or*
. \ Health
Advisories
for
50 Pesticides
CD
CO
401 MS; SW. WC'»
Wnsl-.-ngJorx D C 2
(202) 250-OS44
EJED
EPA
570/
1988.1
-------
Federal Register / Vol. 54. No. 34 / Wednesday. February 22. 1989 / Notices
7599
information to be confidential business
information. Notice of receipt was
published la the Federal Register of
September 14.1988 (53 FR 35S3S). The
submitter voluntarily suspended the
review period to develop data to
address EPA's concerns for ecotoxicity.
The 80-day review period is scheduled
to expire on May 13.1989.
Based on its analysis. EPA finds that
there is a possibility that the substances
submitted for review in this PMN may
be regulated under TSCA. The Agency
requires an extension of the review
period, as authorized by section 5(c) of
TSCA. to Investigate further potential
risk, to examine its regulatory options.
and to prepare the necessary
documents, should regulatory action be
required. Therefore. EPA has
determined that good cause exists to
extend (he review period for an
additional 90 days, to May 13.1989.
FMNs are available for public
inspection in Rm. NE-G004. at the EPA
headquarters, address given above, from
8 a.m. to 4 p.m.. Monday through Fnday.
except legal holidays.
Dated February 8.1989.
fohtt W. Melon.
Director. Chemical Control Division. Office of
Toxic Substances.
[FR Doc. 09-3993 Filed 2-21-89:8.45 am|
•9LLMQ COOC iMP go |j
IWH-FRL-352S-4)
Drinking Water Hearth Advisories
AGENCY: Environmental Protection
Agency (EPA)
ACTION: Notice of availability of
Drinking Water Health Advisories for
pesticides
SUMMARY: This notice announces the
availability of EPA Drinking Water
Health Advisories (HAs) for 50
pesticides. Health Advisories are
available for the following
contaminants:
•\afiuorfen
Airetryn
Ammonium Siilfb
Atrazine
Biivgon
Bromactl
But} late
Carbaryl
CUtrboxin
O.lorarabeo
Ci.iorothalonil
Cv anazire
Dalapon
DiMZinon
I ..VDich loropiopene
Dimethnn
Dmoseb
Diphenamid
Disulfoton
Oiuran
EndothaU
Ethylene thiourea
Fenairuphoi
Fluometuran
Fonofos
Clyphosate
Hexaunom;
MCPA
Maleic hydrazide
Melhomyl
Methyl parathinn
Metolachlor
Metnbuzin
Paraquat
Picloram
Prometon
Pronamide
Propachlor
Propazinc
Propham
Simazine
Tebuthiuron
Terbacil
Terbufoa
2.4.5-Tnchloroptieno \yacutic add
Tnfluralin
These HAs were developed in
conjunction with the National Pesticide
survey sponsored by the EPA Office of
Drinking Water and Office of Pesticide
Programs. The HAs provide information
on the health effects, analytical
methodology, and treatment
for specific contaminants that Would be
useful in dealing with emergency spills
or contamination situations. The HAs
describe nonregulaloiy concentrations
of drinking water contaminants that are
considered protective of adverse health
effects over specific duration* of
exposure. A margin of safety is
incorporated to protect sensitive
members of the population. Health
Advisories are updated «s new
information becomes available.
SUPPLEMENTARY INFORMATION: The
Office of Drinking Water issued draft
HAs for these contaminants on (anuary
8.I9WJ A total of 21 comments were
received. The comments received wcie
reviewed and incorporated where
-
4700. Please refer to accession numbkr
PB88-245931/AS. For copies of the
individual HAs. rather than the en lira
set. contact the EPA Safe Drinking
Water Hotline (800) 428-4791. local (ZOi
382-5533.
For further information contact.
Jennifer Onne. Health Advisory Program
Coordinator. Office of Drinking Water
(WH-450D], U.S. Environmental
Protection Agency. 401M Street. SW..
Washington. DC 20460, or call (202) 382-
7571.
Rebecca W. Hanrner.
Acting Assistant Administrator far Water
IFS Doc. 89-3992 Filed 2-21-afc K4S am]
BUMO cooc ueo-ce-n
FEDERAL DEPOSIT INSURANCE
CORPORATION
Privacy Act of 1974; Amendment to
Existing System of Records
AGENCY: Federal Deposit Insurance
Corporation ("FDIC").
ACTION: Notice of proposed changes to a
system of records: "Attorney—Legal
Intern Applicant System,"
SUMMARY: The proposed amendments
will update the content, use,
retrievability. and retention categories
of Ibis-system of records, add will
clarify existing notification procedures.
This action is being undertaken as part
of a periodic review to revise, as
necessary, the FDIC's systems of
records under the Privacy Act of 1974.
DATE* Comments must be submitted by
Marc* 24.1989. The amendments will
become effective May 8.198R unless a
superseding notice to the contrary is
published before that date.
ADDRESS: Comments should be
addressed to Hoyle L. Robinson.
ExecOive Secretary. Federal Deposit
Insurance Corporation. 55017th Street
N'W,«YVashingUm. DC 20429. or hand-
deluared to Room 6099 at the same
adtJras. Monday through Friday.
between the hours of 9 a.m. and 5pm
FOR frvRTHER INFORMATION CONTACT:
Patli C Fox. Senior Program Attorney
Federal Deposit Insurance Corporation
550 17th Street NW.. Washington. DC
20431 telephone (202) 898-3714
SUPPLEMENTARY INFORMATION: Thu
FDIC's system of records entitled
"Attorney—Legal Intern Applicant
Svslem" is being revised to reflect
diiditions to the system as paftof the
periodic rc\ lew of each FDIC system
records under the Privacy Acfcif 1974, 5
U S.C 5o2a. The categories OBBCords
hdvebcfen changed to inciudanidng
-------
Augustr 1988
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 effectsr 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.
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.
-------
Acifluorfen
August/ 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 5094-66-6 (acid)
62476-59-9 (sodium salt)
Structural Formula
Sodium 5-( 2-chloro-4-(trifluoromethyl)-phenoxy)-2-nltrobenzoate
Synonyms
• Blazer*; Carbofluorfen; RH-6201; Tackle*; Sodium acifluorfen (Meister,
1983).
gaes
• Acifluorfen is used as a selective pre- and post-emergence herbicide to
control weeds and grasses in large-seeded legumes including soybeans
and peanuts (Meister, 1983).
Properties (Windholz et al., 1983; Meister, 1983; CUEMLAB, 198S; WSSA, 1983)
Chemical Formula
Molecular Height
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
(acid)
C14H6ClF3NNa05 (sodium salt)
361.66 (acid)
383.65 (sodium salt)
Off-white solid (acid), brown crystalline
powder/white powder (sodium salt)
124-125«C (sodium salt)
151.5-157'C (acid)
24 mm Hg (45% H20 solution)
>25% (sodium salt) (dimensions not
specified)
-4.85 (acid) (calculated)
Occurrence
8 No information was found in the available literature on the occurrence
of acifluorfen.
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Acifluorfen August, 1988
-3-
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) (Registrtion
Standard Science Chapter for Acifluorfen).
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 (Registration Standard Science Chapter for Acifluorfen).
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 (Registration Standard Science Chapter
for Acifluorfen).
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 (Registration Standard Science
Chapter for Acifluorfen).
0 Greenhouse studies have demonstrated that the uptake of acifluorfen
by rotational crops decreases with aging of residues in soil (Registra-
tion Standard Science Chapter for Acifluorfen).
III. PHARMACOKINETICS
Absorption
0 No information was found in the available literature on the absorption
of acifluorfen.
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Acifluorfen August, 1988
-4-
Distribution
0 No information was found in the available literature on the distribution
of acifluorfen.
Metabolism
0 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.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the 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 LDso 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
mgAg/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
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
-------
Acifluorfen August, 1988
-5-
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.
0 Piccirillo and Bobbins (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 mgAg/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 mgAg/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. Signs 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. (197%) 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 mL/kg/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
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 hyperkeratosis at both dose levels for all formulations.
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Acifluorfen August, 1988
-6-
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.
0 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.
0 In an ocular irritation study (Weatherholtz et al., 1979a), 0.1 mL of
Blazer 25 (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 mg/kg/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 mg/kg/day. In terms of the active ingredient, a NOAEL
of 29.6 mg/kg/day was identified.
0 Barnett (1982) administered Tackle 2S (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,
2,500 or 5,000 ppm. The author indicated that these dietary levels
correspond to average compound intake levels of 0, 1.5, 6.1, 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
-------
Acifluorfen August, 1988
-7-
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 2S (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. Hales 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 mgAg/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
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
-------
Acifluorfen August, 1988
-8-
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. 28 (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
sexes; reductions in albumin and cholesterol; increased absolute and
relative 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-I mice (80/sex/dose)
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
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Acifluorfen August, 1988
-9-
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, Goldenthal et al. (1978b)
administered RH 6201 (a formulation containing sodium acifluorfen) in
the diet to Charles River CO 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 F2 and ?3 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 F2 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
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.
0 In a two-generation reproduction study, Lochry et al. (1986) admini-
stered technical grade Tackle (sodium acifluorfen) of unspecified
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Acifluorfen August, 1988
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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 parameters, 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
8 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 mg/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 2S (a formulation containing
22.4% sodium acifluorfen) 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 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 mgAg/day, lower
average fetal body weight and significantly delayed ossification of
metacarpals and forepaw and hindpaw phalanges were noted. At 180
mgAg/day, there was delayed ossification of caudal vertebrae,
steroebrae 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
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
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Acifluorfen August, 1988
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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.
Mutagenicity
0 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 mg/plate.
0 Brusick (1976) tested RH 6201 (purity not specified) in a mutagenicity
assay using Saccharomyces cersvisiae strain D4 and ^. typhimurium
strains TA 1535, 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 25 (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 McKeon (1981) conducted a primary rat (Fischer 344) hepato-
cyte unscheduled DMA 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 clastogenic events in the bone marrow cells.
• 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.
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.
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Acifluorfen August, 1988
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0 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 28 (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 B6C3F^ 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
mgAg/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,
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-1 mice (80/sex/dose)
for two years in the diet at concentrations that provided dose
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Acifluorfen August, 1988
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levela 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.
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.
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 (up to 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 LQAEL = 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, 1,000 or 10,000),
in accordance with EPA or 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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Acifluorfen August, 1988
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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 mgAg/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 acifluorfen via gavage during days 6 to 19
of gestation.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 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
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Acifluorfen August, 1988
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which the DWEL 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 DWEL.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = (1•25 mg/kg/day) (10kg) = g.13 mg/L (100 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 EPA
or 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 = d-25 mg/kg/day) (70 kg) _ 0.44 mg/L (400 ug/L)
(100) (2 L/day)
where:
1.25 mgAg/day = NOAEL (see Lifetime Health Advisory below).
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
2 L/day = assumed daily water consumption of an adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
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Acifluorfen August, 1988
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The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed. 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-1 mouse dietary study by Goldenthal (1979) was
originally selected to serve as the basis for determination of the DWEL 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 DWEL for acifluorfen is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (1.25 mg/kg/day) = 0.013 mg/kg/day
(100)
where:
1.25 mg/kg/day = NQAEL, based on the absence of mortality and kidney
lesions in rats.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NQAEL from an
animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-013 mg/kg/day) (70 kg) = 0.437 mg/L (40o ug/L)
(2 L/day)
where:
0.013 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day - assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Acifluorfen may be classified in Group B2: probable human carcinogen.
A Lifetime HA is not recommended for acifluorfen. The estimated excess
cancer risk associated with lifetime exposure to drinking water containing
acifluorfen at 437 ug/L is approximately 4 x 10"4 (4.4 x 10"4). This estimate
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Acifluorfen August, 1988
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represents the upper 95% confidence limit from extrapolations (U.S. EPA,
1988) prepared by using the Weibull 82 multistage model (time to death).
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., 1982a) indicated that sodium acifluorfen was car-
cinogenic in B6C3F-| mice. The results 'of the other three studies
(Goldenthal, 1979; Coleman et al., 1978; Barnett et al., 1982b)
provided no evidence of carcinogenicity in two strains of rats and
one strain of mice. Howeverr 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), B6C3F-| 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.
- 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 B6C3F1 mice [NCI, 1978,
1979; both as cited in NAS (1985)]. Although data on nitrofen cannot
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Acifluorfen August, 1988
-18-
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.
0 Evidence has been presented in one carcinogenicity study (Barnett
et al., 1982a) showing that acifluorfen is carcinogenic to mice. Using
the Weibull 82 increased multistage time-to-death with tumor model
(since there is a survival disparity for male mice), the U.S. EPA
(1988) reported a unit risk, q^, estimated in human equivalent
(surface area conversion) as 3.55 x 10~2 (mg/kg/day)~1 [B2J. It was
noted that qi* is an estimate of upper (95%) bound on risk and the
true value of risk is unknown and may be as low as zero.
0 Using this qi* value and assuming that a 70-kg human adult consumes
2 liters of water a day over a 70-year lifespan, the Weibull 82
multistage model estimates that concentrations of 100, 10 and 1 ug
acifluorfen per liter may result in excess cancer risk of 10-4, 10-5
and 10-6, respectively.
0 For comparison purposes drinking water concentrations associated with
an excess risk of 10~6 were 0.2 ug/L, 0.7 ug/L, 7 ug/L, <0.002 ug/L
and <0.002 ug/L for the multihit, one-hit, probit, logit and Weibull
models, respectively.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986a), acifluorfen is classified in
Group B2: probable human carcinogen.
VI. OTHER CRITERIA, GUIDANCE AND 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 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
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Acifluorfen August, 1988
-19-
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.
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, 1988
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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.
MRIO 00122737.
Brusick, D.* 1976. Mutagenicity evaluation of RH-6201. LSI Project No. 2547.
Unpublished study. MRID 00083057.
CFR. 1985. Code of Federal Regulations. July 1, 1985. 40 CFR 180.383.
p. 336.
CHEM1AB. 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, 1988
-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., Koberman, A.M. and Christian, M.S.* 1986. Two-generation Rat
Reproduction Study, Argus Research Laboratories, Inc. Study No.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 DMA 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, 1988
-22-
Schreiner, C.A., M.A. McKenzie and M.A. Mehlman.* 1980. An Ames Salmonella/
mammalian microsome mutagenesis assay for determination of potential
mutagenicity of Tackle 2S MCI0978. Study No. 511-80. Unpublished
study. MRID 00061622.
Schreiner, C., M. Skinner and M. Mehlman.* 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 Water by GC/ECD, January
1986 draft. Available from U.S. EPA's Environmental Monitoring and
Support Laboratory. Cincinnati, OH.
U.S. EPA. 1988. U.S. Environmental Protection Agency. CRAVE carcinogenicity
assessment for lifetime exposure. OPP. June 30.
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.
Whittaker Corporation.* No date, a. Acute oral LDsg 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. 410-0252. Unpublished study. MRID 00061628.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983.
The Merck Index, 10th ed. Rahway, NJ: Merck and Co., Inc.
WSSA. 1983. Herbicide Handbook, 5th Ed. Weed Science Society of America.
Champaign, IL.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August 1988
AMETRYN
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.
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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Ametryn
August 1988
_ 2 —
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 834-12-8
Structural Formula
I
H
2-(Ethylamino)-4-(isopropylamino)-6-(methylthio)-s-triazine
Synonyms
0 N-ethyl-N'-(1-niethylethyl)-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
Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Ametryn has been found in 2 of 1,190 surface water samples analyzed
and in 24 of 560 ground water samples (STORET, 1988). Samples were
collected at 215 surface water locations and 513 ground water
locations, and ametryn was found in 6 states. The 85th percentile of
227.35
Colorless crystals
84 to 85»C
8.4 x 10~7 mm Hg
185 mg/L
-1.72 (calculated)
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Ametryn August 1988
-3-
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. This information is
provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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).
• 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-isopropylamino-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 (t-|/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 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).
• Ametryn1s 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 14c-labeled ametryn orally to
Sprague-Dawley rats. Investigators stated that ametryn was admini-
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Ametryn August 1988
-4-
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.
Distribution
e 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
CO2> When the label wan in the ethyl side chain, 18.1% of the label
appeared as C02> 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
Oliver et al. (1969) studied the excretion of ametryn utilizing
uniformly labeled compound with 14c-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.
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Ametryn August 1988
-5-
Animals
Short-term Exposure
0 The following acute oral LD5Q values for ametryn in rats were
reported: Charles River CD rats, 1,207 rug/kg (males), 1,453 mg/kg
(females) (Grunfeld, 1981); mixed male and female rats (strain not
specified), 1,750 mg/kg (Stenger and Planta, 1961a); male and female
Wistar rats, 1,750 mg/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-ppm 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
rogAg of ametryn to mice 5 days before, on the day of or 2 days after
immunization with sheep erythrocytes (purity 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.
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Ametryn August 1988
-6-
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).
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%
ametryn in polyethylene glycolrsaline (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 LD$Q in
male and female rats was reported to be >3,170 mg/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 Heyer-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, kidney, 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.
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Ametryn August 1988
-7-
Reproductive Effects
0 No information was found in the available literature on the reproduc-
tive effects of ametryn.
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.
0 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-«ssay 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 (up to 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, 1,000 or 10,000),
in accordance with EPA or NAS/ODW guidelines.
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Ametryn August 1988
-8-
L/day = assumed daily water consumption of a child
(1 Vday) or an adult (2 L/day).
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 (9/00o 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 EPA
or NAS/OOW 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
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Ametryn August 1988
-9-
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.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = HO mg/kg/day) (10 kg) (6/7) = 0.86 /L (900 /L)
(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 EPA
or NAS/OCW 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 = (1° "gAg/day) (70 kg) (6/7) = 3 /L (3 000 /L)
(100) (2 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.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/OCW 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-
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Ametryn August 1988
-10-
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. 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.
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
rag/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) - Q.009 mg/kg/d)
(1,000) y *
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 EPA
or 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 mg/kg/day) (70 kg) =0.3 mg/L (300 ug/L)
(2 L/day)
-------
Ametryn August 1988
-11-
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.
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
9 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.
0 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 inade-
quate or no animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
8 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 (#507)
applicable to the determination of certain nitrogen-phosphorus
containing pesticides in water samples. 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.
This method has been validated in a single laboratory, and estimated
detection limits have been determined for the analytes in this method,
including ametryn, the estimated detection limit is 2.0 ug/L (U.S. EPA,
1988).
-------
Ametryn August 1988
-12-
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
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.
0 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.
0 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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Ametryn August 1988
-13-
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.
Domen^oz, R. 1961.* Ametryn: Toxicity in long-term administration. Unpub-
lished study. MRIO 00034838.
Grunfeld, Y. 1981.* Ametryn 80 w.p.: Acute oral toxicity in the rat.
Unpublished study. MRID 00100573.
Kpp»., 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 LD5g 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., 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. Planta.* 1961a. Oral toxicity in rats. Unpublished
study. MRID 00048226.
-------
Ametryn August 1988
-14-
Stenger, P. and V. Planta.* 1961b. Subchronic toxicity test no. 257.
Unpublished study. MRID 00048228.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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. 1988. U.S. EPA Method #507 - Determination of nitrogen and
phosphorus containing pesticides in water by GC/NPD, April, 1988 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: Weed 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
Programs•
-------
August, 1988
AMMONIUM SULFAMATE
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 arc 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.
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.
-------
Ammonium Sulfamate
August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 7773-06-0
Structural Formula
H,N - S - O - NH/
ammonium sulfamate
Synonyms
0 Amicide; Amidosulfate; Ammate; Ammonium amidosulfate; Ammonium
amidosulfonate; Ammonium amidotrioxosulfate; AMS; Fyran 206k; Ikurin.
Uses
0 Herbicide used to control woody plant species.
May be used for poison ivy control in apple and pear orchards
(Meister, 1986).
Properties (Meister, 1986)
H6N203S
114.14
Colorless crystals
131 to 132»C
>1
Negligible «10~7 mm Hg)
684 g/L
Chemical Formula
Molecular Weight
Physical State (25«C)
Boiling Point
Melting Point
Density (20°C)
Vapor Pressure*
Water Solubility*
Specific Gravity
Log Octanol Water/Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
(* WSSA, 1983)
Occurrence
0 No information was found in the available literature on the occurrence
of ammonium sulfamate.
Environmental Fate
a Kbnnai et al. (1974) showed that ammonium sulfamate was very mobile
in soil.
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Ammonium Sulfamate August, 1988
-3-
III. PHABMACOKIMBTICS
Absorption
0 No information was found in the available literature on the absorption
of ammonium sulfamate.
Distribution
0 No information was found in the available literature on the distribution
of ammonium sulfamate.
Metabolism
0 The metabolism of ammonium sulfamate in the urine of dogs was reported
by Bergen and Wiley (1938); however, details of the study were not
clearly defined for assessment.
Excretion
Bergen and Wiley (1938) reported 80 to 84% excretion of sulfamic acid
in the urine of two dogs following oral administration of ammonium
sulfamate in capsules for 5 days.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of ammonium sulfamate in humans.
Animals
Short-term Exposure
0 An oral LD50 of 3,900 mgAg in the rat is reported for ammonium
sulfamate (Meister, 1986).
Dermal/Ocular Effects
0 Five rats received a 20% aqueous solution of ammonium sulfamate (dose
level not specified) on the shaved skin of the back. They were
killed after 16 treatments on the 27th day of the period (Read and
Hueber, 1938). Another five rats received a 50% aqueous solution of
ammonium sulfamate on the shaved skin of the back. These animals
were killed after 11 treatments on the 19th day of the study. It
should be noted that the animals were not prevented from licking the
chemical. Investigators reported that there were no gross pathological
changes of importance in any of the animals. On microscopic patho-
logical examination of the animals, the spleen of 9 of 10 animals had
numerous macrophages with brown pigment. The stomach sections of seven
animals, revealed a brown, granular material in the surface capillaries
of the mucosa.
-------
Ammonium Sulfamate August, 1988
-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 mucose. occasionally contained
yellow-brown granules; in three animals, theze 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 macrophagea filled with hemosiderin. A fourth
spleen section showed m?rked erythrophagia.
Long-term Exposure
• Gupta et al. (1979) reported the results of a 90 days study involving
oral administration of 0, 100, 250 or 500 mg/kg of ammonium sulfamate
to rats (20 animals to each dose group) 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 mgAg* where
body weight was significantly less than controls after the end of
60 days. No significant changes in relative organ weights were
noticed in any group of rats. Merevtological 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 g/kg/day) level of feeding, but growth retardation
and a slight cathartic effect were observed at the 2% (20 gAg/day)
dietary level. No other information was provided by the authors.
• 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 or 500 ppm without any
clinical or nutritional evidence of toxicity. There were no histo-
pathological 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
• Sherman and Stula (1966) reported the results of a three-generation
reproduction study in rats. Rats receiving 0, 350 or 500 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, 1988
-5-
Developaental 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.
Carcinogenicity
0 N'O 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 (up to 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) = /I( ( /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowe9t-Observed-Adverse-Effeet Level
in mgAg bw/day.
BW a assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 data were located in the available literature that were sui- -Le for
deriving a One-day HA value for ammonium sulfamate. It is recommended that
the Longer-term HA value for the 10-fcg 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
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 (20 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, 1988
-6-
Longer-term Health Advisory
The aubchronic 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 mgAg/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 rag/kg ammonium
sulfamate showed lesser weight gain compared to other groups.
Using 250 mgAg/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 mgAg/day) (10 kg) (6/7) = 21<4 /L (20,000 ug/L)
(100) (1 L/day)
where:
250 mgAg/day = NOAEL, based on the absence of hematological and
histopathological changes in rats.
10 kg = assumed body weight of a child.
6/7 = conversion from 6 daya to 7 days.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW 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 m /L (80,000 ug/L)
(100) (2 Vday)
where:
250 mgAg/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 EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
2 L/day = assumed daily water consumption of an adult*
-------
Ammonium Sulfamate August, 1988
-7-
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three-step process. Step 1 determines the Reference Dose
(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 Hater 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. 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 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 mgAg/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 mgAg/day
(1,000)
where:
250 mgAg/day - NOAEL.
6/7 = conversion from 6 days to 7 days.
1,000 = uncertainty factor, chosen in accordance with EPA
or 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, 1988
-8-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - (0.214 mg/kg/day) (70 kg) = 7>5 mg/L (8f000 ug/L)
(2 L/day)
where:
0.214 mgAg/d-.y - 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 (2,000 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 or no 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, 1988
-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, 1988
-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 P.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.
Meister, R., ed. 1986. 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. MRID GS0016-0040.
Rosen, D.E., C.J. Krisher, H. Sherman and E.E. Stula. 1965. Toxicity studies
on ammonium sulfamate. The lexicologist. Fourth Annual Meeting, Williams-
burg, VA. March 8-10.
Sherman/ H. and E. Stula.* 1966. Toxicity studies on ammonium sulfamate.
Unpublished report. MRID GSOO16-0038.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. FDA. 1969. U.S. Food and Drug Administration. Pesticide analytical
manual, Vol. II. Washington, DC.
WSSA. 1983. Weed Science Society of America. Herbicide handbook, 5th ed.
Champaign, IL.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
August/ 1988
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.
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, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1912-24-9
Structural Formula
H
i
H H
2-Chloro-4-ethylamlno-6-isopropylamino-1 , 3, 5-triazine
Synonyms
A At rex; Atranex; Crisatrina; Crisazine; Farmco Atrazine; Griffex;
Shell Atrazine Herbicide; Vectal SC; Gesaprim; Primatol (Meister, 1987)
Uses
0 Atrazine over the past 30 years has been the most heavily used
herbicide in the U.S. It is used for nonselective weed control on
industrial or noncropped land and selective weed control in corn/
sorghum, sugar cane/ pineapple and certain other plants (Meister,
1 987) .
Properties (Meister/ 1987; Windholz, 1976)
Chemical Formula
Molecular Weight
Physical State
Boiling Point (25 mm Hg)
Melting Point
Density (20°)
Vapor Pressure (20°C)
Hater Solubility (22»C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C8H14C1N5
215.72
White, order less, crystalline solid
175 to 177»c
1.187
3.0 x 1Q-? mm Hg
70 rag/L
2 . 3 3 to 2 . 7 1
Occurrence
In a monitoring study of Mississippi River water, atrazine residues
were found at a maximum level of 17 ppb; residues were detected
throughout the year/ with the highest concentrations found in June
or July (Newby and Tweedy, 1976).
Atrazine has been found in 4,123 of 10,942 surface water samples
analyzed and in 343 of 3,208 ground water samples (STORET, 1988).
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Atrazine August, 1988
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Samples were collected at 1/659 surface water locations and 2/510
ground water locations. 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 2/300 ug/L and
in ground water it was 700 ug/L. Atrazine was foound in surface
water of 31 States and in ground water in 13 States. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime/ as they are here.
Therefore/ this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
0 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).
o 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).
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Atrazine August/ 1988
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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).
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).
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).
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).
Ciba-Geigy (1988) recently submitted comments on the atrazine Health
Advisory. These comments included a summary of the results of its
studies on the environmental fate of atrazine. This summary indicated
that laboratory degradation studies showed that atrazine is relatively
stable in the aquatic medium under environmental pH conditions and
indicated that atrazine degraded in soil by photolysis and microbial
processes. The products of degradation are dealkylated metabolites,
hydroxyatrazine and nonextractable (bound) residues. Atrazine and the
dealkylated metabolites are relatively mobile whereas hydroxyatrazine
is immobile.
Ciba-Geigy (1988) also indicated that field dissipation studies
conducted in California, Minnesota and Tennessee show no leaching of
atrazine and metabolites below 6 to 12 inches of soil. The half-
lives of atrazine in soil ranged between 20 to 101 days, except in
Minnesota where degradation was slow. A forestry degradation study
conducted in Oregon showed no adverse effects on either terrestial or
aquatic environments. Also, Bioconcentration studies have shown low
potential for bioaccummulation with a range of 15 to 77X.
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
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Atrazine August/ 1988
-5-
excretion after 72 hours was 20.3% of the administered dose; the
remainder was excreted in urine (65.5%) 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.
0 Bakke et al. (1972) administered single 0.53-mg doses of 14c-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. More than 80% 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).
0 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.
0 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.
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Atrazine August, 1988
-6-
Erickson et al. (1979) dosed Pittman-Moore miniature 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).
Hauswirth (1988) indicated that the rat metabolism studies taken
together are sufficient to show that in the female rat dechlorination
of the triazine ring and N-dealkylation are the major metabolic
pathways. Oxidation of the alkyl substituents appears to be a minor
and secondary metabolic route. The total body half-life is approximately
one and one-half days. Atrazine and/or its metabolites appear to bind'
to red blood cells. Other tissue accumulation does not appear to occur.
Excretion
Urine appears to be the principal route of atrazine excretion in
animals. Following the administration of 0.5 mg doses of 14c-ring-
labeled atrazine by gavage to rats, Bakke et al. (1972) reported that
in 72 hours most of the radioactivity (€5.5%) 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-Oawley 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
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 Schlicher 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. Although it is noted
that the exposure of this patient may have been inclusive to exposure
to other chemicals in addition to atrazine, it is also noted that
atrazine is a skin irritant in animal studies.
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Atrazine August/ 1988
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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. However, these data may not be
representative of the effect of atrazine since the exposed workers
were also exposed to other herbicides.
Animals
Short-term Exposure
0 Acute oral LD5Q values of 3,000 mg/kg in rats and 1,750 mg/kg in
mice have been reported for technical atrazine by Bashmurin (1974);
the purity of the test compound was not specified.
0 Acute oral studies conducted by Ciba-Geigy (1988) with atrazine
(97% a.i.) reflected the following LDsgs: 1,869 mg/kg in rats and
>3,000 mg/kg in mice.
0 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. It is noted that the dose
used in this study was almost 2 x the LDso (Ciba-Geigy, 1988).
0 Gaines and lander (1986) determined the oral 1*050 for adult male and
female rats to be 737 and 672 mg/kg respectively and 2,310 mg/kg for
pups. It is, therefore, noted that young animals are more sensitive
to atrazine than adults. This study also reflected that the dermal
LD50 for adult rats was higher than 2,500 mg/kg.
8 Palmer and Radeleff (1964) administered atrazine as a fluid dilution
or in gelatin capsules to Delaine sheep and dairy cattle (one animal
per dosage group). Two doses of 250 mg/kg atrazine caused death in
both sheep and cattle. Sixteen doses of 100 mg/kg were lethal to the
one sheep tested. At necropsy, degeneration and discoloration of the
adrenal glands and congestion in lungs, liver and kidneys were observed.
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Atrazine August, 1988
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0 Palmer and Radeleff (1969) orally administered atrazine SOW (analysis
of test material not provided) by capsule or by drench to sheep at 5,
10, 25, 50, 100, 250 or 400 mg/kg/day and to cows at 10, 25, 50, 100
or 250 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.
0 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 (ChE) activity•
No other details of this study were reported. Atrazine is not an
organophosphate (OP), therefore, its effect on ChE may not be similar
to the mechanism of ChE inhibition by OPs.
Dermal/Ocular Effects
0 In a primary dermal irritation test in rats, atrazine at 2,800 mg/kg
produced erythema but no systemic effects (Gzheyotskiy et al., 1977).
0 Ciba-Geigy (1988) indicated that the studies it performed reflected
dermal sensitization in rats but not irritation in rabbits' eyes.
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
were necropsied 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.
0 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
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Atrazine August, 1 988
-9-
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.
0 In a study by Ciba-Geigy (1987b) using technical atrazine (97% ai.)/
six-month-old beagle dogs were assigned randomly to four dosage
groups: 0, 15, 150 and 1,000 ppm. These doses correspond to actual
average intake of 0, 0.48, 4.97 and 33.65/33.8 (male/female) mg/kg/day.
Six animals/sex/group were assigned to the control and high dose groups
and four animals/sex/group were assigned to the low- and mid-dose
groups. One mid-dose male, one high-dose male and one high-dose female
had to be sacrificed moribund during the study period. Decreased body
weight gains and food consumption were noted at the high-dose level•
Statistically significant (p <0.05) reductions in erythroid parameters
(red cell count, hemoglobin and hematocrit) in high-dose males were
noted throughout the study as well as mild increases in platelet
counts in both sexes. Slight decreases in total protein and albumin
(p < 0.05) were noted in high-dose males as well as decreased calcium
and chloride in males and increased sodium and glucose in females.
Decrease in absolute heart weight were noted in females and increased
relative liver weight in males of the high-dose group. The mid-dose
females reflected an increase in the absolute heart weight and
heart/brain weight ratios. The most significant effect of atrazine
in this study was reflected in the high-dose animals of both sexes
as discrete mycardial degeneration. Clinical signs associated with
cardiac pathology such as ascites, cachexia, labored/shallow breathing
and abnormal EKG were observed in the group as early as 17 weeks into
the study. Gross pathology reflected severe dilation of the right
atrium and occasionally of the left atrium. These findings were also
noted histopathologically as degenerative atrial myocardium (atrophy
and myolysis). In the mid-dose group, two males and one female
appeared to be affected with the cardiac syndrome but to a much lesser
degree in the intensity of the noted responses. Therefore, the LOAEL
in this study is 4.97 mg/kg/day and the NOAEL is 0.48 mg/kg/day.
0 A two year chronic feeding/oncogenicity study (Ciba-Geigy, 1986) was
recently evaluated by the Agency. In this study, technical atrazine
(98.9% a.i.) was fed to 37 to 38 days-old Sprague-Dawley rats. The
dosage levels used were 0, 10, 70, 500 or 1,000 ppm, equivalent to
0, 0.5, 3.5, 25 or 50 mg/kg/day (using Lehman's conversion factor,
1959). Twenty rats per sex per group were used to measure blood
parameters and clinical chemistries and urinalysis. Fifty rats per
sex per goup were maintained on the treated and control diets for
24 months. An additional 10 rats per sex were placed on control and
high dose (1,000 ppm) diets for a twelve month interim sacrifice and
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Atrazine August, 1988
-10-
another 10 per sex (control and high dose/ 1/000 ppm) for a 13 month
sacrifice (the 1/000 ppm group was placed on control diet for one month
prior to sacrifice). The total number of animals/sex in the control
and HOT groups was 90 and 70 for the 10, 70 and 500 ppm groups. Histo-
pathology was performed on all animals. At the mid- and high-dose/
there was a decrease in mean body weights of males and females.
Survival was decreased in high-dose females but increased in high-dose
males. There were decreases in or.gan-to-body weight ratios in high-dose
animals/ which were probably the result of body weight decreases.
Hyperplastic changes in high-dose males (mammary gland/ bladder and
prostate) and females (myeloid tissue of bone marrow and transitional
epithelium of the kidney) were of questionable toxicologic importance.
There was an increase in retinal degeneration and in centrolobular
necrosis of the liver in high-dose females and an increase in
degeneration of the rectus femoris muscle in high-dose males and
females when compared to controls. Based on decreased body weight
gain/ the LOAEL for non-oncogenic activities in both sexes is
25 mgAg/day and the NOAEL is 3.5 mg/kg/day. However/ oncogenic
activities were noted at 3.5 mg/kg/day (70 ppm) and above as reflected
in the increased incidence of mammary gland tumors in females.
0 A recent 91-week oral feeding/oncogenicity study in mice by Ciba-
Geigy (1987c) has been evaluated by the Agency. In this study,
atrazine (97% ai.) was fed to five-weeks-old CD-1 strain of mice,
weighing 21.0/26.8 grams (female/male). The mice were randomly
assigned to five experimental groups of approximately 60 animals/sex/
group. The dosage tested were 0, 10, 300, 1,500 and 3/000 ppm; these
dosages correspond to actual mean daily intake of 1.4, 38.4, 194.0 and
385.7 mg/kg/day for males, and 1.6, 47.9, 246.9 and 482.7 mg/kg/day
for females. This study shows that there are dose-related effects
at 1,500 ppm or 3,000 ppm atrazine: an increase in cardiac thrombi,
a decrease in the mean body weight gain at 12 and 91 weeks during the
study, and decreases in erythrocyte count/ hematocrit and hemoglobin
concentration. Cardiac thrombi contributed to the deaths of the
group of mice that did not survive to terminal sacrifice. The LOAEL
is set at 1,500 ppm based upon decreases of 23.5% and 11.0% in mean
body weight gain found at 91 weeks in male and female mice, respectively.
Also, an increase in the incidence of cardiac thrombi is found in
female mice in the 1,500 ppm exposure group. None of the above effects
are found at 300 ppm, thus the NOAEL is set at 300 ppm (corresponding
to 38.4 mg/kg/day in males and 47.9 mg/kg/day for females).
Reproductive Effects
0 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 (SOW) 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.
Two litters were produced per generation but parental animals were
chosen from the second litter after weaning for each generation.
Young rats were maintained on the test diets for approximately ten
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Atrazine August, 1988
-11-
weeks in each generation. The third generation pups were sacrificed
after weaning. It is noted that the parental animals of the first
generation were fed only half of the dietary atrazine levels for the
first 3 weeks of exposure. There were no adverse effects of atrazine
on reproduction observed during the course of the three-generation
study. A NOAEL of 100 ppm (5 mg/kg/day) was identified for this
study. However, the usefulness of this study is limited due to the
alteration of the atrazine content of the diet during important
maturation periods of the neonates.
0 A recent two-generations study in rats by Ciba-Geigy (1987a) was
conducted using the 97% ai. technical atrazine. Young rats, 47 to
48 days old were maintained on the control and test diets for 10
weeks before mating. The concentrations used were 0, 10, 50 and
500 ppm (equivalent to 0, 0.5, 2.5 and 25 mg/kg/day using Lehman
conversion factor, 1959). Thirty animals/sex/group were used in each
generation; one litter was produced per generation. The level tested
had no effect on mortality in either generation. Body weight and
body weight gains were significantly depressed (p <0.05) at the
highest dose; however, food consumption was also decreased at this
high-dose level in parental males and females during the premating
period and for the first generation females (FI> on days 0 to 7 of
gestation. No histopathological effects were noted nor other effects
were noted during gross necropsy in either parental generation with
the exception of increased testes relative weight in both generations
at the high dose. In pups of both generation, significant reduction
(p <0.05) in body weight was noted; however, this effect was only
dose-related in the second generation (F2> at both the mid- and
high-dose levels on postnatal day 21. Therefore, maternal toxicity
NOAEL is 2.5 mgAg/day; the reproductive LOAEL is 2.5 mg/kg/day
(reduced pup weight in ?2 generation on postnatal day 21) and the
NOAEL is 0.5 mgAg/day.
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/kg/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 ppm (5 mg/kg/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 rag/kg (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 skeletal development.
No teratogenic effects were observed. The highest dose (1,000 mg/kg)
resulted in 23% maternal mortality and various toxic symptoms. The
100 mg/kg dose had no effect on either dams or embryos and is therefore
the maternal and fetotoxic NOAEL in this study.
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Atrazine August, 1988
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0 In a study by Clba-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/kg/day. Excessive maternal mortality (21/27) was noted at
700 mg/kg/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/kg/day. Developmental toxicity was also present at these dose
levels. Fetal weights were severely reduced at 700 mg/kg/day; delays
in skeletal development occurred at 70 mg/kg/ day/ and a dose-related
runting was noted at 10 mg/kg/day and above. The NOAEL for maternal
toxicity appears to be 10 mg/kg/day/ however/ 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/kg/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/kg/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/kg/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 ppm in the diet
throughout gestation. Based -on an assumed body weight of 300 g and
a daily food consumption of 12 g (Arrington, 1972), these levels
correspond to approximately 0, 2, 4, 8, 12, 16, 20 or 40 mg/kg/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/kg/day for developmental effects. In another phase of
this study, the authors demonstrated that subcutaneous (sc) injections
of 50, 100 or 200 mg/kg atrazine on gestation days 3, 6 and 9 had no
effect on the litter size, while doses of i.800 mg/kg were embryotoxic.
Therefore, a NOAEL of 200 mg/kg 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.
0 Murnik and Nash (1977) reported that feeding 0.01% atrazine to male
Drosophila melanogaster larvae significantly increased the rate of
both dominant and sex-linked 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 Environmental Research Programme of the Commission
of the European Communities. Mutagenic activity was not induced even
when mammalian liver enzymes (S-9) were used; however/ the use of
plant microsomes produced positive results. Also/ in in vivo studies
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Atrazine August, 1988
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in mice/ atrazine induced dominant lethal mutations and increased the
frequency of chromatid 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 stomach.
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 unexposed period.
0 Recently, Spencer (1987) and Dearfield (1988) evaluated several in
vitro and in vivo mutagenicity studies on atrazine that were recently
submitted to the U.S. EPA by Ciba-Geigy. They noted that most of
these studies were inadequate with the exception of the following
three tests: a Salmonella assay; an E. coli reversion assay; and a
Host-Mediated assay. The first two assays were negative for mutagenic
effects; the results of the third assay were equivocal.
0 Ciba-Geigy (1988) indicated that Brusick (1987) evaluated atrazine
mutagenicity and that the weight-of-evidence analysis he used placed
the chemical in a non-mutagenic status. The Agency (Dearfield, 1988)
evaluated Brusick's analysis. It is noted that the use of the weight-
of-evidence approach is not appropriate at the present time. The
in vivo studies by Adler (1980) suggest a positive response. These
findings have not been diminished by other atrazine studies. In
addition, Dearfield (1988) indicated that the scheme used by Brusick
in this analysis is flawed by the lack of calibration of the chemical
test scores to an external standard and by the use of some studies
that are considered inadequate by design to determine the mutagenic
potential of atrazine.
Carcinogenicity
0 Innes et al. (1969) investigated the tumorigenicity of 120 test com-
pounds including atrazine in mice. Two F-j hybrid stocks (C57BL/6 x Anf)
Fj and (C57BL/6 x AKR) F-j 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 ppm atrazine ad libitum in the diet for 18 months. The incidence
of hepatomas, pulmonary tumors, lymphomas and total tumors in atrazine-
treated mice was not significantly different from that in the negative
controls.
0 A two-year feeding/oncogenicity study in rats by Ciba-Geigy (1986)
has been evaluated recently by the Agency. Atrazine (98.9% a.i.)
was fed to 37 to 38 days-old Sprague-Dawley rats. The dosage levels
used were 0, 10, 70, 500 or 1,000 ppm, equivalent to 0, 0.5, 3.5,
25 or 50 mg/kg/day (using Lehman's conversion factor, 1959). The
total number of animals/sex in the control and HOT groups was 90; and
70 animals/sex/group for the 10, 70 and 500 ppm groups. Histopathology
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Atrazine August, 1988
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was performed on all animals. In females, atrazine was associated with
a statistically significant increase in mammary gland fibroadenomas
at 1,000 ppm, in mammary gland adenocarcinomas (including two carcino-
sarcomas at the HDT) at 70, 500 and 1,000 ppm, and in total mammary
gland tumor bearing animals at 1,000 ppm. Each of these increases
was associated with a statistically significant dose-related trend
and was outside of the high end of the historical control range. In
addition, U.S. EPA (1986a) indicated that there was evidence for
decreased latency for mammary gland adenocarcinomas at the 12 month
interim sacrifice that was already submitted by Ciba-Geigy in 1985.
This study was also reported as positive in a briefing paper by
Ciba-Geigy (1987).
0 A recent 91-week oral feeding/oncogenicity study in mice by Ciba-
Geigy (1987c) has been evaluated by the Agency. In this study,
atrazine (97% ai.) was fed to five-weeks-old CD-I mice weighing
21.0/26.8 grams (female/male). The mice were randomly assigned to
five experimental groups of approximately 60 animals/sex/ group.
The dosage tested were 0, 10, 300, 1,500 and 3,000 ppm; these dosages
correspond to actual mean daily intake of 1.4, 38.4, 194.0 and 385.7
mg/kg/day for males, and 1.6, 47.9, 246.9 and 482.7 mg/kg/day for
females. The following kinds of neoplasms were noted in this study:
mammary adenocarcinomas, adrenal adenomas, pulmonary adenomas and
malignant lymphomas. However, no dose-related or statistically
significant increases were observed in the incidences of these
neoplasms. Therefore, atrazine is not considered oncogenic in this
strain of mice.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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 = (HOAEL 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/k9 bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000)
in accordance with EPA or 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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Atrazine August, 1988
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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-kg 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 in the 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/kg/day for maternal toxicity but this
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/kg/day for
maternal toxicity. Thus/ the rabbit appears to be a more sensitive species than
the rat for maternal toxicity/ hence/ the rabbit study with a NOAEL of 1 mg/kg/day
is used in the calculations below.
The Ten-day HA for a 10 kg child is calculated below as follows:
(1 mg/kg/d) x (10kg) = 0., mgL (100 ug/L)
(100) x (1 L/day)
where:
1 mg/kg/day - NOAEL/ based on maternal toxicity evidenced by decreased
body weight gain and food consumption.
10 kg = assumed body weight of a child.
100 = uncertainty factor/ chosen in accordance with EPA or
ODW/NAS guidelines for use with a NOAEL from an animal
study.
1 L/day = 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 0.05 mg/L (50 ug/L)
and the DWEL for a 70-kg adult of 0.2 mg/L (200 ug/L) be used at this time as
conservative estimates 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
(RfO)/ formerly called the Acceptable Daily Intake (ADI). The RED is an esti-
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Atrazine August/ 1988
-16-
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. 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.
Three studies were considered for the development of the Lifetime HA. A
two-year dog feeding study (Woodard/ 1964), a one-year dog feeding study Ciba-
Geigy, 1987b) and a two-year rat oral feeding/oncogenicity study (Ciba-Geigy,
1986).
The first study in dogs (1964) reflected a NOAEL of 0.35 mg/kg/day and a.
LOAEL of 3.5 mg/kg/day that was associated with increased heart and liver
weights in females. The new one-year dog study (1988) reflected a NOAEL of
0.48 mg/kg/day and a LOAEL of 4.97 mg/kg/day based on mild cardiac pathology
intensified at the higher dose tested 33.65/33.8 (male/female) mg/kg/day.
The two-year rat study (Ciba-Geigy, 1986) reflected a NOAEL at 3.5 mg/kg/day
for systemic effect other than oncogenicity; however, this study indicated
that atrazine caused mammary gland tumors at this dose level and above, no
adverse effects were observed at the lowest dose tested, 0.5 mg/kg/day.
The 1964 dog study was initially used for the calculation of the RfD and
the Lifetime HA. However, this study was partially flawed by the lack of
information on the purity of the test material and by the inadequate document-
ation of the hematological data. Therefore, the recent one-year dog study
(Ciba-Geigy/ 1987b), using technical atrazine (97% ai.), is considered as a
more adequate study for the calculation of the RfD and the Lifetime HA. The
NOAEL in this study, 0.48 mg/kg/day, is also supported by the NOAEL of 0.5
mgAg/day in the two-generation reproduction study (Ciba-Geigy, 1987a) and by
the fact that no systemic effects or tumors were noted at this dose level in
the two-year chronic feeding/oncogenicity study in rats (Ciba-Geigy/ 1986).
[Other studies: Woodard Research Corporation (1966) and 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 dog studies for use in calculating the
HA values for atrazine.]
Step 1: Determination of the Reference Dose (RfD)
RfD = 0.48 mg/kg/day = 0.005 mg/kg/day
(100)
(rounded from 0.0048 mg/kg/day)
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Atrazine August/ 1988
-17-
where:
0.48 mg/kg/day = NOAEL, based on the absence of cardiac pathology
or any other/adverse clinical/ hematological, bio-
chemical and histopathological effects in dogs.
100 = uncertainty factor, chosen in accordance with
EPA or ODW/NAS guidelines for use with a NOAEL
from an animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (Q'0048 mg/kg/day) (70 kg) . 0. 168 mg/L (20o ug/L)
(2 L/day)
where:
0.0048 mg/kg/day = RfD (before rounding off to 0.005 mg/kg/day)
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>168 m?/L> <20%) = 0.003 mg/L (3 ug/L)
10
where:
0.168 mg/L = DWEL (before rounding off to 0.2 mg/L)
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 A study submitted by Ciba-Geigy Corporation (1986) in support of the
pesticide registration of atrazine indicated 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 this study.
0 Atrazine was not oncogenic in mice (Ciba-Geigy/ 1987c).
0 Three closely related analogs: propazine/ terbutryn and simazine are
presently classified as Group C oncogens based on an increased incidence
of tumors in the same target tissue (mammary gland) and animal species
(rat) as was noted for atrazine.
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Atrazine August/ 1988
-18-
The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of atrazine.
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.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Toxicity data on atrazine were reviewed by the National Academy of
Sciences (HAS, 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
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, Method
No. 507, applicable to the determination of certain nitrogen-phosphorus
containing pesticides in water samples (U.S. EPA, 1988). 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 has been validated in a single Labora-
tory. The estimated detection limit for the analytes in this method,
including atrazine, is 0.13 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
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Atrazine August/ 1988
-19-
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.
0 Harris and Warren (1964) reported that Amberlite IR-120 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 anion exchange resins. At a pH of 7.2, 45%
removal of atrazine was achieved with Oowex® 2 anion exchange resin
and with H2PO4~ as the exchangeable ion species.
0 Chian 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.
0 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.
0 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' August, 1988
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L. Zaccaro/ A. Abbondandolo and S. Bonatti. 1980. Results of mutagenicity
tests on the herbicide atrazine. Mutat. Res. 74:250. Abstract.
Lykins/ Jr./ B.W./ E.E. Geldreich, J.Q. Adams/ J.C. Ireland and R.M. Clark.
1984. Granular activated carbon for removing nontrihalomethane 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.
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.
Molnar, V. 1971. Symptomatology and pathomorphology of experimental poisoning
with atrazine. Rev. Med. 17:271-274. (English abstract only)
Murnik, 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/ L./ and B.C. Tweedy. 1976. Atrazine residues in major rivers and
tributaries. Unpublished study submitted by Ciba-Geigy Corporation/
Greensboro/ N.C.
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Atrazine August/ 1988
-24-
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. 106.
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.
Schlicher/ 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. Emmond and H.C. Kbrven. 1975. Persistence and
movement of atrazine/ bromacil/ monuron/ and simazine in intermittently-
tilled irrigation ditches. Can. J. Plant Sci. 55:809-816.
Spencer/ H. 1987. Review of several mutagenicity studies on atrazone.
U.S. EPA, Office of Pesticide Programs' review of a Ciba-Geigy data
submission. Accession No. 284052. HRID 402466-01.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency. (Data' file search conducted in March/ 1988).
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.
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. 1988. U.S. Environmental Protection Agency. Method #507 - Determina-
tion of nitrogen and phosphorus containing pesticides in ground water by
GC/NPD. April 14, draft.
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Atrazine August/ 1988
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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.W. 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
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Windholz/ M./ ed. 1976. The Merck index. 9th ed. Rahway, NJ: Merck and
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•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
(BAYGON) Propoxur
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.
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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 114-26-1
Structural Formula
2-( 1-Methylethoxy)-phenol methylcarbamate
Synonyms
0 Baygon Propoxur (proposed common name); Aprocarb; Blattanex;
BAY 39007; Bayer 39007; Pillargon; Propyon; Suncide; Tugon;
QMS 33; Uhden (Meister, 1984).
Uses
A nonfood insecticide used on humans, animals and turf grass
(Meister, 1984).
Propoxur is an insecticide currently registered for use on a variety
of terrestial non-food (ornamentals, lawns, and general outdoor
areas), aquatic non-food (sewage systems and stagnant water), forestry,
domestic outdoor, and indoor sites. Application rates are dependent
on use. For use as a bait, applications include 1.81-18.14 g/1,000 sq
ft. For use on turf, applications include 39.59-85.05 g/1,000 sq ft.
For outdoor uses, applications include 0.046-0.175 Ib/A, 0.375-1.125
Ib/mile with a 300 ft swath, 0.493-2% finished spray, 0.14-0.45/1,000
sq ft, 2.1% D applied directly to ant hills, and 6 strips containing
10% Impr/gypsy month monitoring trap. For treatment of premises,
applications include 0.25-1% finished spray and 0.06-0.28 g/1,000 sq
ft. For indoor treatments, applications include 0.25% G, 0.25-2% P/T,
0.25-1.7% Impr, 1% RTU in traps/bait trays, one fly strip containing
0.4% Impr/1,000 cu ft, contact strips containing 4-10% Impr, shelf
paper containing 0.6-1% Impr, and 0.493-2% finished spray. For
treatment of animals, applications include 0.9% Impr (dab-on), 9.4%
Impr (collars), 8.5-56.7 g/gal (dips), 0.125% RTU (shampoo), and
0.25-0.28% finished spray. Propoxur may be applied with ground or
aerial equipment.
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Baygon August/ 1988
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Propertles (ACGIH, 1984; Meister, 1984; Worthing, 1983; and CHEMLAB, 1985)
Chemical Formula C •) -|H 1
Molecular Weight 209.24
Physical State (at 25°C) White to tan crystalline solid
Boiling Point
Melting Point 91 8C
Density (°C)
Vapor Pressure (120°C) 0.01 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 85 of 624 surface water samples analyzed
and in 0 of 114 ground water samples (STORET, 1988). Samples were
collected at 21 surface water locations and 111 ground water locations.
The 85th percentile of all mean-zero samples was 0. 96 ug/L in surface
water. The maximum concentration found in surface water was also
0.96 ug/L. Baygon was found only in surface water and this finding
was reported only in Michigan and Ohio. This information is provided
to give a general impression of the occurrence of this chemical in
ground and surface waters as reported in the STORET database. The
individual data points retrieved were used as they came from STORET
and have not been confirmed as to their validity. STORET data is
often not valid when individual numbers are used out of the context
of the entire sampling regime/ as they are here. Therefore, this
information can only be used to form an impression of the intensity
and location of sampling for a particular chemical.
Environmental Fate
Ring-labeled 14c-propoxur (radiochemical purity 97%), at 5 ppm,
degrades with a half -life of >28 days in irradiated and non-irradiated
dry sandy loam soil (McNamara and Moore, 1981). Propoxur comprised
75% of the applied in the irradiated soil at 28 days post-treatment.
2-(1-Methylethoxy)phenol comprised <2% in both irradiated and non-
irradiated soil at all sampling intervals. Approximately 1% of the
applied radioactivity was isolated in trapping solutions from the
irradiated samples. Radioactivity was not detected in trapping
solutions from dark control samples.
14c-Propoxur (radiochemical purity >95%), at 0.94-98.6 ug/10 mL/ was
very mobile (Freundlich K^g 0.05-0.30) in sandy loam, silt loam/
and silty clay soil:distilled water slurries (1:5 soil:solution
ratio) based on batch ecuilibrium studies (Lenz and Gronberg, 1980).
Desorption was quite variable; 0-100% (0-5. 9 ug/10 mL) of the adsorbed
was desorbed from the soil.
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Baygon August/ 1988
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0 Aged (28 days) ring- and carbonyl-labeled 14c-propoxur residues
(78-89% propoxur) were very mobile in columns (12-inch length) of silt
loam soil leached with 22.5 inches of water over a 45-day period
(Atwell, 1976). Between 69 and 74% of the recovered 14c-residues were
leached through the columns. The remaining residues were distributed
evenly throughout the columns. In the leachate, 100% of the carbonyl-
labeled 14C-residues and 77% of the ring-labeled 14c-residues were
identified as propoxur; 23% of the ring-labeled 14c-residues were
2-(1-methylethoxyJphenol.
0 Uhaged 14c-propoxur (radlochemical purity >98%), was mobile in
columns of sandy loam and silt loam soil; 70-79% of the applied
radioactivity was leached from the columns (Moellhoff, 1983).
Propoxur was the major 14c-compound in both the leachates and the
columns treated with imaged propoxur. Aged (30 days) ^C-propoxur
residues were slightly mobile in both soils; 72-75% of the applied
radioactivity remained in the soil columns. Propoxur was the
predominant compound in both the soil and leachate after leaching.
Propoxur was very mobile on sand, sandy loam/ sandy clay loam/ silt
loam/ and silty clay soil TLC plates/ with Rf values ranging from
0.70 to 0.89 (Thornton et al., 1976).
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 rag) 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 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).
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Baygon August/ 1988
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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. Zsopropoxyphenol 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.
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 14c or 3n 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
Dawson 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
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Baygon August/ 1988
-6-
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 the 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
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
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Baygon August/ 1988
-7-
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 that time. In severe
cases/ atropine was administered as an 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.
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 LD50 value for technical Baygon (purity not specified)
in male and female Sherman rats was reported to be 83 and 86 mg/kg,
respectively (Gaines, 1969). The oral LDso was 'reported to be 23.5 mg/kg
in mice and 40 mg/kg in guinea pigs (NIOSH, 1987).
• Farbenfabriken Bayer (1961) determined an oral LDso of 100 to 150 mg/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 FB30)
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
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Baygon August/ 1988
-8-
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/ the 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
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. Mb 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 LD5Q 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).
0 Heimann (1982) demonstrated that Baygon (98.8% pure) is not a skin
sensitizer when tested in guinea pigs.
0 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.
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Baygon August/ 1988
-9-
Long-term Exposure
0 Eben and Kimmerle (1973) fed propoxur (98.7% a.i.) to male SPF Wistar
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.
0 Itoot 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 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, 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).
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 for 106 weeks. 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.
No significant inhibition was reported in the examined dose group.
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 (25/sex/dose). 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 assump-
tion that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day
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Baygon August/ 1988
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(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).
Loser (1968b) conducted a 2-year study of Baygon toxicity in beagle
dogs (4/sex/dose). The product, BAY 39007 (technical, 99. 8% pure),
was included in 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,
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Baygon August, 1988
-11-
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
mgAg/day).
0 Bomhard and Loeser (1981) conducted a 2-year feeding study of propoxur
(99.6% a.l.) in SPF CF1/W74 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,
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 mg/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 mg/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).
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Baygon August/ 1988
-12-
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 ppm (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 other fetal measurements were in the
control range. No terata 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 ppm (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-related decrease in
maternal weight gain and fetal weight at the lowest dose tested
(although these effects were not statistically significant).
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 at 10 to 1/500 ug/plate
(with or without microsomal activation).
0 Moriya et al. (1983) tested Baygon at up to 5/000 ug/plate in five
strains of £. 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 j>. 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 SFF strain Bor. WISH)
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
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
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Baygon August/ 1988
-13-
females, respectively). Bladder papillomas were observed in both
males (25/57) 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 Bombard and Loeser (1981) conducted a 2-year feeding study of propoxur
(99.5% a.i.) in SPF CFj/W74 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.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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) = m /L ( /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, 1,000 or 10,000),
in accordance with EPA or 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)
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Baygon August/ 1988
-14-
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/ and since the LOAEL of 0.45 mg/kg (three-
fifths of a 0.75 mg/kg total dose) caused a similar level of ChE inhibition
as the bolus dose of 0.36 mg/kg (ChE inhibition is in the 40% range). There-
fore, the LOAEL, 0.36 mg/kg is used for the calculation below:
The One-day HA for a 10-kg child is calculated as follows:
One-day HA = (0-36 mg/kg/day) (10 kg) = 0.036 mg/L (40 ug/L)
(100) (1 L/day)
where:
0.36 mg/kg/day = LOAEL, based on mild cholinergic signs and 43% in-
hibition of red blood cell ChE in humans 10 minutes
after single oral dose.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA or
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/
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Baygon August/ 1988
-15-
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) resulted in the same level of red blood cell ChE inhibition.
Therefore, the lower dose/ 0.36 mg/kg/day/ is used for calculation of the
Ten-day HA.
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA = (0.36 mg/kg/day) (10 kg) = 0.036 /L (40 /L)
(100) (1 L/day)
where:
0.36 mg/kg/day = LOAEL, based on mild cholinergic signs and 43%
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 with EPA or
NAS/OEW guidelines for use with a LOAEL from a human
study.
1 L/day = assumed daily water consumption of 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 Baygon. It is/ therefore/
recommended that the modified Drinking Water Equivalent Level (OWED 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/ noncarcinogenlc health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
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Baygon August/ 1988
-16-
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. 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 (196Sb) 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, the 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
on the One-day and Ten-day HAs. The 2-year mouse study by Bombard 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/k?/day) = 0.004 mg/kg/day
(100)
(after rounding off from 0.0036 mg/kg/day)
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 EPA or
NAS/ODW guidelines for use with a LOAEL from a human
study.
Step 2: Determination of the Drinking Hater Equivalent Level (DWEL)
DWEL = (0-0036 mg/kg/day) (70 kg) _ Q. 126 mg/L (100 ug/L)
(2 L/day)
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Baygon August, 1988
-17-
where:
0.0036 mg/kg/day = RfD (before rounding off to 0.004 mg/kg/day).
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 = (0.126 mg/L) (20%) = 0.003 mg/L (3 ug/L)
(10)
where:
0.126 mg/L = DWEL (before rounding off to 100 ug/L).
20% = assumed relative source contribution from water.
10 = additional uncertainty factor in accordance with ODW
policy to account for possible carcinogenicity.
Evaluation of Carcinogenic Potential
0 Suberg 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
0 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.
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Baygon August/ 1988
-18-
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 531.1 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 fluores-
cence detector. This method has been validated in a single laboratory.
The limit of detection for this method for Baygon is 1.0 ug/L.
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 August, 1988
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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.
Atwell, S.H. 1976. Leaching characteristics of Baygon on aged soil. Report
No. 50718. Unpublished study received Oct. 21, 1981 under 3125-306;
submitted by Mobay Chemical Corp., Kansas City, MO; CDL:246088-L.
(00085769).
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.
Bombard, E., and E. Loeser.* 1981. Propoxur, the active ingredient of Baygon:
Chronic toxicity study in 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 QMS-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. Thaln 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 mutagenlc 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. Army 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.
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Baygon August, 1988
-20-
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.
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. Institut fur Toxicologie. Unpublished study. MRID 00035412.
Foss, W., and J. Krechniak. 1980. The fate of propoxur in rat. Arch. Toxicol.
4:346-349.
Gaines, 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. S.
Lenz, M.F. and R.R. Gronberg. 1980. Soil adsorption and desorption of
Baygon. Report No. 69016. Unpublished study received Oct. 21, 1981
under 3125-306; submitted by Mobay Chemical Corp., Kansas City, MO;
CDL:246088-N. (00085770).
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.
McNamara, F.T. and K.D. Moore. 1981. Photodecomposition of Baygon on soil.
Report NO. 69476. Unpublished study received Oct. 21, 1981 under 3125-
306; submitted by Mobay Chemical Corp., Kansas City, MO; CDL:246088-F.
(00085765).
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Baygon August, 1988
-21-
Meister, R., ed. 1984. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Moellhoff, E. 1983. Leaching behavior of aged propoxur (Baygon) residues
(laboratory study): RA-162-172B. Unpublished study prepared by Bayer
AG. 12 p. (00149025).
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.
NIOSH. 1987. National Institute for Occupational Safety and Health. Registry
of toxic effects of chemical substances. Sweet/ D.U., ed. Cincinnati/
OH: National Institute for Occupational Safety and Health. Microfiche
edition.
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. HRID 00040447.
Schlueter, G./ and D. Lorke.* 1981. Propoxur/ the active ingredient of
Baygon: Study of embryotox'ic and teratogenic effects on rabbits after
oral administration. Bayer Report No. 10183; MOBAY ACO Report No. 80034.
Bayer AG Institut fur Toxlcologie. Unpublished study. MRID 00100547.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in March/ 1988).
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.
Thornton, J.S., J.B. Hurley and J.J. Obrist. 1976. Soil thin-layer mobility
of twenty-four pesticide chemicals. Mobay Report No. 51016. Prepared
by Mobay Chemical Corp., Kansas City, MO; submitted by Chevron Chemical
Company, Richmond, CA. Ace. No. 259942. Reference 6.
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 #531.1.
Measurement of N-methyl carbamoyloximes and N-methylcarbamates in ground
water by direct aqueous injection HPLC with post column derivatization.
January 1986 draft. Available from EPA's Environmental Monitoring and
Support Laboratory/ Cincinnati, OH 45268.
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.
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Baygon August, 1988
-22-
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.
Worthing, C.R. 1983. The Pesticide Manual: A World Compedium, 7th ed.,
London:BCPC Publishers.
•Confidential Business Information submitted to the Office of Pesticide
Programs
-------
August, 1963
BENTAZON
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.
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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Bentazon
August, 13
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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).
C10H12N203S
240.3
Colorless crystalline powder
137 to 139«C
<0.1 x 10~7 mm Hg (20°C)
500 mg/L
Properties (Worthing, 1983)
Chemical Formula
Molecular Height
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Specific Gravity
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Bentazon was not found in sampling performed at two water supply
stations in the STORET database (STORET, 1988). No information
on the occurrence of bentazon was found in the available literature.
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Bentazon August,
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Environmental Fate
0 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).
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 days' incubation in the dark at 22°C (Drescher, 1973b)
-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). ^C-Bentazon at 1<7 PP™ 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.
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, (14c)bentazon was very mobile in
12 soils, ranging in texture from sand to silty clay loam, with an
Rf value of 1.0.
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.
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).
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, HN, 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, 138S
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III. PHARMACOKINETICS
Absorption
0 Male and female rats (200 to 250 g) given 0.8 mg ^C-bentazon ln 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 administe-
dose (Chasseaud et al./ 1972).
IV. HEALTH EFFECTS
Humans
0 No information on the health effects of bentazon in humans was found
in the available literature.
Animals
H^^HI^B^
Short-term Exposure
0 The oral LDjg of bentazon in the rat was reported to be 2/063 mg/kg
(Meister, 1986).
• LDgg 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).
Dermal/Ocular Effects
• Previously available information was invalidated. Therefore, it was
not possible to utilize that information for the development of the
Health Advisory.
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Bentazon August, 1983
-5-
Lonq-term Exposure
0 Leuschner et al. (1970) reported the effects of bentazon in beagle
dogs. Beagle dogs (three dogs/sex/dose} were given 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).
Reproductive Effects
0 Previously available information was invalidated. Therefore, it was
not possible to utilize that information for the development of the
Health Advisory.
Developmental Effects
0 Previously available information was invalidated. Therefore, it was
not possible to utilize that information for the development of the
Health Advisory.
Mutaqenicity
0 Previously available information was invalidated. Therefore, it was
not possible to utilize that information for the development of the
Health Advisory.
Carcinoqenicity
0 Previously available information was invalidated. Therefore, it was
not possible to utilize that information for the development of the
Health Advisory.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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)
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
. an adult (70 kg).
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:•=--.'.? cr. Auoust, 1038
-6-
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Ifealth Advisory
No data were found in the available literature that were suitable for
determinatia of One-day HA values. It is, therefore, recommended that the
Longer-term HA value for a 10-kg child (0.3 mq/L) 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 (0.3 mq/L) be used at this time as a
conservative estimate of the Ten-day HA.
