Ambient Water Quality Criteria
            Criteria  and  Standards  Division
            Office of Water  Planning  and  Standards
            U.S. Environmental  Protection Agency
            Washington, D.C.

                 3,3'-DICHLOROBENZIDINE  (DCB)


                   Freshwater Aquatic Life

     For freshwater aquatic life, no criterion for any dichlo-

robenzidine can be derived using the Guidelines, and there

are insufficient data to estimate a criterion using other


                   Saltwater Aquatic  Life

     For saltwater aquatic life, no criterion for any dichloro-

benzidine can be derived using the Guidelines, and there

are insufficient data to estimate a criterion using other


                        Human  Health

     For the maximum protection of human health from the

potential carcinogenic effects of exposure to dichlorobenzi-

dine through ingestion of water and contaminated aquatic

organisms, the ambient water concentration is zero.  Concen-

trations of     >o     estimated to result in additional

lifetime cancer risks ranging from no additional risk to

an additional risk of 1 in 100,000 are presented in the

Criterion Formulation section of this document.  The Agency

is considering setting criteria at an interim target risk

level in the range of 10  , 10  , or 10   with corresponding

criteria of 1.69 x 10~2 pg/1, 1.69 x 10~3 jug/l and 1.69

x 10   >jg/l respectively.

     Dichlorobenzidine (4,4'-diamino-3,3'-dichlorobiphenyl)
(DCB) is used in the production of dyes and pigments and
as a curing agent for polyurethanes.  This compound is soluble
in organic solvents, but it is nearly insoluble in water
(Stecher , 1968).
     The affinity of DCB for suspended particulates in water
is not clear; its basic nature suggests that it may be fairly
tightly bound to humic materials in soils  (Radding, et al.
1975).  Soils may be moderate to long term reservoirs.
     DCB has been shown to be a carcinogen in non-human
mammals under controlled laboratory conditions.  Exposure
to DCB results in various types of sarcomas and adenocarcino-
mas.  Tumors have been induced both locally (at the site
of injection) and remotely (multi-system involvement after
feeding) (Rye, et al. 1970).  Experiments show DCB to be
a much less potent carcinogen in animals than the unsubstitut-
ed base (benzidine)   (Rye, et al. 1970) .
     No definitive evidence exists to demonstrate DCB as
a carcinogen in man.  Unlike benzidine, DCB exposure to
humans produced no increases in urinary bladder carcinomas
(Gerarde and Gerarde, 1974; Maclntyre, 1975).  Because of
the dichloro-substitution, this compound may be metabolized
differently from benzidine, which is carcinogenic by virtue
of metabolic activation.
     The molecular formula of 3,3'-dichlorobenzidine  (4,4'-
diamino-3,3'-dichlorobiphenyl) is C-12 H-10 Cl-2 N-2 and
the molecular weight is 253.13  (Stecher, 1968).

     DCB forms brownish needles with a melting point of
132C to 133C (Pollock and Stevens, 1965).  It is readily
soluble in alcohol, benzene, and glacial acetic acid  (Stecher,
1968), slightly soluble in HC1  (Radding, et al. 1975), and
sparingly soluble  in water  (0.7 g/1 at 15  (Stecher, 1968).
When combined with ferric chloride or bleaching powder,
a green color is produced  (Pollock and Stevens, 1965).
     DCB's affinity for water-suspended particulate matter
and soils is uncertain; the fact that it is an organic base
suggests that it may be strongly bound to soil materials
(Radding, et al. 1975) .
     Pyrolysis of DCB  will most likely lead to the release
of HC1.  Because of the halogen substitution, DCB compounds
probably biodegrade at a slower rate than benzidine alone.
The photochemistry of  DCB is unknown.
     Assuming the clean air concentrations of ozone  (2 X
10-9M) and an average  atmospheric concentration of hydroxyl
radicals (3 X 10-15M), the half-life for oxidation of DCB's
by either of these chemical species is on the order of one
and one to ten days, respectively.  Furthermore, assuming
a representative concentration of 10~  M for peroxy radicals
in sunlit oxygenated water, the half-life for oxidation
by these species is approximately 100 days, given the vari-
ability of environmental, atmospheric, and aqueous conditions
(Radding, et al. 1975).

     There are few data available on the bioconcentration,
bioaccumulation, and biomagnification of DCB in the aquatic
environment.  DCB has been shown to be experimentally biocon-
centrated by fish to a significant degree - approximately
1150 fold (Appleton and Sikka, Manuscript).  However, no DCB was detected
in fish sampled from the vicinity of a DCB-contaminated
waste lagoon using analytical methods with sensitivities of
10 to 100 jag/kg (Diachenko, 1978) .


Appleton, H.T./ and H.C. Sikka.  Accumulation, elimination,
and metabolism of dichlorobenzidine in the Bluegill Sunfish.

Diachenko, G. 1978. Personal communication. U.S. Pood and
Drug Administration.

Gerarde, H.W., and D.F. Gerarde. 1974.  Industrial experience
with 3,3'-dichlorobenzidine. Epidemiological study of a
chemical manufacturing plant. Jour. Occup. Med. 16: 322.

Maclntyre, D.I. 1975. Experience of tumors in a British
plant handling 3,3'-dichlorobenzidine. Jour. Occup. Med. 17: 23

Pollock, J.R.A., and R. Stevens, eds. 1965. Dictionary of
organic compounds. Eyre and Spottiswoode, London.

