U.S. Environmental Protection Agency
       Office of Drinking Water
     Criteria &  Standards Division
         Health  Effects Branch

            December,  1990

    PB 89-192280/AS
(replace  this for  old document)
(Please read Instructions on the reverie before complr
Quantification of Toxicological Effects of
Office of Drinking Water
Office of Drinking Water
Criteria and Standards Division
Washington, DC 20460
This document

: PB9 1-143479
Bee. lyyu
is to supersede old document PB89-192280/AS
This document quantifies the drinking water health effects of tetrachloroethylent
through the review of several studies. These studies include animal and humans.
Physical and chemical properties are discussed. Carcinogen! ~i tj of the compcu:
is reviewed and evaluated through various studies.
                                  KEY WORDS AND DOCUMENT ANALYSIS
   health effects,•cancerous, animal studies
   human studies,  cancer  risk, criteria
   document, drinking water,  tetrachloro-
EPA Form 2220-1 
                                 . DISCLAIMER

    This document  has  been  reviewed 1n  accordance with  U.S.  Environmental
Protection  Agency  policy  and approved  for publication.   Mention  of  trade
names or commercial  products does not constitute  endorsement  or recommenda-
tion for use.


      Section 1412(b)(3)(A)  of  the Safe Drinking Water Act, as
 amended in 1986, requires  the  Administrator of the Environmental
 Protection Agency  to  publish Maximum Contaminant Level Goals and
 promulgate National Primary Drinking Water Regulations for each
 contaminant,  which, in the judgment of the Administrator, may
 have  an adverse  effect on  public health and which is known or
^anticipated to occur  in public water systems.  The Maximum
 Contaminant Level  Goal is  nonenforceable and is set at a level at
 which no known or  anticipated  adverse health effects in humans
 occur and which  allows for an  adequate margin of safety.  Factors
 considered in setting the  Maximum Contaminant Level Goal include
 health effects data and sources of exposure in addition to
 drinking water.

      This document provides the health effects basis to support
 establishing values for tetrachloroethylene.  To set these
 values,  data on  pharmacokinetics, human exposure, acute and
 chronic toxicity to animals and humans, epidemiology and
 mechanisms of toxicity were evaluated.  Specific emphasis is
 placed on literature  data  providing dose-response information.
 Thus,  while the  literature search and evaluation performed in
 support of this  document were  comprehensive, only the reports
 considered most  pertinent  in the derivation of the Maximum
 Contaminant Level  Goal are cited in the document.  The
 comprehensive literature search in support of this document
 includes  information  published in the Health Assessment Document,
 its appendix, and  the document "Response to the Issues and Data
 Submissions on Tetrachloroethylene (Perchloroethylene)."

      When adequate health  effects data exist, Health Advisory
 values  for less  than  lifetime  exposures (One-day, Ten-day and
 Longer-term,  approximately 10% of an individual's lifetime) are
 included  in this document.  These values are not used in setting
 the Maximum Contaminant  Level, but serve as informal guidance to
 municipalities and other organizations when emergency spills or
 contamination situations occur.

                                   Michael B. Cook
                                   Office of Drinking Water


      The source documents for background  information used to

 develop this report on the quantification of toxicological

 effects for  tetrachloroethylene  are the U.S. EPA  (1985) health

 assessment document for tetrachloroethylene, its  appendix (1986),

 and  a recent draft  document,  "Response to the Issues and Data

 Submissions  on  Tetrachloroethylene  (Perchloroethylene)"

 (U.S.  EPA, 1990).

      The quantification of toxicological  effects  of a  chemical

 consists of  separate assessments of noncarcinogenic and

 carcinogenic health effects.   Chemicals that do not produce

 carcinogenic effects are believed to have a threshold  dose below

 which  no adverse, noncarcinogenic effects occur,  while

 carcinogens  are assumed to act without a  threshold.


     In  the  quantification of  noncarcinogenic effects, a

Reference Dose  (RfD,  formerly  termed the  Acceptable Daily

Intake),  is  calculated.   The RfD is an estimate of a daily

exposure to  the human population that is  likely to be  without

appreciable  risk of  deleterious  health effects during  a lifetime.

The RfD  is derived  from a  No-Observed-Adverse-Effect Level

 (NOAEL), or  Lowest-Observed-Adverse-Effect Level  (LOAEL),

identified from a subchronic or chronic study and divided by an

uncertainty factor(s).  The RfD is calculated as follows:

     „.*«      fNOAEL or LOAEL)              ,,   ,  ,     .,_,_,,
     RfD = Uncertainty factor(s) = 	 m<3/k<3 body ^ight/day

     Selection of the uncertainty factor to be employed in the

calculation of the RfD is based on professional judgment while

considering the entire database of toxicological effects for the

chemical.  In order to ensure that uncertainty factors are

selected and applied in a consistent manner,  the Office of

Drinking Water (ODW) employs a modification to the guidelines

proposed by the National Academy of Sciences (MAS, 1977, 1980) as


  0  An uncertainty factor of 10 is generally used when good

     chronic or subchronic human exposure data identifying a

     NOAEL are available and are supported by good chronic

     toxicity data in other species.

