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March 31, 1987
CONTROL OF LEGIONELLA IN PLUMBING SYSTEMS
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical methodology
and treatment technology that would be useful in dealing with the contamination
of drinking water. Most of the Health Advisories prepared by the Office of
Drinking Water are for chemical substances. This Health Advisory is different
in that it addresses contamination of drinking water by a microbial pathogen
and examines pathogen control rather than recommending a maximum allowable
exposure level. Thus, for a variety of reasons, the format and contents of
this Health Advisory necessarily vary somewhat from the usual Health Advisory
document.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
This Health Advisory (HA) is based upon information presented in the
Office of Drinking Water's Criteria Document (CD) for Legionella. Individuals
desiring further information should consult the CD. The CD is available for
review at each EPA Regional Office of Drinking Water counterpart (e.g., Water
Supply Branch or Drinking Water Branch), or for a fee from the National
Technical Information Service, U.S. Department of Commerce, 5285 Port Royal
Rd., Springfield, VA 22161, PB # 86-117843/AS. The toll-free number is (800)
336-4700; in the Washington, D. C. area: (703) 487-4650.
INTRODUCTION
Legionellae are bacteria that have been identified as the cause of
legionellosis. Based upon an attack rate of about 1.2 cases of legionellosis
per 10,000 persons per year (Foy et al., 1979), it has been estimated that
more than 25,000 cases of this disease occur annually within the United
States", and are caused primarily by one of the 23 currently recognized species
of the genus Legionella. Most people who have developed Legionnaires Disease,
the pneumonia form of legionellosis, were immunosuppressed or appeared to be
more susceptible because of an underlying illness, heavy smoking, alcoholism,
or age (more than 50 years old). In contrast, while some apparently healthy
individuals have developed Legionnaires Disease, outbreaks involving healthy
people have been limited mostly to the milder non-pneumonia form of the
disease called Pontiac Fever.
Legionellae are widespread in lakes and rivers (Fliermans et al., 1979,
1981). There is some indication that these organisms may be either very
sparse or absent in groundwater (Fliermans et al., 1982; Spino et al., 1984).
Spino et al. ("1984) was unable to isolate legionellae after aeration of
groundwater through a redwood-slat aerator. The possibility that humans may
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Legionella March 31, 1987
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be exposed transiently to legionellae because of their high rate of contact
with water is highly probable, given the high frequency of seropositivity to
legionellae in healthy populations (Wentworth et al., 1984) and the widespread
occurrence of legionellae'in water environments.
In a number of outbreaks of legionellosis that have occurred in the
United States, aerosols of water documented to contain the specific type of
legionellae that was recovered from the patient have been identified as the
vehicle for transmission (Cordes et al., 1981; Stout et al., 1982; Garbe
et al., 1985). It has been hypothesized that legionellae enter buildings
in very low numbers via the treated drinking water. These bacteria may
proliferate in warm water when factors not yet fully determined allow them.
Even when this occurs, as has been shown in numerous buildings, disease
usually does not result. Cases and outbreaks of legionellosis occur only
when aerosols containing legionellae possessing specific virulence factors
(not as-yet determined) are inhaled (possibly ingested) by susceptible
individuals. Foodborne outbreaks or secondary spread have not been reported.
This Health Advisory discusses the control of legionellae in drinking
water. This includes finished water at the treatment facility, the distri-
bution system, and plumbing systems. Plumbing systems include hot water
tanks, taps, showerheads, mixing valves, the faucet aerators, all of which
have been associated with the proliferation of legionellae. This guidance
does not discuss legionellae control for whirlpools, respirators, or heat-
rejection equipment such as cooling towers and air conditioners. These have
all been associated with cases of Legionnaires Disease.
Presence of Legionellae in the Distribution System and Plumbing Systems
Legionellae are found in raw water, in treated waters, and in plumbing
systems (Fliermans et al., 1981; Hsu et al., 1984; Witherell et al., 1984),
but the occurrence and fate of these organisms in the distribution system
between these points are unknown. The organism may survive the treatment
and disinfection process and pass intact through the distribution system.
In addition, opportunities exist for their introduction into the system by
means of broken or corroded piping, repair of existing mains, installation of
new mains, back siphonage and cross connections, any of which may result in
contamination of the water supply. In older distribution systems, especially
those dependent on gravity flow, deterioration of piping may be so severe
that the treated water comes in intimate contact with soil and is subject
to infiltration by surface water. Thus, legionellae may be introduced into
potable water by these routes.
Legionellae surviving initial water treatment may colonize pipe joints
and corroded areas or adhere to the surface or sediment of storage tanks,
especially those constructed of wood. Here, they may find a habitat suitable
for survival and growth (Engelbrecht, 1983). Cul-de-sacs, intermittently
used storage tanks and other sites in which waterflow is absent or restricted
also may be appropriate habitats for legionellae.
New distribution systems or their components that were not appropriately
cleaned and disinfected before being put into use may introduce legionellae
into the system. Although this has not been documented, it may not be
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Legionella March 31, 1987
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coincidence that some of the serious outbreaks of Legionnaires Disease have
occurred in newly-opened institutions or buildings (Haley et al., 1979; Marks
et al., 1979; Helms et al., 1983). Construction activities may have included
intervention into the water supply mains with introduction of contaminated
water or, possibly, disturbance of sediment and sloughing of scale bearing
high concentrations of legionellae by means of hydraulic shock or other
perturbations.
There are numerous reports of legionellae occurring in plumbing systems,
especially in hot water systems. Most of these investigations have been
carried out in hospitals, and many were prompted by outbreaks of nosocomial
(hospital-acquired) Legionnaires Disease. The primary reservoirs in hospitals
are apparently hot water tanks in which water is maintained at temperatures
below 55°C. Legionellae also have been found in showerheads, rubber fittings,
aerator screens, faucet spouts, and other plumbing fixtures. This group of
organisms has also been found in residential plumbing systems such as apartment
buildings and homes (Wadowsky, 1982; Arnow and Weil, 1984), but disease has
not been associated with these findings.
Control at the Water Treatment Facility
Only a few studies have been published on the effectiveness of various
types of treatment for eradicating or reducing legionellae numbers at the
water treatment utility. In one study, Tison and Seidler (1983) examined raw
water and three kinds of distribution water supplies: (1) those treated by
chlorine (free residual 0.2-0.6 mg/L); (2) those treated by sand filtration
and chlorination (free residual O.Q-0.4 mg/L); and (3) those treated by
flocculation, mixed media filtration, and chlorination (free residual
0.5-2.0 mg/L). Legionella were enumerated by direct fluorescent antibody
(DFA) tests and all distribution waters contained about one order of magnitude
fewer Legionella-like cells than did the raw waters, i.e., 10^-104 per liter.
While the evidence suggests that legionellae are common in treated water, the
significance of these results is questionable because the authors were unable
to isolate any legionellae by animal inoculation or culture procedures, and
there are uncertainties about the specificity of the DFA technique used for
legionellae detection.
Most water treatment plants in the United States use chlorine disinfection.
Although extrapolation of laboratory studies to treatment plant situations is
somewhat tenuous, Kuchta et al. (1983) reported that both L_. pneumophila and
L_. micdadei (laboratory-adapted environmental and clinical strains) were much
more resistant to chlorine than was Escherichia coli. At 21°C, pH 7.6, and
0.1 mg/L of free chlorine residual, a 99 percent kill was achieved in less
than one minute for Jj. coli compared to 40 minutes for L. pneumophila. Under
the same conditions, 0.5 mg/L of free chlorine resulted in a 99.9 percent
legionellae kill in about 5 minutes. The contact time for a 99 percent kill
of L_. pneumophila at 4°C was twice as long as it was at 21 °C. The authors
concluded that legionellae can survive low levels of chlorine for rather long
periods of time. In a subsequent study, Kuchta et al. (1984) compared agar-
passaged (laboratory-adapted) and tap water-grown strains of L_. pneumophila
with respect to chlorine.- resistance, and showed that the latter were consid-
erably more resistant. At 0.25 mg/L free residual chlorine, 21°C, and pH
7.6-8.0, a 99 percent kill of agar-passaged L_. pneumophila was usually
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Legionella March 31, 1987
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achieved within 10 minutes, compared to 60 to 90 minutes for tap water-
maintained strains. These data suggest that normal chlorination practices
at treatment facilities may not control legionellae.
In contrast to these data, Hsu et al. (1984) reported that survivals of
L_. pneumophila and _E. coli in various concentrations of chlorine were similar.
In an in vitro study, laboratory-adapted strains of L_. pneumophila Flint 1
serogroup 1 and E. coli B were inoculated into several dilutions of sodium
hypochlorite in sterile tap water, and incubated at 24°C. At 0.2 mg/L residual
chlorine, about an order of magnitude reduction occurred in two hours for both
organisms. Neither organism could be recovered after two hours at concentration?
equal to or exceeding 2.0 mg/L. The pH values were not reported. The reason
for the discrepancy between this study and the Kuchta et al. (1983, 1984)
studies may be due to strain or pH differences.
Control of Legionellae in Plumbing Systems
Chlorine and Heat
Studies on controlling legionellae in plumbing systems have examined
primarily the effectiveness of heat and chlorine. The results of several of
these are described below.
In an attempt to eradicate L. pneumophila from showers in a transplantation
unit experiencing cases of Legionnaires Disease, Tobin et al. (1980) emptied
the hot and cold water tanks and filled them with water containing 50 mg/L
free chlorine. After three hours, this process was repeated. Shower fittings
were removed and held at 65°C for 18 hours before replacement. Legionellae
were not isolated from the shower samples after six months, but were found
again at nine months.
Massanari et al. (1984) controlled a nosocomial outbreak of L_. pneumophila
infection by shock chlorination (15 mg/L) of both hot and cold water supplies
for 12 hours. The system then was flushed and the hot water temperature
raised from 41°C to 64°C for 41 days. These measures significantly reduced
the frequency of positive cultures, but 3/35 of the outlets were still positive.
Thereafter, a continuous-flow proportional chlorination unit was installed
that provided free chlorine levels of 8 and 7.3 mg/L in hot and cold water,
respectively. During the first 16 months of its use, virtually no samples
(N=355) contained L_. pneumophila and no new cases of legionellosis were
identified. The few positive samples were obtained from rooms which had been
vacant for at least 32 days. In this hospital, water is distributed in
copper pipes.
Baird et al. (1984) hyperchlorinated their hospital water supply at a
constant level of 4 mg/L of free chlorine. The rate of nosocomial Legionnaires
Disease decreased by almost two-thirds and the total numbers of legionellae
decreased, but the organisms persisted.
Witherell et al. (1984) attempted to eradicate L_. pneumophila in hospital
plumbing by adding chlorine to the cold water make-up that supplied the hot
water heating system, in proportion to the water demands on the system. This
was to avoid corrosion damage resulting from constant feed chlorination units
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Legionella March 31, 1987
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during periods of low demand. A free chlorine residual of 3.0 mg/L was
maintained in the hot water system for 10 days and then reduced to 1.5 mg/L.
The organism was not detected by direct culture methods subsequent to disin-
fection. The corrosivity of the hot water increased slightly (Langelier
index = -0.3).
Fisher-Hoch et al. (1981) used hypochlorite to obtain a level of
1-2 mg/L of free chlorine at all cold water outlets in Kingston Hospital
where legionellae were present in both cold and hot water. The free chlorine
levels in the hot water could not be maintained above 0.2 mg/L and legionellae
were recoverable at this level. The water temperature was 45°C, which was
warm enough to volatilize the chlorine and cool enough to allow growth of
legionellae. Eradication was accomplished successfully by maintaining the
hot water temperature at 55°-60°C, in addition to disinfection of cold water.
Subsequently, these investigators reported that when a disconnected hot water
tank containing stagnant water was turned on again, L. pneumophila was found
in the water and a case of nosocomial Legionnaires Disease occurred (Fisher-Hoch
et al., 1982). A second disconnected tank which had been drained incompletely
contained a thick brown liquid deposit at the bottom. This deposit contained
5.4 x 108 _L. pneumophila/L. Filling the second tank with water containing
50 mg/L of chlorine for 24 hours followed by descaling did not successfully
eliminate the legionellae. Maintaining a constant water temperature of 70°C
throughout the tank for 1 hour, however, eliminated the organism. Ciesielski
et al. (1984) also noted that legionellae can proliferate in stagnant water
inside hot water tanks.
Dennis et al. (19.82) examined water samples from the plumbing of 52
hotels, none of which was associated with cases of legionellosis. Ten
isolates of L_. pneumophila were obtained from water samples from eight hotels.
Seven of these were from hot water taps or hot-cold mixer showers with water
temperatures ranging from 40° to 54°C at the time of sampling. Evidence that
these temperatures are not sufficient for Legionella control was also provided
by Meenhorst et al, (1983). In their study, guinea pigs exposed to aerosolized
legionellae from contaminated hot tapwater (48°C) contracted pneumonia. The
strain of I,, pneumophila used was isolated from a series of patients in the
Netherlands.
Beam et al. (1984) attempted to control legionellosis outbreaks in two
state development centers for the severely handicapped. In one center, hot
water tanks that were positive for legionellae were heated to 71°C for 72
hours, followed by flushing for 15 minutes. Because of legionellae regrowth,
a monthly heating schedule was established. Subsequently, the chlorine level
was raised from 0.5 mg/L to 2 mg/L. This approach was successful in eradi-
cating legionellae from water sources, but this chlorine level caused leaching
from the iron pipes and consequent discoloration of the water, and was thus
discontinued. Cement liners were installed in the hot water tanks and the
first samples were positive for legionellae. The water temperature was not
reported. Soon after, an outbreak of legionellosis occurred.
Plouffe et al. (1983) examined the relationship between the presence of
J^. pneumophila in potable water, nosocomial Legionnaires Disease, and hot
water temperatures in six buildings. L. pneumophila was found in the hot
water of all four buildings in which hot water was maintained at 43-49°C
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Legionella March 31, 1987
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(110°-120°F), and nosocomial Legionnaires Disease was found in three of these
buildings. No organism and no disease was found in the two buildings where
hot water was maintained at 57-60°C (135°-140°F). When the plumbing system
of one of the buildings experiencing both 1^. pneumophila and Legionnaires
Disease was flushed with 71°C water and the hot water then maintained at
57-60°C, no L^. pneumophila and no new cases of Legionnaires Disease occurred
for at least six months. The authors concluded that colonization and nosocomial
Legionnaires Disease can be ^evented by maintaining the hot water at 57-60°C.
In another attempt to eradicate L. pneumophila and nosocomial Legionnaires
Disease, Yu et al. (1982) raised the temperature in the hot water storage
tanks from 45° to 60°C for 72 hours and flushed 50 showers and 360 faucets
for 20 minutes with the 60°C water to eliminate the organism from the sediment.
A substantial reduction in counts occurred. After three months, colony counts
increased rapidly from four colonies/mL to over 300 colonies/mL and nosocomial
Legionnaires Disease again appeared. The authors concluded that a periodic
schedule of short-term temperature elevation of the hot water system may
control nosocomial Legionnaires Disease.
Stout et al. (1986) tested 75 legionellae isolates for their ability to
withstand high temperatures. Tubes containing buffered yeast extract broth,
sterile water, or hot water tank water plus sediment were inoculated and
placed in 60°C, 70°C or 80°C water baths. At 60°C, four minutes were required
for a one log reduction of L_. pneumophila in the water plus sediment tube.
Approximately 25 minutes were required at this temperature to sterilize a
suspension of _L. pneumophila which contains 1Q8 colonies/mL. The authors
recommend that when flushing distal outlets,, that a flush temperature exceeding
60°C should be maintained for at least 30 minutes.
Muraca et al. (1987) compared the relative efficacies of heat
(608C)," ozone (1-2 mg/L), UV (30,000 uW-scm2 at 254 run) and hyperchlorinatiori
(4-6 mg/L) to eradicate JL. pneumophila in a model plumbing system. Non-turbid
water at 25°C and 43°C and turbid water at 25°C were tested. When samples
were taken of the circulated water, a 5-log kill of a 107 bacteria/mL concen-
tration was achieved with all treatments within six hours. However, it is
noteworthy that heat completely eradicated the Legionella in less than three
hours, whereas UV light had produced its 5-log decrease in 20 minutes and
no further inactivation was seen during the six-hour observation period.
Chlorine and ozone required five hours to effect a similar 5-log decrease and
chlorine achieved complete eradication only in the non-turbid samples during
the six hours, while ozone killed the organisms in both turbid and non-turbid
water in four to five hours.
Ozone Treatment
Edelstein et al. (1982) used ozone in an attempt to eradicate legionellaie
from the potable water supply of an unused wing of a hospital that was known
to be contaminated with bacteria. The results were inconclusive because the
organisms were eliminated from both the experimental wing and the control
wing that was untreated. The latter was thought to be due to excess mechanical
flushing and an unexpected rise in. the chlorine content of the main water supply^
The in vitro susceptibility of L. pneumophila to ozone was on the order of
0.36 mg ozone/L, but was not consistent. The ozone mean residual concentration'
used in the hospital water system was 0.79 mg/L.
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Legionella March 31, 1987
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Ultraviolet Radiation Treatment
Antopol and Ellner (1979) reported that 90 percent of L^. pneumophila
cells in distilled water were killed by 920 microwatt-sec/cm2 of UV radiation.
This could be compared with exposures ranging from 2,100 to 5,000 microwatt-
sec/cm2 for killing of E. coli, Salmonella, Serratia and Pseudomonas. If
the latter values were obtained under the same conditions as those used for
L^. pneumophila, it would indicate that legionellae may be more than twice as
susceptible to UV radiation than are the other organisms.
Gilpin (1984) reported laboratory and field experiments using UV radiation
to inactivate Legionella spp. in standing and recirculating water systems.
Times of exposure to one microwatt/cm2 of UV radiation to produce 90 percent
killing of six species of Legionella ranged from 17 to 44 minutes. A commer-
cial UV apparatus killed 99 percent of the organism in less than 30 seconds
in a three-liter recirculating water system.
In addition, Knudson (1985) reported that when agar plates seeded with
_L« pneumophila were exposed to 240 microwatt/cm2 for 25 seconds or less, a
reduction of six to seven orders of magnitude was observed. However, when
UV-irradiated legionellae were exposed to indirect sunlight for 60 minutes,
the recovery rates were two orders of magnitude greater than those not exposed
to sunlight, due to photoreactivation.
Ethylene Oxide Treatment
Cordes et al. (1981) sterilized Legionella-contaminated showerheads with
ethylene oxide but they were soon recontaminated.
Design of Hot Water Tanks
Legionellae often have been reported in hot water tanks, particularly in
the bottom sediment. The design of these tanks is important in the control
of these bacteria. Most residential hot water tanks are heated from the
bottom near the cold water entrance pipe and are more likely to maintain a
bottom temperature high enough (>55°C) to prevent growth of legionellae.
However, if thermostats in homes have been set low (<55°C) as an energy
conservation measure, growth of legionellae may result. Thermostats for hot
water heaters in hospitals and other health care facilities are usually set
at lower temperatures in conformity with the recommendations of the Joint
Commission on Accreditation of Hospitals that the water temperature be "safe"
(JCAH, 1985). This practice, which is done to prevent scalding of patients
using the hot water, may promote the growth of legionellae. Larger institu-
tional tanks also are heated more often by internal steam coils or by other
heaters located midway from top to bottom of the tank. The water at the
bottom may not be heated sufficiently to kill legionellae. Periodic partial
draining of these tanks from the bottom to eliminate sediment may control
legionellae proliferation. This is especially important, since environmental
microflora in the sediment are known to produce metabolites, possibly including
cysteine, which stimulate legionellae growth (Stout et al., 1985). Removal
from other areas of the plumbing system where water stagnates may also prevent
or control legionellae growth (Stout et al., 1985).
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Legionella March 31, 1987
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Type of Water Fittings
Information on the specific types of gaskets and fittings that support
the colonization of legionellae is not well documented. One study of water
fittings as sources of L_. pneumophila in a hospital plumbing system was
carried out by Colbourne et al. (1984a, 1984b). In well-controlled experiments,
I». pneumophila was isolated from rubber washers and gaskets, but not from
liber ur plastic fittings. The ability of the bacteria to multiply when in
contact with the rubber fittings was demonstrated. When the rubber fittings
were replaced with plastic fittings, 1^. pneumophila could not be isolated up
to one year later. The authors concluded that shower and tap fittings that
support growth of legionellae provide habitats protected from chlorine and
heat. These foci may be seeded constantly or intermittently with legionellae
from hot water tanks or other amplifiers within the distribution system.
When to Control Legionellae in Plumbing Systems
Legionellae are often found in the plumbing systems of hospitals which
have not experienced any cases of Legionnaires Disease. One reason may be
that some strains are more virulent than others. Currently, there is no
practical method for distinguishing the virulent strains from avirulent
strains. For this reason, some experts feel that the mere presence of
legionellae in the absence of the disease is not sufficient grounds to under-
take control measures (Jakubowski et al., 1984). They believe that health
care institutions should focus initially on surveillance for respiratory
illness, especially in high risk patients, rather than to control legionellae
in plumbing systems. If nosocomial legionellosis is identified and environ-
mental strains match patient isolates, then control in plumbing systems is
indicated.
In contrast, Edelstein (1985) states that most authorities would probably
agree that disinfection of a contaminated site is indicated when:
0 it is implicated as a source of an outbreak of Legionnaires Disease
or Pontiac Fever;
0 it is present in a hospital ward housing especially high-risk patients,
such as an organ transplantation unit, regardless of epidemiological
findings; in this case, selective decontamination of certain ward
areas may be feasible; and when
0 it is found in a building which has not been used for some time and
in which the water has stagnated.
Because of the virulence of some of these strains and the fact that at
least 25,000 cases/year or more occur in the U.S., a stronger preventive
approach could also be supported.
In summary, there is no consensus on when measures should be undertaken
to control legionellae in the plumbing system of health care institutions.
Once virulence factors can be identified and virulent strains differentiated
from avirulent strains, routine monitoring of the plumbing system may become
more practical.
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Legionella March 31, 1987
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Until then, the Office of Drinking Water recommends that, on the basis of the
high incidence and mortality rate, health care institutions consider preventive
measures for the control of legionellae in their plumbing systems. 'These
measures could also control other opportunistic pathogens in the system which
might cause nosocomial infections.
Summary
Legionellae are abundant in ambient water, and may survive water treat-
ment, especially since they are relatively resistant to chlorine. Once in
the treated water, they then pass, probably at low levels, through the
distribution system. It is also possible that legionellae enter the distri-
bution system through broken or corroded piping, repair of existing mains,
installation of new mains, back siphonage, and cross connections. When
legionellae enter hot water tanks, they settle to the bottom and, under
certain circumstances, will proliferate. If they proliferate, plumbing
fixtures such as aerators, water fittings, and showerheads may be seeded,
resulting in colonization and growth at these sites.
Inhalation of aerosolized potable water has been suggested from outbreak
investigations as a primary route of infection, although ingestion is also
a possibility. The most susceptible individuals are those with underlying
diseases, especially those involving immunosuppression therapy. In several
outbreaks, however, apparently healthy individuals have developed legionellosis.
Other risk factors include alcohol abuse, surgery and smoking.
In order to reduce legionellae levels in drinking water, the presence of
organic matter and growth of algae and protozoa should be minimized in storage
reservoirs. Moreover, newly-repaired or constructed components of the water
distribution system should be flushed thoroughly and disinfected before being
put into operation. Even after flushing and disinfection, one cannot assume
legionellae have been controlled, since design factors in the distribution
system may impede the efficiency of these measures.
In order to control legionellae growth in hot water plumbing, several
approaches may be considered. Most of the published data have examined the
effectiveness of chlorine and/or heat. The maintenance of free chlorine has
been found effective for controlling legionellae. Shock chlorination also
is effective, but unless free chlorine is maintained within a system, the
organism may reappear. Control probably can be achieved if free chlorine
levels in the hot water are maintained at 8 mg/L, but at this level corrosion
of pipes may occur. In some cases, control may be achieved at 1.5-2 mg/L
free chlorine. Undoubtedly, the level of chlorine found effective will
depend, in part, on the design criteria of the plumbing system. A pertinent
facet in controlling legionellae is the difficulty of controlling batch
chlorination and of maintaining a chlorine residual in hot water. This
problem can be minimized by using a continuous-flow proportional chlorinator
in the hot water system.
Heat shock may eradicate legionellae in hot water tanks, if the temperature
at the bottom of the tank is maintained at 70°C for one hour, but this is a
temporary measure which must be done routinely to be effective. Maintenance
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Legionella March 31, 1987
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of hot water at 55°C or higher apparently controls the organism, while lower
temperatures may not. If legionellae are controlled by heat, care must be
taken to prevent scalding of persons using the water, especially in health
care institutions.
Disinfection of a plumbing system by heat treatment or chlorine treatment
alone may not be as effective as a combination of the two. For example,
growth of legionellae may theoretically be enhanced on the cold water side of
a hot-cold, water mixing valve in a heat-treated plumbing system, a location
where chlorine may be effective.
Effective disinfection of legionellae by ozone, ultraviolet radiation
or ethylene oxide has not been demonstrated by field tests.
In addition to chemical and heat disinfection, other procedures may be
effective in controlling legionellae. Hot water tanks should be designed
to give uniform temperatures throughout. Hot or cold water tanks used
intermittently should be disconnected from the system, drained, flushed,
and disinfected before being reconnected. Hot water tanks should be drained
regularly or at least bled to remove accumulated sludge that may serve as
a substrate for growth of legionellae and other microorganisms. Taps and
showers in unused areas of health care facilities should at least be flushed
before patients are exposed to them. Finally, faucet sieves and aerators,
and rubber washers and gaskets in the plumbing system should be used with
caution, especially in institutions housing physically compromised individuals
and where hot water is maintained at temperatures lower than 55°C.
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Legionella March 31, 1987
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REFERENCES
Antopol, S.C. and P.O. Ellner. 1979. Susceptibility of Legionella pneumophila
to ultraviolet radiation. Appl. Environ. Microbiol. 38:347-348.
Arnow, P.M. and D. Weil. 1984. Legionella pneumophila contamination of
residential tap water. In: C. Thornsberry, A. Balows, J.C. Feeley and
W. Jakubowski, eds. Legionella: Proceedings of the 2nd international
symposium, June 19-23, 1983; Atlanta, GA. Washington, DC: American
Society for Microbiology; pp. 240-241.
Baird, I.M., W. Potts, J. Smiley, N. Click, S. Schleich, C. Connole and
K. Davidson. 1984. Control of endemic nosocomial legionellosis by
hyperchlorination of potable water. In: C. Thornsberry, A. Balows,
J.C. Feeley and W. Jakubowski, eds. Legionella: Proceedings of the 2nd
international symposium, June 19-23, 1983; Atlanta, GA. Washington, DC:
American Society for Microbiology; pp. 333.
Beam, T.R., D. Moreton, T.A. Raab, W. Heaslip, M. Montes, J. Hanrahan,
M. Best and V.L. Yu. 1984. Epidemiology and control of Legionellaceae
in state developmental centers. In: C. Thornsberry, A. Balows,
J.C. Feeley and W. Jakubowski, eds. Legionella: Proceedings of the 2nd
international symposium, June 19-23, 1983; Atlanta, GA. Washington, DC:
American Society for Microbiology; pp. 236-237.
Ciesielski, C.A., M.J. Blaser and W-L.L. Wang. 1984. Role of stagnation and
obstruction of water flow in isolation of Legionella pneumophila from-
hospital plumbing. Appl, Environ. Microbiol. 48:984-987.
Colburne, J.S., M.G. Smith, S.P. Fisher-Hoch and D. Harper. 1984a. Source
of Legionella pneumophila infection in a hospital hot water system:
materials used in water fittings capable of supporting I*, pneumophila
growth. In: C. Thornsberry, A. Balows, J.C. Feeley and W. Jakubowski,
eds. Legionella: Proceedings of the 2nd international symposium,
June 19-23, 1983; Atlanta, GA. Washington, DC: American Society for
Microbiology; pp. 305-307.
Colburne, J.S., D.J. Pratt, M.G. Smith, S.P. Fisher-Hoch and D. Harper. 1984b.
Water fittings as sources of Legionella pneumophila in a hospital plumbing
system. Lancet 1:210-213.
Cordes, L.G., A.M. Wiesenthal, G.W. Gorman, J.P. Phair, H.M. Sommers,
A. Brown, V.L. Yu, M.H. Magnussen, R.D. Meyer, J.S. Wolf, K.N. Shands
and D.W. Fraser. 1981. Isolation of Legionella pneumophila from hospital
shower heads. Ann. Intern. Med. 94(2):195-197.
Dennis, P.J., J.A. Taylor, R.B. Fitzgeorge, C.L.R. Bartlett and G.I. Barrow.
1982. Legionella pneumophirla in wa-ter—plumbing syoteams.—Lancat—I-t-949-951 .
Edelstein, P.H. 1985. Environmental aspects of Legionella. ASM News
51:460-467.
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Legionella March 31, 1987
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Edelstein, P.H., R.E. Whittaker, R.L. Kreiling and C.L. Howell. 1982.
Efficacy of ozone in eradication of Legionella pneumophila from hospital
plumbing fixtures. Appl. Environ. Microbiol. 44:1330-1334.
Engelbrecht, R.S. 1983. Source, treatment, and distribution. In:
P.S. Berger and Y. Argaman, eds. Assessment of microbiology and turbidity
standards for drinking water. EPA 570/9-83-001. U.S. Environmental
Protection Agency, pp. 1-68.
Fisher-Hoch, S.P., C.L.R. Bartlett, J.O'H. Tdbin, M.B. Gillett, A.M. Nelson,
J.E. Pritchard, M.G. Smith, R.A. Swann, J.M. Talbot and J.A. Thomas.
1981. Investigation and control of an outbreak of legionnaires' disease
in a district general hospital. Lancet 1:932-936.
Fisher-Hoch, S.P., M.G. Smith, and J.S, Colbourne, 1982. Legionella pneumo-
phila in hospital hot water cylinders [Letter]. Lancet 1:1073.
Fliermans, C.B., W.B. Cherry, L.H. Orrison and L. Thacker. 1979. Isolation
of Legionella pneumophila from nonepidemic-related aquatic habitats.
Appl. Environ. Microbiol. 37:1239-1242.
Fliermans, C.B., W.B. Cherry, L.H. Orrison, S.J. Smith, D.L. Tison and
D.H. Pope. 1981. Ecological distribution of Legionella pneumophila.
