicable to the determination of organo-
chlorine pesticides in water samples (Standard Methods, 1965).
Specifically, the procedure involves extraction with a mixed solvent,
diethyl ether/hexane or methylene chloride/hexane. The extract is
concentrated by evaporation, and the compounds are separated by gas
chromatography. Detection and measurement are accomplished by the
use of an electron-capture detector. Additional confirmatory identi-
fication can be made through the use of two unlike columns or by mass
spectrome try.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that reverse osmosis (RO), granular-activated
carbon (GAG) adsorption conventional treatment and possibly air
stripping will remove trifluralin from water.
0 U.S. EPA investigated the amenability of a number of compounds, including
trifluralin, to removal by GAG. No system performance data were given.
0 Conventional water treatment techniques of coagulation with alum,
sedimentation and filtration proved to be 100% effective in removing
trifluralin from contaminated water (Nye, 1984).
0 Sanders and Seibert (1983) determined experimentally water solubility,
vapor pressure, Henry's Law Constant and volatilization rates for
trifluralin; 100% of the compound volatilized under laboratory
conditions.
0 Treatment technologies for the removal of trifluralin from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt trifluralin
removal from water must be based on a case-by-case technical evaluation,,
and an assessment of the economics involved.
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Trifluralin August, 1987
-14-
IX. REFERENCES
Anderson, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of
herbicides for possible mutagenic properties. J. Agric. Food Chem.
20:649-656 (cited in U.S. EPA, 1985a).
Day, E., S. West and M. Amundson. 1977. Residues of volatile nitrosamines
in water samples from fields treated with Treflan: Pre-RPAR Review
submission #8. Unpublished study submitted by Elanco Products Company
to the Office of Pesticide Porgrams, Division of Eli Lilly and Company,
Indianapolis, IN.
ELANCO. 1966a.* Eli Lilly and Company. Chronic toxicity studies with
trifluralin. MRID 76447.
ELANCO. 1966b.* Eli Lilly and Company. Teratology studies with trifluralin.
MRID 83647.
ELANCO. 1967.* Eli Lilly and Company. Effects of trifluralin treatment on
reproduction in rats and dogs. MRID 83646.
ELANCO. 1977a.* Eli Lilly and Company. A modified subacute toxicity study
with trifluralin. MRID 134326.
ELANCO. 1977b.* Eli Lilly and Company. Effect of trifluralin treatment on
reproduction in rats and dogs. MRID 83646.
ELANCO. 1980a.* Eli Lilly and Company. The chronic toxicity of compound 36352
(trifluralin) given as a component of the diet to Fischer 344 rats for
two years. MRID 4437.
ELANCO. 1980b.* Eli Lilly and Company. The chronic toxicity of compound 36352
(trifluralin) given as a component of the diet of B6C3F, mice for 24 months.
MRID 4438.
ELANCO. 1983.* Eli Lilly and Company. Genetic toxicology studies with
trifluralin (compound 36352). MRID 126659.
ELANCO. 1984a.* Eli Lilly and Company. Guinea pig sensitization study of
treflan 1OG. A granular formulation (FN-1199) containing 10% trifluralin.
MRID 137468.
ELANCO. 1984b.* Eli Lilly and Company. Teratology study in rabbits (cited in
US EPA, 1985a).
ELANCO. 1985.* Eli Lilly and Company. Special urinalysis study in Fischer
344 rats maintained on diets containing trifluralin (compound 36352)
for 33 months.
Emmerson, J.L. and R.C. Anderson. 1966. Metabolism of trifluralin in the
rat and dog. Toxicol. Appl. Pharmacol. 9:84-97.
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Trifluralin August, 1987
-15-
Golab, T., W. Althaus and H. Wooten. 1978. Fate of !4C-trifluralin in soil.
Unpublished study submitted by ELANCO Products Company, Division of Eli
Lilly and Company, Indianapolis, IN.
Gray, J.E., A. Loh, R.F. Sieck et al. 1982. Laboratory leaching of ethyl-
fluralin. Unpublished study submitted by Elanco Products Company,
Division of Eli Lilly and Company, Indianpolis, IN.
Helmer, J.D., W.S. Johnson and T.W. Waldrep. 1969. Experiment No. WB(F)
9-132: Soil persistence data. Unpublished study submitted by Elanco
Products Company, Division of Eli Lilly and Company, Indianapolis, IN.
Johnson, W. 1977. Determination of trifluralin in agricultural crops and
soil: Procedure No. 5801616. Unpublished study submitted by Elanco
Products Company, Division of Eli Lilly and Company, Indianapolis, IN.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Assoc. Food Drug Off.
Haliani, N. 1976. CIPC and CIPC + PPG-124 interaction study (Exhibit E):
Laboratory No. 97021. Unpublished study prepared by Morse Laboratories,
Inc., submitted by PPG Industries, Barberton, OH.
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Miller, J.H. 1973. Residue Report AGA 2527 - 2nd Report Project No. 120002.
Hosier, J., and D. Saunders. 1978. A hydrolysis study on the herbicide
trifluralin. Unpublished study submitted by Elanco Products Company,
Division of Eli Lilly and Company, Indianapolis, IN.
NAS. 1977. National Academy of Science. Drinking water and health. Vol. I.
Washington, DC: National Academy Press.
NCI. 1978. National Cancer Institute. Bioassay of trifluralin for possible
carcinogenicity. NCI-CG-TR-34.
Nelson, J.O., P.C. Kearney, J.R. Plimmer and P.E. Menzer. 1976. Metabolism
of trifluralin, profluralin and fluchloralin by rat liver microsomes.
Pest. Biochem. Phys. 7:73-82 (cited in U.S. EPA, 1985a).
Nye, J.C. 1984. Treating pesticide-contaminated wastewater development and
evaluation of a system. American Chemical Society.
Parka, S. and J. Tepe. 1969. The disappearance of trifluralin from field
soils. Heed Sci. 17(1):119-122.
Parr, J.F. and S. Smith. 1973. Degradation of trifluralin under laboratory
conditions and soil anaerobiosis. Soil Sci. 115(1):55-63.
RTECS. 1985. Registry of Toxic Effects of Chemical Substances. National
Institute of Occupational Safety and Health. Washington, DC.
-------
Trifluralin August, 1987
-16-
Sanders, P.P. and J.N. Seibert. 1983. A chamber for measuring volatilization
of pesticides from model soil and water disposal systems. Chemosphere.
12(7/8):999-1012.
Standard Methods. 1985. Method 509A, Organochlorine Pesticides, Standard
Methods for the Examination of Water and Waste water, 16th ed. APHA,
AWWA, WPCF.
Stevens, L., C. Collier and D. Woodham. 1970. Monitoring pesticides in
soils from areas of regular, limited, and no pesticide use. Pestic.
Monit. J. 4(3):145-166.
STORET. 1987.
U.S. EPA. 1981a. U.S. Environmental Protection Agency. Summary of reported
incidents involving trifluralin. Pesticide Incident Monitoring System.
Report no. 441. Office of Pesticide Programs, Washington, DC.
U.S. EPA. 1981b. U.S. Environmental Protection Agency. Carcinogenic potency
for trifluralin, including N-nitroso-di-n-propylamine (NDPA) and diethyl-
itrosamine (DENA). Memo from Chao Chen and Bernard Haberman to Marcia
Williams. July 29.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal Regu-
lations. 40 CFR 180.201.
U.S. EPA. 1985a. U.S. Environmental Protection Agency. Pesticide survey
chemical profile. Final report. Contract no. 68-01-6750. Office of
Drinking Water, Washington, DC.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Post phase II regis-
tration standard support team meeting for trifluralin. I. Regulatory
position and rationale. Memo from Robert Ikeda to Registration Standard
Support Team. September 4.
U.S. EPA. 1985c. U.S. Environmental Protection Agency. Trifluralin in
registration standard. Toxicology chapter. Memo from Roland Gessert to
Richard Montfort. June 25.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Draft guidance for the
registration of pesticide products containing trifluralin.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Guidelines for car-
cinogen risk assessment. 51 FR 33992. September 24.
Verhalst, H. 1974. Personal communication to Eli Lilly and Company (cited
in U.S. EPA, 1985a).