Radding, S.B., et al. 1975. Review of the environmental
fate of selected chemicals. U.S. Environ. Prot. Agency,
Washington, D.C.

Rye, W.A., et al. 1970. Facts and myths concerning aromatic
diamine curing agents. Jour. Occup. Med. 12: 211.

Stecher, P.G., ed. 1968. The Merck Index. 8th ed. Merck
and Co., Rahway, N.J.


     There are few data available.on  the  toxicity of  DCB to

freshwater or saltwater organisms.  DCB was  found to  be  acutely

toxic to bluegill sunfish at levels of 0.5 mg/1  or greater in the

water (Sikka, et al. 1978).


     No freshwater or saltwater criterion can be derived for any

dichlorobenzidine using the Guidelines because no Final Chronic

Value for either fish or invertebrate species or a good substi-

tute for either value is available.  There are insufficient data

to estimate a criterion using other procedures.


Sikka,  B.C.,  et al.  1978.   Fate of 3,3'-dichlorobenzidine in

aquatic environment.  U.S. Environ.  Prot. Agency, 600/3-8-068

                 3,3'  DICHLOROBENZIDINE  (DCB)

Mammalian Toxicology and Human. Health Effects

Ingestion from Water

     To date, few systematic measurements of DCB in water

supplies have been undertaken.  In one instance, analysis

of purge wells and seepage water near a waste disposal lagoon

receiving DCB-manufacture wastes showed levels of DCB ranging

from 0.13 to 0.27 mg/1.  High levels of benzidine (up to

2.5 mg/1) were also seen, which may have arisen from photo-

degradation of DCB (Sikka, et al.  1978), since benzidine

is no longer manufactured in the U.S.  Several other dichloro-

benzidine isomers were also detected at levels of 1 to 8

mg/1.  The use of lagoons to handle DCB-containing wastes

might lead to contamination of ground water and pose a threat

to persons relying on nearby wells for drinking water.

     Takemura, et al.  (1965) analyzed the water of the Sumida

River in Tokyo during 1964.   This  river receives the waste

effluents of several dye and pigment factories.  The pres-

ence of DCB was demonstrated by thin layer chromatography.

Although levels of DCB itself were not quantified, colori-

metic analysis revealed that total aromatic amine content

of the water (including benzidine, dichlorobenzidine, ofaiaph-

thylamine, and  /d-naphthylamine) reached levels up to 0.562

mg/1.  The authors suggested that  the presence of the free

amines might be due to chemical reduction of the azo-dyes

by the high levels of H0S and S00  in the river.


Ingestion from Foods
     Few studies have attempted to identify DCB as a contami-
nant of human food.  Since DCB has never had an application
as an agricultural or food chemical, the most .likely source
of dietary DCB would be through consumption of contaminated
     A bioconcentration factor (BCF) relates the concentration
of a chemical in water to the concentration in aquatic organ-
isms, but BCF's are not available for the edible portions
of all four major groups of aquatic organisms consumed in
the United States.  Since data indicate that the BCF for
lipid-soluble compounds is proportional to percent lipids
and the amounts of various species consumed by Americans.
A recent survey on fish and shellfish consumption in the
United States (Cordle, et al. 1978) found that the per capita
consumption is 18.7 g/day.  From the data on the nineteen
major species identified in the survey and data on the fat
content of the edible portion of these species (Sidwell,
et al. 1974), the relative consumption of the four major
groups and the weighted average percent lipids for each
group can be calculated:
                          Consumption       Weighted Average
     Group                (Percent)          Percent Lipids
Freshwater fishes             12                  4.8
Saltwater fishes              61                  2.3
Saltwater molluscs             9                  1.2
Saltwater decapods            18                  1.2
Using the percentages for consumption and lipids for each
of these groups, the weighted average percent lipids is
2.3 for consumed fish and shellfish.


     An average measured steady-state bioconcentration factor

of 500 was obtained for 3,3'dichlorobenzidine in the whole

body of bluegills (Appleton and Sikka, manuscript).   These

fish probably contained about one percent lipids.  An adjustment

factor of 2.3/1.0 = 2.3 can be used to adjust the measured

BCF from the 1.0 percent lipids of the bluegill to the 2.3

percent lipids that is the weighted average for consumed

fish and shellfish.  Thus, the weighted average bioconcentra-

tion factor for 3,3'dichlorobenzidine and the edible portion

of all aquatic organisms consumed by Americans is calculated

to be 500 x 2.3 = 1,150.

     No DCS was detected in fish sampled from the vicinity

of a DCB-contaminated waste lagoon using analytical methods

with sensitivity of 10 - 100 ug/kg (Diachenko, 1978).


