EPA 0553
Health advisories for
lOCs and SOCs
RX000027511
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EPA 0553
RX000027511
April, 1992
ALDRIN
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology, and treatment technology that
would be useful in dealing with the contamination of drinking water. Health Advisories
(HAs) 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. For substances with a carcinogenic potential, chemical
concentration values are correlated with carcinogenic risk estimates by employing a cancer
potency (unit risk) value together with assumptions for lifelong exposure and the ingestion of
water. The cancer unit risk is usually derived from a linearized multistage model with 95%
upper confidence limits providing a low-dose estimate of cancer risk. The cancer risk is
characterized as being an upper limit estimate, that is, the true risk to humans, while not
identifiable, is not likely to exceed the upper limit estimate and in fact may be lower. While
alternative risk modeling approaches may be presented, for example One-hit, Weibull, Logit,
or Probit, the range of risks described by using any of these models 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
carcinogenic risk, the Agency recommends using the linearized multistage model.
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Aldrin
April, 1
II. GENERAL INFORMATION'AND PROPERTIES
CAS No. 309-00-2
Structural Formula
H
Synonyms
Compound 118; ENT 15,949; exo-dimethanonaphthalene; HHDN; hexachloro-
hexahydroendo, l,2,3,4,10,10-hexachloro-l,4,4a,5,8,8a-hexahydro-l,4,5,8-
dimethanonaphthalene; octalene (Sax, 1975; Sax and Lewis, 1987; Sittig, 1981).
Uses
. Aldrin is an organochlorine insecticide that is highly effective against many soil-dwelling
pests. Aldrin is principally used in agriculture at concentrations ranging from 0.5 to 5,
kg/hectare. Other uses include control of termite and ant infestation. Manufacture
and use of aldrin have been discontinued in the United States since 1974 (Clayton and"
Clayton, 1981; Windholz et al., 1983).
Properties (Sittig, 1981; Windholz et al., 1983; Worthing and Walker, 1983; U.S. EPA, 1987).
Chemical Formula
".v/olecular Weight
Physical State
Boiling Point
Meltirg Point
Density (20ฐQ
Vapor Pressure; (20ฐC)
Specific Gravity (20ฐC)
Water Solubility (27ฐC)
Log Octanol/Water
Partition Coefficient
Taste Threshold
.Odor Threshold
Conversion Factor
C12H8C16
364.93
Technical aldrina tan to dark-brown
crystalline solid with a mild chemical odor;
nonflammable; stable between pH 4
: and 8. Pure aldrin - a white powder.
145ฐC at 2 mm Hg
Pure: 104ฐC; technical: 49 to 60ฐC
1.54 g/raL
7.5 x 10'J mm Hg
1.70
0.027 m
3.01
0.017 mg/L water
1 ppm = 14.96 mg/mj at 25ฐC and 760 mm Hg
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Aldrin April, 1992
Occurrence
Aldrin and dieldrin residues are more likely to be found in soil than in water. The
highest dieldrin concentration found in sediment collected from bays and rivers of
North Carolina was 18 ppb, and a study of groundwater in South Carolina showed
aldrin levels of 0.007 ppb (U.S. Army Medical Research and Developmental
Laboratory, 1975).
The present occurrence of aldrin in the environment is not known, although it is
expected to be much lower now than in the past because the manufacture and use
of aldrin were discontinued in 1974.
Environmental Fate
In the soil, aldrin is easily oxidized to dieldrin (Elgar, 1975). The investigator
reported the results of single applications of aldrin to various soils from 12
locations. The proportion of dieldrin in the total soil residue ranged from 60% 1
year after application to 90% after 3 years.
Aldrin is firmly bound to soil, is extremely resistant to leaching and does not
migrate from the point of application (Shell Chemical Company, 1986).
Aldrin and dieldrin have low solubility in water and are strongly sorbed to soil,
particularly to organic matter. Eye (1968) concluded from a study on leachability
that dieldrin residues are not significantly transported-through soil into subsurface
waters. These results were confirmed in studies by ThompJaAi et al. (1970).
Aldrin (technical grade, purity unspecified), applied to subirri-gated soil at 4 pounds
active ingredient per acre (Ib a.i./A), was found to be immobile in a Hager silty clay
loam and in a Lakeland sandy loam soil (Harris, 1969).
, - ~\ *
In a soil-column (52 mm diameter by 60 cm length) leaching study, Foschi et al.
(1970) found that aldrin (purity not specified) did not leach into the 20- to 40-cm
section (or below) after 15 days of soaking ."vith.'approximately 35.4 inches (89.8 cm)
of water in a greenhouse at 26ฐC. Residues of aldrin were detected by gas
chromatography at 3.68 ppm in the top 5-cm section and at 3.07 ppm in the next
15-crn segment. :
. " i')' '
In a laboratory study using labeled material, 2 to; 26% of the applied aldrin was
found to volatilize from various substrates (Lichtenstein'and^Schulz, 1970).
Approximately 19 to 26% of the aldrin volatilized from tapwater, a soil-water
mixture, a pH 7 buffer solution, or glass beads, and 2 to 3% of the pesticide
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Aldrin April,
volatilized from silt loam soil or silt loam soil plus a detergent (0.1% linear
alkylbenzenesulfonate). In another test, fruit flies died when exposed to aldrin
vapors above an aldrin-water mixture containing 25 /j.g aldrin. Approximately 6 /xg
aldrin volatilized in a 24-hour period.
Aldrin residues in samples of rotational crops ranged from 0 to 2.45 ppm; the
highest levels were found in potatoes. The relative concentration of dieldrin in the
recovered residues increased with time; only dieldrin was detected 5 years post-
treatment. Other aldrin degradates were not assayed. Aldrin, at 25 Ib a.i./A,
dissipated in plots of silt loam soil planted to crops each year, declining from 9.77
to 0.05 ppm in 5 years after soil incorporation to a depth of 4 to 5 inches.
Approximately 11% of the total applied dose was present after 5 years; the
percentage of dieldrin found in recovered residues was 97%. Aldrin residues in a
subplot planted to alfalfa as a cover crop dissipated more slowly, declining from
4.04 ppm in 1959 to 0.96 ppra in 1963. During the test period, the-average annual
rainfall was 30.16 inches (76.61 cm), the average annual temperature was 45.3ฐF
(7.4ฐC) and the average temperature for the period from May to October was
62.4ฐF (16.9ฐC) (Lichtenstein and Schulz, 1965).
III. PHARMACOKINETICS
Absorption
Quantitative data on the gastrointestinal (GI) absorption of aldrin by laboratory
animals are not available. However, the results of metabolism studies indicate that
aldrin is slowly absorbed via the portal venous system following oral administration
(U.S. EPA, 1987). In a time-course distribution/metabolism study by Farb et al.
(1973), aldrin was detected in the stomach and small intestine (levels not specified)
of neonatal rats up to 6 days after administration of single oral doses of 10 mg/kg.
In the same group of animals, the concentration of aldrin in the liver increased to
about 13% of the administered dose by the third day.
Ludwig et al. (1964) estimated that approximately 10% of the radio label
administered to two male rats (strain not reported) as wC-labeled aldrin dissolved
in corn oil (43 /ig/rat/day for 90 days) was absorbed by the GI tract.
Results of inhalation studies on human volunteers suggest that up to 50% of
inhaled aldrin vapor (exposure levels not specified) may be retained (Beyermann
and Eckrich, 1973).
_.... Feldmann and Maibach (1974) reported that approximately 8% of the dermally
applied aldrin (in acetone vehicle) was absorbed in humans.
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Aldrin April, 199Z
Distribution
Aldrin is rapidly metabolized to dieldrin. Bioaccumulation of this metabolite in
adipose tissue has been reported (U.S. EPA, 1987). As discussed below, oral
administration of dieldrin at concentrations as high as 160 ppm in rats (Deichmann
et al., 1975) and 70 ppm in dogs (Deichmann et al., 1969), were found to
accumulate in fat. No significant differences were noted between the amount of
insecticide found in lipids and the amount found in the total carcass or abdominal
fat.
Farb et al. (1973) detected aldrin in the kidneys of neonatal rats for up to 72 hours
after administration of single oral doses of 10 mg/kg. In the liver, aldrin
concentrations increased for the first 6 hours following dosing to a maximum of
13% of the administered dose. After 72 hours, aldrin concentration declined to less
than 0.1%. The only metabolite detected in the liver was dieldrin, which was
identified as early as 2 hours post-treatment. A maximum of 31% of the
administered dose was detected as dieldrin in liver tissue 24 hours following dosing.
Ludwig et al. (1964) reported that the ratio of dieldrin to aldrin was 15:1 in the
carcass and 18:1 in the abdominal fat of two male rats given daily oral doses of 4.3
Mg uC-aldrin for 3 months. Approximately 3.60, 1.77 and 1.83% of the total
administered dose of "C was found in the carcass, fat and other tissues,
respectively.
Increasing amounts of dieldrin were noted in the abdominal fat and total
extractable lipids of mice from the F,, F2 and F} generations fed diets containing 5
or 10 pprn aldrin (purity 95%) (doses equivalent to 0.75 and 1.5 mg/kg, respectively,
based on Lehman, 1959). Dieldrin residues in the parents of each generation were
measured after about 260 days on the test diets; dieldrin concentrations in total
lipids were between 115 and 121 ppm in males and between 149 and 159 in females
(Deichmann et al., 1975).
In a study in which six male beagles were given 0.6 mg/kg of aldrin in corn oil daily
for 10 months, dieldrin concentrations in the fat and liver increased during months
1 through 10, reaching levels of 70 and 20 ppm, respectively (Deichmann et al.,
1969). Dieldrin concentrations gradually decreased, and 12 months after aldrin
administration was halted, dieldrin concentrations had dropped to 25;and 6 ppm in
the fat and liver, respectively.
Metabolism
Following absorption, aldrin is readily oxidized to dieldrin in mammals (U.S. EPA.
1987).
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Aldrin April,
Formation of the epoxide (dieldrin) is a major step in the biotransformation of
aldrin. This transformation is probably mediated by mixed-function mono-
oxygenases (aldrin epoxidases), which are found in many organisms including plants
(Mehendale et al., 1972), insects (Krieger and Wilkinson, 1969), fish (Bums, 1976)
and various mammals (Levi and Hodgson, 1985; Davies and Keysell, 1983; Wolff et
al., 1980; Brooks and Harrison, 1966) and humans (McMannus et al., 1984).
Soto and Deichmann (1967) reported that approximately 30% of the aldrin
administered intravenously to dogs was converted to dieldrin within 24 hours of
dosing.
Klein et al. (1968) identified one of the polar metabolites of aldrin in rat urine as
l,l,2,3a,7a-pentachloro-5,6-epoxydecahydro-2,4,7-metheno-3H-cyclopenta[a]-
pentalen-3-one (pentachloroketone).
Excretion
In a study in which male rats were given daily oral doses of 4.3 /ig l4C-aldrin for 3
months, the elimination of radioactivity, expressed as percent of dose administered
weekly per rat, increased from 31% on the second day of dosing to about 100%
during weeks 10 to 12 (Ludwig et al., 1964). Most of the radioactivity was excreted
in the feces. During weeks 10 through 12, fecal excretion accounted for 93 to 94%
of the weekly dose. Urinary excretion was approximately 9%. Biliary excretion was
not measured. The authors reported that after 8 weeks, the amount of radioactivity
being excreted was equivalent to the amount administered, suggesting that an
equilibrium between the intake and storage of aldrin had been established. After
dosing was discontinued, excretion of radioactivity decreased rapidly. Analysis of
urine and feces indicated that predominantly hydrophilic metabolites were excreted.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Acute intoxication following aldrin exposure in humans is characterized by a brief
period of excitation or drowsiness, followed by convulsions, muscle twitching and
coma. Hypothermia generally accompanies death. The majority of individuals
intoxicated with aldrin, however, usually regain consciousness and recover (Hayes,
1982; Jager, 1970).
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Aldrin ' April, 1992.
Jager (1970) reported the acute oral lethal dose of aldrin in an adult male to be 5.0
g (approximately 70 mg/kg, assuming a body weight of 70 kg).
The ingestion of approximately 120 mg (8:2 mg/kg) of aldrin by a 3-year-old female
resulted in collapse and convulsions within 5 minutes and death within 12 hours
(Hayes, 1982).
A 23-year-old male experienced convulsions about 20 minutes after ingesting 25.6
mg/kg aldrin (Spiotta, 1951). After treatment with pentobarbital, his convulsions
ceased. However, he exhibited restlessness, hypothermia, tachycardia and
hypertension for up to 5 days and electroencephalogram (EEG) abnormalities for
up to 6 months after ingestion of aldrin.
Long-term Exposure
*
One male, employed 21 years at a chemical plant and reassigned to the handling of
aldrin concentrate (period and levels of exposure not specified), experienced
involuntary jerking (rapid flexor movement) of his hands and forearms, vomiting
and chronic irritability and insomnia (Hodge et al., 1967). His EEG showed alpha-
wave irregularities, with discharges of slow and sharp waves. After exposure to
aldrin was discontinued, his condition rapidly improved.
Dieldrin (mean 13 ng/g whole milk) was found in the breast milk of women whose
homes were treated annually (or more frequently) with organochlorine pesticides
(Stacey and Tatum, 1985). A correlation between dieldrin levels in the milk and
aldrin treatment of homes was observed. Dieldrin levels in breast milk rose until
the seventh or eighth month after treatment of homes was discontinued. No data
were provided on the health effects of children exposed to dieldrin-contaminated
breast milk.
Animals
Short-term Exposure
The oral median lethal dose (LDj,,) values for aldrin in laboratory animals are as
follows: mice, 44 mg/kg (purity not reported) (Borgmann et al., 1952); rats, 39 to 60
mg/kg (purity not reported) (Gaines, 1969);. guinea pigs, 33 mg/kg (purity not
reported) (Borgmann et al., 1952); female rabbits, 50 to 80 mg/kg (purity, 95%)
(Treon and Cleveland, 1955); and dogs, 65 to 95 rag/kg (purity not reported)
(Borgmann et al., 1952).
Acute toxicity in animals is characterized by increased irritability, salivation,
hyperexcitability, tremors followed by clonic/tonic convulsions, anorexia and loss of
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Aldrin April, 19|
body weight, depression, prostration and death (Borgmann et al., 1952; Hodge et
al., 1967).
Decreased body weight gain and increased mortality were observed in male and
female Osbome-Mendel rats fed 320 ppm (16 mg/kg/day, based on a food
consumption factor of 0.05 from Lehman, 1959) of aldrin (technical grade; 95%
pure) in the diet for 42 days and observed for an additional 14 days (NCI, 1978).
Groups of five rats of each sex were fed diets containing 0, 40, 80, 160 or 320 ppm
aldrin (0, 2, 4, 8 or 16 mg/kg/day, respectively). The No-Observed-Adverse-Effect
Level (NOAEL) for this study was 160 ppm (8 mg/kg/day). This study was a range-
finding study for a long-term carcinogenicity study; therefore, a complete toxicology
profile was not obtained (e.g., biochemical and hematology measurements).
In groups of B6C3Fl mice (five/sex/group) fed aldrin (technical grade; 95% pure) at
concentrations of 0, 2.5, 5, 10, 20, 40 or 80 ppm (0, 0375, 0.75, 1.5? 3, 6 or 12
mg/kg/day, respectively, based on a food consumption factor of 0.15 from Lehman,
1959) in the diet for 42 days, 100% mortality was observed .in the 40- and 80-ppm
(6- and 12-mg/kg/day) groups. One male and one female died in the 20-ppm (3-
mg/kg/day) group; 10 and 20 ppra (1.5 and 3 rag/kg/day) were therefore considered
the NOAEL and Lowest-Observed-Adverse-Effect Level (LOAEL) for this study
(NCI, 1978). This study was a range-finding study for a long-term carcinogenicity
study; therefore, a complete toxicology profile was not obtained.
In a subchronic toxicity study (see the discussion in "Dermal/Ocular Effects" and
"Long-term Exposure" below), Treon and Cleveland (1955) observed 100%
mortality within 2 weeks in groups of male and female Carworth rats (total number
and number/sex not reported) fed 300 ppm aldrin (purity 95%) (15 mg/kg/day,
based on Lehman, 1959). No mortality was noted at lower doses. Administration
of a diet containing 25 ppm aldrin (purity 95%) (0.625 mg/kg/day) to two male and
three female beagle dogs induced fatalities after periods of feeding ranging from 9
to 15 days.
Dermal/Ocular Effects
Treon and Cleveland (1955) reported dermal LD,,, values of 600 to 1,250 mg/kg for
powdered aldrin (purity 95%) in rabbits. The investigators also demonstrated that
the application of powdered aldrin is less toxic to rabbits than aldrin in a vehicle.
The ranges of minimum lethal dosages obtained for aldrin in the powdered form, in
vegetable oil or in Ultraseneฎ (solvent) were 3.5 to 123, 10 to 26, and < 4.8
mg/kg/day, respectively. The rabbits were exposed for 2 hours/day, 5 days/week for
10 weeks.
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Aldrin April, 199Z
Long-term Exposure
Treon and Cleveland (1955) fed male and female Carworth rats (number/dose level
not reported) 2.5, 5, 25, 75 or 300 ppm aldrin (purity 95%) (0.125, 0.25, 1.25, 3.75
or 15 mg/kg/day, respectively, based on Lehman, 1959) in the diet for 6 months;
100% mortality was observed at 15 mg/kg/day within 2 weeks. No differences in
mortality rate were noted among other test groups and the controls. The authors
did not report data on other effects. A NOAEL for the subchronic effects of aldrin
cannot be determined from these data.
Fitzhugh et al. (1964) fed groups of 12 male and 12 female Osborne-Mendel rats
aldrin (purity 99%) in the diet at levels of 0.5, 2, 10, 50, 100 or 150 ppm (0.025, 0.1,
0.5, 2.5, 5.0 and 7.5 mg/kg/day, respectively, based on Lehman, 1959) for 2 years. A
dose-related increase in mortality was observed at a dietary level of 50 ppm or
greater. In addition, significant (p <. 0.05) dose-related increases in relative liver
weights were observed. Histopathological changes observed in the livers of aldrin-
treated rats were chiefly the characteristic "chlorinated insecticide" lesions that
occur only in rodents. These lesions consist of enlarged centrilobular hepatic cells,
with increased cytoplasmic oxyphilia, and peripheral migration of basophilic
granules. The incidence and severity of this nonneoplastic histologic change
increased with increasing dietary level. In rats ingesting amounts of aldrin at 50
ppm or higher, distended and hernorrhagic urinary bladders, enlarged livers and
increased incidences of nephritis were observed. The apparent LOAEL for this
study was 0.5 ppm (0.025 mg/kg/day). A NOAEL was not established.
Deichmann et al. (1970) fed groups of 50 male and 50 female Osbome-Mendel rats
20, 30 or 50 ppm (1, 1.5 or 25 mg/kg/day, respectively, based on Lehman, 1959)
aldrin (technical grade, 95% pure) for 31 months. Groups of 100 rats of each sex
served as controls. Survival and body weight gains were comparable between the
treated and the control groups. Treated animals exhibited tremors and clonic
convulsions. Liver-to-body weight ratios were increased in males fed 30 or 50 ppm.
A moderate increase (not dose-related) in the incidence of hepatic centrilobular
cloudy swelling and necrosis was observed in all aldrin-treated male and female rats,
but not in controls. The LOAEL of 20 ppm (1 mg/kg/day) was established for this
study. A NOAEL was not determined.
Aldrin (technical grade, 95% pure) was administered in the diet at 4 or 8 ppm (0.6
or 12 mg/kg/day, respectively, based on Lehman, 1959) to groups of 50 male mice
and at 3 or 6 ppm (0.45 or 0.90 mg/kg/day, respectively, based on Lehman, 1959) to
groups of 50 female mice for 80 weeks. In a trend test, a significant (p = 0.037),
dose-dependent increase in mortality was observed in females; a similar effect was
not observed in males. A NOAEL was not established because of toxicity at 3 ppm
(0.45 mg/kg/day), the lowest dose tested (NCL 1978).
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Aldrin April, 1
Kitselman and Borgmann (1952) fed groups of seven mongrel dogs of both sexes
(number/sex not specified) 0.2, 0.6 or 2 mg/kg/day of aldrin in medicated meatballs
for up to 228 days. The test material was 99% pure. Dogs administered 2
mg/kg/day exhibited marked body weight loss, and all dogs died between days 60
and 90. No treatment-related effects were observed in dogs receiving 0.2 mg/kg/day
for 190 days or in those administered 0.6 mg/kg/day for 228 days. Based on body
weight loss, 0.6 mg/kg/day and 2 mg/kg/day were considered to be the NOAEL and
LOAEL, respectively, for this study.
In a long-term feeding study by Treon and Cleveland (1955), beagles (two/sex/dose)
fed diets containing 1 or 3 ppm (0.043 to 0.091 or 0.12 to 0.25 mg/kg/day,
respectively, as reported by the authors) aldrin (purity 95%) for 15.6 months gained
weight at rates similar to control dogs. However, at 3 ppm, significant (p < 0.05)
increases in absolute and relative liver weights were noted. Histopathologic
changes, such as fatty degeneration of the liver and vacuolation of renal tubular
cells, were observed in both sexes at the 3-ppm level. At the 1-ppm level, females
exhibited vacuolation of the distal renal tubules. The LOAEL for this study was 1
ppm (0.043 to 0.091 mg/kg/day).
Fitzhugh et al. (1964) administered 0.2, 0.5, 1, 2 or 5 mg/kg/day aldrin (purity 99%)
to 12 mongrel dogs (sexes combined), 6 days/week for periods of up to 25 months. I
Each group consisted of one dog/sex except for the 0.5-mg/kg/day group, which had
one male and three female dogs. All dogs receiving 1, 2 or 5 mg/kg/day died within
49 weeks; the first death occurred on day 22 in a female administered 5 mg/kg/day.
Prior to death, the animals exhibited body weight loss, dehydration and convulsions.
Slight to moderate fatty degeneration was noted in hepatic and renal tubular cells,
and reduced numbers of mature erythroid cells were found in the bone marrow. In
animals receiving 0.5 mg/kg/day, clinical signs of toxicity were limited to convulsions
in one male dog during the 24th month. Dogs in the 0.2-mg/kg/day group exhibited
no adverse effects. The NOAEL in this study appears to be 0.2 mg/kg/day based on
the absence of clinical signs of toxicity, body weight loss and histopathological
changes. However, the adequacy of this study for establishing a NOAEL is
questionable owing to the small number of dogs used.
Reproductive Effects
A dietary level of 12.5 ppm (0.625 mg/kg/day, based on Lehman, 1959) fed over
three generations was reported to cause a reduction in the pregnancy rate of
Carworth rats (Treon and Cleveland, 1955). Groups of rats (number not specified)
were fed aldrin (purity 95%) at concentrations of 0, 2.5, 12.5 or 25.0 ppm (0.125,
0.625 or 125 mg/kg/day, based on Lehman, 1959). No reduction in the number of
live pups/litter was evident in dams fed 0.125, 0.625 or 1.25 mg/kg/day in the diet.
10
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Aldrin April, 1992.
However, the viability of the offspring during lactation was decreased at levels of
0.625 and 1.25 mg/kg/day.
No apparent effects on the fertility or pregnancy rates were evident in mongrel dogs
(two females/group) receiving 0.2, 0.6 or 2.0 mg/kg/day of aldrin (purity 99%) in
medicated meatballs for 1 year (Kitselman, 1953). However, the majority of
apparently healthy pups that were delivered at all levels died within 3 days
postpartum.
In a reproduction study reported by Deichmann et al. (1971), groups of beagles
were administered 0.15 (four females) or 0.3 (four males, three females) mg/kg/day
of aldrin (purity 95%) by capsule 5 days/week for 14 months. Two of the four
females administered 0.15 mg/kg/day failed to achieve estrus during the following 8-
month period in which aldrin feeding had been halted. Similar problems were not
evident in dogs given 0.3 mg/kg/day. During lactation, the viability, of pups from
dams receiving 0.15 or 0.3 mg/kg/day was decreased; 84, 75 and 44% of pups from
dams ingesting 0, 0.15 and 0.3 mg/kg/day, respectively, survived until weaning. The
reduced pup survival may have been due to a prenatal effect or to toxicity
associated with dieldrin in the mothers' milk. Mammary development and milk
production also appeared to be depressed. Some males reportedly exhibited a
depressed sexual drive.
Developmental Effects
Administration of 50 mg/kg of aldrin (purity > 99%) via oral intubation to pregnant
Syrian Golden hamsters on day 7, 8 or 9 of gestation resulted in significant (p <
0.05) increases in fetal death, resorption, anomalies (open eye, webbed foot, cleft
palate and/or lip and fused ribs) and growth retardation (Ottolenghi et al., 1974).
In pregnant CD-I mice given 25 mg/kg of aldrin on day 9 of gestation, significant (p
< 0.01) increases in total anomalies, as well as incidences of open eye, webbed foot
and cleft palate, were also observed. No increases in fetal death or decreases in
fetal body weight were noted in mice, however. No information on maternal
toxicity was presented.
Mutagem'city
In bacterial reverse mutation assays conducted by several investigators, aldrin was
not mutagenic to Salmonella typhimurium (Simmon and Kauhanen, 1978; Cotruvo
et al., 1977; Simmon et al., 1977).
Simmon and Kauhanen (1978) reported that aldrin, at concentrations of 10 to 5,000
jig/plate, did not cause gene conversion in Saccharomyces cerevisiae D3 with or
without Aroclor-induced rat liver microsomes.
11
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Aldrin April, 19,
Georgian (1975) reported that aldrin induced chromosome aberrations in human
lymphocytes in vitro and in rat and mouse bone marrow cells in vivo. The evidence
for a clastogenic response was, however, inconclusive because increases in the
frequency of chromosomal aberrations in the in vivo assays occurred only at
cytotoxic levels. Additionally, chromosome and chromatid gaps, which are unreliable
indicators of damage to genetic material, were included as aberrant figures. The
extent of true chromosomal damage cannot, therefore, be ascertained.
Exposure of primary rat hepatocytes to aldrin at concentrations ranging from 0.5 to
1,000 nmol/mL for 5 to 20 hours did not induce unscheduled DNA synthesis (Probst
et al., 1981).
Carcinogenic! tv
NCI (1978) reported increased combined incidences of follicular ceil adenoma and
carcinoma of the thyroid in both male and female Osbome-Mendel rats (50
animals/sex) fed aldrin (technical grade, 95% pure) at concentrations of 30 or 60
ppm (1.5 and 3 mg/kg/day, respectively, based on Lehman, 1959) for 74 to 80 weeks
compared to controls. The incidences were 4/48,14/38 and 8/38 for males and 3/52,
10/39 and 7/46 for females from the pooled control, low-dose and high-dose groups^
respectively. Differences were significant (p = 0.001) in the low-dose group but nci
in the high-dose group. Significant (p = 0.001) increases in the incidence of
cortical adenoma of the adrenal gland were also observed in the low-dose females,
but this was not considered to be compound-related. Aldrin produced no
significant effect on the mortality of rats of either sex. The authors concluded that
the observed tumors were not associated with treatment. However, the U.S. EPA
(1987) concluded that the occurrence of thyroid tumors should be considered
equivocal evidence of carcinogenicity.
In a carcinogenicity bioassay, aldrin (technical grade, 95% pure) was administered
in the diet at 4 or 8 ppm (0.6 or 1.2 mg/kg/day, based on Lehman, 1959) to 50 male
B6C3Ft mice and at 3 or 6 ppm (0.45 or 0.90 mg/kg/day, based on Lehman, 1959)
to 50 female B6C3F! mice, for 80 weeks. The animals were observed for an
additional 10 to 13 weeks. A significant (p ฃ 0.001) dose-related increase in the
incidence of hepatocellular carcinomas was observed in male but not female mice
fed diets containing 4 or 8 ppm (0.6 or 12 mg/kg/day, respectively) when compared
to matched or pooled controls. Incidences were 3/20 17/92, 16/49 and 25/45 for the
matched control, pooled control, low-dose male and high-dose male groups,
respectively (NCI, 1978).
Davis and Fitzhugh (1962) and Davis (1965) fed 10 ppm (approximately 1.5
mg/kg/day, based on Lehman, 1959) of aldrin (purity not reported) in the diet to a
group of 215 C3HeB/Fe mice ("approximately equally divided by sex") for 2 years:
12
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Aldrin
April, 1992.
the control group consisted of 217 mice. It was reported that the incidence of
hepatic cell adenomas was significantly (p < 0.001) higher in treated mice (35 of
215) than in nine controls (9 of 217). Upon revaluation of the liver lesions, most
of the liver tumors were found to be hepatic carcinomas (Epstein, 1975).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are based upon the identification of adverse health effects
associated with the most sensitive and meaningful noncarcinogenic end point of toricity. The
induction of this effect is related to a particular exposure dose over a specified period of
time, most often determined from the results of an experimental animal study. Traditional
risk characterization methodology for threshold toxicants is applied in HA development. The
general formula is as follows:
(NOAEL or LOAEL) x (BW) ..
= mg/L
(UF)
Uday)
where:
NOAEL or LOAEL
BW
UF
_L/day
= No- or Lowest-Observed-Adverse-Effect Level (in rag/kg bw/day).
= assumed body weight of a child (10 kg) or an adult (70 kg).
= uncertainty factor, (10, 100,1,000 or 10,000), in accordance with
EPA or NAS/OW guidelines.
= 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 determining the One-
day HA for aldrin. The modified Drinking Water Equivalent Level (DWEL) for a 10-kg
child of 0.0003 mg/L (03 fig/L, (calculated below) is recommended for use as a conservative
estimate of the one-day HA for aldrin.
13
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Aldrin April, 199!
For a 10-kg child, the adjusted DWEL is calculated as follows:
DWฃL m (0.00003 mg/kg/day) (10 kg) = Q ^ ^ to QJ
(1 L/day)
where:
0.00003 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
ซ
No suitable information was found in the available literature for determining the Ten-
day HA for aldrin. The modified DWEL for a 10-kg child of 0.0003 mg/L (0.3 ng/L) is
recommended for use as a conservative estimate of the ten-day HA for aldrin.
Longer-term Health Advisory
No suitable information was found in the available literature for determining the
Longer-term HA for aldrin. Several subchronic studies were available. However, either
inadequate study design (e.g., a small number of animals exposed or an inadequate length of
exposure) or inadequate reporting resulted in the disqualification of these studies as the basis
for derivation of the Longer-term HA. The modified DWEL for a 10-kg child of 0.0003 mg/L
(0.3 Mg/L) is recommended for use as a conservative estimate of the longer-term HA for
aldrin.
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 estimate (with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is likely to be without
appreciable risk of deleterious health effects during 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
14
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Aldrin April, 1992-
RfD by the assumed body weight of an adult and divided by the assumed daily water
consumption of an adult. The Lifetime HA in drinking water alone 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 known, probable, or possible human
carcinogen, according to the Agency's classification scheme of carcinogenic potential (U.S.
EPA, 1986), then caution must be exercised in making a decision on how to deal .with
possible lifetime exposure to this substance. For human (A) or probable (B) human
carcinogens, a Lifetime HA is not recommended. For possible (C) human carcinogens, an
additional 10-fold safety factor is used in the calculation of the Lifetime HA. The risk
manager must balance this assessment of carcinogenic potential and the quality of the data
against the likelihood of occurrence and significance of health effects related to
noncarcinogenic end points of toxicity. To assist the risk manager in this process, drinking
water concentrations associated with estimated excess lifetime cancer risks over the range of 1
in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of water/day are pfovided in the
Evaluation of Carcinogenic Potential section.
The study of Fitzhugh et al. (1964) in which groups of Osborne-Mendel rats were fed
diets containing aldrin at concentrations of 0.5, 2, 10, 50, 100 or 150 ppm (approximately
equivalent to 0.025, 0.1, 0.5, 2.5, 5 or 7.5 mg/kg/day, based on Lehman, 1959) for 2 years has
been selected to serve as the basis for the Reference Dose because it was the most
appropriate chronic study found that established a LOAEL for aldrin in this species. In this
study, a LOAEL of 0.5 ppm (0.025 mg/kg/day) was identified, based on an increased
incidence of liver lesions characteristic of exposure to organochlorine pesticides and increased
relative liver weights in animals from all aldrin-treated groups. The liver lesions consisted of
enlarged centrilobular hepatic cells, with increased cytoplasmic oxyphilia and peripheral
migration of basophilic granules. A NOAEL was not established Furthermore, the results of
several other long-term feeding studies in rats using aldrin at higher concentrations (2.5 to 60
ppm; equivalent to 0.13 and 3 mg/kg/day, based on Lehman, 1959) support the results of the
Fitzhugh et al. (1964) study (NCI, 1978; Deichmann et al., 1970; Treon and Cleveland, 1955).
Using the Fitzhugh et al. (1964) study, the Lifetime HA is derived as follows:
Step 1: Determination of the RfD
RfD m (0.025 mg/kg/day) = Q QQQ^ mg/kg/day (rounded to 0.00003 mg/kg/day)
(1,000)
where:
0.025 mg/kg/day = LOAEL, based on an increased incidence in liver lesions and
increased relative liver weights in rats fed aldrin for 2 years.
15
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Aldrin April, 1991
1,000 = uncertainty factor, chosen in accordance with EPA or NAS/OW
guidelines in which a LOAEL from an animal study is employed.
Step 2: Determination of the DWEL
DWEL n (0.00003 mg/kg/day) (70 kg) = Qm mg/L ^
(2 L/day)
where:
0.00003 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed water consumption of a 70-kg adult.
Step 3: Determination of the Lifetime HA
Aldrin has been classified in Group B2: probable human carcinogen (U.S. EPA, 1986a);
thus, a Lifetime HA is not recommended. The estimated excess cancer risk associated with
lifetime exposure to drinking water containing aldrin at 0.88 /xg/L is 4.25 x 10"*. 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
Three adequately conducted long-term carcinogenicity bioassays of aldrin have been
conducted with B6C3Ft, C3HeB/Fe and C3H mice. Based on these studies, there is
sufficient evidence of carcinogenicity for aldrin. Dietary administration of aldrin
induced statistically significant (p < 0.001) increases in hepatocellular carcinomas in
male B6C3F, mice (NCI, 1978), hepatomas in male and female C3HeB/Fe mice
(Davis and Fitzhugh, 1962) and hepatomas in both sexes of C3H mice (Davis, 1965,
as cited in Epstein, 1975). Reevaluation of the hepatomas observed in the latter two
studies showed the hepatomas actually to be hepatocellular carcinomas (Epstein,
1975).
Dietary administration of aldrin increased the combined incidences of follicular cell
adenomas and carcinomas of the thyroid in both male and female Osborne-Mendel
rats; however, the increase was not dose-related'and was significant (p = 0.001) only
at the low dose. This increase was not considered to be treatment-related. It was
concluded that aldrin was not carcinogenic to rats (NCI, 1978).
16
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AJdrin ' April, 1992-
Based on the available data, IARC (1987) concluded that there was limited evidence
for the carcinogenicity of aldrin in animals and inadequate evidence in humans.
lARC's conclusion was based on the occurrence of malignant liver neoplasms in
mice, since one study in rats could not clearly associate the occurrence of thyroid
tumors with aldrin treatment, three additional studies in rats gave negative results
and another rat study was judged to be inadequate. Consequently, IARC classified
aldrin as a Group 3 chemical, a possible human carcinogen.
Applying the criteria described in EPA's guidelines for assessment of carcinogenic
risk (U.S. EPA, I986a), aldrin and its metabolite dieldrin may be classified in Group
B2: probable human carcinogen. This category is for agents for which there is
inadequate evidence of carcinogenicity in human studies and sufficient evidence of.
carcinogenicity in animal studies.
Several carcinogenicity studies have provided evidence that aldrin is carcinogenic in
mice. Three data sets from these studies are adequate for quantitative risk
estimation. Utilizing the linearized multistage model, the U.S. EPA (1987)
performed potency estimates for each of these data sets. The geometric mean of the
potency estimates, (qt*) = 17 (mg/kg/day)'1, was estimated as the cancer potency of
aldrin for the general population.
Using this cancer potency estimate and assuming that a 70-kg human adult consumes
2 liters of water a day over a 70-year lifespan, the linearized multistage model can be
used to estimate that concentrations of 0.2, 0.02 and 0.002 fig liter of aldrin may
result in an excess cancer risk of 10"*, 10"5 and 10"*, respectively.
The linearized multistage model is only one method of estimating carcinogenic risk.
From the data contained in the U.S. EPA (1987) report, it was determined that one
of the three data sets were suitable for determining slope estimates for the probit,
logit, Weibull and gamma-multihit models. The cancer risk estimate (at the upper
95% confidence limit) that can cause one excess cancer per 1,000,000 (10"6) is
associated with exposure to aldrin levels in drinking water of 0.00206 /ug/L for the
multistage model, 0.00356 /ig/L for the probit model, 0.00376 /Ag/L for the logit
model, 0.00356 /xg/L for the Weibull model and 0.00310 ngfL for the multihit model.
Each model is based on different assumptions. Based on the current understanding
of the biological mechanisms of carcinogenesis, the relative accuracy of these models
cannot be predicted
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The U.S. EPA (1980) has established a National Ambient Water Quality Criterion
for aldrin of 0.074 ng/L for human health based on a predicted cancer risk of
17
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Aldrin April, 19J
1:1,000,000. This estimate is based on the ingestion of contaminated drinking water
and aquatic organisms. A criterion of 0.079 ng/L is recommended for ingestion of
contaminated aquatic organisms only.
The Occupational Safety and Health Administration (OSHA) has established an 8-
hour Time-Weighted Average (TWA) atmospheric Permissible Exposure Limit
(PEL) of 0.25 mg/mj with a recommendation to reduce skin absorption of aldrin by
using protective measures (OSHA, 1989).
Aldrin is regulated as a hazardous substance, with a reportable quantity of 1 Ib (0.454
kg) under section 102 of the Comprehensive Environmental Response,
Compensation and Liability Act (CERCLA) for release from vessels and facilities
(U.S. EPA, 1986b).
The American Conference of Governmental and Industrial Hygienists (ACGIH)
established a TWA Threshold Limit Value (TLV) of 0.25 mg/mj for exposure to
aldrin (ACGIH, 1986).
The National Institute for Occupational Safety and Health (NIOSH) recommended a
PEL of 0.25 mg/mj for aldrin (NIOSH, 1988).
VII. ANALYTICAL METHODS
Aldrin has been included in numerous EPA methods for analyzing chlorinated
hydrocarbon pesticides. In earlier procedures aldrin was used as an internal
standard, since it had been rarely detected in the environment, degrading fairly
quickly to dieldrin.
Aldrin can be analyzed by EPA Methods 505 (U.S. EPA, 1988a) and 508 (U.S. EPA,
1988b). Both methods use electron capture gas chromatography for analysis, but
Method 505 is a micro-extraction procedure and Method 508 is one 1-L liquid-liquid
extraction. The detection limit for aldrin is 0.075 mg/L.
VIII. TREATMENT TECHNOLOGIES
Granular-activated carbon (GAC) adsorption and reverse osmosis (RO) can reduce
the levels of aldrin in drinking water supplies.
Lafomara (1978) described case histories of the use of a trailer-mounted spills-
treatment process to remove a number of toxic organic compounds from water,
including aldrin. The treatment trailer, which had a capacity of approximately 0.3
18
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Aldrin April, 1994.
million gal/day (1 million L/day), consisted of three mixed-media filters, each 3.5 feet
(107 cm) in diameter and containing a 2-foot-deep (61-cm) bed of powdered
anthracite on top of a 1.5-foot (46-cm) layer of sand plus three GAG columns, each 7
feet (213 cm) in diameter and containing 5,940 Ib (2,690 kg) of GAG. The
contaminated water (containing 8.5 mg/L of aldrin) was passed through one filter and
one GAG column with a contact time of 17 minutes. After 100,000 gal (380,000 L)
were treated, the effluent contained 0.19 mg/L of aldrin.
Van Dyke et al. (1986) evaluated the ability of a home-use water filter to remove a
number of organic chemicals, including aldrin. The filtering system consisted of a
nonwoven prefilter, a pressed carbon block and a porous polyethylene-fritted core.
The water was supplied at a constant pressure of 50 pounds per square inch (3.5
kg/cm2) gauge. Each run consisted of passing a volume of water equal to 150% of
the filter rated life of 500 gal (1,900 L) through the filter, and analyzing for various
contaminants. Aldrin was present in the influent at a concentration of 68 mg/L, and
the filter removed aldrin below its detection limit of 0.1 mg/L.
Abron and Osburn (1973) investigated polyaraide hollow fiber RO membranes for
use in the removal of aldrin from aqueous solutions. This system was operated at 600
psi, a flow rate of 0.48 gph and a flux rate of 0.077 to 0.082 gpd/ft2. Complete
removal of aldrin with the RO membranes was attained at influent concentrations of
both 4 mg/L and 28.1 mg/L. Aldrin is classified as a membrane-interacting solute.
The aldrin removal may not have been achieved by the reverse osmotic pressure, but
by membrane-solute interaction.
Lykins et al. (1986) tested the effect of disinfection, sand filtration and GAG
adsorption on several organic compounds at a pilot plant in Jefferson Parish, LA.
Four disinfectants were used: chlorine, monochloramine, chlorine dioxide and ozone.
Aldrin was present in the non-disinfected influent at an average concentration of 0.53
ng/L. The disinfection system was designed for a contact time of 30 minutes. Sand
filtration did not reduce the concentration of chlorinated hydrocarbon insecticides
(CHI). About 90 to 93% of the CHI was removed by GAC over the 1-year
operational period. No analytical data are presented on aldrin concentration in
treated effluent.
Data were not found for the removal of aldrin from drinking water by aeration.
However, because of its low vapor pressure, aldrin is not likely to be amenable to
removal by aeration.
19
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Aldrin April, l!
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U.S. Army Medical Research and Developmental Laboratory. 1975. Problem definition
studies on potential environmental pollutants. Technical Report 7509.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality criteria for
aldrin/dieldrin. Report PB81-11730/OWRS. Washington, DC: U.S. EPA, Criteria and
Standards Division.
U.S. EPA. 1986a. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Fed. Reg. 51(185): 33992-34003.
U.S. EPA. 1986b. U.S. Environmental Protection Agency. Effluent guidelines and
standards. 40 CFR, Subchapter N. Section 401.15. Part 401. General provisions, pp. 5,
8.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Carcinogenicity assessment of
aldrin and dieldrin. EPA/600/6-87-006. Washington, DC: U.S. EPA, Office of Health
and Environmental Assessment, Carcinogenesis Assessment Group.
24
-------
Aldrin April,
U.S. EPA. 1988a. U.S. Environmental Protection Agency. Method 505-Analysis of
organochlorine pesticides in water by micro-extraction and gas chromatography, revision
2.1. Cincinnati, OH: Environmental Monitoring and Support Laboratory. December.
U.S. EPA. 1988b. U.S. Environmental Protection Agency. Method 508-Determination of
chlorinated pesticides in water by gas chromatography with an electron-capture detector,
revision 2.1. Cincinnati, OH: Environmental Monitoring and Support Laboratory.
December.
Van Dyke, K., R. Kuennen, J. Stiles, J. Wezeman and J. O'Neal. 1986. Test stand design
and testing for a pressed carbon block water filter. Am. Lab. 18(9):118-132.
Windholz M., S. Budavari, R.F. Bluemetti and E.X. Otterbein, ed 1983. The Merck index--
An encyclopedia of chemicals and drugs, 10th ed. Rahway, NJ: Merck and Company,
Inc.
Wolff, T., H. Green, M-T. Huang, G.T. Miwa and Y.H. Lu. 1980. Aldrin epoxidation
catalyzed by purified rat-liver cytochromes P-450 and P-448. Eur. J. Biochera. 11:545-
551.
Worthing, C.R. and S.B. Walker, eds. 1983. The pesticide manual. In: A world
compendium, 7th ed. Lavenhara, Suffolk, Great Britain: The Lavenham Press Limited,
pp. 6, 191.
25
-------
EPA 0553
KX000027511
April 1992
AMMONIA
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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 Health
Advisories 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 lifelong exposure and the ingestion of water. The
cancer unit risk is usually derived from a linearized multistage model with 95% upper
confidence limits providing a low-dose estimate of cancer risk. 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.
II. GENERAL INFORMATION AND PROPERTIES
Ammonia is a ubiquitous naturally occurring inorganic chemical. In aqueous solution, it
exists in a number of forms that include the following: NH,, NH4OH, NHj.HjO and NH4*. For
purposes of this document, ammonia is used as a general term encompassing NHj and NH4*.
CAS No. 7664-41-7
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Ammonia April 1992,
Structural Formula
HNH
I
H
Ammonia
Synonyms
None.
Uses
Ammonia is used in fertilizers, refrigeration systems and manufacturing processes
(Windholz et al.t 1983).
Used in conjunction with chlorine to form chloramine which is used as drinking
water disinfectant.
Properties (Campbell et al., 1958; Verschueren, 1977; Windholz et al., 1983)
Chemical Formula NH3
Molecular Weight 17.03
Physical State (25 ฐC) Liquid
Boiling Point 33.4ฐC
Melting Point -77.7 ฐC
Density (-33 ฐC, 1 atm) 0.6818
Vapor Pressure (20 ฐC) 8.5 atm
Specific Gravity
Water Solubility (20ฐC) 531 g/L
Log Octanol/Water Partition
Coefficient (log K^,)
Taste Threshold 34 mg/L
Odor threshold (air) 0.04 g/ra3
Conversion Factor
Occurrence
Survey of total ammonia (NH, + NH*4) concentrations in surface waters indicate
an average of 0.18 mg/L in most areas, and 0.5 mg/L in waters near large
metropolitan areas (U.S. EPA, 1979). Levels in ground water are usually low,
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Ammonia April 1992
since ammonia is generally immobile in soil (Feth, 1966). Ammonia is effectively
eliminated when drinking water is chlorinated.
Ammonia is a negligible natural constituent of food, but ammonium compounds
are added in small amounts (0.001 to 2%) to various foods as stabilizers,
leavening agents, flavorings or for other purposes (FASEB, 1974).
Environmental Fate
Ammonia is introduced into the environment through sewage effluents, fertilizer
application, agricultural runoff, drainage from feedlots and through industrial
discharge. Once present in an environment, ammonia participates in the
constant fluxing of nitrogen levels and states of the nitrogen cycle. Processes
related to this cycle are nitrogen fixation, nitrification, denitrification and
ammonification. Ammonia is subject to assimilation by chemotrophic and
phototropic organisms, or may be biologically oxidized (nitrification).
Nonbiological means of moving ammonia within an environment include
diffusion, dilution (in waters), volatilization and sorption to particles.
III. PHARMACOKINETICS
Absorption
Ammonia is produced in humans in the stomach, duodenum, ileum, colon and
feces at an estimated 4,200 mg/kg/day, with the colon and fecal content
contributing about 73%. Of the total amount produced, 4,150 mg/kg/day is
absorbed and 50 rag/kg/day is excreted (Summerskill and Wolpert, 1970).
Conn (1972) administered 9 rag NH4Cl/kg as uncoated tablets to 20 normal
human subjects and 50 cirrhotic patients. Blood ammonia concentration peaked
(mean, 140 /ig NHj/100 mL) at 15 minutes and returned to fasting levels (mean,
105 /ig NHj/100 mL) by 30 minutes in normal human subjects. In 50 patients
with cirrhosis of the liver, however, blood ammonia levels increased from
elevated fasting levels (mean, 155 /ig NHj/100 mL) to higher peak concentrations
(mean, 370 /tg NHj/100 raL) at 15 minutes. This was followed by a slow decrease
in ammonia levels, reflecting impaired hepatic urea synthesis.
Castell and Moore (1971) studied ammonia absorption from the human gut by
direct perfusion in four patients in whom surgical colon bypass procedures had
been performed previously for chronic hepatic encephalopathy. Eight solutions
at pH 5 (acetate buffer) and eight solutions at pH 9 (trihydroxyaminomethane
buffer) with varying NH3-N concentrations between 10 and 150 /tg/mL were
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Ammonia April 1992
infused at 15 mL/rain into a proximal colonic stoma for 20 minutes. In
two patients, 12 infusions were also given at varying pH with both phosphate (pH
5, 6.6 and 8.2) and trihydroxyaminomethane buffer (pH 7, 8, 9). Ammonia
absorption in the human colon showed a positive linear (or slightly curvilinear)
relationship with intraluminal concentrations of ammonia up to 150 /ig/mL.
There was greater absorption of ammonia occurred consistently at all
concentrations, as the pH of the infused fluid was changed from 5 to 9.
Distribution
In healthy individuals, ammonia that is absorbed following oral administration is
mainly converted in the liver to urea; therefore, relatively small amounts reach
systemic circulation (Summerskill and Wolpert, 1970). No other details were
provided.
Metabolism
Much of the ammonia absorbed in the gut is transformed to urea in the liver,
while some is incorporated into tissue proteins (Richards et al., 1968, 1975;
Summerskill and Wolpert, 1970). The transformation of ammonia into tissue
protein varies inversely to the amount of protein consumed in the diet (Kies and
Fox, 1978).
Excretion
Excretion of ammonia following oral administration to humans is modified by
protein intake. Richards et al. (1968, 1975) administered "NJ^Cl orally to
healthy male volunteers fed a normal (70 g of protein in 24 hours) or
protein-restricted, diet (20 g of protein in 24 hours). The total dose of 9.2 to
17.2 mg NH4Cl/kg was administered as five divided doses at 4-hour intervals.
Within 7 days, approximately 70% of the ingested isotope was excreted in the
urine and feces of the test group on the normal diet. In the protein-restricted
diet group, the excretion value was approximately 35% of the ingested isotope.
In uremic patients, excretion of the isotope was comparable to that of healthy,
protein-restricted individuals.
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Ammonia April 1992
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Ingestion of an ammonium hydroxide solution containing 2.4% ammonia
resulted in the death of an adult male. Autopsy revealed a hemorrhagic
esophago-gastro-duodeno-enteritis, with ammonia odor in the stomach
contents (Klendshoj and Rejent, 1966).
Long-term Exposure
No data were found in the available literature on the chronic toricity to
humans of ammonia following oral administration.
Animals
Short-term Exposure
An acute oral LDj,, of 350 mg/kg was reported in rats administered
ammonia (Smyth et al., 1941).
Dermal/Ocular Effects
No studies were found in the available literature on the dermal or ocular
effects of ammonia by oral administration.
Long-term Exposure
No studies were found in the available literature on the long-term effects
of ammonia by oral administration.
Reproductive Effects
No studies were found in the available literature on the reproductive
effects of ammonia by oral administration.
Developmental Effects
No studies were found in the available literature on the developmental
effects of ammonia by oral administration.
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Ammonia April 1992
Mutagenicity
No studies were found in the available literature on the mutagenic effects
of ammonia by oral administration.
Carcinogenic! ty
No studies were found in the available literature on the carcinogenicity of
ammonia by oral administration.
Organoleptic Consideration
Campbell et al. (1958) determined the threshold concentration for
ammonia in redistilled water based on the responses of 21 to 22 judges
participating in "difference tests of the triangle type.11 At ammonia
concentrations of 26, 52 and 105 mg/L, the percentages of correct
identification by the judges were 61.9*. 71.4 and 85.7, respectively.
Defining the threshold concentration as the level at which correct
identification is 50% greater than that expected by chance, the taste
threshold of ammonia was determined to be 34 mg/L. Based on assumed
water consumption of 2 L/day and average body weight of 70 kg, this dose
corresponds to 0.97 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:
(NOAEL or LOAEL^ x (BW = mg/L ( ^tg/L)
(UF) (_ 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/OW guidelines.
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Ammonia April 1992
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 to use in the
determination of the One-day Health Advisory (HA).
Ten-day Health Advisory
No data were found in the available literature that were suitable to use in the
determination of the Ten-day HA.
Longer-term Health Advisory
No data were found in the available literature that were suitable to use in the
determination of the Longer-term HA.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure that is
attributed to drinking water and is considered protective of 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 estimate (with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is likely to be without
appreciable risk of deleterious health effects during 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 in drinking water alone 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 known, probable, or possible human carcinogen,
according to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then
caution must be exercised in making a decision on how to deal with possible lifetime exposure
to this substance. For human (A) or probable (B) human carcinogens, a Lifetime HA is not
recommended. For possible (C) human carcinogens, an additional 10-fold safety factor is used
in the calculation of the Lifetime HA. The risk manager must balance this assessment of
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Ammonia April 1992
carcinogenic potential and the quality of the data against the likelihood of occurrence and
significance of health effects related to noncarcinogenic endpoints of toxicity. To assist the risk
manager in this process, drinking water concentrations associated with estimated excess lifetime
cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of
water/day are provided in the Evaluation of Carcinogenic Potential section.
There are no suitable studies available in the literature for the derivation of a Lifetime
HA for ammonia. Ammonia is produced in humans in the stomach, duodenum, ileum, colon
and feces at an estimated 4,200 mg/kg/day, with the colon and fecal content contributing about
73%. However, ingestion of 2.4% ammonium hydroxide solution (24 g/L) has resulted in the
death of an adult male. Autopsy revealed a hemorrhagic esophago-gastro-duodeno-enteritis,
with ammonia odor in the stomach contents. It appears that the observed toxic effects were due
to the local effects rather than the systemic effects and ammonia at low concentration, per se, is
not very toxic. Therefore, it is recommended that the taste and odor level of 34 mg/L be used
as a guide for the Lifetime HA.
Evaluation of Carcinogenic Potential
The International Agency for Research on Cancer (IARC) has not evaluated the
carcinogenic potential of ammonia.
The weight of evidence that ammonia is a carcinogen has not yet been evaluated
by the EPA. Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), ammonia may be placed in Group D: not
classifiable. This category is for agents with inadequate animal evidence of
carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The American Conference of Governmental Industrial Hygienists (ACGIH)
suggests a Threshold Limit Value (TLV) of 25 ppm (18 mg/m3) as a
Time-Weighted Average (TWA) for an 8-hour work day (ACGIH, 1987-88). A
Short-Term Exposure Limit (STEL) of 35 ppm (27 mg/mj) has been suggested.
VII. ANALYTICAL METHODS
Ammonia is one of the classical water quality monitoring parameters, the
concentration of which is important in eutrophication problems. It has been
monitored at low levels in ambient waters for a long time. EPA Methods 350.1,
350.2, 350.3 are available for the determination of ammonia. Method 350.3 is a
specific ion electrode method, the others are colorimetric/titrimetric/
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Ammonia April 1992
potentiometric. Method 350.1 is an automated procedure. The type of
interferences from other nitrogen compounds such as amines is well documented.
The detection limits available by these methods ranges from 10 to 50 /xgm/L
(U.S. EPA, 1979).
VIII. TREATMENT TECHNOLOGIES
Available data indicate that aeration, reverse osmosis (RO), combination lime
softening and RO, and adsorption by natural zeolite significantly reduce
ammonia concentrations in the drinking water supply.
Powers et al. (1987) developed a mathematical model to examine the operating
parameters that might affect the ammonia removal rate by a batch
aerated-bubble stripping process. The model requires the liquid tank volume,
lime dosing for pH adjustment, steam flow rate dosing and air flow rate to
produce an ammonia concentration profile at a given initial ammonia
concentration. Using this model, the optimum operating conditions for this
stripper are: air flow rate of 110 cfm and a maximum steam flow rate of
700 lb/hr., with a required stripping time of 14 hours. This model predicted a
reduction of the ammonia concentration by 96% from a concentration of
4,000 mg/L as ammonia-nitrogen.
Shpirt (1981) using a bench-scale apparatus and a mathematical model
demonstrated that diffused aeration is a feasible technology for ammonia
reduction. The apparatus consisted of two different types, one giving coarse
bubbles (a section of glass pipe), the other giving fine bubbles (diffuser stone).
The initial concentration of ammonium chloride was in the range of 50 to
100 mg N/L, and the pH was adjusted to 11.5 with a 40% by weight solution of
sodium hydroxide. A series of experiments were performed at different column
heights and air flow rates. The results showed that the overall coefficient of mass
transfer increases with increasing air flow rate and decreases with increasing
depth of diffuser submersion. Furthermore, fine bubble aeration had twice the
efficiency as coarse bubble aeration in removing ammonia.
Houel et al. (1979) studied and reported the removal of ammonia by aeration in
a counter current stripping tower with a cross section of 23 inches by 18 inches
and a total packing depth of 13.83 feet. Air flow rate was maintained at
1,250 cfm and the water flow rate was maintained at 5 gpm. Two types of
packing materials were tested. Type A packing was a 0 J inches thick "egg crate"
type polystyrene sheet, and Type B packing was made from polyvinyl chloride
(PVC) sheets.. Ammonia was present at an influent concentration of 80 mg/L.
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Ammonia April 1992
Type A packing removed an average of 88% of the ammonia, while Type B
packing removed an average of 53%.
Benneworth and Morris (1972) studied the factors affecting ammonia removal by
aeration. They concluded that the rate of ammonia desorption increases rapidly
as the pH rises from 7 to 10.5 and above pH 10.5 no significant improvement can
be obtained. The air requirements are governed by the value of the solubility
coefficient. A theoretical minimum air to liquid ratio was calculated to be
1,400 for 90% ammonia removal.
O'Farrell and Bishop (1976) described a pilot plant which used aeration to
remove ammonia from a second-stage clarifier effluent at pH 10.5. The tower
was operated at a loading rate of a 2 gpm/sq. ft. at an air to water ratio of 2,600.
In general, this system, operated at 45ฐF, removed 56% of the ammonia from an
initial concentration of 9.5 mg/L measured as ammonia-nitrogen.
Terril and Neufeld (1983) reported data from a reverse osmosis unit used to treat
blast-furnace scrubber effluent. The ammonia concentration in the influent was
128 mg/L. The RO unit contained a cellulose acetate membrane (CA) and was
operated at pressures of 350 to 450 psig and a water recovery rate of 70 to 80%.
This system achieved 93% reduction in ammonia levels.
Argo (1984) reported the performance of a lime softening/RO plant for water
reclamation. Potable water was reclaimed from the unchlorinated effluent of an
activated sludge wastewater treatment plant by lime softening, reacidification and
RO. Two 5,000 gpd RO pilot plants consisting of tubular aromatic polyamide
membranes in a 2-1 array configuration were operated in parallel at a flux rate of
7.14 gpd/sq ft and at an applied pressure of 250 psi. This system reduced
ammonia concentration by more than 95% from an average influent
concentration of 15 mg/L.
Blanchard et al. (1984) studied removal of ammonia in a pilot plant using natural
zeolite clinoptilolite, as an ion exchanger. The pilot plant consisted of two
columns operated in series, each 8 inches in diameter and packed with 40 inches
of zeolite, and a flow rate of 12 bed volumes (BV) per hour, for a total BV =
1.16 cu. ft. Ammonia breakthrough occurred after 480 BV at an influent
concentration of 2.63 mg/L. Breakthrough concentration was set at 50 jig/L, after
which the zeolite was regenerated with NaCl at a flow rate of 10 BV/hr.
No data were found for the removal of ammonia from drinking water by
activated carbon adsorption. However, ammonia may not be amenable to
removal by activated carbon adsorption due to its very high solubility and low
molecular weight.
10
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Ammonia . April 1992
IX. REFERENCES
ACGIH. 1987-88. American Conference of Governmental Industrial.Hygienists. Threshold
limit values and biological exposure indices. Cincinnati, OH: American Conference of
Governmental Hygienists, p. 12.
Argo, O.R. 1984. Use of lime clarification and reverse osmosis in water reclamation. Jour.
Water Pollut. Cont. Fed. 56(12): 1,238-1,246.
Benneworth, N.E. and N.G. Morris. 1972. Removal of ammonia by air stripping. Jour. Water
Pollut. Cont. Fed. 71(5):485-492.
Blanchard, G., M. Maunaye and G. Martin. 1984. Removal of heavy metals from waters by
means of natural zeolites. Water Resources 18(12): 1,501-1,507.
Campbell, C.L., R.K. Dawes, S. Deolakkar and M.C. Merritt. 1958. Effects of certain
chemicals in water on the flavor of brewed coffee. Food Res. 23:575-579.
Castell, D.O. and E.W. Moore. 1971. Ammonia absorption from the human colon.
Gastroenterology 60(l):33-42.
Conn, H.O. 1972. Studies of the source and significance of blood ammonia. IV. Early
ammonia peaks after ingestion of ammonium salts. Yale J. Biol. Med. 45:543:549.
FASEB. 1974. Federation of American Societies for Experimental Biology. Evaluation of the
health aspects of certain ammonium salts as food ingredients. FDA contract no. FDA
72-85, NTIS PB-254-532.
Feth, J.H. 1966. Nitrogen compounds in natural water-A review. Water Resour. Res. 2:41-58.
Houel, N., F.H. Pearson and R.E. Selleck. 1979. Air stripping of chloroform from water. Jour.
of Environ. Engr. 105:777-781.
Kies, C. and H.M. Fox. 1978. Urea as a dietary supplement for humans. Adv. Exp. Med. Biol.
105:103-118.
Klendshoj, N.C. and T.A. Rejent. 1966. Tissue levels of some poisoning agents less frequently
encountered. J. Forensic Sci. 11:75-80.
O'Farrell, T.P. and D.F. Bishop. 1976. Conventional tertiary treatment. U.S. Environmental
Protection Agency. EPA-600/2-76-251.
11
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Ammonia ' April 1992
Perry, R.H. and C.H. Chilton. 1973. Chemical engineers handbook, 5th ed. New York, NY:
McGraw Hill Book Co.
Powers, S.E., A.G. Collins, J.K. Edzwald and J.M. Dietrich. 1987. Modeling an aerated bubble
ammonia stripping process. Jour. Water Pollut. Cont. Fed. 59(2):92-100.
Richards, P., B.J. Houghton, A. Metcalfe-Gibson, E.E. Ward and O. Wrong. 1968.
Incorporation of orally administered ammonia into tissue proteins in man: The
influence of diet and uraemia. In: Life Systems, Inc., ed., Nutrition and renal disease.
Edinburgh, Scotland: E&S Livingstone, pp. 93-98.
Richards, P., C.L. Brown, BJ. Houghton and D.M. Wrong. 1975. The incorporation of
ammonia nitrogen into albumin in man: The effects of diet, uremia and growth
hormone. Clin. Nephrol. 3(5):172-179.
Shpirt, E. 1981. Diffused-air stripping of ammonia in advanced wastewater treatment.
Chemistry in Water Reuse 2:497-508.
Smyth, H.F., Jr., J. Seaton and L. Fischer. 1941. The single dose toxicity of some glycols and
derivatives. J. Ind. Hyg. Toricol. 23(6):259-268.
Summerskill, W.H.J. and E. Wolpert. 1970. Ammonia metabolism in the gut. Am. J. Clin.
Nutr. 23(5):633-639.
Terril, M.E. and R.D. Neufeld. 1983. Reverse osmosis of blast-furnace scrubber water.
Environ. Progress 2(2):121-127.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Methods for the chemical analysis of
water and wastes. EPA-600/4-79-020. Environmental Monitoring and Support
Laboratory, ORD. Cincinnati, OH: U.S. EPA.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogenic risk
assessment. Fed. Reg. 51(185):33992-34003.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Development document for effluent
limits, guidelines and standards for the petroleum refinery point-source category. EPA
440/1-79-0146.
Verschueren, K. 1977. Handbook of environmental data on organic chemicals. New York,
NY: Van Nostrand Reinhold Co., p. 96.
12
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Ammonia April 1992
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The Merck index
An encyclopedia of chemicals, drugs and biologicals, 10th ed. Rahway, NJ: Merck and
Co., Inc., p. 498.
13
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EPA 0553
RX000027511 Apnl
ANTIMONY
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology, and treatment technology that would be
useful in dealing with the contamination of drinking water. Health Advisories (HAs) 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.
HAs 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 endpoints
of toxicity. For those substances that are known or probable human carcinogens, according to the
Agency classification scheme (Group A or B), Lifetime Health Advisories are not recommended.
For substances with a carcinogenic potential, chemical concentration values are correlated with
carcinogenic risk estimates by employing a cancer potency (unit risk) value together with
assumptions for lifelong exposure and the ingestion of water. The cancer unit risk is usually
derived from a linearized multistage model with 95% upper confidence limits providing a low-dose
estimate of cancer risk. The cancer risk is characterized as being an upper limit estimate, that is,
the true risk to humans, while not identifiable, is not likely to exceed the upper limit estimate and in
fact may be lower. While alternative risk modeling approaches may be presented, for example One-
hit, Weibull, Logit, or Probit, the range of risks described by using any of these models 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
carcinogenic risk, the Agency recommends using the linearized multistage model.
-------
Antimony Health Advisory
April 1992.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.
Antimony 7440-36-0
Potassium antimony tartrate 2800-74-5
Sodium antimony tartrate none
Sodium antimony bis(pyrocatechol) 2,4-disulfate none
Synonyms
Antimony black; CI77050; Regulus of antimony; Stibium (HSDB, 1988)
Uses
Alloy in semiconductor technology, batteries, antifriction compounds, ammunition,
cable sheathing, flameproofing compounds, ceramics, glass and pottery (Weast et
al., 1986). Also used in type castings for commercial printing.
Properties (ACGIH, 1980; HSDB, 1988; Weast et al., 1986; Windholz et al., 1983)
Antimony
Potassium
antimony
tartrate
Sodium
antimony
tartrate
Sodium antimony
bis(pyrocatechol)
2.4-disulfate
Chemical Formula
Molecular Weight
Physical State (25 ฐC)
Boiling Point
Melting Point
Density (20ฐC)
Vapor Pressure
Water Solubility
Specific Gravity
Log Octanol/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Sb KSbOCAOj
121.75 324.92 308.83
Silver-white, Colorless Transparent
hard, brittle crystals or or whitish
metal white powder scale or
powder
1,635ฐC - -
630ฐC 1008C -
6.691 - -
1 mmHg <3> 886ฐC
Insoluble 8.3 mg/L
6.691 2.6
Sweetish
metallic
Odorless
895.21
Fine crystals
66.7 mg/L Soluble
Occurrence
Antimony (Sb) is a naturally occurring element found as various salts in seawater,
surface water, soils and sediments. Its terrestrial abundance is of the order of 0.7
Mg/g (Brannon and Patrick, 1985). Antimony concentration in the earth's crust is
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Antimony Health Advisory April 199,2.
about 0.2 to 1.0 mg/kg (Rompp, 1979). More than half of the naturally occurring
,Sb in sediments is bound to extractable iron and aluminum (Crecelius et al., 1975).
