ATTACHMENT 7
STATEMENT OF BASIS AND PURPOSE
FOR AN AMENDMENT TO THE
NATIONAL INTERIM PRIMARY^DRINKING WATER REGULATIONS^
ON TRIHALOMETHANES
'AUGUST 1979
OFFICE OF DRINKING WATER
CRITERIA AND STANDARDS DIVISION
ENVIRONMENTAL PROTECTION AGNECY
WASHINGTON, D.C. 20460
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TABLE OF CONTENTS
PAGE
I. SUMMARY 1
II. INTRODUCTION 5
III. THE ROLE OF CHLORINE AND OTHER DISINFECTANTS 10
IV. SOURCES OF TRIHALOMETHANES EXPOSURE 17
V. METABOLISM 22
VI. ACUTE AND CHRONIC HEALTH EFFECTS IN ANIMALS 26
A. Hepatotoxicity
B. Nephrotoxicity
C. Central Nervous System
D. Teratogenicity
E. Mutagenicity
F. Carcinogenicity
VII. HUMAN HEALTH EFFECTS 35
A. NAS Principles of Toxicologioal Evaluation
B. Epidemiologic Studies
•
VIII. MECHANISMS OF TOXICITY 56
IX. RISK ASSESSMENT 60
X. MAXIMUM CONTAMINANT LEVELS 66
XI. REFERENCES 70
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I. Summary
The trihalomethanes (THMs) are a family of organic
compounds, named as derivatives of methane, where three of
the four hydrogen atoms are substituted by a halogen atom.
Although halogens can include fluorine, chlorine, bromine
and iodine, only chlorine and bromine substituents are now
considered for the purpose of this regulation. THMs in
drinking water are produced by the action of the chlorine
added for disinfection or oxidation, with the naturally
occurring organic precursors (e.g., humic or fulvic acids)
commonly found in source waters.
THMs are commonly found in drinking water supplies
throughout the United States. Chloroform has been found at
concentrations ranging from 0.001-0.540 mg/1 andMTTHM)
A
potential concentrations as high as 0.784 mg/1 have been
detected. The concentrations of TTHM increase when raw
water supplies are treated with chlorine for disinfection
and other purposes. TTHM concentrations are indicative of
the presence of other halogenated and oxidized organic
chemicals that are produced in water during chlorination.
People are also exposed to chloroform in the air they
breathe and the food they eat. Analyses of the relative
contribution of chloroform in drinking water, air and food
exposures assumed various levels of exposure based on
monitoring studies. Drinking water may contribute from zero
to more than 90% of the total body burden.
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Chloroform has been shown to be rapidly absorbed on oral
and intraperitoneal administration and subsequently metabol-
ized to carbon dioxide, chloride ion, phosgene, and other
unidentified metabolites. The metabolic profile of chloro-
form in animal species such as mice, rats, and monkeys is
indicated in Table 4 and is qualitatively similar to that in
man.
Mammalian responses to chloroform exposure include:
central nervous system depression, hepatotoxicity, nephro-
toxicity, teratogenicity, and carcinogenicity. These
responses are discernible in mammals after oral and
inhalation exposures to high levels of chloroform ranging
from 30-350 mg/kg; the intensity of response is dependent
upon the dose. Although less toxicological information is
available for the brominated THMs, mutagenicity and
carcinogenicity have been detected in some test systems.
Physiological chemical activity should be greater for the
brominated THMs than for chloroform.
Although short-term toxic responses to THMs in drinking
water are not documented, the potential effects of chronic
exposures to THMs should be a matter of concern. Prolonged
administration of chloroform at relatively high dose levels
(100-138 mg/kg) to rats and mice, manifested oncogenic
effects. Oncogenic effects were not observed at the lowest
dose level (17 mg/kg) in three experiments. Since methods
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do not now exist to establish a threshold no effect level of
exposure to carcinogens, the preceeding data do not imply
that a "safe" level of exposure can be established for
humans.
Human epidemiological evidence is inconclusive, although
positive correlations with some sites have been found in
several studies. There have been 18 retrospective studies
shown in Table 7 that have investigated some aspect of a
relationship between cancer mortality or morbidity and
drinking water variables. Due to various limitations in the
epidemiological methods, in the water quality data, and
problems with the individual studies, the present evidence
cannot lead to a firm conclusion that there is an associaton
between contaminants in drinking water and cancer mortali-
ty/morbidity. Causal relationships cannot be proven on the
basis of results from epidemiological studies. The evidence
from these studies thus far is incomplete and the trends and
patterns of association have not been fully developed. When
viewed collectively, however, the epidemiological studies
provide sufficient evidence for maintaining the hypotheses
that there may be a potential health risk, and that the
positive correlations may be reflecting a causal association
between constituents of drinking water and cancer mortality.
Preliminary risk assessments made by the Science
Advisory Board (SAB), the National Academy of Sciences
(MAS), Tardiff, and EPA's Carcinogen Assessment Group (CAG)
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using different models have estimated the incremental risks
associated with the exposure from chloroform in drinking
water. The exposure to THMs from air and food have not been
included in these computations. The risk estimates associ-
ated with the MCL at the 0.10 mg/1 level are essentially the
same from the MAS and GAG computations (3.4 x 1Q-1* and 4 x
10'^) assuming two liters of water at 0.10 mg/1 chloroform
consumed daily for 70 years.
On the basis of the available toxicological data
summarized in the following report, chloroform has been
shown to be a carcinogen in rodents (mice and--rats) at high
dose levels. Since its metabolic pattern in animals is
qualitatively similar to that in man, it should be suspected
of being a human carcinogen. Epidemiological studies also
suggest a human risk. Therefore, because a potential human
health risk does exist, levels of chloroform in drinking
water should be reduced as much as is technologically and
economically feasible, using methods that will not compromise
protection from waterborne infectious disease transmission.
Although documentation of their toxicity is not so well
established, other THMs should be suspected of posing simi-
lar risks. Because the treatment process that can reduce
drinking water levels of chloroform have about the same
effectiveness in reducing levels of the other THMs, the
proposed regulation is addressed to these substances, as
well.
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II. INTRODUCTION
The extent and significance of organic chemical
contamination of drinking water or drinking water sources
first came to public attention in 1972, when a report,
"Industrial Pollution of the Lower Mississippi River in
Louisiana" was published (EPA, 1972). While this report did
not include quantification of the pollutants found, and was
directed toward locating industrial discharges responsible
for the pollution, the report did include analyses of fin-
ished (treated) drinking water and provided evidence of the
presence of THMs. Subsequently, a more thorough examination
of finished drinking water in the New Orleans area was
carried out, using the most sophisticated analytical methods
available (EPA, 1974). This latter study confirmed the
presence of THMs and many other organic chemicals in
finished drinking water, and furthermore it demonstrated
that one of them, chloroform, was present in high relative
concentrations.
The findings in New Orleans promoted other studies,
primarily for the purpose of determining how widespread and
serious the organic chemical contamination of drinking water
was. Impetus was added by the passage of the Safe Drinking
Water Act (PL 93-523), which directed the EPA to conduct a
comprehensive study of public water supplies and drinking
water sources to determine the nature, extent, sources,
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and means of control of contamination by substances
suspected of being carcinogenic. The National Organics
Reconnaissance Survey of Halogenated Organics (NORS)
(Symons, et al., 1975), or "80 City Study", was aimed
primarily at determining the extent of the presence of four
THMs, chloroform, bromodichloromethane, dibromochloromethane
and bromoform, along with carbon tetrachloride and 1,2-dich-
loroethane, and at determining what effect raw water source
and water treatment practices had on the formation of these
compounds (Table 1). The presence of THMs in finished
drinking water was confirmed, and some trend relating non-
volatile total organic carbon (NVTOC) of the raw-water and
the total trihalomethane (TTHM) was postulated. Chloroform
occurred invariably in water which had been chlorinated,
while it was absent or present at much lower concentrations
in the raw water. Water samples were collected at the
treatment plant in winter and iced for shipment but not
dechlorinated. Thus, those values might approximate minima
for human exposure in the areas selected. Of the various
THMs, chloroform was found at the highest concentrations
(averaging approximately 75 percent of the TTHM), with
progressively less bromodichloromethane, dibromochlorometh-
ane and bromoform being detected.. In some cases chloroform
was found at concentrations greater than 0.300 mg/1; (the
highest value found was 0.540 mg/1). Carbon tetrachloride
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and 1,2-dichloroethane were found at very low concentra-
tions. The concentration of these two components did not
increase after chlorination; therefore, it can be assumed
that these compounds are not related to the chlorination
process.
A Joint Federal/State Survey of Organics and Inorganics
•
in 83 Selected Drinking Water Supplies, carried out by EPA's
Region V (Chicago) provided additional evidence of the ubi-
quitous nature of chloroform and other THMs in chlorinated
drinking water (EPA, 1975). Two conclusions reached in that
study were that raw water relatively free of organic matter
results in finished water that is relatively free of chloro-
form and related halogenated compounds, and that there is a
correlation in some instances between the concentrations of
chloroform, bromodichloromethane, dibromochloromethane and
bromoform in finished water and the amount of organic matter
f
found in raw water.
The National Organics Monitoring Survey (MOMS), directed
by Section 141.40 of the National Interim Primary Drinking
Water Regulations (40 FR 59574, December 24, 1975), was
aimed not only at determining the presence of THMs in addi-
tional water supplies, but also at determining the seasonal
variations in concentration of these substances.
The NOMS sampling included 113 public water systems
designated by the Administrator, and also included analyses
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for approximately 20 specific synthetic organic chemicals
deemed to be candidates of particular concern as well as
analyses of several surrogate group chemical parameters
which are indicators of the total amount of organic con-
tamination. Three phases of this study were completed and
the mean, minimum, and maximum values of chloroform and THMs
in drinking water are reported in Table 1. Phase I analyses
in the NOMS were conducted similarly to the NORS. Phase II
analyses were performed after the THM-producing reactions
were allowed to run to completion. Phase III analyses were
conducted on both dechlorinated samples and on samples that
were allowed to run to completion (terminal). Again chloro-
form was found at the highest concentrations in most cases,
however, in a few cases bromoform was found to be the high-
est concentration of the THMs (0.280 mg/1). The mean con-
centrations of chloroform were 0.043 mg/1, 0.083 mg/1, 0.035
. mg/1, and 0.069 mg/1 for Phase I, II, III (dechlorinated)
and III (terminal), respectively; the mean concentrations
for TTHMs were 0.068 mg/1, 0.117 mg/1, 0.053 mg/1 and 0.100
mg/1 for Phase I, II, III (dechlorinated) and III (termi-
nal), respectively.