  0  An uncertainty factor of 100 is generally used when good

     chronic toxicity data identifying a NOAEL are available for

     one or more animal species (and human data are not

     available),  or when good chronic or subchronic toxicity data

     identifying a LOAEL in humans are available.

  0  An uncertainty factor of 1,000 is generally used when

     limited or incomplete chronic or subchronic toxicity data

      are  available,  or when  good  chronic or subchronic toxicity

      data identify a LOAEL,  but not a NOAEL, for one or more

      animal  species are available,

      The  uncertainty factor  used  for a specific risk assessment

 is  based  principally upon  scientific judgment rather than

 scientific fact  and accounts for  possible intra- and interspecies

 differences.   Additional considerations not incorporated in the

 NAS/ODW guidelines for selection  of an uncertainty factor  include

 the use of a  less-than-lifetime study for deriving an RfD, the

 significance  of  the adverse  health effect, pharmacokinetic

 factors,  and  the counterbalancing of beneficial effects.

      From the RfD,  a Drinking Water Equivalent Level (DWEL) can

 be  calculated.   The DWEL represents a medium-specific  (i.e.,

 drinking  water)  lifetime exposure at which adverse,

 noncarcinogenic  health effects are not anticipated to occur.  The

 DWEL  assumes  100%  exposure from drinking water.  The DWEL

 provides  the  noncarcinogenic health effects basis for

 establishing  a drinking water standard.  For ingestion data,  the

 DWEL  is derived  as follows:
                  x  fBodv weight in kcr)   _         ,   ,
           Drinking water volume in L/day ~

where :

  Body weight = assumed to be 70 kg for an adult.

  Drinking water volume = assumed to be 2 L per day for an  adult


      In addition to the RfD and the  DWEL, Health Advisories  (HAs)

 for exposures of shorter duration  (One-day, Ten-day, and

 Longer-term)  are determined.   The  HA values are used as informal

 guidance to municipalities  and other organizations when emergency

 spills or contamination situations occur.  The HAs are calculated

 using a similar equation to the RfD  and DWEL; however, the NOAELs

 or  LOAELs are identified from  acute  or subchronic studies.   The

 HAs are derived as follows:
             fNQAEL or  LOAEL)  x  (Body weight)    =         .
           {Uncertainty factor(s)) x  (	L/day)   -:	  9
     Using the  above  equation, the following drinking water HAs

are developed for  noncarcinogenic effects:

     1.  One-day HA for  a  10-kg child ingesting 1 L water per day.

     2.  Ten-day HA for  a  10-kg child ingesting l L water per day.

     3.  Longer-term  HA  for a  10-kg child ingesting 1 L water per  day,

     4.  Longer-term  HA  for a  70-kg adult ingesting 2 L water per  day.

     The One-day HA calculated for a 10-kg child assumes a single

acute exposure  to  the chemical and is generally derived from a

study of less than 7  days  duration.  The Ten-day HA assumes a

limited exposure period  of 1 to 2 weeks and is generally derived

from a study of less  than  30 days of duration.  The Longer-term

HA is derived for  both the 10-kg child and a 70-kg adult and

assumes an exposure period of  approximately 7 years (or 10% of an


individual's lifetime).  The Longer-term HA is generally derived

from a study of subchronic duration (exposure for 10% of an

animal's lifetime).

     Quantification of Carcinogenic Effects

     The EPA categorizes the carcinogenic potential of a

chemical, based on the overall weight of evidence, according to

the following scheme:

  0  Group A:  Human Carcinogen.  Sufficient evidence exists from

               epidemiology studies to support causal association

               between exposure to the chemical and human cancer.
  8  Group B:  Probable Human Carcinogen.  Sufficient evidence of

               carcinogenicity in animals with limited (Group Bl)

               or inadequate (Group B2)  evidence in humans.

  0  Group C:  PossibleHuman Carcinogen.  Limited evidence of

               carcinogenicity in animals in the absence of human


  0  Group D:  Not Classified as to Human CarcinQgenicJ.ty.

               Inadequate human and animal evidence of

               carcinogenicity or for which no data are


  0  Group E:  Evidence of Noncarcinoaenicitv for Humans.  No

               evidence of carcinogenicity in at least two

               adequate animals tests in different species or in

               both adequate epidemiologic and animal studies.