Appl. Environ. Microbiol. 41:9-16.
Fliermans, C.B., G.E. Bettinger and A.W. Fynsk. 1982. Treatment of cooling
systems containing high levels of Legionella pneumophila. Water Res.
16:903-909.
Foy, H.M., P.S. Hayes, M.K. Cooney, C.V. Broome, I. Allan and R. Tobe. 1979.
Legionnaires' Disease in a.prepaid medical-care group in Seattle, 1963-756
Lancet 1:767-770.
Garbe, P.L., B.J. Davis, J.S. Weisfeld, L. Markowitz, P. Miner, F. Garrity,
J.M. Barbaree and A.L. Reingold. 1985. Nosocomial Legionnaires Disease:
epidemiologic demonstration of cooling towers as a source. J. Amer.
Med. Assoc. 254:521-524.
Gilpin, R.W. 1984. Laboratory and field applications of ultraviolet light
disinfection on six species of Legionella and other bacteria in water.
In: C. Thornsberry, A. Balows, J.C. Feeley and W. Jakubowski, eds.
Legionella: Proceedings of the 2nd international symposium, June 19-23,
1983; Atlanta, GA. Washington, DC: American Society for Microbiology;
pp. 337-339.
Haley, C.E., M.L. Cohen, J. Halter and R.D. Meyer. 1979. Nosocomial
legionnaires' disease: a continuing common-source epidemic at Wadsworth
Medical Center. Ann. Intern. Med. 90:583-586.
Helms, C.M., R.M. Massanari, R. Zeitler, S. Streed, M.J.R. Gilchrist, N. Hall,
W.J. Hausler, J. Sywassink, W. Johnson, L. Wintermeyer and W.J. Hierholzer.
1983. Legionnaires' disease associated with a hospital water system: a
cluster of 24 nosocomial cases. Ann. Intern. Med. 99(2):172-178.
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Legionella March 31, 1987
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Hsu, S.C., R. Martin and B.B. Wentworth. 1984. Isolation of Legionella
species from drinking water. Appl. Environ. Microbiol. 48:830-832.
Jakubowski, W., C.V. Broome, E.E. Geldreich and A.P. Dufour. 1984. Trans-
mission and control. In: C. Thornsberry, A. Balows, J.C. Feeley and
W. Jakubowski, eds. Legionella: Proceedings of the 2nd international
symposium, June 19-23, 1983; Atlanta, GA. Washington, DC: American
Society for Microbiology; pp. 351-355.
JCAH. 1985. Joint Commission on Accreditation of Hospitals. Accreditation
manual for hospitals; Chapter on plant, technology, and safety management.
JCAH, Chicago.
Knudson, G.B. 1985. Photoreactivation of UV-irradiated Legionella pneumophila
and other Legionella species. Appl. Environ. Microbiol. 49.:975-980.
Kuchta, J.M., S.J. States, J.E. McGlaughlin, R.M. Wadowsky, A.M. McNamara,
R.S. Wolford and R.B. Yee. 1984. Enhanced chlorine resistence of
tapwater grown Legionella pneumophila as compared with agar-passaged
strains. Abstracts, annual meeting of the American Society for Micro-
biology, St. Louis; Paper No. Q2.
Kuchta, J.M., S.J. States, A.M. McNamara, R.M. Wadowsky and R.B. Yee. 19.83.
Susceptibility of Legionella pneumophila to chlorine in tapwater.
Appl. Environ. Microbiol. 46:1134-1139.
Marks, J.S., T.F. Tsai, W.J. Martone, R.C. Baron, J. Kennicott, F.J. Holtzhauer,
I. Baird, D. Fay, J.C. Feeley, G.F. Mallison, D.W. Fraser and T.J. Halpin.
1979. Nosocomial legionnaires' disease in Columbus, Ohio. Ann. Intern.
Med. 90(4):565-569.
Massanari, R.M., C. Helms, R. Zeitler, S. Streed, M. Gilchrist, N. Hall,
W. Hausler, W. Johnson, L. Wintermeyer, J.S. Muhs and W.J. Hierholzer.
1984. Continuous hyperchlorination of potable water system for control
of nosocomial Legionella pneumophila infections. In: C. Thornsberry,
A. Balows, J.C. Feeley and W. Jakubowski, eds. Legionella: Proceedings
of the 2nd international symposium, June 19-23, 1983; Atlanta, GA.
Washington, DC: American Society for Microbiology; pp. 334-336.
Meenhorst, P.L., A.L. Reingold, G.W. Gorman, J.C. Feeley, B.J. van Cronenburg,
C.L.M. Meyer and R. van Furth. 1983. Legionella pneumonia in guinea
pigs exposed to aerosols of concentrated potable water from a hospital
with nosocomial legionnaires' disease. J. Infect. Dis. 147(1):129-132.
Muraca, P, J.E. Stout and V.L. Yu. 1987. Comparative assessment of chlorine,
heat, ozone and UV light for killing Legionella pneumophila within a
model plumbing system. Appl. Environ. Microbiol. 53:447-453.
Plouffe, J.F., L.R. Webster, B. Hackman and M. Macynski. 1983. Hot water
temperature, Ij. pneumophila (LP) and nosocomial legionnaires; disease
(LD). Abstracts, annual meeting of the American Society for Microbiology,
New Orleans; paper L16.
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Legionella March 31, 1987
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Spino, D.F., E.w. Rice and E.E. Geldreich. 1984. Occurrence of Legionella
spp. and other aquatic bacteria in chemically contaminated ground water
treated by aeration. In: C. Thornsberry, A. Balows, J.C. Feeley and
W. Jakubowski, eds. Legionella: Proceedings of the 2nd international
symposium, June 19-23, 1983; Atlanta, GA. Washington, DC: American
Society for Microbiology; pp. 318-320.
Stout, J.E., M.G. Best and V.L. Yu. 1986. Susceptibility of members of the
family Legionellaceae to thermal stress: implications for heat eradication
methods in water distribution systems. Appl. Environ. Microbiol.
52:396-399.
Stout, J., V.L. Yu, R.M. Vickers, J. Zuravleff, M. Best, A. Brown, R.B. Yee
and R. Wadowsky. 1982. Ubiquitousness of Legionella pneumophila in the
water supply of a hospital with endemic Legionnaires Disease. New Eng.
J. Med. 306:466-468.
Stout, J.E., V.L. Yu and M.G. Best. 1985. Ecology of Legionella pneumophila
within water distribution systems. Appl. Environ. Microbiol. 49:221-228.
Tison, D.L. and R.J. Seidler. 1983. Legionella incidence and density in
potable drinking water supplies. Appl. Environ. Microbiol. 45:337-339.
Tobin, J.O'H., J. Beare, M.S. Dunnill, S. Fisher-Hoch, M. French, R.G. Mitchell,
P.J. Morris and M.F. Muers. 1980. Legionnaires' disease in a transplant
unit: isolation of the causative agent from shower baths. Lancet 11:118-121|
Wentworth, B.B., W.A. Chadwick, H.E. Stiefel and D.S. Benge. 1984. Preva-
lence of antibody to various Legionella species in ill and healthy
populations in Michigan. In: C. Thornsberry, A. Balows, J.C. Feeley
and W. Jakubowski, eds. Legionella: Proceedings of the 2nd international
symposium, June 19-23, 1983; Atlanta, GA. Washington, DC: American
Society for Microbiology: pp. 255-257.
Wadowsky, R.M., R.B. Yee, L. Mezmar, E.J. Wing and J.N. Dowling. 1982. Hot
water systems as sources of Legionella pneumophila in hospital and
nonhospital plumbing fixtures. Appl. Environ. Microbiol. 43:1104-1110.
Witherell, L.E., L.A. Orciari, R.W. Duncan, K.M. Stone and J.M. Lawson. 1984.,
Disinfection of hospital hot water systems containing Legionella pneumo-
phila . In: C. Thornsberry, A. Balows, J.C. Feeley and W. Jakubowski,
eds. Legionella: Proceedings of the 2nd international symposium,
June 19-23, 1983; Atlanta, GA. Washington, DC: American Society for
Microbiology: pp. 336-337.
Yu, V.L., M. Best, J. Stout, A. Brown and A. Goetz. 1982. Effectiveness of
intermittent short-term temperature elevation of the hospital water supply
in controlling nosocomial Legionnaires' Disease (LIU.—Ahstrsnts, annual—
meeting of the American Society for Microbiology, Atlanta; paper L19.
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March 31, 1987
BARIUM
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data "describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Barium
March 31, 1987
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This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for barium (U.S.
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd.,
Springfield, VA 22161, PB #86-118031/45. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.
0 Barium —
Barium Chloride — 10361-37-2
Barium Sulfate — 7727-43-7
Synonyms
Barium Sulfate; Barite (Windholz, 1976)
Uses
0 Depending upon the specific compound, barium salts are used for a
number of purposes including drilling mud (Kirkpatrick, 1978), pigment
(Miner, 1969), and as x-ray contrast nedium (Miner, 1969). Other
uses are summarized by Pidgeon (1964).
Properties (Pidgeon, 1964; Preisman, 1964; Miner, 1969; Chilton, 1973;
Kirkpatrick, 1978; Reeves, 1979)
0 The properties of barium compounds vary with the specific compound;
some examples are as follows:
Chemical Formula
Atomic/Molecular Weight
Physical State
Boiling Point
Melting Point
Density (20°C)
Vapor Pressure
Water Solubility (pph)
Log Octanol/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Barium
Barium
Chloride
Ba Bad 2
137.33 208.24
Silver-white solid White solid
1637-1 638°C
729-7308C
3.6 g/cm3
1810 x 10-5 mm Hg
reacts
1560°C
960°C
3,856 g/cm3
31 (0°C)
Barium
Sulfate
BaS04
233.40
Colorless solid
1580°C
4.50 g/cm3
0.000285 (30°C)
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Barium March 31, 1987
-3-
Occurrence
0 Barium is a reactive metal which is not found free in nature but
exists as a number of salts. Barium occurs in nature chiefly as the
mineral barite (33804) and ^-n mucn smaller amounts as witherite
(83003). TJie ">ineral forms are relatively insoluble in water, having
high melting and boiling points and very low vapor pressures (Preisman,
1964'. Barium compounds occur in most geologic materials at levels
of 300-500 ppm. Barium occurs at low levels in most surface and
ground waters with reported levels of less than 340 ug/L. While
barium compounds are used commercially in a number of processes,
contamination of drinking water is usually the result of naturally -
occurring barium and not industrial releases (U.S. EPA, 1987).
0 There are limited survey data on the occurrence of barium in drinking
water. Most supplies contain less than 200 ug/L of barium. Currently,
60 ground water supplies and 1 surface water supply exceed the interim
maximum contaminant level (MCL) of 1,000 ug/L. Barium also occurs in
most foods as a low level contaminant. Based upon the limited infor-
mation available on barium exposure, food is the major source of
barium exposure (U.S. EPA, 1987).
III. PHARMACOKINETICS
Absorption
0 In laboratory animals, the absorption of barium varies with a number
of factors including the species of animal (U.S. EPA, 1985), the
compound tested (McCauley and Washington, 1983), the age of the animal
(Taylor et al., 1962) and the composition of the diet (Lengemann, 1959),
While no definitive human barium absorption studies were found (U.S.
EPA, 19851, barium absorption has been estimated to be approximately
5% in the adult (ICRP, 1973). However, other data (Harrison et al.,
1967) suggest that barium absorption probably is greater than this.
While data in laboratory animals (Lengemann, 1959) suggest that barium
absorption in children may be significantly greater than in adults,
there is currently inadequate information to resolve this issue.
Distribution
0 In the mouse, intravenously injected barium (133BaCl2) is distributed
widely throughout the organism, but is localized principally in the
bone (Dencker et al., 1976).
° Based on autopsy data, barium levels in human bone are relatively
constant and do not appear to increase with age, ranging from an
average value of 7.0 ppm in bone at age 0 to 3 months to an average
of 8.5 ppm at age 33 to 74 years (Sowden and Stitch, 1957).
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Barium March 31, 1987
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Metabolisrn
0 The skeletal metabolism of barium in humans is qualitatively similar
to that of calcium, although the incorporation of these two elements
is quantitatively very different (Bauer et al., 1956,1957).
Excretion
In humans, ingested barium is eliminated principally via fecal excretion
(approximately 72%) following oral exposure (Tipton et al., 1966).
IV. HEALTH EFFECTS
Humans
Acute barium toxicity is associated with hypokalemia and electrocar-
diographic changes as well as other symptoms (Diengott et al., 1964;
Gould et al., 1973; Talwar and Sharma, 1979).
NAS (1977) has concluded that: "The fatal dose of barium chloride for
man has been reported to be about 0.8 - 0.9 g, or 550 - 600 mg of
barium."
Schroeder and Kraemer (1974) concluded that there was a significant
negative correlation between barium,in drinking water and athero-
sclerotic heart disease.
In an epidemiology study, Brenniman et al. (1981) concluded that
there was no statistically significant difference in blood pressure
between those ingesting drinking water containing barium at 7.3 mg/L
as compared to 0.1 mg/L. A concentration of 7.3 mg/L corresponds to
a dose of 0.20 mg/kg/day (assuming a 70-kg adult drinks 2 L per day).
The duration of exposure was not identified.
Animals
Short-term Exposure
The acute oral LD50 of barium varies markedly with species, compound,
age and other factors (U.S. EPA, 1985). For example, the acute oral
LD50 of barium chloride is 220 mg/kg in weanling rats and 132 rag/kg in
adult rats (Tardiff et al., 1980).
Long-term Exposure
Tardiff et al. (1980) exposed rate to barium at 0, 10, 50, or 250
ppm in drinking water for 4, 8 and 13 weeks. The barium concentrations
were approximately 0, 2.75, 13.7 and 66.25 mg/kg/day at the beginning
of the study and 0, 1.7, 6.6 and 31.5 mg/kg/day at the end of the
study. Although the barium body burden increased with increasing
barium dosage, no conclusive signs of barium toxicity were observed
in these animals. A weakness of this study is that, unlike Perry
et al. (1983) below, blood pressure was not measured.
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Barium March 31, 1987
0 Perry et al. (1983) exposed weanling rats to barium at 1, 10 or 100
ppm in drinking water for up to 16 months (average daily barium doses
of 0.051, 0.51 and 5.1 mg/kg, respectively). With the exception of
an increase in blood pressure, there were no signs of toxicity at
any barium dose level. Systolic blood pressure measurements revealed
no increase in pressure in animals exposed to 1 ppm for 16 months,
an increase of 4 mm Hg (p ~.01) in animals exposed to 10 ppm barium
for 16 months, and an increase in systolic pressure of 16 mm Hg (p
<0.001) in animals exposed to 100 ppm barium for 16 months. The
animals in this study were maintained in a special contaminant-free
environment and fed a diet designed to reduce exposure to trace
metals. It is possible that the restricted intake of certain beneficial
metals (e.g., Ca and K) may have predisposed the test animals to the
hypertensive effects of barium (U.S. EPA, 1985).
0 Schroeder and Mitchener (1975a,b) exposed rats and mice to 5 mg/L
barium in drinking water for a lifetime (approximately 0.25 mg/kg/day
for rats and 0.825 mg/kg/day for mice). No compound related adverse
effects were observed. A weakness of this study is that, unlike
Perry et al. (1983) above, blood pressure was not measured.
Reproductive Effects
0 No adequate mammalian study on the potential reproductive effects
of barium was identified (U.S. EPA, 1985).
Developmental Effects
0 No adequate mammalian study on the potential developmental effects of
barium was identified (U.S. EPA, 1985).
Muta^enicity
0 No adequate studies on the mutagenicity of barium were identified
(U.S. EPA, 1985).
Carcinogenicity
0 No adequate studies on the carcinogenicity of barium were identified
(U.S. EPA, 1985).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = „ ( „ ,
(OF) x ( L/day)
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Barium March 31, 1987
-6-
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in rag/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg .
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
The available data are insufficient to develop a One-day HA for barium.
It is recommended that the modified DWEL of 0.51 mg/L (adjusted for a 10-kg
child) be used as the One-day HA for the 10-kg child.
Ten-day Health Advisory
The available data are insufficient to develop a Ten-day HA for barium.
It is recommended that the modified DWEL of 0.51 mg/L (adjusted for a 10-kg
child) be used as the Ten-day HA for the 10-kg child.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for barium.
It is recommended that the DWEL of 1.8 mg/L be used as the Longer-term HA
for the 70-kg adult and the modified DWEL of 0.51 nig/L (adjusted for a 10-kg
child) be used as the Longer-term HA for the 10-kg child.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
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Barium March 31, 1987
-7-
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
Considering the kind, nature and partially contradictory results of the
various barium studies, ODW does not believe that it is appropriate to use
any simplistic formula to determine a Lifetime HA for barium in drinking
water. Rather:
0 No single study, considered alone, is appropriate to calculate a Lifetime
HA for barium.
0 A barium HA must be based on the weight of all the pertinent data,
considered together.
In the Perry et al. (1983) rat drinking water study, 10 ppm barium
(0.51 mg/kg/day) produced a small (4 to 7 mm Hg) but statistically significant
increase in blood pressure by 8 to 16 months; 100 ppm barium (5.1 mg/kg/day)
produced clear hypertension and cardiotoxic effects.
A major shortcoming of the Perry, et al. (1983) study is that the animals
were maintained in a special environment and received both a special diet and
special water, all intended to reduce exposure to trace metals. Because of
the beneficial effects of some metals (i.e., cadmium) and the interactions
of barium with other metals, it is possible that the restricted intake of
other metals may have contributed to the apparent toxicity of barium. In
addition, the results of the Perry, et al. (1983) rat study clearly contradict
the results of the Brenniman, et al. (1981) human study which suggests that
barium in drinking water has no appreciable effect upon blood pressure in
humans, at least at a level of 7.3 ppm (0.20 mg/kg/day) in drinking water.
While the 4 to 7 mm Hg increase in blood pressure observed at 10 ppm
barium (0.51 mg/kg/day) in the Perry, et al. (1983) study may be a compound
related effect, ODW has serious doubts as to whether this 4 to 7 mm Hg increase
in rat blood pressure should be considered an adverse effect in the light of
the negative effects observed in the Brenniman, et al. (1981) human study.
Considering the contradiction between the rat and human data, it was ODW's
judgment that it was not prudent either to ignore the results of Perry et al.
(1983) or to treat the results with the same seriousness they would warrant,
had they been observed in humans.
In ODW's judgment, the most appropriate way to balance the contradiction
between the Perry, et al. (1983) rat study and the Brenniman, et al. (1931)
human study is to use the results of the Perry, et al. (1983) study, with a
reduced uncertainty factor of 10x and to treat the 0.51 mg/kg/day value as
if it were a NOAEL.
Thus, based on the previous discussion, the Lifetime Health Advisory
is derived as follows:
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Barium March 31, 1987
-8-
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.51 mg/kg/day) = 0>051 mg/kg/day
(10)
where:
0.51 mg/kg/day = NOAEL (see discussion above).
10 = uncertainty factor, chosen in accordance with NA3/ODW
guidelines for use with a NOAEL from an animal study•
However, as previously discussed, ODW believes that
an uncertainty factor of 10 is appropriate in this
specific case (see discussion above).
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.051 mg/kg/day) (70 kg) = K8 mg/L (T,800 ug/L,
(2 L/day)
where:
0.051 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (1.8 mg/L) (33%) =1.5 mg/L (1,500 ug/L)
where:
1.8 mg/L = DWEL.
83% = assumed relative source contribution from water (Federal
Register, November 13, 1985).
Evaluation of Carcinogenic Potential
" Due to the absence of toxicological evidence to classify barium as
a potential carcinogen, a quantification of carcinogenic risks for
barium is not appropriate.
0 No information was located in the available literature regarding the
carcinogenic potential of barium in humans nor were any animal studies
found which were adequate to evaluate the carcinogenic potential of
barium.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986) barium is classified in Group 0:
Not classified. This category is for agents with inadequate animal
and human evidence.
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Barium March 31, 1987
-9-
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of barium.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The National Interim Primary Drinking Water Regulations of 1975
established a Maximum Contaminant Level (MCL) drinking water standard
for barium of 1 mg/L (U.S. EPA, 1976).
0 The National Academy of Sciences (NAS, 1982) derived a 1-day Suggested
No-Adverse-Response Level (SNARL) for barium of 6.0 mg/L.
0 The National Academy of Sciences (NAS, 1982) derived a chronic Sug-
gested No-Adverse-Response Level (SNARL) value for barium of 4.7 mg/L.
0 The American Conference of Governmental Industrial Hygienists estab-
lished an occupational threshold limit value (TLV) of 0.5 mg/m3 for
barium nitrate in air (ACGIH, 1980).
8 The USSR standard for waterborne barium is 4 mg/L (NAS, 1977).
0 The OSHA 8-hour time-weighted average exposure limit for soluble
barium compounds is 0.5 mg/m3 in workplace air (OSHA, 1985).
VII. ANALYTICAL METHODS
0 Determination of barium is by atomic absorption (AA) using either
direct aspiration into a flame (U.S. EPA, 1979a) or a furnace technique
(U.S. EPA, 1979b).
0 The direct aspiration AA procedure is a physical method based on the
absorption of radiation at 553.6 nm by barium. The sample is
aspirated into an air-acetylene flame and atomized. A light beam is
directed through the flame into a monochromator, and onto a detector
that measures the amount of light absorbed. Absorbance is proportional
to the concentration of barium in the sample. The detection limit is
100 ug/L using this procedure.
0 The furnace AA procedure is similar to direct aspiration AA except a
furnace, rather than a flame, is used to atomize the sample. The
detection limit is 2 ug/L using this procedure.
VIII. TREATMENT TECHNOLOGIES
Experience indicates that ion exchange, lime softening and reverse
osmosis are effective to remove barium from drinking water. Conven-
tional coagulation/filtration processes are not effective to remove
barium from drinking water (U.S. EPA, 1977).
Weinberg (1973) and Logsdon et al. (1974) reported that ion exchange
softening systems are highly efficient (93 to 98 percent) for reducing
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Barium March 31, 1987
-10-
barium in water, even after water hardness breakthrough. Field
data from two Midwestern full-scale ion exchange softening plants
showed that barium removal was comparable to hardness removal on well
water containing 11-19 mg/L of barium and 225-230 mg/L of hardness as
CaCo3 (BIF, 1970). When these softening units were performing ef-
ficiently and removing all of the hardness from the water, they also
removed all of the barium.
0 Experience indicates that lime softening is very effective in removing
barium from drinking water. Lime softening achieved greater than
90 percent removal in the 10-11 pH range on well water containing
7-8.5 mg/L of naturally occurring barium. Removals decreased below
and above this range. Pilot plant studies conducted at the EPA
Municipal Research Laboratory and full-scale treatment information on
similar types of ground water verified the laboratory data. Pilot
plant test runs on water containing 10-12 mg/L of barium at pH 9.2,
10.5 and 11.6 resulted in removals of 84, 93 and 32 percent, respec-
tively. Grab samples from two full-scale lime softening plants
showed removals of 88 and 95 percent. These plants operated at pH
10.5 and 10.3; the raw water barium concentrations were measured
at 7.5 and 17.4 mg/L, respectively (BIF, 1970).
0 A number of studies indicate that reverse osmosis membranes can remove
more than 90 percent of the barium from drinking water. In an experi-
mental long term study, 99 percent barium removal was obtained using
cellulose acetate membrane at 400-800 psi operating pressures (3IF,
1970). Other studies by Sorg et al. (1980) achieved similar results,
where 95-99 percent removals were obtained by passing water containing
7 mg/L barium through cellulose acetate membranes at 165-180 psi
operating pressures.
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Barium March 31, 1987
-1 1-
IX. REFERENCES
ACGIH. 1980. American Conference of Governmental Industrial Hygienists.
Threshold limit values for chemical substances and physical agents in
the workroom environment with intended changes for 1980. Cincinnati,
Ohio: American Conference of Governmental Industrial Hygienists. p. 35.
Bauer, G.C.H., A. Carlsson and B. Lindquist. 1957. Metabolism of 140ga in
man. Acta. Orth. Scand. 26:241-254.
Bauer, G.C.H., A. Carlsson and B. Lindquist. 1956. A comparative study of
the metabolism of 140Ba and 45Ca in rats. Biochem. J. 63:535-542.
BIF. 1970. Chemicals Used on Treatment of Water and Waste Water Engineering
Data. (Unit of General Signal Corp., Providence, RI) brochure. May.
Brenniman, G.R., W.H. Kojola, P.S. Levy, B.w. Carnow and T. Namekata. 1981.
High barium levels in public drinking water and its association with
elevated blood pressure. Arch. Environ. Health. 36(1):28-32.
Dencker, L., A. Nilsson, C. Ronnback and G. Walinder. 1976. Uptake and
retention of 133ga and 140Ba-T4C>La in mouse tissue. Acta Radiol.
15(4):273-287.
Diengott, D., O. Rozsa, N. Levy and S. Muammar. 1964. Hypokalemia in
barium poisoning. Lancet 2:343-344.
Federal Register, November 13, 1985, Vol. 50, No. 219, pp 46936-47022.
Gould, D.B., M.R. Sorrell and A.D. Luperiello. 1973. Barium sulfide poison-
ing. Arch. Intern. Med. 132:891-894.
Harrison, G.E., T.E.F. Carr and A. Sutton. 1967. Distribution of radioactive
calcium, strontium, barium and radium following intravenous injection
into a healthy man. Int. J. Radiat. Biol. 13(3):235-247.
ICRP. 1973. International Commission on Radiological Protection. Alkaline
earth metabolism in adult man. ICRP Publication 20. Health Phys.
24:125-221.
Kirkpatrick, T. 1978. Barium compounds. In: Kirk-Othmer encyclopedia of
chemical technology, 3rd ed., Vol. 3. New York: John Wiley and Sons.
pp. 463-479.
Lengemann, F.W. 1959. The site of action of lactose in the enhancement of
calcium utilization. J. Nutrition, 69:23-27.
Logsdon, G.S., Sorg, T.J. et al. 1974. Removal of Heavy Metals by Conven-
tional Treatment. Proceedings, 16th Water Quality Conference. Trace
Metals in Water Supplies: Occurrence, Significance and Control.
University of Illinois.
-------
Barium March 31, 1987
-12-
McCauley, P.T., and I.S. Washington. 1983. Barium bioavailability as the
chloride, suifate or carbonate salt in the rat. Drug Chem. Toxicol.
6(2):209-217.
Miner, S. 1969. Air pollution aspects of barium and its compounds. Techni-
cal Report. Bethesda, Md.: Litton Systems, Inc. Contract No. PH-22-
68-25. 69 pp.
NAS. 1977. National Academy of Sciences. Drinking Water and health.
Vol. 1. Washington, D.C.: National Academy Press, pp. 207-305.
National Academy of Sciences. 1982. Drinking Water and Health, Vol. 4.
Washington, D.C.: National Academy Press, pp. 167-170.
OSHA. 1985. Occupational Safety and Health Administration. Code of Federal
Regulations. Title 29 - Labor. Part 1910 - Occupational Safety and
Health Standards. Subpart Z - Toxic and Hazardous Substances. Section
1910.1000 - Air Contaminants. U.S. Government Printing Office,
Washington, DC.
Perry, R.H., and C.H. Chilton. 1973. Chemical engineers' handbook, 5th ed.
New York: McGraw-Hill Book Co. pp. 3-8 - 3-9.
Perry, H.M., S.J. Kopp, M.W. Erlanger and E.F. Perry. 1983. Cardiovascular-
effects of chronic barium ingestion. In: Hemphill, D.D., ed. Trace
substances in environmental health-XVII. Proceedings of University of
Missouri's 17th annual conference on trace substances in environmental
health, Columbia, MO: University of Missouri Press. pp. 155-164.
Pidgeon, L.M. 1964. Barium. In: Kirk-Othmer encylopedia of chemical
technology. 2nd ed. Vol. 3. John Wiley and Sons, New York. pp. 77-80.
Preisman, L. 1964. Barium compounds. In: Kirk-Othmer encylopedia of
chemical technology. 2nd ed. Vol 3. John Wiley and Sons, New York.
pp. 80-98.
Reeves, A.L. 1979. Barium. In: L. Friberg, G.F. Nordberg and V.B. Vouk,
eds. Handbook on the toxicology of metals. Amsterdam: Elsevier/Nortii
Holland Biomedical Press, pp. 321-328.
Schroeder, H.A., and L.A. Kraemer. 1974. Cardiovascular mortality, municipal
water and corrosion. Arch. Environ. Health. 28:303-311.
Schroeder, H.A., and M. Mitchener. 1975a. Life-term effects of mercury,
methyl mercury and nine other trace metals on mice. J. Nutr. 105:452-458,
Schroeder, H.A., and M. Mitchener. 1975b. Life-term studies in rats: effects
of aluminum, barium, beryllium and tungsten. Jo Nutr. 105:421-427.
Sorg, T.J., and Logsdon, G.S. 1980. Treatment technology to meet the interim
primary drinking water regulations for inorganics: Part 5. AWWA.
72(7):411-22.
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Barium March 31, 1987
-13-
Sowden, E.M., and S.R. Stitch. 1957. Trace elements in human tissue. 2.
Estimation of the concentrations of stable strontium and barium in
human bone. Biochem. J. 67:104-109.
Talwar, K.K., and B.K. Sharma. 1979. Myocardial damage due to barium chlond«
poisoning. Indian Heart J. 31(4):244-245.
Tardiff, R.G., M. Robinson and N.S. Ulmer. 1980. Subchronic oral toxicity
of BaCl2 in rats. J. Environ. Path. Tox. 4:267-275.
Taylor, D.M., P.M. Bligh and M.H. Duggan. 1962. 'The absorption of calcium,
strontium, barium and radium from the gastrointestinal tract of the rat.
Biochem. J. 83:25-29.
Tipton, I.H., P.L. Stewart and P.G. Martin. 1966. Trace elements in diets
and excreta. Health Phys. 12:1683-1689.
U.S. EPA. 1976. U.S. Environmental Protection Agency. National interim
primary drinking water regulations. EPA 570/9-75-C03. Washington, D.C.:
U.S. Environmental Protection Agency.
U.S. EPA. 1977. U.S. Environmental Protection Agency. Manual of treatment
techniques for meeting the interim primary drinking water regulations,
revised. U.S. Environmental Protection Agency, EPA-600/8-77-005.