Whittaker, K.F. et al. 1982. Collection and treatment of wastewater
generated by pesticide applicators. EPA-600/2-82-028.
Wiersma, G.B., H. Tai and P.P. Sand. 1972. Pesticide residue levels in
soils, FY 1969—National Soils Monitoring Program. Pestic. Monit. J.
6(3):194-201.
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Trifluralin Au9ust' 1987
-17-
Willis, G.H., R.L. Rogers and L.M. Southwick. 1975. Losses of diuron,
linuron, fenac, and trifluralin in surface drainage water. J. Environ.
Qual. 4(3):399-402.
North, H.M. 1970. The toxicological evaluation of benomyl and trifluralin.
Pesticide Symposia 6th Conference, August, 1966. Halos and Assoc., Miami,
FL. pp. 263-267.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1987
2,4,5-TRICHLOROPHENOXYACETIC ACID
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 acc-urately 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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2,4,5-Trichlorophenoxyacetic Acid
August, 1987
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 93-76-5
Structural Formula
0-CHtCOOH
2,4,5-trichlorophenoxyacetic acid
Synonyms
Uses
2,4,5-T; Brush rhap; Brushtox; BCF-Bushkiller; Dacamine; Decamine 4T;
Oed-Weed; Dinoxol; Envert-T; Estercide t-2 and t-245; Esteron; Fence
rider; Forron; Forst U46; Fortex; Fruitone A; Inverton 245; Line
rider; Phortox; Reddon; Reddox; Spontox; Tippon; Torraona; Transamine;
Tributon; Trinoxol; Trioxon; Veon 245; Verton 2T; VEON; Weedar;
Weedone (Meister, 1983).
0 Salts and esters of 2,4,5-T are widely used to control woody plants
on industrial sites and rangeland. Amine formulations are used
extensively 'for weed control in rice (Meister, 1983).
Properties (BCPC, 1983; Heister, 1983; Windholz et al., 1983; Khan, 1985;
CHEMLAB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C8H503C13
255.49
Crystals
153°C
6.46 x 10-6 am Hg
Solubility of acid is 150 g/L; amine
salts are soluble at 189 g/L (20°C);
esters are insoluble
3.00 (calculated)
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-3-
Occurrence
0 2,4,5-T has been found in 5,009 of 24,516 surface water samples
analyzed and in 360 of 3,238 ground water samples (STORET, 1987).
Samples were collected at 3,967 surface water locations and 2,124
ground water locations, and 2,4,5-T was found in 45 states. The 85th
percentile of all nonzero samples was 0.1 ug/L in surface water and
1 ug/L in ground water sources. The maximum concentration found was
370 ug/L in surface water and 38 ug/L in ground water.
Environmental Fate
0 No information was found in the available literature on the environ-
mental fate of 2,4,5-T.
III. PHARMACOKINETICS
Absorption
0 In a study by Gehring et al. (1973), single oral doses of 5 mg/kg
2,4,5-T were ingested by five male volunteers. Essentially all
the 2,4,5-T was excreted unchanged via the urine, indicating that
gastrointestinal absorption was nearly complete.
0 Fang et al. (1973) administered single doses of 14c-labeled 2,4,5-T
in corn oil by gavage to pregnant and nonpregnant female Wistar rats
at dose levels of 0.17, 4.3 or 41 mg/kg. Expired air, urine, feces,
internal organs and tissues were analyzed for radioactivity. During
the first 24 hours, an average of 75 ±7% of the radioactivity was
excreted in the urine, indicating that at least 75% of the dose had
been absorbed.
0 Piper et al. (1973) administered single oral doses of 14C-labeled
2,4,5-T in corn oil-acetone (9:1) to adult female Sprague-Dawley rats
at dose levels of 5, 50, 100 or 20 mg/kg, and to adult female beagle
dogs at 5 mg/kg. Fecal excretion was 3% at the lowest dose (5 mg/kg)
and increased to 14% at the highest dose (200 mg/kg) in rats. In
dogs given the 5 mg/kg dose, fecal excretion was 20%. These data
indicated that absorption was somewhat dose dependent, but was 80% or
higher at all doses.
Distribution
0 Gehring et al. (1973) administered single oral doses of 5 mg/kg of
2,4,5-T to five male volunteers. Essentially all the 2,4,5-T was
absorbed in the body; 65% of the absorbed dose resided in the plasma
where 98.7% was bound reversibly to protein. The volume of distribution
was 0.097 L/kg. Utilizing the kinetic constants from the single-dose
experiment, the expected concentrations of 2,4,5-T in the plasma
of individuals receiving repeated doses of 2,4,5-T were calculated.
From these calculations, it was determined that the plasma concentra-
tions would essentially reach a plateau value after 3 days. If the
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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daily dose ingested in mg/kg is A0, the concentrations in the plasma
after attaining plateau would range from 12.7 AQ to 22.5 AQ ug/mL
(Gehring et al., 1973).
0 Fang et al. (1973) administered single oral doses of 14c-labeled
2,4,5-T to pregnant and nonpregnant female Wistar rats and internal
organs and tissues were analyzed for radioactivity. Radioactivity
was detected in all tissues, with the highest concentration found in
the kidney. The maximum concentration in all tissues was generally
reached between 6 and 12 hours after administration of the dose
(0.17, 4.3 or 41 mg/kg) by gavage, and then started to decline rapidly.
Radioactivity also was detected in the fetuses and in the milk of the
pregnant rats. The average biological half-life of 2,4,5-T in the
organs was 3.4 hours for the adult rats and 97 hours for the newborn.
0 Piper et al. (1973) administered single oral doses of 5, 50, 100 or
200 mg/kg 2,4,5-T to Sprague-Dawley rats, and found that the apparent
volume of distribution increased with dose, indicating that distribution
of 2,4,5-T in the body was dose-dependent.
Metabolism
0 Gehring et al. (1973) administered single oral doses of 5 mg/kg
2,4,5-T to human volunteers. Essentially all the chemical was
excreted in the urine as parent compound, indicating that there is
little metabolism of 2,4,5-T in humans.
0 Grunow et al. (1971) investigated the metabolism of 2,4,5-T in male
Wistar (AF/Han) rats after receiving single oral doses of 50 mg/kg.
The 2,4,5-T was dissolved in peanut oil and administered by gavage.
Urine was collected for 7 days after dosing and examined by gas
chromatography for 2,4,5-T and its conjugates and metabolites. From
45 to 70% of the administered dose was recovered in urine. In general,
about 10 to 30% of this was as acid-hydrolyzable conjugates, and the
remainder was unchanged 2,4,5-T. Three animals were given doses of
75 mg/kg, and their urine pooled. A metabolite isolated from this
pooled urine was identified as N-(2,4,5-trichlorophenoxy-acetyl)-
glycine.
0 Piper et al. (1973) administered single oral doses of 2,4,5-T to
female Sprague-Dawley rats at dose levels of 5, 50, 100 or 200 m<:/kg.
A small amount of an unidentified metabolite was detected in urine at
the high doses, but not at the lower doses. In adult beagle dogs
given oral doses of 5 mg/kg, three unidentified metabolites were
detected in urine, suggesting a difference in metabolism between rats
and dogs.
0 In a study by Fang et al. (1973) in female Wistar rats, urinalysis
revealed that 90 to 95% of the radioactivity was unchanged 2,4,5-T.
The authors also found three unidentified minor metabolites, two of
which were nonpolar, in the urine.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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Excretion
0 In a study by Gehring et al. (1973), single doses of 5 mg/kg 2,4,5-T
were ingested by five male volunteers. The concentrations of 2,4,5-T
in plasma and its excretion were measured at intervals after ingestion.
The clearances from the plasma, as well as the body, occurred via
apparent first-order rate processes with half-lives of 23.1 and 29.7
hours, respectively. Essentially all the 2,4,5-T was excreted
unchanged via the urine.
• In a study by Fang et al. (1973), 2,4,5-T labeled with 1 4c was orally
administered to pregnant and nonpregnant female Wistar rats at various
dosages, and expired air, urine and feces were analyzed for radio-
activity. During the first 24 hours, 75 ± 7% of the radioactivity
was excreted in the urine and 8.2% was excreted in the feces. No 14C
was found in the expired air. There was no significant difference in the
rate of elimination between the pregnant and nonpregnant rats, or
among the dosages used (0.17, 4.3 and 41 mg/kg). The average biological
half-life of 2,4,5-T in the organs was 3»4-hours for the adult rats
and 97 hours for the newborn.