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III. The Role of Chlorine and Other Disinfectants
All available evidence indicates that chlorination of
drinking water containing naturally occurring organic
chemicals is the major factor in the formation of halogena-
ted organic chemicals, particularly the THMs in finished
drinking water. Chlorinated organic compounds, however, can
also be introduced into drinking water from industrial
outfalls, urban and rural runoff, rainfall, through polluted
air, or from the chlorination in sewage and industrial
wastewater.
Several studies in addition to those mentioned above,
have demonstrated increased THM concentrations in drinking
water. Work by J. J. Rook (1974) in the Netherlands, and
the studies by Bellar, Lichtenberg and Kroner (1974), showed
that chloroform and other halogenated methanes are formed
during the water chlorination process. It should be noted
that these findings came as a result of the development and
application of more sensitive and refined analytical techni-
ques. Recent work by Rook (1974, 1977) has provided some
insight into the organic precursors which might be responsi-
ble for the formation of the THMs. Studies by Sontheimer
and Kuhn (1977) indicate that the THMs may represent only a
portion of the total halogenated products of chlorination of
water. Bunn et al. (1975), have demonstrated that hypo-
chlorite in the presence of bromide and iodide ions but not
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fluoride will react with natural organic matter to produce
all ten possible trihalogenated methanes.
It can be concluded from the above studies and others
that the THMs occur in chlorinated drinking waters, and that
the concentrations of the various THMs are dependent on the
type.and quality of organic precursor substances, the amount
of chlorine used, and the presence of other halogen ions as
well as contact time, temperature and pH.
A number of methods are available for reducing levels of
THMs in drinking water. These options include- modifications
of current treatment practices, such as moving the point of
chlorination, the use of alternative disinfectants such as
chlorine dioxide, chloramines,or ozone, and various methods
that will reduce organic precursor concentrations such as
use of adsorbents like granular activated carbon (GAG).
Two chemicals often mentioned as alternative disinfec-
*
tants, chlorine dioxide and ozone, are both well known as
effective disinfectants and chemical oxidants, and some
history of their practical use in water treatment has been
accumulated particularly in Europe, but also in the United
States.
Chlorine dioxide is usually prepared at the water plant
by the reaction of chlorine (either as gas or as sodium
hypochlorite) with sodium chlorite. Unless an excess of
chlorine is used, there will be unreacted sodium chlorite
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left over from the reaction. When chlorine dioxide reacts
with organic matter in the water, one of the reaction
products is the chlorite ion. Thus, whenever chlorine
dioxide is used to treat water, the presence of chlorite ion
in the treated water can be expected.
EPA is studying the health effects of chlorine dioxide
in water, utilizing several animal species as well as human
volunteers. Studies of the toxicology of chlorine dioxide
and chlorite ion in drinking water reveal considerable
variations. These compounds have been reported to affect
the hematopoietic systems such as oxidative changes in hemo-
globins and and hemolysis of red blood cells. Other bioef-
fects observed include gastrointestinal disturbances. The
preliminary results indicate species variability in biologi-
cal manifestations. Cats and African green monkeys appear
to lie at the extreme ends of the- spectrum from among the
species studied; cats are very sensitive to hematopoietic
effects whereas monkeys were apparently insensitive even at
levels as high as 400 mg/1 (Bull, 1979). An upper limit for
chlorine dioxide by-product exposure is being considered
primarily because of the lack of data concerning the safety
of this material, and particularly its decomposition
products, at higher concentrations (Musil et al., 1963 and
Fridyland and Kagan, 1971. Studies with cats have shown
that chlorite, which is oxidant that can cause anemias, has
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a deleterious effect on red blood cell survival rate at
chlorine dioxide concentrations above 10 mg/1. Preliminary
studies in a small human population did not demonstrate
substantial blood chemistry changes, ex-cept possibly in one
person known to be deficient in glucose-6-phosphotase
dehydrogenase. Lack of sufficient health effects data on
human toxicity for C102 and its by-products prevents
establishment of an MCL at this time, however, work in
progress is expected to provide much additional information
within the coming year. In the meantime, EPA recommends
that monitoring be conducted when chlorine dioxide is used,
and that residual oxidant should not exceed 0.5 mg/1 as
C102.
A preliminary study concerning ozonation of 29 organic
compounds potentially present in water supply sources indi-
cated the formation of a number of products (Cotruvo, Sim-
mon, Spanggord, 1976, 1977). These reaction mixtures were
assayed for mutagenic activity employing 1) five strains of
Salmonella typhimurium (Ames Salmonella/microsome assay);
and 2) mitotic recombination in the yeast Saccharomyces
cerevisiae D3. After very extensive ozonation in water some
of the organic compounds exhibited mutagenic activity in
these systems. Similar more recent studies under extreme
conditions with chlorine dioxide by-products did not exhibit
mutagenic activity (SRI Report).
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Combining ammonia with chlorine to form chloramines has been
called the chloramine process, chloramination, and combined
residual chlorination. The products of this process are
monochloramine, dichloramine or trichloramines (nitrogen
trichloride) depending on the pH and the chlorine to ammonia
ratio. The production of the latter species may contribute
to taste and odor problems in the finished water; however,
chloramination does not reduce the formation of THMs.
Based on the results of numerous investigations, the
comparative disinfectant efficiency of chloramines ranks
last when compared to ozone, chlorine dioxide, hypochlorous
acid acid (HOCL), and hypochlorite ion (OC1-) (NAS, 1977,
1979). Early studies by Butterfield and Waties (1944, 1946,
1948) demonstrated that chloramines required approximately a
100-fold increase in contact time to inactivate coliform
bacteria and enteric pathogens as compared to free available
chlorine at pH 9.5. This work was later confirmed by Kabler
(1953) and by Clarke et al., (1962).
Results with cysts of Entamoeba histolytica and viruses
also confirm the decreased effectiveness of chloramines as
disinfectants. Studies by Fair, et al., (1947) showed that
additional dichloramine is about 60 percent and monochloro-
amine about 22 percent as effective as hypochlorous acid at
pH 4.5 against cysts of E. histolytica. Kelly and Sanderson
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(1960) found that chloramines in the concentration of 1 mg/1
at 25° C required 3 hours at pH 6, or 6 to 8 hours at pH 10
to achieve 99.7 percent inactivation of polio virus. With
0.5 mg/1 free chlorine at pH 7.8, by comparison, inactiva-
tion of 99.99 percent of polio virus can be achieved in
approximately 15 minutes (Liu and McGrowan, 1973). Chlora-
mine treatment finds its widest application in maintenance
of chlorine residuals in the distributing systems. The
human health effects of consuming water treated with
chloramine have not been studied in detail.
Although all of these disinfectants can reduce THM for-
mation, questions have been raised on both their toxicity
and the toxicity of their by-products. Studies are underway
to clarify these matters, and could result in the designa-
tion of maximum permissible levels for certain disinfectants
applied to drinking water.
The use of adsorbents for THM removal has also intro-
duced some unknown factors. Assuming that the adsorption
process is effective for its intended purpose, there is the
possibility that a breakthrough of some of the adsorbed
chemicals may occur, that these substances will be adsorbed
and subsequently slough off to produce intermittent
contamination, or that bacteria and/or toxins will be added
to the water from growth on the adsorbent. All of these
potential effects are controllable in practice, and EPA
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encourages the use of GAG to purify contaminated waters and
to control THM precursors.
Thus, THM concentrations should be reduced, but without
compromising public health from either increased risk of
infectious disease transmission or from the chemicals that
are used. Outbreaks of infectious waterborne disease have
been noted when chlorination systems have been improperly
operated. The alternative control methods outlined prev-
iously are effective, and are also being studied for their
possible side effects. As soon as data become" available,
EPA will make specific recommendations regarding their use.
At the present time, the best approach to reduce THMs in
finished water is to reduce precursors prior'to chlorina-
tion, such as with GAG. This approach has the benefit of
reducing the concentration of many other organic chemicals
in the water as well as to the precursors to THM and other
chlorinated organics. Thus, once the organic chemical
concentrations in the water have been reduced, the chemical
demand for applied disinfectant will be reduced. Thus,
human exposure to all disinfectant chemicals and their de-
gradation products and by-products will be minimized. This
is the intent of the regulation controlling THMs.
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IV. Sources of Trihalomethane Exposure
McConnell et al. (1975), have reported that chloroform
occurs in many common foods and that while some halogenated
compounds in food may result from manufacturing, canning and
pest control practices, chloroform may be introduced as the
result of geochemical processes. Chlorinated compounds are
the halogenated species most prevalent in food, but at least
one food, Limu Kohu, a seaweed or algae eaten in Hawaii,
contains an essential oil which is composed largely of
bromoform (Burreson, et al. 1975).
Chloroform was widely used as an anesthetic in the past,
and, until recently, was a common ingredient in dentifrices
and cough preparations. The Food and Drug Administration
has taken action to halt the use of chloroform in drug
products, cosmetic products, and food-contact articles (41
FR 145026, April 9, 1976). EPA has issued a notice of
*
"rebuttable presumption against registration" of chloroform-
containing pesticides (41 FR 14588, April 6, 1976). Thus,
in addition to drinking water, exposure to some or all of
the THMs is complicated by other environmental sources,
however, exposure from some of these sources is being
reduced.
The relative human chloroform exposures can be estimated
for three major sources of human exposure: atmosphere,
drinking water, and the food supply. The uptake calcula-
tions are based on the fluid intake, respiratory volume, and
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food consumption data for "reference man" as compiled by the
International Commission on Radiological Protection. The
combined uptake for adults from all three sources was
derived by multiplying estimated exposure levels by the
estimated annual intakes and combining the results [ODW
protocol].
Human uptake of chloroform from air, food and drinking
water is given in Table 2. Chloroform and TTHM uptake from
drinking water was estimated by multiplying the chloroform
and THM concentrations from NOMS data (Table -10 by the
average consumption of 2 liters of water per day for the 70
kg adult male, by 365. One hundred per cent absorption of
the amount of chloroform in drinking water is assumed for
these calculations. The total chloroform uptake from water
was estimated as a mean value of 64 mg per year. The
maximum uptake value may be 394 mg per year.
To determine uptake of chloroform from foods, the
concentration of chloroform in each food item in North
American diets was multiplied by the average annual consump-
tion of that food item by adults in the United States (NAS,
1977), and the results were combined again; one hundred per
cent absorption of ingested chloroform was assumed. A
calculated maximum value of about 16 mg of chloroform uptake
per year from total food and a mean value of 9 mg based on
ODW assumptions was obtained.