     The physical properties of DCS (low volatility, large

crystal structure) probably minimize the risk of exposure

of general populations to DCB through inhalation of air

contaminated through industrial processes.  However, inhala-

tion might represent a major source of occupational exposure-

under sub-optimum working conditions.  Akiyama (1970) examined

the exposure of workers to DCB in a pigment plant in Japan

and determined that during the addition of DCB to reaction

vessels for synthesis of DCB pigments, the concentration

of DCB in air reached 2.5 mg/100 m  in 10 minutes of charging

of reaction vessels, and decreased to 0.2 mg/100 m  within

20 minutes.  The distance of the sampling device from the

operation was not specified.  Also, the amount of total

aromatic amines was elevated in exposed personnel (presumably

due to the presence of DCB).  The mean urinary concentration
of aromatic amines in process workers charging the reaction
vessels with DCB and plant  laboratory workers was 20.1 ppm
and 21.1 ppm, respectively.  Levels were only 14.5 ppm in
workers who dried and cracked the pigments, 12.7 ppm in
office clerks, and 13.6 ppm in controls  (medical students).
Although the concentrations detected were highly variable
(i.e., the mean 20.1 ppm from the charging personnel was
derived from data ranging from 48.5 to 10 ppm), it is possible
that the elevated levels result from DCB exposure, since
Akiyama claims that few or no precautions were taken to
prevent exposure, particularly on hot days.  It is uncertain
whether the amines entered the body through respiration
or through dermal absorption.
     Gerarde and Gerarde (1974)  reported on an industrial
process in which both DCB and the DCB diarylide pigments
were manufactured.  Most steps in the process were performed
in closed systems, and the DCB was handled in a salt form
in a slurry (ca. 80 percent water content).  DCB dust was
said not to be a problem.  The possibility that DCB contamina-
tion exposure could, however, occur is indicated by the
statement that "...the floor and accessible surfaces contamin-
ated with the slurry were usually hosed down to prevent
accumulation of dried material...."  Also,  an outbreak of
dermatitis in the plant was attributed to a process change
in DCB production.  In utilizing DCB in pigment production,
the major sources of potential exposure are listed as the
weighing process and charging of the tanks.  Prior to May
1973,  operators wore gloves and  goggles but not dust face


masks.  DCB was manufactured in this plant from 1938 to

1957.  Thereafter, DCB was purchased from an outside supplier.

On-site inspection of three DCB utilizing plants showed

that two of the plants posed relatively low exposure potential

which was due to use of metal reactors and protective arrange-

ments at the point of tank charging.  However, in the third

plant, chemicals were dumped into open reaction vessels

from an elevated platform, posing an enhanced potential

for exposure.  Therefore, a great deal of variability concern-

ing the exposure of individuals to DCB may exist among various



     Because of large particle size and increased usage

of closed systems and protective clothing, dermal absorption

of DCB probably represents a relatively minor route of DCB

exposure in humans at present.  However, Meigs, et al. (1954)

presented some experimental evidence that under certain

environmental conditions favoring moist skin conditions,

such as high relative humidity and high air temperature,

the dermal absorption by humans of benzidine and possibly

other congenors such as DCB may be enhanced.  An experiment

using one rabbit suggested that exposure to DCB may occur

indirectly by reductive metabolism of DCB-based azo pigments,

liberating the free amine In vivo  (Akiyama, 1970).



     Virtually no information exists that quantifies the

degree and rate of absorption of DCB in experimental animals

or in humans, although Meigs, et al. (1954) detected DCB

in the urine of DCB process and manufacturing workers.


     A detailed distributional study of DCB in rats, monkeys,

and dogs given 0.2 mg/kg of   C-DCB by intravenous injection

was reported by Kellner, et al. (1973).  The results indicate

a rather general distribution within the body after a 14-

day observation period with highest levels found in the

liver of all three species.  The bile of monkeys and the

lungs of dogs showed significant levels.


     Little information exists on the metabolism of DCB

in vertebrates.

     DCB metabolites have not been detected in the excreta

of dogs administered DCB by intraperitoneal injection (Sciarini

and Meigs, 1961), or orally  (Gerarde and Gerarde, 1974).

     Kellner, et al. (1973) examined the urine of a rhesus

monkey given 0.2 mg/kg   C-DCB intravenously and found that

in the first four hours following injection, about one-third

of the urinary   C was unchanged DCB, with another third

identified as mono-N-acetyl DCB, based on chrbmatographic

properties.  The remainder of the urinary   C was not recover-

able into ether at pH = 11.  At later intervals, mostly

metabolites were excreted, with non-extractable   C compris-

ing the majority of this material.

     No ortho hydroxy metabolites of DCB were detected in

the urine of human subjects after oral dosing (Gerarde and

Gerarde, 1974).

     Aksamitnaia (1959) reported that prolonged ingestion

(7.5-8.5 months) of small doses or a single large dose of

DCB in rats led to the appearance of four transformation

products, including benzidine and possibly glucuronide conju-

gates.  At best, this conclusion is tenuous because analysis

was done by paper ctiromatography (one solvent system) without

benefit of radiotracer techniques, and the products were

not quantified or further characterized.   DCB was never

detected in the urine in any of the experiments, and products

were not seen until over 7 months of chronic DCB ingestion.

     In a study of the bioconcentration of DCB in bluegill

sunfish, over one-half of the DCB residues in the fish were

in the form of a conjugate which, under very mildly acidic

conditions, hydrolyzed to reform free DCB (Sikka, et al.


     Hirai and Yasuhira (1972) noted that DCB was not oxi-

dized by cytochrome c, whereas benzidine and other deriva-

tives were oxidized.

     The majority of  information available at present suggests

that DCB is resistant to metabolism, with the exception

of certain conjugative mechanisms and possibly certain bio-

activation steps.  Ring chlorination of benzidine probably

blocks ring hydroxylation reactions of DCB for both electronic

and steric reasons (Shriner, et al. 1978).