Longer-term Health Advisory
A 13-weak 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 gi\en 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 meks. At a dose level of
3,000 jpn, overt signs of toxicity, including weight loss and ib health, vere
observed; 1/3 males and 2/3 females died. At 3,000 ppm, all males shoved sign
of o rostatitis. 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:
Longs r-term HA = (2.5 mg/kq/day) (10 kg) = 0.25 mq/L (300 ug/L)
(100) (1 L/day)
whe re :
2.5 mg/kg/day = NOAEL, based on absence of prostatic? ffects in dogj .
10 kg = assumed body weight of IP child.
100 = uncertainty factor, chosen in a
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Bentazon August, 1983
-7-
where:
2.5 mg/kg/day = NOAEL, based on absence of prostatic effects in dogs.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or 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
(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. 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 mq/kq/day) = Q.0025 mg/kg/day
(1,000)
where:
2.5 mg/kg/day = NOAEL, based on the absence of prostatic effects in
dogs.
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Bentazon August, 1988
-8-
1,000 = uncertainty factor, chosen in accordance with EPA
or NAS/OCW guidelines for use with a NOAEL from an
animal study of less-than-lifetime duration.
Step 2: Determination of the Drinking Water Equivalent Level (OWED
DWEL = (0-0025 mg/lcg/day) (70 kg) = 0.0875 mg/L (90 ug/L)
(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 (20 ug/L)
where:
0.0875 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No valid data are available to make a determination of the carcino-
genic potential of bentazon.
0 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 carcinogenic!ty.
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
0 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 chromatography with flame photometric detection. The method
detection limit, for bentazon has not been determined.
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Bentazon August, 1
-9-
VIII. TREATMENT TECHNOLOGIES
0 There is no information available regarding treatment technologies
used to remove bentazon from contaminated water.
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Bentazon August, 1988
-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/ MI.
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-i30propyl-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 Uhiveristy 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-dloxide) 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 2,4-0
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, 1983
-11-
Drescher, 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 AIBA (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-1H-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. Ace. Nbs. 112129, 092225.
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 in 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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May/ 1988).
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Bentazon August, 1983
-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. RfO Work Group.
Worksheet dated April 7.
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.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
-------
August, 1988
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.
For those substances that are knowji 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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Bromacil
August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No: 314-40-9
Structural Formula
5-Bromo-6-methyl-3-(1-methylpropyl)-2,4(1H,3H)-pyrimidinedione
Synonyms
0 Borea; Bromax; Hyvar; Uragan; Urox B; (Meister, 1988).
Uses
0 Herbicide for general weed or brush control in noncrop areas;
particularly useful against perennial grasses (Meister, 1988).
c9H13°2N2Br
261.11
White crystalline solid
158-160«C
8 x 10-4 mm Hg
815 mg/L
Properties (Windholz et al., 1983)
Chemical Formula
Molecular Weight
Physical State (at 25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (100°C)
Specific Gravity
Water Solubility (20°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Bromacil has been found in Florida ground water; a typical positive
was 300 ppb (Cohen et al., 1986).
Bromacil has not been found in any of 3 surface water samples collected
at 2 locations or in any of 841 ground water samples collected at 834
locations (STORET, 1988). This information is provided to give a
general impression of the occurrence of this chemical in ground and
surface waters as reported in the STORET database. The individual
data points retrieved were used as they came from STORET and have not
been confirmed as to their validity. STORET data is often not valid
when individual numbers are used out of the context of the entire
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Broraacil August, 1988
-3-
sampling regime/ as they are here. Therefore, this information can
only be used to form an impression of the intensity and location of
sampling for a particular chemical*
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
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 Imperfect!/ 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 C02 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 CC>2 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 chroraatographic techniques 1*O
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) £10 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 i.2.6 Ib ai/A
(active ingredient/acre) (Bunker et al., 1971; Stecko, 1971).
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Bromacil August/ 1988
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III. PHARMACOKINETICS
Absorption
0 Workers who were exposed to bromacil during production, formulation
and packaging excreted unchanged bromacil and the S-bromo-3-8ec-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.
Distribution
9 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 5-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-raethylpropyl)-6-methyluracil;
5-bromo-3-(2-hydroxy-1-methylpropyl)-6-hydroxymethyluracilf 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-hyaroxyraethyluracil, 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).
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Bromacil August, 1988
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IV. HEALTH EFFECTS
Humans
0 No information was located in the available literature on the health
effects of bromacil in humans.
Animals
0 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
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 one male mongrel dog, a single oral dose in capsules 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 in another dog at doses of 100 or 250 mg/kg given 5 days
apart.
0 Snerman 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 (DuPont, 1962).
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Bromacil August/ 1988
-6-
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).
* 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.
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, 50G or 2,500 ppm bromacil for 90 days. This
corresponds to doses of about 0, 2.5, 25 or 125 rag/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
rag/kg/day was identified as the No-Observed-Adverse-Effect Level
(NOAEL) in this study.
• 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
mgAg/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.
0 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.
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Bromacil August, 1988
-7-
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
mgAg/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 5,000f>pm
females; an increased incidence of centrilobular vacuolation in 250-ppm
males; an increased incidence of scattered hepatocellular necrosis in
5,000-ppm males; and the presence of extravasated erythrocytes in
the hypertrophied hepatocytes of the 1,25CT- and 5,000-ppm 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 testicular 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 FH, 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
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Bromacil August, 1988
-8-
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.
0 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
• In a sex-linked recessive lethal test (Valencia, 1981), Drosophila
melanogaster (Canton-S wild-type stock) were exposed to bromacil in
food at levels of 2, 3, 5 or 2,000 ppm. Bromacil was found to be
weakly mutagenic at the 2,000-ppra 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 Saccharomyees 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 04.
0 Simmon et al. (1977) reported that bromacil was not mutagenic in an
iH 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 £. eerevisiae;
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.
• 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 mg/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 13. melanogaster
(Gopalan and Njage, 1981).
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Bromacil August, 1988
-9-
Carcinogenicity
0 Sherman et al. (1966) fed groups 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.
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 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 (up to 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, 1,000 or 10,000),
in accordance with EPA or 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 acute dosing study in dogs reported by Sherman and Kaplan
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Bromacil August/ 1988
-10-
(1975) is not considered because the bolus treatment with 100 mgAg by capsule
could be an inaccurate indication of the ability of this species to tolerate
more divided doses as in diet or drinking water, as suggested by the 31.2
mgAg/day NOAEL in the 2-year study in dogs by Sherman et al. (1966). Ten-day
HA, derived below, of 5 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 mgAg/day) produced mild pathological changes in the liver, and this
value was identified as a LOAEL. Although a lower oral LOAEL of 100 mgAg
was evident in sheep (Palmer, 1964), this study was not selected for HA
development because of uncertainty in using ruminants in estimating risk to
humans from oral exposure.
Using a LOAEL of 650 mgAg/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 (5,000 ug/L)
(1,000) (1 L/day)
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 EPA or
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 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 (3,000 ug/L)
(100) (1 L/day)
where:
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Bromacil August, 1988
-11-
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 EPA or
NAS/OEW 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) = g.g mg/L (9,000 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.
100 = uncertainty factor, chosen in accordance with EPA or
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. 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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Broraacil August, 1988
-12-
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 12.5 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 12.5 mg/kg/day, the Lifetime HA is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD - (12.5 mg/kg/day) = 0.125 m
g/kg/day
d to 0.1
(100) (rounded to 0.13 mg/kg/day)
where:
12.5 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 EPA or
NAS/OCW guidelines for use with a NOAEL from an
animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-13 mg/kg/day) (70 kg) = 4.55 mg/L (5,000 ug/L)
(2 L/day)
where:
0.13 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 = (4.6 mg/L) (20%) = 0.09 mg/L (90 ug/L)
10
where:
4.6 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.
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Bromacil August, 1988
-13-
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 but not female CD-I 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 EFA's 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 0.088 mg/L was calculated based on an assumed water consumption
of 2 L/day by a 70-kg adult, with 20% contribution from water.
0 The U.S. EPA Office of Pesticide Programs (EPA/OPP) previously
calculated an ADI 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.
9 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, 1988). 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
-------
Bromacil August, 1988
-14-
exchanged to acetone during concentration to a volume of 10 mL or less.
The compounds in the extract are separated by gas chroraatography and
measurement is made with a thermionic bead detector. This method has
been validated in a single laboratory, and estimated detection limits
have been determined for the analytes in the method, including bromacil.
The estimated detection limit is 2.5 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 No information was found in the available literature on treatment
technologies used to remove bromacil from contaminated water.
-------
Bromacil August, 1988
-15-
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
& 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.
-------
Bromacil August, 1988
-16-
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-bronto-3-sec-butyl-6-methy1-
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. Isensee. 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 DMA. Nature 199:
810-811.
Meister, R., ed. 1988. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Moilanen, K.W., and D.G. Crosby. 1974. The photodecomposition of bromacil.
Arch. Environ. Contarn. 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. Hutagenicity testing of some
selected food preservatives, herbicides and insecticides: II Ames Test.
Bangladesh J. Bot. 9(2):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 Hay 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.
-------
Bromacil August, 1988
-17-
Riccio, E., G. Shepherd, A. Pomeroy, K. Mortelmans and M.D. Haters.* 1981.
Comparative studies between the £. cerevisiae 03 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.
I(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.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.
_In_ 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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. U.S. EPA method 507
- Determination of nitrogen- and phosphorus-containing pesticides in
water by gas chromatography with a nitrogen-phosphorus detector (GC/NPD),
April 15, 1988 draft. Available from U.S. EPA's Environmental Monitoring
and Support Laboratory, Cincinnati, OH.
-------
Broraacil August, 1988
-18-
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. MRIO 00143567.
Volk, V.V. 1972. Physico-chemical relationships of soil-pesticide interactions.
In Progress Report, Oregon State Unive'rsity 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.
Wolf, D.C. 1974. Degradation of bromacil, terbacil, 2,4-D and atrazine in
soil and pure culture and their effect on microbial activity* Oiss.
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, O.C., O.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 tnazine and uracil herbicides in soil. Weed Res.
10:18-26.
'Confidential Business Information submitted to the Office of Pesticide
Programs.
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February, 1989
BUTYLATE
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.
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.
-------
Butylate
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 2008-41-5
Structural Formula
February, 1989
-2-
Carbamothioic acid, bis(2-methylpropyl)-, S-ethyl ester
Synonyms
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®+;
Anelda Plus; Genate Plus (Meister, 1988).
Uses
0 Selective preplant herbicide (Meister, 1988).
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)
Log Cctanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
CnH23NOS
217.41
Clear liquid, aromatic odor
138°C
0.9417
1.3 x ID'3 mm Hg
45 mg/L
Occurrence
Butylate has been found in 91 of 836 surface water samples
analyzed and in 2 of 152 ground water samples (STORET, 1988).
Samples were collected at 80 surface water locations and 61 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 138,000 ug/L in ground water sources. The maximum concentration
found was 4.70 ug/L in surface water and 138,000 ug/L in ground
water. This information is provided to give a general impression
of the occurrence of this chemical in ground and surface waters as
-------
Butylate February, 1989
-3-
reported in the STORET database. The individual data points retrieved
were used as they came from STORET and have not been confirmed as to
their validity. STORET data is often not valid when individual
numbers are used out of the context of the entire sampling regime, as
they are here. Therefore, this information can only be used to form
an impression of the intensity and location of sampling for a particular
chemical.
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,
dusobutylformamide, 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 Butylate is slightly mobile to highly mobil-? .n soils ranging in
texture from silty clay loam to gravelly sanj (Gray and Weierich,
1966; Lavy, 1974; Thomas and Holt, 1979; Weidner, 1974).
0 Butylate is fairly volatile; 45 to 50% of He- 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).
0 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).
0 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 PP" £or 28 daYs- Tne 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
0 Data relating specifically to the absorption of butylate were not
located in the available literature; however, some information was
-------
Butylate February, 1989
-4-
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 *
in the low- and high-dose groups, respectively. These results indicate
that butylate is appreciably absorbed from the gastrointestinal Tact
of rats.
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 I400-labeled preparations of butylate (12.3 or 156
mg/kg). Degradation of administered butylate metabolites was also
assessed. Approximately 40% of the administered l^co-butylate was
metabolized by ester cleavage and ^CC^ 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 l^OD-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.
0 S-(l-14c)ethyl-Sutan®, orally administered at about 110 mg Sutan®/kg,
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 l^ohippuric acid, ethyl methyl sulfoxide and
ethyl methyl sulfone was evident.
Excretion
0 Hubbell and Casida (1977) administered 12.3 or 156 mg/kg of
labeled butylate by gavage to adult male Sprague-Dawley rats. Within
24 hours, 60.9 and 64.0% of the administered radioactivity were
expired as C02, 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.
-------
Butylate February, 1989
-5-
A study by Bova et al. (1978) indicates that biotransformation of
S-(l-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 ^CO, as the major product of metabolism,
accounting for 69% of the total administered dose. This rapid pro-
duction of *4C02 may account for the relatively high levels (7.8%) of
14C found in the tissues after 8 days. Urine and feces accounted for
13.9 and 3.2% of the ^C dose, respectively.
Data obtained from a 3-day balance and tissue residue study by Thomas
et al. (1979) show that (l-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 *4C02 accounted for 93.7, 4.0 and
2.0% of the dose, respectively.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of butylate in humans.
Animals
Short-term Exposure
0 The acute oral LD^Q 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 (Wbodard Research Corp., 1967a).
0 Application of butylate technical (85.71% pure) to the eyes of rabbits
resulted in irritation and corneal opacity. No corneal opacity was in
eyes washed after treatment (Raltech, 1979).
-------
Butylate February, 1989
-6-
Long-term Exposure
0 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 (Wbodard Research Corp., 1967b).
0 Dietary feeding of Sutan® Technical and Sutan® Analytical (purities
not specified) to rale Sprague-Dawley rats at dose levels as high as
180 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 (Wbodard Research Corp., 1967c). Hence, 45 mg/kg/day is
identified as a NOAEL.
0 Sutan® Technical (purity not given) was given orally by capsule to
male and female beagle dogs (5/dose/sex) at doses of 5, 25, or 100
mg/kg/day for 12 months. Liver:body weight ratios were increased (P
< 0.05) in males given 25 or 100 mg/kg/day. The effect at the mid
dose mainly reflected an 18% decrease in body weight compared to
controls. Decreased body weights, increased liver weights, and liver
lesions (males only) were observed in males and females treated with
the high dose. The NOAEL is 5 mg/kg/day (Biodynamics, Inc., 1987).
0 Sutan9 Technical (98% pure) was fed in the diet to male and female
Sprague-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. 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. The 10 mg/kg/day dose is considered a NOAEL as
explained on page 9 of this document (Hazelton Laboratories, Inc.,
1978).
8 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.
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Butylate February, 1989
-7-
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-I 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 Twenty five male and 25 female CrlCD(SD)BR rats were fed diets
containing 0, 200, 1,000, or 4,000 ppm (0, 10, 50, or 200 mg/kg/day)
Sutan® Technical (98.2% pure) for 63 days before the animals were
mated (PO group). Two matings were done to produce the Fla and Fib
litters of the first generation. Parental rats obtained from the
first generation (PI group) were mated three times to produce the
F2a, F2b, and F2c litters of the second generation. Assessments
included survival, body and organ weights, reproductive success, and
pathology. Significant (P <0.05) effects at 50 mg/kg/day were increased
liverrbody weight ratios in PO females, decreases in body weights of
PI females and food consumption by PO males, decreased body weights
in F2a pups, and decreased brain weights in Fib male weanlings. At
200 mg/kg/day, increased (P <0.05) incidences of dilated renal pelvis
and retinal folds in Fib rats were evident. The NOAEL is 10 mg/kg/day.
Developmental Effects
0 Sutan® 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
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 of gestation or from day 6 until natural delivery
was without observable effect (NOAEL) on dams and fetuses (Wbodard
Research Corp., 1967d).
0 Administration of Sutan® Technical (99% pure) by gavage to pregnant
rabbits at doses of 0, 10, 100, or 500 mg/kg/day on gestation days 6
through 18 resulted in a NOAEL of 500 mg/kg/day for developmental
effects and 100 mg/kg/day for maternal toxicity. Maternal body
weight was decreased (P < 0.05) with 500 mg/kg/day (Stauffer Chemical
Co., 1987).
Mutagenicity
0 Butylate was not mutagenic in Salmonella typhimurium strains TA1535,
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Butylate February, 1989
-8-
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 ppni,
respectively. Hepatocellular carcinomas were found in 2/69, 3/69,
4/69, 3/70 and 2/70 males given'O 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 TOXIOOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) =
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet 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, 1,000 or 10,000),
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 butylate. It is, therefore,
recommended that the Ten-day HA value (2 mg/L, calculated below) be used
at this time as a conservative estimate of the One-day HA value.
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Butylate Febbruary, 1989
-9-
Ten-day Health Advisory
The teratology study in mice by Wbodard 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 Wdodard
study (1967d), the effect levels in this study are uncertain. Furthermore,
the agent was given in the diet in the Nbodard 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 (2,000 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 EPA or
NAS/OCW guidelines for use with a NOAEL from an animal
study.
Longer-term Health Advisory
The 2-generation reproduction study in rats by Stauffer (1986) has been
selected as the basis for the Longer-term Health Advisory.
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). 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
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Butylate February, 1989
-10-
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. Each generation
in the 2-generation study in rats by Stauffer (1986) can, in effect, also be
considered an assessment of subchronic toxicity. Because this study does
provide a NOAEL (10 mg/kg/day) and a LOAEL (50 mg/kg/day), it is preferred
over the studies by Woodard et al. (1967b,c) in which the data suggest that
higher doses in these studies could have been NOAELs approximating the dose
that is a LOAEL in the Stauffer (1986) study.
Using a NOAEL of 10 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
(10 mg/kg/day) (10 kg)
Longer-term HA = (100) (1 L/day) =1.0 mg/L (1,000 ug/L)
where:
10 mg/kg/day = NOAEL, based on absence of treatment-related effects in a
2-generation reproduction study in rats.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA or
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:
(10 mg/kg/day) (70 kg)
Longer-term HA = (100) (2 L/day) =3.5 mg/L (4,000 mg/L)
where:
10 mg/kg/day = NOAEL, based on absence of treatment-related effects in
a 2-generation reproduction study in rats.
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Butylate February, 1989
-11-
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA or
NAS/ODW guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily 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. 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 12-month chronic toxicity study with Sutan® in dogs by Biodynamics,
Inc. (1987) is selected to serve as the basis for the Lifetime HA value for
butylate. Of the available data, this study is considered to represent the
most sensitive endpoint of toxicity in terms of the lowest NOAEL.
The Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose {RfD)
RfD = (5 mg/kg/day) = Q.05 mg/kg/day (50 ug/kg/day)
(100)
where:
5 mg/kg/day = NOAEL, based on the absence of toxic signs in a
12-month study in dogs
100 = uncertainty factor, chosen in accordance with EPA or
NAS/ODW guidelines for use with a NOAEL from an animal
study.*
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Butylate February, 1989
-12-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.05 tng/kg/day) (70 kg) = 1>8 mg/L (1 800 Ug/L)
(2 L/day)
where:
0.05 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 = (1.8 mg/L)(20%) =0.36 mg/L (360 ug/L)
where:
1.8 mg/L = DWEL.
20% = assumed relative source contribution from water.
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
controls in the 2-year study by Biodynamics (1982) was found, concurrent
control incidence (2.89%) was low compared to laboratory historical
control data (average of 7.07%).
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), butylate may be placed in
Group D: not classified. This category is for substances that show
inadequate evidence of carcinogenicity in animals and 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
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Butylate February, 1989
-13-
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
0 Analysis of butylate is by a gas chromatographic (GC) method appli-
cable to the 5termination of certain nitrogen-phosphorus-containing
pesticides in water samples (U.S. EPA, 1988). 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
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. This
method has been validated in a single laboratory end estimated detection
limits have been determined for the analytes, including butylate. The
estimated detection limit is 0.15 ug/L.
VIII. TREATMENT TECHNOLOGIES
° No information was found in the available literature on treatment
technologies capable of effectively removing butylate from contaminated
water.
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Butylate February, 1989
-14-
IX. REFERENCES
BCPC. 1977. British Crop Protection Council. Pesticide Manual, 5th ed.
Nottingham, England: Boots Company, Ltd., p. 593.
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.
Biodynamics, Inc.* 1987. A twelve month oral toxicity study of Sutan Technical
in dogs. Study No. T-12651. Submitted to Stauffer Chemical Co.,
Richmond, CA. Unpublished final report. MRID 40389101.
Bova, D.L., J.R. DeBaun, J.C. Petersen and J.J. Menn. 1978.* Metabolism of
[ethyl-l^c] 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(l):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(2):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. 1988. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Murnik, M.R. 1976. Mutagenicity of widely used herbicides. Genetics. 83:S5
-------
Butylate February, 1989
-15-
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.
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 Sutan: Accumulate
distribution, and elimination of l^C residues. Unpublished study prepares
by Bionomics, Inc., submitted by Stauffer Chemical Company, Richmond, CA.
Stauffer Chemical Company.* l.-'5a. 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.
Stauffer Chemical Company.* 1986. A 2-Generation Reproduction Study in Rats
with Sutan. Study No. T-11940. Unpublished final report by Stauffer
Chemical Company, Richmond, CA. EPA Acession No. 263612-263614.
Stauffer Chemical Company.* 1987. A teratology study in rabbits with Sutan
Technical. Study No. T-12999. Unpublished final report by Stauffer
Chemical Company, Richmond, CA. MRID 40389102.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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. DeBaun.* 1980. Metabolism of
[l-14C-ethyl] Sutan in the rat: Urinary metabolite identification.
Stauffer Chemical Co., Richmond, CA. Unpublished final report. MRID
00043682.
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Butylate February, 1989
-16-
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.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Residue tolerances
for S-ethyl-diisobutyl thiocarbamate. CFR 180.232. July 1. p. 294.
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. 1988. U.S. Environmental Protection Agency. U.S. EPA method 507
- Determination of nitrogen- and phosphorus-containing pesticides in
water by GC/NPD, April 15, 1988 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.
Wbodard 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.
Wbodard 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 .,
Richmond, CA. Unpublished final report. MRID 000129544.
*Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
CARBARYL
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 noncarcinogenlc end points of toxicity.
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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Carbaryl August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 63-25-2
Chemical Structure Q U
II I
0-C-N-CH,
1-Naphthalenol methylcarbamate
Synonyms
0 Arilate; Bercema 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 C^H^C^N
Molecular Weight 201.22
Physical State (25°C) White crystals
Boiling Point
Melting Point 145°C
Density 1.232 (20°C)
Vapor Pressure (25°C) <4 x 10~5 mm Hg
water Solubility (30°C) 120 mg/L
Log Octanol/Water Partition 0.14
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Carbaryl has been found in 58 of 640 surface water samples analyzed
and in none of 1,541 ground water samples (STORET, 1988). Samples
were collected at 185 surface water locations and 1,409 ground water
locations, and Carbaryl was found in 6 states. The 85th percentile
of all nonzero samples was 260 ug/L in surface water and 0 ug/L in
ground water sources. The maximum concentration found was 180,000
ug/L in surface water and 0 ug/L in ground water. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The 'individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
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Carbaryl August, 1988
-3-
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
0 14c-carbaryl (purity unspecified) at 10 ppm was relatively stable
to hydrolysis in buffered solutions at pH 3 and 6. It hydrolyzed at
pH 9 with a half-life of 3 to 5 hours when incubated at 25°C (Khasawinah
and Holsing, 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.
0 14c-Carbaryl (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 rag. This is about 0.9% of the total exposure, and the authors
interpreted this to mean that dermal absorption was not complete.
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Carbaryl August, 1988
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• Feldmann and Maibach (1974) applied 4 ug/cm2 of 14c-labeled 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.
8 Houston et al. (1974) reported that 14c-carbamyl-labeled carbaryl
administered by gavage to male 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.
0 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
hoursp 5.8% of the label was recovered in the feces, indicating that
about 94.2% had been absorbed.
Distribution
0 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, 0.20 and
0.30; 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
8 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 demethylatlon,
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 (naphthol sulfate) was found in the uterine
leiomyoma.
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Carbaryl August, 1988
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Houston et al. (1974) administered 14C-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.
Excretion
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-naphthol 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 carbaryl/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.
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.
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%.
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
vandekar (1965) investigated the effects of large-scale carbaryl
spraying in a village in Nigeria. No 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.
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Carbaryl August, 1988
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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
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-term Exposure
0 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 1,050 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 LD$Q was calculated to be
988 mg/kg. 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, lacrlmation, 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.
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Carbaryl August, 1988
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0 Weil et al. (1968) fed carbaryl in the diet for 1 week to Harlan-
Wistar albino rats (42-days old) at concentrations yielding ingested
doses of Or 10, 50, 250 or 500 mg/kg/day. Body weight gain was
decreased in animals exposed to 50 mg/kg/day or higher. At 10 mg/kg/day,
ChE activity was not significantly affected in plasma, red blood
cells or brain. At 50 mg/kg/day, plasma ChE was decreased 15 and 18%
and red blood cell ChE was decreased 26 and 47% in females and males,
respectively. At higher doses, larger decreases in plasma and red
blood cell ChE were seen, brain ChE was also decreased (26 and 23% at
250 mg/kg/day and 28 and 33% at 500 mg/kg/day in males and females,
respectively). 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/Ocular 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 LO5Q value was greater
than 4,000 mg/kg for both males and females.
• 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 ing/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
mg/kg/day. The author stated that there were no significant changes
in appetite or weight gain when compared to the controls; 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 mg/kg/day (the highest dose tested).
0 Carbaryl was administered to male rats by gavage at a level of
200 mg/kg, 3 days a week for 90 days (Dikshith et al., 1976). This
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Carbaryl August, 1988
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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
changes were seen in testes, liver and brain (enzymatic determinations)
except for ChE activity, which was inhibited 34% in blood (p <0.001)
and 12% in brain (p <0.05). 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
Op 24, 95 or 414 ppm, which were adjusted to supply ingested doses of
0, 0.45, 1.8 or 7.2 mg/kg/day. Mo 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.
0 Schering (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, hematology
or histologic appearance of tissues. The authors concluded that
5.0 mg/kg/day was a NOAEL in rats.
e 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
mg/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. (1972a) investigated the reproductive effects of carbaryl
in female rats exposed either by gavage or by feeding. Doses of 0,
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Carbaryl August, 1988
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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
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
mg/kg/day from the diet.
0 Murray et al. (1979) studied the reproductive effects of carbaryl
(99% active ingredient) in female CF-1 mice. Carbaryl was admini-
stered 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
mg/kg/day). At the gavage dose of 150 mg/kg/day, the mice gained
less weight and exhibited significant maternal toxicity, including
salivation, ataxia and lethargy, and 10/37 females died during the
experimental period. At 100 mg/kg/day by gavage, a single maternal
death occurred, but weight gain was normal and no other evidence of
maternal toxicity was observed. No maternal deaths or signs of ChE
inhibition were seen among the mice supplied carbaryl in the diet
(1,166 mg/kg/day), although there was a significant (p <0.05) decrease
in body weight gain on days 10 through 15. The incidence of pregnancy,
the average number of live fetuses/litter and the incidence of resorp-
tions were not altered by carbaryl for either route of administration.
Mean fetal body weight and length were significantly (p <0.05) lower
than control values among litters given carbaryl in the diet, but was
not affected among those given carbaryl by gavage. Based on maternal
reproductive effects, the NOAEL in mice was identified as 100 mg/kg/day.
0 In an investigation using Sprague-Dawley rats, carbaryl was admini-
stered by gavage at levels of 1, 10 or 100 mg/kg/day for 3 months
prior to and throughout gestation (Lechner and Abdel-Rahman, 1984).
Carbaryl of formulation grade (purity not specified) was administered
in corn oil. Dams were sacrificed on day 20 for examination. Animals
receiving 100 mg/kg/day showed a significant decrease in weight gain
during the gestational period, occurring primarily in the third week
(days 15 to 20). There was also a slight decrease in the number of
implantation sites and live fetuses per dam after treatment at this
dose level. Fetal weights and body length for all three doses were
within the range of control values. There were no overt signs of
maternal toxicity that suggested ChE inhibition. Based on maternal
weight gain and number of implantations, the NOAEL in this study was
identified as 10 mg/kg/day.
0 Collins et al. (1971) reported 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 0F 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-ppm level. Dose-related decreases were
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Carbaryl August, 1988
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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. (1971) reported the effects of carbaryl in a three-
generation study in gerbils. Carbaryl was fed at dose levels of
Op 2,000, 4,000, 6,000 or 10,000 ppm. Assuming that 1 ppm in the
diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), and
that gerbils are similar to rats in terms of body weight and food
consumption, 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. (1971) 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.
0 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. No 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 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/kg
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
mgAg/day) was identified.
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Carbaryl August, 1988
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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 mg/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-Oawley
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) on day 6 through termination of 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 DMA 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 DMA synthesis, but had no effect on
unscheduled DMA synthesis.
Carcinogenicity
0 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 and 12, occurring in 9, 11, 7, 6 and 11 rats,
respectively.
0 Schering (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.
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Carbaryl August, 1988
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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.
Statistical analyses of the results were not presented.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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) = m /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, 1,000 or 10,000),
in accordance with EPA or 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) = Ie0 mg/L (1,000 ug/L)
(100)(1 L/day)
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Carbaryl August, 1988
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where:
10 rag/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 EPA
or 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
(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. 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 NOAEL 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.
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Carbaryl August, 1988
-14-
Using the NOAEL of 9.6 mg/kg/day, the Lifetime HA for carbaryl is
calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = 0.6 mg/kg/day) = Qf, mg/kg/day
(ioo) y
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 EPA
or 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 (4rooo 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 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.
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Carbaryl August, 1988
-15-
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 (NAS) 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-fcg 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, 1988). In this
method the water sample is filtered and a 400 uL aliquot is injected
into a reverse phase HPLC column. Separation of compounds is achieved
using gradient elution chromatography. After elution from the HPLC
column, the compounds are hydrolyzed with sodium hydroxide. The
methyl amine formed during hydrolysis is reacted with o-phthalaldehyde
(OPA) to form a fluorescent derivative which is detected using a
fluorescence detector. The method detection limit has been estimated
to be 2.0 ug/L for carbaryl.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) adsorption,
ozonation and conventional treatment will 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.
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Carbaryl August, 1988
-16-
Ozonation has been 99% effective in removing carbaryl and its
hydrolysis product, naphthol, from aqueous solution (Shevchenko et al.,
1982). Carbaryl and naphthol 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.
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
wastewater.
A 3-day settling period without any chemical treatment yield a 50%
carbaryl concentration reduction (Holiday and Hardin, 1981).
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 Augustf 1988
-17-
IX. REFERENCES
Andrianova, M.M. and I.V. Alekseev.* 1969. Carcinogenic properties of
Sevinr Maneb, Ciram and Cineb. Vopr. Pitan. 29:71-74. Unpublished
report. MRIO 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. Woodsider J.H. Nair and H.F. Smyth.
1961. Mammalian toxicity of 1-napthyl-N-methylcarbamate (Sevin insecticide)
J. Agr. Food Chem. 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.
Toxicol. 7(1):37-56.
Collins, T.F.X., W.H. Hansen and M.V. Keeler. 1971. The effect of carbaryl
on reproduction of the rat and the gerbil. Toxicol. Appl. Pharmacol.
19:202-216.
Comer, S.W., D.C. Staiff, J.F. Armstrong and H.R. Wolfe. 1975. Exposure of
workers to carbaryl. Bull. Environ. Contain. Toxicol. 13(4):385-391.
Dikshithr 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.
Feldmann, 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.
-------
Carbaryl August, 1988
-18-
Houston, J.B., D.G. Upshall and J.W. Bridges. 1974. Pharmacokinetics and
metabolism of two carbamate insecticides, carbaryl and landrin, in the
rat. Xenobiotica. 5(10):637-648.
IARC. 1976. International Agency for Research on Cancer. IARC monographs
on the evaluation of carcinogenic risk of chemicals to man. Lyon, France:
IARC. 12:37-48.
Khasawinah, 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.
Khasawinah, 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.
Khasawinah, 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-carbamateH214 insecticide chemicals in rats. Unpublished
report. MRID 00049134.
Lechner, O.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 gavage or by dietary inclusion.
Toxicol. Appl. Pharmacol. 51(1):81-89.
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
Rittenhouse, J.R., J.K. Narcisse and R.D. Cavalli.* 1972. Acute oral toxicity
to rats of Orthene in combination with five other cholinesterase-inhibiting
materials. Unpublished report. MRID 00014933.
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, 1988
-19-
Schering, A.G.* 1963. Promecarb (SN 34615): long-term feeding study in rats:
ZK No. 3858. Unpublished report. MRID 00081723.
Shevchenko, M.A.p P.N. Taran and P.V. Marchenko. 1982. Modern methods for
purifying water from pesticides. Soviet Journal of Hater Chemistry and
Technology. 4(4):53-71.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 531.1 -
Measurement of N-methyl carbamoyloximes and N-methylcarbamates in drinking
water by direct aqueous injection HPLC with post column derivatization.
April 15, 1988. Environmental Monitoring and Support Laboratory,
Cincinnati, OH.
Vandekar, M. 1965. Observations of the toxicity of carbaryl, folithion and
3-isopropylphenyl-N-methylcarbamate 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 00118383.
Weil, C.S., M.W. Woodside, C.P. Carpenter and H.F. Smyth. 1971. Current
status of tests of carbaryl for reproductive and teratogenic effects.
Unpublished report. MRID 0008064.
Weil, C.S. 1972a. Comparative study of dietary inclusion versus stomach
intubation on three generations on reproduction, on teratology and on
mutagenesis. Unpublished report. MRID 00125161.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flow column systems. PhD. Thesis, Purdue
University.
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Carbaryl August, 1988
-20-
Whittaker, K. F., J.C. Nye, R.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.
Windholz, 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. MRIO 00086672.
'Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
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.
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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Carboxln August, 1988
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 5234-68-4
Structural Formula
0 H
5,6-Dihydro-2-methyl-N-phenyl-1,4-Oxathiin-3-carboxamide
Synonyms
• Carbathlin; Carboxine; D-715« DCMO; CHOC; F735; Vltavax (Meister,
1983).
Uses
• Systemic fungicide; seed protectant; wood preservative (Meister,
1983).
Properties (Meister, 1983; Windholz et al., 1983; Wo and Shapiro, 1983;
Worthing, 1983; TDB, 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
• Carboxin is rapidly metabolized (oxidized by flavin enzymes found in
fungi mitochondria) in aerobic soil. When applied to soil (aerobic
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Carbox!n August, 1988
-3-
conditions), more than 95$ of Ut« cw.-_-xin h&d degraded within 7
days. The major degradation product tus carboxin sulf oxide, which
represented 31 to 54% of the applied radioactivity at 7 days after
treatment. Several minor degradation products were also formed
(carboxin sulf one, p_-hydroxy carboxin and 14CC>2). 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 sulf oxide is stable in anaerobic soil
(Chin et al., 1972, 1959, 1970s, b; Ozialo and Lacadie, 1978; Dzialo
et al., 1978; Spare, 1979).
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).
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).
In aqueous solution, carboxin was oxidized to carboxin sulfoxide and
carboxin sulfone within 7 days (Chin et al. , 1970a).
III. PHARMACOKINETICS
Absorption
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
Waring (1973) administered single oral doses of carboxin (Vitavax,
6.3 uCi/rat) to fer.j»!e 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.
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
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Carboxin August, 1988
-4-
ppo. 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, b, 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
• 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 previo,:?^, ?e..iale 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 of the ingested dose was provided (PIMS, 1980).
Animals
Short-term Exposure
0 Reagan and Becci ( 1 983) 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 LD5g was reported to be 3,550 mg/kg.
RTECS (1985) reported that the acute oral LDsg for carboxin (purity
not specified) in the rat (age not specified) was 430 mg/kg.
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Carboxin August, 1988
-5-
• 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 n---ogen and
decreased hemoglobin at the 12-week interval in females thu 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-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 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, 1988
-6-
effccts on body weight gain, food consumption, or various hematologicaj.
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 non-neoplastic 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 mgAg/day for
females) based on hepatic effects.
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.
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 mgAg/day) was identified.
0 Holsing (1969b) administered carboxin (technical 0-735; 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
approxim? to 0, 2.5, 5.0 or 15.0 mg/kg/day. No treatment-related
effects * aborted on survival, body weight gain, food consumption,
organ weig..;3, 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.
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Carboxin August, 1988
-7-
a NOAEL of 600 ppm (15 mg/kg/dsy; the higLnst close 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 gp.vr.g;. at doser 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-Dawley rats (20/dose)
on days 6 through 15 of gestation. No compound-related effects were
observed on reproduction, gestation or in skeletal or soft tissue
development. Based on the information presented, a NOAEL of 40
mgAg/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 04. 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.
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Carboxin August, 1988
-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 £• typhimurium strains TA 1535,
1538, 98 and 100.
9 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 0-735; 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 parench/mal cells, were
observed in mice in the 2,500 or 5,000 ppm dose groups (385 and 751
mgAg/day for males; 451 and 912 mg/kg/day for females). In males, the
incidence of pulmo.i^ry 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 Kruskall 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, 1908
-9-
16.7%; in seven 20- to 22-month studies, the incidence of lunc;
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 (up to 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-Ef f ect 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, 1,000 or 10,000),
in accordance with EPA or NAS/OEW
_ 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, 1988
-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. (1962) 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 Cor the 10 -kg child is calculated as follows:
Longer-term HA = (1° "g/kg/day) (10 kg) = K0 ng/L (1,000 ug/L,
(100) (1 L/day)
where:
10 mg/kg/d&y = NOAEL, based on absence of effects on growth, hematology,
blood chemistry, urinalysts, gross pathology and
histopathology in rats exposed to carboxin in the diet
for 90 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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° mg/kg/day) (7° kg) =3.5 mg/L (4,000 ug/L)
(100) (2 L/day)
where:
10 mg/kg/day = NOAEL, based on absence of effects on growt. 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 EPA
or 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
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Carboxin August, 1988
-11-
(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. 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 cirboxin 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) = 0.
(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 s 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 mg/kg/day) (70 kg) =3.5 mg/L (4,000 ug/L)
(2 L/day)
-------
Carhoxln August, 1988
-12-
where?
0.1 mgAg/day = Rfl>«
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination oL 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) repo.t_d 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 HolsJng (1969a) fed Charles River rats carboxin at dietary levels up
to 30 ing/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, 1988
-13-
0 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
" 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, 1988
-14-
IX. REFERENCES
Brusick, D.J., and R.J. Weir.* 1977. Mutagenieity evaluation of 0-735.
CBI 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. Fiial Report. Unpublished study. MRID 00132453.
Byeon, W., H.H. Hyun and S.Y. Lee.* 1978. Mutagenicity of pesticides in the
Salmonella/microsomaJ. 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. 1c*f 4,: 73". -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
stvUes 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, 1988
-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. NRID 00003030.
Jessup, D., G. Gunderson and R. Gail.* 1982. Lifetime carcinogenicity study
in mice {Vitavax): 399-OO^a. Unpublished study. MRID 00114139.
Knickerbocker/ M.* 1977. Teratologic evaluation of Vitavax technical in
Sprague-Oawley rats. Unpublished study. MRID 00003102.
Lacadie, J.A., D.R. Gerecke and R.A. Cardona.* 1978. Vitavax 14C 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 wit 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.
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Carboxln August, 1988
-16-
Smllo, A.R., J.A. Lacadle and B. C&rdona.* 19V/. Photochemical fate of
Vitavax in solution. (Unpublished study submitted by Uniroyal Chemical,
Bethany, Conn. CDL:231932-C.) MRID 00003008.
Spare, W.* 1979. Report: Vitavax microbia.1 metabolism in soil and its effect
on microbes. (Unpublished study prepared by Biospherics., Inc., in
cooperation with United States Testing Co., Inc., submitted by Uhiroyal
Chemical, Bethany, Conn. CDL:098029-A.) NRID 00005540.
TDB. 1985. Toxicology Date. Benk. 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, 1988
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*
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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Chloramben
August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No.; 133-90-4
Structural Formula:
COOH
3-Amino-2-5-dichlorobenzoic acid
Synonyma
• Acp-m-729, Ambiben; Abiben; Amibin; Amoben; Chlorambed; Chlorambene;
NCI-C00055 ornamental weeder; Ornamental weeder; Vegaben; Vegiven
(U.S. EPA, 1985).
Uses
• Preemergent 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
0 Chloramben has been found in 13 of 34 surface water samples anaylzed
and in 1 of 566 ground water samples (STORET, 1988). Samples were
collected at 11 surface water locations and 322 ground water locations,
and Chloramben was found in only 1 state. The 85th percentile of all
non-zero samples was 2.1 ug/L in surface water and 1.7 ug/L in ground
water sources. This information is provided to give a general
impression of the occurrence of this chemical in ground and surface
waters as reported in the STORET database. The individual data
points retrieved were used as they came from STORET and have not been
206.02
Crystals
200-201»C
7 x 10-3 mm Hg (100«C)
700 mg/L
2.32
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Chloramben August, 1988
-3-
confirmed as to their validity. STORET data is often not valid when
individual numbers are used out of the context of the entire sampling
regime, as they are here. Therefore, this information can only be
used to form an impression of the intensity and location of sampling
for a particular chemical.
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 (Registration Standard
Science Chapter for Chloramben).
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
about 30% in 48 hours (Registration Standard Science Chapter for
Chloramben).
• 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 (Registration Standard Science Chapter for
Chloramben).
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 (Registration standard Science Chapter for Chloramben).
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,
or a detoxifying conjugation which goes slowly, if at all (Registration
Standard Science Chapter for Chloramben).
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 (Registration Standard Science
Chapter for Chloramben).
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
(Registration Standard Science Chapter for Chloramben).
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Chloramben August, 1988
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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.
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 LD5o values for Chloramben range from 2,101 mg/kg (Field,
1980) to 5,000 mg/kg (Field and Carter, 1978) in rats; the acute
dermal LDso in rabbits has been reported to be >2,000 (Field and Field,
1980) or >5,000 mg/kg (Field, 1978).
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Chloramben August, 1988
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9 Rees and Re (1978) reported an acute (1 hr) LCsg of >200 mg/L in rat
inhalation studies.
0 Keller (1959) fed male Holtzraan 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
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 mgAg/day),
the highest dose tested, was identified.
Dermal/Ocular Effects
0 Gabriel (1969) applied chloramben (4 to 8 gAg) 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 Bellies (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 and 1,070 mgAg/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-1 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
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Chloramben August, 1988
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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).
0 NCI (1977) administered technical-grade chloramben (90 to 95% active
ingredient) to Osborne-Mendel rats (50/sex/dose) and B6C3F-) 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 ppra. 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 Johnson 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
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Chloramben August, 1988
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consumption, hematology, clinical chemistry, urinalysisr 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
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 mg/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).
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Chloramben August, 1988
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0 Results were positive for the in vitro cytogenic test using Chinese
hamster ovary cells (Galloway and Lebowitz, 1982).
Carcinogenicity
• 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
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. 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.
8 NCI (1977) administered 10,000 or 20,000 ppm technical chloramben
(90-95% 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.
• Johnston and Seibold (1979) reported no evidence of carcinogenic
activity in Sprague-Dawley 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.
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Chloramben August, 1988
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V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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)
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 (3 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 Bellies 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 mg/L (3,000 ug/L)
(100) (1 L/day)
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Chloramben August, 1988
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where:
25 ing/kg/day = NOAEL, based on the absence of systemic toxic effects
in rats fed chloramben for 10 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 Advisories
No data were found in the available literature that were suitable for
the determination of the Longer-terra HA. It is, therefore, recommended that
an adjusted DWEL for a 10 kg child (0.15 mg/L = 200 ug/L) and the DWEL for a
a 70-kg adult (0.525 mg/L = 500 ug/L) be used at this time for the Longer-
term HA values.
Adjusted DWEL = (0.015 mq/kg/day) (10 kg) = 0.15 mg/L (200 ug/L)
(1 L/day)
where:
0.015 mg/kg/day = RfD.
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of 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 Mater 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. If the contaminant is classified as a Group A or B
carcinogen, according to the Agency's classification scheme of carcinogenic
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Chloramben August, 1988
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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, CrliCOBS CD-I 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:
Step 1: Determination of the Reference Dose (RfD)
RfD = (15 mg/kg/day) = Ot015 mg/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 EPA
or 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) = Ot525 m /L (500 /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 rag/L) (20%) = 0.105 mg/L (100 ug/L)
where:
0.525 mg/L = DWEL.
20% = assumed relative source contribution from water.
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Chloramben August, 1988
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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 B6C3F-| mice
(20/gex/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. Johnson 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-1 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
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.
0 Applying the criteria described in EPA'3 guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), chloramben may be classified in
Group 0: not classified. This category is for agents with inadequate
human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 MAS has determined an Acceptable Daily Intake of 0.25 mg/kg/day with
a suggested-no-adverse-effect level of 1.75mg/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 (40 CFR 180.226).
VII. ANALYTICAL METHODS
0 Chloramben may be analyzed using a gas chromatographic (GC) method
applicable to the determination of chlorinated acid, ethers and
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 (EC) gas chromatography. The method detection limit
has not been determined for this compound.
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Chloramben August, 1988
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VIII. 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.
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 10-3 mm Hg at 100°C). Due to its estimated Henry's Coefficient
of 0.15 atm, chloramben probably would not be amenable to aeration or
air stripping.
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Chloramben August, 1988
-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.
Andrawes, N.* 1984. Azniben: Metabolism of 14c -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. LSI Project No. 2595. Final Report. Unpublished study.
MRID 00131187.
Beliles, R.P., and S. Mueller.* 1976. Teratology study in rats: technical
chloramben. LSI Project No. 2577. Final Report. Unpublished study.
MRID 0096618.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.226. July 1, 1985.
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.
Drill, V., S. Friess, H. Hayes et al. (names not specified). * 1982. Retro-
spective audit of the bioassay of chloramben for possible carcinogenicity.
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.* 1978. Oral LDso in rats. Study No. CDC -AM-0 15-78.
MRID 00100318.
Field, W.E.* 1978. Acute dermal application (LDso) — rabbit. Study No.
CDC -AM-0 12-78. Unpublished study. MRID 00100319.
Field, W.* 1980. Oral LD50 in rats: chloramben 10G. Study No. CDC-UC-158.
MRID 00128640.
Field, W., and G. Field.* 1980. 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. — Amiben (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.
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Chloramben August, 1988
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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. Zimraer-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 DMA 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.
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Chloraraben August, 1988
-16-
Rees, D.C. and T.A. Re.* 1978. Inhalation toxicity of amiben sodium salt
3599 in adult Sprague-Oawley rats. Laboratory No. 5764b. unpublished
study. MRID 00100322.
STORET. 1988. STORET Mater Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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, D.C. 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.
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 —
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 45268.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August/ 1988
CHLOROTHALONIL
Health Advisory
Offics of Cri.iking Wetar
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.
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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Chlorothalonil August/ 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1897-45-6
Structural Formula
2,4,5,6-Tetrachloro-1,3-benzenedicarbonitrile
Synonyms
0 Tetrachloroisophthalonitrile; Bravo; Chloroalonil; Chlorthalonil;
Daconil; Exothern; Forturf; Nopcocide N96; Sweep; Termil; TPN; DAC-2787;
T-117-11; DTX-77-0033; DTX-77-0034; DTX-77-0035.
Uses (Meister, 1986)
0 Broad-spectrum fungicide.
Properties (Meister, 1986; CHEMLAB, 1985; Meister, 1983; Windholz et al., 1983)
Chemical Formula CgN2Cl4
Molecular Weight 265.89
Physical State (25°C) White, crystalline solid
Boiling Point 350«C
Melting Point 250 to 251 «C
Density
Vapor Pressure (40°C) 2.0 x 10-6 mm Hg at 25°C
Specific Gravity
Water Solubility (25°C) 1.2 mg/L at 25°C
Octanol/Water Partition 7.62 x 102
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Chlorothalonil has been found in 1 of 4 surface water samples analyzed
from 3 surface water stations and in none of the 633 ground water
samples (STORET, 1988). Samples were collected at 3 surface water
locations and 627 ground water locations; and at the 1 location where
it was found in Missouri, the concentration was 6,500 ug/L. This
information is provided to give a general impression of the occurrence
of this chemical in ground and surface waters as reported in the
STORET database. The individual data points retrieved were used as
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Chlorothalonil August/ 1988
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they came from 3TORET and have not been confirmed as to their
validity. STORET data is often not valid when individual numbers
are used out of the context of the entire sampling regime/ as they
aie here. Therefore/ this information can only be used to form an
impression of the intensity and location of sampling for a particular
chemical.
Environmental Fate
0 Ring-labeled 14c-chlorothalonil/ at 0.5 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 14c-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 At 1,000 ppm, the chlorothalonil degradation product/ 14C-DAC-3701/
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, undated).
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 95°F and 80% of field moisture
capacity (Szalkowski, 1976a). When chlorothalonil (WP) was applied
to nonsterile soils ranging in texture from sand to si^ty ay loam,
at 76 to 100°F and 6% soil moisture, it was degraded with .alf-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 in^.jences
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_37oi
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, undated). 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
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Chlorothalonil August, 1988
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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 chlcrotiialor.il dsgrad&te DAC-37C1 is mobile in sand, loas, 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
the leachate of the sand, loam, 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 14c-chlorothalonil (dose not specified)
orally to albino rats (3/sex; strain not specified). In 48 hours
post-treatment, 60% of the radioactivity was detected in the feces,
suggesting 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 (containing 2.8 uCi) in a 500 mg/kg dose (route
not specified) eliminated 88% of the administered dose unchanged in
the feces over 264 hours, suggesting 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, suggesting 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
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Chlorothalonil August, 1988
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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 14c-chlorothalonil
by oral intubation to CD-1 mice at levels of 0, 1.5, 15 or 105 mg/kg.
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.
8 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 (196Sb) study, only a small amount of the
metabolite, 4-hydroxy-2,5,6-trichloroisophthalonitrile, 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. (1983b) 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 1 1 days, an average
of 88.45% of the administered dose was excreted in the faces, 5.14% in
the urine and 0.32% in expired gases as
Skinner and Stallard (1967) presented results which demonstrated 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-1 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.
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Chlorothalonil August, 1988
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Marciniszyn et al. (1981) reported a study in which single doses of
Hc-chiorothalonil were administered intraduodenally to male Sprague-
Uawley rats at 0.5, 5, 10, 50, 100 or 200 mg/kg. Biliary excretion
of radioactivity was mcnitr-rsd fcr 24 hours. Percent recovery cf
radioactivity was 27.8, 20.7, 16.8, 6.4, 7.8 and 6% for each dose
level, respectively.
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 neck, wrist and forearms. Results
of a patch test conducted with 0.1% Chlorothalonil in acetone were
positive in 7 of 14 subjects. Reactions ranged from a few erythematous
papules to marked ^/ctpular erythema with a brownish hue without
infiltration.
Animals
Short-term Exposure
0 Powers (1965) reported that the acute oral LDsg of Chlorothalonil
(75% wettable powder) in Sprague-Dawley rats was >10 g/kg.
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Chlorothalonil August/ 1988
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0 Doyle and Elsea (1963) reported that the acute oral LDso of DAC-2787
(chlorothalonil), assumed by authors to be 100%, in male Dublin SD
rats was >10 g/kg when the compound was administered in corn oil.
Dermal/Ocular Effects
0 Doyle and Elsea (1963) reported that the dermal LDso of DAC-2787
(assumed by the authors to be 100%), applied as a paste made of
approximately equal parts of DAC-2787 and corn oil, in albino rabbits
was >10 g/kg. At dermal concentrations of 1, 2.15, 4.64 or 10 g/kg
(24-hour exposure on abdominal skin), 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 (assumed
by authors to be 100%, application method unknown) 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
BRAVO 500, a 40% formulation containing 96% pure cl^orothalonil. In
both species, 0.1 mL of the test substance was insx-llxeC into the
conjunctival sac of one eye. Each species displayed mild and transient
ocular irritation as evidenced by corneal opacities that were reversed
by 4 days postinstillation. 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 DAC 2787 (chloro-
thalonil), assumed by authors to be 100%, to Charles River rats
(50/sex/dose) 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 higuest 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.
8 Wilson et al. (1981) administered chlorothalonil (98%) 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
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Chlorothalonil August, 1988
-8-
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 kid-iey weights were also noted
in males at 40, 80, 175 and 375 rag/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
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 mg/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 chlorothalonil (98.4%) to
Charles River CD-I mice (15/sex/dose) for 90 days at dietary concen-
trations of 0, 7.5, 15, 50, 275 or 750 ppm (males: 0, 1.2, 2.5, 8.5,
47.7 or 123.6 mg/kg/day; and females: 0, 1.4, 3.0, 9.8, 51.4 or 141.2
mg/kg/day). No treatment-related effects were noted on survival,
physical condition, body weight, food consumption or gross pathology.
At 750 ppm (141 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 (124 mg/kg/day) and in females
dosed at 275 and 750 ppm (51.0 and 141 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-f 275- ar-i -ppm dose
groups. No other treatment-related histopathological cha.. .5 were
reported. A NOAEL of 15 ppm (2.5 mg/kg/day) is identified for this
study.
Paynter and Murphy (1967) administered OAC 2787 (chlorothalonil) to
beagle dogs (4/sex/dose) for 16 weeks at dietary concentrations of 0,
250r 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 appearancef 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 NOAEL of
750 ppm (18.8 mg/kg/day) is identified.
Hastings et al. (1975) administered 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 Or 0.05, 0.1, 0.2, 0.8, 1.5, 3 or 6 mg/kg/day; Lehman,
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Chlorothalonil Augustf 1988
-9-
1959). No significant differences between treated and control groups
were seen in body weightt food consumptionr mortality or gross patho-
logical changes. Histopathological examination of the kidneys
revealed no demonstrable effects at any dose level. A NOAEL of
120 ppm (6 mg/kg/day) is identified.
0 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
75 mg/kg/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), assumed by authors to be 100%, 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 chlorothalonil,
assumed by authors to be 100%, 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,
compoundrrelated 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 . "^uted and collecting tubules and increased pigment in
the conv_ . .au 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.
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Chlorothalonil Augustf 1988
-10-
0 Tierney et al. (1983) administered technical grade Chlorothalonil
(97.7% pure) 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
intcated that these dietary levels were approximately 0, 119.4.. 251.1 o»
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 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
sguamous 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 Chlorothalonil (98% pure)
by gavage, in aqueous methylcellulose, at doses of 0, 25, 100 or
400 mg/kg/day to Sprague-Dawley Cobs CD rats (25/dose level) on days
6 to 15 of gestation. No co. pound-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 consumption). 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.
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Chlorothalonil August, 1988
-11-
0 Shirasu and Teramoto (1975) administered chlorothalonil (99.3% pure)
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/dayi four of the rin? 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 tera-
togenic effects and a NOAEL of 5 mg/kg/day for maternal toxicity.
Mutagenicity
0 Quinto et al. (1981) reported that chlorothalonil (purity and concen-
trations were not specified) was not mutagenic/ with or without
metabolic activation/ in five tester strains of Salmonella typhimurium.
0 Wei (1982) reported that chlorothalonil (90% pure)/ at concentrations
up to 764 ug/plate/ was not mutagenic in £. 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 (97.8% chlorothalonil)
at concentr--ions 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 £. typhimurium strains TA 1535/ 1537/ 1538/ 98 and 100 and
Escherichia coli WP2 her* and WP2 her'. Chlorothalonil (99.3%) failed
to produce 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 effects of DTX-77-0033 (97.8% chloro-
thalonil) in a DNA repair assay using S_. typhimurium strains TA 1978
and 1538. Chlorothalonil, dissolved in dimethylsulfoxide 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 a commercial preparation of
chlorothalonil (2,500 ppm of active ingredient) 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) di^ not induce mitotic gene conversions.
0 Shirasu et al. (1975) reported that, at concentrations up to 200
ug/disk, chlorothalonil (99.3%) was negative in a rec-assay using
Bacillus subtilis strains H17 and M45.
0 Kouri et al. (1977a) exposed Chinese hamster cells (V-79) and mouse
fibroblast cells (BALB/3T3) in vitro to DTX-77-0034 (chlorothalonil,
97.8% pure) at concentrations 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
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Chlorothalonil August/ 1988
-12-
without metaboljc activation. Chlorothalonil was not mutagenic in
either cell type.
0 Mizens ct al. (19Q3a; reported the results of a micronucleuc test in
Wistar rats, Swiss CFLP mice and Chinese hamsters. Rats were dosed
at 0, 8, 40, 200, 1/000 or 5/000 mg/kg; mice and hamsters received 0,
4, 20, 100, 500 or 2/500 mg/kg. All animals were dosed by gavage and
all received two doses/ 24 hours apart. Technical Chlorothalonil
(98.2% ai) did not induce bone marrow erythrocyte micronuclei in any
of the species tested.
9 Legator (1974) reported the results of an jji vivo cytogenetic test on
DAC-2787 (Chlorothalonil, more than 99% pure) 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 £. typhimurium strains G-46/ TA1530, C-207,
TA1531, C-3076, TA1700, D-3056 and TA1724. Mice (10/dose) received
DAC-2787 (Chlorothalonil, 99% pure) by gavage for 5 days at 6.5
mg/kg/day. The compound did not produce any measurable tiutagenic
response when initially evaluated ir± vitro against Lhe tignt tester
strains of S_. 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 DAC-2787
(Chlorothalonil, 99% pure) 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 signifi-
cant 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 Mizens et al. (19S3b) presented the results of a chromosomal aberration
test in Chinese hamsters. The test animals received two doses of
technical Chlorothalonil/ 24 hours apart, by gavage at concentrations
of 0, 8, 40, 200, 1/000 or 5/000 mg/kg. At 5/000 mg/kg/ a statistically
significant increase in bone marrow chromosomal abnormalities was
observed. However/ the authors concluded that this effect could not
be attributed to Chlorothalonil (98.2% ai) becaus*- the animals exhibited
toxic responses to dosing.
Carcinogenicity
0 NCI (1980) reported the results of a study in which technical-grade
Chlorothalonil (98.5%) 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 doses have been calculated to correspond to approximately
253 and 506 mg/kg/day (Lehman, 1959). Matched controls consisted of
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Chlorothalonil August, 1988
-13-
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 sigr.s 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
(98.5%) 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 mg/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 B6C3F1 mice.
0 Tierney et al. (1983) administered technical-grade chlcro- lonil
(97.7% pure) to Charles River CD-I mice (60/sex/control a.-., 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
forestomach 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 epitheliui.. 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
cortical tubules were noted in all treated groups of male mice.
These changes did not show a dose-response relationship; the increased
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Chlorothalonil August, 1988
-14-
incidence was statistically significant only in the 750 ppm (251
mg/kg/day) group. The authors concluded that the administration of
Chlorothalonil caused an increase in the incidence of primary gastric
tumors and, in rale ra.ce only, caused an increase in the incidence =f
renal tubular neoplasms.
Wilson et al. (1985) gave technical 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 (up to 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 formulas
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 mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 ; a One-day HA for Chlorothalonil. Accordingly, it is
recommended that . .- Longer-term HA value (200 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, 1988
-15-
Ten-day Health Advisory
No information was found in the available literature that was suitable
lor Jetcritilnation of z. Tsr.-day HA for chlorothalcnil. Accordingly, it :'s
recommended that the Longer-term HA value (200 ug/L, calculated below) for a
10-kg child be used at this time as a conservative estimate of the Ten-day
HA value.
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,m?/kg/day}, (1° kq) = 0.15 mg/L (200 ug/L)
(100)(1 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.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
1 L/day = assumed dailj water consumption of a child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = M-5 mg/kg/day) (70 kg) = Q.525 mg/L (500 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.
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Chlorothalonil August/ 1988
-16-
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or MAE/OD17 guidelines for use with a NCAEL frcn 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
(DUEL) 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. 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 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 = <1'5 mg/kg/day) - 0.015 mg/kg/day
(100)
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Chlorothalonil August/ 1988
-17-
where:
1.5 mg/kg/day = NOAEL, based on absence of histopathological changes
in io^s f&d chlcrothalor.il for or.a year.
100 = uncertainty factor, chosen in accordance with EPA
or 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 (500 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
Chlorothalonil is classified in Group B2: probable human carcinogen.
Accordingly, a Lifetime HA is not recommended for Chlorothalonil.
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
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 B6C3F1
mice when administered in the diet, at 403 or 806 Tier/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 CD-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
groups. 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.
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Chlorothalonil August/ 1988
-18-
0 The International Agency for Research on Cancer (IAPC, 1983)
has evaluated the carcinogenic potential of chlorothalonil and
concluded that there is limited evidence of carcinogenicity in
experimental animals.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA/ 1986), 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^j* 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 approac1
However, for completeness, the 10-6 risk numbers for other models
will be given. These values, at the 10-6 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 o,- e NOAEL
of 1.5 mg/kg/day identified in the 2-year dog study (Hols, .g 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).
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, 1988). 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. This
method has been validated in a single laboratory, and estimated
detection limits have been determined for analytes in this method,
including chlorothalonil. The estimated detection limit for chloro-
thalonil is 0.025 ug/L.
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Chlorothalonil
August/ 1988
-19-
VIII. TREATMENT TECHNOLOGIES
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, 1988
' -20-
IX. REFERENCES
Auletta, C.S., and L.F. Rubin.* 1981. Eye irritation studies in monkeys and
ratbita with Bravo 500: Report DS-2787. Unpublished study. MP.ID 00077176.
Blackmore, R.H., and L.D. Shott.* 1968. Final report: three-month feeding
study—rats. Project No. 200-205. Unpublished study. HRID 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.
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Diamond Shamrock Corporation.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.275. July 1, 1985.
CHEMLAB. 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., M. Griselli, M. Giovannetti and R. Barale. 1980. Mutagenicity
of pesticides evaluated by means of gene conversion in Saccharomyces
cerevisiae and Aspergillus 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.
IARC. 1983. International Agency for Research on Cancer. IARC Monographs
on the evaluation of carcinogenic risks of chemicals to humans.
Miscellaneous pesticides. IAPC Monographs, Volume 30. pp. 319-328.
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.
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Chlorothalonil August, 1988
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Kouri, R.E., R. Joglekar and D.P.A. Fabrizio.* 1977a. Activity of
DTX-77-0034 in an iji vitro mammalian cell point mutation assay.
Unpublished study. MRID 00030289.
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; Repair test. Final report.
unpublished study. MRID 00030288.
Kouri/ R.E., A.S. Parmar, J.M. Kuzava et al.* 1977c. Activity of DTX-77-0035
in the Salmonella/microsomal 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 Tokoharaa, 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 14C-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. Willoughby, 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 00127854.
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.
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Chlorothalonil August, 1988
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Paynter, O.E., and J.C. Murphy.* 1967. Final report: 16-week dietary
feeding—dogs. Project No. 200-200. unpublished study. MRID 00057698.
Pollock, G., J. Marciniszyn, J. ICilleen at al.* 1983. Levels of radioactivity
in blood following oral administration of 14c-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.
Quinto, I., G. Martire, G. Vricella, F. Riccardi, A. Perfume, 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 14C-
chlorothalonil (14c-DS-2787) to male mice. Unpublished study. MRID
00137125.
Rodwell, D., M. Mizens, N. Wilson et al.* 1983. A teratology study in rats
with technical chlorothalonil. Unpublished study. MRID 00130733.
Ryer, F.H.* 1966. Radiotracer metabolism study. Unpublished ^tudy.
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, «.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.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
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Szalkowski, M.B.* Undated. Photodegradation and mobility of Daconil and
its major metabolite on soil thin films. Unpublished study submitted
by Diamond Shamrock Agricultural Chemicals, Cleveland, OH.
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Chlorothalonil August, 1988
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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 Qaconil 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-
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Szalkowski, M.B., J.J. Mannion, D.E. Stallard et al.* 1979. Quant-
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2,4,5,6-tetrachloroisophthalonitrile (chlorothalonil, DS-2787) in soil.
Unpublished study submitted by Diamond Shamrock Agricultural Chemi-
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Tierney, W., N. Wilson, J. Killeen et al.* 1983. A chronic dietary study in
mice with technical chlorothalonil. Unpublished study. MRID 00127858.
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for carcinogenic risk assessment; Request for comments. Fed. Reg.
49(227) .-46294-46301. November 23.
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- Determination of chlorinated pesticides in water by GC/LCP . Xpril 15,
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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.
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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.
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Chlorothalonil August, 1988
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Wolfe, A.L., and D.E. Stallard.* 1968a. The fate of DAC-3701 (4-hydroxy-
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nitrile. Unpublished study. MRID 00087254.
•Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
CYANAZINE
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.
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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Cyanazine August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 21725-46-2
Structural Formula
¥ T
-C(CH,)t
N-CH,CH,
H
2-[[4-Chloro-6-(ethylamino)-1,3,5-triazin-2-yl]amino]-2-methylpropanenitrile
Synonyms
• Cyanazine (common name), Bladex, Fortrol, Payze, SD1518, VL19804,
DW3418 and WL19805 (Meister, 1983).
Uses
• Cyanazine is used as a pre- and postemergence herbicide for the
control of annual grasses and broad leaf weeds (U.S. EPA, 1984a).
Properties (U.S. EPA, 1984a; Meister, 1983; CHEMLAB, 1985)
Chemical Formula CgH13ClN6
Molecular Weight 240.7
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 167.5 to 169*C
Density 0.35 (fluffed) to 0.45 (packed) g/cc
Vapor Pressure (20°C) 1.6 x 10-9 to 7.5 x 10-9 mm Hg
Water Solubility (25»C) 171 mg/L
Log Octanol/Water Partition 2.24
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
• Cyanazine has been found in 1,708 of 5,297 surface water samples
analyzed and in 21 of 1,821 ground water samples (STORET, 1988).
Samples were collected at 392 surface water locations and 1,314 ground
water locations. The 85th percentile of all non-zero samples was
4.11 ug/L in surface water and 0.20 ug/L in ground water sources.
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Cyanazine August, 1988
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The maximum concentration found in surface water was 1,300 ug/L and
in ground water it was 3,500 ug/L. Cyanazine was found in surface
water in 7 States and in ground water in 5 States. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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 Nay through August; levels of 2 to 151
ng/L were detected during September through December; and zero levels
were found from January through April (U.S. EPA, 1984a; NAS, 1977).
0 Cyanazine has been found in surface water in Ohio river basins
(Datta, 1984).
0 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).
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 amine, was found only in the air-dried sandy clay loam soil.
0 Freundlich K values were 0.72 for a sandy loam 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 l4C-cyanazine (Osgerby
et al., 1968). No linear correlation was found between organic
matter content and adsorption.