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Table 2. Human Uptake of Chloroform and Trihalomethanes from
Drinking Water, Food and Air
Exposure Levels mg/year
Chemical Drinking Water Food Air*
Mean (Range) Mean (Range) Mean (Range)
Chloroform 64(0.73-343) . 9(2-15.97) 20(0.41-204)
Trihalomethanes 85 (0.73 - 572)
* Calculated from data supplied by Strategies and Air Standards
Division, Office of Air Quality Planning and Standards. Environmental
Protection Agency, Research Triangle Park. The air samples were
collected both from the rural and industrial areas during the years.
1974 - 76.» The mean value was derived from the concentrations
obtained from urban industrialized areas, the minimum value from
the rural area and the maximum value from an urban industrialized
area.
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The calculation for the uptake of chloroform by humans
from ambient air was based upon the assumptions that 63
percent of inhaled chloroform is absorbed, (NAS, 1977); the
volume of air inhaled by an average adult is 8.1 x 10^
liters per year; and 0.02 and 10 ppb (by volume) are the
respective minimum and maximum chloroform concentrations in
urban air. The minimum and maximum values for the annual
uptake of chloroform by an adult were estimated at 0.41 and
204 mg, respectively. Assuming minimum exposures from all
sources, the atmosphere contributes 12 percent, of the total
chloroform, the drinking water contributes 23 percent, and
food is most significant (65/6). Assuming maximum exposures
from all sources, drinking water is the major contributor at
61 percent, with air at 36 percent. Thus, the relative
contribution of drinking water to the total body burden of
chloroform may range from a moderate to a maximum contribu-
tor as the annual exposure from water ranges from nil to 394
mg/year, and from 204 to 0.73 mg/year in ambient air (Table
3).
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Table 3. Uptake of Chloroform for the Adult IJuman from Air,
Water, and Food
Source
Atmosphere
Water
Food Supply
Total
,
Atmosphere
Water
Food Supply
Total
Atmosphere
Water
Food Supply
•*
Total
Adult
mg/yr
Maximum Conditions
204
343
16
563
Minimum Conditions
0.41
0.73
2.00
3.14
Max-Water Min-Air
0.41
343.00
9.00
352.41
Percent
uptake
36
61
3 .
100-. 00
13
23
64
100.00
1
97
2
. 100. 00
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V. Metabolism
Several reports (Brown, et al., 1974; Labigne &
Marchand, 1971; Fry et al., 1972; Paul and Rubenstein, 1963;
Taylor et al. , 1974) have indicated that, chloroform is
rapidly absorbed on oral and intraperitoneal administration
and subsequently metabolized to carbon dioxide and unidenti-
fied metabolites in urine. Species variation in the meta-
bolism of chloroform has been summarized in Table 4. It is
noteworthy that the mouse, a species which shows greater
sensitivity to the oncogenic effect of chloroform (Eschen-
brenner & Miller, 1945; Brown et al. 1974) metabolized
chloroform extensively to carbon dioxide (80£) and unidenti-
fied metabolites (3$) from an oral dose of 60'mg/kg. Rats
also metabolize chloroform to carbon dioxide but to a lesser
extent (66$). In another report, Paul and Rubinstein (1963)
recovered 4 percent carbon dioxide after administering 1484
mg/kg chloroform intradoudenally to rats. The discrepancy
in these two results may be dose related.
Dose related differences in the metabolism of compounds
are known and have recently been reported for the carcinogen
vinyl chloride. Squirrel monkeys, when given 60 mg/kg of
chloroform orally, excreted 97 percent of the dose, with 17
percent as carbon dioxide and 78 percent as chloroform.
Fry, et al. (1972), recovered unmetabolized chloroform
ranging from 17.8-66.6 percent of a 500 mg dose
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Tabld 4. Disposition of Chloroform - Species Variation
ANIMAL
SPECIES
MOUSE
BAT
BAT
BAT
MONKEY
SEX STRAIN
H CBA
CF/LP
C57
M Sprague
Dawley
M Sprague
Dawley
M Squirrel
DOSE
rag/ Jcg
60 po
• 60 po
1484 id
4710 ip
60 po
CHC13
6
20
70
78
METABOLISM (PERCENT)
URINE TOTAL
002 FECES EXCRETION
80 3 93*
66 7 93
0.39
17 2 97
• REFERENCES
Brown et al (1974)
Brown et al 1974
Paul 6 Rubstein (1963) ^
Brown et al (1974)
•Includes radioactivity in carcas,
Po = Orally
id - intradeudenally
ip a intraperitoneal
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of chloroform given to human volunteers during an 8 hour
time period (equivalent to about 7 mg/kg). Since the
metabolism of chemicals is also dependent on age and sex,
the widespread variation in the quantitative disposition of
chloroform in human subjects may be due to the experimental
protocols wherein subjects ranging from 18-50 years of age
were used. Individual variability in the non-homogenous
human population is a major factor.
Metabolic similarities between carbon tetrachloride and
chloroform include the appearance of halide ions in urine
and carbon dioxide in breath. A' related chemical, carbon
tetrachloride, is a common contaminant of the chlorine used
in water disinfection. Carbon tetrachloride also is
metabolized to chloroform in trace amounts, which may in
turn, be biotransformed to carbon dioxide. Both chloroform
and carbon tetrachloride are proven animal carcinogens (see
below). However, this is mentioned because of possible
metabolic production of proximal carcinogens. Toxicity of
carbon tetrachloride, however, has been attributed to a free
radical (CCl^) which is postulated as a metabolic
intermediate. Chloroform appears to be metabolized to form
phosgene (Krishna, 1979).
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Many carcinogens have been reported to form complexes
with proteins, DNA and RNA (Miller & Miller, 1966). In the
case of chloroform, Ilett et al., (1973).reported covalent
bonding of chloroform metabolite(s) to tissue macromolecules
in mice. The covalent bonding increased or decreased when
the animals were pretreated with phenobarbital or piperonyl
butoxide, agents which stimulate or inhibit the metabolism
of foreign compounds by mixed function oxidase enzymes.
This is suggestive of the involvement of chlor'bform
metabolism in these processes. These results may be
interpreted to mean that the potency of an ingested chemical
will be dependent upon its rate of metabolism'to the active
form.
Information regarding the metabolism of bromoform and
other haloforms is not available. However, the structural
similarities of these haloforms with chloroform indicate
that they should also be absorbed by the oral and inhalation
routes of exposure and then metabolized into carbon dioxide
and halide ions. Related halogenated hydrocarbons of the
dihalomethane series (e.g., dichloromethane, dibromoraethane
and bromochloromethane) have been reported (Kubic et al.
1974) to be metabolized to carbon monoxide; the rate of
metabolism of dibromomethane was higher than that of the
dichloromethane.
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VI. Acute and Chronic Health Effects in Animals
Mammalian responses to chloroform include effects on:
the central nervous system, hepatotoxicity, nephrotoxicity,
teratogenicity, and carcinogenicity. Reported oral 1059
values are as follows: for rats, 300 mg/kg (DHEW, 1978); and
for mice, 705 mg/kg (Plaa, et al., 1958).
Jones, et al. (1958), reported the effect of various
oral doses of chloroform on mice 72 hours after exposure:
35 mg/kg — threshold hepatotoxic effect -" minimal
midzonal fatty changes
70 mg/kg — minimal hepatic central fatty infiltration
140 mg/kg — massive hepatic fatty infiltration
350 mg/kg — hepatic centrilobular necrosis
1100 mg/kg -- minimum lethal dose
Acute effects of exposure to chloroform and bromoform
vary among species. Reported lethal doses for chloroform
and bromoform are:
Species Subcutaneous Lethal Dose Values in mg/kg
Mouse LDcQ 704 (Chloroform)
1820 (Bromoform)
Rabbit LD5Q 800 (Chloroform)
410 (Bromoform)
Data on the acute toxicity of dibromochloromethane and
dichlorobromomethane are not available.
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28
Hasegawa (1910) reported dizziness and light intoxication
during 20-minute exposures to chloroform concentrations of
4300-5100 ppm. Repeated exposures up to six days to concen-
trations as low as 920 ppm for 7 minutes resulted in symp-
toms of central nervous system depression (Lehman & Schmidt-
Kehn, 1936). Additional important information has been
submitted to EPA and is discussed below.
Effects of acute and subchronic chloroform exposure on
cholinergic parameters in mouse brain were studied by Vocci,
et al., (1977). Male Swiss Webster ICR mice were gavaged
with single doses of chloroform (30 and 300 mg/kg) -and
sacrificed 15 minutes after administration of chloroform.
In another experiment, the mice were gavaged with 14 or 90
daily doses of chloroform (3 or 30 mg/kg) and sacrificed 18
hours after the last administration. Neither of the above
dosage regimens had any effect on in vitro [3n] choline
uptake in synaptosomes. In another study (ibid) of biosyn-
thesis of acetylcholine in mouse brain, chloroform (30
mg/kg) significantly decreased the [3n] acetylcholine
synthesis (5756 of control). Administration of chloroform (3
mg/kg) for 14 days produced a reduction in [3n] acetylcho-
line (5756 of control) (Vocci, Personal Communication, April
1979).
Chloroform, dichlorobromomethane, chlorodibromomethane
and bromoform, at concentrations of 8 x 10~^ M did not
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29
alter the update of norepinephrine or dopamine into brain
synaptosomes in vitro (Vocci, Personal Communication, April
1979).
D. Teratogenicity
Teratogenic responses to oral dosing of animals with
chloroform were investigated. Rats and rabbits were
administered chloroform at 126 and 50 rag/kg respectively.
No significant fetal deformities were observed (Thompson et
al. (1973). Inhalation of chloroform by Sprague Dawley rats
at 30, 100 and 300 ppm for 7 hours a day, on days 6 through
15 of gestation revealed significant fetal abnormalities
including: acaudia, imperforate anus, subcutaneous edema,
missing ribs and delayed skull ossification (Schwetz et al.
1974).
In an attempt to explain reproductive failure in
laboratory animals, i.e., mice and rabbits, McKinney et al.
(1976) conducted a study using CD-1 mice wherein groups of
mice were given tap water and purified tap water (passed
through a Corning 3508 ORC and a Corning 3508 B demineraliz-
er), respectively. Analysis indicated reduced amounts of
chlorinated compounds in the purified water. The study
could not relate chloroform and other chlorinated organics
in tap water to reproductive failures in laboratory animals,
since the concentrations of chlorinated organics in water
were lowest in those months that reproductive failure was
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30
highest, although there did appear to be small, non-signi-
cant differences in this parameter between the highly
purified and tap water. In a reevaluation involving the
effect of Durham tap water and purified tap water as in the
above study, Chernoff (1977) did not find striking differ-
ences jn the reproductive success of CD-1 mice. No terato-
genic studies on haloforms other than chloroform were
available.