     The excretion of DCB and metabolites following a 0.2

mg/kg intravenous dose of   'C-DCB was studied by Kellner,

et al. (1973) in rats, dogs, and monkeys.  With all species,


measurable elimination had ceased within 7 days of administra-
tion  (Table 1).  Fecal excretion was the predominant route
of elimination  in  rats and dogs, and possibly in monkeys.
     Sciarini and  Meigs  (1961) also noted a preponderance
of fecal elimination of  DCS in dogs.  Finally, Gerarde and
Gerarde (1974)  cite an unpublished study utilizing human
volunteers which concluded that DCB is excreted largely
by the fecal route in man as well as in dogs.
     Insufficient  data is available to assess the ability
of the body to  accumulate significant burdens of DCB through
repeated exposures.
Acute, Sub-acute,  and Chronic Toxicity
     Gaines and Nelson (1977) reported the acute oral toxi-
city of DCB to  male and  female mice.  The LD^Q (mg/kg/day)
for DCB given daily for  seven days was 352 for female mice
(slope = 27.39) and 386  for male mice (slope = 23.15).
The single dose LDcg (mg/kg) was 488 for females and 676
for males.
     Gerarde and Gerarde  (1974) listed results of several
toxicological studies with DCB.  DCB-dihydrochloride failed
to produce skin irritation in rabbits at an unspecified
dose.  An intradermal dose of 700 mg/kg also gave a negative
reaction.   One  hundred mg of DCB-free amine placed in the
conjunctival sac of the  eye of a rabbit gave a negative
reaction,  while 20 mg of DCB dihydrochloride produced ery-
thema, pus, and opacity  of the eye, giving a score of 84
of a possible 110  in one hour according to the method of

                                            TABLE 1

              Excretion of   C-DCB in Rats, Dogs and Monkeys Administered   C-DCB



Monkey 1
Monkey 2


18 + 4
8 + 6
of Total Dose Administered
Feces Cage, feed
Phase IV
79 +12 45 1
84+11 -- 5
46 -- 21
26 -- 20

Balance Residues
0-7 day %

98+12 2
97+8 3
Adapted from Kellner, et al. Arch. Toxicol. 1973.

Draize.  The oral LDcn was given as 7.07.g/kg in albino
rats for DCB free amine, and 3.82 g/kg in male and female
Sprague-Dawley rats for DCB dihydrochloride.  For topical
application to skin, an LD^Q of 8 g/kg in male and female
rats was seen.  Pliss (1959) noted that rats given 120 mg
of DCB subcutaneously exhibited a state of excitation with
short-lived convulsions.
     DCB was acutely toxic to bluegill sunfish at levels
of 0.5 mg/1 or greater in the water (Sikka, et al. 1978).
     No incidence of human fatality resulting from exposure
to DCB has been noted.
     Ten rats exposed to a concentrated atmospheric dust
of DCB dihydrochloride for 14 days showed, upon autopsy,
slight to. moderate pulmonary congestion and one pulmonary
abcess (Gerarde and Gerarde, 1974).  An irritant effect
from HC1 cannot be discounted in the study.
     Freeman, et al. (1973) noted that DCB was cytotoxic
to embryonic rat cells in culture at concentrations of 5
ppm or greater.
     No mortalities were obtained in inhalation studies
where rats were exposed to a concentrated atmosphere of
concentrated DCB dihydrochloride dust for 14 days, or to
355 mg DCB free amine for 2 hours daily for 7 days (Gerarde
and Gerarde, 1974).
     Gerarde and Gerarde (1974) listed the principal reasons
for visits to a company medical clinic by employees working
with DCB.  These were as follows:  1) gastro-intestinal
upset; 2) upper respiratory infection; 3) sore throat; 4)


caustic burns; 5) headache; 6) dizziness; 7) dermatitis.
The only illness apparently directly related to DCS was
dermatitis.  An outbreak of dermatitis was attributed to
a manufacturing process change which led to small amounts
of DCB-free base in the isolated DCB sulfate salt.  Two
cases of acute cystitis were found in the medical record
review of the workers.  One was of infectious origin and
the other related to the presence of renal calculi.  Cysto-
scopic examination of three other workers with urinary system
symptoms revealed two had renal calculi, and another had
cystitis cystica.
Synergistic or Antagonistic Compounds
     No data are available concerning compounds which syner-
gize or antagonize the toxicity of DCB.
     No information is available defining the teratogenic
potential of DCB.  While perhaps not directly relevant to
the question of DCB-induced teratogenesis, several studies
show that DCB can cross the placental barrier and can also
affect developmental systems.
     DCB has been demonstrated to significantly increase
the incidence of leukemia in the offspring of pregnant female
mice given comparatively low doses (ca. 8-10 mg)  of DCB
by subcutaneous injection in the last week of gestation
(Golub, et al. 1974).  This could have been due to post-
natal transfer of DCB to the young through lactation.  How-
ever, transplacental effects of DCB have also been observed.
Shabad, et al. (1972) and Golub (1969)  noted that kidney