0 Grunow et al. (1971) investigated the excretion of 2,4,5-T in male
Wistar (AF/Han) rats after single oral doses of 50 mg/kg. The 2,4,5-T
was dissolved in peanut oil and administered by gavage. From 45 to
70% of the administered dose was recovered in urine within 7 days.
0 Clearance of 14C activity from the plasma and its elimination from
the body of rats and dogs were determined after single oral doses of
labeled 2,4,5-T (Piper et al., 1973). The half-life values for the
clearance of radioactivity from the plasma of Sprague-Dawley (Spartan
strain) rats given doses of 5, 50, 100 or 200 mg/kg were 4.7, 4.2,
19.4 and 25.2 hours, respectively; half lives for elimination from
the body were 13.6, 13.1, 19.3 and 28.9 hours, respectively. Urinary
excretion of unchanged 2,4,5-T accounted for 68 to 93% of the radio-
activity eliminated from the body of the rats. Fecal excretion was
3% at 5 mg/kg, and increased to 14% at 200 mg/kg. These results
indicate that the excretion of 2,4,5-T is altered when large doses
are administered. In adult beagle dogs given doses of 5 mg/kg, the
half-life values for clearance from plasma and elimination from the
body were 77.0 and 86.6 hours, respectively. After 9 days, 11% of
the dose was recovered in urine and 20% was recovered in feces.
IV. HEALTH EFFECTS
0 Technical 2,4,5-T contains traces of the highly toxic compound
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) as an impurity (HAS, 1977).
Preparations of 2,4,5-T formerly contained TCDD at levels of 1 to 80
ppai, a concentration sufficiently high to cause toxic effects that
are characteristic of TCDD. It has not been feasible to completely
eliminate TCDD from technical 2,4,5-T, but HAS (1977) reported it to be
present in commercial 2,4,5-T at less than 0.1 ppm. In the following
sections, the purity of 2,4,5-T or the level of TCDD impurity is
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-6-
given when known. When the generic term "dioxin" is used, no further
information was provided, and the 2,4,5-T is presumed to contain a
variety of dioxin species as well as other phenoxy compounds and
assorted intermediates and breakdown products.
Humans
^•MI^^B
Short-term Exposure
0 No clinical effects were observed in five volunteers who ingested
single oral doses of 5 mg/kg of 2,4,5-T (Gehring et al., 1973).
0 After an explosion in a chemical plant producing 2,4,5-T in 1949,
symptoms in exposed workers included chloracne, nausea, headache,
fatigue, and muscular aches and pains (Zack and Suskind, 1980).
Long-term Exposure
0 The mortality experience in a cohort of 1,926 men who had sprayed
2,4,5-T acid during 1955 to 1971 was followed prospectively from 1972
to 1980. Exposure was generally rather low because the duration of
work had mostly been less than 2 months. In the period 1972 to 1976,
mortality from all natural causes in this group was only 54% of the
expected value (based on age-specific rates for the general population),
and in the next 4-year period, 81% of the expected value. In the
assessment of cancer, mortality allowance was made for 10- and 15-year
periods of latency between the first exposure and the start of the
recording of vital status during the followup. No increase in cancer
mortality was detected, and the distribution of cancer types was
unremarkable. No cases of death from lymphomas or soft tissue sarcomas
were found. It was noted, however, that the study results should be
interpreted with caution due to the small size of the cohort, the low
past exposure, and the brief followup period (Riihimaki et al., 1982).
0 An investigation of the rate of birth malformations in the Northland
region of New Zealand was analyzed with reference to the exposure in
the area to 2,4,5-T, which was applied as frequently as once a month
from 1960 to 1977. The chosen area was divided into sectors rated as
high, intermediate or low, based on the frequency of aerial spraying.
During this period, there were 37,751 babies born in the hospitals in
these sectors. It was estimated that well over 99% of all births
occur in hospitals in this Nortnland area. The epidemic logical
analysis of the birth data gave no evidence that any malformation of
the central nervous system, including spina bifida, was associated
with the spraying of 2,4,5-T. Heart malformations, hypospadias, and
epispadias increased with spraying density, but the increases were
not statistically significant (p >0.05). The only anomaly that
increased in a statistically significant (p <0.05) manner with respect
to the spraying was talipes (club foot) (Hanify et al., 1981).
0 The relationship between the use of 2,4,5-T in Arkansas and the
concurrent incidence of facial clefts in children was studied retro-
spectively. The estimated levels of exposure were determined by
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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categorizing the 75 counties into high, medium and low exposure groups
on the basis of their rice acreage during 6- to 7-year intervals
beginning in 1943. A total of 1,201 cases of cleft lip and/or cleft
palate for the 32 years (until 1974) was detected by screening birth
certificates and hospital records. Facial cleft rates, presented by
sex, race, time period and exposure group, generally rose over time.
No significant differences were found for any race or sex combination.
The investigators concluded that the general increase seen in facial
cleft incidence in the high- and low-exposure groups was attributable
to better case finding rather than maternal exposure to 2,4,5-T
(Nelson et al., 1979).
0 Ott et al. (1980) reported no effects in a survey of 204 workers
engaged in 2,4,5-T production at estimated airborne levels of 0.2 to
0.8 mg/m3 for 1 month to 10 years.
0 Numerous epidemiological studies on the relationship between exposure
to chlorophenoxyacetic acids and cancer induction are reviewed in
U.S. EPA (1985). The conclusion in this review is that there is
"limited" evidence for the carcinogenicity of chlorinated phenoxyacetic
herbicides and/or chlorophenols with chlorinated dibenzodioxin impuri-
ties, primarily based on Swedish case-control studies that associated
induction of soft-tissue sarcomas with exposure to these agents.
Animals
Short-term Exposure
0 The acute oral toxicity of 2,4,5-T was determined in mice, rats and
guinea pigs by Rowe and Hymas (1954) over a 2-week period. The LD5Q
values were 500 mg/kg for rats, 389 mg/kg for mice and 381 mg/kg for
guinea pigs.
0 Drill and Hiratzka (1953) investigated the acute oral toxicity of
2,4,5-T in adult mongrel dogs given single oral doses of 50, 100, 250
or 400 mg/kg by gelatin capsule. Animals were observed for 14 days,
at which time survivors were necropsied. The number of deaths at the
four dose levels were 0/4, 1/4, 1/1 and 1/1, respectively. The LD50
value was estimated to be 100 mg/kg or higher. Marked changes were
not observed in animals that died, effects being limited to weight
loss, slight to moderate stiffness in the hind legs and ataxia (at
the highest doses).
0 Weanling male Wistar rats were fed diets containing 2,4,5-T for 3
weeks to investigate effects on the immune system (Vos et al., 1983).
2,4,5-T (>99% purity, TCDD content not specified) was fed at levels
of 200, 1,000 or 2,500 ppm (approximately 20, 100 or 250 mg/kg/day,
assuming 1 ppm equals 0.1 mg/kg/day in a younger rat by Lehman, 1959).
Following the 3-week feeding period, the animals were sacrificed and
the organs of the immune system, as well as other parameters of
general toxicity, were examined. Even at the lowest dose level of
200 ppm in the diet, 2,4,5-T caused a significant (p <0.05) decrease
in relative kidney weight and a significant (p <0.05) increase in
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-8-
serum igG level, the most sensitive indicators of its effects. In
this study, based on general toxicologic and specific immunologic
effects in the rat, the Lowest-Observed-Adverse-Effect-Level (LOAEL)
was 20 mg/kg/day.
Dermal/Ocular Effects
0 Gearing and Betso (1978) summarized the effects of 2,4,5-T on the skin
and the eye. The dry material is slightly irritating to the skin and
the eye. Highly concentrated solutions may burn the skin with
prolonged or repeated contact and can strongly irritate the eye and
possibly cause corneal damage. Preparations of 2,4,5-T formerly
contained 1 to 80 ppm 2,3,7,8-TCDD, a concentration high enough to
cause chloracne in industrial workers (NAS, 1977).