Sb(III) and Sb(V) forms and methyl antimony compounds have been shown to exist
in natural waters (Andreae et al., 1981; Byrd and Andreae, 1982). Antimony occurs
in seawater at about 0.2 ^g/L (Venugopal and Luckey, 1978). Schroeder et ai.
(1970) found Sb in the air only in a few cities (4 of 58 had 0.42 to 0.85 /zg Sb/m3)
and in a few non-urban areas (3 of 29 with 0.001 to 0.002 /ig Sb/m3) of the United
States. Smoke condensate of cigarettes contains 35 to 60 mg Sb/kg (Gerhardsson
1983).
Ragaini et al. (1977) reported that lead smelting operations in the Kellog Valley of
Idaho resulted in soil Sb concentrations ranging from 5 to 260 jig/g. Crecelius et al.
(1975) reported total Sb concentrations in Puget Sound sediments within 8 to 15 km
from a copper smelter were 2 to 3 times higher than background values. Elevated
concentrations of Sb in sediment have also been noted near the outfall of sewage and
fertilizer facilities (Papakkostides et al., 1975). Sludges used for manuring soils in
Indiana (United States) and collected near Vienna (Austria) from the Danube River
contained between 4 and 22 mg/kg of Sb.
Environmental Fate
Various forms of Sb found in the environment from natural and anthropogenic
sources undergo a complex cycle of chemical interconversion and transfer between
media. Antimony in water may undergo either oxidation or reduction, depending on
pH and other ions present. Soluble forms of Sb (e.g., antimony potassium oxalate
and antimony potassium tartrate) tend to be quite mobile in water, while less soluble
species adsorb to clay or soil particles (Callahan et al., 1979).
Antimony in gaseous, vapor and paniculate forms enters the atmosphere and is
transported via air until it undergoes atmospheric fallout or washout and is deposited
in oceans, estuaries, lakes, rivers, sediments and terrestrial systems. Antimony may
enter the food chain via root uptake by terrestrial plants and via bioaccumulation in
fish- and plant-eating mammals. Antimony deposited in sediment can also be
released to the atmosphere through microbial activity under anaerobic conditions.
Antimony may leach from municipal landfills, sewage sludge, oil-fired plant
incinerator ash and fertilizers to contaminate ground water, surface water and
sediment (Callahan et al., 1979).
Brannon and Patrick (1985) determined that So-amended sediment releases volatile
Sb compounds during anaerobic incubation. Antimony evolution rates of 8 ^g/
mVweek were detected. Release of volatile Sb compounds, presumably stibine, was
most pronounced in the first week of incubation. Two sediment samples released
additional Sb when their overlying water was aerobic. These observations indicate
that release of volatile Sb compounds from anaerobic sediments containing recent
deposits of soluble Sb can occur regardless of the oxygen status of the overlying
water. However; no releases of volatile Sb compounds were noted from sediments
containing no added Sb. These results also suggest that So-amended sediments will
potentially release greater amounts of Sb during water-sediment interactions than
native So-containing sediments (Brannon and Patrick, 1985).
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Antimony Health Advisory April 1992.
ป In a study on leaching of contaminants from a municipal landfill site, Cyr et al.
(1987) found that Sb was present at a level of 0.01 mg/L in the landfill leachate but
was not detected in the background monitoring well. The river sediment near the
landfill site contained 23.9 mg/kg of Sb; however, there was no detectable Sb in the
sediment at the monitoring station further upstream. These results suggest that some
Sb is adsorbed by sediments.
Antimony is only slightly bioaccumulated and has been little studied in aquatic
organisms. Leatherland et al. (1973) found low levels of Sb in various fish and
invertebrates collected off the northwest coast of Africa; Sb was generally present in
higher concentrations in invertebrates than in fish. Aquatic organisms from the
Danube River and Danube Canal in Vienna, Austria, were found to contain only
background levels of antimony (Rehwoldt et al., 1975). Bertine and Goldberg
(1972) obtained similar results in clams, mussels and shrimp.
Antimony may be found as a gas in the form of stibine or its methylated derivatives.
Stibine can be formed by reduction of Sb in the sediments. The elements Sn, Pb,
As, Se and Te, which surround Sb in the periodic table, are subject to
biomethylation, suggesting the possibility of similar biomethylation pathways for Sb
(Parris and Brinkman, 1975, 1976). Stibine is rapidly oxidized in air or oxygenated
waters to form Sb2O3. It is likely then, that most of the stibine formed in the
sediments reacts in the water column to produce the oxide, resulting in
remobilization of Sb.
The methylated forms of Sb are also subject to oxidation. Parris and Brinkman
(1976) estimate the rate of oxidation of trimethylstibine as greater than 10"2/M/sec.
The product of this reaction, (CHj)3SbO, is much more soluble than trimethylstibine,
and, therefore, this oxidation would probably tend to reduce volatility. The rapid
rate of oxidation implies that, if trimethylstibine is formed in natural systems, much
of it would be oxidized before it volatilized and only a small amount of the volatile
antimony compounds formed by either abiotic or biotic mechanisms would be
liberated to the atmosphere.
It has been reported that a species of bacteria, Stibiobaaer senarmontii, utilizes the
energy released by metabolically induced oxidation of Sb (Lyalikova, 1974), but the
distribution and ecological importance of this organism is unknown.
III. PHARMACOKINETICS
Absorption
Antimony is not readily absorbed from the gastrointestinal (GI) tract. Approximately
15% of the administered dose of radioactive Sb (4.4 mg potassium antimony tartrate/
kg plus 7 jig of mSb) was estimated to be absorbed from the intestine of rats. The
animals (30 white rats) were dosed via gastric gavage or intravenous injection
(Moskalev, 1959).
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Antimony Health Advisory April
Gerber et al. (1982) reported that the total-body radioactivity (1.7% of the daily
intake) reached an equilibrium within 4 days in pregnant BALB/c mice following
exposures via diet and intraperitoneal (ip) injection. Assuming a half-life of 6 hours,
7% of the ingested SbCl3 (given as '"SbClj in food) was absorbed in pregnant
BALB/c mice following repeated dosing.
Combined elimination data from cows administered single oral or intravenous (iv)
doses of 124SbCl3 indicated that very little (< 5%) of the orally administered dose
was absorbed via the GI tract of ruminants. About 82% of the orally administered
radiolabel was excreted in the feces (Van Bruwaene et al., 1982).
Very little (9 to 15%) trivalent or pentavalent Sb was absorbed when administered
via oral gavage to 20 to 34 Syrian hamsters. Animals received oral doses containing
1 or 2 /iCi/mL of 124Sb-tartrate. For both valence states, Sb was retained in the
body, with a half-life of less than 1 day (Felicetti et al., 1974).
Distribution
Gerber et al. (1982) indicated that concentrations of mSb (expressed as percent daily
dose per gram tissue) in lung, bone, ovaries and uterus ranged from 0.085 to 0.2%
when given in food to pregnant BALB/c mice. The diet containing mSbC!3 was
started on the day the vaginal plug was observed. The animals were sacrificed 6
days later, and tissue levels of 125Sb were measured. The results were judged by the
authors to be somewhat unreliable due to low levels of radioactivity observed.
Westrick (1953) fed diets containing 0 or 2% Sb,Oj for 7 weeks to five male
Sprague-Dawley rats. Using a mean body weight of 0.18 kg (the mean of reported
initial and final weights) and assuming average food consumption of 12 g/day
(Arlington, 1972), this corresponds to an average daily dose of about 1,100 mg Sb/
kg/day. After 7 weeks, average concentration of Sb in the liver, kidney, heart,
spleen, lung, adrenal and thyroid was 8.9, 6.7, 7.6, 18.9, 14, 67.8 and 88.9 /xg Sb/g
tissue, respectively. A wide variation was observed among replicates of some
tissues.
Gerber et al. (1982) also measured tissue distribution in pregnant BALB/c mice
following ip injection of l25SbCl3 on day 12 of pregnancy. Peak concentrations in
tissues (percent dose per gram tissue) were observed 2 to 6 hours after injection.
Highest tissue levels (approximately 50%) were seen in the intestine and bone
surfaces. Levels in other tissues observed at 2 hours were 1 to 5% in the uterus and
ovary, and 0.01 to 1.0% in the kidneys, liver, spleen, lung, thyroid, blood, muscle,
skin and brain. Low levels (about 0.1%) were measured in the placenta and fetus.
Casals (1972) injected two groups of 10 female mice intramuscularly (im) either with
antimony dextran glucoside (RL-712, 52 mg Sb/kg) or with N-methyl-glucamine-
antimonate (glucantime, 50.3 mg Sb/kg). Animals were sacrificed at various
intervals between 6 hours and 6 weeks after dosing. High levels (100 to 150 Mg) of
Sb were found mainly in the liver and spleen throughout that period; the levels
decreased gradually with time.
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Antimony Health Advisory April 199Z
Rowland (1971) presented a detailed mathematical model of the distribution of Sb in
humans following a single iv injection of 124Sb-labeled potassium antimony tartrate
(PAT) . Four main compartments (blood, liver, skeletal tissue and urine) were
considered by determining radioactivity in the blood and the urine, and via surface
scanning over the liver and over the thigh (as representative of skeletal tissue). The
results from studies on a single patient revealed that 12 minutes after an injection of
48 mg labeled PAT, 12% of the radioactive dose was present in the blood, which
declined to 8 and 6% at 1 and 2 hours post-injection. Another experiment revealed
that 40 to 50% of a 24 mg dose of labeled PAT was taken up by the liver within a
few minutes to 48 hours post-injection. On average, this was followed by a slow
decline with about 50% of the peak value present 16 hours post-treatment. The
content of skeletal tissue reached its peak in 1 hour with a slow decline with 30% of
the dose remaining after 10 days post-treatment. In the urine, 7% of a 48 mg
labeled PAT dose appeared in the first hour followed by 15, 22, 25 and 45% of the
dose at 6 hours and day 1, 2 and 3, respectively.
Leffler and Nordstroem (1984) demonstrated the transfer of Sb from maternal to
fetal blood in three Syrian Golden hamsters intratracheally exposed t;o Sb on days 13
and 15 after fecundation. Concentration gradients of Sb determined in maternal and
fetal blood suggested a possible carrier function of the placenta.
Molokhia and Smith (1969) incubated Sb (trivalent or pentavalent) compounds with
equine whole blood in vitro and found that the erythrocyte membrane was permeable
to trivalent antimony and impermeable to pentavalent antimony. Trivalent Sb bound
to plasma proteins but not to erythrocytes.
Metabolism
Otto and Maren (1950) found large amounts of Sb in erythrocytes following im
injection of stibanose (6 mg Sb(V)/kg) in 11 dogs. Other results (Molokhia and
Smith, 1969; Otto et al., 1947) had shown that Sb(V) does not enter erythrocytes but
that Sb(IH) does, suggesting that the Sb(V) may have been reduced to Sb(III) in
vivo. In contrast, at a lower im dose of 0.5 mg Sb(V)/kg, Otto and Maren (1950)
did not detect Sb accumulation in erythrocytes of dogs. Similarly, iv injection of
either 0.5 or 5 mg Sb(V)/kg did not reveal Sb entry into red cells. The authors
stated that the available data were not sufficient to support a conclusion regarding
possible reduction of Sb(V) to Sb(III).
Goodwin and Page (1943) used polarography to analyze the valence state of
antimony in blood and urine of seven humans injected iv with Sb(V). During the
first 12 hours after the administration of Sb(V) (sodium antimony gluconate
equivalent to 50 Mg Sb), 83.5% (average of three subjects) of the administered dose
was excreted in the urine as Sb(V). Only 2.5% of the administered dose was
excreted as Sb(III) during the same period, indicating that reduction of Sb(V) to
Sb(III) was slight. Otto and Maren (1950) pointed out that some of the Sb(III) found
in urine may have been formed during sample preparation in hydrochloric acid for
polarographic examination.
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Antimony Health Advisory April 1992.
Otto et al. (1947) administered two Sb(m) compounds (lithium antimony thiomalate
and monosodium antimony thioglycollate) im and two Sb(V) compounds (stibanose
and neostibosan) iv to 14 adult male filariasis patients. Antimony concentrations in
red blood cells and plasma were measured colorimetrically. For both Sb(III) and
Sb(V), plasma concentrations were sustained for only a short time (well under 24
hours). For both trivalent compounds, Sb was found largely inside the red blood
cells, with very little in plasma, and the converse was observed for both pentavalent
compounds. The authors concluded that Sb(III) readily enters red blood cells, but
Sb(V) does not.
Excretion
Otto and Maren (1950) reviewed the routes of excretion of parenterally administered
antimony in the mouse, white rat, hamster, guinea pig, rabbit, dog and human.
Trivalent antimony was excreted via the feces and urine. With the exception of the
mouse, pentavalent antimony was excreted primarily in the urine. While the percent
of the dose excreted in the feces was less than 5% for all species tested, the percent
excreted in the urine was approximately 80, 60, 65, 70, 10 and 43% in the white
rat, hamster, guinea pig, rabbit, dog and human, respectively.
Casals (1972) injected (im) female mice either with antimony dextran glucoside (RL-
712, 52 mg Sb/kg) or with N-methyl-glucamine-antimonate (glucantime, 50.3 mg
Sb/kg), and female albino rats with RL-712 (50 mg Sb/kg). In 48 hours, only 12
and 10% of the doses administered were excreted in the urine of mice and rats,
respectively.
Van Bruwaene et al. (1982) administered single oral doses of 124SoCl3 (2.84, 2.72 or
2.00 mCi) to three lactating cows. Since the compound had a specific activity of 3.5
x 10"2 mCi/mmol, the average dose corresponds to 21.1 mg Sb/kg. Total excretion
of Sb in feces was triphasic and totaled about 82% of the dose. Most of the
radioactivity in the feces appeared shortly after dosing (t,/, = 0.91 day). Excretion
in urine and milk was biphasic and totaled 1.1 and 0.008% of the dose in urine and
milk, respectively. Most of the urinary radioactivity appeared in the initial phase (t,/
2 = 0.97 day). Radioactivity in tissues, at 102 days after dosing, totaled 0.024% of
the oral dose. The highest radioactivity was found in the spleen, liver, bone and
skin.
Lippincott et al. (1947) administered potassium antimony tartrate (18 mg Sb/mL to
77 patients over 29 days, for a total of 576 mg/patient) or ruadin (8.7 mg Sb/mL to
33 patients over 25 days, for a total of 566 mg/patient) parenterally to humans as
treatment for infection with Schistosoma japonicum. The average 24-hour excretion
of Sb in urine of subjects given potassium antimony tartrate ranged from 12 to 25%,
whereas excretion in the group given ruadin ranged from 17 to 42%. The combined
excretion of Sb in urine and feces within 48 hours was approximately 55 % of the
administered dose.
In the Otto et al. (1947) study in adult males, most (21.6 to 70% of the daily dose)
of the Sb administered daily was excreted in urine in 24 hours with only low levels
(0.8 to 8.4% of the daily dose) present in feces. Pentavalent antimony was excreted
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Antimony Health Advisory April 199,2,
in urine more rapidly than trivalent antimony (up to 52% in 24 hours as opposed to
15% in 24 hours for trivalent).
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Kaplan and Korff (1937) briefly reviewed several reports of "food poisoning" that
were traced to Sb extracted from enamel-coated vessels by acid contents (lemonade).
Symptoms were not detailed, but acute attacks of vomiting occurred in at least one
case. The amount of Sb ingested was not reported. Tests performed by the authors
indicated that 0.5 to 2.6 mg of Sb, an amount equal to about one-fourth of an emetic
dose, could be extracted from 200 g of sauerkraut.
Miller (1982) reported a fatality due to Sb poisoning. The patient was administered
two or three oral doses of James powder (each dose containing 66 mg of Sb) as
treatment for headache, kidney trouble and fever. The total dose was 132 to 198 mg
of Sb (1 to 1.5 mg/kg of body weight). The treatment resulted in severe vomiting
and diarrhea lasting for 18 hours and, finally, death.
Jolliffe (1985) reported that sodium stibogluconate ("Pentostam"), given iv in a
standard daily dose of 600 mg Sb(V) for 10 days to 16 British soldiers with
cutaneous leishmaniasis, did not adversely affect either glomerular or renal function.
Schroeder et al. (1946) reported the effect of trivalent and pentavalent antimony
compounds on the electrocardiogram of human patients being treated for
schistosomiasis. Sodium antimony bis(pyrocatechol-2,4-disulfonate) (Stibophen NF
or fuadin) was given im and potassium antimony tartrate was given iv daily or on
alternate days for about 1 month. Assuming an average body weight of 70 kg,
average daily doses ranged from 0.24 to 0.89 mg Sb/kg/day. Examination of 315
electrocardiograms (EKGs) from 100 patients revealed that the EKGs were not
indicative of cardiac damage or serious impairment of cardiac function.
Rugemalila (1980) reported two deaths due to parenteral antimony (astiban)
intoxication. The first case involved a 4-year-old girl with a history of periodic
fevers. She was administered (route not specified) 100 mg astiban (stibocaptate) for
active schistosomiasis (about 2 mg Sb(UI)/kg), followed by a second dose 2 weeks
later. The second case involved a 70-year-old woman who was put on weekly
injections of 320 mg astiban (about 2 mg Sb(III)/kg) to control hookworm. Both
patients died shortly after the second dose.
Long-term Exposure
Oliver (1933) examined six adult males who had worked in a Sb smelter for 2 to 13
years. The workers had received considerable exposure to Sb as evidenced by the
presence of Sb in the feces (an average of 47.5 mg total/day) but not in the urine.
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Antimony Health Advisory April L99JJ.
No signs of adverse effects were identified, including cardiac, kidney or bladder
effects, general health and hematology.
Brieger et al. (1954) examined workmen in a plant where antimony trisulfide was
used in the manufacture of grinding wheels. Antimony levels ranged from 0.58 to
5.5 mg/m3 (equal to approximately 0.4 mg/kg antimony trisulfide). In the workers
studied, 14 of 113 had blood pressures that were > 150/90 mmHg, and 37 of 75
showed significant changes in their EKGs, mostly in the T-wave. Ulcers were
detected in 7 of 111 exposed persons (63/1,000) as compared with 15/1,000 in the
total plant population. No other disorders suspected of being related to Sb exposure
were observed.
Chulay et al. (1985) observed EKG changes in 59 Kenyan patients dosed orally with
10, 20 or 40 to 60 mg/kg/day Sb(V) (sodium stibogluconate) for leishmaniasis.
Dose-related increases in EKG abnormalities were found following 65 courses of Sb
treatment which lasted for 4 months. The incidences of EKG abnormalities were
22% (2/9 patients) at 10 mg Sb/kg/day; 52% (25/48 patients) at 20 to 30 mg Sb/kg/
day; and 100% (8/8 patients) at 40 to 60 mg Sb/kg/day. Furthermore, the frequency
of EKG abnormalities increased with the duration of treatment.
Belyaeva (1967) presented evidence suggestive of possible adverse effects of Sb in
female workers employed in an Sb plant. The female workers in the Sb plant
showed increased incidence of spontaneous late abortions (12.5%) when compared to
female workers working under similar conditions but not exposed to Sb dust (4.1 %).
Doll (1985) compared mortality due to lung cancer with mortality due to other
causes in a British factory manufacturing antimony oxide. According to the authors,
the compound was manufactured before World War II, but no records were available
before 1961 either for dust measurements or for males who terminated employment.
An increase in lung cancer (Standardized Mortality Ratio = 186) was observed for
men first employed prior to 1961. Antimony oxide dust concentrations have been
greatly reduced since the 1960s.
Potkonjak and Pavlovich (1983) reported clinical findings for 51 male workers (ages
31 to 54) exposed for 9 to 31 years to dust containing a mixture of antimony trioxide
(ranging from 39 to 89%) and antimony pentoxide (ranging from 2 to 8%) in an
antimony smelting plant. Characteristic changes observed in the smelters were
described as a form of pneumoconiosis simplex or antimoniosis. The symptoms
observed included chronic coughing, conjunctivitis, orange-colored staining of
frontal tooth surface, chronic bronchitis, chronic emphysema, inactive tuberculosis
and pleural adhesions. "Antimony dermatitis" characterized by vesicular or pustular
lesions was seen in more than half the exposed workers.
Animals
Short-term Exposure
The acute oral LDM values for potassium antimony tartrate (tartar emetic) in mice
and rats range from 115 to 600 mg Sb/kg (Bradley and Fredrick, 1941; HSDB,
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Antimony. Health Advisory April 1992,
1988), whereas an oral LDM of 15 mg Sb/kg has been reported for rabbits (HSDB,
1988). The iv and ip LDM values for Sb and its various salts in mice, rats, guinea
pigs and rabbits are generally somewhat lower, ranging from 11 to 329 mg Sb/kg
(Bradley and Fredrick, 1941; Ercoli, 1968; Ghaleb et al., 1979; Girgis et al., 1965;
HSDB, 1988).
Flury (1927) determined emetic doses of six Sb compounds (antimony trioxide,
antimony pentoxide, sodium antimonate, potassium antimonate, sodium meta-
antimonate and potassium antimony tartrate) in dogs. Potassium antimony tartrate
was the most effective compound, causing emesis at 33 mg/kg (about 12 mg Sb/kg).
A dose of 16 mg/kg (about 6 mg Sb/kg) produced no apparent effect.
Cats appear to be more sensitive than dogs to the emetic effect of potassium
antimony tartrate. Flury (1927) observed emesis in three cats given 11.5 or 14.3
mg/kg (4.3 or 5.4 mg Sb/kg) in 50 mL of water via stomach tube. One cat dosed
with 6.9 mg/kg (2.6 mg Sb/kg) showed no apparent response.
Flury (1927) fed high doses of five Sb compounds to rats (one per chemical) for 9
days. Each rat received daily doses increasing from 100 mg to 2 or 3 g. Doses up
to 2 g/day of antimony trioxide or antimony pentoxide or up to 3 g/day of sodium
meta-antimonate caused no adverse effects. Potassium antimony tartrate was found
to be toxic, however, causing death after the daily dose was increased to 500 mg
(about 1,000 mg/kg Sb/kg) on day 7. Potassium antimonate produced adverse
effects at dose levels of 2 g/day, but recovery was rapid when dosing ceased.
Pribyl (1927) investigated the toxicity and effect on nitrogen metabolism in four
rabbits given 15 mg potassium antimony tartrate/kg/day (given in a milk plus sugar
solution) over a 7- to 22-day period. This corresponds to a dose of 5.6 mg Sb/kg/
day. Nonprotein nitrogen, urea nitrogen, and ammonia nitrogen were measured in
the blood and urine of each animal before and after exposure. A small rise (10 to
13%) in nonprotein nitrogen in blood and urine was observed (no p value given);
this was partly due to an increase in urea nitrogen. Mean urine ammonia nitrogen
was also slightly increased (7%, no p value given). The author interpreted these
increased nitrogen levels in blood and urine as evidence of increased protein
catabolism. Gross and microscopic examination showed hemorrhagic lesions in the
stomach and small intestine, liver atrophy with fat accumulation and congestion, and
hemorrhage in the kidney cortex, with some tubular necrosis. This study suggested
a Lowest-Observed-Adverse-Effect Level (LOAEL) of 5.6 mg/kg/day based on
minimal histological injury in tissues.
Dermal/Ocular Effects
No information was found in the available literature on the dermal/ ocular effects of
Sb in experimental animals. However, an accumulated dose of 30 g of Sb (iv
injections of potassium antimony tartarate, 87.5 g in 86 days) produced leukoderma
and rough, dull, bumpy, granular skin in an 18-year-old male patient
(Christopherson, 1921).
10
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Antimony Health Advisory April 199ฃ,
Fifty-one male workers, exposed for 9 to 31 years to dust containing a mixture of
antimony trioxide and antimony pentoxide in an antimony smelting plant, were
examined for clinical effects. Thirty-two of the 51 had "antimony dermatitis,"
characterized by vascular or pustular lesions (Potkonjak and Pavlovich, 1983).
Long-term Exposure
Flury (1927) exposed rats (two per test group) to potassium antimony tartrate,
potassium antimonate, antimony trioxide, or antimony pentoxide in food. Doses
began at 0.1 mg/day and were steadily increased over the course of 107 days to a
final level of 4 mg/day. No toxic effects were observed, and growth was unaffected
except for a stimulation of growth at low doses.
In another study, Flury (1927) tested higher doses of potassium antimony tartrate,
antimony trioxide, and sodium meta-antimonate in food for 131 days. Groups of
two rats were exposed to doses of the first two compounds beginning at 1 mg/day
for 45 days, and then increasing over the course of 86 days to 200 mg/day, with the
dose being increased steadily over the course of exposure. The third compound was
given in doses from 3 to 1,000 mg/day in a similar pattern. No effects were seen,
even at the highest doses, for antimony trioxide and sodium meta-antimonate, but
potassium antimony tartrate caused a systemic deterioration and death at 200 mg/
day. This corresponds to about 485 mg Sb/kg/day, based on a mean body weight of
155 g reported by the author..
In a 91-day oral toxicity study in male/female Wistar rats, Bombard et al. (1982)
reported that two Sb-containing pigments produced no effects on behavior, food
consumption, growth, mortality, hematological and clinical data, or organ weights.
The pigments were nickel rutile yellow [(Tio ggSb0.0jNioa7J)OJ and chrome rutile
yellow [(Tio.ปซSbo.
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Antimony Health Advisory April 199Z
females during the first year, but did result in weight loss in males after 18 months
(p < 0.025) and decreased weight gain in females measured at 12 and 18 months (p
<0.005). In females, Sb shortened the median and 75% life spans by 49 to 86 days
whereas males were only minimally affected. Upon necropsy, histologic
examination of the liver revealed no significant differences in the incidence or
degree of fatty degeneration between controls (22.2%) and animals fed Sb (16.4%).
This study suggests a LOAEL of 0.5 mg Sb/kg/day based on minimal liver fatty
degeneration and decreases in weight and longevity (an average mouse weight of 40
g, and 4 mL water consumption/day containing 5 ppm Sb is assumed).
Schroeder et ai. (1970) administered potassium antimony tartrate (0 or 5 mg Sb/L)
in drinking water to groups of at least 50 male and 50 female Long-Evans rats from
the time of weaning until death. This corresponds to an average daily dose of 0.35
mg Sb/kg/day, based on authors' calculations. Mean longevity in days ฑ standard
error (SE) was 1,160ฑ27.8 for control males, 1,304ฑ36.0 for control females,
999ฑ7.8 for treated males, and 1,092ฑ30.0 for treated females. Antimony had a
negligible effect on body weight. Serum cholesterol levels were increased in male
rats (97.6ฑ4.9 mg/100 mL in dosed males versus 77.5ฑ2.1 mg/100 mL in controls)
and decreased in female rats (97.0ฑ5.6 mg/100 mL in dosed females versus
116.0ฑ6.0 in controls) when compared to control animals. Fasting blood glucose
levels were not significantly different in either males or females, but nonfasting
blood glucose levels were lower in both males (94.5 ฑ6.2 mg/100 mL in dosed
males versus 134.4ฑ5.1 mg/100 mL in controls) and females (82.5 ฑ7.0 mg/100
mL in dosed females versus 114.2ฑ5.4 mg/100 mL in controls). No significant
effects of Sb on glucosuria, proteinuria, heart weight or heart/body weight ratio
were observed. There was no evidence that Sb induced carcinogenesis. Deposition
of Sb in kidney, liver, heart, lung and spleen also was observed (mean range 10.14
to 17.67 /xg/g, value variation range 1.7 to 60.1 ng/g; no Sb was detected in control
samples). Antimony accumulated in the soft tissues with age (from 279 to 1,070
days); pooled samples showed a tendency to increase in concentration (p < 0.05),
with a correlation coefficient of 0.525. This study identified a LOAEL of 0.35 mg
Sb/kg/day based on decreased longevity, and altered blood glucose and serum
cholesterol levels.
Reproductive Effects
Belyaeva (1967) investigated the reproductive effects of antimony trioxide in rats
following repeated inhalation exposures to 250 mg/m1 of SbO, dust over a 2-month
period. Sterility and fewer offspring were noted in dosed rats when compared to the
control group.
Hodgson et al. (1927) injected female rabbits with 7 to 17 10-mg doses of sodium
antimony tartrate (2.2 mg Sb/kg) or 9 to 16 50-mg doses of an unknown organic
antimony compound over 16 to 38 days, and English white mice (male and female)
with 30 to 39 doses of 10 mg of another unknown organic antimony salt over 60 to
77 days. In general, in the female rabbits and mice, contraception, abortion and
fetal damage (details not specified) occurred; in male mice, the antimony salt did not
cause sterility.
12
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Antimony Health Advisory April 1992,
Developmental Effects
James et al. (1966) fed antimony potassium taitrate to four yearling ewes at a dose
level of 2 mg/kg of body weight for 45 days or throughout gestation. All ewes fed
Sb gave birth to normal, full-term lambs. No adverse effects were noted in ewes at
necropsy.
Casals (1972) reported the absence of any abnormalities in Wistar rat fetuses whose
mothers were exposed to the pentavalent antimonial drug RL-712 (antimony dextran
glycoside) during gestation. Five im injections of 125 or 250 mg Sb/kg were
administered between days 8 and 14 of gestation.
Mutagenicitv
Kanematsu and Kada (1978) and Kanematsu et al. (1980) determined that antimony
trichloride, antimony pentachloride and antimony trioxide were mutagenic in the
Bacillus subtilis (H17 and M45) rec-assay. An improved rec-assay procedure was
employed to reveal the DNA damaging capacity of the three Sb compounds.
Potassium antimony tartrate and sodium antimony tartrate induced chromosomal
aberrations in cultured human leukocytes (Paton and Allison, 1972; Hashem and
Shaw Id, 1976). Piperazine antimony tartrate and potassium antimony tartrate
induced chromosomal aberrations in bone marrow cells of rats injected ip (El Nahas
etal., 1982).
Carcinogenicitv
Schroeder et al. (1968), as described previously in the Longer-term Exposure
section, studied the effect of lifetime exposure to Sb on tumor frequency in mice.
Tumors (benign and malignant) were found in 34.8% of control animals (no
explanation was given for the high tumor incidence in controls) and 18.8% of the
So-treated animals. The authors concluded that Sb exposure had no effect on the
incidence or type of spontaneous benign or malignant tumors.
Schroeder et al. (1970), as described previously (in the Longer-term Exposure
section) studied the effect of lifetime exposure to antimony on tumor frequency in
rats. No significant effects of Sb exposure on tumor frequency were observed in
either male or female rats.
Watt (1983) reported that antimony trioxide induces fibrosis and neoplasms in female
rats when inhaled at levels close to the Threshold Limit Value (TLV). Female CDF
rats and S-l miniature swine were exposed to antimony trioxide dust at 1.6 ฑ1.5 mg/
mj (as Sb) or 4.2ฑ3.2 mg/m3 (as Sb) for 6 hours/day, 5 days/week, for
approximately 1 year. The lungs of exposed animals (rats and swine) were mottled
and heavier than the lungs of unexposed animals. Primary lung neoplasms were seen
in rats but not in swine. Most of the neoplasms were seen in the higher dose group
and were either scirrhous carcinomas, squamous cell carcinomas or bronchoalveolar
adenomas. The incidence and/or severity of the response was related to the
exposure time and the exposure level.
13
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Antimony Health Advisory
April 199*.
Groth et al. (1986) exposed three groups (90 males and 90 females/ group) of albino
rats via inhalation to SbjO, (mean Time-Weighted Average (TWA) = 45.0 and 46.0
mg Sb2Oj/mJ in two chambers), to Sb ore concentrate (mean TWA = 36.0 and 40.1
mg Sb ore/m3 in two chambers) or to filtered air. Histopathological examinations
revealed the presence of lung neoplasms (squamous cell carcinomas,
bronchioloalveolar adenomas, bronchioloalveolar carcinomas or scirrhous
carcinomas) in 27% of the females in the SbjO, group and 25% of the females in the
Sb ore concentrate group. No lung tumors were found in the male rats or control
females.
V.
QUANTIFICATION OF TOXICQLOGICAL EFFECTS
Health Advisories (HAs) are based upon the identification of adverse health effects
associated with the most sensitive and meaningful noncarcinogenic end point of toxicity. The
induction of this effect is related to a particular exposure dose over a specified period of time, most
often determined from the results of an experimental animal study. Traditional risk characterization
methodology for threshold toxicants is applied in HA development. The general formula is as
follows:
where:
NOAEL
or
LOAEL
BW
UF(s)
No-Observed-Adverse-Effect Level (the exposure dose in mg/kg/ bw/
day)
Lowest-Observed-Adverse-Effect Level (the exposure dose in mg/kg bw/
day)
Assumed body weight of protected individual (10 kg for child or 70 kg
for adult)
Uncertainty factors, based upon quality and nature of data (10, 100,
1,000, or 10,000 in accordance with EPA or NAS/OW guidelines)
L/day = Assumed water consumption (1 L/day for child or 2 L/day for adult)
One-day Health Advisory
No information was found in the available literature that was suitable for determination of a
One-day Health Advisory (HA) for Sb. Accordingly, it is recommended that the Drinking Water
Equivalent Level (DWEL) (10 /tg/L, calculated below) for a 10-kg child be used at this time as a
conservative estimate of the One-day HA.
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Antimony Health Advisory April 1992.
Ten-dav Health Advisory
The study by Pribyl (1927) was evaluated as the basis for calculation of the Ten-day HA for
a 10- kg child. However, this study appeared dated and offered very little study details upon which
a confident analysis could be made. In the absence of appropriate data, it is recommended that the
DWEL of 10 jig/L be used as a conservative estimate for a 10-day exposure in children.
Longer-term Health Advisory
No information was found in the available literature that was suitable for determination of
the Longer-term HA value for Sb. It is, therefore, recommended that the DWEL (0.01 mg/L,
calculated below) for a 70-kg adult be used as a conservative estimate for a longer-term exposure.
Since the Reference Dose (RfD) and DWEL were based on a lifetime study in rodents and a large
safety factor (1,000) was incorporated into their derivation, it can be assumed that the DWEL will
more than adequately protect both adult and children over longer-term exposure.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure that is attributed to
drinking water and is considered protective of noncarcinogenic adverse health effects over a lifetime
exposure. The Lifetime HA is derived in a three-step process. Step 1 determines the RfD,
formerly called the Acceptable Daily Intake (ADI). The RfD is an estimate (with uncertainty
spanning perhaps an order of magnitude) of a daily exposure to the human population (including
sensitive subgroups) that is likely to be without appreciable risk of deleterious health effects during
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 in
drinking water alone 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 known, probable, or possible carcinogen, according to
the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then caution must
be exercised in making a decision on -how to deal with possible lifetime exposure to this substance.
For human (A) or probable (B) human carcinogens, a Lifetime HA is not recommended. For
possible (C) human carcinogens, an additional 10-fold safety factor is used in the calculation of the
Lifetime HA. The risk manager must balance this assessment of carcinogenic potential and the
quality of the data against the likelihood of occurrence and significance of health effects related to
noncarcinogenic endpoints of toxic ity. To assist the risk manager in this process, drinking water
concentrations associated with estimated excess lifetime cancer risks over the range of 1 in 10,000
to 1 in 1 ,000,000 for the 70-kg adult drinking 2 L of water/day are provided in the Evaluation of
Carcinogenic Potential section.
15
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(ฐ'35 kg/day) = ฐ-00035 mg S^S/day (rounded to 0.0004 mg Sb/kg/day)
Antimony Health Advisory April 1992.
The study by Schroeder et al. (1970) has been selected to serve as the basis for calculation
of the Lifetime HA for Sb because it involved lifetime exposure of rats to potassium antimony
tartrate (the most toxic of the common antimony compounds) given in drinking water. This study
identified a LOAEL of 0.35 mg/kg/day on the basis of decreased longevity and altered blood levels
of glucose and cholesterol.
Using a LOAEL of 0.35 mg Sb/kg/day, the Lifetime HA is calculated as follows:
Step 1: Determination of RfD
RfD
where:
0.35 mg Sb/kg/day = LOAEL, based on decreased longevity and altered blood glucose
and cholesterol in rats exposed to potassium antimony tartrate in
drinking water for a lifetime (Schroeder et al., 1970).
1 ,000 = uncertainty factor. This uncertainty factor was chosen in
accordance with EPA or NAS/OW guidelines for use with a
LOAEL from an animal study.
Step 2: Determination of Drinking Water Equivalent Level (DWEL)
DWEL = (0.0004 mg Sb/kg/day) (70 kg) = Q Q14 fflg $b/L (rQunded ^ {Q ^
a (2 L/day)
where:
0.0004 mg Sb/kg/day = RfD.
2 L/day = assumed water consumption of 70-kg adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA - (0.014 mg Sb/L) (20%) 0.0028 mg Sb/L (3 /xg Sb/L)
where:
0.014 mg Sb/L = DWEL
20% = assumed relative source contribution from water.
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Antimony Health Advisory April 1992,
Evaluation of Carcinogenic Potential
No evidence of carcinogenicity was found in two lifetime studies ia which mice and rats
were supplied with drinking water containing 0 or 5 ppm Sb (as potassium antimony tartrate)
from the time of weaning until death (Schroeder et al., 1968, 1970).
Primary lung neoplasia was observed in rats exposed via inhalation to antimony trioxide dust
at 1.6ฑ1.5 mg/m3 (as Sb) or 4.2ฑ3.2 mg/m3 (as Sb) for 6 hours/day, 5 days/week for
approximately 1 year (Watt, 1983). Since no systemic neoplasia was evident and no
absorption data were presented, these data cannot be utilized in assessing the potential
carcinogenicity of a soluble form of Sb in drinking water.
Groth et al. (1986) exposed rats to antimony ore and antimony trioxide via inhalation.
Histopathological examinations revealed the presence of lung neoplasms (squamous cell
carcinomas, bronchioloalveolar adenomas, bronchioloalveolar carcinomas or scirrous
carcinomas) in 27% of females dosed with SbjOs and 25% of the females in the Sb ore
category. No lung rumors were found in male rats or control females.
Applying the criteria described in EPA's guidelines for assessment of carcinogenic risk
(U.S. EPA, 1986), Sb may be classified in Group D: not classifiable. This group is for
substances with inadequate human and animal evidence of carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The American Conference of Governmental Industrial Hygienists recommends a
TWA of 0.5 mg Sb/m3 (HSDB, 1988).
VII. ANALYTICAL METHODS
Methods for metal analysis involve spectroscopy, either emission or absorption. In all of the
methods the metal is dissolved and thermally excited. All elements when excited emit or
absorb light frequencies characteristic of that element. Most metal spectroscopy is done in
the ultra-violet and x-ray regions.
Direct Aspiration Atomic Absorption Spectroscopy (AA). In this technique, the
dissolved metals are aspirated into a flame source, and excited to the point that the
metals are dispersed to a mono-atomic state, a light source whose cathode is the
metal of interest passes through the flame, the resulting absorption of light by the
element of interest is directly proportional to concentration. The disadvantages of
this technique is the one at a time determination of the metals and the insensitivity of
the technique. This is EPA Method 204. 1 with a detection limit of 50
Graphite Furnace Atomic Absorption (GFAA). In this technique a specific amount
of liquid is dried on the thermal source, effecting a concentration step. The sample
is electro-thermally excited. This technique has great sensitivity. This is EPA
Method 204.2 with a detection limit of 0.8 /*g/L.
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Antimony Health Advisory April 199iL
Inductively Coupled Plasma Atomic Emission Spectroscopy. This method uses
aspiration of a liquid sample, but the flame is actually a plasma torch of Argon
excited to super hot levels by radio-frequency (RF) radiation. Here the metals are
excited to the levels where they emit radiation. By using classical dispersion grating
optics a large number of metals can be determined simultaneously. This is EPA
Method 200.7 with a detection limit of 8 /ig/L.
Inductively Coupled Plasma Mass Spectrometry (TCP/MS). In this case the
excitation is again by an RF plasma, but the excited atoms are then interfaced into a
mass spectrometer. Quantitation is achieved by computerized software programs
similar to those in other EPA MS organic methods. This is EPA Method 6020 with
detection limit of 0.02 /xg/L.
VIII. TREATMENT TECHNOLOGIES
The available literature indicates that conventional coagulation/ filtration will possibly
remove Sb from contaminated drinking water. If granular activated.carbon (GAC)
adsorption is added as a post-treatment, then the removal efficiency will be
improved.
Hannah et al. (1977) tested the effectiveness of a pilot plant utilizing coagulation/
filtration or excess lime treatment in removing Sb. The plant consisted of a rapid
mix designed for a capacity of 4 ppm, a flocculator, a sedimentation basin, and dual-
media filtration. Antimony was present in the influent at a concentration of 0.6 mg/
L. Hydrated lime was added at a dose of 415 mg/L and a pH of 11.5; ferric
chloride was added at a dose of 40 mg/L and a pH of 6.2; and alum was added at a
dose of 220 mg/L and a pH of 6.4. Excess lime treatment produced an Sb reduction
of 28 percent. The ferric chloride coagulation produced an antimony reduction of 65
percent, and the alum coagulation produced an Sb reduction of 62 percent.
Hannah et al. (1977) also reported the results of using GAC adsorption as a post-
treatment to the conventional coagulation/filtration mentioned above, following dual-
media filtration. Using the above So-containing effluents from the coagulation/
filtration process, two GAC columns, operated in parallel and designated as "old"
and "new," were tested. The "old" GAC had been in use for several months before
this evaluation was made. When the lime coagulation effluent was processed,
through the "old" GAC column and additional 36% of the antimony was removed for
a total of 64%, while the "new" GAC column removed an additional 24% for a total
of 52%. When the ferric chloride coagulation effluent was processed through the
"old" GAC column an additional 7% of the antimony was removed for a total of
72%, and the "new" column removed no additional antimony. When the alum
coagulation effluent was processed, through the "old" GAC column an additional
13% of the antimony was removed for a total of 75%, while the "new" GAC column
removed an additional'9%, for a total of 71%.
18
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Antimony Health Advisory April 1992.
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Antimony Health Advisory April
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Westrick, M.L. 1953. Physiologic responses attending administration of antimony, alone or with
simultaneous injections of thyroxin. Proc. Soc. Exp. Biol. Med. 82:56-60.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The Merck indexAn
encyclopedia of chemicals, drugs, and biologicals, 10th ed. Rahway, NJ: Merck and
Company, Inc., p. 715.
23
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EPA 0553
HX000027511 APnl-
BERYLLIUM
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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 understand-
ing 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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Beryllium April. 1992.
II. GENERAL INFORMATION AND PROPERTIES
Information is available on a variety of beryllium compounds that have a broad range of
solubilities in water. In the development of Health Advisories (HAs) for beryllium, data on the
more soluble beryllium salts are considered to be the most relevant for two reasons: first,
insoluble beryllium salts and ores are not considered likely to be present in public drinking
water systems; second, there is currently no evidence to suggest that insoluble beryllium salts or
ores can be absorbed by the gastrointestinal (GI) systems of animals or humans.
CAS Nos.
Beryllium - 7440-41-7
Beryllium chloride - 7787-47-5
Beryllium sulfate - 13510-49-1
Beryllium oxide - 1304-56-9
Structural Formulas
Beryllium carbonate: BeCO3
Beryllium ortho-phosphate: Bej(PO4)2
Beryllium chloride: BeClj
Beryllium oxide: BeO
Beryllium sulfate: BeSO4
Synonyms
Glucinium. Precious forms of beryl: emerald, aquamarine.
Uses
Beryllium is used in high-performance products in metallurgical, aerospace and
nuclear technologies because of its unique combination of properties, such as an
unusually high melting point, high modulus of elasticity, extreme hardness, low
coefficient of thermal expansion and a high stifmess-to-weight ratio. Also, because
beryllium has a low atomic weight, it is highly permeable to X-rays, and thin sheets
are commonly used as windows for X-ray tubes (U.S. EPA, 1988).
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Beryllium
April. 199Z
Properties (Windholz et al., 1976; Weast et al., 1986)
The properties of inorganic beryllium compounds vary with the specific compound;
some examples follow:
Beryllium
Sulfate
Beryllium
Chloride
Beryllium
Oxide
BeSO4
105.07
Crystals
550-600ฐC
2.443
BeCl2
79.93
Whitish
crystals
482.3 ฐC
399.2ฐC
1.90
BeO
25.01
Light
amorphous
powder
2,530ฐC
Beryllium
Carbonate
BeCO,
69
White
powder
Very
soluble
0.2 mg/L (30ฐC)
Chemical Formula
Molecular Weight
Physical State (at 25ฐC)
Boiling Point (at 25 mm Hg)
Melting Point
Density
Vapor Pressure
Water Solubility
Octanol/Water Partition
Coefficient (log K^)
Taste Threshold (water)
Odor Threshold (water)
Odor Threshold (air)
Conversion Factor
Occurrence
Beryllium is a naturally occurring element found in the earth's crust at an average
concentration of 2.5 pprn. Beryllium is found chiefly as the minerals beryl
(Be,Al2Si6Ol8), bromellite (BeO), chrysoberyl (BeAljO4) and beryilonite (NaBePO4)
(Weast et al., 1986).
Many common beryllium compounds (for example, the chloride and nitrate) are
readily soluble in water. Others, such as the sulfate complex, are only moderately
soluble, and the carbonate and hydroxide compounds are almost insoluble in cold
water. Ionic beryllium is not likely to be found in natural waters except in trace
amounts, because, in the normal pH range of these waters, the oxides and
hydroxides formed at pH 5-8 are relatively insoluble (U.S. EPA, 1988).
In general, beryllium concentrations are well below 1 jtg/L in surface, ground, and
rain waters (Callahan et al., 1979),
* Described as insoluble in cold water and decomposing in hot water (Weast et al., 1986).
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Beryllium April, 1992.
Environmental Fate
There are currently no data available on the environmental fate of beryllium.
Callahan et al. (1979) have predicted that because beryllium salts commonly found
in natural waters have very low aqueous solubilities, they are probably precipitated
or adsorbed onto solids soon after introduction to the aqueous environment.
Complexing agents may solubilize beryllium into its ionic form, but ambient water
quality data suggest that concentrations of this element in heavily polluted waters
are quite low. Thus, beryllium in natural water systems is found predominantly in
paniculate rather than dissolved form.
III. PHARMACOKINETICS
Absorption
The available absorption data indicate that the uptake of orally administered
beryllium sulfate is extremely poor. Water solubility data on this compound
indicate that it is relatively insoluble at lower water temperatures, but decomposes
to BeSO4.4HjO at higher water temperatures. Thus, unless stomach acids enhance
dissolution of beryllium sulfate, it is unlikely that a significant concentration of
beryllium ion would reach the GI tract.
Single oral radiolabeled doses of beryllium were administered as a single dose to
mice, rats, monkeys and dogs (Furchner et al., 1973). Based on a weighted average
of the 2-day cumulative urinary excretion data, absorption of beryllium was
estimated to be 0.6%.
In studies of guinea pigs fed 10 or 30 rag/day of beryllium sulfate (approximately
13.3 or 40 mg/kg/day, based on the assumptions of Lehman, 1959), the amount of
beryllium absorbed was 0.006% of that ingested (approximately 0.08 or 0.24
mg/kg/day) (Hyslop et al., 1943).
Reeves (1965) added 6.6 or 66.6 /ig/day of beryllium as beryllium sulfate to the
drinking water of male Sprague-Dawley rats for 24 weeks. Of the total amount of
beryllium administered during the study, more than 99% of that recovered was
found in the feces. Total body uptake was estimated to be less than 1% for both
groups, with most of this found in the skeleton at levels that were independent of
dose. The author speculated that most of the beryllium in intestinal Quid was
probably precipitated as a phosphate, which would account for the observed low
levels of absorption and uptake observed.
Exposure of 160- to 180-gram Fischer 344 rats to 447 /*g/raj of beryllium oxide at a
concentration of 447 /ig/raj for 1 hour resulted in the incorporation of 0.2 /xg of
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Beryllium April, 199/L
beryllium into the lung tissue (Hart et al., 1984). Assuming a respiration rate of
0.0057 m3/hour, the percent uptake by inhalation equals 0.27(0.0057 m3 x 447 /ig/ra3),
or 7.8%.
Distribution
In studies of orally administered beryllium compounds, the organs that retained
significant amounts of beryllium were the skeleton, liver and kidney. When Reeves
(1965) administered beryllium sulfate in the drinking water of rats for 24 weeks,
beryllium levels were highest in the GI tract and the skeleton, with somewhat lower
levels in the blood and the liver.
In vitro studies have indicated that when beryllium sulfate solutions are incubated
with artificial or natural human serum, beryllium orthophosphate and hydroxide
precipitates are found in these fluids (Reeves and Vorwald, 1961).
Metabolism
Excretion
No information was found in the available literature on the metaboh'sm of
beryllium.
In rats, the route of administration influences the route of excretion of beryllium.
Fecal excretion is the major route if beryllium is administered orally or via
inhalation. Urinary excretion is the major route following intravenous
administration.
Approximately 80% and 76%, respectively, of a 6.6 or 66.6 /xg/day oral dose of
beryllium (administered via drinking water as beryllium sulfate) was recovered in
the feces of rats during a 24-week study, this was more than 99% of the total
recovered dose. Less than 1% was recovered from the urine at either dose
(Reeves, 1965).
Rhoads and Sanders (1985) showed that nearly all of the beryllium cleared from the
lungs of rats administered beryllium oxide via inhalation was excreted in the feces.
Urinary excretion was the major route following parenteral administration of
beryllium as beryllium chloride (very water soluble compound) in rats (Furchner et
al., 1973).
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Beryllium April, 1992.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
No information was found in the available literature on the human health effects of
short-terra oral exposure to beryllium compounds.
Long-term Exposure
No information was found in the available literature on the human health effects of
longer-term oral exposure to beryllium compounds.
Carcinoeenicitv
Evidence of the carcinogenic'potential of beryllium and its compounds in humans is
considered to be limited, whereas, in animals there is sufficient evidence of
carcinogenicity (IARC, 1987).
Animals
Short-term Exposure
In general, beryllium compounds administered orally produce less acute toxicity in
animals than those administered by other routes, probably because beryllium salts
are poorly absorbed from the GI tract (Schroeder and Mitchener, 1975b).
One of the earliest observed effects of short-term exposure to beryllium was the
development of rickets in young rats fed diets containing beryllium carbonate
(Guyatt et al., 1933; Businco, 1940).
Guyatt et al. (1933) reported that 21- to 24-day-old rats fed diets containing
beryllium carbonate at five dosage levels from 0.125% to 2.0% developed rickets
after 3 weeks. The effects were dose dependent, with the lowest dose (1.25 g/kg of
BeCO, diet, 163 rag/kg of Be diet or approximately 163 rag/kg/day of Be) resulting
in a mild case of rickets, while higher doses (2.0% or approximately 261 rng/kg/day
of Be) resulted in almost a complete lack of calcification of the long bones (the
femur and the tibia).
Businco (1940) conducted a series of experiments in which young rats (strain not
specified) were fed beryllium carbonate mixed in milk at doses of 0.06 g/day on days
0 to 14; 0.16 g/day on days 15 to 34 and 0.24 g/day on days 35 to 83. A
Time-Weighted Average (TWA) dose of 0.19 g/day (about 700 mg/kg/day of Be)
was estimated by the U.S. EPA (1988) based on the data provided by the author.
No effects on body weight or general appearance were observed when animals were
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Beryllium Aprili
fed 0.06 g/day of beryllium carbonate (about 260 rag/kg/day of Be). After 40 days
of exposure, a 25% weight loss was observed in treated rats compared with controls.
A reduction in weight of greater than 50% was seen in treated animals at the end
of the study (day 83). In addition, histologjcal and radiographic examination of the
long bones (femur, tibia, and fibula) and vertebrae revealed typical rachitic lesions.
Dermal/Ocular Effects
No data have been located on the dermal or ocular effects of exposure to beryllium
compounds.
Long-term Exposure
In the limited information available on the toxic effects of long-term oral exposure
to beryllium sulfate, there is no clear evidence of adverse effects other than slight
depressions in body weight.
Schroeder and Mitchener (1975a) administered 5 mg/L of beryllium as beryllium
sulfate in the drinking water of Long-Evans rats (52 per sex) until natural death.
Based on data provided in the study, this level corresponds to approximately 0.538
mg/kg/day of Be. Gross and microscopic pathological changes were evaluated and
clinical chemistry and urine analyses were performed. No treatment-related effects
were observed in any parameter tested. There was a slight depression in growth of
male rats from 2 to 6 months of age. In a similar study by Schroeder and
Mitchener (1975b), 54 male and 54 female Charles River (CD) mice were
administered about 1 mg/kg/day of beryllium as beryllium sulfate in their drinking
water. A slight decrease in body weight in females 6 to 8 months of age (p <
0.025, p < 0.01, respectively) and a slight general increase in male body weight were
noted. No other treatment-related effects were found.
Cox et al. (1975) fed rats 0, 5, 50 or 500 ppm of beryllium sulfate in the diet for 2
years. Based on the assumptions of Lehman (1959), these dietary levels are
equivalent to dosage levels of approximately 0, 0.25, 2.5 and 25 mg/kg/day of Be. A
slight decrease in body weight in the high-dose group was reported.
Reproductive Effects
No information was found in the available literature about the reproductive effects
of oral, dermal or inhalation exposure to beryllium.
Developmental Effects
No information was found in the available literature on the developmental effects
of oral exposure to beryllium.
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Beryllium April, 1992,
Tsujii and Hoshishima (1979) observed mild neurotoric effects in the offspring of
pregnant CFW mice administered 140 rag/day of beryllium sulfate (approximately
0.0047 rng/kg/day of Be) by intraperitoneal injection. The mice were given this dose
for 3 consecutive days and eight times every other day for a total of 11 days during
pregnancy.
Mutagcnicity
Simmon (1979a) reported beryllium sulfate to be negative in rautagenic response in
Salmonella typhimurium strains TA1535, TA1536, TA1537, TA98 and TA100, with
and without S-9 metabolic activation. Rosenkranz and Poirier (1979) also reported
negative results in this assay (the Ames test).
Beryllium sulfate was also negative when tested for mitotic recombination activity in
Saccharomyces cereviseae D3 (Simmon, 1979b).
Simmon et al. (1979) reported that beryllium sulfate was not rautagenic in the
host-mediated assay with S. typhimurium strains TA1530, TA1535 and TA1538 and
Saccharomyces cerevisiae D3. Mice were injected intramuscularly with 25 rag/kg
beryllium sulfate or administered an oral dose of 1,200 rag of beryllium sulfate.
Four hours after the treatment, microorganisms were recovered from the peritoneal
cavity and plated for mutant colonies. Mutation frequencies were not significantly
increased for either strain.
Larramendy et al. (1981) demonstrated that beryllium sulfate was clastogenic in
cultured human lymphocytes. This study showed a six-fold increase in aberration
frequency, which was primarily due to an increase in DNA breaks. This study also
demonstrated the potential of beryllium sulfate to induce sister-chromatid
exchanges (SCEs) in both cultured human lymphocytes and Syrian hamster embryo
cells.
Kanematsu et al. (1980) found beryllium sulfate to be weakly rautagenic in the
recombination assay. Inhibition of growth because of DNA damage was observed
in Bacillus subtilis strains H17 (rec*) and M45 (rec). Similar results were also
obtained by Kada et al. (1980).
Rosenkranz and Poirier (1979) reported negative results in the Pol assay, which
measured the ability of cells deficient in their ability to repair DNA damage to
grow after exposure to beryllium sulfate (250 /tg). Escherichia coli strains pol A*
zndpol A* were used in this study, both in the presence and absence of an S-9
activation system.
Beryllium sulfate was reported to produce negative results in the DNA repair test
using rat primary hepatocyte cultures (Williams et al., 1982) and in the mitotic
recombination assay using the yeast Saccharomyces cerevisiae D, (Simmon, 1979b).
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Beryllium April, 199JL
Kubinski et al. (1981) reported that ionic beryllium induced the formation of DNA
protein complexes (adducts) when E. coli cells and Ehrlich ascites cells were treated
with radioactive DNA in the presence of 30 rnM beryllium.
Skilleter et al. (1983), using synchronized rat liver-derived epithelial cell cultures
(BL9L), found that beryllium sulfate blocked the cell cycle at the Gv phase and
caused inhibition of cell division.
Beryllium ion (2+) has been reported to increase the raisincorporation of
nucleotides during DNA polymerization (Luke et al., 1975). Beryllium chloride
increased the error frequency in the incorporation of nucleotide bases into DNA
(Sirover and Loeb, 1976). Nucleotide base substitutions induced by beryllium
chloride were found to be two- to three-fold greater than those found
spontaneously in the lacl system (Zakour and Glickman, 1984).
Carcinogenicitv
Beryllium compounds administered by injection or inhalation can induce malignant
tumors in laboratory animals. The two types of cancers observed using these routes
were lung cancer and osteosarcoraa.
Gardner and Heslington (1946) reported induction of osteosarcoma following
intravenous injections of zinc beryllium oxide into rabbits. (Zinc oxide, zinc silicate
and silicic acid were all 'noncarcinogenic by this route.) Since then, numerous other
studies (e.g., Cloudman et al., 1949; Nash, 1950; Dutra and Largent, 1950; Barnes et
al., 1950; Hoagland et al., 1950) have demonstrated that many different beryllium
compounds, including beryllium metal, are tumorigenic when administered
intravenously.
Barnes et al. (1950) injected beryllium into the ear veins of 4-kg rabbits twice
weekly for 5 weeks. The beryllium was administered as an aqueous suspension of
particles of 5 fi or less in diameter of zinc beryllium silicate (an ore of questionable
solubility); the suspension contained a total of 72 mg of beryllium. The TWA daily
dose over the 120-week period of the experiment was 0.0021 mg/kg/day of Be. No
adjustment was made for a less than lifetime observation period because other
studies have indicated that osteosarcomas almost always develop within 2 years of
exposure (the lifespan of the test animals was 6 years). Four of nine animals
surviving 32 weeks or longer developed bone tumors. Osteosarcomas did not
develop in animals injected with zinc silicate.
The only study reporting the development of osteosarcoma following inhalation of
beryllium was that of Dutra et al. (1951). In this study, one of six rabbits exposed
to an aerosol of beryllium oxide at 6 mg/mj for 25 hours/week for 13 months
developed malignancies,
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Beryllium April, 1992,
Pulmonary tumors were produced in 18 of 19 rats that survived a 17-month
exposure to 15 rag/mj of beryl ore (Wagner et ah, 1969).
Reeves et al. (1967) exposed 150 rats to beryllium sulfate in an aerosol at a
beryllium concentration of 3425 /ig/m3 for 35 hours/week for a 72-week period.
The initial response appeared 4 weeks after the first exposure and consisted of
hyperplasia of the pulmonary epithelium, which progressed to metaplasia, anaplasia
and lung cancer. The first tumors were discovered after 9 months of exposure, and
the incidence of tumors was 100% at 13 months. All of the tumors were alveolar
adenocarcinomas.
Schroeder and Mitchener (1975a,b) reported a slight, statistically nonsignificant
increase in the incidence of lymphoma leukernias in female mice and a slightly
higher but still nonsignificant increase in the incidence of grossly observed tumors
in male rats given beryllium sulfate at a concentration of 5 mg/L (approximately
0.538 mg/kg/day) in drinking water over a lifetime.
In an unpublished 2-year study which has not been peer reviewed, Cox et al. (1975)
administered beryllium sulfate at Be dietary levels of 0, 5, 50 and 5dO ppm
beryllium (as beryllium sulfate) in the diet to Wistar albino rats (50/sex/group).
(These levels are equivalent to approximately 0, 025, 2.5 and 25 rag/kg bw/day
based on the dietary assumptions of Lehman, 1959). Mild body-weight depression
. was observed at the highest dosage level. Reticulum cell sarcomas in the lung were
seen in all dose groups and in controls, and similar lesions were seen in lymph
nodes, bone marrow and abdominal organs. The incidence of lung reticulum cell
sarcoma was higher in males than females, and was statistically significant in males
at the lowest two doses but not at the highest dose (tumor incidences were 10, 17,
16 and 12 at 0, 5, 50 and 500 ppra, respectively). This study is considered to be
suggestive of a carcinogenic response to ingested beryllium, but the lack of a
statistically significant response at the highest dose level severely limits its
interpretation as a positive study. In addition, a more recent abstract (Morgareidge
et al., 1977) based on the same study concludes that this dietary regimen had no
effect on tumorigenesis.
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:
(NOAEL or LOAEL^ x (EW) = rag/L (
(UF)(_Uday)
10
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Beryllium April, 199Z
where:
NOAEL or LOAEL = No- or Lowest-Observed-Aciverse-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/OW guidelines.
L/day = assumed daily water consumption of a child (1 L/day)
or an adult (2 L/day).
One-day Health Advisory
Studies containing exposure duration data appropriate for the calculation of a One-day
HA could not be located in the available literature. It is recommended, therefore, that the
Ten-day HA of 30 mg/L be used as an estimate of exposure for a One-day HA for a 10-kg child.
Ten-day Health Advisory
The study by Businco (1940) has been selected to serve as the basis of the Ten-day HA
for the 10-kg child because it was conducted for a more appropriate exposure duration than
other available studies using the oral route. In this study, 0.06 g of beryllium carbonate (0.26
g/kg of Be) was fed to 30- to 40-g young rats for 14 days. This resulted in no adverse effects on
body weight or general appearance. No other toxicity end points were measured for this time
period or dose; however, animals fed higher doses showed a dose-related decrease in the rate of
body-weight gain and a decrease in calcification and development of long bones. A No-
Observed-Adverse-Effect-Level (NOAEL) for beryllium of 260 mg/kg/day was established based
on the absence of adverse effects on body weight and gross toxicological effects in this study.
The Ten-day HA for the 10-kg child is calculated as follows:
Ten-day HA = (260 mg/kg/dav> (10 k$ = 26 nig/L (rounded to 30,000 /tg/L)
(100) (1 L/day)
where:
260 mg/kg/day = NOAEL, based on the absence of adverse effects on
body weight gain and bone development in rats fed
beryllium carbonate for 2 weeks (Businco, 1940).
10 kg = assumed body weight of a child.
11
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Beryllium April, 1992.
100 = uncertainty factor chosen in accordance with EPA and
NAS/OW guidelines in which a NOAEL from an
animal study is employed.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The study by Businco (1940) was selected to serve as the basis for the Longer-term HA
because of its exposure duration and because it provides information on an important toxicity
end point for beryllium exposure (i.e., the development of rickets). In this study, young rats (30
to 40 g) were fed increasing doses of beryllium carbonate for 83 days. Rats received daily doses
of 60 mg of beryllium carbonate (7.8 rag Be) on days 0 to 14, 160 mg (20.8 rag Be) on days 15
to 34 and 240 mg (31.2 rag Be) on days 35 to 83. The estimated average body weight of the
treated animals was 55.5 g. A TWA dose of 0.191 g/rat/day (3.4 g/kg) of beryllium carbonate
can be estimated based on the information provided by the authors. The 3.4 g/kg-dose (443 mg
Be) resulted in a greater than 50% reduction in body weight gain and decreased development
and calcification of the long bones. Therefore, 443 mg/kg/day of beryllium is considered to be
the Lowest-Observed-Adverse-Effect Level (LOAEL). No other adverse effects were reported
in this study.
No other available study was found to be more appropriate for the calculation of the
Longer-term HA. Therefore, the LOAEL of 443 mg/kg/day from the Businco (1940) study is
recommended for the calculation of the Longer-term HA.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-terra HA = C443 me/ke/dav') (10 kg) = 4.4 rag/L (rounded to 4,000 /tg/L)
(1,000) (1 L/day)
where:
443 rag/kg/day = LOAEL, based on suppressed body weight gain and
development of rickets in rats fed BeCO, for 83 days
(Businco, 1940).
10 kg = assumed body weight of a child.
1,000 = uncertainty factor chosen in accordance with EPA and
NAS/OW guidelines in which a LOAEL from an animal
study is employed.
1 L/day = assumed daily water consumption of a child.
12
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Beryllium April, 1992.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = C443 mg/kg/day") (70 kg) = 15-5 mg/L (rounded to 20,000
(1,000) (2 L/day)
where:
443 mg/kg/day = LOAEL, based on suppressed body weight gain and
development of rickets in rats fed BeCOj for 83 days
(Businco, 1940).
70 kg = assumed body weight of an adult.
1,000 = uncertainty factor chosen in accordance with EPA and
NAS/OW guidelines in which a LOAEL from an animal
study is employed.
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 estimate of a daily exposure level 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 known, probable or possible human carcinogen,
according to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then
caution must be exercised in making a decision on how to deal with possible lifetime exposure
to this substance. For human (A) or probable human (B) carcinogens, a Lifetime HA is not
recommended. For possible human carcinogens (C), an additional 10-fold safety factor is used
in the calculation of the Lifetime HA. The risk manager must balance this assessment of
carcinogenic potential and the quality of the data against the likelihood of occurrence and
13
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Beryllium April, 199&
significance of health effects related-to noncarcinogenic end points of toxicity. To assist the risk
manager in this process, drinking water concentrations associated with estimated excess cancer
risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of water per
day are provided in the Evaluation of Carcinogenic Potential section.