tissue  taken from embryos of pregnant female mice  treated
with DCB exhibited altered behavior in organ culture,  including
increased survival and hyperplastic changes in epithelium
not seen in controls.
     The degree of exposure of pregnant women to DCB  is
probably low.  The work force involved in the manufacture
and utilization of DCB is predominantly or totally male.
Maclntyre (1975) lists 5 women, all between the ages  of 20
and 34 years, as having been DCB service or production workers
in a plant in Great Britain as opposed to 217 men.  Therefore,
such exposure may be  through consumption of contaminated
fish or water, from DCB impurity in various DCB-based  pigment
products, or in vivo  release of DCB from absorbed  on  ingested
pigments.  There is little evidence for any of these  occur-
rences at present.
     Garner, et al. (1975) compared the relative mutageni-
city of benzidine, DCB, and other analogs in the bacterial
mutagenesis system developed by Ames, et al. (1973),  utili-
zing the Salmonella typhimurium tester strain TA 1598, an
indicator of frame shift mutagenesis.  The relevant data
are summarized in Table 2.  These results show that DCB
is considerably more  potent as a frameshift mutagen in this
system  than is benzidine.  Also, a low degree of mutation
is elicited by DCB but not by benzidine in the absence of
the S-9 activation enzyme system.

                              TABLE  2

               Mutagenicity  of  DCB  in  the  Ames  Assay


3,3' -dichlorobenzidine
sulfate salt, technical grade


Dimethyl sulfoxide (control)

^ig chemical/

2 Revertants/
S-9 plate
+ 3360
+ 7520
+ 5490
+ 8350
-f- 430
+ 640
+ 16
'Adapted  from Garner,  et  al.  Cancer  Lett.  1:39,  1975.

"S9  is  the NADPH-fortified  rat  liver activation  enzyme preparation.
 + signifies preparation  present;  -, preparation absent.

     Similar observations were made by Lazear and Louis

(1977), utilizing an enzyme activation system obtained from

the livers of male mice and Ames tester strain TA98  (an

indicator of frameshift mutation).  As before, DCB was much

more mutagenic than benzidine and, unlike benzidine, retained

an appreciable mutagenic activity without the liver enzymes.

DCB was also slightly mutagenic towards tester strain TA100,

indicating base-pair substitution mutation,


     Stula, et al. (1975) maintained 50 male and 50 female

rats on a dietary level of DCB of 1000 mg/kg.  The average

50 percent survival was 356 days with average days on the

test of 349 days for females and 353 days for males.  The

range of days on the test was 118-486 days for males and

143-488 days for females.  The rats were 38 days old at

the start of the assay and were apparently autopsied at

time of death or after 486-488 days (not specified).  The

results of this study are listed in Table 3.

                                TABLE 3

              Induction of Cancer in Male and Female Rats
                        by 1000 ppm.Dietary DCB
No. of Cancers
Type of Cancer

Mammary adenocarcinoma
Zymbal gland
-Adapted from Stula, et al. Toxicol. Appl. Pharmacol. 1975.
3Significantly greater than controls at p  0.05
 The number of animals examined histologically was 44 each for male
 and female.

     In addition to the cancers listed in Table 3, the occur-
rence of malignant lymphoma was elevated over controls but
not at statistically significant (p<0.05) levels.  No
bladder cancer was noted.
     In a recent study, Stula, et al. (1978) reported
on the induction of both papillary transitional cell carcinomas
of the urinary bladder and hepatic carcinomas in female
beagle dogs.  An oral dose of 100 mg DCB was administered
to the experimental animals, 3 times per week for 6 weeks,
then 5 times per week continuously for periods up to 7.1
years.  DCB was found to be carcinogenic at statistically
significant levels (p < .025).  The incidences of hepatic
carcinomas were 4/5 and 0/6 in DCB-treated and control groups,
respectively.  The incidences of urinary bladder carcinomas
were 5/5 and 0/6, respectively.
     For 12 months, 6 times weekly, Pliss (1959) added 0.5
to 1.0 ml of a 4.4% suspension of DCB to the feed of rats
(mixed sex) of a strain assumed by Pliss to have a low sponta-
neous tumor rate.  Each rat received a total dose of 4.53 g.
Neoplasms were detected in 22 of 29  (75.8 percent) surviving
animals.  Tumors, primarily carcinomas, were observed in
a broad spectrum of organs including mammary gland, Zymbal
glands (sebaceous gland of the external auditory meatus),
bladder, skin, small intestine, liver, thyroid gland, kidney,
hematopoietic (lymphatic) system, and salivary glands.
     An assay of DCB carcinogenicity was also done with
mice (Pliss, 1959).  The mice received 0.1 ml of a 1.1 percent
DCB suspension in their food for 10 months, receiving a

total dose of 127.5 to 135 mg DCB.  Hepatic tumors were
found in 4 of 18 mice surviving after 18.5 months (22.2
percent).  A sebaceous gland carcinoma and a lung adenoma
were also seen.
     The Pliss studies show that DCB may possess carcinogenic
activity in both rats and mice.  However, the massive and
apparently acutely-toxic dose levels employed, the uncertain
purity of the commercial product used, the virtual lack
of dose-response data, and the lack of adequate controls
limit the studies' utility for assessing human health hazards.
     Carcinogenicity assays were performed using rats and
mice which received DCB by subcutaneous injection (Pliss,
1959; Pliss, 1963), but are not considered here because
of the irrelevancy of the subcutaneous route of administration
to human exposure.
     Griswold, et al.  (1968) examined the potency of cancer
induction by DCB, benzidine, and other compounds, using
induction of mammary cancer in young female Sprague-Dawley
rats as the major  index.  Forty day-old female Sprague-Dawley
rats were given 30 mg of DCB every 3 days for 30 days by
gavage and were then observed for 9 months-  Under the condi-
tions of this assay, DCB was ineffective as a mammary carcino-
gen but benzidine was highly effec.tive at lower doses.
     Sellakumar, et al.  (1969) maintained male and female
hamsters on a diet containing 0.1 percent  (1000 ppm) of
DCB.  The duration of the study was not specified.  With
30 animals of each sex, no cancer was observed.  However,