U.S. EPA. 1979a. U.S. Environmental Protection Agency. Method 208.1.
Atomic Absorption, direct aspiration, In: Methods for Chemical Analysis
of Water and Wastes. EPA-600/4-79-020, March.
U.S. EPA. 1979b. U.S. Environmental Protection Agency. Method 208.2.
Atomic Absorption, furnace technique, In: Methods for Chemical Analysis
of Water and Wastes. EPA-600/4-79-020, March.
U.S. EPA. 1983. U.S. Environmental Protection Agency. Barium occurrence in
drinking water, food, and air. Office of Drinking Water.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Draft health effects
criteria document for barium. CSD, Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogenic risk assessment. Federal Register. 51(185):33992-34003.
September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Estimated national
occurrence and exposure to barium in public drinking water supplies.
CSD. Office of Drinking Water.
Weinberg, L.M. 1973. Report of analytical evaluation and treatability
study. For Wight Consulting Engineers on Lake Zurich Water Well *5.
CHEMED Corp., Dearborn Environmental Engineers, July.
Windholz, M., ed. 1976. The Merck-Index: An encyclopedia of chemicals and
drugs, 9th ed. Rahway, NJ: Merck and Co., Inc. p. 995.
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March 31, 1S87
CADMIUM
Health Advisory Draft
Office of Drinking Water
U.S. Environmental Protection Agene/
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse healtn effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assis.t Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Cadmium
March 31, 1987
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This Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for cadmium (U.S.,
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd.,
Springfield, VA 22161, ?B #86-117942/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.
Cadmium — 7440-43-9
Cadmium Chloride — 10108-64-2
Cadmium Oxide — 1306-19-0
0 Cadmium is used for a number of purposes including the following
(Stubbs, 1978): batteries, electroplating, stabilizer, pigments,
and as an alloy with other metals.
Properties (Schindler, 1967; Weast, 1971; IARC, 1976; Parker, 1978)
0 The properties of cadmium compounds vary with the specific compound;
some examples are as follows:
Chemical Formula
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure (400°C)
Water Solubility
Log Octanol/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Cadmium
Cd
Atomic/Molecular Weight 112.40
Cadmium
Chloride
CdCl2
183.32
Soft white solid Solid
765°C
320.9°C 568°C
8.642 g/cm3 4.047 g/cm3
1.4 mmHg
Soluble
Cadmium
Oxide
CdO
128.40
Solid
1,559°C
900°C
8.15 g/cm3
Insoluble
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Cadmium March 31, 1987
-3-
Occurrence
0 Cadmium is a naturally-occurring metallic element, present in most of
the earth's crust at levels below 1 ppm. Cadmium is commercially
obtained as a byproduct during the production of zinc. Commercial
uses of cadmium and its compounds include metal plating, electronics,
paints, and pigments. Cadmium is released to the environment during
its uses and from other commercial activities. However, these releases
have not resulted in the contamination of ground and surface waters
(U.S. EPA, 1987).
0 Naturally occurring levels of cadmium in surface and ground water
normally fall in the range of 1-10 ug/L. State monitoring data have
reported that 21 ground water supplies and 4 surface water supplies
currently exceed 10 ug/L. Cadmium occurs at low levels in food and
air. The FDA Total Diet Study reports that adults currently receive
34 ug/day of cadmium from their diets. Based upon this information
food appears to be the major route of exposure for cadmium (U.S. EPA,
1987).
0 Cadmium is found in both cigarettes and cigarette smoke and as the
absorption of inhaled cadmium can approach levels as high as 96%
(CEC, 1978), smoking can account for a substantial fraction of the
body burden of cadmium (Ellis et al., 1979)
III. PHARMACOKINETICS
Absorption
0 The absorption of cadmium following oral administration to laboratory
animals, and presumably humans, is modified by many factors including
dose (Engstrom and Nordberg, 1979), age (Kostial et al., 1983), diet
(Suzuki et al., 1969) and by the presence of other metals such as
calcium (Washko and Cousins, 1976).
0 Cadmium does not readily cross the skin (CEC, 1978).
0 Cadmium is very readily absorbed following inhalation; as much as 96%
of the cadmium deposited in the lungs may be absorbed (CEC, 1978).
Distribution
0 In both rats (Sabbioni et al., 1978) and humans (Sumunio et al.,
1975), cadmium distributes throughout the body and accumulates in
the kidney and liver where it may attain levels 10 to 100 times
greater than those of other tissues.
Metabolism
Whole cadmium is not metabolized to other compounds as is the typical
organic drinking water contaminant; once within the body, cadmium
readily combines with the the low molecular weight protein(s)
metallothionein (Foulkes, 1982).
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Cadmium March 31, 1987
-4-
Excretion
Once absorbed, cadmium is eliminated in humans principally via the
urine (U.S. EPA, 1985).
Cadmium is eliminated very slowly in humans; a half-life for elimina-
tion of cadmium has been estimated to be 10 to 33 years (Ellis et al.,
1979). The long half-life of cadmium in humans is principally accounted
for by the marked accumulation and retention of cadmium in the kidney
and liver (Friberg et al., 1974)
In humans, average body retention of radiolabelled cadmium chloride,
measured one to five weeks post exposure, was approximately 4.6%
(McLellan et al., 1978).
IV. HEALTH EFFECTS
Humans
In humans, the symptoms of cadmium toxicity following acute exposure
include nausea, vomiting, diarrhea, muscular cramps and salivation
(Arena, 1963). In the case of severe intoxication, sensory distur-
bances, liver injury and convulsions may result, which, in fatal
intoxications, are followed by shock and/or renal failure and cardio-
pulmonary depression (CEC, 1978).
1 The estimated acute lethal dose of cadmium is 350 to 35,000 mg for a
70-kg adult (CEC, 1978).
' For emesis, the NOAEL for cadmium in adults is 0.043 mg/kg/day followin:
an acute oral exposure to cadmium salts (Lauwerys, 1979).
Chronic non-occupational oral exposure to very high levels of cadmium
has resulted in such adverse health effects as the Itai-Itai disease
observed in Japan (principally in multiparous women), which is
characterized by pain, osteomalacia, osteoporosis, proteinuria,
glucosuria, and anemia (U.S. EPA, 1985).
' While it has been suggested that cadmium may play a role in hyper-
tension, there is considerable uncertainty concerning what, if any,
role cadmium may play in this disease. (Perry et al., 1977a and b;
Kopp et al., 1982, 1983).
0 Renal toxicity (e.g. proteinuria) following low level chronic oral
exposure to cadmium is believed to be the most sensitive manifestation
of cadmium toxicity (CEC, 1978; U.S. EPA, 1985). It has been estimated
that the concentration at which 10% of the population is likely to
display signs of renal dysfunction is 180 to 220 ug Cd/g renal cortex.
Individuals with values over 285 ug/g usually display signs of renal
dysfunction (U.S. EPA, 1985).
3 Friberg et al. (1974) hypothesized that renal damage may occur when,
over a 50 year period, a person's daily cadmium intake equals or
exceeds 0.352 mg/day.
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Cadmium March 31, 1987
-5-
Animals
Short-term Exposure
0 The acute oral LDjQ of cadmium compounds in the rat varies with the
compound and ranges from 16 mg/kg for cadmium cyanide to > 5,000 mg/kg
for cadmium sulfide (CEC, 1978).
0 Toxic effects, resulting from oral exposure to various cadmium compounds,
have been observed in a variety of animal tissues (U.S. EPA, 1985)
including the nervous system (Gabbiani et al., 1967), kidney (CEC,
1978), liver (Stowe et al., 1972), bone (Larsson and Piscator, 1971),
hematopoietic system (Stowe et al., 1972), cardiovascular system
(Kopp et al., 1978) and immune system (Koller, 1973).
Long-term Exposure
0 Cadmium-induced renal toxicity (e.g. proteinuria) has been observed
in animals in the absence of renal histopathology (CEC, 1978).
0 In a 24-week male rat drinking water study, animals exposed to 2.15
and 6.44 mg cadmium/kg/day developed a significant level of proteinuria
(P <0.05), while animals exposed to the lowest level tested, 0.84 mg
cadmium/kg/day (NOAfiL), did not develop proteinuria (Kotsonis and
).
12 month rat drinking water study, no adverse effects were
observed in animals exposed to 0.008, 0.035, 0.181, 0.361 or 0.375
(NOEL) mg cadmiumAg/day . However, at three months, the animals
exposed to the highest level tested, 3.04 mg cadmium/kg/day, developed
anemia and did not gain weight normally (Decker et al., 1958).
Reproductive Effects
0 In a rat oral study, cadmium was administerted at 0, 0.1, 1.0 and
10.0 mg cad mi urn/kg/day (as CdC^) respectively, to groups of male and
female adult rats for six weeks; males and females were mated for
three weeks, and cadmium was administered during the mating period;
pregnant females were given cadmium during the gestation period. In
the 10 mg/kg group, the number of total implants and live fetuses
decreased significantly (p <0.05) while resorptions increased signifi-
cantly (p <0.01); fetuses showed decreased body weight (p <0.05) and
delayed ossification of the sternebrae and caudal vertebrae. No
effects were observed at 0.1 or 1.0 mg cadmium/kg/day (Sutou et al.,
198°).
Developmental Effects
8 In a rat drinking water study, fetal growth retardation was observed
in animals whose dams were exposed to 100 mg cadmium/L but not in
those exposed to 0.1 or 10 mg cadmium/L during gestation (Ahokas
et al., 1980).
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Cadmiam March 31, 1987
-6-
Mutagenicity
0 While cadmium has been observed to cause chromosomal aberrations in
several in vitro studies (e.g., Watanabe et al., 1979, and Di Paolo
and Casto, 1979), strong evidence of mutagenic effects following oral
ingestion is not available (U.S. EPA, 1985).
Carcinogenicity
0 Cadmium and cadmium compounds have been shown to induce sarcomas at
local injection sites (Haddow et al., 1964; Gunn et al., 1967). In
addition, cadmium chloride administered to rats by aerosol for 18
months has produced lung tumors (Takenaka et al., 1983). These data
are not believed relevant to the consumption of cadmium in drinking
water (U.S. EPA, 1985).
0 Although cancers of the prostate and lung have been noted in cadmium
smelter workers in an epidemiological study (Lenten et al., 1976),
evidence regarding the carcinogenicity of cadmium in humans following
oral exposure is largely conjectural (U.S. EPA, 1985).
0 No evidence of cadmium oncogenicity has been found in chronic oral
animal studies (Schroeder et al., 1965; Kanisawa and Schroeder,
1969; Loser, 1980).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = /L ( „ j
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
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Cadmium March 31, 1987
-7-
One-day Health Advisory
The study by Lauwerys (1979) was selected to serve as the basis for the
One-day HA for cadmium. In this study, the NOAEL for cadmium-induced emesis
in adult humans following a single dose of cadmium was 0.043 mg cadmium/kg/day.
This study was selected because it is of appropriate duration and was conducted
in the most appropriate species, humans; more suitable data are not available.
The HA for a 10-kg child is calculated as follows:
One-day HA = (0.043 mg/kg/day) (10 kg) = 0.043 mg/L (43 ug/L)
(10) (1 L/day)
where:
0.043 mg cadmium/kg/day = NOAEL for emesis following acute exposure
to adults (Lauwerys, 1979).
10 kg = assumed body weight of a child.
10 = uncertainty factor, chosen in accordance with
NAS/ODW guidelines for use with a NOAEL from
a human study.
1 L/day = assumed daily water consumption of a child.
Ten-day Health Advisory
A 24-week oral exposure study in rats (Kotsonis and Klaassen, 1978) was
considered for use as the basis of the Ten-day HA. In this study a NOAEL of
0.84 mgAg/day was identified for proteinuria. If this NOAEL and an uncertainty
factor of 100 were used, the Ten-day HA value would be 0.08 mg/L. This value
is not markedly different from the One-day HA of 0.043 mg/L (based on a study
which demonstrated cadmium-induced emesis in adult humans). However, since
the Ten-day HA value of 0.08 mg/L would be greater than the One-day HA value,
it is recommended that the more conservative One-day HA of 0.043 mg/L (43 ug/L)
be used as the Ten-day HA.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for
cadmium. It is recommended that the DWEL of 18 ug/L be used as the Longer-
term HA for the 70-kg adult and the modified DWEL of 5 ug/L (adjusted for a
10-kg child) be used as the Longer-term HA for the 10-kg child.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
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Cadmium March 31, 1987
-8-
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that mediam, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
There are no adequate oral exposure studies in humans which provide a
NOAEL for the chronic effects of cadmium. Friberg et al. (1974) concluded
that the critical concentration of cadmium in the renal cortex of humans
associated with renal dysfunction is 200 ug/g wet weight; this is supported
by the recent reassessment by Kjellstrom et al. (1984). The 200 ug/g critical
concentration was based on a comprehensive review of evidence from animal
experiments and from analyses of kidneys from workers occupationally exposed
to cadmium. The 200 ug/g value is probably the most widely accepted estimate
of the critical concentration for renal dysfunction (NA3, 1977; CEC, 1978).
However, Roels et al. (1983) reported that the critical concentration in the
human renal cortex is 216 ug/g tissue wet weight and that less than 10% of
occupationally exposed males may develop renal dysfunction at this concentration,
Several models have been proposed to estimate the daily intake (exposure)
of cadmium required to produce the critical concentration in the renal cortex.
Each model has inherent limitations. Friberg et al. (1974) estimated that a
daily cadmium intake of 0.352 mg/day for 50 years would result in a renal
cortex concentration of 200 ug/g. This model assumes 4.5% absorption of the
daily oral dose and 0.01% excretion per day of the total body burden, both
reasonable estimates. Thus, 0.352 mg of cadmium per day in a 70-kg adult
(0.005 mgAg/day) is a reasonable estimate of the daily cadmium intake that
would result in renal dysfunction. In that the Friberg et al., (1974) value
of 0.005 mg/kg/day is associated with renal dysfunction, 0.005 mg/kg/day is a
LOAEL value which normally would require that an uncertainty factor of 100 be
used. However, considering the relatively low level of uncertainty concerning
cadmium toxicity in this case, it is judged that an uncertainty factor of 100
is unreasonably high and that an uncertainty factor of 10 is more appropriate.
As previously discussed, the study by Friberg et al., (1974) is the
most appropriate from which to derive the Lifetime Health Advisory. From
these results, a LOAEL of 0.005 mg/kg was identified. Using this LOAEL, the
Lifetime Health Advisory is derived as follows:
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Cadmium March 31, 1987
-9-
Step 1: Determination of the Reference Dose (RfD)
RfD = (0.005 mg/kg/day) = 0.0005 mgAg/day
(10)
where:
0.005 mgAg/day = LOAEL based on renal dysfunction in humans.
10 = uncertainty factor; this uncertainty factor, while
smaller than would normally be required by NA3/ODW
guidelines, was judged to best reflect the uncertainty
concerning cadmium toxicity in humans.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0*0005 mgAg/day) (70 kg) _ 0.018 mg/L (18 ug/L)
(2 L/day)
where:
0.0005 mgAg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (0.018 mg/L) (25%) = 0.005 mg/L (5 ug/L)
where:
0.018 mg/L = DWEL.
25% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 A quantitative evaluation of the carcinogenicity of cadmium has not
been conducted since there is no conclusive evidence that cadmium is
carcinogenic following oral exposure.
0 U.S. EPA has recommended that cadmium not be considered a suspect
human carcinogen for the purpose of calculating an ambient water
quality criterion (U.S. EPA, 1980).
0 Based on exposure to cadmium via inhalation, IARC (1982) has
classified cadmium and certain cadmium compounds in Group 2B: Limited
evidence for carcinogenicity in humans, sufficient evidence for
carcinogenicity in animals.
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Cadmium March 31, 1987
-10-
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), cadmium, on the basis of inhalation
data, may be classified in Group B1: Probable human carcinogen. This
category is for agents for which there is inadequate evidence from
human studies and sufficient evidence from animal studies. However,
as there are inadequate data to conclude that cadmium is carcinogenic
via ingestion, cadmium is dealt with here as Group D: Not classified.
This category is for agents with inadequate animal evidence of
carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE, AND STANDARDS
0 The National Academy of Sciences (NAS, 1982) has calculated a one-day
Suggested No-Adverse Response Level (SNARL) of 0.150 mg/L for cadmium
in drinking water for 70-kg adults.
0 The National Academy of Sciences (NAS, 1982) has calculated a seven-day
SNARL of 0.021 mg/L of drinking water for 70-kg adults.
0 The National Academy of Sciences (NAS, 1982) has calculated a chronic
exposure SNARL of 0.005 mg/L for cadmium in drinking water for 70-kg
adults; this value is based on the assumption that water contributes
20% of the daily cadmium intake.
0 A FAO/WHO expert committee has proposed a provisional tolerable weekly
standard of no more than 57.1 to 71.4 ug Cd/week (WHO, 1972).
0 The World Health Organization (WHO, 1984) has recommended that the
concentration of cadmium in drinking water not exceed 0.005 mg/L.
0 The Commission of the European Communities (CEC, 1975) has recommended
that the concentration of cadmium in drinking water not exceed O.OOE mg/1.
0 The current U.S. EPA primary drinking water standard for cadmium is
0.010 mg/L of drinking water (U.S. EPA, 1976).
0 The recommended threshold limit values (TLVs) for cadmium dusts,
salts and oxide fumes for occupational eight hour time-weighted
average exposure is 0.05 mg/m3 (ACGIH, 1980).
0 The OSHA 8-hour time-weighted average exposure limit for cadmium fume
is 0.1 mg/m3 in workplace air; the acceptable ceiling concentration
for cadmium fume is 0.3 mg/m3 (OSHA, 1985).
0 The OSHA 8-hour time-weighted average exposure limit for cadmium dust
is 0.2 mg/m3 in workplace air; the acceptable ceiling concentration
for cadmium dust is 0.6 mg/m3 (OSHA, 1985).
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Cadmium March 31, 1987
-1 1-
VII. ANALYTICAL METHODS
0 Determination of cadmium is by atomic absorption (AA) using either
direct aspiration into a flame (U.S. EPA, 1979a) or a furnace technique
(U.S. EPA, 19795).
0 The direct aspiration AA procedure is a physical method based on the
absorption of radiation at 228.8 nm by cadmium. The sample is
aspirated into an air-acetylene flame and atomized. A light beam is
directed through the flame into a monochromator and onto a detector
that measures the amount of light absorbed. Absorbance is proportional
to the concentration of cadmium in the sample. The detection limit
is 5 ug/L using this procedure.
0 The furnace AA procedure is similar to direct aspiration AA except a
furnace, rather than a flame, is used to atomize the sample. The
detection limit is 0.1 ug/L using this procedure.
VIII. TREATMENT TECHNOLOGIES
0 Effective removal of cadmium from source waters may be achieved with
treatment methods such as coagulation with alum or iron salts, lime
softening, ion exchange and reverse osmosis. Laboratory experiments
and pilot plant studies indicate that the effectiveness of cadmium
removal by coagulation is pH dependent. Ferric sulfate coagulation
studies on river water containing 0.3 mg/L of cadmium showed removals
to increase from 20 % at pH 7.2 to above 90 % at pH 8 and above.
Alum coagulation results on river water also increased with pH, but
the data indicated that, above pH 8, removals may depend on che
turbidity of the raw water. In some tests with low turbidity water
(1-10 jtu), removals decreased as the pH increased (U.S. EPA, 1978).
0 Experience indicates that lime softening is capable of achieving
cadmium removal from water greater than 98 \ in the pH range in
well water containing 0.3 mg/L of cadmium. Removals equally as good
were obtained at pH 11.2-11.3 when the initial cadmium concentration
was increased up to 10 mg/L (U.S. EPA, 1978).
0 There are limited performance data on the use of ion exchange as a
treatment method for removal of cadmium from drinking water. The
plating industry uses ion exchange for reducing cadmium in wastewaters
and other wastewater streams studied have successfully used ion
exchange for removing cadmium (Lindstedt et al., 1976; Nippon, 1976;
Amax, 1977; Laszlo, 1977; Ameron, 1978). However, there is one report
of 99 % removal efficiency for cadmium from drinking water using a
home ion exchange softener (Personal communications, Ciccone
Engineering, V.J. from Culligan Co., 1982). Tap water spiked
with 0.10 mg/L of cadmium chloride and used as feed water to a cation
exchanger on the sodium cycle produced product water with a cadmium
level less than 0.01 mg/L.
-------
Cadmium March 31, 1987
-12-
Experience indicates that reverse osmosis can effectively remove
cadmium from drinking waters. A study by Mixon (1973) showed a 90
and 9.8% cadmium removal, respectively, from 0.10 mg/L and 0.98 mg/L
spiked water samples, using three laboratory-scale cellulose acetate
membranes operated at 400 psi. No difference in cadmium rejection
was noted when barium, chromium, copper, lead and zinc were introduced.
Another study by Hindin et al. (data) indicated a 70 percent removal
for cadmium concentrations of 0.097, 0.959 and 9.25 mg/L using a
laboratory size reverse osmosis cellulose acetate cell. A study
performed by Huxstep (1982) in Florida related to inorganic con-
taminant removal from potable water by reverse osmosis resulted in a
96-98 % removal of cadmium.
Protection against cadmium from corrosion of water distribution
systems, in general, may be achieved by a number of methods including
pH adjustment, addition of lime, increasing alkalinity, or addition
of phosphates or silicates. The extent and type of treatment
selection is dependent on the characteristics of the water and the
compatibility of existing treatment with regard to various materials
used to convey the water through the distribution system.
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Cadmium March 31, 1987
-13-
IX. REFERENCES
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Documentation of the threshold limit values, 4th ed. Cincinnati, OH:
American Conference of Governmental Industrial Hygienists, pp. 59-61.
Ahokas, R.A., P.V. Dilts and E.B. LaHaye. 1980. Cadmium-induced fetal
growth retardation: protective effect of excess dietary zinc. Am. J.
Obstet. Gynecol. 136:216-226.
Amax, Inc. 1977. Removal of metal ions from wastewater. U.S. Patent
4,025,430, submitted January 12, 1976. May 24.
Ameron, Inc. 1978. System for removal of toxic heavy metals from drinJcing
water. U.S. Patent 4,096,064, submitted April 5, 1976. June 20.
Arena, J.M. 1963. Poisoning: chemistry, symptoms and treatment. Spring-
field, IL: Charles C. Thomas, p. 127.
CEC. 1975. Commission of the European Communities. Proposal for a council
directive relating to the quality of water human consumption. J. Official
European Communities. 18:2-17.
CEC. 1978. Commission of the European Communities. Criteria (dose/effect
relationships") for cadmium. Oxford: Permagon Press, pp. 1-198.
Decker, C.F., R.U. Byerrum and C.A. Hoppert. 1957. A study of the distribution.
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66:140-145.
DiPaolo, J.A., and B.C. Casto. 1979. Quantitative studies of in vitro
morphological transformation of Syrian hamster cells by inorganic
metal salts. Cancer Res. 39:1008-1013.
Ellis, K.J., D. Vartsky, I. Zanzi, S.H. Cohn and S. Yasumura. 1979. Cadmium:
in vivo measurement in smokers and nonsmokers. Science. 205:323-325.
Engstrom, B., and G.F. Nordberg. 1979. Dose dependence of gastrointestinal
absorption and biological half-time of cadmium in mice. Toxicology.
13:215-222.
Foulkes, B.C., ed. 1982. Biological roles of metallothionein. New York:
Elsevier/North-Holland.
Friberg, L., M. Piscator, G.F. Nordberg and T. Kjellstrom. 1974. Cadmium
in the environment, 2nd ed. Boca Raton, Florida: CRC Press Inc.
Gabbiani, G., A. Gregory and D. Baic. 1967. Cadmium-induced selective
lesions of sensory ganglia. J. Neuropath. Exp. Neur. 26:498-506.
Gunn, S.A., T.C. Gould and W.A.D. Anderson. 1967. Specific response of
mesenchymal tissue to carcinogenesis by cadmium, Arch. Pathol.
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Cadmium March 31, 1987
-14-
Haddow, A., F.J.C. Roe, C.E. Dukes and B.C.V. Mitchley. 1964. Cadmium
neoplasia: Sarcomata at the site of injection of cadmium sulphate in
rats and mice. Brit. J. Cancer. ,18:667-673.
Hindin, E., G.H. Dunstan et al. Water reclamation by reverse osmosis.
Bulletin 310, Washington State University.
Huxstep, M.R. 1982. Inorganic contaminant removal from potable water by
reverse osmosis (Task 49AS, Treatment of Small Community Water Supplies
by Reverse Osmosis). Charlottee Harbor (FL) Water Association, Inc.,
Progress Report, January 1 - March 31, 1982. U.S. Environmental
Protection Agency.
IARC. 1976. International Agency for Research on Cancer. Monographs on tne
evaluation of carcinogenic risk of chemicals to man. Cadmium, nickel,
some epoxides, miscellaneous industrial chemicals and general considera-
tions on volatile anesthetics, Vol. 11. Lyon: International Agency for
Research on Cancer, pp. 39-74.
IARC. 1982. International Agency for Research on Cancer. IARC monographs on
the evaluation of the carcinogenic risk of chemicals to humans. Suppl.
4: 133-135.
Kanisawa, M., and H.A. Schroeder. 1969. Life term studies on the effect
of trace elements on spontaneous tumors in mice and rats. Cancer Res.
29:892-895.
Kjellstrom, T., C.G. Blinder and L. Friberg. 1984. Conceptual problems in
establishing the critical concentration of cadmium in human kidney
cortex. Env. Res. 33:284-295.
Roller, L.D. 1973. Immunosuppression produced by lead, cadmium and mercury,,
Am. J. Vet. Res. 34:1457-1458.
Kopp, S.J., V.W. Fisher, M. Erlanger, E.F. Perry and H.M. Perry. 1978.
Electrocardiographical, biochemical and morphological effects of chronic
low level cadmium feeding on rat heart. Proc. Soc. Exp. Biol. Med.
159:339-345.
Kopp, S.J., T. Glonek, H.M. Perry, M. Erlanger and E.F. Ferry. 1982.
Cardiovascular actions of cadmium at environmental exposure levels.
Science. 217:837-839.
Kopp, S.J., H.M. Perry, E.F. Perry and M. Erlanger. 1983. Cardiac physio-
logic and tissue metabolic changes following chronic low-level cadmium
and cadmium plus lead ingestion in the rat. Toxicol. Appl. Pharmacol.
69:149-160.
Kostial, K., I. Simonovic, I. Rabar, M. Blanusa and M. Landeka. 1983.
Age and intestinal retention of mercury and cadmium in rats. Environ.
Res. 31:111-115.
-------
Cadmium March 31, 1987
-15-
Kotsonis, F.N., and C.D. Klaassen. 1978. The relationship of metallonthionein
to the toxicity of cadmium after prolonged oral administration to rats.
Toxicol. Appl. Pharmacol. 46:39-54.
Larsson, S.E., and M. Piscator. 1971. Effect of cadmium on skeletal tissue
in normal and cadmium-deficient rats. Isr. J. Med. Sci. 7:495-498.
Laszlo, M. 1977. Process for removing heavy metals from fluid media.
U.S. Patent 4,060,410, submitted July 7, 1975. November 29.
Lauwerys, R. 1979. Cadmium in man. In: Webb, ed. The chemistry, biochem-
istry and biology of cadmium. Elsevier/North Holland Biomedical Press,
pp. 433-453.
Lemen, R.A., J.S. Lee, J.K. Wagoner and H.P. Blejer. 1976. Cancer mortality
among cadmium production workers. Ann. NY Acad. Sci. 271:273-279.
Loser, E. 1980. A 2-year oral carcinogenicity study with cadmium on rats.
Cancer Lett. 9:191-198.
Linstedt, K.D., C.P. Houck et al. 1971. Trace element removals in advanced
wastewater treatment processes. Journal WPCP. 43(7):1507-13.
McLellan, J.S., P.R. Flanagan, M.J. Chamberlain and L.S. Valberg. 1978.
Measurement of dietary cadmium absorption in humans. J. Toxicol. Environ.
Health. 4:131-138.
Mixon, F.O. 1973. Removal of toxic metals from water by reverse osmosis.
R&D Progress Report No. 889. U.S. Department of Interior, Office of
Saline Water.
NAS. 1977. National Academy of Sciences. Drinking Water and Health. Volume 1.
Washington, DC: National Academy Press, p. 939.
NAS. 1982. National Academy of Sciences. Drinking Water and Health. Volume 4.
Safe Drinking Water Committee. Washington, D.C.: National Academy Press,
pp. 170-174.
Nippon Electric Co., Ltd. 1976. Improvements in or relating to the extraction
of heavy metals from industrial wastewaters. British Patent 1,457,528,
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OSHA. 1985. Occupational Safety and Health Administration. Code of Federal
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1910.1000 - Air Contaminants. U.S. Government Printing Office,
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Parker, P.O. 1978. Cadmium compounds. In: Kirk-Othmer, encyclopedia of
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-------
Cadmium March 31, 1987
-16-
Perry, H.M., M. Erlanger and E.F. Perry. 1977a. Elevated systolic pressure
following chronic low-level cadmium feeding. Am. J. Physiol. 232:H114-
H121.
Perry, H.M., M. Erlanger and E.F. Perry. 1977b. Hypertension following
chronic, very low dose cadmium feeding. Proc. Soc. Exp. Biol. Med.
156:173-176.
Personal communication between V.J. Ciccone Engineers and Culligan, August 4,
1982.
Ribelin, W.E. 1963. Atrophy of rat testis as index of chemical toxicity.
Arch. Pathol. 75:229-235.
Roels, R., R. Lauwerys and A.N. Dardenne. 1983. The critical level of
cadmium in human renal cortex: a re-evaluation. Toxicol. Letters.
15:357-360.
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Cadmium toxicity studies under long term-low level exposure (LLE) con-
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Equilibrium concepts in natural water systems. Adv. in Chem. Series 67,
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Schroeder, H.A., J.J. Balassa and W.H. Vinton. 1965. Chromium, cadmium and
lead in rats: effects on life span, tumors and tissue levels. J. Nutr.
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Stowe, H.Do, M. Wilson and R.A. Goyer. 1972. Clinical and morphologic
effects of oral cadmium toxicity in rabbits. Arch. Pathol. 94:389-405=
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Sumino, K., K. Hayakawa, T. Shibata and S. Kitamura. 1975. Heavy metals in
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tility, teratogenicity and dominant lethal tests in rats and administered
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upon the retention rate of orally administered 115Cd C12 in mice with
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Health. 7:155-162.