The study by Schroeder and Mitchener (1975a) has been selected to serve as the basis
for the Lifetime HA because currently it is the only available published study using an
appropriate route and exposure duration. In this study, male and female rats were administered
5 ppm of beryllium in their drinking water for a lifetime. The only significant effect was a slight
reduction in body weight in males from 2 to 6 months of age. A NOAEL of 0.538 rng/kg/day
for beryllium was calculated by multiplying the 5-ppra dose (5 rag/L) by an average water
consumption rate of 0.035 L/day for rats and dividing the product by the average rat body
weight (0.325 g) given in this study. It should be noted, however, that several weaknesses have
been identified in this study, such as the presence of other trace elements and minerals
(including chromium) in the drinking water, the use of nonrandornized animals and the
administration of only one dose.
The unpublished study of Cox et al. (1975) (also reported in an abstract by Morgareidge
et al., 1977) could also be used to calculate the reference dose (RfD) for beryllium. In this
study, rats were exposed to the 0, 5, 50 or 500 ppm (approximately 0, 025, 2.5 or 25 rag/kg/day)
of beryllium in the diet for 2 years. The only toxic effect confirmed was a slight decrease in
body weight in the highest dose group. A NOAEL of 25 mg/kg/day for beryllium can be
identified based on the results of this study. Since this study is unpublished and presumably has
not been peer reviewed, it cannot be used in the calculation of a Drinking Water Equivalent
Level (DWEL) for beryllium. In addition, there is some question about the finding of
pulmonary reticulum cell sarcoma at the two lowest doses. However, until a more detailed
review of the Cox et al. (1975) study is conducted, the Agency-verified RfD (U.S. EPA, 1988)
based on the Schroeder and Mitchener (1975a) study is recommended for use in calculating the
DWEL.
Step 1: Determination of the RfD
RfD = (0.538 mg/kg/dav) = 0.005 mg/kg/day
(100)
where:
0.538 mg/kg/day = adjusted NOAEL, based on the absence of effects in rats given
BeSO4 in drinking water over a lifetime (Schroeder and Mitchener,
1975a).
100 = uncertainty factor chosen in accordance with EPA and NAS/OW
guidelines in which a NOAEL from an animal study is employed.
14
-------
Beryllium April, 199jL
Step 2: Determination of the DWEL
DWEL = C0.005 me/kg/dav) (70 kg) = 0.175 rng/L (rounded to 0.2 rag/L)*
(2 Uday)
where:
0.005 rag/kg/day = RฃD.
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
Beryllium is classified by the U.S. EPA as Group B2: probable human carcinogen.
Therefore, the determination of a Lifetime HA is not recommended.
The estimated excess cancer risk associated with lifetime exposure to drinking water
containing 175 /ig/L beryllium is approximately 2.5 x 10"2. This estimate represents the upper
95% confidence limit from extrapolations prepared by U.S. EPA's Human Health Assessment
Group using the linearized multistage model. The actual risk is unlikely to exceed this value.
Evaluation of Carcinogenic Potential
IARC (1987) has classified beryllium and beryllium compounds in Group 2A:
probable human carcinogen. This group includes chemicals for which there is
limited evidence of carcinogenicity in humans and sufficient evidence in animals.
Using the U.S. EPA (1986) Guidelines for Carcinogen Risk Assessment, beryllium
is classified as a Group B2 carcinogen. This classification is used for compounds in
which there is sufficient evidence of carcinogenicity in animal studies and
inadequate evidence in human studies.
The U.S. EPA (1987) has calculated an inhalation carcinogenic potency factor for
beryllium of 2.4 x 10"J (/ig/m1)"1 based on the epidemiological study by Wagoner et
al. (1980) and the industrial hygiene reviews by NIOSH (1972) and Eisenbud and
Lisson (1983). These data have been combined arithmetically to estimate a
plausible upper-bound incremental lifetime cancer risk of 2 x 10"J for exposure to
air containing 1 pg/m} of beryllium.
Because beryllium administered via inhalation and injection is carcinogenic in
animals, it may be carcinogenic when administered orally as well. Thus, oral
?This is the DWEL based on the verified RfD; however, as discussed previously, it may be
appropriate to the RfD.
15
-------
Beryllium April, 199Z
exposure to beryllium in drinking water may represent a carcinogenic risk to
humans.
The U.S. EPA (1988) has derived an oral carcinogenic potency factor for beryllium .
based on the results of intravenous infusion studies in which osteosarcomas were
induced in rabbits. Based on these data, a human carcinogenic potency factor of
1,843 (mg/kg/day)'1 for intravenous infusion of beryllium was identified. After
adjusting for an oral absorption efficiency of 0.6% (Furchner et al. 1973), an oral
potency factor of 11 (mg/kg/day)'1 is obtained (U.S. EPA, 1988).
An oral carcinogenic potency estimate can be derived by extrapolation or the
inhalation slope factor based on human data, after adjusting for relative absorption
efficiency (McGinnis, 1988). Absorption from the GI tract has been reported to be
approximately 0.6% (Furchner et al. 1973), and absorption from the lungs was
approximately 7.8% (Hart et al. 1984). The relative difference in absorption
efficiency is 0.6/7.8, or 0.077. Therefore, the extrapolation-based oral carcinogenic
potency estimate equals 2.4 (rag/ra3)'1 x 70 kg/20rn3/day x 0.077, or 0.65
(mg/kg/day)'1.
If the geometric mean of the slope factors based on animal inhalation studies (U.S.
EPA, 1987) is adjusted without surface-area correction for a 7Q-kg person inhaling
20 raVday, and a relative absorption factor between the inhalation and oral routes
of 0.77 is applied, an oral carcinogenic potency factor of 12.7 rag/kg/day'1 is derived
(McGinnis, 1988).
A carcinogenic potency estimate of 43 (rag/kg/day)'1, based on extrapolations from
the Schroeder and Mitchener (197Sa) drinking water study in rats, has been verified
by the Agency-wide CRAVE workgroup (U.S. EPA, 1991), despite the lack of a
significant tumorigenic response. Since no significant response was detected, this
estimate is an upper-bound value; that is, the risk is not expected to be greater, but
may be less, than the derived value. The limited evidence of carcinogenicity in the
Cox et al. (1975) study provides further support for this conclusion. When a
linearized multistage model is applied to the Schroeder and Mitchener (1975a)
animal data, the resulting criteria are 0.8, 0.08 and 0.008 /tg/L, with corresponding
carcinogenic risk levels of 10"*, 10"5 and 10"*, respectively, for a 70-kg man
consuming 2 L/day of drinking water.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The World Health Organization (WHO) has not established a guideline for
drinking water quality for beryllium (IRPTC, 1987).
National regulation by OSHA (1985) established a Permissible Exposure Limit
(PEL) 8-hour TWA of 2 /xg/rn1, an acceptable ceiling limit of 5 pg/m1 and an
acceptable maximum peak above ceiling of 25 jtg/raj for 30 minutes.
IS
-------
Beryllium April, 1993,
Advisories issued by various agencies for beryllium in the air are as follows:
NIOSH (1972) recommended an occupational exposure limit of 0.5 /zg/ra3; ACGIH
(1987) recommended a TWA-Threshold Limit Value (TWA-TLV) of 0.002 mg/m3.
The U.S. EPA (1980) established ambient water quality criteria of 68 ng/L for the
consumption of 2 L of ambient water and fish, and 1,170 ng/L for the consumption
of aquatic organisms, to correspond with a risk level of 10"s.
VII. ANALYTICAL METHODS
Methods available to analyze beryllium are summarized below. The bulk of the
methods available for metal analysis involve spectroscopy, either emission or
absorption. In all of these methods, the metal is dissolved and thermally excited.
When excited, all elements emit or absorb light frequencies characteristic of that
element. Most metal spectroscopy is done in the ultraviolet and x-ray regions.
Direct Aspiration Atomic Absorption Spectroscopv CAA). In this technique (EPA
Method 210.1), dissolved metals are aspirated into a flame source and excited to
the point of dispersion into a mono-atomic state; a light source whose cathode is
the metal of interest passes through the flame, and the resulting absorption of light
by the element of interest is directly proportional to its concentration.
Disadvantages of this technique include the inability to analyze more than one
metal at a time and the insensitiviry of the technique. The detection limit is 0.25
Graphite Furnace Atomic Absorption (GFAA). This technique differs from AA, in
that a specific amount of liquid is dried on the thermal source, effecting a
concentration step. The sample is then electrothermally excited. This technique is
very sensitive, but is still a tedious one-raetal-at-a-time determination.
Inductively Coupled Plasma Atomic Emission Spectroscopv. This method utilizes
aspiration of a liquid sample, but the flame is actually a plasma torch of Argon
excited to super hot levels by radio-frequency radiation. In this technique, the
metals are excited sufficiently to emit radiation. By using classical dispersion-
grating optics, a large number of metals can be analyzed simultaneously. Currently,
the utility of this technique is limited by lack of sensitivity for drinking water
Maximum Contaminant Levels (MCLs), but sample preparation techniques under
development that will overcome this problem should be available after March of
1989.
Inductively Coupled Plasma Mass Spectrometrv (TCP/MS). This technique (EPA
Method 602.0) is the most expensive of those listed in this section, but it may be
the most effective and cost-effective if used for all U.S. EPA programs seeking
lower detection levels and/or MCLs. In this technique excitation is effected using a
radio-frequency plasma, but the excited atoms are interfaced into a mass
spectrometer. Quantitation is achieved by computerized software programs similar
17
-------
Beryllium April, 199.2.
to those used in other U.S. EPA MS organic methods. The main problem with
TCP/MS is metal deposition on the interface, resulting in giving the possibility of
"memory" when an extremely low-concentration sample is run after a high-
concentration sample. However, normal analytical steps can be used to resolve this
problem. The detection limit of TCP/MS is 0.1 /xg/L.
VIII. TREATMENT TECHNOLOGIES
The available information indicates that reverse osmosis (RO) systems and
conventional coagulation/filtration will remove beryllium in drinking water.
Fox and Sorg (1987) described the effectiveness of using an RO unit as a
point-of-use device for removing beryllium. A laboratory-size RO unit containing a
spiral-wound polyamide membrane was operated at a pressure of 42 ฑ 2 psi. This
system removed 97.7% of the beryllium from a 0.043-rag/L influent.
Hannah et al. (1977) tested the effectiveness of a pilot plant utilizing
coagulation/filtration or excess lime treatment in removing beryllium. The plant
consisted of a rapid mix- designed for a capacity of 4 gpm, a flocculater, a
sedimentation basin and dual-media filtration. Beryllium was present in the
influent at a concentration of 0.1 rng/L. Hydrated lime was added at a dose of 415
rag/L and a pH of 11.5; ferric chloride was added at a dose of 40 rag/L and a pH of
62', and alum was added at a dose of 220 rag/L and a pH of 6.4. Excess lime
treatment reduced the concentration of beryllium by 99.4%. Ferric chloride
coagulation reduced the beryllium concentration by 94%, while alum coagulation
reduced the beryllium concentration by 98.1%.
Hannah et al. (1977) also reported the results of using granular-activated carbon
(GAC) adsorption as a post-treatment to the conventional coagulation/filtration
mentioned above. Using the beryllium-containing.effluents from the
coagulation/filtration process above, two GAC columns, operated in parallel and
designated as "old" and "new," were tested. The "old" GAC was in use for several
months before an evaluation of its performance was made. When the lime
coagulation effluent was processed through the "new" or the "old" GAC columns, an
additional 0.1% of the beryllium was removed, for a total of 99.5%. When the
ferric chloride coagulation effluent was processed through the "old" GAC column,
an additional 1.4% of the beryllium was removed, for a total of 98.7%; the "new"
GAC column removed an additional 4.9% for a total of 98.9%. When the alum
coagulation effluent was processed through the "old" or the "new" GAC column an
additional 0.8% of the beryllium was removed, for a total of 98.9%.
18
-------
Beryllium April, 1992,
IX. REFERENCES
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limit values and biological exposure indices for 1987-1988. Cincinnati, OH: ACHIH,
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Chemical Co., Inc., Milwaukee, WI.
Barnes, J.M., F.A. Denz and H.A. Sissons. 1950. Beryllium bone sarcomata in rabbits. Br. J.
Cancer 4:212-222.
Branion, H.D., B.L. Guyatt and H.D. Kay. 1931. Beryllium rickets. J. Biol. Chera. (Science
Proceedings XXV) 92:11.
Businco, L. 1940. The rickets-producing effect of beryllium carbonate. Ind. Med. Rev.
11:417-442. (In Italian; translation.)
Callahan, M.A., M.W. Slimak, N.W. Gabel, I.P. May, C.F. Fowler, J.R. Freed, P'. Jennings, R.L
Durfee, F.C. Whitmore, B. Maestri, W.R. Mabey, B.R. Holt, and C. Gould. 1979.
Water related environmental fate of 129 priority pollutants. Publication no. EPA
440/4-79-029. Washington, DC: U.S. Environmental Protection Agency.
Cloudraan, A.M., D. Vining, S. Barkulis, and JJ. Nickson. 1949. Bone changes observed
following intravenous injections of beryllium. Am. J. Pathol. 25:810-811.
Cox, G.E., D.E. Bailey and K. Morgareidge. Food and Drug Research Laboratories, Inc. 1975.
Chronic feeding studies with beryllium sulfate in rats. Draft final report. Pittsburgh,
PA. Aluminum Company of America.
Dutra, F.R. and EJ. Largent, 1950. Osteosarcoma induced by beryllium oxide. Am. J. Pathol.
26:197-209.
Dutra, F.R., EJ. Largent and J.L, Roth. 1951. Osteogenic sarcoma after inhalation of
beryllium oxide. Arch. Pathol. 51:473-479.
Eisenbud, M. and J. Lisson. 1983. Epidemiological aspects of beryllium-induced nonmalignant
lung disease: A 30-year update. J. Occup. Med. 25:196-202.
Fox, K.R. and TJ. Sorg. 1987. Controlling arsenic, fluoride and uranium by point-of-use
treatment J. AWWA. 79(10):81-84.
Furchner, I.E., C.R. Richmond and J.E. London. 1973. Comparativeraetabolism of
radionuclides in mammals. VIE. Retention of beryllium in the mouse, rat, monkey and
dog. Health Phys. 24:293-300.
19
-------
Beryllium April, 1992*
Gardner, L.U. and HJF. Heslington. 1946. Osteosarcorna from intravenous beryllium
compounds in rabbits. Fed. Proc. 5:221.
Guyatt, G.L., H.D. Kay and H.D. Branion. 1933. Beryllium rickets. J. Nutr. 6:313-324.
Hannah, S.A., M. Jelus and J.M. Cohen. 1977. Removal of uncommon trace metals by physical
and chemical treatment processes. J. Water Pollut. Cont. Fed. 49:2297-2309.
Hart, B.A., A.G. Harmsen, R.B. Low and R. Emerson. 1984. Biochemical, cytological and
histological alterations in rat lung following acute beryllium aerosol exposure. Toxicol.
Appl. Pharmacol. 75:454-465.
Hawley, G.G. 1981. The condensed chemical dictionary, 10th ed. New York, NY: Van
Nostrand Reinhold Co.
Hoagland, M.B., R.S. Grier, and M.B. Hood 1950. Beryllium and growth. I.
Beryllium-induced osteogenic sarcomata. Cancer Res. 10:629-635.
Hyslop, F., E.D. Palmes, W.C. Alford, A.R. Monaco and L.T. Fairhall. 1943. The toxicity of
beryllium. NIH Bull. no. 181. Washington DC: National Institutes of Health, p. 56.
I ARC. 1987. International Agency for Research on Cancer. IARC monographs on the
evaluation of carcinogenic risks to humans. Suppl. 7. Beryllium and beryllium
compounds. Lyon, France: IARC, pp. 127-128.
IRPTC. 1987. International Register of Potentially Toxic Chemicals. IRPTC data profile on
beryllium. Geneva, Switzerland: United Nations Environment Programme.
Kada, T., K. Hirano and Y. Shirasu. 1980. Environmental chemical mutagens by the Rec-assav
system with Bacillus subtilis. Chem. Mutagens. 6:149-173.
Kanematsu, N., M. Hara and T. Kada. 1980. Rec assay and mutageniciry studies on metal
compounds. Mutat. Res. 77:109-116.
Kay, H.D. and D.I. Skill 1934. CLXDC Beryllium rickets IL The prevention and cure of
beryllium rickets. Biochem. J. 28:1222-1227.
Kubinski, H., G.E. Gutzke and Z.O. Kubinski. 1981. DNA-cell-binding (DCB) assay for
suspected carcinogens and mutagens. Mutat. Res. 89:95-136.
Larraraendy, ML., N.C Popescu and J.A. diPaolo. 1981. Induction by inorganic metal salts of
sister chromatid exchanges and chromosome aberrations in human and Syrian hamster
cell strains. Environ. Mutagen. 3:597-606.
Lehman, A. 1959. Appraisal of the safety of chemicals in foods, drugs and cosmetics.
Association of Food and Drug Officials of the United States.
20
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Beryllium . April, 1992.
Luke, M.Z., L Hamilton and T.C. Hollocher. 1975. Beryllium-induced raisincorporation by a
DNA polymerase. Biochem. Biophys. Res. Comm. 62:497-501.
McGinnis, P. 1988. Review of beryllium file for CRAVE Summary, Task #29. Memorandum
to L. Papa, Environmental Criteria and Assessment Office, U.S. Environmental
Protection Agency. Cincinnati, Ohio. Syracuse Research Corporation, December 28
1988.
Morgareidge, K., G.E. Cox, D.E. Bailey and MA. Gallo. 1977. Chronic oral toxicity of
beryllium in the rat. Toxicol. Appl. Pharmacol. 41:204-205.
Nash, P. 1950. Experimental production of malignant tumors by beryllium. Lancet. 1:519.
NIOSH. 1972. National Institute for Occupational Safety and Health. Criteria for a
recommended standard occupational exposure to beryllium. Washington, DC:
Department of Health, Education and Welfare, p. 72-10269.
OSHA. 1985. Occupational Safety and Health Administration. Occupational standards and
permissible exposure limits. Code of Federal Regulations 29:1910.1000.'
Perry, R.H. and C.H. Chilton. 1973. Chemical engineers handbook, 5th ed. New York, NY:
McGraw Hill Book Co.
Reeves, A.L. 1965. The absorption of beryllium from the gastrointestinal tract. Arch. Environ.
Health 11:209-214.
Reeves, A.L., D. Deitch and AJ. Vorwald. 1967. Beryllium carcinogenesis. I. Inhalation
exposure of rats to beryllium sulfate aerosol. Cancer Res. 27:439-445.
Reeves, A.L. and AJ. Vorwald 1961. The humoral transport of beryllium. J. Occup. Med.
3:567-571.
Rhoads, K. and C.L. Sanders. 1985. Lung clearance, translocation and acute toxicity of arsenic,
beryllium, cadmium, cobalt, lead, selenium, vanadium and ytterbium oxides following
deposition in rat lung. Environ. Res. 36:359-378.
Rosenkranz, H.S. and LA. Poirier. 1979. Evaluation of the rautagenicity and DNA-modifying
activity of carcinogens and noncarcinogens in microbial systems. J. Natl. Cancer Inst.
62:873-892. '
Schroeder, H.A. and M. Mitchener. 1975a. Life-term studies in rats: Effects of aluminum,
barium, beryllium and tungsten. J. Nutr. 105:421-427.
Schroeder, H.A. and M Mitchener. 1975b. Life-term effects of mercury, methyl mercury and
nine other trace metals on mice. J. Nutr. 105:452-458.
21
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Beryllium April, 199,2.
Scott, J.K., W.F. Neuraan and R. Allen. 1950. The effect of added carrier on the distribution
and excretion of soluble 7Be. J. Biol. Chera. 182:291-298.
Simmon, V.F. 1979a. In vitro mutagenicity assays of chemical carcinogens and related
compounds with Salmonella typhimurium. J. Natl. Cancer Inst. 62:893-899.
Simmon, V.F. 1979b. In vitro assays for recornbinogenic activity of chemical carcinogens and
related compounds with Saccharomyces cerevisiae D3. J. Natl. Cancer Inst. 62:901-909.
Simmon, V.F., H.S. Rosenkranz, E. Zeiger and LA. Poirier. 1979. Mutagenic activity of
chemical carcinogens and related compounds in the intraperitoneal host-mediated assay.
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Sirover, MA. and LA. Loeb. 1976. Metal-induced infidelity during DNA synthesis. Proc.
Natl. Acad. Sci. 73:2331-2335.
Skilleter, D.N., RJ. Price and R.F. Legg. 1983. Specific GrS phase cell block by beryllium as
demonstrated by cytofluorometric analysis. Biochern. J. 216:773-776.
Tsujii, H. and K. Hoshishiraa. 1979. The effect of the administration of trace amounts of
metals to pregnant mice upon the behavior and learning of their offspring. Shinshu
Daigaku Nogakuba Kiyo. 16:13-27.
U.S. EPA. 1991. U.S. Environmental Protection Agency. Integrated Risk Information System
(IRIS) online. Cincinnati, Ohio: U.S. EPA Office of Health and Environmental
Assessment, Environmental Criteria and Assessment Office.
U.S. EPA. 1988. U.S. Environmental Protection Agency. Drinking water criteria document for
beryllium. Draft. ECAO-CIN-0003. Cincinnati, OH; U.S. EPA Environmental Criteria
and Assessment Office.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Health assessment document for
beryllium. Document no. EPA 600/8-84/-026F. Research Triangle Park, NC: U.S. EPA
Office of Health and Environmental Assessment, Environmental Criteria and
Assessment Office.
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U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality criteria
document for beryllium. Prepared by the Office of Health and Environmental
Assessment, Environmental Criteria and Assessment Office. Document no. EPA
440/5-80-024. Washington, DC: Office of Water Regulations and Standards.
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environmental assessment. Vol. II. MEG charts and background information.
Document no. EPA 600/17-77-136b. Washington, DC: U.S. EPA.
Wagner, W.D., D.H. Groth, J.L. Holtz, G.E. Madden and HJE. Stokinger. 1969. Comparative
inhalation toxicity of beryllium ores, bertrandite and beryl, with production of
pulmonary tumors by beryl. Toxicol. Appl. Pharmacol. 15:10-29.
Wagoner, J.K., P.F. Infante and D.L. Bayliss. 1980. Beryllium: An etiologic agent in the
induction of lung cancer, nonneoplastic respiratory disease among industrially exposed
workers. Environ. Res. 21:15-34.
Weast, R.C. 1982. CRC handbook of chemistry and physics, 62nd ed. Boca Raton, FL: CRC
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1985-1986, 66th ed. Boca Raton, FI: CRC Press, Inc.
Williams, G.M., M.F. Laspia and V.C. Dunkel. 1982. Reliability of hepatocyte primary culture.
Mutat. Res. 97:359-370.
Windholz M., S. Budavari, L.Y. Strqumtsos and MJ*. Fertig, eds. 1976. The Merck index. An
encyclopedia of chemicals and drugs, 9* ed. Rahway, NJ: Merck and Co., Inc.
Zakour, R.A. and B.W. Glickman. 1984. Metal-induced mutagenesis in the lacl gene of
Escherichia coli. Mutat. Res. 126:9-18.
23
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EPA 0553
EX000027511
April 1992
BORON
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology, and treatment technology that
would be useful in dealing with the contamination of drinking water. Health Advisories (HAs)
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.
HAs 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 endpoints of toxicity. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or B), Lifetime
Health Advisories are not recommended. For substances with a carcinogenic potential,
chemical concentration values are correlated with carcinogenic risk estimates by employing a
cancer potency (unit risk) value together with assumptions for lifelong exposure and the
ingestion of water. The cancer unit risk is usually derived from a linearized multistage model
with 95% upper confidence limits providing a low-dose estimate of cancer risk. The cancer
risk is characterized as being an upper limit estimate, that is, the true risk to humans, while
not identifiable, isjiot likely to exceed the upper limit estimate and in fact may be lower.
While alternative risk modeling approaches may be presented, for example one-hit, Weibull,
logit, or probit, the range of risks described by using any of these models 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
carcinogenic risk, the Agency recommends using the linearized multistage model.
II. GENERAL INFORMATION AND PROPERTIES
CAS Nos.
Boron 7440-42-8
Boric acid 10043-35-3
Sodium tetraborate 1303-96-4
1
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Boron
April 1992
Structural Formula
Not applicable.
Synonyms
Boron: none
Boric acid: orthoboric acid
Sodium tetraborate (as the decahydrate): borax
Uses
Elemental boron (B) and its carbides are used in composite structural
materials, in high-temperature abrasives, in special purpose alloys and in steel-
making (Hawley, 1981). Boric acid and borates are used in glass manufacture,
cleaners, wood and leather preservatives, flame retardants, cosmetic products
and neutron absorbers for nuclear installations (Sittig, 1981). Borax is also
used as an insecticide for cockroaches and black carpet beetles (Windholz et
al., 1983). Boron halides are used as catalysts in the manufacture of
magnesium alloy products, metal refining, magnesium solder fluxes, rocket
fuels, and detoxification of nitrosamine-contaminated wastes (U.S. EPA, 1975).
Boron hydrides are used as reductants to control heavy metal discharges in
wastewater, as catalysts and in jet and rocket fuels (U.S. EPA, 1975). They are
also used as household insecticides.
Properties (Windholz et al., 1983)
Boron Boric acid
Borax
Chemical Formula
Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Specific Gravity
Log Octanol/Water
Partition
Coefficient (log
Taste Threshold
Odor Threshold
Conversion Factor
B
10.81
Solid
2,300ฐC
2.35
Insoluble
2.45
H3BO3
61.84
Solid
171ฐC
1 g/18 mL (cold)
1.43
3813
Solid
320ฐ
75ฐ (rapid heating)
1.73
1 g/16 mL (cold)
1.73
NA
NA
35.27
-------
Boron April 1992
Occurrence
The most common forms of boron in nature are boric acid and sodium
tetraborate.
Boron, a naturally occurring element, is found in soil at an average concentration
of 10 mg/kg (Weast, 1988), in ocean waters at a concentration of 4.6 mg/L
(Weast, 1988) and in freshwater at a concentration of 0.01 /ig/g (U.S. EPA,
1980).
The total daily boron intake in normal human diets has been reported to range
from 2.1 to 4.3 mg/day (Zook and Lehman, 1965) and from 1.3 to 4.4 mg/day
(Hamilton and Minski, 1972).
Boron found in coal, oil shale and geothermal fluids contributes to
environmental pollution (White, 1980). Excluding borate production, an
estimated 1,000 to 4,000 tons of boron are released to the environment each year
(U.S. EPA, 1975).
Municipal sludge and industrial wastewater contribute to the release of boron to
soil, ground water, and estuaries (Hemphill et al., 1981).
Environmental Fate
No information regarding the environmental fate of boron was located in the
literature.
III. PHARMACOKINETICS
Absorption
Jansen et al. (1984b) administered a single oral dose of 750 mg of boric acid to
each of six human volunteers (approximately 1.9 mg/kg of boron; average body
weight 70.8 kg). At least 93.9% of the dose was absorbed from the
gastrointestinal (GI) tract, as measured by the recovery of boric acid in urine
after 96 hours.
Kent and McCance (1941) orally administered a total of 352 mg of boron (as
boric acid) to each of two women over a 3-day period (approximately 2.35 mg/kg/
day, body weight 50 kg). At least 93 to 94% of the boric acid was absorbed from
the GI tract, as measured by recovery of boric acid in urine within a week.
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Boron April 1992
Draize and Kelley (1959) estimated that urinary excretion of boron accounted
for 50 to 66% of the quantities of boric acid (17.1 to 119.9 mg/kg/day of boron)
administered orally to rabbits on four consecutive days.
Draize and Kelley (1959) reported no increase in the urinary excretion of boron
by a human volunteer following a 4-hour exposure of the forearm to 15 g of
boric acid (approximately 37.5 mg/kg of boron, body weight 70 kg) wetted with
physiological saline. Results indicated that little or no dermal absorption of
boron occurred through the intact skin.
Mulinos et al. (1953) applied a commercial talcum powder containing 5% boric
acid to the skin of six infants with no (one infant) or varying degrees (mild, one
infant; moderate, two infants; marked, one infant) of diaper rash. Boric acid was
not detected in the urine of any of the infants before powder application.
Twenty-four hours after powder application, boric acid was detected in the urine
of the infants with moderate to marked rashes, and it persisted in the urine for
at least 48 hours. No boric acid was detected in the urine of infants with no rash
or mild diaper rash, indicating that little or no absorption occurred through the
intact skin. No quantitative data on absorption were presented in this study.
Vignec and Ellis (1954) applied a powder containing 5% boric acid 7 to 10 times
daily for at least 1 month to infants varying in age from 1.25 to 10 months. The
infants were exposed to a calculated dose of approximately 2.33 g boric acid/
infant/day. Boron concentrations in the blood and urine were determined
colorimetrically. In the test group containing 12 infants, blood boron levels
varied from 0.01 mg to 0.14 mg/100 mL (0.04 ฑ 0.05 mg/100 mL, mean ฑ SD),
and the urine levels varied from 0.02 mg to 0.44 mg/100 mL (0.16 ฑ 0.14 mg/100
mL, mean +. SD). Corresponding values in the 12 control infants were 0 to 0.19
mg/100 mL (0.10 ฑ 0.05 mg/100 mL, mean ฑ SD) for blood and 0 to 0.3 mg/100
mL (0.08 ฑ 0.07 mg/100 mL, mean SD) for urine. Twelve other infants tested
developed diaper rashes varying in severity from mild erythema to moderately
severe diaper rash with blood levels of boron ranging from 0 to 0.15 mg/100 mL
(0.03 ฑ 0.04 mg/100 mL, mean ฑ SD). Similar boron levels in the serum and
urine of control and test infants indicate.that no remarkable topical absorption
of.boric acid occurred in this study.
Draize and Kelley (1959) applied 5% aqueous boric acid (35 mg/kg of boron) to
the intact or damaged skin of rabbits for 1.5 hours a day for 4 days. Net urinary
excretion of boric acid amounted to 0.25, 1.3 or 3.7 mg/kg of boron, depending
on whether the application had been made to intact, abraded or burnt and
partially denuded skin, respectively. Application of a wet boric acid powder (700
mg/kg of boron) resulted in net urinary excretion of boric acid amounting to
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Boron April 1992
0.14, 0.98 or 22.2 mg/kg of boron, depending on whether the application had
been made to intact, abraded or burnt and partially denuded skin, respectively.
Distribution
Locksley and Sweet (1954) reported that borate (administered in borax)
concentrations in the tissues of normal mice were directly proportional to the
dosage of borax administered (1.8 to 71.0 mg/kg of boron) for at least 2 hours
after intraperitoneal injection. After equilibrium was achieved in
nephrectomized mice, about 90% of the borate was distributed uniformly
throughout the body water, and about 10% was bound in the intracellular
compartment.
Grella et al. (1976) described transplacental distribution of boric acid in humans.
A 34-week pregnant female accidentally swallowed 70 g of boric acid
(approximately 245 mg/kg of boron assuming a 50-kg body weight). A fetus
delivered 2 hours later by Cesarean section died shortly afterward from
cardiovascular failure.
Metabolism
No information on the metabolism of boron was found in the literature.
Excretion
Jansen et al. (1984b) administered single oral doses of boric acid (approximately
1.9 mg /kg of boron) to six male human volunteers aged 30 to 58 years (mean
47.3 years), over 93% of the dose was excreted in the urine within 96 hours of
dosing.
Jansen et al. (1984a) administered via a 20-minute intravenous infusion a total
dose of boric acid of 570 to 620 mg (corresponding to approximately 1.4 to 1.5
mg/kg of boron for a 70-kg male) to each of eight male adults (22 to 28 years
old). Within 120 hours of administration, 98.7% of the dose was excreted in the
urine.
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Boron April 1992
IV. HEALTH EFFECTS
Humans
Short-term Exposure
The lowest oral lethal dose (LD^) of boric acid in humans is reported to
be 640 mg/kg (112 mg/kg of boron), and the dermal LD^, is 8,600 mg/kg
(1,505 mg/kg of boron); the intravenous LD^ for boric acid is reported to
be 29 mg/kg (5.1 mg/kg of boron) (Stokinger, 1981).
Six fatalities occurred within 3.5 days after eight infants ingested formula
prepared with a 2.5% aqueous solution of boric acid. In one case (3.49 kg
body weight), it was estimated that the amount present in the whole body
was between 1 and 2 g of boric acid at death, 45 hours after ingestion
(50.1 mg/kg of boron, for ingestion of 1 g boric acid). The other infants
also may have ingested less than 3 g of boric acid (Young et al., 1949).
Five fatalities were reported among 11 infants who drank formula
prepared with 2.5% aqueous boric acid. The total amount of boric acid
ingested by each of the five infants who died was estimated to range from
4.5 to 14.1 g; ingestion by the six survivors was estimated to range from
2.0 to 4.5 g. It was not specified whether ingestion was limited to one
dose in all cases. For one surviving female infant, who apparently
consumed only one dose, the estimated ingestion was 3 g of boric acid
(165 mg/kg of boron in an infant weighing 3.2 kg) (Wong et al., 1964).
A 5-month-old child (6.1 kg body weight) ingested 1.83 g of a boric acid
solution (53 mg/kg of boron). Except for vomiting 1 hour later and a
slight lethargy at about 12 hours after ingestion, the child was alert and
well 36 hours after ingestion. Based on the serum level of boric acid, a
total body boric acid level of 491 mg (i.e., 85.9 mg boron, or 14 mg/kg)
was estimated at 7 hours after ingestion (i.e., 14 mg/kg of boron) (Martin,
1971).
A 109-kg 44-year-old female ingested 14 g of boric acid (i.e., 22.5 mg/kg
of boron). The patient exhibited erythema, desquamation, elevated liver
function tests and central nervous system (CNS) involvement. The patient
was discharged after 2 weeks in the hospital (Schillinger et al., 1982).
Linden et al. (1986) reported four cases of nonfatal ingestion of boric
acid. Two adult females ingested 298 g of a 99% boric acid-containing
insecticide and 80 g of boric acid in a fungicide, respectively (presumably
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Boron April 1992
1.0 and 0.28 g/kg of boron, respectively, assuming a 50-kg body weight).
The doses ingested by the two other subjects were not clearly specified.
All four subjects recovered, and the two adults presented no systemic
signs of toxicity following release from the hospital.
Six male volunteers aged 30-58 years (mean 47.3) received single oral
doses of boric acid (Jansen et al. 1984b). Three volunteers ingested 750
mg boric acid dissolved in 100 mL of water (1.9 mg/kg of boron based on
an assumed 70-kg body weight). Three other volunteers swallowed 24.95
g to 49.6 g of a commercial water-emulsifying ointment containing 2.97%
(w/w) boric acid (total amount of boric acid consumed by the individuals
was 0.740 g to 1.473 g; 1.8 mg to 3.6 mg/kg of boron). No adverse health
effects were reported in any of the individuals following a single ingestion
of 1.8 to 3.6 mg/kg boron during the 96-hour observation period.
In another study (Jansen et al., 1984a), eight 22- to 28-year-old male
volunteers were given 20-minute intravenous infusions of 21 mg/mL boric
acid in sterile water; the boric acid was infused at a rate of 28.52 to 31.0
mg/min, resulting in a total dose of 570 to 620 mg (1.4 to 1.5 mg/kg) of
boron, for a 70-kg adult. Within 120 hours of treatment, 98.7% of the
boron dose was excreted in the urine. This indicated that excretion of
boron was nearly complete, and that boron exhibited no tendency to
accumulate in the tissues. None of the volunteers reported any
discomfort following the infusion. The No-Observed-Adverse-Effect Level
(NOAEL) in this study was 1.4 to 1.5 mg/kg of boron.
Long-term Exposure
Following ingestion of borax (present in pacifiers coated with a borax-
honey mixture) over of period of several weeks, two infants suffered
seizures and exhibited neurological symptoms. One infant (4.5 months
old) also had erythema of the scalp, trunk and limbs; scanty hair and
increased cellularity in bone marrow aspirates. Other clinical findings for
, the second child (9 months old) were unremarkable. The following
ingestions of borax were estimated: an accumulated total of 125 mg over
=v a 12-week period for the first child and 9 g over 5 weeks for the second
child (approximately 66 and 13 mg/kg/day of boron, respectively) (Gordon
et al., 1973).
In a human nutrition study using basal diets supplying 0.25 mg/day of
boron, Nielsen et al. (1987) reported that supplementing the basal diets of
12 postmenopausal women with 3 mg/day of boron for 119 days (an
approximate total of 0.065 mg/kg/day of boron for a 50-kg woman)
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Boron April 1992
reduced calcium and magnesium excretion in the urine and elevated
serum steroid levels. The effect appeared to be enhanced by low
magnesium levels in the diet.
Animals
Short-term Exposure
Acute oral LDJO values for boric acid in mice and rats range from 466 to
900 mg/kg of boron, and the acute oral LDJO values for borax in rats range
from 396 to 689 mg/kg of boron (Pfeiffer et al, 1945; Wang et al., 1984;
Weir and Fisher, 1972).
Three of six dogs dosed orally with boric acid at 350 mg/kg of boron (as
boric acid) died within 48 hours of exposure (Pfeiffer et al., 1945).
Immature rats (five groups of 20 to 24) of unspecified weight were
administered boric acid in their drinking water at levels of 1 g/L, 2.5 g/L
or higher unspecified concentrations (approximately 22.8, 57 or more mg/
kg/day of boron in animals weighing 100 g and consuming 12 to 14 mI7
day of water) for at least 30 days. Growth was not inhibited in animals
receiving the lowest dose, but growth was inhibited in all other animals
after 20 or 30 days. There were no significant findings in hematology or
in gross and microscopic pathology (Pfeiffer et al., 1945).
Seal and Weeth (1980) administered 0, 150 or 300 mg of boron (as borax)
per liter of drinking water to 45 male Long-Evans rats for 70 days. The
basal diet contained approximately 54 ^g boron/g of feed. The total
intake of boron of the treated rats was 23.7 and 47.4 mg/kg/day,
respectively (assuming a body weight of 0.35 kg, a fluid intake of 0.049 L/
day and a daily food consumption equal to 5% of the body weight). Both
doses of borax produced significant (p < 0.05) decreases in body weight;
in the weights of the testes, seminal vesicles, spleen and right femur and
in the levels of plasma triglycerides. In addition, spermatogenesis was
impaired in animals receiving the highest dose. The Lowest-Observed-
Adverse-Effect Level (LOAEL) in this study was 23.7 mg/kg/day of boron.
Forbes and Mitchell (1957) administered 19 (unsupplemented control
diet), 73,104 or 198 mg/kg/day of a boric acid-supplemented diet to 48
male Sprague-Dawley weanling rats; based on the assumptions of Lehman
(1959), these doses correspond to approximately 3.8, 14.6, 20.8 or 39.6 mg/
kg/day of boron. At the end of 8 weeks, animals in the two higher-dose
groups had significantly (p = 0.05) lower body weights than the untreated
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Boron April 1992
controls; no effect was observed in animals receiving 73-mg/kg of boron.
However, the percentage of the boron dose retained in soft tissue and
skeleton was inversely proportional to the dietary boron level, ranging
from 45% in the 19-mg/kg group to 16% in the 198-mg/kg group. Based
on the results of this study, the suggested NOAEL is 14.6 mg/kg/day.
Green et al. (1973) observed no adverse effects in six 30-day-old rats
administered 75 ppm borax (0.2 mg/kg/day of boron).
Dermal/Ocular Effects
Roudabush et al. (1965) studied the dermal irritation potential of borax
and boric acid. Ten mL of 5% aqueous borate (5.7 g/L of boron) and 5
mL of 10% aqueous boric acid (17.5 g B/L) were applied under occlusion
to the clipped, intact and abraded skin of rabbits and guinea pigs. Sites
were scored for irritation at 24 and 72 hours. Both borax and boric acid
were found to be mild-to-moderate irritants.
In an unpublished study sponsored by the Cosmetics, Toiletries and
Fragrances Association (CTFA) and reviewed by an American College of
Toxicology expert panel (ACT, 1983), a bath preparation containing 0.4%
boric acid (0.7 g/L of boron) was tested for dermal irritation in albino
rabbits. The test solution was applied undiluted and under occlusion to
the shaved intact skin of each animal for 24 hours. Of the nine animals
tested, eight experienced mild-to-moderate erythema.
In an unpublished CTFA-sponsored study reviewed by ACT (1983), a bath
preparation containing 0.4% boric acid (0.7 g/L of boron) was tested for
ocular irritation in albino rabbits. The eyes of the rabbits were rinsed
with warm water 4 seconds after instillation of the test material. The
product was found to be moderately irritating.
Long-term Exposure
In a subchronic study conducted by the National Toxicology Program
(NTP, 1987), groups of 10 male and 10 female B6C3F, mice were fed
diets containing 0, 1,200, 2,500, 5,000,10,000 or 20,000 ppm of boric acid
for 13 weeks (approximately 0, 34, 68,136, 272 or 544 mg/kg/day of boron
for males and 0, 47, 94, 188, 376 or 752 mg/kg/day of boron for females,
respectively, based on reported average values for feed consumption by
controls on week 4 of the experiment). Over 60% mortality was observed
at the highest dose level, and 10% mortality was observed among males
dosed with 10,000 ppm boric acid. At doses of 5,000 ppm or higher,
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Boron April 1992
degeneration or atrophy of the seminiferous tubules was observed in
males, and weight gain was suppressed in animals of both sexes.
Extramedullary hematopoiesis of the spleen of minimal to mild severity
was observed in all dosed groups. The lowest dose tested (1,200 ppm)
was determined to be the LOAEL for this study.
Weir and Fisher (1972) administered borax or boric acid in the diet to
Sprague-Dawley rats (10/sex/dose) for 90 days at doses of 0, 52.5, 175, 525,
1,750 or 5,250 ppm as boron equivalents; these doses corresponded to
approximately 0, 2.6, 8.8, 26, 88 or 260 mg/kg/day of boron, respectively,
assuming a food consumption equivalent to 5% of the body weight as
recommended by Lehman (1959). Both borax and boric acid produced
100% mortality in the highest-dose group and complete atrophy of the
testes in all males fed diets containing 1,750 ppm. At 1,750, both
compounds produced significant (p < 0.05) decreases in body weight and
in the mean weights of the liver, kidney, spleen and testes.
At lower doses, changes in organ weights were inconsistent. At 52.5 ppm,
borax produced increases in the mean weights of the spleen, kidneys and
ovaries in females, and boric acid produced an increase in the mean
weight of the liver in both males and females. Male rats fed 175 ppm
exhibited increased kidney weights. These changes, however, were not
observed in animals fed 525 ppm of either compound. Microscopic
examination revealed partial testicular atrophy in four males fed borax at
525 ppm and in one male fed boric acid at 525 ppm. Although 52.5 ppm
would appear to be a LOAEL in this study, changes in organ weights were
inconsistent at this level of dietary boron and no histopathological data
were provided.
Weir and Fisher (1972) administered a diet containing 0, 17.5, 175 or
1,750 ppm of boron in the form of borax or boric acid to groups of five
young male and five young female beagle dogs for 90 days (corresponding
to approximately 0, 0.44, 4.4 or 44 mg/kg body weight/day, assuming a
daily food intake of 2.5% of the body weight as recommended by Lehman
(1959). Except for one death in a male dog at the 1,750-ppm boron level
(as borax), dogs fed both boron compounds were normal in appearance,
behavior, elimination, body weights and food consumption. At 1,750 ppm
of boron, both compounds produced significant (p < 0.05) decreases in
thyroid- and testes-to-body weight ratios and severe testicular atrophy with
degeneration of the spermatogenic epithelium in all male dogs. At 175
ppm of boron (as boric acid), a decrease in the testes-to-body weight ratio
was observed. This effect, however, was not accompanied by histologica!
changes. At 17.5 ppm of boron (as borax), the LOAEL for this study, the
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Boron ' April 1992
spleen-to-body weight ratio in males was significantly (p < 0.05) elevated.
No histological alterations were observed in dogs fed 175 ppm (or less) of
boron (as boric acid); histological data for borax at lower doses were not
reported.
Settimi et al. (1982) administered 3 g/L of borax in the drinking water to
male Wistar rats for 14 weeks (approximately 20.8 mg/kg/day of boron,
assuming a body weight of 0.350 kg and an average water consumption of
21.4 mL as reported in the study data. Significant (p < 0.05) elevations in
RNA concentration and in succinate dehydrogenase and acid proteinase
activities were reported in the brains of the treated rats. Significant
decreases (p < 0.05) in reduced nicotinamide adenine dinucleotide
phosphate (NADPH) cytochrome reductase activity and in the content of
cytochromes b, and P-450 were observed in the liver. No effect on body
weights or on liver, kidney or testes weights was observed after exposure
to boron. Histopathology was not performed. Because only one dose
level was studied, a NOAEL or LOAEL cannot be identified.
Wang et al. (1984) administered 0, 1, 5, 50 or 500 mg/L of boron (as
borax) in the drinking water to rats of an unspecified strain for up to 198
days. The treatment groups consisted of 5 rats/sex/group, and the control
group consisted of 10 rats/sex. The doses given corresponded to
approximately 0, 0.055, 0.280, 2.8 or 28 mg/kg/day of boron (assuming an
average body weight of 350 g and a water consumption rate of 19i5 mL/
day (water intakes of 22 mL for males and 17 mL for females were
reported at week 26 of the study). Although changes were reported in the
pancreas-to-body weight ratios of all treated animals, detailed results were
not provided. Changes included a significant decrease in the pancreas-to-
body weight ratio in female rats on day 98, and a significant increase in
the pancreas-to-body ratio in males on day 198. However, no
histopathologic changes were noted.
In a lifetime study, Schroeder and Mitchener (1975) administered 5 ppm
of boron (as sodium metabo rate) in the drinking water to 162 male and
female Charles River CD Swiss mice; this dose corresponded to
approximately 8.1 mg/kg/day of boron, assuming a water intake of 5.5 mL/
day and an average body weight of 34 g (based on the reported weights at
90 days). No effects were observed in body weights or longevity, the only
parameters studied. Consequently, the NOAEL in this study was 8.1 mg/
kg/day.
Weir and Fisher (1972) fed groups of four young male and four young
female dogs diets containing 0, 58, 117 or 350 ppm of boron (as borax or
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Boron April 1992
boric acid) for 2 years; these doses corresponded to approximately 0, 1.5,
2.9 or 8.8 mg/kg/day of boron, assuming a daily food intake of 2.5% of the
body weight as recommended by Lehman (1958). An additional group of
dogs was subsequently fed a diet containing 1,170 ppm of boron as borax
(approximately 29 mg/kg/day) for 38 weeks. At this dose, severe testicular
atrophy and spermatogenic arrest were evident within 26 weeks of
treatment. At the lower doses, there were no apparent effects on body
weight, food consumption, organ weights, organ-to-body weight ratios or
clinical parameters. Gross and microscopic findings were comparable to
those found in the controls. This study defines a NOAEL of 8.8 mg/kg/
day of boron.
Weir and Fisher (1972) fed groups of 35 male and 35 female
SpragueDawley rats 0, 117, 350 or 1,170 ppm of boron as borax or boric
acid in the diet for 2-years; these doses corresponded to approximately 0,
5.9, 17 J or 58.5 mg/kg/day of boron, assuming a daily food consumption
of 5% of the body weight as recommended by Lehman (1959). At 1,170
ppm, both animals fed borax and animals fed boric acid exhibited
decreased food consumption during the first 13 weeks of study and
suppressed growth throughout the study. The weights of the testes and
the testes-to-body weight ratio were significantly (p < 0.05) decreased,
and the brain- and thyroid-to-body weight ratios were also significantly (p
< 0.05) decreased. In addition, the seminiferous epithelium was
atrophied and the size of the tubules in the testes was decreased. No
treatment-related effects were observed in rats treated with 350 or 117
ppm boron as borax or boric acid. Therefore the LOAEL in this study
was 1,170 ppm (58.5 mg/kg/day of boron) based on the findings noted
above. A NOAEL of 350 ppm (17.5 mg/kg/day of boron) was identified in
this study.
Reproductive Effects
In two acute exposure studies, adult male rats were dosed orally with 0 or
200 mg/kg of boric acid in a single administration, or dosed with 0, 250,
500, 1000, or 2000 mg/kg in a single administration. Equivalent doses of
boron were 0 or 41 mg/kg or 0, 41, 82,164 or 328 mg/kg, respectively. No
definite histologic changes were detected in animals given 250 or 500
mg/kg of boric acid. In animals dosed with 1,000 or 2,000 mg/kg of boric
acid, adverse effects to the testes and sperm were noted (Linder et al.,
1990).
Male rats administered 9,000 ppm of boric acid (equivalent to 74
mg/kg/day of boron based on the assumed weight of an adult rat) in their
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Boron April 1992
diet for 28 days exhibited testicular lesions. These lesions were expressed
as an inhibition of spermiation followed by exfoliation of germ cells and
pachytene-cell death. After 28 days, extreme epithelial disorganization
and germ cell loss were noted. It was also determined that boron
exposure reduced basal testosterone levels after 4 days of dosing and
levels remained low during dosing (Treinen and Chapin, 1991).
Dixon et al. (1976) administered 0, 0.3, 1.0 or 6.0 mg/L of boron (as
borax) in the drinking water to male Sprague-Dawley rats (10/group) for
up to 90 days (corresponding to approximately 0.042, 0.14 or 0.840 mg/kg/
day of boron assuming a water consumption rate of 35 mL/day and body
weight of 0.250 kg). There were no observable reproductive effects, no
changes in serum chemistry or weight of the body, testes, prostate, or
seminal vesicles. Fructose, zinc and acid phosphatase levels in the
prostate were also unchanged. No effects on male fertility were observed,
suggesting a NOAEL of 0.840 mg/kg/day (the highest dose tested).
Silaev et al. (1977) administered a single daily oral dose of 1 g/kg of boric
acid (175 mg/kg/day of boron) to 12 experimental and six control sexually
mature male albino rats for 2 weeks. Vacuolation and granulation of the
cytoplasm as well as an almost total absence of nuclear chromatin were
observed in the spermatids of most seminiferous tubules. In some
seminiferous tubules, an appreciable reduction in tubular diameter and a
complete absence of germinative cells were observed
Krasovskii et al. (1976) administered 0, 0.3, 1.0 or 6.0 mg/L of boron (as
boric acid) in drinking water to random-bred white male rats (number not
specified) for 6 months approximately 0.015, 0.05 or 0.3 mg/kg/day of
boron, based on the authors' calculations. No significant toxic effects
were observed at 0.3 mg/L of boron. A significant decrease in mobility
time (p < 0.01) and number of spermatozoa (p < 0.01) was observed in
animals in the two higher-dose groups, the intensity of these effects
increased in dose-related fashion. In addition, at the highest dose level,
there was a significant (p < 0.01).decrease in the organ-to-body weight
ratio for gonads and a significant (p < 0.01) increase in serum levels of
blood aldolase. This study defines a NOAEL and a LOAEL of 0.02 and
0.05 mg/kg/day of boron, respectively. The value of this data is
questionable, however, since these results have not been repeated
elsewhere.
Dixon et al. (1976, 1979) administered, 0, 500, 1,000 or 2,000 mg/kg of
boron (as borax) in the diet to male Sprague-Dawley rats (18/group) for
60 days (approximately 25, 50 or 100 mg/kg/day of boron assuming a food
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Boron April 1992
consumption of 5% of the body weight, Lehman, 1959). Testicular and
plasma levels of boron were elevated in a dose-dependent fashion at 30
and 60 days of treatment. Significant (p < 0.05) decreases in the weights
of the liver, testes and epididymis were observed at the 1,000- and 2,000-
mg/kg dietary levels. Seminiferous-tubule diameter was significantly (p <
0.05) decreased in a dose-dependent manner in all treatment groups;
however, loss of germinal cell elements was observable only at the 1,000-
or 2,000-mg/kg dietary level. Aplasia was complete at the highest dose.
The activities of hyaluronidase (H) and sorbitol dehydrogenase (SDH),
testicular enzymes associated with postmeiotic spermatogenic cells, were
significantly (p < 0.05) decreased in apparent dose-related fashion for all
treatment levels. The activity of a related testicular enzyme, lactic acid
dehydrogenase (isozyme-X), was decreased at the 500-mg/kg dietary level
but was significantly (p < 0.05) decreased only at the two higher-dose
levels. Plasma levels of follicle-stimulating hormone (FSH) were
significantly (p < 0.05) elevated in a dose-dependent fashion at all dose
levels. Serial mating studies revealed reduced fertility without change in
copulatory behavior only at the two higher-dose levels. Based upon the
finding of a dose-dependent tubular germinal aplasia which was reversible
at low doses, this study defined a NOAEL of 25 mg/kg/day of boron.
Weir and Fisher (1972), in a multigeneration reproduction study,
administered 0, 117, 350 or 1,170 ppm boron (as borax or boric acid) in
the diet to groups of 8 male and 16 female Sprague-Dawley rats; these
doses corresponded to approximately 0, 5.9, 17.5 or 585 mg/kg/day of
boron, assuming a food consumption of 5% of the body weigh as
recommended by Lehman (1959). No adverse effects on reproduction or
gross pathology were observed in the rats dosed with 5.9 or 17.5 mg/kg/
day of boron as borax or boric acid. Animals in test groups fed 58.5 mg/
kg/day of boron1 as borax or boric acid were found to be sterile. In
addition, males exhibited a lack of spermatozoa in their atrophied testes,
and females exhibited decreased ovulation in the majority of the ovaries
examined. This study identified a NOAEL of 175 mg/kg/day of boron.
Developmental Effects
. In the multigenerational study conducted by Weir and Fisher (1972), in
which female rats were administered 117 to 1,170 ppm of boron, no
reduction in the number of living offspring and no physical defects were
found.
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Boron April 1992
Mutagenicity
Boric acid either was not mutagenic (Iyer and Szybalski, 1958; Szybalski,
1958) or produced equivocal results (Demerec et al., 1951) in Escherichia
coli Sd-4.
Sodium borate and boric acid did not cause gene mutations in the
Salmonella typhimurium preincubation assay with and without rat liver S9
(Benson et al., 1984).
Similarly, boric acid, both with and without rat and Syrian hamster liver
S9 preparations, was negative in an interlaboratory evaluation of the S.
typhimurium preincubation assay (Haworth et al., 1983).
Boric acid did not induce forward mutations in L5178Y mouse lymphoma
cells under nonactivated or S9-activated conditions. Similarly, boric acid
did not induce chromosome aberrations or increase the frequency of sister
chromatid exchanges in Chinese hamster ovary cells under nonactivated or
S9-activated conditions (NTP, 1987).
Boron was reported to induce mitotic suppression and chromosomal
abnormalities in Papaver somniferum; no details were provided (Sopova et
al., 1981).
Carcinogenicity
Groups of 50 male and 50 female B6C3F, mice were fed diets containing
0, 2,500 or 5,000 ppm boric acid for 103 weeks (approximately 0, 21.9 or
43.8 mg/kg/day of boron, assuming a daily food intake of 5% of the body
weight). Mortality in the dosed animals was significantly higher in the
treated males (40 and 56% at 2,500 and 5,000 ppm, respectively) than in
controls (18%). Body weight gain and food consumption were lower in
both sexes. No treatment-related clinical signs of carcinogenicity were
observed. At 5,000 ppm, boric acid caused an increased incidence of
testicular atrophy and interstitial hyperplasia in males. Under the
conditions of this study, there was no evidence of carcinogenicity for boric
. acid in male or female mice (NTP, 1987).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs), are based upon the identification of adverse health effects
associated with the most sensitive and meaningful noncarcinogenic endpoint of toxicity. The
15
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Boron
April 1992
induction of this effect is related to a particular exposure dose over a specified period of time,
most often determined from the results of an experimental animal study. Traditional risk
characterization methodology for threshold toxicants is applied in HA development. The
general formula is as follows:
UA (NOAEL or LOAEL) (BW) _ .
HA = -: ; '.: L = mg/L (rounded to ug/L)
(UF) ( L/day) * V ^ '
where:
NOAEL =
No-Observed-Adverse-Effect Level (the exposure dose in mg/kg
bw/day).
or
LOAEL = Lowest-Observed-Adverse-Effect Level (the exposure dose in mg/
kg bw/day).
BW = assumed body weight of protected individual (10 kg for child or
70 kg for adult).
UF(s) = uncertainty factors, based upon quality and nature of data (10,
100, 1,000, or 10,000 in accordance with EPA or NAS/OW
guidelines).
L/day = assumed water consumption (1 L/day for child or 2 L/day for
adult).
One-day Health Advisory
Although no ideal acute toxicity study was found for calculating the One-day Health
Advisory (HA) for boron, the study of Jansen et al. (1984b) has been selected as providing the
best available data. In this study, no adverse effects were reported following oral ingestion of
3.6 mg/kg of boron, as boric acid, by each of six human, males. Other studies on accidental
ingestion of boric acid solution reported deaths in six out of eight infants (Young et al., 1949)
ingesting 50 to 100 mg/kg of boron; or in five of eleven infants ingesting 200 to 700 mg/kg of
boron (Wong et al., 1964). In one study (Martin, 1971), a five-month-old child (6.1 kg)
ingested approximately 1.83 g of boric acid (14 mg/kg) of boron; the child vomited 1 hour later
and exhibited slight lethargy for about 12 hours. The child was alert and well 36 hours after
ingestion.
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Boron ' April 1992
The NOAEL from the Jansen et al., (1984b) study (3.6 mg/kg/day) was used to
calculate the One-day HA for boron. The One-day HA for the 10-kg child is calculated as
follows:
One-day HA - (3-6 mg/kg/day) (10 kg) =
7 (10) (1 L/day) &
where:
3.6 mg/kg 1 day = NOAEL based on absence of adverse effects in human adults
ingesting boric acid.
10 kg = assumed weight of child.
10 = Uncertainty factor; this uncertainty factor was chosen in
accordance with EPA or NAS/OW guidelines in which a
NOAEL from a human study is employed.
1 L/day = assumed water consumption by a 10-kg child.
Ten-day Health Advisory
The available data are insufficient to develop a Ten-day HA for boron. Therefore, it
is recommended that the Longer-term HA for a 10-kg child, 0.9 mg/L, be used as a
conservative estimate of the Ten-day HA.
Longer-term Health Advisory
The study conducted by Weir and Fisher (1972) involving dietary ingestion of boric
acid and borax for 90 days in dogs was selected as the basis for calculation of the Longer-term
HA for boron. In this study, dogs were administered 0, 1.5, 2.9, 8.8 or 29 mg/kg/day of boron
as borax or boric acid. This study defined a NOAEL of 8.8 mg/kg/day, based on the absence of
reproductive effects, gross pathology, and testicular histopathology. Higher doses of boron
(23.7 and 29 mg/kg/day) have been shown to produce testicular atrophy in Long-Evans rats
(Seal and Weeth, 1980) and in beagle dogs (Weir and Fisher, 1972), respectively. The results
of subchronic studies in rats and dogs (Weir and Fisher, 1972), although well-conducted, were
not considered because histopathology was not completely reported at the lower doses. In
addition, the organ-weight changes observed in the rats in lower-dose groups were
inconsistent. The NTP (1987) study reported extramedullary hematopoiesis at all doses (34 to
752 mg/kg/day of boron) in mice administered boric acid for 13 weeks. No NOAEL was
defined in this study and, at the LOAEL of 34 mg/kg/day, mild spleen pathology was evident.
17
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Boron
April 1992
Therefore, the NOAEL identified in the Weir and Fisher (1972) study, 8.8 mg/kg/day, was
used to calculate the Longer-term HA for boron.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA
6
<8'8
kg) =
(100) (1 L/day)
= 08g
(rounded to 0.9 mg/L)
v '& '
where:
8.8 mg/kg/day
100
10kg
1 L/day
NOAEL, based on the absence of reproductive effects, or
gross pathology and testicular histopathology, in a chronic
2-year study in dogs exposed to borax or boric acid.
uncertainty factor, chosen in accordance with EPA or
NAS/OW guidelines for use with a chronic animal study.
assumed weight of a child.
assumed water consumption of a 10-kg child.
The Longer-term HA for a 70-kg adult is calculated as follows:
Longer-term HA = (8-8 mg/kg/day) (70 kg) = ed
6 (100) (2 L/day) & v & '
where:
8.8 mg/kg/day =
100 =
70kg =
2 L/day =
NOAEL, based on the absence of reproductive effects, or gross
pathology in a chronic 2-year study, in dogs exposed to borax or
boric acid.
uncertainty factor, chosen in accordance with EPA or NAS/OW
guidelines for use with a chronic animal study.
assumed weight of an adult.
assumed water consumption of an adult.
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Boron April 1992
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 (RED), formerly called the Acceptable Daily Intake (ADI).
The RfD is an estimate (with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is likely to be without
appreciable risk of deleterious health effects during 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 in drinking water alone 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 known, probable, or possible carcinogen, according
to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then caution
must be exercised in making a decision on how to deal with possible lifetime exposure to this
substance.
For human (A) or probable (B) human carcinogens, a Lifetime HA is not
recommended. For possible (C) human carcinogens, an additional 10-fold safety factor is used
in the calculation of the Lifetime HA. The risk manager must balance this assessment of
carcinogenic potential and the quality of the data against the likelihood of occurrence and
significance of health effects related to noncarcinogenic endpoints of toxicity. To assist the
risk manager in this process, drinking water concentrations associated with estimated excess
lifetime cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult
drinking 2 L of water/day are provided in the Evaluation of Carcinogenic Potential section.
The chronic toxicity study in dogs by Weir and Fisher (1972) was selected to serve as
the basis for calculation of the Lifetime HA for boron. Other chronic studies were
considered, including a chronic study with rats (Weir and Fisher, 1972) and a lifetime study
with mice (Schroeder and Mitchener, 1975). However, limitations or deficiencies in these
studies precluded their use in the development of the Lifetime HA for boron.
Weir and Fisher (1972) reported a NOAEL for boron of 8.8 mg/kg/day in dogs fed
boric acid in the diet for 2 years. This NOAEL was selected because there is clear evidence
that the dog is more sensitive than the rat to boron.
19
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Boron April 1992
Schroeder and Mitchener (1975) reported no effects on longevity or body weight in
mice administered 8.1 mg/kg/day of boron (as metaborate) for a lifetime. This study was not
selected because a limited number of parameters were tested and because only one dose level
was tested.
In the study selected as a basis for development of the Lifetime HA for boron (Weir
and Fisher, 1972), dogs were administered 0, 1.5, 2.9, 8.8 or 29 mg/kg/day of boron as borax or
boric acid for 2 years. This study defined a NOAEL for boron of 8.8 mg/kg/day based on the
absence of either testicular atrophy or spermatogenic arrest at this level of exposure.
Using the NOAEL from this study, identified in the chronic dog study by Weir and
Fisher the Lifetime HA is derived as follows:
Step 1: Determination of the RfD
RfD = (8.8 mg/kg/day) = Q 08g mg/kg/day (rounded to 0.09 mg/kg/day)
(100)
where:
8.8 mg/kg/day = NOAEL, based on the absence of gross on light-microscopic
pathology in dogs exposed to borax or boric acid in the diet for 2
years.
100 = uncertainty factor, chosen in accordance with EPA or NAS/OW
guidelines for use with a NOAEL from a chronic animal study.
Step 2: Determination of the DWEL
DWEL . (0.09 mg/kg/day) (70 kg) = ^
(2 L/day) & ^ & '
where:
0.09 mg/kg/day = RfD.
70 kg = assumed weight of an adult.
2 L/day = assumed water consumption of a 70-kg adult.
20
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Boron April 1992
Step 3: Determination of the Lifetime HA
Lifetime HA = (3.15 mg/L) (20%) = 0.63 mg/L (0.6 mg/L)
where:
3.15 mg/L = DWEL.
0.20 = assumed relative source contribution.
Evaluation of Carcinogenic Potential
According to the U.S. EPA classification scheme for carcinogenic potential
(U.S. EPA, 1986), boron is classified as Group D: not classifiable.
No quantitative assessment of excess cancer risk attributable to boron has been
reported.
The NTP (1987) conducted a bioassay with B6C3Ft mice fed diets containing
boric acid at levels of 0, 2,500 or 5,000 ppm for 103 weeks. No evidence of
carcinogenicity was found in male or female mice.
Boric acid has not been found to be mutagenic in bacterial and mammalian cell
assays.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The U.S. EPA has established tolerances for total boron of 30 ppm in or on
cottonseed and of 8 ppm in or on citrus fruits (U.S. EPA, 1987b).
VII. ANALYTICAL METHODS
Methods for metal analysis involve spectroscopy, either emission or absorption.
In all of these methods, the metal is dissolved and then thermally excited. All
elements when excited emit or absorb light frequencies characteristic of that
element. Most metal spectroscopy is done in the ultraviolet and x-ray regions.
Direct Aspiration Atomic Absorption Spectroscopy (AA). In this technique,
dissolved metals are aspirated into a flame source and excited to the point of
the dispersion into a mono-atomic state; a light source whose cathode is the
metal of interest passes through the flame, and the resulting absorption of light
21
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Boron April 1992
by the element of interest is directly proportional to its concentration.
Disadvantages of this technique include the inability to analyze more than one
metal at a time and the insensitivity of the technique.
Graphite Furnace Atomic Absorption (GFAA). In this technique, a specific
amount of liquid is dried on the thermal source, effecting a concentration step.
The sample is then electrothermally excited. This technique is very sensitive.
Inductively Coupled Plasma Atomic Emission Spectroscopv. This method (EPA
Method 200.7) utilizes aspiration of a liquid sample, but the flame is actually a
plasma torch of Argon excited to super hot levels by radio- frequency radiation.
In this technique, the metals are excited sufficiently to emit radiation. By using
classical dispersion-grating optics, a large number of metals can be analyzed
simultaneously. This technique has a detection limit of 1.25
Inductively Coupled Plasma Mass Spcctrometrv (ICP/MSV In this technique,
the excitation is effected using a radio- frequency plasma, but the excited atoms
are interfaced into a mass spectrometer. Quantitation is achieved by
computerized software programs similar to those used in other U.S. EPA MS
organic methods.