at 0.3 percent dietary DCB, 4 transitional cell bladder
carcinomas, some liver tumors and diffuse chronic intrahep-
atic obstructing cholangitus were seen.  At 0.1 percent
in the diet of benzidine, many liver tumors were obtained
but no bladder cancer was found.
     DCB was also found to produce transformation in cultured
rat embryo cells infected with Rauscher leukemia virus  (Freeman,
et al. 1973).   The index of transformation was the development
of macroscopic foci of spindle cells, lacking polar orientation
and contact inhibition.  Cells from typical foci were tumori-
genic when transplanted into newborn Fisher rats.  Transplant-
ability of DCB-transformed cells was not specified.  DCB-
induced transformation was seen at concentration of 5 ppm
in the medium, but not at 1 ppm.  Levels of 10 ppm or higher
were cytotoxic.  The in vitro test system detected transfor-
mation-activity in 6 of 7 aromatic amines characterized
as active ir\ vivo carcinogens, 1 of 2 aromatic amines classed
as weak iri vivo carcinogens, and 0 of 3 aromatic amines
classed as non-carcinogenic in vivo.  DCB was classed by
the authors as a weak carcinogen.
     The history of human industrial experience with DCB
has been summarized and analyzed by Gerarde and Gerarde
(1974) and Rye, et al.  (1970) in the United States; by Maclntyre
(1975) and Gadian  (1975) in Great Britain; and by Akiyama
(1970) in Japan.  The concensus of these authors, achieved
through epidemiological studies, is that there is no evidence
that DCB itself has induced bladder cancer, the characteristic
lesion induced by benzidine, naphthylamine, and other carcino-
genic aromatic amines used in the dye and pigment industry.

The case for DCS carcinogenicity has been made largely on
the basis of its structural similarity to benzidine and
its tumorogenicity in several species of animals  (Maclntyre,
1975).  One problem associated with epidemiological studies
of DCB effects in humans is that the population which has
been exposed only to DCB is small.  Many workers have also
handled benzidine or other carcinogens.  Also, the character-
istic latency period for induction of bladder cancer by
chemicals is quite long, exceeding 16 years for benzidine
(Haley, 1975), and may not have elapsed for many workers.
Finally, most of these studies have focused solely upon
bladder cancer as the disease of interest.  As discussed
below, this approach may be misleading and fallacious in
view of the pattern of DCB carcinogenesis in animals and
the nature of cancer observed in DCB process workers.
     Gadian  (1975) examined the health records of 59 workers
at a dyestuff plant in Great Britain who were exposed from
1953 through 1973 to DCB only and compared them to those
working with both benzidine and DCB, and to unexposed popula-
tions.  This time was justified as the average latency period
for chemically-induced bladder cancer in humans (ca. 18
years).  It was calculated that the DCB process worker was
actually exposed to DCB for a maximum of 10 hours per work
week.  Men whose total DCB exposure was less then 245 hours
(6 months full time work) were excluded from the study,
leaving 35 segregated DCB workers.  These 35 workers, repre-
senting a total of 68,505 hours of DCB exposure, had no
urinary tract tumors, no other tumors, and two deaths from
other causes  (coronary thrombosis, cerebral hemorrhage).

In contrast, among 14 mixed benzidine and DCS workers with
16,200 hours exposure (approximately 60 percent worked with
benzidine, 40 percent worked with DCB), three men developed
tumors of the bladder, and one man developed carcinoma of
the bronchus.  One death from coronary thrombosis occurred.
Since the use of benzidine ceased in 1964, the mixed group
had a longer time to develop tumors than the DCB-segregated
group.  Therefore, the DCB-alone hours worked during the
same period  (1953-1964)  as the mixed group was 31,945 hours.
These results, while admitting that the population studied
was small, were taken as evidence that DCB can be safely
used if the provisions of the Carcinogenic Substances Regula-
tions are observed.
     Maclntyre (1975) also surveyed the health history of
a DCB-utilizing plant in Great Britain.  It was noted that
the vast majority  (209 out of 217) of production and service
workers had received first exposure to DCB less than 20
years before the time of the report, indicating that the
latent period for tumor formation might not have elapsed.
Only 3 of the 217 exposed workers were deceased.  The causes
of death were amytrophic lateral sclerosis (age 55 years,
15 years of DCB exposure, 39 years since first exposed),
carcinoma of the lung (age-61 years, 1 year of DCB exposure,
12 years since first exposed), and pneumonia (age - 70 years,
10 years of DCB exposure, 43 years since first exposed).
Three other employees who had not been exposed to DCB died
of bronchial carcinoma.   All employees exposed to DCB since
1965 have received cytological testing twice yearly, with