Long-term Exposure
0 Drill and Hiratzka (1953) investigated the subchronic toxicity of
2,4,5-T in adult mongrel dogs. One or two dogs of each sex per group
were fed capsules in food containing 0, 2, 5, 10 or 20 mg/kg 2,4,5-T,
5 days per week for 13 weeks. Animals were weighed twice weekly, and
blood was taken on days 0, 30 and 90. Upon death or completion of
the study, animals were necropsied with histological examination of
a number of tissues. No deaths occurred at doses of 10 mg/kg/day or
less, but 4/4 animals receiving 20 mg/kg/day died. No effects on
body weight, hematology and pathology were seen except in animals
that died. The No-Observed-Adverse-Effect-Level (NOAEL) was identified
as 10 mg/kg/day.
0 McCollister and Kociba (1970) examined the effects of 2,4,5-T admini-
stered in the diet for 90 days to male and female Sprague-Dawley rats
(Spartan strain). The 2,4,5-T (99.5% pure, <0.5 ppm dioxin) was
included in the diet at levels corresponding to doses of 0, 3, 10, 30
or 100 mg/kg/day. Five animals of each sex were used at each dose
level. At the conclusion of the study, necropsy, urinalyses, blood
counts and clinical chemistry assays were performed. There was no
mortality in any group. At 100 mg/kg, animals of both sexes had
depressed (p <0.05) body weight gain, a slight but significant
(p <0.05) decrease in food intake and elevated (p <0.05) serum alkaline
phosphatase (AP) levels. Necropsy revealed paleness and an accentuated
lobular pattern of the liver, with some inconsistent hepatocellular
swelling. Males (but not females) had slightly elevated serum glutamic-
pyruvic transaminase (SGPT) levels, and slight decreases in red blood
cell counts and in hemoglobin. Males given 100 mg/kg/day had increased
(p <0.05) kidney/body and liver/body weights. At the 30 mg/kg/day
dose level, males exhibited increased (p <0.05) liver/body, kidney/body,
and kidney weights. Females given 30 mg/kg/day had sligntly but
significantly (p <0.05) elevated AP and SGPT levels, but the authors
felt that the clinical significance of these latter findings was
doubtful. No effects observed at the 3 or 10 mg/kg dose level were
considered to be related to the intake of 2,4,5-T. From this study,
a NOAEL of 10 mg/kg/day and a LOAEL of 30 mg/kg/day were identified.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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0 Groups of Sprague-Dawley rats (50/sex/level) were maintained on diets
supplying 3, 10 or 30 mg/kg/day of 2,4,5-T for 2 years (Kociba et al.,
1979). The 2,4,5-T was approximately 99% pure, containing 1.3% (w/w)
other phenoxy acid impurities. Dioxins were not detected, the limit
of detection for TCDD being 0.33 ppb. An interim sacrifice was
performed on an additionally included group of 10 animals of each sex
at 118 to 119 days. Control groups included 86 animals of each sex.
The highest dose level was associated with some degree of toxicity,
including a decrease in body weight gain (p <0.05 in females) and an
increase in relative kidney weight (p <0.05 in males). Increases
(p <0.05) in the volume of urine excreted and in the urinary excretion
of coproporphyrin and uroporphyrin were also observed at this dose
level. Increased (p <0.05) morphological changes were observed in the
kidney, liver and lungs of animals administered 30 mg/kg/day. The
kidney changes involved primarily the presence (p <0.05) of mineralized
deposits in the renal pelvis in females. Effects noted at the 10 mg/kg
dose level were primarily an increased (p <0.05) incidence of miner-
alized deposits in the renal pelvis in females. During the early
phase of the study there was an increase (p- <0.05) in urinary excretion
of coproporphyrin in males. At the lowest dose level (3 mg/kg),
there were no changes that were considered to be related to treatment
throughout the 2-year period. From this study in rats, a NQAEL of
3 mg/kg/day was identified.
Reproductive Effects
0 Male and female Sprague-Dawley rats (F0) were fed lab chow containing
2,4,5-T «0.03 ppb TCDD) to provide dose levels of 0, 3, 10 or 30
mg/kg/day for 90 days and then were bred (Smith et al., 1981). At
day 21 of lactation, pups were randomly selected for the following
generation (Fj) and the rest were necropsied. Subsequent ma tings were
conducted to produce F2, F3a and F3o litters, successive generations
being fed from weaning on the appropriate test or control diet.
Fertility was decreased (p <0.05) in the matings of the F3b litters in
the group given 10 mg/kg/day. Postnatal survival was significantly
(p <0.05) decreased in the F2 litters of the 10 mg/kg group and in the
F1, F2 and F3 litters of the 30 mg/kg group. A significant decrease
(p
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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Developmental Effects
0 Sparschu et al. (1971) tested 2,4,5-T (commercial grade, 0.5 ppm TCDD)
at levels of 50 or 100 mg/kg/day in pregnant rats (strain not specified)
on either days 6 to 15 (50 mg/kg) or days 6 to 10 (100 rag/kg) of
gestation. The 2,4,5-T was administered by oral intubation in a
solution of Methocel, and controls were given an appropriate volume
of Methocel. At the 50 mg/kg dose, there was a slightly higher
incidence of delayed ossification of the skull bones, but this was
not considered a teratogenic response. The 100 mg/kg dose (administered
on days 6 to 10) was toxic to the dams and caused a high incidence of
maternal deaths (only 4 of the 25 pregnant rats survived). Of these,
three had complete early resorptions, and one had a litter of 13
viable fetuses that showed toxic effects (not further described) but
no terata. From these data for maternal effects, a NOAEL of 50 mg/kg
and a LOAEL of 100 mg/kg were identified. Also identified were a
NOAEL of 100 mg/kg for teratogenicity and a LOAEL of 50 mg/kg for
fetotoxicity.
0 A sample of 2,4,5-T (technical grade) containing 0.5 ppm TCDD as well
as other phenoxy compounds was administered to CD-1 rats by oral
intubation on days 6 through 15 of gestation at dose levels of 10,
21.5, 46.4 or 80 mg/kg/day (Courtney and Moore, 1971). Examination
of offspring revealed that the sample was not teratogenic at these
dose levels. There was a significant (p <0.05) increase in fetal
mortality at the 80 mg/kg/day dose levels (the maternal LD4Q). In
two 2,4,5-T-treated fetuses, mild gastrointestinal hemorrhages were
observed as a fetotoxic effect. Kidney anomalies were also slightly
increased with the effect most pronounced at the 80 mg/kg level, but
the number of litters examined was too small to evaluate this observa-
tion. In a separate study, rats were administered 50 mg/kg/day in an
identical protocol, but in this case they were allowed to litter, and
the neonates were examined and weighed on day 1 and followed for 21
days. Postnatal growth and development were comparable to that of
the control animals. A NOAEL of 46.4 mg/kg/day for both fetotoxicity
and teratogenicity in the CD-1 rat was identified from these data.
0 Sprague-Dawley rats (50/group) and New Zealand White rabbits (20/group)
were given oral doses (gavage for rats, capsules for rabbits) of
2,4,5-T (containing 0.5 ppm TCDD) during gestation (Emerson et al.,
1971). The rats received daily doses of 1, 3, 6, 12 or 24 mg/kg on
days 6 through 15, while the raobits were administered 10, 20 or 40
mg/kg on days 6 through 18 of gestation. In both species, animals
were observed daily, weighed periodically and subjected to Cesarean
section prior to parturition. Rabbit pups were kept for observation
for 24 hours and then sacrificed. There were no observable adverse
effects in dams of either species treated with the 2,4,5-T. Litter
size, number of fetal resorptions, birth weights and sex ratios all
appeared to be unaffected in the treated groups. Detailed visceral
and skeletal examinations were performed on the control and high-dose
groups for each species, and no embryotoxic or teratogenic effects
were revealed. A NOAEL for fetotoxic and maternal effects identified
from this study was 24 mg/kg/day for the rat and 40 mg/kg/day for the
rabbit.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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0 Several different samples of 2,4,5-T (containing <0.5 ppra TCDD) were
tested in pregnant Wistar rats by daily oral administration on days
6 through 15 of gestation at dose levels between 25 and 150 mg/kg/day
(Khera and McKinley, 1972). In some cases, fetuses were removed by
Cesarean section for examination; some animals were allowed to litter,
and the offspring were observed for up to 12 weeks. At doses of
100 rag/kg, there was an increase (p <0.05) in fetal mortality and an
increase (p <0.05> in skeletal anomalies; a visceral anomaly was noted
(dilatation of the renal pelvis), which was slightly increased over
the control level, but was not statistically significant (p >0.05).