VIII. TREATMENT TECHNOLOGIES
Available data indicate that reverse osmosis (RO), electrodiaiysis (ED/EDR),
lime softening, ion exchange (IX) and granular activated carbon (GAC)
adsorption can significantly reduce boron levels in contaminated drinking water
supplies.
Jarrett (1978) reported that an RO treatment designed for bakery use reduced
boron levels by 60% (from an initial concentration of 0.3 mg/L). This system
was designed to treat 20,000 gpd of water at an operating pressure of 600 psi
and a water recovery rate of 75%. The source water was well water with a total
hardness of 1,830 mg/L.
Folster et al. (1980) reported boron removal by RO and ED/EDR treatment
plants in San Jon and Alamogordo, New Mexico. The RO systems in both
locations consisted of hollow-fiber (HF) and spiral-wound (SW) configuration
membranes. The HF RO was operated at 515 psi and a water recovery rate of
78%, while the SW RO was operated at 430 psi and a water recovery rate of
79%. The ED/EDR plants were operated at a water recovery rate of 85%. At
the San Jon plant, boron was present in the influent at an initial concentration
of 0.8 mg/L. ED/EDR treatment removed 20% of the boron, while HF RO
22
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Boron April 1992
treatment removed 15% of the boron. The SW RO system did not remove any
boron. At the Alamogordo site, ED/EDR treatment removed 50% of the boron
(from an initial concentration of 0.01 mg/L), and SW RO treatment removed at
least 94% of the boron (from an initial concentration of 0.09 mg/L). No
removal was achieved by HF RO.
. Potable water was reclaimed from unchlorinated effluent by an activated sludge
wastewater treatment plant using lime softening, reacidification and RO at
Water Factory-21 (WF 21) in Orange County, California (Argo, 1984). The
influent was softened at pH 11. The softened water was carbonated to pH 8.0,
filtered and passed through an RO system. Two 5,000 gpd RO units with
tubular aromatic polyamide membranes arranged in a 2-1 array configuration
were used. The average flux rate was 7.14 gpd/sq ft at 250 psi applied pressure.
The lime/RO process removed 80% of the boron (from an initial average
influent concentration of 0.55 mg/L).
Wong (1984) studied the use of two different IX resins to reduce boron levels in
water from a utility plant, thereby reclaiming it for potable use. A boron-
specific anion exchange resin (IRA-743) and a strong base anion exchange resin
were tested. Both resins tested removed boron to below 0.1 mg/L from an
initial concentration of 10 mg/L. The IX columns contained 50 ft, of resin and
were operated at a loading rate of 2 gpm/ft3. Both types of resins were
regenerated with 4 Ib/ft3 of a sodium hydroxide solution. IRA-743 provided an
exchange capacity of 0.125 Ib boron/ft3 resin, while the strong base anion resin
provided an exchange capacity of 0.36 Ib boron/ft3 resin.
Choi and Chen (1979) developed batch adsorption isotherms for a variety of
commercially available GAC systems. Regardless of the characteristics of the
background solution, a boron removal efficiency of approximately 90% can be
achieved with Filtrasorb* carbon at a dose of 25 g/L if the initial boron
concentration does not exceed 5 mg/L.
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Boron April 1992
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Windholz, M., S. Budavari, R.F. Bluminetti and E.X. Ottenbein, eds. 1983. The Merck index
An encyclopedia of chemicals, drugs and biologicals, 10th ed. Rahway, NJ: Merck
and Company, Inc., pp. 1,324-1,325.
Wong, J.M. 1984. Boron control in power plant reclaimed water for potable reuse. Environ.
Prog. 3(1):5-11.
Wong, L.C., M.D. Heimbach, D.R. Truscott and B.D. Duncan. 1964. Boric acid poisoning:
Report of 11 cases. Can. Med. Assoc. J. 90:1,018-1,023.
Young, E.G., R.P. Smith and O.C. Macintosh. 1949. Boric acid as a poison. Report of six
accidental deaths in infants. Can. Med. Assoc. J. 61:447-450.
Zook, E.G. and J. Lehman. 1965. Total diet study: Content of ten minerals-aluminum,
calcium, phosphorus, sodium, potassium, boron, cooper, iron, manganese and
magnesium. J. Assoc. Off. Agric. Chem. 48:850-855.
28
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SPA 0553
RX000027511
April 1992
CHLORPYRIFOS
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology, 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 endpoints 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 lifelong exposure and the ingestion of water. The cancer unit risk
is usually derived from a linearized 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, ibgjt, 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.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 2921-88-2
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Chlorpyrifos April 1992
Structural Formula
3 J~^
Chlorpyrifos
Synonyms
Brodan; Chlorpyrifos; Chlorpyrifos-ethyl; Detmol U.A.; Dowco 179; Dursban;
Dursban F; Ent 27311; Eradex; Ethion, dry; Lorsban; NA 2783 (Dot); OMS-0971;
Phosphorothioic acid, O,O-diethyl O-(3,5,6-trichloro-2-pyridyl) ester; Pyrinex.
Uses
Chlorpyrifos is a broad-spectrum insecticide with many uses. An estimated 7 to
11 million pounds of Chlorpyrifos are produced each year in the United States for
domestic use. Of the total domestic Chlorpyrifos usage, 57% is applied to corn
and 5 to 6% to cotton. Commercial pest control and lawn and garden services
consume 20 to 22% of the annual Chlorpyrifos usage, followed by domestic
household and lawn and garden application (9 to 13%).
Properties (Kenaga, 1980; Windholz et al., 1983; Worthing, 1987)
Chemical Formula
Molecular Weight 350.57
Physical State (25ฐC) White, granular crystals
Boiling Point
Melting Point 41 to 43.5ฐC
Density
Vapor Pressure (258C) 1.87 x 10's mm Hg
Water Solubility (25ฐC) 2 ppm
Specific Gravity
Log Octanol/Water Partition 4:99
Coefficient (log K^)
Taste Threshold (water)
Odor Threshold (water)
Conversion Factor
(ppm air as mg/m3)
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Chlorpyrifos April 1992
Occurrence
Based on a 3-year, 20-city nationwide study conducted by the Food and Drug
Administration, Gartrell et al. (1985) estimated that the average daily intake of
chlorpyrifos from food and beverages (including water) is approximately 0.001 to
0.005 /zg/kg. Contaminated grain and cereal products and garden fruits were the
food groups through which exposure occurred.
Environmental Fate
Chlorpyrifos hydrolyzes readily in water, its rate of hydrolysis increases with
temperature (Worthing, 1987). When mixed with distilled water (pH 6.5) or
pasture water (pH 7 to 9), chlorpyrifos levels dropped an average of 25% at 24ฐC
and 48% at 38ฐC within 8 hours (Schaefer and Dupras, Jr., 1969).
Chlorpyrifos administered at a rate of 3.4 kg chlorpyrifos/hectare (ha) dissipated
fairly rapidly in sand and organic muck soils with respective half-lives of 2 and 8
weeks in the top 15 cm of soil (Chapman and Harris, 1980). Low levels of
chlorpyrifos (2 to 3% of the amount applied) remained in both soils for up to 2
years. 3,5,6-Trichloro-2-pyridinol was the primary degradation product, reaching
maximum concentrations of 13 and 39% of the chlorpyrifos applied to the sand
and muck soils, respectively. The concentrations of the oxygen analog of
chloropyrifos were ฃ 0.004 ppm in all samples. Chlorpyrifos (EC, eraulsifiable
concentrate), applied at 4 kg chlorpyrifos/ha to turf grass, dissipated rapidly with
a half-life of <14 days in the soil and turf cover (Sears and Chapman, 1979).
Movement of chlorpyrifos from the turf into the soil was minimal (<18% of the
recovered chlorpyrifos at any time during the study).
Breakdown of chlorpyrifos in soil primarily results from microbial metabolism
(Miles et al., 1979). Chlorpyrifos (10 ppm) is degraded more rapidly in sandy
loam soil (half-life, <1 week) than in organic soil (half-life, 25 weeks). In
sterilized soils, the half-life for chlorpyrifos is >17 weeks. Half-lives of 11 to 141
days were reported in another study in soils ranging in texture from loamy sand
to clay (Bidlack, 1979). 3,5,6-Trichloro-2-methoxypyridine and two unidentified
minor metabolites of chlorpyrifos were recovered after a 1-year incubation
period; most of the radiocarbon, however, was recovered as 14CO2 with small
amounts incorporated into soil organic matter.
After 30 days of aerobic aging of soil, uC-chlorpyrifos degraded with half-lives of
15 days in loam and 58 days in clay soils. The half-lives in treated and
anaerobically incubated loam and clay were 39 and 51 days, respectively. The
major degradation product formed was 3,5,6-trichloro-2-pyridinol. Degradation
of this compound was very slow. Evolution of 14CO2 was insignificant, and
incorporation of UC into the soil organic matter was slow. Relatively low levels
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Chlorpyrifos April 1992
(<3%) of 3,5,6-trichloro-2-methoxypyridine and two unidentified metabolites
were present in small amounts of the samples (Bidlack, 1979).
Chlorpyrifos was immobile in loamy sand and sandy loam soil; only 2 to 13% of
the applied radioactivity leached out of the zone of application with 6 to 10
inches of water. Mobility in beach sand was low. After leaching with 10 inches
of water, 78.9% of the surface-applied Chlorpyrifos remained in the top inch of
sand (Dow Chemical Company, 1972).
Chlorpyrifos was not persistent in pond water treated at 0,05 Ib active ingredient
per acre (a.i./a) or in a polluted aquatic environment treated at 0.23 to 0.27 Ib
a.i./a (Schaefer and Dupras, 1970; Madder, 1977). The rates of decline were not
determined, and losses to underlying segments were not investigated.
Chlorpyrifos applied to pond or rice floodwater as a slow-release formulation
(chlorinated polyethylene pellets) exhibited no patterns of decline in 22 weeks
(Nelson and Evans, 1973). The concentration of Chlorpyrifos was extremely
variable in the top 1 inch of pond sediment and rice plot soil; however, there was
a clear trend toward the partitioning of Chlorpyrifos from water onto soil and
sediments.
14C-Chlorpyrifos residues found in wheat, soybeans and beets planted 119 days
after treatment of loamy sand soil with 14C-chlorpyrifos at 2 Ib a.i./acre amounted
to 0.31, 0.31 and 0.03 ppm Chlorpyrifos equivalents, respectively. Chlorpyrifos
was largely degraded in the soil before the crops were planted, however, and the
plant residues consisted primarily of unidentified UC residues. Residues in wheat
and soybeans concentrated in the vegetative portions of the plants (Bauriedel et
al., 1976).
III. PHARMACOKINETTCS
Absorption
Chlorpyrifos (unlabeled, 99.8% pure) was readily absorbed from the
gastrointestinal (GI) tract in six men given a single oral dose at 0.5 mg/kg (Nolan
et al., 1984). Absorption was estimated to be approximately 70% over a 5-day
period. Blood Chlorpyrifos levels remained low (<30 ng/mL) throughout the
study. Mean blood concentrations of the principal metabolite of Chlorpyrifos,
3,5,6-trichloro-2pyridinol (3,5,6-TCP), peaked at 0.93 /ig/mL 6 hours after
ingestion. There was a 1- to 2-hour delay in the absorption of the oral dose.
Approximately 90% of a single oral dose of 50 mg ^Cl-chlorpyrifos/kg (in corn
oil) was absorbed from the GI tract of male Wistar rats within 2 to 3 days after
dosing (Smith et al., 1967).
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Chlorpyrifos April 1992
A single oral dose of uC-chlorpyrifos (2,6-ring-labeled) (19.1 mg/kg, 23.5 /xCi/mg;
or 6.9 mg/kg, 48.4 /iCi/mg) was absorbed rapidly by male Sprague-Dawley rats
(Dow Chemical Company, 1972). At least 73% of the l4C-chlorpyrifos was
absorbed by the rats within 3 days after administration. Only 1.6 to 2.5% of the
administered radioactivity remained in the tissues and carcass at 72 hours
postdosing; blood I4C levels peaked 1 to 3 hours after dosing, accounting for 3 to
6% of the ingested 14C-chlorpyrifos. About 68 to 76%, 5 to 15%, and 0.15 to
0.63% of the administered 14C was eliminated in the urine, feces and expired air,
respectively, within 72 hours. The authors reported that absorption may have
been slightly reduced in some animals as a result of predose starvation and
frequent bleeding at 2-hour intervals.
Less than 3% of single doses of analytical grade (unlabeled, 99.8% pure)
chlorpyrifos (5.0 mg/kg, dissolved in dipropylene glycol methyl ether or methylene
chloride) was absorbed 7 days after dermal application to six men (Nolan et al.,
1984). Blood levels of 3,5,6-TCP, which were used to determine absorption and
clearance rates of chlorpyrifos, peaked at 0.063 /ig/mL 24 hours post-dosing. The
average half-life for the appearance of 3,5-6-TCP in the blood was 22.5 hours.
Distribution
Because of the rapid elimination of chlorpyrifos and its metabolites following
administration of a single oral dose of 0.5 mg/kg to six men (Nolan et al., 1984),
they are not expected to accumulate to any appreciable extent in humans.
The highest levels of radioactivity in male Wistar rats given a single oral dosage
of MCl-chlorpyrifos (50 mg/kg) were recovered at 4 hours post-dosing in the
kidneys, liver, lung and fat (0.0924, 0.0690, 0.406 and 0317 mmol radioactive
equivalents/kg tissue, respectively) (Smith et al., 1967). Radioactivity was
eliminated rapidly from the liver (tw, 10 hours), kidney (tw, 12 hours) and
muscle (tw, 16 hours) but was retained for a longer period of time by fat tissue
(tw, 62 hours).
Tissue 14C residue levels were low (<1 ppm) 72 hours after male Sprague-Dawley
rats were given a single oral dosage of [2,6-l4Cpyridyl]chlorpyrifos (19.1 mg/kg;
23.5 /xCi/rag) (Dow Chemical Company, 1972). Fat and intestines contained the
highest levels of radioactivity (approximately 0.757 and 0.363 ppra, respectively);
brain 14C residue concentrations were <0.010 ppm.
Metabolism
Very low levels (<30 mg/mL) of unchanged chlorpyrifos were found in the blood
and no parent compound was recovered in the urine during the 5 days after six
men were given a single oral dose (0.5 mg/kg) of the pesticide (Nolan et al.,
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Chlorpyrifos April 1992
1984). Most of the chlorpyrifos was converted to 3,5,6-TCP; however, other
metabolites were not identified.
In studies ot" workers exposed occupationally to chlorpyrifos, several urinary
metabolites of the insecticide were identified by gas chromatography. O,O-
Diethyl phosphate was found in 96% of the urine specimens, and O,O-diethyl
phosphorothionate was recovered from 28% of the samples (Hayes et al., 1980;
Lores and Bradway, 1977). Hayes et al. (1980) reported that 8-hour exposure
levels were <27.6 /zg/m3 for workers exposed to spray emulsions of Dursban E-2
containing 23.5% chlorpyrifos.
Two major metabolites, 36Cl-3,5,6-trichloro-2-pyridyl phosphate and MCl-3,5,6-
TCP, were recovered from the urine and feces of male Wistar rats administered a
single oral dose of 50 mg 36Cl-chlorpyrifos/kg (in corn oil) (Smith et al., 1967).
Male Sprague-Dawley rats given a single oral dose of 14C-chlorpyrifos (19.1
mg/kg, 23.5 /zCi/mg) excreted 3,5,6-TCP as the major metabolite and another
unidentified compound in the urine (Dow Chemical Company, 1972). A total of
1% of the 14C recovered in expired air, almost all of which was UC-CO2,
suggesting that some cleavage of the pyridyl ring had occurred.
In an in vitro study using rat hepatic microsomes, 14C-chlorpyrifos (10 mg/mL,
10.6 mCi/mmol) was readily metabolized to 3,5,6-TCP (Dow Chemical Company,
1972). No other metabolites were found. The reaction was NADPH-dependent,
and binding of chlorpyrifos to microsoraes occurred prior to catabolism. These
findings were also noted in studies by Sultatos et al. (1981, 1982, 1985) and
Sultatos and Murphy (1983). They indicated that chlorpyrifos may be
metabolized by a glutathione-mediated process. Male Charles River Swiss mice
injected with chlorpyrifos (70 mg/kg) displayed a "moderate but transient"
depletion of hepatic glutathione (Sultatos et al., 1982).
Mostafa et al. (1983) reported that the in vivo alkylating activities of l-14C-ethyl-
labeled chlorpyrifos were high following intraperitoneal injection of 5- or 15-
rng/kg doses in male mice (strain not given). Labeled 7-ethylguanine found in
hepatic RNA hydrolysates measured approximately 5.5 x 10 3% of the
administered radioactivity. The two major unidentified radioactive peaks
associated with hepatic DNA hydrolysates corresponded to 3 x 10"% and 2.3 x
10"3% of the applied 14C dose. The authors reported that the total incorporation
of 14C into mouse liver nucleic acids was greater for RNA than for DNA. In
addition, the degree of 14C-incorporation appeared to be dose related.
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Chlorpyrifos April 1992
Excretion
Within 5 days after ingesting a single oral dose of chlorpyrifos (0.5 mg/kg), six
men eliminated an average of 70% of the administered insecticide via the urine,
with a urinary elimination half-life of 27 hours (Nolan et al., 1984). Fecal
elimination of chlorpyrifos and/or its metabolites was not measured.
Smith et al. (1967) reported that approximately 90% of a radioactive dose of MC1-
chlorpyrifos (50 mg/kg) administered orally to male Wistar rats was excreted in
the urine within 2 to 3 days. The remaining 10% was eliminated in the feces.
Approximately 68 to 70%, 14 to 15%, and 0.15 to 0.39% of a single oral dose of
14C-chlorpyrifos (19.1 mg/kg, 23.5 ^Ci/mg) administered to two male Sprague-
Dawley rats were eliminated in the urine, feces and exhaled air, respectively,
within 72 hours after dosing (Dow Chemical Company, 1972). Thus, the urine
provided the primary route of elimination for the insecticide and/or its
metabolites.
Essentially all of the 3% of a dermal dose (5.0 mg/kg) of chlorpyrifos absorbed
by male volunteers was eliminated in the urine within 7 days post-administration
(Nolan et al., 1984). An elimination half-life of 27 hours was reported.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Plasma cholinesterase (ChE) activity was depressed to about 15% of
predose levels following administration of a single oral dose of 0.5 mg
chlorpyrifos/kg to six men (Nolan et al., 1984). Enzyme activity returned
to near-normal (i.e., 80 to 90% of predose levels) within 4 weeks. No
other signs or symptoms of toxicity were observed during the 30-day post-
treatment period.
In a study by Dow Chemical Company (1972), 16 human male volunteers
(four/dose) received 0, 0.014, 0.03 or 0.10 rag chlorpyrifos/kg/day (in
capsule form) for 28, 28. 21 or 9 days, respectively. The high-dose
treatment (0.10 mg/kg/day) was discontinued after 9 days due to a runny
nose and blurred vision in one individual. The authors did not state why
administration of the 0.03 mg/lcg dose was terminated on day 21. Mean
plasma ChE activity in the high-dose (0.10 mg/kg) group was inhibited by
about 30% when compared to the mean control value (p < 0.05) and by
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Chlorpyrifos April 1992
about 65% when compared to baseline (i.e., pretreatment) levels. In the
group receiving 0.03 mg/kg/day doses, plasma ChE activity averaged about
70% of pretreatment levels and 87% of concurrent control values;
however, these differences were not statistically significant. Plasma ChE
activity was comparable for low-dose and control individuals. Plasma
ChE activities of all affected persons returned to pretreatment levels
within 4 weeks after administration of test material was terminated. No
effect on erythrocyte ChE activity was observed at any dose. This study
identified a No-Observed-Adverse-Effect Level (NOAEL) of 0.03 mg/kg/
day and a Lowest-Observed-Adverse-Effect Level (LOAEL) of 0.1
mg/kg/day based on the absence or presence of decreased plasma ChE
activity.
Five office workers exposed to chlorpyrifos in the air (levels not reported)
for 5 to 21 hours over a 3-day period had significantly (p <0.01) reduced
eythrocyte ChE levels 1 month after exposure, when compared to values
obtained 4 months post-exposure (Hodgson et al., 1986). Erythrocyte
ChE activity measured on the first day after exposure was estimated to be
approximately 33% of the 4-month value. Physical examinations, nerve
conduction studies, and routine blood and urine tests were normal for all
but one worker, who developed numbness and tingling in the fingertips of
both hands 3 weeks after exposure. Most of the individuals complained
of fatigue, weakness and anxiety and experienced diarrhea, abdominal
pain and nausea within hours and also during the first 3 weeks after
exposure to chlorpyrifos. Symptoms were resolved by 4 weeks, and no
chlorpyrifos was detected in the office air 2 weeks after the initial
exposure period.
A 42-year-old man who ingested approximately 300 mg chlorpyrifos/kg
was comatose and showed acute signs of cholinergic toxicity through day
17. Longer term neurological effects (leg weakness, reduced or abolished
tendon reflex, reduced or lost vibration sense, and muscle denervation)
were present from day 40 and became progressively worse with time
(Lotti et al., 1986). Blood concentration of chlorpyrifos dropped in an
exponential manner from 680 nmol/L on day 3 to 49 nraol/L on day 10;
none was detected 13 days after ingestion. Blood ChE, plasma
butylcholinesterase and lymphocyte neuropathy target esterase (Nit)
activity levels were markedly depressed on day 30 but began to increase
thereafter, through day 90. Inhibition of NTE preceded the development
of polyneuropathy.
An 11-day-old boy, exposed to chlorpyrifos in the home, became lethargic
and cyanotic prior to respiratory arrest (Dunphy et al., 1980). The infant
was resuscitated but remained limp and relatively unresponsive to stimuli.
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Chlorpyrifos April 1992
His red blood cell ChE activity level was about 50% below normal. After
8 days, the infant appeared well but ChE activity was not measured.
Exposure was probably via both the oral and cutaneous routes, since
chlorpyrifos residues were found on dish towels, food preparation surfaces
and the infant's clothing. Direct inhalation exposure may also have
occurred since the house reportedly smelled strongly of insecticides when
the baby was taken to the hospital.
Insecticide-related signs or symptoms of toxicity were not observed in any
of six men who received a single dermal application of 5.0 mg
chlorpyrifos/kg (Nolan et al., 1984). Mean plasma ChE activity was
depressed slightly (to about 13% of predose levels) but did not exhibit a
consistent pattern among the individual volunteers.
Seven human adults (sex not reported) were exposed dermally, by patch
tests, to 1.0, 1.5, 3.0, 5.0 or 7.5 mg chlorpyrifos/kg; the total exposure
areas ranged from 2.25 to 13 JO in2, and the length of exposure was 12
hours (Dow Chemical Company, 1972). No skin irritation was observed
in any of the subjects, and both erythrocyte and plasma ChE levels
remained unchanged throughout the experimental period. In addition, no
morphological alterations were observed in lymphocytes obtained from
exposed sites. The data indicate that low levels of chlorpyrifos do not
present a significant toxicity hazard from acute skin exposure.
Plasma ChE activities in a group of seven adult humans (sex not
reported) decreased by about 30% following multiple 12-hour dermal
exposures to chlorpyrifos (Dow Chemical Company, 1972). During the
first test period, individuals received three applications of 25 mg
chlorpyrifos/kg, and in the second experiment, each subject received
applications of 5 mg chlorpyrifos/kg. No other effects, including dermal
irritation, were observed. ChE activity levels returned to normal within 7
to 9 days after the final exposure.
Spray workers exposed to 0.5% chlorpyrifos emulsion in field trials for
malaria control snowed decreased plasma and erythrocyte ChE activity
levels (Eliason et al., 1969). In this study, five of seven sprayers showed
more than a 50% reduction in ChE within 2 weeks after spraying began.
In a study by Ludwig et al. (1970), groups of two to three human
volunteers were exposed to one of several thermal aerosols containing
chlorpyrifos. Exposures of 3 to 8 minutes at concentrations of about 0.8
/im/m3 produced no significant changes in ChE levels. This concentration
is similar to the application rate recommended in thermal fogging.
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Chlorpyrifos April 1992
Long-term Exposure
Plasma ChE was significantly (p <0.001) inhibited in a group of 17
workers exposed occupationally to a Time-Weighted Average (TWA) of
7.54 mg chlorpyrifos/m3 for 8 hours/day, 5 days/week, for 2 years, when
compared to age- and sex-matched controls (Hayes et al., 1980). Most
workers experienced headaches and complained of aggravated nasal or
respiratory problems. General physical examinations were normal,
however, as were erythrocyte ChE activity levels.
Two groups of machine-operating farm workers (numbers not reported)
exposed daily to a granular insecticide containing 5% chlorpyrifos were
examined over a 2-year period (Majczakowa et al., 1985). These tractor
drivers and feeder operators were in contact with insecticide
concentrations not exceeding 0.015 and 0.040 mg/m3, respectively, based
on samples periodically analyzed from breathing areas of workers. The
authors reported that up to 2 mg chlorpyrifos were recovered from the
workers' hands at various sampling intervals. Average potential exposure
at work to chlorpyrifos was estimated to be 0.373 mg/hour for the first
year of the study and 0.034 mg/hour for the second year. No signs of
toxicity or changes in blood ChE activity were observed.
Animals
Short-term Exposure
An oral LDJO of 152 mg/kg was reported for female mice and 169 mg/kg
for female rats given chlorpyrifos by intubation in soy bean oil (details of
the chlorpyrifos formulation were not provided) (Berteau and Deen,
1978). Oral LD50 values for male and female rats ranged from 118 to 245
mg/kg; no significant sex-related differences were observed (Gaines, 1969;
McCollister et al., 1974). The acute oral LD^ for male guinea pigs was
504 mg/kg, and no deaths were noted in male and female rabbits dosed
with 1,000 mg/kg (McCollister et al., 1974).
In a study conducted by Dow Chemical Company (1972), each of three
rhesus monkeys (sex not specified) was given a single oral dose of 3.5 mg
chlorpyrifos/kg. Erythrocyte ChE levels were 60% below pretreatment
levels at 4 hours post-dosing but increased to 66, SO and 82% of baseline
values at 8, 24 and 48 hours, respectively. Plasma ChE levels were more
severely affected and were only 6, 8, 14 and 30% of baseline values at the
respective sampling times tested above.
10
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Chlorpyritbs April 1992
Two rhesus monkeys (sex not reported) given a single oral dose of 2
mg/kg/day chlorpyrifos for 3 consecutive days showed no clinical signs of
toxicity (Dow Chemical Company, 1972). A sharp decrease (15 to 25% of
control values) in plasma ChE activity was observed 24 hours after the
initial dosing. An additional 5% reduction was observed after
administration of the second and third doses. Erythrocyte ChE activity
levels dropped only slightly during the first day; greater reductions (to 60
to 65% of control levels) were observed on the second and third days of
the study.
In a range-finding study conducted by Dow Chemical Company (1972),
pairs of beagle dogs consuming a diet containing 0.6 ppm (0.015
mg/kg/day, based on Lehman, 1959) chlorpyrifos for 12 days showed no
changes in either plasma or erythrocyte ChE activity. When the chemical
was administered for 28 days at a dietary concentration of 2 ppm (0.1
mg/kg/day), the plasma ChE activity in one female was reduced by 50%
within 7 days after the study began. In another study, dogs fed 6, 20 or
60 ppm (0.15, 0.5 and 1.5 mg/kg/day)..chlorpyrifos for 35 days showed
reduced plasma ChE activity to 42%, 25% and 17% of pretreatment
values, respectively; however, erythrocyte and brain ChE activities did not
change. From these two studies, it was concluded that the NOAEL was
0.015 mg/kg/day for dogs exposed orally to chlorpyrifos.
Symptoms of severe ChE inhibition developed in beagle dogs (two/sex/
group) fed 2,000 (50 mg/kg/day) or 600 (15 mg/kg/day) ppm chlorpyrifos
in the diet for 5 and 16 days, respectively (Dow Chemical Company, 1972;
conversions based on Lehman, 1959). These dogs were taken off their
respective diets and placed on a 200-ppm diet. Additional groups of dogs
consumed a 200-ppm (5-mg/kg/day) diet for up to 45 days or a 20- or 60-
pprn (0.5 or 1.5 mg/kg/day) diet for up to 88 days. Slowed growth was
observed in all males and in females consuming 200 ppm chlorpyrifos.
Plasma and erythrocyte ChEs were depressed in all groups of animals.
Brain ChE activity was decreased in both sexes receiving 200 ppm but
only in females consuming the 60-ppra diet. Gross and histological
examination of tissues was normal in all dogs. This study indentified a
brain ChE-depression NOAEL of 0.5 mg/kg/day and a LOAEL of 1.5
rag/kg/day for beagles of both sexes.
Acute dermal and inhalation exposures to chlorpyrifos (in 65% xylene)
were as toxic to mice and rats as were oral exposures. A dermal LD50
value of 202 mg/kg for rats was reported by Gaines (1969), and inhalation
LCjo values of 152 and 169 mg/kg were reported for female mice and rats,
respectively, by Berteau and Deen (1978).
11
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Chlorpyritbs April 1992
In a study by Berteau and Dean (1978), groups of 16 mature female
NAMRU mice (30 to 35 g) inhaled a 65% xylene aerosol cloud containing
the equivalent of 0.1 to 50 mg chlorpyrifos/kg for 27 to 50 minutes. Dose-
related decreases in plasma acetylcholinesterase were observed; enzyme
activity was 55% of predosing activity following exposure to 0.2 mg/kg and
less than 10% after exposure to 50 mg/kg.
Groups of 10 male and 10 female Sprague-Dawley rats that inhaled an
aerosol cloud containing 5 mg chlorpyrifos/L for an unspecified amount of
time exhibited lachrymation, slight nasal discharge and gasping during
exposure (Dow Chemical Company, 1972). Animals appeared normal
during the 14-day post-inhalation period, and postmortem examination of
tissues revealed no gross pathological changes.
Dermal/Ocular Effects
Chlorpyrifos (0.5 mL of a 24% solution) was applied to the intact and
abraded skin of six New Zealand albino rabbits (sex and age not
reported) (Dow Chemical Company, 1972). Animals were exposed to the
test material for 24 hours. Moderate to severe erythema developed on all
exposed areas; slight necrosis was observed on four of the intact areas and
five of the abraded areas. All exposed skin areas had some degree of
edema. Reactions of intact and abraded skin of three additional rabbits
exposed to chlorpyrifos (24% in solution) for 6 hours included slight
erythema, slight edema and slight necrosis within 10, 30 to 60, and 90 to
210 minutes, respectively.
Instillation of chlorpyrifos (0.1 mL of a 24% solution) into the
conjunctival sac of the right eye of six New Zealand albino rabbits (sex
and age not reported) produced conjunctival redness, iritis and corneal
injury in all treated eyes (Dow Chemical Company, 1972).
No skin or eye irritation developed in any of the 40 adult male and
female mongrel dogs (number/sex not reported) or 85 puppies dipped
repeatedly in 0.0125, 0.025, 0.05 or 0.10% chlorpyrifos solutions (Dow
Chemical Company, 1972). Adults were dipped three to six times at 15-
or 30-day intervals; puppies (6 to 8 weeks old) were dipped up to three
times in the 0.025% solution but only once in the 0.05% solution.
Percutaneous injections of 1.0, 2.0 or 3.98 g chlorpyrifos (as a 25%
solution) into groups of four albino rabbits (sex not reported) induced
slight to moderate erythema, swelling, and necrosis (Dow Chemical
Company, 1972). One mid-dose rabbit died 3 days after exposure, and
. three high-dose animals died within 6 to 9 days post-dosing.
12
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Chlorpyrifos April 1992
Long-term Exposure
Groups of 20 albino rats (10/sex/dose) were maintained on diets
containing 10, 30, 100 or 300 ppm chlorpyrifos (approximately 0.5, 1.5, 5
or 15 mg/kg, respectively, based on Lehman, 1959) for 90 days (Dow
Chemical Company, 1972). Another group of rats that received 1,000
ppm (50 mg/kg) chlorpyrifos in the feed was included in this study, but
due to high mortality, this group was terminated after 28 days. Plasma
and erythrocyte ChE activity levels were depressed in a dose-related
manner, at 1,000 ppm the average plasma ChE activity for both sexes was
less than 1% of the control value. Brain ChE activity also was reduced to
about 30%, 20% and 10% of control values in animals consuming 100,
300 and 1;000 ppm chlorpyrifos, respectively. Exposure to 0.5 mg/kg/day
for 90 days caused a 3 to 7% reduction in brain ChE activity, and 1.5-
mg/kg/day doses reduced brain ChE activity by 19 to 22% after 90 days
(neither was significantly different from control values at p = 0.05). The
animals dosed at 1,000 ppm exhibited signs of severe ChE depression
(e.g., tremors, bloody noses, circling and backing, ulceration of the cornea
and nostrils), decreased food consumption, significant weight loss and
increased mortality. Rats consuming the 300-ppm feed experienced
tremors, slight diuresis and slight growth retardation. Consumption of the
three lowest doses produced no signs of toxicity. The NOAEL based on
reduced brain ChE activity was 0.5 mg/kg/day.
In a 91-day study conducted by Dow Chemical Company (1972), groups
of 20 albino rats (10/sex/dose) fed 3.0 or 10.0 mg chlorpyrifos/kg/day
exhibited reduced plasma and erythrocyte ChE activities (35 to 58% and
14 to 26% of control values, respectively) and showed slight to severe
signs of ChE inhibition and toxicity (e.g., hunched appearance, tremors,
weight loss). Rats consuming a 03- or 1.0-mg/kg/day diet had depressed
plasma and erythrocyte ChE levels. Male rats given 03 mg
chlorpyrifos/kg/day had reduced body weight gains. No adverse effects
were observed at the 0.03- or 0.1-mg/kg/day dose levels. Survival was not
affected at any exposure level, and ChE activities returned to normal
within 1 to 2 weeks after withdrawal of the test compound from the diet.
This study identified a NOAEL of 0.1 mg/kg/day and a LOAEL of 03
mg/kg/day (for male rats).
Albino rats (20/sex/group) consuming dietary levels of 0.03, 0.15 or 0.75
mg chlorpyrifos/kg/day for 6 months showed no significant clinical or
histological signs or symptoms of organophosphate poisoning (Dow
Chemical Company, 1972). Animals ingesting the high-dose feed
exhibited reduced plasma and erythrocyte ChE activities (i.e., 35 to 60%
and 50% of control values, respectively). Brain ChE activity was not
13
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Chlorpyrifos April 1992
affected at any treatment level. This study identified a NOAEL of 0.15
mg chlorpyrifos/kg/day.
A group of 20 Sprague-Dawley rats (4 weeks old. sexes combined) were
fed diets containing 100 ppm chlorpyrifos (5 mg/kg/day) for 1 year.
Seventeen animals also consumed a diet supplemented with both
chlorpyrifos (100 ppm) and corn oil (final concentration, 20%) (Buchet et
al., 1977; conversions based on Lehman, 1959). Control animals
consumed nonsupplemented or corn oil-only supplemented diets (19 and
18, respectively). All rats grew normally and had normal levels of
hormone sensitive lipase, lipoprotein lipase, serum fatty acids (free and
total), serum glycerol and serum cholesterol (total). Total glycerol and
cholesterol content of the aorta in test animals were also comparable to
controls, but total aortic fatty acids were increased in animals consuming
both chlorpyrifos and 20% oil in the diet. Total blood ChE activity was
reduced by 40% in the chlorpyrifos-exposed group on the normal feed
and by 60% in those animals on the fat-enriched regimen. A LOAEL of
5 mg/kg/day, based on reduced ChE activity and increased total aortic
fatty acids, was identified in this study.
In a study conducted by McCollister et al. (1974), groups of 7-week-old
Sherman rats (25/sex/dose) fed a diet containing 1.0 or 3.0 mg
chlorpyrifos/kg/day for 2 years exhibited significantly (p < 0.05) depressed
plasma ChE activity levels. Erythrocyte ChE activities were depressed (p
< 0.05) by approximately 67% and 85% of control values in rats fed the
1.0- and 3.0-mg/kg/day diets, respectively; brain ChE activity was
significantly (p < 0.05) reduced (to about 57% of controls) in the high-
dose animals only. Effects on ChE activity were reversible when
consumption of a chlorpyrifos-free diet was resumed. ChE activity in
animals fed 0.1 mg/kg/day was comparable to control values. No clinical
signs of toxicity were observed at any dose. A NOAEL of 0.1 rag
chlorpyrifos/kg/day based on plasma ChE activity, and a NOAEL of 3.0
mg/kg/day based on systemic effects were established in this study.
Groups of three or. four rhesus monkeys (males and females combined)
that received chlorpyrifos by gavage at doses of 0.08, 0.4 or 2.0 mg/kg/day
for 6 months showed no significant compound-related clinical effects
compared to control animals (Dow Chemical Company, 1972). The only
evidence of exposure to chlorpyrifos was reduced plasma and erythrocyte
ChE activities in the mid- and high-dose monkeys (significance levels not
reported). Midbrain and cerebrum ChE values were not affected in any
group. Histological examination revealed that the liver and kidney
showed no abnormalities in any of the animals. A NOAEL of 0.08
14
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Chlorpyrifos April 1992
mg/kg/day was identified, based on the absence of inhibition of ChE at
this dose.
Plasma ChE activity was depressed by 40 to 75% of pretest or control
values in groups of three or four male and female beagle dogs (aged 10 to
11 months) given 0.1, 1.0 or 3.0 mg chlorpyrifos/kg/day in the diet for 1 to
2 years (McCollister et al., 1974). Brain ChE activity was slightly
depressed (by 8 to 21% of the control value) at the highest dose level;
however, these decreases were not statistically significant (p < 0.05).
Both plasma and brain ChE activities returned to normal when dosed
animals were placed on the control diet. No significant health effects
were observed at any dose (0.01 and 0.03 mg/kg/day) using the following
criteria: mortality, body weight, food intake, hematological and clinical
chemistry parameters, organ weight, tumor incidence, and gross and
histopathologicar examination of tissues. The only notable difference was
a statistically significant (p < 0.05) increase in the mean liver-to-body
weight ratio of high-dose males administered chlorpyrifos for 2 years.
The NOAEL for dogs identified in this study was 0.03 mg chlorpyrifos/kg/
day based on plasma ChE activity levels and was 3.0 mg/kg/day based on
systemic effects.
Reproductive Effects
In a three-generation reproduction study, groups of 15 male and 15
female Sprague-Dawley albino rats that received up to 1.0 mg
chlorpyrifos/kg/day in the feed showed no adverse reproductive or
postnatal effects, as judged by fertility, gestation, viability and lactation
indices (Dow Chemical Company, 1972). Litter size, pup weight and sex
ratios of offspring from treated rats also were unaffected by exposure to
the test compound. In addition, ingestion of chlorpyrifos (0.03, 0.1 or 0.3
. mg/kg/day by the first-generation rats and 0.1, 03 or 1.0 mg/kg/day by the
second- and third-generation rats) had no adverse effects on survival,
body weight gains and food consumption of either male or female
parents. Third-generation rats (both sexes) consuming the 1.0-mg/kg/day
diet had depressed plasma and erythrocyte ChE activities, as did females
given feed containing 0.3 mg chlorpyrifos/kg/day. It was concluded that
the reproductive NOAEL from this study is 0.1 mg/kg/day.
Dow Chemical Company (1972) reported that multiple exposures to
chlorpyrifos (0.025, 0.05 or 0.10% solutions) via dipping produced no
maternal toxicity in mongrel dogs and had no effect on gestation or
parturition. Twelve dogs were dipped one to four times at 15- or 30-day
intervals. Animals were cither not pregnant or up to 58 days pregnant at
the time of the first dip (average gestation period, 63 ฑ 7 days).
15
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Chlorpyrifos April 1992
Everett (1982) studied the effects of dermal applications of a test material
consisting of 43.2% chlorpyrifos on 185 Holstein bulls. No other dosing
information was provided. In the 172 bulls that did not exhibit severe
toxic effects, semen production initially was depressed but returned to
normal within 6 months post-treatment. Normal semen production was
measured for at least 12 months post-treatment. Six bulls became very
sick following exposure to chlorpyrifos, and their semen production did
not return to normal within 12 months. The remaining seven bulls died
after being treated with the insecticide.
Developmental Effects
In a developmental study conducted by Deacon et al. (1980), 40 to
47 pregnant CF-1 mice were dosed, by gavage, with 1, 10 or 25 mg
chlorpyrifos/kg/day on gestation days 6 through 15. Fifty-one animals
were used in the control group. Severe maternal toxicity, including
mortality, clinical signs of ChE inhibition, and significant (p < 0.05)
decreases in maternal body weight gains and food and water consumption
were reported at 25 mg/kg/day. Plasma and erythrocyte ChE levels were
significantly (p < 0.05) reduced at all dose levels tested when compared
with controls. Developmental toxic effects were reported at 25 mg/kg/day.
The findings included significant reductions in fetal body weight and
crown-rump lengths. Exencephaly was noted in four fetuses from three
litters of mice dosed with 25 mg/kg/day (nonsignificant) and in five fetuses
from five litters in the 1-mg/kg/day dose group (significant at p < 0.05).
Significant increases (p < 0.05) in the incidences of delayed ossification of
the skull and sternebrae were also reported at the highest dose level. The
incidence of sternebrae abnormalities was high (p < 0.05) among fetuses
bom to dams in the lowest dose group (1 mg/kg/dayj. These results were
not repeatable, however, when additional groups of 35 to 41 CF-1 mice
were given 0, 0.1, 1 or 10 mg chlorpyrifos/kg by gavage on gestation days
6 through 15 (Deacon et al., 1980). Thus, although the first study
suggests a fetal LOAEL of 1 mg/kg/day based on reduced ChE activity
and adverse developmental effects, the data are equivocal due to the lack
of any significant response in the second group of test animals. From
both studies combined, the NOAEL appears to be 0.1 mg/kg/day.
. Chlorpyrifos was not teratogenic in rats, as judged by external, skeletal
and visceral examination of second-litter fetuses from third-generation
Sprague-Dawley female rats administered the insecticide by gavage at 1.0
mg/kg/day on gestation days 6 through 15 (Dow Chemical Company,
1972). Parental females received chlorpyrifos in the diet at levels of 0.1,
0.3 or 1.0 mg/kg/day for the rest of their lives. Maternal weight gains and
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Chlorpyrifos April 1992
food consumption, corpora lutea, resorptions, fetus viability, pup weights
and sex ratios also appeared to be unaffected.
In the dog reproduction study by Dow Chemical Company (1972), 58 of
85 (68%) pups born to mongrel dogs that had been dipped repeatedly in
chlorpyrifos (0.025, 0.05 or 0.10% solutions) died before 8 weeks of age.
Only one puppy death was attributed to chlorpyrifos (due to symptoms
suggestive of organophosphate toxicity). No other details were given.
Mutagenicity
Chlorpyrifos was not mutagenic in Salmonella typhimurium or Escherichia
coli either in the absence or presence of both mammalian and plant
microsomes (Gentile et al., 1982; Moriya et al., 1983; Shirasu et al., 1976,
Waters et al., 1982). Similarly, chlorpyrifos did not induce reverse
mutations in Zea mays (Gentile et al., 1982; Seehy et al., 1984) or cause
sex-linked recessive lethal mutations in Drosophila melanogaster (Waters
et al., 1982).
Evidence of in vivo induction of micronucleated polychromatic
erythrocytes in mouse bone marrow cells following intraperitoneal or oral
administration of chlorpyrifos have been reported (Amer and Fahmy,
1982). Chlorpyrifos induced mitotic suppression and increased the
frequency of chromosome aberrations in Vtcia faba (Abdou and Abdei-
Wahab, 1985). It produced clastogenic effects in barley (Kaur and
Grover, 1985).
Chlorpyrifos, without exogeneous metabolic activation, was genotoxic in
DNA polymerase I-deficient E. coli and recombination-deficient S.
typhimurium (Waters et al., 1982). However, inconsistent results have
been seen in the Bacillus subtilis rec-assay. Waters et al. (1982) reported
a positive response, but Shirasu et al. (1976) and Kada et al. (1980) found
no genotoxic activity. Chlorpyrifos was not recombinogenic in
Saccharomyces cerevisiae D3 either with or without rat liver microsomes
(Waters et al., 1982) or in S. cerevisiae D4 with or without mammalian and
plant microsomes (Gentile et al., 1982). Unscheduled DNA synthesis was
not increased in chlorpyrifos-treated human lung fibroblast (Waters et al.,
1982). Chlorpyrifos was judged negative for the induction of sister
chromatid exchanges in Chinese hamster ovary cells, chick embryos
(Muscarella et al., 1984). and the LAZ-007 human lymphoid cell line
(Sobti et al., 1982).
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Chlorpyrifos April 1992
Carcinogenic! tv
Chlorpyrifos was not tumorigenic in a chronic feeding study in which 7-
week-old Sherman rats were administered up to 3.0 mg chlorpyrifos/kg/
day for 2 years (McCollister et al., 1974). Thirty rats/sex/dose were used;
an additional five to seven rats/sex/group killed for interim gross and
microscopic pathological examinations were normal throughout the
experimental period. In this study too few animals were included to fully
assess the carcinogenicity of Chlorpyrifos in rats. Since there was no
evidence of toxicity at tested doses, a minimal toxic dose (MTD) may not
have been used.
No excess tumors developed in groups of 10- to 11-month-old beagle dogs
(three or four/sex/group) fed 0.01, 0.03, 0.1, 1.0 or 3.0 mg chlorpyrifos/kg
in the diet for 1 to 2 years (McCollister et al., 1974). In addition, gross
and microscopic examination of tissues was normal for all dose levels
throughout the experimental period. However, this study is considered of
limited usefulness in providing data to assess the carcinogenic potential of
Chlorpyrifos: the study was relatively short (i.e., less-than-lifetime for
dogs), and the animals may not have been tested at MTD.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day, Longer-term,
and Lifetime exposures if adequate data are available that identify a sensitive noncarcinogenic
end point of toxicity. The HAs for noncarcinogenic toxicants are derived using the following
formula:
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level in rag/kg bw/day.
BW = assumed body weight of a child (10 kg) or an adult (70 kg).
UF = uncertainty factors (10. 100, 1.000, or 10,000), in accordance with EPA or
NAS/OW guidelines.
_ L/day = assumed daily water consumption of a child (1 L/day) or an adult 2 L/
day).
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Chlorpyrifos April 1992
One-day Health Advisory
No suitable information was found in the available literature for determining the One-
day Health Advisory (HA) for chlorpyrifos. The Ten-day HA for a child, 30 /xg/L, calculated
below, is recommended for use as a conservative estimate for a 1-day exposure to chlorpyrifos.
Ten-day Health Advisory
The human oral-exposure study by Dow Chemical Company (1972) has been selected to
serve as the basis for the Ten-day HA because it contains adequate human data of appropriate
length. In this study, groups of four healthy adult men were administered chlorpyrifos in
capsule form at doses of 0, 0.014, 0.03 or 0.10 mg/kg for 28, 28, 21 or 9 days, respectively.
Adverse health effects were observed only in the high^dose (0.10 mg/kg) group; at this exposure
level, plasma ChE activity was reduced by approximately 65% when compared to control values.
One individual in this group also experienced a runny nose and blurred vision. A NOAEL of
0.03 mg/kg/day was identified from this study.
Another study by Dow Chemical Company (1972), in which a NOAEL of 0.015 mg/kg
was reported for beagle dogs consuming chlorpyrifos in the feed for 12 days, was also
considered for the calculation of the Ten-day HA. However, this value was not used because
the available human data were within one order of magnitude of the animal data. Therefore,
the results of this dog study support the human data used to calculate the Ten-day HA but are
not the most appropriate for the derivation of this HA.
The Ten-day HA for the 10-kg child is calculated as follows:
Ten-day HA - (ฐ^ (?^yff ^ - 0-03 mg/L - (rounded to 30 Mg/L)
where:
0;03 mg/kg/day = NOAEL, based on the absence of decreased plasma ChE activity in male
human subjects exposed to chlorpyrifos via the oral route for 21 days
(Dow Chemical Company, 1972).
10 kg = assumed weight of child.
10 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines
in which a NOAEL from a human study is employed.
1 L/day = assumed water consumption of 10-kg child.
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Chlorpyrifos April 1992
Longer-term Health Advisory
The study by Dow Chemical Company (1972) in which humans received daily oral doses
or" chlorpyrifos tor up to 28 consecutive days has been selected to serve as the basis for the
Longer-term HA because it contains human data that identify a NOAEL and a LOAEL.
Although several adequate subchronic animal studies were available, data from these studies
indicate that humans are more sensitive to chlorpyrifos following oral administration than other
species. Using human data would, therefore, allow for a more conservative estimate of a
Longer-term HA for a 10 kg child. The following studies used plasma and erythrocyte ChE
activity levels as the basis for NOAEL and LOAEL values. Two 3-month feeding studies with
rats identified NOAELs of 0.1 and 1.5 mg/kg/day; respective LOAELs were 0.3 and 5 mg/kg/day
(Dow Chemical Company, 1972). A 6-month study in which rats received chlorpyrifos in the
diet identified a NOAEL of 0.15 mg/kg/day and a LOAEL of 0.75 mg/kg/day (Dow Chemical
Company, 1972). Finally, monkeys exposed via oral gavage to chlorpyrifos for 6 months showed
no adverse effects at 0.08 mg/kg/day doses but had reduced ChE activities at 0.4 mg/kg/day
(Dow Chemical Company, 1972). The human study, also conducted by Dow Chemical Company
(1972), identified a NOAEL of 0.03 mg/kg/day and a LOAEL of 0.1 mg/kg/day based on the
absence or presence of reduced plasma ChE activity following administration of chlorpyrifos (in
capsules) to groups of four healthy male adults for 9 and 21 days, respectively.
The Longer-term HA for the 10-kg child is calculated as follows:
Longer-term HA = (0-03 nig/kg/day) (10 kg) = QM mg/L (rounded t(j 3Q
(10) (1 L/day)
where:
0.03 mg/kg/day = NOAEL, based on the absence of decreased plasma ChE activity in male
human subjects exposed to chlorpyrifos via the oral route for 21 days
(Dow Chemical Company, 1972).
10 kg = assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines
in which a NOAEL from a human study is employed.
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 = (ft03 f ^ <70 kg) . 0.105 mg/L (rounded to 100 Mg/L)
(10) (2 L/day)
20
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Chlorpyrifos April 1992
where:
0.03 mg/kg/day = NOAEL, based on the absence of decreased plasma ChE activity in male
human subjects exposed to chlorpyrifos via the oral route for 21 days
(Dow Chemical Company, 1972).
70 kg = assumed body weight of an adult.
10 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines
in which a NOAEL from a human study is employed.
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 estimate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious health effects during 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 in drinking water alone 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 known, probable, or possible human carcinogen,
according to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then
caution must be exercised in making a decision on how to deal with possible lifetime exposure
to this substance. For human (A) or probable human (B) carcinogens, a Lifetime HA is not
recommended. For possible human carcinogens (C), an additional 10-fold safety factor is used
to calculate the Lifetime HA. The risk manager must balance this assessment of carcinogenic
potential and the quality of the data against the likelihood of occurrence and significance of
health effects related to noncarcinogenic endpoints of toxicity. To assist the risk manager in
this process, drinking water concentrations associated with estimated excess lifetime cancer risks
over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of water/day are
provided in the Evaluation of Carcinogenic Potential section.
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Chlorpyrifos April 1992
The oral-exposure study with humans by Dow Chemical Company (1972) has been
selected to serve as the basis for the Lifetime HA because it contains adequate human exposure
data that are supported by animal data. A NOAEL of 0.03 mg/kg/day for humans was
identified in this study. Three animal studies also were considered for the calculation of the
Reference Dose (RfD), Drinking Water Equivalent Level (DWEL), and Lifetime HA. The
chronic (2-year) feeding study in rats (McCollister et al., 1974) was judged unacceptable due to
insufficient numbers of animals (25/sex/dose with an additional 5 to 7/sex/dose for interim
sacrifices). The other animal studies, which included the reproduction study with rats by Dow
Chemical Company (1972) and the chronic feeding study with dogs by McCollister et al. (1974),
were considered adequate. However, the human oral-exposure study by Dow Chemical
Company (1972) was judged most appropriate for the calculation of the Lifetime HA for
chlorpyrifos. Although this human study was coreclassified as supplementary because only four
males/dose were used, when considered with the available experimental data in animals, the
NOAEL for ChE inhibition in humans (0.03 mg/kg/day) appeared comparable to that in rats
(0.1 mg/kg/day) and dogs (0.03 mg/kg/day).
Using the human study (Dow Chemical Company, 1972), the Lifetime HA is derived as
follows:
Step 1: Determination of the RfD
m (0.03 mg/kg/day) =
(10) * & '
where:
0.03 mg/kg/day = NOAEL, based on the absence of decreased plasma ChE activity
in male human subjects exposed to chlorpyrifos via the oral route
for 21 days (Dow Chemical Company, 1972).
10 = uncertainty factor, chosen in accordance with EPA or NAS/OW
guidelines in which a NOEL from a human study is employed.
Step 2: Determination of the DWEL
T/TrT (0.003 mg/kg/day) x (70 kg) n ,nc , . , t inn n.
DWEL = .1 ?ฐLLL i ฐL = 0.105 mg/L (rounded to 100 ug/L)
(2 L/day)
where:
0.003 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
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Chlorpyritbs April 1992
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime HA
Lifetime HA = (0.1 mg/L) (20%) = 0.02 mg/L (20 /zg/L)
where:
0.105 mg/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
No evidence of carcinogenicity was found in rats or dogs fed up to 3.0 mg
chlorpyrifos/kg/day for 1 to 2 years (McColIister et al., 1974).
The International Agency for Research on Cancer (IARC) has not evaluated the
carcinogenic potential of chlorpyrifos.
Applying the criteria in EPA's guidelines for assessment of carcinogenic risk
(U.S. EPA, 1986), chlorpyrifos may be classified in Group D: not classifiable.
This category is for agents with inadequate animal evidence for carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE. AND STANDARDS
The American Conference of Governmental Industrial Hygienists recommends a
Threshold Limit Value-Time-Weighted Average of 0.2 rag/m3 and a Short-term
Exposure Limit of 0.6 mg/m3 for dermal exposures (ACGEH, 1988).
The Food and Agriculture Organization/World Health Organization (FAO/
WHO, 1984) Acceptable Daily Intake for chlorpyrifos is 0.01 mg/kg/day (Gartrell
et al., 1985).
VII. ANALYTICAL METHODS
Chlorpyrifos is one of the phosphorus-containing pesticides. The relative ease
with which this pesticide can be monitored by element-specific detectors has
usually led to its inclusion in pesticide monitoring studies. Chlorpyrifos can be
analyzed by Methods 622 (U.S. EPA, 1982) and 507 (U.S. EPA, 1988). In both
methods a liter of sample is extracted with methylene chloride, the solvent is then
exchanged for hexane or methyl tertiary butyl ether. Analysis is by an element-
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Chlorpyrifos April 1992
specific thermionic nitrogen-phosphorus detector, which allows even relatively
"dirty" samples to be analyzed with no clean-up. The estimated detection limit
for this residue is 0.3 /ig/L.
VIII. TREATMENT TECHNOLOGIES
There is presently no information available describing treatment technologies
capable of removing chlorpyrifos from contaminated drinking water supplies.
Chlorpyrifos may be amenable to removal by activated carbon adsorption due to
its solubility.
Chlorpyrifos is probably not amenable to removal by aeration on the basis of its
Henry's Coefficient value.
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Chlorpyrifos April 1992
IX. REFERENCES
Abdou, RiF. and M.A. Abdel-Wahab. 1985. Cytological and developmental effects of certain
insecticides in Vicia faba. Int. Pest Control. 27:123-125.
ACGIH. 1988. American Conference of Governmental Industrial Hygienists. Threshold limit
values and biological exposure indices for 1987-1988. 2nd printing. Cincinnati, OH:
ACGHI, p. 16.
Amer, S.M. and M.A. Fahmy. 1982. Cytogenetic effects of pesticides. I. Induction of
micronuclei in mouse bone marrow by the insecticide Dursban. Mutat. Res. 101:247-
255.
Bauriedel, W.R., R.L. McKellar and J.H. Miller. 1976. A rotational crop study using UC-
iabeled chlorpyrifos: GHrC 876. Dow Chemical U.S.A., Midland, MI." Unpublished
study. Contract No. 464-448. Publication No. CDL:224341-A.
Berteau, P.E. and W.A. Deen. 1978. A comparison of oral and inhalation toxicities of four
insecticides to mice and rats. Bull. Environ. Contam. Toxicol. 19:113-120.
Bidlack, H.D. 1979. Degradation of chlorpyrifos in soil under aerobic, aerobic/anaerobic and
anaerobic conditions: method GH-C 1258. Dow Chemical U.S.A., Midland, MI."
Unpublished study. Contract No. 464448. Publication No. CDL:241547-A.
Buchet, J.P., R. Lauwerys and R. Roels. 1977. Long term exposure to organophosphorous
pesticides and lipid metabolism in the rat. Bull. Environ. Contam. Toxicol. 17:75-183.
Chapman, R.A. and C.R. Harris. 1980. Persistence of chlorpyrifos in a mineral and an organic
soil. J. Environ. Sci. Health B15:39-46.
Deacon, M.M., J.S. Murray, M.K. Pilny, K.S. Rao, D.A. Dittenber, T.R. Hanley, Jr. and J.A.
John. 1980. Embryotoxicity and fetotoxicity of orally administered chlorpyrifos in mice.
Toxicol. Appl. Pharmacol. 54:31-40.
Dow Chemical Company.' 1972. Product literature. Dowco 179. No. 112118. Washington,
DC: U.S. Environmental Protection Agency.
Dunphy, J., M. Kesselbrenner, A. Stevens, B. Vlec and R.J. Jackson. 1980. Pesticide poisoning
in an infant - California. MMWR. 29:254-255.
"Confidential Business Information. Submitted to the EPA Office of Pesticide Programs.
25
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Chlorpyritbs April 1992
Eliason, D.A., M.F. Cranmer, D.L. von Windeguth, J.W. Kilpatrick, I.E. Suggs and H.F. Schoof.
1969. Dursban premises applications and their effect on the cholinesterase levels of
spraymen. Mosq. News 29:591-595.
Everett, R.W. 1982. Effect of Dorsban 44 on semen output of Holstein bulls. J. Dairy Sci.
65:1781-1794.
FAO/WHO. 1984. Food and Agriculture Organization of the United Nations/World Health
Organization. Guide to codex recommendations concerning pesticide residues. Part 2:
Maximum limits for pesticide residues. Rome: FAO/WHO.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol. 14:515-5344.
Gartrell, M.J., J.C. Craun, D.S. Podrebarac and E.L. Gunderson. 1985. Pesticides, selected
elements, and other chemicals in adult total diet samples, October 1979 - September
1980. J. Assoc. Off. Anal. Chem. 68:1,184/1,197.
Gentile, J.M., GJ. Gentile, J. Bultman, R. Sechriest, E.D. Wagner, and MJ. Plewa. 1982. An
evaluation of the genotoxic properties of insecticides following plant and animal
activation. Mutat. Res. 101:19-29.
Hayes, A.L., R.A. Wise and F.W. Weir. 1980. Assessment of occupational exposure to
organophosphates in pest control operators. Am. Ind. Hyg. Assoc. J. 41:568-575.
Hodgson, MJ., G.D. Block and D.K. Parkinson. 1986. Organophosphate poisoning in office
workers. J. Occup. Med. 28:434-437.
Kada, T., K. Hirano and Y. Shirasu. 1980. Screening of environmental chemical mutagens by
the rec-assay system with Bacillus subtilis. In: de Serres, F.T. and A. Hollaender, eds.
Chemical mutagens: Principles and methods for their detection. Vol. VI. New York:
Plenum, pp. 149-174.
Kaur, P. and I.S. Grover. 1985. Cytologica! effects of some organo-phosphorus pesticides. I.
Mitotic effects. Cytologia 50:187-197.
Kenaga, E.E. 1980. Correlation of bioconcentration factors of chemicals in aquatic and
terrestrial organisms with their physical and chemical properties. Environ. Sci. Technol.
14:553-556.-
Lehman, A. 1959. Appraisal of the safety of chemicals in foods, drugs, and cosmetics.
Association of Food and Drug Officials of the United States.
Lores, E.M. and D.E. Bradway. 1977. Extraction and recovery of organophosphorus
metabolites from urine using an anion exchange resin. J. Agric. Food Chem. 25:75-79.
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Chlorpyritbs April 1992
Lotti. M., A. Moretto, R. Zoppellari, R. Dainese, N. Rizzuto and G. Barusco. 1986. Inhibition
of lymphocytic neuropathy target esterase predicts the development of organophosphate-
induced delayed polyneuropathy. Arch. Toxicol. 59:176-179.
Ludwig, P.D., DJ. Kilian, HJ. Dishburger and H.N. Edwards. 1970. Results of human
exposure to thermal aerosols containing Dursban" insecticide. Mosq. News 30:346-354.
Madder, DJ. 1977. The disappearance from efficacy in and effect on nontarget organisms of
diflubenzuron, methoprene and chlorpyrifos in a lentic ecosystem. Master's thesis.
University f Manitoba.
Majczakowa, W., H. Badach, Z. Soczewinska-Klepacka and A. Molocznik. 1985. Evaluation of
the conditions of work while using pesticide in the granular form - Dursban 5G.
Medycyna Wiejska. 20(4):269-278. (In Polish; summary in English.)
McCollister, S.B., R.J. Kociba, C.G. Humiston, D.D. McCollister and PJ. Gehring. 1974.
Studies of the acute and long-term oral toxicity of chlorpyrifos (0,0-diethyl-0(3,5,6-
trichloro-2-pyridyl)phosphorothioate). Food Cosmet. Toxicol. 12:46-61.
Miles, J.R.W., C.M. Tu and C.R. Harris. 1979. Persistence of eight organo phosphorus
insecticides in sterile and nonsterile mineral and organic soils. Bull. Environ. Contam.
Toxicol. 22:312-318.
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.
Mostafa, I.Y., Y.M. Adam and S.M.A.D. Zayed. 1983. Bioalkylation of nucleic acids in mice
by insecticides. I. Alkylation of liver RNA and DNA by chlorpyrifos. Z. Naturforsch.
38c:461-464.
Muscarella, D.E., J.F. Keown and S.E. Bloom. 1984. Evaluation of the genotoxic and
embryotoxic potential of chlorpyrifos and its metabolites in vivo and in vitro. Environ.
Mutagen. 6:13-23.
Nelson, J.H. and E.S. Evans, Jr. 1973. Field evaluation of the larvicidal effectiveness, effects on
nontarget species and environmental residues of a slow-release polymer formulation of
chlorpyrifos: March-October 1973. Study No. 44-022-73/75. U.S. Army Environmental
Hygiene Agency.
Nolan, R.J., D.L. Rick, N.L. Freshour and J.H. Saunders. 1984. Chlorpyrifos:
Pharmacokinetics in human volunteers. Toxicol. Appl. Pharmacol. 73:8-15.
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Chlorpyrifos April 1992
Schaefer, C.H. and E. Dupras, Jr. 1969. The effects of water quality, temperature and light on
the stability of organophosphorus larvicides used for mosquito control. Proc. Papers
37th Ann. Conf. Calif. Mosq. Control Assoc., pp. 67-75.
Schaefer, C.H. and E.F. Dupras, Jr. 1970. Factors affecting the stability of Dursban in polluted
waters. J. Econ. Entomol. 63:701-705.
Sears, M.K and R.A. Chapman. 1979. Persistence and movement of four insecticides applied
to turfgrass. J. Econ. Entomol. 72:272-274.
Seehy, M.A., M. Moussa and E. Badr. 1984. Induction of reverse mutation in waxy locus of
Zea mays pollen grain by pesticides. Egypt. J. Genet. Cytol. 13:137-142.
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.
Smith, G.N., B.S.. Watson and F.S. Fischer. 1967. Investigations on Dursban insecticide.
Metabolism of [J6Cl]0,0-diethyl 0-3,5,6-trichloro-2-pyridyl phosphorothioate in rats. J.
Agric. Food Chem. 15:132-138.
Sobti, R.C., A. Krishan and C.D. Pfaffenberger. 1982. Cytokinetic and cytogenetic effects of
some agricultural chemicals on human lymphoid cells in vitro: organophosphates.
Mutat. Res. 102:89-102.
Sultatos, L.G., L.G. Costa and S.D. Murphy. 1981. The role of glutathione in the
detoxification of chlorpyrifos and methyl chlorpyrifos in mice. Pharmacologist 23:214.
Sultatos, L.G., L.G. Costa and S.D. Murphy. 1982. Factors involved in the differential acute
toxicity of the insecticides chlorphyrifos and methyl chlorpyrifos in mice. Toxicol. Appl.
Pharmacol. 65:144-152. '
Sultatos, L.G., L.D. Minor and S.D. Murphy. 1985. Metabolic activation of phosphorothioate
pesticides: Role of the liver. J. Pharmacol. Exp. Then 232:624-628.
Sultatos, L.G. and S.D. Murphy. 1983. Hepatic microsoraal detoxification of the
organophosphates paraoxon and chlorpyrifos in the mouse. Drug Metab. Dispos. 11:232-
238.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Method 622 - The determination of
organophosphorus pesticides in industrial and municipal wastewater. Cincinnati, OH:
Environmental Monitoring and Support Laboratory. January.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Fed. Reg. 51(85):33992-34003 September 24.
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Chlorpyrifos April 1992
U.S. EPA. 1988. U.S. Environmental Protection Agency. Method 507--The determination of
nitrogen- and phosphorus-containing pesticides in water by gas chromatography with a
nitrogen-phosphorus detector. Cincinnati, OH: Environmental Monitoring and Support
Laboratory. December.
Waters, M.D., S.S. Sandhu, V.F. Simmon, K.E. Mortelmans, A.D. Mitchell, T.A. Jorgenson,
D.C.L. Jones, R. Valencia and N.E. Garrett. 1982. Study of pesticide genotoxicity.