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Takenaka, S., H. Oldiges, H. Konig, 0. Hochrainer and G. Oberdorstar. 1983.
Carcinogenic!ty of cadmium chloride aerosols in W rats. JNCI. 70:367-373,
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U.S. EPA. 1978. U.S. Environmental Protection Agency. Manual of treatment
techniques for meeting the interim primary drinking water regulations,
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U.S. EPA. 1979a. U.S. Environmental Protection Agency. Water Method 213.1.
Atomic Absorption, direct aspiration. In: Methods for chemical analysis
of water and wastes. EPA-60/4-79-020, March.
U.S. EPA. 1979b. U.S. Environmental Protection Agency. Method 213.2,
Atomic Absorption, furnace technique. In: Methods for chemical analysis
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drinking water criteria document on cadmium. Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51(185):33992-34003.
September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Occurrence of cadmium
in public water supplies. CSD. Office of Drinking Water.
Washko, P.W., and R.J. Cousins. 1976. Metabolism of 109Cd in rats fed normal
and low-calcium diets. J. Tox. Environ. Health. 1:1055-1066.
Watanabe, T., T. Shimada and A. Endo. 1979. Mutagenic effects of cadmi am
on mammalian oocyte chromosomes. Mutation Res. 67:349-356.
Weast, R.C., ed. 1971. CRC handbook of chemistry and physics, 52nd ed.
Cleveland, OH: The Chemical Rubber Co.
WHO. 1972. World Health Organization. Evaluation of certain food additives
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WHO. 1984. World Health Organization. Guidelines for drinking water quality
— recommendations. Volume 1. Geneva: World Health Organization.
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March 31, 1987
CHROMIUM
Health Advisory
Office of Drinking Water
U.S. Ehvironmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Chromium March 31, 1987
-2-
This Health Advisory is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for chromium (U.S.
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd., Spring-
field, VA 22161, PB #86-113072/AS. The toll-free number is (800) 336-4700;
in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS NO.
0 Chromium — 7440-47-3
Chromium (III) Chloride — 10025-73-7
Chromic Acid, Dtpotassium Salt — 7789-00-6
Synonyms
0 None
Uses
Chromium and its salts have a variety of uses including the following
(for additional information see Hartford, 1979):
0 Hexavalent chromium compounds are used widely in industry for chrome
alloy and chromium metal production, for metal finishing and corrosion
control (Love, 1947) and as mordants in the textile industry (Her,
1954).
0 Chromium salts are used as anticorrosive agents in cooling waters, in.
the leather tanning industry, in the manufacture of catalysts, in
pigments and paints, and in fungicides and wood preservatives (Hartford,
1979).
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Cn r om i urn
March 31, 1987
-3-
Properties (Hem, 1970; Weast, 1971; Windholz, 1976}
" The properties of chromium compounds vary with the specific compound;
some examples are as follows:
Chromium
Chromium (III)
Chloride
Chromic Acid,
Dipotassium Salt
Chemical Formula
Atomic/Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Log OctanoI/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Occurrence
Cr
51.996
blue-white solid
2,642°C
1,900°C
7. 1 4 gm/cm3
0.5 ug/L
CrCl3
122.90
solid
83 °C
194.20
solid
968. 3 °C
2.76 g/cm3 (15°C) 2.732 g/cm3 (18°C)
inslouble 62.9 g/100 mL (20°C)
Chromium is a relatively rare, naturally occurring element in the
earth's crust. Chromium occurs in most rocks and minerals at levels
of 200 ppm. A few minerals contain chromium at levels of 2-3,000
ppm. Chromium is not mined in the U.S. commercially; it is imported.
Chromium is released to the environment during industrial activities.
However, current data suggest that surface and ground water levels of
chromium are the result of naturally-occurring chromium leaching from
mineral deposits. Soluble chromium has been reported to occur in
surface waters at levels up to 84 ug/L and in ground water at levels
of 50 ug/L (U.S. EPA, 1987).
Federal surveys of surface and ground water drinking water supplies
have reported that most supplies contain less than 5 ug/L. Currently,
17 ground water supplies and one surface water supply exceed the
interim standard of 50 ug/L (U.S. EPA, 1987).
III. PHARMACOKINETICS
Absorption
In general, with the exception of the Cr III glucose tolerance factor
(GTF), Cr VI is more readily absorbed than Cr III:
0 In humans and experimental animals, gastrointestinal absorption of
inorganic salts of Cr III is low (from 0.5% to 3°s). However, Cr VI
and organic complexes of Cr III are more readily absorbed (approxi-
mately 2% to 10% for Cr VI and 10% to 25% for organic complexes of
Cr III) (U.S. EPA, 1985).
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Chromium March 31, 1987
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0 In humans administered 20 ng of Cr III as 51CrCl3 in water, 0.5% of the
dose was recovered in the urine, indicating little absorption (Donald-
son and Berreras, 1966). In rats, Mertz et al. (1965) reported 2% to
3% absorption of Cr(III) based on total body counting of animals admin-
istered ^CrCl^ by intubation at doses ranging from 1.5 to 100 ug/kg.
0 GTF, an organic complex of Cr III with nicotinic acid and an ami no
acid that is found in brewer's yeast, was absorbed in rats at 10% to
25% of the administered dose (Mertz, 1976; Mertz et al., 1978).
0 An estimate of 2.1% absorption of Cr VI based on recovery in urine was
reported for humans administered 20 ng of Na251Cr04 in water (Donaldson
and Barreras, 1966).
0 Rats administered drinking water containing 25 mg/L Cr III as chromic
chloride had 12.5 times greater tissue levels of chromium than rats
whose drinking water contained 25 mg/L Cr VI as potassium chromate
(Mackenzie et al., 1958).
Distribution
Depending on the particular compound (e.g., GTF) Cr III and Cr VI
differ in their distribution within an organism; in general Cr III crosses
membranes much more slowly than Cr VI (U.S. EPA, 1985):
0 Chromium circulates in the plasma primarily.in a nondiffusible form.
A small fraction (9% to 12%) is in a more diffusible form which is
filtered and partially reabsorbed in the kidney (Collins et al.,
1961). An approximate plasma half-life of 6 hours for 51cr III in
rats was reported by Hopkins (196S) after intravenous administration
of either 0.1 or 1.0 ug/kg.
0 Cr III has an affinity for iron-binding proteins (Gray and Sterling,
1950; Hopkins and Schwarz, 1964).
0 The spleen and kidneys were shown to have the highest concentrations
of chromium when rats were administered Cr III as chromium chloride
in intravenous doses of 0.1 or 1.0 ug/kg (Hopkins, 1965). Similar
results were reported by Mackenzie et al. (1958) when rats received
drinking water containing 25 mg/L of either Cr III as chromic
chloride or Cr VI as potassium chromate. The calculated doses were
1.87 mg/kg/day for males and 2.41 mg/kg/day for females.
0 The placenta appears to be highly selective in its permeability to
the various forms of chromium. Inorganic Cr III administered as
51CrCl3 (chromium chloride) intravenously or by stomach intubation
does not cross the placental barrier to an appreciable extent in rats
(Mertz et al., 1969). However, Cr III administered by stomach
intubation to pregnant rats in the form of GTF (obtained from yeast)
is recovered readily from the fetus (Mertz and Roginski, 1971). The
dosages in these two studies were unspecified.
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Chromium March 31, 1987
-5-
0 Cr VI traverses biological membranes by diffusion or facilitated
transport, possibly via an anion transport system (Alexander at al.,
1982). It is reduced to Cr III intracellularly by the cytochrome
P-450 system in the presence of NADPH. Cr ill reacts with nucleophilic
ligands and cellular macrocnolecules (Gruber and Jennette, 1978).
Metabolism
0 The metabolism of chromium in mammalian species is not well under-
stood and is complicated by the presence of the two oxidation states,
Cr III and Cr VI (U.S. EPA, 1985).
Excretion
The kidney appears to be the principal route of excretion of chromium
compounds:
0 The urinary system is the major excretory route for absorbed chromium,
accounting for 80% or more of chromium excretion (Kraintz and Talmage,
1952). Very little is known about the form in vhich chromium is
excreted.
0 After intravenous administration, chromium is also excreted in the
feces, although reports in the literature vary considerably on the
percentage. Hopkins (1965) reported that 0.5% to 1.7% of the initial
dose of Cr III was excreted in the feces of rats eight hours after
intravenous administration of 51CrCl3 at 0.1 ug/100 g.
IV. HEALTH EFFECTS
Humans
In general, Cr VI compounds are more toxic than Cr III compounds:
0 The toxicity of chromium has been attributed primarily to Cr VI, which
has been shown to produce liver and kidney damage, internal hemorrhage,
dermatitis and respiratory problems. The immediate symptoms are
generally nausea, repeated vomiting and diarrhea (U.S. EPA, 1985).
0 Doses of 0.5 to 1.5 g of K2Cr2O7 have been fatal in humans. The
estimated LDLO for K^r2O7 in children is 26 mg/kg (Cr VI at 9.2
mg/kg) (NIOSH, 1983).
0 Subchronic and chronic dermal exposure to Cr VT in the form of chromic
acid may cause contact dermatitis and ulceration of the skin (Barrows,
1978). For example, Denton et al. (1954) reported information on an
individual who was patch-tested on three occasions with 0.005%
potassium dichromate solution and the filtrate of two cement samples
which contained 0.0001% and 0.0004% Cr VI. The individual repeatedly
showed a positive erythematous, edematous, papulovesicular patch-test
reaction to each test solution.
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Chromium «arch 31, 1987
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0 Chronic inhalation of dust or air containing Cr VI may cause respira-
tory problems including perforated or ulcerated nasal septa and
decreased spirometric values (U.S. EPA, 1985). For example, Bloomfield
and Blum (1928) reported perforated/ulcerated nasal septa and inflamed
nasal mucosa in workers exposed to chromic acid (Cr VI) (0.1 to 5.6
mg/m3 air) for one week to three years.
Animals
Short-term Exposure
In general, Cr VI compounds are more toxic than Cr III compounds:
0 The oral LD5g for various salts of Cr III range from 600 to 2,600 mg/kg
(Smyth et al., 1969).
0 The oral LD50 of Cr VI (as Na2Cr2O7) in rats is 19.8 mg/kg (NIOSH, 1983),
0 Rats were exposed to drinking water containing Cr VI (K2CrC>4) at levels
of both 80 and 134 rag Cr VI/L for 60 days (8.3 and 14.4 mg Cr Vl/kg/day
respectively) without adverse effect (Gross and Heller, 1946).
Long-term Exposure
0 In a one year rat drinking water study, consumption of water containing
0 to 25 mg/L of either Cr III (CrCl3) or Cr VI (K2Cr04) (0 to 1.87
mgAg/day for male rats and 0 to 2.41 mg/kg/day for female rats)
produced no significant differences in weight gain, appearance or
pathological changes in the blood or other tissues (Mackenzie et al.,
1958). NOAELs of 1.87 mg/kg/day (males) and 2.41 mg/k-j/day (females)
can be identified from the results of this study.
0 In a rat drinking water study in which 5 mg/L Cr III (about 0.42
mg/kg/day) was administered from the time of weaning until death, no
adverse effects were observed (Schroeder et al., 1965). A NOAEL of
0.42 mg/kg/day can be identified from the results of this study.
0 In a four year female dog drinking water study (five dose groups with
two animals per group), Cr VI (K2CrO5) at 0.45 to 11.2 mg/L (0.012 to
0.30 mg/kg Cr VI) was without effect in terms of changes in physical
condition, food consumption, growth rate, organ weights, urinalysis
results and hematological analyses. Therefore, a NOAEL of 0.30 mg/kg/day
can be identified from the results of this study (Anwar et al., 1961).
Reproductive Effects
0 No information was found in the available literature on the reproductive
effects of chromium.
Developmental Effects
0 No information was found in the available literature on the develop™
mental effects of chromium.
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Chromium March 31, 1987
-7-
Mutagenicity
0 The genotoxic effects of chromium are well documented both in in vivo
and In vitro studies. The pathway by which chromium exerts these
effects is believed to involve penetration of the cell membrane by
Cr VI, followed by intracellular reduction to Cr III. Extracellular
Cr III crosses the cell membrane, but less efficiently. Once inside
the cell, Cr III can form tight complexes with DNA, accounting for
its mutagenic potential (U.S. EPA, 1985).
0 Compounds of both Cr III and Cr VI increase non-complementary nucleo-
tide incorporation into DNA (Raffetto et al., 1977; Ma j one and Rensi,
1979), with Cr VI producing effects at lower doses. Exposure of cells
from rat liver and kidney to Cr VI leads to increased cross-linking
in DNA. Petrilli and De Flora (1978) reported positive Ames tests
for Cr VI. However, Cr III exerted no effect at relatively high
concentrations (presumably because of its inability to penetrate
cells). Similar results were reported by Gentile et al. (1981).
0 The difficulty of observing mutagenic effects of Cr III may be related
to its slight uptake by cells under most conditions. Warren et al.
(1931), studied the mutagenicity of a series of hexacoordinate Cr III
compounds and concluded that, in the proper ligand environment, the
metal possesses considerable genetic toxicity.
Carcinogenicity
There is inadeqaate evidence to determine whether or not oral exposure
to chromium can lead to cancer:
0 No increase in tumor rates over that of the control animals was
observed in rats exposed rats to Cr III (chromium oxide pigments)
at 293, 586 or 1,466 mg/kg/day in the diet for two years (Ivankovic
and Preussmann, 1975).
0 While the carcinogenicity of inhaled Cr VI is well established for
occupational exposure of humans (Hayes et al., 1979), the effects
are observed only in the respiratory passages and in the lungs, and
are believed to have no bearing on carcinogenic risk following oral
exposure to the metal (U.S. EPA, 1985).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-terni (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = or LOAEL) X (BW) = _ „ ( _ „ ,
(UF) x ( L/day)
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Chromiu-n March 31, 1987
-8-
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
In considering the toxicity of chromium compounds, it is important to
realize that chromium III is an essential nutrient required in trace quantities
for normal glucose metabolism - i.e. GTF. Some forms of chromium may also be
important,in the metabolism of lipids and other carbohydrates (U.S. EPA, 1935).
The Health Advisories will be determined on the basis of the effects of
Cr VI measured as total chromium. Separate Health Advisories will not be
established for Cr III for the following reasons:
1. Based on the work of Schroeder and Lee (1975), there is reason to
believe that oxidizing agents (e.g. due to chlorination of water)
may accelerate the normal conversion of Cr III to Cr VI at the
point of consumption (i.e., the tap).
2. Health Advisories based on total chromium will allow for the possible
conversion of Cr III to Cr VI.
3. As discussed in this document, there is reason to believe that Cr
VI is more toxic than Cr III. Thus Health Advisories base! on th»
effects of Cr VI will conservatively protect against the toxic
effects of any Cr III not converted to Cr VI.
One-day Health Advisory
The available data are insufficient to develop a One-day HA for chromium.
It is recommended that the Ten-day HA of 1.4 mg/L be used as the One-day HA
for the 10 kg child.
Ten-day Health Advisory
Gross and Heller (1946) exposed both male and female rats for 60 days
to drinking water containing K2Cr04 at either 300 or 500 mg/L (Cr VI at
80 mg/L and 134 mg/L, respectively). Using reported average body weights of
270 and 260 g, respectively, and assuming consumption of 28 mL water per day,
tne average ingested doses of Cr VI are calculated to be 8.3 and 14.4 "ijAg/day,
respectively. After two months, the rats receiving Cr VI at 8.3 mg/kg/day were
described as normal. A "slight roughness of coat" «M-> noted in rats receiving
14.4 mg/kg/day, but this is not considered to be an adverse health effect; this
observation is not associated with other adverse health effects. Therefore,
14.4 mg/kg/day represents the NOAEL for Cr VI in this study.
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Chromium March 31, 1987
-9-
The Ten-day HA for a 10-kg child is calculated as follows:
Ten-day HA = H 4.4 mg/kg/day) (10 kg) = 1.4 mg/Ij (1400 ug/L)
(100) (1 L/day)
where:
14.4 mg/kg/day = NOAEL based on the absence of adverse effects in
rats exposed to chromium in drinking water.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
Mackenzie et al. (1958) studied the effects of chronic ingestion of
Cr III and Cr VI in rats. Both male and female Sprague Dawley rats (34 days
old) were supplied with drinking water containing Cr as CrCl3 (Cr III) or as
K^CrO^ (Cr VI) in a series of doses up to 25 mg/L for a period of one year.
Assuming an average weight of 375 g for males and 290 g for females, and an
average daily water intake of 28 mL (Arrington, 1972), the average dose for
males and females receiving 25 mg/L is calculate-1 to be 1.87 and-2.41 mg
Cr Vl/kg/day, respectively. No significant adverse effects on appearance,
weight gain, food consumption or blood chemistry were noted at any of the
dose levels. However, the animals receiving the highest dose (25 mg/L) of
Cr VI showe? da approximate 20% reduction in water consumption.
Cr VI at 2.41 mg/kg/day was identified as the NOAEL in this study. The
Longer-term HAs are calculated as follows:
For a 10-kg child:
Longer-term HA = (2.41 mg/kg/day) (10 kg) = Q 24 /L (240 /T ,
(100M1 L/day)
where:
2.41 mg/kg/day = NOAEL based on the absence of adverse effects in
rats exposed to chromium in drinking water.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
1 L/day = assumed daily water consumption of a child.
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Chromium March 31, 1987
-10-
For a 70-kg adult:
Longer-term HA = (2.41 mg/kg/day) (70 kg) = 0<84 /L (840 /L }
(100) (2 L/day)
where:
2.41 mg/kg/day = NOAEL based on the absence of adverse effects in
rats exposed to chromium in drinking water.
70 kg = assumed body weight of an adult.
100 = uncertainty factor, chosen in accordance with NAS/ODW
s for use with a NOAEL from an animal study.
2 1,/day = assumed daily water consumption of an adult.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual1 s total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adversu 'vjilth effects over a lifetime exposure. The Lifetime HA
is derived vi t three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of -ieleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinog'eni. • 'le-tlth effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consu.no ti D-L jf an
adult. Tie Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative sonrr-i contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by MacKenzie et al. (1958) (described under the Lo-vjer-term HA)
is considered appropriate to serve as the basis for the Lifetime HA. The
Anwar et al. (1961) study was not selected because only two animals per dose
group were used.
j th-i NOAEL of 2.41 mg/kg/day, the Lifetime HA is derived as follows:
St-sp 1: Determination of the Reference Dose (RfD)
RfD = (2.41 mg/kg/day) = Q.0048 mg/kg/day
(100) (5)
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Chromium March 31, 1987
-1 1-
where:
2.41 mg/kg/day = NOAEL based upon the absence of adverse effects in
rats exposed to chromium in drinking water.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from -vi i-iinal study.
5 = additional uncertainty factor to compensate for less-
than-lifetime exposure.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0-0048 mg/kg) (70 kg) = ., 70 „
(2 L/day)
where:
0.0048 mrj/k'j = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of Lifetino Health Advisory
Lifetime HA = (170 ug/L) (71%) = 120 ug/L
where:
170 ug/L = DWEL.
71% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 There is no evidence of carcinogenic effects following oral exposure
to chromium. Thus, no assessments for carcinogenic risks from oral
exposure to chromium have been conducted. Inhalation of chromium,
however, is associated with an increased frequency of lung cancer in
humans.
0 EPA's CAG has estimated the lifetime cancer risk due to >\ constant
exposure to air containing 1 ug/m3 of elemental chromium to be
1.2 x 10~2 (U.S. EPA, 1983).
0 Based on exposure to chromium via inhalation, IARC (1982) has classified
chromium and certain chromium compounds in Group * (Chromium VI);
sufficient evidence for carcinogenicity in humans and ani.nal?.
0 Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), chromium may be classified in
Group A: Human carcinogen. This category is for agents for which
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Chromium March 31, 1987
-12-
there is sufficient evidence to support the causal associatioa '.i-jtwtjen
exixDsure to- the agents and cancer. However, as there are inadequate
data to conclude that chromium is carcinogenic via ingestion, chromium
is dealt with here as Groan D: Not classified. This category is for
agents with inadequate animal evidence of carcinogenicity .
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
Recommended or established standards for chromium in the United States
include:
0 50 ug Cr VI per litur for drinking water (U.S. PHS, 1962).
0 50 ug total chromium per liter for drinking water (NAS, 1974;
U.S. EPA, 1976).
0 1 ug/m3 for carcinogenic forms of Cr VI in workplace air (NIOSH, 1975)
0 25 ug/m3 TWA or 50 ug/m3 ceiling for non-carcinogenic forms of Cr VI
in workplace air (NIOSH, 1975).
0 The recourieri ]-.• 1 Ambient water quality criterion for Cr VI is 50 ug/L
(U.S. EPA, 1980).
0 An estimated adequate and safe intake for chraaium of 50 to 200 ug/day
for adults has been established (NAS, 1980a,b). This range is based
on the absence of sujns of chromium deficiency in the major portion
of the U.S. population which consumes a.-i 'U--rage of 60 ug of chromium
per day.
0 The OSHA 8-hour time-weighted average exposure limit for ciir- >irL'i>a,
soluble chromic, and chromous salts as chromium is 0.5 mg/m3 (03HA,
1985).
VII. ANALYTICAL METHODS
0 Determination of chromium is by atomic absoro!;; m (AA) using either
direct aspiration into a flame (U.S. EPA, 1979a) or a furnace technique
(U.S. EPA, 1979b).
0 The direct aspiration AA procedure is a physical method based on the
absorption of radiation at 357.9 nm by chromium. Tha s-inple is
aspirated into an air-acetylene flame and atomized. A 1i
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Chromium March 31, 1987
-13-
VIII. TREATMENT TECHNOLOGIES
0 The treatment technologies that are available to remove chromium from
water include coagulation/filtration, lime softening, ion exchange and
reverse osmosis. However, the type of treatment that may be applied
is dependent on the species of chromium present.
0 Laboratory and pilot plant studies indicated that using ferric sulfate,
Cr III removals were near 100 percent in the range of pH 6.5 to 9.5.
Alum was less effective between pH 7.5 and 8.5, with removals around
90 percent or better. Above and below this pH range, removals were
slightly lower, 80 to 90 percent. In removing Cr VI, laboratory and
pilot-plant tests confirmed that of the three coagulants used, only
ferrous sulfate was effective. With alum and ferric sulfate, Cr VI
removals did not exceed 30 percent. By comparison, ferrous sulfate
coagulation achieved 90 percent removal or better (U.S. EPA, 1977).
0 Results of jar and pilot-plant tests indicate that Cr III removal
efficiencies with lime softening were approximately 72 percent at pH
8.5 to 9.5 and greater than 99 percent at pH 11 to 11.5. Results
with Cr VI in the same tests in all cases were less than 10 percent
(U.S. EPA, 1977; Sorg, 1979).
0 Since Cr in occurs in cationic species and Cr VI in anionic species,
a cation exchanger in series with an anion exchanger may be required
for their removal. Removal of Cr VI from sewage (Sorg, 1979), industrial
wastewater, rinse waters from chromium plating operations (Miller
and Mindler, 1978), cooling tower blowdown (Richardson et al., 1963;
Miller and Mindler, 1978), and metal recovery (Sussman et al., 1945)
has been demonstrated. Laboratory tests on a simulated Arizona well
water (TDS 174 mg/L, pH 7.85) having 0.019 mg/L of Cr VI showed a
breakthrough of Cr VI at roughly 12,000 bed volumes (U.S. EPA, 1932).
Reports concerning industrial wastewater treatment indicate that ion
exchange can successfully remove Cr III to below 0.05 mg/L (Patterson,
1975). Strong acid cationic resins have been used for removing Cr
III as a contaminant from metal plating rinse waters and from chromate
treated cooling waters. Vendor information indicates that operating
pH levels of between 6 and 8 are adequate for C~ III removal with pH
above 7 being slightly better than pH below 7 (Rohm and Haas Co.,
1980). Ion exchange softening using standard strong acid synthetic
resins operating in the sodium cycle should effectively remove Cr
III with 90 percent or greater efficiency (Amore, 1977). in tests
of home softeners with tap water spiked with 1 mg/L of chromium
nitrate, the chromium content continued to be reduced to 0.020 mg/L
after 192 cycles, at which point the test was discontinued.
0 Reverse osmosis (RO) membranes can efficiently remove from 82 to 99
percent of the chromium in a feed water source (Fox, no date; Mixon,
1973; Johnston et al., 1978). Pilot plant tests using both cellulose
acetate and hollow fiber (polyamide) membranes demonstrated their
effectiveness in removing both Cr III and Cr VI. Cr III removal
ranged from 90 to 98 percent and Cr VI removal ranged from 82 to 97
percent. Slightly better removal was achieved with the hollow fiber
than with the cellulose acetate membranes (Fox, no date).
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Chromium March 31, 1987
-14-
IX. REFERENCES
Alexander, J., J. Aseth and T. Norseth. 1982. Uptake of chromium by rat
liver mitochondria. Toxicol. 24:115-122.
Amore, F. 1977. Technical Letter 20: Removal of Water Supply Contaminants
-- Chromium. Illinois State Water Survey, 1977.
Anwar, R.A., R.F. Langham, C.A. Hoppert, B.V. Alfredson and R.U. Byerrum.
1961. Chronic toxicity studies: III. Chronic toxicity of cadmium and
chromium in dogs. Arch. Environ. Health. 3:456-460.
Arrington, L.R. 1972. The laboratory animals. In: Introductory laboratory
animal science. The breeding, care and management of experimental
animals. Interstate Printers and Publishers, Danville, IL. pp. 9-11.
Burrows, D. 1978. Chromium and the skin. Br. J. Dermatol. 99:587-595.
Collins, R.J., P.O. Fronm and W.D. Collings. 1961. Chromium excretion in
the dog. Am. J. Physiol. 201:795-798.
Davids, H.W., and M. Lieber. 1951. Underground water contamination by chromium
wastes. Water Sewage Works. 98:528-534.
Donaldson, R.M., Jr., and R.F. Barreras. 1966. Intestinal absorption of
trace quantities of chromium. J. Lab. Clin. Med. 68:484-493.
Fox, K.R. (No Date). Removal of inorganic contaminants from drinking water
by reverse osmosis. U.S. Environmental Protection Agency (unpublished).
Gentile, J.M., K. Hyde and J. Schubert. 1981. Chromium genotoxicity as
influenced by complexation and rate effects. Toxicol. Lett. 7:439-445.
Gray, S.J., and K. Sterling. 1950. The tagging of red cells and plasma
proteins with radioactive chromium. J. Clin. Invest. 29:1604-1613.
Gross, W.G., and V.G. Heller. 1946. Chromates in animal nutrition. J. Ind,
Hyg. Toxicol. 28:52-56.
Gruber, J.E., and K.W. Jennette. 1978. Metabolism of the carcinogen chromate
by rat liver microsomes. Biochem. Biophys. Res. Commun. 82:700-706.
Hartford, W.H. 1979. Chromium compounds. In; M. Grayson and D. Eckroth,
eds. Kirk-Othmer encyclopedia of chemical technology, Vol. 6. New
York, NY: John Wiley and Sons. pp. 82-120.
Hayes, R.B., A.M. Lilienfeld and L.M. Snell. 1979. Mortality in chromium
chemical production workers: a prospective study. Int. J. Epidemiol.
8:365-374.
Hem, J.D. 1970. Study and interpretation of the chemical characteristics of
natural water, 2nd ed. U.S. Geological Survey Water-Supply Paper 1473.
p. 199.
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Chromium March 31, 1987
-15-
Hopkins, L.L. 1965. Distribution in the rat of physiological amounts of
injected Cr51(III) with time. Am. J. Physiol. 209:731-735.
Hopkins, L.L., and K. Schwarz. 1964. Ch
specifically siderophilin. Biochem. Biophys. Acta
IARC. 1982. International Agency for Research on Cancer. IARC monographs on
the evaluation of the carcinogenic risk of chemicals to humans. Suppl.
4: 133-135.
Her, R.K. 1954. Process for the production of Verner type chromium complexes.
U.S. Patent No. 2,683,156.
Ivankovic, S., and R. Preussman. 1975. Absence of toxic and carcinogenic
effects after administration of high doses of chromic oxide pigment in
subacute and long term feeding experiments in rats. Food Cosmet. Toxicol.
1 3:347-351.
Johnston, J.K., and H.S. Lim. 1978. Removal of persistent contaminants from
municipal effluents by reverse osmosis. Environmental Protection Service,
Environment Canada.
Kraintz, L., and R.V. Talmage. 1952. Distribution of radioactivity following
intravenous administration of trivalent chromium-51 in the rat and
rabbit. Proc. Soc. Exp. Biol. Med. 81:490-492.
Love, C.H. 1947. German production of some of the more important inorganic
pigments. Washington, DC: Hobart Publishing Co. pp. 47-63.
MacKenzie, R.D., R.U. Byerrum, C.F. Decker, C.A. Hoppert and R.F. Langham.
1958. Chronic toxicity studies. II. Hexavalent and trivalent chromium
administered in drinking water to rats. AMA Arch. Ind. Health. 18:232-234.
Majone, F., and D. Rensi. 1979. Mitotic alterations, chromosome aberrations
and sister chromatid exchanges induced by hexavalent and trivalent
chromium on mammalian cells in vitro. Caryologia. 32:379-392.
Mertz, W., E.E. Roginski and R.C. Reba. 1965. Biological activity and fate
of trace quantities of intravenous chromium (III) in the rat. Am. J.
Physiol. 109:489-494.
Mertz, W., E.E. Roginski, F.J. Feldman and D.E. Thurman. 1969. Dependence
of chromium transfer into the rat embryo on the chemical form. J. Nutr.
99:363-367.
Mertz, W., and E.E. Roginski. 1971. Chromium metabolism: The glucose
tolerance factor. In; W. Mertz and W.E. Cornatzer, eds. Newer trace
elements in nutrition. New York, NY: Marcel Dekker. pp. 123-151.
Mertz, W. 1976. Chromium and its relation to carbohydrate metabolism. In;
R.E. Burch and J.F. Sullivan, eds. Symposium on trace elements. Med.
Clin. North Am. 60:739-744.