E. Mutagenicity
The THMs (chloroform, bromodichloromethane, dibromo-
chloromethane, dibromochloromethane and bromoform) were
assayed _iri vitro for mutagenic activity using strains of
Salmonella typhimurium (TA 100 & TA 1535). The assays were
conducted in desiccators to allow each compound to volatil-
ize so that only the vapor phase came in contact with
bacteria on the petri dishes. The activation system was
»
tested and found not to be required for the bromohalometh-
anes since they were positive in the absence of activation.
The results obtained were as follows: (a) chloroform was not
mutagenic in TA 100 with or without activation, nor in TA
1535 without activation; (b) bromodichloromethane was
mutagenic in TA 100 without activation, with a doubling dose
of approximately 25 microliters; (c) dibromochloromethane
was mutagenic in TA 100 without metabolic activation, with a
doubling dose of approximately 3.5 microliters; (d) bromo-
form was mutagenic in TA 100 without metabolic activation,
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31
with a doubling dose of approximately 25 microliters, and
was also mutagenic in TA 1535 with metabolic activation,
with a doubling dose of approximately 100 microliters
(Tardiff, 1976). All three compounds demonstrating
mutagenic activity did so in a dose-response mode. For
certain classes of compounds, except for many chlorinated
hydrocarbons (Ames, 1973) the Ames test which utilizes
Salmonella typhimurium bacteria correlates highly (90
percent) with the ^n vivo carcinogenicity bioassay.
F. Carcinogenicity
Prolonged administration of chloroform at relatively
high dose levels to animals, specifically mice and rats,
manifested oncogenic effects. The investigation conducted
by Eschenbrenner and Miller (1945) produced hepatomas in
female mice (strain A) given repeated dosages ranging from
0.145 to 2.32 mg of chloroform for a period of only four
months. Minimum doses of 593 mg/kg chloroform per day
(total of 30 doses) produced tumors in all of the surviving
animals.
In a recent bioassay (NCI, 1976) linking chloroform with
oncogenicity, rats and mice of both sexes were fed doses of
chloroform ranging from 90 to 200 (rats), and 138-477 (mice)
mg/kg. In this study, the lowest dose for observed carcino-
genic effect (kidney epithelial tumors) in male rats was 100
mg/kg and for mice 138 mg/kg administered to the animals for
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32
a total period of 78 weeks. A related halogenated hydorcar-
bon, carbon tetrachloride, was carcinogenic in Osborne
Mendel rats and in B6C3F1 mice at dosages ranging from 57 to
160 rag/kg and 1250 to 2500 mg/kg, respectively. The
incidence of hepatocellular tumors formed in these animals
at both dose levels almost approached one hundred percent
(Table 5). The percent survival in mice treated with
chloroform and carbon tetrachloride is depicted in Table 6.
Almost all the animals on treatment with carbon tetrachlor-
ide died between 91-92 weeks whereas with chloroform
treatment at both dose levels, 73 and 46 percent of the
animals survived. Miklashevskii et al. (1966) fed chloro-
form to rats at 0.4 mg/kg apparently for 5 months and
detected no histopathlogical abnormalities after this
treatment. A recent study on the carcinogenic effect of
chloroform at dose levels of 17 mg/kg/day and 60 mg/kg/day
was conducted by Roe (1976), utilizing the rat (Sprague-
Dawley), the beagle dog and four strains of mice (ICC Swiss,
C57B1, CVA and CF/1). Comparison with the NCI study (1976)
indicates that the number of animals and the duration of the
experiment were essentially similar; the major differences
were the dosages, which were lower than in the NCI study,
and the vehicle, which was toothpaste. The only finding of
neoplasia was an excess of tumors of the renal cortex in the
male ICI-Swiss mice at a dose level of 60 mg/kg/day.
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33
Table 5. Comparison of Hepatocellular Carcinoma Incidence in
Chloroform and Carbon Tetrachloride-Treated Mice
Animal Group
Males Controls
Low Dose
High Dose
Females Controls
Low Dose
High Dose
Chloroform
Carbon Tetrachloride
5777"
18/50
44/45
1/80
36/45
39/41
5/77
49/49
47/48
1/80
40/40
43/45
Table 6. Comparison of Survival of Chloroform and Carbon
Tetrachloride - Treated Mice
Chloroform •
Initial
Animal
Males
Group
—i B
Controls
Low Dose
High Dose
Females Controls
Low Dose
High Dose
No.
77
50
50
80
50
50
78
Weeks
53
43
41
71
43
36
90
Weeks
38
37
35
65
36
11
Carbon Tetrachloride
Initial
No.
77
50
50
80
50
50
78
Weeks
53
11
2
i
71
10
4
91-92
Weeks
38
0
0
65
0
1
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34
However, animals fed 17 mg/kg/day of chloroform showed no
incidence of renal carcinoma.
Some renal tumors were also seen in control animals in a
later study. The negative results observed in the dog
experiment may be explained on the basis that either the
animals were not exposed for a suitable length of time (i.e.
duration of life span) or that an insufficient number of
animals were tested, or that this species may not have been
responsive to the oncogenic effect of chloroform. The
negative results of the rat study may be explained on the
basis of lack of strain sensitivity. Based on the extra-
polation from the NCI study, the dose was too low to produce
an effect in so few animals (Cueto, NCI, 1979).
Much less information is available on the carcinogeni-
city of bromohalomethanes. Preliminary results from the
strain A mouse pulmonary tumor induction technique (Theiss
et al., 1977) indicated that bromoform produced a positive
pulmonary adenoma response while chloroform did not. Other
studies (Poirier, et al., 1975) indicated that in several
instances brominated compounds exhibited more carcinogenic
activity than their chlorinated analogs in the pulmonary
adenoma bioassay.
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35
VII. Human Health Effects
A. NAS Principles of Toxicological Evaluation
The recent NAS (1977) report entitled "Drinking Water
and Health" identified several principles for assessing the
irreversible human effects of long and continued low dose
exposure to carcinogenic substances.
Principle 1; Effects in animals, properly qualified, are
applicable to man.
Principle 2; Methods do not now exist to establish a
threshold for long-term effects of toxic agents.
Principle 3: The exposure of experimental animals to
toxic agents in high doses is a necessary and valid
method of discovering possible carcinogenic hazards in
man.
Principle 4: Materials should be assessed in terms of
human risk, rather than as "safe" or "unsafe".
On the basis of studies in animals and human toxicological
data the NAS (1977) has recommended that strict criteria
should be applied for establishing exposure limits to
chloroform.
The National Institute for Occupational Safety and
Health has recommended that the occupational exposure to
chloroform should not exceed 2 ppm determined as time-
weighted average exposure for up to a 10 hour work day.
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36
The human health effects as observed in accidental,
habitual, and occupational exposures appear to indicate that
the effects produced by exposure to chloroform are similar
to those found in experimental animals. These include
effects on the central nervous system, liver, and kidney.
The" symptoms observed (.Storms, 1973) in a 14 year old
patient following an accidental exposure to an unknown
amount of chloroform included cyanosis, difficulty in
breathing and unconsciousness. Liver function tests
measured by serum enzyme levels four days after ingestion
indicated high levels of SCOT, SGPT, and LDH. The authors
also noted damage to the cerebellum characterized by an
instability of gait and a slight tremor on finger-to-nose
testing. The symptoms disappeared in two weeks.
Several cases of habitual chloroform use have also been
recorded by Heilbrunn et al. (19^5). A case study of
interest was a 33 year old male who had habitually inhaled
chloroform for 12 years. The subject showed psychiatric and
neurological symptoms including restlessness, hallucina-
tions, convulsions, dysarthria, ataxia, and tremors of the
tongue and fingers.
Lunt (1953) reported that delayed chloroform poisoning
in obstetric patients, anaesthetized with chloroform is
characterized by renal dysfunction as indicated by: albumin,
red blood cells, and pus in the urine. Chloroform exposure
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37
of humans by inhalation was studied by Lehman and Schmidt-
Kehl (1936). Ten different concentrations of chloroform
were used and the chloroform concentrations were determined
by the alkaline hydrolysis method. Exposure at concentra-
tions of 7 ppm for 7 minutes and at all higher levels up to
3000 ppm caused symptoms of central nervous system
depression.
Desalva et al. (1975) studied the effects of chloroform
in humans; the subjects were given dentifrice containing
3.U56 chloroform and raouthwash with 0.43* chloroform for 1 to
5 years. No hepatotoxic effects were observed at estimated
daily ingestion of 0.3 to 0.96 mg/kg chloroform. Reversible
hepatotoxic effects were manifested at 23 to 27 mg/kg/day
chloroform ingested for 10 years in a study conducted by
Wallace (1959).
B. Epidemiologic Studies
By August 1979, 18 epidemiological studies, and addi-
tional unpublished reports discussed possible relationships
between cancer mortality and morbidity and drinking water
supplies. The results of the studies are shown in Table 7
in the approximate chronological order of completion. The
table shows the statistically significant results of analy-
sis by anatomical site. The statistically significant posi-
tive results are denoted by "M" for males and "F" to females
and the statistically significant negative results are
denoted by "-" before the "M" or "F".
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39
Five of the studies were published through August 1979.
All of the studies were retrospective in design; sixteen
were correlation studies, and four used a case-control
approach. Four studies utilized cancer morbidity or
incidence rather than mortality as a measure of disease
frequency. The studies vary in sample size, cancer sites
considered, factors selected as possible explanatory
variables, parameters selected as indicators of water
quality, and in the statistical techniques used for
analysis, so caution must be used in comparing the results
of one study with the results of another study.
There are several problems which make the results
difficult to interpret: 1) there is limited water quality
data on organics and other contaminants in the finished
drinking water, and the data which exist cover less than
five years; and 2) the water quality data are often from -
geographic areas other than those (usually counties)
reporting cancer mortality data.
The water quality data are recent, and Lt. is not known
to what extent they reflect past exposure to THMs. This is
important, since the latent period for most types of cancer
is measured in decades. Comparison of the various study
results is difficult also because of the different
approaches used.
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40
In general, retrospective epidemiological studies are a
useful methodological tool in i hypothesis generation.
The results from these studies, when viewed collectively,
can provide some insight into the postulation of causal
relationships which then need to be tested further, using
epidemiological designs such as case-control or cohort
studies, for documentation.