all tests proving negative.  A 1974 meeting of occupational
physicians is also cited, stating that in Europe approximately
1000 persons have been exposed to OCB with a zero incidence
of bladder cancer.
     Gerarde and Gerarde  (1974) reported the results of
an epidemiological study of workers exposed to DCB in manufac-
ture and utilization in a plant in the United States.  A
survey of the number of DCB-exposed workers who developed
neoplasms and the type of neoplasm was presented.  These
included lung cancer (2 workers), leukemia-bone marrow (1),
lipoma  (6), rectum-papilloma  (3), sigmoid colon carcinoma
(2), prostate carcinoma (1), breast muscle myoblastoma (1),
skin basal cell epithelioma (1).   A total of 17 workers
of the total of 207 workers surveyed had developed neoplasma.
     The etiology of bladder cancer was discussed and the
data treated, using several epidemiological and statistical
approaches.  According to this approach, if DCB were as
potent as benzidine as a bladder carcinogen, a total of
22 cases of bladder cancer out of 163 DCB production workers
would have been observed, whereas none were seen.  The possible
induction by DCB of tumors at sites other than the bladder
was not considered.

     Based upon existing data, .there is little doubt that
DCS is carcinogenic in several animal species including
rats, mice, hamsters, and dogs.   According to current method-
ology, the experimental evidence serves as an indication
that a potential carcinogenic risk is posed to man.  DCS
induces tumors in a variety of tissues in animals, with
mammary, hematopoietic, and skin (Zymbal gland)  tissue being
the most affected.  Many of the tumors,have been character-
ized as malignant.

                    CRITERION FORMULATION
Existing Guidelines and Standards
     The American Conference of Governmental Industrial
Hygienists (1977) has recommended that no exposure to DCB
by any route should be permitted, because of a demonstrated
high carcinogenic response in animals.  Strict regulations
have recently been- promulgated by the Occupational Safety
and Health Administration to minimize or eliminate occupation-
al exposure to DCB  (CFR,-1977).  To date, no standards have
been placed on permissable levels of DCB in the environment
or in food.
Current Levels of Exposure and Special Groups at Risk
     It is estimated that between 250 and 2500 workers re-
ceive exposure to DCB in the U.S., compared to 62 for benzi-
dine (Fishbein, 1977) .  Given the stringent precautions
which ipust be taken in the manufacture and use of DCB, the
level of exposure may be minimal at present, although no
data is available.  However, past exposure of individuals
working without benefit of protective measures must present
a cause for concern. In addition, the general population
may receive exposure to DCB through contaminated drinking
water or food  (fish), although there is no significant evi-
dence for this at the present.
     Additional groups that may be at risk include worker
in the printing or graphic arts professions handling the
DCB-based azo pigments.  DCB may be present as an impurity
in the pigments, and there is some evidence that DCB may
be metabolically liberated from the azo pigment.  More infor-
mation on the level of exposure to the pigments in needed.

Basis and Derivation of Criterion

     The safe dose of DCB in water was calculated from the

carcinogenicity assays, using a linear, non-threshold mathema-

tic model described in the appendix.  The calculation assumes

a risk of 1 in 100,000 of developing cancer as a result

of daily consumption of 2 liters of water and DCB-contaminat-

ed fish having a bioconcentration factor of 1150.  Although

several carcinogenicity studies are available for use in

calculating a criterion for DCB in drinking water, only

the work of Stula and coworkers (1975, 1978) was considered,

since the studies by Pliss (1959, 1963) lack appropriate

control data.  Mor;e specifically, the data on induction

of papillary transitional cell carcinomas of the urinary

bladder and hepatic carcinomas in female beagle dogs  (Stula,

et al. 1978) were chosen as a base for the calculations.

Based on these data, a DCB criterion of 1.69 x 10   jug/1

is judged to be adequate to protect the population consuming

the water.  This dose is low from an occupational viewpoint

and should justify efforts to eliminate exposure of workers

to DCB.           ;

     Under the Consent Decree in NRDC vs. Train, criteria

are to state "recommended maximum permissible concentrations

(including where appropriate, zero) consistent with the

protection of aquatic organisms, human health, and recrea-

tional activities."  DCB is suspected of being a human car-

cinogen.  Because there is no recognized safe concentration

for a human carciriogen, the recommended concentration of

DCB in water for maximum protection of human health is zero.

         Because attaining a zero concentration level may be

    infeasible in some cases and in order to assist the Agency

    and States in the possible future development of water qual-

    ity regulations, the concentrations of DCB corresponding

    to several incremental lifetime cancer risk levels have

    been estimated.  A cancer risk level provides an estimate

    of the additional incidence of cancer that may be expected

    in an exposed population.  A risk of 10"  for example, indi-

    cates a probability of one additional case of cancer for

    every 100,000 people exposed, a risk of 10   indicates one

    additional case of cancer for every million people exposed,

    and so forth.

         In the Federal Register notice of availability of draft

    ambient water quality criteria, EPA stated that it is consid-

    ering setting criteria at an interim target risk level of

    10~5, 10~6 or 10~7 :as shown in the table below.
Exposure Assumptions           Risk Levels and Corresponding Criteria^ '

                                  10f 7         1~6        10~5

2 liters of drinking water     0   1.69 x 10~4  1.69 x 10~3 1.69 x 10~2
and consumption of 18.7
grams of fish and shellfish  (2)

                               0   1,
                                     jug/1         jug/1        jug/1
Consumption of fish            0   1.85 x 10 4  1.85 x 10~3 1.85 x 10 2
shellfish only.