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Chromium March 31, 1987
-16-
Mertz, W., R.A. Anderson, W.R. Wolf and E.E. Roginski. 1978. Progress in
chromium nutrition research. In: M. Kirchgessner, ed. Trace element
metabolism in man and animals. Proc. Third Int. Symp. Freising, July,
1977. pp. 272-278.
Miller, W.S., and A.B. Mindler. 1978. Ion exchange separation of metal ions
from water and waste waters. Permutit R&D Center.
Miller, W.S. 1978. Removal and recovery of chromates from cooling tower
blowdown. In: Ion Exchange for Pollution Control, Vol. I. CRC Press,
Inc.
Mixon, F.O. 1973. The removal of toxic metals from water by reverse osmosis.
U.S. Department of the Interior, INT-OSWRDPR-73-899.
NAS. 1974. National Academy of Sciences. Water quality criteria 1972.
Washington, DC: National Academy Press, p. 62.
NAS. 1980a. National Academy of Sciences. Recommended dietary allowances,
9th rev. ed. Washington, DC: National Academy Press, pp. 159-161.
NAS. 1980b. National Academy of Sciences. Drinking Water and Health.
Volume 3. Washington, DC: National Academy Press, pp. 266, 364-369,
374-375.
NIOSH. 1975. National Institute for Occupational Safety and Health. Occu-
pational exposure to chromium VI. Criteria document HEW(NIOSH). 76-129.
NIOSH. 1983. National Institute for Occupational Safety and Health. Registry
of Toxic Effects of Chemical Substances (RTECS). Vol. 2, p. 72.
OSHA. 1985. Occupational Safety and Health Administration. Code of Federal
Regulations. Title 29 - Labor. Part 1910 - Occupational Safety and
Health Standards. Subpart Z - Toxic and Hazardous Substances. Section
1910.1000 - Air Contaminants. U.S. Government Printing Office,
Washington, DC.
Patterson, J.w. 1975. Wastewater Treatment Technology. Ann Arbor Science
Publisher, Inc.
Petrilli, F.L., and S. De Flora. 1978. Oxidation of inactive trivalent
chromium to the mutagenic hexavalent form. Mutat. Res. 58:167-173.
Raffetto, G., S. Parodi, C. Parodi, M. DeFarrari, R. Troiano and G. Brambilla..
1977. Direct interaction with cellular targets as the mechanism for
chromium carcinogenesis. Tumori. 63:503-512.
Richardson, E.W., E.D. Stobbe et al. 1968. Ion exchange traps chromates for
reuse. Environmental Science and Technology. 2(11):1006-16.
Rohm and Haas Co. 1980. Amber lite Ion Exchange Resins. Technical Bulletins
for IR-120 and IRA-402.
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Chromium March 31, 1987
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Schroeder, H.A., J.J. Balassa and W.H. Vinton, Jr. 1965. Chromium, cadmium
and lead in rats: Effects on life span, tumors and tissue levels.
J. Nutr. 86:51-66.
Schroeder, D.C., and G.F. Lee. 1975. Potential transformations of chromium
in natural waters. Water Air Soil Pollut. 4:355-365.
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J.S. Nycum. 1969. Range finding toxicity data: List VII. Am. Ind.
Hyg. Assoc. Journal. 30:470.
Sorg, T.J. 1979. Treatment technology to meet the interim primary drinking
water regulations for inorganics: part 4. JAWWA. 71(8):454-66.
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Industrial and Engineering Chemistry. 37(7):618-22.
U.S. EPA. 1976. U.S. Environmental Protection Agency. National interim
primary drinking water regulations. EPA 570/9-76-003. Washington, DC:
pp. 63-64.
U.S. EPA. 1977. U.S. Environmental Protection Agency. Manual of treatment
techniques for meeting the interim primary drinking water regulations,
revised. EPA-600/8-77-005.
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Absorption, direct aspiration. In: Methods for Chemical Analysis of
Water and Wastes. EPA-600/4-79-020, March, 1979.
U.S. EPA. 1979b. U.S. Environmental Protection Agency. Method 218.2. Atomic
Absorption, furnace technique. In: Methods for Chemical Analysis of
Water and Wastes. EPA-600/4-79-020, March, 1979.
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quality criteria for chromium. EPA 440/5-80-035. Washington, D.C.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Personal communication.
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document for chromium. Review Draft. EPA 600/8-82-014A. Washington, D.C,
U.S. EPA. 1985. U.S. Environmental Protection Agency. Health Effects
Criteria Document for C' romium. Criteria and Standards Division. Office
of Drinking Water. Washington, DC.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51(185):33992-34003.
September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Estimated national
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CSD. Office of Drinking Water.
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Chromium March 31, 1987
-18-
U.S. PHS. 1962. U.S. Public Health Service. Drinking water standards.
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drugs, 9th ed. Rahway, NJ: Merck and Co., Inc. pp. 228-289.
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March 31, 1987
CYANIDE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Cyanide March 31, 1987
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This Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for cyanide (U.S.
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd.,
Springfield, VA 22161, PB # 86-117793/AS. The toll free number is (800)
336-4700; in the Washington, D.C. area: (703) 437-4650.
II. GENERAL INFORMATION AND PROPERTIES
0 Cyanides are a group of organic and inorganic compounds that contain
the cyano (CN) group. Free cyanide is defined as the sum of cyanide
present as HCN and as CN~. The organic cyanides are called nitriles
and few of them dissociate to yield CN~ or HCN. In this Health
Advisory, only a few widely used industrial inorganic cyanides will
be discussed.
CAS No.
0 Hydrogen Cyanide — 74-90-8
Sodium Cyanide — 143- 33-9
Potassium Cyanide — 151-50-8
Synonyms
0 Hydrogen Cyanide: Prussic Acid
Uses (U.S. EPA, 1985)
0 Cyanide is used in rat and pest poisons, silver and metal polishes,
photographic solutions, fumigating products, in the production of
various resins such as acrylates, methyl acrylate and nitriles and
in electroplating.
0 Although there are a number of organic and inorganic compounds that
contain the CN group, in this document only a few widely used indus-
trial inorganic cyanides will be considered.
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Cyanide
March 31, 1987
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Properties (Weast, 1980; Towill et al., 1978)
0 The properties of cyanide compounds vary with the specific compound;
some examples are as follows:
Chemical Formula
Molecular Weight
Physical State
Hydrogen
Cyanide
HCN
27.03
colorless gas
or liquid
25.70°C
-13.24°C
0.688 (20°C)
Boiling Point
Melting Point
Densi ty {g/cm 3)
Vapor Pressure —
Water Solubility (g/100 mL) miscible
Octanol/Water 0.66
Partition Coefficient
Taste Threshold
Odor Threshold
Conversion Factor 1.123
Sodium
Cyanide
NaCN
49.01
colorless solid
14.96°C
563.7°C
1.60-1.62
48 (10°C)
-0.44
2.037
Potassium
Cyanide
KCN
65.12
colorless solid
634.5°C
1.553 (20°C)
71.6 (20°C)
2.707
Occurrence
In 1978, cyanide production in the United States exceeded 700 million
pounds. Cyanide wastes are released from the pyrolysis of natural and
synthetic materials (Towill et al., 1973).
Despite numerous potential sources of pollution, cyanide is relatively
uncommon in U.S. drinking water. In 1970, a survey of 969 water
supplies failed to reveal cyanide concentrations above 0.2 mg/L. Of
2,595 samples examined, the highest cyanide concentration found was
8 ug/L and the average concentration was 0.09 ug/L (McCabe et al.,
1970).
III. PHARMACOKINETICS
Absorption
Free cyanides are absorbed readily from the lungs, the gastrointestinal
tract and the skin by animals and humans. Inhalation exposure to
HCN provides the most rapid route of entry (U.S. EPA, 1985).
Dogs treated with KCN at single gavage doses equivalent to 20, 50
and 100 mg HCN (1.57, 4.42 or 8.40 mg HCN/kg bw) absorbed 72%, 24%
and 16.6%, respectively, through the GI tract (Gettler and Baine, 1938)
The dogs died within 155, 21 and 8 minutes, respectively, after dosing.
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Cyanide March 31, 1987
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Distribution
0 Once cyanide is absorbed, it is distributed rapidly by the blood
throughout the body. Distribution patterns vary with the route of
exposure (U.S. EPA, 1985).
0 High levels of cyanide were found in brains and livers of 3 human
subjects who ingested fatal doses of cyanide (Gettler and Baine, 1938).
0 In rabbits, intramuscular injection of HCN gave higher levels of CN
in blood and tissues than did KCN administration (Ballantyne et al.,
1972).
0 When radiolabeled KCN (5 mgAg) was administered orally to rats over
24 hours, a rapid decline of radioactivity from whole blood and plasma
was observed with a small increase in the levels in erythrocytes
(Farooqui and Ahmed, 1982). The majority of the radioactivity in the
erythrocytes (94.3%) was found in the hemolysate rather than the
membranes. The heme fraction contained 70% of the radioactivity while
14 to 25% and 5 to 10% were found in the globin and cell membrane,
respectively.
0 Cyanide does not accumulate in blood and tissues following chronic
exposure. Virtually no cyanide was found in the plasma or kidneys of
rats treated with dietary concentrations of 100 and 300 ppm (mgAg
diet) for two years (Howard and Hanzal, 1955). Low levels were found
in erythrocytes (mean of 1.97 ug). Increased levels of~ thiocyanate
were found in plasma (1123 ug), erythrocytes (246 ug), liver (665 ug)
and kidney (1188 ug).
0 Yamamoto et al. (1982) found that rats on oral (gavage) exposure to
cyanide (NaCN) (7 and 21 mg CN/kg/bw) showed higher levels of cyanide
in lungs and liver compared to blood. On inhalation exposure to HCN
at concentrations of 356 and 1,180 ppm (392 and 1,298 mg/m3), concen-
tration in the lungs exceeded that in the blood.
Metabolism
0 Cyanide is detoxified by an intramitochondrial enzyme, rhodanese,
which catalyzes the transfer of sulfur from a donor to cyanide to form
the less toxic thiocyanate. Rhodanese is widely distributed throughout
the body; high doses are found in the liver (U.S. EPA, 1985).
0 Other minor detoxification pathways include spontaneous reaction with
cystine to form 2-imino-4-thiozolidene carboxylic acid and with hydroxy-
cobalamine to form cyanocobalamine - i.e. vitamin B-|2 (U.S. EPA, 1985) =
Excretion
The major route of cyanide elimination is as the thiocyanate in the
urine, although some cyanide enters the metabolism of one-carbon
compounds and C02 is eliminated in expired air. A small amount of
HCN is eliminated in expired air (U.S. EPA, 1985).
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Cyanide
March 31, 1987
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Rats eliminated 80% of subcutaneously-injected cyanide as thiocyanate
in the urine, while 15% was eliminated as urinary 2-imino-thiozolidine
carboxylic acid (Wood and Cooley, 1956).
A man who had ingested 3 to 5 g KCN (at least 1 .2 g HCN was present in
the blood 3 hours later) eliminated a total of 237 mg thiocyanate in
72-hour urine (Liebowitz and Schwartz, 1948).
IV. HEALTH EFFECTS
Humans
The enzyme cytochrome oxidase enables cells to utilize oxygen.
Cyanide inhibits this enzyme thus resulting in effective cellular
anoxia (U.S. EPA, 1985).
Acute exposure to cyanide by the oral route has usually occurred from
suicide attempts (NIOSH, 1976). Signs of acute poisoning by cyanide
include rapid breathing, gasping, tremors, convulsions and death
(DiPalma, 1971).
The fatal oral doses of cyanide compounds range from 50 to 200 mg (0.7
to 2.9 mg CN'/kg 'bw) (U.S. EPA, 1985). Within 20 minutes of ingestion
of fatal doses, events progress from hyperventilation, vomiting,
unconsciousness, convulsions, rapid and irregular heart rate, gasping,
vascular collapse and cyanosis, to death.
Although data regarding chronic oral exposure of humans to HCN, KCN
or NaCN are not available, there are a number of reports on the etiology
of thyroid disorders and neuropathies characterized by optic atrophy,
nerve deafness and spinal ataxia in people living in certain tropical
areas of Africa where the staple diet consists largely of cassava.
Cassava contains a high level of the cyanogenic glycoside, linamarin,
which releases cyanide on metabolism or acid hydrolysis in vivo
(Osuntokun et al., 1969; Makene and Wilson, 1972).
Case studies and epidemiological studies of case-hardeners, electro-
platers, metal polishers, photographic material workers and HCN
fumigators have revealed effects in workers typical of sublethal
cyanide poisoning, including headache, dizziness and thyroid enlarge-
ment (NIOSH, 1976).
Animals
Short-term Exposure
The acute oral LD5Q for KCN was 10 mgAg (4 mg CN~/*g) in rats (Hayes,
1967; Gaines, 1969) and 8.5 mg KCN/kg (3.4 mg CN~/*g) in mice (Sheehy
and Way, 1968). The LD5Q of intraperitoneally administered NaCN for
mice was 3.2 mg CN~/kg (Kruszyna et al., 1982).
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Cyanide March 31, 1987
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Mice administered 1 or 2 mg KCN/kg (0.4 or 0.8 mg/CN~/kg) intra-
peritoneally showed minimal or no effects, while 3 to 5 mg KCN/kg
(1.2-2.0 mg CN~/kg) resulted in signs of toxicity (convulsions, agi-
tation) (Isom et al., 1982). A dose of 6 mg KCN/kg (2.4 mg CN~/*g)
resulted in 20% mortality.
Doses that are fatal to one species may be harmless to others. An
oral dose of 3.8 mg KCN/kg (1.5 mg CN-/kg) was fatal to a dog in 155
minutes (Gettler and Baine, 1938) but a higher dose of 8 mg KCN/kg
(3.2 mg CN~/kg), equal to the ^050 in mice, had only minimal effects
on guinea pigs (Basu, 1983).
Rats tolerated higher doses of cyanide (80 mg CN~/kg bw/day) when
mixed in the diet (Kreutler et al., 1978) than when administered by
gavage (4.0 mg CN~/kg bw) (Ferguson, 1962).
Rats tolerated 25 daily doses of 1 0 mg KCN/kg bw (4 mg CN-/kg bw)
when the chemical was mixed in the diet; each of these doses was
equal to the acute oral LDsg (Hayes, 1967).
Rats tolerated higher oral doses of KCN (approximately 30 mg KCN/kg
bw/day or 1 2 mg CN~/kg bw/day for 21 days) when administered in
drinking water (Palmer and Olsen, 1979) than when KCN was administered,
in a bolus (approximately 10 mg/kg bw KCN; 4.5 mg CN~/kg bw) by
gavage with water as the vehicle (Hayes, 1967; Gaines, 1969).
Rats receiving approximately 12 mg CN-/kg bw/day for 21 days in
drinking water had significantly increased liver weights compared
with controls, while rats receiving approximately 8 mg CN~/kg bw/day
in the diet did not (Palmer and Olsen, 1979).
Beagle dogs consuming 3 mg CN-/kg bw/day in the diet for 30 days showed
no clinical signs of toxicity, effects on body weight, hematology or
histopathologic lesions (American Cyanamid Co., 1959).
Long-term Exposure
Animals can tolerate higher doses of cyanide when administered in the
diet or in the drinking water during longer-term exposures (Hayes,
1967; Palmer and Olsen, 1979) than as a bolus dose by gavage.
Pigs (sows) maintained on diets containing cyanide (30.3, 276.6
and 520.7 mg CN~/kg diet) throughout gestation and lactation
showed hyperplasia of kidney glomerular cells and accumulation
of colloid and morphological changes in follicular cells of the
thyroid (Tewe and Maner, 1981b). (See also discussion under
Developmental Effects, below.)
Weanling rats maintained on a diet containing 1,500 ppm KCN for 11.5
months (approximately 30 mg CN-/kg bw/day) had a significantly reduced
body weight gain, increased excretion of thiocyanate at 4 months
and at 11 months, decreased plasma thyroxine levels, and decreased
thyroxine secretion rates at 4 months (Philbrick et al., 1979). The
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Cyanide March 31, 1987
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effects appeared to be greater in the animals on the vitamin B-j 2~ and
methionine-deficient diet. There were no definitive histopathologic
lesions in the optic or CNS tissues, thyroid or sciatic tissues;
however, vacu'olization and myelin degeneration were observed in
spinal cord sections.
0 Dogs receiving _>. 0.27 mg CN-/k9 bw/day, administered in a capsule for
15 months, had degenerative changes in ganglion cells of the CNS
(Hertting et al., 1960). These effects may be due to the fact that
CN- was administered in a capsule (similar to a bolus dose by gavage).
0 Rats maintained for 104 weeks on diets that had been fumigated with
HCN to give average dose levels of 76 mg/kg diet and 190 mg/kg diet
(i.e., approximately 3.6 and 7.5 mg CN~/kg bw/day for males and
4.6 and 10.8 mg CN~/kg bw/day for females) resulted in no effects
clinically or histologically (Howard and Hanzel, 1955). The only
effects of treatment were increased CN~ levels in the red blood
cells, increased thiocyanate levels in the plasma, red blood cells,
liver and kidneys of animals from both treatment groups.
Reproductive Effects
0 No effects were seen on the reproductive performance of pregnant rats
fed 500 mg CN~/k9 diet (KCN) through gestation and lactation (Tewe
and Maner, 1981a). Offspring that were continued on the test diet
after weaning consumed less food and grew at a significantly reduced
rate compared to control offspring.
Developmental Effects
0 Severe teratogenic effects were seen in hamsters administered cyanide
by subcutaneously implanted osmotic minipumps that delivered cyanide
at a rate of 3.3-3.4 mg CN-/kg bw/hour (79.2-81.6 mg CW/kg bw/day)
from day 6-9 of gestation (Doherty et al., 1982).
0 Piglets born to pigs maintained on diets containing cyanide (30.3,
276.6 and 520.7 mg CN~/^9 diet) throughout gestation and lactation
showed reduced organ-to-body weight ratios of the thyroid, spleen and
heart in the high and/or medium dose groups relative to the low-dose
group piglets (Tewe and Maner, 1981b). (See also discussion under
Long-term Exposure, above.)
Mutagenicity
0 Potassium cyanide was not mutagenic in Salmonella typhimurium with or
without metabolic activation (De Flora, 1981).
0 A study using HCN gas reported marginally mutagenic activity in
S_._ typhimurium strain TA100 (Kushi, 1983). Addition of S-9 mix
decreased the mutagenic activity.
0 Cyanide was negative in a modified rec assay in Bacillus subtilis
(Karube et al., 1981).
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Cyanide March 31, 1987
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Carcinogenicity
0 No information was located in the available literature on the
carcinogenicity of cyanides.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula;
HA = JNOAEL or LOAEL) x (BW) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effeet-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/aay = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
The available data are insufficient to develop a One-day HA for cyanide.
It is recommended that the modified DWEL of 0.22 mg/L (adjusted for the 10-kg
child) be used as the One-day HA for the 10-kg child.
Ten-day Health Advisory
While the study by Palmer and Olsen (1979) was considered as the basis
for the Ten-day HA, it is recommended that the modified DWEL of 0.22 mg/L
(adjusted for a 10-kg child) be used as the Ten-day HA for the 10-kg child.
The NOAEL observed by Howard and Hanzal (1955) in a two-year rat study (which
serves as the basis for the DWEL and Lifetime HA) was 10.8 mgAg/day, in good
general agreement with the NOAEL of 8 mg/kg/day observed by Palmer and Olsen
(1979) in a 21-day rat study. As the NOAELs in the two studies were little
different and as greater confidence was placed in the Howard and Hanzal (1955)
study, it was determined that it was appropriate to use the modified DWEL as
the Ten-day HA.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for cyanide
It is recommended that the DWEL of 0.77 mg/L be used as the Longer-term HA
for the 70-kg adult and the modified DEWL of 0.22 mg.L (adjusted for a 10-kg
child) be used as the Longer-term HA for the 10-kg child.
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Cyanide March 31, 1987
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Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study of Howard and Hanzal (1955) has been selected to serve as the
basis for the DWEL and Lifetime HA. In this study, rats were maintained for
104 weeks on diets that had been fumigated with HCN to give average dose
levels of 76 or 190 mg/kg diet (i.e., approximately 3.6 and 7.5 mg/kg/day for
male rats and 4.6 and 10.8 for female rats). No clinical or histological
effects were observed at either dose level.
Using the NOAEL of 10.8 mg/kg/day, the DWEL and Lifetime HA are derived
as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (10.8 mg/kg/day) = 0.022 mgAg/day *
(100) (5) ^ y/
* NB: The RfD is in good general agreement with the observation (NIOSH, 1976)
that 1 mg HCN/m^ is without effect in humans via inhalation.
where:
10.8 mgAg/day = NOAEL for absence of clinical and histological effects
in rats exposed to HCN in the diet for 104 weeks.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
5 = additional uncertainty factor selected to allow for
possibly greater absorption of cyanide from water
than from the diet.
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Cyanide March 31, 1987
-10-
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.022 mg/kg/day) (70 kg) = 0>77 mg/L (770 ug/L)
(2 L/day)
where:
0.02 mq/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
The DWEL of 770 ug/L assumes that 100% of the exposure to cyanide is
via drinking water. It is probable, however, that exposure occurs via other
routes. Therefore, if one assumes that drinking water contributes 20% of daily
exposure to cyanide/ then the Lifetime Health Advisory would be 154 ug/L.
The Lifetime HA is calculated as follows:
Lifetime HA = (770 ug/L) (20%) = 154 ug/L
where:
770 ug/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 There is no available information pertaining to the carcinogenicity
of cyanides.
0 IARC has not calculated the carcinogenic potential of cyanides«
0 Applying the criteria described in EPA's final guidelines for
assessment of carcinogenic risk (U.S. EPA, 1986), cyanide may be
classified in Group D: Not classified. This category is for agents
with inadequate human and animal evidence of carcinogenicity.
i
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The ambient water quality criterion for cyanide has been proposed at
3.77 mg/L assuming that a 70 kg human consumes 2 L of water and 6.5 g
of fish per day with a bioconcentration factor of 1.0 (U.S. EPA, 1982).
0 The U.S. PHS (1962) recommended that concentrations of cyanide in
water supplies not exceed 0.2 mg/L in order to protect human health.
The U.S. PHS (1962) also recommended that concentrations in drinking
water be kept below 0.01 mg/L since this level or lower can be achieved
by proper treatment.
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Cyanide March 31, 1987
-11-
0 ACGIH (1980) has recommended a TLV for alkali cyanides in workroom air
of 5 mg CN-/m3'
0 NIOSH has recommended a TLV of 5 mg/m^ for CN~ which was adopted by
OSHA (1981).
VII. ANALYTICAL METHODS
0 Determination of cyanide is by volumetric titration or colorimetry
(U.S. EPA, 1979). The cyanide as hydrocyanic acid (HCN) is released
from cyanide complexes by means of a reflux-distillation operation
and absorbed in a scrubber containing sodium hydroxide solution. The
cyanide ion in the absorbing solution is then determined by volumetric
titration or colorimetrically. The titration procedure uses a standard
solution of silver nitrate and an indicator. The detection limit is
1 mg/L. In the colorimetric measurement, the cyanide is converted to
cyanogen chloride, a reagent is added to form a colored complex and
the absorbance is measured. The detection limit is 20 ug/L.
VIII. TREATMENT TECHNOLOGIES
0 Several treatment technologies for the removal of cyanide are avail-
able, although most of what has been reported in the literature
involves wastewater applications. The treatment of high concentrations
of cyanide (and cyanide complexes) in industrial waste streams and
mine drainage runoff has been studied extensively, but only limited
information is available on reductions of low cyanide levels in
drinking water supplies.
0 The general treatment technologies that may be practical for reducing
cyanide levels in drinking water include oxidation by chlorine or
ozone, ion exchange and reverse osmosis.
0 Oxidation by chlorine may be the cheapest and most practical method to
remove cyanide from the water. In addition to the removal of cyanide,
chlorine oxidation may cause secondary beneficiary effects. These
include disinfection of the water, oxidation of iron and manganese,
oxidation of taste and odor causing compounds. Practical experience
in the wastewater industry and the laboratory indicate that chlorine
oxidation is capable of removing 99% or more of the cyanide from the
water (Gott, 1978; Smith et al., 1980).
0 Oxidation by ozone may be used to destroy cyanide in water if the
formation of trihalomethanes is to be avoided. The use of ozone
oxidation for cyanide is a relatively new technique. Laboratory and
pilot studies by Cullivan (Cullivan, no date) indicated that complete
destruction of cyanide in water required a 1.3 to 1 ozone to cyanide
molar ratio. These results were achieved with 5 to 15 mg/L of cyanide
in the influent water. However, the results of study indicated that
at cyanide concentrations of less than 5 mg/L, the rate of destruction
is decreased.
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Cyanide March 31, 1987
-12-
Although reverse osmosis and ion exchange can reduce cyanide levels in
the water, their application may not be practical in the economical
sense if cyanide is the only contaminant to be removed. Experience
by Moore (1976) and Trachtenberg et al. (1979) indicates that a
well-designed ion exchange facility can remove over 99% of the cyanide
present in the water. Pilot plant studies performed by Rosehart
(1973) treating mine-waters by reverse osmosis, resulted in cyanide
removal ranging from 28.7 to 81.6 percent, respectively.
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Cyanide March 31, 1987
-13-
IX. REFERENCED
ACGIH. 1980. American Conference of Governmental Industrial Hygienists.
Documentation of the threshold limit values for substances in workroom
air, 4th ed., with supplements through 1981. Cincinnati, OH. pp. 109-110.
American Cyanamid Co. 1959. Report on sodium cyanide: 30-day repeated
feedings to dogs. Central Med. Dept. Report Number 59-14.
Ballantyne, B., J. Bright, D.W. Swanston and P. Williams. 1972. Toxicity
and distribution of free cyanides given intramuscularly. Med. Sci. Law.
12:209-219.
Basu, T.K. 1983. High-dose ascorbic acid decreases detoxification of cyanide
derived from amygdalin (laetrile): studies in guinea pigs. Can. J.
Physiol. Pharmacol. 61(11):1426-1430.
Cullivan, B.M. No date . Industrial Toxics Oxidation: An Ozone-Chlorine Compar-
ison. Presented at the 33rd Purdue Industrial Waste Conference.
De Flora, S. 1981. Study of 106 organic and inorganic compounds in the
Salmonella/microsome test. Carcinogenesis. 2(4):283-298.
DiPalma, J.R., ed. 1971. Noxious gases and vapors: I. Carbon monoxide,
cyanides, methemoglobin and sulfhemoglobin. In; Drill's Pharmacology
in Medicine. McGraw-Hill Book Co., NY. pp. 1189-1205.
Doherty, P.A., V.H. Perm and R.P. Smith. 1982. Congenital malformations
induced by infusion of sodium cyanide in the golden hamster. Toxicol.
Appl. Pharmacol. 64:456-464.
Farooqui, M.Y.H., and A.E. Ahmed. 1982. Molecular interaction of acrylonitrile
and potassium cyanide with rat blood. Chem. Biol. Interact. 38:145-159.
Ferguson, H.C. 1962. Dilution of dose and acute oral toxicity. Toxicol.
Appl. Pharmacol. 4:759-762.
Gaines, T.B. 1969. Acute toxicity of pesticides. Toxicol. Appl. Pharmacol.
14:515-534.
Gettler, A.O., and J.O. Baine, 1938. The toxicology of cyanide. Am. J. Med.
Sci. 195:182-198.
Gott, R.D. 1978. Development of waste water treatment at the Climax Mine.
Mining Congress Journal 64(4):28-34.
Hayes, W.T. 1967. The 90-dose LD50 and a chronicity factor as measurer of
toxicity. Toxicol. Appl. Pharmacol. 11:327-335.
Hertting, G., 0. Kraupp, E. Schnetz and S. Wieketich. 1960. Untersuchungen
uber die Folgen einer chronischen Verabreichung akut toxischer Dosen von
Natriumcyanid an Hunden. Acta. Pharmacol. Toxicol. 17:27-43.
-------
Cyanide March 31, 1987
-14-
Howard, J.W., and R.F. Hanzal. 1955. Chronic toxicity for rats of food
treated with hydrogen cyanide. J. Agric. Food Chem. 3s325-329.
Isom, G.E., G.E. Burrows and J.L. Way. 1982. Effect of oxygen on the
antagonism of cyanide intoxication-cytochrome oxidase, in vivo.
Toxicol. Appl. Pharmacol. 65(2):250-256.
Karube, I., T. Matsunaga, T. Nakahara, S. Suzuki and T. Kata. 1981. Pre-
liminary screening of mutagens with a microbial sensor. Anal. Chem.
53(7):1024-1026.
Kreutler, P.A., V. Varbanov, W. Goodman, G. Olaya and J.B. Stanbury. 1978.
Interactions of protein deficiency, cyanide and theocyanate on thyroid
function in neonatal and adult rats. Am. J. Clin. Nutrit. 31:282-289.
Kruszyna, R., H. Kruszyna and R.P. Smith. 1982. Comparison of hydroxylamine,
4-dimethylaminophenol and nitrite protection against cyanide poisoning
in mice. Arch. Toxicol. 49:191-202.
Kushi, A., T. Matsumoto and D. Yoshida. 1983. Mutagen from the gaseous
phase of protein pyrolyzate. Agric. Biol. Chem. 47(9):1979-1982.
Liebowitz, D., and H. Schwartz. 1948. Cyanide poisoning: Report of a case
with recovery. Am. J. Clin. Pathol. 18:965-970.
Makene, W.J., and J. Wilson. 1972. Biochemical studies in Tanzanian patients
with ataxic tropical neuropathy. J. Neurol. Neurosurg. Psychiatry.
35:31-33.
McCabe, L.J., J.M. Symons, R.D. Lee and G.G. Robeck. 1970. Survey of com-
munity water supply systems. J. AWWA. 62:670-687,
Moore, F.L. 1976. An improved ion exchange resin method for removal and
recovery of zinc cyanide and cyanide from electroplating wastes.
J. Environ. Sci. Health. 7:459-467.
NIOSH. 1976. National Institute for Occupational Safety and Health. Criteria
for a recommended standard...occupational exposure to hydrogen cyanide and
cyanide salts (NaCN, KCN and Ca(CN)2). NIOSH Publ. No. 77-108. Dept.
Health, Educ. & Welfare. U.S. Govt. Printing Office, Washington, D.C.
OSHA. 1981. Occupational Safety and Health Administration. General Industry
OSHA Safety and Health Standards (29 CFR 1910). OSHA 2206. U.S. Dept.
of Labor, Washington, D.C.