When the evidence from all studies is weighed, an
emphasis can be placed not only on the statistical signifi-
cance of single correlation coefficients but on their
consistency and patterns. When more than one independent
study shows positive associations for site-specific cancers,
then the association may not be due to chance alone. When
the association is verified by consistent results across all
four sex-race groups (white male, non-white male, white
female, non-white female), the association is more likely to
be used due to the variable considered and the evidence
should be viewed more seriously. The studies done so far
suggest the appropriateness of concern.
There is much evidence (both epidemiological and experi-
mental) that most human cancers result from a combination of
causes (Weisburger, 1977). Etiologic factors (e.g. smoking
as a cause of lung cancer, soot as a cause of scrotal cancer
in chimney sweeps) that result in increased relative risk
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41
greater than 5, were among the first to be discovered. The
etiologic factors associated with cancers of gastrointesti-
nal and urinary tract are more difficult to isolate from
epidemiological studies because of the lower incidence and
mortality rates, the interaction of environmental causes,
and site»specific differences. The increased relative risk
of populations exposed to most factors suspected of being
associated with gastrointestinal and urinary cancers are
less than three. Effects as small as, or smaller than,
these, are difficult to detect or quantify.
A number of the epidemiologic studies relating "water
quality" to cancer did not define the water quality
parameter by chemical constituents but instead compared
cancers in persons who used water from different sources.
Among the first of these was an investigation by Page,
Talbot, and Harris (1974). The study considered Louisiana
parish (county) cancer mortality rates for 1950-69, for
total cancers and various selected cancer sites, and related
these to the percentage of the parish populations drinking
water from the Mississippi River, which is known to be con-
taminated by many organic chemicals (Laseter, 1972). The
variables controlled were the rural-urban character of the
parish, median income, population density, and proportion of
population employed in the petroleum, chemical, and mining
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42
industries. An unweighted regression analysis showed a
positive correlation between drinking water and total cancer
(excluding cancer of the lung, urinary tract, GI tract, and
liver), and then separately for cancer of the gastrointes-
tinal organs and lung cancer. These investigations suggest-
ed an association between cancer mortality rates and use of
drinking water from the Mississippi.
Meinhardt, et al. (1975), commenting on the Page-Harris
report, looked at the cancer mortality gradient by apparent
"dose" of river water and concluded that there was a random
distribution of-high and low cancer mortality rates among
the river water consumers along the lengths of the Missouri
and Mississippi River systems.
Subsequent reports by Page and Harris (1975, 1976) on
the "Relation Between Cancer Mortality and Drinking Water in
Louisiana" utilized explanatory variables and cancer sites
similar to those in the first study; relationships for all
four sex-race groups were considered. Positive regression
coefficients for the water variable that were found
statistically significant were:
Total cancer sites: WM, NWM, NWF
All other than lung: WM
Urinary Tract: WM, NWF
Gastrointestinal: WM, NWM, WF, NWF
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13
Tarone and Gart (1975) reviewed the Page-Harris work and
included an additional variable, elevation above sea level.
By using a weighted regression analysis for four race-sex
groups, statistically significant, positive correlations
were found between the water variable and total cancer and
lung cancer mortality for white males (WM), non-white males
(NWM), and non-white females (NWF). The correlations were
not statistically significant for while females (WF) for the
same sites. Thus, there was a lack of consistency across
the four sex-race groups for the aforementioned cancer
sites.
Vasilenko and Magno (1975) conducted an ecological study
in New Jersey and determined the relation between water
source and age-adjusted cancer mortality from lung, stomach
and urinary tract cancer of white females. Water quality
•
was estimated from the ratio of the number of households
served by public systems and private water companies to the
number served by individual wells. Positive associations
were found for lung and stomach cancer.
DeRouen and Diem (1975) also reviewed the relationship
of cancer mortality in Louisiana and the Mississippi River
as the drinking water source looking at ethnic variables as
a possible confounding factor. By dividing Louisiana into a
northern and southern section, they were able to mimic an
ethnic division of the population. Many of the variables
-------
(urban-rural characteristics, median income, employment
characteristics, and elevation above sea level) included in
the previous studies were omitted. The water variable was
handled differently by the investigators. Population groups
were dichotomized into those who obtained none of the water
from the Mississippi River, and those who obtained some or
all from the river. The results showe a positive
relationship between cancer mortality and drinking water,
for gastrointestinal cancer. The cancer mortality rates for
southern parishes of Louisiana whose source of drinking
water is the Mississippi River are higher than in the
southern parishes whose source of drinking water is not the
Mississippi River for the following:
Stomach: NWF Cervix: NWF
Rectum: WM Lung: NWF
Large Intestine: WF, NWF Total Cancer: NWF
The cancer mortality rates tend to be higher for the
southern parishes with river water use than northern for
river water parishes for cancer of the urinary tract,
gastrointestinal tract, and the lung.
In another set of analyses and comments, DeRouen and
Diem (1975) discuss the problems associated with interpreta-
tion of regression coefficients as they relate to the Page
and Harris Report, particularly the problem of making inter-
ferences from correlational studies. They concluded that
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45
inconsistencies such as the failure to see the same rela-
tionships for all sex-race groups reduces the credibility of
the hypothesis of a causal relationship between water source
and cancer risk.
An analysis was done by McCabe (1975) of EPA using the
50 (of a total of 80) NORS cities with a 1950 population
greater than 25,000 and 70 percent or more of the city's
population^ receiving water comparable to that sampled by
EPA. McCabe showed a statistically significant correlation
between the chloroform concentrations in the drinking water
and the cancer mortality rate by city for all cancers
combined.
In a second analysis by McCabe using water quality data
from Region V, correlations between chloroform and TTHMs and
total cancer mortality were not positive. When the same -
correlations were done using Region V plus NORS data for
chloroform and total trihalogenated methane concentration
levels, a positive statistically significant result was
obtained.
Several epidemiological studies have been conducted in
the Ohio River area. Buncher (1975) conducted a study of 88
counties (in Ohio, bordering the Ohio River) of which 14
used the Ohio River as a drinking water source. Buncher
reports no significant relationship with drinking water from
the Ohio River and the higher cancer mortality rates. There
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46
was a weak positive correlation between the chloroform
concentration in 23 cities and the cancer mortality rate for
all cancer sites in white males. Similar results were found
in 77 cities (59 with surface water supplies) between
chloroform concentrations and pancreatic cancer mortality in
white females. For cities that accounted for more than 70
percent of the county population, there was a significant
correlation between chloroform concentration and bladder
cancer mortality rates for both white males and white
females.
As a follow up on the Buncher study, a study by Kuzma,
et al. (1977), considered the 88 Ohio counties, classified
as either ground water or surface water counties based on
the source of the drinking water used by a majority of the
county residents. A two-stage analysis was performed and no
•
statistically significant results were shown between the
drinking water from the Ohio River and cancer mortality
rates. However, rates for stomach, bladder, and total
cancers were higher for white males in counties served by
surface water supplies (probably chlorinated) than in
counties served by ground water supplies (probably not
chlorinated).
Reiches, et al. (1976), re-examined the Ohio data using
a different methodology. Correlations between the surface
drinking water variable and cancer mortality rates for
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47
stomach cancer and total cancers for both white males and
females were statistically significant. The correlations
between the drinking water variable and cancer mortality
rates of the pancreas, bladder, esophagus, gastrointestinal
tract, and urinary organs were significant for white males
only.
Although several studies defined the water quality para-
meter by chlorination or levels of chloroform, only one
study has considered the relationships of cancer with all
THMs, both collectively and separately. Cantor et al.
(1978) studied the correlation of cancer mortality at six-
teen anatomical sites with the presence of concentration
levels for each THM and TTHM in drinking water for whites.
Counties were grouped according to the percentage of the
county population served by the sampled water supply. In
both sexes, there was a positive dose-response gradient of
increasing correlation between trihalomethane concentration
and bladder cancer. The correlation was stronger for bromo-
form than with chloroform. There was a negative correlation
in white females of stomach cancer with total THM levels.
Kidney cancer in white males showed a positive correlation
with chloroform levels. Lung cancer in white females showed
a positive correlation with THM levels. Among white males
non-Hodgkins' lymphoma showed a positive correlation with
bromoform. A positive dose-response was observed between
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48
brain cancer mortality (in both sexes) with increasing use
of water containing chloroform, but the associations were
not strong.
Alavanja, et al. (1976) conducted a retrospective, case-
control study of female cancer mortality and its relation-
•
ship to drinking water chlori'nation in seven selected New
York counties. A statistically significant association was
found between a region being served from a chlorinated
drinking water supply and combined gastrointestinal and
urinary tract cancer mortality rates in that region. There
was also a higher mortality for the summed gastrointestinal
and urinary cancer in urban areas served by chlorinated
surface or ground drinking water supplies than in urban
areas served by nonchlorinated supplies, however, the
results should be viewed cautiously due to the small numbers
in the sample.
Alavanja (1977) expanded this study and included
gastrointestinal and urinary cancer deaths. Results showed
that males living in the chlorinated water areas of three
counties and females living in the chlorinated water areas
of two counties were at greater risk of gastrointestinal and
urinary tract cancer mortality than individuals living in
the non-chlorinated areas. Alavanja (1978) did a second
study (shown on Table 7), which expanded the first to
nineteen counties in New York and several specific cancer
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49
sites. Statistically significant positive associations were
found for males and lung cancer and for females and
pancreatic cancer. Statistically significant positive
associations were found for both males and females and
cancer of the large intestine, combined gastrointestinal,
and all cancers.
Kruse (1977) conducted a retrospective, case control
study of white males and females in Washington County,
Maryland. The relationship between mortality and morbidity
from liver (including biliary passages) and kidney cancer in
areas supplied by chlorinated public water supplies was
analyzed. While there was a higher incidence of liver can-
cer among the exposed groupj i.e.^the group which consumed
chlorinated drinking water, the correlations were not
statistically significant. It should be noted that the
sample size was small and that fewer than 50 cases each of
liver cancer and kidney cancer were counted.
Salg (1977) also conducted a retrospective study of var-
ious cancer mortality rates and drinking water from a var-
iety of sources and receiving different types of treatment
in 346 counties in seven states in the Ohio River Valley
Basin. She compared mortality rates for white and non-white
males and females using weighted regression analyses, sur-
face water usage showed weak but statistically significant
associations between chlorinated water supplies (regardless
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50
of source) and the following cancers: for white males -
esophagus, respiratory organs, large intestine, rectum,
bladder, other urinary organs and lymphosarcoma and
reticulosarcoma; for white females - breast'and rectum, and
for non-white females -esophagus and larynx. Rectal cancer
showed positive correlations across all race-sex groups. It
should be noted that the test of significance utilized for
this study was p < 0.10, which is less stringent than that
used in other studies.