Basic Life Sci. 21:275-326.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The Merck indexAn
encyclopedia of chemicals, drugs, and biologicals, 10th ed. Rahway, NJ: Merck and
Company, Inc., pp. 309-310.
Worthing, C.R. and S.B. Walker, eds. 1987. The pesticide manual, 8th ed. Lavenham, Suffolk:
The British Crop Protection Council, pp. 179-180.
29
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EPA 0553
RX000027511
April 199ฃ
ISOPHORONE
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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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ISOPHORONE April 1997.
II. GENERAL INFORMATION AND PROPERTIES
. CAS No. 78-59-1
Structural Formula
Svnonvms (CAS, 1988)
3,5,5-Trimethyl-2-cyclohexen-l-one; l,l,3-Trimethyl-3-cyclohexen-5-one;
Trimethylcyclohexenone, Isoacetophorone
Uses (NIOSH, 1978)
Isophorone is used primarily as a solvent for many types of lacquers, including vinylic
coating resins. It is also used as a solvent/cosolvent for polyvinyl and nitrocellulose
resins, pesticides, herbicides, fats, oils and gums. Isophorone is used as a chemical
intermediate in the manufacture of other solvents and plant growth retardants. It
has also been used as a repellent to stop woodpeckers from damaging utility poles.
Properties (Amoore and Hautala, 1983; Union Carbide, 1968; Veith et al., 1980; Hawley, 1981)
Chemical Formula
Molecular Weight 138.21
Physical State Liquid
Boiling Point 215.28C
Melting Point -8.1ฐC
Density (20ฐC) 0.923 g/mL
Vapor Pressure (25ฐC) 0.44 mm Hg
Specific Gravity 0.923
Water Solubility (20ฐC) 12 g/L
Log Octanol/Water Partition 1.67
Coefficient (log K,,)
Taste Threshold
Odor Threshold (air) 0.2 ppm (peppermint-like)
Conversion Factors (25ฐC) ppm = 5.65 mg/m3
mg/m3 = 0.18 ppm
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ISOPHORONE April
Occurrence
Production figures for isophorone were not available. The greatest potential for
exposure is probably in the workplace. NIOSH (1978) estimated that over
1.5 million persons were occupationally exposed to isophorone, primarily via
inhalation and dermal contact.
Isophorone has been detected in drinking water. The highest concentration of
isophorone reported in finished drinking water was 9.5 /ig/L (U.S. EPA, 1975).
Lesser concentrations of isophorone have been found in drinking water in Cincinnati
(0.02 /ig/L) and New Orleans (1.5 to 2.9 /ig/L) (U.S. EPA, 1974). Trace quantities
(<0.01 ppb) of isophorone have been reported in the Delaware River near an
industrial area (Sheldon and Hites 1978), in the wastewater from a tire
manufacturing plant (Jungclaus et al., 1976) and in effluents from latex and chemical
plants (Shackelford and Keith, 1976).
Environmental Fate
Little has been published with regard to the environmental fate of isophorone. In
aqueous solutions, isophorone is converted by sunlight into three different tricyclic
diketodimers (Jennings, 1965). The significance of this reaction in reducing the
concentration of isophorone in surface water is unknown. Isophorone is degraded'by
microorganisms in both domestic wastewater and in synthetic saltwater (Price, et al.,
1974).
III. PHARMACOKINETICS
Absorption
No data were found on how much isophorone is absorbed by any route of
administration in animals or humans. However, evidence of systemic toxicity
following oral administration indicates that absorption does occur.
Distribution
No data were found on the distribution of isophorone to specific tissues by any route
of exposure in humans or animals.
Metabolism
Dutertre-Catella et al. (1978) identified metabolites in Wistar rats and New Zealand
rabbits that received a single oral dose of isophorone at 1 g/kg bw in olive oil by
gavage. Metabolites included:
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ISOPHORONE April 199Z
5,5-dimethyl-2-cyclohexen-l-one-3-carboxylic acid, thought to be formed from
methyl oxidation.
isophorol (3,5,5-trimethyl-2-cyclohexen-l-ol), found as the glucuronide conjugate
and formed by reduction of the ketone.
- dihydroisophorone (3,5,5-trimethylcyclohexanone) resulting from the
hydrogenation of the cyclohexene double bond.
Excretion
Following oral administration of isophorone (1 g/kg bw), rabbits partially eliminated
isophorone unchanged in expired air and in the urine. The remainder was oxidized
to a carboxylic acid (oxidation of methyl group) or reduced to isophorol (ketone
reduction), which was eliminated in the urine as a glucuronide conjugate (Patty,
1982).
No other information on excretion of isophorone and its metabolites'in humans or
animals was found in the literature.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Isophorone vapors produce irritation to the eyes, nose and throat of unacclimatized
subjects after 15-minute exposures to 25 ppm (125 mg/m3) (Silverman et al., 1946).
The majority of the 12 volunteers were not uncomfortable at 10 ppm isophorone (50
mg/m3).
Union Carbide (1963) reported that volunteers exposed to isophorone (via
inhalation) found levels of 200 ppm (1,000 mg/m3) to be intolerable even for
exposure durations of 1 minute, as were exposures of 40 ppm (200 mg/m3) for 4
minutes. Subjects reported eye, nose and throat irritation, headaches, dizziness and
nausea at concentrations of 40 and 85 ppm (200 and 425 mg/m3) and faintness,
drunkenness and a feeling of suffocation at levels of 200 and 400 ppm (1,000 mg/m3
and 2,000 mg/m3) isophorone.
Silverman et al. (1946) reported that unacclimatized persons (12 males and 12
females) found 25 ppm of isophorone (15 minutes) caused irritation of eyes, nose
and throat.
N1OSH (1978) reported that occupational workers complained of fatigue and malaise
(not specified) after 1 month of working at levels of 5 to 8 ppm (25 to 40 mg/m3)
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ISOPHORONE April 1992
isophorone. After workroom levels were reduced to 1 to 4 ppm (5 to 20 mg/m3)
through improvements in ventilation, no further complaints were received.
Long-term Exposure
Workers in a screen printing plant were routinely exposed to several organic solvents
in air, including isophorone (Samimi, 1982). Workers were reported to have
time-weighted average (TWA) exposures to isophorone of 8.3 to 23 ppm. No reports
of worker complaints were mentioned.
Animals
Short-term Exposure
The oral LD^ of isophorone has been reported to be 2.0 and 2.37 g/kg bw for mice
and rats, respectively (Patty, 1982).
A skin penetration LDj<, of 1.39 g/kg (in rabbits) was reported in a technical data
booklet (Union Carbide, 1968). This result was obtained from a 24-hour covered
skin contact with isophorone. No details on experimental protocol such as the
number of animals exposed were presented.
When rats were exposed for 8 hours in an isophorone-saturated atmosphere
(approximately 580 ppm), 1/6 animals died (Union Carbide, 1963).
Groups of 10 male Wistar rats were exposed to 5-17.8 mg/L (8,853, 150 ppm) of
isophorone for 4 hours. An LCX of 7.0 mg/L (1,240 ppm) was calculated (Hazleton
Laboratories, 1965).
At low concentrations, isophorone caused respiratory irritation which was measured
by the reflex decrease in respiratory rate. Six male Swiss OF, mice were exposed to
4 to 100 ppm (23 to 565 mg/m3) of isophorone for 5 minutes (De Ceaurriz et al.,
1981). The concentration of isophorone causing a 50% decrease in respiratory rate
was 28 ppm (157 mg/m3).
The neurobehavioral effects of isophorone were quantified by measuring the
duration of immobility in a "behavioral despair" swimming test (De Ceaurriz et al.,
1984). Mice (10 male Swiss OF, mice) were exposed to 89 to 137 ppm (500 to 770
mg/m3) isophorone for 4 hours. The concentration that produced a 50% decrease in
immobility (IDjo) was 110 ppm (620 mg/m3).
Male and female CD rats (10/sex) were exposed via inhalation to isophorone for 4
weeks. The average measured concentration of isophorone was 0.208 mg/L (37 ppm)
5 days/week for 6 hours/day. These animals were compared with an unexposed
control group (Hazleton Laboratories, 1968). Mild transient nasal bleeding was
reported in treated rats. No other abnormal clinical signs were found. Body weight
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ISOPHORONE April 1992.
gain was reduced significantly in male rats. Absolute and relative liver weights were
also reduced in males. No exposure-related effects were found by gross necropsy.
Hematologic parameters were not affected by treatment.
Ten male Wistar rats and male and female guinea pigs (10/sex/group) were exposed
via inhalation to isophorone for 6 weeks. The concentrations ranged from 25,500
ppm (1,402,830 mg/m3) 8 hours/day, 5 days/week. No effects were seen in animals
exposed to 25 ppm (140 mg/m3) (Smyth et al., 1942). No distinction was made
between rats and guinea pigs, since the toxic effects found in both species were
similar. Histopathologic changes were found in animals that survived exposure to
>50 ppm (280 mg/m3). These included effects on the lung (congestion), kidney
(dilation of Bowman's capsule, cloudy swelling), liver (congestion) and spleen
(congestion). Deaths were reported at concentrations of >100 ppm (565 mg/m3).
Histologic examination of animals that died from isophorone exposure revealed
severe lung and kidney damage. Reduced body growth was noted in animals exposed
to ~100 ppm (565 mg/m3). Other effects, reported only at 500 ppm (2,830 mg/m3),
included excretion of albumin in the urine, hematologic changes, conjunctivitis and
nasal irritation. Thus, the NOAEL and LOAEL are identified as 25-and 50 ppm,
respectively.
Dermal/Ocular Effects
Isophorone had a weak irritant action on rabbit and guinea pig skin. The dermal
LDM was reported to be 1,390 mg/kg bw (Union Carbide, 1968) and 1,500 mg/kg
(Patty, 1982) for rabbits. Moderate skin irritation was observed when isophorone
was held in contact with guinea pig skin for 24 hours.
Isophorone produced corneal opacity, inflammation of eyelids, discharge and
conjunctiva when administered to the eyes of rabbits (Truhaut et al., 1972).
Carpenter and Smyth (1946) reported moderate corneal injury and ocular burns
when undiluted isophorone was administered to the eyes of rabbits.
Long-term Exposure
Nor-Am Agricultural Products (1972a) conducted a 90-day feeding study with
isophorone in rats. CFE rats (20/sex/group) were fed isophorone in their daily diet
for 90 days at levels of 0, 750, 1,500 or 3,000 ppm. Based on the actual feed intake,
the doses are equivalent to 0, 57, 102.5 or 233.8 mg/kg/day in the males and 0, 78.9,
163.8 or 311.8 mg/kg/day in the females respectively. No compound-related effects
were observed in the female animals. However, at 3,000 ppm, the male rats suffered
a significant decrease (8 to 11%) in body weight gain from week 6 through 11 (p
<0,01). Thus the study identifies a LOAEL of 3,000 ppm and a NOAEL of 1,500
ppra (234 and 103 mg/kg/day, respectively).
Beagle dogs (four/sex/group) were administered isophorone for 90 days at doses of 0,
35, 75 or 150 mg/kg bw/day in gelatin capsules (Nor-Am Agricultural Products,
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ISOPHORONE April 199,2.
1972b). No compound-related effects were observed either clinically or
histopathologically. The NOAEL for this study is identified as 150 mg/kg/day.
In a 13-week study (NTP, 1986), F344/N rats (13/sex/group) were administered
isophorone by gavage at levels of 0, 62.5, 125, 250, 500 or 1,000 mg/kg bw per day 5
days/week, (equivalent to average daily doses of 0, 45, 89, 179, 357 or 714 mg/kg/day).
One female rat that received the highest dose died. No compound-related gross or
microscopic effects were observed in either sex. Examination of the kidneys of
high-dose and control animals confirmed a lack of nephrotoxicity. Rats receiving
1,000 mg/kg were reported to be sluggish and lethargic after dosing, suggesting CNS
effects. These effects indicate that the LOAEL is less than 1,000 mg/kg/day
(equivalent to an average daily dose of 714 mg/kg/day), and the NOAEL is less than
500 mg/kg (equivalent to an average daily dose of 357 mg/kg/day).
In a 13-week study (NTP, 1986) of B6C3F, mice (10/sex/group) administered
isophorone by gavage (5 days/week) at doses of 0, 62, 125, 250, 500 or 1,000 mg/kg
bw. Three of 10 females administered 1,000 mg/kg isophorone died before the end
of the study. Investigators considered the deaths to be compound-related. No
compound-related gross or microscopic effects were observed in either sex, and
examination of the kidneys of high dose and control animals confirmed a lack of
nephrotoxicity. Significant mortality at the high dose indicates that the LOAEL is
below 1,000 mg/kg (equivalent to an average daily dose of 714 mg/kg/day) and the
NOAEL is below 500 mg/kg (equivalent to an average daily dose of 357 mg/kg/day).
A 2-year bioassay was conducted by NTP (1986) on F344/N rats (50/sex/group).
Isophorone was administered by gavage at doses of 0, 250, or 500 mg/kg bw 5
days/week for 2 years. These doses are equivalent to average daily doses of 0, 179,
or 357 mg/kg/day. The overall incidence of nephropathy was similar between dosed
and vehicle control male rats. (Dosed male rats exhibited increased mineralization
of the kidney tubules [control, 1/50; low dose, 31/50; high dose, 20/50], and epithelial
hyperplasia of the renal pelvis [control, 0/50; low dose, 5/50; high dose 5/50]).
Nephropathy incidence in female rats was somewhat increased (low dose, 39/50; high
dose, 32/50) compared to female vehicle controls (21/50). Based on the conditions of
this study, the LOAEL was 179 rag/kg/day (the lowest dose tested).
B6C3F, mice (50/sex/group) were administered 0, 250 or 500 mg/kg bw isophorone by
gavage 5 days/week for 2 years (NTP, 1986). Dosed male mice showed an increased
incidence of hepatic coagulative necrosis (control 3/48; low dose, 10/50; high dose,
10/50) and hepatomagaly (control, 23/48; low dose, 39/50; high dose, 37/50) when
compared to controls. These effects in female mice were comparable in treated and
control animals. Chronic focal inflammation was also observed at increased
incidences in dosed male mice. Based on the conditions of this study, the LOAEL
was 250 mg/kg (equivalent to an average daily dose of 179 mg/kg/day).
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ISOPHORONE April 1991
Reproductive Effects
No reports of studies on reproductive effects associated with exposure to isophorone
were found in the literature.
Developmental Effects
An inhalation teratology study was conducted with pregnant F344/N rats and CD-I
mice (22/group) at isophorone levels of 0, 25, 50 or 115 ppra for 6 hours/day on days
6 to 15 of gestation (Bio/dynamics, 1984). No statistically significant differences
among control and treatment groups were found for any of the fetal external, visceral
or skeletal parameters in either rats or mice. However, maternal toxicity was noted
at 115 ppm in both rats and mice as evidenced by statistically significant differences
in mean body weight and food consumption of the dams. Rats also had increased
alopecia and genital staining. The NOAEL for this study in rats and mice appears to
be 50 ppm (2,500 mg/m3).
Mutagenicitv
Isophorone showed a weak mutagenic response in the L5178Y tk+/tk- Mouse
Lymphoma Forward Mutation Assay in the absence of S9 mix (McGregor et
al., 1988). Isophorone was not mutagenic in Salmonella strains TA100, TA1535 or
TA98, with or without activation (NTP, 1986; Cheh, 1986). However, when
isophorone was chlorinated under conditions similar to those used in wastewater
chlorination, mutagenic activity by the Ames/Salmonella assay using strain TA100 was
increased (Cheh, 1986).
There are conflicting results as to whether isophorone is mutagenic in mammalian
cells in vitro. When tested at concentrations of 4,001,200 /ig/mL, positive results
were obtained in the mouse lymphoma L5178Y/TK ฑ assay in the absence of
metabolic activation (Nil', 1986). Two subsequent replications of this study by
McGregor et al. (1988) verified these results. However, when this assay was
performed in another laboratory, using similar concentrations of isophorone
(1,301^300 Mg/mL), the results were negative, with or without activation (O'Donoghue
et al., 1988).
Isophorone induced sister-chromatid exchanges in Chinese hamster ovary cells in the
absence (NTP, 1986), but not the presence of, Arochlor 1254-induced rat liver S9 mix
(O'Donoghue et al., 1988). In addition, isophorone did not induce chromosomal
aberrations in Chinese hamster ovary cells, with or without S9 mix (NTP, 1986), or
mouse bone marrow assay in vivo (O'Donoghue et al, 1988).
Carcinogenicitv
F344/N rats (50/sex/group) were administered isophorone by gavage at doses of 0,
250 or 500 mg/kg bw 5 days/week for 2 years (NTP, 1986). These doses were
8
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ISOPHORONE April 199JL
equivalent to an average daily dose of 179 and 357 mg/kg/day. Dosed males showed
a slight increased incidence of renal tubular cell adenomas and adenocarcinomas
over the vehicle controls [controls, 0/50; low-dose, 3/50 (6%); high-dose, 3/50 (6%)].
The authors considered that these tumors were significant, since the historical
incidence for such tumors at this laboratory was 4/1,091 (0.4%). Dosed male rats
also had an increased incidence of epithelial hyperplasia of the renal pelvis, and
among the high-dose males, 4/50 animals also exhibited tubular cell hyperplasia.
High-dose males had a significantly higher incidence of carcinomas of the preputial
gland than did the low-dose or control males (controls and low-dose, 0/50; high-dose,
5/50). However, the prepuce is among those tissues examined microscopically only
when a neoplasm is visible to the prosector. Therefore, the actual incidence of all
types of proliferative lesions of the prepuce is not known, since only five high-dose
males and two low-dose females were sampled for histological examination, and the
diagnosis or actual occurrence of preputial gland tumors has been sporadic in vehicle
controls in previous National Toxicology Program (NTP) studies. (The NTP
historical control values for preputial gland tumors ranged from 0 to 7, and 5 were
observed in corn oil vehicle controls in one previous comparable NTP study in the
same laboratory.) Therefore, the NTP concluded that isophorone exposure produced
some evidence of carcinogicity in male rats but not in females.
Isophorone was administered by gavage to B6C3F, mice (50/sex/group) at doses of 0,
250 or 500 mg/kg bw 5 days/week for 2 years (NTP, 1986). These doses were
equivalent to an average daily dose of 179 and 357 mg/kg/day. In the high-dose male
mice, isophorone exposure appears to be associated with a statistically significant
increase in hepatocellular adenomas or carcinomas compared with low-dose and
control male mice (control, 18/48 [38%]; low-dose, 18/50 (36%); high-dose, 29/50
[58%]). High-dose male mice also showed an increase in mesenchymal tumors
(fibroma, fibrosarcoma, neurofibrosarcoma or sarcoma) of the integumentary system
(control, 6/48; low-dose, 8/50; high-dose, 14/50). There was an increase in
lymphomas or leukemias in the low-dose male mice but not in the high-dose group.
However, the survival of male mice was low (final rates: control, 16/50; low-dose,
16/50; high dose, 19/50). The NTP, owing to the reduced survival, analyzed the data
with Life Table Test, and found no significant elevation or trends for these sites.
Although the unadjusted tests showed significant elevations, the survival was so low,
including early deaths and high control losses, the results in male mice are
considered by NTP and the U.S. Environmental Protection Agency (EPA) as
"equivocal." No treatment-related neoplasms were observed in female mice.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day, Longer-term
(approximately 7 years) and Lifetime exposures if adequate data are available that identify a
sensitive noncarcinogenic end point of toxicity. The HAs for noncarcinogenic toxicants are
derived using the following formula:
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ISOPHORONE April 1992.
(NQAEL or LOAEL) x
(UF) ( 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/OW guidelines.
L/day = Assumed daily water consumption of a
child (1 L/day) or an adult (2 IVday).
One-day Health Advisory
No data were found in the available literature that were suitable for the determination
of the One-day HA value for isophorone. Therefore, it is recommended that the
Longer-term HA value for a 10-kg child (15 mg/L, calculated below) be used as a
conservative estimate for a One-day exposure.
Ten-day Health Advisory
No data were found in the available literature that were suitable for the determination
of the Ten-day HA value for isophorone. Therefore, it is recommended that the Longer-term
HA value for a 10-kg child (15 mg/L, calculated below) be used as a conservative estimate of
a Ten-day exposure.
Longer-term Health Advisory
The 90-day subchronic study in dogs (Nor-Am Agricultural Products, 1972b) was
selected as the basis for the Longer-term HA for isophorone. In the study, a NOAEL of
150 mg/kg/day was identified. The dietary study in rats (Nor-Am Agricultural Products,
1972a) identified a lower NOAEL of 103 mg/kg/day, had been considered as the basis for the
Longer-term HA. However, in the subchronic study in the dog, isophorone was administered
via capsule which provided for a better control of the dose. The actual amount of isophorone
ingested by rats in the dietary study might be less than the amount assumed (partly due to
evaporation), making the NOAEL uncertain and the dog study identified a higher NOAEL,
making it the most appropriate study.
10
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ISOPHORONE April 199Z
The Longer-term HA for a 10-kg child is calculated as follows:
(150 mg/kg/day) (10 ke)
Longer-term HA = (100) (1 L/day) = 15 mg/L O5-000 ^S71-)
150 mg/kg/day = NOAEL, based on the 90-day dog study (Nor-Am
Agricultural Products, 1972b).
10 kg = Assumed body weight of a child.
100 = Uncertainty factor, chosen in accordance with EPA
or NAS/OW guidelines for use with a NOAEL from
a study in animals.
1 L/day = Assumed daily, water consumption of a child.
Since isophorone has been classified by EPA as a Group C carcinogen based on limited
evidence of carcinogenicity in animals, the chemical is considered to have potential to cause
cancer in humans. Therefore, the Longer-term HA is kept at 15 mg/L and not be rounded
up to 20 mg/L. For prudent purposes, the Longer-term HA for a child (15 mg/kg/day,
calculated above) is also recommended as the Longer-term HA for 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 estimate (with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is likely to be without
appreciable risk of deleterious health effects during 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 in drinking water alone 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 assured.
If the contaminant is classified as a known, probable, or possible human carcinogen,
according to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986),
11
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ISOPHORONE April 1992.
then caution must be exercised in making a decision on how to deal with possible lifetime
exposure to this substance. For human (A) or probable (B) human carcinogens, a Lifetime
HA is not recommended. For possible (C) human carcinogens, an additional 10-fold safety
factor is used in the calculation of the Lifetime HA. The risk manager must balance this
assessment of carcinogenic potential and the quality of the data against the likelihood of
occurrence and significance of health effects related to noncarcinogenic endpoints of toxicity.
To assist the risk manager in this process, drinking water concentrations associated with
estimated excess lifetime cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the
70-kg adult drinking 2 L of water/day are provided in the Evaluation of Carcinogenic
Potential section.
The subchronic study in dogs (Nor-Am Agricultural Products, 1972b) has been selected
as the basis for the RfD for isophorone. In this study, male and female dogs were fed
isophorone in capsules at 0, 35, 75 or 150 mg/kg/day for 90 days. All animals survived the
study with no signs of adverse effect. A NOAEL of 150 mg/kg/day was identified. The NTP
(1986) 2-year bioassay in rats and mice demonstrating systemic effects at an average daily
dose of 179 mg/kg/day supports the selection.
As described above, the RfD was calculated using the NOAEL of 150 mg/kg/day from
the subchronic dog study (Nor-Am Agricultural Products, 1972b).
Step 1: Determination of the Reference Dose (RfD)
RfD = *(l"ooof ^^ = ฐ'15 MSfc&tey (rounded to 0.2 mg/kg/day)
where:
150 mg/kg/day = NOAEL, based on the subchronic study in dogs.
1,000 = Uncertainty factor, chosen in accordance with EPA or
NAS/OW guidelines for use with a NOAEL from a
subchronic study in animals.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
I/CT (0.2 mg/kg/dav) (70 kg) _ n ^/w^ ^
DWEL = (2L/day) = mg/^( ' /*8^
where:
0.2 rng/kg/day = RfD.
70 kg = Assumed body weight of an adult.
2 L/day = Assumed daily water consumption of an adult.
12
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ISOPHORONE April
Step 3: Determination of the Lifetime Health Advisory
where:
7 mg/kg/day = DWEL.
20% = Assumed relative source contribution from water.
10 = Additional safety factor based on OW policy to account
for possible carcinogenicity of Group C carcinogens.
Evaluation of Carcinogenic Potential
The International Agency for Research on Cancer (IARC) has not evaluated the
carcinogenic potential of isophorone.
Applying the criteria described in EPA's guidelines for assessment of carcinogenic
risk (U.S. EPA, 1986), isophorone has been classified by EPA in Group C: possible
human carcinogen. This category applies to agents for which there is limited
evidence of carcinogenicity from animal studies in the absence of human data.
In the NTP 2-year gavage study, the report concluded that there is some evidence of
carcinogenicity in male rats due to an increased incidence of relatively rare renal
tubular cell tumors at 250 and 500 mg/kg/day and rare preputial gland carcinomas at
500 mg/kg/day. Based on the combined incidences of preputial gland and renal
tubular cell tumors in male rats, the carcinogenic potency (ql*) was estimated by
EPA to be 4 x 10'J per (mg/kg/day) using the multistage model:
The Office of Water used the following mathematical models for comparison of the
oncogenic lifetime risk for a 70-kg adult. The cancer risk estimates (95% upper
limit) that may cause one excess cancer in 1,000,000 (10~*) population is associated
with exposure to isophorone levels in drinking water of 9 /ig/L (multistage), 3 /ig/L
(one hit), 0.3 /ig/L (logit), 0.1 /ig/L (Weibull), 0.05 /ig/L (multihit) and 220 /tg/L
(probit).
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
Based on reports of eye, nose and throat irritation, fatigue and malaise at levels
below 10 ppm (NIOSH, 1978; Union Carbide, 1963), the American Conference of
Governmental Industrial Hygienists (ACGIH) has set a Threshold Limit Value
(TLV) of 5 ppm with the stipulation that 5 ppm is also a ceiling concentration to
prevent eye and throat irritation. OSHA (1989) has lowered the isophorone
13
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ISOPHORONE April 199Z.
standard from 25 ppm to 4 ppm for an 8-hour Time-Weighted Average (TWA).
NIOSH currently recommends a permissible exposure limit of 4 ppm (23 mg/m}) as
a concentration for an 8-hour TWA.
VII. ANALYTICAL METHODS
Isophorone can be analyzed by either EPA Method 609 (U.S. EPA, 1984a) or 625
(U.S. EPA, 1984b), depending on the detection level sought. In the Methods, 1L of
sample is extracted with methylene chloride and concentrated to ImL. For Method
609, the solvent must be exchanged to hexane, with analysis by Flame-Ionization
Gas Chromatography. Method 625 is the Semi-volatile Gas Chromatography Mass
Spectrometry procedure. A detection limit of 5.7 ^.g/L can be reached with Method
609.
VIII. TREATMENT TECHNOLOGIES
Available data indicate that lime softening, granular activated carbon (GAC)
adsorption and chemical oxidation will significantly reduce levels of isophorone in
drinking water supplies.
McCarty et al. (1982) reported the removal of isophorone at Water Factory 21, a
wastewater treatment plant in Orange County, CA, using an excess-lime softening
treatment process, followed by GAC. The lime dose was 350 to 400 mg/L as CaO
to raise the pH to 11. The softened water was recarbonated to a pH of 8.0, filtered,
and chlorinated. Isophorone was present at an initial concentration of 0.3 Mg/L.
The GAC columns contained Filtrasorb 300 carbon and were each designed for a
capacity of 1 million gallons per day, a hydraulic flow rate of 4.9 gallons/min/sq ft
and an empty bed contact time (EBCT) of 30 minutes. Lime softening alone
reduced isphorone concentration to 0.05 /ig/L and GAC further reduced isophorone
to below 0.05 jig/L.
Borup and Middlebrooks (1987) conducted tests to determine the feasibility of
treating contaminated water by a UV light-catalyzed oxidation process with
hydrogen peroxide (H2O2) as an oxidant. Oxidation with 250 jig/L of hydrogen
peroxide and UV irradiation with an intensity of 1,210 microwatts/cm2 for 60
minutes reduced initial isophorone concentrations of 62 mg/L to below 0.05 mg/L.
Data were not found for the removal of isophorone from drinking water by
aeration. However, isophorone is probably only slightly amenable to removal by
aeration due to its moderate Henry's Coefficient value, high solubility and low vapor
pressure.
14
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ISOPHORONE April 199Z.
IX. REFERENCES
Amoore, I.E. and E. Hautala. 1983. Odor as an aid to chemical safety: Odor threshold limit
values and volatilities for 214 industrial chemicals in air and water dilution. J. Appl.
Tpncol. 3:272-290.
Bio/dynamics. 1984. Inhalation teratology study in rats and mice. Final Report 3223772.
Unpublished study performed by Bio/dynamics Inc., East Millstone, NJ, for Exxon
Biomedical Science, East Millstone, NJ. OTS Section 4 submission Doc ID 40-855049.
Microfiche No. OTS 0507224.
Borup, M.B. and EJ. Middlebrooks. 1987. Photocatalyzed oxidation of toxic organics.
Water Science and Technology 19:381-390.
Carpenter, C.P. and H.F. Smyth Jr. 1946. Chemical burns of the rabbit cornea. Am. J.
Ophthalmol. 29:1,363-72.
CAS. 1988. Chemical Abstract Service Online Registry File. August 3.
Cheh, A.M. 1986. Mutagen production by chlorination of methylated alpha, beta-
unsaturated ketones. Mutat. Res. 169 No. 1-2:1-9.
De Ceaurriz, J.C., J.C. Micillion, P. Bonnet and J.P. Guenier. 1981. Sensory irritation caused
by various industrial airborne chemicals. Toxicol. Lett. (AMST) 9(2):137-143.
De Ceaurriz, J.C., J.C. Micillino, B. Barignac, P. Bonnet, J. Muller and J.P. Guenier. 1984.
Quantitative evaluation of sensory irritating and neurobehavioral properties of aliphatic
ketones in mice. Fd. Chem. Tox. 22(7):545-549.
Dutertre-Catella, H., P. Nguyen, Q. Dang and R. Truhaut. 1978. Metabolic transformation
of the 3,5,5-trimethyl-2-cyclohexene-l-one (isophorone). Toxicol. Eur. Res. 1:209-216.
Hawley, G.G., 1981. The condensed chemical dictionary, 10th ed. New York, NY: Van
Nostrand Reinhold Co. p. 581.
Hazleton Laboratories. 1968. Assessment and comparison of subacute inhalation toxicities of
three ketones. Final Report. Prepared by Hazleton Laboratories, Inc. Falls Church,
VA for Exxon Chem Amers. Houston, TX. OTS 8d submission Doc ID. 878210935,
Microfiche No. 206267.
Hazleton Laboratories. 1965. UCX determination, acute inhalation exposure - rats. Final
Report. Prepared by Hazleton Laboratories, Inc., Falls Church, VA for Exxon Chem.
Amers., Houston, TX. OTS 8d submission Doc. ID. 878210933, Microfiche No. 206267.
Jennings, P. 1965. Photochemistry of isophorone. Diss. Abstr. 26:698.
15
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ISOPHORONE April 1992,
Jungclaus, G., L Games and R. Kites. 1976. Identification of trace organic compounds in
tire manufacturing plant wastes. Anal. Chera. 48:1,894-1,896.
McCarty, D.L., D. Argo, M. Reinhard, J. Graydon, N. Goodman, and M. Aieta. 1982.
Performance of Water Factory 21 in removing priority pollutants. Proc. Water Reuse
Symposium: 2,325-2,349.
McGregor, D.B., A. Brown, P. Cattanach, I. Edwards, D. McBride, C. Riach and
WJ. Caspary. 1988. Responses of the L5178Y tk+/tk- mouse lymphoma cell forward
mutation assay: III. 72 coded chemicals. Environ. Molec. Mut. 12:85-154.
NIOSH. 1978. National Institute for Occupational Safety and Health. Criteria for a
recommended standard, occupational exposure to ketones. DHEW (NIOSH)
Publication No. 78-173. pp. 44, 75, 176.
Nor-Am Agricultural Products. 1972a. 90-Day subchronic toxicity of isophorone in the rat.
MRID No. 00123976. Available from EPA. Write to FOI, EPA, Washington, D.C.
20460.
Nor-Am Agricultural Products. 1972b. 90-Day subchronic toxicity of isophorone in the dog.
MRID No. 00123977. Available from EPA. Write to FOI, EPA, Washington, D.C.
20460.
NTP. 1986. National Toxicology Program. Toxicology and carcinogenesis studies of
isophorone (CAS No. 78-59-1) in F344/N rats and B6C3F, mice (gavage studies). NTP
TR-291. NIH Publication No. 86-2547.
O'Donoghue, J.L, S.R. Haworth, R.D. Curren, P.E. Kirby, T. Lawlor, EJ. Moran, R.D.
Phillips, D.L Putnam, A.M. Rogers, R.S. Slesinski. 1988. Mutagenicity studies on
ketone solvents: methyl ethyl ketone, methyl isobutyl ketone and isophorone. Mutat.
Res. 206(2):149-161.
OSHA. 1989. Air contaminants. Final Rule. U.S. Department of Labor. Occupational
Safety and Health Administration. Code of Federal Regulations. 29 CFR 1910. Fed.
Reg. 54(12):2941.
Patty, F.A. 1982. Patty's industrial hygiene and toxicology, 3rd rev. ed. Vol. 2C:4,786-4,787.
Price, K., G. Waggy and R. Conway. 1974. Brine shrimp bioassay and BOD of
petrochemicals. J. Water Pollut. Control Fed. 46:63-77.
Samimi, B. 1982. Exposure to isophorone and other organic solvents in a screen plant. Am.
Ind. Hyg. Assoc. J. 43:43-48.
Shackelford, W. and L. Keith. 1976. Frequency of organic compounds identified in water.
U.S. EPA 600/4-76-062. Athens, GA: U.S. Environmental Protection Agency.
16
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ISOPHORONE April 1992,
Sheldon, L. and R. Hites. 1978. Organic compounds in the Delaware River. Environ. Sci.
Technol. 12:1,188-1,194.
Silverman, L., H. Schulte and M. First. 1946. Further studies on sensory response to certain
industrial solvent vapors. J. Ind. Hyg. Toxicol. 28:262-266.
Smyth, H., J. Seaton and L Fischer. 1942. Response of guinea pigs and rats to repeated
inhalation of vapor of mesityl oxide and isophorone. J. Ind. Hyg. Toxicol. 24:46-50.
Truhaut, R., H. Dutertre-Catella and P. Nguyen. 1972. Study of the toricity of an industrial
solvent, isophorone. Irritating capacity with regard to the skin and the mucosae. J.
Eur. Toxicol. 5:31-37.
Union Carbide Corporation. 1968. Ketones booklet F-419771. New York, NY: Union
Carbide Corporation. 21 p.
Union Carbide Corporation. 1963. Toxicological studies-isophorone summary data sheet.
New York, N.Y.: Ind. Med. Toxicol. Dept, Union Carbide Corporation:
U.S. EPA. 1974. U.S. Environmental Protection Agency. Draft analytical report: New
Orleans area water supply study. Region IV, Dallas, TX. Surveillance and Analysis
Division, Lower Mississippi River facility, Slidell, LA.
U.S. EPA. 1975. U.S. Environmental Protection Agency. Preliminary assessment of
suspected carcinogens in drinking water. Report to Congress. Washington, DC: U.S.
Environmental Protection Agency.
U.S. EPA. 1984a. U.S. Environmental Protection Agency. U.S. EPA Method 609 -
Nitroaromatics and isophorone. 40 CFR part 136, Oct 26.
U.S. EPA. 1984b. U.S. Environmental Protection Agency. U.S. EPA Method 625 -
Base/neutrals and acids. 40 CFR part 136, Oct 26.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Fed. Reg. 51(185):33,992-34,003.
Veith, G.D., KJ. Macek, S.R. Petrocelli and J. Carroll. 1980. An evaluation of using
partition coefficients and water solubility to estimate bioconcentration factors for
organic chemicals in fish. ASTM STP 707. Aquatic Toxicology. In: Easton J.G. et al.,
eds. Amer. Soc. Test Mater, pp. 116-129.
17
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EPA 0553 April. 199.2
RX000027511
MALATfflON
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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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Malathion April. 1992
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 121-75-5
Structural Formula
S CH2-C-0-C2H5
CHj-0-P-S-CH -C - "
O.O-Dimethvl S-d.Z-dicarfaethoxvethvDphosphorodithioate
Synonyms
Calmathion, Carbetox, Carbofos, Celthion, Chemathion, Cimexan, Cythion,
Dorthion, Emmatox, Ethiolacar, Etiol, Extermathion, Forthion, Fosfothion,
Fosfotion, Fyfanon, Hilthion, Karbofos, Kypfos, Lucathion, Malacide, Malafor,
Malamar, Malaphele, Malathiazol, Malathon, Malathyl, Malatol, Malatox, Maldison,
Maltox, Mercaptothion, Phosphothion, Prioderrn, Sadofos, Sumitox, Vetiol, Zithion,
Compound 4049 (U.S. EPA, 1988a; Meister, 1987; IARC, 1983).
Uses
Malathion is an organophosphorous insecticide used extensively to control a wide
variety of insects and mites. It is used particularly where a high degree of safety to
mammals is desired (Meister, 1987).
An estimated 15 to 20 million pounds of the active ingredient are used annually,
based on 1985 and 1986 data (U.S. EPA, 1988a). Industrial, commercial and
government applications constitute 40% of the annual use in the United States, and
approximately 33% is used in and around the home. Direct application to
agricultural crops accounts for 12% of use, and agricultural noncrop use is 15% of
the total.
Properties (Windholz et al., 1983; Worthing and Walker, 1987; U.S. EPA, 1988a)
Chemical Formula C10H,9O6PS2
Molecular Weight 330.36
Physical State (at 25ฐC) liquid
Boiling Point (at 0.7 mm Hg) 156-157ฐC
Melting Point (ฐC) 2.9
Density (at 20ฐC) 1.20 g/mL
Vapor Pressure (at 30ฐC) 0.00004 mm Hg
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Malathion April, 1992.
Specific Gravity
Water Solubility (at 20ฐC) 145 mg/L
Log Octanol/Water Partition 2.89
Coefficient (log K^)
Taste Threshold
Odor Threshold
Conversion Factor 1 ppm = 1 mg/m3
Occurrence
In a national surface water monitoring program (1976 to 1980), malathion was
found in 0.3% of the samples, with a maximum value of 0.18 ppm (Carey and Kutz,
1985). Malathion was not detected in sediments.
Ambient air concentrations of malathion as a result of crop treatment were 0.01 to
0.60 mg/m3 in work areas, and 0.001 mg/m3 in an agricultural community (Hayes,
1971).
In the National Soils Monitoring Program, malathion was detected at levels of 0.08
to 0.19 mg/kg in cropland soil (Carey et al., 1978, 1979).
In FDA total diet studies conducted from 1982 to 1986, no tolerance-exceeding
residues of malathion were observed (U.S. EPA, 1988a).
Environmental Fate
Malathion at 10 mg/L was added to raw river water in closed glass containers (at
room temperature and exposed to natural and artificial light) (Eichelberger and
Lichtenberg, 1971). Within 4 weeks, malathion residues were not detectable.
Malathion is very mobile in loamy sand and sandy loam soils (LaFleur, 1979).
Adsorption ratios determined in Norfolk loamy sand and Cecil sandy loam soils
treated with malathion at 10, 20, 40 or 80 mmol/kg were 0.73 to 0.95, and
desorption ratios were 0.75 to 0.95.
It is not possible to fully assess the environmental fate of malathion because
acceptable data are lacking. Based on theoretical calculations, however, application
of malathion to land could result in aquatic concentrations of 0.3 mg/L owing to
surface runoff and aerial drift (U.S. EPA, 1988a).
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Malathion April. 1992_
III. PHARMACOKINETICS
Absorption
Malathion is absorbed through the gastrointestinal tract, the respiratory tract and
the skin, as shown by the rapid onset of postexposure symptoms of poisoning (NAS,
1977; WHO, 1982).
Wester et al. (1983) determined the percutaneous absorption of 5 mg/cm2 malathion
repeatedly administered to the ventral forearm skin of male volunteers for 14 days.
Monitoring of urine indicated that dermal absorption of malathion on day 1 (4.5%)
was not significantly different from dermal absorption on day 8 (3.5%). Feldman
and Maibach (1970, 1974) reported absorption of 7.8% and 8.2% after application
of a single dose of malathion at 4 jig/cm2 to the ventral forearm of human
volunteers.
Distribution
Distribution of malathion is general, and low concentrations are found in many
tissues (NAS, 1977; WHO, 1982).
Morgade and Barquet (1982) reported the analyses of tissues obtained at the
autopsy of a suicide victim (times of ingestion, death and autopsy were unavailable).
Malathion was detected primarily in adipose tissue (78 ppra), kidney (17 ppm),
brain (5 ppm), spleen (1 ppm) and bile (1 ppm). Malaoxon, the main active
metabolite, was detected primarily in adipose tissue (8 ppm), brain (1 ppm) and
kidney (1 ppm). Malathion monocarboxylic acid was detected primarily in bile (221
ppm), kidney (106 ppm) and spleen (59 ppm).
Metabolism
Malathion is a phosphorodithioate. Its oxidized analogue, the phosphate malaoxon,
is an anticholinesterase agent. The conversion of malathion to malaoxon is carried
out by the liver microsomal monooxygenase system. Competing with the activation
of malathion are phosphatase and carboxylesterase enzymes which degrade
malathion to less toxic metabolites such as malathion monocarboxylic acid and
dicarboxylic acid, various phosphoric acids and the O-demethylated product. The
degradation rate of malaoxon exceeds its activation rate, so there is generally little
accumulation of malaoxon in mammalian systems (NAS, 1977).
Excretion
Malathion is rapidly metabolized in mammals, and its metabolites are excreted
mostly in the urine within 24 hours (IARC, 1983; WHO, 1982).
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Malathion April. 199,2-
Urinary excretion of l4C-malathion after intravenous administration was
approximately 90%, with a half-life of 3 hours (Feldman and Maibach, 1974).
Dermal administration of uC-malathion resulted in peak urinary excretion of
radioactivity between 12 and 24 hours of the first day. The excretion half-life of
dermally applied malathion was 30 hours (Wester et al., 1983).
In the Health and Nutrition Examination Survey II (1976-1980) conducted by the
National Center for Health Statistics, the malathion metabolites, malathion
dicarboxylic acid and monocarboxylic acid, were detected in 1.1% and 0.5%,
respectively, of human urine samples collected from individuals 12 to 74 years of
age (Carey and Kutz, 1985).
IV. HEALTH EFFECTS
Humans
Short-term Exposure
The acute toxicity of malathion is due to inhibition of acetylcholinesterase at nerve
endings, leading to an accumulation of endogenous acetylcholine. The effects are
manifested by muscarinic, nicotinic and central nervous system symptoms. The
cause of death is primarily respiratory failure (IARC, 1983).
Technical malathion may contain organophosphorous impurities that diminish the
organism's ability to detoxify malathion. Isomalathion, trimethyl phosphothioates
and other organophosphate impurities have been shown to potentiate the toxicity of
malathion (WHO, 1982).
Moeller and Rider (1962) administered malathion (of unspecified purity) in
capsules to human male volunteers at 8 mg/day for 32 days, 16 mg/day for 47 days
and 24 mg/day for 56 days (five subjects/treatment group). Assuming a body weight
of 70 kg, the doses were 0.11, 0.23 and 034 mg/kg/day. There was no significant
depression of plasma or erythrocyte cholinesterase activity in the 0.11- and
0.23-mg/kg/day treatment groups (maximum cholinesterase inhibition of
approximately 10%). At the highest dose, plasma and erythrocyte cholinesterase
were depressed a maximum of 25% by about 3 weeks after the end of treatment.
There were no clinical signs of poisoning. Thus, the NOAEL in humans for this
study was 0.23 mg/kg/day and the LOAEL was 0.34 mg/kg/day.
Between 1981 and 1985, malathion was the third most common cause of pesticide
illness in California. However, it ranked only eleventh as a cause of hospitalized
pesticide illnesses reported between 1982 and 1986 (U.S. EPA, 1988a; Brown et al..
1989).
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Maiathion April. 199JL
Thirty-five cases of poisoning by ingestion of unspecified amounts of malathion in
India were examined (Chhabra et al., 1970). Pinpoint pupils, cyanosis, excessive
salivation and pulmonary edema were observed. Electrocardiographic abnormalities
were observed in 37% of the cases. Histopathological examination at autopsy
indicated damage to the myocardium.
A number of studies of workers spraying malathion have been conducted in Haiti
(Miller and Shah, 1982; Warren et al., 1985), India (Gupta et al., 1980; Siddiqui et
al., 1980) and Pakistan (Baker et al., 1978; Miller and Shah, 1982). Weakness,
headache, nausea and mild eye irritation were reported. Miller and Shah (1982)
measured plasma and erythrocyte cholinesterase levels in 68 malathion sprayers. In
12 workers, plasma cholinesterase levels ranged from 31.8 to 48.6% of unexposed
controls, and erythrocyte cholinesterase levels ranged from 51.2 to 65.4% of
controls. Lower levels of cholinesterase inhibition were reported in workers using
better safety procedures.
In two cases of attempted suicide by raalathion ingestion (one victim ingested
approximately one cup; the amount ingested by the second victim was unknown),
serum cholinesterase of both patients was completely inhibited for 2 or 3 days
following admission to the hospital; recovery to normal levels occurred in 15 to 20
days following intensive atropine and 2-PAM therapy. No followup was reported
after release from the hospital (Hanna and Choo-Kang, 1983).
Long-term Exposure
Although raalathion has been used as an insecticide for many years, no reports of
chronic or delayed toxicity to humans resulting from long-term exposure were
located. There is no documented case of peripheral neuropathy due to malathion
exposure, even though other organophosphates have been shown to have this effect
(IARC, 1983).
Animals
Short-term Exposure
In mammals, acute oral LDjg values have been reported as 1,000 to 1,375 mg/kg in
Sherman rats (Gaines, 1969) and 1,680 mg/kg in mice (Berteau and Deen, 1978).
Oral LDjg values in Wistar rats were age dependent. Values were 209 mg/kg at 1 to
2 days, 707 mg/kg at 6 to 7 days, 1,085 mg/kg at 12 to 13 days and 1,806 mg/kg at 17
to 18 days of age (Mendoza and Shields, 1977).
Male rats (five/dose) fed diets containing 0, 100 or 500 ppm of 98% pure technical
malathion (approximately 0, 0.5 and 25 mg/kg/day, based on the dietary assumptions
of Lehman, 1959) for 8 weeks had no significant depression of whole-blood
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Malathion April, 1993.
cholinesterase levels (Frawley et al., 1957). The dose of 25 mg/kg/day, the highest
dose tested, may be considered a NOAEL for rats in this study.
In a 33-day study, rats (10/dose) were fed diets containing malathion at levels
resulting in doses of 0, 10, 90 or 470 mg/kg/day (Golz and Shaffer, 1955). There
were no gross signs of toxicity related to treatment. Cholinesterase activity was
determined in plasma, erythrocytes, brain and liver. Significant inhibition of
erythrocyte Cholinesterase was observed at 90 and 470 mg/kg/day, and significant
inhibition of plasma Cholinesterase was observed at 470 mg/kg/day. The NOAEL
for rats in this study was 10 mg/kg/day, and the LOAEL was 90 mg/kg/day.
Other organophosphorus insecticides can potentiate the toxicity of malathion when
appropriate levels are fed in the diet. The toxicity of malathion was increased as
much as 10-fold by feeding o-ethyl-o-p-nitrophenyl phenylphosphothioate (EPN),
parathion, azinphosmethyl, Folex and triorthotolyl phosphate (TOTP). The
potential toxicity was correlated with the inhibition of liver aliesterases (Su et al.,
1971).
Dermal/Ocular Effects
A single dermal application of 4,444 mg/kg of malathion in an emulsifiable
concentrate killed 4 out of 10 Sherman rats (Gaines, 1969).
Technical malathion is nonsensitizing and only mildly irritating to the eyes and skin
(U.S. EPA, 1988a).
Long-term Exposure
Mixed-breed dogs (one of each sex/group) were administered diets containing 0, 25,
100 or 250 ppm of 98% pure technical malathion (approximately 0, 0.6, 2.5 and 6.3
mg/kg/day based on the dietary assumptions of Lehman, 1959) for 12 weeks
(Frawley et al., 1957). No significant depression of plasma Cholinesterase was
observed. The highest dose tested, 6.3 mg/kg/day, caused significant erythrocyte
Cholinesterase inhibition (maximum of 25%). The level of statistical significance
was not provided by the authors. Thus, 2-5 mg/kg/day (100 ppm in feed) may be
considered a NOAEL, and 6.3 mg/kg/day (250 ppm in feed) may be considered a
LOAEL for dogs in this study. The small number of animals in each treatment
group reduces the degree of confidence in the results of the study.
Albino rats (20 to 30 rats/group) were fed 90% pure technical malathion in the diet
for 2 years at 0, 6, 70 or 350 mg/kg/day (Hazelton and Holland, 1953; Golz and
Shaffer, 1955). At 6 mg/kg/day, erythrocyte, plasma and brain Cholinesterase were
inhibited by 10 to 30% of control levels. At 70 mg/kg/day, 60 to 95% inhibition of
erythrocyte Cholinesterase was observed, and at 350 mg/kg/day, 60 to 95% inhibition
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Malathion April. 1992.
of erythrocyte, plasma and brain cholinesterase was reported. A NOAEL for
cholinesterase inhibition was not established, but the LOAEL was 6 mg/kg/day
based on the conditions of this study. The methods and results of this study were
not completely reported.
Pure-bred beagle dogs were administered malathion (95% purity) by capsule at
doses of 0, 62.5, 125 or 250 mg/kg/day (six dogs/sex/group) for 12 months (Tegeris
Laboratories, 1987). Creatinine, serum alanine aminotransferase (SGPT) and blood
urea nitrogen (BUN) were decreased in the high-dose animals. Elevated liver and
kidney weights were observed at all doses, generally in a dose-related manner.
Erythrocyte and plasma cholinesterase were inhibited at all dose levels. Erythrocyte
cholinesterase was inhibited at about 75% of the control activity at all doses. A
NOAEL for cholinesterase inhibition was not established, and the LOAEL was 62.5
mg/kg/day.
Reproductive Effects
In a two-generation continuous feeding study (Kalow and Marton, 1961), Wistar
rats were fed 240 mg/kg/day of 95% pure technical grade malathion. Males and
females were bred after 10 weeks of treatment. Survival of the pups was
significantly reduced at 7 and 21 days after birth, and growth retardation was
observed 9 weeks after birth. Because adverse effects were observed at the only
dose tested, a NOAEL cannot be established for this study.
Developmental Effects
Kimbrough and Gaines (1968) administered malathion intraperitoneally at 600 and
900 mg/kg to pregnant Sherman rats on day 11 of gestation. There were no
significant differences between the malathion-treated females and the controls
relative to dead fetuses per litter, resorptions, average weight of fetuses, average
weight of placenta or malformations of fetuses.
Khera et al. (1978) administered technical malathion (purity not specified) at 50,
100, 200 and 300 mg/kg to female Wistar rats by gastric intubation on days 6
through 15 of gestation. There was no evidence of fetal or maternal toxicity.
Mutagenidty
Malathion produced negative results in the Ames assay, with and without S9
activation, in Salmonella typhimurium tester strains TA97a, TA98 and TA100, at
concentrations up to 5 x 10'J M (Pednekar et al., 1987). Tester strains TA1536,
TA1537, TA1538, TA98 and TA100 were negative in another test, but malathion
was positive at the highest assay concentration (300 mg/plate) for TA1535 and for
Bacillus subtilis TKJ6321 strain (Shiau et al., 1980).
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Malathion April. 1992-
In Drosophila melanogaster, malathion did not induce point mutations when
administered at 50 ppm in culture media. It was also ineffective in producing .total
or partial sex-chromosome losses and nondisjunction (Velazquez et al., 1987).
In a test using human hematopoietic cell cultures, malathion inhibited cell growth in
proportion to the doses of 50, 100, 200 and 400 mg/mL. In studies of chromosomal
aberrations, however, the incidence of metaphases with aberrations did not increase
(Huang, 1973). However, Wiszkowska et al. (1986) reported chromosomal
aberrations in lymphocyte cultures exposed to 10, 40 or 70 /xg/mL of malathion.
Malathion induced increases in the number of aberrant cells in the bone marrow of
Syrian hamsters treated in vivo at 240 and 2,400 rng/kg, but the increases were not
dose related (Dzwonkowska and Huebner, 1986).
Malathion was positive in the mouse micronucleus test after cutaneous exposure at
120, 240 and 480 mg/kg, but the effect was not dose related (Dulout et al., 1982).
In mice treated with raalathion intraperitoneally, subchromatid and chromatid-type
aberrations in bone marrow cells were induced at 230 mg/kg, but net at 115 mg/kg
(Dulout et al., 1983). In male mice receiving an intraperitoneal injection of
malathion at 300 mg/kg, there was no increase in chromosome aberrations in bone
marrow cells and spermatogonia (Degraeve and Moutschen, 1984).
Malathion did not induce dominant lethal effects in mice when administered in the
diet for 7 weeks at unspecified doses (Jorgenson et ah, 1976) or when injected
intraperitoneally (single dose) at 300 mg/kg (Degraeve and Moutschen, 1984).
Carcinogenicitv
In National Cancer Institute bioassays (NCI, 1978, 1979a), technical grade
malathion was administered to Osbome-Mendel rats, Fischer 344 rats and B6C3F,
mice. The Osborne-Mendel rats (50/sex/group) were administered Time-Weighted
Average (TWA) doses of 0, 4,700 or 8,150 ppm malathion in the diet for 80 weeks,
and subsequently observed for 29 to 33 weeks. Feed consumption rates were not
provided, but using the dietary assumptions of Lehman (1959), the doses could be
estimated at 0, 235 and 408 mg/kg/day. There was no significant evidence of
carcinogenicity (NCI, 1978; Huff et al., 1985; U.S. EPA, 1988a). The IARC
Working Group (IARC, 1983) noted that the duration of treatment was 80 weeks,
which is not a rodent lifetime.
Fischer 344 rats (50/sex/group) were fed 95% pure malathion in the diet at
concentrations of 0, 2,000 and 4,000 ppm for 103 weeks (NCI, 1979a). They were
then observed for an additional 2 to 3 weeks. Feed consumption rates were not
provided, but using the dietary assumptions of Lehman (1959), the doses could be
estimated at 0, 100 and 200 mg/kg/day. No treatment-related evidence of
carcinogenicity was observed (NCI, 1979a; Huff et al., 1985; U.S. EPA, 1988a).
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Malathion April, 199Z.
Charles River B6C3F, mice (50/sex/group) were administered 95% pure raalathion
in the diet at concentrations of 0, 8,000 and 16,000 ppra for 80 weeks and observed
14 to 15 weeks (NCI, 1978). Feed consumption rates were not provided, but using
the dietary assumptions of Lehman (1959), the doses could be estimated at 0, 1,200
or 2,400 mg/kg/day. A significant increase in hepatocellular tumors (combined
carcinomas plus neoplastic nodules) was observed in the high-dose male group; the
increase was not statistically significant under a time-adjusted analysis. NCI (1978)
noted that historically the incidence of hepatocellular carcinoma in controls in this
strain of mouse is often higher than that observed in the high-dose male group in
this study. NCI (1978) and IARC (1983) concluded that no significant evidence of
carcinogenicity was found. IARC (1983) and U.S. EPA (1988a) noted that duration
of treatment was only 80 weeks. U.S. EPA (1988a) noted a dose-related trend in
the incidence of hepatocellular carcinoma (p = 0.019) and increased incidence of
hepatocellular carcinoma in the high-dose males (p = 0.031); however, these
findings were considered questionable.
NCI conducted oncogenicity studies on the raalathion metabolite, roalaoxon, in the
Fischer 344 rat and the B6C3F! mouse (NCI, 1979b). Fischer 344 rats
(50/sex/group) were administered 0, 500 or 1,000 ppm of malaoxon in the diet for
103 weeks and were observed for 1 to 2 additional weeks. Feed consumption rates
were not provided, but using the dietary assumptions of Lehman (1959), the doses
could be estimated at 0, 25 or 50 mg/kg/day. No significant increase in tumor
incidence was observed in male rats (NCL 1979b; Huff et al., 1985; U.S. EPA,
1988a). A significant increase in combined thyroid C-cell adenomas and carcinomas
in female rats was observed; however, this was considered to be of questionable
biologic relevance since historical data showed that the incidence of these tumors
was higher than that of the controls in this study (NCL 1979b). In a re-examination
of the study, Huff et al. (1985) concluded there was equivocal evidence of
carcinogenicity in rats with regard to thyroid C-celi adenomas and carcinomas.
Charles River B6C3Ft mice were administered malaoxon via dietary doses of 0, 500
or 1,000 ppra for 103 weeks and were observed for 1 to 2 additional weeks. Feed
consumption rates were not provided, but using the dietary assumptions of Lehman
(1959), the doses could be estimated at 0, 75 or 150 rag/kg/day. No evidence of
carcinogenicity of malaoxon in the mice was observed (NCI, 1979b; U.S. EPA,
1988a).
The above data in mice and rats were also evaluated by the Office of Pesticide
Programs (U.S. EPA, 1990) and found to be flawed for the purpose of adequately
assessing the carcinogenic potential of this chemical. This was determined because,
among other things, the maximum tolerated dose may not have been reached in
these studies.
10
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Malathion April, 199^
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 toricity. The HAs for noncarcinogenic toxicants are derived
using the following formula:
(NOAEL or LOAEL) x (BW) ...
* (UF) (_ Uday) = mg/L (rฐUnded tO
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level (in
rag/kg bw/day).
BW = assumed body weight of a child (10 kg)'or an adult
(70 kg).
UF = uncertainty factor, (10, 100, 1,000 or 10,000), in
accordance with EPA or NAS/OW 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 that are suitable for calculating the One-day HA. It is
recommended that the adult DWEL adjusted for a 10-kg child, 200 jig/L, calculated as
described in the Longer-term HA for the 10-kg child, be used as a conservative estimate of
the One-day HA.
Ten-day Health Advisory
No studies were located that are suitable for calculating the Ten-day HA. Khera et al.
(1978) administered technical malathion to pregnant Wistar rats at levels of 50,100, 200 or
300 mg/kg/day during days 6 to 15 of gestation. Although Khera et al. (1978) did not find any
evidence of fetal or maternal toricity, this study is not suitable for determining a NOAEL
value because of lack of testing for acetylcholinesterase inhibition, a relevant endpoint for
malathion. As reported in Golz and Schaffer (1955), significant inhibition of erythrocyte
cholinesterase was observed in rats administered malathion for 33 days at doses of 90 or 470
mg/kg/day. Thus, some inhibition of acetylcholinesterase also may have occurred, and gone
unobserved, in the Khera et al. (1978) study, making it unsuitable for risk assessment. It is
therefore recommended that the adult DWEL adjusted for a 10-kg child, 200 /ig/L, calculated
11
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Malathion April, 199Z,
as described in the Longer-term HA for the 10-kg child, be used as a conservative estimate of
the Ten-day HA.
Longer-term Health Advisory
The study in humans by Moeller and Rider (1962) used for determining the DWEL has
also been selected as the basis for the Longer-term HA. As described in the next section, a
NOAEL of 0.23 mg/kg/day and a LOAEL of 0.34 rag/kg/day for erythrocyte cholinesterase
inhibition in humans were identified in this study.
The Longer-term HA for a 10-kg child is calculated as follows:
Longer-term HA = ^"f^ (*ฐ kg) = 0.23 mg/L (rounded to 200 MgflL)
(iU) (1 iv day)
where:
0.23 mg/kg/day = NOAEL, based on the absence of cholinesterase depression in
humans exposed orally to malathion for 32 to 56 days.
10 kg = assumed weight of a child.
10 = uncertainty factor, this uncertainty factor was chosen in
accordance with EPA or NAS/OW guidelines in which a
NOAEL from a human study was employed.
1 L/day = assumed water consumption of a 10-kg child.
The DWEL of 0.7 mg/L (700 /ig/L) derived for the Lifetime HA (see next section) is
used for the Longer-term HA for the 70-kg adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure that is
attributed to drinking water and is considered protective of 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 estimate of a daily exposure level 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
12
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Malathion April, 199i.
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 known, probable or possible human carcinogen, according to the
Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then caution must
be exercised in making a decision on how to deal with possible lifetime exposure to this
substance. For human (A) or probable human (B) carcinogens, a Lifetime HA is not
recommended. For possible human carcinogens (C), an additional 10-fold safety factor is
used in the calculation of the Lifetime HA The risk manager must balance this assessment
of carcinogenic potential and the quality of the data against the likelihood of occurrence and
significance of health effects related to noncarcinogenic end points of toxicity. To assist the
risk manager in this process, drinking water concentrations associated with estimated excess
cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of
water per day are provided in the Evaluation of Carcinogenic Potential section.
The study by Moeller and Rider (1962) has been selected as the basis for the Lifetime
HA. Human males were treated orally with malathion in capsules at 0.11, 0.23 and
0.34 mg/kg/day for 32 to 56 days. There were no clinical signs of poisoning, and at the two
lowest doses, 0.11 and 0.23 mg/kg/day, there was no inhibition of plasma or erythrocyte
cholinesterase. Administration of 034 mg/kg/day resulted in a maximum inhibition of plasma
and erythrocyte cholinesterase of 30% of pretest values. Thus, the NOAEL was
0.23 mg/kg/day and the LOAEL was 034 mg/kg/day, under the conditions of this study.
Chronic animal studies were evaluated for calculating the Lifetime HA, but were not
judged to be suitable. Golz and Shaffer (1955) identified a LOAEL of 6 mg/kg/day for
cholinesterase inhibition in rats administered malathion for 2 years. The methods and results
were not adequately reported, and a NOAEL was not identified. Tegeris Laboratories (1987)
identified a LOAEL of 62J mg/kg/day for cholinesterase inhibition and systemic effects in
dogs administered malathion for 1 year. A NOAEL was not identified. In contrast to the
above animal studies, the Moeller and Rider (1962) 56-day study in humans identified a
NOAEL and LOAEL of 023 and 034 mg/kg/day, respectively. In addition, because it was
performed in humans, the Moeller and Rider (1962) study provides a more sensitive indicator
of the potential for cholinesterase inhibition in humans following oral exposure to malathion.
Using this study, the Lifetime HA is calculated as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD , (0.23 mg/kg/day) = Q Q23 mg/kg/day (rounded to 0.02 rag/kg/day)
where:
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Malathion April, 199JL
0.23 mg/kg/day = NOAEL, based on the absence of cholinesterase
depression in humans exposed orally to raalathion for
32 to 56 days.
10 = uncertainty factor, this uncertainty factor was chosen in
accordance with EPA or NAS/OW guidelines in which
a NOAEL from a human study was employed.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL m (0.02 mg/kg/day) (70 kg) = Q ? mg/L (rounded ^ ^ ^
(2 L/day)
where:
0.023 mg/kg/day = RfD (before rounding).
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime HA
Lifetime HA = (0.7 mg/L) (20%) = 0.14 rag/L (rounded to 100 Mg/L)
where:
0.81 rag/kg/day = DWEL.
(20%) = . assumed relative source contribution for a drinking
water disinfectant.
Evaluation of Carcinogenic Potential
Oncogenicity studies of raalathion were negative in rats (NCI, 1978, 1979a; U.S.
EPA, 1988). A study in mice was judged negative by NCI (NCI, 1978; Huff et al.,
1985), but the data were considered equivocal by U.S. EPA (1988a). Another study
was required
An oncogenicity study of malaoxon was equivocal in rats (NCI, 1979b; Huff et al.,
1985; U.S. EPA, 1988a) and negative in mice (NCI, 1979b; Huff et al., 1985; U.S.
EPA, 1988a).
14
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Malathion April, 199Z,
The Office of Pesticide Programs reevaluated the available oncogenicity studies in
rats and mice (U.S. EPA, 1990) and concluded that the oncogenic potential of this
chemical remains undetermined from the available data (Group D).
IARC (1983) has classified malathion and raalaoxon in Group 3: chemicals that
cannot be classified as to their carcinogenicity for humans.
The carcinogenic potential of malathion has been evaluated by the U.S. EPA.
Applying the criteria described in EPA's guidelines for assessment of carcinogenic
risk (U.S. EPA, 1986), malathion may be classified in Group D: not classifiable.
This category is for agents with inadequate animal or no evidence of carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The Safe Drinking Water Committee of the National Academy of Sciences
calculated a Suggested-No-Adverse-Effect Level (SNARL) in drinking water for
malathion of 0.16 rag/L, based on the Moeller and Rider (1962) study (NAS, 1977).
The current exposure criteria for airborne malathion in the United States are
15 mg/mj, 10-hour TWA (NIOSH, 1985) and 10 mg/m3 (ACGffl, 1986).
Tolerances for malathion in a variety of raw agricultural commodities have been
established. For most fruits and vegetables, the level is 8 ppra (U.S. EPA, 1988a).
VII. ANALYTICAL METHODS
Malathion is one of the class of phosphorodithioate. pesticides. The relative ease
with which this pesticide can be monitored by element-specific detectors has led to
its inclusion in many pesticide-monitoring studies. Malathion can be analyzed by
EPA Methods, 622 (U.S. EPA, 1982) and 507 (U.S. EPA, 1988b). In both methods,
a solvent is exchanged for hexane or methyl tertiary butyl ether (MTBE). Analysis
is by an element-specific thermionic nitrogen-phosphorus detector, which allows
even relatively "dirty" samples to be analyzed with no cleanup. The estimated
detection limit for malathion is Q25 jig/L.
VIII. TREATMENT TECHNOLOGIES
Available data indicate that ozone oxidation and activated carbon adsorption can
significantly reduce malathion levels in drinking water.