0 14C-Cyanazine readily moved through 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 14C-cyanazine was inter-
mediately mobile on sandy clay loam and of low mobility on loamy
sand soil thin-layer chroraatography (TLC) plates (Rf 0.36 and 0.20,
respectively) (McMinn and Standen, 1981). Aerobically and anaerobically
aged l^c-cyanazine residues, primarily the amide degradate (SD 20258),
were intermediately mobile to mobile on sandy clay loam soil TLC plates.
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Cyanazine August, 1988
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Aged 14c-cyanazine residues readily leached through columns containing
sand (47.8% of applied compound), loamy sand (69.7% of applied compound)
and sandy loam (26.9% of applied compound) soils eluted with 20 cm of
water (Eadsforth, 1984). The amide degradation product (SD 20258)
was predominant in the leachate from the sandy soil (45% of radio-
activity in leachate); the acid degradate (SO 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 residues).
III. PHARMACOKINETICS
Absorption
0 Studies by 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 rag/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 major metabolic
pathways of cyanazine in the rat and cow involved: (1) conversion of
the cyano group to an amide to form 2-chloro-4-ethylamino-6-(1-amido-
1-methylethylamino)-s-thiazine; (2) N-deethylation to form 2-chloro-4-
amino-6-(1-cyano1-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-amino-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
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Cyanazine August, 1988
-5-
2-hydroxy-4-ethylamino-6-(1-carboxy-1-methylethylamino)-s-triazine
and 2-hydroxy-4-amino-6-(1,carboxy-1-methylamino)-s-triazine,
respectively (Shell, 1969).
Studies by Shell Chemical Company (1969) and Hutson et al. (1970) in
rats with ring-labeled and side-chain-labeled cyanazine (cyano-14cr
isopropyl-14c and ethylamino-14C) indicated that only the ethylamino-Hc
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.
Crayford and Hutson (1972) identified 5 metabolites in urine of
rats, an additional 2 (total 7) in feces and 4 metabolites in bile.
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-^arboxy-1-methylamino)-s-triazine
(DW4385) and as 2-hydroxy-4-amino-6-{1-carboxy-1-methylethylamino)-
s-triazine) (DW4394). Approximately 91% of compound OW4385 and 84%
of compound DW4394 were recovered unchanged from urine and feces.
Excretion
Orally administered low doses of cyanazine (described above) were
rapidly excreted in the urine and feces of rats and dogs (Shell,
1969; Hutson et al., 1970; Crayford and Hutson, 1972).
In rats treated with 1 to 4 mg/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, 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, 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, 1969).
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Cyanazine August, 1988
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IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of cyanazine in humans.
Animals
Short-term Exposure
0 Acute oral LDgg values reported for rats range from 149 to 835 mg/kg
(SRI, 1967b; NIOSH, 1987; Young and Adaraik, 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 LDsgs of 182, 380 and 141 mg/kg
for the rat, mouse and rabbit, respectively.
0 The acute dermal 1.050 in rabbits treated with technical cyanazine
(purity unspecified) was >2,000 mg/kg (SRI, 1967a; Young and Adamik,
1979c); in rats, the LDso was >1,200 mg/kg (97% a.i.) (Walker et al.,
1974).
0 The acute inhalation LCso for cyanazine dust (% a.i. not specified)
in rats was >2.28 mg/L/hr (Bishop, 1976) (thus cyanazine would be
classified in 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 mgAg 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 this 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.
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Cyanazine August, 1988
-7-
The LOAEL in this study appeared to be 0.05 mg/kg/day (lowest dose
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 at 100 mg (Young and Adamik,
1979a) and slight skin irritation at 2,000 mg (Young and Adamik,
1979c) 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;
Walker et al., 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
ernesis 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 et al. (1968) was
performed using 0.1, 1.0 or 100 ppm (equivalent to 0.005, 0.05 or
0.5 mgAg/day; Lehman, 1959) of technical cyanazine (purity not speci-
fied: 97% or 75% a.i.) 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
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Cyanazine August, 1988
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consisted of six animals/sex and received empty gelatin capsules.
Frequent ernesis within 1 hour of dosing was observed throughout the
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-
mgAg/day dosage group. In addition, the reported data were limited
to a summary report.
8 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 mg/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 mgAg/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 in the diet of
rats (Lehman, 1959); however, convulsions were noted in the rats
3 months after the study initiation and throughout the remainder of
the 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, 1988
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0 A recent one-year feeding study in dogs using atrazine (98% a.i.)
by Dickie (1986) has been evaluated by the Agency. Five experimental
groups of 6 animals/sex/ group were exposed to the following doses:
0, 10, 25, 100 or 200 ppm. These doses were equivalent to actual
consumption of 0, 0.27, 0.68, 3.20 or 6.11 mg/kg/day for males and
0, 0.28, 0.72, 3.02 or 6. 39 mg/kg/day for females, respectively. No
systemic toxicity was noted at 10 or 25 ppm. However, dose-related
decreases in body weight and body weight gains were noted at 100 and
200 ppm, as well as elevated platelet counts, reduced levels of total
protein, albumin and calcium in both sexes. At these two high -dose
levels, there were also slight but not statistically significant
decreases in spleen weights, and increases in relative liver weights
in females, and increases in liver weights and decreases in testes
weights in males. Other noted changes in organ weights (i.e. , heart,
lung and kidneys) were not considered significant at these dose
levels since they were not consistent with changes in the absolute
and relative organ weight values. No gross or microscopic
findings related to treatment were noted. Thus, in this study, the
LOAEL is 3.1 mg/kg/day and the NOAEL is 0.7 mg/kg/day.
Reproductive Effects
0 A three-generation reproduction study in Long-Evans rats (Eisenlord
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
based on the dietary assumptions of Lehman, 1959) 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 F^ female weanlings.
0 In a repeat two -generation reproduction study in Sprague -Dawley rats
(MIL Research Laboratory, 1987), atrazine (100% a.i.) was administered
in feed at 0, 25, 75, 150 or 250 ppm.- These doses are equivalent to
actual food consumption of 0, 1.8, 5.3, 11.1 or 18. 5 mg/kg/day,
respectively; however, these values changed during lactation to 0, 3.8,
11.2, 23.0 or 37.1 mg/kg/day, respectively. Dose-related decreases in
pups' viability and body weights were noted at 75 ppm and above,
therefore, the NOAEL for reproduction may be 25 ppm (3.8 mg/kg/day).
However, this level, 25 ppm, (equivalent to 1 . 8 mg/kg/day during
non-lactating periods of the study) may be considered as the LOAEL
for parental animals due to the noted decreases in body weight and
food consumption at this level.
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,
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Cyanazine August, 1988
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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 anophthalmia/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 (98.5% 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% Methocel
emulsion as a vehicle. Maternal body weight reductions during dosing
were noted at the 10- and 25-mg/kg/day levels. Diaphragmatic hernia
associated with liver protrusion, microphthalmia and anophthalmia
were observed at the 25 mg/kg/day dose level. A teratogenic NOAEL
could not be determined from this study at 10 mg/kg/day and a maternal
toxicity NOAEL at 2.5 mg/kg/day.
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
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/kg/day (Bui, 1985a).
0 An additional study in Sprague-Dawley rats (Shell, 1983) did not
reflect any maternal or developmental toxicity at the highest dose
tested, 30 mg/kg/day.
Mutagenicity
0 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. (1974a) 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
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Cyanazine August, 1988
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8- and 24-hour intervals after oral dosing with 50 or 100 rag/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, 1984b).
0 Dean et al. (1974b) 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).
Carcinogenicity
0 Cyanazine was not determined to have a carcinogenic potential in a
2-year mouse study (Shell, 1981).
• 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.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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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Cyanazine August, 1988
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BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 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-
Dawley rats (Shell, 1983).
Using 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 m /L (100 u /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.
100 = uncertainty factor, chosen in accordance with EPA
or 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 0.02 mg/L (20 ug/L) be used for a 10-kg
child as a conservative estimate for the Longer-term HA value and the DWEL
of 0.07 mg/L (70 ug/L), calculated below, be used for a 70-kg adult.
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Cyanazine August, 1988
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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. 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.
Five chronic studies were available for evaluation: (Da 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 Long-term Exposure Section); (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; and recently, (5) a 1-year feeding study in dogs
(Dickie, 1986) with a NOAEL of 25 ppm (approximately 0.7 mg/kg/day based on
actual mean food consumption of males and females).
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, the 2-year dog study (Walker et al., 1970a)
was used previously for the Lifetime HA calculations, using a NOAEL of
1.25 mg/kg/day. This study was of marginal acceptability because only a
summary report was available for evaluation and histopathological data were
missing for 1/4 females at 1.25 mg/kg/day. However, this NOAEL was
supported by a similar NOAEL from a 13-week rat subchronic feeding study by
Walker and Stevenson (1968b). Thus using this NOAEL from both studies (i.e.,
the 2-year dog study and 13-week rat study) and applying a large uncertainty
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Cyanazine August, 1988
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factor of 1,000-fold was appropriate for the calculation of the RfD and
the Lifetime HA (in the absence of more adequate chronic studies at that time)
At the present time, the new 1-year dog feeding study by Dickie (1986) is a
more adequate study for these calculations.
The NOAEL of 25 ppm (equivalent to 0.7 mg/kg/day based on mean actual
food consumption in male and female dogs) from the Dickie study (1986) is
used in the calculation of the RfD. In this 1-year study, thirty male and
female dogs (6 animals/sex/dose) were fed diets containing 0, 10, 25, 100 or
200 ppm cyanazine (98% a.i.). These doses were equivalent to actual mean
intakes of 0, 0.27, 0.68, 3.20 or 6.11 mg/kg/day in males and 0, 0.28, 0.72,
3.02 or 6.39 mg/kg/day in females, respectively. No systemic toxicity was
noted at 25 ppm (0.7 mg/kg/day) in both sexes. The LOAEL was 100 ppm
(3.1 mg/kg/day) based on reduced body weights and body weight gains, elevated
platelet counts/ and reduced levels of total protein, albumin and calcium in
males and females. There were also slight, not statistically significant,
decreases in spleen weights and increases in liver weights in the females and
increases in liver weights and decreases in testes weights in the males. No
gross or microscopic findings related to treatment were noted.
Using a NOAEL of 0.7 mg/kg/day, the Lifetime HA is calculated as
follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.7 mg/kg/day) = 0.002 mg/kg/day
(300)
where:
0.7 mg/kg/day = NOAEL based on absence of toxicity in the dog
during the one-year feeding exposure.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
3 = modifying factor used to compensate for the lack of
a chronic rat study as required by the Office of
Pesticide Programs.
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.
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Cyanazine August, 1988
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Step 3: Determination of the lifetime Health Advisory
Lifetime HA = (0.07 mg/L) (20%) = 0.014 mg/L (10 ug/L)
where:
0.07 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., 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.
0 Cyanazine is a chloro-s-triazine derivative that has a chemical
structure analagous to atrazine, propazine and simazine. These three
analogs were found to significantly (p <0.05) increase the incidence
of mammary tumors in rats and they are classified as group C oncogens.
Based on structure-activity relationship, cyanazine may reflect a
similar pattern of 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 cyanazine.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), cyanazine may be classified in
Group 0: 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, 1985) 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. Method #4 (U.S. EPA, 1986b). In this method, 1 L of sample
is extracted 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. Using
this method, the estimated detection limit for cyanazine is 0.3 ug/L.
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Cyanazine August, 1988
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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, 1988
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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). MRIO #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., B.C. 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. MRIO #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., K.R. Senner, B.D. Perquin and S.M.A. Doak.* 1974a. 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)
Dean, B.J., E. Thorpe and D.E. Stevenson.* 1974b. 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)
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.
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Cyanazine August, 1988
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Dickie, B.C. 1986. One-year oral feeding study in dogs with the triazine
herbicide - cyanazine. Study #6160-104 and addendum #107F (unpublished
study performed by Hazleton Laboratories for duPont deNemours & Co.).
MRID 40081901 and 40229001.
Eisenlord, G., G.S. Loquvam and S. Leung.* 1969. Results of reproduction
study of rats fed diets containing 50 15418 over three generations:
Report No. 47. (Unpublished study received on unknown date under 9G0844;
prepared by Hine 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 studies 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-
(1-methyl-1-cyanoethylamino)-s-triazine in the rat. J. Agric. Food. Chera.
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 (SD-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:070584-A). 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.
-------
Cyanazine August, 1988
-19-
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC.: National Academy Press.
NIOSH. 1987. 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)
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 atrazine: A maize-
microbe bioassay. Mutat. Res. 38:287-292.
Shell.* 1969. Metabolism of cyanazone (Unpublished study submitted by Shell
Chemical Company). MRID #00032348. (Cited in U.S. EPA, 1984b)
Shell.* 1981. Two-year oncogenicity study in the mouse. (Unpublished report
submitted under pesticide petition number 9F2232, EPA accession number
247295 to -298).
Shell.* 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 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 and -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
#960844, 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 SD-15418 (technical cyanazine). Submitted
by Shell Chemical Co., Washington, DC., Pesticide Petition #9G0844,
Accession #91460. (Cited in U.S. EPA, 1984b)
STORET. 1988.. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
U.S. EPA. 1984a. U.S. Environmental Protection Agency. Draft health and
environmental effects profile for cyanazine. Cincinnati, OH: Environmental
Criteria and Assessment Office.
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Cyanazine August, 1988
-20-
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. 1985. U.S. Environmental Protection Agency. 40 CFR. 180.307.
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 #4
- Determination of pesticides in ground water by HPLC/UV, January, 1986
draft. Available from U.S. EPA's Environmental Protection Monitoring and
Support Laboratory, Cincinnati, Ohio 45268.
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. 00 07.6 9.
(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. Kbdama, E. Thorpe and A.B. Wilson. 1974.
Toxicological studies with the 1,3,5-triazine herbicide cyanazine.
Pestic. Sci. 5(2):153-159.
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Cyanazine August, 1988
-21-
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.
WIL Research Laboratories. 1987. Two-generation rat reproduction study,
WIL #93001, August 12 (unpublished study submitted by duPont).
MRID # 403600-01.
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.
Young, S.M., and E.R. Adamik.* 1979a. Acute eye irritation study in rabbits
with SD 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.)
MRID #00026427. (Cited in U.S. EPA, 1984b)
Young, S.M., and E.R. Adamik.* 1979b. Acute oral toxicity study in rats
with SD 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.)
MRID #00026424. (Cited in U.S. EPA, 1984b)
Young, S.M., and E.R. Adamik.* 1979c. Acute dermal toxicity study in rabbits
with SD 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.)
MRID #00026425. (Cited in U.S. EPA, 1984b)
Young, S.M., and E.R. Adamik.* 1979d. Delayed contact in hypersensitivity
study in guinea pigs with SD 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 Office of Pesticide
Programs
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August, 1988
DCPA (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
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.
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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DCPA (Dacthal)
August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1861-32-1
Structural Formula
Synonyms
Dimethyl tetrachloroterephthalate
Uses
2,3,5,6-Tetrachlorodimethyl-1,4-benzenedicarboxylic acid; DCPA;
Chlorothal; Dacthalor; DAC; DAC-4; DAC-893; DCP (Meister, 1983)
0 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
0 Dacthal has been found in 386 of 1,995 surface water samples analyzed
and in 12 of 982 ground water samples (STORET, 1988). Samples were
collected at 584 surface water locations and 844 ground water locations,
331.99
Crystals
156°C
2.5 x 10-6 mm Hg at 25«C
0.5 mg/L at 25»C
6.8 x 103
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DCPA (Dacthal) August, 1988
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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. This information is
provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime/ as they are here.
Therefore/ this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
0 In aqueous solutions/ dacthal is not very photolabile with a half-
life of greater than one week. Dacthal is also stable versus soil
photolysis (Registrant Standard Science chapter for dacthal).
0 Soil metabolism of dacthal proceeds with a half-life of greater than
2 to 3 weeks. Degradation rate is affected by temperature. No
degradation of dacthal has been observed in sterile soils (half-life
of 1/590 days) (Registrant Standard Science chapter for dacthal).
0 Degradation products of dacthal include monomethyltetrachlorotere-
phthalate (MTP) and tetrachloroterephthalic acid (TTA) (Registrant
Standard Science chapter for dacthal).
0 TTA has been shown to be very mobile in soils/ whereas dacthal is not
(Registrant Standard Science chapter for dacthal).
III. PHARMACOKINETICS
Absorption
0 Tusing (1963) reported that 3 humans receiving single oral doses of
pure dacthal (25 mg) excreted up to 6% of the doses in the urine as
metabolites over a 3-day period. When 3 other humans were administered
50 mg doses of dacthal/ approximately 12% was excreted in the urine
over a similar time period, indicating that at least 12% of a 50 mg
dose was 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/
about 97% of the administered doses were eliminated as the parent
compound in the feces by 96 hours. Approximately 3% of dacthal was
converted to the monomethylester of tetrachloroterephthalic 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 tetrachloro-
terephthalic 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.
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DCPA (Dacthal) August, 1988
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Distribution
8 Skinner and Stallard (1963) reported that following a single oral
dose of dacthal (100 or 1,000 ing/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
(dacthal containing 1.1% of the monomethyl ester and 1.7% of the
tetrachlorophthalate) at 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 Tusing (1963) reported that humans who took single oral doses of pure
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%).
0 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 (Tusing, 1963), 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
Tusing (1963) reported that pure dacthal, administered as single
25 mg or 50 mg oral doses to volunteer subjects (3 at each dose), did
not cause any observable effects. Assuming 70 kg body weight, these
amounts correspond to doses of 0.36 or 0.71 mg/kg. Hemograms, liver,
kidney and urine analyses from the six human volunteers were normal.
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DCPA (Dacthal) August, 1988
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Animals
Short-term Exposure
0 The acute oral LD$Q for male and female albino rats (Spartan strain)
was reported to be greater than 12,500 mg/kg (Wazeter et al., 1974a).
0 The acute oral LDso 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 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).
0 Keller (1961) reported that oral administration (by capsule) of
800 mg/kg/day of DCPA (88.5% active ingredient) 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 1*050 value for albino rabbits was reported to be
greater than 10,000 mg/kg (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 White
rabbits (five/sex) in a paste form. Desquamation (which ranged from
very slight to slight) and very slight erythema were observed. There
was no pathology noted, and dacthal caused only slight irritation in
some animals*
0 A single application of 3.0 mg 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 Goldenthal 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,
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DCPA (Dacthal) August, 1988
-6-
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 Dieterich (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
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 mg/kg/day for females; the highest dose tested) was identified
for this study.
0 Paynter and Kundzin (1963b) 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 mg/kg/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 0, 50 or 500 mg/kg/day.
This study reported an evaluation of data collected on the second
parental generation (?2) and through weaning of the first litter
(F2a)> The authors reported that a second litter (?2b' 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. Mb other adverse reproductive effects were observed.
0 Paynter and Kundzin (1963a) 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
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DCPA (Dacthal) August, 1988
-7-
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.
Developmental Effects
0 Powers (1964) fed pregnant New Zealand white 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 mg/kg/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).
Mutagenicity
0 No significant increase in mutation frequency was observed in Droso-
phila melanogaster larvae that had been fed media containing 0.1 to
10 mH dacthal (Paradi and Lovenyak, 1981).
0 Dacthal had no mutagenic activity, with or without activation, in
Salmonella assays (Auletta et al., 1977), in iji vivo cytogenetic
tests (Kouri et al./ 1977b), in DNA repair tests (Auletta and Kuzava,
1977) or in dominant lethal tests (Kouri et al./ 1977a).
Carcinogenicity
0 Paynter and Kundzin (1963b) 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 (up to 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)
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
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DCPA (Dacthal) August, 1988
-8-
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 Tusing (1963) 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 1 0-kg child
(80 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 mgAg/day) (10 kg) _ 75 mg/L (80,000 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 da c thai for 28 days.
1 0 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 were found in the available literature that were suitable for
deriving the Longer-term HA value for dacthal. It is, therefore, recommended
that the Drinking Water Equivalent Level (DWEL) of 20 mg/L (20,000 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, 5.0 mg/L (5,000 ug/L), be used for the
Longer-term HA value for a child.
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DCPA (Dacthal) August, 1988
-9-
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and LS 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. 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 study in rats by Paynter and Kundzin (1963b) 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
(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 EPA
or 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/kg/day) (70 kg) = 17.5 mg/L (2000 ug/L)
(2 L/day)
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DCPA (Dacthal) August, 1988
-10-
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 (4,000 ug/L)
where:
17.5 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Paynter and Kundzin (1963b) 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/ 1986), dacthal may be classified in
Group 0: not classified. This category is for substances with inade-
quate animal evidence of carcinogenic!ty.
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 pesticide in water samples
(U.S. EPA, 1988). 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. This method has been
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DCPA (Dacthal) August, 1988
-11-
validated in a single laboratory and estimated detection limits have
been determined for the analytes in the method, including dacthal.
The estimated detection limit is 0.02 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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DCPA (Dacthal) August, 1988
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IX. REFERENCES
Auletta, A., A. Parmar and J. Kuzava.* 1977. Activity of DTX-0003 in the
Salmonella/mlcrosomal assay for bacterial mutagenicity. Microb. Assoc.
Rpt. DS-002. Unpublished study. MRID 00100774.
Auletta, A., and J. Kuzava.* 1977. Activity of DTX-77-0005 in a test for
differential inhibition of repair deficient and repair competent strains
of Salmonella typhimurium. Microb. Assoc. Rpt. DS-0001. Unpublished
study. MRID 00100776.
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.
Goldenthal, E.I., F.X. Wazeter, 0. Jessup et al.* 1977. 90-day toxicity
study in rats. MRID 00100773.
Hazleton, L.N., and W.H. Dieterich.* 1963. Final report: Two-year dietary
feeding - dogs. Unpublished study. MRID 00083584.
Johnson, D., J. Myer and A. Olafsson.* 1981. Acute dermal toxicity (LDsg)
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. 28- day dietary feeding study - rats.
Unpublished study. MRID 00083571.
Kouri, R., A. Parmar, J. Kuzava et al.* 1977a. Activity of DTX-77-0004 in
the dominant lethal assay in rodents for mutagenicity. Microb. Assoc.
Proj. No. T1077. Final Report. Unpublished study. MRID 00100775.
Kouri, R., A. Parmar, J. Kuzava et al.* 1977b. The activity of DTX-77-0006
in the in vivo cytogenetic assay in rodents for mutagenicity. Microb.
Assoc. Proj. No. T1083. Unpublished study. MRID 00107907.
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.
Paradi, E., and M. Lovenyak. 1981. Studies on genetical effect of pesticides
in Drosophila melanogaster. Acta Biol. Sci. Hung. 32:119-122.
-------
DCPA (Dacthal) August, 1988
-13-
Paynter, O.E., and M. Kundzin.* 1963a. Reproduction study - albino rats.
MRID 00083578.
Paynter, O.E., and M. Kundzin.* 1963b. Two year dietary administration - rats.
Final Report. MRID 00083577.
Paynter, O.E., and M. Kundzin.* 1964. Reproductive study - rats. Second
phase: Project No. 200. 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.
MRID 00083579.
STORET. 1988. STORE! Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Tusing, T.W.* 1963. Oral administration - humans. MRID 00083583.
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. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51 ( 1 85) : 33992-34003. Septem-
ber 24.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 515.1 -
Determination of chlorinated pesticides in water by GC/ECD, April 1 5,
1988 draft. Available from U.S. EPA's Environmental Monitoring and
Support Laboratory, Cincinnati, Ohio.
Wazeter, F.X., E.I. Goldenthal and W.P. Dean.* 1974a. Acute oral toxicity
male and female albino rats. Unpublished study. MRID 00031872.
Wazeter, F.x. , E.I. Goldenthal and W.P. Dean.* 1974b. Acute oral toxicity
(LD50) in beagle dogs. Unpublished study. MRID 00031873.
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 Pesticide
Programs.
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February, 1989
DALAPON
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 ovpr 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.
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 February, 1989
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 75-99-0
Structural Formula
CH3CCI2COOH
(2,2-Cicnloropropionic acid)
Synonyms
Dalapon (ANSI, BSI, WSSA), DPA, a_ fapon 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: Reinert and Rogers, 1987)
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
Water Solubility (25°C) 50,000 mg/L
Log Octanol/Vbter Partition 1.47
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, 1988).
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Dalapon February, 1989
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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).
9 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 salt 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 highly
mobile compound (Warren, 1954; Helling, 1971; Kenaga, 1974) and should
be readily leachable from soils, field data show 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 February, 1989
-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 ppai. Based on the observations that dalapon decomposition .s
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. PHARMAOOKINETICS
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
0 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 humans, data in cattle (Redemann and Hanaker, 1959)
suggest that dechlorination may be involved in the metabolism of
dalapon.
Excretion
Available information suggests that at least 50% of orally admini-
stered dalapon is eliminated via the kidneys in dogs and humans
(Hoerger, 1969).
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Dalapon February, 1989
-5-
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 :alapon in humans
was found in the available literature.
Animals
Short-term Exposure
0 The sodium salt of dalapon is relatively nontoxic, with an oral LDsg
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 Concentrated sodium dalapon solutions have been found to be irritating
to the skin and eyes of rabbits (Paynter et al., 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/kg/day. The
No-Observed-Adverse-Effect Level (NOAEL) in this study was identified
as 11.5 mg/kg/day based on increases in kidney weight at higher
doses. (See discussion under Longer-term Health Advisory below.)
a
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 al., 1960).
0 With the exception of an increase in kidney weight in male rats,
sodium dalapon (65% pure) was without effect in a 2-year dietary study
(Paynter et al., 1960); the NOAEL in this study was 15 mg/kg/day.
(See discussion under Longer-term Health Advisory below.)
Reproductive Effects
0 Administered 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 February, -1989
-6-
Developmental Effects
0 Sodium dalapon (purity not specified) was not teratogenic in the rat
at doses as high as 2,000 mgAg/day (Bnerson et al., 1971; Thomson
et al., 1971). In the study by Emerson et al, sodium dalapon was
administered orally to pregnant rats over a 10-day period (day 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 decrease 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). This study
identified a LOAEL of 500 mg/kg/day.
0 Dalapon (sodium/magnessium salt, purity 99.3%) was administered
orally (gavage) to New Zealand white female rabbits on days 6
through 13 of gestation at doses of 30-, 100-, and 300 mg/kg/day.
Nineteen animals were treated at the high- dose (300 mg/kg/day) and
sixteen animals per group were used as controls as well as the low-,
mid- and high- dose groups (BASF, 1987). Administration of dalapon
at the mid- and high- dose elicited maternal toxicity: significant
decreases in the body weight gain and food consumption. At the high
dose, decreased fetus weight, decreased water consumption, enlarged
spleen and even abortions (three abortions in 19 dams) were noted.
Administration of dalapon at the low dose (30 mg/kg /day), however,
did not cause any adverse effect. This study identified a NOAEL of
30 mg/kg/day for dalapon Na/Mg salt.
Mutagenicity
0 Dalapon was not mutagenic in a variety of organisms including Salmonella
tvphimurium, Escherichia coli, T4 bacteriophage, Streptomyces soelicolor
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.
Care inogen ic i ty
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 TOXIOOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of toxicity.
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Dalapon February, 1989
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The HAs for noncarcinogeruc toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) X (BW) = mg/L ( ^^
(UF) x ( L/day) ^ ^'
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mo/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10.000),
in accordance with EPA or 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 (3 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 Emerson et al. (1971) had been considered
to serve as the basis for determination.of the Ten-day HA for a 10-kg
child. However, in this study, the chemical purity was not specified, and
the lowest dose administered (500 mg/kg/day, LOAEL) had maternal toxicity.
A recent teratology study with rabbits (BASF, 1987) was selected to serve as
the basis for the 10-day HA. This study specified the purity of the chemical
and provided a lower LOAEL (100 mg/kg/day), and a NOAEL of 30 mg/kg/day. In
this study, groups of inseminated New Zealand white rabbits were given oral
doses of 0, 30-, 100- and 300 mg/kg/day dalapon (Dowpon M: Na/Mg salt, purity
99.3%) on days'6 through 18 of gestation. Significant decreases in maternal
body weight, in food and water consumption were noted in the mid- and high-
dose groups. Fetal body weight from dams given the high- dose was also
decreased. However, no adverse effects were noted in the low- dose group,
and a NOAEL of 30 mg/kg/day was identified. Standards for dalapon are
commonly expressed in terms of the acid rather than the salt. Thus, it is
necessary to convert the NOAEL for the Na/Mg salt, 30 mg/kg/day to the
equivalent value for the acid. It is assumed that Dowpon M was approximately
5:1 mixture of sodium and magnessium salts.
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Dalapon February, 1989
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The NOAEL for dalapon as acid = (30 mq/kq/day)(143}(7) _ 26.5mg/kg/day
<165J(5) + (308.5)(1)
whers:
30 mg/kg/day = NOAEL for the Na/Mg salt
143 = formula weight of dalapon as acid
163 = molecular-weight of sodium dalapon
308.5 = molecular weight of magnessium dalapon
7 = total number of dalapon acid moiety in Dowpon
5 = number of sodium dalapon in Dowpon M
1 = number of magnessium dalapon in Dowpon M
(Each magnessium binds two acid moieties)
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA = (25.5 mg/kg/dayHlO kg) =2.7 mg/L (3,000 ug/LJ
(100)(1 L/day)
where:
26.5 mg/kg/day = NOAEL of dalapon as acid based on body weight
decreases in dams
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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Dalapon February, 1989
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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 dl. (1960) rat
dietary studies together, the 15 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 = (IS mg/kg/day) (0.65) (143) = 3 ma/ka/dav
165
where:
15 mg/kg/day = NOAEL for 65% pure sodium dalapon.
0.65 = purity of sodium dalapon used in determining NOAEL.
143 = 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/kq/day) (10 kg) = 0.26 mg/L (300 ug/L)
^ (100) (3) (1 L/day)
where:
8 mg/kg/day = NOAEL based on kidney weight increases in male rats.
10 kg - assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/OCW
guidelines for use with a NOAEL from animal study*
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Dalapon February, 1989
-10-
3 = additional uncertainty factor of 3 to allow for
deficiencies in the quality of the data base.
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 = (8 mg/kg/day) (70 kg) =0.9 mg/L (900 ug/L)
(100) (3) (2 L/day) * y/
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.
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 affects 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 DUEL 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. 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.
As discussed under Longer-term HA above, 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:
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Dalapon February, 1989
-11-
Step 1: Determination of the Reference Dose (RfD)
RfD = (8 mg/kg/day) = 0.03 mg/kg/day
(100) (3)
where:
8 mgAg/day = NQAEL for 100% dalapon as acid.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL front an animal study.
3 = additional uncertainty factor of 3 to allow for
deficiencies in the quality of the data base.
Step 2: Determination of the Drinking Hater Equivalent Level (DWEL)
DWEL = (0.03 mg/kg/day) (70 kg) = 0.9 mg/L (900 ug/L)
(2 L/day) W
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 = (0.9 mg/L) (20%) = 0.18 mg/L (200 ug/L)
where:
0.9 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 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.
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Dalapon February, 1989
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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. ANALmCAL 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, 1936b). 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 determiner'
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 February, 1989
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IX. REFERENCES
BASF Corporation. 1987. Dalapon: Oral (gavage) teratogenicity study in
the rabbit. BASF Corporation, Chemicals Division, Germany. MRID No.
40125701.
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.
Qnerson, 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).
HoIston, J.T., and W.E. Loomis. 1956. Leaching and decomposition of
2,2-dichloropropionic acid in several Iowa soils. Meeds. 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. Tbxicological and residue data useful in the environ-
mental safety evaluation of dalapon. Residue Rev. 53:109-151.
Kurinnyi, A.I., N.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.
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Dalapon February. 1989
-14-
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).
Reinert, K.H., and J.H. Rodgers. 1987. Fate and persistence of aquatic
herbicides. Rev. Environ. Contamination and Tox. pp. 61-98. V. 98.
Smith, G.N., M.E. Getzendaner and A.M. 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 Chamical Co., Midland, MI (cited in Kenaga, 1974).
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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 S3 -
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., llth Ann. Meeting,
Fargo, ND (cited in Kenaga, 1974).
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-, 1 983
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.
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 noncarcinogenlc end points of toxicity.
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,
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 333-41-5
Structural Formula
0,0-Diethyl-0-(2-isopropyl-4-methyl-6-pyrimidinyl)phosphorothiote
Synonyms
0 Antigal; AG-500; Basudin; Bazudln; Ciazinon; Ducutox; Dassitox;
Dazzel; Dianon; Diater; Oiaterr-Fos; Diazajet; Diazide; Diazitol;
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 C12H2103^3?
Molecular Weight 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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DiazLnon August, 1988
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Occurrence
0 Diazinon has been found in 6,026 of 22,291 surface water samples
analyzed and in 74 of 3,633 ground water samples (STORET, 1987).
Samples were collected at 3,555 surface water locations and 2,835
ground water locations, and diazinon was found in 46 states. The
85th percentile of all nonzero samples was 0.01 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. This information is provided to give a general impression
of the occurrence of this chemical in ground and surface waters as
reported in the STORET database. The individual data points retrieved
were used as they came from STORET and have not been confirmed as to
their validity. STORET data is often not valid when individual
numbers are used out of the context of the entire sampling regime,
as they are here. Therefore, this information can only be used to
form an impression of the intensity and location of sampling for a
particular chemical.
Environmental Fate
e 14c-oiazinon (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 1*C-radioactive material balance was 87-89% at
the 0 hour and 84% at all other experimental points.
e 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-1'-raethyl)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 258C, 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 INHP. 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 14C residues
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Diazinon August, 1988
-4-
(15.1% after 166 days) corresponded to IMHP breakdown. Mo other major
metabolites were found. Radioactivity in the H2S04 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).
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%.
9 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
9 The metabolism of diazinon 14C-labeled in the pyrimidine ring was
investigated in Wistar rats (200 g) after administration by stomach
tube (Hucke 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 In vitro in rat liver
mlcrosomes obtained from adult male rats (Nakatsugawa et al., 1969).
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Diazinon August, 1983
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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.
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 Wedin 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 ChE
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
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Diazinon August, 1983
-6-
plasma ChE 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).
Long-term Exposure
8 No information was found in the available literature on the long-term
health effects of diazinon in humans.
Animals
Short-term Exposure
8 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 DeProspo
(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 LD5Q value was calculated to be 395.6 mg/kg.
8 Hazelette (1984) investigated the effects of dietary hypercholesteremia
(HCHOL) on sensitivity to diazinon in inbred male C56BL/6J mice. The
LDso of diazinon in HCHOL 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 HCHOL animals. It was concluded that HCHOL resulted in an increase
in susceptibility to, and toxicity of, diazinon.
8 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.
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Diazinon August, 1983
-7-
0 The effect of diazinon (tech. grade) 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
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 Davles and Holub (1980a) compared the subacute toxicity of diazinon
(approximately 99%) 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 for 7 days, 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.
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Diazinon August, 1988
-8-
Long-term Exposure
o Female Wiscar rats were fed a semipurified diet containing 0 or 0.1
to 15 ppm diazinon 99%) for up to 92 days with no visible toxic
effects (Davies and Holub, 1980b). 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 cholines-
terase (ChE) activities. Plasma ChE was judged most sensitive. A
NOAEL of 0.1 ppm, 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.
8 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-tng/kg dose level had significant weight loss and signs of dehydra-
tion, emaciation, pale skin coloration and an unthrifty hair coat.
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.
0 Barnett and Rung (1980) fed Charles River CD-I mice diazinon (87.6%)
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-
liferatlve 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 (25%) 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,
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Diazinon August, 1988
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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 67%, 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 ing/kg/day).
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 F^ animals, and histopathology of F^ animals. All
findings were reported to be normal, but there were no detailed data
provided. A NQAEL of 8 ppm (0.4 mg/kg/day), the highest dose tested,
was identified in this study.
0 Diazinon was administered by gavage at dose levels of O/ 7, 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 statistically 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 resor'bed 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.
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Diazinon August:, 1983
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Developmental Effects
0 Diazinon at dose levels of 7, 25 or 100 rag/kg was administered by
gavage 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
carboxymethyl cellulose (CMC) and a group of controls was given 0.2%
CMC only. At the 100-tng/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. 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 mg/kg/day is identified.
0 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
mg/kg/day, the highest dose tested.
Mutagenicity
0 Four strains of Salmonella typhlmurium were used to assay ihe muta-
genic potential of diazinon (Marshall et al., 1976). Negative
results were found by these investigators as well.
0 The mutagenicity of diazinon was tested in bacterial reversion-assay
systems with five strains of Salmonella typhimurium and one strain of
Escherichia coll (Moriya et al., 1983). No evidence of mutagenic
activity was noted in any of the test systems.
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 mg/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.
Carcinogenicity
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.
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Diazinon August, 1988
-11-
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 noted 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.
0 Charles River CD-I mice were fed diazinon (87.6%) in the diet at
levels of 4, 20 or 100 ppm for 18 months (males) or 19 months (females)
(Barnett and Rung, 1980). Assuming that 1 ppm in the diet of mice is
equivalent to 0.15 mg/kg/day/ this corresponds to doses of about 0.6,
3 or 1 5 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, prollferative or neoplastlc lesions due to the admin -
istration of diazinon, and the study was judged to be negative with
respect to carcinogenic! ty.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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:
HA = (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, 1,000 or 10,000),
in accordance with EPA or NAS/OOW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
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Diazinon August, 1988
-12-
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.
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-term 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/ Ooull 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) = 0.02 /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 EPA
or NAS/OCW 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 Woodard 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 (Doull 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.
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Diazinon August, 1988
-13-
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) = 0.005 mg/L (5.0 ug/L)
(100) M L/day) * *
where:
0.05 mg/kg/day = NOAEL, based on absence of ChE inhibition in monkeys
given diazinon orally for 52 weeks.
10 kg = assumed body weight of a child.
100 = uncertainty factor/ chosen in accordance with EPA
or 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 = (0-05 mg/kg/day) (70 kg) „ Q.0175 rag/L (20 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 EPA
or 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
(DUEL) 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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Diazinon August, 1988
-14-
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. 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 MAs since toxicological end points and numbers of
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
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) a 0.00315 mg/L (3 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 (0.6 ug/L)
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Diazinon August, 1988
-15-
where:
0.00315 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Two studies on the carcinogenlcity of diazinon in mice have been
reported (NCI, 1979; Barnett and Rung, 1980). Neither study revealed
any evidence of carcinogenicity.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of diazinon.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), diazinon may be classified in
Group E: evidence of non-carcinogenlcity for humans. This category
is for substances that show no evidence of carcinogenicity in at
least two adequate animal tests or in both epidemiclogic and animal
studies.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The NAS (1977) has calculated an ADI of 0.002 rag/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
9 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, 1988). 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.
This method has been validated in a single laboratory, and estimated
detection limits have been determined for the analytes in the method,
including diazinon. The estimated detection limit is 0.25 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
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Diazinon August, 1983
-16-
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/ft^/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.
GAC is effective for diazinon removal. Dennis and Kobylinski (1983)
and tennis 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.
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.
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 al., 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, 1988
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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.
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August, 1988
DICAMBA
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.
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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Dicamba August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-00-9
Structural Formula
2-Methoxy-3,6-dichlorobenzoic Acid
Synonyms
0 Banes, Banex, Banlen, Banvel 0, Banvel, Brush buster, Dianat, Oianate,
Dicambe, Mediben, Mondak, MOBA, 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 (Berg, 1986).
Properties (Berg, 1986; CHEMLAB, 1985; Windholz et al. , 1983; Worthing,
1983; WSSA, 1983)
Chemical Formula
Molecular Weight 221.04
Physical State (at 25°C) Crystals
Boiling Point
Melting Point 114 to 116°C
Density
Vapor Pressure (20°C) 3.41 x 10~5 mm Hg
Specific Gravity
Water Solubility (20°C) 4,500 mg/L at 25°C
Log Octanol/Water Partition 3.67 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Dicamba has been found in 262 of 806 surface water samples analyzed
and in 2 of 230 ground water samples (STORET, 1988). Samples were
collected at 151 surface water locations and 192 ground water locations;
dicamba was found in 9 states. The 85th percentile of all non-zero
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Dicamba August/ 1988
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samples was 0.15 ug/L in surface water and 0.07 ug/L in ground water.
The maximum concentration found in surface water was 3.3 ug/L, while
in ground water it was 0.07 ug/L. This information is provided to
give a general impression of the occurrence of this chemical in ground
and surface waters as reported in the SIORET database. The individual
data points retrieved were used as they came from STORET and have not
been confirmed as to their validity. STORET data is often not valid
when individual numbers are used out of the context of the entire
sampling regime, as they are here. Therefore/ this information can
only be used to form an impression of the intensity and location of
sampling for a particular chemical.
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).
8 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).
9 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/acre (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 lb/A/ respectively
(Salman et al., 1972).
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Dicamba August/ 1988
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III. PHARMACOKINETICS
Absorption
0 Atallah and Yu (1980) reported that mice, rats, rabbits and dogs
administered single oral doses of 14C-dicamba (99% purity, approxi-
mately 100 mg/kg) excreted an average of 85% of the administered
dose in urine, as the parent compound, in 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
unchanged 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 (Atallah and Yu, 1980). They found that tissue
levels were low and that dicamba did not accumulate in mammalian
tissues.
0 Tye and Engel (1967) also found low levels of dicamba in kidneys, liver
and blood. These data also indicate that dicamba does not accumulate.
Metabolism
0 The metabolism of Hc-dicamba (99% purity) was investigated in mice,
rats, rabbits and dogs after administration of approximately 100 mg/kg
per os (Atallah and Yu, 1980). Between 90 to 99% of the dicamba was
recovered unchanged in the urine of all four species. 3,6-Dichloro-
2-nydroxybenzoic acid (DCHBA, a metabolite) was not detected in any
urine sample at a level greater than 1% of the dose. There was also
a small amount of unknown metabolites totaling about 1%.
Excretion
Atallah and Yu (1980) investigated the excretion of Hc-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 administered
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 9%) 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.
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Dicamba August/ 1988
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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•
Animals
Short-term Exposure
0 Reported acute oral LDjQ values for technical dicamba [85.8% active
ingredient (ai)] range from 757 to 1,414 mg/kg (Witherup et al.,
1962) in rats. The acute oral 1*050 in Rice has been reported to be
>4,640 mg/kg (Witherup et al., 1962) and 316 mg/kg in hens (Roberts
et al., 1983).
0 An acute inhalation LC50 of >200 mg/L was reported in Spartan strain
rats (Wazeter et al., 1973).
0 The neurotoxic effects of dicamba in hens were studied by Roberts
et al. (1983). Technical dicamba (86.82%ai) 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 (inability to stand) 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 Charles River CD strain were fed diets
containing 658 to 23,500 ppm of technical dicamba (85.8% ai) 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 to 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 Wazeter et al. (1974) reported an acute LD50 for technical Banvel
(85.8% ai) of >2000 mg/kg in dermal studies on New Zealand White
rabbits (administration vehicle not stated).
0 Heenehan et al. (1978) studied the sensitization potential of technical
dicamba (86.8% ai) in Hartley 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 judged to cause
moderate dermal sensitization.
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Dicamba August/ 1988
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0 Technical dicaraba (86.8% ai) 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 (Dean et al., 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.
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
8 Laveglia et al. (1981) fed CD rats (20/sex/dose) technical dicamba
(86.8% ai) 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. However, there
was 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 decrease in
body weight was lower (p <0.05) at week 13 when compared to controls.
At this dose, there was also a decrease (p <0.05) in the wet weight
of the kidney. In addition, an increase (p <0.01) in liver weight,
when expressed as % of body weight, was reported. 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 (corresponding to doses of
0, 3.8, 12, 37.3, 119 or 364 mg/kg/day) for 15 weeks (Edson and Sand-
erson, 1965). Following treatment, general behavior, physical appear-
ance, 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 (37 mg/kg/day) or
less. Relative liver-to-body weight ratios increased (p value not
specified) at 1,000 and 3,162 ppm (119 and 364 mg/kg/day). Based on
these data, the authors identified a NOAEL of 316 ppm (37 mg/kg/day).
0 Davis et al. (1962) fed beagle dogs (three/sex/dose) technical dicamba
(90% ai) 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. Although
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 and histopathology was done only
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Dicamba August/ 1988
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on the heart, lung, liver and kidney. Based on this marginal infor-
mation, a NOAEL or LOAEL will not be identified.
0 Sprague-Dawley rats (32/sex/dose) were fed technical dicamba (90% ai)
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
reported no adverse effects upon survival, body weight, food consump-
tion, organ weight, hematologic values or histology at the dose
levels tested. Since no data were presented for evaluation of pharma-
cologic effects, gross pathology, urinalysis or clinical chemistry
and incomplete histological data were presented, a NOAEL could not be
determined for this study due to insufficient data.
0 Dicamba (Technical Reference Standard, 86.8% ai) was administered for
1 year to male and female beagle dogs (4/sex/dose) in their diet at
mean dosage levels of 0, 2, 11 or 52 mg/kg/day (Blair, 1986). There
were no chemical-related changes in behavior, food consumption, body
weight, clinical hematology, biochemistry or urinalysis. In addition,
no compound-specific effects on gross or microscopic pathology were
observed. Based upon these data, a NOAEL of 52 mg/kg/day can be
identified.
Reproductive Effects
0 Charles River CD rats (20 females or 10 males/dose) were fed diets
containing technical dicamba (87.2% ai) in doses of 0, 5, 50, 100,
250 or 500 ppm through three generations (Witherup et al., 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% ai) was administered per os to pregnant New
Zealand White rabbits (23-27/dose) at doses of 0, 0.5, 1, 3, 10 or
20 mg/kg/day from days 6 through 18 of gestation (Wazeter et al./
1977). 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 significant and
identified a developmental toxicity NOAEL of 10 mg/kg/day. However/
the EPA/OPP identified a maternal and fetotoxic NOAEL of 3 mg/kg/day,
based on a reduction in body weights and increased post-implantation
losses at the highest dose.
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 throught 19 of gestation (Smith et al., 1981). No maternal
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Dicamba August, 1988
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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 fetotoxicity
or developmental effects were observed at the dose levels tested.
Based on these findings, a NOAEL for- maternal toxicity of 160 mgAg/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 typhimurium
(TA 98, TA 100, TA 1535, TA 1537 and TA 1538) or Escherichia coli
(WP2 her} either with or without metabolic activation.
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 and the data was not
presented in the English summary. Accordingly, the significance of
this report is unknown.
Carcinogenicity
0 Sprague-Dawley rats (32/sex/dose) were administered dicamba (90% ai)
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/Lj
(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, 1,000 or 10,000),
in accordance with EPA and NAS/ODW guidelines.
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Dicamba August/ 1988
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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 Wazeter et al. (1977) 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 % ai) by gastric
intubation at dosage levels of 0, 0.5, 1, 3, 10 or 20 mg/kg/day from days 6
through 18 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/kg/day dose group. Based
on these data, a maternal and fetal toxicity NOAEL of 3 mg/kg/day is identified.
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 37 mg/kg/day, respectively) that were higher than the
NOAEL (3 mg/kg/day) of the rabbit study (Wazeter et al./ 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 (0.03 mg/kg/day)
be used at this time as the basis for the Longer-term HA. As a result, the
Longer-term HA for the 10-kg child is 0.3 mg/L (300 ug/L) and for the 70-kg
adult is 1.0 mg/L (1,000 ug/L).
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Dicamba August, 1988
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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. 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.
Although the 1-year dog study by Blair (1986) identifies a NOAEL of
52 mg/kg/day, this NOAEL will not be used to calculate the RfD. This decision
is based on the facts that the IRDC study did not address reproductive and
developmental toxicity and that the study of Wazeter et al. (1977) identifies,
according to EPA/OPP, a maternal and fetotoxic NOAEL of 3 mg/kg/day. Since
the reproductive and developmental effects appear to be seen at the lower
exposure levels and to protect against potential fetotoxic effects, the RfD
will be calculated using the NOAEL of 3 mg/kg/day.
The Lifetime HA is derived from this NOAEL as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = 3 mg/kg/day = Q.03 mg/kg/day
(100)
where:
3 mg/kg/day = NOAEL based on the absence of fetotoxic effects.
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)
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Dicamba August, 1988
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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 = (1.05 mg/L) (20%) =0.2 mg/L (200 ug/L)
where:
1.05 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 One study on the carcinogenicity of dicamba in rats has been reported
(Davis et al., 1962). Although it revealed no evidence of carcinogen-
icity, it is limited in scope.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenicity of dicamba.
0 Applying the criteria described in EPA's 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, 1985).
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, 1988). In this method, approximately 1 liter of
sample is acidified and the compounds are extracted with ethyl ether
using a separatory funnel. The derivatives are hydrolysed with
potassium hydroxide and extraneous organic material is removed by a
solvent wash. After acidification, the acids are extracted and con-
verted to their methyl esters using diazomethane as the derivatizing
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Dicamba August, 1988
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agent. Excess reagent is removed and the esters are determined by
electron capture GC. This method has been validated in a single
laboratory and estimated detection limits have been determined for
the analytes in this method. The estimated detection limit for
dicamba is 0.081 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate granular-activated carbon (GAG) 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, 1988
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IX. REFERENCES
Atallah, Y.H., and C.C. Yu.* 1980. Comparative pharmacokinetics and
metabolism of dicamba in mice, rats, rabbits and dogs. Unpublished
Study. 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.
Blair, M.* 1986. Dicamba - One year dietary toxicity study in dogs. IRDC
report no. 163-696 (Dated December 19, 1986). Unpublished study.
EPA accession number 55947-3.
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. Jolley, K.L. Stemmer et al.* 1962. The feeding for two
years of the herbicide 2-methoxy-3,6-dichlorobenzoic acid to rats and
dogs. Unpublished study. MRID 00028248.
Dean, W.P., E.I. Goldenthal, D.C. Jessup et al.* 1979. Three-week dermal
toxicity study in rabbits. IRDC No. 163-620. MRID 00128090.
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. Braun.* 1978. A dermal sensitization
study in guinea pigs. Compound: Banvel 45, Banvel technical: Project
No. 5055-78. Unpublished study. 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.
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Dicamba August/ 1988
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Kurinnyi, A.I., M.A. Pilinskaya, I.V. German and T.S. L'vova. 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 and L. Brewer.** 1981. Thirteen-week dietary
toxicity study in rats with dicamba. IRDC Mo. 163-671. unpublished
Study. MRID 00128093.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs
and cosmetics. Assoc. Food Drug Off. U.S.
Moriya, M., T. Ohta, K. Watanabe, T. Miyazawa, K. Kato and Y. Shirasu. 1983.
Further mutagenicity studies on pesticides in bacterial reversion assay
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NAS. 1977. National Academy of Sciences. Drinking Hater and Health. Vol. 1.
Washington, DC: National Academy of Science Press.
Roberts, N., C. Fairley, C. Fish et al.* 1983. The acute oral toxicity
(LD^Q) and neurotoxic effects of dicamba in the domestic hen. HRC Report
No. 24/8355. Unpublished study. 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
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soils of Texas. Weed Sci. 21:393-396.
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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.
Smith, A.E. 1973a. Degradation of dicamba in prairie soils. Weed Res.
13:373-378.
Smith, A.E. 1973b. Transformation of dicamba in Regina heavy clay.
J. Agric. Food Chem. 21:708-710.
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products 3,6-dichlorosalicylic acid in prairie soils. J. Agric. Food Chem.
22:601-605.
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Dicamba August/ 1988
-15-
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Dicamba August/ 1988
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•Confidential Business Information submitted to the Office of Pesticide
Programs•
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August, 1988
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.
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 prob-it
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 r 3-Dichloropropene August, 1 988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 542-75-6
Structural Formula
C1CH2 H C1CH2 Cl
^ I X /
c = c c = c
/ \ / \
H Cl H H
(trans) (cis)
1 , 3-Dichloropropene
(approximately 46% trans/42% cis)
Synonyms
e Dichloro-1,3-propene; 1,3-dichloro-1-^propene; cis/trans-1, 3-dichloro-
propene; 1,3-D; DCP; D-D (approximately 28% cis/27% trans)
Uses
0 DCP is the active ingredient in Telone*, a registered trademark of
the Dow Chemical Company.
0 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; Clayton and Clayton, 1981)
Chemical Formula
Molecular 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 Octano I/Water Partition 25
Coefficient
Conversion Factor (25°C)(air) 1 mg/L = 220 ppm; 1 ppm = 4.54 mg/m3
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1,3-Dichloropropene August, 1988
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Occurrence
0 In California (Maddy et al., 1982), 54 wells 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).
0 Monitoring data from New York have shown positive results for DCP in
ground water (U.S. EPA, 1986b).
0 In deep well sampling in southern California (65 to 1,200 foot depths),
no DCP was detected. In shallow wells (3 to 4 meters) around potato
fields in Suffolk County, NY, DCP was detected up to 138 days after
application (OPP, 1988).
0 DCP has been found in 41 of 1,088 surface water samples analyzed and
in 10 of 3,949 ground water samples (STORET, 1988). Samples were
collected in 800 surface water locations and 2,506 ground water
locations; DCP was found in 13 states. The range of concentrations
found in ground water was 0.2 ug/L to 90 ug/L. The 85th percentile
of all non-zero samples was 1.3 ug/L in surface water and 3.4 ug/L
in ground water. This information is provided to give a general
impression of the occurrence of this chemical in ground and surface
waters as reported in the STORET database. The individual data
points retrieved were used as they came from STORET and have not been
confirmed as to their validity. STORET data is often not valid when
individual numbers are used out of the context of the entire sampling
regime, as they are here. Therefore, this information can only be
used to form an impression of the intensity and location of sampling
for a particular chemical.
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.
0 DCP hydrolyses as a function of temperature not as a function of pH.
At 10»C, the half-life is 51 days while at 20°C it is 10 to 13 days.
Chloroallyl alcohol is the main hydrolytic degradate. Some phololysis
of DCP does occur (OPP, 1988).
0 In laboratory aerobic soil metabolism studies, DCP degrades to
chloroallyl alcohol in 20 to 30 days where soil pH is between 5.0 and
7.0, the temperature is between 15 and 20°C and the organic matter
content is from 1.5 to 11.6 percent in sandy loam or clay soils. In
anaerobic soil metabolism studies, DCP degrades to chloroallyl alcohol
to less than 8 percent in 30 days. For anaerobic aquatic metabolism
studies, the half-life was reported to be about 20 days at pHs of
6.9 to 7.5 (OPP, 1988).
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1,3-Dichloropropene August, 1988
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Zn a field dissipation study done in the Netherlands, DCP (220-250 Ib/A)
injected into the soil at 9 to 19 cm depths was found to move rapidly
downward over a 2 week period. In a similar study in Delano, CA, DCP
was injected at 1,310 1/ha to 1,638 1/ha (=lb/A) to 81 cm. Samples
at 14 days noted the presence of DCP (up to 0.5 ppm) at all depths to
8 feet (OPP, 1988).
III. PHARMACOKINETICS
Absorption
0 Toxicity studies indicate that DCP is absorbed from skin, respiratory
and gastrointestinal systems (Clayton and Clayton, 1981).
0 Oral administration of DCP in rats resulted in approximately 90%
absorption of the administered dose (Hutson et al., 1971).
Distribution
0 Radiolabeled 14C D-D (55% DCP) was administered orally in arachis
oil in rats. After 4 days, most of the administered dose, based on
measured radioactivity, was recovered primarily 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-chloropropr2-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 derivitive (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.
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.
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1,3-Dichloropropene August, 1988
-5-
IV. HEALTH EFFECTS
Humans
0 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).
0 Inhalation of DCP at high vapor concentrations resulted in gasping,
refusal to breathe, coughing, substernal pain and extreme respiratory
distress at vapor concentrations over 1,500 ppm (Gosselin et al.,
1976).
0 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 and >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 administered by
gavage as a 10% solution in corn oil to rats. The oral I^gs in male
and female rats were 713 and 740 mg/kg, respectively (Torkelson and
Oyen, 1977). In another study, the oral 1*059 ^n tne mouse for both
males and females was 640 mg/kg (Toyoshima et al., 1978).
Dermal/Ocular Effects
0 The percutaneous LD50s for male and female mice dosed with DCP 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).
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 mg/kg (Dow, 1978).
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.
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1,3-Oichloropropene August, 1988
-6-
The only effect noted was slight cloudy swelling of renal tubular
epithelium 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, or 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 or 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 and 30 mg/kg/day doses, the relative weight of the
kidney of males was higher than controls. The authors conclude that
the no-toxic-effect level for DCP was between 3 and 10 mg/kg/day.
The actual No-Observed-Adverse-Effect-Level (NOAEL) was 3 mg/kg/day
(Til et al, 1973). This is the only study that can be used to develop
a reference dose. However, because the design does not ideally
address drinking water, a modifying factor will be used.
0 The National Toxicology Program (NTP, 1985) evaluated the chronic
toxicity and carcinogenicity of Telone II in rat's 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 rag/kg) and 50 male and female B6C3Fj mice (doses 0, 50
or 100 mg/kg) were gavaged with Telone II in corn oil, 3 days per
week up to 104 weeks. Ancillary 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
inadequate due to the deaths of vehicle control animals. Many
chronic toxicity parameters (hematology/ clinical chemistry) were not
determined. The DCP used in the NTP study had a different stabilizer
from the current Telone II.
0 Scott et al. (1987) exposed groups of male and female B6C3F1 mice
(70 animals/sex/exposure concentration) to vapors of Telone II* soil
fumigant for 6 hours/day, 5 days/week for up to 24 months at 0, 5,
20 or 60 ppm. Urinary bladder effects including hyperplasia of
bladder epithelium were noted in both sexes at 20 and 60 ppm. Hyper-
trophy and hyperplasia of the nasal respiratory mucosa were observed
in most 60 ppm exposed mice of both sexes and in 20 ppm exposed
females. Hyperplasia of the epithelial lining of the nonglandular
portion of the stomach was observed in 60 ppm exposed males.
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1,3-Dichloropropene August, 1988
-7-
0 Lomax et al. (1987) exposed groups of 70 male and female Fischer 344
rats to vapors of Telone II* soil fumigant for 6 hours/day, 5 days/week
for up to 24 months at targeted concentrations of 0, 5, 20 or 60 ppm.
The NOAEL was 20 ppm. The highest dose caused histopathological
changes in nasal tissue as well as a decrease in body weight gain
during the first year of this study. Males and females exposed to
60 ppm showed decreased thickness and erosions of the nasal epithelium
as well as minimal submucosa fibrosis.
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 or 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., 1980).
0 Breslin et al. (1987) exposed by inhalation groups (F0) of 30 males
and 40 females for 10 weeks, 6 hours/day, 5 days/week to Telone II*
at concentrations of 0, 10, 30 and 90 ppm prior to breeding. Exposure
was increased to 7 days/week during breeding at weeks 11 to 13.
Exposure of FI male and female parents to Telone II* began after
weaning (approximately week 32 of the study) and continued for
12 weeks (5 days/week and 6 hours/day). The NOAEL for reproductive
effects in the study was i.90 ppm, the highest dose tested. Conception
indices of females were somewhat reduced in the FI and ?2 generations.
At 90 ppm, both males and females developed hyperplasia of respiratory
epithelium and focal degeneration of olefactory tissue. Decreased
body weight was observed in males and females exposed to 90 ppm.
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
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1,3-Dichloropropene August, 1988
-8-
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 OCP. 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 TA10Q 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 since studies
by DeLorenzo et al. (1975) reported positive results in TA1978 (with
and without metabolic activation) for a commercial mixture of DCP and
a mixture of pure cis- and trans-DCP.
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., 1978).
0 Sudo et al. (1978) tested DCP in a reverse mutation assay with
£. coli B/r Wp2 with negative results.
0 DCP was negative for reverse mutation in the mouse host-mediated test
with _S_. typhimurium G46 in studies by Shirasu et al. (1976) and Sudo
et al. (1978).
Carcinogenicity
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).
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1,3-Dichloropropene August, 1988
-9-
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). Incidences of neoplasms
were not significantly increased in male mice (NTP, 1985).
0 Thirty female Ha:ICR Swiss mice received weekly subcutaneous injections
of cis-DCP. The dose was 3 mg OOP/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 Ouuren, 1979).
0 Scott et al. (1987) exposed groups of male and female B6C3F1 mice (70
animals/sex/dose) to vapors of Telone II* for 6 hours/day, 5 days/week
for up to 24 months at 0, 5, 20 or 60 ppm. The only tumorigenic
effect was an increased incidence in benign lung tumors (bronchlo-
loalveolar adenomas) in the 60 ppm exposed males. There were no
tumorigenic effects in the lower-dose males or at any of the doses in
females.
0 Lomax et al. (1987) exposed Fisher 344 rats (70 rats/sex/dose) to
vapors of Telone II* (0, 5, 20 and 60 ppm). The two year exposure by
inhalation did not result in increases in tumor incidence.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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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1,3-Dichloropropene August, 1988
-10-
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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,
calculated below) be used as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
The Til et al. (1973) 13 weeks subchronic gavage study in rats has been
selected to serve as the basis for calculating the 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 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/kcf/day) (10 kg) = 0.03 /L (30 u /L)
(100) (10) (1 L/day)
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 EPA or
NAS/OOH guidelines for use with a NOAEL from an animal
Study.
10 = modifying factor, selected since this was the only useful
gavage study available and classified as supplementary
data. Also there were considerable toxicological data gaps.
1 L/day = assumed daily water consumption of a child.
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1f3-Dichloropropene August, 1988
-11-
For a 70-kg adult:
Longer-term HA = (3.0 mg/kg/day) (70 kg) = 0.105 mg/L (100 ug/L)
(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 EPA or
NAS/ODW guidelines for use with a NOAEL from an animal
study.
10 = modifying factor, selected since this was the only useful
gavage study available and classified as supplementary
data. Also there were considerable toxicological data gaps.
2 L/day = assumed daily water consumption of an adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three-step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). Prom the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed dally 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. 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 DWEL.
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:
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1,3-Dichloropropene August, 1988
-12-
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 EPA or
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
gavage study available and classified as supplementary
data. Also there were considerable toxicological data gaps.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-0003 mg/kg/day) (70 kg) = .011 /L (10 /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 B2, probable human carcinogen. The estimated cancer risk associated
with lifetime exposure to drinking water containing DCP at 10 ug/L is
approximately 5.0 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.
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1,3-Dichloropropene August, 1988
-13-
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 qj*
value (NCI, 1985) consisted of the 2-year study data excluding the
ancillary studies data. The ancilllary 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/m^) as a Threshold Limit Value for
DCP (Clayton and Clayton, 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.
While an estimated detection limit has not been- calculated for the
two isomers of 1,3-dichloropropene, work done with 1,1-dichloropropene
would indicate a range for 1,3-DCP of 0.02 to 0.05 ug/L.
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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1r3-Dichloropropene August, 1988
-14-
IX. REFERENCES
Breslin, W.J., H.D. Kirk, C.M. Streeter, J.F. Quast and J.R. Szabo.* 1987.
Telone II soil fumigant: two-generation inhalation reproduction study
in Fischer 344 rats. Prepared by Health and Environmental Sciences USA.
Submitted by Dow Chemical Company, Midland, MI. MRID 403124.
Brooks, T.M., B.J. Dean and A.S. Wright.* 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. MRID 61059.
Clark, D., D. Blair and S. Cassidy.* 1980. A 10 week .inhalation study of
mating behavior, fertility and toxlcity 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.
Clayton, G.D. and F.E. Clayton. 1981. Patty's Industrial hygiene and toxi-
cology. 3rd ed., New York, NY: John Wiley and Sons, Inc. Vol. 2B,
pp. 3573-3577.
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. 90-Day 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., Midland, MI. MRID 115213.
-------
3-Dichloropropene August, 1988
-15-
Flessel, P.* 1977. Letter dated Apr. 8, 1977: Subject: Mutagen testing
program, mutagenlc activity of Telone II in the Ames Salmonella assay*
Prepared by Calif. Oept. 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, MO: The Williams
and Wilkins Co., p. 120.
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. Fundam. Appl. Toxicol.
8:562-570.
Hutson, O.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.
Lomax, L.G., L.L. Calhoun, W.T. Stott and L.E. Frauson.* 1987. Telone II*
soil fumigant: 2-year inhalation chronic toxicity-oncogenicity study in
rats. Prepared by Health and Environmental Sciences, USA. Submitted
by Dow Chemical Company, Midland, MI. MRID 403122.
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.
OPP. 1988. Office of Pesticide Programs, Exposure Assessment Branch, USEPA,
Washington, DC.
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. Hine. 1980. Mutagenicity of 2- and 3-carbon halo-
genated compounds in Salmonella/mammalian microsome test. Environmental
Mutagenesis. 2:59-66.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
-------
3-Dichloropropene August, 1988
-16-
Stott, W.T., K.A. Johnson, L.L. Calhoun, S.K. Weiss and L.E. Frauson.* 1987.
Telone II* soil fumigant: 2-year inhalation chronic toxicity-oncogenicity
study in mice. Prepared by Health and Environmental Sciences, USA. Sub-
mitted by Dow Chemical Company, Midland, MI. MRID 403123.
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. Spanjers, V. J. Feron and P.J. Reuzel. 1973.* Sub-chronic
(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 as
determined by repeated exposure of laboratory animals. American Industrial
Hygiene Association Journal. 38:217-223. Dow Chemical U.S.A., Midland, MI.
MRID 39686.
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.
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1,3-Dichloropropene August, 1988
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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. Dimick.* 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/
HI: MRID 117052.
*Confidential Business Information submitted to the Office of Pesticide
Programs.
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August/ 1988
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.
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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Dieldrin
August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 60-57-1
Structural Formula
Dieldrin; 3,4,5,6,9,9-hexachloro-1a,2,2a,3,6,6a,7,7a-octahydro-
2,7:3,6-dimethanonaphth[2,3-b]oxirene (Windholz, 1983).
Synonyms
0 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 (Neister, 1983).
Properties (NAS, 1977; Weast and Astle, 1982; Windholz, 1983)
Chemical Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor pressure (20°C)
Water Solubility (258C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold (water)
Conversion Factor
C12H8C160
380.93
Crystals
175 to 176«C
3. 1 x 10-6 mm Hg
0.25 mg/L
0.04 mg/L
Occurrence
Oieldrin has been found in 7,320 of 50,473 surface water samples
analyzed and in 223 of 5,443 ground water samples (STORET, 1988).
Samples were collected at 9,021 surface water locations and 4,131
ground water locations, and Oieldrin 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. This information is provided to give a
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Dieldrin August, 1988
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general impression of the occurrence of this chemical in ground and
surface waters as reported in the STORET database. The individual
data points retrieved were used as they came from STORET and have
not been confirmed as to their validity. STORET data is often not
valid when individual numbers are used out of the context of the
entire sampling regime, as they are here. Therefore, this information
can only be used to form an impression of the intensity and location
of sampling for a particular chemical.