Mah, et al. (1977), conducted a retrospective study of
the white population in the Los Angeles County area of the
relationship between cancer mortality and morbidity and the
chlorinated drinking water supply. They did not reveal any
trends and showed no significant relationships for either
cancer mortality or morbidity. The authors pointed out
several methodological problems, including the diluting
effect of migration into the area covered by this study.
Hogan et al. (1979) also utilized the NORS and Region V
data sets and applied various statistical procedures to the
data in order to determine the effects of using different
statistical models. Their results were similar to previous
studies showing a positive correlation between rectal-intes-
tinal and bladder cancer mortality rates and chloroform
levels in drinking water when weighted regression analysis
were applied. However, as the authors pointed out, "the
marked extent to which these results were dependent on (1)
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51
the weighting scheme adopted in the analysis, (2) the pre-
sumed appropriateness of the data, and (3) the characteris-
tics of the statistical model, was also clearly illustra-
ted."
Wilkins (1978) conducted a case-control study in
Washington County, Maryland and investigated the association
•
between liver, kidney and bladder cancer and chlorinated
water source. A positive correlation was found for female
liver cancer and male bladder cancer and the chlorinated
drinking water source. Due to small numbers .of cases the
outcome of this study should be viewed with suspicion.
Rafferty (1979) studied associations between drinking
water quality in North Carolina communities and cancer
mortality rates. The drinking water supplies were
characterized by domestic and/or industrial contribution.
No significant positive association were found.
Tuthill and Moore (1978) investigated the association
between cancer mortality rates and parameters of water
quality for Massachusetts community public water supplies.
The average annual chlorine dose was one of the independent
water characteristics. Simple correlations showed that the
average chlorine dose level in the water was negatively
associated with female buccal cancer, and positively assoc-
iated with female esophageal and male respiratory cancers.
Occupation, population mobility, and other demographic
variables were controlled.
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52
In summary, many but not all of the studies have found
positive correlations between some characteristics of
drinking water and various cancer mortality/morbidity rates.
However, these correlations are dependent upon the selection
and appropriateness of the data, the weighting scheme and
extrapolation in the analysis, and the characteristics of
the statistical model. Because of these dependencies the
quantitative, causal interpretation of results generated
from an indirect or ecological study should be viewed as
tenuous for the primary purpose of generating hypotheses and
even questionable in most cases.
It is important in the evaluation process to consider
the results from other epidemiological studies as they
develop hypothesis of potential causal associations between
cancer mortality and other agents. For example, the
confounding factors of diet, occupation, and smoking all
have been suggested as potential causative agents of bladder
cancer, Cole (1972). Therefore, any epidemilogical study
that investigates the possible association between bladder
cancer and drinking water should be designed to avoid the
problems that result in confounding of the data. None of
the studies completed thus far have obtained data on or
controlled for diet; several studies have attempted to
control for occupational exposure (Page and Harris, 1971* and
1975; Cantor, et al., 1978; Tuthill and Moore, 1978);
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53
only the studies by Kruse (1977) and Wilkins (1978) obtained
smoking data. Only a few studies considered four sex-race
groups (the number of non-whites is too small in some of the
geographic areas) and of those studies only a few showed
consistent patterns of associaton of specific cancer sites.
e.g., Salg (1977)-rectum. Several studies which considered
only white populations found positive correlation coeffi-
cients for both sexes: Kuzma (1977) - stomach; De Rouen
(1975) - intestine, stomach and bladder; Buncher (1975) -
bladder; Reiches (1976) - stomach; Cantor (1978) - bladder;
Hogan (1979) - intestine and bladder; and Alavanja (1978) -
intestine. Only a few studies defined the water quality
variable by the chloroform concentrations (McCabe, 1975;
Buncher, 1975; Cantor et al., 1977; Hogan et al., 1977;
Alvanja, 1978), and by the THM concentrations (Cantor et al.
1977).
Of particular interest are possible correlations of
liver and kidney cancer rates with drinking water, since the
animal exposure data indicate that hepatocellular carcinomas
and hepatic modular hyperplasias have been observed in
B6C3F1 strains of mice after life time exposure to chloro-
form. Several of the preliminary studies grouped the cancer
sites for the anatomical systems, e.g. , gastrointestinal and
urinary organs, in order to increase the sample size. One
of the studies (Cantor, 1978) which considered site-specific
cancer mortality showed a positive association between
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drinking water and cancer of the kidneys in white males.
The absence of any positive association between drinking
water and liver cancer mortality may be due in part to small
sample sizes, very low incidence of the disease, or because
the exposure levels of contaminants in trace amounts over a
lifetime may be below a no-effect level (Weisburger, 1977).
The incremental- increase may be too small to measure for
statistical significance. On the other hand, many scien-
tists believe that the specific site in which cancer appears
in animal tests need not necessarily be the same site in
which the cancer is likely to appear in humans.
Thus, the evidence is incomplete and the trends and
patterns of association have not been fully developed. As
stated previously, a causal relationship cannot be estab-
lished by correlation studies. When viewed collectively,
the epidemiological studies completed thus far provide
evidence for maintaining a hypothesis that there may be a
health risk and that the positive correlations may be due to
an association between some constituents of drinking water
and cancer mortality. The animal test data alone provide a
firm basis for policy decision making. Additional epidemio-
logical studies may provide evidence regarding the strength
of the associations and the possibility of a causal rela-
tionship between drinking water and cancer mortality, and
thus provide a stronger basis for further regulatory action.
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55
The NAS Epidemiology Subcommittee of the Safe Drinking
Water Committee reviewed the first thirteen of the aforemen-
tioned eighteen studies. In the report, "Epidemiological
Studies of Cancer Frequency and Certain Organic Constituents
of Drinking Water — A Review of Recent Literature Published
and Unpublished," September 1978, the Committee reached^the
following conclusions, which are consistent with EPA. Among
the group of studies that characterized water quality by
actual measurements, the results suggest:
that higher concentrations of THMs in drinking water may
be associated with an increased frequency of cancer of
the bladder. The results do not establish causality,
and the quantitative estimates of increased or decreased
risk are extremely crude. The positive association
found for bladder cancer was small and had a large
margin of error; not only statistical, but much more
importantly, because of the very nature of the studies.
Further research is being conducted with more definitive
analytical studies. A large case-control bladder cancer
study with 3,000 cases and 6,000 controls is being conducted
by the National Cancer Institute (NCI). Three other
case-control colon cancer studies are being conducted in
Louisiana, Pennsylvania, and Utah. The results of these
studies may provide more solid evidence to answer the
question of possible associations between water quality and
increased incidence of bladder and colon cancer.
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56
VIII. Mechanism of Toxicity
Biologic responses upon exposure of mammals to chloro-
form include effects on the central nervous system resulting
in narcosis, hepatotoxicity, nephrotoxicity, teratogenicity
and carcinogenicity. Elucidation of the mechanism of
toxicity of chloroform and related compounds has been
attempted by several researchers.
Scholler (1968) and McLean (1970) observed that
phenobarbital pretreatment of rats caused an increase in
liver necrosis after administration of chloroform. Later,
Brown, et al. (1972*) reported that exposure of rats to an
atmosphere containing chloroform (0.5%) for 2 hour markedly
decreases glutathione (GSH) concentration in the liver when
the animals have been pretreated with phenobarbital. In an
attempt to further elucidate the role of GSH in chloroform-
induced hepatotoxicity, Docks and Krishna (1976) injected
chloroform into rats pretreated with microsomal enzyme
inducers - phenobarbital, 3-methylcholanthrene, acetone and
isopropanol. A dose of chloroform as little as 0.2 mg/kg
decreased liver GSH levels and caused centrilobular necrosis
within 24 hours in phenobarbital pre-treated rats. At a
dose of 0.05 ml/kg, chloroform did not decrease liver GSH or
cause liver necrosis. When the rats were not pretreated
with phenobarbital, a chloroform dose of 0.2 ml/kg caused
neither GSH depletion nor necrosis. In this connection, it
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57
is interesting to note that cysteine, which is a precursor
of GSH and a common amino acid in one's diet, protected the
liver from the hepatotoxicity produced by chloroform. The
animals were also protected from the hepatotoxic effect by
pretreatment with cystamine, not a precursor of GSH, thus
suggestive, of a mechanism other than of GSH depletion in the
hepatotoxicity of CHC13.
Earlier reports by Ilett, et al. (1973) suggested the
possibility of another mechanism involving the formation of
an active metabolite of chloroform responsible- for the
chloroform-induced hepatotoxicity. This study correlated
the renal and hepatic necrosis with covalent binding of
chloroform metabolites to tissue macromolecule. Bioactiva-
tion of xenobiotics including chloroform, involves mixed
function enzymes; the NADPH cytochrome reductase-cytochrome
P-450 coupled systems. Sipes, et al. (1972) studied the
bioactivation of carbon tetrachloride, chloroform and
bromotrichloromethane utilizing I4c-labeled compounds and
rat liver microsomes. The covalent binding of radiolabel to
microsomal protein was used as a measure of conversion of
the compounds to reactive intermediates. The authors con-
cluded that cytochrome P-450 is the site of bioactivation of
these three compounds rather than NADPH cytochrome C reduc-
tase. CC14 bioactivation proceeds by cytochrome P-450
dependent reductive pathways, while CHCl^ activation,
proceeds by cytochrome P-450 dependent oxidative pathways.
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58
The isolation and identification of an active metabolite of
chloroform supposedly responsible for toxicity was attempted
by Pohl and his co-workers (1977). 2-oxithiazolidine-4-
carboxylic-acid, an in vitro metabolite of chloroform, and
presumably formed by the reaction of cysteine and phosgene
(COC^), was isolated and characterized. When the
incubation was conducted in an atmosphere of [1°0] 02,
the trapped COC12 contained [^0]. These findings
suggest that C-H bond of CHC13 is oxidized by a cytochrome
P-H50 mono-oxygenase to produce trichloromethanol which
spontaneously dehydrochlorinates to phosgene. The electro-
philic phosgene could react with water to form carbon
dioxide, a known metabolite of CHClj in vitro and in vivo
or with microsomes to yield a covalently bound product. The
i.n vitro oxidation of chloroform and its relationship to
chloroform toxicity has been further substantiated by the
studies wherein deuterated chloroform was used. Pohl and
Krishna (1978) reported that CDCl^ was metabolized slower
than chloroform suggesting that the cleavage of C-H bond of
chloroform is the rate determining step in the enzymatic
process. The observation that CDC13 is less hepatotoxic
than CHC13 indicates that the cleavage of the C-H bond is
also the critical step in the process leading to CHCl^
induced hepatotoxicity. The finding that CDC13 depletes
less glutathione in the liver of rats than CHCl^ suggests
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59
the active metabolite phosgene is responsible for the
depletion of glutathione.