Sharma et al. (1987) used activated charcoal to remove malathion from saline water
with an initial concentration of 1.6 g of malathion/L. The activated charcoal
adsorption capacity for malathion was found to be 117 mg malathion/g charcoal.
15
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Malathion April, 199.2,
Zeff and Harris (1984) evaluated the applicability of the UV-ozonation process for
water reuse from industrial wastewater at a model P-602 (Ultrox*) pilot plant in
Lathrop, California. Preliminary bench tests were conducted to provide operating
guidelines for the initial pilot plant tests. Malathion was present in the influent at a
concentration of 0.5 mg/L. The bench results indicated that the pilot plant should
operate with a retention time of 2 to 5 hours, and an ozone dosage of 2 to 5 g/L at
an ozone concentration of 2 to 3 weight percent. Under these conditions, the
concentration of raalathion in the effluent reached nondetectable levels in 2 hours.
Using the operating conditions derived from the bench-scale work, the pilot plant
was put into operation. The pilot plant removed 93% of the influent 24 mg/L
raalathion when operated at a 12 L/min wastewater flow and an ozone dosage of
3.1 g/L. Lime pretreatment increased the removal to 995%.
Data were not found for the removal of raalathion from drinking water by aeration.
However, because of its low vapor pressure, malathion probably is not amenable to
removal by aeration.
16
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Maiathion April, 1995.
IX. REFERENCES
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Carey, A.E., J.A. Gowen, H. Tai, W.G. Mitchell and G.B. Wiersma. 1979. Pesticide residue
levels in soils and crops from 37 states, 1972. National Soils Monitoring Program (IV).
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Carey, A.E., J.A. Gowen, H. Tai, W.G. Mitchell and G.B. Wiersma. 1978. Pesticide residue
levels in soils and crops, 1971. National Soils Monitoring Program (III). Pestic. Monit.
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Chhabra, M.L., G.C. Sepha, S.R. Jain, R.R. Bhagwat and J.D. Khandekar. 1970. E.C.G. and
necropsy changes in organophosphorus compound (rnalathion) poisoning. Indian J.
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Degraeve, N. and J. Moutschen. 1984. Genetic and cytogenic effects induced in the mouse
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Dulout, F.N, O.A. Olivero, H. von Guradze and M.C. Pastori. 1982. Cytogenic effect of
rnalathion assessed by the micronucleus test. Mut. Res. 105:413-416.
Dulout, F.N., M.C. Pastori and O.A. Olivero. 1983. Malathion-induced chromosomal
aberrations in bone-marrow cells of mice: dose-response relationships. Mut. Res.
122:163-167.
Dzwonkowska, A. and H. Huebner. 1986. Induction of chromosomal aberrations in the
Syrian hamster by insecticides tested in vivo. Arch. Toxicol. 58:152-156.
17
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Malathion April, 1992,
Eichelberger, J.W. and JJ. Lichtenberg. 1971. Persistence of pesticides in river water.
Environ. Sci. Technol. 5:541-544.
Feldraan, R J. and H.I. Maibach. 1974. Percutaneous penetration of some pesticides and
herbicides in man. Toxicol. Appl. Pharmacol. 28:126-132.
Feldraan, RJ. and H.I. Maibach. 1970. Absorption of some organic compounds through the
skin in man. J. Invest. Dermatol. 54:399-404.
Frawley, J.P., H.N. Fuyat, E.G. Hagan, J.R. Blake and O.G. Fitzhugh. 1957. Marked
potentiation in mammalian toxicity from simultaneous administration of two
anticholinesterase compounds. J. Pharmacol. Exp. Ther. 121:96-106.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol. 14:515-534.
Golz, H.H. and C.B. Shaffer. 1955. Malathion: summary of pharmacology and toxicology.
Central Medical Department, American Cyanamid Co.
Gupta, S.K., M.K. Pandga, J.P. Jani and S.K Kashyup. 1980. Health risks in ultra-low
volume (ULV) aerial spray of malathion for mosquito control. J. Environ. Sci. Health.
B15:287-294.
Hanna, WJ. and E. Choo-Kang. 1983. Malathion poisoning: A report of 2 cases. West
Indian Med. J. 32:109-111.
Hayes, W.J., Jr. 1971. Studies on exposure during the use of anti-cholinesterase pesticides.
Bull. World Health Organ. 44:277-288.
Hazelton, L.W. and E.G. Holland. 1953. Toxicity of malathion: Summary of mammalian
investigations. AMA Arch. Indus. Hyg. Occup. Med. 8:399-405.
Huang, C.C. 1973. Effect on growth but not on chromosomes of the mammalian cells after
treatment with three organophosphorus insecticides. Proc. Soc. Exp. Biol. Med.
142:36-40.
Huff, J.E., R. Bates, S.L. Eustis, J.K. Haseman and E.E. McConnell. 1985. Malathion and
malaoxon: histopathology reexamination of the National Cancer Institute's
carcinogenesis studies. Environ. Res. 37:154-173.
IARC. 1983. International Agency for Research on Cancer. IARC monographs on the
evaluation of carcinogenic risk of chemicals to humans. Vol. 30. Lyon, France: IARC.
pp. 103-129.
18
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Malathion April, 199JL
Jorgenson, TA^ CJ. Rushbrook and G.W. Newell. 1976. In vivo mutagenesis investigations
of ten commercial pesticides. Toxicol. Appl. Pharmacol. 37:109 (abstract).
Kalow, W. and A. Marton. 1961. Second-generation toxicity of raalathion in rats. Nature
192(4801):464-465.
Khera, K.S., C. Whalen and G. Trivett. 1978. Teratogenicity studies on linuron, malathion
and methoxychlor in rats. Toxicol. Appl. Pharmacol. 45:435-444.
Kimbrough, R.D. and T.B. Gaines. 1968. Effect of organic phosphorus compounds and
alkylating agents on the rat fetus. Arch. Environ. Health. 16:805-808.
LaFleur, K. 1979. Sorption of pesticides by model soils and agronomic soils: Rates and
equilibria. Soil Science 127:94-101.
Lehman, AJ. 1959. Appraisal of the safety of chemicals in foods, drugs, and cosmetics.
Association of Food and Drug Officials of the United States, Quarterly Bulletin.
Meister, R.T., ed. 1987. Farm Chemicals Handbook. Willoughby, OH: Meister Publishing
Co.
Mendoza, C.E. and J.B. Shields. 1977. Effects on esterases and comparison of IM and LDSO
values of malathion in suckling rats. Bull. Environ. Contain. Toxicol. 17:9-15.
Miller, S. and MA. Shah. 1982. Cholinesterase activities of workers exposed to
organophosphorus insecticides in Pakistan and Haiti and an evaluation of the tintometric
method. J. Environ. Sci. Health. B17(2):125-142.
Moeller, H.C. and J.A. Rider. 1962. Plasma and red blood cell cholinesterase activity as
indications of the threshold of incipient toxicity of ethyl-p-nitrophenyl
thionobenzenephosphonate (EPN) and malathion in human beings. Toxicol. Appl.
Pharmacol. 4:123-130.
Morgade, C. and A. Barquet. 1982. Body distribution of malathion and its metabolites in a
fatal poisoning by ingestion. J. Toxicol. Environ. Health 10:321-325.
NAS. 1977. National Academy of Sciences. Drinking Water and Health. Vol. 1.
Washington, DO National Academy Press, p. 622.
NCI. 1979a. National Cancer Institute. Bioassay of malathion for possible carcinogenicity.
U.S. Department of Health, Education & Welfare. (Tech. Rep. Ser. No. 192; DHEW
Publ. No (Nffl) 79-1748.) Washington, DC.
19
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Malathion April,
NCI. 19795. National Cancer Institute. Bioassay of raalathion for possible carcinogenicity.
U.S. Department of Health, Education & Welfare. (Tech. Rep. Ser. No. 135; DHEW
Publ. No. (NIH) 79-1390.) Washington, DC
NCI. 1978. National Cancer Institute. Bioassay of raalathion for possible carcinogenicity.
U.S. Department of Health, Education & Welfare. (Tech. Rep. Ser. No. 24; DHEW
Publ. No. (NIH) 78-824.) Washington, DC.
NIOSH. 1985. National Institute for Occupational Safety and Health. NIOSH Pocket
Guide to Chemical Hazards. U.S. Department of Health and Human Services. Public
Health Service. Printing (DHEW [NIOSH] Pub. No. 78-120). Washington, DC
Pednekar, M.D., S.R. Gandhi and MS. Netrawali. 1987. Evaluation of mutagenic activities
of endosulfan, phosalone, malathion and permethrin, before and after metabolic
activation, in the Ames Salmonella test. Bull. Environ. Contam. Toxicol. 38:925-933.
Sharma, S.R., H.S. Rathore and S.R. Ahmed. 1987. Studies on removal of raalathion from
water by means of activated charcoal. Ecotoxicology and Environmental Safety 14:22-29.
Shiau, S.Y., R.A. Huff, B.C. Wells and I.C. Felkner. 1980. Mutagenicity and DNA damaging
activity for several pesticides tested with Bacillus subtilis mutants. Mut. Res. 71:169-179.
Siddiqui, M.KJ., T.D. Seth and MC. Saxena. 1980. Acetyicholinesterase activity in red
blood cells of healthy, diseased and exposed persons. Indian J. Med. Sci. 34:289-292.
Su, M.Q., F.K. Kinoshita, J.P. Frawley and KJP. DuBois. 1971. Comparative inhibition of
aliesterases and cholinesterase in rats fed eighteen organophosphorus insecticides.
Toxicol. Appl. Pharmacol. 20:241-249.
Tegeris Laboratories, Inc.1 1987. One-year oral toxicity study in purebred beagle dogs with
AC6.601. Laboratory study number 85010. MRTD 401885-01. February 10,1987.
U.S. EPA. 1990. U.S. Environmental Protection Agency. 1990. Carcinogenicity peer review
of malathion. Office of Pesticide Programs. April 12.
U.S. EPA. 1988a. U.S. Environmental Protection Agency. Guidance for the reregistration
of pesticide products containing malathion as the active ingredient. Washington, DC:
Office of Pesticides and Toxic Substances. January.
U.S. EPA. 1988b. U.S. Environmental Protection Agency. Method 507, the determination
of nitrogen and phosphorus containing pesticides in water by gas chromatography with a
'Confidential Business Information submitted to the Office of Pesticide Programs.
20
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Malathion April, 1992-
nitrogen-phosphorus detector. Cincinnati, OH: Environmental Monitoring and Support
Laboratory. December.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Fed. Reg. 51:33992-34003.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Method 622, the determination of
organophosphorus pesticides in industrial and municipal wastewater. Cincinnati, OH:
Environmental Monitoring and Support Laboratory.
Velazquez, A., A. Creus, N. Xamena and R. Marcos. 1987. Lack of rautagenicity of the
organophosphorus insecticide malathion in Drosophila melanogaster. Environ. Mutagen.
9:343-348.
Warren, M., H.C. Spencer, F.C. Churchill, VJ. Francois, R. Hippolyte and MA. Staiger.
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monitoring of urinary metabolites and blood cholinesterase levels. Bull. World Health
Org. 63:353-360.
Wester, R.C., H.I. Mailbach, DA.W. Bucks and R.H. Guy. 1983. Malathion percutaneous
absorption after repeated administration to man. Toricol. Appl. Pharmacol. 68:116-119.
Windholz, M., S. Budavari, R.F. Blumetti and E.S. Otterbein, eds. 1983. The Merck
indexAn encyclopedia of chemicals, drugs, and biologicals, 10th ed. Rahway, NJ:
Merck and Co., Inc. p. 813.
Wiszkowska, H., I. Kulamowicz, A. Malinowska and Z. Walter. 1986. The effect of
malathion on RNA polymerase activity of cell nuclei and transcription products in
lymphocyte culture. Environ. Res. 41:372-377.
WHO. 1982. World Health Organization. Recommended health-based limits in
occupational exposure to pesticides: Malathion. WHO Technical Report Series No.
677. pp. 13-37.
Worthing, CR. and S.B. Walker, eds. 1987. The pesticide manual A world compendium,
8th ed. The British Crop Protection Council. United Kingdom: Thornton Health.
Zeff, J.D. and JA. Harris. 1984. Chemistry and application of ozone and ultraviolet light for
water reuse-pilot plant demonstration. Proceedings, Industrial Waste Conference.
38:105-116.
21
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EPA 0553
RX000027511
April,
p-NITROPHENOL
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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-Nitrophenol April, 1992.
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 100-02-7
Structural Formula
OH
Synonyms
4-Nitrophenol; 4-hydroxynitrobenzene; phenol,4-nitro;
p-hydroxynitrobenzene; para-nitrophenol (Windholz et al., 1989).
Uses
/7-Nitrophenol is used in the manufacture of the pesticides, ethyl and methyl
parathions and n-acetyl-p-arainophenol. Other uses include production of dyes,
fungicides and leather preservatives. It is also used as an analytical indicator
(Windholz et al., 1989; Hawley, 1987; NIOSH, 1981).
Properties (NIOSH, 1981)
Chemical Formula CjHsNOj
Molecular Weight 139.11
Physical State colorless to slightly
yellow crystals
Boiling Point 279ฐC (decomposes)
Melting Point 113-114ฐC
Density (20ฐC) 1.479 g/cmj
Vapor Pressure (146ฐC) 2.2 mm Hg
Specific Gravity (20ฐC) 1.479
Water Solubility (25ฐC) 16 g/L
Log Octanol/Water Partition
Coefficient 1.91
Taste Threshold 43.4 rag/L
Odor Threshold (water) 2.5 rag/L
Conversion Factor 1 rng/rn3 = 0.18 ppm
(in air at 25ฐC) (rounded from
0.176 ppm)
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P-Nitrophenol April, 1992.
Occurrence
Alber et al. (1989) detected /7-nitrophenol in rain and snow at concentrations
ranging from 0.49 jig/L to 17.1 /xg/L. p-Nitrophenol has also been detected in the
secondary effluent from publicly owned treatment works and in industrial effluents
in surface waters (Ellis et al., 1982). Alber et al. (1989) attributed the high
concentrations of /7-nitrophenol and other phenols in rain and snow to
photochemical reactions of aromatic hydrocarbons, NOX and OH' radicals.
Lokke (1985) attributed the presence ofp-nitrophenol in Danish soils to leaching
from chemical waste disposal sites.
Environmental Fate
p-Nitrophenol is soluble in water and is not removed by aeration (Batchelder,
1975). The bioconcentration of p-nitrophenol in fish and other aquatic life is
considered insignificant based on its low water solubility and low octanol/water
partition coefficient (< 100) (Leo et al., 1971).
The biodegradation ofp-nitrophenol in lake and river waters is faster at higher p-
nitrophenol concentrations than at lower concentrations (Zaidi et al., 1989; Siragusa
and DeLaune, 1986). Zaidi et al. (1989) conducted a study to assess the factors
limiting biodegradation of chemicals present at low concentrations. p-Nitrophenol
at 50, 75 and 100 /ig/L was extensively decomposed in lake water inoculated with
Corynebacterium sp. but not in the uninoculated water. However, the presence of
this bacterium resulted in the mineralization of only 35% of p-nitrophenol present
at a concentration of 26 /tg/L. Addition of 250 mg of cyclohexiraide, an inhibitor of
microbial growth, increased biodegradation to 75% within 70 hours. This effect of
cyclohexirnide was absent at higher levels of p-nitrophenol and in uninoculated
waters. Contaminants may be of limited impact if the compound is found at low
concentrations or if bacterial inhibitors are present in the polluted environment. In
other studies (Spain et al., 1980; Spain and Van Veld, 1983), pre-exposure of
coastal sediment to various compounds affected their biodegradation rates upon re-
exposure of the sediment to the same compounds.
^-Nitrophenol was totally decomposed by soil microflora within 16 days at 25 ฐC
(Alexander and Lustigman, 1966). The half-life for degradation of/;-nitrophenol
was 10 times higher under anaerobic conditions than under aerobic conditions.
This was due to the absence or very small number of microorganisms in the subsoil
and to the higher acidity of the subsoil (pH 4.7) compared with the topsoil (pH 5.4)
(Lokke, 1985). The half-life for 2 rag/kg of ^-nitrophenol was 0.7 to 1.2 days in the
topsoil under aerobic conditions and 14 days under anaerobic conditions. In the
subsoil, the half-life of p-nitrophenol was 40 days under aerobic conditions, but
under anaerobic conditions the degradation of ^-nitrophenol was minimal.
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P-Nitrophenol April, 1992.
The sorption and desorption of p-nitrophenol in sediment from a Louisiana gulf
coast estuary were studied under aerobic and anaerobic conditions (Siragusa and
DeLaune, 1986). The results showed that decomposition of p-nitrophenol,
measured as release of radiolabeled carbon dioxide, was several orders of
magnitude faster under aerobic than under anaerobic conditions (1.04 x 10"J versus
2.95 x 10's /xg/day/g dry sediment, respectively). The results suggested thatp-
nitrophenol is rapidly desorbed from the water column to the bottom sediment,
where it may persist for years owing to the slower rate of biodegradation under
anaerobic conditions in estuarine sediment than occurs under aerobic conditions.
Moreover, the equilibration rate of p-nitrophenol sorption-desorption did not affect
the re-release of p-nitrophenol.
The results from a biodegradability laboratory test carried out by Tabak et al.
(1981) on 114 organic priority pollutants showed rapid acclimation and complete
biodegradation of p-nitrophenol. The static culture flask-screening method utilizing
biochemical oxygen demand (BOD) dilution water was used. Five dr 10 rag/L of p-
nitrophenol and 5 mg/L of yeast extract as the synthetic medium were incubated for
7 days under static conditions at 25ฐC in the dark. Three weekly subcultures were
done with final incorporation of wastewater as the microbial inoculum.
The biodegradation of p-nitrophenol was compared in three laboratory test systems
and in the field utilizing freshwater from a pond (Spain et al., 1984). The water
samples were treated simultaneously with 0.2 mg/L of radiolabeled p-nitrophenol
and incubated at 19ฐC. The results showed that p-nitrophenol was biodegraded
more slowly by microorganisms acclimated in the laboratory than by those found in
the pond. The acclimation periods in the laboratory systems that contained
sediment were similar to those in the pond, indicating that the results from
laboratory test systems can be compared with those from the field.
III. PHARMACOKINETICS
Absorption
Limited information is available on the gastrointestinal absorption of p-nitrophenol
following oral administration. Robinson et al. (1951) administered 200 mg/kgp-
nitrophenol in water to rabbits via a stomach tube. It was concluded that p-
nitrophenol was rapidly absorbed from the gastrointestinal tract following oral
administration in rabbits, since it was completely excreted within 1 day.
Based on the results of a study using 60- to 80-day-old hairless mice (SKH-hr-1
strain-), Jetzer et al. (1988) concluded that p-nitrophenol is absorbed through the
skin, and that the rate of permeation is temperature-dependent. The results
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P-Nitrophenol April, 1992.
showed that permeation rates decreased with decreasing temperatures, but were
higher in non-isothermal studies than in low-temperature isothermal studies.
Distribution
No information was available on the tissue distribution of p-nitrophenol following
oral administration. Groups of 200- to 350-gram male Wistar rats were
intravenously (1.6, 4.0 or 8.0 rag/kg) and intra-arterially (4.0 and 8.0 mg/kg) injected
with p-nitrophenol (Machida et al., 1982). p-nitrophenol rapidly equilibrated
between red blood cells and plasma. This equilibrium was so rapid that
p-nitrophenol-glucuronide and -sulfate could be identified 1 minute after dosing.
The ratio of p-nitrophenol concentrations in red blood cells and plasma was the
same for all dosagess at the 2-, 4- and 8-minute sample collection intervals.
Metabolism
p-Nitrophenol undergoes both phase I (reduction) and phase II (glucuronidation
and sulfation) metabolic transformation in the liver. In addition to sulfation and
glucuronidation, p-nitrophenol undergoes reduction to panz-amino-phenol
(Robinson et al., 1951). Rabbits (age and sex not reported) were orally
administered 200 mg/kg of p-nitrophenol in a water suspension, and 24-hour urine
samples were collected. Excretion was complete in 1 day. Most of the dose (82 to
92%) was excreted as unchanged p-nitrophenol, and 11 to 19% was excreted as the
reduced ami no compound. Of the parent chemical, 59 to 62% was conjugated and
excreted as glucuronide, and 13 to 21% was excreted as organic sulfates.
Meerman et al. (1987) reported that there was no sex difference in conjugation of
p-nitrophenol in the rat. Male and female Wistar rats (60 days old) were injected
via the lateral tail vein with 60 /imol/kg of p-nitrophenol. Both p-nitrophenol-
sulfate and -glucuronide were detected in the 24-hour urine samples collected.
Both female and male rats excreted the same amount of p-nitrophenol-sulfate and
-glucuronide (35 and 37% of the administered dose, respectively).
p-Nitrophenol also undergoes extrahepatic metabolism. Machida et al. (1982)
found that formation of p-nitrophenol-glucuronide occurred in the lungs and
kidneys of 200- to 350-gram male Wistar rats administered p-nitrophenol. The rats
were administered the p-nitrophenol via three routes as follows: 1.6, 4 and 8 mg/kg
in physiological saline solution were given intravenously through the left femoral
vein, 4 and 8 rag/kg were injected intra-arterially and 1.6 and 4 mg/kg were injected
into the hepatic portal vein. Administration of 4mg/kg of p-nitrophenol via the
femoral vein and via the portal vein produced different p-nitrophenol plasma
concentrations. Similarly, the area under the curve produced by intravenous
femoral vein administration (AUV^) and the area under the curve produced by
portal vein injection (AUV ) at 1.6 and 4 rag/kg were significantly (p < 0.05)
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P-Nitrophenol April, 199Z
different. No significant difference was seen between AUC^, values and the areas
under the curve produced by intra-arterial dosing. Nevertheless, pulmonary
metabolism ofp-nitrophenol was not ruled out. The hepatic metabolic clearance
rate for the femoral dose was 17 to 21 mL/min, which corresponded to the hepatic
blood flow rate (10-20 mL/min). This suggested that both sulfation and
glucuronidation ofp-nitrophenol are limited by hepatic blood flow.
Livers isolated from 20- to 30-gram male Swiss mice were perfused in situ with 1 to
100 pM p-nitrophenol (Sultatos and Minor, 1985). Preliminary experiments showed
that less than 1% of the p-nitrophenol appeared in the bile as unchanged
compound or metabolites, indicating that biliary excretion is not a major pathway.
In animals perfused with up to 4 yiM. p-nitrophenol, the sulfate conjugate was the
only metabolite. However, in animals perfused with p-netrophenal at
concentrations of 15 /iM or higher, there was more unchanged p-nitrophenol and
p-nitrophenol-glucuronide than p-nitrophenol-sulfate in the bile. The steady-state
concentrations of p-nitrophenol-glucuronide and sulfate were reached at 25 and
15 /xM, respectively. This suggested that sulfation is kinetically more favorable than
glucuronidation at the lower concentrations, while the reverse is true at higher
concentrations.
In perfused livers from female Sprague-Dawley rats (140 to 150 g), 0.06 mM
p-nitrophenol was rapidly hydroxylated to 4-nitrocatechol; the latter competed with
p-nitrophenol for conjugation with glucuronic acid and sulfate (Reinke and Moyer,
1985). The oxidation ofp-nitrophenol occurred via the reduction of nicotinamide-
adenine dinucleotide phosphate (NADP) at pH 7.1-7.4.
Excretion
The in vivo Specific Activity Difference Ratio (SADR) technique was used to
measure the excretory metabolites ofp-nitrophenol in rat kidneys (Tremaine et al.,
1984). Male Sprague-Dawley rats (335 to 400 g) were given continuous infusions of
a solution containing 2 /imol l4C-p-nitrophenol/min/kg via the jugular vein; the
infusions were given at a constant rate (between 0.02 and 0.08 mL/min). Either p-
nitrophenol-glucuronide or p-nitrophenol-sulfate was infused at 0.3 /xmol/min/kg.
Urine and arterial blood samples were collected. The mean p-nitrophenol plasma
concentration was 31 jiM. A steady state was reached, indicating first-order
elimination kinetics for each of the infused compounds. During steady-state
conditions, 75% of the infused radiolabel was recovered in urine. In the urine, 99%
of the radiolabel was found in p-nitrophenol-sulfate and p-nitrophenol-glucuronide.
The rat kidney formed both of these conjugates at equal rates. No unchanged
14C-p-nitrophenol was identified in the urine. The renal clearance for p-nitrophenol
in the rat was 6.4 mL plasma/min/kg/kidney, which is 1.6 times the glomerular
filtration rate. It was reported that the kidney accounted for a minimum of 20% of
the endogenously formed conjugates ofp-nitrophenol. The bile does not appear to
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P-Nitrophenol April, 1992.
account for the significant amounts of p-nitrophenol excretion in the rat (0.9% of
the infused radiolabel was recovered in the bile).
Robinson et al., 1951 administered 200 mg/kg ofp-nitrophenol to rabbits via a
stomach tube. The compound was exclusively and completely excreted in the 24-
hour urine sample. Most of the dose (82 to 92%) was excreted as unchanged p-
nitrophenol, and 11 to 19% was excreted as the reduced amino compound. Of the
unchanged p-nitrophenol, 59 to 62% was conjugated and excreted as glucuronide,
and 13 to 21% was excreted as organic sulfates.
Gorge et al., 1987 studied the excretion and metabolism of p-nitrophenol in two
frog species. The frogs were injected with 3 to 5 mg/kg p-nitrophenol (2 to 3
/xCi/animal) into the dorsal lymphatic sac through the thigh muscle. Within 24
hours, both frogs excreted 90 to 95% of the administered dose, p-nitrophenol was
metabolized primarily by glucuronidation (12.5%) and sulfation (46%). Smaller
amounts of /vnitrophenol were reduced to p-nitrocatechol and then conjugated.
The kinetics of excretion fit a two-compartment model. Unchanged p-nitrophenol
accounted for 35% of the excreted dose.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
No information was found in the available literature regarding short-term exposure
to p-nitrophenol in humans.
Long-term Exposure
Nitro-amino compounds are thought to have nephrotoxic effects. Yoshida et al.
(1989) conducted a clinical cross-sectional study with 62 male workers (aged 20 to
64 years) exposed to aromatic nitro-amino compounds, including ^-nitrophenol, in a
chemical factory. The urinary enzyme activities of workers were measured as an
index of renal damage; 27 office workers served as controls. The duration of
exposure ranged from 3.6 to 18 years for both exposed and control groups.
Information on smoking and alcohol consumption was obtained through a
questionnaire. The environmental concentrations of the nitro-amino compounds
were not measured directly, but the investigators identified a probable exposure
level of 0.3 mg/mj. This value was based on a study conducted by the authors in
which workers were exposed to 0.38 mg/m3 p-nitrochlorobenzene. In these workers,
the authors found 0.5 mg of urinary diazo metabolites per mg of creatinine. In the
present study, the exposed workers also excreted diazo metabolites at 0.5 mg/mg
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P-Nitrophenol April, 1992,
creatinine. In addition, for two decades the amounts handled by the workers were
constant. Therefore, these workers were probably exposed to aromatic nitro-amino
compounds at concentrations greater than 03 mg/m3, since the aromatic nitro-
amino compounds can be absorbed through the skin and the respiratory tract of
workers. In exposed workers who were 50 years of age or older, there was a
significant increase (p < 0.05) in urinary n-acetyl-beta-
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P-Nitrophenol April, 1992.
transient, increases in the rate and depth of respiration when administered to the
normally innervated common carotid artery. The effect was absent when the same
dose was injected into a denervated common carotid artery. The investigators
concluded that p-nitrophenol and other nitro compounds act indirectly to stimulate
the carotid chemoreceptors, resulting in changes in respiration. A similar
experiment was conducted on the aortic chemoreceptors in the dog. Thirty dogs, 6
to 24 kg, were given 0.2-2.0 mL of a 0.5% solution of p-nitrophenol intra-arterially
through the left common carotid artery; injection of 5 rag of .p-nitrophenol led to an
immediate, but transient stimulation of respiration, producing an indirect, specific
effect on the aortic chemoreceptors (Shen, 1962).
In a subchronic inhalation toxicity study, Sprague-Dawley rats (15/sex, 6 to 7 weeks
of age) were exposed to 0, 1, 5 and 30 rag/m3 of p-nitrophenol dust for 6 hours a
day, 5 days a week, for 4 weeks (Hazleton Laboratories America, Inc., 1983). A
control group was exposed to filtered air. Exposure to/7-nitrophenol caused no
deaths, and no consistent exposure-related effects were noted in hematology values,
clinical chemistry values, gross examination and histopathology, body or organ
weights, or ophthalmoscopic examination.
Dermal/Ocular Effects
/7-Nitrophenol is absorbed from the skin and causes strong skin irritation, but it
does not sensitize skin. Its dermal LDM in guinea pigs is > 1 g/kg. Repeated 10-
day skin applications to guinea pigs caused staining of skin within 24 hours and
caused decreased weight gain within 10 days (Eastman Kodak Company, 1980).
Six 2.5- to 3.0-kg New Zealand White albino rabbits were tested for p-nitrophenol
induced skin irritation. A mass of 0.5 +. 0.007 g of p-nitrophenol moistened with
saline was applied to each of two intact and two abraded sites of the rabbit ear for
24 hours. By the first day of application, the skin stained yellow on all four treated
sites of all six animals. Scarring occurred in four rabbits by day 14. In three of
these four rabbits, dark brown discoloration of the skin was seen prior to the
appearance of scabs or scars (Branch, 1983b).
/>-Nitrophenol (70 rag in 0.1 raL) was instilled into the conjunctival sac of the right
eye of each of six 2.2- to 3.0-kg New Zealand albino rabbits. The untreated eye of
each rabbit served as the control (Branch, 1983c). Moderate-to-severe corneal
cloudiness was seen through day 21 in five animals. Periocular skin and/or fur of all
five animals was stained yellow throughout the study. Blistered conjunctivae were
seen during the first 3 days of application. Scabbing and alopecia of the upper and
lower eyelids were seen in one female rabbit. Corneal neovascularization was found
in five animals from day 7 to day 21. Based on the nature of the corneal
involvement, /7-nitrophenol was classified as corrosive to the eyes.
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P-Nitrophenol April, 1992
Long-term Exposure
Adult Sprague-Dawley rats (20/sex/group) were given p-nitrophenol by gavage daily
for 13 weeks (Hazleton Laboratories America, Inc., 1989). The doses were selected
based on an initial pilot study with six rats/sex/dose, in which 0, 1.0, 10, 50, or 100
mg/kg/day /7-nitrophenol was administered for 4 weeks. No adverse effects were
seen at any dose, so the No-Observed-Adverse-Effect Level (NOAEL) was 100
mg/kg/day. On the basis of the findings from this inital study, the doses chosen
were 25, 70, and 140 rng/kg/day of p-nitrophenol. The dosages were administered
by gavage in water, the control group was administered water only. End points
measured included: clinical observations, body weights, food consumption,
hematology and clinical chemistry values, ophthalmoscopic examination, organ
weights and histopathology. Death occurred in all groups, but was significant
(p < 0.05) only at the high dose in both sexes. Deaths in the high-dose group
might have been caused by the acute toxicity ofp-nitrophenol, including
raethernoglobinemia. In mid-dose females and all high-dose rats, the liver, kidneys,
and lungs were stained dark. The authors suggested that this affect might be due to
respiratory stress. The NOAEL for this study was 25 mg/kg/day, and the Lowest-
Observed-Adverse-Effect Level (LOAEL) was 70 mg/kg/day based on the increased
mortality in both sexes and histopathological findings of low-to-moderate congestion
of the liver, kidneys, and lungs.
The only chronic toxicity study on p-nitrophenol was an abstract obtained from the
U.S.S.R. The data provided information that was qualitative but inadequate to
make quantitative assessment (Makhinya, 1969). Chronic administration (probably
oral, dose not mentioned) of p-nitrophenol to warm-blooded animals adversely
affected "neurohumoral regulation." Higher doses of p-nitrophenol caused gastritis,
enteritis, colitis, hepatitis, neuritis and hyperplasia of the spleen, and inhibition of
oxidation processes.
Reproductive Effects
Plasterer et al. (1985) evaluated the effects of prenatal exposure to selected
chemicals in a screening protocol designed to detect reproductive effects, p-
nitrophenol disserved in com oil was administered via gavage at 0 or 400 mg/kg/day
to groups of 50 pregnant CD-I mice during gestational days 7 to 14. A limited
number of end points were studied: for maternal toxicity, they were mortality and
body weight; for developmental toxicity, they were number of live and dead pups
per litter, pup body weight on days 1 and 3 and gross structural abnormalities on
day 1. Evidence of maternal toxicity included a significantly lower survival rate
(80%) among treated animals, as well as a significantly (p < 0.05) decreased weight
gain (18.7 g versus 22.8 g in the control group). Developmental toxicity was
manifested as a slight (nonsignificant) reduction in the number of live pups per
litter (9.8 versus 10.8 in the control group on day 1).
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P-Nitrophenol April, 199Z
Angerhofer (1985) conducted a two-generation dermal study to investigate the
reproductive effects of p-nitrophenol (a leather fungicide incorporated into combat
boots), p-nitrophenol dissolved in 100% ethanol was applied along the dorsal body
line over a 4- by 10-cra area; the doses were 0 (ethanol control), 0 (saline control),
50, 100 and 250 rag/kg/day, 5 days/week, applied to groups of 24 female and 12
male Sprague-Dawley rats. The F0 generation was exposed for 140 days before
mating. The females continued to be exposed through breeding, gestation, and
lactation. The Ft generation was exposed for 168 days after weaning; the females
again were exposed through breeding, gestation, and lactation. Toxic signs included
a dose-related pattern of dermal irritation (erythema, scaling, and cracking). No
reproductive effects were observed.
Developmental Effects
Kavlock (1990) evaluated the developmental toxicity of p-nitrophenol in a structure-
activity relationship study of phenols, p-nitrophenol (dissolved in a mixture of
water, Tween 20ฎ, propylene glycol, and ethanol in a ratio of 4:4:1:1) was
administered via gavage at 0, 100, 333, 667, or 1,000 rag/kg on gestational day 11 to
groups of 12 to 13 Sprague-Dawley rats. Maternal toxicity end points included
overt signs of toxicity, mortality, body weight and implantation scars at the end of
weaning. Developmental toxicity end points included viability, weight of offspring
at postnatal days 1, 3 and 6, overt malformations, and perinatal loss. Mortality
among dams exposed to 667 and 1,000 mg/kg of />-nitrophenol was 3/13 and 4/12,
respectively (versus 0 in all other groups). At 333, 667, and 1,000 rag/kg, the litter
size on postnatal days 1 and 6 decreased nonsignificantly. The NOAEL was 100
mg/kg, and the LOAEL was 333 rag/kg.
Mutagenicity
p-Nitrophenol (10 to 500 jig/plate) was not rautagenic in the Salmonella
typhimurium/mammaAian rnicrosome plate incorporation assay (Ames test)
performed by McCann et al. (1975).
As part of the National Toxicology Program (NTP), o- and p-nitrophenols were
among 250 coded compounds evaluated by independent laboratories in the
preincubation modification to the standard Ames test (Haworth et al., 1983). Over
a concentration range of 10 to 3333 /ig/plate (para), there was no evidence of a
mutagenic response in S. typhimurium TA1535, TA1537, TA98, or TA100 either in
the absence or presence of liver S9 preparations from Aroclor 1254-induced rats or
hamsters. Similarly, 1 to 100 fig/plate of p-nitrophenol was negative in
preincubation Ames tests conducted with 5. typhimurium TA98 and TA100 (Suzuki
et al., 1983). The inclusion of the co-mutagen norharman into the S9 mix did not
cause an increase in mutant colony counts of either strain exposed to the three
isoraers. The limited data presented by Kawai et al. (1987) support the earlier
findings that p-nitrophenol is not mutagenic under preincubation conditions.
11
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P-Nitrophenol April, 1992,
Under the conditions of the nitro reduction preincubation Ames test performed by
Dellarco and Prival (1989), ? assayed to cytotoric levels, was not
.mutagenic in the presence of rat liver S9-mix containing flavin mononucleotide
(FMN). It was noteworthy that the investigators presented data demonstrating the
ability of FMN to facilitate the nitro reduction of two nonmutagenic nitro
compounds (i.e., nitrobenzene and p-nitrophenol). It was concluded that the lack of
a rnutagenic response by p-nitrophenol in this well-conducted study was not related
to a failure of FMN to reduce this compound.
Nonactivated p-nitrophenol (125 to 2,000 jig/plate) was not genotoric in DNA
repair-deficient strains of 5. typhimurium or Escherichia coli (Rashid and Mumma,
1986).
Major data gaps exist in this genetic toxicology evaluation of p-nitrophenol (i.e., no
mammalian cell gene mutation or cytogenetic assays were found in the available
literature); however, the Ames test has an established record of correctly identifying
mutagenic nitroaromatics as carcinogens (Klopman et al., 1990). Accordingly,
Tennant et al. (1990) have used the weight of evidence from well-conducted
Salmonella gene mutation assays showing no indication of mutagenesis to predict
that p-nitrophenol will not be a carcinogen in the rodent carcinogenicity bioassay
that is currently being performed by the NTP.
Carcinogenicitv
Only one carcinogenicity study was found in the literature. Thirty-one mice
received single biweekly 25-^iL dermal applications of a 20% dioxane solution (2.5
mg) of p-nitrophenol applied to the back for 12 weeks. No skin tumors were
observed during the study (Boutwell and Bosch, 1959).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day, Longer-term
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:
TTA (NOAEL or LOAEL) x (bw) ... . .4
HA = (UF) x (_ L/day) = mg/L (rฐunded tO
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P-Nitrophenol April, 1992.
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/OW guidelines.
L/day = assumed daily water consumption of a child (1 L/day) or
an adult (2 L/day).
One-day Health Advisory
Only acute LDjo studies of p-nitrophenol were available. The acute oral LDK values in
rats ranged from 50 to 230 mg/kg (Eastman Kodak Company, 1980; Vernot et al., 1977;
Branch, 1983a). In mice, the oral LDj,, values ranged from 200 to 470 mg/kg (Vernot et al.,
1977; Eastman Kodak, 1980). Clinical signs of toxicity observed in these studies included
salivation, lethargy, ptosis, prostration, convulsions, dyspnea, and death. In rabbits, the oral
MLD was 600 to 900 mg/kg (Monsanto Chemical Co., 1956). The intraperitoneal LD^ in rats
was 50 to 97 mg/kg, whereas that in mice was 50 mg/kg (Von Oettingen, 1941; Eastman
Kodak Company, 1980). In a developmental toxicity study by Kavlock (1990), Sprague-
Dawley rats were not affected when exposed to a single oral dose of 100 mg/kg p-nitrophenol
(lowest dose tested) on day 11 of gestation. Although this dose, 100 mg/kg, appears to be a
NOAEL in this study, the acute oral LD^ in rats was also reported to vary between 50 and
230 mg/kg. As noted, an LD^ of 50 mg/kg is half the dose considered to be a NOAEL in the
Kavlock (1990) study. Therefore, in the absence of adequate short-term studies to calculate a
One-day HA for a 10-kg child, it is recommended that the Longer-term HA for a 10-kg child
calculated below, 0.833 mg/L (rounded to 800 pg/L), be used as a conservative estimate of the
One-day HA value for p-nitrophenol.
Ten-day Health Advisory
No adequate short-term study was available to calculate a Ten-day HA for
p-nitrophenol. The range-finding experiment performed to determine the doses for the study
by Hazleton (1989) did not result in any toxicity at the highest dose used (100 mg/kg) after 4
weeks of exposure. However, because of the design of this study, these data are inadequate
for use in calculating the Ten-day HA, especially since the acute oral LDj,, in rats may be as
low as 50 mg/kg. Therefore, it is recommended that the Longer-term HA for a 10-kg child
below, 0.833 mg/L (rounded to 800 /ig/L), be used as a conservative estimate of the Ten-day
HA for p-nitrophenol.
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P-Nitrophenol April, 1992.
Longer-term Health Advisory
The subchronic study reported by Hazleton Laboratories America, Inc. (1989) in rats
was used to determine the Longer-term HA. ^-nitrophenol was administered daily for 13
weeks to Sprague-Dawley rats of both sexes at 0, 25, 70, and 140 mg/kg. Mortality occurred
at the mid and high doses but was significant only at the high dose. In raid-dose females and
all high-dose rats, p-nitrophenol caused dark discoloration of the lungs, livers and kidneys.
This effect was thought to be an indirect result of respiratory stress. The NOAEL in this
study was 25 mg/kg/day, and the LOAEL value was 70 mg/kg/day based on lung, liver, and
kidney effects and on the unscheduled mortalities.
The Longer-term HA for the 10-kg child is calculated as follows:
Longer-term HA = (25 mg/kft/dav) (10 kg) =. 0.833 rag/L (rounded to 800 Mg/L)
(300) (1 L/day)
where:
25 mg/kg/day = NOAEL, based on the absence of clinical signs and only
minimal to moderate congestion noted in the
histopathological examination (Hazleton Laboratories,
1989).
10 kg = assumed weight of a child.
300 = uncertainty factor, this 100-fold uncertainty factor was
chosen in accordance with EPA or NAS/OW guidelines
for use of a NOAEL from an animal study. An extra
three-fold UF was used for lack of adequate
reproductive/developmental data.
1 L/day = assumed water consumption of a 10-kg child.
The Longer-term HA for the 70-kg adult is calculated as follows:
Longer-term HA = ^/ffi8??]/'7? k^ = 2.916 mg/L (rounded to 3,000
where:
25 mg/kg/day = NOAEL, based on the absence of clinical signs and only
minimal to moderate congestion noted in the
histopathological examination (Hazleton Laboratories,
1989).
14
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P-Nitrophenol April, 1992.
70 kg = assumed weight of an adult.
300 = uncertainty factor, this 100-fold uncertainty factor was
chosen in accordance with EPA or NAS/OW guidelines
for use of a NOAEL from an animal study. An extra 3-
fold UF was used for lack of adequate
reproductive/developmental data.
2 L/day = assumed water consumption of a 70-kg adult.
Note that the Longer-term HA for the 70-kg adult, 2.9 rag/L, is also close to the water-
odor threshold ofp-nitrophenol of 2.5 mg/L.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure that is
attributed to drinking water and is considered protective of noncarcinogenic adverse health
effects over a lifetime exposure. The Lifetime HA is derived in a three-step process. Step 1
determines the Reference Dose (RfD), formerly called the Acceptable Daily Intake (ADI).
The RfD is an 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 known, probable or possible carcinogen, according to the
Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then caution must
be exercised in making a decision on how to deal with possible lifetime exposure to this
substance. For human (A) or probable human (B) carcinogens, a Lifetime HA is not
recommended. For possible human carcinogens (C), an additional 10-fold safety factor is
used in the calculation of the Lifetime HA. The risk manager must balance this assessment
of carcinogenic potential and the quality of the data against the likelihood of occurrence and
significance of health effects related to noncarcinogenic end points of toxicity. To assist the
risk manager in this process, drinking water concentrations associated with estimated excess
lifetime cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult
drinking 2 L of water per day are provided in the Evaluation of Carcinogenic Potential
section.
The study by Hazleton Laboratory America, Inc. (1989) has been selected to serve as
the basis for calculation of the Reference Dose (RfD). In this study, adult Sprague-Dawley
15
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P-Nitrophenol April, 199Z
rats (20/sex/group) were dosed daily for 13 weeks with p-nitrophenol at dosages of 25, 70, and
140 mg/kg/day. The end points measured included clinical observations, body weights, food
consumption, hematology and clinical chemistry values, ophthalmoscopic examination, organ
weights, and histopathology. The NOAEL was 25 rag/kg/day based on the absence of clinical
signs, and the LOAEL was 70 mg/kg/day based on miniraal-to-moderate congestion of the
liver, kidneys, and lungs. Other studies were not available in the literature.
Using this study, the Lifetime HA is derived as follows:
Step 1: Determination of the RfD
RfD = (25 mY = Qm rag/kg/day
(3,000)
where:
25 rag/kg/day = NOAEL, based on the absence of clinical signs and only
minimal to moderate congestion noted in the histopathologjcal
examination (Hazleton Laboratories, 1989).
3,000 = uncertainty factor, this uncertainty factor was chosen in
accordance with EPA or NAS/OW guide- lines for use of a
NOAEL from an animal study of less-than-lifetime duration.
An extra three-fold UF was used for lack of adequate
reproductive/developmental data.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-008 mfc/kg/day) (70 kg) = Q 2g mg/L (rounded to 300 Mg/L)
(2 L/day)
where:
0.008 mg/kg/day = RfD.
70 kg = assumed weight of an adult.
2 L/day = assumed water consumption of a 70-kg adult.
Step 3: Determination of the Lifetime HA
Lifetime HA = 0.28 mg/L x 20% = 0.056 mg/L (rounded to 60 /tg/L)
16
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P-Nitrophenol April, 1992.
where:
0.28 mg/L = DWEL.
20% = assumed relative source contribution from drinking water.
Evaluation of Carcinogenic Potential
Only one dermal toxicity study of/7-nitrophenol was found in the literature.
Thirty-one mice received single biweekly 25-/iL applications of a 20% dioxane
solution (2.5 mg) of p-nitrophenol applied to the back for 12 weeks. No skin tumors
were observed throughout the duration of the study (Boutwell and Bosch, 1959).
Applying the criteria described by the U.S. EPA (U.S. EPA, 1986), ^-nitrophenol
may be classified in Group D: not classifiable. This category is for agents with
inadequate evidence of carcinogenicity in animals.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
A taste threshold of 43.4 rag/L was reported by National Academy of Sciences
(1982), and a water-odor threshold of 2.5 mg/L was reported in the Handbook of
environmental data on organic chemicals (Verschueren, 1983).
VII. ANALYTICAL METHODS
Gas chromatography (GC) and high-performance liquid chromatography (HPLC)
have been used to identify and quantify /r-nitrophenol in environmental samples and
in human urine (Kirby et al., 1979; Alber et ah, 1989). The detection limits for GC
and HPLC were 0.5 /zg/L. However, Alber et al. (1989) reported that the resolution
of the GC method was better than that obtained from HPLC analysis.
VTIL TREATMENT TECHNOLOGIES
To be prepared by the U.S. EPA.
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P-Nitrophenol April, 199Z.
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Jetzer, W.E., S.Y.E. Hou, A.S. Huq, N. Duraiswamy, N.F.H. Ho and G.L. Flynn. 1988.
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aromatic nitro compounds. Japn. J. Ind. Health 29:34-54.
Kirby, K.W., J.E. Keiser, J. Groene and E.F. Slach. 1979. Confirmation of /wa-nitrophenol
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759.
Klopman, G., M.R. Frierson and H.S. Rosenkrantz. 1990. The structural basis of the
mutagenicity of chemicals in Salmonella typhimurium: The Gene-Tox database. Mutat.
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Leo, A., C. Hanch and D. Elkins. 1971. Partition coefficients and their uses. Chem. Rev.
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Lokke, H. 1985. Degradation of 4-nitrophenol in two Danish soils. Environ. Pollut. (series
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Machida, M., Y. Morita, M. Hayashi and S. Awazu. 1982. Pharmacokinetic evidence for the
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P-Nitrophenol April, 1992-
Makhinya, A.P. 1969. Comparative hygienic and sanitary-toxicological studies on nitrophenol
isoraers in relation to their normalization in drinking waters. Prom. Zagryazgneniya
Vodoemov. 9:84-85. Chemical Abstract 72:4723/c.
McCann, J., E. Choi, E. Yamasaki and B.N. Ames. 1975. Detection of carcinogens as
mutagens in the Sabnonella/micTOsome test: Assay of 300 chemicals. Proc. Nat Acad
Sci. USA 72:5135-5139.
Meerman, J.H.N., C. Nijland and GJ. Mulder. 1987. Sex differences in sulfation and
glucuronidation of phenol, 4-nitrophenol and n-hydroxy-2-acetylaminofluorine in the rat
in vivo. Biochera. Pharmacol. 36(16): 2605-2608.
Monsanto Chemical Co. 1956. A certificate of analysis and summary of toxicological
investigation of ^ara-nitrophenol. Project no. Y-56-56. Report no. 86-890000356.
Washington, DC: U.S. Environmental Protection Agency.
National Academy of Sciences. 1982. Drinking Water and Health, vol. 4. Washington, DC:
National Academy Press, p. 230.
NIOSH. 1981. National Institute for Occupational Safety and Health Information profiles
on potential occupational hazards: Nitrophenols. 2nd draft Center for Chemical
Hazard Assessment, Syracuse Research Corporation. TR81-536. Cincinnati, OH:
NIOSH.
Plasterer, M.R., W.S. Bradshaw, G.M. Booth and M.V. Carter. 1985. Developmental toxicity
of nine selected compounds following prenatal exposure in the mouse: Naphthalene, p-
nitrophenol, sodium selenite, dimethylphthalate, ethylenethiourea and four glycol ether
derivatives. J. Tox. Environ. Health 15:25-38.
Rashid, K.A. and R.O. Mumraa. 1986. Screening pesticides for their ability to damage
bacterial DNA. J. Environ. Sci. Health 4:319-334.
Reinke, L.A. and MJ. Moyer. 1985. ^-nitrophenol hydroxylation: A microsomal oxidation
which is highly inducible by ethanol. Drug Metabol. Disp. 13:548-552.
Robinson, D., J.N. Smith and R.T. Williams. 1951. Studies in detoxication 39. Nitro
compounds, (a) The metabolism of o-, m-, and^-nitrophenols in the rabbit, (b) The
glucuronides of the mononitrophenols and observations of the anomalous optical
rotations of triacetyl beta-o-nitrophcnylglucuronide and its methyl ester. Biochera. J.
50:221-227.
Shen, T.C.R. 1962. The stimulating effect of dinitro-ort/io-cresol, dinitro-phenol zndpara-
nitrophenol on the aortic chemoreceptors in dogs. Arch. Int. Pharmacodyn. 140(3-
4):521-527.
20
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P-Nitrophenol April, 1992.
Shen, T.C.R. and W.H. Hauss. 1939. Influence of dinitro-phenol 1,2,4-dinitro-ort/w-cresol
and/?ara-nitrophenol upon the carotid sinus chemoreceptors of the dog. Arch. Int.
Pharmacodyn. 63:251-258.
Siragusa, G.R. and R.D. DeLaune. 1986. Mineralization and sorption of p-nitrophenol in
estuarine sediment. Environ. Tox. Chem. 5:175-178.
Spain, J.C., P.H. Pritchard and A.W. Bourquin. 1980. Effects of adaptation on
biodegradation rates in sediment/water cores from estuarine and freshwater
environments. Appl. Environ. Microbiol. 40:726-734.
Spain, J.C. and P.A. Van Veld. 1983. Adaptation of natural microbial communities to
degradation of xenobiotic compounds: Effects of concentration, exposure time,
inoculum and chemical structure. Appl. Environ. Microbiol. 45:428-435.
Spain, J.C., P.A. Van Veld, CA. Monti, P.H. Pritchard and C.R. Gripe. 1984: Comparison of
p-nitrophenol biodegradation in field and laboratory test systems. Appl. Environ.
Microbiol. 48(5):944-950.
Sultatos, L.G. and L.D. Minor. 1985. Biotransformation of paraxon and/7-nitrophenol by
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22
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EPA 0553
RX000027511
. April 1992
PHENOL
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology 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 Health
Advisories 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 lifelong exposure and the ingestion of water. The
cancer unit risk is usually derived from a linearized multistage model with 95% upper
confidence limits providing a low-dose estimate of cancer risk. The cancer 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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Phenol April 1992
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 108-95-2
Structural Formula
Phenol
Synonyms
Carbolic acid; hydroxybenzene; oxybenzehe; phenic acid; phenyl hydroxide;
phenylic acid (Windholz et al., 1983).
Uses
Phenol is used as a disinfectant, antiseptic, bactericide, and antimicrobial agent.
It is also used in the manufacture of resins and medical and industrial organic
compounds and dyes. In addition, phenol serves as a solvent for petroleum
refining and as a reagent in chemical analysis (Windholz et al., 1983).
Properties (Leonardos et al., 1969; U.S. EPA, 1980; Windholz et al., 1983)
Chemical Formula
Molecular Weight 94.11
Physical State (25ฐC) Colorless, acicular crystals or
white crystals
Boiling Point 182ฐC
Melting Point 43 ฐC (40.85 ฐC for ultrapure
material)
Density 1.071 g/mL
Vapor Pressure (25 ฐC) 03513 mm Hg
Specific Gravity (water = 1 ) 1 .0722 at 20/24 ฐC
Water Solubility (16 ฐC) 66.7 g/L
Log Octanol Water Partition
Coefficient (log K,,,,)
Odor Threshold 0.05 ppm (air); 1.0 mg/L (water)
Taste Threshold 0.3 mg/L
Conversion Factor (25ฐC, 1 mg/raj = 0.263 ppm
760 mm Hg) 1 ppm = 3.84 mg/mj
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Phenol April 1992
Occurrence
The Organic Monitoring Survey Program (U.S. EPA, 1980) qualitatively detected
phenol (i.e., carbolic acid) in 2 of 110 raw water supplies; the compound was
identified by gas-liquid chromatography and mass spectrometry, but no
quantification was made. The National Commission on Water Quality found that
the annual mean concentration of phenol from the lower Mississippi River was
1.5 pg/L (U.S. EPA, 1980).
In studies conducted from 1972 through 1977, water samples from Lake Huron,
the Detroit River and the St. Clair River were analyzed for phenol (Konasewich
et al., 1978). Phenol levels varied widely, from nondetectable to 24 mg/L;
concentrations of phenol in the Detroit River ranged from <0.5 to 5 /xg/L.
Traces of phenol (undetectable to 100 jig/L) have been reported in other rivers in
the United States (Elder et al., 1981; Jungclaus et al., 1978; Sheldon and Hites,
1979).
Phenol has been found in trace quantities (30 ppt) in urban/ suburban air in
Columbus, OH; however, its concentration has been found to be considerably
higher (520 to 44,000 ppt) in industrial areas of the United States (Brodzinsky
and Singh, 1982). Some of the phenol in the air might be from automobile use;
Kuwata et al. (1980) has reported phenol at an average concentration of 0.29
ppm in auto exhaust. Phenol was also found in particulate matter in the air, at a
mean concentration of 52 ppt, during a smog episode in West Covina, CA
(Cronn et al., 1977).
Environmental Fate
On an aqueous environment in which microbes are present (e.g., activated sludge
or pond or river water), phenol degrades completely within one to several days
(Kincannon et al., 1983; Petrasek et al., 1983; Richards and Shieh, 1986; Tabak et
al., 1981); in natural estuarine water, phenol was estimated to persist for as long
as 72 days (Lee and Ryan, 1979). Phenol has been found to degrade completely
within 2 to 5 days when mixed with wet, nonsterile soil in air (Rizet et al., 1977;
Walker, 1954); however, anaerobic degradation of phenol was only 20% of
aerobic degradation after 40 days (Baker and Mayfield, 1980). Rees and King
(1981) found that high concentrations of phenol (>3 g/L) will inhibit microbial
biodegradation by destroying the organisms. In contrast to these laboratory
results, detectable concentrations of phenol have been found in water taken from
aquifers 15 to 18 months after accidental spills had occurred (Baker et al., 1978;
Delfino and Dube, 1976).
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Phenol April 19921
Phenol, at ppb to ppm concentrations in water, may be destroyed (photolyzed) by
sunlight or converted to odorous chlorophenols if the water is chlorinated
(Wajon et al., 1982). Phenol does not evaporate from water (Aaberg et al., 1983)
and is only moderately volatile as a pure material or on dry soil (Ehrlich, 1982).
A review of the available literature revealed no information concerning the
bioaccumulation of phenol by aquatic microorganisms or by aquatic invertebrates
or vertebrates (Aaberg et al., 1983; Branson, 1978; Kobayashi et al., 1979).
Although the fate of phenol released into the atmosphere at low levels has not
been studied comprehensively, it has been calculated that its half-life, by reaction
with sunlight- generated hydroxyl radicals, is 0.7 to 14 hours; this value is
dependent on the amount of other pollutants present (Finlayson-Pitts and Pitts,
1986; Hendry and Kenley, 1979). The nitrate radical, formed in the air from
pollutants and sunlight, also reacts rapidly with phenol (Carter et al., 1981) with a
calculated half-life of 0.28 to 12.5 minutes, depending on the nitrate radical
concentrations (Finlayson-Pitts and Pitts, 1986). Other atmospheric processes,
including direct photolysis and adsorption on particulates, appear to be insignifi-
cant to the environmental fate of phenol (Callahan et al., 1979).
III. PHARMACOKINETICS
Absorption
Absorption of a small single diluted oral dose of phenol by humans and rats is
high; approximately 85 to 100% of phenol administered by this route is absorbed
within 24 hours. Retention of inhaled phenol ranges between 60 and 88% for
humans exposed to vapors for 7 hours. Dermal absorption of phenol vapors is
proportional to the concentration of the vapor used, with about one-third of the
phenol concentration in the air being absorbed through the skin of adult men
and women. Regardless of route, absorption generally is rapid, as evidenced by
the development of symptoms of toxicity within minutes after administration of,
or exposure to, phenol.
Single oral doses of 14C-phenol (0.01 mg/kg, 2.7 pCif person) were readily
absorbed by the gastrointestinal tract of three healthy men (Capel et al., 1972).
Approximately 85 to 98% of the radioactivity administered was excreted in the
urine within 24 hours; fecal 14C was not measured.
Absorption of phenol by three female Wistar rats was high (91 to 100%)
following administration of a single oral dose of 25 mg l4C-phenol/kg
(5.0 /iCi/animal) (Capel et al., 1972). Radioactivity was recovered in urine only.
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Phenol April 1992
In a series of 12 experiments conducted by Piotrowski (1971), a group of seven
healthy human volunteers (six men, one woman) inhaled approximately 5, 10 or
25 mg phenol/m3 for 8-hour intervals, with two 30-minute breaks at 2.5 and 5.5
hours after exposure began. (Actual measured exposures were between 4.8 and
26.1 mg/m3.). To avoid absorption of phenol vapor through the skin, subjects
remained outside the exposure chamber and inhaled the air from inside through
a face mask connected to the interior of the chamber. Analysis of urine collected
during exposure and 24 hours after termination of each experiment showed that
essentially all (99%) of the phenol inhaled was excreted. Thus, absorption of
phenol via the inhalation route is rapid and nearly complete.
On the basis of dermal absorption/vapor exposure data, Piotrowski (1971)
calculated an absorption coefficient of 0.35 m3/hr, indicating that for each hour
human subjects were exposed to phenol vapors (5, 10 or 25 mg/mj for 7 hours),
they absorbed, through the skin, the amount of phenol contained in 0.35 m3 of
air. Clothing provided minimal protection against dermal absorption of phenol
vapors.
Distribution
Deichmann (1944) recovered phenol and/or its metabolites in all major organs
and tissues from groups of five albino rabbits (sex and strain not given).killed 1
to 3 minutes after receiving oral doses of 500 mg phenol/kg. The liver contained
the highest levels of phenol (up to 304 /zg/g tissue), with most (up to 293 jig/g)
recovered as "free" phenol and the remainder (up to 31 jig/g) as conjugates of
phenol. Intermediate levels of up to 171, 126, 103 and 71 /*g/g were found in the
lungs, blood, brain and spinal cord, and kidneys, respectively. Muscle tissue
contained totals of 8 to 38 /tg/g, and urine taken from the bladder contained 17
to 29 /ig/g concentrations of phenol and related compounds.
Metabolism
Four .metabolites were recovered from the urine of three healthy adult men
following administration of single oral doses of 0.01 mg 14C-phenol/kg
(2.7 /iCi/person) (Capel et al., 1972). Radiochromato-graphic scans of urine
samples showed peaks that corresponded to phenyl sulfate, phenyl glucuronide,
quinol monosulfate and quinol monbglucuronide. The data indicate that humans
can conjugate phenol with sulfate and glucuronic acid and oxidize it to quinol.
Approximately 65 and 35% of the radioactivity recovered from the urine of two
female rhesus monkeys given single oral doses of l*C-phenol (50 mg/kg,
10 Ci/animal) were associated with phenyl sulfate and phenyl glucuronide,
respectively (Capel et al., 1972). Female squirrel and capuchin monkeys
metabolized a 25-mg/kg dose of MC-phenol (8 /xCi/animal). The data indicate
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Phenol April 1992
that rhesus monkeys and humans metabolize phenol in a similar manner, with
phenyl sulfate as the primary urinary metabolite; in contrast, squirrel and
capuchin monkeys eliminate mostly phenyl glucuronide, with phenyl sulfate and
quinol glucuronide as secondary metabolites.
Capel et al. (1972) reported that 30 female ICI mice given single oral doses of
l4C-phenol (25 mg/kg, 2.5 /xCi/ animal) excreted the same four metabolites
(phenyl sulfate, phenyl glucuronide, quinol monosulfate and quinol
monoglucuronide) that humans dosed orally excreted.
Excretion
The primary route of elimination for phenol is the urine. Capel et al. (1972)
reported that 24 hours after ingesting single oral doses of 0.01 mg 14C-phenol/kg
(2.7 /xCi), three healthy adult men excreted an average of 90% (range, 85 to
98%) of the administered radioactivity in urine. Feces were not collected.
Elimination of a single oral dose of 14C-phenol (25 to 50 mg/kg) was slower from
female rhesus and squirrel monkeys than from humans; these monkeys excreted,
respectively, approximately 43 and 31% of the administered radioactivity in the
urine within 24 hours post-dosing (Capel et al., 1972). In contrast, the one
female capuchin monkey used in this study excreted 73% of the 14C dose (8 /xCi)
in the urine within 24 hours. Fecal excretion of radioactivity from any animal
was not measured.
Of the seven rodent species examined by Capel et al. (1972), female rats
eliminated the greatest amount (95%) of a 25-mg/kg oral dose of 14C-phenol in
the urine during the first 24-hour post-dosing collection period. Two female
golden hamsters eliminated 73 and 78% of the administered radioactivity
(5 /iCi/animal) in the urine within 24 hours post-dosing. The elimination of 14C
was comparable (64 to 68%) for 30 ICI female mice and 4 English guinea pigs
given single oral doses of 25 mg/kg 14C-phenol (2.5 to 2.7 pCi/animal). The
24-hour urinary excretion of phenol metabolites was lowest for female Egyptian
jerboas (47%) and female Steppe lemmings (40%). The authors did not collect
feces.
In a series of 12 experiments described earlier (Piotrowski, 1971), approximately
60 to 88% of the phenol to which individuals were exposed was retained by the
body, and 99% of this was excreted in urine within 24 hours after exposure was
terminated.
Exhaled air of albino rabbits (five/group) (sex not specified) contained between
100 and 700 fig phenol at 15, 90. 120, 150 or 360 minutes after administration of
a single oral dose of 500 mg phcnol/kg to each animal (Deichmann, 1944). The
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Phenol April 1992
highest concentrations were measured at 2 hours post-dosing, and the total .
amount of phenol recovered in the expired air accounted for less than 0.1% of
that administered.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
The rapid systemic distribution of phenol following acute exposure
produces cardiac arrhythmia, wide fluctuations in blood pressure,
respiratory distress, reduced body temperature and death due most often
to respiratory failure (Bruce et al., 1987). Ingestion of even small
amounts of phenol causes severe burns of the mouth and esophagus, as
well as abdominal pain. The human oral LDLO for phenol was estimated
to be 140 mg/kg (Bruce et al., 1987):
An epidemiologjcal study (Baker et al., 1978) described the outbreak of
human illness resulting from the contamination of underground drinking
water from an accidental spill of phenol near East Troy, WI, during July
1974. In this retrospective study, 17 individuals with an average age of
21.7 years were estimated to have ingested 10 to 240 mg
phenol/person/day over approximately 1 month, primarily during July and
August 1974. The authors' exposure estimate cannot be supported by the
data. The exposed individuals experienced diarrhea, dark urine, mouth
sores and burning of the mouth during that period. Testing of well water
1 week after the spill revealed "phenol" concentrations between 0.21 and
126 mg/L; however, it is not clear whether the authors actually measured
pure phenol or total phenolic compounds. Over the next 6 months the
highest phenol concentration was 1,130 mg/L. Physical examination of
exposed individuals 6 months after the accident revealed no. significant
adverse health effects when compared to unexposed controls in terms of
incidence of skin rash, mouth lesions, conjunctivitis, and abnormal
sensation. Routine blood chemistry, urinalysis (including phenol levels)
and liver function tests were normal. A true Lowest-Observed-Adverse-
Effect Level (LOAEL) for phenol following oral exposure in humans
cannot be determined from this study because of uncertain exposure
levels.
Truppman and Ellenby (1979) reported that 10 of 43 (23%) patients
undergoing phenol face peels developed major cardiac arrhythmias when
50% or more of the face was painted in less than 30 minutes; the surface
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Phenol April 19921
area of treated skin was also positively correlated with cardiac
abnormalities in rapidly treated patients. In contrast, no arrhythmias
occurred when the same areas of facial skin were painted in a 60-minute
period. Cardiac abnormalities were not correlated with age of patient,
dosage of intravenous analgesics used during treatment, nasal oxygen
administration, phenol formulation (i.e., a 50% [v/v] saponated aqueous
mixture or a 46% [w/v] nonsaponated glycerine-water mixture), or prior
cardiac status (only 1 of the 10 affected individuals had a previous cardiac
history). The data indicate that the duration and extent of dermal
exposure to phenol are closely related to the development of systemic
toxic effects, particularly cardiac arrhythmias.