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) (MacXay 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 (Hayes, 1974); hence, it was absorbed.
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.
The 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 the concentration of dieldrin in the blood
increased as fat was mobilized (Heath and Vandekar, 1964).
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Dieldrin August, 1988
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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).
Animals
Short-term Exposure
8 RTECS (1985) reported the acute oral LD$Q 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/kg, respectively.
Dermal/Ocular Effects
0 Aldrin or Dieldrin (dry powder) applied to rabbit skin for 2 h/day,
5 days/week for 10 weeks had no discernible effects (IPCS, 1987).
Long-term Exposure
8 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. Male 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 Mongrel 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
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Dieldrin August, 1988
-5-
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 convulsions.
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).
0 Groups of beagle dogs (five/sex/dose) were treated daily by capsule
with dieldrin (>99% purity) at 0.0, 0.005 or 0.05 mg/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 phosphatase activity 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 Oieldrin (>99% pure) was administered to CF1 mice of both sexes
(30/sex/dose) 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 (18-20/dose) 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 (2.99% 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
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Dieldrin August, 1988
-6-
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
no effect was seen on fetal survival or weight.
0 Dieldrin (87% pure) was not found to be teratogenic in the CD rats and
CD-1 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-rag/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
receiving 6.0 mg/kg/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
mg/kg/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).
9 Dieldrin significantly decreased the mitotic index and increased
chromosome abnormalities in STS mice bone marrow cells in an in 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
0 A dose-related increase in the incidence of hepatocellular carcinomas
was observed in B6C3F-| mice, with the incidence in the high-dose
males being significantly (p = 0.025) 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 mg/kg/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 mg/kg/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).
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Dieldrin August, 1988
-7-
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day/ ten-day,
longer-term (up to 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) „ m /L ( u /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effe.ct 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, 1,000 or 10,000),
in accordance with EPA or 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 DWEL for a 10-kg child (0.0005 mg/L) 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) 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) 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
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Dieldrin August/ 1988
-8-
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
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. 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 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 = 0.00005 mg/kg/day
100
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 EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
Step 2: Determination of the Drinking Water Equivalent (DWEL)
DWEL = (0.00005 mg/kg/day)(70 kg) = 0.00175 mg/L (2 ug/L)
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.
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Dieldrin August, 1988
-9-
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.
Evaluation of Carcinogenic Potential
o Applying the criteria described in EPA's final guidelines for
assessment of carcinogenic risk (U.S. EPA, 1986), dieldrin may be
classified in Group B2: probable human carcinogen.
o 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, q.j* = 16 (mg/kg/day)~1, was estimated as the potency for
the general population (U.S. EPA, 1987).
o Using this q-| * 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.
o 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, lo'git, 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.
o IARC (1982) concluded that there is limited evidence that dieldrin is
carcinogenic in laboratory animals.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
o ACGIH (1984) has established a short-term exposure limit (STEL) of
0.75 mg/m^ and an 8-hour Threshold Limit Value (TLV)-TWA exposure
0.25 mg/m3 for dieldrin.
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Dieldrin August/ 1988
-10-
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
(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).
Residue tolerances ranging from 0.02 to 0.1 ppm have been established
for dieldrin in or on agricultural commodities (U.S. EPA, 1985).
WHO
(1982) established guidance of 0.03 ug dieldrin/L in drinking water.
VII. ANALYTICAL METHODS
0 Determination of dieldrin is by Method 508, Determination of Chlori-
nated Pesticides in Water by Gas Chromatography with an Electron
Capture Detector (U.S. EPA/ 1988). In this procedure/ a measured
volume of sample of approximately 1 liter is solvent extracted with
methylene chloride by mechanical shaking in a separatory funnel or
mechanical tumbling in a bottle. The methylene chloride extract is
isolated/ dried and concentrated to a volume of 5 mL after solvent
substitution with methyl tert-butyl ether (MTBE). Chromatographic
conditions are described which permit the separation and measurement
of the analytes in the extract by GC with an electron capture detector
(ECD). This method has been validated in a single laboratory/ and
estimated detection limits have been determined for the analytes in
the method/ including dieldrin/ that the estimated detection limit
is 0.02 ug/L.
VIII. TREATMENT TECHNOLOGIES
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+%.
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.
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).
Pirbazari and Weber (1983) determined adsorption isotherms using GAC
on dieldrin in water solutions. Resin adsorption was also found to
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Dieldrin August/ 1988
-11-
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.
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, 1988
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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. 139.
Baldwin, M.K. and J. Robinson. 1972. A comparison of the metabolism of
HEOO (Dieldrin} in CFj 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
Organization.
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, 1988
-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: Heister
Publishing Company.
NAS. 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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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(185):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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 508. Determi-
nation of chlorinated pesticides in water by gas chromatography with an
electron capture detector. Available from EPA's Environmental Monitoring
and Support Laboratory, Cincinnati, Ohio.
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Dieldrin August, 1988
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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.
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.
Uhedited final draft.
Windholz, M. 1983. The Merck index. 10th ed. Rahway, NJ: Merck and Co., Inc.
pp. 450-451.
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August, 1988
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.
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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Dimethrin August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
Structural Formula
2,4-Dimethylbenzyl-2,2-dimethyl-3 (2-methylpropenyl) -cyclopropane carboxylate
Synonyms
0 EOT 21,170; Ghrysanthemumic acid; 2,4-Dimethylbenzylester.
Uses
c Insecticide for use in ponds and swamps as a mosquito larvicide
(Meister, 1986).
Properties
Chemical Formula C18H24°2
Molecular Weight 286.39 (Ambrose, 1964)
Physical State (25°C) Amber liquid
Boiling Point 175°C
Melting Point —
Density —
Vapor Pressure (25°C) —
Specific Gravity 0.98
Water Solubility (25°C) Insoluble (further details not provided)
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, 1988
-3-
III. PHRRMROOiQNBnCS
Absorption
0 In a preliminary metabolic study by Ambrose (1964), four rabbits were
given 5 mL/kg (5 mg/kg) 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 chrysanthemumic acid and the
glucuronic ester of 2,4-dimethyl benaoic 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
Humans
No information on the health effects of dimethrin in humans was found
in the available literature.
Animals
Short-term Exposure
• *Rie acute oral LDgg 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 observed 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, 1988
5 days per week for 3 weeks. This corresponds to an average daily
dose of 7 rag/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
" 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
" 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, 1988
-5-
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/kg/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 NQAEL of 300 rag/kg was identified from this study.
0 As described in a review by Cohen and Grasso (1981), dimethrin has
been implicated as a hypolipidemic agent and causes an increase in
hepatic peroxisone proliferation . Dietary administration of certain
hypolipidemic agents to rodents has resulted in the induction of
liver carcinomas.
Reproductive Effects
9 No information on the reproductive effects of dimethrin was found in
the available literature.
Developmental Effects
0 No information on the 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.
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Dimethrin August, 1988
Carcinogenicity
0 No information on the carcinogenicity 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 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 = (roAEL or LOAEL) x (BW) = _ ,^/L ( _ ug/L)
(UF) x ( L/day) *' *'
where:
NOAEL or TnATT. = 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, 1,000 or 10,000),
in accordance with EPA or 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 (10 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 values for dimethrin. It is therefore
recommended that the Longer-term HA for a 10-kg child (10 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,
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Dimethrin August, 1988
-7-
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.
Kidney-to-body weight ratios were also significantly higher in those groups.
Histqpathological examinations revealed dose-related morphological changes in
the liver occurring at dose levels as low as 0.6%. A NDAEL of 0.2% dimethrin
(120 mg/kg/day for males; 130 mg/kg/day for females) was identified in this
study.
Using a NQAEL of 120 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (120 mg/kg/day) (10 kg) = 12 ,ng/L (io,000 ug/L)
(100) (1 L/day)
where:
120 mg/kg/day = NQAEL, 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 EPA
or NAS/ODW guidelines for use with a NQAEL from an
animal study.
1 L/day = assumed daily water consumption of a child.
Using a NQAEL of 120 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = (120 mg/kg/day) (70 kg) = 42 mg/L (40,000 ug/L)
(100) (2 L/day)
Where:
120 mg/kg/day = NQAEL, 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 EPA
or NAS/ODW guidelines for use with a NQAEL 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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Dimethrin August, 1988
-8-
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 LQAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). Fran 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 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 NQAEL 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) = 0.3 mg/kg/day
where:
300 mg/kg/day = NQAEL, 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 EPA
or NAS/ODW guidelines for use with a NQAEL 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 ^/^ (io,000 ug/L)
(2 L/day)
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Dimethrin August, 1988
-9-
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.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (10.5 mg/L) (20%) = 2.1 mg/L (2,000 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 mg/L apparently exceeds the
water solubility of dimethrin (insoluble).
Evaluation of Carcinogenic Potential
0 No information on the carcinogenicity 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 maybe classified in
Group D: not classified. This category is for substances with
inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
' 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 Dimethrinaisoa cyclopropane carboxviate pesticide which, as such, can
be analyzed by EPA Method #616 (U.S. EPA, 1984). This method covers
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Dimethrin August, 1988
-10-
Cft) pesticides such as cyclpprate and resmethrin, which are chemically
identical to dinethrin. The method is similar to other 600 series
methods in that 1 liter of sample is extrcated with methylene chloride
and reduced to 1 mL or less. Analysis is by flame-ionization gas
chromatography (FID/GC). A cleanup procedure is provided in case
interferences are noted.
* While method #616 can be used for monitoring dimethrin, the analyst
should demonstrate precision and accuracy data for this compound
before proceeding. An estimated detection limit (EDL) should also be
determined as specified by regulation (CFR, 1984). Ihe detection
limit should fall into the range of 20 to 50 ug/L.
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, 1988
-11-
IX.
Ambrose, A.M. 1964. Toxicologic studies on pyrethrin-type esters of chrysan-
themumic acid II. Chrysanthemumic acid, 2,4-dimethylbenzyl ester.
Toxicol. Appl. Pharmacol. 6:112-120.
CFR. 1984. Code of Federal Regulations. 40 CFR Part 136, Appendix B.
Federal Register Vol. 49, No. 209. October 26.
CFR. 1985. Oode 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. 1984. U.S. Environmental Protection Agency. EPA Method #616 -
Determination of CHO Compounds. Available from EPA's Environmental
Monitoring and Support Laboratory, Cincinnati, Ohio.
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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August, 1988
DINOSEB
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.
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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Dinoseb August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 88-85-7
Structural Formula
NO,
2-sec-butyl-4,6-dlnitrophenol
Synonyms
• DNBP, dinitro, dinoseb (BSI, ISO, WSSA); dlnosebe (France); Basanite
(BASF Wyandotte); Caldon, Chemox General, Chemox PE, Chemsect DNBP,
DN-289 (product discontinued), Dinitro, Dinitro-3, Dinitro General,
Dynamite (Drexel Chemical); Elgetol 318, Gebutox, He1-Fira (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 (Meister, 1984).
Uses
• Dinoseb is used as a herbicide, desiccant and dormant fruit spray
(Meister, 1984).
Properties (WSSA, 1983)
Chemical Formula CjgH^^Os
Molecular Weight 240
Physical State (room temp.) Dark amber crystals
Boiling Point
Melting Point 32«C
Density (*C) 1.2647 (45'C)
Vapor Pressure (151°C) 1 mm Hg
Specific Gravity
Water Solubility 52 mg/L (25«C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
• Dinoseb has been found in 1 of 89 surface water samples analyzed and
in none of 1,270 ground water samples (STORET, 1988). Samples were
collected at 89 surface water locations and 1,184 ground water locations'
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Dinoseb August, 1988
-3-
and dinoseb was found in Ohio. The 85th percentile of all nonzero
samples was 1 ug/L in surface water. This information is provided
to give a general impression of the occurrence of this chemical in
ground and surface waters as reported in the STORET database. The
individual data points retrieved were used as they came from STORET
and have not been confirmed as to their validity. STORET data is
often not valid when individual numbers are used out of the context
of the entire sampling regime, as they are here. Therefore, this
information can only be used to form an impression of the intensity
and location of sampling for a particular chemical.
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).
° In water with natural sunlight, dinoseb had a half-life of 14 to 18
days; with artificial light, it had a half-life of 42 to 58 days
(Dinoseb Task Force, 1985b).
0 With soil thin-layer chromatography plates, dinoseb was intermediate
to very mobile in silt loam, sand, sandy loam and silty clay loam
(Dinoseb Task Force, 1985c).
9 Soil adsorption studies gave a K^ of less than 5 for four soils: a
silt loam, sand, sandy loam and silty clay loam, with organic matter
content of 0.8 to 3% (Dinoseb Task Force, 1985d).
III. PHARMACOKINETICS
Absorption
0 Following oral administration of dinoseb to rats (Bandal and Casida,
1972) and mice (Gibson and Rao, 1973) (specific means of administration
not specified), approximately 25% of the administered dose appeared in
the feces. However, following intraperitoneal (ip) administration in
the mouse, approximately 40% appeared in the feces, thus suggesting
to Gibson and Rao (1973) that dinoseb is initially completely absorbed
following oral administration with subsequent secretion into the gut.
Distribution
0 Following oral administration of dinoseb in the mouse (specific means
of administration not specified), no appreciable amounts accumulated
in the blood, liver or kidney (Gibson and Rao, 1973).
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Dinoseb August, 1988
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Metabolism
0 while the metabolism of dinoseb has not been completely characterized,
a number of metabolites have been identified including: 2-(2-butyric
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
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 on the long-term health effects of dinoseb in humans
was found in the available literature.
Animals
Short-term Exposure
0 In rats and mice, the acute oral LDsg 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 other information on the dermal or ocular effects of dinoseb in
animals was found in the available literature.
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 males 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
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Dinoseb August, 1988
-5-
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.
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 increased liver weight
at intermediate doses, the No-Observed-Adverse-Effect Level (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 two-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).
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Dinoseb August, 1988
-6-
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).
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 were
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.
Mutagenicity
0 With the exception of an increase in DMA 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; Moriya et al., 1983).
Carcinogenicity
9 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).
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Dinoaeb August, 1988
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V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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, 1,000 or 10,000),
in accordance with EPA or 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/kg/day) (10 kg) = Ot3 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.
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Dinoseb August, 1988
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10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 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 two-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 = d-0 mg/kg/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 EPA
or 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'° mg/kg/day) (70 kg) = 0.035 mg/L (40 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 EPA
or 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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Dinoseb August/ 1988
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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
(DWZL) 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. 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 mg/kg/day
(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 2 years.
1,000 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW 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) = Of035 mg/L (40 ug/L)
(2 L/day)
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Dinoseb August, 1988
-10-
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)
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
mg/kg/day (Hazleton Labs, 1977).
8 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 inadequate
human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
• Tolerances have been established for dinoseb (CFR, 1985) at
0.1 ppm on a wide variety of agricultural commodities.
• 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
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Dinoseb August, 1988
-11-
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 aeration.
However, limited data suggest that aeration would not be effective in
the removal of dinoseb from drinking water (ESE, 1984).
0 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.
8 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, 1988
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IX. REFERENCES
Bandal, S.K. and J.E. Casida. 1972. Metabolism and photoalteration of
2-sec-butyl-4,6-dinitrophenol (oNBP 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. In: 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.
Oinoseb 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.
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Dinoseb August, 1988
-13-
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.
Froslie, A. and O. 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. Fd. Cosmet. Toxicol. 11:45-52.
Hall, L., R. Under, 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. Needs, 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 (DNBP). 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, U.K. McElroy and A. Curley. 1982.
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. Watanabe, K. Kato and Y. Shirasu. 1983. Further
mutagenicity studies on pesticides in bacterial reversion assay systems.
Mutat. 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.
-------
Oinoseb August, 1988
-14-
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.
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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
U.S. EPA. 1985. U.S. EPA Method 615 - Chlorinated Phenoxy Acids. Fed. Reg.
50:50701. October 4.
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 ed.
Champaign, IL.
Confidential Business Information submitted to the Office of Pesticide
Programs•
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August, 1988
DIPHENAMID
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.
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, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 957-51-7
Structural Formula
HC-C-N(CH,),
NjN-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).
C16H17ON
239.30
White crystalline solid
135°C
260 mg/L
Properties (Windholz et al., 1983)
Chemical Formula
Molecular Weight
Physical State (at 25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (25°C)
Specific Gravity
Water Solubility (27°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Diphenamid has not been found in the 3 surface water samples collected
at 2 locations or in 678 groundwater samples taken at 676 locations
(STORET, 1988).
Environmental Fate
0 Diphenamid is stable to hydrolysis at pH 5, 7 and 9 tor 7, 12 and
10 days, respectively, at elevated temperature (49°C or 120°F)
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Diphenamid August, 1988
-3-
(NOR-AM, 1986).
0 Diphenamid is intermediately mobile (class 3) on silt loam and silty
clay Icam soil TLC plates; or. .sandy loam, it is in class 5, indicating
that it would leach readily in this soil (Helling and Turner, 1963).
III. PHARMACOKINETICS
Absorption
0 McMahon and Sullivan, (1965) administered 50 mg/kg diphenamid-carbonyl
14C to rats by 3 different routes: oral, intraperitoneal (ip) or
subcutaneous (sc) injection. Diphenamid was well absorbed by each of
the three routes with peak blood levels being greatest for ip,
followed by sc and oral. The half lives of diphenamid in blood for
each of the three routes were comparable.
Distribution
0 No information was found in the available literature on the distri-
bution of diphenamid.
Metabolism
0 In the McMahon and Sullivan (1965) study described above, the main
route metabolism of diphenamid was reported as N-dealkylation to
nondiphenamid. The nondiphenamid is excreated as the N-glucura-
nide. p-Hytroxylation was reported as a minor rate of metabolism.
Excretion
No specific information of the excretion of diphenamid was found in
the available literature. However, in the above study, the authors
evaluated the urinary metabolites of diphenamid and reported that
diphenamid was readily absorbed and metabolized into excretable
metabolites.
IV. HEALTH EFFECTS
Humans
No information was found in the available literature on the health
effects of diphenamid in humans.
Animals
Short-term Exposure
0 RTECS (1987) (attached) reported the acute oral U^g values in the
rat, mouse, dog, monkey and rabbit to be 685, 600, 1,000, 1,000 and
1,500 mg/kg, respectively.
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Diphenamid August, 1988
-4-
DermaI/Ocular Effects
0 Weddon and Brown (1976) applied Enide SOW (a 93% wettable powder
formulation of diphencmid) to the intact or abraded skin of New
Zealand rabbits (two/sex/dose) for 24 hours at 0, 200,. 1,000 or
2,000 mg/kg. Neither skin irritation nor systemic-effects were
observed in any of the exposed animals.
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 (SGOT) 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 Charles River 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 and a LOAEL of 30 mg/kg/day
are identified by this study.
Reproductive Effects
0 In a three-generation reproduction study, Woodard et al. (1966a)
administered technical diphenamid to Charles River 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 (Fg, F1b, F2b) at any dose tested. Weanlings of the F3b
generation dosed with 30 mg/kg/day showed reversible liver changes,
including slight congestion, glycogen depletion fin-" 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. (I966a) 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.
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Diphenamid August, 1988
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Mutagenicity
0 Moriya et ai. (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_. subtilis or in reversion assays
with J^. coli or J3. typhimurium.
Carcinogenicity
0 In a 2-year feeding study in rats by Hollingsworth et al. (1966),
diphenamid was administered to Charles River 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 (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of tr-:icity.
The HAs for noncarcinogenic toxicants are derived using the folio., g formula:
HA = (NOAEL or LOAEL) x (BW) _ mg/Ii ( 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, 1,000 or 10,000),
in accordance with EPA or 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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Diphenamid August, 1988
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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'tiirie as a conservative
estimate of the One-day HA value.
For a 10-kg child, the adjusted DWEL is calculated as follows:
DWEL = (0«03 mgAg/day) (10 *g) =0.3 mg/L
(1 L/day)
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 advene, r -ninogenic health effects would not be expected to occur.
The DWEL is der ' - n the multiplication of the RfD by the assumed body
weight of an adi. .id 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
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Diphenamid August, 1988
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value of 20% is assumed. 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
(.he risks associated with lifetime exposure to this 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
nig/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.
(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 mgAg/day) = 0>03 nig/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 EPA
or 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 mgAg/day) (70 kg) = ,
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Diphenamid August, 1988
-8-
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.
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, 1988). In this method, approxi-
mately 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. This method has been
validated in a single laboratory, and the estimated detection limit
for the analytes, including diphenamid, is 0.6 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAC) adsorp-
tion will remove diphenamid from water.
0 Whi£taker (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
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Diphenamid August, 1988
-9-
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.
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, 1988
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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.
McMahon, R.E., and H.R. Sullivan. 1965. The Metabolism of the Herbicide
Diphenamid in Rats. Biochem.Pharmacol. 14:1085-1092.
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. 1987. 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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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. 1988. U.S. Environmental Protection Agency. U.S. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
water by GC/NPD, April 15, 1988 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
-------
Diphenamid August, 1988
-11-
University.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The
Merck Index, 10th ed. Rahway, N.T: Merck and Co., Inc.
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.
-------
August, 1988
DISULFOTON
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.
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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Disulfoton August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 298-04-4
Structural Formula
0, 0-Diethyl-S-[2-(ethylthio) -ethyl] , phosphorodlthioate
Synonyms
0 Disulfoton; Disyston; Disystox; Dithiodemeton; Bayer 19639; Di-syston;
Ethyl thiometon; Frumin AL; M-74 (Meister, 1983).
Uses
0 Systemic insecticide-acaricide (Meister, 1983).
Properties ( Meister , 1983; Windholz et al., 1983)
Chemical Formula
Molecular Weight 274.38
Physical State (at 25«C) Pale yellow liquid
Boiling Point 108°C (0.01 mm Hg); 132 to 133°C (1.5 mm Hg)
Melting Point
Density (20«C) 1.144
Vapor Pressure (at 20°C) 1.8 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 368 surface water samples
and none of 1,182 ground water samples analyzed (STORET, 1988).
The concentration of this surface water sample was 0.34 ug/L.
Samples were collected at 93 surface water locations and 1,080 ground
water locations. This was also the maximum concentration detected.
Disulfoton was detected only in California. This information is
provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
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Disulfoton August, 1 988
-3-
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
0 Disulf oton 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 sulf oxide and disulfoton sulfone were more
mobile in sandy loam, clay loam and silty clay loam soils than the
parent compound. Aging 32p-aisulfoton prior to elution increased the
adsorption to 10 to 20 times that of unaged 32p -disulfoton, meaning
that the degradates are probably less mobile than the parent compound.
Mobility of disulfoton in soil appears to decrease as organic matter
content and cation exchange capacity (EC) increase (Kawamori et al.,
1971a, 1971b).
0 Insecticidal disulfoton residues are mobile in a silt loam soil as
determined by a mosquito bioassay (Lichtenstein et al., 1966). Based
on soil thin-layer chromatography (TLC) plates, disulfoton has low
mobility in sand (Rf 0.18), sandy loam (Rf 0.16) and silt loam
(Rf 0.11 and 0.33), and intermediate mobility in a sandy clay loam
soil {Rf 0.39) (Thornton et al., 1976).
0 When applied to subirrigated soil columns at 20 Ib active ingredient
(a.i.) per acre, disulfoton exhibited slight upward mobility in a
Hagerstown silty clay loam and a Lakeland sandy loam soil (Mobay
Chemical Corporation, 1972). Disulfoton (6 Ib/gal EC), applied at
4 Ib ai/A to sloping field plots (1 inch/ft) of sandy loam, silt loam
and highly organic silt loam soils, was slightly mobile in runoff
water. Disulfoton concentrations measured about 1.6% of applied
amounts over a 28-day period in which 1.5 to 2.5 inches of irrigation
was applied (Flint et al., 1970).
0 Disulfoton (granular, G) dissipates rapidly in field plots of sandy
loam soil treated at 2 kg/ha (incorporated to a depth of 10 cm),
with a half -life of 1 week, and 90% loss after 5 weeks. Disulfoton
sulf oxide has a half -life of 8 to 10 weeks, while disulfoton sulfone
remains fairly stable over a 42-week period. Disulfoton sulfone was
detected at a depth of 20 cm (Suett, 1975). Disulfoton residues
dissipated with half -lives of 1 to 6 months in muck-sand, silt loam,
and clay soils treated with disulfoton 10% G or 6 Ib/gal EC at 10 ppm.
Dissipation from the upper 6 inches was enhanced by increasing amounts
of rainfall (Chemagro Corporation, 1969; Loeffler, 1969; Mobay Chemical
Corporation 1964).
0 Disulfoton is likely to be found in runoff water and sediment from
treated and cultivated fields. In monitoring studies conducted in
1973 and 1974, disulfoton was found at average concentrations of
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Disulfoton August, 1988
-4-
13.8 ppb in sediment samples taken from tailwater pits receiving
irrigation and rainfall runoff water from cultivated silt loam corn
fields. The maximum concentration in sediment samples was 32.7 ppb.
The compound was also detected in soil samples from a tailwater pit
draining silt loam corn and sorghum fields at an average concentration
of 11 ppb. Sediment samples in tailwater pits draining sorghum
fields contained disulfoton at a concentration of 117.2 ppb (Kadoum
and Mock, 1978).
III. PHARMACOKINETICS
Absorption
0 Puhl and Fredrickson (1975) administered by gavage single oral
doses of disulfoton-o-ethyl-1-14C (99% purity) to Sprague-Dawley
rats (12/sex/dose). Hales received 1.2 mg/kg and females received
0.2 mgAg. 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. Hales 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 intraperitoneally
(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
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Disulfoton August, 1988
-5-
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), 97.8% 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.
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, 1974; 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 LD5Q values ranged from 8.9 to 12.7 mg/kg
(Bombinski and Dubois, 1957; Crawford and Anderson, 1973).
0 Mihail (1978) reported acute oral LDsg values of 7.0 rag/kg and
8.2 mg/kg in male and female NMRI mice, respectively.
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Disulfoton August, 1988
-6-
0 Hixson (1982) reported that the acute oral LD$Q of disulfoton (97.8%
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 rag/kg on two occasions, 22 days apart.
The study also employed groups of 5, 10 and 5 hens for positive
controls, antidote controls and negative controls, respectively.
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.
0 Taylor (1965) reported the results of a demyelination study in which
white Leghorn hens (six/dose) were administered disulfoton in the diet
for 30 days at concentrations of 0, 2, 10 or 25 ppm, these dietary
levels are equivalent to 0, 0*1, 0.6 and 1.5 lag/kg/day. The author
indicated that no evidence of demyelination was observed in any of
the tissues examined. Based on this information, a NOAEL of 1.5
mgAg/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 LD5Q 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
• Hayes (1983) presented the results of a 23-month feeding study in
which CO-1 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
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Disulfoton August, 1988
-7-
group of animals received disulfoton 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 mgAg/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;
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 mgAg/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 mg/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
(97.91% 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
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Disulfoton August, 1988
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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)
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/kg/day) (lowest dose
tested).
Reproductive Effects
0 Taylor (1966) conducted a three-generation reproduction study in
which albino Holtzraan 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 matings. Histopathologically, F;jb
litters at 10 ppm (0.5 mg/kg/day) exhibited cloudy swelling and fatty
infiltration of the liver (both sexes), mild nephropathy in kidneys
(both sexes) and juvenile hypoplasia of the testes. No histopatho-
logical 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 Fat 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 (32 to 42%) 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
analyses, incomplete necropsy report and insufficient histopathology
data).
Developmental Effects
0 Lamb and Hixson (1983) conducted a study in which CD 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 fetuses
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Disulfoton August, 1988
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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
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 ma.ternal 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 coli strains WP 2, WP 2uvrA, CM 571,
CM 611, HP 67 and WP 12. These tests were performed without metabolic
activation; however, demeton-S-methyl sulphoxide, the major metabolite
of disulfoton, was also mutagenic in these microbial tests (U.S. EPA,
1984a).
• Simmon (1979) presented the results of an unscheduled DMA 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.
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Disulfoton August, 1988
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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-1 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
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
(97.91% 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 (up to 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, 1,000 or 10,000),
in accordance with EPA or NAS/ODW guidelines.
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Disulfoton August, 1988
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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 10-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.
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 mgAg/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 10-kg child is
calculated as follows:
Ten-day HA = (0-1 mg/kg/day) (10 kg) = 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.
100 = uncertainty factor chosen in accordance with EPA
or 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 mg/kg/day) for
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Disulfoton August, 1988
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69 weeks, then 5.0 ppm (0.125 mg/kg/day) for weeks 70 to 72, and finally
8.0 ppm (0.2 mgAg/day) from weeks 73 to 104. Exposure to 2.0 ppm (0.05
mgAg/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:
Longer-term HA = (0-025 mg/kg/day) (10 kg) = 0.0025 mg/L (3 ug/L)
(100) (1 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.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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/Jcg/day) (70 kg) = 0.0088 /L (9 /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 EPA
or 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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Disulfoton August, 1988
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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 (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. 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. 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 studies by Hayes (1985) and Carpy at 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)
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Disulfoton August, 1988
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where:
0.04 mg/kg/day = LOAEL, based on ChE inhibition and optic nerve
degeneration in rats exposed to disulfoton in the
diet for 2 years.
1,000 = uncertainty factor, chosen in accordance with EPA
or 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)
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, 1986), 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).
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Disulfoton August, 1988
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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 subchronic
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.
0 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
mgAg/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 (Method #507, U.S. EPA, 1988}. 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 has been validated in a single
laboratory and the estimated detection limit for the analytes in this
method, including disulfoton, is 0.3 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, 1988
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Hayes, R.H.* 1983. Oncogenicity study of disulfoton technical on mice.
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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 PMC Corp.,
Philadelphia, PA. CDL:099350-A. MRID 00028625.
STORET. 1988. STORE! Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.
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Disulfoton August, 1988
-19-
Taylor, R.E.* 1965. Letter sent to Chemagro Corporation dated Jan. 5, 1965:
Report on demyelination studies on hens. Report No. 15107. Uipublished
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: Oi-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.* 1984. 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. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1988. U.S. Environmental Protection Agency. U.S. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
water by GC/NPD, April 15 draft. Available from U.S. EPA's Environ-
mental Monitoring and Support Laboratory, Cincinnati, OH 45268.
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, 1988
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.
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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS NO. 330-54-1
Structural Formula
M I
0
N-C-N(CH,),
N1 -{ 3,4-Dichlorophenyl) -N,N-dimethylurea
Synonyms
• Cekiuron, Crisuror., Dailon, Diater, Di-on, Direx 4L, Ditox-800,
Diurex, Diurol, Drexel Diuron 4L, Dynex, Farmco Diuron, Karroex,
Unidron, Vonduron. Discontinued names: Sup'r Flo, Urox 'D' (Meister,
1988).
Uses
0 Herbicide. Diuron at low rates as a selective Herbicide to control
germinating broadleaf, grass weeds in numerous crops such as sugarcane,
pineapple, canebernes, alfalfa, grapes, cotton/ and peppermint.
General weed Killer at higher rates. As soil sterilant, more persistent
and preferred over monuron on lighter soil and/or in areas of heavy
rainfall. Drexel Diuron 4L, Farmco Diuron Flowable control weeds in
wheat, barley, citrus, vineyards, bananas, pineapple, cotton,
and noncrop areas (Meister, 1988).
Properties (Meister, 1988; Windholz et al., 1983)
Chemical Formula
Molecular Weight 233.10
Physical State (at 25°C) White crystalline solid
Boiling Point
Melting Point 158-159»C
Density
Vapor Pressure (5C«C) 3.1 x 10-6 mm Hg
Specific Gravity
Water Solubility (25«C) 42 mg/L
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Diuron has been found in none of the 25 surface water samples analyzed
and in none of 1,337 ground water samples (STORET, 1988). Samples were
collected at 22 surface water locations and 1,292 ground water locations.
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Diuron August, 1988
-3-
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).
Environmental Fate
0 Radiolabelqd 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 29eC and moisture levels at approximately field capacity
(Walker and Roberts, 1978; Elder, 1978). 3,4-Dichloroaniline (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,b; 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 cation exchange capacity (CEC). Soil
texture apparently is not, by itself, a major factor governing the
mobility of diuron in soil.
9 In a study using radiolabeled material, the diuron degradation products
(96% pure) had K^ values of 66 and 115 in silty clay loam soils,
indicating that they are relatively immobile or less mobile than diuron
(Elder, 1978).
9 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).
Diuron may have the potential to leach into ground water.
0 Phytotoxic residues persisted for up to 12 months in soils ranging in
texture from sand to silty clay loam to 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 Bardsley, 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
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Diuron August, 1988
-4-
and/or irrigation water. Three degradation products (DCPMU, DCPU,
and OCA) were identified in soil (planted with cotton) that had
received multiple applications of diuron (80% wettable powder totaling
5 to 5.7 lb/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
Duseja, 1973b; Bowtner 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. Diuron 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. PHARMACOKINETIC5
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 0.625 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 of 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-dichlorophenyl)-urea,
and may also have contained some 3,4-dichloroaniline. No unaltered
diuron was detected.
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Diuron August, 1988
-5-
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 ma]or
metabolite was N-(3,4-dichlorophenyl)-urea. Small amounts of
N-(3,4-dichlorophenyl)-N1 -methylurea, 3,4-dichloroanaline,
3,4-dichlorophenol and unmetabolized diuron also were detected.
Excretion
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
tne 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 3,400 mg/kg have been
reported in albino rats by Boyd and Krupa (1970), NIOSH (1987) 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 LD$Q was 3,400 mg/kg. Survivors sacrificed after
14 days showed large and dark-colored spleens with numerous foci of
blood.
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 were
noted in both the spleen and bone marrow.
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Diuron August, 1988
-6-
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 9,000 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,
400 ppm (2CJ mg/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 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 technical diuron at doses of 1, 2.5, 5 or 10
mg/kg to intact abraded skin of New Zealand White rabbits for 24 hours.
Adverse effects were not detected in exposed animals.
8 In studies conducted by DuPont
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Diuron August, 1988
-7-
0 Hodge et al. (1967) fed diuron (80% wettable powder) to groups of
Charles River rats (20/sex/dose) for 90 days at dietary levels of 0,
250 or 2,500 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,
12.5 or 125 mg/kg/day. At 2,500 ppm, both males and females ate less
and gained Less weight than did controls. There was a slight decrease
in red blood cell count, more pronounced in females than in urinalysis
and histology were evaluated* No effects were observed at 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.
e 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 (80ft a.i.) in the diet at dose levels of 0, 25, 125,
250 or 1,250 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 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. Hematological
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 nemosiderosis 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 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 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 0 and 125 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 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 F2b and F3a litters. A LOAEL
of 125 ppm (6.25 mg/kg/day) was identified.
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Diuron August, 1988
-8-
Developmental Effects
0 Khera et al. (1979) administered by gavage a formulation containing
80% diuron at dose levels of 125, 250 or 500 mg of formulation per
kg body weight to 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, i.e. wavy ribs, extra ribs and delayed ossification.
The incidence of wavy ribs was statistically significant at 250 rag/kg
(p <0.05) 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
• Andersen et al. (1972) reported that diuron did not exhibit mutagenic
activity in T$ bacteriophage test systems (100 ug/plate) or in tests
with eight histidine-requinng 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 £•
typhimurium (with or without metabolic activation), in a Chinese
hamster ovary/hypoxanthine guanine phosphoribosyl-transferase (CHO/HGPRT)
forward gene mutation test or in unscheduled DMA synthesis tests in
primary rat hepatocytes. However, in an in vivo cytogenetic test in
rats, diuron was observed to cause clastogenic effects.
Carcinogenicity
• 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, 6.25, 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.
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Diuron August, 1988
-9-
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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 ( /LJ
(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, 1,000 or 10,000),
in accordance with EPA or 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 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 0, 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 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-Day HA = d25 mg/kg/day) (10kg) (0.80) = 1.0 mg/L (1,000 ug/L)
(1,000) (1 L/day)
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Diuron August, 1988
-10-
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.
O.N80 = correction factor to account for 80% active ingredient.
1,000 = uncertainty factor, chosen in accordance with EPA
or 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:
Longer-term HA = (2.5 mg/kg/day) (10 kg) =0.25 mg/L (300 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 EPA
or 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 = (2.5 mg/kg/day) (70 kg) = 0.875 mg/L (900 ug/L)
(100) (2 L/day)
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Diuron August, 1988
-11-
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.
70vkg = assumed body weight of an adult*
100 = uncertainty factor, chosen in accordance with EPA
or 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. 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 estimating the Lifetime HA for diuron. In
this study, dogs (2 to 3/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 HA is calculated as
follows:
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Diuron August, 1988
-12-
Step 1: Determination of the Reference Dose (RfD)
RfD = (0-625 mg/kg/dav) = 0.002 mg/kg/day
where:
v
0.625 mg/kg/day = NOAEL, based on absence of hematological effects in
dogs exposed to diuron via the diet for 2 years.
100 = uncertainty factor chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
3 = additional uncertainty factor. This factor is used
to account for a lack of adequate chronic toxicity
studies in the data base preventing establishment
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 u /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 (10 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/lose) 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 Agency for Research on Cancer has not evaluated the
carcinogenic potential of diuron.
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Diuron August, 1988
-13-
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 A structurally related analogue of diuron (i.e., linuron) appears to
reflect soiqe oncogenic activity. In addition, a Russian study by
Rubenchik et al. (1973) reported gastric carcinomas and pancreatic
adenomas in rats (strain not designated) given diuron at 450 mg/kg/day
for 22 months. However, the actual data for the study are 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 mg/kg from a dog study and an uncertainty factor of
300 has been calculated (U.S. EPA, 1986b).
0 Residue tolerances that range from 0.1 to 7 ppm have been established
for diuron in or on agricul tural commodities (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 the measurement is
conducted with an ultraviolet (UV) detector. The method has been
validated in a single laboratory and the estimated detection limit
for diuron is 0.07 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%.
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Diuron August/ 1988
-14-
Conventional water treatment techniques of coagulation with ferric
sulfate, 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.
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
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 an individual technology 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, 1988
-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. 1978b. 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-(p_-chlorophenyl)-
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.W. Evans and R.C. Rhodes. 1965. Disappearance of diuron in
cotton field soils. Ir± 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(l):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/ 1988
-16-
DuPont.* 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 DMA 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.v Degradation of specifically labeled diuron in soil and
availability of its residues to oats. Doctoral dissertation. Honolulu/
HI: 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. ^n_ Herbicide contamination of surface runoff waters. Utah State
University, pp. 33-35, 38-43. EPA-R2-73-266; project no. 13030 FDJj
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):311-314. 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. In_ 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, 1988
-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 text/ prepared by Ciba-Geigy,
AG, submitted by Shell Chemical Company, Washington, DC.
Hance, R.J. 1965a. 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)-N1,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. In_ 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. Contain. Toxicol. 22:522-529.
-------
Diuron August, 1988
-18-
Larson, K.A.* 1976. Acute dermal toxicity—Diuron. Unpublished study.
MRID 00017795.
Larson, K.A., and J.H. Schaefer.* 1976. Bye irritation study using the
pesticide diuron. For Colorado International Corporation. Unpublished
Study. MRID 00017797.
v
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, araetryne, 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, HI. 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. 1988. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
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. 1987. National Institute for Occupational Safety and Health.
Registry of Toxic Effects of Chemical Substances (RTECS). Microfiche
edition. July.
Rubenich, B.L., N.E. Botsman, G.P. Gorman and L.I. Loevskaya. 1973.
Relation between the chemical structure and carcinogenic activity
of urea derivatives. Cukalogiya (Kiev) 4:10-16.
Spiridonov, Y.Y., V.S. Skhiladze and G.S. Spiridonova. 1972. The effects of
diuron and monuron in a meadow-bog soil of the moist subtropics of
Adzhariia. Subtrop. Crops. (1):150-155.
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study. MRID 00028006.
Taylor, R.E.* 1976b. Primary skin irritation study. Project T1002.
Unpublished study. MRID 00028007.
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Diuron August/ 1988
-19-
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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carcinogen risk assessment. Fed. Reg. 51(185):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.
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- 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.
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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
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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.
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*Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
ETHYLENE THIOUREA
Health Advisory
Office of Ctinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Ctinkinq
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 nonregulatosy
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.
For those substances that are known or probable human carcinogens, according
to the Agency classification scheme (Qroup A or B), Lifetime HAs are not
recommended. The chemical concentration values for Qroup 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, Waibull, 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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Ethylene Thiourea Ajqust, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
Ethylene thiourea (ETU) is no longer used in commerce but is a common
degradation product of the ethylene bisdithiocarbamate (EBDC) pesticides.
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
V
rv
" - NH
2-Imidazolidinethione
Synonyms
0 ETU
Uses
0 Degradation product of several EBDC pesticides.
Properties
Chemical Formula
Molecular Weight 102. 2
Riysical State (25°C) Wiite crystals
Boiling Point —
Melting Point 203°
Density —
\fepor Pressure —
Specific Gravity —
Water Solubility (30°C) 20 a/L
Log Octanol/Vfeter Partition —
Coefficient
Taste Threshold —
Odor Threshold —
Occurrence
ETU was not found in sampling performed at 264 ground water stations,
according to the STORET database (STORET, 1988).
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Ethylene Ihiourea August, 1988
-3-
Environmental Fate
0 Ethylene thiourea can be degraded by photolysis (U.S. EPAf 1982).
0 14C-Ethylene thiourea was intermediately mobile (R^ 0.61) to very
mobile (Rf 1.00) 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 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). The ethylene
bisdithiocarbamates (EBDCs) are Generally unstable in the presence of
moisture and oxygen, and the EBDCs decompose rapidly in water 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 ETU (U.S. EPA, 1982).
III. PHARMACOKINETICS
Absorption
0 Allen et al. (1978) reported a very high rate of absorption of 14C-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 mgAgr 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 mgAg by gastric intubation, total tissue distribution
at 48 hours was apnroxijnately 25% of the administered 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 Ihiourea August, 1988
-4-
Metabolism
0 Iverson et al. (1980) identified the 24-hour urinary metabolites of
14(>ETU orally administered to male Spraque-Dawley rats at 4 me/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 EBECs rapidly.
ETU and ethylene bisdiisothiocyanato sulfide (EBIS) are the major
metabolites formed (U.S. EPA, 1982). Approximately 7% of an EBDC
dose is converted to ETU in vivo in the rat and 2% in the mouse
(Nelson, 1987; Jordan and Neal, 1979).
Excretion
0 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.
0 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 suitable information was found in the available literature on the
health effects of ETU in humans.
Animals
Short-term Exposure
0 The acute oral LD5Q for ETU 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 eguivalent to
approximately 0.1 mg/kg/day (lehman, 1959), these levels correspond
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Ethylene Thiourea August, 1988
-5-
to doses of about 5, 10, 50 or 75 mg/kg/day. Four hours after the
injection of ^1, uptake was decreased significantly in rats that had
ingested ETU at 500 or 750 ppm for all time periods. At 24 hours after
13ll injection, uptake was significantly decreased in rats that had
ingested 100, 500 or 750 ppm for all time periods. Histoloqically,
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.
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 ware
given by gavage to mice (10/group, sex not specified). Body wsiqht
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 mq/kq was
determined.
In a study by Freudenthal et al. (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 younq rat
is equivalent to 0.1 mgAg (Lehman, 1959), these levels correspond to
doses of about 0, 0.1, 0.5, 2.5, 12.5 or 62.5 mq/kq. Thyroid function,
food consumption, body weight qain 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 thyroqlobulin.
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 TQ, 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.
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.
In a 60-day study, which was a continuation of the above study by
Freudenthal et al. (1977), 14/40 rats in the 625-ppm group died.
Thyroid hyperolasia 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
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Ethylene Ihiourea Auqust, 1988
-6-
dose grcxip when exposure was continued to 90 days. Thus, the NDAEL
for this study is presumed to be 25 ppm, or 2.5 mg/kg.
Dermal/Ocular Effects
0 No information was found in the above literature on the dermal/ocular
effects of ETU.
Long-term Exposure
0 Freudenthal et al. (1977) described alterations in thyroid function
and changes in thyroid morphology when Sprague-Dawley rats were
administered ETU (96.8% pure) in the diet at levels of 1 to 625 pan
(approximately 0.1 to 62.5 mgAg/day based on Lehman, 1959) for up to
90 days. The NDAEL was reported to be 19.5 mg/kg/day at week 1 and
12.5 mgAg/day at week 12.
0 Graham and Hansen (1972) measured 131I 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 aoproximately 0.05 mq/kq/day (Lehman,
1959), these dosages are equivalent to approximately 2.5, 5, 25 and
37.5 mq/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 mgAg/day) for all feeding periods. At 24 hours
after l^I 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 mgAg» the NDAEL 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 0, 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 mq/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 ware 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. Ihe NDAEL for thyroid effects in this study was 25 ppm
(approx iraately 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 par
in their diet for 18 months (approximately 8.75 and 17.5 mq/kg/day,
based on Lehman, 1959). An increased incidence (significance not
specified) of simple goiter was also reported in all treatment groups.
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Ethylene Ihiourea August, 1988
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0 In a 2-year study by Graham et al. (1975), Charles River rats were
fed ETU (purity not specified) in the diet at 0, 5, 25, 125, 250 or
500 ppm (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). Ihere was an increased iodine (131I) 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 mq/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 (100% pure) at 10, 20,
40 or 80 mgAg/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 weiaht 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 Wiera (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, meningorrhagia, meninqorrhea,
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/kq, and at
10 mg/kg there was only a retardation of parietal ossification and of
cerebellar Rarkinje-cell migration. Retarded parietal ossification
was the only abnormality seen at 5 mq/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
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Ethvlene Ihiourea August, 1988
-8-
NOAEL for maternal toxicity was 40 mg/kg/day, and the NOAEL for
developmental effects was 5 nigAg/day.
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 brainrbody 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. Ihe results
of this study are not useful for determining a NOAEL or LOAEL.
Dose-related central nervous system (CNS) lesions in Wistar rat
fetuses were reported by ttiera 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. Wane 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;
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 administration
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 teratogenic effects in rats.
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. Ffowever, fetuses from cats in a
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Ethylene Ihiourea August, 1988
-9-
moribund state subsequent to ETU toxicosis (30 to 120 mg/kq dosage
groups) did show a high incidence (11/35) of malformations including
colobona, umbilical hernia, spina bifida and cleft palate. Maternal
toxicity and death were observed at dose levels of 10 mo/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 NDAEL 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 defec,ts of the
skeletal system (micrcgnathia, micronelia, oligodactyly, kyphosis)
and the CNS (hydrocephalus, encephalocele), as well as cleft palate
were noted in a majority of fetuses at this dose level. Nb clear
evidence of teratogenicity was seen in groups of rats administered
dose levels of 5 to 40 mg/kg/day. fb 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 ware
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 Schupbach and Hummler (1977) reported that ETU induced mutations of
the base-pair substitution type in S. typhimurium TA 1530 j.n_ vitro as
well as in a host-mediated assay. In the host-mediated assay, a
dose of 6,000 mgAg (^050 = 5,400 mgAg) 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
administrations of 700, 1,850 or 6,000 mq/kg to Swiss albino mice;
it was concluded that ETU does not induce any chromosomal anomalies
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Ethylene Ihiourea August, 1988
-10-
in the bone marrow. No dominant-lethal effect was observed after
single oral doses of 500, 1,000 or 3,500 mg/kq were qiven to male
mice.
Carcincqenicity
0 Graham et al. (1975) reported that ETU was a follicular thyroid
carcinogen in male and female Charles River rats that ware 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).
8 In a survey of several compounds for tumorigenicity, 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 siqnificant
(p <0.01) increases in hepatcmas (14/16 or 18/18 for males and 18/18
or 9/16 for females) and in total tumor incidence. Pulmonary tumors
and lymphcmas were also investigated, but ware 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 bv 'administration of a control
diet for 6 months. Assuming that 1 ppm in the diet of older rats is
equivalent to approximately 0.05 mq/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 TOXICDDDGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcincgenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NDAEL or LOAEL) x (BW) = m* ( ug/L)
(UF) x ( Vday)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
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Ethylene Ihiourea August, 1988
-11-
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 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 (0.3 mq/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 eguals 0.1 mq/kg/day, based on Lehman, 1959). Toxic
effects on thyroid function and morphology were observed after 30 days'
exposure to 125 ppm or greater. Nb 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 ^n_ utero at 5 mq/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/kq.
Therefore, 2.5 mgAg was selected as a conservative NOAEL for deriving the
HA.
Using the NDAEL 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 (300 uaA)
(100) (1 L/day)
where:
2.5 mg/kq/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 EPA
or NAS/ODW guidelines for use with a NDAEL frcm 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, 131I
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.
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Ethylene Thiourea August, 1988
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Assuming that 1 ppm in the diet of older rats is eouivalent 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
seen at the 125-, 250- and 500-ppm levels. Uptake of 1-^1 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 mg/kg/day) (10 kg) = 0.125 mgA (100 ugA)
(100) (1 L/day)
where:
1.25 mgAq/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 EPA
or NAS/OCW 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 nq/kq/day) (70 kg) = 0.44 mgA (400 ugA)
(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 EPA
or NAS/ODW 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-
cincgenic adverse health effects over a lifetime exposure. The Lifetime HA
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Ethylene Thiourea August, 1988
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is derived in a three-step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). Ihe 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 he 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, noncarcincgenic 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. Ihe Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC frcm drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed. 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 hvperplasia 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 = (0.25mqAg/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 EPA
or 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.
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Ethylene Thiourea August, 1988
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Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.00003 mg/kg/day) (70 kg) = 0.00105 wg/L (1 ugA)
(2 L/day)
where:
0.00003 mg/kg/day = Rf D.
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, 198ft)), ETU is classified in Qroup 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 ug/L is approximately 4.3 x 10~*>.
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 hepatcmas 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 Qiarles River rats that were fed the compound (ourity 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 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 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 hepatcmas and in total tumor incidence. ND
other dose level was tested. (Pulmonary tumors and lymphcrtias were
also investigated in this study.)
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Ethylene Ihiourea August, 1988
-15-
Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986b), ETU may be classified in
Group B2: probable human carcinogen based on sufficient evidence
from animal studies.
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 4, 2.4 and 0.24 ug/L for 10"4, 10~5 and 1(T6 cancer risk
levels, respectively.
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 Ethylene thiourea is analyzed by a nitrogen-phosphorus detector/gas
chromatographic method as described in Method #6 (U.S. EPA, 1988).
In this procedure the ETU sample is mixed with ammonium chloride and
potassium fluoride and passed through an exchange column (Extrelut).
The ETU is then eluted with methylene chloride, concentrated for
exchange with ethyl acetate to a volume of 5 mL. The method describes
conditions which permit the separation and measurement of ETU by GC
with a nitrogen-phosphorus detector. This method has been validated
by a single laboratory. The estimated detection limit for ETU by
Method #6 is 5 ug/L.
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 frcm drinking
water by activated carbon adsorption. However, since ethylene thiourea
has a high solubility and is hydrophilic, treatment with activated
carbon probably would not be effective.
0 No data were found on the removal of ethylene thiourea frcm 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 frcm drinking
water by aeration. Since vapor pressure data are unavailable, Henry's
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Ethylene Ihiourea August, 1988
-16-
Coeffielent, and thus the effectiveness of aeration, cannot be
estimated. However, the high melting point and the high solubility
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 Ihiourea August, 1988
-17-
IX. REFERENCES
Allen, J.R., J.P. ^fon Miller and J.L. Seymour. 1978. Absorption, tissue
distribution and excretion of l^C ethylenethiourea by the Rhesus monkey
and rat. Res. Comm. Chan. Path. Fharmacol. 20:109-115.
Arnold, D.L., D.R. Krewski, D.B. Junkins , P.P. McGuire, C.A. Moodie and
I.C. Munro. 1983. Reversibility of ethylenethiourea-induced thyroid
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. 5:821-834.
Freudenthal, R.I., G. Kerchner, R. Parsing and R. Baron. 1977. Dietary
subacute toxicity of ethylene thiourea in the laboratory rat. J. Fnv.
Bath. Toxicol. 1:147-161.
Graham, S.L. and W.H. Hansen. 1972. Effects of short-term administration
of ethylene thiourea upon thyroid function of the rat. Bull. Environ.
Contain. Toxicol. 7(1): 19-25.
Graham, S.L., W.H. Hansen, K.J. Davis and C.H. Perry. 1973. Effects of
one-year administration of ethylenethiourea upon the thyroid of the rat.
J. Agr. Food Cnem. 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. Valerio, L. Petrucelli, L. Fishbein, E.R. Hart
and A.J. Pallotta. 1969. Bioassay of pesticides and industrial chemicals
for tumorigenicity 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.
Hiarmacol. 52:16-21.
Jordan, L.W., and R.A. Neal. 1979. Examination of the in vivo metabolism of
maneb and zineb to ethylenethiorea (ETU) in mice. Bull. Environ. Contam.
Toxicol. 22:271-277.
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 in vivo. Food Cosmet. Toxicol.
20:273-278.
-------
Ethylene Thiourea August, 1988
-18-
Khera, K.S. 1984. Ethylenethiourea-induced hindpaw deformities in mice and
effects of metabolic modifiers on their occurrence. J. Toxicol. Environ.
Health. 13:747-756.
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 orogeny
survival following ethylenethiourea treatment 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. WLlloughby, OH: Meister
Publishing Go.
Nelson, S.S. 1987. Bioconversion of mancozeb to ETU in rats. Rohm and Haas
Technical Report No. 31C-87-24. Submitted to EPA. MRID 40301101.
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. Ishn. 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.
Schupbach, 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 mutaqenic
metabolite of ethylenebis-dithiocarbamate. Mut. Res. 26:189-191.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Teramoto, S., R. Saito and Y. Shirasu. 1980. Teratogenic effects of combined
administration of ethylenethiourea and nitrite in mice. Teratology.
21:71-78.
Ulland, B.M., J.H. Vfeisburger, 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.
-------
Ethylene Ihiourea August, 1988
-19-
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.
U.S.'EPA. 1986a. U.S. Environmental Protection Aqency. Final report.
Task 2: Environmental Fate and Exposure Assessment. June 10.
U.S. EPA. 1986b. U.S. Environmental Protection Aqency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method #6 - Determi-
nation of Ethylene Thiourea (ETU) in Ground water by Gas Chromatography
with a Nitrogen-Riosphorus Detector. Available fron U.S. EPA's Environ-
mental 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, 1988
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.
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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Endothall August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 145-73-3
Structural Formula
7-Oxabicyclo-(2,2,1)-heptane-2,3-dicarboxylic acid
Synonyms
0 1,2-dicarboxy3,6-endoxocyclohexane; Aquathol K; Hydrothol 191;
Hydrothol 191 Granular; Endothall Truf Herbicide; Herbicide 273;
Des-i-cate; Accelerate (Meister, 1988).
Uses
0 Preemergence and postemergence herbicide, defoliant, dessicant,
aquatic algicide, growth regulator (Meister, 1988).
Properties (Reinert and Rogers, 1984; Neely and Hackay, 1982; Chiou et al.,
1977; Carlson et al., 1978; Simsiman et al., 1976)
Chemical Formula C8H10°5
Molecular Weight 186.06
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 144°C to the anhydride
Density
Vapor Pressure (25°C) Negligible
Specific Gravity
Water Solubility (25°C) 100 g/L (acid monohydrate)
1.228 g/L (dipotassium salt)
Log Octanol/Water Partition 1.91 (acid)
Coefficient 1.36 (dipotassium salt)
Taste Threshold —
Odor Threshold
Conversion Factor
Occurrence
0 Endothall was not found in any of the 3 surface water or 604 ground
water samples analyzed that were taken from 2 surface water locations
and 600 ground water locations (STORET, 1988). This information is
provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
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Endothall August, 1988
-3-
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
8 The low KQW values indicate that endothall would not significantly
partition to sediments. KOC values for equilibrium sorption studies
using the dipotassium salt were 110 and 138 for sidiment-water systems
from a small eutrophic pond and oligomesotrophic reservoir, respec-
tively (Reinert and Rodgers, 1984). An overall K_ value of 0.958 was
calculated from this study which compares favorably with 1C values
for the acid ranging from 0.41 to 0.9 calculated from a flask system
containing Lake Tomahawk water and sediment (Simsiman and Chesters,
1975). A K_ value of approximately 0.4 was calculated from the data
presented by Sikka and Rice (1973) in which a Syracuse, NY farm pond
was treated with dipotassium endothall. Therefore, sorption would
not be considered a significant environmental fate process for
endothall in the environments studied.
0 Endothall has been shown to not significantly bioconcentrate. In
laboratory and field studies, consistently low endothall levels have
been observed. A BCF for endothall in mosquito fish (Gambusia affinis)
of 10 was observed in a modified Metcalf model ecosystem (Isensee,
1976). In a field study by Serns (1977), a 5 mg/L dipotassium
endothall concentration resulted in BCF values in bluegills ranging
from 0.003 to 0.008. After 72 hr, fish flesh residues were not
detectable. Endothall residues in caged bluegills were consistently
below the minimum detectable level of 0.1 mg/kg in a rservoir study
by Reinert and Rodgers (1986). Similar results were seen after an
application of the diamine salt (Walker, 1963). Comparable fish BCF
values calculated from regression equations were 0.65 (Neely et al.,
1974) and 1.05 (Chiou et al., 1977).
0 Some organisms will exhibit temporary endothall residues that exceed
the water concentration by more than an order of magnitude. Isensee
(1976) observed BCF values of 150 for the water flea, 63 for green
alga (Oedogonium), and 36 for a snail (Physa); however, the endothall
concentrations with the organisms were transient and were not passed
along trophic levels. A BCF of 0.73 was calculated for the dipotas-
sium salt of endothall in duckweed (Lemna minor) using the endothall
KQW and the regression equation found in Lockhart et al. (1983).
0 Volatilization, hydrolysis, and oxidation are not significant fate
processes affecting the persistence of endothall in aquatic environ-
ments (Reinert and Rodgers, 1984). Endothall is also not subject to
photochemical degradation. In a laboratory study using the disodium
salt of endothall, no degradation was observed when a lamp of 254 nm
wavelength was employed (Mitchell, 1961).
9 Biotransformation and biodegradation are the dominant fate processes
affecting the persistence of endothall in aquatic environments
(Simsiman and Chesters, 1975; Holmberg and Lee, 1976; Simsiman et al.,
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Endothall August, 1988
-4-
1976). A biotransfonoation half-life of 8.35 d was observed in a
shake-flask study using three [14c]endothall concentrations and water
from an oligomesotrophic reservoir (Reinert et al./ 1986). Overall
aqueous decay rates are considered a good estimate of endothall
biotransformation in that other fate processes are insignificant.
Keckemet (1980) observed an aqueous endothall half-life of about 6.7 d
after a review of the literature. The persistence of both the dipotas-
sium and amine salts was less than 7 d in Gatan Lake. Panama Canal
(Gangstad, 1983a). A half-life for the dipotassium salt of 7.3 d was
calculated from field studies using farm ponds (Yeo, 1970). Reinert
et al. (1985) observed 4.1 d endothall half-life in 133 L plastic
greenhouse pools containing water, sediment, and Eurasian watermilfoil*
Dipotassium endothall half-lives in a marginally treated north Texas
reservoir ranged from 1.1 to 1.2 days (Rodgers et al., 1984; Reinert
and Rodgers, 1986). The results presented in Holmberg and Lee (1976)
compare favorably with the above endothall half-lives. A 4.1 d half-
life was calculated from a Wisconsin pond treated with dipotassium
endothall.
Endothall persistence in sediments ranged from 0 to 7 d in work
reported by Keckemet (1980) and less than 4 d after a nominal 2 mg/L
dosage in a Texas reservoir (Rodgers et al., 1984; Reinert and Rodgers,
1986). Gangstad (1983a) observed endothall persistence in Gatan Lake
sediment <3 d for the dipotassium salt, but >21 d for the amine salt
when treated with 2 mg/L. In a pond study, Langeland and Warner
(1986) observed an overall endothall (Aquathol® K) persistence of
26 d after an initial concentration of 2.1 mg/L.
III. PHARMACOKINETICS
Absorption
0 Few data exist regarding endothall pharmacokinetics in mammals.
Soo et al. (1967) performed pharmacokinetic experiments with male
and female Wistar rats. Taking ^4C levels in urine and exhaled air
as a crude estimate of absorption, approximately 10% of a 5 mg/kg oral
dose of 14C-labeled (at carbons 1 or 2 of the ring) 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.
Endothall measurements in this study were as 14C label.
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
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Endothall August, 1988
-5-
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
9 The metabolism of endothall is not known to be characterized.
Excretion
Soo et al. (1967) described excretion as follows:
0 Clearance of C-endothall was biphasic in the stomach (^1/2 = 2.2 and
14.2 hours) and kidney (t£ = 1.6 and 34.6 hours) and monopnasic
in the intestine and liver (tj = 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 faces.
Urinary excretion accounted for approximately 7% of the dose, and
approximately 3% of the radioactive label was recovered in expired
carbon dioxide.
0 Recycling of endothall by biliary circulation was ruled out as a
major excretory route because 14c activity in liver was small relative
to that in the original dose.
0 Approximately 20% of the dose excreted in the feces was unchanged
endothall with the remaining radioactivity presumed to be an endothall
conjugate.
0 Soo et al. (1967) also found no radioactivity in pups from lactating
dams given oral doses of 14c-endothall.
IV. HEALTH EFFECTS
Humans
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).
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Endothall August, 1988
-6-
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 Dig/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 acid 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.,
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 heraorrhagic 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.
0 Mine male dogs (one dog/dose) were dosed orally with capsules containing
1 to 50 mg disodium endothalI/kg/day (0.8 to 40 mg endothall ion/kg/day)
for 6 weeks (Brieger, 1953b). All dogs that were administered 20 to
50 mg disodium endothal1/kg/day died within 11 days. Vomiting and diarrhea
were observed in the group given 20 mg disodium endothal1/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
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, conjunctival irritation
and iridic congestion. Furthermore, technical endothall apparently
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Endothall August, 1988
-7-
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 test
generation. Pups in the 4-mg/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.
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 ion/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
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Endothall August, 1988
-8-
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).
0 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
0 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-term (up to 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).
*Because the test material in the various toxicity studies was salt or acid
forms of endothall, the HAs described herein are expressed in terms of
endothall ion.
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Endothall August, 1988
-9-
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 mg endothall ion/kg/day, based on maternal toxicity)•
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA = (8 mg/kg/day) (10 kg) = 0.8 /L (800 u /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 EPA or
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
There is concluded to be insufficient data for calculation of a Longer-
term HA. Therefore, the DWEL adjusted for a 10-kg 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 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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Endothall August/ 1988
-10-
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
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. 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/kg/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/kg/day, the Lifetime HA for endothall is calculated
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfU = (2 mg/kg/day) = 0.02 mg/kg/day
(100)
where:
2 mg/kg/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•
100 = uncertainty factor, chosen in accordance with EPA or
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/dav) (70 kg) = 0.7 mg/L (700 ug/L)
(2 L/day)
where:
0.02 mg/kg/day = RfD.
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Endothall August, 1 988
-11-
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 (100 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 0 (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).
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 Endothall is a dicarboxy1lie herbicide used to control aquatic
vegetation. It presents a difficult analytical challenge since its
highly polar structure does not allow simple solvent extraction from
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Endothall August, 1988
-12-
water. It lacks any elements which would allow use of the selective
detectors commonly used in pesticide analysis. To obtain any
sensitivity chemical derivatization is required both for detection
and concentration of the extract.
A draft method has been developed by EMSL-CI (U.S. EPA, 1988) which
allows a small volume of sample to be derivatized with pentafluoro-
phenylhydrazine (PFPH), and cleaned up by solid phase absorbant-
extraction. Analysis is by electron-capture gas chromatography•
The single laboratory estimated detection limit is 9.1 ug/L. The
single operator recovery and precision for this method in tap water
is 69% and 4%, respectively.
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, 1988
-13-
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. EPA Accession No. 246012.
Brieger, H.* 1953b. Endothall, long term oral toxicity test—rats. EPA
Pesticide Petition No. 6G0503, redesignated No. 7F0570, 1966. EPA
Accession No. 246012.
Carlson, R., R. Whitaker and A. Landskov. 1978. Endothall. Chapter 31.
Iii G. Zweig and J. Sherma, eds. Analytical methods for pesticides and
plant growth. New York: Academic Press, pp. 327-340.
Chiou, C.T., V.H. Freed, D.W. Schmedding and M. Manes. 1982. Partitioning
of organic compounds in octanol-water systems. Environ. Sci. Technol.
16:4-9.
CFR. 1977. Code of Federal Regulations. 40 CFR 180.293.
CFR. 1979. Code of Federal Regulations. 21 CFR 193.180. April 1, 1979.
Gangstad, E.O. 1983. Herbicidal, environmental, and health effects of
endothall. OCE-NRM-23. Office, Chief of Engineers, U.S. Army Corps of
Engineers, Washington, DC, 25 pp.
Goldstein, F. 1952. Cutaneous and intravenous toxicity of endothall
(disodium-3-endohexahydrophthalic acid). Pharmacol. Exp. Ther. 11:349.
Holmberg, D.J., and G.F. Lee. 1976. Effects and persistence of endothall in
the aquatic environment. J. Water Poll. Control Fed. 48:2738-2746.
Isensee, A.R. 1976. Variability of aquatic model ecosystem-derived data.
Int. J. Environ. Stud. 10:35-41.
Keckemet, 0. 1980. Endothall-potassium and environment. Proc. 1980 Brit.
Crop Protection Conf. Weeds, 8 pp.
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. EPA Accession No. 24601.
Langeland, K.A., and J.P. Warner. 1986. Persistence of diquat, endothall
and fluridone in ponds. J. Aquat. Plant Mgmt. 24:43-46.
-------
Endothall August, 1988
-14-
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. EPA
Accession No. 245680.
Lockhart, W.L., B.N. Billeck, G.G.E. de March and D.C.G. Muir. 1983. Uptake
and toxicity of organic compounds: Studies with an aquatic macrophyte
(Lemna minor). In: W.E. Bishop, R.E. Cardwell and B.B. Heidolph, eds.
Aquatic Toxicology and Hazard Assessment: Sixth Symposium. ASTM STP 802.
ASTM, Philadelphia, PA, pp. 460-468.
Meister, R., ed. 1988. Farm Chemicals Handbook. Willoughby, OH. Meister
Publishing Company.
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. EPA
Accession No. 244126.
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, MO.
Mitchell, L.C. 1961. The effects of ultraviolet light (2537A) on 141
pesticide chemicals by paper chromatography. J. Assoc. Off. Analyt.
Chem. 44:643-712.