Osuntokun, B.O., G.L. Monekosso and J. Wilson. 1969. Relationship of a
degenerative tropical neuropathy to diet, report of a field study. Br.
Med. J. 1:547-550.
Palmer, I.S., and O.E. Olson. 1979. Partial prevention by cyanide of selenium
poisoning in rats. Biochem. Biophys. Res. Commun. 90(4):1379-1386.
-------
Cyanide March 31, 1987
-15-
Philbrick, D.J., J.B. Hopkins, D.C. Hill, J.C. Alexander and R.G. Thomson.
1979. Effects of prolonged cyanide and thiocyanate feeding in rats.
J. Toxicol. Environ. Health. 5:579-592.
Rosehart, R.G. 1973. Mine water purification by reverse osmosis. Can. J.
Chem. Eng. 51(12):788-789.
Sheehy, M., and J.L. Way. 1968. Effect of oxygen on cyanide intoxication.
III. Mithridate. J. Pharmacol. Exp. Ther. 161:163-168.
Smith, R., M.S. Siebert and W.H.J. Hattingh. 1980. Removal of inorganic
pollutants from waste water during reclamation for potable reuse. Water
SA. 6(2):92-95.
Tewe, O.O., and J.H. Maner. 1981a. Long-term and carry-over effect of dietary
inorganic cyanide (KCN) in the life cycle performance and metabolism of
rats. Toxicol. Appl. Pharmacol. 58(1):1-7.
Tewe, O.O., and J.H. Mkner. 1981b. Performance and pathophysiological changes
in pregnant pigs fed cassava diets containing different levels of cyanide.
Res. Vet. Sci. 30(2):147-151.
Towill, L.E., J.S. Arury, B.L. Whitfield, E.B. Lewis, E.L. Galyan and A.S
Hammone. 1978. Reviews of the environmental fate of pollutants: V.
Cyanide. U.S. EPA Report No. EPA 600/1-78-027. Health Effects Research
Laboratory, Office of Research and Development, U.S EPA Cincinnati, OH.
Available through NTIS, Order No. PB 289920; Springfield, VA.
Trachtenberg, J.J., and M.A. Murphy. 1979. Removal of iron cyanide complexes
from waste water utilizing and ion exchange process. Light Metals J.
U.S. EPA. 1979. U.S. Environmental Protection Agency. Method 335.2. Titri-
metric; Spectrophotometric. In: Methods for Chemical Analysis of Water
and Wastes. EPA600/4-79-020, March 1979.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Ambient water quality
criteria for cyanides, with errata for ambient water quality criteria
documents dated June 9, 1981 (Updated: February 23, 1982). Environmental
Criteria and Assessment Office. Cincinnati, OH. NTIS PB 81-117483.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Health effects
criteria document for cyanide. Office of Drinking Water.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Final guidelines
for carcinogen risk assessment. Federal Register. 51(185):33992-34003.
September 24, 1986.
U.S. PHS. 1962. U.S. Public Health Service. Drinking water standards.
U.S. Govt. Printing Office, Washington, D.C. PHS Publ. No. 956.
Weast, R.C., ed. 1980. CRC handbook of chemistry and physics. 61st ed.
CRC Press, Inc., Boca Raton, FL. pp. B-98, B-133, B-147.
-------
Cyanide March 31, 1987
-16-
Wood, J.L., and S.L. Cooley. 1956. Detoxication of cyanide by cystine.
J. Biol. Chem. 218:449-457.
Yamamoto, K., Y. Yamamoto, H. Hattori and T. Samori, 1982. Effects of routes
of administration on the cyanide concentration distribution in the various
organs of cyanide-intoxicated rats. Tohoku J. Exp. Med. 137:73-78.
-------
March 31 , 1987
MERCURY
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAS are subject to
change as new information becomes available,
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group \ or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer anit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dosi= estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Mercury
March 31, 1937
-2-
This Health Advisory (HA) is based on information presented in the Officl
of Drinking Water's Health Effects Criteria Document (CD) for Mercury (U.S.
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd.f
Springfield, VA 22161, PB # 86-117827/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS Nos.
Mercury — 7439-97-5
Mercury (II) Chloride — 7437-94-7
Mercury (II) Sulfate — 7783-36-0
Synonyms
0 Mercury (II) chloride: mercuric bichloride; mercury perchloride.
Mercury (II) sulfate: mercuric sulfate.
Uses (U.S. EPA, 1985)
0 While this document is concerned with the toxic effects of ionic
mercury, it is metallic mercury that has the most uses. Some uses
of metallic mercury include use as a cathode in the electrolytic
preparation of chlorine and caustic soda, and in electrical apparatuse--,
dental amalgams, catalysts and in pulp and paper manufacture.
Properties (Weast, 1971)
0 The properties of inorganic mercury compounds vary with the specieLc
compound; some examples are as follows:
Chemical Formula
Mercury
Hg
Atomic/Molecular Weight 200.59
Mercury (II)
Chloride
HgCl2
271.49
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Log Octano I/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Silver liquid White powder
356.53°C
-38.87°C
13.5939
Insoluble
302°C
276°C
5.44
6.9 g/100
Mercury (II)
Sulfate
HgS04
296.65
White powder
Decomposes
6.47
(20°C) Decomposes
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Mercury March 31, 1937
-3-
Occurrence
Mercury, although a relatively rare element, is ubiquitous in the
earth's crust, occurring at levels from 10 to 500 ppb as a sulfide,
chloride or oxide. However, mercury can form organic compounds that
can bioaccumulate in the food chain and become a significant toxico-
logical concern. Only a small fraction of mercury in ground and
surface waters occurs in the organic form (U.S. EPA, 1987).
The majority of mercury used commercially in the United States is
imported. These commercial Uses have resulted in releases of mercury
and its compounds to surface waters. Naturally occurring levels of
mercury in ground and surface water are less than 0.5 ug/L, although
higher levels may occur in ground water from local mineral deposits.
Ground water surveys have found mercury at levels above 0.5 ug/L in
15 to 30% of wells tested. Surface water surveys report that about
20% of surface waters exceed 0.5 ug/L. State compliance data report
that 16 ground water and 16 surface water wells currently exceed the
maximum contaminant level of 2 ug/L (U.S. EPA, 1987).
III. PHARMACOKINETICS
Absorption
It is estimated that between 7 and 15% of orally administered inorganic
mercury is absorbed by humans (Rahola et al., 1971; Task Group on Metal
Accumulation, 1973).
Distribution
0 Rothstein and Hayes (1960) administered a single c^ose of 203Hg (as
Hg(N03)2/' °'2 '"SA?) by intravenous injection to seven male Wistar
rats. Distribution of mercury was primarily to kidney, liver, blood,
skin and muscle. Other tissues contained only fractional percentages
of the administered dose. In general, each tissue except the kidney
showed a maximum value four hours or one day post-treatment, followed
by rapid clearance. The kidney continued to accumulate mercury with
maximum concentrations reached at 6 to 1 5 days. For example, after
four hours, only 9% of the body burden of mercury was found in the
kidney; by the fifteenth day post-treatment 86% of the remaining
mercury was in the kidney.
0 Jugo (1976) administered single intravenous injections of 203^g (ag
HgCl2; 0.15 Tig/kg) to 2- or 21-week old female albino rats (strain
not specified). Approximately 28 and 51% of the administered dose
was present in the kidneys of the 2- and 21-week old rats, respectively,
after 144 hours. Approximately 9% of the dose was present in the liver
of 2-week old rats,- less than 1% was present in the liver of older
rats. In both groups of rats, the blood and brain contained less
than 1% of the administered dose.
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Mercury March 31, 1987
-4-
Metabolism
No information was found in the available literature on the metabolism
of inorganic mercury.
Excretion
Rahola at al. (1971) administered single oral doses of protein bound
methyl mercury (14 ug/subject) and inorganic mercury (6 ug/subject)
to human volunteers. Approximately 85% of the administered inorganic
mercury was eliminated in the feces within 4 to 5 days; only about
0.2% was excreted in the urine. After 50 days the daily excretion of
inorganic mercury in the urine and feces was about 0.02% of the admin:i
stered dose. Approximately 6% of the administered dose of methyl
mercury was eliminated in the feces within 3 to 4 days; excretion in
the urine was negligible at first, but increased slowly. After 100
days, 20% of the daily excretion of mercury was via the urine.
Rothstein and Hayes (1960) reported on the excretion of mercury in
rats administered single intravenous injections of Hg (as HgfNO^),;
0.2 mg/kg). These authors indicated that the clearance of mercury
from rats occurred in three phases: a rapid phase invoking 35% of
the dose lasting for a few days; a slower phase involving 50% of the
dose with a half-time of 30 days,- and a slow phase involving 15% of
the dose with a half-time of approximately 100 days. Since mercury
was found to accumulate in the kidney in the first few days following
dosing, the two slow phases of excretion represent primarily clearance
from the kidney.
IV. HEALTH EFFECTS
Humans
Short-term Exposure
Gleason, et al. (1957) estimated that the lethal oral dose for mercuri<
salts in humans is 1 to 4 g (equivalent to 14 to 57 mg/kg body weight)
Ingestion of a dose of 1.5 g of mercuric chloride (HgCl2) produced
emesis after 5 minutes, followed by severe abdominal pain with a
brief period of loss of consciousness (Pesce et al., 1977).
Long-term Exposure
No information was found in the available literature on the human
health effects of long-term exposure to inorganic mercury.
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Mercury March 31, 1987
-5-
Animals
Short-term Exposure
0 Male and female Brown-Norway rats (varying numbers per dose group)
were given subcutaneous injections of mercuric chloride, three times
per week for a maximum of 12 weeks. The dose levels administered
were 0, 0.05, 0.1, 0.25, 0.5, 1.0 or 2.0 mg/kg/injection. Rats that
received doses of 0.1 mg/kg/injection or higher developed renal
disease characterized by antiglomerular basement membrane antibodies
and the appearance of deposits in the glomerular tufts and in the
small renal arteries. Proteinurea and a nephrotic syndrome were also
observed in these rats. Based on these results, a NOAEL of 0.05
mg/kg/injection is identified (Druet et al., 1978).
Long-term Exposure
0 Male and female rats (strain not specified; 20 to 24/group) were
' administered mercury (as mercuric acetate) in the diet for up to 2
years at concentrations of 0, 0.5, 2.5, 10, 40 or 160 ppm. Assuming
that 1 ppm in the diet of rats is equivalent to 0.05 mg/kg/day (Lehman,
1959), these dose levels correspond to 0, 0.025, 0.125, 0.50, 2.0 or
8.0 mg/kg/day. At the highest dose level (3.0 mg/kg/day), body
weight was slightly depressed in male rats only (statistical significance
not specified). Kidney weights were significantly increased (p < 0.05)
in the 2.0 and 8.0 mg/kg/day dose groups. Pathological changes
originating in the proximal convoluted tubules were also noted at
these dose levels, with more severe effects in females than in males.
Based on these results, a NOAEL of 10 ppm (0.5 mg/kg/day) is identifier!.
A number of deficiencies limit the usefulness of this study. These
deficiencies include the small number of animals surviving past 18
months, lack of information on the number of animals in each group
having detectable pathological changes and the absence of statistical
analysis of body weight changes in males (Fitzhugh et al., 1950).
Reproductive Effects
0 No information was found in the available litsature on the reproductive
effects of inorganic mercury.
Developmental Effects
0 Oral dosing of Syrian golden hamsters with mercuric acetate on day 3
of gestation at levels ranging from 4 to 100 mg/kg produced a dose-
related response in numbers of resorptions and abnormal embryos.
While these findings were evident at the 4 mg/kg dose lovel, the
percentage of change was not significantly different from controls
at this low level (Gale, 1974).
Mutagenicity
0 No evidence is currently available to indicate that the mercuric
salts are mutagenic.
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Mercury March 31, 1987
-6-
•
Carcinogenic!ty
0 No evidence was found in the available literature on the carcinogenic!ty
of inorganic mercury.
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity,
The HAs for noncarcinogenic toxicants are derived using the following formula:
where:
HA = (NOAEL or LOAEL) x (BW) = __ mg/L ( __ ug/L >
(UF) x ( ___ L/day)
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Sffeet-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
The available data are insufficient to develop a One-day HA for mercury.
It is, therefore, recommended that the modified DWEL (1.58 ug/L) be used at
this time as a conservative estimate of the One-day HA value.
Ten-day Health Advisory
The available data are insufficient to develop a Ten-day HA for mercury<
It is, therefore, recommended that the modified DWEL (1.58 ug/L) be used at
this time as a conservative estimate of the Ten-day HA value.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for mercurv,
It is, therefore, recommended that the modified DWEL (1.58 uq/L) be used at
this time as a conservative estimate of the Longer-term HA value for the
10-kg child.
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Mercury March 31, 1987
-7-
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
The study by Druet et al. (1978) has been selected to serve as the basis
for the Lifetime Health Advisory. In this study, Brown-Norway rats were given
subcutaneous injections of mercuric chloride, three times per week for up to
12 weeks at dose levels of 0, 0.05, 0.1, 0.25, 0.5, 1.0 or 2.0 mg/kg/injection.
Kidney damage, characterized by proteinurea and a nephrotic syndrome, was
observed in rats that received doses of 0.1 mg/kg/injection or higher.
Based on these results, a NOAEL of 0.05 mg/kg/injection is identified.
Using this NOAEL, the Drinking Water Eguivalent Level and Lifetime Health
Advisory are derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (100) (0.05 mg/kg/injection) (0.739) (36) = 0.153 ug/kg/day
(10) (84 days) (1,000)
-------
Mercury March 31, 1987,
-8-
where-.
0.05 mg/kg/injection = NOAEL for absence of renal effects in rats.
36 = number of doses.
0.739 = percentage of mercury in mercuric chloride.
84 days = exposure period.
1,000 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study
of less-than-lifetime duration.
100/10 = assumed subcutaneous absorption factor relative
to ingestion.
Step 2: Determination of the Drinking Water Equivalent Level (DWEL)
DWEL = (0.158 ug/kg/day)(70 kg) = 5<5 ug/L
(2 L/day)
where:
0.158 ug/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of. an adult.
Step 3: Determination of the Lifetime Health Advisory
Lifetime HA = (5.5 ug/day) (20%) = 1 .1 ug/L
where:
5.5 ug/L = DWEL.
20% = assumed relative source contribution from water.
Evaluation of Carcinogenic Potential
0 The International Agency for Research on Cancer has not evaluated the
carcinogenic potential of mercury.
0 Applying the criteria described in EPA's guidelines for assessment of
carcinogenic risk (U.S. EPA, 1986), mercury may be classified in
Group D: not classified. This group is for substances with inadequate
animal evidence of carcinogenicity.
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Mercury March 31, 1987
-1 1-
IX. REFERENCES
Druet, P., E. Druet, F. Potdevin and C. Sapin. 1978. Immune type glomerulo-
nephritis induced by HgCL2 in the brown Norway rat. Ann. Immunol.
1290:777-792.
Ebersole, G. , and J.T. O'Connor. 1972. The removal of mercury from water by
conventional water treatment processes. Presented at 92nd Annual
Conference, American Water Works Association, Chicago, IL, June.
Fitzhugh, O.G., A. Nelson, E. Laug and F. Junze. 1950. Chronic oral
toxicants of mercuric-phenyl and mercuric salts. Arch. Ind. Occup. Med,
2:433-442.
Gale, T.F. 1974. Embryopathic effects of different routes of administration
of mercuric acetate in the hamster. Environ. Res. 8:207-213.
Gleason, M.N., R.E. Gosselin and H.C. Hodge. 1957. Clinical Toxicology of
Commercial Products. Baltimore, MD: Williams and wilkins Co., p. 154.
Jugo, S. 1976. Retention and distribution of HgCl, in suckling and- adult-
rats. Health Physics. 30:240-241.
Lehman, A.J. 1959. Appraisal of the safety of chemicals in foods, drugs and
cosmetics. Assoc. Food Drug Off. U.S., Q. Bull.
Logsdon, G.S., and J.M. Symons. 1973. Mercury removal by conventional water
treatment techniques. J. Amer. Water Works Assoc. 65(8) :554-562.
Pesce, A.J., I. Hanenson and K. Sethi. 1977. B% microglobulinuria in a
patient with nephrotoxicity secondary to mercuric chloride ingestion.
Clin. Toxicol. 1 1 (3) : 309-31 5.
Rahola, T. , T. Hattula, A. Korlainen and J.K. Miettinen. 1971. The biologi-
cal halftime of inorganic mercury (Hg'"*") in man. Scand. J. Clin. Invest.
27{suppl. 116) :77. (Abstract)
Rothstein, A., and A.D. Hayes. 1960. The metabolism of mercury in the rat
studied by isotope techniques. J. Pharmacol. 130:166-176.
Sigworth, E.S., and S.3. Smith. 1972. Adsorption of inorganic compounds by
activated carbon. J. Amer. Water Works Assoc. 64(6) : 386-91 .
Sorg, T.J. 1977. Manual of treatment techniques for meeting the interim
primary drinking water standards. U.S. Environmental Protection Agency,
EPA-600/8-77-005.
Sorg, T.J. 1979. Treatment technology to meet the interim primary drinking
water regulations for organics: Part 4. J. Amer. Water Works Assoc.
71 :454-66.
Task Group on Metal Accumulation. 1973. Accumulation of toxic metals with
special reference to their absorption, excretion and biological halftimes.
Environ. Phys. Biochem. 3:65-67.
-------
Mercury March 31, 1987
-12-
Theim, L., D. Badorek et al. 1976. Removal of mercury from drinking water
using activated carbon. J. Amer. Water Works Assoc. need volume 445-51 .
U.S. EPA. 1973. U.S. Environmental Protection Agency. Water Quality Criteria,
1972. Ecol. Res. Ser. Rep. Comm. of Water Quality Criteria. MAS, U.S.
GPO, Washington, DC. EPA R3/73/033.
U.S. EPA. 1979a. U.S. Environmental Protection Agency. Method 245.1. Manual
cold vapor technique. In: Methods for Chemical Analysis of Water and
Wastes, SPA-600/4-79-020.
U.S, EPA. 1979b. U.S. Environmental Protection Agency. Method 245.2.
Automated cold vapor technique. In: Methods for Chemical Analysis of
Water and Wastes, EPA-600/4-79-020.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality-
criteria for mercury. EPA 440/5-80-05b. Office of Water Regulations
and Standards, Washington, DC.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Drinking water
criteria document for mercury (draft report). Office of Drinking Water,
Washington, DC.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for car-
cinogen risk assessment. Fed. Reg. 51(185) :33992-34003. September 24.
U.S. EPA. 1987. U.S. Environmental Protection Agency. Estimated national
occurrence and exposure to mercury in public drinking water supplies.
Criteria and Standards Division. Office of Drinking Water, Washington, n •":,
Weast, R.C., ed. 1971. CRC handbook of chemistry and physics, 52nd ed.
Cleveland, OH: The Chemical Rubber Co.
WHO. 1971. World Health Organization. International standards for drinking
water. Geneva, Switzerland.
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Mercury
March 31, 1937
-9-
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
°- The U.S. EPA has recommended an ambient water quality criterion for
the protection of health of 144 ng/L (U.S. EPA, 1980) and for drinking
water of 2 ug/L (U.S. EPA, 1973).
0 A WHO expert group has recommended an international standard for
mercury in drinking water at 1 ug Hg/L (WHO, 1971).
VII. ANALYTICAL METHODS
0 Determination of mercury is by flameless atomic absorption using
either a manual cold vapor technique (U.S. EPA, 1979a) or an automated
cold vapor technique (U.S. SPA, 1979b).
0 The flameless atomic absorption procedure is a physical method based
on the absorption of radiation at 253.7 rnn by mercury vapor. The
mercury is reduced to the elemental state and aerated from solution in
a closed system. The mercury vapor passes through a cell positioned
in the light path of an atomic absorption spectronhotorneter. Absorbance
is measured as a function of mercury concentration. The detection
limit for mercury is 0.2 ug/L using either the manual or automated
technique.
VIII. TREATMENT TECHNOLOGIES
° Laboratory and pilot plant studies indicate that coagulation/filtration
is moderately effective in removal of inorganic mercury from drinking
water. Ferric sulfate coagulation achieved 66% removal at pH 7 and
97% removal at pH 8 from water containing 0.05 mg/L of inorganic
mercury. Alum coagulation was shown to be much less effective: 47%
of the mercury was removed at pH 7 and 33% at pH 8. • It has been
found that coagulation/filtration is less effective for removal of
organic mercury. However, the mercury removal efficiency of existing
coagulation/filtration systems can be improved by the addition of
powdered activated carbon (PAC) to the raw water influent. Laboratory
tests by Logsdon and Symons (1973) have shown that each milligram
per liter of PAC added removes 0.0001 mg/L of either inorganic or
organic mercury.
0 Lime softening is essentially ineffective for removal of organic
mercury taut moderately effective for removal of inorganic mercury,
depending on the pH of the water. Laboratory studies by Logsdon and
Symons (1973) have shown that in the 10.7-11.4 pH range, lime softening
removed 60 to 80% of the inorganic mercury, wheraas only about 30%
removal was achieved at oH 9.4.
The use of activated carbon as a process to remove mercury from
drinking water has been reported by various investigators (Logsdon
and Syrnons, 1973; Sigworth and Smith, 1972; Sorg, 1979: Theim et al.,
1976). Laboratory tests were performed by pumping solutions of tap
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Mercury March 31, 1987
-10-
water and either soluble inorganic or organic mercury through columns
of granular activated carbon for extended periods of time. The
results showed that «0 to 99% of the mercury may be removed from the
water by this technology (Sigworth, et al. 1972; Logsdon and Symons,
1973) .
Limited pilot-plant studies have been reported by Sorg (1977) on the
use of reverse osmosis for mercury removal. One study investigating
the removal of heavy metals, pesticides and other toxic chemicals
from secondary wastewater effluent resulted in inorganic and organic
mercury removals of 82 and 83%, respectively. Another test involved
a hollow fiber membrane with raw water flow of 1.25 gpm, 170 to 200
psi, and 40 to 50% water recovery. The spiral wound membrane system
showed a 25% mercury removal, while the hollow fiber system efficien'-,
removal was 79 to 81%.
Several preliminary ion exchange experiments have been carried out
by Ebersole and O'Connor (1972) to investigate organic and inorganic
mercury removal from drinking water. These studies showed that as
much as 98% of inorganic mercury added to distilled water could be
removed by cation and anion exchange resins operated in series.
Although these experiments were very preliminary, the results indi-
cated that ion exchange may be an effective method for inorganic
mercury removal.
-------
March 31, 1987
NICKEL
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored by the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Nickel
March 31, 1987
-2-
This Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for Nickel (U.S.
EPA, 1985). The HA and CD formats are similar for easy reference. Individuals
desiring further information on the toxicological data base or rationale for
risk characterization should consult the CD. The CD is available for review
at each EPA Regional Office of Drinking Water counterpart (e.g., Water Supply
Branch or Drinking Water Branch), or for a fee from the National Technical
Information Service, U.S. Department of Commerce, 5285 Port Royal Rd.,
Springfield, VA 22161, PB # 86-1 17801/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. area: (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS No.
0 Nickel — 7440-02-0
Nickel Chloride -- 7718-54-9
Nickel Oxide — 1313-99-1
Synonyms
0 Nickel Chloride: Nickelous Chloride
Nickel Oxide: Bunsenite
Uses (U.S. EPA, 1985)
0 While this document is concerned with the toxic effects of ionic nickel,
it is metallic nickel which has the most uses. Some uses of metallic
nickel include use in the manufacture of stainless steel, various
other alloys and in electroplating.
Properties (Weast, 1971)
0 The properties of nickel compounds vary with the specific compound;
some examples are as follows:
Chemical Formula
Atomic/Molecular Weight
Physical State
Boiling Point
Melting Point
Density
Vapor Pressure
Water Solubility
Log OctanoI/Water
Partition Coefficient
Taste Threshold
Odor Threshold
Nickel
Ni
58.71
silver metal
2,732°C
1,453°C
8.90
insoluble
Nickel
Chloride
NiCl2
129.62
yellow solid
973°C (sublimes)
1,001°C
3.55
Nickel
Oxide
NiO
74.71
green-black solid
1,990°C
6.67
64.2 g/100cc (20°C) --
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Nickel March 31, 1987
-3-
Occurrence
0 Nickel is a metallic element which is not found free in nature but
exists as a number of salts. Nickel compounds are found in most
geologic materials at levels up to several thousand ppm. Nickel
occurs at low levels in most surface and ground waters. Because
nickel compounds are relatively insoluble, the levels of nickel in
most surface and ground waters are less than 100 ppb. Since nickel
compounds are used commercially in a number of industries, contamination
of drinking water is the result of naturally occurring nickel compounds
proliferated during industrial activities (U.S. EPA, 1979a; 1983a).
0 Nickel is a component of some plumbing material. When pipes and other
materials corrode, nickel can be released to drinking water. However,
available information suggests that releases from this source are small
(U.S. EPA, 1979a; 1983a).
0 There are limited survey data on the occurrence of nickel in drinking
water. Based upon these data, most supplies contain less than 40 ug/L
of nickel. The highest level reported for a drinking water supply was
490 ug/L. Nickel also occurs at low levels in food. Based upon the
limited information available, diet is the major source of nickel
exposure with water making only a minor contribution (U.S. EPA, 1979a;
1983a).
Ill. PHARMACOKINETICS
The major routes of nickel intake for both humans and animals are
inhalation and ingestion, and to a lesser extent percutaneous
absorption. The extent of nickel absorption is dependent not only
on the concentration of inhaled or ingested nickel, but also on the
chemical and physical forms of nickel (U.S. EPA, 1985). Since inhala-
tion and percutaneous exposures are not relevant to drinking water,
emphasis will be placed on studies using the oral route of exposure.
Absorption
Very little of the nickel ingested in food is absorbed. Total dietary
intake of nickel ranges from 107 to 900 ug/day with average values of
160-500 ug daily (U.S. EPA, 1985); about 1-10% of this is absorbed
(Horak and Sunderman, 1973).
In rats, intubation of 6%i in dilute acid solutions resulted in 3-5%
absorption of radiolabelled nickel (Ho and Furst, 1973).
There was no uptake of nickel in rats chronically exposed to drinking
water at levels of 5 ppm over the lifetime of an animal (Schroeder
et al., 1974).
Transplacental transfer of nickel to the fetus takes place in both
humans and animals. Newborn rats of mothers fed 1000 ppm Ni in
the diet showed whole body levels of 22-30 ppm nickel (Phatak and
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Nickel March 31, 1987
-4-
Patwardhan, 1950). Also, similar levels of nickel (0.04-2.8 ppm)
were seen in the liver, heart and muscle of fetuses as were seen in
adult humans (Casey and Robinson, 1978).
0 Absorption from inhalation exposure to nickel carbonyl is both rapid
and extensive. Sunderman and Selin (1968) exposed rats to nickel
carbonyl at 100 mg Ni/L of air for 15 minutes. It was estimated that
half of the inhaled amount was initially absorbed. On the other
hand, inhalation exposure to insoluble particulate nickel (e.g., the
oxide or the subsulfide) results in very little absorption.
Distribution
0 The tissue distribution in animals orally exposed to Ni is dependent
upon the concentration of the compound. Calves fed supplemental
nickel in the diet at levels of 62.5, 250 or 1000 ppm showed somewhat
elevated levels of nickel in pancreas, testes and bone at 250 ppm;
pronounced increases were seen in these tissues at 1000 ppm (O'Dell
et al., 1971).
0 Weanling rats exposed to nickel (acetate) in diet up to levels of
1000 ppm showed increased levels of nickel in kidney, liver, heart
and testes as nickel concentration was increased, with the greatest
accumulation in the kidneys (Whanger, 1973).
Metabolism
0 Serum albumin is the main carrier protein for nickel in the sera of
humans, rabbits, rats and bovine species. In the sera of rabbits and
humans the nickel-rich metalloproteins ^ •]-macroglobulin (nickeloplasmin)
and 9.5 S C<"1-glycoprotein, respectively have been identified (NAS,
1975).
Excretion
The main excretory route of absorbed nickel in humans and animals
appears to be the urine (Ho and Furst, 1973) with biliary excretion
also occurring in experimental animals (Onkelinx et al., 1973). The
deposition of nickel in hair of humans also appears to be an excretory
mechanism (Nechay and Sunderman, 1973). Unabsorbed dietary nickel is
excreted in the feces.
IV. HEALTH EFFECTS
Humans
No clinical or epidemiologic studies dealing with the toxicity of
nickel following oral exposure were found in the available literature.
The toxicity of nickel to humans and animals is a function of the
chemical form of the element and the route of exposure. There has
been a suggestion of a correlation between chronic inhalation exposure
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Nickel March 31, 1987
-5-
to nickel carbonyl and respiratory tract cancer from epidemiclogical
studies which have been confirmed in experimental animals. Dermatitis
(nickel itch) is another frequent effect of exposure to nickel (SPA,
1983b). However, these data are not pertinent to the effects due to
ingestion of nickel in drinking water.
Animals
Short-term Exposure
0 The oral LD50 values converted to mg nickel/kg bw range from 105 mg/kg
bw for nickel chloride in male rats to 186 mg/kg for nickelocene in
mice (U.S. EPA, 1985).
0 Nickel chloride administered orally to rats at doses of 0.5 to 5.0
mg/kg/day for 2 to 4 weeks led to a significantly decreased thyroid
absorption of iodine (Lestrovoi et al., 1974).
0 Nickel acetate in the diet of weanling OSU brown rats for six weeks at
concentrations of 100, 500 or 1000 ppm (i.e., 10, 50 or 100 mg Ni/kg
bw) resulted in a significantly reduced weight gain at 500 ppm; rats
exposed to 1000 ppm lost weight. At 500 and 1000 ppm, there was a
dose-related decrease in blood hemoglobin concentration, packed cell
volume and plasma alkaline phosphate activity. Cytochrome oxidase
activity was decreased significantly (p< 0.005) in both heart and liver
in the high-dose group. Iron concentration was increased significantly
(p< 0.05) in red blood cells, heart, kidney, liver and testes in the
1000 ppm group; elevated levels of iron concentration also were seen
in the 500 ppm group. No significant effects were seen on body weights,
mineral content and enzyme activity in the 100 ppm group in comparison
with control levels. The 100 ppm (10 mg Ni/kg bw) is considered a
NOAEL while 500 ppm (50 mg Ni/kg bw) is a LOAEL (Whanger, 1973).