In the experiments involving the isolation and
characterization of metabolites of chloroform, the evidence
for the metabolism of chloroform to phosgene i£ vitro, by
the oxidative pathway was present. Recent research has
indicated the possibility of formation of phosgene in vivo.
Pohl, et al. (1979), isolated and characterized 2-oxo-thia-
zolidine-H-carboxylic acid from the liver of .rats pretreated
with cysteine carboxylic acid after a dose of chloroform
and/or deutrated chloroform. In these experiments, deuter-
ated chloroform yielded less amount of metabolite, confirm-
ing once again the specificity of the cytochrome P-450
dependent enzymes in the mediation of oxidative dehalogena-
tion of chloroform and its toxicity.
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60
IX. Risk Assessment
The establishment of chloroform as an animal carcinogen,
plus the epidemiological data and mutagenesis data on THMs,
show that a potential human risk exists from the consumption
of THMs, but these data do not quantify the risk. Methods
have been developed to estimate the level of risk, based on
an assumption that there is no threshold level for the
action of a carcinogen. The state-of-the-art at the present
time is such that no experimental tools can accurately
define the absolute numbers of excess cancer "deaths
attributable to chloroform in drinking water. Due to the
biological variability and a number of assumptions required,
each of the risk-estimating procedures leads to a different
value. There is wide variation among these estimates and
their interpretation.
The EPA Science Advisory Board (SAB)(1975), using the
highest levels of chloroform then reported in drinking water
by the-NORS data (0.300 mg/1) and assuming a maximum daily
intake of 4 liters of water for a 70 kg man, attempted to
estimate the risk. The estimates were based on the
Eschenbrenner and Miller (1945) animal data, which them-
selves are subject to great variability since the experi-
ments used only 5 animals per sex per dose. Using a linear
extrapolation of the animal data over more than 2 orders of
magnitude dose from mice to humans at the 0.300 mg/1 concen-
tration level, the lifetime incidence of liver tumors in man
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61
were estimated to range from 0 to .001 (95% confidence
limits) or 0 to 100 x 10~5 in a lifetime. This rate may
be compared with the lifetime incidence of 260 x 10~5 for
malignancy of liver derived from the data of the Third
National Cancer Survey (1976). This estimate would range
from zero to approximately 4056 of the observed incidence of
liver cancer in the United States that may be attributable
to exposure to chloroform in drinking water at the 0.300
mg/1 level. It should be noted that this value is at the
upper limit of the confidence interval and the linear non-
threshold dose-effect model allows an estimate of maximal
risk where a risk has actually been observed. Most other
models would yield lower estimates. The SAB, however, also
stated that a more reasonable assumption would yield lower
estimates of the risk.
Tardiff (1976) using four different models, calculated
the maximum risks from chloroform ingestion via tap water.
Using a margin of safety of 5000 applied to the minimum
effect animal dose, i.e., the Weil conjecture, the "safe"
level was calculated to be 0.2 mg/kg/day. Using the log-
probit model and the slope recommended by Mantel and Bryan,
the conclusion reached was that at a maximum daily dose of
0.01 mg/kg the risk would be between 0.016 and 0.683 cancers
per million exposed population per year. Using the identi-
cal data, but with the experimental slope of the dose
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62
response curve as found in the mice as opposed to the slope
of the one in the previous calculation, the conclusion
reached was that a maximum daily dose of 0.01 mg/kg would
produce less than one tumor per billion population per
lifetime. Using the linear, or one hit model, usually
considered to be the most conservative, a risk estimate of
between 0.42 and 0.84 cancers per million population per
year was calculated to result from a dosage level of 0.01
mg/kg/day. The two step model produced an estimated maximum
risk of between 0.267 and 0.283 cancers per million popula-
tion per year at a dose level of 0.01 mg/kg/day.
In the National Academy of Sciences (1977) report on
"Drinking Water and Health," lifetime risks were estimated
from the more recent, and much more extensive NCI animal
data using a multi-stage model.
For a concentration of chloroform at 1 ug/liter the
estimated incremental lifetime cancer risk would fall at
approximately 1.7 x 10~6 per microgram per liter at the
upper 95% confidence limit, assuming 70 year daily consump-
tion of water at that level. Assuming lifetime exposure at
the standard of 0.10 mg/1 level in drinking water the
incremental risk would be 3.4 x 1Q-4 assuming two liters
of water at 0.10 mg/1 consumed daily for 70 years.
In evaluating the risk estimates, it is important to
compare the calculated maximum risk with the current cancer
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63
mortality data. Both liver and kidney cancer are rare
diseases in the U.S. « 5 per 100,000 population per year).
The standardized mortality rates in the U.S. for white
males and females combined are 52.5 per million per year for
liver cancer and 29.2 per million per year for kidney
carcinoma.
Based on the various risk estimates, Tardiff (1976)
calculated that the percent of the annual cancer mortality
attributable to chloroform in drinking water could be 1.60J
and 1.44$ for liver and kidney cancer respectively assuming
the maximum exposure levels. Applying these percentages to
the actual cancer mortality rates, the number of cancer
deaths per year would be 168 from liver carcinoma or 84 from
kidney carcinoma; an estimated maximum of 252 cancer deaths
per year attributable to chloroform in drinking water.
Reitz, Gehring, and Park (1978) discussed EPA's
procedures in estimating risk. They stated that EPA
"seriously overestimates the actual potential of chloroform
.... (for) two major reasons." These are: (1) The mechanism
through which chloroform exerts its toxicity, and (2) reli-
iance on the NCI bioassay protcols which call for high doses
of chloroform, and by not conducting studies at lower doses
which usually induce relatively less carcinogenicity, there
is a likelihood of ignoring a possible detoxification mecha-
nism which protects test animals until they are overwhelmed
by very large doses. They also suggest that an experiment
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64
to evaluate the carcinogenicity of chloroform at lower doses
must be performed before high/low dose extrapolations can be
performed. Definitive data do not exist to prove or to dis-
prove the above claims.
The authors indicated that EPA's proposed standard for
THMs of 0.10 mg/1 in drinking water supplies was based on
the carcinogenic risk estimates. It should be pointed out
the EPA's proposed standard for THM was based upon that
feasibility of achieving the TTHM concentration in drinking
water, as well as the potential adverse health effects.
EPA's Office of Water Planning and Standards and Office
of Research and Development with EPA's Carcinogen Assessment
Group, developed a risk estimate in the draft document,
"Chloroform - The Consent Decree Ambient Water Quality
Criteria Document" (1979). The method used assumed consump-
tion of 2 liters/per day of drinking water and 18.7 gm/per
day of fish and shellfish. The lifetime risk estimates for
excess cancers ranged from 10~5f 10~6, and 10~7 with
corresponding consumption of 2.1 ug/1, 0.21 ug/1 and 0.021
ug/1, respectively. The difference in these risk estima-
tions may be explained by the assumption of daily fish
consumption as well as other exposure sources. Without the
fish consumption, the equivalent concentrations are 4.8 ug/1
and 0.48 ug/1 for estimated cancer risk of 1 x 10~5 and 1
x 10~6, respectively. When this estimate is computed for
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65
the concentration of 0.10 mg/1 for levels in drinking water,
the incremental risk would be 4.0 x 10-1* assuming two
liters of water at 0.10 mg/1 was consumed daily for 70
years.
At an assumed lifetime exposure of 2 liters of water per
day at 0.10 mg/1 chloroform the risk reduction to the
impacted population was estimated as a range of approxi-
mately 200-500 total cases. It should be noted however,
that these average exposure levels in the impacted popula-
tion may result in overestimates of the risk-in light of the
facts that: 1) The computations are based upon lifetime
exposures. In actuality the proposed interim standard will
likely be reduced in the future as technologically feasible,
and, therefore, the lifetime exposure values will be less.
2) The interim standard encourages maximum reduction obtain-
able using current technology. A much lower average expo-
sure is likely in the future because technology will most
likely improve and result in greater exposure reductions.
On the other hand, these may be underestimated because they
are based upon toxicity exposure data from chloroform, which
is only a portion of the TTHMs, which are only a portion of
the by-products of the chlorination process; therefore, the
magnitude of the contribution to the risk of the other THMs,
which in some cases contribute significantly to TTHMs, is
unknown. The exposure to THMs from air and food have not
been included in these computations.
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66
X. Seclected Maximum Contaminant Levels (MCLs)
Since a risk to the public exists from exposure to TTHMs
and other chlorination by-products in drinking water, the
potential for that risk should be reduced as much as is
technologically and economically feasible without increasing
the risk of microbiological contamination. This can be
accomplished by several means, and the Safe Drinking Water
Act (PL 93-523) provides two major regulatory avenues - 1)
the establishment of an MCL, or 2) the institution of a
treatment requirement.
EPA has determined that the establishment of an MCL in
the Interim Primary Drinking Water Regulations, along with
monitoring requirements, is the most effective and immediate
approach to reducing the levels of THMs in drinking water.
The Administrator has determined that monitoring is both
technically and economically feasible (refer to "Economic
Impact Analysis of a Trihalomethane Regulation for Drinking
Water," EPA, 1977). Measures taken to reduce the THM
concentrations will concurrently provide the additional
benefit of reducing human exposure to the other undefined
by-products of chlorination and possibly other synthetic
organic contaminants.
Since it is known that chlorination of water is primari-
ly responsible for the relatively high levels of THMs in
drinking water, modifications in the chlorination process,
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67
the substitution of other disinfectants, and the use of
adsorbents and other technologies to remove precursor
chemicals are possible approaches to control. The optimal
approach would be to reduce organic precursor concentrations
prior to addition of a disinfectant in order to reduce
disinfectant demand and minimize all by-products.
Use of a chlorine residual in a less active form such as
combined chlorine or chloramine will significantly reduce
THM formation; however, chloramines are much less potent
disinfectants than free chlorine, and therefore, this
approach must only be used after careful consideration, and
assurance of maintenance of excellent biological quality.
The two chemicals most often mentioned as substitute
.disinfectants, ozone and chlorine dioxide, are both well
known as effective disinfectants and chemical oxidants. The
issues of the biological effects and toxicity of these
disinfectants and their by-products are being clarified by
studies underway. In the meantime, EPA recommended that the
residual total oxidant levels after application of chlorine
dioxide should be limited to 0.5 milligram per liter.