Neely, W.B., and 0. Mackay. 1982. An evaluative model for estimating environ-
mental fate. In: K.L. Dickson, A.W. Maki and J. Cairns, Jr., eds.
Modeling the Fate of Chemicals in the Aquatic Environment. Ann Arbor
Science Publishers, Inc., Ann Arbor, MI, pp. 127-143.
Neely, W.B., D.R. Branson and G.E. Blau. 1974. Partition coefficient to
measure bioconcentration potential of organic chemicals in fish. Environ.
Sci. Tech. 8:1113-1115.
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
4581-EIE. Summary data on acute oral toxicity and dermal irritation in
rabbits (Endothall). EPA Accession No. 244125.
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). EPA Accession No. 246012.
Reinert, K.H., and J.H. Rodgers, Jr. 1984. Influence of sediment types on
the sorption of endothall. Bull. Environ. Contain. Toxicol. 32:557-564.
-------
Endothall August, 1 988
-15-
Reinert, K.H., and J.H. Rodgers, Jr. 1986. Validation trial of predictive
fate models using an aquatic herbicide (endothall). Environ. Toxicol.
Chem. 5:449-461.
Reinert, K.H., J.H. Rodgers, Jr./ H.L. Hinman and T.J. Leslie. 1985. Coml-
partmentalization and persistence of endothall in experimental pools.
Ecotox Inviron. Safety. 10:86-96.
Rodgers, J.H., Jr., K.H. Reinert and H.L. Hinman. 1984. Water quality
monitoring in conjunction with the Pat Mayse Lake aquatic plant management
program. In: Proceedings, 18th Annual Meeting, Aquatic Plant Control,
Research Program. November 14-17, 1983. Misc. Paper A-84-1. Raleigh,
NC, U.S. ARmy Corps of Engineers, pp. 17-24.
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. 246012.
Science Applications, Inc.* 1982. A dose range-finding teratology study of
endothall techni'cal 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.
Serns, S.L. 1977. Effects of dipotassium endothall on rooted aquatics and
adult and first generation bluegills. Water Res. Bull. 13:71-80.
Sikka, H.C., and C.P. Rice. 1973. Persistence of endothall in the aquatic
environment as determined by gas-liquid chromatography. J. Agric. Food
Chem. 21:842-846
Simsiman, G.V., and G. Chesters. 1975. Persistence of endothall in the
aquatic environment. Water, Air, Soil Poll. 4:399-413
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. Metabolism of 14c-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)
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
-------
Endothall August, 1988
-16-
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 York: Marcel Oekker, Inc., pp>
815-833.
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. 1988. U.S. Environmental Protection Agency. Draft test method
XXX for endothall. Environmental Monitoring and Support Laboratory,
Cincinnati, Ohio. June, 1988.
Vigfusson, N.V. 1981. Evaluation of the mutagenic potential of Aquathol K
by induction of sister chromatid exchanges in human lymphocytes in vitro.
Report to Municipality of Metropolitan Seattle, Seattle, WA. EPA
Accession No. 245680.
Walker, C.R. 1963. Endothall derivatives as aquatic herbicides in fishery
habitats. Weeds. 11:226-232.
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 aiiJ G.B. Wilson. 1956. Cytological and genetical
effects of the defoliant endothall. J. Hered. 47:151-154.
Yeo, R.R. 1970. Dissipation of endothall and effects on aquatic weeds and
fish. Weed Sci. 18:282-284.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
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.
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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Fenamiphos
August, 1988
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 22224-92-6
Structural Formula
-2-
CH,S
0 H
0-P-N-CH(CH,),
OC>HS
(1-Methylethyl)-ethyl-3-methyl-4-(methylthio)phenyl-phosphoramidate
Synonyms
o Nemacur; B 68138; Bay 68138; Bayer 68138; ENT 27572; Phenamiphos
(Meister, 1983).
Uses
0 Systemic nematicide (Meister, 1983).
C13H22°3NSP
303 (calculated)
Brown, waxy solid
49.2«C
7.5 x 10~7 mm Hg
400 mg/L
Properties (Meister, 1983)
Chemical Formula
Molecular Weight
Physical State (at 25«C)
Boiling Point
Melting Point
Density
Vapor Pressure (30°C)
Water Solubility (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Fenamiphos has not been detected in 664 ground water samples analyzed
from 659 locations (STORET, 1988). No surface water locations were
tested.
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Fenamiphos August, 1988
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Environmental Fate
0 Ring-labeled 14c-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 14C-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/cra2, 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 compound 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.
0 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.
0 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 lb/galIon 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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Fenamiphos August, 1988
-4-
8 Aged (30 days) 14C-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 Gronberg (1969) administered 14c-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 mgAg
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 42% or 50% of the administered radioactivity was expired
as C02, respectively. Thirty-eight to 40% of the ethyl-14C labels
were in urine and 5% in feces, respectively. Eighty percent of the
methylthio-3H label was found in urine. The majority of the admini-
stered 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, including any data on accidental
poisoning (U.S. EPA, 1979).
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Fenamiphos August, 1988
-5-
Animals
Short-term Exposure
0 NIOSH (1987) reported the acute oral LD50 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. This corresponded 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.
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
(82% a.i.) 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 0.05 mg/kg/day (Lehman, 1959), this corresponds to doses of 0, 0.2,
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Fenamiphos August, 1988
-6-
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 mgAg/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 mgAg/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 a 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 plasma and 16%
in red blood cells in females. At 5 .ppm, ChE in plasma was inhibited
44% and 41%,and red blood cell ChE was' inhibited 26% and 22% (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 20 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
mgAg/day), lowest dose tested, was identified, but not a NOAEL.
• 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
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. 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
weight in males. No compound-related effects were observed in any of
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Fenamiphos August, 1988
-7-
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).
0 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 ppm 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 mgAg/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 mgAg/day was identified.
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Fenamiphos August, 1988
-8-
Mutagenicity
0 Herbold (1979) reported that fenamlphos 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 NMRI mice (Herbold and Lorke,
1980)| acute oral doses of 5 mg/kg did not produce mutagenic effects.
Carcinogenicity
0 Hayes et al. (1982) administered fenamiphos (90% purity) for 20 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 (up to 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, 1,000 or 10,000),
in accordance with EPA or 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, 1988
-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 mg/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 mgAg or less, respectively. Chain fusion of sternebrae were observed in
the 1.0 mgAg 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:
Ten-day HA = (0»1 mg/kg/day) (10 kg) (0.88) = Q.009 mg/L (9 ug/L)
(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 EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
1 L/day = assumed 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
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Fenamiphos August, 1988
-10-
level (0.125 mgAg/day). No significant effect was seen at 2 ppm or less
(0.05 mgAg/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 Kimmerle
(1968) identified a NOAEL of 0.2 mgAg/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 mgAg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-terra HA = (0'05 nig/kg/day) (10 kg) = 0.005 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 EPA
or NAS/OOH 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) = Q.018 mg/L (20 ug/L)
(100) (2 L/day)
where:
0.05 mg/kg/day - NOAEL, based on absence of significant cholinesterase
inhibition in uogs exposed to fenamiphos via the diet
for 3 months.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or 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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Fenamiphos August, 1988
-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
(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. 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. 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/kcr/dav) = 0.00025 mg/kg/day
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 EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
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Fenamiphos August, 1 988
-12-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL - 0-00025 mg/kg/day) (70 kg) = 0.009 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.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.009 mg/L) (20%) = 0.0018 mg/L (2 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.
0 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 carcinogenicity.
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 mgAg/day (U.S. EPA, 1985).
0 The World Health Organization (WHO) calculated a TADI of 0.0003
mgAg/day for fenamiphos (Vettorazzi and Van den Hurk, 1985).
VII. ANALYTICAL METHODS
0 Analysis of fenamiphos is by a gas chromatographic (GC) method
applicable to the determination of certain nitrogen-phosphorus-
containing pesticides in water samples (Method #507, U.S. EPA, 1988).
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Fenamiphos August, 1988
-13-
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 has been validated in a
single laboratory. The estimated detection limits for analytes with
this method, including fenamiphos, is 1.0 ug/L.
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, 1988
-14-
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 Cheraagro Corp., Kansas City, MO; CDL:091690-H. MRID 00067117.
Crawford, C., and R. Anderson.* 1973. The eye and skin irritancy of Nemacur
3 Ibs/gal spray concentrate to rabbits. Report No. 37549. Unpublished
study. MRZD 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, 1983
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. Mallicoat.* 1982. Technical fenamiphos
oncogenicity study in mice. Report No. 3037. Unpublished study.
MRID 00098614.
Herbold, B.* 1979. Nemacur: Salmone1la/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, 1988
-15-
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics* Assoc. Food Drug Offt
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. 1987. National Institute for Occupational Safety and Health. Registry
of toxic effects of chemical substances (RTECS). Microfiche edition. July.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Tweedy, B.G., 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. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.349, p. 324. July 1, 1985.
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Fenamiphos August, 1988
-16-
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. 1988. U.S. Environmental Protection Agency. Method #507 -
Determination of nitrogen and phosphorus containing pesticides in water
by GC/NPD. April 15 draft. Available from U.S. EPA's Environmental
Monitoring and Support Laboratory, Cincinnati, Ohio 45268.
Vettorazzi, G., and G.W. Van den Hurk. 1985. Pesticides reference index,
JMPR 1961-1984. p. 10.
^Confidential Business Information submitted to the Office of Pesticide
Programs
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August, 1988
FLUOMETURON
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.
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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Fluometuron August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 2164-17-2
Structural Formula *J
H-N-C-NtCH,),
N,N-Dimethyl-N-(3-(trifluoromethylJphenyl)-urea
Synonyms
0 C 2059; Cotoron; Cottonex; (Meister, 1988).
Uses
0 Preemergence and postemergence control of annual grasses and broadleaves
in cotton and sugarcane (Meister, 1988).
Properties (Windholz et al., 1983; CHEMLAB, 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 10~7 mm 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 156 ground water samples analyzed
from 125 locations or in 14 surface water samples analyzed from 14
locations (STORET, 1988). This information is provided to give a
general impression of the occurrence of this chemical in ground and
surface waters as reported in the STORET database. The individual
data points retrieved were used as they came from STORET and have not
been confirmed as to their validity. STORET data is often not valid
when individual numbers are used out of the context of the entire
sampling regime, as they are here. Therefore, this information can
only be used to form an impression of the intensity and location of
sampling for a particular chemical.
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Fluometuron August, 1988
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Environmental Fate
» 14c-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 chromatography (TLC) tests of soil (Helling, 1971;
Helling et al., 1971).
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 Fogleman (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 rag/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 Fogleman (1967) detected radioactivity in the liver, kidneys,
adrenals, pituitary, red blood cells, blood plasma and spleen 72 hours
after oral administration of He-labeled fluometuron at dose levels of
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Fluometuron August, 1988
-4-
50 or 500 mg/kg in rats. The highest concentration was detected in
red blood cells.
Metabolism
0 Boyd and Fogleman (1967) concluded that, by thin-layer chromatographic
analysis, the urine of rats in their study contained m-trifluoromethyl-
aniline, desmethyl-fluometuron, demethylated fluometuron, hydroxylated
desmethyl-fluoraeturon, 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
Boyd and Fogleman (1967) reported that urinary excretion of radio-
active label peaked at 24 hours after administration of 14c-fluometuron
(50 mg/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 Fogleman (1967), fecal excretion of fluometuroii
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.
Animals
Short-term Exposure
0 NIOSH (1985) reported the acute oral LD^g values of fluometuron as
6,416, 2,500, 900 and 810 rag/kg in the rat, rabbit, mouse and guinea
pig, respectively.
0 Sachsse and Bathe (1975) reported an acute oral LD5Q value of
4,636 mg/kg for both male and female Tif RA1 rats.
0 Foglemann (1964a) reported the acute oral LD$Q values for CFW albino
mice as 2,300 mg/kg in females and 900 mg/kg in males.
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Fluometuron August, 1988
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Dermal/Ocular Effects
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.
0 Fogleman (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.
9 Technical fluometuron was not found to be an eye irritant in rabbits
(Fogleman, 1964c).
Long-term Exposure
0 Fogleman (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 Fogleman (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
15mgAg/day (NOAEL).
0 In the NCI (1980) study, B6C3F1 mice and F344 rats (10 of each sex)
were given fluometuron (>99% pure) in the diet for 90 days at 250,
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Fluometuron August, 1 988
-6-
500, 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 Halpighian corpuscles
and depletion of lymphocytic elements. Body weight gain was reduced
(>10%) in male and female mice given more than 2,000 ppm. Assuming
that 1 ppm in the diet equals 0.05 mg/kg/day in the rat and 0.15
mg/kg/day in the mouse (Lehman, 1959), 1,000 ppm (NOAEL) corresponds
to 50 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 were 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 0.05 mg/kg/day in the rat (Lehman, 1959),
a Lowest-Observed-Adverse-Effect Level (LuAEL) of 250 ppm (12.5
mg/kg/day) is suggested in this study.
0 No noncarcinogenic effects (survival, body weight and pathological
changes) in B6C3F1 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.
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Fluometuron August/ 1988
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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.
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 <0.0.5 at the low and mid doses; insufficient number of fetuses for
statistical analysis at the high dose). A LQAEL 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 mg/kg bw) given as a
single oral dose of an aqueous suspension by gavage resulted in a
strong inhibition of mouse testicular DMA 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-J 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 B6C3Fj mice, an increased incidence
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Fluometuron August, 1988
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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%J. NCI (1980)
concluded that additional testing was needed because of equivocal
findings for male mice and because both rats and mice may have been
able to tolerate higher doses. The NOAELs identified for rats and
mice are 12.5 and 75 mg/kg/day, respectively.
Chronic feeding studies with technical fluometuron in rats and mice
are ongoing to satisfy OFF data requirements.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) X (BW) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet Level
in mg/kg bw/day.
x.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UP = uncertainty factor (10, 100, 1,000 or 10,000) in
accordance with EPA or NAS/OON 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 (2 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 (1 984) was not selected because a NOAEL was not
identified. It is therefore recommended that the Longer-term HA value for a
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Fluometuron August, 1988
-9-
10-kg child (2 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 Fogleman (1 965b) has been selected
to serve as the basis for the Longer-term HA value for fluometuron. In this
study, dogs given technical fluometuron at dose levels of 0, 1.5, 15 or 150
mg/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 Fogleman (1965a) and NCI
(1980) were not selected because the 15 mg/kg/day NOAEL in the Fogleman
(1965b) study was below the lowest dose of 75 mg/kg/day in the Fogleman
(1965a) study where effects were noted, and pathological changes in spleen
found with the dietary level of 12.5 mg/kg/day in the repeat NCI (1980)
90-day study in rats were not found with this level 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 Fogleman (1965a) study and 12.5 mg/kg/day (estimated) in the
NCI (1980) carcinogenic!ty bioassay were NOAELs, it is concluded that 15
mg/kg/day would be consistent with a NOAEL 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 Fogleman (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) . 1t5 /L (2,000 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 EPA or
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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Fluometuron August, 1988
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The longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (15 mg/kg/day) (70 kg) . 5.3 mg/L (5 000 ug/L)
(100) (2 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.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA or
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, noncarclnogenic 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. 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 NCI (1980) carcinogenic!ty 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 12.5 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
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Fluometuron August, 1 988
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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 B6C3F-| mice was not considered because higher dose levels
(500 and 1,000 ppm, estimated as 75 and 150 mg/kg/day) were used.
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 = (12.5 mg/kg/day) = Q.0125 mg/kg/day
(100)(10) (rounded to 0.013 mg/kg/day)
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 EPA or
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)
DWEL - (0-013 mg/kg/day) (70 kg) = 0.455 mg/L (500 ug/L)
(2 Vday)
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 Health Advisory
Lifetime HA = (0.46 mg/L) (20%) = 0.09 mg/L (90 ug/L)
where:
0.46 mg/L = DWEL.
20% = assumed relative source contribution from water.
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Fluometuron August, 1988
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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 carcinogenicity for humans.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986a), 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 carhamate and
urea pesticides in water samples (U.S. EPA, 1986b). 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. This method has been validated in a
single laboratory and estimated detection limits have been determined
for the analytes in the method, including fluometuron. The estimated
detection limit is 0.10 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAG) adsorption
will remove fluometuron from water.
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Fluometuron August, 1988
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Whittaker (1980) experimentally determined adsorption isotherms for
fluometuron on GAG.
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.
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, 1988
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IX. REFERENCES
Arhur, A., and V. Triana.* 1984. Teratology study (with fluometuron) in
rabbits. Ciba-Geigy Corporation. Report No. 217-84. Unpublished
study. MRID 842096.
Boyd, V.F., and R.W. Fogleman.* 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.
HRID 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(41:428-433.
Dunkel, V.C., and V.F. Simmon. 1980. Mutagenic activity of chemicals
previously tested for carcinogenicity in the National Cancer Institute
bioassay. PROGRAM. IARC. Sci. Publ. 27:283-302.
Fogleman, 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 00019093.
Fogleman/ R.W.* 1964b. Compound C-2059 80 HP-repeated rabbit dermal toxicity.
AME Associates for CIBA Corporation. Project No. 20-0242. Research
Project CF-740. Unpublished study. MRID 00018593.
Fogleman, R.W.* 1964c. Compound C-2059 Technical — Acute eye irritation —
Rabbits. AME Associates for CIBA Corporation. Project No. 20-042.
Unpublished study. MRID 0019032.
Fogleman, R.W.* 1965a. Cotoran — 90-day feeding rats. AME Associates for
CIBA Corporation. Project No. 20-042. Unpublished study. MRID 00019034.
Fogleman, 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, D.* 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.
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Fluometuron August, 1988
-15-
Guth, J. A.., H. Geissbuehler and L. Ebner. 1969. Dissipation of urea
herbicides in soil. Heded. 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. 1988. 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 mitotic gene conversion in Saccharomyces cerevisiae. Mutat. Res.
22:111-120.
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Fluoraeturon August, 1988
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Siglin, J.C., P.J. Becci and R.A. Parent.* 1981. Primary skin irritation in
rabbits (EPA - FIFRA): FDRL: Study Mo. 6817A. Food and Drug Research
Laboratories for Ciba-Geigy. Unpublished study. MRID 00068040.
STORKT. 1988. STORET Water Quality Pile. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
TIB. 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. 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 #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.
Whittaker, K.F. 1980. Adsorption of selected pesticides by activated carbon
using isotherm and continuous flov 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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August, 1988
FONOFOS
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.
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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Fonofos August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 944-22-9
Structural Formula
O-Ethyl-S-phenylethylphosphonodithioate
Synonyms
0 Difonate; Difonatal; Dyfonate®; ENT 25, 796; Fonophos; Stauffer N2790
(Meister, 1983).
Uses
0 Soil insecticide (Meister, 1983).
Properties (Windholz et al., 1983; TOB, 1985)
Chemical Formula C^QH^5OS2P
Molecular Weight 246.32
Physical State (25»C) Light yellow liquid
Boiling Point 130°C
Melting Point
Vapor Pressure (25°C) 2.1 x 10-4 mm Hg
Specific Gravity (20«C) 1.154
Water Solubility (25«C) Practically insoluble
Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
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 (Kadoum and Mock, 1978).
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, 1988
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0 Fonofos has ben found in 194 of 3,399 surface water samples analyzed
and in 2 of 570 ground water samples (STORET, 1988). Samples were
collected from 183 surface water locations and 456 ground water
locations. The 85th percentile of all non-zero samples was 0.32 ug/L
in surface water and 8,900 ug/L in ground water sources. The maximum
concentration found in surface water was 33 ug/L and in ground water
it was 8,900 ug/L. Fonofos was found in surface water in Iowa,
Illinois, Michigan and Ohio, and in ground water in Alabama. This
information is provided to give a general impression of the occurrence
of this chemical in ground and surface waters as reported in the
STORET database. The individual data points retrieved were used as
they came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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 (McBain 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, 0-ethylethane phosphonic acid, 0-ethyl 0-methylethyl phosphonate,
diphenyl disulfide, methylphenyl sulfoxide, and methylphenyl sulfone
(Hoffman et al., 1973; Hoffman and Ross, 1971). The soil fungus,
Rhizopus japonicus rapidly degraded 14c-fonofos to yield dyfoxon,
thiophenol, ethylethoxy phosphonic acid and methylphenyl sulfoxide
(Lichtenstein et al., 1977).
0 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 14c-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 14c-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 Hc-fonofos applied
volatilized from soil water (a suspension of fine sand in tapwater
or tapwater alone; 1% volatilized from a silt loam soil alone).
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Fonofos August, 1988
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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).
In the field, fonofos dissipated with a half-life of 28 to 40 days
when either a 10% G or a 4 Ib/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).
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 mg/kg. 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 rag/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-14c-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
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Fonofos August, 1988
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terminal metabolites, including: 0-ethylethane phosphonothioic acid,
0-ethylethane phosphonic acid, and 0-conjugates of 3- and 4-(hydrox-
phenyl)methyl sulfone. HcBain 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 label
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.
Hoffman et al. (1971) dosed rats orally with 14c-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 19-year old
woman who ate pancakes prepared with ingredients containing fonofos.
No quantitative 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.
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Fonofos August, 1988
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Animals
Short-'term Exposure
0 Reported values for the oral LDsg 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 mgAg (Horton, 1966a,b; Dean, 1977).
0 Horton (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. These levels were stated by the authors to
be equivalent to 0 and 0.2 mg/kg/day. Plasma and red blood cell
cholinesterase (ChE) were measured at 2 and 4 weeks; organ weights,
brain ChE 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 mgAg/day).
0 In a demyelination study, groups of 10 adult hens each received
technical fonofos (99.8% pure) in the diet for 46 days (Woodard and
Woodard, 1966). Levels fed were equivalent to 0, 2, 6.32 or 20
mgAg/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 LD^Q values of fonofos for the rabbit (both sexes)
ranged from 121 to 147 mg/kg (Horton, 1966a,b). However, Dean (1977)
determined a different LDso in rabbits: 25 mgAg for females and
100 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 the intact or
abraded skin of New Zealand rabbits (five/sex/dose; the five animals
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Fonofos August, 1988
-7-
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 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.
Only values for abraded skin were reported. 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 NOAEL 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 difonate at 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. These levels were stated by
the authors to be equivalent toO, 0.4, 1.5 or 6 mg/kg/day. Dogs
showed increased lacrimation and salivation plus convulsions (at
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Ponofos August, 1988
-8-
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 inhibition 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 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/kg/day), the lowest dose tested, was identified.
0 Pure-bred beagle dogs were fed fonofos in the diet for 2 years
(Woodard 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, aach 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) (Banerjee
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.
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Fonofos August/ 1988
-9-
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
the diet of a rat is equivalent to 0.05 mg/kg/day (Lehman, 1959),
this corresponds 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 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 mgAg/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. Based
on these findings, the NOAEL for fetotoxicity identified in this
study was 2 mg/kg/day. The teratogenic NOAEL identified for this
study was 8 mg/kg/day.
Mutagenicity
e Fonofos, with or without metabolic activation, was not mutagenic in
each of five microbial assay systems (the Ames (Salmonella typhimurium)
test; reverse mutation in an Escherichia coli strain; mitotic recombi-
nation 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 (WI-38) cells
(Simmon, 1979).
Carcinogenicity
0 Groups of 30 male and 30 female CO 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 (Banerjee 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 (up to 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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Ponofos August, 1988
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HA = (NOAEL or LOAEL) x (BW) = mq/I| / UCJ/Ij)
(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).
UP = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or NAS/ODW 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 adjusted DWEL for a 10-kg child of 0.02 mg/L (20 ug/L)
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 adjusted DWEL for a 10-kg child of 0.02 mg/L (20 ug/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 fonofos. It is therefore
recommended that the adjusted DWEL for a 10-kg child of 0.02 mg/L (20 ug/L)
be used at this time as a conservative estimate of the Longer-term HA and
that the DWEL of 0.07 mg/kg (70 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 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
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Fonofos August, 1988
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which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the Rf0 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. 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 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 EPA
or 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 /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.
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Fonofos August, 1988
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Step 3: Determination of the Lifetime Health Advisory
lifetime HA = (0.07 mg/L) (20%) = 0.014 mg/L (10 ug/L)
where:
0.07 mg/L = DWEL.
20% = assumed relative source contribution from water.
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 mg/kg/day) for 2 years (Banerjee 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 0: 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 (Method #507, U.S. EPA, 1988). 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-
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Fonofos August, 1988
-13-
phosphorus detector. The method detection limit has not been deter-
mined 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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Fonof os August, 1 988
-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.
Ahmed, J. and F.O. Morrison. 1972. Longevity of residues of four organo-
phosphate insecticides in soil. Phytoprotection. 53(2-3) : 71-74.
Banerjee, 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. Lor be r. 1986. Monitoring ground water for
pesticides in the U.S.A. Evaluation of Pesticides in Ground Water,
American Chemical Society Symposium Series #315.
Dean, W.P.* 1977. Acute oral and dermal toxicity (LDsg) in male and female
albino rats. Study No. 153-047. International Research and Development
Corporation. Unpublished study. MRIDS 00059860, 00059856 and 00059857.
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
Wilkins. p. 413
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, Richmond, 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 LDso - rats; acute dermal toxicity -
rabbits; acute eye irritation - rabbits. Technical Report T-986. Stauffer
Chemical Company. Unpublished study. MRID 00090806.
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.
-------
Fonofos August, 1988
-15-
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 Chera. 26(1):45-50.
Kiigemagi, U. and L.C. Terriere. 1971. The persistence of Zinophos and
Dyfonate in soil. Bull. Environ. Contam. 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. Kbrte 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 O-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. Contam. 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(Part 2):A197.
Minor, J., J. Downs, G. Zwicker et al.* 1982. A teratology study in CD-1
mice with Dyfonate technical T-10192. Final report. Stauffer Chemical
Company. Unpublished study. MRID 00118423.
Schulz, K.R. and E.P. Lichtenstein'. 1971. Field studies on the persistence
and movement of Dyfonate in soil. J. Econ. Entomology. 64(1):283-287.
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Fonofos August, 1988
-16-
Simmon, V.P. 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.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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. 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. 1988. U.S. Environmental Protection Agency. U.S. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, April 15 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, Ohio 45268.
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.
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 00082233.
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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Auyusii,
GLiTPHOSATE
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.
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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Glyphosate August, 1 988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1071-83-6
Structural Formula
0 0
I
-P-OH
HO-C-CH2-N-CH2
I I
H OH
Glycine, N-(Phosphonomethyl)
Synonyms
Uses
0 Active ingredient (glyphosate) in Rodeo®, Roundup®, Landmaster®,
Shakle®, Roundup L & G®, Polado®.
0 Herbicide for control of grasses, broad leaved weeds and woody brush
(U.S. EPA, 1986b).
Properties (Meister, 1983)
Chemical Formula' C3HgN05P
Molecular Weight 169.07
Physical State (25°C) White crystalline solid
Boiling Point
Melting Point 200°C (decomposes)*
Density 0.5 g/mL (bulk density of dried technical)*
Vapor Pressure
Water Solubility 11.6 g/L*
Log Octanol/Water Partition 0.0006-0.0017*
Coefficient
Taste Threshold
Odor Threshold —
Conversion Factor
* Monsanto, 12/13/87
Occurrence
0 No glyphosate was detected in 6 surface water samples or 98 ground-
water samples taken from 3 surface water stations and 97 groundwater
stations, respectively (STORET, 1988).
Environmental Fate
0 Biodegradation is considered the major fate process affecting
glyphosate persistence in aquatic environments (Reinert and Rodgers,
1987). Glyphosate is biodegraded aerobicaly and anaerobically by
microorganisms present in soil, water, hydrosoil and activated sludge.
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Glyphosate August, 1988
-3-
14c-Glyphosate (94% glyphosate, 5.9% ami name thylphosphonic acid) and
aminoraethylphosphonic acid were stable in sterile buffered water at
pH 3, 6, and 9 during 32 days of incubation in the dark at 5 and 35°C
(Brightwell and Malik, 1978).
14c-Glyphosate ( 94% glyphosate, 5. 9% aminomethylphosphonic acid) was
adsorbed to Drummer silty clay loam, Ray silt, Spinks sandy loam,
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).
(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)
14c -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 for 30 days. Glyphosate residue
levels were not detectable (<0.025 ppm) in milk and eggs from cows
and chickens on diets containing glyphosate.
Distribution/Metabolism
0 The distribution, metabolism and retention of glyphosate by the
tissues appear to be minimal since 60 to 90% of a single oral dose
is rapidly eliminated unchanged in the feces (Duerson and Sipes,
1987).
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.
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Glyphosate August, 1988
-4-
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
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
Glyphosate, a widely utilized herbicide, was evaluated for acute
irritation, cumulative irritation, photoirritation and allergic and
photoallergic contact potential in 346 volunteers. The herbicide was
less irritant than a standard liquid dishwashing detergent and a
general all purpose cleaner. There was no evidence for the induction
of photoirritation, allergic or photoallergic contact dermatitis.
Ten percent glyphosate in water is proposed as a diagnostic patch
test concentration (Maibach, 1986).
Animals
Short-term Exposure
0 An oral LD50 cf 5,600 tag/kg in the rat is reported for glyphosate
(Monsanto Company, 1985).
o Bababunmi et al. (1978) reported that daily intraperitoneal admini-
stration of 15, 30, 45 or 60 mg/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 LD5g for glyphosate in the rabbit was reported to be
>5,000 mg/kg (Monsanto Company, 1985).
Long-term Exposure
0 In subchronic studies reported by the Weed Science Society of America
(1983), technical-grade glyphosate was fed to rats and dogs at dietary
levels of 200, 600 or 2,000 ppm 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 histopathological changes. No other
details or data were provided.
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Glyphosate August, 1988
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0 Bio/dynamics/ Inc. (1981a) 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.
0 Glyphosate was administered to mice at dietary levels of 5,000, 10,000
and 50,000 ppm for 3 months. Decreased body weight gains were observed
in the high-dose group. Nb treatment-related effects in pathologic
or histopathologic evaluations were observed. The no-effect level
was considered to be 10,000 ppm (Monsanto Company, 1985).
0 Beagle dogs were fed glyphosate at dietary concentrations of 30, 100
or 300 ppm for 2 years. No evidence of treatment-related chronic or
carcinogenic effects were observed. The no-effect level was considered
to be 300 ppm (Monsanto Company, 1985).
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 FQ generation) were administered test diets
for 60 days prior to breeding. Administration was continued through
mating, gestation and lactation for two successive litters (?iar
Fife). 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 F^ 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 administered to pregnant rabbits by gavage 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.
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Glyphosate August/ 1988
-6-
0 No teratogenic effects were observed in the offspring of rats admini-
stered glyphosate by gavage at dosage levels of 300, 1,000 and 3,500
mgAg/day on days 6 through 19 of gestation. Toxic effects were noted
in the high-dose treated animals and their offspring (Monsanto Company,
1985).
Mutagenicity
0 The Monsanto Company (1985) 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 Njagi and Gopalan (1980) found that glyphosate 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 months, 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
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 (up to 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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Glyphosate August, 1988
-7-
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UP = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 glyphosate. It is, therefore,
recommended that the Ten-day HA value of 20 mg/L be used at this 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 serve 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 mgAg/day on days 6 through 27 of gestation showed effects at
350 mg/kg/day; however, no treatment-related effects were reported at lower
dose levels. The No-Observed-Adverse-Effect-Level (NOAEL) identified in
this study is, therefore, 175 mgAg/day.
Using a NOAEL of 175 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = (175 mg/kg/day) (10 kg) =17.5 mg/L (20,000 ug/L)
(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 EPA
or NAS/OEW guidelines for use with a NOAEL from an
animal study.
1 L/day = assumed daily water consumption of a child.
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Glyphosate August, 1988
-8-
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 (1 mg/L) and 4 mg/L for
an adult 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). 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. 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
three generations. Even though no compound-related changes in the reproductive
indices were observed when compared to controls at a dose level of 30 mg/kg/day,
there were pathological changes of renal focal tubular dilation in male F^b
weanling rats at this level. Therefore, the lower dose level of 10 mg/kg/day
was identified as the NOAEL.
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>1 mg/kg/day
(100)
where:
10 mg/kg/day = NOAEL, based on absence of renal focal tubular
dilation in rats.
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Glyphosate August, 1988
-9-
100 = uncertainty factor, chosen in accordance with EPA
or 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 /L (4/00o 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 (800 ug/L)
where:
4.0 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
o 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, 1986b)].
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
resulting from the use of irrigation water following applications of
glyphosate herbicide around aquatic sites (U.S. EPA, 1985).
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Glyphosate August/ 1988
-10-
VII. ANALYTICAL METHODS
0 A method for glyphosate has been developed for ODW by EMSL-CI
(U.S. EPA, 1988). Other methods are available for glyphosate.
However/ this modification has been directed toward water/ since
glyphosate is used as an aquatic herbicide and the media of concern
is drinking water. In this procedure a water sample is filtered
and a samll aliquot (200 ul) is injected into a referse phase high
pressure liquid chromatograph column. Separation is by isocratic
elution. Post-column derivatization involves oxidation of the
glyphosate to glycine with reaction to o-phthalaldehyde to form a
sensitive fluorophore. The method detection limit for this procedure
for single laboratory validation is 6 ug/L for tap water, and 9 ug/L
for ground water. Across four concentrations with 8 data points for
each type of water (tap water, ground water), recoveries ranged from
95 to 110%, with a relative standard deviation ranging from 2 to 12%.
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, 1988
-11-
IX. REFERENCES
Bababunmi, E.A., 0.0. Olorunsogo and 0. Bassir. 1978. Toxicology of glypho-
sate in rats and mice. Toxicol. Appl. Pharmacol. 45(1):319-320.
Bio/dynamics, Inc.* 1981a. A lifetime feeding study of glyphosate (Roundup
Technical)in rats. 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 77-2063 for Monsanto Co., St. Louis, MO.
EPA accession no. 245909. (Unpublished report) Cited in Monsanto, 1985.
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.
Duerson, C.R., and I. Glenn Sipes. 1987. Absorption of glyphosate in the
male Fischer rat. The Toxicologist. 7(1):Abstract 1986.
Maibach, H.I. 1986. Irritation, sensitization, photoirritation and photo-
sensitization assays with a glyphosate herbicide. Contact Dermatitis.
15:152-156.
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. 1983. Rodeo herbicide: Toxicological and environmental
properties. 800 N. Lindbergh Blvd., St. Louis, MO. Rodeo Herbicide
Bulletin No. 1.
Monsanto Company. 1985. Material safety data sheet, glyphosate technical.
800 N. Lindbergh Blvd., St. Louis, MO. MSDS No. 107-183-6.
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, O.O. 1981. Inhibition of energy-dependent transhydrogenase
reaction by N-(phosphonomethyl)glycine in isolated rat liver mitochondria.
Toxicol. Lett. 10:91-95.
-------
Glyphosate August, 1988
-12-
Olorunsogo, 0.0., 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.
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, 0.0., E.A. Bababunmi and 0. Bassir. 1979a. Effect of glyphosate
on rat liver mitochondria in vivo. Bull. Environ. Contain. Toxicol.
22:357-364.
Olorunsogo, 0.0., E.A. Bababunmi and 0. Bassir. 1979b. The inhibitory effect
of N-(phosphonomethyl)glycine in vivo on energy-dependent, phosphate-
induced swelling of isolated rat liver mitochondria. Toxicol. Lett.
4:303-306.
Reinert/ K.H., and J.H. Rodgers. 1987. Fate and persistence of aquatic
herbicides. Rev. Environ. Contain. Toxicol. 98:61-98.
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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
U.S. EPA. 1985. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.364. July 1.
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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. U.S. EPA Draft
Method — Analysis of glyphosate in drinking water by direct aqueous
HPLC injection with post-column derivatization. EMSL-CI, Cincinnati,
Ohio. June, 1988.
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, 1988
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.
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 carcinooens
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 state^ values. Excess cancer risk
estimates may also be calculated using the One-hit, Weibull, .Legit 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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Hexazinone August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No.; 51235-04-2
Structural Formula:
3-Cyclohexyl-6-(dimethylamino)-l methyl-l,3,5-triazine-2,4(lH,3H)-dione;
Synonyms
0 \felpar; 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, drainaqe ditches, and
sugar and alfalfa (Kennedy, 1984).
Properties (Kennedy, 1984; CHEMLAB, 1985)
Chemical Formula C11H20902N3
Molecular Weight 252
Fhysical State (25°C) Wiite crystalline solid
Roiling Point —
Melting Point 115-117°C
Density —
\fenor Pressure (86°C) 2 x 1(T7 ran ftj
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 not been found in any surface or ground water
samples analyzed from 9 samples taken at 2 locations or 6 samples
from 6 locations, respectively (STORET, 1988).
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Hexazinone August, 1988
-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. !4C-Hexazinone residues
had a half-life of about 25 weeks. Of the extractable ^4C residues,
half was present as parent compound and/or 3-cyclchexyl-l-methyl-6-
methylamino-l,3,5-triazine-2,4-(lH,3H)-<:lione. Also present were
3-(4-hydroxycyclohexyl)-6-(dime thy lamino)-! -methyl -l-(lH,3H)-d lone
and the triazine trione (Rhodes, 1975b).
0 A soil column leaching study used !4C-hexazinone , half of which was
aged for 30 days and applied to Flanagan silt loam and Fallsinrjton
.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, 197 5b).
0 A field soil leaching study indicated that 140-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 egualed 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 dissioation study using a Kevport 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). In a separate study with Keyport silt
loam, seme leaching of the parent compound to a depth of 12 to 18
inches was observed (Holt, 1979).
III. PHARMAOOKINETICS
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 August, 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 (125mq/ka) hexazinone in
the diet to male rats for 17 days. This was followed by a single dose
of 18.3 mg/300 g (61 mgAg) 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 ^C-hexazinone were found in any body
tissues when the animals were administered >14 mo/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-hydroxycyclohexvl)-6-(dimethylamino)l-methyl-
l,3,5-triazine-2,4-(lH,3H)-dione (metabolite A); 3-cyclohexyl-6-
(methylamino)-l-methyl-l,3,5-triazine-2,4-(lH, 3H)-dione (metabolite B);
and 3-( 4-hydroxvcyclohexyl)-6-(methylamino)-1-methyl-l,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 in the urine or the feces. Similar results were reported
by Rapisarda (1982).
Excretion
Rapisarda (1982) and Rhodes et al. (1978) reported that excretion of
l^c-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, 1988
-5-
IV. HEALTH EFFECTS
Humans
The Itesticide 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 wcman 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
a^ort-term Exposure
0 Reported oral LD$Q values for technical -qrade hexazinone in rats ranqe
from 1,690 to >7,500 mq/kq (Matarese, 1977; Cbshiell and Hinckle,
1980; Kennedy, 1984).
Henry (1975) and Kennedy (1984) reported the oral LDsg value of
technical-grade hexazinone in beagle dogs to be >3,400 mq/kg.
Reported oral LD5Q values for hexazinone in guinea piqs range from
800 to 860 rag/kg (Dale, 1973; Kennedy, 1984).
Kennedy (1984) studied the response of male rats to repeated oral
doses of hexazinone (89 or 98% active ingredient). Groups of six
rats were intubated with hexazinone, 0 or 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.
In an 8-week range-finding study (Kennedy and Kaplan, 1984), Charles
River CD-I mice (10/sex/dose) 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 eouals
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 mq/kg/day. Nb
differences ware observed in general behavior and appearance, mortality,
body weights, food consumption or calculated food efficiency between
control and exposed groups. No gross pathologic lesions ware 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
Nb-Observed-Adverse-Effect Level (NOAEL) of 2,500 ppm (375.0 mq/kg/day)
was identified by the authors.
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Hexazinone August, 1988
-6-
Dermal/Ocular Effects
0 In an acute dermal toxicity test performed by McAlack (1976), up to
7,500 mgAg of a 24% aqueous solution of hexazinone (reported to be
1,875 mq/kg of active ingredient) was applied ccclusively for 24
hours to the shaved backs and trunks of male albino rabbits. No
deaths were observed throughout a 14-day observation period. Nb
other symptoms were reported.
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/kq) 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
toxicitv. 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 mgAq/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 EHwards (1977) applied 6,000 mqAg hexazinone as a 63% solution occlu-
sively to the shaved backs and trunks of male albino rabbits. All
rabbits showed moderate skin irritation which cleared 7 days after
cessation of treatment. ND 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. Ihe test material was found to be nonirritating
and nonsensitizing at 48 hours post-application.
0 Using a 10% solution, Goodman (1976) repeated the Morrow (1972) study
with guinea pigs and observed no irritation or sensitization.
0 C&shiell 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. Wien
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 throuqh 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, 1988
-7-
Lonq-term Exposure
0 In a 90-day feedinq study, Sherman et al. (1973) fed beaqle 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 mq/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 NDAEL of 1,000 ppm
(25 mg/kg/day) and a Lowest-Observed-Adverse-Effect Level (LOAEL) of
5,000 ppm (125 mq/kg/day) were identified.
0 In a 90-day feeding study (Kennedy and Kaplan, 1984), Crl-CD rats
(10/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 mq/kg/day (Lehman, 1959), these levels correspond to
about 0, 10, 50 or 250 mg/kg/day. Hematoloqical 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,
corresponded to 0, 11, 60 or 160 mgAg/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.
Ihe 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 mq/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 siqnificant increases
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Hexazinone August, 1988
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in relative liver weight. Sporadic occurrence of statistically
significant changes in hematological effects were considered by
the authors to be unrelated to hexazinone treatment. Histoloqically,
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/dcse) 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 mgAg/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 ware 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 <0.05).
Male rats fed 2,500 ppm had slightly elevated leukocyte counts with
a greater proportion of eosinophils. Male rats fed either 1,000 or
2,500 ppm displayed decreased alkaline phosphatase activity. Statisti-
cally significant effects on organ weights included elevated relative
lung weights in males fed 1,000 ppm; lower kidney and lower relative
liver and heart weights in males fed 2,500 ppm; increased liver and
spleen weights in females fed 200 pom; and elevated stomach and
relative brain weights in females fed 2,500 ppm. At necropsy, gross'
pathologic findings were similar among all groups. Changes attributed
to hexazinone were not apparent in any of the tissues evaluated
microscopically. The authors identified 200 ppm (10 mg/kg/day) as
the NOAEL. However, the increased liver and spleen weights observed
in females would indicate that 200 ppm might be more appropriately
identified as a LOAEL.
Reproductive Effects
0 In a one-generation reproduction study (Kennedy and Kaplan, 1984),
Crl-CD rats (10/sex/dose) received hexazinone (>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 eguals 0.05 mg/kg/day (Lehman,
1959), this corresponds to approximately 6, 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. Ihe 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 (DuRant, 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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Hexazi none August, 1988
-9-
mg/kg/day, assuming the above dietary assumptions for a rat). Pollowing
90 days of feeding, 20 rats/sex/dose were selected to serve as the
parental (FQ) generation. Reproductive parameters tested included
the number of matings, number of pregnancies and number of pups per
litter. Rips were weighed at weaning, and one male and female were
selected fron 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). Ffowever, 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 Tag/kg/day following the previously stated
dietary 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 KDAEL 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 (14-17/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. Eregnancy 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 mgAg/day for maternal toxicity, fetal toxicity, and
teratogenicity can be identified.
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Hexazinone August, 19R8
-10-
Mutaqenicity
0 The ability of hexazinone to induce unscheduled DMA 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 DMA synthesis was observed.
0 \falachos 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/faiL); no significant increases in
clastogenic activity were seen at 1.58, 3.94 and 7.93 mM (0.4, 1.0
and 2.0 mgAiL). With S-9 activation, significant increases in aber-
rations were noted only at a concentration of 15.85 mM (4.0 mg/mL).
Concentrations above these yielded no analyzable metaphase cells due
to cytotoxicity.
0 In a study designed to evaluate the clastogenic potential of hexazinone
in rat bone marrow cells (Farrow et al., 1982), Spraque-Dawley CD rats
(12/sex/dose) were given a single dose of 0, 100, 300 or 1,000 mg/kg
of hexazinone by gavage (vehicle not reported), No statistically
significant increases in the freguency 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 TA100 at concentrations up
to 7,000 ug/plate. The compound was not found to be rautagenic, with
or without S-9 activation (Duftjnt, 1979).
Carcinogenicity
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 centnlobular
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 Crl-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 TDXIOOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 7 years) and lifetime exposures if adeguate data are
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Hexazinone August, 1988
-11-
available that identify a sensitive noncarcinoaenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = ^/L ( uqA)
(UP) x ( L/day)
where:
NOAEL or LQAEL = 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, 1,000 or 10,000),
in accordance with EPA or 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 lonqer-term HA value of 3 mq/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
Ihe study reported by Kennedy and Kaplan (19R4) in which nreqnant rabbits
(14-17/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). Ihe 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 not selected to derive a Ten-day value.
It is, therefore, recommended that the Longer-term HA value of 3 mg/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 (4/sex/dose) 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,
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Hexazinone August, 198ft
-12-
lowered albumin/globulin ratios and slightly elevated liver weights were
observed at the highest dose. A NDAEL of 1,000 ppm (25 mq/kg/day) and a
LQAEL of 5,000 ppm (125 mg/kg/day) were identified. This NOAEL is generally
supported by a 90-day rat feeding study that reported a NOAEL of 50 mq/kg/day
(Kennedy and Kaplan, 1984). Effects in dogs exposed to hexazinone at 50
mg/kg/day have not been reported.
Using a NDAEL of 25 mg/kg/day, the Longer-term HA for a 10-kg child
is calculated as follows:
Longer-term HA = (25 mgAg/day) (10 kg) = 2.5 mgA (3,000 ug/L)
(100) (1 L/day)
where:
25 mg/kg/day = NDAEL, 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 EPA
or NAS/ODW guidelines for use with a NDAEL fron 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/kq/day) (70 kg) = 8.75 ^/L (9,000 UqA)
(100) (2 L/day)
where:
25 mg/kg/day = NDAEL, 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 EPA
or NAS/ODW guidelines for use with a NDAEL frcm 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-
cincgenic 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). Ihe RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
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Hexazinone August, 1988
-13-
appreciable risk of deleterious effects over a lifetime, and is derived fron
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). Fran the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A CWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure frcm 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 frcm drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed. 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. ND 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) = Q.03 mg/kq/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 EPA
or 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/kq/day) (70 kg) - 1.05 mq/L (1,000 uq/L)
(2 L/day)
where:
0.03 mg/kg/day = RfD.
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Hexazinone August, 1988
-14-
70 kg = assumed body weight of an adult.
2 [/day = assumed daily water consumption of an adult.
Step 3: Determination of Lifetime Health Advisory
Lifetime HA = (1.05 mg/L) (20%) = 0.21 mg/L (200 uq/L)
where:
1.05 mg/L = EWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 No evidence of carcincgenicity 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 D: not
classified. This category is for substances with inadequate or no
animal evidence of carcincqenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 Residue tolerances range from 0.5 to 5.0 opm 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 qas 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
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 chromatoqraphy,
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 August, 1988
-15-
IX. REFERENCES
CHEMLAB. 1985. The Chemical Information System, CIS, Inc. Baltimore, MD.
Dale, N.* 1973. Oral LD$Q 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 11)50 test in rats—EPA proposed
guidelines. Haskell Laboratory Report No. 943-80, unpublished study.
MRID 00062980.
DuEbnt.* 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-l-methyl-6-dimethylamino (INA 3674, hexazinone).
Accession No. 97323.
Edwards, D.F.* 1977. Acute skin absorption test on rabbits LD5Q. Haskell
Laboratory Report No. 841-77, unpublished study. MRID 00091140.
Farrow, M, T. Cartina, M. Zito et. al.* 1982. In vivo bone marrow cytoqenetic
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 DMA 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 testicide 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 Prcqrams. 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(lH,3H)dione, 3-cyclohexyl-l-
methyl(-6-dimethylamino (INA-3674). One-year Interim Report. Haskell
Laboratory Report No. 585-75. MRID 00078045.
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Hexazinone August, 1988
-16-
Kennedy, G.L. 1984. Acute and environmental toxicity studies with hexazinone.
Fund. Appl. Ibxicol. 4:603-611.
Kennedy, G.L. , and A.M. Kaplan. 1984. Chronic toxicity, reproductive, and
teratogenic studies of hexazinone. Fund. Appl. Toxicol. 4:960-971.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, druqs, and
cosmetics. Assoc. Food Drug Off. of the U.S.
Matarese, C.* 1977. Oral LD5Q test. Haskell Laboratory Report No. 1037-77,
unpublished study. MRID 0011477.
McAlack, J.W.* 1976. Skin absorption LDso« 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 ouinea
pigs. Haskell Laboratory Report No. 489-72, unpublished study. MRID
00104978.
Morrow, R.* 1973. Skin absorption toxicity ALD and skin irritancy test.
Haskell Laboratory Report No. 503-73, unpublished study. MRID 00104974.
Rapisarda, C.* 1982. Metabolism of 14C-labeled hexazinone in the rat. E.I.
duPont de Nemours & Co. Document No. AMR-79-82, unpublished study.
Accession No. 247874.
Fhodes, Robert C. 1975a. Studies with "\felpar" weed killer in water.
Biochemicals Department Experimental Station, E. I. duPont de Nemours
& Co., Inc., Wilmington, DE.
Rhodes, Robert C. 1975b. Decomposition of "^felpar" weed killer in soil.
Biochemicals Department Experimental Station, E. I. duPont de Nemours
& Go., Inc., Wilmington, DE.
Rhodes, Robert C. 1975c. Mobility and adsorption studies with "
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 \felpar (R) waed
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. Ihree month feeding study in
dogs with sym-triazine-2,4(lH,3H)-dione, 3-cyclohexyl-l-methy(-6-dimethyl-
amino-(INA-3674). Haskell Laboratory Report No. 408-73. MRID
SIORET. 1988. SIORET Vbter Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
-------
Hexazinone August, 1988
-17-
U.S. EPA. 1981. U.S. Environmental Protection Agency. Pesticide Incident
Monitoring System. Office of Pesticide Programs, Washington, DC.
February-
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. Gate of Federal
Regulations. 40 CFR 180.396.
U.S. EPA. 1985b. 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.
VLachos, D, J. Martenis and A. Hbrst.* 1982. In vitro assay for chromosome
aberrations in Chinese Hamster Ovary (CHO) cells. Haskell Laboratory
Report No. 768-82, unpublished study. MRID 00130709.
*Confidential Business Information submitted to the Office of Ifesticide
Programs.
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August, 1988
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>
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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Maleic Hydrazide August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 123-33-1
Structural Formula
OH
0
^N
NH
1r2-Dihydro-3,6-pyridazinedione
Synonyms
0 Antergon; Chemform; De-Sprout; MH; Retard; Slo-Gro; Sucker-Stuff;
(Meister, 1983).
Uses
• Plant growth retardant (Meister, 1983).
Properties (Meister, .1983; CHEMLAB, 1985; TOB, 1985)
Chemical Formula C4H402N2
Molecular Weight 112.09
Physical State (25«C) Crystalline solid
Boiling Point
Melting Point 292°C
Density 1.60
Vapor Pressure (50«C) 0 mm Hg
Specific Gravity
Water Solubility (25«C) 6,000 mg/L
Log Octanol/Water Partition -3.67 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
• No information was found in the available literature on the occurrence
of maleic hydrazide.
Environmental Fate
• 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 (Registration Standard Science Chapter for Maleic Hydrazide;
WSSA, 1983).
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Maleic Hydrazide August, 1988
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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 (Registration Standard Science Chapter
for Maleic Hydrazide; 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 (Registration Standard Science Chapter
for Maleic Hydrazide; 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, suggesting that 88% had been absorbed.
Distribution
0 Kennedy and Keplinger (1971) administered 14C-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 suggest 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 ethereal 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, 1988
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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 mg/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 LD5Q values were calculated to be 6,300 mg/kg for males, 6,680
mg/kg 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
(Uhiroyal Chemical, 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 salts (purities not specified) of maleic
hydrazide (Tate, 1951). The diethanolamine salt showed an LDsg
value of 2,350 mg/kg, while the LDsg 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, 1988
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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.
• 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 given),
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,
interstitial pneumonia, loss of hair and significant reduction in
weight gain compared with the controls (at 4 months, controls had
gained 30%; those fed maleic hydrazide 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.
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Maleic Hydrazide August, 1988
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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 year. 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 hydrazide in the diet of
rats. On this basis, a Lowest-Observed-Adverse-Effect 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(SO)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 rag/kg/day. No adverse effects on reproductive indices were
observed at any dietary level. However, increased mortality was
observed in F^ parents that received 30,000 ppm. Also at this dose
level, body weights were reduced in Fg parents during growth and
reproduction and in FI 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.
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Maleic Hydrazide August, 1988
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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, fetotoxrcity
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. An 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
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
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Maleic Hydrazide August, 1988
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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/Uay 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
100 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 mg, was not mutagenic in
Salmonella typhimurium (TA 1530). However, two formulations (MH-30
and Royal MH), at 50, 100 and 200 uL (undiluted), were highly mutagenic
in this system.
0 Moriya et al. (1983) reported that maleic hydrazide was not mutagenic
in five strains of S. typhimurium.
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Maleic Hydrazide August, 1988
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0 Ercegovich and Rashid (1977) observed a weak mutagenic response with
maleic hydrazide (purity not specified) in five strains of S_. 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 mg/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 normochromatic 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
chroraated 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/raL) 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.
Carci nogeni city
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
(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 maximum tolerated doses. No evidence
of increased tumor frequency was detected in gross or histologic
examination.
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Maleic Hydrazide August, 1988
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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 rag/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 rag/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 Uhiroyal 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.
Uhiroyal Chemical (1971) reported a 2-year study 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
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
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Maleic Hydrazide August, 1988
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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 (up to 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, 1,000 or 10,000),
in accordance with EPA or 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-OEA salt. For this reason, studies involving MH-DEA have not been consid-
ered as candidates in calculating HA values for maleic hydrazide.
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.
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Maleic Hydrazide August, 1988
-12-
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-mg/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 NOAEL of 100 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) = 10 m /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 EPA
or 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 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
NOAEL values of 2,500 and 500 mg/kg/day, respectively. These studies have
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Maleic Hydrazide August, 1988
-13-
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 Hater Equivalent Level (DUEL) of 18 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. 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 maleic
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 mg/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:
Step 1: Determination of the Reference Dose (RfD)
(500 mg/kg/day) = 0.5 mg/kg/day
(1,000) y
where:
500 mg/kg/day = LOAEL, based on proteinuria in rats exposed to maleic
hydrazide in the diet for 28 months.
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Maleic Hydrazide August, 1988
-14-
1,000 = uncertainty factor, chosen in accordance with EPA
or NAS/OEW guidelines for use with a LOAEL from an
animal study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-5 mgAg/day) (70 kg) =17.5 mg/L (18,000 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 (4,000 ug/L)
where:
1 7. 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 orally 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.
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.
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Maleic Hydrazide August, 1988
-15-
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 alkaline solution
to drive off volatile basic interferences. Distillation with zinc
and nitrogen sweep expels hydrazine liberated from maleic hydrazide.
Hydrazine is reacted in acid solution with p-dimethylaminobenzaldehyde
to form a yellow compound which 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, 1988
-16-
IX. REFERENCES
Aldridge, D.* 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.O. Passey, W.S. Bullough,
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. MRID 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.6. 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.
42:1101-1114.
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 Uhiroyal 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.
-------
Maleic Hydrazide August, 1988
-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. 15:122. 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.
Morlya, 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.
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.
Sabharwal, 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.
-------
Maleic Hydrazide August, 1988
-18-
Shapiro, R.* 1977a. Acute oral toxiclty: 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
DMA-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.D.* 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, J., I. Schmeltz and D. Hoffmann. 1979. Hydrazines as mutagens in a
histidine-requirin^ 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.
U.S. FDA. 1975. U.S. Food and Drug Administration. Pesticide analytical
manual. Vol. II. Washington, DC.
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Male.ic Hydrazide August, 1988
-19-
Van Der Heijden, C.A., E.M. Den Tonkelaar, J.M. Garbis-Berkvens and G.J. Van
Esch. 1981. Maleic hydrazide, carcinogenic!ty 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, 1988
MCPA
(4-Chloro-2-Methylphenoxy)-Acetic 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 enforceaole Federal standards. The HAS are subject to
change as new nformation 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
exposjres based on data describing noncarcinogenic end points of toxicity.
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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MCPA August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 94-74-6
Structural Formula
•OCHjCOOH
(4-Chloro-2-methylphenoxy)-acetic acid
Synonyms
0 MCPA; MCP; Agroxone; Hormotuho; Metaxon.
Uses
0 MCPA is a hornone-type herbicide used to control annual and perennial
weeds in cereals, grassland and turf (Hayes, 1982).
Properties (CHEMLAB, 1985; Meister, 1983)
Chemical Formula
i-lolecular Weignt 200.63
Physical State (25°C) Light brown solid
Boiling Point.
Melring Point 118 to 119°C
Vapor Pressure (25°C)
Density (25°C) 1.56
Water Solubility 825 mg/L (20° C)
Log Octanol/Water Partition 2.07 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
MCPA has been found in 4 of 18 surface water samples analyzed and in
none of 118 ground water samples (STORET, 1988). Samples were collected
at 13 surface water locations and 117 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 range of concentrations was
0.04 to 0.54 utj/L. This information is provided to give a general
impression of the occurrence of this chemical in ground and surface
waters as reported in the STORET database. The individual data points
retrieved were, used as they came from STORET and have not been confirmed
as to their validity. STORET data is often not valid when individual
numbers are used out of the context of the entire sampling regime, as they
are here. Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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MCPA August, 1988
-3-
0 MCPA nas been detected in groundwater in Montana. The highest level found
was 5.5 ppb (5.5 ug/L).
1 Frank et al. (1979) detected MCPA residues (1.1 to 1000 ppb) in 2 of
237 wells in Ontario, Canada, between 1969 and 1978.
Environmental Fate
0 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
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 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-o-cresol, is formed at low levels
(1.3% or less) within 1 day of treatment.
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.
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MCPA August, 1988
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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 Ylitalo (1979) treated rats with 8 mg of 14OMCPA [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.
° Elo and Ylitalo (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/kcj) 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, 1988
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IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 Case reports of attempted suicide by ingestion of MCPA have been
published (Jones et al., 1967; Johnson et al, 1965; Geldmacher et
al., 1966). Symptoms included pinpoint pupils, diminished/absent
reflexes, low blood pressure, spasms, unconsciousness, and death.
Dose estimates were not reported.
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. The farmer exhibited reversible aplastic anemia,
muscular weakness, hemorrhagic gastritis and signs of slight liver
damage that were later followed by pancytopenia of all of the myeloid
cell lines. In a followup study of the exposed farmer, Timonen and
Palva (1930) reported the occurrence of acute myelomonocytic 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 Gurd et al. (1965) reported an acute oral LDjg value for MCPA (purity
not specified) of 560 mg/kg in mice. Oral LDjg's for MCPA in mice
of 550 mg/kg and 700 mg/kg/day in rats were reported in RTECS (1985).
0 Elo et al. (1932) 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 1*C-MCPA,
14C-PABA, 14C-sucrose, 14c-antipyrine and lodinated 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 the
iso-octyl ester of MCPA (purity not specified) at doses of 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 LDjQ values for MCPA (purity not
specified) in rabbits of 4.8 g/kg for males and 3.4 g/kg for females.
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MCPA August, 1988
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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
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 NCPA (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 (l0/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), these levels correspond 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
retardation and elevated relative kidney weights at 400 ppm (20
mgAg/day) or more. 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 technical (considered to
be 100% a.i.) in the diet of Charles-River CD rats (10/sex/dose) for
3 months. Doses were reported as 0, 4, 8 or 16 mg/kg/day; the concen-
tration 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-tobody 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 nig/kg/day
were identified by this study.
0 Holsing and Kundzin (1968) administered oral doses of MCPA technical
to Charles-River CD rats at dose levels of 0, 25, 50, and 100 mg/kg/day
for 13 weeks. Cytopathological changes in the liver and kidneys were
observed at all doses. Kidney effects included focal hyperplasia of
the epithelial 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
beagle 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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MCPA August, 1988
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0 He11wig (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 mg/kg (Lehman, 1959), these levels
correspond to doses of 0, 0.15, 0.75 or 3.75 mg/kg/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
kidneys 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.) by capsule at 0, 25, 50 or 75 mg/kg/day to beagle dogs
(three/sex/dose) for 13 weeks. Histopathological changes and altera-
tions 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 transanunase, serum-oxaloacetic transaminase,
alkaline phosphatase and serum bilirubin. Histopathological alterations
were seen in the liver, kidney, lymph nodes, testes, prostate and
bone narrow. 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. (1955) 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 tne diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels
correspond 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/kj/day). At 400 ppm (20 mg/kg/day) or
greater, there was a reduction in numbers of red blood cells, hemo-
globin 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), these levels
correspond to doses of 0, 2.5, 7.5 or 22.5 mg/kg/day. Body weight
depression was observed in the F1 and F2 generations at the two
highest doses. A NOAEL of 22.5 mg/kg/day was identified for reproductive
function, and a NOAEL of 2.5 mg/kg/day was identified for fetoxtoxicity
(depressed weight gain).
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MCPA August, 1988
-8-
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
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/kg/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 Lovell (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 TOO 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
(TH 100, TA 9b, TA 1535, TA 1537, TA 1538) or in Escherichia coli
(WP2 her) either with or without -netabolic activation.
0 In studies conaucted by Magnusson et ai. (1977), there were no
effects on chromosome disjunction, loss or exchange in Drosophila
fed MCPA at 250 or 500 ppm.
° 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, 10~4,
10~3M, 1 hour) with and without activation.
Carcinogenicity
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).
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MCPA August, 1988
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V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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.
Brt = assumes body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 tne 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-kg 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
Several reproductive/developmental toxicity studies have been performed
in which rats or raboits 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 toxicity observed in these studies was a reduction
in body weight in rats exposed to 75 mg/kg (Irvine, 1980). Estimates of mater-
nal NOAELs range from 30 to 125 mg/kg/day (Irvine, 1980; Irvine et al, 1980).
In contrast, fetotoxicity has been observed at dose levels as low as 7*5
mg/kg/day (MacKenzie, 1986). These studies were judged to be inadequate for
evaluating the toxicity of MCPA from acute oral exposure, especially with
respect to kidney toxicity observed after longer durations of exposure.
It is therefore recommended that the Longer-term HA value for a 10-kg 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
Hendriksen, 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
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MCPA August, 1988
-10-
25 mg/kg/day (Holsing, 1968) and 3 mg/kg/day (Reuzel and Hendriksen, 1986).
Renal toxicity is not unique to dogs and has been observed in rats after
90-day exposure at dose levels of 20 mg/kg/day (Verschuuren et al., 1975) and
25 mg/kg/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 Hendriksen (1980). In these studies, oral doses of 0, 3, 12 or
48 mg/kg/day, and 0, 0.3, 1 or 12 mg/kg/day, respectively, were administered
to dogs for 13 weeks. Increases in blood urea, SGPT and creatinine levels
were observed at dose levels as low as 3 mg/kg/day; low prostatic weight and
mucopurulent conjunctivitis were observed at higher dose levels. A NOAEL of
1 mg/kg/day was identified by these studies.
Using a NOAEL 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 kg) _ 0.1 mg/L (100 ug/L)
(100) (1 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.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA of
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 mg/kg/day) (70 kg) = 0>4 mg/L (400 ug/L)
(100) (2 L/day)
where:
1.0 mgAg/day = NOAEL, based on the absence of renal effects in dogs
exposed to NCPA in the diet for 90 days.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA or
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-
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MCPA August, 1988
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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. 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 tor tne 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 EPA or
NAS/ODW guidelines for use with a NOAEL from an
animal study.
3 = additional uncertainty factor, used to account for
the incomplete database on chronic toxicity.
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MCPA August, 1988
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Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.0005 mg/kg/day) (70 kg) = 0.02 mg/L (20 ug/L)
(2 L/day)
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 = (0.02 mg/L) (20%) = 0.004 mg/L (4 ug/L)
where:
0.02 = 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) classified
the potential carcinogenicity of MCPA in both humans and laboratory
animals as 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. Tnis 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/kg/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
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MCPA August, 1988
-13-
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
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.
VIII. TREATMENT TECHNOLOGIES
0 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, 1988
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Helling, C.S. 1971. Pesticide mobility in soils. II. Application of soil
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MacKenzie, K.M. 1986. Two-generation reproductive study with MCPA in rats.
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of proanide, yalan, MCPA and 2,4-D in rice irrigation systems of the
Kuban River. Agrokhimiya (Agricultural Chemistry). 3:95-106.
STORET. 1988. STORET Mater Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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MCPA August, 1988
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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, F., 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. Peltonen. 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.M. den Tonkelaar. 1975. Short-term oral
and dermal toxicity of MCPA and MCPP. Toxicology. 3:349-359.
*Confidential Business Information submitted to the Office of Pesticide
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August/ 1988
METHYL PARATHION
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
cpntamination 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.
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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Methyl Parathion
August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 298-00-0
Structural Formula
-PCOCH,),
0,0-Dimethyl-0-(4-nitrophenyl) phosphorothioic acid
Synonyms
0 Metaphos; Dimethyl parathion; Folidol M; Metocide; Penncap M; Sinafid
M-48; Wofatox; Cekuraethion; Devithion; Drexel Methyl Parathion 4E;
E601; Fosferno M50; Gearfos; Parataf; Partron-M; Tekwaisa; Parathion-
methyl (Meister, 1988).
Uses
0 A restricted-use pesticide for control of various insects of economic
importance; especially effective for boll weevil control (Meister, 1988.
Properties (Hawley, 1981; Meister, 1988; 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 (25°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
263.23
white crystalline solid
35 to 36»C
0.97 x 10~5 mm Hg
55 to 60 mg/L
3.11 (calculated)
Methyl parathion has been found in 1,070 of 27,082 surface water
samples analyzed and in 8 of 2,836 ground water samples (STORET,
1988). Samples were collected at 3, 558 surface water locations and
2,111 ground water locations, and methyl parathion was found in 20
states. The 85th percentile of all nonzero samples was 1.18 ug/L
in surface water and 0.05 ug/L in ground water sources. The maximum
concentration found was 13 ug/L in surface water and 0.05 ug/L in
ground water. This information is provided to give a general impression
of the occurrence of this chemical in ground and surface waters as
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Methyl Parathion August, 1988
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reported in the STORET database. The individual data points retrieved
were used as they came from STORET and have not been confirmed as to
their validity. STORET data is often not valid when individual
numbers are used out of the context of the entire sampling regime/ as
they are here. Therefore, this information can only be used to form
an impression of the intensity and location of sampling for a particular
chemical.
Environmental Fate
0 Methyl parathion (99% pure) at 10 ppra 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, silty 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.