Long-term Exposure
* Nickel added to the diet of mice resulted in reduced body weight gain
in females at a dietary concentration of 1100 ppm nickel and reduced
body weight gain in both males and females at 1600 ppm (Weber and
Reid, 1969b).
0 Studies in chicks (Weber and Reid, 1968a; Ling and Leach, 1979) and
calves (O'Dell et al., 1970) have shown adverse effects at dietary
levels ranging from 250 to 700 mg Ni/kg diet.
0 Nickel (as nickel chloride) administered to rats at a concentration
of 225 ppm in drinking water (17.6 mg Ni/kg bw) for four months led
to a significant reduction in body weight (p< 0.05) compared with
controls (Clary, 1975). Daily urinary volume and urinary zinc and
calcium concentrations were reduced significantly. Also, at sacrifice,
serum lipid and cholestrol concentrations were reduced significantly
(p< 0.05).
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Nickel March 31, 1987
-6-
Daily doses of 25 mg/kg bw of nickel sulphate administered by oral
intubation to male rats for 120 days caused degenerative cellular
changes in the liver and kidney (von Waltschewa et al., 1972). In
the treated rats, testes were smaller than in controls. Other testicular
changes included interstitial cell proliferation, transparent vessel
walls, reduced number of spermatozoa and their precursors and decreased
concentrations of succinodehydrogenase and steroid 3-a-dehydrogenase.
Rats fed a diet containing nickel acetate at concentrations of 0.1 to
10% (16.6-166 mg Mi/kg bw) for 10-190 days led to a high rate of
mortality, hypoplasia of bone marrow, thymus and spleen, progressive
renal tubular degeneration, mural exudative pulmonary alveolar lesions
and noninflammatory lysis of pancreatic exocrine cells (Ashrof and
Sybers, 1974).
In a chronic study with mice fed a diet devoid of cadmium and Low in
other metals with 5 ppm nickel added to their drinking water (approxi-
mately 0.85 mg Ni/kg bw/day) no significant effects were observed.
Only body weights of animals dying after one year were depressed by
4% to 13% over controls (Schroeder et al., 1964).
The mean body weights of both male and female rats were reduced
significantly (p<0.025) compared to controls at 18 months in a study
where rats were administered 5 ppm nickel (average daily dose estimated
to be 0.41 mg Ni/kg bw) in drinking water for life (Schroeder et al.,
1974). Lifespan was not affected. Histopathology revealed an increased1
incidence (p<0.025) of focal myocardial fibrosis (13.3%) in the experi-
mental group compared to the control.
In a two-year feeding study with beagle dogs administered nickel
sulfate hexahydrate at dietary levels of 0, 100, 1,000 or 2,500 pom
(0, 3, 29 or 70 mg Ni/kg bw), no significant effects on body weight,
hematology, urinalysis, organ-to-body weight ratios or histopathology
were noted at 100 or 1,000 ppm. At 2,500 ppm, body weight gain was
depressed, hemoglobin and hematocrit values tended to be lower and
kidney- and liver-to-body weight ratios were significantly higher
(p <0.05). Pathological changes in the lungs and granulocytic hyper-
plasia of the bone also were noted in the high dose group. Based on
these findings, the NOAEL from this study is 1,000 ppm (29 mg/kg bw)
(Ambrose et al., 1976).
In a two-year feeding study in rats given 0, 100, 1,000 or 2,500 ppm
nickel sulfate in milk (0, 5, 50 and 125 mg Ni/kg bw), no significant
effects were reported at 100 ppm (Ambrose et al., 1976). Body weight
was reduced significantly (p<0.05) in both male and female rats fed
2,500 ppm nickel when compared with controls. At 1,000 ppm, body
weight also was reduced in both sexes. Animals fed 1,000 or 2,500 ppm
nickel diets had significantly higher (p<0.05) heart-to-body weight
ratios and significantly lower liver-to-body weight ratios than
controls. The 1,000 ppm (50 mg/kg bw) represents a LOAEL from this
study and 100 ppm (5 mg/kg bw) is a NOAEL.
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Nickel March 31, 1987
-7-
Reproductive Effects
0 In a three-generation reproduction study in rats, nickel sulfate
hexahydrate fed at levels of 0, 250, 500 or 1,000 ppm (0, 12.5, 25 or
50 mg Ni/kg/day) led to a slight decrease in adult body weight at
mating and weaning in the 1000 ppm group over controls. Fertility,
gestation, viability and lactation indices were not affected. The
body weights of weanlings from the 1,000 ppm group were reduced in all
generations. The incidence of stillborn pups was 19%, 12% and 15% in
the F-|a and 4%, 20% and 25% in the F-|b generations in the 250, 500 and
1000 ppm groups, respectively, compared to 4% and 2% in the control
F1a and F-JJ-, generations. Elevated incidence of fetal mortality was not
observed in the F2 and F3 generations (Ambrose et al., 1976).
0 In another three-generation reproduction study, rats were provided
drinking water containing 5 ppm nickel (salt not specified, estimated
total daily dose was 0.43 mg/kg) (Schroeder and Mitchener, 1971).
Neonatal mortality was increased significantly (p <0.025) in all
generations of exposed rats compared to controls. The number of
runts were increased significantly in the first (p <0.025) and third
(p <0.0001) generations. Average litter size was reduced somewhat
in the F3 generation. In this study, the diet was found to be deficient
in trace metals (particularly chromium).
0 No significant differences were observed in the litter size and
initial body weight of pups when male and female rats were fed diets
containing 250, 500 or 1,000 ppm nickel (daily dose of 10, 20 or 40 mg
Ni/kg bw) for 8 weeks before breeding and continuing through lactation
(Phatak and Patwardhan, 1950).
Developmental Effects
0 Transplacental transfer of nickel is well documented in laboratory
animals (U.S.EPA, 1985).
0 In a three-generation reproduction study in rats (Ambrose et al., 1976)
(described above) no evidence of teratogenicity was seen in weanlings
of rats fed nickel sulfate hexahydrate at levels of 0, 250, 500 or
1,000 ppm (0, 12.5, 25 or 50 mg Ni/kg/day).
Mutagenicity
0 Nickel chloride was not mutagenic in Escherichia coli and Bacillus
subtilis (U.S. EPA, 1985).
0 Nickel chloride and nickel sulfate were mutagenic or weakly mutagenic
in eukaryotic test systems (U.S. EPA, 1985).
Nickel induced chromosomal aberrations in cultured mammalian cells
and sister chromatid exchanges in both cultured mammalian cells and
in human lymphocytes (U.S. EPA, 1985).
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Nickel March 31, 1987
-8-
Carcinogenicity
0 It has been demonstrated that the incidence of respiratory tract
cancers in nickel refinery workers is statistically significantly
elevated (NIOSH, 1977; IARC, 1976; NAS, 1975); these data are not,
however, relevant to the consumption of nickel in drinking water.
0 Repeated i.p. injections of nickel acetate at a dose of 360 mg/kg
have induced lung carcinomas in mice (Stoner et al., 1976). This is
not, however, relevant to the consumption of nickel in drinking water.
0 No evidence of carcinogenicity has been found in those chronic studies
in which nickel was administered orally to laboratory animals (Schroedsr
et al., 1964, 1974; Schroeder and Mitchner, 1975).
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a sensitive noncarcinogenic end point of toxicity.
The HAs for noncarcinogenic toxicants are derived using the following formula:
HA = (NOAEL or LOAEL) x (BW) = mg/L ( ug/L)
(UF) x ( L/day)
where:
NOAEL or LOAEL = No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
BW = assumed body weight of a child (10 kg) or
an adult (70 kg).
UF = uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
One-day Health Advisory
The available data are insufficient to develop a One-day HA for nickel.
It is recommended that the Ten-day HA of 1.0 mg/L be used as the One-day HA
for the 10-kg child.
Ten-day Health Advisory
The study by Whanger (1973) has been selected for the derivation of a
Ten-day HA. Dose-response relationships were observed in this 6-week dietary
study defining a NOAEL for nickel of 100 ppm in diet (10 mg/kg bw/day) and a
LOAEL of 500 ppm in diet (50 mg/kg bw/day). The biological endpoints included
body weight gain, hematology parameters and cytochrome oxidase activity.
The Ten-day HA for Ni for a 10-kg child is calculated as follows:
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Nickel March 31, 1987
-9-
Ten-day HA = (10 mg/kg/day) (10 kg) = T.0 mg/L (1,000 ug/L)
(1 L/day) (100)
where:
10 mg Ni/kg bw/day = NOAEL for absence of effects on weight gain,
hematology parameters and cytochrome oxidase
activity in rats following 6-week oral exposure.
10 kg = assumed body weight of a child.
100 = uncertainty factor, chosen in accordance with
NAS/ODW guidelines for use with a NOAEL from an
animal study.
1 L/day = assumed daily water consumption of a child.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for nickel.
It is recommended that the OWEL of 0.35 mg/L be used as the Longer-term HA
for the 70-kg adult and the modified DWEL of 0.1 mg/L (adjusted for a JO-kg
child) be used as the Longer-term HA for the 10-kg child.
The Agency is in the process of reviewing a draft report of a 90-day
gavage study in rats (Mayhew, 1987). The final report is expected to be
available in July or August, 1987. After the official final report has been
reviewed and considered, it may serve as the basis for a longer-term health
advisory.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
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Nickel March 31, 1987
-10-
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
Because of various problems with the two teratogenicity/reproductive
toxicity studies of Schroeder and Mitchner (1971) and Ambrose et al.,
(1976), the two-year rat feeding study of Ambrose et al., (1976) is used for
the derivation of the Lifetime HA. In this study, rats were given 0, 100,
1,000 or 2,500 ppm nickel sulfate (approximate daily dose was 0, 5, 50 or
125 mg Ni/kg bw) in their diet. No significant effects were reported at
100 ppm. Body weight was reduced significantly (p <0.05) in both male and
female rats fed 2500 ppm nickel compared to controls. At 1000 ppm also, the
body weight was reduced for the male and female rats. The NOAEL identified
in this study is 100 ppm (5 mg/kg bw) .
Using this NOAEL, the Lifetime Health Advisory is derived as follows:
Step 1: Determination of the Reference Dose (RfD)
RfD = (5 mg/kg/day) = 0.01 mg/kg/day
(100) (5) y/ ^ y
where:
5 mg/kg/day = NOAEL for absence of effects on weight gain in rats.
100 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with a NOAEL from an animal study.
5 = additional uncertainty factor selected to allow for
possibly greater absorption of nickel from water than
from the diet.
Step 2: Determination of the Drinking Water Equivalent (DWEL)
DWEL = (0-01 mg/kg/day) (70 kg) = 0>35 mg/L (350 ug/L)
(2 L/day)
where:
0.01 mg/kg/day = RfD.
70 kg = assumed body weight of an adult.
2 L/day = assumed daily water consumption of an adult.
Step 3: Determination of the Lifetime Health Advisory
The DWEL of 350 ug/L assumes 100% of the exposure to nickel occurs via
drinking water. The available data indicate that the estimated intake of
nickel from food and air are 400 ug/day and 0.6 ug/day (negligible), respec-
tively. Factoring in these data on human exposure, a Lifetime HA of
0.150 mg/L (150 ug/L) would result.
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Nickel March 31, 1987
-11-
Evaluation of Carcinogenic Potential
0 Nickel has not been shown to be carcinogenic through oral exposure.
Data are not available concerning the potential carcinogenic effects
of ingested nickel compounds in humans.
0 A relative high degree of evidence exists to demonstrate that certain
nickel compounds or mixtures of nickel compounds are carcinogenic to
humans via inhalation. Nickel refinery dust and nickel subsulfide
(which is believed to be the major nickel component of the refinery
dust) are classified in Group A: Human carcinogen, based on the EPA
final guidelines for assessment of carcinogen risk (U.S. EPA, 1986).
In the case of nickel carbonyl, while there is insufficient evidence
from epidemiological studies, there is sufficient evidence from
animal studies to classify it in Group B2: Probable human carcinogen.
0 Based upon an evaluation of the carcinogenic potential of nickel from
inhalation and intramuscular injection, IARC has concluded that nickel
and certain nickel compounds are group 2A chemicals (IARC, 1976).
However, at the present time there is insufficient evidence to classify
nickel as a carcinogen following oral exposure.
0 Applying the criteria described in EPA's final guidelines for assess-
ment of carcinogenic risk (U.S. EPA, 1986), nickel via inhalation or
intramuscular injection may be classified in Group B: Probable human
carcinogen. This category is for agents for which there is inadequate
evidence from human studies and sufficient evidence from animal
studies. However, as there are inadequate data to conclude that
nickel is carcinogenic via ingestion, nickel is dealt with here as
Group D: Not classifiable as to human carcinogenicity. This category
is for agents with inadequate human and animal evidence of carcino-
genicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
ACGIH (1983) has established a TWA-TLV of 1.0 mg Ni/m3 for metallic
nickel salts and 0.1 mg Ni/m3 for soluble nickel salts.
0 The NIOSH (1977) criterion for occupational exposure to nickel is a
TWA of 15 ug Ni/m3.
0 EPA (U.S. EPA, 1980; 1982) derived an ADI of 1.46 mg Ni/day and
established an ambient water quality criterion of 0.632 mg Ni/L.
VII. ANALYTICAL METHODS
Determination of nickel is by atomic absorption (AA) using either
direct aspiration into a flame (U.S. EPA, 1979b) or a furnace technique
(U.S. EPA, 1979c).
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Nickel March 31, 1987
-12-
0 The direct aspiration AA procedure is a physical method based on the
absorption of radiation at 232.0 nm by nickel. The sample is aspirated
into an air-acetylene flame and atomized. A light beam is directed
through the flame into a monochromator and onto a detector that
measures the amount of light absorbed. Absorbance is proportional to
the concentration of nickel in the sample. The detection limit is
40 ug/L using this procedure.
0 The furnace AA procedure is similar to direct aspiration AA except a
furnace, rather than a flame, is used to atomize the sample. The
detection limit is 1 ug/L using this procedure.
VIII. TREATMENT TECHNOLOGIES
0 Treatment techniques that may be capable of removing nickel from
drinking water include lime softening, ion exchange and reverse osmosis.
Conventional coagulation is moderately effective in removing nickel
from drinking water. Although the removal of nickel from drinking
water supplies by these technologies have not been extensively studied,
some information is available from waste water technology surveys.
0 Gulp et al. (1978) reported excellent removal of nickel with lime
softening, ranging from 90.9 to 99.9 percent, for wastewater with
nickel concentrations from 5 mg/L to 160 mg/L. Maruyama et al.
(1975) reported removal efficiencies of 95 percent with low lime
softening (260 mg/L lime dosage) and 98 percent with high lime softening
(600 mg/L dosage) from domestic wastewater containing 5 mg/L of nickel.
0 Cation exchange has been used extensively in the plating industry to
recover nickel. Normally, these operations have employed cation resins
in the hydrogen cycle because of the need to recover both acid and
metal for recycle. Nickel was eluted with sulfuric acid, 6 to 10 Ib
H^SO^/ft^ of resin in 10 percent solution. The reported efficiencies
of removing nickel from plating industry wastewater are 96 to 100
percent (Keramida and Etzel, 1982).
0 Reverse osmosis (RO) membranes have been tested and shown to remove
nickel effectively from source water. A laboratory scale study evalu-
ating the performance of cellulose acetate membrane with plating rinse
showed that cellulose acetate has a rejection efficiency for Ni+2 Of
99.6 percent. Other membranes are commercially available: cellulose
acetate butyrate, nylon hollow fibers, polyurethanes (Golomb, 1972).
These membranes, however, have not been tested for their efficiencies
to remove nickel. The cellulose acetate membrane was field tested on
a small industrial automatic plating line. The wastewater nickel
concentration was varied: 1,700 mg/L, 50 mg/L, 12 mg/L. Tests by
Golomb (1974) have shown that cellulose acetate membrane can be used
to remove effectively 99+ percent of nickel from the waste rinse streams.
0 Pilot plant studies evaluating the efficiency of coagulation indicated
that alum was only 25-45 percent effective to remove nickel from water
(Maruyama et al., 1975; Hannah et al., 1977).
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Nickel March 31, 1987
-13-
Another study by Uillson (1978), investigating the removal on trace
metals from tap water and municipal wastewater, determined the effi-
ciency of calcium hydroxide proved to be 91.3 percent effective in
removing nickel from tap water and 63.3 percent effective in removing
nickel from wastewater at a pH of 9.5.
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Nickel March 31, 1987
-14-
IX. REFERENCES
ACGIH. 1983. American Conference of Governmental Industrial Hygienists.
TLVs. Threshold limit values for chemical substances and physical agents
in the work environment with intended changes for 1983-84. Cincinnati,
OH. p. 27.
Ambrose, A.M., P.S. Larson, J.R. Borzelleca and G. R. Hennigar, Jr. 1976.
Long-term toxicologic assessment of nickel in rats and dogs. J. Food
Sci. Technol. 13:181-187.
Ashrof, M., and H.D. Sybers. 1974. Lysis of pancreatic exocrine cells and
other lesions in rats fed nickel acetate. Amer. J. Pathol. 74:102a.
Casey, C.E. and M.F. Robinson. 1978. Copper, manganese, zinc, nickel,
cadmium, and lead in human foetal tissues. Br. J. Nutr. 39:639-646.
Clary, J.J. 1975. Nickel chloride - induced metabolic changes in the rat
and guinea pig. Toxicol. Appl. Pharmacol. 31:55-65.
Gulp, R.J., G.M. Wesner et al. 1978. Handbook of Advanced Wastewater
Treatment. 2nd. Van Nostrand Reinhold Co.
Golomb, A. 1972. Application of reverse osmosis to electroplating waste
treatment. Plating 59(4):316-19.
Golomb, A. 1974. Application of reverse osmosis to electroplating waste
treatment. Plating 61(5):432-42.
Hannah, S.A,, M. Telus and J.M. Cohen. 1977. Removal of uncommon trace
metals by physical and chemical treatment processes. Journal WPCF
49(11):2297-309.
Ho, W., and A. Furst. 1973. Nickel excretion by rats following a single
treatment. Proc. West. Pharmacol. Soc. 16:245-248.
Horak, E., and P.M. Sunderman, Jr. 1973. Fecal nickel excretion by healthy
adults. CLin. Chem. 19:429-430.
IARC. 1976. International Agency for Research on Cancer. Nickel and nickel
compounds. IARC Monographs. 2:75-112.
Keramida, V., and J.E. Etzel. 1982. Treatment of metal plating wastewater
with a disposable ion exchange material. In: Proceedings of the 37th
Industrial Waste Conference. Purdue University.
Lestrovoi, A.P., A.I. Itskova and I.N. Eliseev. 1974. Effect of nickel on
the iodine fixation of the thyroid gland when administered perorally and
by inhalation. Gig. Sanit. 10:105-106.
Ling, J.R., and R.M. Leach. 1979. Studies on nickel metabolism: Interaction
with other mineral elements. Poultry Sci. 58(3):591-596.
-------
Nickel March 31, 1987
-15-
Maruyama, T., S.A. Hannah and J.M. Cohen, 1975. Metal removal by physical
and chemical treatment processes. Journal WPCF 47(5):962-75.
Mayhew, D.A. 1987. Ninety-day gavag« study in albino rats using nickel.
Draft final report by American Biogenics Corp., Decatur, IL.
NAS. 1975. National Academy of Sciences. Nickel. National Academy of
Sciences Committee on Medical and Biological Effects of Environmental
Pollutants. Washington, DC.
i
Nechay, M.W., and F.W. Sunderman, Jr. 1973. Measurements of nickel in hair by
atomic absorption spectrometry. Ann. Clin. Lab. Sci. 3:30-35.
Nillson, R. 1978. Removal of metals by chemical treatment of municipal
waste water. Water Research. 5:51-60.
NIOSH. 1977. National Institute of Occupational Safety and Health. Criteria
for a recommended standard.. .occupational exposure to inorganic nickel.
NIOSH Publ. No. 77-164. Washington, DC.
O'Dell, G.D., W.J. Miller, A, King, S.L. Moore and D.M. Blackmon. 1971.
Effect of dietary nickel level on excretion and nickel content of tissues
in male calves. J. Anim. Sci. 32:769-7730.
Onkelinx, C. 1973. Compartmental analysis of the metabolism of 6%i(II) in
rats and rabbits. Res. Comm. Chem. Pathol. Pharmacol. 6:663.
Phatak, S.S., and V.N. Patwardhan. 1950. Toxicity of nickel. J. Sci. Ind. Res
9B:70-76.
Schroeder, H.A., J.J. Balassa and W.H. Vintin, Jr. 1964. Chromium, lead,
cadmium, nickel and titanium in mice: Effect on mortality, tumors and
tissue levels. J. Nutr. 83:239-250.
Schroeder, H.A., and M. Mitchener. 1971. Toxic effects of trace elements on
the reproduction of mice and rats. Arch. Environ. Health. 23:102-106.
Schroeder, H.A., M. Mitchener and A.P.Mason. 1974. Life-term effects of
nickel in rats: survival, tumors, interactions with trace elements and
tissue levels. J. Nutr. 104:239-243.
Schroeder, H.A., and M. Mitchener. 1975. Life-term effects of mercury, methyl
mercury and none other trace metals on mice. J. Nutr. 105:452-458.
Stoner, G.D., M.B. Shimkin, M.C. Troxell, T.L. Thompson and L.S. Terry.
1976. Test for carcinogenicity of metallic compounds by the pulmonary
tumor response in strain A mice. Cancer Res. 36:1744-1747.
Sunderman, F.W., Jr., and C.E. Selin. 1968. The metabolism of nickel-63
carbonyl. Toxicol. Appl. Pharmacol. 12:207.
U.S. EPA. 1979a. U.S. Environmental Protection Agency. Water related
environmental fate of 129 priority pollutants. Office of Water Planning
and Standards. EPA-440/4-79-029.
-------
Nickel March 31, 1987
-16-
U.S. EPA. 1979b. U.S. Environmental Protection Agency. Method 249.1.
Atomic Absorption, direct aspiration. In: Methods for Chemical Analysis
of Water and Wastes. EPA-600/4-79-020.
U.S. EPA. 1979c. U.S. Environmental Protection Agency. Method 249.2.
Atomic Absorption, furnace technique. In: Methods for Chemical Analysis
of Water and Wastes. EPA-600/4-79-020.
U.S. EPA. 1980. U.S. Environmental Protection Agency. Ambient water quality
criteria document for nickel. Environmental Criteria and Assessment
Office, Cincinnati, OH. EPA 440/4-80-060. NTIS PB 81-117715.
U.S. EPA. 1982. U.S. Environmental Protection Agency. Errata for ambient
water quality criteria documents. February 23. p. 14.
U.S. EPA. 1983a. U.S. Environmental Protection Agency. Nickel occurrence
in drinking water, food and air. Office of Drinking Water.
U.S. EPA. 1983b. U.S. Environmental Protection Agency. Health assessment
document for nickel. Office of Research and Development. Environmental
Criteria and Assessment Office. Research
Triangle Park, NC. EPA-600/8-83-012.
U.S. EPA. 1985. U.S. Environmental Protection Agency. Drinking water
criteria document for nickel. Environmental Criteria and Assessment
Office, Cincinnati, OH. EOA-600/X-84-193-1.
U.S. EPA. 1986. U.S. Environmental Protection Agency. Guidelines for
carcinogen risk assessment. Federal Register. 51 (1 85): 33992-34003.
September 24.
Von Waltschewa, W., M. Slatewa and I. Michailow. 1972. Hodenveranderungen
bei weissen Ratten durch chronische Verabreichung von Nickel sulfat.
(Testicular changes due to long-term administration of nickel sulphate
in rats.) Exp. Pathol. 6:116-120. (Ger. with Eng. Abstr.)
Weast, R.C., ed. 1971. CRC handbook of chemistry and physics, 52nd ed.
Cleveland, OH: The Chemical Rubber Co.
Weber, C.W., and B.L. Reid. 1969a. Nickel toxicity in young growing chicks.
J. Nutr. 95:612-616.
Weber, C.W., and B.L. Reid. 1969b. Nickel toxicity in young growing mice.
J. Anim. Sci. 28:620-623.
Whanger, P.O. 1973. Effects of dietary nickel on enzyme activities and
mineral content in rats. Toxicol. Appl. Pharmacol. 25:323-331.
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March 31, 1987
NITRATE/NITRITE
Health Advisory
Office of Drinking Water
U.S. Environmental Protection Agency
I. INTRODUCTION
The Health Advisory (HA) Program, sponsored fay the Office of Drinking
Water (ODW), provides information on the health effects, analytical method-
ology and treatment technology that would be useful in dealing with the
contamination of drinking water. Health Advisories describe nonregulatory
concentrations of drinking water contaminants at which adverse health effects
would not be anticipated to occur over specific exposure durations. Health
Advisories contain a margin of safety to protect sensitive members of the
population.
Health Advisories serve as informal technical guidance to assist Federal,
State and local officials responsible for protecting public health when
emergency spills or contamination situations occur. They are not to be
construed as legally enforceable Federal standards. The HAs are subject to
change as new information becomes available.
Health Advisories are developed for One-day, Ten-day, Longer-term
(approximately 7 years, or 10% of an individual's lifetime) and Lifetime
exposures based on data describing noncarcinogenic end points of toxicity.
Health Advisories do not quantitatively incorporate any potential carcinogenic
risk from such exposure. For those substances that are known or probable
human carcinogens, according to the Agency classification scheme (Group A or
B), Lifetime HAs are not recommended. The chemical concentration values for
Group A or B carcinogens are correlated with carcinogenic risk estimates by
employing a cancer potency (unit risk) value together with assumptions for
lifetime exposure and the consumption of drinking water. The cancer unit
risk is usually derived from the linear multistage model with 95% upper
confidence limits. This provides a low-dose estimate of cancer risk to
humans that is considered unlikely to pose a carcinogenic risk in excess
of the stated values. Excess cancer risk estimates may also be calculated
using the One-hit, Weibull, Logit or Probit models. There is no current
understanding of the biological mechanisms involved in cancer to suggest that
any one of these models is able to predict risk more accurately than another.
Because each model is based on differing assumptions, the estimates that are
derived can differ by several orders of magnitude.
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Nitrate/Nitrite March 31, 1987
-2-
This Health Advisory (HA) is based on information presented in the Office
of Drinking Water's Health Effects Criteria Document (CD) for nitrate and
nitrite (U.S. EPA, 1985). The HA and CD formats are similar for easy reference.
Individuals desiring further information on the toxicological data base or
rationale for risk characterization should consult the CD. The CD is available
for review at each EPA Regional Office of Drinking Water counterpart (e.g.,
Water Supply Branch or Drinking Water Branch), or for a fee from the National
Technical Information Service, U.S. Department of Commerce, 5285 Port Royal
Rd., Springfield, VA 22161, PB #86-117959/AS. The toll-free number is (800)
336-4700; in the Washington, D.C. areas (703) 487-4650.
II. GENERAL INFORMATION AND PROPERTIES
CAS tio.
0 Potassium Nitrate — 7757-79-1
Potassium Nitrite -- 7758-09-0
Synonyms
0 Potassium Nitrate: Saltpeter (Windholz, 1976)
Uses
Among other uses, nitrate and nitrite have a variety of uses including
the following (U.S. EPA, 1985):
0 The major use of nitrate is in inorgaiic fertilizers.
0 Nitrate is used in the manufacture of explosives, glassmaking and as
a heat-transfer fluid and a heat-storage medium for solar heating
applications.
0 Both nitrate and nitrite are used in curing meats.
Properties (Weast, 1974)
e The properties of nitrate and nitrite compounds vary with the specific
compound; some examples are as follows:
Potassium Potassium
Nitrate Nitrite
Chemical Formula KNO3 KNO2
Molecular Weight 101.11 85.11
Physical State solid ... _ solid
Boiling Point 400°C (decomposes) decomposes
Melting Point 334°C 440°C
Density 2.109 (16 C) 1.915
Vapor Pressure ' —
Water Solubility (0°C) 13.3 g/100cc 281 g/100cc
Log Octanol/Water
Partition Coefficient
Taste Threshold —
Odor Threshold —
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Nitrate/Nitrite
Occurrence
March 31, 1987
-3-
Nitrate and nitrite are naturally occurring inorganic ions which make
up part of the nitrogen cycle. Wastes containing organic nitrogen
enter the soil and are decomposed first to ammonia which is subsequently
oxidized to nitrite and nitrate. Because nitrite is easily oxidized
to form nitrate, nitrate predominates in ground and surface waters.
Nitrate then is taken up by plants during their growth and converted
back to organic form. Levels of nitrate in water can be raised as
the result of the contamination by nitrogen containing fertilizers or
human and animal wastes. Nitrate and nitrite ions are very mobile in
soil and readily move with ground water (U.S. EPA, 1987).
Surveys of naturally occurring levels of nitrate and nitrite in ground
and surface water have found that levels normally do not exceed 1 to 2
mg/L for nitrate and 0.1 mg/L for nitrite. Surface waters generally
contain lower levels of nitrate and nitrite than ground water.
Nitrate has been included in a number of drinking water surveys.
Nitrates occur at levels of less than 1 mg/L in most surface and
ground water supplies. Nitrates occur at levels exceeding 5 mg/L in
about 3% of surface waters and 6% of ground waters. Currently, 40
surface water supplies and 568 ground water supplies exceed the
nitrate MCL of 10 mg/L. Systems which exceed the MCL are usually
contaminated by nitrates from the use of fertilizers or from animal
wastes or septic systems. Nitrite levels have not been surveyed in
drinking water supplies but are expected to be much lower than 1 mg/L
(U.S. EPA, 1987).
Nitrates occur naturally in a number of foods, particularly vegetables.
Nitrates also are added to meat products as a preservative. For
adults, the major source of nitrates appears to be from dietary
sources. For infants, water appears to be the major source of exposure
(U.S. EPA, 1987).
III. PHARMACOKINETICS
Absorption
Both nitrate and nitrite are readily and completely absorbed following
oral administration:
0 Nitrate is absorbed by active transport from the upper small intestine
and nitrite is absorbed by diffusion across the gastric mucosa and
also through the wall of the intestinal tract (U.S. EPA, 1985).