(1)  Calculated by applying a modified "one hit" extrapola-
tion model described in the FR 15926, 1979.  Appropriate
bioassay data used in the calculation of the model are pre-
sented in Appendix 1.  Since the extrapolation model is
linear to low doses, the additional lifetime risk is directly
proportional to the water concentration.  Therefore, water
concentrations corresponding to other risk levels can be
derived by multiplying or dividing one of the risk levels
and corresponding water concentrations shown in the table
by factors such as 10, 100, 1,000, and so forth.
(2)  Ninety-one percent of DCB exposure results from the
consumption of aquatic organisms which exhibit an average
bioconcentration potential of 1150 fold.  The remaining
9 percent of DCB exposure results from drinking water.
     Concentration levels were derived assuming a lifetime
exposure to various amounts of DCB,  (1) occurring from the
consumption of both drinking water and aquatic life grown
in water containing the corresponding DCB concentrations
and,  (2) occurring solely from consumption of aquatic life
grown in the waters containing the corresponding DCB concen-
     Although total exposure information for DCB is discussed
and an estimate of the contributions from other sources
of exposure can be made, this data will not be factored
into the ambient water quality criteria formulation because
of the tenuous estimates.  The criteria presented, therefore,
assume an incremental risk from ambient water exposure only.

                          APPENDIX I
            Summary and Conclusions Regarding the
      Carcinogenicity of  3,3'-Dichlorobenzidine  (DCS)*
     3,3'-Dichlorobenzidine (DCB) is used as an intermediate
in the synthesis of dyes and pigments.  It is structurally
related to carcinogenic aromatic amines, which have been
used in the dye and pigment industries.
     Five epidemiological studies of employees handling
DCB in chemical plants in the United States, Great Britain,
and Japan have provided no evidence of DCB-induced cancers.
However, investigative problems associated with these studies,
such as too short a follow-up time and small sample size,
make them unreliable as the sole basis for making conclusions
about human cancer risks from DCB.
     DCB has induced carcinomas in three species of experimen-
tal animals receiving oral doses of the chemical.  Dogs
(female) developed papillary transitional cell carcinomas
of the urinary bladder and hepatocellular carcinomas.  Ham-
sters developed transitional cell bladder carcinomas, liver
cell and cholangiomatous tumors.  Rats developed mammary
adenocarcinomas (male and female), granulocytic leukemia
(males), and Zymbal gland carcinomas  (males).
     Two studies of the mutagenicity of DCB  showed that
it was mutagenic in two Salmonella typhimurium tester strains
(TA 1538, TA 98) in the presence and absence of an S-9 liver
enzyme system.  DCB also transformed cultured rat embryo
cells and the  transformed cells were tumorigenic when trans-
planted into newborn rats.


     The carcinogenic, mutagenic, and transforming activities

of DCB in laboratory organisms and its chemical similarity

to benzidine, a human bladder carcinogen, are strong evidence

that it is likely to be a human carcinogen.

     The water quality criterion for DCB is based on the

induction of papillary transitional cell carcinomas of the

urinary bladder and hepatic carcinomas in female beagle

dogs, given an,oral dose of 100 mg 3,3'-dichloroben2idine,

3 times per week for 6 weeks, then 5 times per week continuously

for up to 7.1 years (Stula, et al. 1978). Dose-response

data for dogs were selected because dogs developed urinary

bladder tumors, as do humans, when exposed to certain aromatic

amines.  The concentration of DCB in water, calculated to

                                       c               2
keep the lifetime cancer risk below 10  , is 1.69 x 10

micrograms per liter.   .
     *This summary has been prepared and approved by the

      Carcinogens Assessment Group, EPA, on June 15, 1979.

                  Summary of Pertinent Data

     The water quality criterion for DCB is based on the

induction of papillary transitional cell carcinomas of the

urinary bladder and hepatic carcinomas in female beagle

dogs, given an oral dose of 100 mg DCS, three times per

week for six weeks, then five times per week continuously

for periods up to 7.1 years (Stula, et al. 1978).  The inci-

dence of urinary bladder carcinomas observed in DCB-treated

dogs was 5/5 as compared to 0/6 in the control group.  The

incidences'of hepatic carcinomas were 4/5 and 0/6 in DCB-

treated and control groups, respectively.  The criterion

was calculated from the following parameters:

     n.   = 4.5* (urinary bladder carcinomas)
     Ntu = 5
     nfc^ = 4  (hepatic carcinoma)
     N^ = 5               d (timeweighted average)
     n   = 0                  concentration) =7.36 mg/kg/day
     >C  = 6               F = 0.0187 kg
     Le  = 7.1 yrs.        R = 1150
     le  = 7.1 yrs.        W = 11.391 kg
     L   = 8.65 yrs.

Based on these parameters, the one-hit slope (BH) is 1.036

(mg/kg/day)    for urinary bladder carcinomas and 0.724 (mg/kg/day)~

for hepatic carcinomas.  The resulting water concentration

for DCB, calculated to keep the individual lifetime cancer
risk below 10  , is 1.69 x 10   micrograms per liter.
    *This underestimate of the true number, based on Burkson's

     correction factor, was chosen in order to obtain a

     finite mathematical estimate.


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