The survival of the progeny was not affected up to doses of 100 mg/kg,
and in only one trial was there a low average litter size and viability.
This effect was not duplicated in a repeat test with the same sample.
At the 25 and 50 mg/kg dose levels, significant (p <0.05) differences
from controls were not apparent. With respect to fetotoxicity, this
study identified a NOAEL of 50 mg/kg/day in the rat.
0 The teratogenic effects of 2,4,5-T were examined in golden Syrian
hamsters after oral dosing (by gavage) on days 6 through 10 of gestation
at dose levels of 20, 40, 80 or 100 mg/kg/day (Collins et al., 1971).
Four samples of 2,4,5-T with dioxin levels of 45, 2.9, 0.5 or 0.1 ppm
were administered. Three samples, which had no detectable dioxin
(based on TCDD), were also tested. The 2,4,5-T samples induced fetal
death and terata. The incidence of effects increased with increasing
content of the TCDD impurity. 2,4,5-T with no detectable dioxin
produced no malformations below the 100 mg/kg dose level. Using the
data from the 2,4,5-T samples with no detectable dioxin, a NOAEL of
80 mg/kg/day for the hamster was identified.
• Behavioral effects resulting from in utero exposure to 2,4,5-T were
examined in Long-Evans rats after single oral doses were administered
during gestation (Crampton and Rogers, 1983). The sample of 2,4,5-T
contained <0.03 ppm TCDD. Novelty response abnormalities were
detected after single doses as low as 6 mg/kg were administered on
day 8 of gestation. Examination of the brain in the affected offspring
failed to reveal any changes of a qualitative or quantitative structural
nature in various areas of the brain. With respect to behavioral
effects, the LOAEL for this study is 6 mg/kg.
0 The teratogenic effects of technical 2,4,5-T (TCDD content 0.1 ppm)
were studied using large numbers of pregnant mice of C57BL/6, C3H/He,
BALB/c and A/uAX inbred strains and CD-1 stock (Gaines et al., 1975).
Dose-response curves were determined for the incidence of cleft
palate, embryo lethality and fetal growth retardation. These deter-
minations were replicated 6 to 10 times for each inbred strain and
35 times for the CD-1. The number of litters studied ranged from 236
for BALB/c mice to 1,485 for CD-1 mice. Treatment was by gavage on
days 6 to 14 of pregnancy, and dose levels of 2,4,5-T ranged from 15
to 120 mg/kg/day. The lowest dose tested in the A/JAX was 15 mg/kg,
and this dose was teratogenic. The other strains and CD-1 demonstrated
teratogenicity at 30 mg/kg, the lowest dose tested. There were
significant (p <0.05) differences in sensitivities among the strains
for the parameters measured. Based on this study in the mouse, the
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
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LOAEL for teratogenic effects is 15 mg/kg/day for the A/JAX strain
and 30 mg/kg/day for the other strains.
0 Neubert and Dillmann (1972) studied the effects of 2,4,5-T in pregnant
NMRI mice. Three samples of 2,4,5-T were utilized: one had <0.02 ppm
dioxin, and was considered "dioxin-free"; a second sample had a dioxin
content of 0.05 ± 0.02 ppm; and the third sample had an undetermined
dioxin content. The 2,4,5-T was administered by gavage on days 6
through 15 of gestation at dose levels from 8 to 120 mg/kg/day.
Fetuses were removed on day 18 and examined. Cleft palate frequency
exceeding (p <0.05) that of the controls was observed with doses
higher than 30 mg/kg with all samples. Reductions (p <0.05) in fetal
weight were observed with all samples tested at doses as low as 10 to
15 mg/kg. There was no clear increase in embryo lethality over that
of controls at these lower doses. With the purest sample of 2,4,5-T,
single oral doses of 150 to 300 mg/kg were capable of producing
significant (p <0.05) incidences of cleft palate. The maximal terato-
genic effect was seen when the 2,4,5-T was administered on days 12 to
13 of gestation. Based on the data obtained with the purest sample
of 2,4,5-T, the teratogenic NOAEL is 15 mg/kg/day and the fetotoxic
NOAEL is 8 mg/kg/day.
0 Roll (1971) examined the teratogenic effects of 2,4,5-T in NMRI-Han
mice after oral administration on days 6 to 15 of gestation at dose
levels of 0, 20, 35, 60, 90 or 130 mg/kg/day. The 2,4,5-T sample had
a purity of 99.6%, with a dioxin content of <0.01 ppm (measured by
the DOW method), or 0.05 ± 0.02 ppm (measured by the U.S. Food and
Drug Administration (FDA) method). Peanut oil was used as the vehicle.
Animals were sacrificed on day 18 and examined for defects. Fetal
weight was significantly (p <0.05) lower than control at all doses.
Resorptions were significantly (p <0.05) increased at 60 mg/kg and
above. The incidence of cleft palates was significantly (p <0.05)
higher at 35 mg/kg and higher, but there was no effect at 20 mg/kg.
There were also dose-dependent increases in ossification defects of
sternum and various other bones. The authors concluded that 2,4,5-T
alone (independent of TCDD contamination) was teratogenic in mice,
and that the teratogenic NOAEL in this strain was 20 mg/kg/day. In
view of the significantly (p <0.05) lower fetal weight at 20 mg/kg/day,
this level may also be considered the LOAEL for fetotoxicity.
0 No teratogenic effects were observed in the offspring of female
rhesus monkeys that were given oral doses of 0.05, 1.0 or 10.0 mg
2,4,5-T (containing 0.05 ppm TCDD)/kg/day in capsules during gestation
days 22 through 38. Neither was toxicity evident in the mothers
(Dougherty et al., 1976).
Mutagenicity
0 At 250 and 1,000 ppm 2,4,5-T (with no detectable TCDD), mutation
rate was significantly (p <0.05) increased at the higher dose in the
sex-linked recessive lethal test in Drosophila as carried out by
Majumdar and Golia (1974). The sex-linked test was not affected by
920 or 1,804 ppm of the sodium salt of 2,4,5-T at pH 6.8 in a study
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-13-
carried out by Vogel and Chandler (1974). Although they found no
cytogenetic effects in Drosophila, Magnusson et al. (1977) concluded
that 1,000 ppm 2,4,5-T «0.1 ppm TCDD) did cause an increase (p <0.05)
in the number of recessive lethals compared to the controls. Rasmusson
and Svahlin (1978) treated Drosophila larvae to food containing 100
and 200 ppm 2,4,5-T; survival was low at 200 ppm, but 2,4,5-T had
no observable effect on somatic mutational activity.
• Anderson et al. (1972) found that neither 2,4,5-T nor its butyric
acid form showed any mutagenic action when tested on histidine-
requiring mutants of Salmonella typhimurium.
0 Buselmaier et al. (1972) found that intraperitoneal injection of
2,4,5-T (dioxin levels not given) had no effect in the host-mediated
assay (500 mg/kg) or in the dominant lethal test (100 mg/kg) with
NMRI mice. Styles (1973), likewise, found no increase in back mutation
rates with the serum of rats treated orally with 2,4,5-T in the
host-mediated assay with Salmonella typhimurium (dosages and purity
of the samples not given).
0 Shirasu et al. (1976) found that 2,4,5-T did not induce mitotic gene
conversion in a diploid strain of Saccharomyces cerevisiae. When the
pH of the treatment solution was less than 4.5, Zetterberg (1978)
found that 2,4,5-T was mutagenic in haploid, DNA-repair-defective
J3. cerevisiae.
0 Jenssen and Renberg (1976) investigated the cytogenetic effects of
2,4,5-T in mice by examining the ability of the herbicide to induce
micronuclei formation in the erythrocytes of mouse bone marrow. CBA
mice were treated at 8 to 1 0 weeks of age (20 to 30 g) with a single
intraperitoneal injection of 100 mg/kg of 2,4,5-T (<1 ppm TCDD) dis-
solved in Tween 80 and physiological saline. Cytogenetic examination
at 24 hours and 7 days after treatment showed no detectable increase
in micronuclei in the erythrocytes compared to controls. A weak
toxic effect on the mitotic activity was indicated, as judged by a
decrease in the percentage of polychromatic erythrocytes.