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Methyl Parathion August, 1988
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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-iabeled
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 faces was low, never exceeding 10% of the
dose. This indicated that absorption was at least 90% complete.
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
8 Hollingworth et al. (1967) gave 32p-iabeled 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 desmethyl
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 Neal 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 <0.01). The activity of weanling rat liver (2.7 units)
was intermediate between these two. In the case of adult CF-1 mice,
the activity of female liver (3.2 units) was significantly greater
(p <0.05) than that of the males (2.3 units). The activity of young
adult male guinea pig liver extracts was 5.6 units. The authors noted
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Methyl Parathion August, 1988
-5-
that these differences in metabolic detoxification rates correlated
with the sex and species differences in susceptibility to the acute
oral toxic effects of this family of compounds.
Nakatsugawa et al. (1968) investigated the degradation of methyl
parathion using liver microsomes from adult male rats and rabbits
(strains not specified). Metabolism occurred by two oxidative pathways:
activation of the phosphorus-sulfur bond to the phosphorus-oxygen
analog, and cleavage at the aryl phosphothioate bond to yield p-nitro-
phenol. These reactions occurred only in the presence of oxygen and
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-labeled 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
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 10 mg 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
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Methyl Parathion August, 1988
-6-
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, -:it the authors
cautioned that it cannot be assumed that they are less hazardous than
the ultra-low-volume spray residues.
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.
Thirty-two different dosages were evaluated by Rider et al. (1969),
ranging from 1 to 19 mg/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 mg/day. At this point, the dose was
increased by 1.0 mg/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.
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 rag/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.
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 mg/day, a mean maximum depression
of 37% occurred. Based on their criteria of 20 to 25% average
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Methyl Parathion August, 1988
-7-
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 ing/day corresponds to a No-Observed-Adverse-Effect Level
(NOAEL) of 0.31 mg/kg/day, and the 30 mg/day dose corresponds to 0.43
mgAg/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.
Animals
Short-term Exposure
0 Reported oral LD50 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-1 mice, respectively (Haley et al.,
1975); 30 mg/kg in male ddY mice (Isshiki et al., 1983); 18.0 and
8. 9 mgAg in male and female Sprague-Dawley rats, respectively (Sabol,
1985); and 9.2 mg/kg in rats of unreported strain (Galal et al., 1977).
9 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-LD5Q 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).
9 Daly et al. (1979) administered methyl parathion (technical, 93.65%
active ingredient) to Charles River CD-1 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
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Methyl Parathion August, 1988
-8-
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
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 LD$Q dose of active S^ 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 NZISI 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 imounobiological
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 LDsg of 67 mg/kg for methyl parathion
in male and female Sherman rats.
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Methyl Parathion August, 1988
-9-
0 Galloway (1984a,b) studied the skin and eye irritation properties of
methyl 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.
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 LD5g
was 1,200 mg/kg. The ED$Q was 950 mg/kg for cholinesterase inhibition
and 550 mg/kg for acetylcholinesterase inhibition.
Long-term Exposure
0 Daly and Rinehart (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).
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Methyl Parathion August, 1988
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Tegeris and Underwood (1978) investigated the toxicity of methyl
parathion (94.32% a.i.) 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 rag/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 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.
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.
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 B6C3FL
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
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Methyl Parathion August, 1988
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about 5.2 or 11.5 mg/kg/day, respectively. Females were fed at the
original levels throughout. Mortality 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.
Daly et al. (1984) conducted a chronic feeding study of methyl
parathion (93.65% active ingredient) in Sprague-Dawley (CD) 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
and quantitative tests for neurotoxicity. Qphthalmoscopic 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,
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Methyl Parathion August, 1988
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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 rag/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 Fla, Flb and F2a groups at
30 ppm, and in weanlings of the F^a. group at 10 ppm. At 30 ppm,
there was also a reduction in fertility of the F2b dams at the second
mating; the first mating resulted in 100% of the animals having
litters, while at the second mating, only 41% had litters. Animals
exposed to 10 ppm methyl parathion 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 15 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
FL dams at the 25-ppm dose level. A slight decrease in body weight
was noted in F^a 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 20
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 acetyltransferase (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,
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Methyl Parathion August, 1988
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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 prenatally 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
behaviors of the offspring. The fetotoxic LOAEL for this study was
1.0 mg/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 15 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.
0 Fuchs et al. (1976) reported a study in which oral administration of
methyl parathion to pregnant Wistar rats on either days 5 to 9, 11 to
15, or 11 to 19 of gestation resulted in growth retardation of
offspring and increased resorptions at 3 mgAg* The NOAEL was 1
mg/kg.
Mutaqenicity
9 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
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Methyl Parathion August, 1988
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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 LD50.
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 E_. 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.
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 B35M 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 07 of Saccharomyces
cerevisiae), based on mitotic recombination (in 03), and mitotic
crossing over, mitotic gene conversion, and reverse mutation (in 07).
0 In a study for dominant lethality in mice by Jorgenson et al. (1976),
males (20 per dose group) were given methyl parathion in the diet for
7 weeks at 3 dose levels (not reported). Positive controls given
triethylene melamine and untreated controls were also studied.
Following treatment, each male was mated to 2 adult females weekly
for 8 weeks. Methyl parathion was ineffective in this test.
Carcinogenicity
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.
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Methyl Parathion August, 1988
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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 pprn, 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 (up to 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) = m /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, 1,000 or 10,000),
in accordance with EPA or 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 (0.3 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.
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Methyl Parathion August, 1988
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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 mg/kg/day) (10 kg) = 0.31 mg/L (300 ug/L)
(10) (1 L/day)
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 EPA or
NAS/ODW 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 mgAg/day was identified,
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 10-kg child is
calculated as follows:
Longer-term HA - (0*3 mq/kg/day) (10 kg) a 0.03 mg/L (30 ug/L)
(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 EPA or
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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Methyl Parathion August, 1988
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Using a NOAEL of 0.3 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
Longer-term HA = (0-3 mg/kg/day) (70 kg) = 0.10 mg/L (100 ug/L)
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.
70 kg = assumed body weight of an adult.
100 =• uncertainty factor, chosen in accordance with EPA or
NAS/ODW 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). 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. 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
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Methyl Parathion August, 1988
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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
mg/lcg/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) = Q.00025 mg/kg/day
(100) y * *
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.
100 = uncertainty factor, chosen in accordance with EPA
or 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/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).
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Methyl Parathion August, 1988
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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
mgAg/day, using a NOAEL of 0.043 mg/kg/day in humans (Rider et al.,
1969) and an uncertainty factor of 10 (NAS, 1977). From this ADI,
NAS calculated a chronic Suggested-No-Adverse-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.
0 The U.S. EPA Office of Pesticide Program (EPA/OPP) previously calcu-
lated a provisional ADI (PADI) of 0.0015 mgAg/day, based on a NOAEL
of 0.3 mgAg/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 mgAg/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/m3 (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
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Methyl Parathion August, 1988
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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 methyl parathion, 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.
0 Under laboratory conditions, GAC removed 99+% of methyl parathion
(Whittaker et al., 1982).
0 Reverse osmosis is a promising treatment method for methyl parathion-
contaminated water. Chian (1975) reported 99.5% removal efficiency
for two types of membrane 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, as breakthrough of methyl parathion would probably
occur once the adsorption potential of the membrane was exhausted.
0 Oxidation with ozone and chlorine may be possible in the treatment of
methyl parathion.
0 Oxidation with 4.5 and 9.5 mg/L ozone reduced the methyl parathion by
95 to 99%. The same removal efficiency was achieved with 1 and 2 mg/L
chlorine (Gabovich and Kurennoy, 1974).
0 Ozonation with 0.32 mg ozone/mg methyl parathion reduced methyl
parathion in drinking water by 90 to 95% (Shevchenko et al., 1982).
0 Oxidation degradation by either ozone or chlorine produces several
degradation products, whose environmental toxic impact should be
evaluated prior to selecting oxidative degradation for treatment of
methyl parathion-contaminated water (Shevchenko et al., 1982).
0 Aeration does not seem to be a practical technique for removing
methyl parathion from potable water (Saunders and Sieber, 1983).
0 Treatment technologies for the removal of methyl parathion from water
are available and have been reported to be effective. However,
selection of individual or combinations of technologies for methyl
parathion removal from water must be based on a case-by-case technical
evaluation, and an assessment of the economics involved.
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Methyl Parathion August, 1988
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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)
Agcheou* 1983. Persistence and release rate of Penncap M insecticide in
water and hydrosoil: Project No. WT-5-82. Unpublished study.
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Fan, A., J.C. Street and R.M. Nelson. 1978. Immunosuppression in mice
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Haley, T.J., J.H. Farmer, J.R. Harmon and K.L. Dooley. 1975. Estimation of
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Hawley, G.G. 1981. The Condensed Chemical Dictionary, 10th ed. NY: Van
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Hollingworth, R.M., R.L. Metcalf and I.R. Fukuto. 1967. The selectivity of
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Jorgenson, T.A., C.J. Rushbrook, and G.W. Newell. 1976. In vivo mutagenesis
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Lehman, A.j. 1959. Appraisal of the safety of chemicals in foods, drugs and
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Methyl Parathion August, 1988
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cides—soil leaching. Unpublished study.
Rashid, K.A., and R.O. Mununa. 1984. Genotoxicity of methyl parathion in
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Riccio, E., G. Shepherd/ A. Pomeroy, K. Mortelmans and M.D. Waters. 1981.
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Methyl Parathion August/ 1983
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August, 1988
METHOMYL
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.
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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Methomyl August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 16752-77-5
Structural Formula
0
CHj-C'N-0-C-N-CH,
S-CH, H
S-Methyl-N [ (methylcarbamoyl ) oxy ] -thioacetimidate
Synonyms
0 Dupont Insecticide 1179; Dupont 1179; Insecticide 1,179; Insecticide
1179; IN 1179, Lannate; Hesomile; Nudrin; SD 14999; WL 18236 (Meister,
1983).
Uses
0 Methomyl 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
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"^ mm Hg
Specific Gravity 1.29
Water Solubility (25°C) 10,000 mg/L
Log Octanol/Water Partition -3.56
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Methomyl has been found in 1 of 423 surface water samples analyzed
at a concentration of 2 ug/L, and was not found in 1,374 ground water
samples analyzed (STORET, 1988). Samples were collected at 145
surface water locations and 1,326 ground water locations. This
information is provided to give a general impression of the occurrence
of this chemical in ground and surface waters as reported in the
STORET database. The individual data points retrieved were used as
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Methorny1 August, 1988
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they came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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).
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 virtually
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 1-14C-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,
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Methomyl August, 1988
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purity not specified) to rats and analyzed 13 major tissues for residues
at 1 and 3 days after dosing. Only 10% of the label was present in
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 methomyl 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 1-14C-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, 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
Baron (1971) stated that within 72 hours after receiving a single
oral dose of 1-14C-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.
Harvey (1974) reported that 75% of an oral dose of 1-14C-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
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Methomyl August, 1988
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air. In contrast to other carbamates, sulfur-containing metabolites
were not found in the urine.
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/k
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Methomyl August, 1988
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the plant was 2 years. Packaging workers had the highest rate of
adverse symptoms: small pupils (46%), nausea and vomiting (46%),
blurred vision (46%) and increased salivation (27%). Biomedical
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 LDgg reported for methomyl in the fasted male and female
rat ranged from 17 to 25 mg/kg (Bedo and Cieleszky, 1980; Dashiell
and Kennedy, 1984; Kaplan and Sherman, 1977). The oral LD5Q in the
nonfasted rat was 40 mg/kg (Dashiell and Kennedy, 1984). Clinical signs
in rats included chewing motions, profuse salivation, lacrimation,
bulging eyes, fasciculations and tremors characteristic of ChE inhibition.
0 The acute oral LD50 for methomyl in the mouse ranged from 27 to 35 mg/kg
(Boulton et al., 1971; El-Sebae et al., 1979).
0 The oral LOgg in hens was 28 mg/kg and in Japanese quail, 34 mg/kg.
(Kaplan and Sherman, 1977).
0 The 4-hour inhalation LC-0 of methomyl in rats was 300 mg/m . Animals
showed the typical signs of ChE inhibition, including salivation,
lacrimation and tremors (ACGIH, 1984).
0 Bedo and Cieleszky (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 tremors in
rats, and brain ChE levels were decreased. In the liver, mixed-function
oxidase and glucose-6-phosphatase activity, and levels of glycogen
and vitamin A were unaffected. Apparently, dose levels of 2 or 3 mg/kg
did not produce these effects, but did result in increased activities
of chyrcotrypsin, lipase, and amylase in pancreatic juice.
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 Cieleszky (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 levels
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Methomyl August, 1988
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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).
0 Kaplan and Sherman (1977) administered methomyl (90% pure) to six
male Charles River 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 histopathologic 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
tolerated 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 Rabbits (5/sex) 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 Charles
River-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), these levels correspond 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 values were
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Nethonyl August, 1988
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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 marrow. Microscopic examination of all other tissues showed
no consistent abnormalities. Based on these observations, this study
identified a NOAEL of 50 ppm (2.5 mg/kg/day) and a LOAEL of 250 ppm
(12.5
0 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 1 3 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),
these levels correspond 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 methomyl were found during or at the conclusion of the study.
Based on these data, a NOAEL of 10 mg/kg/day was identified.
0 Homan 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, but absolute kidney weights were not significantly
increased. Red blood cell ChE activity was elevated at the high dose
levels, but plasma and brain ChE levels were normal at all dose
levels. Histopathological 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), these levels correspond 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 NOAEL of 100 ppm (5 mg/kg/day).
0 Kaplan and Sherman (1977) reported a 22-month dietary feeding study
in which Charles River-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), these levels correspond to doses of about
0, 2.5, 5, 10 or 20 mg/k9~/day. Mortality data were not reported. At
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Methomyl August, 1988
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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
red blood cells. • This study identified a LOAEL of 200 ppra
(10 mg/kg/day) and a NOAEL of 100 ppm (5 mq/kq/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), these levels correspond 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 urine analyses 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 3
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, incoor-
dination 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 methomyl (99%
purity) 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), these levels correspond to doses of about 0, 7.5, 15
or 120 mg/kg/day. Survival was significantly reduced (no 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.
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Methorny1 August, 1988
-10-
0 Haskell Laboratories (1981) reported a 2-year study of methomyl (99%
purity) in rats. Charles River C-D rats (100/sex/dose) were fed
methomyl in the diet at dose levels of 0, 50, 100 or 400 ppm for
up to 2 years. Assuming 1 ppm in the diet is equivalent to 0.05
rag/kg/day (Lehman, 1959), these levels correspond to doses of 0,
2.5, 5 or 20 mg/kg/day. Effects seen in the 400 ppm group include
reduced weight gain in both sexes, and in females, lower erythrocyte
counts, hemoglobin values and hematocrits. Blood and brain cholinester-
ase levels were within the range as were other clinical chemistry
parameters. A NOAEL of 5 mg/kg/day and a LOAEL of 20 mg/kg/day were
idnetified from this study.
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.
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 F^ 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), these
levels correspond to doses of about 0, 1.5 or 3 mg/kg/day. One-third
of the fetuses were stained with Alizarin Red S and examined for
skeletal defects. 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 100, 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
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Methorny1 August, 1988
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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 Haskell Laboratories (1981) report on a study in which Charles River
C-D rats (100/sex/group) were administered methomyl (99% pure) in the
diet at dietary levels of 0, 50, 100 or 400 ppm for two years. Gross
and histological examination revealed no increase in tumor incidence
in either sex.
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
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 (up to 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, 1,000 or 10,000),
in accordance with EPA or 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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Methomyl August, 1988
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One-Day Health Advisory
No information found in the available literature was suitable for deter-
mination of the One-day HA value for methorny1. It is, therefore, recommended
that the Longer-term HA (0.3 rag/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 administered methomyl via
gavage at doses 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, equivalent to a dose of 5 mg/kg/day (Kaplan and
Sherman, 1977; Bedo and Cieleszky, 1980). A similar NOAEL, 2.5 mg/kg/day,
was found for lifetime studies in dogs (Kaplan and Sherman, 1977). No con-
trolled 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).
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 acute studies were judged to be inadequate for the
basis of the Ten-day HA value. Therefore, it is recommeded that the Longer-
term HA 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
The onset of subchronic or chronic methomyl toxicity appears to occur at
doses similar to those that cause acute toxicity. Acute ChE inhibition
has been reported to occur at doses as low as 5.1 mg/kg/day for rats exposed
via gavage, and human fatalities from ingestion of approximate methomyl doses
of 12 mg/kg in bread and 13 mg/kg in drinks have been reported (Liddle et al,
1979; Araki et al., 1982). 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). Because of the severity of the toxic
effects, the most conservative estimate of the NOAEL (2.5 mg/kg/day from the
chronic study in dogs; Kaplan and Sherman, 1977), was selected as the basis
for the Longer-term HA. In this study, beagle dogs were exposed to methomyl
in the diet at approximate doses of 0, 1.25, 2.5, 10, or 25 mg/kg/day for
2 years. Dogs receiving 1.25 or 2.5 mg/kg/day showed no evidence of toxic
effects while those receiving the two highest doses exhibited histopathological
changes in the kidney and spleen. Based on a NOAEL of 2.5 mg/kg/day, the
Longer-term HA for the child and adult are calculated as follows:
Longer-term HA = (2.5 mq/kq/day) (10kg) = o.3 mg/L (300 ug/L)
child (100)(1 L/day)
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Methorny1 August, 1988
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where:
2.5 mg/kg/day = NOAEL, based on absence of effects on blood chemistry
(including ChE activity), hematology, urinanalysis,
histopathology or body weight in dogs exposed via
the diet for 2 years.
10 kg = assumed body weight of a child
1 L/day = assumed daily water consumption of child
100 = uncertainty factor, chosen in accordance with
EPA or NAS/ODW guidelines for use with a NOAEL from
an animal study.
Longer-term HA = (2.5 mg/kg/day) (70kg) = 0.9 mg/L (900 ug/L)
adult (100)(2 L/day)
where:
2.5 mg/kg/day = NOAEL based on absence of effects on blood chemistry
(including ChE activity), hematology, urinanalysis,
histopathology or body weight in dogs exposed via
the diet for 2 years.
70 kg = assumed body weight of an adult
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
estimate of a daily exposure to the human population that is likely to be
without appreciable risk of deleterious effects over a lifetime, and is
derived from the NOAEL (or LOAEL), identified from a chronic (or subchronic)
study, divided by an uncertainty factor(s). From the RfD, a Drinking Water
Equivalent Level (DWEL) can be determined (Step 2). A DWEL is a medium-specific
(i.e., drinking water) lifetime exposure level, assuming 100% exposure from
that medium, at which adverse, noncarcinogenic health effects would not be
expected to occur. The DWEL is derived from the multiplication of the RfD by
the assumed body weight of an adult and divided by the assumed daily water
consumption of an adult. The Lifetime HA is determined in Step 3 by factoring
in other sources of exposure, the relative source contribution (RSC). The
RSC from drinking water is based on actual exposure data or, if data are not
available, a value of 20% is assumed. 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
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Methomyl August, 1988
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assessing the risks associated with lifetime exposure to this chemical.
Chronic exposure to methorny1 in the diet induces renal toxicity in rats
and dogs. Rats exposed to 900 ppm (20 mg/kg/day) for 22 months exhibited
kidney tubular hypertrophy and vacuolation of the eptithelial cells, and
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).
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). 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 this 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 histopathological 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) = 0.025 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 EPA or
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) = 0.9 mg/L (900 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.
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Methomyl August, 1988
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Step 3: Determination of a Lifetime Health Advisory, rounded to one significant
figure
Lifetime HA = (0.9 mg/L) (20%) =0.2 mg/L (200 ug/L)
where:
0.9 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; Haskell Laboratories, 1981; 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 methorny1.
0 Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986), methomyl may be classified
in Group E: no evidence of carcinogenicity. This group is used 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. Two year studies in rats and mice have not
revealed any evidence of carcinogenicity (Kaplan and Sherman, 1977;
Haskell Laboratories, 1981; Hazelton Laboratories, 1981).
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, which is 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
for methomyl (Vettorazzi and Van den Hurk, 1985), based on the same chronic
toxicity data, but using a larger uncertainty factor than that used to
derive the RfD.
0 ACGIH (1984) has adopted a threshold limit value (TLV) of 0.2 mg/m3
as a time-weighted average exposure to methomyl for an 8-hour day.
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Methomyl August, 1988
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VII. ANALYTICAL METHODS
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, 1988). 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.
This method has been validated in a single laboratory, and the estimated
detection limit for methomyl is 0.5 ug/L.
VIII. TREATMENT TECHNOLOGIES
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.
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.
Treatment technologies for the removal of methomyl from water are
available and have been reported to be effective (Whittaker, 1980).
However, the selection of an individual technology or a combination
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, 1988
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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-
l-^c-methorny 1 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. 4(5):320-325.
Dorough, H.W. 1977. Metabolism of carbamate insecticides. Available from
the National Technical 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. Khali1 and S. El-Fiki. 1979. Comparative
selective toxicity of some insecticides to insects and mammals. Proc. Br.
Crop Prot. Conf. Pests and Diseases, 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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Methomyl August, 1988
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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-((methylcarbamoyl)oxy)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 14C-methomyl 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.
Haskell Laboratories. 1981. Long-term feeding study in rats with methomyl;
INX-1179 (Final Report; Haskell Laboratory Report No. 235-81)
Hazelton Laboratories.* 1981. Final report: 104-week chronic toxicity and
carcinogenicity study in mice. Project No. 201-510. Unpublished study.
EPA ACC# 070241.
Homan, E.R., R.R. Haronpot 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
on 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.
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Methomyl August, 1988
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Liddle, J.A., R.D. Kimbrough, L.L. Needham, R.E. Cline, A.L. Smrek, L.W. Yert
and D.D. Bayse. 1979. A fatal episode of accidental methomyl poisoning.
Clin. Toxicol. 15:159-167.
McAlack, J.W.* 1973* 10-day subacute exposure of rabbit skin to lannate (R)
insecticide: Haskell Lab report No. 24-73. Unpublished study.
MRID 00007032.
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.
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.L. and B. Reiff. 1973. Effects of oximes on the 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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003. Septem-
ber 24.
U.S. EPA. 1988. 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, ECAO, Cincinnati, Ohio.
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.
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Methomyl August, 1988
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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.
Woodside, M.D., L.R. DePass and J.B. Reid.* 1978. UC 45650: Results of
feeding in the diet of rats for 7 days: Proj 41-102. Unpublished study.
MRID 00044880.
^Confidential Business Information submitted to the Office of Pesticide
programs.
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August, 1988
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.
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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Metolachlor August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS NO. 51218-45-2
Structural Formula
2-Chloro-N- ( 2-ethyl-6-methylphenyl )-N- ( 2-methoxy-l-methylethyl ) acetamide
Synonyms
0 o-Acetanilide ; 2-chloro-6 ' -ethyl-N- ( 2-methoxy-l-methylphenyl ) ;
Dual*; Bleep®; Metetilachlor; Pimagram; Primextra; CGA-24705.
Uses (Meister, 1986)
0 Selective herbicide for pre-emergence and preplant incorporated weed
control in corn, soybeans, peanuts, grain sorghum, pod crops, cotton,
saf flower, woody ornamentals, sunflowers and flax.
Properties (Meister, 1986; Ciba-Geigy, 1977; Windholz et al., 1983; Worthing,
1983)
Chemical Formula
Molecular Weight 283.46
Physical State White to tan liquid
Boiling Point 100°C (at 0.001 mm Hg)
Melting Point
Density
Vapor Pressure (20°C) 1.3 x 10~5 mm Hg
Specific Gravity —
Water Solubility (20°C) 530 mg/L
Octanol/Water partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
Metolachlor has been found in 2,091 of 4,161 surface water samples
analyzed and in 13 of 596 ground water samples (STORET, 1988).
Samples were collected at 332 surface water locations and 551 ground
water locations, and metolachlor was found in 7 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 139 ug/L in surface water and 11.0 ug/L in ground water. This
information is provided to give a general impression of the occurrence
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Metolachlor August, 1988
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of this chemical in ground and surface waters as reported in the
STORET database. The individual data points retrieved were used as
they came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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
0 l^c-Metolachlor (test substance uncharacterized), at approximately
4.6 Ib active ingredient (ai)/acre (A), degraded with a half-life
of 48-144 hours at 39-44°C when irradiated with artificial light
(uncharacterized) and 6-8 days (approximately 3,000 Langley units)
at 50-55°C when irradiated with natural sunlight (Aziz, 1974). The
degradate, N-propen-l-ol-2-yl-N-chloroacetyl-2-methyl-6-ethylaniline
(CGA-41638), was approximately 4-6% of the applied in both the
artificial light- and natural sunlight-irradiated samples; three
unknown degradates were also isolated. Nonextractable ^-^C-residues
accounted for approximately 40% of the applied, and volatiles accounted
for approximately 7-10% of the applied after 168 hours of irradiation
with artificial light or 8 days of irradiation with natural sunlight.
(purity unspecified at approximately 8 ppm was
essentially stable in loamy sand soil, over a period of 64 days
(Kaiser, 1974). Serilization of the soil had no appreciable effect
on degradation. The soil was maintained at approximately 60% of
field capacity (temperature unspecified).
14c-Metolachlor (purity unspecified) at 1-10 ppm adsorbed to sandy
clay loam, loam, and two sand soils with Freundlich adsorption
constants (K) ranging from 1.54 to 10 ug/g indicating mobility in
these soils (Burkhard, 1978). Except for one sand soil, as soil
organic matter content increased, adsorption increased. In loam soil
l^C-metolachlor desorbed with K values of 4.87 and 3.57 after 1 and
3 days, respectively. Adsorption and desorption occurred at about
the same rate.
Aged (30 days) c-metolachlor (purity unspecified) residues were
mobile in columns of loamy sand soil with approximately 26% of applied
radioactivity leached from the columns, and approximately 87% of the
applied radioactivity remaining in the soil (Dupre, 1974a). Radio-
activity was concentrated (approximately 60% of applied) in the top
3 inches of soil.
l^c-Hetolachlor (purity unspecified) residues were mobile in sandy
loam, sand, and silt loam soils (detected to the 12-inch depth in
12-inch columns) when leached with 20 inches of water (Houseworth,
1973?). l^c-Metolachlor residues were immobile in peat soil with
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Metolachlor August, 1988
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approximately 98.8% of the applied radioactivity detected in the top
1 inch of soil. The leachate from the sandy loam, sand, loam, and
silt loam soils contained 36.4, 20.9, 4.0, and 0.4% of the applied
radioactivity, respectively.
!4C-Metolachlor (purity unspecified) residues at 1 Ib ai/A dissipated
from a loamy sand soil (8° slope) in the leachate (0.25% of applied
radioactivity), and the runoff water (3.15% of applied) and sediment
(1.41% of applied) after the application of approximately 1.52 inches
of simulated rainfall in 7 days (Dupre, 1974b). Of the applied
rainfall, approximately 85% was collected as runoff. Greater than
85% and 5% of the applied radioactivity was detected in the treated
and untreated soils, respectively. Total recovery of radioactivity
was 95.8% of applied.
Freundlich adsorption K values of 0.58, 1.46, and 7.83 were calculated
for 14C-metolachlor (test substance uncharacterized) in sand, silt
loam, and sandy loam soils, respectively (Harvey and Jordan, 1978).
Adsorption was positively correlated to soil organic matter content.
Lettuce planted 14 weeks posttreatment and harvested at 26 weeks
contained 0.025 ppm of 14C-metolachlor residues (Newby, 1979).
Lettuce planted 41 weeks posttreatment and sampled at 13 and 15 weeks
contained 0.144 and 0.065 ppm of 14C-metolachlor residues, respectively.
Soil at about 40 and 56 weeks posttreatment cotained approximately
70-80 and 95% nonextractable 14C-metolachlor residues, respectively.
Residues of phenyl-labeled 14C-metolachlor were taken up by greenhouse-
grown, rotational winter wheat planted 168 days after a 2.0 Ib ai/A
application to a silt loam soil (Sumner and Cassidy, 1974d). Total
radioactivity was <0.15 ppm in forage sampled 189 to 238 days post-
treatment, and 0.60 and 0.03 ppm in straw and grain, respectively,
harvested 273 days posttreatment.
Residues of phenyl-labeled 14C-metolachlor were taken up by greenhouse-
grown, rotational oats planted 270 days after a 2.0 Ib ai/A application
to a silt loam soil (Sumner and Cassidy, 1974c). Total radioactivity
was <0.17 ppm in forage sampled 300-330 days posttreatment, and 0.27
and 0.05 ppm in straw and grain, respectively, harvested 375 days
posttreatment.
14C-Metolachlor residues were detected in the whole plant (0.04 ppm),
tops «0.09 ppm), and roots (<0.06 ppm) of rotational carrots planted
9 months after an application of 14C-metolachlor (purity unspecified)
at 2 Ib ai/A to silt loam soil (Sumner and Cassidy, 1974a). Total
residues of 14C-metolachlor in the silt loam soil were 0.26 ppm in
the 0-3 inch depth at carrot planting, and 0.04-0.35 ppm in the
0-9 inch profile in the subsequent samplings through harvest.
14C-Metolachlor residues were detected in the whole plant «0.17 ppm),
stalks (0.07 ppm), beans (0.04 ppm), meal (0.05 ppm), and oil (<0.01 ppm)
of rotational soybeans planted 9 months after an application of
14c-metolachlor (purity unspecified) at 2 Ib ai/A to silt loam soil
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or
-5-
(Sumner and Cassidy, 1974b). Total residues of 14C-metolachlor in the
silt loam soil were 0.26 ppm in the 0-3 inch depth at soybean planting,
and 0.06-0.35 ppm in the 0-9 inch profile in the subsequent samplings
through harvest.
Photodeqradation studies in water: One study (Aziz and Kahrs, 1974;
Aziz and Kahrs, 1975) cannot be validated because the experimental
conditions were referenced, rather than described, and the reference
was not available for review. In addition, this study would not fulfill
data requirements because the test substance was uncharacterized, the
incubation temperatures were not reported, it was not reported if the
test solutions were buffered or if wavelengths <290 nm were filtered
out, the conditions under which dark controls were maintained were
not reported, the natural sunlight conditions were not provided, the
intensity and wavelength distribution of the artificial light source
were not provided, and degradates that comprised >10% of the applied
that were detected in the test solution exposed to artificial light
were not characterized.
Leaching and adsorption/desorption studies; Four studies were reviewed
and considered to be scientifically valid. The first study (Burkhard,
1978) partially fulfills data requirements by providing information
on the adsorption of parent metolachlor. However, the study was not
conducted in a 0.01 N calcium ion solution, desorption data were
reported for only one of the four soils tested, and the majority of
the study was conducted on foreign soils. The second study (Dupre,
1974a) partially fulfills data requirements by providing information
on the mobility of aged (30 days) metolachlor residues. However, the
purity of the test substance was unspecified, K^ values were not reported,
and l^C-metolachlor residues were not characterized. The third study
(Houseworth, 1973a) partially fulfills data requirements by providing
information on the mobility of metolachlor residues. However, the
purity of the test substance was unspecified, and K
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-6-
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-l-methylethyl) (MET-003) and MET-004.
Excretion
When treated with 14C-metolachlor (approximately 31 mg/kg 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,
cyanosis, hypothermia, collapse, convulsions, diarrhea, gastrointestinal
irritation, jaundice, weakness, nausea, shock, sweating, vomiting, CNS
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
The acute oral LDso of technical metolachlor in the rat was reported
to be 2,780 mgAg (95% confidence range of 2,180 to 3,545 mgAg;
Bathe, 1973).
Technical metolachlor in corn oil was shown to be emetic in beagle
dogs, precluding the establishment of an LDso (Roche, 1974). However,
an "emetic dose" of 19 + 9.7 mgAg was established.
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-7-
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 No-Observed-Adverse-Effect Level
(NOAEL) in this study.
Dermal/Ocular Effects
0 The LDso of technical metolachlor in the rabbit when tested by the
unabraded dermal route is greater than 10,000 mg/kg (Bach, 1974).
0 Sachsse (1973b) evaluated the dermal irritation potential of technical
metolachlor (3/sex/dose) on the Russian strain rabbits (weighing 2 to
3 kg). The chemical was applied to abraded and unabraded skin for
24 hours and then observed for periods up to 72 hours. The results
demonstrated that technical metolachlor is non-irritating to rabbit
skin.
0 Sachsse and Ullman (1977) studied skin sensitization in the albino
guinea pig by the intradermal-injection method. Technical metolachlor
(CG-24705) 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. The authors concluded that technical
metolachlor is a skin sensitizer.
0 A study of eye irritation by technical metolachlor in the Russian
strain rabbit (3/sex/dose) 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% ai) 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)
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-8-
0 Tisdel et al. (1982) presented the results of a study in which
metolachlor (95% ai) was administered to Charles River CD-I 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
were observed in terms of physical appearance, food consumption,
hematology, 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, the significance of which is uncertain. 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.
Based on this data, a NOAEL of 300 ppm (15 mg/kg/day) is identified.
Reproductive Effects
0 A three-generation rat reproduction study was reported by Adler and
Smith (1978). Targeted dietary exposures for metolachlor (96.5% ai)
were 0, 30, 300 or 1,000 ppm. The 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
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Metolachlor August, 1988
-9-
administered technical-grade metolachlor (purity not specified) at
dietary doses of 0, 30, 300 or 1,000 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
(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 Fia and F2a litters. Food consumption was
reduced significantly for FI 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 FI parental males and females at 1,000 ppm and increased
thyroid-to-body weight and thyroid-to-brain weight in FI 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% ai) 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 cp-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 (95.4% pure) in New Zealand White rabbits (16/dose). The
compound was administered via stomach tube as a suspension in aqueous
0.75% hydroxymethylcellulose at levels of 0, 36, 120 or 360 mg/kg/day.
Single oral doses 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 significant developmental or fetotoxic
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Metolachlor August, 1988
-10-
effects were observed in the 319 fetuses, pups or late resorptions
evaluated from all dose groups. Thus, a minimal NOAEL of 360 mg/kg/da.
for fetotoxicity and a NOAEL of 36 mg/kg/day for maternal toxicity
were identified.
Mutaqenicity
0 Technical metolachlor (purity not specified) was tested in the Ames
Salmonella test system, using S. typhimurium strains TA1535, TA1537,
TA98 and TA100 (Arni 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 Fritz (1976) reported the results of a dominant lethal study in male
albino NMRI-derived mice 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.
Carcinoqenicity
0 Marias et al. (1977) presented the results of a study in which
technical-grade metolachlor was administered to Charles River CD-I
mice (50/sex/dose) at dietary concentrations of 0, 30, 300 or 3,000 ppi
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. Two samples of
metolachlor (99.9% ai and 96.5% ai) were received. The 99.9% sample
was used in dietary formulations for the first 33 weeks and the 96.5%
pure sample was used for the rest of the study. Results of this
study indicated no evidence of oncogenicity in either sex.
0 Tisdel et al. (1982) presented the results of a study in which
metolachlor (95% ai) was administered to Charles River CD-I 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 Ciba-Geigy (1979) 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
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Metolachlor August, 1988
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tunors 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.
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. 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 (up to 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) - m/L i uq/L)
(UP) x ( L/day) * *
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 Longer-Term HA value for the 10 kg child (2.0 mg/L, calculated
below) be used at this time as a conservative estimate of the One-day HA
value.
Ten-day Health Advisory
There were no satisfactory studies found in the available literature to
use for the calculation of the Ten-day HA. It is, therefore, recommended
that the Drinking Water Equivalent Level (DWEL), of adjusted for a 10-kg
child 2 mg/L (2,000 mgA)f be used for the Ten-day HA.
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Metolachlor August, 1988
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Lonqer-Term Health Advisory
No studies were found in the literature that were suitable for deriving
the Longer-term HA value for metolachlor. In the lifetime study of Tisdel
et al. (1983), a NOAEL of 15 mg/kg/day was found in rats. This NOAEL was
similar to that NOAEL (13.7 mgAg/day) found by Goldenthal et al. (1979).
However, Goldenthal et al. (1979) used only one dog of each sex per dose group
(0, 13.7, 22.7 or 40.2 mg/kg/day), while Tisdel et al. (1983) used 70/sex/dose
at 01.5, 15 or 150 mg/kg/day. Accordingly, more confidence can be placed in
the Tisdel et al. study. It is, therefore, recommended that the DWEL of 5.0
mg/L (5,000 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, 2.0 mg/L (2,000
ug/L), be used for the Longer-term HA 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. 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 mgAg/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. Based on the
decreased body weight gain at 3,000 ppm (150 mg/kg/day), a NOAEL of 300 ppm
(15 mg/kg/day) was identified.
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Metolachlor August, 1988
-13-
The Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD):
KfD = 15 mg/kq/day = 0.15 mg/kg/day
where:
15 mgAg/day = NOAEL, 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 EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study.
Step 2: Determination of the Drinking water Equivalent Level (DWEL)
DWEL = (0.15 mq/kq/day)(70 kg) = 5.25 mg/L (5,000 ug/L)
(2 L/day) y y
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 = (5.25 mq/L) (20%) = O.l mg/L (100 ug/L)
where:
5.25 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 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, both conducted using rats, showed an
increased tumor incidence related to treatment. Ciba-Geigy (1979)
reported a statistically significant increase in primary liver tumors
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August, iy. a
-14-
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
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).
The International Agency for Research on Cancer has not evaluated the
carcinogenicity of metolachlor.
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
applicable to the determination of certain nitrogen-phosphorus
containing pesticides in water samples (U.S. EPA, 1988). 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. This method has been validated in a single
laboratory, and estimated detection have been determined for the
methods using this method. The estimated detection limit for
metolachlor is 0.75 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.
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Metolachlor August, 1988
-15-
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
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, 1988
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IX. REFERENCES
Adler, G.L. , and S.H. Smith.* 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. MRID 00015632.
Arni, P., and D. Muller.* 1976. Salmonella/mammalian-microsorae mutagenicity
test with CGA 24705. Test for mutagenic properties in bacteria. PH 2.632.
Received January 19, 1977 under 7F1913. MRID 00015397.
Aziz, S.A.* 1974. Photolysis of CGA-24705 on soil slides under natural and
artificial sunlight conditions: Report No. GAAC-74102. Unpublished
study received on unknown date under 5F1606; submitted by Ciba-Geigy
Corp., Greensboro, NC; CDL:094385-J. MRID 00016301.
Aziz, S.A. , and R.A. Kahrs.* 1974. Photolysis of CGA-24705 in aqueous solution
under natural and artificial sunlight conditions: Report No. GAAC-74041.
Unpublished study received Sept. 26, 1974 under 5F1606; submitted by
Ciba-Geigy Corp., Greensboro, NC; CDL:094385-D. MRID 00016300.
Aziz, S.A. , and R.A. Kahrs.* 1975. Photolysis of CGA-24705 in aqueous solution
— additional information: Report No. GAAC-75021. Unpublished study
received on unknown date under 5F1606; submitted by Ciba-Geigy Corp. ,
Greensboro, NC; CDL:094385-M. MRID 00016302.
Bach, K.J.* 1974. Acute dermal LDso of CGA-24705- Technical in rabbits:
Affiliated Medical Research, Inc. Contract No. 120-2255-34. Received
September 26, 1974 under 5G1553. Unpublished study. MRID 00015526.
Bathe, R. 1973.* Acute oral LDso of technical CGA-24705 in the rat: Project
No. Siss 2979. Received September 26, 1974 under 5G1553. Unpublished
study. MRID 00015523.
Burkhard, N.* 1978. Adsorption and desorption of metolachlor (Dual) in
various soil types: Project Report 45/78. Unpublished study received
July 23, 1981 under 100-587; prepared by Ciba-Geigy, Ltd., Switzerland,
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:245627-D. MRID 00078291.
Ciba-Geigy Corporation.* 1977. Section A General Chemistry, unpublished
study received January 19, 1977 under 7F1913. MRID 00015392.
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 00130776.
CFR. 1985. Code of Federal Regulations. 40 CFR 180.368. July 1, 1985.
Coquet, B., L. Galland, D. Guyot, X. Pouillet and J.L Rounaud.* 1974. Three-
month oral toxicity study trial of CGA 24705 in the dog. IC-DREB-R 740119.
Received September 26, 1974 under 5G1553. Unpublished study. MRID 0005247V,
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Metolachlor August, 1988
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Dupre, G.D.* 1974a. Leaching characteristics of 14C-CGA-24705 and its degra-
dation products following aging in sandy loam soil under greenhouse
conditions: Report No. 73021-6. Unpublished study received Sept. 26,
1974 under 5G1553; prepared by Bio/dynamics, Inc., submitted by Ciba-Geigy
Corp., Greensboro, NC; CDL:094222-C. MRID 00015657.
Dupre, G.D.* 1974b. Runoff characteristics of 14C-CGA-24705 applied to sandy
loam soil under greenhouse conditions: Report No. 73022-1. Unpublished
study received Sept. 26, 1974 under 5G1553; prepared by Bio/dynamics,
Inc., submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:094222-D.
MRID 00015658.
Fritz, H.* 1976a. Reproduction study on CGA-24705 Tech. Rat: Segment II test
for teratogenic or embryotoxic effects: PH 2.632. Unpublished study,
including addendum, received January 19, 1977 under 7F1913. MRID 00015396.
Fritz, H.* 1976b. 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 00015630.
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. MRID 00016631.
Hambock, H.* 1974a. 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. MRID 00039193.
Hambock, H.* 1974b. Project 12/74: Addendum to Project 7/74: Metabolism of
CGA 24705 in the rat: AC 2.52. Unpublished study. MRID 00015425.
Hambock, H.* 1974c. Project 1/74: Distribution, degradation and excretion of
CGA 24705 in the rat: AC 2.52. MRID 00039192.
Harvey, R.F., and G.L. Jordan.* 1978. Comparative study of the biological
and physical properties of acetanilide herbicides in soil. Unpublished
study received May 3, 1979 under 43142-1; prepared by Univ. of Wisconsin.
Submitted by Boots Hercules Agrochemicals Co., Wilmington, DE; CDL:098274-H.
MRID 00031328.
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. Chenu Eng. 88:88-89.
Houseworth, L.D.* 1973?. Report on parent leaching studies for CGA-24705:
Report No. 1. Unpublished study received Sept. 26, 1974 under 5G1553;
prepared by Univ. of Missouri, Dept. of Plant Pathology, submitted by
Ciba-Geigy Corp., Greensboro, NC; CDL:094222-E. MRID 00015659.
-------
Metolachlor August, 1988
-18-
Jessup, D.C., R.J. Arceo, F.L. Estes et al.* 1979. 6-month chronic oral
toxicity study in beagle dogs: IRDC No. 382-054. unpublished study.
MRID 00032174.
Kaiser, F.E.* 1974. Soil degradation study of Gba-Geigy (sic) 14C-CGA-24705.
Unublished study received Mar. 27, 1975 under 5F1606; prepared by
Analytical Biochemistry Laboratories, Inc., submitted by Ciba-Geigy
Corp., Greensboro, NC; CDL:094376-N. MRID 00016296.
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. MRID 00041283.
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.
MRID 00084003.
Meister, R., ed. 1986. Farm Chemicals Handbook. Willoughby, OH: Meister
Publishing Co.
Newby, L. 1979.* Dual rotational studies: Additional information: Report
No. ABR-79091. Unpublished study received Aug. 1, 1979 under 100583;
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:238899-A. MRID 00015538.
Roche, W.J.* 1974. Affiliated Medical Research, Inc. Emetic dose 50 in
beagle dogs: Affiliated Medical Research, Inc. Contract No. 120-2255-34.
Received September 26, 1974, Greensboro, NC. MRID 00015525.
Ross, R.H., and K. Balu. 1985. Summary of metolachlor water monitoring for
1979-July 1985. Report #EIR-85024. Submitted by Safety Evaluation Dept.,
Agricultural Division, Ciba-Geigy Corp., Greensboro, NC. Accession No.
260602. (NO MRID)
Sachsse, K.* 1973a. Irritation of technical CGA-24705 in the rabbit eye:
Project No. Siss 2979. MRID 00015528.
Sachsse, K.* 1973b. Skin irritation in the rabbit after single application
of Technical CGA-24705. Project No. Siss 2979. Unpublished study.
MRID 00015530.
Sachsse, K., and L. Ullmann.* 1977. Skin sensitizing (contact allergenic)
effect in guinea pigs of Technical CGA-24705. Project No. Siss 5726.
Unpublished study. MRID 00015631.
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. MRID 00080897.
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Metolachlor August, 1988
-19-
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Sumner, D.D./ and J.E. Cassidy.* 1974a. The uptake of phenyl-14C-CGA-24705
and its aged soil degradation products in rotation carrots: Report No.
GAAC-74112. Unpublished study received June 30, 1978 under 100-583;
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:234216-H. MRID 00022882.
Sumner, D.D., and J.E. Cassidy.* 1974b. The uptake of phenyl-14C-CGA-24705
and its aged soil degradation products in rotation soybeans: Report No.
GAAC-74113. Unpublished study received June 30, 1978 under 100-583;
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:234216-G. MRID 00022881.
Sumner, D.D., and J.E. Cassidy.* 1974c. The uptake of phenyl-14C-CGA-24705
and its aged soil degradation products in rotation oats: Report No.
GAAC-74085. Unpublished study received July 23, 1981 under 100-587;
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:245627-K. MRID 00022883.
Sumner, D.D., and J.E. Cassidy.* 1974d. The uptake of phenyl-14C-CGA-24705
and its aged soil degradation products in rotation wheat: Report No.
GAAC-74071. Unpublished study received Sept. 26, 1974 under 5F1606;
submitted by Ciba-Geigy Corp., Greensboro, NC; CDL:094385-H. MRID 00022878.
Tisdel, M., P. MacWilliams, R. Dahlgren et al. 1982. Carcinogenicity study
with metolachlor in albino mice. Hazelton Raltech study No. 79020.
Unpublished study. MRID 0011759.
Tisdel, M., T. Jackson, P. MacWilliams et al.* 1983. Two-year chronic oral
toxicity and oncogenicity study with metolachlor in albino rats: Raltech
study No. 80030. Final report. Unpublished study. MRID 00129377.
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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. U.S. EPA Method 507
- Determination of nitrogen and phosphorus containing pesticides in
water by GC/NPD, April 15, 1988 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, 1988
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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 1988
METRIBUZIN
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.
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 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 21087-64-9
Structural Formula
4-Amino-6-(1r1-dimethylethyl)-3-methylthio-1,2,4-triazin-5(4H)-one
Synonyms
0 Bayer 6159; Bayer 6443H; Bayer 94337; Lexone; Sencor; Sencoral;
Sencorer; Sencorex
Uses
0 Herbicide used for the control of a large number of grass and broadleaf
weeds infesting agricultural crops (Meister, 1983).
Properties (CHEMLAB, 1985)
Chemical Formula CgH 14(^48
Molecular Weight 214.28
Physical State (at 25°C) white crystalline solid
Boiling Point
Melting Point 125-126°C
Density
Vapor Pressure (25°C) 10~5 mmHg (20°C)
Specific Gravity
Water Solubility (25°C) 1,200 mg/L
Log Octanol/Water Partition -5.00 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor ~
Occurrence
0 Metribuzin has been found in 936 of 4,651 surface water samples
analyzed and in none of 416 ground water samples (STORET, 1988).
These samples were collected at 376 surface water locations and
293 ground water locations; metribuzin was found in 14 states. The
85th percentile of all nonzero samples was 4.79 ug/L in surface water.
The maximum concentration found in surface water was 34.45 ug/L.
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Metribuzin August 1988
-3-
This information is provided to give a general impression of the
occurrence of this chemical in ground and surface waters as reported
in the STORET database. The individual data points retrieved were
used as they came from STORET and have not been confirmed as to their
validity. STORET data is often not valid when individual numbers
are used out of the context of the entire sampling regime, as they
are here. Therefore, this information can only be used to form an
impression of the intensity and location of sampling for a particular
chemical.
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).
0 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-triazin-5(4H)-one (DA) and parent
compound, respectively. A substantial portion of the applied radio-
activity (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 ppm 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).
o 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
Kg 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).
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Metribuzin August 1988
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0 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.
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-Metribuzin 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 (Obrist 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, metribuzin 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 Chemical 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, et al, 1972) to evaluate absorption, distri-
bution and metabolites. Analysis of blood samples showed a peak
level at 4 hours.
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Hetribuzin August 1988
-5-
Distribution
0 No information was found in the available literature on the distribution
of metribuzin. Khasawinah et al. (1972) did not provide information
on distribution of metribuzin in the dog study.
Metabolism
0 No information was found in the available literature on the metabolism
of metribuzin. Khasawinah et al. (1972) study in dogs did not elaborate
on metabolism in metribuzin.
Excretion
0 Khasawinah et al. (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 LD5Q values
following the administration of technical metribuzin to guinea pigs
and rats as 245 and 1,090 rag/kg, respectively, for male animals,
and 274 and 1,206 rag/kg, respectively, for females.
0 Mobay Chemical (1978) reported the acute oral LDsg 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.
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Hetribuzin August 1988
-6-
Long-term Exposure
0 Loser et al. (1969) administered metribuzin to Wistar rats (15/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 the dietary assumptions
of Lehman, 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 (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 Rienter (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 the dietary assumptions of
Lehman, 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 Mohr (1974) reported that dietary concentrations of 25,
35, 100 or 300 ppm 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 the dietary assumptions of Lehman, 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 the dietary assumptions of Lehman, 1959) for 24 months.
Following treatment, food consumption, general behavior and appearance,
clinical chemistry, hematology, urinalysis, body and organ weights
and histopathology were evaluated. No toxicologic effects were
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Metribuzin August 1988
-7-
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 hyperglycemia
and temporary hypercholesterolemia were noted in animals administered
1,500 ppm (3 7.5 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
the dietary assumptions of Lehman, 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.
Developmental Effects
0 Uhger 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/kg/day or less. Mo treatment-related effects were
reported at any dose level in fetuses based on gross, soft tissue and
skeletal examinations.
• 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.
Carcinogenicity
0 Hayes et al. (1981) conducted studies in which technical metribuzin
was administered in the diet to albino CD-1 mice (50/sex/dose) at 200,
800 or 3,200 ppm (approximately 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
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Metribuzin August 1988
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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 (up to 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 / uq/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, 1,000 or 10,000),
in accordance with EPA or 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.
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 (5,000 mg/L)
(100) (1 L/day)
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Hetribuzin August 1988
-9-
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 EPA
or 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 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,
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 mq/kg/day) (10 kg) = 0.25 mg/L (300 ug/L)
(100) (1 L/day) * y
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 EPA
or 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 2.5 mg/kg/day, the Longer-term HA for a 70-kg adult is
calculated as follows:
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Metribuzin August 1988
-10-
Longer-term HA = (2.5 mq/kg/day) (70 kg) = Q.875 mg/L (900 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 EPA
or 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. 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 mgAg/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
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Metribuzin August 1988
-11-
mice; however/ this value exceeded the LOAEL (37.5 mg/kg/day) reported by Loser
and Mlrea (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 metribuzin via
the diet for 24 months.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODH 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 (900 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 (200 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, location,
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.
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Metribuzin August 1988
-12-
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).
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 separator/ 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. This method has been
validated in a single laboratory, and estimated detection limits have
been determined for the analytes, including metribuzin; the estimated
detection limit is 0.15 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular-activated carbon (GAC) 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 metribuzin from contaminated wastewater (Whittaker,
et al., 1982). One type of GAC showed breakthrough for metribuzin
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Metribuzin August 1988
-13-
(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 GAG.
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'
Treatment technologies for the removal of metribuzin from water are
available and have been reported to be effective. However/ selection
of an individual technology or combination 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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Metribuzin August 1988
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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, HO: p.
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. American Chemical Society,
(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 oryzal'in 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. ^n_ 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.
Ford, 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. Mallicoat 1981. Metribuzin (R) (Sencor)
oncogenicity study in mice: 80050. Unpublished study. MRID 00087795.
Houseworth, L.D. and B.G. 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 1988
-15-
Inukai/ H. and A. lyatomi.* 1977. Bay 94337: Mutagenicity test on bacterial
systems: Report no. 67; 54127. Unpublished study. MRID 0008677Q.
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.
Khasawinah/ A. M., 0. R. Flint/ H. R. Shaw/ and D. D. Cox (1972). The
metabolic fate of carbonyl cih-SENCOR in Dogs. Unpublished study
submitted by Moay Chemical Corp., MRID 00045264.
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 Chemagro 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. Mawdesley-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 experiment): Report no. 4887; Report no. 41814.
Unpublished study. MRID 00061260.
Loser/ E. and U. Mohr.* 1974. Bay 94337: Chronic toxicity studies on rats
(two-year feeding experiment): Report no. 4888; Report no. 41816.
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 mouse: Report nos. 4942 and 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 nos. 6110 and 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.
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Metribuzin August 1988
-16-
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.
Obrist, J.J. and J.S. Thornton. 1979. Soil thin-layer mobility of Baycor
(TM), Baytan, Drydene and Peropa'l (TM). Unpublished study prepared in
cooperation with Agricultural Consultants, Inc., submitted by Mobay
Chemical Corp., Kansas City, MO.
Pither, K.M. ahd 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. Heed
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.
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Metribuzin August 1988
-17-
Sharon, H. and G.R. Stephenson. 1976. Behavior and fate of metribuzin.
Heed Sci. 24(2):153-160. Submitter report no. 49127. In unpublished
study submitted by Mobay Chemical Corp., Kansas City, MO.
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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.G. 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 507
- 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, 1988
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 whichoadverse 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.
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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Paraquat
August/ 1988
-2-
II. GENE \L INFORMATION AND PROPERTIES
Par juat, with a chemical name 1,1'-dimethyl-4,4'-dipyridinium
ion, is resent mostly as the dichloride salt (CAS No. 1910-42-5) or
as the c nethyl sulfate salt (CAS No. 2074-50-2, molecular weight 408.48)
(Meister 1988). Contents discussed below pertain to paraquat dichloride.
CAS No. 1910-42-5
Structur 1 Formula
CH,-N
2CI
Synonyms
Uses
1,1'-Dimethyl-4,4'-bipyridinium-dichloride
T-5, Actor, Cekuquat, Crisquat, Herboxone, Dextrone, Dexuron, Esgram,
;ramocil, Gramoxone, Gramuron, Goldquat 276, Herbaxon (Agro Qiimicas),
ierboxone, Mofisal, Osaquat Super, Paracol, Paracote, Pathclear,
Teeglone, Priglone, Simpar, Sweep, Terraklene,'Totacol, Total,
'oxer, Weedol. Discontinued name: Ortho (Meister, 1988).
0 .ontact herbicide and desiccant used for desiccation of seed crops,
or noncrop and industrial weed control in bearing and nonbearing
ruit orchards, shade trees, and ornamentals, for defoliation and
esiccation of cotton, for harvest aid in soybeans, sugarcane, guar,
nd sunflowers, for pasture renovation, for use in "no-till" or before
lanting or crop emergence, dormant alfalfa and clover, directed
pray, and for killing potato vines. Paraquat is also effective for
radication of weeds on rubber plantations and coffee plantations and
gainst paddy bund (Meister, 1988).
Propert: s (ACGIH, 1980; Meister, 1988; CHEMIAB, 1985; TDB, 1985)
hemical Formula
olecular Weight.
•hysical State
oiling Point
elting Point
ansity
apor Pressure
pecific Gravity
ater Solubility
og Octanol/Water Partition
Coefficient
aste Threshold
dor Threshold
onversion Factor
C12H14N2.2C1
257.18
Colorless to yellow crystalline solid
175 to 180°C
1.24 g/mL (20«C)
No measurable vapor pressure
1.24 at 20°C/20»C
Very soluble
2.44 (calculated)
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Paraquat August/. 1988
-3-
Occurrence
0 Paraquat was undetectable in 843 ground water samples analyzed
(STORET* 1988). Samples were collected at 813 ground water locations.
No surface water samples were collected for analysis.
Environmental Fate
o 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., 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.
0 14c-Paraquat (test substance uncharacterized) was immobile in silt
loam 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* 1968).
0 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).
0 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, 1988
-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 14c-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, ™C-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 1*C-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 major
residue was unchanged paraquat.
Metabolism
0 Paraquat dichloride or paraquat dimethyl sulfate (radiochemical
purity: 99.3 to 99.8%), labeled with 14C in either methyl groups or
in the ring, was poorly absorbed from the gastrointestinal tract of a
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Paraquat August, 1988
-5-
cow (Leahey et al., 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, 14c-labeled) 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, respectively. (Clark, 1965). Signs of toxicity included
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Paraquat August, 1988
-6-
respiratory distress and cyanosis among rats and guinea pigs, blood-
stained droppings among the hens, and muscular weakness, incoordi nation
and frequent vomiting of frothy secretion among the cats.
0 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 rag/kg and 80 to 90 mg paraquat ion/kg for rats (FDA, 1970).
0 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, 1977b). 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 conjunctival sac of one eye in each of six male New Zealand
White rabbits (Bullock, 1977a). 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 Lowest-Observed-Adverse-Effect Level (LOAEL), and 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 paraquat ion for male
dogs and 0, 0.48, 1.00 or 1.58 mg paraquat ion/kg/day for females.
Clinical and behavioral abnormalities, 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 (LOAEL) 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 cation/kg/day)
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Paraquat August/ 1988
-7-
0 Technical paraquat dichloride (32.7% cation) was fed to Alderley
Park mice (60/sex/dose) for 97-99 weeks at levels of 0, 12.5, 37.5
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 cation/kg. The animals were observed for
toxic signs, and body weights, food consumption and utilization;
urinalysis, gross pathology and histopathology were also evaluated.
Renal tubular degeneration in the males, weight loss and decreased
food intake in the females, were observed in the 37.5 ppm dose group.
Based on these findings, a NOAEL of 12.5 ppm cation (1.87 mg/kg/day)
was identified for both sexes.
0 Fischer 344 rats (70/sex/dose) were fed diets containing 0, 25, 75
or 150 ppm of technical paraquat (32.7% cation) for 113 to 117 weeks
(males) and 122 to 126 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 cation/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. Based on these
results, an approximate NOAEL of 25 ppm (1.25 mg cation/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 genera-
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 (FQ, F-\ 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, 1988
-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., 1977). 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. (1978) 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, LOAEL)
and above. Based on these findings, the developmental and maternal
NOAEL of 1 mg paraquat ion/kg/day was identified.
Mutagenicity
0 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
blood) (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 (£124 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 et al., 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 ppra for the initial 35 weeks and then
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Paraquat August, 1988
-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. The results show 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 when pooled, the skin tumors (squamous cell carcinomas
in the head region including ear, nasal cavity, oral cavity and skin)
appeared to provide some equivocal evidence of .carcinogenicity in the
high-dose males. However, the observed skin- tumors were low in
incidence and occurred at different sites in the head region. The
different tumor sites were considered to be anatomically different
and inappropriate for pooling by an independent pathology evaluation
review (Experimental Pathology Laboratory, Inc.) as well as by the
second EPA Review of Paraquat (1988). Thus, the tumors could not be
associated with oral exposure to the compound.
0 Paraquat dichloride (98% pure) fed to JCR:ICR mice (60/sex/group) at
0, 2, 10, 30 or 100 ppm for 104 weeks did not induce statistically
significant dose-related oncogenic responses (Toyoshima et al.,
1982a). At 100 ppm, increased mortality and statistically significant
changes in hematology, clinical chemistry and organ weights were
observed in both sexes, indicating an MTD was reached. The experiment
identifies a NOAEL of 30 ppm.
0 Paraquat dichloride (98% pure) fed to
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Paraquat August, 1988
-10-
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet 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, 1,000 or 10,000),
in accordance with EPA or 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., 1978) 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
ion/kg/day. Another adequate study of comparable duration reported a NOAEL
of 5 mg/kg/day for developmental effects, and a NOAEL of 1 mg/kg/day for
maternal effects supports this selection (Hodge et al., 1977).
Using a NOAEL of 1 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
Ten-day HA = d mg/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 EPA or
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, 1988
-11-
Longer-term Health Advisory
No subchronic 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 Hater Equivalent Level (DWEL) of 0.2 mg/L
(200 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, 0.05 mg/L (50 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. 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 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 ion/kg/day) was identified
based on the absence of hematological, biochemical, gross pathological and
histological 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 (Utchfield et al., 1981).
Step 1: Determination of the Reference Dose (RfD)
RfD = (0'45 mg ion/kg/day) = 0.0045 mg/kg/day
(100)
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Paraquat August, 1988
-12-
where:
0.45 mg ion/kg/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
EPA or 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) = 0.16 m /L (200 /L)
(2 L/day)
where:
0.0045 mg/kg/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%) = 0.03 mg/L (30 ug/L)
where:
0.16 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 Based on the recent EPA Peer Review of the data on paraquat (U.S.
EPA, 1988a), the chemical is classified in Group E, evidence of
noncarcinogenicity for humans (U.S. EPA, 1986). The chemical provided,
at most, only equivocal evidence for skin tumors at high dose in male
rats (but not in the females) in one study, but no evidence for
carcinogenicity in three other studies in rats and mice. The observed
skin tumors were low in incidence, occurred at different sites in the
head region, and the tumor incidence was significantly different from
controls only when pooled from all sites in the head region. Since
pooling of the different tumor sites was considered by pathologists
to be inappropriate, the tumors could not be associated with
oral exposure to the compound.
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of paraquat.
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Paraquat August, 1988
-13-
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 The American Conference of Governmental Hygienists has presented a
threshold limit value of 0.1 mg/m3 for paraquat of respirable particle
sizes (ACGIH, 1980).
VII. ANALYTICAL METHODS
0 Paraquat is analyzed by a Draft EPA method (U.S. EPA, 1988b) for
drinking water. In this procedure a measured quantity of sample
is extracted with a solid phase adsorbant. After elution from the
adsorbant tube, the paraquat residue is determined by high performance
liquid chromatography with an ultraviolet detector (HPLC/UV). The
method has a single laboratory validation maximum detection limit
(MDL). For paraquat, the MDL is 0.80 ug/L.
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
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 August, 1988
-14-
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, 1988
-15-
IX. REFERENCES
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Documentation of the threshold limit values for substances in workroom
air, 4th ed. Cincinnati, OH: ACGIH.
Benigni, R., H. 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.* 1977a. S-1103: 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.
Bullock, C.H.* 1977b. S-1104: 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.
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 00152691.
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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 LDsg - rat. Project No.
stated. Chambers/ GA. HRID GS 0262-003. Cited in U.S. EPA, 1985.
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 chroma tography. 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. Weight and J. Wilson. 1977. 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. Weight and J. Wilson. 1978. 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 in 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.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food and Drug Off. U.S., Q. Bull.
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Paraquat August, 1988
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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 generations. 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, H.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. Djebbar. 1986. Removal of paraquat, diquat and
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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/1325S, September 24. Cited in U.S. EPA,
1985. MRID GS 0262-004.
Meister, R. , ed. 1988. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
NAS. 1977. National Academy of Sciences. Drinking water and health.
Washington, DC: National Academy Press.
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California sunlight. Unpublished submission by Chevron Chemical Co.
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Parry, J.M. 1977. The use of yeast cultures for the detection of environmental
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dichloride: A cytogenetics study in human lymphocytes in vitro. Imperial
Chemical Industries PLC, England. Report No. CTL/P/1351, September 3.
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and residues in pigs using 14c -ring-labelled paraquat. Huntingdon Research
Centre, England. Project No. AR 2692 A, October 1. MRID 00028599.
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TDB. 1985. Toxicology Data Bank. MEDLARS II. National Library of Medicine's
National Interactive Retrieval Service.
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U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
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of paraquat. Memo from Reto Engler to Robert Taylor, June 28.
U.S. EPA. 1988b. U.S. Environmental Protection Agency. EPA Draft Method
"Determination of diquat and paraquat in drinking water by high performance
liquid chromatography with an ultraviolet detector." This method is
available from EPA's Environmental Monitoring and Support Laboratory,
Cincinnati, Ohio.
Upton, B.P., P. Hendley and M.W. Skidmore.* 1985. Paraquat: Hydrolytic
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Richmond, CA.
Weber, J.B., T.M. Ward and S.B. Weed. 1986. Adsorption and desorption of
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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
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Worthing, C.R. 1983. The pesticide manual. Published by the British Crop
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'Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
PICLORAM
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.
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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Picloram
August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-02-01
Structural Formula
(4-amino-3,5,6-trichloropicolinic acid)
Synonyms
0 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 (35°C)
Specific Gravity
Water Solubility
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Picloram has been found in 420 of 744 surface water samples analyzed
and in 3 of 64 ground water samples (STORET, 1988). Samples were
collected at 135 surface water locations and 30 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 0.02 ug/L in
ground water sources. The maximum concentration found was 4.6 ug/L
in surface water and 0.02 ug/L in ground water. This information
is provided to give a general impression of the occurrence of this
241.6
White powder
Decomposes
215°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, 1988
-3-
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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 CO2? 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.
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Picloram August, 1988
-4-
0 A 500-kg Holstein cow was administered 5 ppm 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
1^c-picloram (dose not specified), where 95% of the dose was absorbed
(Dow, 1983).
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 1^C-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, omentum 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 C02 was detected in expired air of rats given C-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 CO2 (Nolan et al., 1980).
One Holstein cow administered 5 ppm picloram (approximately 0.23 mg/kg/day)
in feed for 4 consecutive days excreted more than 98% of the dose in
the urine (Fisher et al., 1965).
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Picloram August, 1988
-5-
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
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.
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
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 LD5Q values ranging from
2,000 to 4,000 mg/kg (NIOSH, 1980; Dow 1983).
0 In a 7-day feeding study by Dow (198la), picloram was fed to female dogs
(one/dose) at dose levels of 400, 800 or 1,600 mg/kg bw/day. Picloram
was acutely toxic (emesis, loss of body weight) 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 7- to 14-day study by Dow (1981b)f beagle dogs (one dog per
dose) 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.
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Picloram August, 1988
-6-
0 In a 32-day feeding study by Dow (1980), 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.
Dermal/Ocular Effects
0 Most formulations of picloram have been evaluated for the potential
to produce skin sensitization reactions in humans. Dow (1983) 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). When the tri-
isopropanolamine 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 mg/kg/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 mg/kg/day) for males. Therefore, the 7-mg/kg/day dose level
was considered to be a NOAEL.
0 In a 13-week feeding study, CDF Fischer 344 rats (15/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-mg/kg/day dose levels. The NOAEL in this
study was identified as 50 mg/kg/day.
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 females and males at all dose levels tested.
0 Picloram (94%) was fed to Fischer 344 rats (50/sex/dose) for two years
in the diet at dose levels of 0, 20, 60 or 200 mg/kg (Dow, 1986).