In a similar study by Gross (1984), cardiac arrhythmias occurred in 39%
(21 of 54) of patients who underwent chemical exfoliation of the face and
neck simultaneously and in 22% (22 of 100) of subjects who had their
faces and necks treated 24 hours apart. The chemical preparation used in
this study contained approximately 50% phenol. Serum phenol levels were
monitored throughout these procedures in a total of 21 individuals; single
measurements were taken in an additional 15 subjects. For all but four
patients, serum phenol levels increased as the surface area of treated skin
increased. However, no relationship was observed between serum phenol
and irregular heart rhythm. The data suggest that age, sex and cardiac
history were not important in predicting susceptibility to cardiac
arrhythmias in patients whose skin is treated with phenol.
Animals
Short-term Exposure
Flickinger {1976) estimated an oral LDM for phenol in rats (sex and strain
not given) of 650 mg/kg/day. The oral LDy, for phenol in Wistar rats was
530 or 540 rag/kg when 2 to 10% aqueous phenol was used and 340 mg/kg
when 20% aqueous phenol was used (Deichmann and Witherup, 1944).
Mortality tests were carried out in groups of 20 or 30 Wistar rats of
differing ages by administering a single oral dose of phenol at 600 mg/kg
to each animal (Deichmann and Witherup, 1944). After exposure, death
occurred in 90% of the 10-day-old rats within 12 to 24 hours; in 30% of
the 5-weekold rats within 30 to 90 minutes; and in 60% of the adult rats
within 30 to 65 minutes. Thus, young (weanling) and older rats appear to
be more susceptible to a single oral dose of phenol than the 5-week-old
rats (Deichmann and Witherup, 1944).
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Phenol April 1992
Dermal/Ocular Effects
A dermal LD50 of 850 rag/kg (95% confidence limits, 600 to 1,200 mg/kg)
for rabbits (strain and sex not specified) exposed to phenol for a
maximum of 12 hours was reported by Flickinger (1976). The
percutaneous LD50 for phenol in adult female Alderly Park rats was 625
mL/kg (approximately 650 mg/kg) (Conning and Hayes, 1970).
Upon application of 100 mg phenol into the eyes of male albino rabbits,
the conjunctivae became inflamed, the cornea became opaque and the
animals experienced marked discomfort (Flickinger, 1976). At 24 hours
after exposure, there was severe conjunctivitis, iritis, corneal opacification
occluding most of the iris and corneal ulceration over the entire corneal
surface. By day 14, the exposed eyes exhibited keratoconus and pannus
formation.
Ocular and nasal irritation was observed in rats subjected to inhalation of
aerosolized aqueous phenol at 900 mg/m3 for 8 hours (236 ppm phenol/8
hours) (Flickinger, 1976).
Long-term Exposure
No effects on liver, kidneys or any other organs were observed in 5- to
6-week-old B6C3Fi mice (10/sex/dose) administered up to 10,000 ppm
phenol (2,000 mg/kg) in drinking water for 13 weeks (NCI, 1980). (Other
groups of mice [10/sex/dose] received 100, 300, 1,000 or 3,000 ppm phenol
[20, 60, 200 or 600 mg/kg/day, respectively] in drinking water.) Dose
conversion was done according to Tatken and Lewis (1983). Control
animals (10/sex) received only tapwater. Survival was 100% for all doses
and feed consumption among all treatment groups was comparable to
controls. However, water consumption dropped to 60 and 20% that of
controls for mice administered 3,000 and 10,000 ppm phenol, respectively.
Gains in mean body weight in all treated groups (except the high-dose
group) were comparable to or greater than those in controls. In mice
receiving 10,000 ppm phenol, weight gain was 80% less than control
values for males and 33% less for females. The greatest differences in
body weight occurred during the first half of the study. Microscopic
examination revealed no compound-related pathology. This study
identified a No-Observed-Adverse-Effect Level (NOAEL) of 360
mg/kg/day (adjusted for water consumption) for mice receiving phenol in
drinking water.
In a 90-day oral exposure study (NCI, 1980), groups of 10 male and 10
female Fischer 344 rats received 0. 100, 300, 1,000, 3,000, or 10,000 ppm
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Phenol April 1992
phenol (0, 13, 38, 125. 375 or 1,350 mg/kg, respectively) in drinking water.
Dose conversion was done according to Tatken and Lewis (1983).
Survival was 100% for all groups. Mean body weight gains for rats
receiving 10,000 ppm were depressed 16% for males and 26% for females
in comparison to controls. No other differences in body weights were
observed. Rats rejected the high-dose water; males and females
consumed 67 and 50% less water, respectively, than same-sex control rats.
No compound-related gross or microscopic alterations were detected at
necropsy. A NOAEL of 375 mg/kg/day was established in this study.
Body weight gain and water consumption were reduced in B6C3Ft mice
(50/sex/dose) administered either 2,500 or 5,000 ppm phenol (500 and
1,000 mg/kg/day, respectively) in drinking water for 103 weeks (NCI,
1980). High- and low-dose groups' water consumption was depressed by
75 and 50 to 60%, respectively. Therefore, the actual delivered dose for
the high- and low-dose groups translates into 750 and 275 mg/kg/day
(Tatken and Lewis, 1983). Animals were 5 to 6 weeks old at the start of
the study. No other clinical or histopatho-logical signs related to the
consumption of phenol were observed. Mortality also was not affected in
either treatment group; approximately 84 to 96% of the male and 80 to
84% of the female mice (including controls) survived until the end of the
2-year study. This study identified a LOAEL of 275 mg/kg/day, based on
depressed body weight gain.
Ingestion of 2,500 or 5,000 ppm phenol (313 and 625 mg/kg/day,
respectively, based on dose conversions done according to Tatken and
Lewis, 1983) for approximately 2 years was not toxic to Fischer 344 rats
(50/sex/dose) (NCI, 1980). Mean body weights of males and females from
the high-dose group were somewhat lower than those of control rats;
however, these were not reported as being statistically significant. Food
consumption among treated and control animals was comparable, but
water consumption by the low- and high-dose groups was reduced by 10
and 20%, respectively. Mortality was not affected by consumption of
either level of phenol; after 104 to 105 weeks of treatment, 44 to 60% of
the males (including controls) and 74 to 78% of the females (including
controls) were still alive. The inflammatory, degenerative and
hyperplastic lesions seen in treated animals were similar in number and
kind to those that occur naturally in aged F344 rats. The highest dose
tested, 625 mg/kg/day, was identified as the NOAEL and was based on
the absence of any significant adverse health effects.
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Reproductive Effects
In a series of two-, three-, and five-generation reproduction studies by
Heller and Pursell (1938), no deleterious effects on growth, reproduction,
and rearing of young were observed in rats (number and strain not
reported) consuming 15 to 5,000 ppm phenol (1.9 to 625 mg/kg/day based
on conversion factors of Tatken and Lewis, 1983) in drinking water. In
contrast, growth was stunted in dams and pups at concentations of 7,000,
8,000, 10,000 and 12,000 ppm phenol; reproductive capabilities were
retarded and absent in animals ingesting the two highest doses,
respectively. Mortality was increased in both parents and offspring at
concentrations of 8,000 ppm and higher, despite a drop in water
consumption at these dose levels. The authors also noted that mothers
receiving 10,000 ppm phenol neglected their young. Although not
specifically designed to examine the reproductive or developmental
toxicity of phenol, this study indicated a lack of structural abnormalites
following oral exposure to the test material. The study also identified a
NOAEL of 625 mg/kg/day.
Developmental Effects
Price et al. (1986) tested the effects of oral administration of phenol in
CD-I mice. Pregnant mice were given 70, 140 or 280 mg/kg on gestation
days 6 through 15; no other details concerning administration of the test
material were given. Treatments produced dose-related incidences of
maternal tremors, ataxia, lethargy, and irritability; the incidence of these
symptoms was reported as being statistically significant at the highest
dose, but no p level was given. In addition to exhibiting clinical signs of
toxicity, 11% of the mice (4/35) dosed with 280 mg/kg died; fetuses from
this group had reduced body weights and an increased incidence (p level
not specified) of cleft palate. The highest NOAEL in this study, based on
the absence of systemic and fetal toxicity, was 140 mg/kg/day.
Jones-Price and coworkers (1983) reported that phenol administered by
gavage to groups of 20 to 22 pregnant CD-I rats on gestation days
6 through 15 produced no maternal effects at 30, 60 or 120 mg/kg/day.
An increase in resorptions was noted for ail phenol-treated groups when
compared with controls, but the effects were statistically significant only at
the 30- and 60-mg/kg/day levels; thus, the rate of resorptions was not
dose-related. Fetal body weights decreased with increasing dose levels,
and the values at 120 mg/kg were significantly lower (p < 0.001) than
controls. No structural abnormalities were noted at any dose level tested.
The highest NOAEL in this study was 60 mg/kg/day; the LOAEL, based
on highly significantly reduced fetal body weights, was 120 mg/kg/day.
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Phenol April 1992
Mutagenicity
The weight of evidence suggests that phenol is not mutagenic in
Salmonella typhimurium (DeMarini et al., 1987; Gocke et al., 1981;
Haworth et al., 1983); or Drosophila melanogaster (Gocke et al., 1981).
However, rat liver S9-activated phenol was mutagenic in L5178Y mouse
lymphoma cells (Tennant et al., 1987).
Phenol was clastogenic in Asperigillus nidulans cultured Chinese hamster
ovary (CHO) cells (Tennant et al., 1987) and in vivo in male mice (Lowe
et al., 1987) but was negative in mouse bone marrow micronuclei assays
(Gocke et al., 1981; Lowe et al., 1987).
Houk and DeMarini (1988) reported that phenol induced prophage
induction in Escherichia coli WP2r Phenol did not cause DNA strand
breaks in mouse lymphoma L5178YS (Pellack-Walker and Blumer, 1986);
however, Garberg and Bolcsfoldi (1985) listed phenol as positive at this
endpoint in mouse lymphoma cells. In vitro, phenol increased the
frequency of sister chromatid exchange (SCE) in CHO cells (Tennant et
al., 1987) and human lymphocytes (Morirnoto and Wolff, 1980); however,
in vivo mouse studies with phenol were negative for SCE induction (Lowe
et al., 1987).
Carcinogenicity
The most recent chronic oral toxicity study on phenol was performed by
the National Cancer Institute (NCI, 1980). A total of 200 B6C3Fi mice
(50/sex/dose) were administered 2,500 or 5,000 ppm (500 and 1,000
mg/kg/day, respectively, based on conversions of Tatken and Lewis, 1983)
phenol (98.47% pure) in drinking water for 103 weeks. Animals were 3
to 4 weeks old at the start of the study. Control mice (50/sex) received
tapwater only. Histopathological examination revealed no phenol-related
toxic or carcinogenic efffects in mice under the conditions of this
experiment. In addition, statistical analysis (by Cochran-Armitage and
Fisher exact tests) indicated that no tumor at any site could be clearly
associated with the administration of phenol in this bioassay.
In the chronic exposure study by NCI (1980), Fischer 344 rats
(50/sex/dose) were administered drinking water containing 2,500 or 5,000
ppm phenol (98.47% pure). These doses are equivalent to 335 and 625
mg/kg/day, respectively (based on conversions of Tatken and Lewis, 1983).
Matched controls (50/sex) received tapwater. Animals were treated for
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Phenol April 1992
103 weeks and observed for an additional 2 weeks. Under the conditions
of this bioassay, phenol was not carcinogenic in F344 rats of either sex.
Salaman and Glendenning (1957) reported that "S" strain albino mice
(four groups of 20 mice each) showed strong tumor-promoting activity
after initiation with 0.3 mg dimethylbenzoylanthracene (DMBA) and
subsequent, repeated skin applications of 20% phenol (w/v in acetone) for
24 to 32 weeks. This concentration of phenol produced significant skin
damage and was weakly carcinogenic when applied alone. A 5% solution
of phenol had a moderate promoting effect but was not carcinogenic
without prior initiation.
In tumor-promotion studies conducted by Van Duuren et al. (1968),
20 female ICR/Ha Swiss mice were treated first with a single dose of
benzo(a)pyrene, then with applications of 3 mg phenol in acetone, three
times/week, for 1 year. Four animals developed papillomas, and one had
developed a squamous carcinoma at the end of the treatment period. No
tumors developed in any of the control groups (i.e., animals treated only
with initiator, phenol, or acetone, and animals given no treatment).
In subsequent experiments on cocarcinogenesis, 20 female ICR/Ha Swiss
mice received dermal applications of 5 pg benzo(a)pyrene and 3 mg
phenol (in acetone) three times/week for approximately 66 weeks (460
days) (Van Duuren et al., 1971). At the end of the study, three mice had
developed papillomas, and one had developed a carcinoma. No tumors
were observed in the group receiving phenol alone. However, among the
animals exposed only to benzo(a)pyrene, eight had developed papillomas,
and one had developed a squamous carcinoma. On the basis of these
data, it was concluded that phenol was not a cocarcinogen; rather, at the
dose used in Swiss mice, phenol slightly inhibited the tumorgenic response
normally exhibited by benzo(a)pyrene. The partial inhibitory effect of
phenol on the carcinogenic activity of benzo(a)pyrene was confirmed by
Van Duuren and Goldschmidt (1976) and Van Duuren et al. (1973).
However, these latter studies did not evaluate the effects of solvents and,
in some cases, of pretreatment with either DMBA or benzo(a)pyrene.
In Vitro Cvtotoxicitv
Phenol was found to be cytotoxic to BF-2 cells derived from blue gill sun
fish (Babich and Borenfreund, 1987), rat lung epithelial cells (Li, 1986),
V79/4 Chinese hamster lung fibroblasts (Hunt et al., 1987), baby hamster
kidney fibroblasts (Tyas, 1978), and human amnion epithelial cells
(Eichhorn et al., 1987).
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Phenol April 1992
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:
(NOAEL or LOAEL) x fBW> _ ,
(UF) (_ IVday) = _ mg/L (
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/OW 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 determining the
One-day Health Advisory (HA) for phenol. The modified Drinking Water Equivalent Level
(DWEL) of 6,000 ngfL for a 10-kg child, calculated below, is recommended for use as a
conservative estimate for a one-day exposure.
Ten-day Health Advisory
No suitable information was found in the available literature for determining the
Ten-day HA for phenol. The modified DWEL of 6,000 /xg/L for a 10-kg child, calculated below,
is recommended for use as a conservative estimate for a 10-day exposure.
Longer-term Health Advisory ,
Two adequate subchronic oral exposure studies with phenol have been conducted (NCI,
1980). These 90-day drinking water studies identified NOAELs of 360 and 375 mg/kg/day for
mice and rats, respectively, based on the absence of systemic toxicity. However, since exposure
of the fetus is determined by maternal exposure, and since significant fetal effects without
14
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Phenol April 1992
maternal toxicity have been observed in rats given 120 mg phenol/kg/day (on gestation days 6
through 15) (Jones-Price et al., 1983), this dose level must be considered in calculating the
Longer-term HA. It is, therefore, recommended that the modified DWEL of 6,000 /xg/L for a
10-kg child be used as a conservative estimate for Longer-term exposure for a child.
., .... . r,,,/r:T (0.6 mg/kg/dav) (10 kg) , n ,,-fw. n^
Modihed DWEL = J (\\IA \ = " m&** (ฐ'000 Mg/L)
where:
0.6 mg/kg/day = RfD (see below).
10 kg = assumed body weight of a child.
1 L/day = assumed daily water consumption of a child.
The DWEL of 20,000 /xg/L, calculated below, should be used as a conservative estimate
for the Longer-term HA value for the 70-kg adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure that is
attributed to drinking water and is considered protective of 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 estimate (with uncertainty spanning perhaps an order of magnitude) of a daily
exposure to the human population (including sensitive subgroups) that is likely to be without
appreciable risk of deleterious health effects during 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 in drinking water alone 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 known, probable, or possible human carcinogen,
according to the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986), then
caution must be exercised in making a decision on how to deal with possible lifetime exposure .
to this substance. For human (A) or probable (B) human carcinogens, a Lifetime HA is not
recommended. For possible (C) human carcinogens, an additional 10-fold safety factor is used
in the calculation of the Lifetime HA. The risk manager must balance this assessment of
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Phenol April 19921
carcinogenic potential and the quality of the data against the likelihood of occurrence and
significance of health effects related to noncarcinogenic endpoints of toxicity. To assist the risk
manager in this process, drinking water concentrations associated with estimated excess lifetime
cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of
water/day are provided in the Evaluation of Carcinogenic Potential section.
The developmental toxicity study in rats by Jones-Price et al. (1983) has been selected to
serve as the basis for the Lifetime HA because it is an adequately conducted study in which the
LOAEL for fetal toxicity, 120 mg/kg/day, was lower than the NOAELs from other adequate
subchronic, chronic and reproductive/developmental studies (i.e., 140 to 625 mg/kg/day for rats
and mice, respectively) (Heller and Pursell, 1938; NCI, 1980; Price et al., 1986). These
subchronic and chronic studies also suggest that the rat is more sensitive to oral exposure to
phenol than mice (Jones-Price et al., 1983; NCI, 1980; Price et al., 1986). The epidemiological
study by Baker et al. (1978) was considered inadequate for estimation of a NOAEL because the
authors' exposure estimate was not supported by the data.
In the study by Jones-Price et al. (1983), groups of 20 to 22 female CD-I rats were given,
by gavage, 30, 60 or 120 mg phenol/kg/day on gestation days 6 through 15. No maternal effects
were observed at any dose, but reductions in fetal body weights were dose-related and
significantly lower (p < 0.001) than control values at the 120-mg/kg/day dose level. No
structural abnormalities were noted at any dose level. Results of this study suggest a NOAEL
and LOAEL for developmental effects of 60 and 120 mg/kg/day, respectively.
Using the Jones-Price et al. (1983) study, the Lifetime HA is derived as follows:
Step 1: Determination of the RfD
RfD = Jnno^ = ^ ras/'ic8/day
where: .
60 mg/kg/day = NOAEL, based on the absence of
developmental effects in fetuses of rats
exposed to phenol by gavage on gestation
days 6 through 15 (Jones-Price et al., 1983).
100 = uncertainty factor, chosen in accordance
with EPA's proposed developmental toxicity
guidelines.
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Phenol . - April 1992
Step 2: Determination of the DWEL
DWEL= (0'6 "gffiPV70 kg) = 21 mg^L (rounded to 20,000
where:
0.6 mg/kg/day = verified 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 HA
Lifetime HA = (21 mg/L) (20%) = 4.2 mg/L (4,000 Mg/L)
where:
21 mg/kg/day = DWEL.
20% = assumed RSC from water.
Evaluation of Carcinogenic Potential
IARC has not evaluated the carcinogenic potential of phenol.
Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), phenol may be classified in Group D: not
classifiable. This category is for agents with inadequate animal evidence of
carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE. AND STANDARDS
The American Conference of Governmental Industrial Hygienists recommends a
dermal Time-Weighted Average Threshold Limit Value (TWA-TLV) of 5 ppm
phenol, which is equivalent to an inhalation exposure of 19 mg phenol/m3
(AGGIH, 1988).
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Phenol April 1992'
The National Institute of Occupational Safety and Health/Occupational Safety
and Health Administration (NIOSH/OSHA) standard for phenol exposure is 20
mg/m3 (approximately 5 ppm) in air, averaged over an 8-hour work shift
(NIOSH/OSHA, 1981). NIOSH also recommends a ceiling of 60 mg phenol/m3
for a 15-minute exposure (NIOSH/OSHA, 1981).
The Public Health Service, U.S. Department of Health, Education and Welfare,
established a standard of 1 /zg/L phenols in drinking water. This was based on
the taste and odor threshold as a result of chlorination of phenols (U.S. DHEW,
1962).
VII. ANALYTICAL METHODS
Depending on the detection level desired, several methods are available for the
determination of phenol.
The classical methods for quantifying phenol, EPA Methods 420.1 and 420.2,
determine total phenols. To determine specific phenols, EPA Methods 604 or
625 should be used (U.S. EPA, 1984a,b). In these procedures the typical 1L of
sample is extracted with methylene chloride, and the extract reduced to ImL.
Analysis in Method 604 is by Flame lonization-Gas Chromatography (FID/GC).
Method 625 is a Gas Chromatographic-Mass Spectrometry Method for
determining semi-volatile base/neutrals and acids, phenol being determined in the
acid fraction. The method calls for using dueterated phenol (D-5 phenol) as a
surrogate. The detection limit for phenol by Method 604 is 0.14 /ig/L and by
Method 625 it is 1.5 /xg/L. Method 604 has an optional derivatization procedure
utilizing pentafluorobenzylbromide. The resulting derivatives of many phenols
have much lower detection levels utilizing Electron Capture-Gas Chromatography
(EC/GC).
VIII. TREATMENT TECHNOLOGIES
Available data indicate that granular activated carbon (GAC) adsorption will
significantly reduce phenol levels in drinking water.
GAC adsorption isotherms and pilot plant tests were conducted to evaluate GAC
performance in removing phenol (Royer, 1980). Freundlich isotherm constants
for phenol were determined to be 133 mg/L ""po,"' for K and 0.299 for 1/n. The
pilot plant consisted of four GAC columns operated in series, each 3 inches in
diameter and packed with 40 inches of carbon. A contact time of 18.7 minutes
produced a 95% phenol reduction from an influent concentration of 938 mg/L.
The carbon usage rate was 44.5 Ib GAC'1.000 gal.
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Phenol April 1992
Zogorski et al. (1976) studied the kinetics of the adsorption phenomenon for
phenols in a batch reactor. The GAC initial concentration was kept constant at
500 mg/L. The adsorption of phenolic compounds by GAC was extremely rapid.
Approximately 60 to 80% of the adsorption occurred within the first hour of
contact.
Kim et al. (1986) reported a maximum phenol adsorption capacity of 0.13 g
phenol/g carbon at an influent concentration of 40 mg/L by two 2-stage pilot
scale, anaerobic GAC reactor units operated in series. Each stage consisted of
two reactors: a fixed-bed (once- through mode) containing 0.5 inch Rasching
rings and a fluidized GAC bed with recirculation. Each reactor was designed for
an empty bed contact time (EBCT) of 24 hours. The performance of this system
was evaluated in terms of phenol removal by adsorption, biomass production and
biogas production.
Data were not found for the removal of phenol from drinking water by aeration.
However, phenol may be slightly amenable to removal by aeration due to its
moderate Henry's Coefficient value.
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Phenol April 1992
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. 28:172-178.
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Phenol April 1992
Price, C.J., T.A. Ledoux, J.R. Reed, P.W. Fisher, L.L. Paschke, M.C. Marr and C.A. KimmeK
1986. Teratological evaluation of phenol in rats and mice. Teratology 33:92C-93C.
Rees, J.F. and J.W. King. 1981. The dynamics of anaerobic phenol biodegradation in lower
greensand. Technol. Bioechnol. 31:306-310.
Richards, D J. and W.K. Shieh. 1986. Biological fate of organic priority pollutants in the
aquatic environment. Water Res. 20:1077-1090.
Rizet, M., J. Mallevialle and J.-C. Cournante. 1977. Pilot plant investigation of the evolution of
various pollutants during artificial recharge of an aquifer by a basin. Prog. Water
Technol. 9:203-215.
Royer, M.D. 1980. Pilot-plant and laboratory prediction of the per-formance of a mobile,
full-scale granular activated carbon system for treating hazardous materials
spill-contaminated water. Proceedings of the 1980 National Conference on Control of
Hazardous Material Spills.
Salaman, M.H. and O.M. Glendenning. 1957. Tumor promotion in mouse skin by sclerosing
agents. Br. J. Cancer 11:434-444.
Sheldon, L.S. and R.A. Kites. 1979. Sources and movement of organic chemicals in the
Delaware River. Environ. Sci. Technol. 13:574-579.
Tabak, H., S.A. Quave, C.I. Mashni and E.F. Barth. 1981. Biodegradability studies with
organic priority pollutant compounds. J. Water Pollut. Contr. Fed. 53:1503.
Tatken, R.L. and RJ. Lewis, Sr., eds. 1983. Registry of toxic effects of chemical substances,
vol. 3. Cincinnati, OH: NIOSH.
Tennant, R.W., B.H. Margolin, M.D. Shelby, E. Zeiger, T.K. Haseman, T. Spalding, W.
Caspary, M. Resnick, S. Stasiewicz, B. Anderson and R. Minor. 1987. Prediction of
chemical carcinogenicity in rodents from in vitro genetic toxicity assays. Science
236:933-941.
Truppman, E.S. and J.D. EUenby. 1979. Major electrocardiographic changes during chemical
face peeling. Plast. Reconstr. Surg. 63:44-48.
Tyas, MJ. 1978. A histochemical study of the effect of phenol on the mitochondria and
lysosomes of cultured cells. Histochem. J. 10:333-342.
U.S. DHEW. 1962. U.S. Dept. of Health, Education and Welfare. Public Health Service
drinking water standards. Rockville, MD: Public Health Service, p. 51.
24
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Phenol April 1992
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality criteria
document for phenol. EPA 440/5-80-066. NTIS PB81- 117772. Washington, DC: U.S.
Environmental Protection Agency.
U.S. EPA. 1984a. U.S. Environmental Protection Agency. EPA Method 604 - Phenols. 40
CFR Part 136. October 26.
U.S. EPA. 1984b. U.S. Environmental Protection Agency. EPA Method 625 - Base/neutrals
and acids. 40 CFR Part 136. October 26.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for carcinogen risk
assessment. Fed. Reg. 51(185):33991-34003. September 24.
Van Duuren, B.L. and B.M. Goldschmidt. 1976. Cocarcinogenic and tumor-promoting agent in
tobacco carcinogenesis. J. Natl. Cancer Inst. 56:1,237-1,242.
Van Duuren, B.L., B.M. Goldschmidt, C. Katz, S. Melchionne and A. Sivak. 1971.
Cocarcinogenesis studies on mouse skin and inhibition of tumor production. J. Natl.
Cancer Inst. 46:1,039-1,044.
Van Duuren, B.L., C. Katz and B.M. Goldschmidt. 1973. Cocarcinogenic agents in tobacco
carcinogenesis. J. Natl. Cancer Inst. 51:703-705.
Van Duuren, B.L., A. Sivak, L. Langseth, B.M. Goldschmidt and A. Segal. 1968. Initiators and
promoters in tobacco carcinogenesis. In: Wynder, E. and D. Hoffman, eds. Toward a
less harmful cigarette. NCI Monograph 28. Bethesda, MD: National Cancer Institute,
pp. 173-180.
Wajon, I.E., D.H. Rosenblatt and E.P. Burrows. 1982. Oxidation of phenol and hydroquinone
by chlorine dioxide. Environ. Sci. Technol. 16:396- 402.
Walker, N. 1954. Preliminary observations on the decomposition of chlorophenols in soil.
Plant Soil 5:194-204.
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encyclopedia of chemicals, drugs, and biotogicals, 10th ed. Rahway, NJ: Merck and
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Zogorski, J.S., S.D. Faust and J.S. Haas. 1976. The kinetics of adsorption of phenols by
granular activated carbon. J. of Colloid and Interface Science 55(2):329-341.
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EPA 0553
RX000027511
April 1992
SILVER
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology and treatment technology that would
be useful in dealing with the contamination of drinking water. Health Advisories (HAs)
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 the 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.
HAs 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. For substances with a carcinogenic potential, chemical concentration values are
correlated with carcinogenic risk estimates by employing a cancer potency (unit risk) value
together with assumptions for lifelong exposure and the ingestion of water. The cancer unit risk
is usually derived from a linearized multistage model with 95% upper confidence limits
providing a low-dose estimate of cancer risk. The cancer risk is characterized as being an upper
limit estimate, that is, the true risk to humans, while not identifiable, is not likely to exceed the
upper limit estimate and in fact may be lower. While alternative risk modeling approaches may
be presented, for example one-hit, Weibull, logit, or probit, the range of risks described by using
any of these models 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 carcinogenic risk, the Agency recommends using the
linearized multistage model.
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Silver April 1992
II. GENERAL INFORMATION AND PROPERTIES
CAS No. 7440-22-4
Structural Formula
Not applicable.
Synonyms
Argentum; argentum crede; C I 77820; collargol; L-3; shell silver; silber
(German); silver atom; silver colloidal (Weast, 1977; Stokinger, 1981).
Uses
Silver is used in coinage; the manufacture of tableware, jewelry and ornaments;
electroplating; the manufacture of solder, brazing alloys and high-capacity silver-
zinc and silver-cadmium batteries; the processing of food and beverages; inks and
dyes; etching of ivory and electrical contacts. Silver is also used as a drinking
water disinfectant and as a catalyst in hydrogenation and oxidation. It is
extensively used in photographic processing, in mirror production and in dental
alloys. Silver salts are used in the treatment of warts and burns.
In addition to metallic silver, compounds of silver used are silver oxide, silver
acetate, silver bromide, silver chloride, silver cyanide, silver iodate, silver iodide,
silver nitrate and silver sulfate (Stokinger, 1981). The use of dilute silver nitrate
solution (AgNO3) for prophylaxis against ophthalmia neonatorum is still a
routine requirement in some States (Harvey, 1985). Silver oxide is used in the
purification of drinking water because of its toxicity to bacteria and other
potentially pathogenic microorganisms (Budavari et al., 1989).
Properties (Weast, 1977; Hawley, 1981)
Chemical Formula Ag
Molecular Weight 107.868
Physical State Soft, ductile, malleable, lustrous white metal,
face-centered cubic structure
Boiling Point 2.212ฐC
Melting Point 960i5ฐC
Density 10.50 g/mL at 20ฐC
Vapor Pressure
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Silver April 1992
Water Solubility Silver is insoluble in water and soluble in fused
alkali hydroxides, peroxides and cyanides, hot
sulfuric and nitric acids. Most silver salts have
limited solubility in water; the solubilities of silver
nitrate and silver acetate are 122 and 1.02 g/100 mL
water, respectively.
Specific Gravity 10.5 at 20ฐC
Occurrence
Silver is a naturally occurring rare element, with concentrations averaging 0.1
/ig/g in the Earth's crust (Weast, 1977). Silver has a low solubility in water (0.1
to 10 mg/L, depending on the pH and the chloride concentration) (Hem, 1970).
The average concentration of silver in seawater is 0.15 to 0.3 ng/kg (Weast, 1977).
Drinking water contains extremely low concentrations of silver. It is reported
that the concentration of silver in 380 samples of finished drinking water from
the United States ranged from less than 1 up to 5 /ig/L (Nordberg and
Gerhardsson, 1988). The atmospheric concentration of silver over Southern
California is of the order of 2 ng/m3 (Bruland et al., 1974). It has been estimated
that the emission of silver iodide crystals during cloud seeding result in a silver
concentration in the air of about 0.1 ng/m3 (Nordberg and Gerhardsson,
1988). Soil concentrations of silver vary greatly by geological location. Granite
igneous rocks in Nevada contain up to 50 mg/kg silver, and coal fly ash may
contain up to 15 mg/kg silver (Nordberg and Gerhardsson, 1988).
Trace amounts of silver are found in natural and finished waters originating from
natural sources and from industrial waste. Data from 1,577 samples of well and
surface waters from 130 points in the United States showed detectable silver
concentrations (0.1 /ig/L or greater) in only 104 samples. The concentrations
ranged from 0.1 to 38 /ig/L: the median was 2.6 /ig/L (Kopp and Kroner, 1967).
The highest concentrations were noted in the St. Lawrence and Colorado Rivers
(Durum and Hafty, 1961).
Chemical analysis of finished water from public supplies of the 100 largest U.S.
cities revealed trace quantities of silver as high as 7 /ig/L, with a median
concentration of 2.3 /ig/L (Durfor and Becker, 1964). In another survey of
finished water, silver was found in 6.1% of 380 samples, with concentrations
ranging from 0.3 to 5 /ig/L (mean was 2.2 /ig/L) (Kopp and Kroner, 1967).
Environmental Fate
' Adsorption appears to be the dominant process leading to partitioning of silver
into the sediments. Silver concentrations in lake sediments and nearby soils were
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Silver April 1992
found to correlate with organic content. Silver concentrations in sediments were
100 times the concentration of overlying waters (Freeman, 1977). The affinity of
silver to three clays decreased in the following order: montmorillonite > illite >
kaolinite. However, silver has a stronger affinity to MnO, and Fe(OH)3 (Kharkar
et al., 1968). Sediments from Colorado streams showed a high correlation (r =
0.913) between silver and manganese content. Iron oxides also adsorb silver but
play a secondary role to MnO2 (Chao and Anderson, 1974; Dyck, 1968). Almost
all of the silver adsorbed to MhO, and other sediments was released on contact
with seawater, indicating that silver transported in the paniculate phase of the
water column may be released in the marine or estuarine environment.
Several studies have shown that silver bioaccumulates in aquatic organisms at
relatively low concentrations because most of its compounds are sparingly soluble
in water. Bioconcentration ratios ranging from 10 to 100 were reported for
largemouth bass and bluegill (Coleman and Cearley, 1974; Cearley, 1971).
Activated sludge microbes bioaccumulate silver at about 100 times the
concentration of silver present in solution (Chin, 1973). Freeman (1977)
reported fluctuations in silver concentrations in plankton that were closely
correlated to changes in lake water concentration, while benthic species showed
fluctuations more closely correlated to concentrations in the sediments.
Soluble silver compounds that ionize readily are quite toxic to fish. The
ranges from 0.2 mg Ag/L for young eels to 0.003 mg Ag/L for salmon fry
(Terhaar et al., 1972). Silver complex, as in the thiosulfate that occurs in
photographic processing effluents, however, is at most, only slightly toxic to fish,
the 96-hr LCj,, is greater than 250 mg Ag (in the thiosulfate complex) (Bard, et
al., 1976).
III. PHARMACOKINETICS
Absorption
Dequidt et al. (1974) administered colloidal silver to Wistar rats orally at 1.68
g/kg for 4 days or 0.42 g/kg for 12 days. The rats absorbed about 2 and 5% of
the administered dose, respectively.
Very little absorption was observed in rats (strain and sex not given)
administered carrier-free radioactive silver (< 1 /xg; 1 /iCi) intragastrically by
stomach tube (Scott and Hamilton, 1950). Within 4 days after dosing, about 99
and 0.18% of the original dose was eliminated in the feces and urine,
respectively. Total tissue distribution amounted to 0.835% of the administered
dose.
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Silver April 1992
Furchner et al. (1966, 1968) administered uฐAg-labeled silver nitrate via the oral
and intravenous (iv) routes to female RF mice (0.25 /xCi, oral; 0.25 to 0.26 pCi,
iv), male Sprague-Dawley rats (0.5 /iCi via either route), beagle dogs (0.6 //Ci,
oral: 0.4 pCi, iv), and Macacca mulatto monkeys (0.6 jzCi via either route) (dose
of Ag in mg not stated). The total body burden and persistence of silver
(calculated as an equilibrium factor) increased in proportion to species size and
was greater with the iv route than with the oral route in all of the tested species.
Very little silver was absorbed from the gut; in all species, cumulative excretion
ranged between 90 and 99% on the second day after oral ingestion. The extent
of absorption was found to be directly proportional to the transit time through
the gut in these species.
In studies evaluating therapeutic uses of silver in humans and the effects of
occupational exposure to silver, silver was readily absorbed following ingestion or
inhalation, especially when the silver ingested was in the colloidal form (Hill and
Pillsbury, 1939; Newton and Holmes, 1966; Dequidt et al., 1974).
Distribution
Silver was found to be present in almost all human body tissues and to
accumulate over the lifetime. Tipton and Cook (1963) performed postmortem
analyses of several metals in the tissues of 150 adults who had died
instantaneously and who had spent their lives in the United States. The highest
concentrations of silver were found in the thyroid, followed by the skin, liver,
adrenal, intestine (sigmoid colon) and other tissues.
In a study involving 30 American males, minute amounts of silver have been
found spectrophotometrically in the blood, brain, liver, lung, rib bone and
intestines but not in the urine, kidney, heart, spleen, muscle, long bones and
stomach (Kehoe-et al., 1940).
A 47-year-old woman developed argyria after she had ingested an excess amount
(not specified) of an oral anti-smoking remedy containing 6 mg silver acetate per
lozenge over a period of 6 months (East et al., 1980). The tissue distribution of
silver was evaluated by neutron activation analysis. No evidence of silver was
found in or between epidermal cells. The estimated total body silver content of
the subject was 6.4+2 g after the woman had been taking the lozenges for 2-1/2
years. Silver retention was measured following ingestion of tracer silver acetate
(4.5 mg) containing 4.43 jtCi 110Ag. After approximately (he first week, 18% of
the radioactive silver tracer was retained and this retention remained constant up
to 30 weeks. The blood level of radioactivity 2 hours after administration was
low (0.00045% of the dose/mL). On days 1, 2, 4 and 7, urinary silver excretion in
a 12-hour overnight sample was 2.90, 2.87, 1.94 and 2.40 x 10"*% of the dose in
the whole sample, respectively. Silver was detected in various tissues;
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Silver April 1992
concentrations ranged between 71.7 +3.7 /zg/g (in the skin biopsy sample) and
0.34 +_ 0.04 /ig/g (in pubic hair). Uptake of silver by the skin was substantial
(about 8,000 times normal uptake) following ingestion of the oral silver
preparation.
Whole-body radiation monitoring was carried out on a 29-year-old man who had
accidentally inhaled an unknown amount of dust containing uฐAg from an
experimental nuclear reactor (Newton and Holmes, 1966). Radiation monitoring
during the first 155 days after exposure revealed widespread radioactivity in the
body, with approximately 25% of the total radiation concentrated near a location
corresponding to the liver.
Creasey and Moffat (1973) reported silver granule (30- to 90-nm diameter)
deposition in the kidneys of mature rats (strain not given) after 5 weeks of silver
nitrate ingestion in drinking water (0.15%). However, weanling rats exhibiting a
similar distribution of silver had a much slower deposition time (12 to 14 weeks
until visualization by either the naked eye or by light microscopy), and the
granules were much smaller (< 30 nm in diameter) than those found in the adult
animals.
Sprague-Dawley rats were maintained for up to 60 weeks on drinking water
containing 6, 12, or 24 mM silver nitrate (Walker, 1971). Silver deposition was
observed within 6 weeks following treatment with 12 mM (1,296 mg silver/L)
silver nitrate solution in the basement membranes of the glomerulus, colon and
liver. After 12 or more weeks, the choroid plexus, thyroid acinar and basement
membranes of the skin surface, urinary bladder and prostatic acini showed silver
deposition. Silver deposition continued to appear in animals restored to silver-
free tapwater for 4 weeks following ingestion of 12 mM silver nitrate in drinking
water for 10 weeks. No silver deposition was observed microscopically in rats
maintained on 6 mM silver nitrate (648 mg/L of silver) in drinking water for 12
weeks. In another study, silver was found to accumulate in the glomerular
basement membrane of mice exposed to 6 mM silver nitrate in drinking water for
a period of 21 weeks (Day et al., 1976).
Olcott (1947) administered silver in drinking water (as a 1:1,000 solution of either
silver nitrate or silver chloride) to albino rats from weaning until death. Deposits
of silver granules were found in the homogeneous layer of the membrane of
Bruch, in the membrane underlying the epithelium of the ciliary body and in the
outer layer of the optic nerve and its vessels as they enter the eyeball.
One day after iv administration of radioactive silver to rats, most of the
radioactivity was detected in the liver and spleen of the treated animals
(Anghileri, 1969; Gammill et al., 1950). Silver concentrations decreased gradually
thereafter.
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Silver April 1992
Metabolism
No information on the metabolism of silver was found in the available literature.
Excretion
About 90 to 99% of the 110Ag administered orally (as silver nitrate) to male
Sprague-Dawley rats, female RF rats, male beagle dogs and Macacca mulatto
monkeys was eliminated in the feces; small amounts were eliminated in the urine
(Furchner et al., 1968). A similar elimination pattern was detected in rats
following iv administration of llฐAg (Gregus and Klaassen, 1986; Scott and
Hamilton, 1950). Most of the radioactivity found in the feces was eliminated via
the bile (Tichy et al., 1986; Gregus and Klaassen, 1986; Klaassen, 1979).
Biliary excretion of silver was investigated in male Sprague-Dawley rats after iv
administration of 110Ag as silver nitrate (10 to 50 /zCi/kg, 2 mL/kg) (Gregus and
Klaassen, 1986). Two hours after injection of 0.01, 0.03, 0.1 or 0.3 mg/kg metal
ion, the biliary concentration of silver as a percentage of the administered dose
was about 26, 45, 32 and 35%, respectively. The fraction of the dose excreted
was not markedly altered by the administered dose. Further review of the metal
concentration ratios in the bile, liver and plasma indicated that silver was highly
concentrated in the bile relative to the plasma, and that it tended to accumulate
in the liver.
A marked variation in biliary excretion was observed in different species
administered uฐAg as silver nitrate, in a single iv injection at 0.1 mg/kg of silver
over a period of 2 hours (Klaassen, 1979). Thirty minutes after treatment, male
Sprague-Dawley rats excreted silver into the bile at a rate of 0.25 /xg/min/kg, New
Zealand White male rabbits excreted 0.05 ng/min/kg and mongrel male dogs
excreted 0.005 /ig/min/kg. The concentration of silver in the plasma was
markedly lower in the dog than in the rat or rabbit (at 2 hours, the 110Ag
concentration in the plasma of the dog, rabbit and rat was 0.03, 0.1 and
0.3 /ig/mL, respectively), indicating a larger volume of distribution in the dog.
This variation does not appear to be attributable to differences in the transfer of
silver from plasma to liver, but rather from liver to bile. The species with the
lowest biliary excretion rate (the dog) had the highest liver concentration of silver
(rat 1.24, rabbit 2.13 and dog 2.9 /xg silver/g liver). In all species, the
concentration of silver in the bile was greater than that in plasma with no
observable dose gradient, thereby indicating an active transport process and a
saturable mechanism.
Whole-body radiation monitoring was carried out on a 29-year-old man who
accidentally inhaled an unknown amount of dust containing 110Ag (Newton and
Holmes, 1966) from an experimental nuclear reactor. Labeled silver was not
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Silver April 1992
detected in the urine samples taken during the first 54 days but was detected in
bulked fecal samples up to about day 300. The inhaled silver was eliminated
from the body (data for the first 2 days were not available) following a biphasic
exponential decay curve that corresponded to a two-compartment model. The
effective half-life for the first phase was about 1 day; the half-life identified for
the slower second (terminal) phase was 43 days. Earlier data on the inhalation
exposure of female beagle dogs to radioactive silver fit a three-compartment
model with initial, intermediate and terminal half-life phases of 1.7, 8.4 and 40
days (Phalen and Morrow, 1973).
Thirty silver workers (average age 45 years) exposed to 1 to 100 /zg/m3 of silver in
the air (based on a 2-month environmental monitoring), for a mean exposure
duration of 20 years, had a mean blood silver concentration of 0.011 /xg/mL
(more than 80% had levels ranging from 0.006 to 0.026 /xg/mL) (Di Vincenzo et
al., 1985). Matched controls (n = 35) not exposed to silver showed no detectable
blood silver. Silver in urine samples was generally not detected in workers or in
matched controls. Average fecal excretion of silver in the exposed workers
(n = 29) and the controls (n = 25) was about 15 and 1.5 ttg/g, respectively.
IV. HEALTH EFFECTS
Humans
No experimental studies providing data on the adverse health effects of silver
ingestion in humans were found in the available literature.
Short-term Exposure
Hill and Pillsbury (1939) reported that accidental ingestion of large doses
of silver nitrate resulted in abdominal pain, diarrhea, vomiting, corrosion
of the gastrointestinal tract, shock, convulsions and death. A fatal single
oral dose of silver nitrate was estimated to be 10 g (about 143 mg/kg for a
70-kg person).
Long-term Exposure
In the early part of this century, inorganic silver salts and colloidal silver
preparations (e.g., silver arsphenamine) were used as therapeutic agents,
especially in the treatment of syphilis. The single adverse effect resulting
from the therapeutic use of silver asphenamine administered by injection
was argyria (also called argyrosis or argyrism), which is characterized by
permanent discoloration and darkening of the skin on exposure to light.
In the early stages of argyria, the discoloration is observed in the
8
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Silver April 1992
conjunctiva or the nail bed, the mucous membranes of the mouth and
pharynx, the lips and many internal organs (Hill and Pillsbury, 1939). Of
the 19 reported cases of syphilitic patients, 4 developed argyria after
receiving, a total of 6.3 to 10.2 g silver arsphenamine (14.5% silver) via
injection. Although the smallest dose inducing argyria corresponded to
0.91 g of metallic silver, the actual dose is likely to be higher because
exposure to silver via ingestion of food and water and via consumption of
other silver-containing therapeutic agents prior to the study were not
taken into account. Furthermore, this estimate is conservative because
the observed response occurred in a small fraction of studied individuals.
These individuals may have been sensitive because they were in the late
stages of syphilis.
Gaul and Staud (1935) reviewed several studies pertaining to argyria in
humans resulting from the use of organic and colloidal silver medication.
From 1914 to 1928, 13 cases of argyria were reported, and 27 additional
cases were reported from 1928 to 1935. Approximately 30% of the cases
occurred in persons under 40 years of age, and 20% occurred in children
under 10 years of age. Argyria developed in 65% of the cases following
administration of medication via the pharyngeal and intranasal routes,
and 35% occurred following administration via the oral route. The
duration of treatment ranged from 1 month to 11 years (about 1 mg
silver/kg from oral treatment with argyrol, a mild silver-protein complex
and one of three medications used).
Clinical and therapeutic data were presented for 10 males, aged 23 to 64
years, and for one woman, aged 49 years, who were treated with iv
injections of silver arsphenamine over a 2- to 9.75-year period (Gaul and
Staud, 1935). The total dose of silver arsphenamine injected (31 to 100
injections) ranged from 4.1 to 20 g (0.13 to 0.475 g/injection). In some
patients, argyria developed after a total dose of 4, 7 or 8 g of silver
arsphenamine; in other patients, argyria did not develop until 10, 15 or 20
g had been injected. The degree of discoloration was directly
proportional to the amount of injected silver arsphenamine and to the
duration of exposure. The authors also studied 10 cases of generalized
argyria by biospectrometric analyses of biopsies and concluded that
argyria becomes clinically apparent after an equivalent of about 8 g of
silver arsphenamine is retainted in the body. The authors referred to a
previous study in which they established that the quantity of silver in a
biopsy specimen was directly proportional to the dosage of silver
arsphenamine. From the data presented, silver arsphenamine
administered in about 40 doses (0.18 to 0.21 g/injection) over 2 to 4 years
to the males in this study corresponded to an average total dose of 8 g *
silver arsphenamine.
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Silver April 1992
Blumberg and Carey (1934) reported the case of a 33-year-old woman
who had ingested capsules containing 16 mg of silver nitrate three
times/day during alternate weeks for a period of 1 year. The only effects
observed were skin pigmentation and possible asthenia and anemia.
Spectrographic analyses of blood samples revealed marked argyria; the
woman's blood silver level was 0.5 mg/L 1 week after cessation of silver
ingestion; only a small decrease was observed in blood silver levels 3
months after silver ingestion had ceased. Heavy traces of silver in the
skin, moderate amounts in the urine and feces and trace amounts in the
saliva were also reported in samples tested 3 months after ingestion of the
capsules was stopped. The total amount of silver ingested during the 1-
year period was calculated to be 6.4 g (about 35 mg/day of silver, on
alternate weeks over 1 year), which was identified as the lowest observed
cosmetic effect level for this study.
Kent and McCance (1941) reported that a woman who had washed out
her nose for many years with an organic silver preparation suffered from
generalized argyria. The weekly intake of silver via food ingestion over
the 3 consecutive weeks of experimentation was 0.05, 0 and 0.7 mg. No
urinary excretion of silver was detected during these weeks. However,
weekly fecal excretion over the 3-week period was 1.3, 1.5 and 2.3 mg.
Thus, the woman was in negative silver balance (-1.25, -1.5 and -1.6
mg/week), which was not attributed to any physiologic effect but was due,
perhaps, to desquamation of silver-containing mucous membrane cells of
the mouth, nose and gastrointestinal tract. This study did not provide any
other information from which a dose-effect relationship could be derived.
Animals
Short-term Exposure
A single oral dose of 420 mg/kg of silver colloid did not result in any
deaths in rats. Mortality was noted only after repeated daily oral
ingestion of 1,680 mg/kg for 4 days (Dequidt et al., 1974).
Diplock et al. (1967) administered 1,500 mg/L of silver acetate (970 mg/L
of silver) in drinking water to groups of nine weanb'ng Norwegian hooded
rats fed a basal vitamin E-deficient diet. All of the rats (mean age 55
days) died with liver necrosis 2 to 4 weeks after the addition of silver
acetate to the water. In another group, 1.0 ppm of selenium was added
to the diet; 4 of the 9 rats (mean age 74 days) died with liver necrosis. In
another group 120 ppm of vitamin E was added to the diet. All nine rats'
in this group had normal livers at necropsy on day 86 (50 days of silver
exposure). A No-Observed-Adverse-Effect Level (NOAEL) and a
10
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Silver April 1992
Lowest-Observed-Adverse-Effect Level (LOAEL) could not be
established based on these data.
Bunyan et al. (1968) reported that no liver necrosis occurred in
Norwegian hooded rats (about 5 weeks old) fed a vitamin E-deficient diet
and 1,000 mg/L of silver acetate (650 mg/L of silver) in drinking water
(duration not stated). When the selenium content of the diet was
reduced by omitting yeast from the diet (so that 8.3% casein was the only
protein in the diet) liver necrosis was induced by as little as 130 mg/L of
silver acetate (85 mg/L of silver) in drinking water. Vitamin E at 120
mg/L reportedly prevented liver necrosis. A NOAEL and LOAEL were
not identified.
Grasso et al. (1969) observed a rapidly fatal liver necrosis beginning on
day 14 after addition of silver acetate to the diet (130 to 1,000 ppm silver
acetate; 4 to 33 mg/kg/day, based on Lehman, 1959) or to the drinking
water (1,500 ppm silver acetate; 97.5 mg/kg/day assuming a 200-g rat
consumes 20 mL/day of water) of vitamin E-deficient weanling Norwegian
hooded rats. LOAELs for silver administered in the diet or drinking
water were not identified.
Van Vleet (1976) reported that four weanling swine fed a diet containing
adequate selenium and vitamin E and 0.5% silver acetate (3,250 ppm of
silver, or 130 mg/kg/day based on the assumptions of Lehman, 1959) for 4
weeks developed anorexia, diarrhea and growth depression; three of the
four pigs died. Hepatic lesions in all four pig* were consistent with
hepatosis dietetica. No lesions developed in pigs fed 0.2% silver acetate
(1,300 ppm of silver, or 52 mg/kg/day based on the assumptions of
Lehman, 1959). Vitamin E (100 ITJ/kg diet) but not selenium (1 ppm)
supplementation in the diet (two pigs/group) prevented development of
lesions and mortality. The LOAEL and NOAEL were 130 and 52
mg/kg/day of silver, respectively.
Wagner et al. (1975) administered silver for 52 days as silver acetate in
drinking water at 0, 76 or 751 mg/L to groups of ten 21-day-old Holtzman
strain rats fed a vitamin E-deficient and low-selenium (0.02-ppm) diet. A
similar regimen was given to vitamin E-deficient rats whose diets
contained added selenium (0.5 ppm). In 4 of the rats fed the vitamin E-
deficient and low selenium diet, 751 mg/L of silver caused severe growth
depression and death within 39 to 46 days. Although histologjcal
examinations were not conducted, no gross evidence of liver necrosis was
observed. Dietary selenium improved growth and survival among the rats
given 751 mg/L of silver and completely overcame the growth depression
in rats fed 76 mg/L of silver the concentration of silver in the liver was
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Silver April 1992
increased in rats fed selenium. The activity of liver glutathione
peroxidase in the selenium-supplemented group was reduced to 30% of
control in the at 76-mg/L group and to 4% of control in the 751-mg/L
group.
The enzyme activity was not reduced in erythrocytes and was not detected
in the livers of rats fed the low-selenium (0.02 ppm) basal diet.
Administration of 751 mg/L of silver in drinking water for 15 weeks in
rats fed a diet supplemented with vitamin E (100 lU/kg) and selenium
(0.5 ppm as sodium selenite) decreased glutathione peroxidase activity in
the liver, erythrocytes and kidneys to 5, 37 and 38% of the controls fed
the same diet, respectively. In addition, a 15% decrease in body weight
was seen in treated rats. It was postulated that the antagonism of silver
and selenium is exerted through an effect on the biosynthesis of the
selenium-containing enzyme glutathione peroxidase. The 15-week study
(with vitamin E and selenium) suggests a LOAEL for silver of less than
751 ppm (114.2 mg/kg/day of silver) in rats.
In a review of the literature, Ganther (1980) suggested that silver toxicity,
which is suppressed by the addition of low levels of vitamin E or selenium
to the diet, may be due to a silver-induced deficiency of selenium that
makes it unavailable for glutathione peroxidase synthesis. The
effectiveness of selenium in reducing toxicity is attributed to overcoming
this deficiency. Vitamin E, however, although effective in overcoming
silver- induced growth depression, did not prevent the depression of
glutathione peroxidase activity. Thus, it appeared that the growth
depression was not caused by the silver- induced decrease in glutathione
peroxidase activity.
Dermal/Ocular Effects
Rungby (1986) applied three drops of a 0.66% (42 ppm silver) silver
nitrate solution to the right eye of male Wistar rats. Forty-five days after
treatment, the rats were killed. Silver deposits were found in the cornea
and conjuctiva and scattered in the cells of the outermost part of the
anterior corneal epithelium; heavy deposits were found in Bowman's
layer, reticular fibers of the corneal stroma, Descemet's membrane and
the posterior corneal epithelium. No lexicological effects were reported.
A NOAEL and LOAEL were not established.
Long-term Exposure
A study by Walker (1971) showed that silver deposits could be found
within the glomerular basement membranes in the kidneys of Sprague-
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Silver April 1992
Dawley rats after the rats ingested 12 mM AgNOj (assuming a 200-g rat
consumes 20 mL fluid/day; 12 mM AgNO3 or 1,296 mg silver/L equals
130 mg/kg/day) for 4, 6, 8, 10, 12, 16, 25 or 60 weeks. Rats remained in
excellent clinical condition. Between the 76th and 81st weeks of ingesting
this concentration of AgNO3, however, a rapid deterioration in clinical
condition was noted. By contrast, rats given 6 mM AgNO3 (65
mg/kg/day) exhibited no difference in appearance, behavior or fluid
consumption compared with control animals. After 12 weeks, no
prominent silver deposits were noted, and the experiment was
discontinued. In rats given 24 mM AgNO3, a large decline in fluid intake
occurred in the first week. The rats were poorly groomed, listless and
continued to demonstrate reduced fluid intake during the second week, at
which time the experiment was discontinued. A LOAEL of 130
mg/kg/day was identified. A NOAEL was not established because,
although no effects were observed at 65 mg/kg/day, the exposure was
halted after 12 weeks. Therefore, no comparison between this group and
the nine rats receiving 130 rag/kg/day could be made.
As noted by Walker (1971), silver deposits were present in the basement
membrane of the kidneys of mice administered 6 rnM silver nitrate (65
mg/kg/day of silver) for 12 days to 14 weeks (Day et al., 1976). No
toxicity other than a slight reduction in water intake was noted. The
NOAEL was 65 mg/kg/day.
Olcott (1948) reported that a 1:1,000 solution of silver nitrate in drinking
water (63.5 mg/kg/day of silver, assuming a 200-g rat consumes 20 mL
water/day) given to rats for 218 days induced intense pigmentation of
many tissues, most notably the basement membrane of the glomeruli, the
walls of the vessels between the kidney tubules, the portal vein and other
parts of the liver, the choroid plexus of the brain, the choroid layer of the
eye and the thyroid g'and. No shortening of lifespan or reduction of body
weight occurred. A NOAEL of 63.5 mg/kg/day was identified.
Olcott (1947) examined the eyes in life and from tissue sections at
necropsy of 139 albino rats administered a 1:1,000 solution of silver
nitrate in drinking water (63.5 mg/kg/day of silver, assuming a 200-g rat
consumes 20 mL water/day). A treatment duration of 218 days (3.2 g
silver total) resulted in eyes that were slightly gray (stage 1); a duration of
373 days (5.7 g silver total) resulted in eyes more gray than pink (stage 2);
a duration of 447 days (6.8 g silver total) resulted in dark but translucent
eyes (stage 3), and a duration of 553 days (9.4 g silver total) resulted in
opaque eyes (stage 4). Histologjcal sections showed few, but definite,
granules in the membranes of Bruch at stage 1. At stage 4, the
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Silver April 1992
membrane was almost uniformly black. No toxicity was observed. A
NOAEL of 63.5 mg/kg/day was identified.
Olcott (1950) dosed rats (>9 months of age) with a 1:1,000 solution of
silver nitrate (63.5 mg/kg/day of silver, assuming a 200-g rat consumes
20 mL water/day). The end point of hypertrophy was defined as a ratio
of the left ventricle weight to the final body weight of the rat, expressed
as the weight of the left ventricle per 100 g of body weight ^0.2. A total
of 233 rats were autopsied. The study author reported that a large
number of spontaneous deaths occurred because of advanced pulmonary
lesions, but these lesions were not attributed to silver ingestion. After
long-term silver ingestion (duration not stated), a statistically significant
(no p value, Chi = 3.13) hypertrophy of the left ventricle was noted in
29% of rats given silver nitrate compared to 12% of rats given water.
The authors postulated that this finding was indicative of vascular
hypertension produced when silver deposition caused a thickening of the
basement membrane of the renal glomeruli. A NOAEL and LOAEL
were not identified because the length of exposure was not reported.
Reproductive Effects
No information on the reproductive effects of silver was found in the
available literature.
Developmental Effects
Robkin et al. (1973) reported that the concentration of silver in the livers
of 12 anencephalic human fetuses was higher (0.75 ฑ 0.15 mg/kg) than
those found in 9 premature infants (0.68 ฑ 0.22 mg/kg), 12 fetuses
obtained through therapeutic abortions (0.23 ฑ 0.05 mg/kg) or 14
spontaneously aborted fetuses (0.21 + 0.05 rag/kg). The authors were not
able to determine if the higher concentrations of silver in anencephalic
fetuses was associated with the malformation or if the concentrations
were related to fetal age.
Rungby et al. (1987) treated Wistar rat pups from two litters with
subcutaneous injections of silver lactate monohydrate; two pups from each
litter received daily injections of 0.10, 0.20 or 0.35 mg during weeks 1, 2,
or 3 to 4, respectively. The authors reported that hyppocampal tissues
from the treated fetuses contained significantly (p < 0.05) smaller
pyramidal cells; they speculated that the findings suggest that pyramidal
cells are the first elements in the hyppocampus to show signs of silver
toxicity, or that these cells are selective sites for silver neurotoxicity.
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Silver April 1992
Mutagenicity
Silver nitrate (5 x 10"6 to 10"s%), in the absence of exogenous metabolic
activation, did not increase the frequency of reversion to streptomycin
nondependence in Escherichia coli (Demerec et al., 1951).
Similarly, AgNO3 (0.1 /xM) was neither mutagenic in E. coli WP2 nor
comutagenic in ultraviolet-irradiated cultures of E. coli WP2 (Rossman
and Molina, 1986).
Silver chloride (0.05 M), at a single nonactivated dose, caused equivalent
inhibition of recombinational repair-proficient (H-17) and repair-deficient
(M-45) strains of Bacillus subtilis (Nishioka, 1975).
Carcinogenicity
Local sarcomas have been induced after subcutaneous implantation of
foils and discs of silver and other noble metals. However, Furst (1979,
1981) concluded that intraperitoneal and subcutaneous implants were
invalid indicators of carcinogenicity because of the phenomenon of
solid-state carcinogenesis, which results in local fibrosarcomas even with
insoluble solids such as smooth ivory and plastic but not with
crumbled tin.
Schmaehl and Steinhoff (1960) reported that colloidal silver injected
subcutaneously into rats resulted in tumors in 8 of 26 rats surviving longer
than 14 months. In six of the eight rats, the tumor was at the
subcutaneous injection site. In 700 untreated rats, the rate of
spontaneous tumor formation was 1 to 3%; no vehicle control was
reported.
Furst and Schlauder (1977) suspended silver powder in trioctanoin and
gave it once each month by intramuscular injection to Fischer 344 rats
(50/sex/group). The dose given was 5 mg each for five treatments and 10
mg each for five more treatments, for a total of 75 mg silver. An inert
material was used as the vehicle control, and cadmium was used as a
positive control. No fibrosarcomas appeared at the injection site in silver-
treated rats. Injection site sarcomas were found only in the vehicle-
control (1/50) and cadmium-treated (30/50) rats. The latent period in the
vehicle-control group was 19 months, and the latent period in the
cadmium-treated group was as short as 4 months. The authors concluded
that silver was not tumorigenic when administered as a finely divided
powder.
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Silver April 1992
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:
MA - (NOAEL or LOAEL) (BW) _
^ " (UF) (_ IVday) - _ rng^ ( Mg/L)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect Level (the exposure dose
in mg/kg bw/day).
BW = assumed body weight of a child (10 kg) or an adult (70 kg).
UF(s) = uncertainty factors based upon quality and nature of data (10,
100, 1,000 or 10,000), in accordance with EPA or NAS/OW
guidelines.
L/day = assumed daily water consumption (1 L/day/for child or 2 L/day/for
adult).
One-day Health Advisory
No data suitable for determining the One-day Health Advisory (HA) for silver was
found in the available literature. Acute toxicity studies such as Dequidt et al. (1974) provide
data on lethal doses but do not provide the dose-response data required to calculate the HA.
The cosmetic Drinking Water Equivalent Level (DWEL) for silver of 0.2 mg/L calculated below
is recommended for use as a conservative estimate of the One-day HA for a child or an adult.
Ten-day Health Advisory
No data suitable for determining the Ten-day HA for silver was found in the available
literature. Short-term studies have provided information on lethal doses, but have not provided
NOAELs or LOAELs. The cosmetic DWEL for silver of 0.2 mg/L calculated below is
recommended for use as a conservative estimate of the Ten-day HA for a child or adult.
Longer-term Health Advisory
No data suitable for determining the Longer-term HA for silver were found in the
available literature. The data found were not sufficient to establish a NOAEL and LOAEL.
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Silver April 1992
The cosmetic DWEL for silver of 0.2 mg/L calculated below is recommended as a conservative
estimate of the Longer-term HA for an adult or 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 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 estimate of a daily exposure level 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 known,
probable or possible human carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution must be exercised in making a decision
on how to deal with possible lifetime exposure to this substance. For human (A) or probable
human (B) carcinogens, a Lifetime HA is not recommended. For possible human carcinogens
(C), an additional 10-fold safety factor is used in the calculation of the Lifetime HA. The risk
manager must balance this assessment of carcinogenic potential and the quality of the data
against the likelihood of occurrence and significance of health effects related to noncarcinogenic
end points of toxicity. To assist the risk manager in this process, drinking water concentrations
associated with estimated excess cancer risks over the range of 1 in 10,000 to 1 in 1,000,000 for
the 70-kg adult drinking 2 L of water per day are provided in the Evaluation of Carcinogenic
Potential section.
Specific information on adverse health effects in humans or animals of oral exposure to
graded levels of silver needed for setting Lifetime Health Advisory levels, was not found in the
available literature. Two reports of argyria following ingestion of silver were found (Blumberg
and Carey, 1934; East et al., 1980), but these reports are of single cases with inadequate
documentation of the dose rate. On the other hand, valuable clinical and therapeutic data were
presented on human cases of argyria by Gaul and Staud (1935, as cited in U.S. EPA, 1980) and
by Hill and Pillsbury (1939, as cited in U.S. EPA, 1980). These data indicate that about 0.9 to
1.5 g of silver administered over a period of 1 to 2 years as iv injections will cause argyria in
patients. However, the individuals are likely to have been sensitive because they were in the
late stages of syphilis. Furthermore, many other patients received similar dose regimen without
developing argyria.
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Silver April 1992
Derivation of the Lifetime Health Advisory for the Cosmetic Effect of Silver
A Lifetime HA based on cosmetic effects is calculated assuming that 1 g of iv silver will
produce mild argyria in the most sensitive individuals (Gaul and Staud, 1935; Hill and Pillsbury,
1939). Assuming a 4% absorption rate (Furchner et al., 1968) following oral exposure, the 1-g iv
dose corresponds to an oral dose of 25 g (1 g/ 0.04 = 25 g). This dose is then averaged over a
lifetime, assumed to be 70 years:
70 years x 25 g = 978 /ig/day
25,550 days
Based on an adult body weight of 70 kg, this corresponds to the Lowest-Observed-
Cosmetic-Effect Level of 14 jig/kg/day (978 /zg/day/70 kg = 14 /xg/kg/day). Using 14 /ig/kg/day as
the Lowest-Observed-Cosmetic-Effect level for silver, a Lifetime HA for the cosmetic effect of
silver is calculated as follows:
Step 1: Determination of the Cosmetic RfD
14 ^g/kg/day
Cosmetic RfD = 3 =4.7 /zg/kg/day (rounded to 5 /zg/kg/day)
where:
14 /zg/kg/day = Lowest-Observed-Cosmetic-Effect Level based on argyria.
3 = Uncertainty factor: An uncertainty factor of 3 is used to estimate
an RfD associated with lifetime exposure. This uncertainty factor
was applied for the following reasons: First, a 10-fold uncertainty
factor is usually applied to human data to account for intraspecies
variability. However, since this calculation has already included
sensitive individuals, a 10-fold uncertainty factor is not warranted.
Second, an uncertainty factor less than 10 (i.e., 3) is sufficiently
protective since the estimated dose causing argyria within 1-3
years is being apportioned over a lifetime, and the effect is based
on argyria which is considered a cosmetic effect, and not an
adverse health effect.
Step 2: Determination of the Cosmetic DWEL
5 ug/kg/dav x 70 kg
Cosmetic DWEL = 2 L/day = 175 iig/L (rounded to 200 /zg/L)
18
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Silver April 1992
where:
5 /zg/kg/day = cosmetic RfD.
70 kg = assumed weight of an adult.