Carcinogenici ty
0 Znnes et al. (1969) investigated the potential carcinogenic effects
of 2,4,5-T in two hybrid strains of mice derived by breeding SPF
C57BL/6 female mice to either C3H/Anf or AKR males. Beginning at
6 days of age, 2,4,5-T was administered by gavage in 0.5% gelatin to
a group of 72 mice at a dose level of 21.5 mg/kg/day. This was
reported to be the maximum tolerated dose. At 28 days of age, the
2,4,5-T was added to the diet at a level of 60 ppm, corresponding to
a dose of about 9 mg/kg/day (assuming that 1 ppm equals 0.15 mg/kg/day
in the diet from Lehman, 1959). This dose was fed for 18 months, at
which time the study was terminated. All animals were necropsied and
the tissues were examined both grossly and microscopically. There
were no significant (p >0.05) increases in tumors in either strain of
treated mice.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-14-
0 A lifetime study using oral administration of 2,4,5-T in both sexes
of two strains of mice, C3Hf and XVII/G, was performed by Muranyi-
Kovacs et al. (1976). The 2,4,5-T, which contained less than 0.05
ppm of dioxins, was administered in the water (1,000 mg/L) for 2
months beginning at 6 weeks of age, and thereafter in the diet at
80 ppm (12 mg/kg/day) until death or when the mice were sacrificed _iri
extremis. In the treated C3Hf mice there was a significant (p <0.03)
increase in the incidence of total tumors found in female mice and a
significant (p <0.001) increase in total nonincidental tumors in each
sex, which the authors interpreted as life-threatening. No signifi-
cant (p >0.05) difference was found in the XVIZ/G strain between the
treated and control mice. The authors felt that 2,4,5-T demonstrated
carcinogenic potential in the C3Hf strain, but that additional studies
in other strains and in other species of animals needed to be performed
before a reliable conclusion with respect to carcinogenicity could be
made.
0 Groups of Sprague-Dawley rats (50 each of males and females) were
maintained on diets supplying 3, 10 or 30 mg/kg/day of 2,4,5-T for 2
years (Kociba et al., 1979). The 2,4,5-T was approximately 99% pure,
containing 1.3% (w/w) other phenoxy acid impurities. Dioxins were
not detected, the limit of detection for TCDD being 0.33 ppb. An
interim sacrifice was performed on an additionally included group of
10 animals of each sex at 118 to 119 days. Control groups included
86 animals of each sex. At the end of the 2-year period, there was
no significant (p >0.05) increase in tumor incidence in any treated
group compared to the control for either male or female animals.
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, 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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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-15-
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-T. The study in humans
by Gehring et al. (1973) was not selected because observations of the subjects
were reported simply as clinical effects without further details. The
behavioral study in rats by Crampton and Rogers (1983) was not selected
because the interpretation of altered novelty response behavior in the absence
of other toxic signs needs further investigation before definitive conclusions
can be made. It is therefore recommended that the Ten-day HA value for a
10-kg child (0.8 mg/L, calculated below) be used at this time as a conservative
estimate of the One-day HA value.
Ten-day Health Advisory
The study by Neubert and Dillman (1972) has been selected to serve as
the basis for determination of the Ten-day HA value for 2,4,5-T. This
developmental study in rats identified a NOAEL of 8 mg/kg/day and a LOAEL
of 15 mg/kg/day, based on reduced body weights in pups from dams exposed on
days 6 to 15 of gestation. This LOAEL is supported by a number of other
developmental studies in rodents that identified LOAELs ranging from 15 to
100.mg/kg/day (Roll, 1971; Sparschu et al., 1971; Xhera and McKinley, 1972;
Gaines et al., 1975). In the 21-day feeding study in rats by Vos et al.
(1983), a LOAEL of 20 mg/kg/day was identified based on effects on kidney
weight and the immune system. The 8 mg/kg/day NOAEL for fetal effects selected
from the Neubert and Dillman (1972) study may not be applicable to a 10-kg
child; however, the assumptions for a 10-kg child are used with this NOAEL
in this case since, although a NOAEL was not found in the 21-day study by
Vos et al. (1983) where the observed effects are applicable to a 10-kg child,
the LOAEL of 20 mg/kg/day is 2.5 times higher than the NOAEL used for the
Ten-day HA.
Using a NOAEL of 8 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (a "g/*g/day) (10 kg) = 0.8 mg/L (800 ug/L)
(100) (1 L/day) y *
where:
8 mg/kg/day = NOAEL, based on absence of maternal or fetal effects in
rats exposed by gavage on days 6 to 15 of gestation.
10 kg a assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOA2L from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The reproduction study by Smith et al. (1981, 1978) has been selected
to serve as the basis for the Longer-term HA value for 2,4,5-T because the
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-16-
reduction in neonatal survival over multiple generations is concluded to be
relevant to the Longer-term HA for a 10-kg child. The MOAEL identified was
3 mg/kg/day, and the LOAEL was 10 mg/kg/day. Other possible selections have
a higher NOAEL [10 mg/kg/day in the 90-day feeding study in rats by McCollister
and Kociba (1970) and the 90-day oral treatment study in dogs by Drill and
Hiratzka (1953)].
Using a NOAEL of 3 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (3 mg/kg/day) (10 kg) = 0.3 mg/L (300 ug/L)
(100) (1 L/day)
where:
3 mg/kg/day = NOAEL, based on absence of adverse effects in neonatal rats
in the three-generation reproduction study in rats given
2,4,5-T in the diet.
10 kg a assumed body weight of a child.
100 o 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 a 70-kg adult is calculated as follows:
Longer-term HA = (3 mg/kg/day) (70 kg) = 1.05 mg/L (1,050 ug/L)
(100) (2 L/day)
where:
3 mg/kg/day = NOAEL, based on absence of adverse effects in neonatal rats
in a three-generation reproduction study in rats given
2,4,5-T in the diet.
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 st:dy.
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
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-17-
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 Kbciba et al. (1979) has been selected to serve as the
basis for the Lifetime HA value for 2,4,5-T. In this study, rats were fed
2,4,5-T in the diet for 2 years. Based on observations of effects of 2,4,5-T
on various biochemical parameters in addition to gross and microscopic obser-
vations related to general toxicity in the rats, this study identified a
NOAEL of 3 mg/kg/day and a LOAEL of 10 mg/kg/day. This study is supported by
the three-generation rat study (Smith et al., 1981, 1978) that identified a
NOAEL of 3 mg/k9/day.
Using this study, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (3.0 mg/kg/day) . 0<003 mg/kg/day
(100) (10)
where:
3.0 mg/kg/day = NOAEL, based on absence of adverse effects on the
kidneys, liver and lungs of rats exposed to 2,4,5-T
in the diet for 2 years.
100 a uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
10 = modifying factor used by U.S. EPA Office of Pesticide
Programs to account for data gaps (chronic feeding
study in dogs) which does not make it possible to
establish the most sensitive end point for 2,4,5-T.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.003 mg/kg/day) (70 kg) = 0.,05 ng/L (105 ug/L)
(2 L/day)
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-18-
vhere:
0.003 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 - (0.105 mg/L) (20%) - 0.021 mg/L (21 ug/L)
where:
0.105 mg/L - DWEL.
20% o assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Chronic feeding studies with 2,4,5-T in Sprague-Dawley rats (Kociba
et al., 1979) and C57BL/6 x C3H/Anf, C57BL/6 x AKR and XVII/G strains
of mice (Innes et al., 1969; Muranyi-Kovacs, et al; 1976) were
negative for carcinogenic effects. A chronic feeding study with
2,4,5-T in C3Hf mice was inconclusive (Muranyi-Kovacs et al., 1976).
0 IARC (1982) concluded that the carcinogenicity of 2,4,5-T is indeter-
minant (Group 3, inadequate evidence in animals and humans).
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), 2,4,5-T may be classified in
Group D: not classified. This category is for agents with inadequate
animal evidence of carcinogenicity.