There was a significant dose-related increase in size and altered
tictorial properties of centrolobular hepatocytes and increased rela-
tive liver weights in males and females dosed at 60 and 200 mg/kg/day.
The LOAEL based on histologic changes in the liver is 60 mg/kg/day
and the NOAEL is 20 mg/kg/day.
0 Osborne-Mendel rats receiving picloram at 370 or 740 mg/kg/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.
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Picloram August, 1988
-7-
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 12 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 Fia 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 P-j^ generation. The second generation (F2a an<* F3b^
was derived from F2b animals after 110 days of age. Two weanlings
per sex per dose from litters of each generation were observed for
gross pathology. Gross pathology was also assessed 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 FJJ-J generation which was fed 75 mg/kg/day dose. No
other effects were noted. Based on these results, a NOAEL of 25 mg/kg/day
was identified.
Developmental Effects
0 In the McCollister et al. (1967) study described above, the F-|C, 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.
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 treat-
ment; 14 maternal deaths occurred between days 8 and 17 of gestation
in these dose groups. Evidence of retarded fetal development as
reflected by an increase in unossified fifth sternebrae, was observed
in all treatment groups but did not occur in a dose-related manner;
The occurrence of bilateral accessory ribs was increased significantly
in fetuses of dams given 1,000 mg/kg for 10 days during gestation.
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Picloram August, 1988
-8-
At this dose level, there was also maternal toxicity . Since adverse
effects occurred at all doses, 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).
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. In
addition, 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
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Picloram August, 1988
-9-
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 same study, NCI (1978), 50 male and 50 female mice received
picloram at 208 or 417, and 361 or 723 mg/kg/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 B6C3F-| mice.
Dow (1986) retested picloram (94% pure) in a 2-year chronic feeding/
oncogenicity study in Fisher 344 rats. Rats (50/sex/dose) were fed 0,
20, 60 or 200 mg/kg/day. Oncogen!c 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 (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = mg/L ( u /L)
(UF) x ( L/day)
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, 1,000 or 10,000),
in accordance with EPA or 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
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Picloram August, 1988
-10-
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 (1981b) has been selected to serve
as the basis for the Ten-day HA value for pic lor am 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 7-day study
in 'dogs by Dow (198la) with a NOAEL of 400 mg/kg/day and a 32-day study in
mice by Dow (1980) 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/kg/day) (10 kg) = 2Q mg/L (20,000 ug/1)
(100) (1 L/day)
where:
200 mg/kg/day = NOAEL based on the absence of reduced feed intake in
beagle dogs exposed to pic lor am for 7 to 14 days.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 pic lor am 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 trans ami nase 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 10-kg child is calculated as
follows:
where:
Longer-term HA = (7 mgAg/day) (10) = 0.7 mg/L (700 Ug/L)
(100) (1 L/day)
7 mg/kg/day = NOAEL, based on the absence of relative and absolute
liver weight changes.
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Pi dor am August, 1988
-11-
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or NAS/OOW 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 = (7 mg/kg/day) (70) =2.45 mg/L (2,000 ug/L)
(100) (2 L/day)
where:
7 mg/kg/day = NOAEL, based on the absence of relative and absolute
liver weight changes.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or 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. 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
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Pi dor am August, 1988
-12-
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
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. The results of this
study were chosen, even though they reflect less than lifetime exposure, because
the results indicate that the dog is more sensitive than the rat in a long term
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.
100 = uncertainty factor, chosen in accordance with EPA
or 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 (2,000 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 (500 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
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Picloram August, 1988
-13-
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. Results of the rat study are
questionable since possible cross-contamination may have occurred.
The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of picloram.
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).
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, 1988). 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 deriva-
tizing agent. Excess reagent is removed, and the esters are determined
by electron-capture (EC) gas chromatography. This method has been
validated in a single laboratory, and estimated detection limits have
been determined for analytes in this method, including picloram. The
estimated detection limit is 0.14 ug/1.
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Picloram August, 1988
-14-
VIII. TREATMENT TECHNOLOGIES
0 No information was found on treatment technologies capable of
effectively removing picloram from contaminated water.
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Picloram August, L988
-15-
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):447-45l. 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.* 1980. Picloram: Results of a 32-day toxicity tolerance in feed in
BgC^F^ mice. Dow Chemical Laboratories, Midland Michigan. (No Dow
number). Received May 16, 1980. EPA Accession No. 247156.
Dow.* 1981a. Results of range-finding study of picloram (4-amino-3,5,6-
trichloropicolinic acid) administered orally in gelatin capsules to
beagle dogs. Dow Chemical, Texas. TXT:K- 38323(24).
Dow.* 1981b. Results of a short-term palatability study of picloram (4-amino-
3,5,6-trichloropicolinic acid) fed in the diet to beagle dogs.
Dow Chemical, Texas. TXT: K-38323(25).
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 Fischer 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.D. 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.
-------
Picloram August, 1988
-16-
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 1^c-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, Ml. 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, 1988
-17-
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.
NAS. 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(1):36-42.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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 1988
-18-
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 pi dor am. 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. 1988. U.S. Environmental Protection Agency. U.S. EPA Method 515.1
- Determination of chlorinated acids in water by GC/ECD, April 14, 1988
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.
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August, 1988
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.
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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1610-18-0
Structural Formula
OCH,
H
H
H
2,4-bis(isopropylamino)-6-methoxy-s-triazine
Synonyms
0 Prometon; Gesafram 50; Ontracic 800; Primatol 25E; Pramitol; Methoxy-
propazine (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
Molecular Weight
Physical State (25°C)
Boiling Point
Melting Point
Density
Vapor Pressure (20°C)
Specific Gravity
Water Solubility (20°C)
Log Octanol/Water Partition
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
C10H19N50
225.34
White crystals
91 to 92°C
1.088 g/cm3
2.3 x 10"6 mm. Hg
750 mg/L
-1.06 (calculated)
Occurrence
Prometon has been found in 386 of 1,419 surface water samples analyzed
and in 36 of 746 ground water samples (STORET, 1988). Samples were
collected at 250 surface water locations and 250 ground water locations.
and prometon 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. This informa-
tion is provided to give a general impression of the occurrence of
this chemical in ground and surface waters as reported in the STORET
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Frometon August, 1988
-3-
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
0 Prometon residues resulting from agricultural practice have been detected
in California ground waters at 0.21-80 ppb (Eiden, 1987).
Environmental Fate
0 Prometon is stable to hydrolysis at pH 5, 7, and 9 at 25°C foj 40
days (Ciba-Geigy, 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 chromatography (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
Absorption
0 Prometon is rapidly absorbed from the gastrointestinal tract. Based
on the radioactivity recovered in the urine and faces, prometon is
completely absorbed within 72 hours in the rat (Bakke et al., 1967).
Distribution
0 Seventy-two hours after intragastric intubation of 1^C-prometon in
rats, no detectable levels of radioactivity were found in any of
the tissues examined (Bakke et al., 1967).
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Prometon August, 1988
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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, 1971).
0 Based on the metabolites formed, triazine ring cleavage apparently
does not occur during prometon metabolism (Ciba-Geigy, 1971).
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 LD5Q value for prometon ranges from 1,750 to 2,980 mg/kg
in rats and is 2,160 mg/kg in mice (Heister, 1983; NIOSH, 1985).
The acute inhalation LCso value in rats is >3.6 mg/L for 4 hours
(Meister, 1983).
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.
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Prometon August, 1988
-5-
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 mq/kg/day). At 300 ppm and less, the body
weight and food consumption for both males and females were comparable
to those of the controls. The NOAEL and LOAEL identified in this
study are 300 and 1,000 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, 1976).
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. 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
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Prometon August, 1988
-6-
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 Plorek 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
mgAg/day (98% a.i.) in corn oil. Maternal weight gain was depressed in
rats administered 120 and 360 mg/k9/day • A teratogenic NOAEL of 360
nig/kg/day (the highest dose tested) and a maternal-toxicity NOAEL of
36 mg/kg/day were identified.
8 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/k9/day
(the highest dose tested) and a maternal-toxicity NOAEL of 3.5 mgAg/day
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 (up to 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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Prometon August, 1988
-7-
UF = uncertainty factor (10, 100, 1,000, or 10,000)
in accordance with EPA or 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 prometon. It is therefore
recommended that the DUEL value adjusted for the 10-kg child (0.2 mg/L,
calculated below) be used at this time as a conservative estimate of the
One-day HA value.
Ten-day Health Advisory
Prometon has been the subject of several acute toxicity assays including
four-week feeding studies in rats and dogs, and developmental toxicity studies
with rats and rabbits in which pregnant animals were dosed for 10 days during
gestation (Killeen et al., 1976a; Florek et al., 1981; Killeen et al., 1976b;
Lightkep et al., 1982). The only toxic effect consistently observed in these
studies was reduced weight gain in treated animals. Although of appropriate
duration, these studies were judged unacceptable for deriving the Ten-day HA;
fluctuations in weight gain may not be an appropriately sensitive end point
of toxicity for the basis of the HA. For this reason, it is recommended that
the OWEL, adjusted for a 10-kg child (0.2 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 (1982), rats were fed a diet
containing 0, 10, 50, 100 or 300 ppm prometon (0, 0.5, 2.5, 5 or 15 mg/kg/day)
for 90 days. A NOAEL of 15 mg/kg/day (the highest dose tested) was identified,
Rats were also tested in a 4-week feeding study in which a a NOAEL of 100
mg/kg and a LOAEL of 300 mq/Kq/day were identified (Killeen et al. 1976a).
Although the NOAEL from the subchronic study can be used as a basis for the
Longer-term HA, lower NOAELs have been identified in 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). It is therefore recommended that the DWEL (0.5 mg/L,
calculated below) and the DWEL, adjusted for the 10-kg child (0.2 mg/L,
calculated below) be used as conservative estimates of the Longer-term HA
for the adult and child, respectively.
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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Prometon August, 1988
-8-
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. 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. Effects on body weight gain have been observed
in acute studies at doses as low as 120 mg/kg/day for rats (Florek et al.,
1981), 25 mg/kg/day for dogs (Killeen et al., I976b), and 24.5 mg/kg/day for
rabbits (Lightkep et al., 1982). One subchronic study was available (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 150
mg/kg was identified from the 4-week feeding study by Killeen et al. (1976a).
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)
RfD = (15 mgAg/day) = Q.QIS
(1,000)
where:
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 EPA or
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*015 mgAg/day) (70 kg) = 0.5 mg/L (50o ug/L)
(2 L/day)
where:
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Prometon August, 1988
-9-
0.015 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
The DWEL, modified for the 10-kg child and rounded to one significant figure,
is derived as follows:
DWEL = (0.015 mgAg/day) (10 kg) = 0>2 mg/L (200 ug/L)
child (1 L/day)
where:
0.015 mg/kg/day = RfD.
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of a child.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = 0.5 mg/L x 20% = 0.1 mg/L (100 ug/L)
where:
0.5 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.
II. ANALYTICAL METHODS
0 Analysis of prometon is by a gas chromatographic (GC) method appli-
cable to the determination of certain nitrogen-phosphorus containing
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Prometon August, 1988
-10-
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. This method has been validated in a single laboratory,
and the estimated detection limit for prometon is 0.3 ug/L.
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 an individual technology or a combination 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, 1988
-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.* 1971. Metabolism of s-triazine herbicides. Unpublished study.
EPA Accession No. 55672.
Ciba-Geigy.* 1976. Acute toxicity studies with prometon technical (97%) -
primary skin irritation test - albino rabbits. Industrial Bio-Text
Laboratories, Inc. IBT No. 8530-09308. Unpublished study. EPA Accession
No. 231815.
Ciba-Geigy. 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. 1985b. Photolysis of prometon 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. 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. 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. 1986*. Field disposition studies in California, Nebraska and
New York. Preprared by Daniel Sumner. 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.D. Christin, M.S. Christin and E. Marshall.* 1981. Teratogen-
icity 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, 1988
-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. Munulkin 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. Contam. Toxicol. 22(4/5):561-566.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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. Reg. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. U.S. EPA Method #507 -
Determination of nitrogen and phosphorus containing pesticides in 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.
-------
August, 1988
PRONAMIDE
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.
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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 23950-58-5
Structural Formula
TV 0 H CH3
Vg-.U-C.CH
I
CH,
3,5-Dichloro(N-1 , 1-dimethyl-2-propynyl)benzamide
Synonyms
0 Kerb®; Kerb® SOW; Propyzamide; RH315 (Meister, 1983).
Uses
0 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).
Properties (NIOSH, 1985; TDB, 1985)
Chemical Formula
Molecular Weight
Physical State (25°C)
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 Pronamide has been found in 20 of 391 ground water samples analyzed,
all in California (STORET, 1988). No surface water samples were
collected. Ground water samples were collected 391 locations.
The concentration of all samples found was 1.0 ug/L. STORET contains
no information on surface water sampling for pronamide. This informa-
tion is provided to give a general impression of the occurrence of
this chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
256.14
White crystals
154 to 156°C
8.5 x 10~5 mm Hg
0.48 gm/cc
15 ppm (mg/L)
3.05 to 3.27
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Pronamide August, 1988
-3-
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
o 14c-pronamide (100% radiopurity) at 1.5 ppm hydrolyzes very slowly
(10% of applied material) in sterile, deionized water buffered to
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).
0 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 C02 evolution of bacteria and
fungi (Lashen, 1970).
0 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.
0 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,b; Kostowska et al., 1978;
Walker, 1978; Zandvoort et al., 1979). Data are insufficient to
determine the effect, if any, of meteorological conditions or the
role leaching may play in pronamide dissipation.
0 The environmental fate of pronamide is the subject of several unpub-
lished, undated reports (Cummings and Yih; Fisher and Cummings; Rohm
and Haas; Satterthwaite and Fisher; Yih).
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Pronamide August, 1938
-4-
III. PHARMACOKINETICS
Absorption
0 No information on the absorption of pronamide was found in the
available literature. Data on the systemic toxicity of oral doses
indicate that pronamide is absorbed following ingestion.
Distribution
0 No information on the distribution of pronamide was found in the
available literature.
Metabolism
8 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) (Yin and Swithenbank,
undated). The major metabolite in the feces was 2-(3,5-dichlorophenyl)-
4,4-dimethyl-5-hydroxymethyloxazoline (15%), and the major metabolites
in the urine were
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Pronamide August, 1988
-5-
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, 1970; Regel, 1972). No signs of irritation were observed
by Powers (1970). Twenty-four hours after exposure, Regel (1972)
observed erythema, which subsided at 72 hours.
0 Powers (1970) 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.
Long-term Exposure
0 Charles River CD 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). These levels correspond to 0, 2.5, 7.5,
22.5, 67.5 or 202.5 mg/kg/day, respectively, assuming 1 ppm in the
diet of rats is equivalent to 0.05 mg/kg/day (Lehman, 1959). Signifi-
cant 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-ObservedAdverse-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, 1,350 or 4,050 ppm pronamide (100% a.i.) for 3 months (Larson
and Borzelleca, 19675)t These levels correspond to approximate doses
of 0, 11, 34 or 101 mg/kg/day, assuming 1 ppm in the. diet of dogs 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
34 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 the diet of
dogs 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.
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Pronamide August, 1988
-6-
0 Smith (1974) administered Kerb® (97% a.i.) to 6-week-old (C57 BL16 x
C3H Anf)FI male and female mice (100/sex/dose), for 78 weeks at dietary
pronamide concentrations of 0, 1,000 or 2,000 ppm (0, 150 or 300 mg/kg/day,
assuming 1 ppm in the diet of mice is equivalent to 0.15 mg/kg/day;
Lehman, 1959). Male and female mice that ingested 2,000 ppm gained
significantly 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, 13, 70, 329 or 2,260 ppm (0, 2, 10,
49 or 340 mg/kg/day, assuming 1 ppm in the diet of mice is equivalent
0.15 mg/kg/day; Lehman, 1959) for up to 24 months. Another group was
fed 2,500 ppm 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 329 ppm (49 mg/kg/day) was identified.
Reproductive Effects
0 Costlow and Kane (1985) administered pronamide (technical, 97% pure)
to New Zealand White rabbits (18/dose) at doses of 0, 5, 20 or 80
mg/kg/day during gestation days 7 to 19. An increased incidence of.
gross and microscopic liver lesions, one maternal 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. 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 observed at higher doses.
0 In a three-generation reproduction study, 20 to 25 albino CD rats were
fed a diet containing pronamide (RH-315; 97% a.i.) 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, these levels correspond
to doses of 0, 1.5, 5 or 15 mg/lfig/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 con-
clusions. Based on this information a NOAEL of 300 ppm (15 mg/kg/day,
the highest dose tested) was identified.
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Pronamide August, 1988
-7-
Developmental Effects
0 Costlow and Kane (1985) administered pronamide (technical, 97% pure)
to New Zealand White rabbits (13/dose) at doses of 0, 5, 20 or 80
mg/kg/day (technical, 97% pure) during gestation days 7 to 19. 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 80 mg/kg/day based on the absence of
developmental effects at any dose tested. The NOAELs for developmental
effects and maternal toxicity, described above, are 20 and 5 mg/kg/day•
respectively.
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, 1971). No
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. Individual animal data
were not available to validate these conclusions, and, therefore,
the study was not validated. 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
(Fabrizio, 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 B6C3F1 mice (Newberne et al., 1982), technical
pronamide (97%) was fed to the animals (63 animals/dose) at target doses
of 0, 20, 100, 500 or 2,500 ppm. Actual measured levels were 0, 13,
70, 329, or 2262 ppm (0, 2, 10, 49 or 340 mg/kg/day assuming 1 ppm in
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Pronamide August, 1988
-8-
the diet of mice is equivalent to 0.15 mg/kg/day). Another group was
fed a target dose of 2,500 ppm pronamide 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 control
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 target 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 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 the diet of rats is equiva-
lent 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 (up to 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, 1,000 or 10,000)
in accordance with EPA or 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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Pronamide August, 1988
-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 pronamide. It is therefore
recommended that the Longer-term HA value of 0.8 mg/L (800 ug/L) be used at
this time as a conservative estimate of the One-day HA value for pronamide.
Ten-day Health Advisory
Toxicity from acute exposure to pronamide has been assessed in three
reproductive/developmental toxicity studies. No effects were observed in
rats exposed to pronamide via gavage (Vogin, 1972) or in feed (Larson and
Borzelleca, !967b) 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 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. Minor 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 used to derive
the Longer-term HA (7.5 mgAg/day; Larson and Borzelleca, 1970b), indicating
little difference may exist between doses which cause acute and chronic
toxicity. It is therefore recommended that the Longer-term HA value of 0.8
mg/L (800 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 both acute and chronic administra-
tion 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 dogs to 90 mg/kg/day
(Larson and Borzelleca, 1967b), and chronic feeding of 300 and 375 mg/kg/day
to mice (Smith, 1974; Newberne et al, 1982). Data on rats are equivocal. Signs
of hepatoxicity from subchronic exposure (increased liver weight after 90
days' exposure to 7.5 mgAg/day; Larson and Borzelleca, 1970a) were not observed
in a three generation study with doses as high as 15 mg/kg/day (Larson and
Borzelleca, 1970c).
These data indicate that there may be little difference between doses which
cause acute, subchronic or chronic toxicity. Therefore, it is recommended that
the basis for the Lifetime HA be used to derive a conservative estimate of the
Longer-term HA. In this study, beagle dogs fed a diet containing pronamide
a dose levels of 0, 30, 100 or 300 ppm (0, 0.75, 2.5 or 7.5 mg/kg/day) for
two years showed no adverse effects at any of the doses tested. A NOAEL of
7.5 (the highest dose tested) was identified in this study.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term = (7.5 mgAg/dayH 10 kg) = 0.8 mg/L
child (100)(1 L/day)
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Pronamide August, 1988
-10-
where :
7.5 mg/kg/day = NOAEL based on the absence of adverse effects in
dogs administered pronamide in the diet for 2 years.
10 kg = assumed body weight of a child
1 L/day = assumed daily water consumption of a child.
100 = Uncertainty factor, chosen in accordance with
NAS/ODW guidelines for use with a NOAEL from
an animal study.
The Longer-term HA for a 70 -kg adult is calculated as follows:
Longer-term = (7.5 ing Ag/day ) ( 70 kg) = 3.0 mg/L
adult (100) (2 L/day)
where:
7.5 mg/kg/day = NOAEL based on the absence of adverse effects in
dogs administered pronamide in the diet for 2 years.
70 kg = assumed body weight of an adult
2 L/day = assumed daily water consumption of an adult.
100 = uncertainty factor, chosen in accordance
with EPA or NAS/ODW guidelines for use with
a NOAEL from an animal study.
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
(RiD), 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. If the contaminant is classified as a Group A or B
carcinogen, according to the Agency's classification scheme of carcinogenic
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Pronamide August, 1988
-11-
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), mouse (Newberne et al. ,
1982), and dog (Larson and Borzelleca, 1970b). 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 deficiencies in the rat study (data on individual
animals not provided/ insufficient number of animals routinely monitored for
clinical chemistry parameters) 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 consideration, 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, 0.75, 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 mgAg/day) = Q.075 mg/kg/day
(100)
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 EPA
or 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 mgAg/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, 1988
-12-
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (2.6 mg/L) (20%) = Q.05 mg/L (50 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, 1986), pronamide may be classified
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 mg/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 (GO method appli-
cable to the determination of certain nitrogen-phosphorus containing
pesticides in water samples (U.S. EPA, 1988). 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. This method has been validated in a single laboratory,
and the estimated detection limit for pronamide is 0.76 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) report 99% removal efficiency of chlorinated
pesticides by a thin-film composite polyamide membrane operating at a
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Pronamide August, 1988
-13-
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.
-------
Pronamide August, 1988
-14-
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-14C-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 ^4C-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(l):83-36.
Hance, R.J., P.D. Smith, T.H. Byast and E.G. Cotterill. 1978a. Effects of
cultivation on the persistence and phytotoxicity 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.
Kostowska, 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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Fronamide August, 1988
-15-
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, 1988
-16-
Powers, M.B.* 1970a. Final Report - Acute Oral - Rats; Draize Eye - Rabbits;
Acute Dermal - Rabbits; Acute Inhalation Exposure - Rats. 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, Marysvilie, 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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.
U.S. EPA. 1985b. U.S. Environmental Protection Agency. Code of Federal
Regulations. 40 CFR 180.106. p. 252. July 1, 1985.
-------
Pronamide August, 1988
-17-
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. 1988. U.S. Environmental Protection Agency. U.S. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, 4/15/88. Available from U.S. EPA's Environmental
Monitoring and Support Laboratory, Cincinnati, OH.
Vogin, E.E.* 1971. 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.
Yih, 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, 1988
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.
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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1918-16-7
Structural Formula
2-chloro-N-isopropylacetinilide
Synonyms/Common Formulations
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 C^H^CINO
Molecular Weight 211.69
Physical State (room temp.) White crystalline solid
Boiling Point 110°C at 0.03 mm HG
Melting Point 77 to 78°C
Density (25°C) 1.13 g/mL
Vapor Pressure (25°C) 7.9 x 10~5 mm Hg
Specific Gravity 1.242
Water Solubility (20°C) 580 mg/L
Log Octanol/Water Partition 2.30
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Propachlor has been found in 34 of 1,690 surface water samples
analyzed and in 2 of 99 ground water samples (STORET, 1988). Samples
were collected at 475 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. This
information is provided to give a general impression of the occurrence
of this chemical in ground and surface waters as reported in the
STORET database. The individual data points retrieved were used as
they came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
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Propachlor August, 1988
-3-
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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,
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 1^C-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. 1^C-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.
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Propachlor August, 1988
-4-
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).
0 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
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).
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 LD5Q 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 LD5Q 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
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Propachlor August, 1988
-5-
(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
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 - the highest dose tested (37.5 mg/kg/day) was identified.
Dermal/Ocular Effects
0 The acute dermal LDsg 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 1.050 in rabbits (380 mg/kg); the lowest LD50 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,
conjunctival 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).
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Propachlor August, 1988
-6-
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
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 mgfkg/day (the highest
dose tested).
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Propachlor August, 1988
-7-
Mutagenicity
0 Technical propachlor was not genotoxic in assays of Salmonella
typhimurium 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 by intraperitoneal
injection at dose levels of 0.05, 0.2 or 1.0 mg/kg to F344 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 at levels up to 60 ug/ml. Propachlor was cytotoxic
at levels of 40 ug/ml and higher (Flowers, 1985). Primary rat
hepatocytes exposed to 0.1 to 5,000 ug/mL technical-grade propachlor
showed no effect on unscheduled DNA synthesis when compared to controls
(Steinmetz and Mirsalis, 1984).
Carcinogenicity
8 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 (up to 7 years) and lifetime exposures if adequate data are
available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) X (BW) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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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Propachlor August, 1988
-8-
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
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 contrasted by a developmental 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). 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) = o.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 EPA
or 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
While there are two 90-day studies suitable for developing 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 since acute data indicate toxicological concern at low
levels (Miller, 1983). The resulting Longer-term HA thus becomes 0.1 mg/L
and 0.5 mg/L for a 10-kg child and a 70-kg adult, respectively.
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Propachlor August, 1988
-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 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. 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.
While there are no suitable studies to calculate a Lifetime HA, 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 since it is the only
data available of extended duration. 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 (RfD)
RfD = (13.3 mg/kg/day) = 0.013 mg/kg/day (10 ug/kg/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 EPA
or 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 mgAg/day) (70 kg) = 0.46 mg/L (500 ug/L)
(2 L/day)
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Propachlor August, 1988
-10-
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 Health Advisory
Lifetime HA = (0.46 mg/L) (20%) = 0.092 mg/L (90 ug/L)
where:
0.46 mg/L = DWEL.
20% = assumed relative source contribution from water.
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 probable 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 ADZ 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
0 Analysis of propachlor is by a gas chromatographic (GC) method
applicable to the determination of certain chlorinated pesticides
in water samples (U.S. EPA, 1988). 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.
This method has been validated in a single laboratory, and the
estimated detection limit for analytes such as propachlor is 0.5 ug/L.
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Propachlor August, 1988
-11-
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
(7.9 x 10~5 mm Hg at 25°C). Propachlor probably would not be amenable
to aeration or air stripping because its Henry's Coefficient is
approximately 3.8 x 10~9 m^ atm. mol'l. 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). .
- 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 1988
-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 in vivo rat bone marrow cytogenetics assay
(SR 84-180): Final Report: SR~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 and
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
-------
Propachlor August, 1988
-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 Water Research Division. December.
Monsanto Company.* Undated. Toxicology. Summary of studies 241666-C through
241666-E. Unpublished study. MRID 25527.
NAS. 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 00152865.
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(7):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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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Propachlor August, 1988
-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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 508 - Deter-
mination of chlorinated pesticides in ground water by GC/ECD, 4/15/88
draft. Available from U.S. EPA's Environmental Monitoring and Support
Laboratory, Cincinnati, OH.
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
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August/ 1988
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.
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, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 139-40-2
Structural Formula
ci
H
H H
6-Chloro-N,N1-bis(1-methylethyl)-1-3,5-triazine-2,4-diamine
Synonyms
• 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; CHEMIAB, 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 10~8 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 33 of 1,097 surface water samples
analyzed and in 15 of 906 ground water samples (STORET, 1988).
Samples were collected at 244 surface water locations and 607 ground
water locations. The 85th percentile of all non-zero samples was
2.3 ug/L in surface water and 0.2 ug/L in ground water sources. The
maximum concentration found was 13 ug/L in surface water and 300 ug/L
in ground water. Propazine was found in five States in surface water-
and in four States in ground water. This information is provided' to
give a general impression of the occurrence of this chemical in
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Propazine August, 1988
-3-
ground and surface waters as reported in the STORET database. The
individual data points retrieved were used as they came from STORET
and have not been confirmed as to their validity. STORET data is
often not valid when individual numbers are used out of the context
of the entire sampling regime, as they are here. Therefore, this
information can only be used to form an impression of the intensity
and location of sampling for a particular chemical.
0 Propazine was detected in ground water in California at trace levels
«0.1 ppb) (U.S.G.S., 1985).
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 C02 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/gin.
• 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.
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Propazine August, 1988
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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.
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 1^C-propazine (41 to
56 rag/kg) to rats by gastric intubation. Radioactivity in a variety
of tissues was observed to decrease from an average of 46.7 ppra 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
0 Eighteen metabolites of propazine have been identified in the urine of
rats given single oral doses of 14C-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 14C-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.
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Propazine August, 1988
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Animals
Short-term Exposure
0 The reported acute oral 1*059 values for propazine (purity not specified)
were >5,000 mg/kg in mice (Stenger and Kindler, 1963b) and 1,200 mg/kg in
guinea pigs (NIOSH, 1987).
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$Q 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 arabic suspension of propazine (0.4 mL/animal) was applied
once a day for 5 consecutive days to shaved and intact skin of five
rats then washed away 3 hours after application.
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.
Long-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 the assumption that 1 ppm in
the diet is equivalent to 0.05 mg/kg/day (Lehman, 1959), these doses
correspond to 0, 2.5, 10 or 50 mg/kg/day. No compound-related changes
were observed in appearance, general behavior, hematology, clinical
chemistry, urinalysis, gross pathology and histopathology. There was
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Propazine August, 1988
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a 12% reduction (p <0.01) in body weight of females at 1rOOO ppm (50
mgAg/day) at the end of the study. Based on body weight loss, a
NOAEL of 200 ppm (10 mg/kg/day) and a Lowest-Observed-Adverse-Effect
Level (LOAEL) of 1,000 ppm (50 mg/kg/day) were identified.
0 Geigy (1960) dosed rats (12/sex/dose) of an unspecified strain with
propazine (50% a.i.) by stomach tube for 90 days at 0, 250 or 2,500
nig/kg/day (a.i.) or for 180 days at 0 or 250 mg/kg/day (a.i.). In the
90-day study, a reduction in body weight and feed consumption were
reported at 2,500 mg/kg/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 consump-
tion, gross appearance and behavior, mortality, gross pathology and
histopathology. This study identified a NOAEL 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
mgAg/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
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Propazine August, 1988
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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
decrease in mean pup weights for all generations except F-|a. 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 DMA 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 AKRJF} or (C57BL/6 x C3H/ANf)Ff for 18 months at a dose level of
0 or 46.4 mg/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.
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Propazine August, 1988
-8-
0 Jessup et al. (1980b) fed CO 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
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 CD 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 0.15 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 (up to 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, 1,000 or 10,000),
in accordance with EPA or 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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Propazine August, 1988
-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 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.
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 = MO mg/kg/day) (10 kg) = 1<0 mg/L (1,000 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 EPA or
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 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/kg/day) (10 kg) = 0<5 m /L (500 u /L)
(100) (1 L/day)
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Propazine August, 1988
-10-
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.
10 kg = assumed body weight of a child.
100 = uncertainty factor/ chosen in accordance with EPA or
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 - (5 mg/kg/day) (70 kg) = 1>75 m /L (2,000 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 EPA or
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
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Propazine August, 1988
-11-
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed. 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
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) = Q.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 EPA or
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)
DWEL = (0*02 mg/kg/day) (70 kg) = 0.70 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.
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Propazine August, 1988
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Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0*70 mq/L) (20%) = Q.014 mg/L (10 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.
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 rag/kg/day (Jessup et al«,
1980a) or in an 18-month feeding study in mice at a dose of 46.4
ing/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).
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 (1986b) has established residue tolerances of 0.25 ppm
for propazine in or on various agricultural commodities (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
mg/kg/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 mgAg/day and a
relative source contribution factor of 20%.
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Propazine August, 1988
-13-
I. ANALYTICAL METHODS
0 Analysis of propazine is by a gas chromatographic (GC) using method #507,
a method applicable to the determination of certain nitrogen-phosphorus
containing pesticides in water samples (U.S. EPA, 1988). 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. This method has been validated in a single laboratory and the
limit of detection for propazine was 0.13 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, 1988
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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. 15(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 embryo toxic effects. Experiment No. 227642.
Unpublished study. MRID 00087879.
Geigy, S.A.* 1960. Chronic toxicity of propazine 50 HP. Unpublished study.
MRID 00111671.
Hayes, W.J. 1982. Pesticides studied in man. Baltimore, MD: Williams and
Wilkins. p. 564.
IPC.* 1984. Industrie Prodotti Chimici. Atrazine product chemistry data.
Unpublished compilation. MRID 00141156.
Innes, J. , B. Ulland, M.G. Valerio, 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.
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 and L.J. Ackerman.* 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 chemicals 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. 1987. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances (RTECS). Microfiche edition.
July.
-------
Propazine August, 1988
-15-
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:
subchronische toxizitat-rratte, haut. Unpublished study, including German
text. MRID 00111677.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Strasser, F.* 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. 51(185):33992-34003. September 24.
U.S. EPA. 1986b. 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. EPA. 1988. U.S. Environmental Protection Agency. U.S. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
ground water by GC/NPD, April 15 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH. 45268.
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Propazine August, 1988
-16-
U.S.G.S. 1985. U.S. Geological Survey. Regional assessment project. C. Eiden.
Wazeter, P., 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, P., 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•
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August, 1988
PROPHAM
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.
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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Prop-.am August, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 122-42-9
Structural Formula
0
•N-C-0-CH(CH3)2
H
Phenyl 1-methylethyl carhamate; isopropyl-N-phenylcarbamate;
Isoproopyl carbanilate.
Synonyms
0 IPC, IFC, Ban-Hoe/ Beet-Kleen, Chera-Hoe, Premalox, Triherbide-IPC,
Tuberite (Meister, 1988).
Uses
0 Pre- and postemergence herbicide for control of weeds in alfalfa,
clover/ flax/ lettuce/ safflower/ spinach/ sugar-beets, lentils and
peas and on fallow land. Prevents cell division. Acts on meristematic
tissue (Meister, 1988).
Properties (Meister, 1988; Cohen, 1984; CHEMLAB, 1985; TDB, 1985)
Chemical Formula C10H13°2N
Molecular Weight 179.21
Physical State (25»C) White crystals
Boiling Point (at 25 mm Hg)
Melting Point 87°C
Density 1.09 g/mL (20'C)
Vapor Pressure (25°C) (sublimes slowly at 25°C)
Specific Gravity (20°C/20«C) 1.09
Water Solubility (25°C) 250 mg/L
Log Octanol/Water Partition 1.22 (calculated)
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor —
Occurrence
0 Propham has b»en found in 1 of 392 surface water samples analyzed
and was undetictable in 583 ground water samples (STORET/ 1988).
Samples were collected at 131 surface water locations and 572 ground
water locations/ and propham was found in Texas. The 85th percent!le
of the sample and the maximum concentration found in surface water
was 2 ug/L. This information is provided to give a general impression
of the occurrence of this chemical in ground and surface waters as
reported in the STORET database. The individual data points retrieved
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Propham August, 1988
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were used as they came from STORET and have not been confirmed as to
their validity. STORET data is often not valid when individual
numbers are used out of the context of the entire sampling regime, as
they are here. Therefore, this information can only be used to form
an impression of the intensity and location of sampling for a particular
chemical.
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), 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 14c-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
loam 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).
o 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
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Propham August, 1988
-4-
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 Wiedraann, 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
the 0- to 6-inch depth of silt loam soil (Wiedmann and Pensyl, 1981).
Ring-labeled 14c-propham (formulated as ChemHoe 135) applied at 4 Ib
a.i./A dissipated with a half-life of <7 days in the upper 3 inches
of silt loam soil treated in November, 1981 (Wiedmann et al., 1982).
The second half-life occurred approximately 133 days post-treatment.
III. PHARMACOKINETICS
Absorption
0 After oral administration of 1,100 rag/kg 14c-isopropyl-labeled propham
(99% a.i.) to rats (1,100 mg/kg), 88% of the label appeared in urine
within 4 days. After oral doses of 1,100'mg/kg of 14C-phenyl-labeled
propham, 96% was excreted in urine and 2% was excreted in feces
(Chen, 1979).
0 Fang et al. (1972) reported that in rats given oral doses (ranging
from less than 4 mg/kg to 200 mg/kg) of 14C-propham (99% a.i.)
80 to 85% was excreted in urine and 5% was expired in air, indicating
that propham is well absorbed (85 to 98%) from the gastrointestinal
tract.
Distribution
0 Chen (1979) administered single oral doses of ^c-phenyl- or
14c-isopropyl-labeled propham (1,100 mg/kg 99% a.i.) to rats. Trace
amounts of both 14c-phenyl- or 14c-isopropyl-labeled (0.5 to 1.2%)
propham were present in the liver, kidneys, muscle and carcass after
48 hours.
0 Paulson and Jacobsen (1974) administered single oral doses of
14c-propham (100 mg/kg 99% a.i.) to goats. Six hours later, only low
levels (0.2%) were detectable in milk.
Metabolism
0 Chen (1979) administered single oral doses of 14C-phenyl-labeled
propham (1,100 mg/kg 99% a.i.) to rats by gavage. Most of the dose
(96%) was excreted in urine as metabolites. The primary metabolites
identified were the sulfate ester conjugate and the glucuronide
conjugate of isopropyl 4-hydroxycarbanilate, which accounted for 78
and 1.3%, respectively, of the total primary metabolites recovered.
Similar studies in rats (single oral dose of 100 mg/kg) by Paulson
et al. (1972) support the rapid metabolism and excretion of propham.
In these studies a third metabolite (the sulfate ester of 4-hydroxy-
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Propham August, 1988
-5-
acetanilide) and a fourth (unidentified) metabolite were found to
account for 12.3% and 8.9%, respectively, of the total metabolites
detected in urine. The data demonstrate that ring hydroxylation at
the 4-position and subsequent conjugation as well as hydrolysis and
subsequent tl-acetylation occurred prior to excretion.
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 14c-propham (100
mg/kg; 99% a.i.) was excreted in the urine and feces, respectively
(Chen, 1979; Paulson et al., 1972).
Pang et al. (1972) reported that after oral administration of ring-
or chain-14c-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 rugae
and irregular thickening of the stomach. The acute oral LD5Q values
in male and female rats were reported to be 3,000 ± 232 mg/kg and
2,360 ±118 mg/kg, respectively. A No-Observed-Adverse-Effect Level
(NOAEL) cannot bQ derived from the study because the doses used were
too high, and adverse effects were found at all doses tested.
0 Brown and Gross (1949) reported that when a single dose of 1,136
mg/kg propham (purity not specified) was administered orally to 8
Sprague-Dawley rats, no adverse effects were observed. A dose of
2,174 mg/kg resulted in periods of light anesthesia, a higher dose
of 3348 mg/kg resulted in light anesthesia and death (one death in
6 tested). Deep anesthesia was produced when 4,425 mg/kg of propham
was administered to rats with subsequent death of 38% of the test
animals (3 deaths in 8 animals tested).
0 The acute inhalation LCso value in albino rats was reported to
be 10.71 mg/L (or 10,710 mg/m3, PPG Industries, 1978).
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Propham August, 1988
-6-
Dermal/Ocular Effects
0 The acute dermal LD$Q 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).
Long-term Exposure
0 Tisdel et al. (1979) fed Sprague-Dawley rats ( 30/sex/dose) propham
(technical grade, purity not specified) in the diet at 0, 250, 1,000
or 2,000 ppm for 91 days. Assuming that 1 ppm in the diet of rats is
equivalent to 0.05 mg/kg/ day (Lehman, 1959), these levels are equivalent
to 0, 12.5, 50 or 100 mg/kg/day. Following treatment, body weight,
organ weight, growth, clinical chemistry, gross pathology and histo-
pathology were evaluated. No effects were reported at 1,000 ppm
( 50 mg/kg/day) or lower in any parameters measured. At the highest
dose (2,000 ppm or 100 mg/kg/day) there was a significant increase in
spleen weight (p <0.05) and in spleen-to-body weight ratio (p <0.01)
in males, and a 70% inhibition of plasma cholinesterase (p <0.01) in
females at 45 days. Based on the above data, a NOAEL of 1,000 ppm
(50 mg/kg/day) was identified.
Reproductive Effects
0 In a report of a three-generation rat reproduction study, Ravert
(1978) reported data from the ?2 to weaning of the F2b generation.
Sprague-Dawley rats (10 males or 20 females/ dose) were administered
technical grade propham (purity not specified) in the diet at dose
levels of 0, 87.5, 250, 750 or 1,500 ppm for 9 weeks prior to breeding
for each parental generation. Assuming that 1 ppm in the diet of
rats is equivalent to 0.05 mg/kg/day (Lehman, 1959), these levels are
equivalent to 0, 4.4, 12.5, 37.5 or 75 mg/kg/day. It was not clear
whether the test animals were also fed propham-containing diets
during pregnancies or through weaning of offspring. No effects were
reported on fertility, mortality or pup development at any dose level
tested.
Developmental Effects
0 Ravert and Parke ( 1 977) administered technical propham (purity not
specified) by gavage to pregnant Sprague-Dawley rats (16 to 20/dose) ,
at levels of 0, 37.6, 376 or 1,879 mg/kg/day on days 6 through
15 of gestation. End points that were monitored included maternal
and fetal body weight and the number of corpora lutea, implants, live
fetuses and dead fetuses. Fetuses were also examined for soft-tissue
and skeletal anomalies. The only effects detected were reduced
maternal and fetal body weights and higher resorption rates at the
highest dose tested (1,879 mg/kg) and increased incidences of incomplete
ossification of the parietal and frontal bones of the skull at 375.8
and 1,879 mg/kg. An apparent NOAEL appears to be 37.6 mg/kg/day.
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Propham August, 1988
-7-
However, in this experiment, the high dose (1,879 mg/kg/day is too
high (i.e., one-half of the LD50); nearly two-thirds of the pregnant
rats at this dose died prior to scheduled sacrifice. Further, the dose
intervals are also relatively large. Therefore, a reliable NOAEL can
not be determined accurately due to the large difference in dosages
tested and the marginal effect noted at 376 mg/kg/day (For more
information on the developmental effects, see Worthing, 1979).
Mutagenicity
0 Using the Ames Salmonella test, Margard (1978) reported that propham
(purity not specified, 1,000 ug/plate) did not show any indications
of mutagenic activity either with or without activation.
0 When propham (100 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
0 Znnes et al. (1969) administered propham to (C57BL/6 X C3H/ANf) or
(C7BL/6 X AKR) mice (18/sex) in the diet at 0 or 560 ppm for 18 months.
According to the author, this corresponds to a dose of about 0 to
215 mg/kg. 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 dietary assumptions appropriate for guinea
pigs are also appropriate for hamsters and 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.
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Propham August, 1988
-8-
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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»Ef feet 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, 1,000 or 10,000),
in accordance with EPA or 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 propham. It is, therefore,
recommended that the Longer-term HA value for a 10 -kg child, 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 Longer-term HA of 5 mg/L for a 10 -kg child, calculated below, is
used for the Ten-day HA since the apparent NOAEL (37.6 mg/kg/day) in the
teratology study by Ravert and Parke (1977) was not necessarily the highest
NOAEL, due to the large difference between the doses selected (a ten-fold
difference between 37.6 and 376 mg/kg/day).
Longer-term Health Advisory
The study by Tisdel et al. (1979) has been selected to serve as the
basis for the Longer-term HA value for propham. In this study, rats were fed
propham in the diet for 91 days. At 100 mg/kg/day, plasma cholinesterase was
inhibited (70%) and spleen-to-body weight ratios were increased. No effects
were observed at 50 mg/kg/day. This NOAEL is supported by the NOAEL of 75
mg/kg/day identified in the three-generation reproduction study in rats by
Ravert (1978).
Using a NOAEL of 50 mg/kg/day, the Longer-term HA for a 10 -kg child is
calculated as follows:
Longer-term HA = (50 mg/kg/day) (10 kg) =5.0 mg/L (5,000 ug/L)
(100) (1 L/day)
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Propham August/ 1988
-9-
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.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 = (50 mg/kg/day) (70 kg) = j 7. 5 mg/L (20,000 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 EPA
or 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. If the contaminant is classified as a Group A or B
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Propham August, 1988
-10-
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 91-day study by Tisdel et al. (1979), which
identified a NOAEL of 50 mg/kg/day and was selected to serve as the basis for
the Longer-term HA, has also been selected for deriving the Lifetime HA.
Using this study, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (50 mg/kg/day) = 0.02 mg/kg/day (rounded from
M'OOO) (3) 0.017 mg/kg/day)
where:
50 mg/kg/day = NOAEL, based on the absence of any cholinesterase
inhibition or effects on organ weights in rats fed
propham in the diet for 91 days.
1,000 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study of less-than-lifetime duration.
3 = additional uncertainty factor. This factor is used
to account for a lack of adequate chronic toxicity
studies in the data base, preventing establishment
of the most sensitive toxicological end point.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.017 mg/kg/day) (70 kg) = Q.595 mg/L (600 uq/L)
(2 L/day)
where:
0.017 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 = (0.595 mg/L) (20%) = 0.12 mg/L (100 ug/L)
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Propham August, 1988
-11-
where:
0.595 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer (IARC, 1976) evaluated
propham and concluded that the carcinogenic potential is currently
cannot be determined.
0 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.
VTI. ANALYTICAL METHOD
0 Analysis of propham is done by Method #4 of the NFS survey methods,
"Determination of Pesticides in Groundwater by High Performance Liquid
Chromatography with an Ultraviolet Detector" (U.S. EPA, 1986b). In
this method a 1 liter sample is extracted with methylene chloride,
reduced to 5 ml with methanol substitution and analysis by HPLC with
a UV detector. The method is being validated in a single laboratory
and the estimated detection limit for propham is 0.75 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that granular activated carbon (GAG) 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.
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Propham August, 1988
-12-
The studies cited above indicate that GAC adsorption is the most
promising treatment technique for the removal of propham from water.
However, selection of an individual technology 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.
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Propham August, 1988
-13-
IX. REFERENCES
Brown, J.H. and P. Gross.* 1949. Acute toxicity study of isopropyl n-phenyl
carbamate. Uhpublished study. MRIO 00075264.
CHEMLAB. 1985. The Chemical Information System, CIS, Inc., Bethesda, MD.
Chen, Y.* 1979. Summary of animal metabolism of IPC. Uhpublished study.
MRID 00115438.
Cohen, S.Z. 1984. List of potential groundwater contaminants. Memorandum
to I. Pomerantz. Washington, D.C.: U.S. Environmental Protection Agency.
August 28.
Fang, S.C., E. Fallin, M.L. Montgomery et al.* 1972. Metabolic studies of
14C-labeled propham and chloropropham in female rats. Unpublished study.
MRIO 00037854.
Fang, S.C. and E. Fallin. 1974. Metabolic studies of 14c-labeled propham
and chloropropham in the female rat. Pest. Biochem. Physiol. 4:1-11.
Friedrick, U. and G. Nass. 1983. Evaluation of a mutation test using S49
mouse lymphoma cells and monitoring simultaneously the induction of
dexamethasone resistance, 6-thioguanine resistance and ouabain resistance.
Mutat. Res. 110:147-162.
Gusik, F.F.* 1976. Photolysis of carbon 14 ring-labeled isopropyl carbanilate
(IPC) in water. Uhpublished study received Sept. 17, 1979 under 748-224;
submitted by PPG Industries, Inc., Barberton, OH; CDL:240988-C. MRIO
00115466.
Hardies, D.E.* 1979. Metabolism of isopropyl carbanilate on a Wooster silt
loam soil: BR 21422. Unpublished study received Sept. 17, 1979 under
748-224; submitted by PPG Industries, Inc., Barberton, OH; COL:240988-1.
MRIO 00115472.
Hardies, D.E.' and D.Y. Studer.* 1979&. Metabolism of isopropyl carbanilate
on a Woodburn silt loam soil: BR 21448. Unpublished study received
Sept. 17, 1979 under 748-224; submitted by PPG Industries, Inc., Barberton,
OH; CDL:240998-F. MRID 00115469.
Hardies, D.E. and D.Y. Studer.* 1979b. Metabolism of isopropyl carbanilate
on an Altvan loam soil: BR 21531. Unpublished study received Sept. 17,
1979 under 748-224; submitted by PPG Industries, Inc., Barberton, OH;
COL:240988-H. MRID 00115471.
Hardies, D.E. and D.Y. Studer.* 1979c. Metabolism of isopropyl carbanilate
on a Hanford sandy loam: BR 21566. Unpublished study received
Sept. 17, 1979 under 748-224; submitted by PPG Industries, Inc., Barberton,
OH; CDL:240988-G. MRID 00115470.
Hardies, D.E. and D.Y. Studer.* 1979d. Absorption of isopropyl carbanilate
on five soil types: BR 21590. Unpublished study received Sept. 17,
1979 under 748-224; submitted by PPG Industries, Inc., Barberton, OH;
CDL:240987-C. MRID 00038945.
-------
Propham August, 1 988
-14-
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. 21(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. Valerio, L. PetruceHi, 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. 1988. 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 reproduction 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.
-------
Prophara August/ 1988
-15-
Ryan, A.J. 1971. The metabolism of carbamate pesticides. CRC Grit. Rev.
Toxicol. 1:33-51.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
TDB. 1985. Toxicology Data Bank. Medlars II. National Library of Medicine's
National Interactive Retrieval Service.
Terrell, Y. , and G. Park.* 1977. Acute oral toxicity in rats (IPC technical).
Unpublished study. MRID 00115421.
Tisdel, M., G. Rao, G. Thomson et al.* 1979. IPC (propham) subchronic oral
dosing study in rats. Unpublished study. MRID 00128777.
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 #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.
Van Esch, G.J., and R. Kroes. 1972. Long-term toxicity studies of chloro-
propham and propham in mice and hamsters. Food Cosmet. Toxicol.
10:373-381.
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.
Wiedmann, J.L., and J. Pensyl.* 1981. Dissipation of IPC and PPG-124 in soil
treated with ChemHoe 135 — Spring 1980: BR 22412. Unpublished study
received Dec. 20, 1982 under 748-224; submitted by PPG Industries, Inc.,
Barberton, OH; CDL:249100-A. MRID 00121299.
Wiedmann, J.L., D. Mattle, D.R. Coffman and J. Pensyl.* 1982. Determination
of IPC and PPG-124 residues in soil treated with carbon 14 labeled
Chemhoe 135: BR 22882. Unpublished study received Dec. 20, 1982 under
748-224; submitted by PPG Industries, Inc., Barberton, OH; CDL:249100-B.
MRID 00121300.
Worthing, C.R. 1979. Pesticide manual, 6th ed. British Crop Protection
Council, Worcestershire, England.
Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
SIMAZINE
Healtii 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.
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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Simazine August, 1988
-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.
CAS__Np_._ 122-34-9
Strucjuraj Formula
Cl
c H n
2-Chloro-4,6-bis(ethylamino)-l,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; Reinert and
Rogers, 1987)
Chemical Formula
Molecular Weight 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 io-9 mm Hg
Water Solubility (20°) 3.5 mg/L
Log Octanol/Water Partition 2.51
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
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Simazine August, 1988
-3-
Occurrence
0 Simazine has been found in 922 of 5,873 burface water samples analyzed
and in 202 of 2,654 qround water samples (STORET, 1988). Samples
were collected at 6?6 surface water locations and 2,128 ground water
locations. The 85th percent!le of all non-zero samples was 2.18 ug/L
in surface water and 1.60 uq/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/L. Simazine was found in surface water in 16
states and in ground water in 8 states. This information is provided
to give a general impression of the occurrence of this chemical in
ground and surface waters as reported in the STORET database. The
individual data points retrieved were used as they came from STORET
and have not been confirmed as to their validity. STORET data is
often not valid when individual numbers are used out of the context
of the entire sampling regime, as they are here. Therefore, this
information can only be used to form an impression of the intensity
and location of sampling for a particular chemical.
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),
Envi r o nment_a_1__F_ate_
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).
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-4,6 bis(amino)-s-triazine, and several unidentified polar
-.ompounds 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).
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 included G-28279, 2-chloro-4,6-
bis(amino)-s-triazine, 2-hydroxy-4,6-bis(ethyl ami no)-s-triazine, and
2-hydroxy-4-ethylamino-6-amino-s-triazine.
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
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Simazine August, 1988
-4-
equilibrium) studies. Using batch equilibrium tests, Kd 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 d
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 (Kd
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, 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 and Turner, 1968; Helling,
1971).
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 (Harris, 1967; Ivey and Andrews, 1965;
Rodgers, 1968). 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 (Mattson et al., 1969; Martin et al.,
1975; Walker, 1976). The degradate, 2-chloro-4-ethylamino-6-amino-s-
triazine (G-28279) was detected at the 0- to 6-inch depcr •- d at the
6- to 12-inch depth (Mattson et al . , 1969; Martin et al
.,
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 (Larsen et al.,
1966; Flanagan et al., 1968; Kahrs, 1969; LeBaron, 1970; Smith et al.,
1975; Kahrs, 1977).
A recent review of Simazine environmental fate is provided by Reinert
and Rodgers (1987). This review may provide additional information
to the data summarized in the above sections.
III. PHARMACOKINETICS
Absorption
o
Orr and Simoneaux (1986) studied the metabolism of 14C-simazine
(uniformly labeled in the triazine ring) in groups of five male and
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Simazine August, 1988
-5-
five female Sprague-Dawley rats following oral administration in
single doses at 0.5 or 200 mg/kg. At the low dose, 51 to 62% of the
dose was eliminated as simazine equivalents in the urine and about
12% was found in the tissues, suggesting that about 63 to 74% (i.e.,
0.315 to 0.37 mg/kg) of the dose was absorbed. At the high dose,
only about 22% (i.e.v 44 mg/kg) of the dose was found in urine and
2% (4 mg/kg) in the tissues of males or females. A third group of
animals received daily single oral doses of unlabeled simazine at
0.5 mg/kg/day for 14 days followed by a single l^C dose. Elimination
of l^C in urine ranged from 59 to 66% of the dose, and 8% was found
in the tissues. Bakke and Robbins (1968) reported that, in goats and
sheep, from 67 to 77% of a dose (the dose was not reported in this
abstract) of 14C-simazine (given orally in gelatin capsules) was
excreted in urine. This suggests absorption was around 70%.
0 Bakke and Robbins (1968) reported that in goats and sheep, from 67 to
77% of a dose of ^C-simazine (given orally in gelatin capsules) was
excreted in urine. This suggests that absorption was approximately 70%.
DJ_s_tn_byti_on_
0 Orr and Simoneaux (1986) reported that 7 days following oral admini-
stration of l^C-simazine to male and female Sprague-Dawley rats at
0.5 or 200 mg/kg, the highest 14C residue levels, expressed as simazine
equivalents found in the red blood cells, were 16.3 to 19.9 ppm at
the high dose and 0.18 to 0.23 ppm at the low dose. Residue levels
in red blood cells were slightT-y higher in males than in females.
Lower concentrations were found in the other tissues ranging from 0.0
to 0.16 ppm at the low dose and from 0.78 to 5.2 ppm at the high
dose. Relatively higher residues were found in the liver and kidney
of rats receiving the low dose (0.1 to 0.16 ppm), with lower levels
found in fat, plasma and bone (0.0 to 0.03 ppm). A similar pattern
was observed in rats receiving the high dose, except that the spleen
contained higher 1*C residues (4.1 to 5.2 ppm) than the liver and
kidney (2.9 to 4.0 ppm). Residue levels in tissues of animals receiving
the low 1*C Jose following 14 days of repeated dosing were generally
lower than those in rats receiving a single low dose, except in red
blood cells. The authors suggested that repeated dosing with unlabeled
simazine significantly reduced the number of available binding sites
in all tissues, except the red blood cells.
0 In rats receiving the high dose, 14C residue levels, expressed as
ppm, were 29-fold (kidney) to 516-fold (spleen) higher than those in
animals receiving the low dose (Orr and Simoneaux, 1976).
M_ejj_bp_l_1_sjn
0 Simoneaux and Sy (1971) conducted a preliminary investigation on the
metaboli . :* ^-K ring-labeled simazine found in the urine of female
white rat.,. The rats were dosed once orally at 1.5 mg/kg, and the
urine eliminated within 24 hours was collected and analyzed by thin-
layer chromatography and thin-layer electrophoresis. The metabolites
identified by comparison to authentic compounds were conjugated
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Simazine August, 1988
-6-
mercapturates of the following 3 metabolites: hydroxysimazine,
2-hydroxy-4-amino-6-ethylamino-s-triazine, and 2-hydroxy-4,6-diamino-
s-triazine. These metabolites accounted for 6.8, 6.1 and 14% of the
administered radioactivity, respectively. When the urine was hydrolized
with performic acid (oxidizing agent) to cleave sulfur-carbon bonds
of the conjugated mercapturates, prior to thin-layer chromatography,
about 8.1, 7.7 and 31.3% of the radiolabel was detected as the three
identified metabolites, respectively. Approximately half of the
radioactivity in the urine was not identified.
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
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.
Guddewar and Dauterman (1979) purified glutathione S-transferase
61-fold from mouse liver. This enzyme conjugates chloro-s-triazine
herbicides. A chloro group at the C2-position was found to be
necessary for conjugation to occur. When this 2-chloro was replaced
by a methylmercapto group, no conjugation occurred. The reaction
rate decreased with triazine analogs lacking the alkyl side-chains.
Atrazine was conjugated faster than either simazine or propazine.
Similar results were obtained by Bohme and Bar (1967), 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 nng-labeled expound 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-dealkylation,
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.
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Simazine August, 1988
-7-
Excretion
0 Orr and Simoneaux (1986) studied the metabolism of ^C-cimazine
(uniformly labeled in the triazine ring) in Sprague-Dawley rats
following oral administration in single doses. Males and females
receiving 0.5 mg/kg eliminated 50.5 and 62.1% of the dose, respectively,
in the urine, and 19.1 and 13.3% of the dose in the feces, 7 days
after dosing. In animals receiving a single dose of l^C-simazine
following repeated dosing at 0.5 mg/kg/day for 14 days with unlabeled
simazine, males eliminated 58.5 and 24.5% of the dose in the urine
and feces, respectively, whereas females eliminated 66.0 and 17.8% of
the dose. A third group of males and females receiving a single dose
of 200 mg/kg eliminated about 21% of the dose in the urine and about
2.0% in the feces. The elimination pattern with time after dosing
was biphasic with rates being faster in females than males. Most of
the radioactivity was eliminated in 72 hours with a half-life of 9 to
15 hours. Elimination of the remaining radioactivity occurred at a
slower rate with half-life values ranging from 21 to 32 hours.
0 In a preliminary study with two female white rats dosed orally at
1.5 mg/kg ring-labeled l^C-simazine, Simoneaux and Sy (1971) found
less than 0.05% of the dose being eliminated as 14C-carbon dioxide
96 hours after dosing. Approximately 49.3% of the dose was eliminated
in the urine and 40.8% in the feces.
0 Bakke and Robbins (1968) studied the excretion of simazine in goats
and sheep using triazines labeled with *4C on the ring or on the
ethyl ami no sidechains. No *4C02 was detected from animals that
received the ring-labeled compounds, which suggested that the triazine
ring was not metabolized. Approximately 67 to 77% of the administered
ring-labeled activity was found in the urine, and 13 to 25% was found
in the feces. About 0.16 ppm of ^C residues were present in the
milk immediately after dosing, but the level decreased to 0.04 ppm
within 48 hours.
0 it. John et al. (1965) fed unlabeled simazine (5 pp.:.) to a lactating
cow for 3 days. Urine and milk were collected and analyzed during
the feeding period and for 3 days thereafter. No simazine was detected
in the milk (sensitivity of method approximately 0.1 ppm), and only
about 1% of the administered simazine was excreted in the urine as
the parent compound. Milk, excreta and body tissues were not analyzed
for simazine metabolites.
0 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.
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Simazine August, 1988
-8-
IV. HEALTH EFFECTS
Humans
There were 124 cases of contact dermatitis noted Dy Yelizarov (1977)
in the Soviet Union among workers manufacturing simazine and propazine.
Mild cases lasting 3 or 4 days involved pale pink erythema and slight
edema. Serious cases lasfiny 7 to 10 days involved greater erythema
and edema and a vesiculopapular reaction that sometimes progressed to
the formation of bullae.
Animals
Oral LDso 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).
Mazaev (1964) 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 l°o iC.cs 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% acti've 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 a: J
sheep by drench or capsule to 10 doses at 25, 50, 100 or 250 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 SOW (purity not stated) at either 31 daily doses of 50
mg/kg or 14 daily 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).
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Simazine August, 1988
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0 The acute inhalation LC^g 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).
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 caused transient
inflammation of conjunctivae, but was not irritating to the iris or
cornea (USDA, 1984).
Long-term j_X-Pj)_su_re.
° 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 t" saminase
(SCOT) and creatinine and increased cholesterol and inor~ ic 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.
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Simazine August, 1988
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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.
Assuming that 1 ppm in the diet of dogs is equivalent to 0.025
mg/kg/day, 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.
0 Dshurov (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. The 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 weed's. 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 whether the
authors considered the 50% formulation when providing the dosage
levels.
Reproductive j/fects
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 SOU
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 Fja litters. After weaning
first litters, parents were remated to produce F^ 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 ?2b litters.
F2b weanli—'s were fed the same dietary levels of simazine (0, 50
or 100 pp ' ~^b rats were mated to produce F3a and F3b 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 approxi-
mately 5 mg/kg/day.
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Simazine August, 1988
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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 spennatogenesis.
0
Simazine (97% a.i.) was administered by gavage to a group of 19 New
Zealand White rabbits (2.5 to 4.5 kg) at daily doses of 5, 75 or
200 mg/kg for days 7 through 19 of presumed gestation (Ciba-Geigy,
1984). The herbicide was delivered as a suspension in 3% cornstarch
containing 0.5 percent Tween 80. A control group was given the
vehicle only. One death was observed in each group; however, death
of the low-dosed dam was assumed to be due to a dosing accident and
not compound-related. Maternal toxicity at 75 and 200 mg/kg was
indicated by abortions, tremors, decreased motor control and activity,
ataxia (in one high-dosed animal), few or no stools, anorexia, weight
loss, and decreased body weight gain. Embryotoxicity was not oberved,
but fetal toxicity, which was believed to be the consequence of
maternal toxicity, was evident in the intermediate- and high-dose
groups as indicated by decreased numbers of viable fetuses. Fetal
toxicity at 200 mg/kg was also reflected by reduced fetal body weights,
increased occurrence of floating and fully formed ribs, and decreased
ossification of the patellae. No malformations were associated with
any dose level of simazine, and at 5 mg/kg the herbicide exerted
neither toxic nor teratogenic effects. The authors concluded that at
doses of 75 mg/kg/day and higher, simazine was very toxic to fetuses
and dams but was neither embryotoxic nor teratogenic.
Chen (1981) studied the teratogenic effects of simazine in rats. The
herbicide was administered by gastric intubation to pregnant rats
from the 6th to 15th day of gestation at dose levels of 78, 312,
1,250 and 2,500 mg/kg bw/day. Ossification of the sternum and cranium
was delayed at all four dose levels. At doses of 312 mg/kg and above,
body weight gain was inhibited in the dams, the number of live fetuses
was reduced, and the rate of fetal resorption increased. At the higher
dose level (2,500 mg/kg), simazine caused delayed fetal development.
The author concluded that simazine was toxic to rat embryos but was
not teratogenic.
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/rn^ via inhalation for 1 to 3 hours/day on days 7
through 14 of gestation.
Simazine has shown negative results in a variety of microbial
mutagenicity assay systems including tests with the following
organisms: SalrnoneJJj. J^J^JJLMTJM (Simmons et al., 1979;
Eisenbeis et al., 198T; Anderson e't al., 1972); Escherichia
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Simazine August, 1988
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coli (Simmons et al., 1979; Fahng, 1974); BaciUus subtilis
(Yimmons et al., 1979); Serratia marcescens (F^hrig,"T9T2fy;""and
l^L^lL^^-Si c_er_evi_sj_ae ~(S~i mmon s"e t ~a 1., 1979).
0 Simazine induced lethal mutations in the sex-linked recessive lethal
test using the fruitfly Pj_o_s_QJ>_hjJ_a. jjejjnogaster (Valencia, 1981).
In a study reported by Mur~nTk and Nash (l97T), simazine increased
X-linked lethals when injected into male D. ine1_an_og_a_st_er, 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).
Ca r c i n o g e n i c i ty
0 Simazine was not tumorigenie 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 120 compounds, male and female mice of two hybrid
strains (C57BL/6 x C3H/Anf)Fi and (C57BL/6 x AKR)Fj were exposed to
simazine (purity not stated) at- the maximum tolerated dose of 215 mg/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 (I960) conducted a 2-year dietary study in
Charles River rats administered simazine 50W (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,
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).
0 An interim report by Ciba-Geigy Corporation (1987) on a new chronic
feeding/oncogenicity study in Sprague-Dawley rats indicates that
simazine may cause mammary gland tumors. However, when this study is
complete, an evaluation of these data will be performed.
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Simazine August, 1988
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v- QUANT I F I CAT! ON OF TOX I COL06I_CA_L_ EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-oay,
longer-term (apprcximatelj 7 years) an- lifetirr.2 exposures if adequits data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
where:
HA = NL _or L_L__x_. . __ mg/L ( _ ug/L)
T * L __ 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, 1,000 or 10,000),
in accordance with EPA and NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
On e -day He a 1 t_h_
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 the Ten-day HA of 0.5 mg/L (500 ug/L) calculated below be used at this
time as a conservative estimate of the One-day HA value.
l£T~day _H_e_a_lt_h_^dvi sory
No suitable studies were found in the available literature for the deter-
mination of the Ten-day HA value for simazine, with the exception of a rabbit
teratology study by Ciba-Geigy (1984. In this study, simazine was administered
by gavage to female rabbits during gestation days 7 to 19 (13 days). A NOAEL
of 5 mg/kg/day for maternal toxicity was established and a LOAEL of 75 mg/kg/day
was noted based on abortions and tremors, decreased motor control and activity,
anorexia and weight loss. The NOAEL of 5 mg/kg/day in the rabbit is selected
for use as the basis for calculation of the Ten-day HA.
The Ten-day HA for a 10-kg child is calculated a? f'-iiov-s:
Ten-day HA = (5.0 mg /k g/dayJ (10 k gj = 0.5 mg/L (500 ug/L)
where:
5.0 mg/kg/day = NOAEL, based on the absence of adverse effects in
female rabbits given simazine by gavage during days
7 to 19 of gestation.
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Simazine August, 1988
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100 = uncertainty factor, chosen in accordance with EPA
or NAS/ODml guidelines for use with a NOAEL from an
animal study.
10 kg = assumed body weight of a child.
1 L/day = assumed water consumption of a 10-kg child.
Longer-term Health Ad vi sory
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
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 10-kg child and that
the DWEL of 0.2 mg/L (200 ug/L) be used for a 70-kg adult.