0 Following oral administration, both nitrate and nitrite are readily
and completely absorbed: both 13N05 and 1^O^ were completely
absorbed within ten minutes after administration of 10 to 100 mgAg
in mice (Parks et al., 1981). Similar results for nitrate (dose not
• specified) in rats were reported by Witter.et al. (1979).
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Nitrate/Nitrite March 31, 1987
-4-
Distribution
Both nitrate and nitrite readily distribute throughout the tissues
but do not bioaccumulate:
0 Rapid, homogeneous distribution of nitrate (dose unspecified) was
observed in rats 45 to 60 minutes after dosing by gavage (Witter
et al., 1979) .
0 Both 1%03~ and 1 ^NC^" achieved transient equilibrium in mice within
five minutes after intratracheal administration of 10 to 100 mg/kg
(Parks et al., 1981). Equilibrium between the intravascular and
extravascular compartments of rabbits was reached within five minutes
after injection of either radiochemical into rabbits.
0 Nitrate secretion in saliva by humans was reported by Spiegelhalder
et al. (1976) after ingestion of vegetables and vegetable juices.
Secretion of nitrate by the gastric mucosa in rats was observed by
Bloomfield et al. (1962) following intraperitoneal doses of sodium
nitrate ranging from 60 to 200 mg/kg.
0 In rats, nitrite has been shown to cross the placenta (Shuval and
Greuner, 1977).
0 No evidence was found for bioaccumulation of nitrate or nitrite in
any tissue (U.S. EPA, 1985).
Metabolism
While nitrate is not directly metabolized to other compounds in humans,
nitrate is metabolized by bacteria in humans - particularly infants - to nitri t .
which, by reacting with hemoglobin, can markedly decrease the ability of blood
to carry oxygen to the tissues:
0 While there is no evidence that mammals metabolize nitrate into other
compounds (Parks et al.,1981), the bacteria found in human saliva
and the stomach can reduce nitrate to nitrite (U.S. EPA, 1985).
0 Due to decreased acidity (increased pH), particularly in the stomach
of the bottle-fed infant, bacteria capable of reducing nitrate to
nitrite may proliferate in the stomach thus leading to an increased
formation of nitrite in infants 3 months old or less (U.S. EPA, 1985) .
0 Nitrite reacts with the hemoglobin (the chemical responsible for the
ability of blood to transport oxygen to the tissues) in erythrocytes
to form methemoglobin which is unable to transport oxygen (Parks
et al., 1981).
0 The enzyme methemoglobin reductase converts methemoglobin to hemoglobin
and nitrate, thus, reversing the process induced by nitrite (Smith and
Beutler, 1966).
0 Bacteria in the saliva reduce 5% of absorbed nitrate into nitrite
(Spiegelhalder et al., ^.
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Nitrate/Nitrite March 31, 1987
-5-
In animals:
0 Gruener et al. (1973) observed that the activity of methemoglobin
reductase in rat fetuses was nearly ten times higher than that of
adult rats.
0 Nitrite in the stomach can react with secondary amines and other amine
substrates to form N-nitroso compounds that may be oncogenic (Sander
et al., 1968; Oshima and Bartsch, 1981). Vitamin C and vitamin E
can inhibit the formation of nitrosamines (Archer et al., 1975;
Kamm et al., 1977).
Excretion
0 Nitrate is readily excreted by the kidneys (U.S. EPA, 1985).
0 In humans, about 25% of the nitrate absorbed is secreted in saliva
(Spiegelhalder et al., 1976). •
0 While it has been suggested that appreciable amounts of nitrate are
eliminated in human (Donahoe, 1949) and cows milk (Davison et al.,
1964), there are inadequate data to support this conclusion.
0 The half-life for elimination of nitrite in dogs, sheep and Shetland
ponies (0.5-0.6 hrs) is too rapid to be accounted for by renal excre-
tion, thus suggesting that metabolism may be significant (Schneider
and Yeary, 1975).
IV. HEALTH EFFECT^
Humans
The lethal dose of potassium nitrate for an adult ranges from 54 to
462 mgAg; the lethal dose of sodium nitrite ranges from 32 to 154
(Burden, 1961).
The toxicity of nitrate in humans is due to the reduction of nitrate
to nitrite. By reacting with hemoglobin, nitrite forms methemoglobin
which will not transport oxygen to the tissues and thus can lead to
asphyxia (see Metabolism, above) (U.S. EPA, 1985).
The normal methemoglobin level in humans has been shown to range
between 1 and 2\ (Shuval and Greuner, 1977). A level greater than
3% is defined as methemoglobinemia. However, there is a consistent
elevation of the methemoglobin concentration in pregnant women from
the 14th week through delivery (Skrivan, 1971).
Walton (1951) published a survey by the American Public Health Associ-
ation which found that more than 278 cases of cyanosis in infants
were associated with nitrate-contaminated water. No cases of cyanosis
in infants were associated with water containing 10 mg/L or less of
nitrate-nitrogen. See also the discussion under Ten-day HA, below.
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Nitrate/Nitrite March 31, 1987
-6-
Winton et al. (1971) compared methemoglobin levels with nitrate ion
intake in 111 infants younger than six months old. Only three infants
had methemoglobin levels above 2.9%. They were the youngest of five
infants who had received more than 10 mgAg/day of nitrate ion.
Craun et al. (1981) conducted an epidemiologic study of 102 children
aged one to eight years in Washington County, Illinois. Of the study
subjects, 64 consumed water with high nitrate levels (22 to 111 mg/L
nitrate-nitrogen) and 38 consumed water with low nitrate levels (less
than 10 mg/L nitrate-nitrogen). Ingestion of water containing 22 to
111 mg/L nitrate-nitrogen did not produce abnormal mean methemoglobin
levels and was not related to increased methemoglobin levels in com-
parison to controls. See also the discussion under Ten-day HA, below*
Hegesh and Shiloah (1982) demonstrated that nitrites were synthesized
in infants with acute diarrhea. See also the discussion under Ten-day
HA, below.
In pregnant woman, the level of methemoglobin increases from the normal
methemoglobin level (between 0.5 and 2.5% of total hemoglobin) to a
maximum, 10.5%, at the 30th week of pregnancy and subsequently declines
to normal after delivery (Skrivan, 1971). Thus, pregnant women may
be more sensitive to the induction of clinical methemoglobinemia by
nitrite at approximately the 30th week of pregnancy.
Animals
Short-term Exposure
In the rabbit and rat, acute oral LDsg values for potassium nitrate
of 1,166 mgAg and 1,986 mgAg/ respectively, have been reported
(Windholz, 1976; WHO, 1962). The acute oral LD50 of sodium nitrate
in the rabbit has been reported to be 1,955 mgAg (Windholz, 1976).
In the rat, reported acute oral LD50 values for sodium nitrate range
from 46 to 120 mgAg (Druckery et al., 1963; Imaizumi et al., 1980;
Windholz, 1976; WHO, 1962).
Unlike humans, in which nitrite toxicity relates to the formation of
methemoglobin (see Metabolism, above), the immediate toxic effect of
nitrite in some species (e.g. the horse) is due to nitrite induced
vasodilation which results in cardiovascular collapse and shock
(U.S. EPA, 1985).
In a three week mouse drinking water study, elevated methemoglobin
levels were observed in 50-day-old mice administered nitrite ion (as
sodium nitrite) at levels of 133 and 178 mgAg/day but not at 88
(Shuval and Greuner, 1977).
Long-term Exposure
In a six month rat feeding study, 2,500 mg nitrate/kg/day produced a
marked diuretic effect within two months when compared with rats fed
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Nitrate/Nitrite March 31, 1987
-7-
equimolar levels of sodium chloride; 250 mg nitrate/kg/day caused no
diuresis and is identified as the NOAEL in this study (Fritsch
et al., 1980).
0 In a six month rat feeding study, both 250 and 2,500 mg nitrate/kg/day
as well as 25 and 250 mg nitrite/kg/day induced hemorrhagic areas in the
spleen (Fritsch et al. , 1980). Therefore, 250 mg nitrate/kg/day and
25 mg nitrite/kg/day are identified as LOAELs in this study.
0 Two long-term studies using ICR mice reported increases in amyloidosis
(starchy deposits) and hemosiderosis after ingestion of very high doses
of sodium nitrate (2,500 and 5,000 mg nitrate/kg/day: Sugiyama et al.,
1979) and sodium nitrite (208, 416 and 833 mg nitrite/kg/day: Inai, et
al., 1979). LOAELs of 2,500 mg nitrate/kg/day and 208 mg nitrite/kg/day
can be identified from the results of these studies.
Reproductive Effects
0 In a developmental toxicity study reported by Globus and Samuel (1976)
(described below) no evidence of sodium nitrite-induced adverse
reproductive effects was observed.
Developmental Effects
0 Groups of mice were intubated with sodium nitrite at 16.7 mg/kg/day
on days 0 through 14, 16 or 18 of gestation (Globus and Samuel, 1978).
Analysis of fetal livers indicated that maternally administered sodium
nitrite stimulated fetal hepatic erythropoiesis. No evidence of a
nitrite related effect upon fetal mortality, resorptions, mean weight,
number of offspring or incidence of skeletal malformation was observed.
0 The nitrosation of amides or amines in the stomach produces N-nitroso
compounds which may pass through the placenta to exert teratogenic or
fetotoxic effects (Ivankovic, 1979; Teramoto et al., 1980).
Mutagenicity
0 Both sodium nitrite and sodium nitrate were negative in host-mediated
assays in mice (FDA, 1972a and b). Other host mediated assays did
not find sodium nitrite to be mutagenic in mice (Couch and Friedman,
1975) or in either rats or mice (Whong et al., 1979).
0 Dominant lethal gene tests in rats were negative for both sodium
nitrate and nitrite (FDA, 1972a and b); a cytogenetic assay in rat
bone marrow cells was also negative for both compounds.
0 Kodama et al. (1976) reported that sodium nitrite induced mutations
to azaquanine resistance in cultured FM3A cells (a C3H mouse mammary
carcinoma cell line). Sodium nitrite was mutagenic in Salmonella
typhimurium (FDA, 1972a,b) and _E. coli Sd-4 (Hussain and Ehrenberg,
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Nitrate/ 31, 1987
-8-
Carcino^enicity
0 It was judged (U.S. EPA, 1985) that the available animal studies
(Newberne 1978, 1979; Maekawa et al., 1982) provided inconclusive
evidence regarding the carcinogenicity of nitrate and nitrite
administered orally in the absence of nitrosatable compounds .
0 Many studies have documented carcinogenesis (ouuit and prenatal) in
which both nitrite and nitrosatable compounds were orally administered
to animals (NAS, 1981); tumors were induced in many organs including
the stomach, esophagus and nasal cavity.
0 More than 120 N-nitroso compounds have been tested for carcinogenicity
and greater than 75% of these compounds have been shown to be carcino-
genic (Shank and Magee, 1981). These compounds have been demonstrated
to be carcinogenic in at least 22 species and carcinogenic transpla-
centally in at least five species (Schmahl and Habs, 1980). All
species tested have shown tumor formation following treatment with
at least one of the N-nitroso compounds tested. Tumors have been
induced in every organ and tissue and most cell types. While organ
specificity is observed within a species even after administration by
different routes, clear differences in target tissue have been noted
between species (Lijinsky et al., 1975).
V. QUANTIFICATION OF- TOXICOLOGICAL EFFECTS
Health Advisories (HAs) are generally determined for One-day, Ten-day,
Longer-term (approximately 7 years) and Lifetime exposures if adequate data
are available that identify a. sensitive noncarcinogenic end point of toxicity «
The HAs for noncarcinogenic toxicants are derived using the following formulas
HA = INOAEL or LOAEL) x (BW) = mg/L ( u /L)
(UF) x ( L/day)
where:
NOAEL or LOAEL - No- or Lowest-Observed-Adverse-Effect-Level
in mg/kg bw/day.
•
BW » assumed body weight of a child (10 kg) or
an adult (70 kg).
UF » uncertainty factor (10, 100 or 1,000), in
accordance with NAS/ODW guidelines.
L/day = assumed daily water consumption of a child
(1 L/day) or an adult (2 L/day).
The available data suggest that calculation of the HA values for nitrate/
nitrite should: • •
8 Recognize the newborn infant as the population group at greatest
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Nitrate/Nitrite March 31, 1987
-9-
0 Recognize and consider the conversion of orally ingested nitrate to
nitrite.
0 Utilize human data wherever possible due to the extreme species varia-
tion (both quantitative and qualitative) observed in nitrate and
nitrite toxicity.
HA values are presented below for a 4-kg infant (assumed to consume
0.64 L of formula per day) and a 70-kg adult. Normally, HAs are determined
for the 1 0-kg child and the 70-kg adult. However, newborn infants (assumed
to weigh 4 kg) are the population subgroup at greatest risk and thus HAs are
provided for the 4-kg infant.
While no separate HAs for the 10 kg child are provided, the HAs for the
70-kg adult will be protective for all age groups other than the 4-kg infant,
in that they are based upon data obtained in children (Craun et al., 1981).
Nitrate is toxic because it is converted to nitrite and thus the toxicity
of nitrate and nitrite must be additive. Thus, nitrate and nitrite cannot be
considered independently. Each HA is presented in terms of both mg nitrate-
nitrogen/L drinking water and mg nitrite-nitrogen/L drinking water. Appropriate
use of these values requires information on both the nitrate and nitrite
content of drinking water so that a total "effective" nitrate concentration
can be calculated and used as follows :
0 The "effective" nitrate-nitrogen concentration (mg/L) for all age
groups is equal to nitrate-nitrogen + 10x nitrite-nitrogen.
0 The "effective" nitrate-nitrogen concentration (mg/L) should not
exceed the appropriate nitrate standard for the appropriate group
(4-kg infant or 70-kg adult) or exposure period.
One-day Health Advisory
The available data are insufficient to develop One-day HAs for nitrate
and nitrite. The Ten-day HA should be protective of one-day exposures.
Ten-day Health Advisory
Populations other than the 4-kg infant:
Craun et al. (1981) conducted an epidemiologic study of 102 children
aged one to eight years in Washington County, Illinois. Of the study subjects,
64 consumed water with high nitrate levels (22 to 111 mg/L nitrate-nitrogen)
and 38 consumed water with low nitrate levels (less than 10 mg/L nitrate-
nitrogen). Ingestion of water containing 22 to 111 mg/L nitrate-nitrogen
did not result in abnormal mean methemoglobin levels and was not related to
increased methemoglobin levels in comparison to controls. In the entire
study group of 102 children, only five had methemoglobin levels greater than
2% (maximum of 3.1% in a child from the low exposure group).
For a 70-kg adult and all age groups other than the 4-kg infant, the
Ten-day nitrate HA value is 111 mg/L nitrate-nitrogen, the NOAEL observed by
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Nitrate/Nitrite March 31, 1987
-10-
Craun et al. (1981). Since the study was based on observations in humans
and since the most sensitive subgroup (i.e., infants) is considered separately,
no uncertainty factor has been employed in deriving the Ten-day nitrate HA
from the NOAEL.
There are no studies that provide a direct measure of the NOAEL for
nitrite in children. The Ten-day nitrite HA for a 70 kg adult and all other
age groups other than the 4 kg infant can be calculated from the NO?,EL for
nitrate, assuming 10% conversion of nitrate to nitrite, as follows:
(111 mg/L nitrate-nitrogen)(0.10) = 11 mg/L nitrite-nitrogen
where:
111 mg/L = NOAEL for nitrate based on the absence of methemoglobinemia
in children.
0.10 = assumed 10% conversion of nitrate to nitrite by 10-kg
child.
For a 4-kg infant:
Walton (1951) published a survey by the American Public Health Asso-
ciation which found more than 278 cases of cyanosis in infants that were
definitely associated with consumption of nitrate-contaminated water by the
infant or the nursing mother. No cases associated with water containing 10
mg/L or less of nitrate-nitrogen were found. As previously noted, Hegesh
and Shiloah (1982) demonstrated that nitrites were synthesized in infants
with acute diarrhea. Nitrites are responsible for methemoglobinemia and
thus it is possible that infants with diarrhea may be the population most
sensitive to the toxic effects of both nitrate and nitrite. As diarrhea is
relatively common in infants, it is believed that at least some of the infants
noted in Walton (1951) had diarrhea (U.S. EPA, 1985). Thus it was concluded
that Walton (1951) could serve as a basis for the protection of all infants
including those with diarrhea.
Based on the previous discussion, the Ten-day nitrate HA for 4-kg infants
is 10 mg/L nitrate-nitrogen, the NOAEL for methemoglobinemia observed by
Walton (1951). Studes by Donahoe (1949), Winton, et al. (1971) and Toussaint
and Wurkert (1982) support this HA.
No study provides a direct measure of the NOAEL for nitrite in infants.
However, the Ten-day nitrite HA for the 4-kg infant can be calculated from
the NOAEL for nitrate as follows:
(10 mg/L nitrate-nitrogen)(100%) = 1 mg/L nitrite-nitrogen
10
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Nitrate/Nitrite March 31, 1987
-11-
where:
10 mg/L = NOAEL for nitrate-nitrogen based on the absence of methemo-
globinemia in infants.
100% = assumed 100% conversion of nitrate to nitrite by 4-kg
infant.
10 = uncertainty factor, chosen in accordance with NAS/ODW
guidelines for use with data from a study in humans.
Longer-term Health Advisory
The available data are insufficient to develop Longer-term HAs for
nitrate and nitrite. However, for both nitrate and nitrite, it is judged that
the Ten-day HA for the 4-kg infant will offer protection against the formation
of methemoglobin induced by the ingestion of either nitrate or nitrite in all
age groups.
Lifetime Health Advisory
The Lifetime HA represents that portion of an individual's total exposure
that is attributed to drinking water and is considered protective of noncar-
cinogenic adverse health effects over a lifetime exposure. The Lifetime HA
is derived in a three step process. Step 1 determines the Reference Dose
(RfD), formerly called the Acceptable Daily Intake (ADI). The RfD is an esti-
mate of a daily exposure to the Human population that is likely to be without
appreciable risk of deleterious effects over a lifetime, and is derived from
the NOAEL (or LOAEL), identified from a chronic (or subchronic) study, divided
by an uncertainty" factor(s). From the RfD, a Drinking Water Equivalent Level
(DWEL) can be determined (Step 2). A DWEL is a medium-specific (i.e., drinking
water) lifetime exposure level, assuming 100% exposure from that medium, at
which adverse, noncarcinogenic health effects would not be expected to occur.
The DWEL is derived from the multiplication of the RfD by the assumed body
weight of an adult and divided by the assumed daily water consumption of an
adult. The Lifetime HA is determined in Step 3 by factoring in other sources
of exposure, the relative source contribution (RSC). The RSC from drinking
water is based on actual exposure data or, if data are not available, a
value of 20% is assumed for synthetic organic chemicals and a value of 10%
is assumed for inorganic chemicals. If the contaminant is classified as a
Group A or B carcinogen, according to the Agency's classification scheme of
carcinogenic potential (U.S. EPA, 1986), then caution should be exercised in
assessing the risks associated with lifetime exposure to this chemical.
No suitable studies for calculation of a Lifetime Health Advisory were
located. However, for both nitrate and nitrite, it is judged that the Ten-day
HA for the 4-kg infant (10 mg/L nitrate-nitrogen and 1 mg/L nitrite-nitrogen)
will offer protection againisTTthe^ formation of methemoglobin induced by the
ingestion of either nitrate or nitrite in all age groups.
As previously discussed, the 4-kg infant is the most sensitive member of
the population with respect to the formation of methemoglobin induced by
either nitrite directly or by the in vivo reduction of nitrate to nitrite.
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Nitrate/Nitrite March 31, 1987
-12-
In addition, as the 4-kg infant ages, e.g., to a 10-kg child, the sensitivity
to the effects of methemoglobin as well as the amount of nitrate reduced to
nitrite decrease, thus rendering the older child and the adult less sensitive
to the effects of both nitrate and nitrite. Thus, it has been concluded that
the Ten-day HA for the 4-kg infant for both nitrate and nitrite (10 mg/L
nitrate-nitrogen and 1 mg/L nitrite-nitrogen) will offer adequate protection
against methemoglobin formation in all other age groups as well.
Evaluation of Carcinogenic Potential
0 A number of animal studies provided inconclusive evidence regarding
the carcinogenicity of nitrate and nitrite administered in the absence
of nitrosatable compounds (U.S. EPA, 1985).
° Applying the criteria described in EPA's guidelines for assessment
of carcinogenic risk (U.S. EPA, 1986), both nitrate and nitrite may
be classified in Group D: Not classified. This category is for
agents with inadequate animal evidence of carcinogenicity.
VI. OTHER CRITERIA, GUIDANCE AND STANDARDS
0 The interim Maximum Contaminant Level for nitrate-nitrogen is 10 mg/L
(U.S. EPA, 1976b).
0 The U.S. Public Health Service recommended a' limit of 10 mg/L nitrate-
nitrogen or 45 mg/L nitrate ion (U.S. PHS, 1962).
0 The Committee on Water Quality Criteria of the National Academy of
Sciences recommended that nitrate-nitrogen concentration-in public
water supplies not exceed 10 mg/L and nitrite-nitrogen not exceed
1 mg/L (NAS, 1972).
0 The EPA Quality Criteria for Water (U.S. EPA, 1976a) suggested that
the maximum concentrations of nitrate-nitrogen and nitrite-nitrogen in
domestic water supplies not exceed 10 mg/L and 1 mg/L, respectively.
VII. ANALYTICAL METHODS
Determination of nitrite alone, or nitrite and nitrate combined, is
by colorimetry or spectrophotometry (U.S. EPA, 1979a;b). In these
methods, a sample is passed through a column containing granulated
copper-cadmium to reduce nitrate to nitrite. The nitrite (that
which was originally present plus reduced nitrate) is determined by
diazotizing with sulfanilamide and coupling with N-(1-naphthyl)-
ethylenediamine dihydrochloride to form a highly colored azo dye
which then is measured colorimetrically or spectrophotometrically.
Separate, rather than combined, nitrate-nitrite values are obtained
by carrying out the procedure first with, and then without, the copper-
cadmium reduction step. The applicable range of the colorimetric and
spectrophotometrie methods is 0.05 to 10 mg/L nitrate-nitrogen and
0.01 to 1 mg/L nitrite-nitrogen, respectively.
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Nitrate/Nitrite
March 31, 1987
-13-
An alternative reduction procedure may be used (U.S. EPA, 1979c).
this method, nitrate is reduced to nitrite with hydrazine sulfate.
The applicable range of this method is 0.01 to 10 mg/L.
In
VIII. TREATMENT TECHNOLOGIES
0 Ion exchange and reverse osmosis are the practical methods currently
in use to remove nitrates from water. Conventional coagulation and
lime softening are not effective treatment methods for the removal of
this contaminant (U.S. EPA, 1977; Laverentz, 1974).
0 The application of ion exchange resins for nitrate removal has a well
established history and is recognized as a practical treatment for
drinking water systems (U.S. EPA, 1977; Gillies, 1978; Sorg, 1978;
Sorg, 1980).
0 Laboratory experiments and pilot plant studies have shown that some
strong base and weak base ion exchange resins are nitrate selective
and can reduce the nitrate concentration from as high as 50 mg/L
(as N) to 0.5 mg/L (Holzmacher, 1971; Gregg, 1973; Korngold, 1973;
Gaundett, 1975; Kuelow et al., 1975). One full-scale ion exchange
plant has been operating successfully on Long Island, New York, since
1974. This plant lowers the nitrate level of 20-30 mg/L in the raw
water to 0.5 mg/L. The finished water is a blend of treated and raw
. water and contains about 5 mg/L of nitrate (as N). Other installations
removing nitrate include a 40,000 gpd plant at Curryville, Pennsylvania
and the 2,500 gpd plant in the Virgin Islands.
0 An important feature of the commercial nitrate -ion exchange resin
is that nitrate is not the most preferred ion in jthe exchange but
rather the sulfate ion. However, field studTes™by Guter (1982) in
McFarland, California have shown that nitrates can be removed effec-
tively in the presence of sulfates as high as 380 mg/L.
0 Although reverse osmosis (RO) systems have not been installed to remove
specifically nitrates, removal efficiencies of 67 to 95% (high pressure)
have been demonstrated. There are two plants currently operating
which can provide data on nitrate removal. Laverentz (1974) reported
that in Greenfield, Iowa, nitrate is reduced from 0.2 mg/L NO3-N to
0.014 ng/L NOj-N. In San Diego Country Estates, Romona, California,
the nitrate is reduced from 12.4 mg/L NO3-N to 4.2 mg/L N03-N.
However, there are considerable experimental field data obtained
when cellulose acetate was the only commercial membrane as well as
more recent field tests that indicate nitrate rejection ranges for
cellulose acetate membranes from 70 to 80% (Sourirajan, 1977), 80 to
90% (Metcalf and Eddy, Inc., 1979), and 58 to 86% (Weber, 1972).
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Nitrate/Nitrite March 31, 1987
-14-
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of nitrite with ascorbate and its relation to nitrosamine formation.
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Bloomfield, R.A., J.R. Hersey, C.W. Welsch, G.B. Garner and M.E. Muhrer.
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Burden, E.H.W.J. 1961. The toxicology of nitrates and nitrites with particu-
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Couch, D.B., and M.A. Friedman. 1975. Interactive mutagenicity of sodium
nitrite, dimethylamine, methylurea and ethylurea. Mutat. Res.
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Craun, G.F., D.G. Greathouse and D.H. Gunderson. 1981. Methemoglobin levels
in young children consuming high nitrate well water in the United States..
Int. J. Epidemiol. 10:309-317.
Davison, K.L., W. Hansel, L. Crook, K. McEntee and M.J. Wright. 1964.
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Druckery, H., D. Steinhoff, H. Beuthner, H. Schneider and P. Klarner. 1963.
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FDA. -1972a. Food and Drug Administration. Stanford Research Institute.
Study of mutagenic effects of sodium nitrate (71-7). Menlo Park, CA.
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Contract FDA 71-267. Rept. No. FDABF-GRAS-084. 103 pp.
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Gaundett, R.B. 1975. Nitrate Removal from Water by Ion Exchange. Water
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Gillies, M.T. 1978. Drinking Water Detoxification. Noyes Data Corporation.
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Gregg, J.C. 1973. Nitrate Removal at Water Treatment Plant. Civ. Eng.
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Gruener, N., H.I. Shuval, K. Behroozi, S. Cohen and H. Shechter. 1973.
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Guter, G.A. 1982. Removal of nitrate from contaminated water supplies for
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Hegesh, E., and J. Shiloah. 1982. Blood nitrates and infantile methemo-
globinemia. Clinica Chimica Acta. 125:107-115.
Holzmacher, R.G. 1971. Nitrate removal from a ground water supply. Water
Sewage Works. 118(7):210-213.
Hussain, S., and L. Ehrenberg. 1974. Mutagenicity of primary amines combined
with nitrite. Mutation Res. 26:419-422.
Imaizumi, S., I. Tyuma, K. Imai, H. Kosaka and Y. Ueda. 1980. In vivo
studies on methemoglobin formation by sodium nitrite. Int. Arch. Occup.
Environ. Health. 45:97-104.
Inai, K., Y. Aoki and S. Tokuoka. 1979. Chronic toxicity of sodium nitrite
in mice, with reference to its tumorigenicity. Gann. 70:203-208.
Inui, N., Y. Nishi, M.M. Hasegawa, M. Taketumi, M. Yamamoto and A. Tanimura.
1980. Induction of 8-azaguanine-resistant mutation and neoplastic trans-
formation of hamster embryonic cells by coadministration of sodium
nitrite and aminopyrine. J. Cancer Res. Clin. Oncol. 97:119-128.
Ivankovic, S. 1979. Teratogenic and carcinogenic effects of some chemicals
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Natl. Cancer Inst. Monogr. 51:103-115.
Kamm, J.J., T. Dashman, H. Newmark and W.J. Mergens. 1977. Inhibition of
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Keulow, R.W., K.L. Kropp, J. Withered and J.M. Symons. 1975. Nitrate removal
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Kodama, F., M. Umeda and T. Tsutsui. 1976. Mutagenic effect of sodium
nitrite on cultured mouse cells. Mutat. Res. 40:119-124.
Korngold, E. 1973, Removal of nitrates from potable water by ion exchange.
Water, Air, Soil Pollut. 2:15-22.
Laverentz, D.L. 1974. Economic feasibility of desalting systems for municipal
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Lijinsky, W., G.M. Singer and H.W. Taylor. 1975. Carcinogenic N-nitroso
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Maekawa, A., T. Ogiu, H. Onodera et al. 1982. Carcinogenicity studies of
sodium nitrite and sodium nitrate in F344 rats. Food Cosmet. Tox.
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Metcalf and Eddy, Inc. 1979. Wastewater Engineering: Treatment, disposal
reuse, 2nd ed. McGraw-Hill Co.
NAS. 1972. National Academy of Sciences. Water quality criteria. National
Academy Press. Washington, DC. EPA R3-73-033, 1973.
NAS. 1978. National Academy of Sciences. Nitrates: an environmental assess-
ment. National Academy Press. Washington, DC.
NAS. 1981. National Academy of Sciences. The health effects of nitrate,
nitroso compounds. National Academy Press. Washington, DC,
P.M. 1978. Dietary nitrite in the rat. Final Report or. Contract
FDA-74-2181, Food and Drug Administration, Public Health Service, U.S.
Department of Health, Education and Welfare, Rockville, MD.
Newberne, P.M. 1979. Nitrite promotes lymphoma incidence in rats. Science.
204: 1079-1081 .
Ohshima, H., and H. Bartsch. 1981. Quantitative estimation of endogenous
nitrosation in humans by monitoring N-nitrosoproline excreted in the
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Parks, N.J., K.A. Krohn, C.A. Mathis, J.H. Chasko, K.R. Geiger, M.E. Gregor
and N.F. Peek. 1981. Nitrogen-13-labeled nitrite and nitrate: Distri-
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Sander, J., and F. Schweinsberg. 1972. Interrelationships between nitrate,
nitrite and carcinogenic N-nitroso-compounds. 1. Communication:
nitrate, nitrite and nitrosable amino-compounds in food and drugs,
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Infektionsk. Hyg. Abt. 1: Orig. Reihe B 156:299-340. (In German;
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Schneider, N.R., and R.A. Yeary . 1975. Nitrite and nitrate pharmacokinetics
in the dog,-sheep, and pony. Am. J. Vet. Res. 36:941-947.
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