The National Organics Monitoring Survey found that the
mean total trihalomethane (TTHM) concentrations in the
drinking water systems evaluated were approximately 0.068,
0.117, 0.053 and 0.100 mg/1 for Phase 1, II, III (dechlori-
nated) and III (terminal) respectively, with the highest
levels of 0.78U mg/1 in Phase II (refer to Table 1).
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68
It is reasonable to assume that the various calculated
risk estimates fop chloroform indicate a potential risk to
public health. It is possible that a percentage of the
total number of liver and/or kidney cancers are attributable
to exposure of chloroform in drinking water, although it is
most likely that drinking water exposure would interact with
a number of other variables such as smoking and diet as
effect modifiers in a multifactorial manner. It is also
likely that the other by-products of chlorination also
present a potential risk.
Thus, based upon a number of risk extrapolations
assuming various levels of exposure to chloroform in
drinking water, it has been estimated that such exposures
may cause an annual excess of cancers in the U.S. population
(ranging from 0 to several hundred). At higher levels of
exposure of chloroform (> 0.300 mg/1) the cancer risk
estimates are even higher.
The reduction of TTHMs to an MCL level of 0.10 mg/1
would reduce the unnecessary and excessive exposure to these
potential human carcinogens, mutagens, and chronic
toxicants, and other effects. At the same time, measures
taken to reduce THM levels (such as the use of adsorbents)
will concurrently result in reduction of human exposure to
other contaminants in drinking water.
Since it is economically and technologically feasible to
reduce the THM levels in drinking water, and since benefits
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69
are achieved by reducing the health risks of exposure, EPA
has decided to establish the MCL at 0.10 mg/1 as the initial
feasible step in a phased, regulatory approach. As more
data become available from implementation experience, and
toxicology and epidemiology, standards are expected to
become more restrictive. In the meantime, EPA and the
States should continue to take steps as necessary on a
case-by-case basis to provide adequate protection for the
delivery of safe drinking water to the public, by minimizing
the amounts of toxic chemicals in the water.
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XI. References
Alavanja, M., et al., 1978. "An Epidemiological Study
of Cancer Mortality and Trihalomethanes in Drinking
Water in 19 New York State Counties." U.S. EPA, Office
of Research and Development, Health Effects Research
Laboratory, Cincinnati, Ohio, Unpublished.
Alavanja, M., Goldstein, I., and Susser, M., 1977.
"Case Control Study of Gastrointestinal Cancer Mortality
in Seven Selected New York Counties in Relation to
Drinking Water Chlorination." U.S. EPA, Office of
Research and Development, Health Effects Research
Laboratory, Cincinnati, Ohio, Unpublished.
Alvanja, M., Goldstein, I., and Susser, M., 1976.
"Report of Control Study of Cancer Deaths in Four
Selected New York Counties in Relation to-Drinking Water
Chlorination." U.S. EPA, Office of Research and
Development, Health Effects Research Laboratory,
Cincinnati, Ohio, Unpublished Draft.
American Conference on Industrial Government Hygienists,
1975. "Threshold Limit Values for Chemical Substance in
Workroom Air." Cincinnati, Ohio.
Ames, B.N., W.E. Durston, E. Yamaski, and F.D. Lee,
1973. "Carcinogens are Mutagens: A Simple Test System
Combining Liver Homogenates for Activation and Bacteria
for Detection." Proc. Nat. Academy of Science, USA, 70;
2281-2286.
Bellar, R.A., Lichienberg, J.J., and Kroner, R.C., 1974.
"The Occurrence of Organohalides in Finished Drinking
Waters." Journal of American Water Works Association,
66:
Brenniman, G., et al., 1979. "Relationship Between
Cancer Mortality and Chlorinated Drinking Water."
University of Illinois, Unpublished Draft.
Brown, D.M., Langley, P.P., Smith, D., and Taylor, D.C.,
1974. "Metabolism of Chloroform I. The Metabolism of
[C] Chloroform by Different Species." Xenobiotica M;
155-163.
Buncher, C.R., 1975. "Cincinnati Drinking Water - An
Epidemiologic Study of Cancer Rates." University of
Cincinnati Medical Center, Cincinnati, Ohio.
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Bunn, W.W., Haas, B.B., Deane, E.R., and Kleopfer, R.D.,
1975. "Formation of Trihalomethanes by Chlorination of
Surface Water." Environmental Letters 10; 205-213.
Burreson, G.J., Moore, R.E., and Roller, P.P., 1975.
"Volatile Halogen Compounds in the Algae Aspargopis
Taxiformis." J. Agric. Food Chem. 2jl: 856-861.
Butterfield, C.T. and Wat'tie, S., 1944. "Influence of
pH and Temperature on the Survival of Coliforms and
Enteric Pathogens When Exposed to Free Chlorine."
Public Health Rep. 5_8_: 1837-1866.
Butterfield, C.T. and Wattie, S., 1946. "Influence of
pH and Temperature on the Survival of Coliforms and
Enteric Pathogens When Exposed to Chloramine." Public
Health Rep. 6J_: 157-192.
Canter, K., Hoover, R., Mason, T., and McCabe, L., 1978.
"Association of Cancer Mortality with Halomethanes." J.
of National Cancer Inst. 6J_: 979-985.
Chernoff, 1977. "Personal Communication."
Clarke, N.A. and Kabler, P.K., 1954. "Inactivation of
Purified Coxsackie Virus in Water by Chlorine." Am. J.
Hyg. 59: 119-127.
Cotruvo, J.A., Simmon, V.F. and Spanggord, R.J., 1977.
"Investigation of Mutagenic Effects of Products of
Ozonation Reactions in Water." Annals of the New York
Academy of Sciences, in press, and "Personal
Communications."
Cole, P.C., Hoover, R., and Friedell, G.H., 1972.
"Occupation and Cancer of the Lower Urinary Tract."
Cancer, 29: 1250-60.
DeRouen, T.A. and Diem, J.E., 1975. "Ethnic,
Geographical Difference in Cancer Mortality in
Louisiana." Tolane University, School of Public Health
and Tropical Medicine, Unpublished.
DeRouen, T.A. and Diem, J.E., 1975. "The New Orleans
Drinking Water Controversy: A Statistical Perspective."
American Journal of Public Health, 6_5_: (No. 10): 1060.
DeSalva, S., Volpe, A., Leigh, G., and Regan, T., 1975.
"Long-Term Safety Studies of a Chloroform-Containing
Dentrifice and Mouth Rinse in Man." Food Cosmet.
Toxicol. 13: 529-532.
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72
Dewey, W., Personal Communication, Medical College of
Virginia.
Docks, E.L. and Krishna, G., 1976. "The Role of
Glutathione in Chloroform Induced Hepatotoxicity."
Experimental and Molecular Pathology 24; 13-22.
EPA, 1979. "Chloroform - The Consent Decree Ambient
Water Quality Criteria Document." EPA, Office of Water
Planning and Standards. Unpublished Draft.
. •
EPA, April 1976. "Industrial Pollution of the Lower
Mississippi River in Louisiana." Region VI, Dallas,
Texas.
EPA, 1974. "Analytical Report, New Orleans Water
Supply Study." Region VI, Dallas, Texas.
EPA, 1975. "A Report - Assessment of Health Risk from
Organics in Drinking Water by an Ad Hoc Study Group to
the Hazardous Materials Advisory Committee." Science
Advisory Board. Unpublished.
EPA, June 1975. "Region V Joint Federal/State Survey of
Organics and Inorganics in Selected Drinking Water
Supplies." Region V, Chicago, Illinois.
EPA, 1976. "Risk Evaluation of Chloroform." Office of
Pesticides Programs, Criteria and Evaluation Division.
Unpublished.
EPA, 1977. "Chloroform Risk Assessment in Drinking
Water." Cancer Assessment Group. Unpublished.
EPA, August 1977. "Economic Impact Analysis of a
Trihalomethane Regulation for Drinking Water," prepared
by Temple, Barker, and Sloan, Inc. for EPA, Office of
Drinking Water.
EPA, June 1976. "Interim Treatment Guide for the
Chloroform and other Trihalomethanes." EPA, Water
Supply Research Division, Municipal Environmental
Research Laboratory, Cincinnati, Ohio.
Eschenbrenner, A.B. and Miller, E., 19^5. "Introduction
of Hepatomas in Mice by Repeated Oral Administration of
Chloroform with Observation on Sex Differences." J.
Nat'l. Cancer Inst. 5_: 251-255.
Fry, F.J., Taylor, T., and Hathaway, E.D., 1972.
"Pulmonary Elimination of Chloroform and Its Metabolites
in Man." Arch. Int. Pharmacodyn. 196: 98-111.
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73
Hefferman, P., 1977, Personal Communication.
Heilbrunn, G., Liebert, E. , and Szanto, P.B., 1945.
"Chronic Chloroform Poisoning - Clinical and
Pathological Report of a Case." Arch. Neurol. Psych.
53: 68-72.
Hogan, M.D., Chi, P., Hoel, D.G., and Mitchell, T.J.,
1979. "Association Between Chloroform Levels in
Finished Drinking Water Supplies and Various
Site-Specific Cancer Mortality Rates." J. of Env. Path.
& Tox. 2: 873-887.
Ilett, K.F., Reid, W.P., Sipes, I.G. , and Krishna, G.,
1973. "Chloroform Toxicity in Mice: Correlation of
Renal and Hepatic Necrosis with Covalent.Binding of
Metabilities to Tissue Macromolecules." Exptl.
Molec. Pathol. 19: 215-229.
Jones, W.M., Marguis, G., and Stephen, C.R., 1958.
"Hepatoxicity of Inhalation Anesthetic Drugs."
Anesthesiology _!£: 715-23.
Kelly, S.M. and Sanderson, W.W., 1960. "The Effect of
Chlorine in Water on Enteric Viruses. II. The Effect
of Combined Chlorine on Poliomyelitis and Coxsackie
Viruses." Am. J. Public Health, 5_0: 14-20.
Kubic, V.L., Anders, M.W., Engel, R.R., Barlow, C.H.,
and Caughey, W.S., 1974. "Metabolism of Dihalomethanes
to Carbon Monoxide. In Vivo Studies." Drug Metabolism
and Disposition 2: 53-57.
Kruse, C.W., 1977. "Chlorination of Public Water
Supplies and Cancer - Washington County, Maryland
Experience." A Final Report, John Hopkins University,
School of Hygiene and Public Health to the Office of
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