L i fet i me Health Ad vi spry
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 consume*•on of an
adult. The Lifetime HA is determined in Step 3 by factoring in <. er sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is Dased on actual exposure data or, if data are not available, a
value of 20% is assumed. 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 nated for 10 days, resulting in Fja litters.
After weaning first litters, parents were remated to produce FID 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 ^2b weanlings were then divided into 50- and 100-ppm dosage
groups. F2b rats were mated to produce F3a and f^b 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 5 mq/kq/dc-;..
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Simazine August, 1988
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It is important to note that, in this study, rats in the Fg generation were
exposed to simazine at the high dose (100 ppm) only. However, considering that
the FI and F? generations treated with ICO 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 ng/kg/day is used for calculation of
the RfD.
Step 1: Determination of the Reference Dose (RfD)
RfD = 1419/kg/day. = 0.005 mg/kg/day
HTrffifr)
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 EPA or
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 (DUEL)
DWEL = 10.005 jnq/kg/day) (70 kg) = Otl75 mg/L (2oo 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%) = 0.0035 mg/L (4 ug/L)
where:
0.175 mg/L = DWEL
20% = Assumed relative source contribution from water.
10 = additional uncertainty factor, according to ODW policy,
to account for possible carcinogenicity.
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Simazine August, 1988
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E va1uat 1 on of C a rcj nogen i c Potent i a1
0 An interim report by Ciba-Geigy Corporation (1987) on a new chronic
Teeaing/oncogemcity study in Sprague-Dawley rats indicates that
simazine reflects some oncogenic activity in the mammary glands as
previously noted with atrazine and propazine. Apparently this
positive study has been completed recently and submitted to the
Office of Pesticide Programs. An evaluation of this study will be
performed in the near future. Neither the study in mice by Innes et
al. (1969) nor the study in rats by Hazelton Laboratories (1960) is
considered adequate for assessment of the carcinogenicity of this
substance. However, the Hazelton rat study (I960) still reflected a
potential positive oncogenic effect in the same target organ, the
mammary glands.
0 Simazine is a chloro-s-triazine derivative, with a chemical structure
analogous to atrazine and propazine. Both of these 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.
0 Applying the criteria described in EPA's guidelines for the assessment
of carcinogenic risk (U.S. EPA, 1986a), simazine may be preliminarily
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 Programs completes
a peer review of the weight of the evidence for simazine and its
analogs.
VI. OTHER CRITERIA, GUI DANCE 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 appli-
cation to growing aquatic weeds (U.S. FDA, 1979).
0 Residue tolerances have been established for simazine alone and the
combined residues of simazine and its metabolites 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 pesti-
cides in water samples (U.S. EPA, 1988). In this method, Method #507,
approximately 1 L of sample is extracted with methylene chloride.
The extract is concentrated and the compounds are separated using
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Simazine August, 1988
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caplllary column GC. Measurement is made using a nitrogen-phosphorus
detector. Tne method detection limit has not been determined for
the analytes in this method, including simazine. The estimated
detection limit is 0.075 ug/L.
VIII« 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 ineffective in
removing simazine (Miltner and Fronk, 1985a). 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
the simazine levels in the water from the Maumee River (Baker, 1983).
Miltner 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,
although no quantitative data on simazine concentrations were reported.
0 Rees and Au (1979) reported that an adsorption column containing XAD-2
resin removed 81 to 95% of the simazine in spiked tap water.
0 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.
0 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
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Simazine August, 1988
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interval. Ozonation of spiked, distilled water resulted in a 92%
removal of sinazine. Oxidation of spiked, distilled water with
hydrogen peroxide obtained a 19 to 42% removal of simazine, and in
spiked Ohio River water, a smaller doss^e cvc-r a shortor tinie 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 August, 1988
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IX. REFERENCES
Andersen, A.M. 1968. The inactivation of simazine and linuron in soil by
charcoal. Weed Res. 8:58-fiO.
Anderson, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of herbi-
cides for possible mutagenic properties. J. Agric. Food Chem. 20:649-656.
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., 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.
Chen, B.Q. 1981. Experimental studies on toxicity and teratogenicity of
simazine. Chung Hua Fang I Hsueh Tsa Chih. 15(2):83-85.
Ciba-Geigy Corporation. 1980. 21-Day subacute dermal toxicity study in
rabbits. Bio-Research Laboratories, Project No. 12017. January 14.
MRID 00057567.
Ciba-Geigy Corporation. 1984. A teratology study of simazine technical in
New Zealnd White rabbits. MRID -o. 161407, Accession Nos. 252938 and
252938 (a 1986 amendment).
Ciba-Geigy Corporation. 1987. Rat chronic feeding/oncogenicity study (in
progress). MRID No. 401514-0, Accession No. 40154.
Cohen, S.Z., C. Eiden and M.N. Lorber. 1986. Monitoring ground water for
pesticides in the IKS.A. In_: Evaluation of pesticides in ground water,
American Chemical Society Symposium Series. #31E.
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]
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Simazine August, 1988
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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.
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Gold, B., K. Balu and A. Hofberg. 1973. Hydrolysis of simazine in aqueous
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00025441, 00025442, 00042793 and 00080626.
Helling, C.S. 1971. Pesticide mobility in soils: II. Applications of soil
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Helling, C.S., and B.C. Turner. 1968. Pesticide mobility: Determination by
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Innes, J.R.M., B.M. Ulland, M.G. Valeric et al. 1969. Bioassay of pesticides
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J. Natl. Cancer Inst. 42:1101-1114.
Ivey, M.J., and H. Andrews. 1965. Leaching of simazine, atrazine, diuron,
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Kahrs, R.A. 196.. ' Oi.cermination 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.
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Simazine August, 1988
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Kahrs, R.A. 1977. Simazine lakes—1975 EUP Program: Status Report—1977:
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Keller, A. 1978. Degradation of simazine (Gesatop) in soil under aerobic-
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6-isopropylamino-^-triazine (cyprazine) in the rat. J. Agric. Food Chem.
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Larsen, H., D.L. Sutton, A.R. Eaton et al. 1966. Summary of residue studies—
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LeBaron, H.M. 1970. Fate of simazine in the aquatic environment: Report No.
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Martin, V., L. Motko, B. Gold et al. 1975. Simazine residue tests: AG-A
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Mazaev, V.T. 1964. Experimental determination of the maximum permissible
concentrations of cyanuric acid, simazine, and a 2-hydroxy derivative of
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Miltner, R.J., and C.A. Fronk. 1985a. Treatment of synthetic organic contami-
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Miltner, R.J., and C.A. Fronk. 1985b. Treatment of synthetic organic contami-
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Monsanto Company, (date not available). The soil dissipation of glyphosate,
alachlor, atrazine and simazine herbicides. Unpublished study.
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Simazine August, 1988
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Murnik, M.R., and C.L. Nash. 1977. Mutagenicity of the triazlne herbicides
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4,6-bis(ethylamine)-s-triazine. Ph.D. Thesis. University of California.
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Simazine August, 1988
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Simazine August, 1988
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Confidential Business Information submitted to the Office of Pesticide
Programs.
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August, 1988
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.
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
modsls. 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, 1988
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 34014-18-1
Structural Formula M _
'
CK3
// \\
H3c - C C C — N—C-NH-CH3
1 \X 0
CH3 b °
N-[5-(1,1-Dimethyl ethyl)-1,3,4-thiadiazol-2-yl]-N,N1-dimethylurea
Synonyms
0 Combine; Herbic; Graslan; Perflan; Spike.
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 Cgl^gO^S
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) 2 x 10~^ mm Hg
Specific Gravity
Water Solubility (25°C, pH 7) 2,500 mg/L
Log Octanol/Water Partition
Coefficient 1.79
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 No occurrence data has been found in the STORET database (STORET, 1988)
Environmental Fate
0 Tebuthiuron is resistant to hydrolysis. 14C-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 (Mosier and
Saunders, 1976).
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Tebuthiuron August, 1988
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After 23 days of irradiation with artificial light (20-W black light),
tebuthiuron accounted for 87 to 89% of the applied radioactivity in
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 in 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 14C-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).
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Tebuthiuron August, 1988
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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,
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 mean levels of 2.7 and
6.2 ppm for the 100- and 200-ppm 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/kg. 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 dimethylethy1
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 mg/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 dog, 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.
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Tebuthiuron August, 1988
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IV. HEALTH EFFECTS
Humans
No information on the health effects of tebuthiuron in humans was
found in the available literature.
Animals
Short-term Exposure
0 Todd et al. (1974) reported the acute oral 1>D$Q 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 Todd et al. (1972a) supplied 40 weanling Sprague-Dawley rats (105-146g)
with food containing tebuthiuron (purity not stated) at levels of
2,500 ppm for 15 days. At various time periods, 5 rats were necropsied
and evaluated. 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/kq/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 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 Todd et al. (1974) 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,
but there was slight transient hyperemia of the conjunctiva.
All eyes were normal by the end of the 7-day test period.
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Tebuthiuron August, 1988
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Long-term Exposure
o 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/kg/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/kg/day tebuthiuron showed diffuse vacuolation of
the pancreatic acinar cells. The degree of this change ranged from
slight to moderate, but the effect was not associated with necrosis
or with the presence of an inflammatory response. One female rat
receiving 100 mg/kg/day tebuthiuron showed very slight pancreatic
changes. Based on these results, a No-Observed-Adverse-Effect-Level
(NOAEL) of 40 mg/kg/day and a Lowest-Observed-Adverse-Effect-Level
(LOAEL) of 100 mg/kg/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/kg/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/kg females also showed increased spleen-
to-body weight ratios. Histopathological findings were unremarkable.
The LOAEL was identified as 25 rag/kg, based on increased thyroid-to-
body weight ratios and increased alkaline phosphatase values. A
NOAEL of 12.5 mg/kg was identified.
o Todd et al. (I976a) administered tebuthiuron (purity > 97%)
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
80 mg/kg/day. Physical appearance, behavior, food intake, body
weight gain and mortality were recorded. Hematologic and blood
chemistry values were obtained throughout the study; urinalysis was
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Tebuthiuron August, 1988
-7-
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 ppm in the diet of a mouse is equivalent to 0.150
ing/kg/day; therefore, these dietary levels correspond to approximately
60, 120 or 240 mg/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.
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, 7, 14 or 28 mg/kg/day, based on actual food
consumption) for a period of 101 days preceding two breeding trials.
First generation (F1) 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 FQ males per treatment
group. In addition, representative F1a and F2a weanlings and F1 adults
were necropsied and given histopathologic examinations after live-phase
observations were completed. No changes in the efficiency of food
utilization (EFO) were noted during the F0 growth period, but during
the FI growth period, a statistically significant (p £0.05) depression
in cumulative (124 days) EFU values occurred in both male and female
rats receiving 28 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 F-| generation receiving 14 or 28 mg/kg/day; mean
body weight was depressed significantly (p £0.05) only in the high-
dose females. In the 7 mgAg/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
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Tebuthiuron August, 1988
-8-
found in any offspring. In adult males from the FQ generation, no
dose-related histologic lesions were found, and sperm morphology and
sperraatogenesis were normal. A LOAEL of 14 mg/kg/day was determined
for a lower rate of body weight gain during the 101-day pre-mating
period in F1 females, and a NOAEL of 7 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 ppm, 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.
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 exhibit unscheduled DNA synthesis.
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
typhimurium (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. (I976a) administered tebuthiuron (purity > 97%) 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.
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Tebuthiuron August, 1988
-9-
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.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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, 1,000 or 10,000),
in accordance.with EPA or 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, 2.5 mg/L (3,000 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 Todd et al. (1975) has been selected to serve as the basis
for the Ten-day HA value for tebuthiuron because the NOAEL in the Dutch-Belted
rabbit was the lowest end point observed in a short-term developmental study.
This study identified a NOAEL of 25 mg/kg/day (the highest dose tested) based
on an absence of maternal toxicity. In another developmental study in rats
by Todd et al. (1972d), a NOAEL of 90 mg/kg/day (the highest dose tested) was
recorded. Since it is unknown whether the rabbit or the rat is more sensitive,
the lower NOAEL was conservatively chosen in deriving the 10-day HA.
Using a NOAEL of 25 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
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Tebuthiuron August, 1988
-10-
Ten-day HA = (25 mg/kg/day) (10 kg) = 3.5 mg/L (3,000 ug/L)
(100) (1 L/day)
where:
25 mg/kg/day = NOAEL, based on the absence of maternal toxicity in
Dutch-Belted rabbits exposed to tebuthiuron by diet.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with EPA
or 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 two-generation reproduction study in rats (Hoyt et al., 1981) has
been chosen to serve as the basis for the Longer-term HA for tebuthiuron.
In this study, four groups of Wistar rats (25/sex) were fed tebuthiuron
at 0, 7, 14, and 28 mg/kg/day in the diet for 101 days (Fo 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. No adverse
effects were reported in this study except for a lower body weight gain
during premating period in F]_ females at the dietary levels of 14 and
28 mg/kg. The NOAEL was identified as 7 mg/kg/day. The chronic study
by Todd et al. (1976b) in mice was not selected because the weight loss
and vacuolization 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. The subchronic (90-day) feeding
study in beagle dogs reported by Todd et al. (1972) was not selected
because the NOAEL (12.5 mg/kg/day) identified in that study was significantly
higher than that of the rat study. In addition, the duration of the rat
study was more appropriate for the derivation of a Longer-term HA.
Using the NOAEL of 7 mg/kg/day, the Longer-term HA for a 10-kg child is
calculated as follows:
Longer-term HA = (7 mg/kg/day) d-0 *g) =0.7 mg/L (700 ug/L)
(100) (1 L/day)
where:
7 mg/kg/day = NOAEL, based on effects on the rate of weight gain in
rats exposed to tebuthiuron in the diet for 101 days.
100 = uncertainty factor, chosen in accordance with EPA
or 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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Tebuthiuron August, 1988
-11-
The Longer-term HA for the 70-kg adult is calculated as follows:
Longer-term HA = (7 mg/kg/day) (70 kg) =2.45 mg/L (2000 ug/L)
(100) (2 L/day)
where:
7.0 mg/kg/day = NOAEL, based on effects on the rate of weight gain in
rats exposed to tebuthiuron in the diet for 101 days.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA
or 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. 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 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, 7,
14 or 28 mg/kg/day in the diet for 101 days (F0 rats) or 121 days (F-, 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 Fia 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 14 and 28 mg/kg. The NOAEL was identified as 7 mg/kg/day.
The chronic study by Todd et al. (1976b) in mice was not selected because the
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Tebuthiuron August, 1988
-12-
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 7 mg/kg/day, the Lifetime HA is calculated 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 effects on the rate of weight gain in
rats exposed to tebuthiuron in the diet for 101 days.
100 = uncertainty factor, chosen in accordance with EPA
or 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 kg) = 2.45 mg/L (2,000 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 m9/L) (20%> = 0.49 mg/L (500 ug/L)
where:
2.45 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 tebuthiuron.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), tebuthiuron may be classified
in Group D: not classifiable as to human carcinogenicity. This
category is for substances with inadequate human and animal evidence
of carcinogenicity.
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Tebuthiuron August, 1988
-13-
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, 1988). 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. This method has been validated in a single
laboratory, and estimated detection limit for the analytes in the
method, such as tebuthiuron, is 1.3 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, 1988
-14-
IX. REFERENCES
Adams, E., J. Magnussen, j. Emmerson et al.* 1982* Radiocarbon levels in the
milk of lactating rats given I^C-tebuthiuron (compound 75503) in the diet.
Eli Lilly and Company, Greenfield, IN. Unpublished study. MRID 00106081.
Berard, D.F.* 1977. ^C-Tebuthiuron degradation study in anaerobic soil.
Prepared and submitted by Eli Lilly and Co., Greenfield, IN.
MRID 00900098.
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.
MRIO 000416090.
Day, E.W.* 1976a. Laboratory soil leaching studies with 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-I.
MRID 00020782.
Day, E.W.* 1976b. Aged soil leaching study with herbicide tebuthiuron. Unpub-
lished study received Feb. 18, 1977 under 1471-109; submitted by Elanco
Products Co., Div. of Eli Lilly and Co., Indianapolis, IN. CDL:095854-J.
MRID 00020783.
Elanco 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. Sieck, 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.
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.
-------
Tebuthiuron August, 1988
-15-
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.
Hosier, 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 Elanco 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
iri 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.* 1976b. 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.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
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.
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Tebuthiuron August, 1988
-16-
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, P.O. Gossett and D.M. Norton.*
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. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1988. U.S. Environmental Protection Agency. EPA Method #507
- Determination of nitrogen and phosphorus containing pesticides in
water by GC/NPD, April 15, 1988 draft. Available from U.S. EPA's
Environmental Monitoring and Support Laboratory, Cincinnati, OH.
'Confidential Business Information submitted to the Office of Pesticide
Programs
-------
August/ 1988
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.
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 lowdose 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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Terbacil August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 5902-51-2
Structural Formula
5-Chloro-3-(1,1-dimethylethyl)-6-methyl-2,4(lHr3H)-pyrimidinedione or
3-tert -Butyl-5-chloro-6-methyluracil
Synonyms
0 Sinbar; Geonter (Meister, 1988).
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, 1988).
Properties (Meister, 1988)
Chemical Formula CgH 1302^01
Molecular Height 216.65
Physical State (at 25°C) White crystals
Boiling Point (at 25 mm Hg)
Melting Point 175-177°C
Density 1.34 g/mL (25°C)
Vapor Pressure (29. 5«C) 4.8 x 10-7 mm Hg
Specific Gravity 1.34 (25/25«C)
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, 1988). No information was found in available
literature on the occurrence of terbacil.
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Terbacil August, 1988
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Environmental Fate
o 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., 1970). ^c^rerbacil 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, !4C-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 label 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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Terbacil August, 1988
-4-
0 Phytotoxic 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. However, evidence of systemic toxicity indicates that
terbacil is absorbed following ingestion.
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 determine a lethal dose of terbacil in dogs
because repeated ernesis 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.
0 In rats (details not available), the LD50 was between 5,000 and 7,500
mgAg (Sherman, 1965). At 2,250 mg/kg, inactivity, weight loss and
incoordination were noted.
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Terbacil August/ 1988
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Dermal/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
mg/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 acetone:dioxane 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 abraded dorsal skin and the others receiving intradermal injections.
After 2 weeks/ 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 ragAg/day.
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. 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, 1988
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750 to 1,125 lag/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
mgAg/day. Female mice from the 187-tng/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.
Wazeter et al. (1967b) administered terbacil (80% a.i.) in the diet
to Charles River CD 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.
Due to an outbreak of respiratory congestion observed in all study
groups at week 27, all animals were placed on antibiotic treatment
(tetracycline hydrochloride), and the therapy was successful. 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
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
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Terbacil August/ 1 988
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(SGOT)/ serum glutamate pyrurate transaminase (SGPT) and cholesterol.
No adverse compoundrrelated 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 LOAEL of 6.25 mg/kg (250 ppm) and a NOAEL of 1.25 mg/kg/day
(50 ppm).
Reproductive Effects
0 wazeter et al. (1967a) administered terbacil (80% a.i.) in the diet
to male and female Charles River CD 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 beha-
vior, 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 or 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 LOAEL of 12.5 mg/kg/day
and a NOAEL of 2.5 mg/kg/day.
Developmental Effects
0 E.I. OuPont (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
of Of 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
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Terbacil August, 1988
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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 LOAEL of 600 mg/kg/day and a NOAEL of 200 mg/kg/day.
0 Culik et al. (1980) administered terbacil (96.6% a.i.) in the feed to
female Charles River CD 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 implanta-
tions 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 (lowest dose tested) 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
DMA synthesis in primary cultures of rat hepatocytes (0.01 and 1.0 uM),
did not exhibit 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,
since genetic assays did not indicate the induction of chromosomal
breakage or loss.
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Terbacil August, 1988
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Carcinogenicity
• 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. No increased incidence of cancer in the treated animals
was found.
0 Wazeter et al. (1967b) administered terbacil (80% a.i.) in the diet
to Charles River 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). No evidence of compound-related carcinogenic
effects was noted.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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)
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 terbacil. It is, therefore,
recommended that the Ten-day HA value for a 10-kg child, 0.25 mg/L (300 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, 1988
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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 terbacil.
It identifies a LOAEL of 12.5 rag/L, based on a reduced body weight gain in the
males in all three generations, and a NOAEL of 2.5 mg/kg/day. 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 = (2.5 mg/kg/day) (10 kg) = 0.25 mg/L (300 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 EPA or
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 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 (300 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 EPA
or 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/ 1988
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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 (900 ug/L)
(100) (2 L/day)
where:
2.5 mg/kg/day = NOAEL, based on absence of reduced body weight gain
in male rats.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
2 L/day = assumed daily water consumption of an adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three-step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study/ divided
by an uncertainty factor(s). 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. 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 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, 1988
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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) = 0.013 mg/kg/day (rounded from
(100) 0.0125 mgAg/day)
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 EPA
or 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 mg/kg/day) (70 kg) = 0.44 mg/L (40o ug/L)
(2 L/day)
where:
0.0125 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 = (0.44 mg/L) (20%) = 0.09 mg/L (90 ug/L)
where:
0.44 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 terbacil.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), terbacil may be classified in
Group E: evidence of noncarcinogenicity for 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. Studies by Goldenthal et al. (1981) and Wazeter et al.
(1967b) reported no induction of any carcinogenic effect in mice or
rats, respectively, administered terbacil in the diet for 2 years.
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Terbacil August/ 1988
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VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 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 (GC) method applicable
to the determination of certain organonitrogen pesticides in water
samples (U.S. EPA, 1987). 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 the
measurement is made with a thermionic bead detector. This method
has been validated in a single laboratory, and the estimated detection
limit for terbacil is 4.5 ug/L.
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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Terbacil August, 1988
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IX. REFERENCES
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apple orchards. Bulletin No. 863. Washington State University, College
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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, movement, and
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soils. In; Land disposal of hazardous wastes. U.S. Environmental
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Doughty, C.C. 1978. Terbacil phytotoxicity and quackgrass (Agropyron repens)
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Gardiner, J.A.* Undated a. Examination of 14c-terbacil-treated soil for the
possible presence of 5-chlorouracil. Unpublished study submitted by
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Gardiner, J.A.* 1964. Laboratory exposure of 2-14c-terbacil to moisture,
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Gardiner, J.A., R.C. Rhodes, J.B. Adams, Jr. and E.J. Soboezenski. 1969.
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Isom, W.H., H.P. Ford, M.P. Lavalleye and L.S. Jordan.* 1969. Persistence
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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.
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several herbicides in a soil cropped to sugarcane. J. Agric. Univ.
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Liu, L.C., H. Cibes-Viade and F.K.S. Kbo. 1971. Adsorption of atrazine and
terbacil by soils. J. Agric. Univ. Puerto Rico 55(4):451-460.
Mansell, R.S., O.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
movement 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. 1988. 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(54):S54. Abstract.
Paynter, O.F.* 1966. Final report. Acute oral toxicity study in dogs.
Haskell Laboratory for Toxicology and Industrial Medicines, Newark, OE.
Unpublished Study. MRID 00012206.
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Terbacil August, 1988
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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. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental -Protection Agency (data file search conducted in May, 1988).
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. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Fed. Reg. 51(185):33992-34003.
September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Draft document.
Method 507. Determination of nitrogen and phosphorus-containing pesti-
cides in water by gas chromatography with a nitrogen-phosphorus detector.
Available from U.S. Environmental Protection Agency, Environmental
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the albino rat. IRDC No. 125-100. International Research and Development
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•Confidential Business Information submitted to the Office of Pesticide
Programs•
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August, 1988
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.
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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P - S - CHi - S - C - CH
Terbufos August, 1988
-2-
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 13071-79-9
Structural Formula
CH3CH2O
CH3CH2O
S-[[(1,1-Dimethylethyl)thio]raethyl]0,0-diethyl phosphorodithioate
Synonyms
0 Counter; Contraven (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 CgH2i(>2PS3
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 1.1
Water Solubility (258C) 15 ing/L
Log Octanol/Water Partition 595
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Technical 87 to 97% pure
Occurrence
o Terbufos has been found in 134 of 2,016 surface water samples
analyzed and in 0 of 283 ground water samples (STORET, 1988). The
85th percentile of all nonzero samples was .10 ug/L in surface water
with a maximum concentration found was 2.25 ug/L. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
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Terbufos August, 1988
-3-
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
Environmental Fate
0 14C Terbufos hydrolyzes rapidly in buffer solutions (Miller and
Jenney, 1973). At a concentration of 4.6 ppm, the hydrolosis half-
lives were 4.5, 5.5 and 8.5 days at pH 5, 7, and 9, respectively.
Principal hydrolysis products were identified as formaldehyde
(accounting for 50-70% of the applied radioactivity), t-butyl
mercaptan, and 0,0-diethylphosphorodithioic acid.
o 14c Terbufos photodegrades rapidly on silica gel glass surfaces
(Miller and Jenney, 1973). Approximately 30% of the radioactivity
was recovered as terbufos within 8 days.
0 Terbufos degrades with a half-life of about 10 days in a silt loam
soil under aerobic conditions (USEPA, 1977). Terbufos residues
dissipate fairly rapidly in the field. In Illinois, residues from
application of terbufos (Counter 15-G) at 1 Ib ai/A dissipated
to 0.31 ppm within 40 days; residues were non-detectable (<0.05 ppm)
by day 135 (Steller et al., 1973a). In Colorado, residues from
a similar application of terbufos declined from 5.9 ppm to 1.56 ppm
100 days later; residues of 0.15 ppm were present at 309 days
after application
0 14C Terbufos residues are immobile in four different soil types:
sand and sandy loam from New Jersey, Wisconsin silt loam and a
North Dakota silt clay (Hui, 1973). Terbufos residues were slightly
more mobile when sandy loam soil was aged for 30 days before leaching
it with 22.5 inches of water - 18% of the applied radioactivity was
detected in the 3.5 to 7.0 inch layer. Formaldehyde was the only
degradate identified (Steller et al., 1973b)
III. PHARMACOKINETICS
Absorption
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 (phosphorylated metabolites), resulting from a single
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Terbufos August, 1988
-4-
oral dose of technical 1^C-terbufos (0.8 mg/kg) 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 Wistar rats at a dose level
of 0.8 mg/kg by gavage. Of all the radioactivity recovered in 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.
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
flagger, 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 significant absorption by any of the exposed workers.
For example, all urinary alkyl phosphate analyses were negative
(detection level, 0.1 ppm). Plasma and red blood cell cholinesterase
values of the exposed workers showed no significant (95% confidence
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Terbufos August, 1988
-5-
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 LDgQ 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 (25 female/25 male; 10/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:
respiratory depression, piloerection, clonic convulsions, exophthalmus,
ptosis, lacrimation, hemorrhage and decreased motor activity.
0 Consultox Laboratories (1975) reported that the acute oral LDg0 value of
technical-grade (86%) terbufos in male Wistar rats was 1.5 mg/kg.
Terbufos was administered by gavage in doses of 0.50 to 2.5 mg/kg to a
total of 50 male rats (10/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, disorientation, chromodocryorrhea,
piloerection and body tremors.
0 American Cyanamid (1972a) reported acute oral LDgg 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 or weight.
0 American Cyanamid (1972b) reported additional acute oral LD50 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 (only dose tested) technical terbufos for 28 days.
Red blood cell ChE was not inhibited at the dose tested.
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Terbufos August, 1988
-6-
0 Tegeris Labs (1987) fed dogs (6/sex/dose) technical terbufos
(purity 89.6%) in a corn oil vehicle (capsule) at doses of 0, 1.25,
2.5, 5.0, and 15 ug/kg/day for 28 days in order to define a plasma
cholinesterase NOAEL, which was not previously defined in a one
year study conducted by American Cyanamid (1986). Based upon the
evaluation of plasma, RBC and brain chloinesterase (ChE) activities
at 1, 2, and 4 weeks during a 28 day dosing regime, a plasma
ChE activity LOAEL (ranging from 10-20% depression) was determined
to be 2.5 ug/kg/day in male and female dogs. A plasma ChE NOAEL
was identified at 1.25 ug/kg/day.
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 rag/kg to the shaved, intact or
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
eosinophils 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 mg/kg
terbufos to their shaved backs. Dermal contact with terbufos was
maintained for 24 hours. The dermal LD5Q value was 1.0 mg/kg. An
acute dermal test with 96.7% terbufos resulted in an LD50 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.
Long-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
(20/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, 0.046 or
0.09 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, and
erythrocyte ChE. Brain ChE was analyzed at the study's termination.
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 statisti-
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Terbufos August, 1988
-7-
cally 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 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. No other histological lesions
were found to be compound-related. The systemic NOAEL based on
absence of liver effects was determined to be 0.02 mg/kg in this
study.
0 Biodynamics Inc. (1987) administered terbufos (purity 89.6%) to
Charles River CD rats (30/sex/dose) in a corn oil vehicle at
doses of 0, 0.125, 0.5 and 1.0 ppm (estimated doses of 0.006,
0.025, and 0.05 mg/kg/dayf based on Lehman (1959)) for one
year. The systemic NOAEL was identified at 0.05 mg/kg/day, the
highest dose tested. The cholinesterase NOAEL was defined at
0.025 mg/kg/day and the ChE LOAEL was observed at 0.05 mg/kg/day
based on plasma and brain cholinesterase decreases (10-30%).
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%.
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 (sixty/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). At
the high dose (0.1 to 0.4 mg/kg/day), there was a noticeable decrease
in mean body weight and mean food consumption. Mortality rates were
24 and 27% (males and females, respectively) at the high dose, 19%
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Terbufos August, 1988
—8—
(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/kg/day).
0 McConnell (1983) conducted a follow-up pathology study of the results
obtained from Rapp et al. 1974. Effects reported were gastric ulcera-
tion and/or erosion of glandular and nonglandular stomach mucosa in
high-dose rats. No similar effect was seen in low and 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.
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 kg, 5 to 6 months old) at doses of 0, 0.015, 0.120, 0.240 and
0.480 mg/kg/day for 1 year. The high doses were eventually reduced
to 0.090 and 0.060 mg/kg/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/kg/day
were found dead. At the two highest doses (0.240 and 0.480 mg/kg/day),
decreased body weights and food consumption were observed. Mean '
erythrocytic parameters of high-dose males and females were signifi-
cantly reduced at 3 months but not at 6 months or at termination of
the study. Plasma ChE activity was significantly inhibited to 55% of
controls at 0.015 mg/kg/day. Slight inhibition of erythrocyte ChE
activity occurred at 0.120 mg/kg/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/kg/day in males and 0.090 mg/kg/day in females based on lack
of erthrocyte ChE inhibition.
0 American Cyanamid (1986) administered terbufos (purity 89.6%)
in corn oil vehicle (capsule) to beagle dogs (6/sex/dose) at doses
of 0, 0.015, 0.06, 0.09 and 0.12 mg/kg/day for one year. The
major effect of terbufos was upon cholinesterase activity. There
was a substantial depression in male and female dog plasma
cholinesterase activity (30-60%) in all treatment groups compared
against controls at week 13 through 52. RBC ChE activity in all
dogs was moderately inhibited (20%) in both mid and high doses.
Brain ChE inhibition was more variable but depression in ChE
activity was apparent at mid and high doses. There was no evidence
of compound-related effect upon mean organ weights or organ-body or
brain weight ratios nor upon histopathology of non-neoplastic
lesions. A plasma ChE NOAEL could not be established. The LOAEL for
this study was identified as 0.015 mg/kg/day, the lowest dose tested.
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Terbufos August, 1988
-9-
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 females/dose, weighing 185.6 g) for a period of 6 months at levels
of 0, 0.25 and 1 ppm. 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 (FQ) was dosed for
60 days. A reproductive LOAEL of 0.05 rag/kg was determined based on
an increased percentage of litters with dead offspring in each of
3 generations as compared to controls. A NOAEL of 0.0125 mg/kg was
identified.
Developmental Effects
0 MacKenzie (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/kg/day on days 7 to 19 of gestation. Repro-
ductive 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 mg/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. No doses led to teratogenic effects. Slightly decreased
maternal 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 toxic at dose
levels of 0.10 and 0.20 mg/kg/day. A maternal NOAEL of 0.05 mg/kg/day,
the lowest dose tested, was identified.
-------
Terbufos August, 1988
-10-
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/raL (ppm) terbufos (purity not specified)
did not cause any significant increase in the frequencies of chromosomal
aberrations. Only a concentration of 100 nL/mL proved to be cytotoxic.
0 Allen et al. (1983) conducted mutagenicity tests with terbufos
(87.8% a.i.) in the presence of S-9 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 S-9 activation at doses of 50, 42, 33, 25, 10 and 5 ug/ml.
Terbufos proved to be cytotoxic at 75 to 100 ug/mL with activation
and at 50 to 70 mg/raL without activation. There were no increases
in the frequency of chromosomal aberrations. The authors concluded
that terbufos reflected a negative mutagenic potential.
0 Godek et 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 Rapp et al. (1974a) administered technical terbufos in the diet to
groups of mice (75/sex/dose) at levels of 0, 0.5, 2.0 and 8.0 ppm
for 18 months. These doses correspond to 0.075, 0.3 and 1.2 mg/kg/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
attributable to terbufos were observed in these four organs.
0 Rapp et al. (1974b) administered technical terbufos in. the diet to
groups of Long-Evans rats (60/sex/dose) at levels of 0, 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). There were no
indications of tumorigenic effects at any dose tested.
0 McConnell (1983) conducted a follow-up pathology evaluation of the
results obtained from Rapp et al (1974b) and concluded that the
compound had no effect on tumorigenesis.
-------
Terbufos August, 1988
-11-
o Tegeris Labs, Inc. (1986) administered terbufos (purity 89.6%)
in the diet to groups of Charles River CD-I mice (65/sex/dose)
at levels of 0, 3, 6, and 12 ppm for 18 months. These doses
correspond to 0.45, 0.90, and 1.8 mg/kg/day based on Lehman (1959).
At these dose levels there was no indication of oncogenic effect.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for one-day, ten-day,
longer-term (up to 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, 1,000 or 10,000),
in accordance with EPA or 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 from day 6-15
of gestation showed no clinical signs of toxicity in the adult animals and no
reproductive or teratogenic 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. In addition, the Rodwell (1985) study is of a most appropriate duration
(10 days) for deriving a 10-day HA.
Using a NOAEL of 0.05 mg/kg/day, the Ten-day HA for a 10-kg child is
calculated as follows:
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Terbufos August, 1988
-12-
Ten-day HA = (0.05 mg/kg/day) (10 kg) = 0>005 /L (5 /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 = uncertainty factor, chosen in accordance with EPA
or 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 28-day feeding study in beagle dogs by Tegeris Labs (1987) has
been selected to serve as the basis for the Longer-term HA value
for terbufos. In this study, dogs wer« administered terbufos in a
corn oil vehicle (capsule) at doses of 1.25, 2.5, 5.0 and 15 ug/kg/day.
A plasma cholinsterase NOAEL, which was not previously defined in a
one year dog study conducted by American Cyanamid (1986), was defined at
1.25 ug/kg/day in the Tegeris Lab (1987) study. The plasma ChE LOAEL
was determined to be 2.5 ug/kg/day based on decreased plasma ChE
activity in male and female dogs. Other studies of longer duration were
not selected for a number of reasons. First, the Tegeris Labs (1987)
study was designed to define the plasma ChE NOAEL which was not achieved
in the one year study conducted by American Cyanamid (1986). It is a
follow-up study employing the same parameters as the American Cyanamid
(1986) study, only the doses are lower. The results of the two studies,
evaluated together, adequately define the chronic endpoint of ChE activity
depression. In other studies, the ChE NOAEL is either an order of
magnitude larger than the NOAEL defined by Tegeris Labs (1987) or only
a LOAEL has been defined. For example, Daly et al. (1979) identified
a NOAEL an order of magnitude higher as is also the case in Shellen-
berger (1986). Rapp et al. (1974b) and American Cyanamid (1986)
were rejected since these studies identified LOAELs only. Morgareidge
et al. (1973) was rejected because there was a high degree of
uncertainty associated with the actual dose consumed by the test
animals.
Using a NOAEL of 0.0013 mg/kg/day; the Longer-term HA is calculated
as follows:
Longer-term HA = (0.0013 mg/kg/day) (10 kg) = 0.0013 mg/L (1.0 ug/L)
(10) (1 L/day)
-------
Terbufos August, 1988
-13-
where:
0.0013 mg/kg/day = NOAEL, based on absence of inhibition of plasma
cholinesterase in dogs exposed to terbufos 28 days.
10 kg = assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study. An uncertainty factor of 10 is
chosen since the endpoint in question is cholin-
esterase inhibition which was not accompanied
by any adverse systemic effects.
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 = (0.0013 mg/kg/day) (70 kg) = 0.0045 mg/L/day (5 ug/L)
(10) (2 L/day)
where:
0.0013 mg/kg/day = NOAEL, based on absence of cholinesterase inhibition
in dogs exposed to terbufos for 28 days.
70 kg = assumed body weight of an adult.
10 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study. An uncertainty factor of 10 is chosen
since the endpoint in question is cholinesterase
inhibition which was not accompanied by any adverse
systemic effects.
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
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
-------
Terbufos August, 1988
-14-
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. 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 28-day feeding study in beagle dogs by Tegeris Labs (1987) has
been selected to serve as the basis for the Longer-term HA value
for terbufos. In this study dogs were administered terbufos in a
corn oil vehicle (capsule) at doses of 1.25, 2.5, 5.0 and 15 ug/kg/day.
A plasma cholinsterase NOAEL, which was not previously defined in a
one year dog study conducted by American Cyanamid (1986), was defined at
1.25 ug/kg/day in the Tegeris Lab (1987) study. The plasma ChE LOAEL
was determined to be 2.5 ug/kg/day based on decreased plasma ChE
activity in male and female dogs. Other studies of longer duration were
not selected for a number of reasons. First, the Tegeris Labs (1987)
study was designed to define the plasma ChE NOAEL which was not achieved
in the one year study conducted by American Cyanamid (1986). It is a
follow-up study employing the same parameters as the American Cyanamid
(1986) study, only the doses are lower. The results of the two studies,
evaluated together, adequately define the chronic endpoint of ChE activity
depression. In other studies, the ChE NOAEL is either an order of
magnitude larger than the NOAEL defined by Tegeris Labs (1987) or only
a LOAEL has been defined. For example, Daly et al. (1979) identified
a NOAEL an order of magnitude higher as is also the case in Shelien-
be rger (1986). Rapp et al. (1974b) and American Cyanamid (1986)
were rejected since these studies identified LOAELs only. Morgareidge
et al. (1973) was rejected because there was a high degree of
uncertainty associated with the actual dose consumed by the test
animals.
Using this study, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.0013 mg/kg/day) „ 0.00013 mg/kg/day
(10)
where:
0.0013 mg/kg/day = NOAEL, based on absence of inhibition of cholin-
esterase in dogs exposed to terbufos for 28 days.
10 = uncertainty factor, chosen in accordance with EPA
or NAS/ODW guidelines for use with a NOAEL from an
animal study. An uncertainty factor of 10 was chosen
since the endpoint in question is cholinesterase
inhibition which was not accompanied by any adverse
systemic effects.
-------
Terbufos August, 1988
-15-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.00013 mg/kg/day) (70 kg) = 0.0045 mg/L/day (5.0 ug/L)
(2 L/day)
where:
0.00013 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 = (0.0045 mg/L) (20%) = 0.0009 mg/L (0.9 ug/L)
where:
0.0045 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
carcinogenic 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 D: not classified. This category is for substances with
inadequate human and animal evidence of carcinogenicity. Although
two carcinogenicity studies are available for at least two animals,
these studies have been considered flawed by the Agency.
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 terbufos is by a gas chromatographic (GC) method applicable
to the determination of certain nitrogen-phosphorus containing pesti-
cides in water samples (U.S. EPA, 1988). In this method, approximately
1 liter of sample is extracted with methylene chloride. The extract
-------
Terbufos August, 1988
-16-
is concentrated and the compounds are separated using capillary
collumn QC. Measurement is made using a nitrogen-phosphorus detector.
This method has been validated in a single laboratory and estimated
detection limits for analytes such as terbufos are estimated at
0.5 ug/1.
VIII. TREATMENT TECHNOLOGIES
0 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 1 2 a tin.
-------
Terbufos August, 1988
-17-
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.
American Cyanamid Company.* 1986. One-year toxicity study in purebread
beagle dogs with AC 92,100. Unpublished data. Accession # 263678-
263680.
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.
Biodynamics Inc.* 1987. A One-year dietary toxicity study with AC 92,100
in rats. Unpublished data. EPA Accession # 400986.
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
worker exposure to terbufos during planting operations of corn. Arch.
Environ. Contarn. Toxicol. 15(1):113-120.
Godek, E., 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.
Hui, T. * 1973. Counter soil insecticide: soil leaching studies of 92,100:
PD-M 10:455-483. Final report. Unpublished study received May 1, 1974
under 4F1496; submitted by American Cyanamid Co., Princeton, N.J.
MRID 87693
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., Q. Bull.
-------
Terbufos August, 1988
-18-
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.
Heister, R.T., ed. 1986. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Miller, P. and K. Jenney.* 1973. CL 92, 100 Counter insecticide: metabolic
studies of 14C-labeled CL 92, 100 in hydrolytic and photolytic
environments: PD-M 10:959-1007. Progress report, Apr. 5, 1973 - Oct. 15,
1973. Unpublished study received May 1, 1974 under 4F1496; submitted
by American Cyanamid Co., Princeton, N.J. MRID 87694
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 41139.
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.
Parke, 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 worker study with
aerial application of Counter 15-G. Report No. C-2370. Unpublished study.
MRID 137760.
Rapp, R., A. Tebaldi, N. Wilson et al.* 1974a. An 18 month carcinogenicity
study of AC 92,100 in mice: Project No.7lR-728. Unpublished study
by Biodynamic, Inc. MRID 85170.
Rapp, W., N. Wilson, M. Mannion et al.* 1974b. 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.
Rodwe11, 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.
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Terbufos August, 1988
-19-
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.
Steller, W., A. Schopbach, p. Ogg et al.* 1973a. Counter 15-G: total residues
of Counter (CL 92,100) and its metabolites in soil: Report No. C-480.
Unpublished study received May 1, 1974 under 4F1496; submitted by
American Cyanamid Co., Princeton, N.J. MRID 87706.
Steller, W., P. Ogg, and H. Van Scoik.* 1973b. Counter 15-G: residues of
Counter (CL92,100) and its individual toxic metabolites in soil:
Report No. C-385. Unpublished study received May 1, 1974 under
4F1496; submitted by American Cyanamid Co., Princeton, N.J. MRID 87708.
STORET. 1988. STORET Water Quality File. Office of Water. U.S. Environ-
mental Protection Agency (data file search conducted in May, 1988).
Tegeris Laboratories, Inc.* 1986. Chronic dietary toxicity and oncogenicity
study with AC 92,100 (Terbufos) in mice. Unpublished study. EPA
Accession # 400986.
Tegeris Laboratories, Inc.* 1987. 28-day oral toxicity study in the dog
with AC 92,100. Unpublished study. EPA Accession # 4037401-4037402.
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. 1977. The Degradation of Selected Pesticides in Soil: A Review
of the Published Literature. Municipal Environmental Research
Laboratory, Cincinnati, Ohio. EPA 600/9-77-022.
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.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 507 -
Determination of nitrogen and phosphorous containing pesticides in water
by GC/NPD, April 15, 1988 draft. Available from U.S. EPA's Environmental
Monitoring and Support Laboratory, Cincinnati, OH.
Windholz, M., S. Budvari, R.F. Blumetti and E.S. Otterbein. 1983. The Merck
Index, 10th ed. Rahway, NJ: Merck and Company.
Confidential Business Information.
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August, 1988
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 accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
-------
2,4,5-Trichlorophenoxyacetic Acid
II. GENERAL INFORMATION AND PROPERTIES
August/ 1988
-2-
CAS No. 93-76-5
Structural Formula
-CH,COOH
2,4,5-trichlorophenoxyacetic acid
Synonyms
9 2,4,5-T; Dacamine; Ded-Weed; Fencerider; Forron; Inverton 245; Linerider;
T-Nox; U-46 (Meister, 1988).
Uses
9 The use of 2,4,5-T in the United States has been cancelled since 1985.
Some or all applications may be classified by the U.S. EPA as Restricted
Use Pesticides (Meister, 1988).
Properties (BCPC, 1983; Meister, 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
C8H503Cl3
255-49
Crystals
1539C
6.46 x 10-6 mm Hg
Solubility of acid is 150 g/L; amine
salts are soluble at 189 g/L (208C);
esters are insoluble
3.00 (calculated)
Occurrence
9 2,4,5-T has been found in 4,021 of 21,616 surface water samples
analyzed and in 99 of 2,905 ground water samples (STORET, 1988).
Samples were collected at 3,838 surface water locations and 1,853
-------
2,4,5-Trichlorophenoxyacetic Acid August/ 1988
-3-
ground water locations/ and 2,4,5-T was found in 44 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. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime/ as they are here.
Therefore/ this information can only be used to form an impression of
the intensity and location of sampling for a particular chemical.
Environmental Fate
0 Salts dissociate in aqueous media/ and esters are rapidly hydrolized
to acids by enzymatic action in animals/ plants/ and soil (Loos/ 1975).
0 The phenoxy ether link is stable in plants and animals/ but is cleaved
by bacterial action in soil and in the rumen of cattle and sheep (Leng/
1977). The rate of cleavage is inhibited by substitution of a third
chlorine at the meta position on the ring and is sterically hindered
by the angular methyl group in the 2-propionic acid side-chain.
0 The first step in degradation of chlorophenoxy acids and chlorophenols
occurs by hydroxylation at the 4-position with shift of the 4-chloro
group to the 3- or 5-position (NIH shift). This reaction also is
inhibited by the presence of a m-chloro substituent; thus/ residues
of 2/4,5-trichlorophenol appeared in the milk of cows and in the
liver and kidney of cattle maintained on feed containing 100 or 300 ppm
2/4,5-T, respectively (Leng, 1977).
0 The rate of microbial degradation of chlorophenoxyalkanoic acids in
the environment depends on a number of factors, including sunlight,
temperature, moisture and organic matter in soil/ and whether the
organisms were adapted by repeated treatment (Loos/ 1975; NRCC, 1978;
Bovey and Young/ 1980). Under field conditions favorable for degra-
dation, 2,4-D disappears within about 2 to 3 weeks, whereas other
phenoxy herbicides disappear more slowly (Loos, 1975). The average
half-lives in a laboratory study were 4, 10, 17 and 20 days for 2,4-D,
dichloroprop, fenoprop, and 2,4,5-T, respectively/ when added as
dimethylamine salts at about 5 ppm in three soils containing partially
decomposed litter obtained from two forest sites and one grassland
site in Oklahoma (Altom and Stritzke/ 1973).
0 Photodegradation is rapid when phenoxy herbicides are exposed to
sunlight in water containing dissolved organic matter, and the
resultant chlorophenols are photolyzed even more rapidly than the
parent acids (NRCC, 1978; Bovey and Young, 1980). The dioxin impurity
TCDD is also degraded rapidly by sunlight, particularly in the presence
of its herbicide carrier on the surface of leaves or in the presence
of other hydrogen donors dissolved in water (NRCC, 1978).
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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Plants treated with phenoxy herbicides rapidly hydrolyze esters and
salts to the parent acids, which form ether-soluble conjugates with
amino acids or water-soluble conjugates with sugars, depending on the
relative susceptibility or resistance of the plants to the herbicides
(Loos, 1975).
III. PHARMACOKINETICS •
Absorption
0 In a study by Gehring et al. (1973), single oral doses of 5 rag/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; 65% of the absorbed dose remained 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
daily dose ingested in mg/kg is AQ, 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
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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(0.17, 4.3 or 41 mg/kg) by gavage, and then declined rapidly. Radio-
activity 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 which received 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 mg/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.
Excretion
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 14C 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
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
-6-
was excreted in the urine and 8.2% was excreted in the feces. No
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
clearance 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 radioactivity eliminated
from the body of the rats. Fecal excretion was 3% at 5 mg/kg, and
increased to 14% at 200 mgAg- 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. By 9 days post-treatment, 11% of the dose
had been recovered in urine and 20% had been 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 (MAS, 1977).
Preparations of 2,4,5-T formerly contained TCDD at levels of 1 to 80
ppm, 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 MAS (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
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
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).
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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 Northland area. The epidemiological
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
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.
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2,4,5-Trichlorophenoxyacetic Acid August/ 1988
-8-
Numerous epidemiclogical 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 LD50
values were 500 mg/kg for rats, 389 mgAg 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 LDsg
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
serum ZgG 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 Gehring 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 (HAS, 1977).
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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 rag/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 mgAg/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 and kidney/body
weight ratios and kidney weights. Females given 30 mg/kg/day had
slightly 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 treatment-related effects were observed at
the 3 or 10 mg/kg dose level. From this study, a NOAEL of 10 mg/kg/day
and a LOAEL of 30 mg/kg/day were identified.
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
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2,4,SVTrichlorophenoxyacetic Acid August, 1988
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deposits in the renal pelvis in females. Effects noted at the 10 rag/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 rag/kg),
there were no changes that were considered to be related to treatment
throughout the 2-year period. From this study in rats, a NOAEL of
3 mgAg/day was identified.
Reproductive Effects
0 Male and female Sprague-Dawley rats (Fg) were fed lab chow containing
2,4,5-T «0.03 ppb TCDO) to provide dose levels of 0, 3, 10 or 30
mgAg/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 (F-|) and the rest were necropsied. Subsequent matings were
conducted to produce F2, F3a and F3b 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 F^ 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
F.J, F2 and F3 litters of the 30 mg/kg group. A significant decrease
(p <0.05) in relative thymus weight was seen only in the F3b generation
of the 30 mg/kg group, but the relative liver weights of weanlings
was significantly (p <0.05) increased in the F2, F3a and F3b litters
of this dosage group. Smith et al. (1981) concluded that dose levels
of 2,4,5-T that were sufficiently high to cause signs of toxicity in
neonates had no effect on the reproductive capacity of the rats,
except for a tendency toward a reduction of postnatal survival at a
dose of 30 mg/kg. Reproduction was not impaired at the lowest dose
of 3 mg/kg. The apparent NOAEL with respect to reproductive capacity
and fetotoxic effects in this study is 3 mg/kg/day, taking into
consideration the report, on apparently the same study, by Smith et
al. (1978) who noted a significant (p <0.05) decrease in F-j (10 and
30 mg/kg on days 14 and 21) and F3 (3 mg/kg on day 14, and 10 and 30
mg/kg on day 21) litters and concluded that there was no effect of
2,4,5-T on rat reproduction except for a tendency toward a reduction
in neonatal survival at 10 and 30 mg/kg.
Developmental Effects
0 Sparschu et al. (1971) tested 2,4,5-T (commercial grade, 0.5 ppm TCOO)
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 mg/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
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
-11-
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 LD^)* 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 rabbits 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.
0 Several different samples of 2,4,5-T (containing <0.5 ppm TCDD) were
tested in pregnant Wistar rats by daily oral administration on days
6 through 15 of gestation at dose levels between 2 5 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 mg/kg, there was an increase (p <0.05) in fetal mortality and an
increase (p <0.05) in skeletal variations; 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.
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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This effect was not duplicated in a repeat test with the same sample.
At the 25 and 50 ragAg 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 2,4,5-7 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 mgAg/day for the hamster was identified.
0 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 (latency, ambulation,
rearing, grooming, and defecation in open field testing) abnormalities
in offspring 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/JAX 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 mgAg/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
LOAEL for teratogenic effects is 15 mg/kg/day for the A/JAX strain
and 30 mg/kg/day for the other strains.
o 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
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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 mgAg. 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.
9 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 DON 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
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 with 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.
0 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.
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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).
• Shirasu et al. (1976) found that 2,4,5^1 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, DMA-repair-defective
£. 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 10 weeks of age (20 to 30 g) with a single
intraperitoneal injection of 100 mg/kg of 2,4,5-T «1 ppra 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.
Carcinogenicity
0 Innes 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 mgAg/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.
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 in
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 XVII/G strain between the
treated and control mice. The authors felt that although 2,4,5-T
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
-15-
demonstrated carcinogenic potential in the C3Hf strain, 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 madei
Groups of Sprague-Dawley rats (50 males and 50 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) = 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 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.
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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; Khera 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 = (8 mg/kg/day) (10 kg) = O.g mg/L (800 ug/L)
(100) (1 L/day)
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 = 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 reproduction study by Smith et al. (1978, 1981) has been selected
to serve as the basis for the Longer-term HA value for 2,4,5-T because the
reduction in neonatal survival over multiple generations is concluded to be
relevant to the Longer-term HA for a 10-kg child. The NOAEL identified was
3 mgAg/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 Kbciba (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) = Q.3 mg/L (300 ug/L)
(100) (1 L/day)
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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 = 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) = ^05 mg/L (, 000 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 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
(DUEL) 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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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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 Kociba 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
2,4,5-7 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 toxiclty in the rats, this study identified a
NOAEL of 3 mgAg/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/kg/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) = 0.01 mg/kg/day
(100) (3)
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= uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
3 = 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 Hater Equivalent Level (DWEL)
DWEL = (0-01 mg/kg/day) (70 kg) = 0.35 /L (400 /L)
(2 L/day)
where:
0.01 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 = (0.35 mg/L) (20%) = 0.07 mg/L (70 ug/L)
where:
0.35 mg/L = DWEL.
20% = assumed relative source contribution from water.
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2,4,5-Trichlorophenoxyacetic Acid August/ 1988
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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-TCOO
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 rag/kg/day, based on the
results of a rat chronic oral NOAEL of 3 mgAg/day with an uncertainty
factor of 300 (used because of a data gap).
0 The National Academy of Sciences (NAS, 1977) has calculated an ADI
of 0.1 mg/kg/day, using a NOAEL of 10 mg/kg/day (identified in a
90-day feeding study in dogs) and an uncertainty factor of 100. A
chronic Suggested-No-Adverse-Effeet-Level (SNARL) of 0.7 mg/L was
calculated based on the ADI 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-TWA) 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).
VII. ANALYTICAL METHODS
0 Determination of 2,4,5-T is by a liquid-liquid extraction gas
chromatographic procedure (U.S. EPA, 1978; Standard Methods, 1985).
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2,4,5-Trichlorophenoxyacetic Acid August/ 1988
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Specifically, the procedure involves the extraction of chlorophenoxy
acids and their esters from an acidified water sample with ethyl
ether. The esters are hydrolyzed to acids and extraneous organic
material is removed by a solvent wash. The acids are converted to
methyl esters which are extracted from the aqueous phase. Separation
and identification of the esters is made by gas chromatography.
Detection and measurement is accomplished by an electron capture,
mcirocoulometric 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 chromatograph with an
electron capture detector results in an approximate detection limit
of 10 ng/L for 2,4,5-T.
VIII. 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.
0 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).
0 Treatment technologies for the removal of 2,4,5yr 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, 1988
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Anderson, K.J., E.G. Leighty and M.T. Takahashi. 1972. Evaluation of
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BCPC. 1983. British Crop Protection Council. The pesticide manual. A
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Courtney, K.D. and J.A. Moore. 1971. Teratology studies with 2,4,5-tri-
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Crampton, M.A. and L.J. Rogers. 1983. Low doses of 2,4,5-trichlorophenoxy-
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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.
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Innes, J. R.M., B.M. Ulland, M.G. Valerio, L. Pe truce Hi, L. Fishbein, E.R. Hart,
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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) : 2 91 -2 99.
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. , O.J. Keyes, R.W. Lisowe, R.P. Kalnins, 0.0. Dittenber, C.E. Wade,
S.J. Gorzinski, 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 and Drug Officials of the United States.
Leng, M.L. 1977. Comparative metabolism of phenoxy herbicides in animals.
In: G.W. Ivie, H.W. Dorough, eds. , Fate of Pesticides in Large Animals.
Academic Press: New York, pp. 53-76.
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Loos, M.A. 1975. Phenoxyalkanoic Acids. In: P.C. Kearney, D.D. Kaufman,
eds., Herbicides Chemistry, Degradation and Mode of Action, 2nd ed.,
vol. 1. Marcel Dekker: New York, pp. 1-128.
Magnusson, 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.
Meister, R., ed. 1988. Farm chemicals handbook. Willoughby, OH: Meister
Publishing Company.
Muranyi-Kovacs, I., G. Rudali and J. Imbert. 1976. Bioassay of 2,4,5-tri-
chlorophenoxyacetic acid for carcinogenicity in mice. Br. J. Cancer.
33:626-633.
NAS. 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.
NRCC. 1978. National Research Council of Canada. Phenoxy herbicides -
Their effects on environmental quality. NRC No. 16075. NRCC CNRC
Publications: Ottawa, Canada, 440 pp.
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):47-50.
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.
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2,4,5-Trichlorophenoxyacetic Acid August, 1988
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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.
Riihiraalci, 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
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Rebeck, G.G., K.A. Dostal, 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
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Smith, F.A., B.A. Schwetz, F.J. Murray, A.A. Crawford, J.A. John, R.J. Kociba
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February, 1989
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 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.
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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Trifluralin February, 1989
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II. GENERAL INFORMATION AND PROPERTIES
CAS No. 1582-09-8
Structural Formula
N(CH2CH2CH,)2
CF3
alpha, alpha, alpha-Trifluoro-2,6-dinitro-N,N-dipropyl-p-toluidine
Synonyms
0 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 (Meister, 1983; U.S. EPA, 1985b)
Chemical Formula
Molecular Weight 335.2
Physical State (25°C) Orange, crystalline solid
Boiling Point 139 to 140°C
Melting Point 46 to 49°C
Density
Vapor Pressure (25°C) 1.1 x 10~4 mm Hg
Specific Gravity
Water Solubility (25°C) >1 mg/L
Log Octanol/Water Partition 4.69
Coefficient
Taste Threshold
Odor Threshold
Conversion Factor
Occurrence
0 Trifluralin does not have a strong 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 February, 1989
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0 Trifluralin has been found in 172 of 2,047 surface water samples
analyzed and in 1 of 507 ground water samples (STORET, 1988). Samples
were collected at 249 surface water locations and 386 ground water
locations, and trifluralin was found in 7 states. The 85th percentile
of all nonzero samples was 0.54 ug/L in surface water. This information
is provided to give a general impression of the occurrence of this
chemical in ground and surface waters as reported in the STORET
database. The individual data points retrieved were used as they
came from STORET and have not been confirmed as to their validity.
STORET data is often not valid when individual numbers are used out
of the context of the entire sampling regime, as they are here.
Therefore, this information can only be used to form an impression
of the intensity and location of sampling for a particular chemical.
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 (Maliani, 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 nitrogen gas). Autoclaving
and flooding the soil decreased the degradation rate of the compound
(Parr and Smith, 1973).
o 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 14f, 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).
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Trifluralin February, 1989
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0 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 ppm in fields treated with trifluralin at various
rates for 1, 2, 3 or 4 consecutive years (Parka and Tepe, 1969).
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 ppn. 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 (GI) 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 enterohepatic circulation.
Distribution
0 No information was found in the available literature on the distri-
bution of trifluralin.
Metabolism
0 Pour metabolites of trifluralin were identified in rats. Twelve rats
were given 100 mg/kg ^CF -trifluralin in corn O1i 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 iri 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).
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Trifluralin February, 1989
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Excretion
Rats given an oral dose (100 mg/kg) of ^CF^-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 (Emnerson and
Anderson, 1966).
IV. HEALTH EFFECTS
Humans
Short-term Exposure
0 The Pesticide Incident Monitoring System database indicated 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, 198la).
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, as cited in U.S. EPA, 1985a) 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
LDso values have been reported: mice >5 g/kg; rats >10 g/kg; dogs,
rabbits and chickens >2 g/kg (Meister, 1983; RTECS, 1985).
0 An inhalation LCso value of a liquid formulation containing 41%
trifluralin (species not stated) of >2.44 mg/L/hour was reported
(U.S. EPA, 1985c). No other information was available.
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Trifluralin February, 1989
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Dermal/Ocular Effects
0 The results of a primary dermal-irritation study in the rabbit were
negative 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 con-
tact sensitization developed during this time (Negelski et al., 1984).
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 (BLANCO, 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 41.2% liquid to rabbit eyes, 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-derived 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 contin-
uously 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
(Worth et al., 1977)
0 In a 90-day study, male F344 rats were fed trifluralin at 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 dietary levels of 0, 50, 200, 800, 3,200 and 6,400 ppm trifluralin,
respectively (Emmerson et al., 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 treatment groups receiving 10 mg/kg/d or greater. Other effects
included increased aspartate transaminase, urinary calcium, inorganic
phosphorus and magnesium at levels >160 mg/kg/day. A NOAEL of 2.5
mg/kg/day and a Lowest-Obsevred-Adverse-Effect Level (LOAEL) of
10 mg/kg/d can be identified from this study.
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Trifluralin February, 1989
-7-
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 (Worth et al., 1966c).
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 oE 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 noted 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 (Emmerson and Pierce,
1980). This study appears to identify a NOAEL of 0.08% trifluralin
(30 to 37 mg/kg/day).
0 B6C3Fi 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 (Emmerson and Owen, 1980). No effects were noted at the
low-dose level (40 mg/kg/day).
0 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
0 In a four-generation reproduction study (Worth et al., 1966b), rats
were given 0, 200 or 2,000 ppm trifluralin in the diet (0, 10 or
100 mg/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
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Trifluralin February, 1989
-8-
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 (Worth et al., 1966b). Dogs (two/sex/dose at
400 ppn and three/sex/dose at 1,000 ppm) 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, 1984). No adverse
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
identified as 225 mg/kg/day.
0 Rabbits (number not specified) exposed to 225 or 500 mg/kg/day
trifluralin during pregnancy exhibited anorexia and cachexia and
aborted litters (U.S. EPA, 1985c). Fetotoxicity as evidenced by
decreased fetal weight and size was observed at the high-dose level.
These effects were not observed at the 100 mg/kg/day level.
0 In a rabbit teratology study, a total of 32 mated females were given
225, 450 or 1,000 mg/kg/day trifluralin by gavage (Wbrth et al.,
1966a). 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 (ELANOO, 1984).
The 500 and 800 mg/kg/day levels resulted in decreased live births,
cardiomegaly and wavy ribs in the progeny. No effects on progeny were
observed at 225 mg/kg/day or less (ELANCO, 1984).
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 (ELANOO, 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
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Trifluralin February, 1989
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negative when tested at levels of 50 to 800 ing/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 B6C3Fi mice 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 3,744 ppm (males) or 2,740 or 5,192 ppm
(females) for 78 weeks and observed ror an additional 13 weeks after
exposure. A significant dose-related increase in hepatocellular
carcinomas 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 carcinomas 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 tag/kg/day (Lehman,
1959), these doses correspond to approximately 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, B6C3Fj mice were given 0, 563, 2,250 or
4,500 ppm trifluralin. These doses correspond to 40, 180 or 420
mg/kg/day (Lehman, 1959) in the diet (Emmerson and Owen, 1980).
Levels of a nitrosamine contaminant of trifluralin, NDPA, were below
the 0.01-ppn analytical detection limit. A total of 40 animals/sex/
treatment group was used and 60/sex for controls. 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 glomerulo-
nephritis 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; (corresponding to 30, 128 or 272 mg/kg/day
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Trifluralin February, 1989
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for males and 37, 154 or 336 mg/kg/day for females) in the diet for
2 years (Bmmerson and Pierce, 1980). 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 (up to 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)
(UP) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100, 1,000 or 10,000),
in accordance with EPA or 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 trifluralin. Therefore, it is
recommended that a modified DWEL (0.03 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 DWEL is calculated as follows:
DWEL = (0.0025 mg/kg/day) (10 kg) = 0.025 mg/L <3
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.
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Trifluralin February, 1989
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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 DWEL (0.03 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 (0.03 mg/L) for a 10-kg child be used as a
conservative estimate for a Longer-term exposure for a child and the DWEL
(0.1 mg/L; calculated below) be used for the 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. 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 Emmerson et al. (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 animals receiving 10 mg/kg/day or greater. A NOAEL
of 2.5 mg/kg/day was identified. Other longer-term studies report NOAELs at higher
doses.
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.0025 mg/kg/day (0.003 mg/kg/day)
(1,000)
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Trifluralin February, 1989
-12-
where:
2.5 mg/kg/day = NOAEL, based on absence of increased urinary globulins
in rats consuming a trifluralin diet for 3 months.
1,000 = uncertainty factor, chosen in accordance with EPA
or 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 = (0.003 mg/kg/day) (70 kg) = 0.105 mg/L (100 ugA)
(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 = (0.105 mg/L) (20%) = Q.0021 mg/L (2 ug/L)
where:
0.105 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 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.
e In an NCI (1978) study of female B6C3Fi mice, significant dose-related
increases in hepatocellular carcinomas and alveolar adenomas were
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 of kidney, urinary bladder
and thyroid tumors (Emmerson and Pierce, 1980).
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Trifluralin February, 1989
-13-
0 The evidence from the Bnmerson and Pierce (1980) and NCI (1978) studies
indicates that trifluralin has carcinogenic potential. Based on the
results of the Emmerson and Pierce (1980)' study, the U.S. EPA Carcinogen
Assessment Group (CAG) has prepared a quantitative risk estimate of
trifluralin exposure (U.S. EPA, 1981b). The CAG estimated a potency
factor (qj*) 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 per 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 ranging from 0.05 to 2.0 ppm trifluralin have been
established for a variety of agricultural commodities (U.S. EPA,
1985).
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
0 Determination of trifluralin is by Method 508 (U.S. EPA, 1987). In
this procedure, a measured volume of sample of approximately 1 liter
is solvent-extracted with methylene chloride by mechanical shaking in
a separatory funnel or mechanical tumbling in a bottle. The methylene
chloride extract is isolated, dried and concentrated to a volume of
5 mL after solvent substitution with methyl tert-butyl ether (MTBE).
Chromatographic conditions are described which permit the separation
and measurement of the analytes in the extract by GC with an electron
capture detector (ECD). An alternative manual liquid-liquid extraction
method using separatory funnels is also described. This method has
been validated in a single laboratory, and the estimated detection
limit determined for the analytes in this method, including trifluralin,
is 0.025 ug/L-
VIII. TREATMENT TECHNOLOGIES
0 Available data indicate that reverse osmosis (RO), granular-activated
carbon (GAC) 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 GAC. 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).
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Trifluralin February, 1989
-14-
Sanders and Seibert (1983) determined experimentally water solubility,
vapor pressure, Henry's Lav Constant and volatilization rates for
trifluralihr 100% of the compound volatilized under laboratory
conditions.
Treatment technologies for the removal of trifluralin from water are
available and have been reported to be effective. However, selection
of individual technology or combination 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 February, 1989
-15-
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