0 The Carcinogen Assessment Group (CAG) of the U.S. EPA classified
chlorophenoxyacetic acids and/or chlorophenols containing 2,3,7,8-TCDD
in IARC category 2A (probably carcinogenic in humans on the basis
of limited evidence in humans), but a quantitative cancer risk estimate
only for 2,3,7,8-TCDD itself was made. The CAG considered the human
evidence for the carcinogenicity of 2,3,7,8-TCDD alone to be "inadequate"
because of the difficulty in attributing observed effects solely to
the presence of 2,3,7,8-TCDD, which occurs as an impurity in the
phenoxyacetic acids and chlorophenols (U.S. EPA, 1985).
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The U.S. EPA/Office of Pesticide Programs has calculated a Provisional
Acceptable Daily Intake (PADI) value of 0.003 mg/kg/day, based on the
results of a rat chronic oral NOAEL of 3 mg/kg/day with an uncertainty
factor of 1,000 (used because of data gaps).
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2,4, 5-Trichlorophenoxyacetic Acid August, 1987
*
-19-
0 The National Academy of Sciences (HAS, 1977) has calculated an ADI
of 0.1 mg/kg/day, using a NOAEL of 1 0 mg/kg/day (identified in a
90-day feeding study in dogs) and an uncertainty factor of 100. A
chronic Suggested-No-Ad verse-Effect-Level (SNARL) of 0.7 mg/L was
calculated based on the ADZ of 0.1 mg/kg/day.
0 The American Conference of Governmental Industrial Hygienists (ACGIH,
1981) has recommended a Threshold Limit Value-Time-Weighted Average
(TLV-THA) of 10 mg/m3 and a Threshold Limit Value-Short-Term Exposure
Limit (TLV-STEL) of 20 mg/m3.
0 The ADI recommended by the World Health Organization is 0 to
0.03 mg/kg (Vettorazzi and van den Hurk, 1983).
VTI. ANALYTICAL METHODS
0 Determination of 2,4,5-T is by a liquid-liquid extraction gas
chroma tographic 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 that are extracted from the aqueous phase. Separation
and identification of the esters is made by gas chroma tography.
Detection and measurement are 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 instrumentation
used. Typically, using a 1-L sample and a gas chroma tograph with
an electron-capture detector results in an approximate detection
limit of 10 ng/L for 2,4,5-T.
TREATMENT TECHNOLOGIES _
0 Available data indicate that granular-activated carbon (GAG) and
powdered-activated carbon (PAC) adsorption will effectively remove
2,4,5-T from water.
0 Robeck et al. (1965) experimentally determined adsorption isotherms
for the butoxy ethanol ester of 2,4,5-T on PAC. Based on these
results, it was calculated that 14 mg/L PAC would be required to
remove 90% of 2,4,5-T, while 44 mg/L PAC would be required to remove
99% of 2,4,5-T (Pershe and Goss, 1979; Robeck et al. , 1965).
0 Robeck et al. (1965) reported the results of a GAC column operating
under pilot plant conditions. At a flow rate of 0.5 gpm/ft3, 99+%
of 2,4,5-T was removed. By comparison, treatment with 5 to 20 mg/L
PAC removed 80 to 95% of the same concentration of 2,4,5-T.
0 In a laboratory study conducted with an exchange resin, Rees and Au
(1979) reported 89±2% removal efficiency of 2,4,5-T from contaminated
water by adsorption onto synthetic resins.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-20-
Conventional water treatment technique of coagulation with alum,
sedimentation and sand filtration removed 63% of the 2,4,5-T ester
present in spiked river water (Robeck et al., 1965).
Treatment technologies for the removal of 2,4,5-T from water are
available and have been reported to be effective. However, selection
of individual or combinations of technologies to attempt 2,4,5-T
removal from water must be based on a case-by-case technical evaluation,
and an assessment of the economics involved.
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2,4,5-Trichlorophenoxyacetic Acid August, 1987
-21-
IX. REFERENCES
ACGIH. 1981. American Conference of Governmental Industrial Hygenists.
Threshold limit values for chemical substances and physical agents in
the workroom environment. Cincinnati, OH: ACGIH, p. 27.
Anderson, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of
herbicides for possible mutagenic properties. J. Agric. Food Chem.
20:649-656.
BCPC. 1983. British Crop Protection Council, The pesticide manual. A
world compendium, 7th ed. (C. R. Worthing, ed.). 2,4,5-T, p. 11120.
Buselmaier, W., G. Roehrborn and P. Propping. 1972. Mutagenicity investi-
gations with pesticides in the host-mediated assay and the dominant
lethal test in mice. Biol. Zentralbl. 91:311-325.
CHEMLAB. 1985. The chemical information system,, CIS, Inc., Bethesda, MD.
«
Collins, T.F.X., G.H. Williams and G.C. Gray. 1971. Teratogenic studies
with 2,4,5-T and 2,4-D in the hamster. Bull. Environ. Contain,, Toxicol.
6(6):559-67.
Courtney, K.D. and J.A. Moore. 1971. Teratology studies with 2,4,5-tri-
chlorophenoxyacetic acid and 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Toxicol. Appl. Pharmacol. 20:396-403.
Crampton, M.A. and L.J. Rogers. 1983. Low doses of 2,4,5-trichlorophenoxy-
acetic acid are behaviorally teratogenic in rats. Experientia. 39:891-2.
Dougherty, W.J., F. Coulston and L. Golberg. 1976. The evaluation of the
teratogenic effects of 2,4,5-trichlorophenoxyacetic acid in the Rhesus
monkey. Environ. Qual. Saf. 5:89-96.
Drill, V.A. and T. Hiratzka. 1953. Toxicity of 2,4-dichlorophenoxyacetic
acid and 2,4,5-trichlorophenoxyacetic acid. A report on their acute and
chronic toxicity in dogs. Arch. Ind. Hyg. Occup. Med. 7:61-67.
Emerson, J.L., D.J. Thompson, R.J. Strebing, C.G. Gerbig and V.B. Robinson.
1971. Teratogenic studies on 2,4,5-trichlorophenoxyacetic acid in the
rat and rabbit. Food Cosmet. Toxicol. 9:395-404
Fang, S.C., E. Fallin, M.L. Montgomery and V.H. Freed. 1973. Metabolism and
distribution of 2,4,5-trichlorophenoxyacet.4 acid in female rats. Toxicol.
Appl. Pharmacol. 24(4):555-563.
Gaines, T.B., J.F. Holson, C.J. Nelson and H.J. Schumacher. 1975. Analysis
of strain differences in sensitivity and reproducibility of results in
assessing 2,4,5-T teratogenicity in mice. Toxicol. Appl. Pharmacol.
33:174-175. Abstract No. 30.
Gehring, P.j. and J.E. Betso. 1978. Phenoxy acids: Effects and fate in
mammals. Ecol. Bull. 27:122-133.
-------
2,4,5-Trichlorophenoxyacetic Acid August, 1987
-22-
Gehring, P.J., C.G. Krammer, B.A. Schwetz, J.Q. Rose, V.K. Rowe and J.S. Zimmer.
1973. The fate of 2, 4,5-trichlorophenoxyacetic acid (2,4,5-T) following
oral administration to man. Toxicol. Appl. Pharmacol. 25(3):441.
Grunow, W., C. Bohme and B. Budczies. 1971. Renal excretion of 2,4,5-T by
rats. Food Cosmet. Toxicol. 9:667-670.
Hanify, J.A., P. Metcalf, C.L. Nobbs and K.J. Worsley. 1981. Aerial spraying
of 2,4,5-T and human birth malformation: An epidemiological investigation.
Science. 21 2(4492) :349-351 .
IARC. 1982. International Agency for Research on Cancer. IARC monographs
on the evaluation of carcinogenic risk of chemicals to man. Lyon, France:
IARC, Suppl. 4.
Innes, J.R.M., B.M. Ulland, M.G. Valeric, L. Petrucelli, L. Fishbein, E.R. Hart,
A.J. Pallotta, R.R. Bates, H.L. Fa Ik, J.J. Gart, H. Klein, I. Mitchell
and J. Peters. 1969. Bioassay of pesticides and industrial chemicals
for tumor igenicity in mice: A preliminary note. J. Natl. Cancer Inst.