2 L/day = assumed water consumption of a 70-kg adult.
The DWEL is derived on the assumption that 100% of the silver intake comes from
drinking water. As estimated by the World Health Organization (WHO, 1984), the upper-
bound intake of silver from food is 20 to 80 jig/day and is essentially negligible from air.
Therefore, the Lifetime HA for the cosmetic effect of silver can be calculated by subtracting the
amount of silver obtained in food.
Step 3: Lifetime HA for the Cosmetic Effect of Silver
Lifetime HA for Cosmetic Effect =
(0.005 me/kg/dav^ (70 kg) - 0.08 mg/dav = 0.135 mg/L (rounded to 0.1 mg/L)
2 L/day
Thus, a concentration of silver in water of 100 pg/L or 0.1 mg/L is considered protective
of the cosmetic effect of silver (argyria) for the general population.
Evaluation of Carcinogenic Potential
Applying the criteria described in the U.S. EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), silver has been classified in Group D: not classified. This
category is for agents with inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
The U.S. EPA had originally regulated silver with an MCL of 50 /ig/L.
However, since silver caused only argyria, a cosmetic effect, the U.S. EPA (1991)
replaced the primary standard of 50 jig/L with a value of 100 /xg/L as the
secondary standard.
VII. ANALYTICAL METHODS
Most of the methods available for silver analysis involve atomic absorption
spectroscopy. In these methods, the metal is dissolved and thermally
excited. When excited, the metal absorbs light frequencies characteristic of that
19
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Silver April 1992
element. In addition, colorimetric (dithizone) methods (American Public Health
Association, 1976) and inductively coupled plasma atomic emission spectroscopy
can be used to analyze silver (CFR, 1987)
Direct Aspiration Atomic Absorption Spectroscopy (AA). In this technique the
dissolved metal (silver) is aspirated into a flame source and excited to the point
of dispersion into a mono-atomic state; a light source whose cathode is the metal
of interest passes through the flame and the resulting absorption of light by the
element of interest is directly proportional to concentration. Disadvantages of
this technique include the inability to analyze more than one metal at a time and
the insensitivity of the technique.
Graphite Furnace Atomic Absorption (GFAA). This technique differs from AA
in that a specific amount of liquid is dried on the thermal source, effecting a
concentration step. The sample is electrothermally excited. This technique has
great sensitivity, but it is still a tedious one-metal-at-a-time determination.
VIII. TREATMENT TECHNOLOGIES
Available data indicate that the following treatment technologies have been
effective in removing silver from drinking water supplies: direct filtration,
coagulation/ filtration, lime softening, and reverse osmosis (RO).
Sorg et al. (1978) conducted laboratory-jar tests without coagulants using Ohio
river water spiked with an initial silver concentration of 0.15 mg/L. Direct
filtration of this water with the pH adjusted with soda ash from 6.8 to 9.4
reduced the silver concentration by 49 to 56%.
Sorg et al. (1978) also ran laboratory-jar tests to evaluate the effect of pH on
silver removal using Ohio river water and found that pH had no significant effect
on silver removal by either alum or iron coagulation in the pH 6 to 8 range.
Removal of silver by alum coagulation decreased above pH 8. Poor floe
formation was postulated as the cause of this decrease. Further jar tests showed
that alum removal of silver was directly proportional to the turbidity of the
water. Iron coagulation tests did not show a turbidity effect; 55 to 65% of the
silver was removed from river water in these tests.
Hannah et al. (1977) tested the effectiveness of a pilot plant utilizing
coagulation/filtration or excess lime treatment followed by granular-activated
carbon (GAC) adsorption in removing silver. The plant consisted of a rapid mix,
flocculator, sedimentation basin and dual-media filtration. Silver was present at
an initial concentration of 0.5 mg/L. Ferric chloride was added at a dose of 40
mg/L and a pH of 6.2; alum was added at a dose of 415 mg/L and a pH of 11.5.
Excess lime treatment removed 97.1% of the silver, and GAC adsorption
20
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Silver April 1992
increased the overall removal efficiency to 98%. Coagulation/filtration with
ferric chloride removed 98.2% of the silver, and GAC adsorption increased the
overall removal efficiency to 99%. Coagulation with alum removed 96.9% of the
silver and GAC adsorption increased the overall removal efficiency to 99%.
McCarty et al. (1982) reported the removal of silver by Water Factory 21, a
wastewater treatment plant in Orange County, CA, using an excess-lime
softening treatment process, followed by RO. The Lime (350 to 400 mg/L as
calcium oxide) was added to raise the pH to 11. The clarified water was
carbonated to pH 8.0, filtered and chlorinated. Silver was present at an initial
concentration of 1.6 mg/L. The RO unit waj> composed of spiral-wound cellulose
acetate membranes. Lime softening alone removed 76% of the silver, and the
RO unit alone removed 63% of the silver, with an overall 90% silver reduction
by lime softening/RO treatment.
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Silver April 1992
IX. REFERENCES
American Public Health Association. 1976. Standard methods for the examination of water
and wastewater, 14th ed. Washington, D.C.
Anderson, J.B., E.A. Jenne and T.T. Chao. 1973. The sorption of silver by poorly crystallized
manganese oxides. Geochim. Cosmochim. Acta 37:611-622. Reviewed in U.S. EPA.
1979. U.S. Environmental Protection Agency. Water related environmental fate of 129
priority pollutants. Vol. I. EPA-440/4-79-29a. Washington, DC: U.S. EPA, p. 17-1.
Anghileri, L.J. 1969. Studies on the in vivo breakdown of insoluble halides. Acta Isot.
9:347-356.
Bard, C.C., JJ. Murphy, D.L. Stone and C.J. Terhaar. 1976. Silver in photoprocessing effluents.
Jr. Water Pollution Control Fed. 489(2) 389-394.
Blumberg, H. and T.N. Carey. 1934. Detection of unsuspected and obscure argyria by the
spectrographic demonstration of high blood silver. J. Am. Med. Assoc. 103:1,521-1,524.
Bruland, K.W., K. Bertine, M. Koide and E.D. Goldberg. 1974. History of metal pollution in
the Southern California coastal zone. Environ. Sci. #1, Technol. 8:425-432.
Budavari, S., M.J. O'Neil, A. Smith and P.E. Heckelman. 1989. The Merck Index, llth ed.
Rahway, NJ: Merck and Co., Inc.
Bunyan, J., A.T. Diplock, M.A. Cawthome and J. Green. 1968. Vitamin E and stress.
Nutritional effects of dietary stress with silver in vitamin E-deficient chicks and rats. Br.
J. Nutr. 22:165-182.
Carson, B.L. and I.C. Smith. 1977. Silver, an appraisal of environmental exposure. In: Trace
metals in the environment, vol. 2, silver. Ann Arbor, MI: Ann Arbor Science.
Cearley, I.E. University of Oklahoma. 1971. Toxicity and bioconcentration of cadmium,
chromium and silver in Micropteris salmoides and Lepomuis macrocluiers. Ph.D.
thesis. (Abstract only.) Diss. Abstr. 328:5281.
Chao, T.T. and BJ. Anderson. 1974. The scavenging of silver by manganese and iron oxides in
stream sediments collected from two drainage areas of Colorado. Chem. Geol. 14:159-
166.
Chin, Y. 1973. Recovery of heavy metals by microbes. Ph.D. thesis. London, Ontario:
University of Western Ontario School of Engineering, pp. 18-20.
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CFR. 1987. Code of Federal Regulations. Inductively-coupled plasma atomic emission
spectrometric method for trace element analysis of water and wastes, method 200.7.
Appendix C to part 136, pp. 512-523.
Coleman, R.L. and J.E. Cearley. 1974. Silver toxicity and accumulation in largemouth bass and
bluegill. Bull. Environ. Contam. Toxicol. 12(1):53-61. Reviewed in U.S. EPA. 1979.
U.S. Environmental Protection Agency. Water related environmental fate of 129
priority pollutants. Vol. I. EPA-440/4-79-29a. Washington, DC: U.S. EPA, pp. 17-1 to
17-9.
Cotton, F.A. and G. Wilkinson. 1972. Advanced inorganic che'mistry. New York, NY:
Interscience Publishers, pp. 1,044-1,052. Reviewed in U.S. EPA. 1979. U.S.
Environmental Protection Agency. Water related environmental fate of 129 priority
pollutants. Vol. I. EPA-440/479-29a. Washington, DC: U.S. EPA, pp. 17-1 to 17-9.
Creasey, M. and D.B. Moffat. 1973. The deposition of ingested silver in the rat kidney at
different ages. Experientia 29:326-327.
Danscher, G. 1981. Light and electron microscopic localization of silver in biological tissue.
Histochemistry 71:177-186.
Day, W.A., J.S. Hunt and A.R. McGiven. 1976. Silver deposition in mouse glomeruli.
Pathology 8:201-204.
Demerec, M., G. Bertani and J. Flint. 1951. A survey of chemicals for mutagenic action on E.
co/i. Am. Nat. 85:119-136.
Dequidt, J., P. Vasseur and J. Gromez-Potentier. 1974. Etude toxicologique experimentale de
quelques derives argentiques. I. Localisation et elimination. Bull. Soc. Pharm. Lille
1:23-35. [Experimental toxicological study of some silver derivatives] (in French).
Diplock, A.T., J. Green, J. Bunyan, D. McHale and I.R. Muthy. 1967. Vitamin E and stress.
The metabolism of D-alpha tocopherol in the rat under dietary stress with silver. Br. J.
Nutr. 21:115-125.
Di Vincenzo, G.D., C. J. Giordano and L.S. Schrieves. 1985. Biologic monitoring of workers
exposed to silver. Int. Arch. Occup. Environ. Health 56:207-215.
Durfor, C.N. and E. Becker. 1964. Public water supplies of the 100 largest cities in the United
States, 1962. U.S. Geological Survey Paper 1812. Washington, DC: U.S. Government
Printing Office.
Durum, W.H. and J. Hafty. 1961. Occurrence of minor elements in water. U.S. Geological
Survey Circular 445. Washington, DC: National Academy of Sciences.
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Silver April 1992
Dyck. W. 1968. Adsorption and coprecipitation of silver on hydrous ferric oxides. Can. J.
Chem. 46:1441-1444.
East, B.W., K. Boddy, E.D. Williams. D. Maclntyre and A.L.C. McLay. 1980. Silver retention,
total body silver and tissue silver concentrations in argyria associated with exposure to
an anti-smoking remedy containing silver acetate. Clin. Exp. Dermatol. 5:305-311.
Freeman, R.A. 1977. The ecological kinetics of silver in an alpine lake eco-system. Second
ASTM symposium on aquatic toxicology, Oct. 31 to Nov. 1, 1977. (Preprint only.)
Cleveland, OH: Reviewed in U.S. EPA. 1979. U.S. Environmental Protection Agency.
Water related environmental fate of 129 priority pollutants. Vol. I. EPA-440/479-29a.
Washington, DC: U.S. EPA, pp. 17-1 to 17-9.
Furchner, I.E., G.A. Drake and C.R. Richmond. 1966. Retention of silver-110 by mice.
University of California, Las Alamos: U.S. Atomic Energy Commission, Science
Laboratory, pp. 186-190.
Furchner, I.E., C.R. Richmond and G.A. Drake. 1968. Comparative metabolism of
radionuclides in mammals. IV. Retention of silver-110 in the mouse, rat, monkey and
dog. Health Physics 15:505-514.
Furst, A. 1981. Bioassay of metals for carcinogenesis: Whole animals. Environ. Health
Perspect. 40:83-92.
Furst, A. 1979. Problems in metal carcinogenesis. In: Kharasch, N., ed. Trace metals in heart
and disease. New York, NY: Raven Press, pp. 83-92.
Furst, A. and M.C. Schlauder. 1977. Inactivity of two noble metals as carcinogens. J. Environ.
Pathol. Toxicol. 1:51-57.
Gammill, J.C., B. Wheeler, E.L. Carothers and P.F. Hahn. 1950. Distribution of radioactive
silver colloids in tissues of rodents following injection by various routes. Proc. Soc. Exp.
Biol. Med. 74:691-695.
Ganther, H.E. 1980. Interactions of vitamin E and selenium with mercury and silver. Ann.
N.Y. Acad. Sci. 355: 212-225.
Gaul, L.E. and A.H. Staud. 1935. Clinical spectroscopy. Seventy cases of generalized argyrosis
following organic and colloidal silver medication, including a biospectrometric analysis
of ten cases. J. Am. Med. Assoc. 104:1,387-1,390.
Grasso, P., R. Abraham, R. Hendy, A.T. Diplock, L. Goldberg and J. Green. 1969. The role of
dietary silver in the production of liver necrosis in vitamin E-deficient rats. Exp. Mol.
Pathol. 11:186-199.
24
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Silver April 1992
Gregus, Z., C.D. KJaassen. 1986. Disposition of metals in rats: a comparative study of fecal,
urinary and biliary excretion and tissue distribution of eighteen metals. Toxicol. Appl.
Pharmacol. 85:24-38.
Hannah, S.A., M. Jelus and J.M. Cohen. 1977. Removal of uncommon trace metals by physical
and chemical treatment processes. J. Water Pol. Con. Fed. 49(11):2,297-2,309.
Harvey, S.C. 1985. Antiseptics and disinfectants, fungicides, ectoparasiticides. In: Oilman,
A.G., L.S. Goodman, T.W. Rail and F. Murad, eds. Pharmacological basis of
therapeutics, 7th ed. New York, NY: MacMillan Publishing Co.
Hawley, G.G. 1981. The condensed chemical dictionary, 10th ed. New York, NY: Van
Nostrand Reinhold Co.
Hem, J.D. 1970. Study and interpretation of the chemical characteristics of natural waters.
U.S. Geological Survey Paper 1473. Washington, DC: U.S. Geological Survey, pp. 202-
203.
Hill, W.R. and D.M. Pillsbury. 1939. Argyria, the pharmacology of silver. Baltimore, MD:
Williams and Wilkins Company.
Jackson, W.F. and B.R. Duling. 1983. Toxic effects of silver-silver chloride electrodes on
vascular smooth muscles. Circulation Res. 53(1): 105-108.
Just, J. and A. Szniolis. 1938. Germicidal properties of silver in water. J. Am. Water Works
Assoc. 18(4):492-506.
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concentration of certain trace metals in biological materials. J. Nutr. 19:579-592.
Kent, N.L. and R.A. McCance. 1941. The absorption and excretion of "minor" elements by man.
1. Silver, gold, lithium, boron and vanadium. Biochem. J. 35:837-844.
Kharkar, D.P., K.K. Turekian and K.K. Bertine. 1968. Stream supply of dissolved silver,
molybdenum, antimony, selenium, chromium, cobalt, rubidium and cesium to the
oceans. Geochim. Cosmochim. Acta 32:285-298.
Klaassen, C.D. 1979. Biliary excretion of silver in the rat, rabbit and dog. Toxicol. Appl.
Pharmacol. 50:49-55.
Kopp, J.F. and R.C. Kroner. 1967. Trace metals in waters of the United States. A five-year
summary of trace metals in rivers and lakes of the United States. (October 1, 1962 to .
September 30, 1967.) Cincinnati, OH: U.S. Department of Interior, Federal Water
Pollution Control Administration, Division of Pollution Surveillance.
25
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Silver April 1992
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.
Luoma, S.N. and E.A. Jenne. 1977. The availability of sediment-bound cobalt, silver and zinc
to a deposit-feeding clam. In: Drucher, H. and R.E. Wilding, eds. Biological
implications of metals in the environment. Richland, WA: Hanford Life Sciences
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McCabe, L.J., J.M. Symons, R.D. Lee and G.G. Robeck. 1970. Survey of community water
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National Academy of Sciences. Drinking water and health. Washington, DC: National
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McCarty, P.L., D. Argo, M. Reinhard, J. Giaydon, N. Goodman and M. Aieta. 1982.
Performance of Water Factory 21 in removing priority pollutants. Proc. Water Reuse
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NAS. 1977. National Academy of Sciences. Drinking water and health. Washington, DC:
National Academy of Sciences.
Newton, D. and A. Holmes. 1966. A case of accidental inhalation of zinc-65 and silver-110.
Radiat. Res. 29:403-412.
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rats ingesting silver salts. Arch. Pathol. 49:138-149.
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animal. Am. J. Path. 24:813-833.
Olcott, C.T. 1947. Experimental argyrosis. III. Pigmentation of the eyes of rats following
ingestion of silver during long periods of time. Am. J. Path. 23:783-789.
Phalen, R.F. and P.E. Morrow. 1973. Experimental inhalation of metallic silver. Health Phys.
24: 509-518.
Robkin, M.A., D.R. Swanson and T.H. Shepard. 1973. Trace metal concentrations in human
fetal livers. Trans. Am. Nucl. Soc. 17:97.
26
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Silver April 1992
Rossrnan. T.G. and M. Molina. 1986. The genetic toxicology of metal compounds: II.
Enhancement of ultraviolet light-induced mutagenesis in Escherichia coli WP2. Environ.
Mutagen. 8:263-271.
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of experimental animals. Exp. Eye Res. 42:93-94.
Rungby, J., L. Slomianka, G. Danscher, A.H. Andersen and M.J. West. 1987. A quantitative
evaluation of the neurotoxic effect of silver on the volumes of the components of the
developing rat hippocampus. Toxicol. 43:261-268.
Schmaehl, D. and D. Steinhoff. 1960. Versuche zur krebserzejigung mit kolloidalen silber-und
goldlo'sungen an ratten. Z. Krebsforsch. 63:586-591. [Experimental carcinogenesis in
rats with colloidal silver and gold solutions] (in German).
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primary drinking water regulations for inorganics: Part 3. J. Am. Water Works Assoc.
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Patty's industrial hygiene and toxicology, 3rd ed., vol. 2A. New York, NY: John Wiley
and Sons, pp. 1,881-1,894.
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Tichy, P., J. Rosina, K. Blaha Jr. and M. Cikrt. 1986. Biliary excretion of "ฐAg and its kinetics
in the isolated perfused liver in rats. J. Hyg. Epidemiol. Microbiol. Immunol. 30:145-
148.
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U.S. EPA. 1988. Limiting values of radionuclides intake and air concentration and dose
conversion factors for inhalation, submersion and ingestion. Federal Guidance Report
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189. September 1988.
27
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Silver April 1992
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assessment. Fed Reg. 51(185): 33,992-34,003.
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1 to 17-9.
U.S. EPA. 1975. U.S. Environmental Protection Agency. Chemical analysis of interstate
carrier water supply systems. EPA-430/9-75-005. Washington, DC. Reviewed in NAS.
1977. National Academy of Sciences. Drinking water and health. Washington, DC:
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DC: U.S. Department of Health, Education and Welfare, Public Health Service.
Van Vleet, J.F. 1976. Induction of lesions of selenium-vitamin E deficiency in pigs fed silver.
Am. J. Vet. Res. 37:1,415-1,420.
Wagner, P.A., W.G. Hoekstro and H.E. Ganther. 1975. Alleviation of silver toxicity by selenite
in the rat in relation to tissue glutathione peroxidase. Proc. Soc. Exp. Biol. Med.
148:1,106-1,110.
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11:201-204.
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Pathol. 52:589-593.
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CRC Press, p. 239.
WHO, 1984. Guidelines for drinking water quality, vol. 2. Health criteria and other supporting
information. World Health Organization. Geneva, pp. 141-144.
28
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EPA 0553
~RX000027511
April 199.2.
THALLIUM
Drinking Water Health Advisory
Office of Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory Program, sponsored by the Office of Water (OW), provides
information on the health effects, analytical methodology, and treatment technology that would
be useful in dealing with the contamination of drinking water. Health Advisories (HAs) 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.
HAs 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
endpoints of toxicity. For those substances that are known or probable human carcinogens,
according to the Agency classification scheme (Group A or B), Lifetime Health Advisories are
not recommended. For substances with a carcinogenic potential, chemical concentration values
are correlated with carcinogenic risk estimates by employing a cancer potency (unit risk) value
together with assumptions for lifelong exposure and the ingestion of water. The cancer unit risk
is usually derived from a linearized multistage model with 95% upper confidence limits providing
a low-dose estimate of cancer risk. The cancer risk is characterized as being an upper limit
estimate, that is, the true risk to humans, while not identifiable, is not likely to exceed the upper
limit estimate and in fact may be lower. While alternative risk modeling approaches may be
presented, for example One-hit, Weibull, Logit, or Probit, the range of risks described by using
any of these models 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 carcinogenic risk, the Agency recommends using the linearized
multistage model.
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Thallium Health Advisory April 1992.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.
Thallium 7440-28-0
Thallium chloride 7791-12-0
Thallium sulfate 7446-18-6
Structural Formula
TI
Synonyms
Thallium Tl
Thallium chloride thallous chloride
Thallium sulfate eccothal, thallous sulfate
Uses
Thallium salts are used to manufacture crystals, imitation jewelry, optical systems, and
fiberglass. Thallium is also used to produce pigments, corrosion-resistant alloys,
catalysts, low-temperature thermometers, photoelectric cells, scintillation counters, and
other electronic equipment (Manzo and Sabbioni, 1988).
^Tl has been used in myocardial imaging (Atkins et al., 1977); nonradioactive isomers
of thallium are used in high-temperature superconductors (Waldrop, 1988).
Properties (Windholz et al., 1983)
The properties of thallium compounds vary with the specific compounds; examples
follow:
Thallium Thallium
Thallium chloride sulfate
Chemical Formula Tl T1C1 T12SO4
Atomic/Molecular Weight 204.38 239.85 504.85
Physical State Bluish-white metal White powder
Boiling Point 1457ฐC
Melting Point 303.5ฐC 430ฐC 632ฐC
Density 11.85 7.0 6.77
Vapor Pressure
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Thallium Health Advisory April 1992.
Water Solubility Insoluble 3.8 Mg/mL 2.70 g/100 mL
cold water at 0ฐC; 4.87 g/100
mL at 20ฐC
Specific Gravity 11.85 7.0 6.77
Log Octanol/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Occurrence
Thallium (Tl) is a naturally occurring element; mean concentration in the earth's crust
is 1 ppm (Smith and Carson, 1977).
In addition to the metal, thallium also exists in two oxidation states,' Tl*1 and Tl*3.
Generally, inorganic Tl*1 compounds are more stable than their Tl*3 analogs.
However, Tl*3 can also form stable complexes with many ligands (Manzo and Sabbioni,
1988).
Thallium occurs in small amounts in all living organisms; natural levels in plants are
reported to be between 0.01 and 3,800 ppm ash weight, 0.5 ppm being typical for most
species (Sabbioni and Manzo, 1980).
Cationic and neutral Tl*3 compounds are present in seawater and freshwater, but
concentrations have not been reported (Ridley et al., 1977).
Thallium metal is obtained from the flue dust of copper, zinc and lead smelters and
as a byproduct of cadmium and sulfuric acid production. It has been estimated that
about 1,600 tons (worldwide) of thallium are released to the environment annually
from emission sources, from refineries (as impurities), and from other products (Smith
and Carson, 1977). Iron and steel production and cement industries are sources of
thallium emission to the environment (Manzo and Sabbioni, 1988). It is estimated that
140 tons of thallium are released from plants in the United States (Smith and Carson,
1977).
Sulfide components of coal are particularly rich in thallium, and coal burning is
regarded as a major source of thallium emission to the environment (Smith and
Carson, 1977). Average thallium content in coal varies from 0.05 to 3 ppm; however,
some brown coal has levels up to 100 ppm.
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Thallium Health Advisory April
Environmental Fate
Thallium can be removed from solution by adsorption (onto clay minerals),
bioaccumulation or precipitation as sulfide (in reducing environments). Most of the
ligands common to aerobic waters form soluble salts with thallium, so that
precipitation is not important under oxidizing conditions (Callahan and Slimak, 1979).
Thallium is strongly adsorbed by montmorillonitic clays. Magorian et al. (1974)
demonstrated that a 1-g/L suspension of the clay hectorite could remove 97% of a 100-
/xg/L concentration of thallium (species not given) within 24 hours. Similarly, a 1-mg/L
concentration of thallium was reduced to 115 Mg/L, and a 10-/zg/L solution was
reduced to below 1 /ig/L. The above values are for pH 8.1; adsorption is not as
effective at pH 4.0.
Mathis and Kevern (1975), in a study of heavy metal cycling in a lake in southwestern
Michigan, were able to detect thallium only in the sediments. Thallium levels in the
water, plankton and fish were below the limits of detection.
The alga Ulothrix sp. was able to concentrate thallium by a factor of 127 to 220 within
1 hour; in comparison, the concentration factors for 2.7-hour exposures were 114 for
lead, 30 for cadmium, 80 for zinc and 313 for copper (Magorian et al., 1974).
III. PHARMACOKINETICS
Absorption
Thallium is readily absorbed by humans and laboratory animals following oral, dermal
or intratracheal administration (Barclay et al., 1953; Lie et al., 1960; Munch, 1934;
Shaw, 1933).
Barclay et al. (1953) reported that Tl was readily absorbed from the gastrointestinal
(GI) tract of a female cancer patient administered 1.8 tig "Tl NO3 (4 ng Tl/kg)
following oral administration. In addition to radioactive thallium, the woman also
received single oral doses of 45 mg thallium sulfate (approximately 0.80 mg Tl/kg)
every 3 days for a total of five doses. Blood "Tl levels peaked 2 hours after dosing
and amounted to approximately 3% of the administered radioactivity. At 48 hours
post-dosing, blood "*n concentrations were equivalent to approximately 1.5% of the
dose. The patient excreted 15.3% of the administered radioactivity in the urine within
5.5 days after dosing and eliminated an average of 3.2% of the body burden of Tl for
24 days (i.e., until death). Only 0.4% of the administered "Tl was eliminated in the
feces during the first 3 days after compound ingestion.
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Thallium Health Advisory April 199Z
Male Wistar-derived rats were given one dose of "Tl, as thallous nitrate, by six routes:
oral (767 /ig Tl/kg), intravenous (iv) (38 jig Tl/kg), intramuscular (im) (96 /ig Tl/kg),
subcutaneous (96 ng Tl/kg), intratracheal (123 /ig Tl/kg) and intraperitoneal (ip) (146
Mg Tl/kg) (Lie et al., 1960). The body burden of "Tl, as percent of dose, was similar
for all routes of exposure and was found to decay with a single exponential function,
which extrapolated to 100% at zero time, regardless of route of administration. The
author concluded that thallium is completely absorbed from the GI tract.
Based on the recovery of thallium from the urine, Shaw (1933) reported that at least
61.6% of an oral dose of thallium sulfate (25 mg Tl/kg) was absorbed from the GI
tract of a dog.
Thallium can be absorbed through the skin, as evidenced by the toxicity of topically
applied thallium ointments. Munch (1934) reviewed 51 case histories of women
treated for thallium poisoning following external application of an ointment containing
thallous acetate at a concentration of 3 to 8%. In 29 cases, between 2 and 24 ounces
of the ointment (approximately 53 to 636 mg Tl/kg for a 5.5% ointment and a 50-kg
woman) had been applied an unspecified number of times. By several weeks after
application, neurological and GI symptoms and alopecia were observed.
Distribution
In mice (Andre et al., 1960), rats (Downs et al., 1960; Lie et al., 1960; Sabbioni et al.,
1980, 1982) and humans (Barclay et al., 1953; Talas et al., 1983), Tl distributes
throughout the body, including the kidneys, liver, gonads and brain; highest levels are
found in the kidney, where Tl levels may be two- to four-fold higher than in other
tissues. Several animal studies suggest that Tl also undergoes rapid transplacental
transfer in mice (Andre et al., 1960; Olsen and Jonsen, 1982) and rats (Gibson and
Becker, 1970; Sabbioni et al., 1982; Ziskoven et al., 1983).
Talas et al. (1983) administered to each of five female and five male patients an iv
tracer dose of 201T1, as thallium chloride (less than or equal to 106 and 130 ng/kg for
male and female patients, respectively). Levels of radioactivity were serially assayed
in plasma for up to 1 day after dosing, and the data were analyzed in terms of a two-
compartment model. For the central compartment, the distribution volume was found
to be 0.26 L/kg, which is similar to the adult's extracellular volume of about 0.2 L/kg.
Distribution into the peripheral compartment was very fast (Tw = 3.9 min). The total
volume of distribution was found to be 4.23 L/kg, which is much larger than the
average total body water volume in humans (0.6 L/kg). This large observed total
volume of distribution is consistent with a migration of Tl into the intracellular space.
The average terminal half-life of Tl was determined to be 2.15 days.
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Thallium Health Advisory April 1992.
Barclay et al. (1953) also studied the tissue distribution of ^Tl in the female cancer
patient (45.4 kg) who had been dosed orally for 15 days until a total dose of 1.8 tig of
204TlNOj was achieved (approximately 4 ng Tl/kg). Following the death of the patient
24 days after the initial administration of 2ฐ*T1, analysis of tissue radioactivity indicated
that Tl was widely and unevenly distributed throughout the organs. It was estimated
that radioactivity in tissues amounted to 45% of the dose. Levels of Tl in tissues were
represented on a relative scale, as percent radioactive Tl per gram of the average body
distribution per gram body weight (i.e., the total body radioactivity at death/weight of
patient in grams). Highest tissue levels were found in scalp hair (420% per gram),
followed by renal papilla (354%), renal cortex (268%), heart (236%) and spleen
(200%). Intermediate levels were found in adrenal medulla (157%), pancreas (129%),
liver (125%), rib marrow (124%) and adrenal cortex (109%). Lower values were
found in nervous tissue (70 to 13%), ovary and capsule (54%) and abdominal
subcutaneous fat (6.7%).
Valence state had essentially no effect on organ or intracellular distribution of Tl. At
16 hours after oral administration of 3.15 mg MtTl (Tl*1 or Tl*3) (15.75 mg/kg) to
groups of five male Sprague-Dawley rats (190 to 200 g) the kidneys, liver, and testes
contained approximately 6 to 7, 3 to 4, and 4%, respectively, of the radioactive dose;
the salivary glands, heart and brain each contained <.!% (Sabbioni et al., 1980).
About 36 to 42% of the total radioactivity in kidney homogenates was recovered in the
cytosol, 23 to 27% was in the nuclei, and 15 to 19% was in the liposomes;
mitochondrial and microsomal fractions each contained approximately 10% of the total
kidney 201T1. Only trace amounts of MtTl (i.e., a combined total of <1% of the
administered radioactivity) were recovered from the six organs listed above after a
group of five male rats were dosed orally with M1Tl-labeled dimethyl thallium bromide
in olive oil (3.15 mg Tl/rat). Although organ distribution differed, the pattern of
intracellular distribution of dimethylthallium was the same as that of Tl*1 and Tl*3.
Lie et al. (1960) administered 204T1, as thallium nitrate, by the oral route (767 /ig Tl/kg)
and by five other routes, as indicated above, to groups of four or five male Wistar rats.
In orally dosed animals, over the first 7 days, the highest levels of Tl per gram of tissue
were found in kidney (4.71% of the body burden per gram of tissue), followed in
decreasing order by salivary glands (1.08%), testes (0.88%), muscle (0.78%), bone
(0.74%), GI tract (0.62%), spleen (0.56%), heart (0.54%), liver (0.52%), respiratory
system (0.49%), hair (0.37%) skin (0.37%) and brain (0.27%). Except for hair, which
accounted for 60% of the body burden after 21 days, the relative concentrations of Tl
in tissues were largely independent of time.
Thallium (administered as 10 ppm thallous sulfate, equivalent to 740 /xg/kg bw/day)
concentrated in the testes of 10 male Wistar rats, who ingested the compound in
drinking water for 60 days (Formigli et al., 1986). Testes of treated rats contained an
average of 6.3 /xg Tl/g.
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Thallium Health Advisory April 1992
In a study by Downs et al. (1960), Wistar rats (of unspecified sex and number) were
fed a diet containing thallous acetate at a concentration of 0.003% (approximately 1.4
mg Tl/kg bw/day assuming 15 g of diet/day and a body weight of 250 g). After 63 days
on the diet, the highest levels of Tl were found in kidney (24 /ig Tl/g wet tissue),
followed by liver, bone, spleen, lung and brain. Levels in these other tissues decreased
from 16 jig Tl/g wet tissue in liver to 5 Tl/g wet tissue in brain.
The highest concentrations of Tl (60 /xg/g tissue) were found in the kidney and small
intestine of four pregnant COBS rats 72 hours after each animal was given a single
oral (gavage) dose of 10 mg Tl/kg on gestation day 17 (Sabbioni et al., 1982). Liver,
brain and placenta tissue contained approximately 18 to 19 jig/g; salivary glands and
blood had average thallium concentrations of 10.6 and 2.0 p.g/g, respectively. Thallium
was also found in the maternal brain (18.4/ig Tl/g tissue) and in the fetal brain (10.1
Mg Tl/g).
Ziskoven et al. (1983) detected Tl in maternal kidney and brain and in fetuses of 20
pregnant Wistar rats and 16 pregnant Kissleg mice within 10 minutes after the animals
were administered thallous sulfate orally at a dose level of 8 mg Tl/kg. Concentrations
of Tl between 1 and 4 x 10"4 M were measured in maternal kidneys at 1 through 50
hours post-dosing. Maternal brain levels of Tl increased slowly, peaking at 3 x 1CCS M
Tl 60 hours after compound administration. Fetal tissue contained up to 7 x 10"3 M
Tl at 1 hour and within 4 hours after dosing, Tl concentrations peaked at 4 x 10"3 M.
No statistical differences existed between species.
In a study by Gibson and Becker (1970), pregnant Sprague-Dawley rats were infused
continuously on day 20 of gestation with ""Tl, as thallous sulfate, at doses of 0.2, 0.4,
0.8,1.6, 3.2 or 6.4 mg/min/kg. These infusion rates correspond to 0.16, 0.32, 0.64,1.3,
2.6 and 5.2 mg Tl/min/kg. Thirty-two minutes after the initiation of infusion, maternal
blood Tl levels were approximately 15 times higher than those in fetal blood, on a
weight basis, at the lowest dose, and approximately 30 times higher at the highest dose
(e.g., approximately 450 nmoles/mL and 15 to 16 nmoles/g in maternal and fetal blood,
respectively).
Compared to other tissues, the highest levels of radioactivity (3.35% of the
administered dose/g wet tissue) were recovered from the kidneys of four female COBS
rats 4 hours after ip injection of 2 tig mi\+l (50 /iCi/rat) on gestation day 13 (Sabbioni
et al., 1982). All other tissues contained <0.60%/g; cerebellum, brain and blood each
contained
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Thallium Health Advisory April 1992.
Distribution of radioactive thallium was comparable among groups of five male
Sprague-Dawley rats (190 to 200 g) following ip administration of 0.00004, 2, 20, or
2000 Mg ^'TTVanimal (0.21 ng to 10.3 mg Tl/kg; 50 jiCi/animal) (Sabbioni et al., 1980).
At 16 hours post-dosing, the muscles, kidneys and testes contained approximately 28
to 41, 5 to 6, and 2 to 3%, respectively, of the administered radioactivity. Very low
levels (<1%) of M1T1 were recovered from lung, heart, brain and blood tissue.
Stomach and small intestine levels of "'Tl peaked 2 hours after administration of 2 /ig
Tl, accounting for about 0.7 and 1.5% of the original dose, respectively. "'Tl persisted
in the kidneys of rats in the 2-^tg/animal group, with approximately 2.5% of the
radioactive dose (per gram of tissue) recovered at 192 hours after compound
administration. Brain and heart levels of "'Tl increased slightly between 2 and 192
hours post-dosing in the l-/ig group, but 1 g of either tissue contained no more than
1% of the administered test material. Time-course distribution data for other
experimental groups were not reported. Between 25 and 50% of total homogenate
"'Tl of liver and kidney cells of all dosed rats was recovered from the cytosolic and
nuclear fractions. Thus, tissue and intracellular distribution of Tl are not dose-
dependent, and bioaccumulation is not likely to occur, based on experimental exposure
levels.
Metabolism
Sabbioni et al. (1980), in their metabolism study with rats, recovered equivalent
amounts of "'Tl from several tissues (kidneys, liver, testes, lung, heart, muscle, brain
and blood) and subcellular fractions of liver and testes homogenates (nuclei,
mitochondria, lysosomes, microsomes andcytosol) following oral administration of 3.15
mg Tl*' or Tl*3 (15.75 mg/kg) to groups of five male Sprague-Dawley rats. The
authors note, however, that the form and valence state of the thallium recovered were
not discernable.
Excretion
Barclay et al. (1953) studied the excretion of thallium by a patient female cancer
patient dosed orally with 1.8 jig "TlNOj (4 Atg Tl/kg) and then 45 mg thallium sulfate
(36 mg Tl/kg) every 3 days for a total of five doses. In 5.5 days, the patient had
excreted 15.3% of the radioactive dose in urine; in 3 days, she excreted only 0.4% in
the feces. Approximately 3.2% of the amount of Tl in the body was excreted per day.
Based on the daily excretion of Tl, it was calculated that at the time of death, 24 days
after dosing, residual radioactivity in tissues amounted to 45% of the administered
dose.
Lehmann and Favari (1985) administered thallous sulfate by gavage at a dose of 12.35
mg/kg (approximately 10 mg Tl/kg) to female Wistar rats (4 to 8/group). By day 8
8
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Thallium Health Advisory April 1992.
post-dosing, 32% of the dose was eliminated in feces and 21% in urine. Lie et al.
(I960) reported that, in Wistar rats dosed orally with ""Tl as thallous nitrate (767 /ig
Tl/kg), the ratio of fecal to urinary excretion of thallium increased from about 2 to 5
between days 2 and 16 after dosing, and decreased thereafter to about 3 by day 20.
The ratios of fecal to urinary excretion were similar, regardless of whether dosing was
oral, iv, ip or im.
Most of a single oral dose of thallous sulfate (25 mg TVkg) administered to one dog
was eliminated via the urine (Shaw, 1933). Excretion of Tl was 32 and 61.6% of the
dose at 3 and 36 days after dosing, respectively. No data for excretion in feces were
given.
The biological half-life of Tl in rats is reported to be 3.3 days (Lie et al., 1960) and 7.3
days (Lehmann and Favari, 1985). A mean whole-body half-life of 9.8 days for Tl has
been reported in humans dosed iv with MtTl (Atkins et al., 1977).
IV. HEALTH EFFECTS
Humans
Short-term Exposure
The minimum lethal oral dose for Tl in adult humans is estimated to be in the range
of 0.2 to 1.0 g (i.e., approximately 3 to 14 mg Tl/kg for a 70-kg adult) (Clayton and
Clayton, 1981; Grunfeld and Hinostroza, 1964; Heath et al., 1983; Moeschlin, 1980).
Oral administration (in what the authors reported appeared to be deliberate
poisonings) of approximately 0.72 g Tl, as the acetate, to one 56- and one 60-year-old
male in two or three divided doses (i.e., a total of approximately 10.3 mg Tl/kg/person)
produced neurological symptoms (paresthesia of the hands and feet, facial nerve
weakness, visual impairment) and death within 1 month of symptom development
(Cavanagh et al., 1974). A smaller dose of thallium (0.24 g Tl, 3.4 mg Tl/kg),
administered orally to a 26-year-old male, produced chest pain, vomiting, weakness
and paresthesia in both feet. Alopecia appeared 19 days after the onset of symptoms.
Within several weeks, hair regrowth was seen and health returned to normal
(Cavanagh et al., 1974). Postmortem examinations of the two older subjects revealed
minimal neuronal degeneration in the sciatic nerve and spinal cord of one patient and
marked and extensive distal degeneration of several types of nerve fibers and early
degenerative changes in nerve fibers of the oculomotor muscles in the second patient.
Peripheral neuropathy was present in all three individuals. Pulmonary congestion and
edema also were observed, as were centrilobular congestion, necrosis and fatty changes
of the liver. Thallium was recovered from the tissues/remains of the first two
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Thallium Health Advisory April 1992,
individuals and in the urine of the third subject; however, no control data were
reported for comparison.
Several cases of accidental (oral) Tl poisoning have been reported. Ingestion of 0.49
to 1.2 g Tl (7.0 to 17 mg Tl/kg), as the sulfate, by 20- to 25-year old males produced
a wide range of neurological symptoms, ataxia, alopecia, impairment of renal function,
tachycardia, high blood pressure, coma and death (De Groot et al., 1985; Grunfield
and Hinostroza, 1964; Moeschlin, 1980; Moeschlin and Condrau, 1950). Similar
symptoms and death occurred in a 20-year-old man who consumed approximately 3.8
to 7.6 g thallium (54 to 109 mg Tl/kg) as the nitrate salt (Davis et al., 1980).
Ingestion of an unspecified number of Tl-coated peanuts by a 3-year-old boy produced
ptosis, dysarthria, restricted movement of the left leg, and alopecia of the scalp. The
patient eventually recovered, but a residual, slightly winded gait and minimal left ankle
weakness persisted (Taber, 1964).
Munch (1934) reviewed 51 case histories of women treated for neurological and
gastrointestinal symptoms and/or alopecia following external application of a depilatory
ointment containing thallous acetate at a concentration of 3 to 8%. In 29 cases (where
doses could be estimated), between 2 and 24 ounces of the ointment (approximately
53 to 636 mg Tl/kg for a 5.5% ointment and a 50-kg woman) had been applied an
unspecified number of times.
Long-term Exposure
No information was found in the available literature.
Animals
Short-term Exposure
The acute oral LDM values for various Tl compounds in mice and rats range from 16
mg Tl/kg for thallium sulfate in rats to 46 mg Tl/kg for thallium sulfate in mice
(Danilewicz et al., 1979; Downs et al., 1960; NIOSH, 1985; Tikhova, 1964; Truhaut,
1959).
Intraperitoneal administration of thallic chloride tetrahydrate in single doses of 50,100
or 200 mg/kg (26.7, 53.4 or 106.8 mg Tl/kg, respectively) to Sprague-Dawley rats
produced ultrastructural changes in mitochondria, increased number of autophagic
lysosomes and structural changes in the endoplasmic reticulum of the liver (Woods and
Fowler, 1986).
10
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Thallium Health Advisory April 1992.
Subcutaneous administration of thallium acetate in two to three weekly doses of 10 to
15 mg/kg (i.e., 7.8 to 11.6 mg Tl/kg) for up to 16 days to Sprague-Dawley rats
produced mitochondrial granules in kidney and liver cells and lipofuscin bodies in
brain neurons. Acute necrosis was seen in the mesencephalon of two out of four rats
(Herman and Bensch, 1967).
Dermal/Ocular Effects
No information was found in the available literature.
Long-term Exposure
Thallous acetate administered in the diet of 100-g young Wistar rats (five/sex/dose) for
15 weeks at levels of 5, 15 or 50 mg/kg of diet (i.e., 0.4, 1.2 or 3.9 mg Tl/kg bw/day,
respectively) produced high mortality: 20, 40 and 100%, respectively, for the three
dosage groups. Mortality in the control group was also high: 40% by week 15.
Approximately 80 and 60% of the male and female rats, respectively, in a fourth group
of test animals (five/sex) given a diet containing 30 mg thallous acetate/kg feed
(approximately 2.4 mg Tl/kg bw/day) for 63 days died by week 8 of treatment; 70% (7/
10) of the controls for this group survived the 9-week experiment. The only significant
finding at necropsy was moderate to marked alopecia in rats fed a dose of 1.2 mg Tl/
kg bw/day or greater. No histopathologic changes were observed (Downs et al., 1960).
Due to the high mortality rate among all groups, a true no-effect level could not be
established.
In a comparison study, thallic oxide was administered in the diet to groups of five 100-
g young Wistar rats for 15 weeks at levels of 20, 35, 50, 100 and 500 mg/kg diet (i.e.,
1.8, 3.1, 4.5, 9.0 and 44.8 mg Tl/kg bw, respectively). High mortality resulted at all
dose levels: 20% in untreated controls, 20 and 60% at the two lower doses, and 100%
at the three higher doses. Necropsy of survivors showed a significant (p .<0.05)
elevation in kidney weights for females at the two lower test doses and for males
receiving the lowest dose. Alopecia was observed at all dose levels. No
histopathologic changes were found in lungs, liver, kidneys and brain (Downs et al.,
I960). Again, the high mortality rate precluded identification of a true health effect
level.
Thallium sulfate administered in the drinking water of 80 female Sprague-Dawley rats
at 10 mg Tl/kg (i.e., about 1.4 mg Tl/kg/day) for 36 weeks produced 21% mortality and
a 20% incidence of alopecia. Abnormal electrophysiological parameters (e.g., decrease
in amplitude of motor action potentials) were observed in 10 of 16 rats. Histological
examination of the sciatic nerve from six rats revealed Wallerian degeneration of
scattered fibers and vacuolization and lamination of the myelin sheath of about 10%
of the fibers (Manzo- et al., 1983). Because the study employed only one dose level.
11
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Thallium Health Advisory April 1992*
and because the one dose used caused a high number of deaths, a no-effect level could
not be established.
Thallium sulfate administered by gavage to Sprague-Dawley rats (20/sex/dpse) at levels
of 0.01, 0.05 or 0.25 mg/kg bw/day (i.e., 8.1, 40.5 or 202 /xg Tl/kg bw/day, respectively)
for 90 days resulted in significant (p <0.05) changes in serum enzymes and electrolytes
(Stoltz et al., 1986). In males, elevations in the enzymatic activities of serum glutamic-
oxaloacetic transaminase (SGOT) and lactic dehydrogenase (LDH) were observed.
Serum levels of sodium and calcium were significantly (p <0.05) elevated at the two
higher dose levels. In females, dose-related increases in sodium were apparent at the
end of treatment, and SGOT and LDH were transiently elevated at 30 days of
treatment. No histopathologic changes were found at the light microscopic level.
Electron microscopy was not performed. A No-Observed-Adverse-Effect Level
(NOAEL) of 0.20 mg/kg/day was established, based on the absence of gross and light
microscopic histopathology.
Subcutaneous administration of thallium acetate to 15 Sprague-Dawley rats as an
initial injection of 10 to 20 mg/kg (i.e., 7.8 to 15 mg Tl/kg) followed by weekly
injections of 5 mg/kg (i.e., 3.9 mg Tl/kg) for up to 24 weeks produced ultrastructural
damage in kidneys, liver and brain. Although no changes were visible at the light -
microscopic level, electron microscopic examination revealed mitochondria! changes
in kidney and liver and lipofuscin bodies in brain neurons (Herman and Bensch, 1967).
A no-effect level was not identified due to uncertainty in dosing protocols.
Reproductive Effects
Ten male Wistar rats exposed to 10 ppm thallous sulfate (approximately 740 /zg Tl/kg/
day) in drinking water for 60 days exhibited adverse effects on sperm cell maturation
and mo til icy. In addition, histological examinations showed alterations in the
epithelium of the seminiferous tubules and in the Sertoli cells. The testicular Tl
concentration was 6.3 /ig/g in treated rats versus less than 0.08 ng/g in controls (p
<0.01) (Formigli et al., 1986). The use of only one dose precluded establishment of
a NOAEL and Lowest-Observed-Adverse-Effect Level (LOAEL) for the reproductive
effects of Tl.
Developmental Effects
In a study by Claussen et al. (1981), pregnant NMRI mice (29 to 36/ dose level) were
dosed by gavage with thallous acetate or thallous chloride at daily doses of 0, 3 or 6
mg/kg (corresponding to 0, 2.3, and 4.7 mg Tl/kg/day for the acetate, and 0, 2.6 and
5.1 mg Tl/kg/day for the chloride, respectively) on days 6 through 15 of gestation. In
groups receiving thalious chloride, a slight increase in postimplantation loss and
postnatal mortality was noted only at the 6-mg/kg dose level. For groups receiving
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Thallium Health Advisory April 1992.
thallous acetate, slightly increased incidences of cleft palate were observed at 3 and
6 mg/kg; a slight reduction in fetal weights was also noted at 6 mg/kg. Thus, this study
established a NOAEL and LOAEL of 3 and 6 mg/kg, respectively, for developmental
toxicity of thallous chloride; in contrast, all levels of thallous acetate had some type
of adverse effect on fetal development.
Claussen et al. (1981) also administered thallous acetate or thallous chloride to groups
of 9 to 24 pregnant Wistar rats by gavage at doses of 0, 3, 4.5 or 6 mg/kg (i.e., 0, 2.3,
3.5 and 4.7 mg Tl/kg/day for the acetate, and 0, 2.6, 3.8 and 5.1 mg Tl/kg/day for the
chloride, respectively) on days 6 through 15 of gestation produced 100% maternal
mortality in the two highest dose groups. Lower doses of 3 mg/kg (2.3 or 2.6 mg Tl/kg/
day) were associated with developmental effects including wavy ribs and dumbbell-
shaped sternebrae in fetuses, and a slight increase in postnatal mortality. All doses,
therefore, were associated with adverse developmental effects.
Group of six pregnant Sprague-Dawley rats were administered thallous sulfate ip at
2.5 mg/kg (2.0 mg Tl/kg/day) on days 8 to 10 or 12 to 14 of gestation or at 10 mg/kg
(8.1 mg Tl/kg/day) on days 12 to 14 of gestation. All thallium-treated groups had
litters with reduced fetal body weights. Increased incidences of unossified vertebral
bodies were noted in fetuses of both test groups dosed on days 12 to 14 of gestation.
Signs of maternal toxicity were noted in all Tl treatment groups -- diarrhea, lethargy,
irritability and body hair loss. However, incidences of hydronephrosis were increased
only in fetal groups receiving 2.5 mg/kg (Gibson and Becker, 1970). The data show
that administration of either 2.0 or 8.1 mg Tl/kg/day produces adverse developmental
effects in fetal rats.
Postnatal intravenous administration of thallous sulfate to 6- and 9-day-old pups
(number not specified) at doses of 20 and 40 jig/g (16 and 32 mg Tl/kg, respectively)
produced defective calcification and severely hypoplastic columnar cartilage of the long
bones by day 18. However, the dose causing these effects was not specified (Nogami
and Terashima, 1973).
Cultured 10.5-day-old rat embryos (number not reported), when exposed to 3, 10, 30
or 100 Mg Tl/mL, exhibited dose-related growth retardation at all dose levels, with a
complete growth inhibition at 100 jig/mL. At 3 ng/mL, no gross abnormalities were
evident, but cytotoxicity to the central nervous system was revealed microscopically
(Anschutz et al., 1981). Thus, all doses had some detrimental effect on developing rat
fetuses.
Mutagenicitv
Reverse mutation assays in Salmonella typhimurium and Escherichia coli were negative
(Claussen et al., 1981; Kanematsu et al., 1980). Although the results were uniformly
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Thallium Health Advisory April 1992.
negative, limitations of the Ames assay for detection of metal-induced mutagenesis
make the results in Salmonella inconclusive (McCann et al., 1975). Similar problems
may be encountered with the results in E. coli.
Dose-dependent increases in chromosome and chromatid breaks were observed in rat
embryo fibroblasts incubated in the presence of 10^ to 10"5 M thallium carbonate for
24 hours (Zasukhina et al., 1981).
Thallium was positive in a rat dominant lethal assay, using animals dosed orally with
thallium carbonate solution (5 x 10"* to 5 x 10* mg/kg/body weight, daily for 8 months).
The significance of these results is difficult to assess due to incomplete reporting of
both the results and the protocol (Zasukhina et al., 1981, 1983).
Incubation of the Bacillus subtilis DNA repair-competent strain, H17 (rec*), and the
DNA repair-deficient strain, M45(rec"), with a single dose (0.001 M) of nonactivated
thallium nitrate caused preferential inhibition of M45, indicating that DNA damage
had occurred (Kanematsu et al., 1980).
Thallium carbonate produced decreased levels of vaccinia virus reactivation, and a
dose-dependent increase in strand breaks in C57BL/6 and CBA mouse embryo cells
and rat embryo fibroblasts, following incubation with 10"4 to 10"* M thallium carbonate
(Zasukhina et al., 1981, 1983).
No significant increases in sister chromatid exchange were observed in bone marrow
cells of Chinese hamsters administered two oral doses of 5 or 10 mg/kg thallium
chloride in a 24-hour period (Claussen et al., 1981).
Thallium acetate at 0.1 and 0.2 mM produced a significant (p < 0.01) and dose-related
enhancement of viral-induced cell transformation of Syrian hamster embryo cells
incubated in the presence of 0.025, 0.05, 0.1 or 0.2 mM thallium acetate (Casto et al.,
1979).
Carcinogenicitv
No studies on the carcinogenicity of thallium were identified.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories are based upon the identification of adverse health effects associated
with the most sensitive and meaningful noncarcinogenic end point of toxicity. The induction of
this effect is related to a particular exposure dose over a specified period of time, most often
determined from the results of an experimental animal study. Traditional risk characterization
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Thallium Health Advisory April 1992.
methodology for threshold toxicants is applied in HA development. The general formula is as
follows:
(NOAEL or LOAEL) (BW)
' '
where:
NOAEL = No-Observed-Adverse-Effect Level (the exposure dose in mg/kg bw/day).
or
LOAEL = Lowest-Observed-Adverse-Effect Level (the exposure dose in mg/kg bw/day).
BW = assumed body weight of protected individual (10 kg for child or 70 kg for
adult).
UF(s) = uncertainty factors, based upon quality and nature of data (10, 100, 1,000, or
10,000 in accordance with EPA or NAS/OW guidelines).
L/day = assumed water consumption (1 L/day for child or 2 L/day for adult).
One-day Health Advisory
The available data are insufficient to develop a One-day HA for Tl. It is recommended
that the Longer-term HA for the 10-kg child, 7.0 /ig Tl/L, be used as a conservative estimate of
the One-day HA value.
Ten-day Health Advisory
The available data are insufficient to develop a Ten-day HA for Tl. It is recommended
that the Longer-term HA for the 10-kg child, 7.0 pg Tl/L, be used as a conservative estimate of
the Ten-day HA value.
Longer-term Health Advisory
The study of Stoltz et al. (1986) has been selected to serve as the basis for the Longer-
term HA for Tl. In this study, Sprague-Dawley rats (20/sex/group) were dosed by gavage with
0.008, 0.040 or 0.20 mg Tl/kg/day (dose based on the percent of Tl in T12SO4) for 13 weeks. Gross
pathologic and light microscopic examination of organs and tissues did not reveal any significant
treatment-related effects. Apparent instances of alopecia were not confirmed following light
microscopic examination of hair follicles for possible damage. Treated rats also displayed
moderate changes in blood chemistry, which included increases in SGOT, LDH, and sodium
15
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Thallium Health Advisory April 1992.
levels. However, although the enzymatic activities were significantly elevated (p < 0.05) with
respect to vehicle-treated controls, it was not possible to ascertain whether the effects were dose-
related due to scattering of data points. Furthermore, no gross or histopathologic lesions were
reported that correlated with the blood chemistry changes. Thus, based on the absence of gross
or light microscopic histopathologic changes, the dose of 0.20 mg Tl/kg/day is considered a
NOAEL.
Other studies identified toxic effects of Tl but were not used because NOAELs were not
established and/or effects were seen at higher doses than the NOAEL of 0.20 mg Tl/kg/day
reported by Stoltz et al. (1986). These studies are described below.
True NOAEL and LOAEL values could not be established in the Downs et al. (1960)
study in which Wistar rats maintained for up to 15 weeks on a Tl-containing diet developed
alopecia; the significance of the results cannot be assessed due to the high mortality seen in all
dose levels and controls. Furthermore, tests for testicular toxicity and neurotoxiqity, which were
observed in other studies (Formigli et al., 1986; Herman and Bensch, 1967), were not conducted.
Thus, the study by Downs et al. (1960) is not suitable for derivation of the Longer-term HA.
Manzo et al. (1983) reported axonal destruction and increased mortality in rats
administered 1.4 mg Tl/kg/day in the drinking water for 36 weeks, and Formigli et al. (1986)
reported decreased sperm motility, increased numbers of immature sperm cells, and cytoplasmic
vacuolization in Sertoli cells in rats administered 0.74 mg TL/kg/day in the drinking water for 60
days. Because both studies were performed at only one dose level, and consequently no data on
the dose-response relationship are available for extrapolating effects to lower dose levels, neither
study is suitable for derivation of the Longer-term HA.
Herman and Bensch (1967) reported histopathological changes in the kidney, liver and
brain of rats dosed subcutaneously with Tl. Following an initial subcutaneous injection (precise
dose unspecified, in the range of 7.8 to 15.5 mg Tl/kg), rats were administered weekly
subcutaneous injections of 3.9 mg Tl/kg, generally once each week (approximately 0.6 mg Tl/kg/
day), for up to 24 weeks. The uncertainty in the dosing protocol, the relatively high levels of Tl
administered in each weekly dose, and the difficulties in extrapolating between subcutaneous and
oral administration make this study unsuitable for derivation of HA values.
In view of the uncertainty associated with the selected NOAEL value of 0.20 mg Tl/kg/
day(Stoltz et al., 1986), an uncertainty factor of 3 has been introduced into the calculations of the
Longer-term HA to account for inadequate testing of other endpoints of toxicity. This factor of
3 is in addition to the factor of 100 for use with a NOAEL for an animal study.
The Longer-term HA for a 10-kg child is calculated as follows:
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Thallium Health Advisory April 1992.
Longer-term HA . "g) 0.0066 mg/L (rounded * 7
where:
0.20 mg Tl/kg/day = NOAEL, based on the absence of gross or light-microscopic histopathologic
changes in rats exposed to thallous sulfate, by gavage, for 90 days (Stoltz et
al., 1986).
10 kg = assumed weight of a child.
100 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines for
use with a NOAEL from an animal study.
3 = extra uncertainty factor to account for inadequate testing of other species,
endpoints, and uncertainties with associated critical study.
1 L/day = assumed water consumption of a 10-kg child.
The Longer-term HA for a 70-kg adult is calculated as follows:
T ... (0.20 mg Tl/kg/day) (70 kg) nfY,i r, , A * > ~>r\ * \
Longer-term HA = ^ - 2 - ฐL LLฑ - ฐL = 0.023 mg/L (rounded to 20 iig/L)
* (100) (3) (2 L/day) * V ** '
where:
0.20 mg Tl/kg/day = NOAEL, based on the absence of gross or light-microscopic histopathologic
changes in rats exposed to thallous sulfate, by gavage, for 90 days (Stoltz et
al., 1986).
70 kg = assumed weight of an adult.
100 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines for
use with a NOAEL from an animal study.
3 = extra uncertainty factor to account for inadequate testing of other species,
endpoints, and uncertainties associated with a critical study.
2 L/day = assumed water consumption of a 70-kg adult.
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Thallium Health Advisory April 1992,
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
estimate (with uncertainty spanning perhaps an order of magnitude) of a daily exposure to the
human population (including sensitive subgroups) that is likely to be without appreciable risk of
deleterious health effects during 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 in drinking water
alone 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 known, probable or possible carcinogen, according to
the Agency's classification scheme of carcinogenic potential (U.S. EPA, 1986c), then caution must
be exercised in making a decision on how to deal with possible lifetime exposure to this substance.
For human (A) or probable (B) human carcinogens, a Lifetime HA is not recommended. For
possible (C) human carcinogens, an additional 10-fold safety factor is used in the calculation of
the Lifetime HA. The risk manager must balance this assessment of carcinogenic potential and
the quality of the data against the likelihood of occurrence and significance of health effects
related to noncarcinogenic endpoints of toxicity. To assist the risk manager in this process,
drinking water concentrations associated with estimated excess lifetime cancer risks over the range
of 1 in 10,000 to 1 in 1,000,000 for the 70-kg adult drinking 2 L of water/day are provided in the
Evaluation of Carcinogenic Potential section.
No chronic study was found to determine the Lifetime HA for Tl. The 90-day study with
rats by Stoltz et al. (1986) has been selected for calculation of the Lifetime HA for Tl. In this
study, Sprague-Dawley rats (20/sex/group) were dosed with T12SO4 by gavage at doses of 0.008,
0.040 or 0.20 mg Tl/kg/day for 13 weeks. Moderate dose-related changes were observed in some
blood chemistry parameters: increased SCOT, LDH and sodium levels, and decreased blood
sugar levels. The only grossly observed finding at necropsy thought to be treatment-related was
alopecia, especially in female rats; however, microscopic evaluations did not reveal any
histopathologic alterations. Furthermore, no gross or histopathologic lesions were observed that
correlated with the blood chemistry changes. Based on the results of this study the dose of 0.20
mg Tl/kg/day is considered a NOAJEL.
Using the study of Stoltz et al. (1986), the DWEL is derived as follows:
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Thallium Health Advisory April 199Z
Step 1: Determination of the Reference Dose (RfD)
RfD = (0-2ฐ = 0.066 ug Tl/kg/day (rounded to 0.07 ug Tl/kg/day)
where:
0.20 mg Tl/kg/day = NOAEL, based on the absence of gross or light-microscopic histopathologic
changes in rats exposed to thallous sulfate, by gavage, for 90 days (Stoltz et
al., 1986).
1,000 = uncertainty factor, chosen in accordance with EPA or NAS/OW guidelines for
use with a NOAEL for an animal study of less-than-lifetime duration.
3 = additional uncertainty factor to account for inadequate testing of other
species, endpoints of toxicity, and uncertainties associated with the critical
study.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL __ (0.07 ug Tl/kg/day) (70 kg) = ^ ^ m (rQunded to 2
(2 L/day) \
where:
0.07 Mg Tl/kg/day = RfD.
70 kg = assumed weight of an adult.
2 L/day = assumed water consumption of a 70-kg adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (2.0 jig TVL) (20% ) = 0.4 ug Tl/L
where:
2.0 Mg Tl/L = DWEL.
20% = assumed relative source contribution from water.
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Thallium Health Advisory April 1992.
Evaluation of Carcinogenic Potential
The carcinogenic potential of Tl has not been evaluated by the U.S. Environmental
Protection Agency (U.S. EPA) or the International Agency for Research on Cancer
(IARC).
No quantitative assessment of excess cancer risk due to Tl has been reported.
Thallium was found to be genotoxic in in vitro assays with mammalian cells.
VI. OTHER CRITERIA. GUIDANCE AND STANDARDS
Thallium salts are designated as a hazardous substance under Section 311(b)(2)(A) of the
Federal Water Pollution Control Act and further regulated by the Clean Water Act
Amendments of 1977 and 1978. These regulations apply to discharges of these substances
(U.S. EPA, 1986a).
Reportable quantity of thallium salts, when discharged into or upon the navigable waters
and adjoining shorelines of the United States, is 1,000 pounds (454 kg) (U.S. EPA,
1986b).
VII. ANALYTICAL METHODS
Methods for metal analysis involve spectroscopy, either emission or absorption. In all of
the methods, the metal is dissolved and thermally excited. All elements when excited
emit or absorb light frequencies characteristic of that element. Most metal spectroscopy
is done in the ultra-violet and x-ray regions.
Direct Aspiration Atomic Absorption Spectroscopy (AA). In this technique the dissolved
metals are aspirated into a flame source, and excited to the point that the metals are
dispersed to a mono-atomic state, a light source whose cathode is the metal of interest
passes through the flame, the resulting absorption of light by the element of interest is
directly proportional to concentration. The disadvantages of this technique is the one at
a time determination of the metals and the insensitivity of the technique. This is EPA
method number 279.1 with a detection limit of 25
Graphite Furnace Atomic Absorption (GFAA). In this technique a specific amount of
liquid is dried on the thermal source, effecting a concentration step. The sample is
electrothermally excited. This technique is very sensitive. This is EPA method number
286.2 with a detection limit of 0.25 /ig/L.
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Thallium Health Advisory April 1992
Inductively Coupled Plasma Atomic Emission Spectroscopv. This method utilizes
aspiration of a liquid sample, but the flame is actually a plasma torch of argon excited to
super hot levels by radio-frequency (RF) radiation. Metals are excited to levels where
they emit radiation. By using classical dispersion grating optics a large number of metals
can be determined simultaneously. This is EPA method 200.7 with a detection limit of
10/ig/L.
Inductively Coupled Plasma Mass Spectrometrv (ICP/MS). In this case the excitation is
again by an RF plasma, but the excited atoms are then interfaced into a mass
spectrometer. Quantitation is achieved by computerized software programs similar to
those in other EPA MS organic methods. This is EPA method number 6020 with a
detection limit of 0;03 /ig/L.
VIII. TREATMENT TECHNOLOGIES
The available literature indicates that conventional coagulation/filtration processes will
possibly remove Tl from contaminated drinking water. If granular activated carbon
(GAG) is added as a post-treatment, then the removal efficiency will be improved.
Hannah et al. (1977) tested the effectiveness of a pilot plant utilizing coagulation/filtration
and excess lime treatment in removing Tl. The plant consisted of a rapid mix designed
for a capacity of 4 ppm, a flocculator, a sedimentation basin and a dual-media filtration.
Thallium was present in the influent at a concentration of 0.5 mg/L. Ferric chloride was
added at a dose of 40 mg/L and at a pH of 6.2; alum was added at a dose of 220 mg/L
and at a pH of 6.4; and hydrated lime was added at a dose 415 mg/L and at a pH of 11.5.
Excess lime treatment produced a Tl reduction of 60%. The alum coagulation produced
a Tl reduction of 31%, and the ferric chloride coagulation produced a Tl reduction of
30%.
Hannah et al. (1977) also reported the results of using GAG adsorption as a post-
treatment to the conventional coagulation/filtration mentioned above, following dual-
media filtration. Using the above Tl-containing effluents from the coagulation/ filtration
process, two GAG columns, operated in parallel and designed as "old" and "new" were
tested. The "old" GAG had been in use for several months before this evaluation was
made. When the lime coagulation effluent was processed through the "old" GAG column,
an additional 24% of the Tl was removed for a total of 84%, and the "new" GAG column
removed an additional 12% for a total of 72%. When the ferric chloride coagulation
effluent was processed through the "old" column, an additional 17% of the Tl was
removed for a total of 47%, and the "new" GAG column removed an additional 15% for
a total of 45%. When the alum coagulation effluent was processed through the "old"
GAG column, an additional 21% of the Tl was removed for a total of 50%, and the "new"
GAG column removed an additional 8% for a total of 39%.
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Thallium Health Advisory April 1992.
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