Jenssen, D. and L. Renberg. 1976. Distribution and cytogenetic test of
2,4, -D and 2,4,5-T phenoxyacetic acids in mouse blood tissues. Chem.
Biol. -Interact. 14(3-4): 291 -299.
Khan, M.A.Q. 1985. Personal communication to Environmental Criteria and
Assessment Office, U.S. Environmental Protection Agency, Cincinnati, OH.
January.
Khera, K.S. and W.P. McKinley. 1972. Pre- and postnatal studies on 2, 4,5-
trichlorophenoxyacetic acid, 2, 4,-dichlorophenoxyacetic acid and their
derivatives in rats. Toxicol. Appl. Pharmacol. 22:14-28.
Kociba, R.J. , D.J. Keyes, R.W. Li so we, R.P. Kalnins, D.D. Dittenber, C.E. Wade,
S.J. Gorzinskl, N.H. Mahle and B.A. Schwetz. 1979. Results of a two-year
chronic toxicity and oncogenic study of rats ingesting diets containing
2, 4,5-trichlorophenoxyacetic acid (2,4,5-T). Food Cosmet. Toxicol.
17:205-221.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Association of Food end Drug Officials of the united States.
Magnus son, J. , C. Ramel and A. Eriksson. 1977. Mutagenic effects of chlori-
nated phenoxyacetic acids in Drosophila melanogaster. Hereditas.
87(1):121-123.
Majumdar, S.K. and J.K. Golia. 1974. Mutation test of 2, 4,5-trichlorophen-
oxyacetic acid on Drosophila melanogaster. Can. J. Genet. Cytol.
16(2): 465-466.
McCollister, S.B. and R.J. Kociba. 1970. Results of 90-day dietary feeding
study on 2, 4,5-trichlorophenoxyacetic acid. Unpublished study by Dow
Chemical. MRID 00092151.
-------
2,4,5-Trichlorophenoxyacetic Acid August, 1987
-23-
Meister, R., ed. 1983. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Muranyi-Kovacs, I., G. Rudali and J. Zmbert. 1976. Bioassay of 2,4,5-tri-
chlorophenoxyacetic acid for carcinogenicity in mice. Br. J. Cancer.
33:626-633.
HAS. 1977. National Academy of Sciences. Drinking water and health, Vol. 1.
Washington, DC: National Academy Press.
Nelson, C.J., J.F. Holson, H.G. Green and D.W. Gaylor. 1979. Retrospective
study of the relationship between agricultural use of 2,4,5-T and cleft
palate occurrence in Arkansas. Teratology. 19:(3)377-384.
Neubert, D. and I. Dillmann. 1972. Embryotoxic effects in mice treated with
2,4,5-trichlorophenoxyacetic acid and 2,3,7,8-tetrachlorodibenzo-p-dioxin.
Naunyn-Schmiedeberg's Arch. Pharmacol. 272:243-264.
Ott, M.G., B.B. Holder and R.D. Olson. 1980. A mortality analysis of
employees engaged in the manufacture of 2,4,5-trichlorophenoxyacetic
acid. J. Occup. Med. 22(1):47-SO.
Pershe, E.R. and J. Goss. 1979. Uses of powdered and granular activated
carbon in water treatment. J. New Eng. Water Works Assoc. (9):254-286.
Piper, W.N., J.Q. Rose, M.L. Leng and P.J. Gehring. 1973. The fate of
2,4,5-trichlorophenoxyacetic acid (2,4,5-T) following oral administra-
tion to rats and dogs. Toxicol. Appl. Pharmacol. 26:339-351.
Rasmusson, B. and H. Svahlin. 1978. Mutagenicity tests of 2,4-dichloro-
phenoxyacetic acid and 2,4,5-trichlorophenoxyacetic acid in genetically
stable and unstable strains of Drosophila melanogaster. Ecol. Bull.
27:190-192.
Rees, G.A.V. and L. Au. 1979. Use of XAD-2 macroreticular resin for the
recovery of ambient trace levels of pesticides and industrial organic
pollutants from water. Bull. Environ. Contain. Toxicol. 22(4/5): 561-566.
Riihimaki, V., S. Asp and S. Hernberg. 1982. Mortality of 2,4-dichloro-
phenoxyacetic acid and 2,4,5-trichlorophenoxyacetic acid herbicide
applicators in Finland: first report of an ongoing prospective cohort
study. Scand. u. Work Environ. Health. 8(1):37-42.
Robeck, G.G., K.A. Costal, J.M. Cohen and J.F. Kreissl. 1965. Effectiveness
of water treatment processes in pesticide removal. .J. Am. Water Works
Assoc. (2):181-199.
Roll, R. 1971. Studies of the teratogenic effect of 2,4,5-T in mice. Food
Cosmet. Toxicol. 9(5):671-676.
Rowe, V.K. and T.A. Hymas. 1954. Summary of toxicological information on
2,4-D and 2,4,5-T type herbicides and an evaluation of the hazards to
livestock associated with their use. Am. J. Vet. Res. 15:622-629.
-------
2,4,5-Trichlorophenoxyacetic Acid August, 1987
-24-
Shirasu, Y., M. Noriya, K. Kato, A. Furuhashi and T. Kada. 1976. Mutagenicity
screening of pesticides in the microbial system. Mutat. Res. 40:19-30.
Smith, F.A., B.A. Schwetz, F.J. Murray, A.A. Crawford, J.A. John, R.J. Kociba
and C.J. Humiston. 1978. Three-generation study of rats ingesting
2,4,5-trichlorophenoxyacetic acid in the diet. Toxicol. Appl. Pharmacol.
45:293 (Abst.).
Smith, F.A., F.J. Murray, J.A. John, K.D. Nitschke, R.J. Kociba and B.A.
Schwetz. 1981. Three-generation reproduction study of rats ingesting
2,4,5-trichlorophenoxyacetic acid in the diet. Food Cosmet. Toxicol.
19:41-45.
Sparschu, G.L., F.L. Dunn, R.W. Lisowe and V.K. Rowe. 1971. Study of the
effects of high levels of 2,4,5-trichlorophenoxyacetic acid on fetal
development in the rat. Food Cosmet. Toxicol. 9:527-530.
Standard Methods. 1985. Method 509B, chlorinated phenoxy acid herbicides.
Standard Methods for the Examination of Water and Wastewater, 16th ed.
APHA, AWWA, WPCF.
STORET. 1987.
Styles, J.A. 1973. Cytotoxic effects of various pesticides in vivo and J.n
vitro. Mutat. Res. 21(1):50-51.
U.S. EPA. 1978. U.S. Environmental Protection Agency. Method for chloro-
phenoxy acid herbicides in drinking water. Methods for organochlorine
pesticides and chlorophenoxy acid herbicides in drinking water and raw
source water. Office of Drinking Water. Washington, DC. Interim
draft, July.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Health assessment
document for polychlorinated dibenzo-p-dioxins. Office of Health and
Environmental Assessment. Cincinnati, OH. EPA/600/8-84/014F.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for car-
cinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
Vettorazzi, G. and G.W. van den Hurd, eds. 1985. Pesticides reference index.
J.M.P.R. p. 41.
Vogel, E. and J.L.R. Chandler. 1974. Mutagenicity testing of cyclamate and
some pesticides in Drosophila melanogaster. Experientia. 30(6):621-623.
Vos, J.G., E.I. Krajnc, P.K. Beekhof and M.J. van Logten. 1983. Methods for
testing immune effects of toxic chemicals: Evaluation of the immunotoxicity
of various pesticides in the rat. Pestic. Chem.: Hum. Welfare Environ.,
Proc. 5th Int. Congr. Pestic. Chem. 3:497-504.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck index—an encyclopedia of chemicals and drugs, loth ed. Rahway, NJ:
Merck and Company, Inc.
-------
2,4,5-Trichlorophenoxyacetic Acid August, 1987
-25-
Zack, J.A. and R.R. Suskind. 1980. The mortality experience of workers
exposed to tetrachlorodibenzodioxin in a trichlorophenol process accident.
J. Occup. Med. 22(1 ):11-14.
Zetterberg, G. 1978. Genetic effects of phenoxy acids on microorganismso
Ecol. Bull. 27:193-204.
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