                               ^ e _iri

      If toxicological evidence  leads to the classification of the

 contaminant as a known,  probable,  or possible human carcinogen,

 mathematical models are  used to calculate the estimated excess

 cancer risk associated with the ingestion of the contaminant in

 drinking water.   The data used  in  these estimates usually come

 from lifetime exposure studies  in  animals.  In order to predict

 the  risk for humans for  animal  data, animal doses must be

 converted to equivalent  human doses.  This conversion includes

 correction for noncontinuous exposure, less-than-lifetime

 studies,  and for differences in size.  The factor that

 compensates for the size differences is the cube root of the

 ratio of  the animal and  human body weights.  It is assumed that

 the  average adult human  body weight is 70 kg and that the average
 water consumption of an  adult human is 2 liters of water per day.

      For  contaminants with a carcinogenic potential, chemical

 levels  are  correlated with a carcinogenic risk estimate by

 employing a cancer  potency (unit risk) value together with the

 assumption  for lifetime  exposure via ingestion of water.  The

 cancer  unit risk is usually derived from a linearized multistage

 model with  a 95% upper confidence  limit and provides a low-dose

 estimate; that is,  the true risk to humans, while not

 identifiable,  is not likely to  exceed the upper limit estimate

 and,  in fact,  may be lower.   Excess cancer risk estimates may

 also be calculated  using other  models such as the one-hit,

Weibull,  logit,  and probit.   There is little basis in the 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 any others.  Because each model is based upon

 differing assumptions, the estimates that are derived for each

 model  can differ  by several orders of magnitude.

     The scientific database used to calculate and support the

 setting  of  cancer risk rate levels has an inherent uncertainty

 due to the  systematic and random errors in scientific

 measurement.  In  most cases, only studies using experimental

 animals  have  been performed.  Thus, there is uncertainty when the

 data are extrapolated to humans.  When developing cancer risk

 rate levels,  several  other areas of uncertainty exists, such as

 the incomplete knowledge concerning the health effects of
 contaminants  in drinking water; the impact of the experimental

 animal's age, sex, and species; the nature of the target organ

 system(s) examined; and  the actual rate of exposure of the

 internal targets  in experimental animals or humans.

 Dose-response data usually are available only for high levels of

 exposure, not for the lower levels of exposure closer to where a

 standard may be set.  When there is exposure to more than one

 contaminant, additional  uncertainty results from a lack of

 information about possible synergistic or antagonistic effects.


     Acute or chronic exposure to tetrachloroethylene can cause

liver,  kidney and  CNS toxicity in a variety of species including


 man.   Tetrachloroethylene vapors  are  irritating to mucous
 membranes,  eyes  and skin.   The  most serious effects of tetra-
 ehloroethylene exposure (severe CNS depression/death) occur when
 high  tetrachloroethylene concentrations are inhales or large
 doses are  administered via gavage.  Such effects are unlikely to
 occur due  to  tetrachloroethylene  exposure via drinking water.

      Although tetrachloroethylene toxicity has been studied in
 organ systems of many  species,  only limited data are available on
 the most sensitive  end point  of toxicity for the chronic
 ingestion  of  tetrachloroethylene.  Assessment of tetra-
 chloroethylene toxicity must  be gleaned from inhalation studies
 and studies of acute/subchronic ingestion.  Man may be the most
 sensitive  species with respect  to the CNS effects of acute
 tetrachloroethylene inhalation  (Stewart et al., 1970).  Liver and
 kidney toxicity  from tetrachloroethylene exposure, often observed
 in experimental  animals has not been studied in detail for
 humans.  Several studies indicate that mice are more sensitive to
 tetrachloroethylene liver  and kidney toxicity than are rats
 (Schumann et  al., 1980;  NTP,  1985; NCI, 1977).  Guinea pigs
 suffer hepatotoxic  effects at concentrations for which no changes
were observed in rats,  rabbits  and monkeys (Rowe et al., 1952).
No direct comparison has been made between the sensitivity of
guinea pigs and mice to  tetrachloroethylene, but similar toxic
effects increase in  liver  weight)  were observed in chronic
studies of mice exposed  to 200  ppm (1,360 mg/m3)  for 7 hours/day
for 236 day for 8 months (Kylin et al., 1965; approximately
                                MM 0 mm

 equivalent to  160  mg/kg/day,  see Appendix) and chronic studies of

 guinea  pigs exposed to 100 ppm  (678 mg/m3) for 7 hours/day for

 236  days  (Rowe,  1952;  approximately equivalent to 63 mg/kg/day,

 see  Appendix).
      Observations  in  Humans

      Inhalation  exposure  to  tetrachloroethylene has been studied

 in man under  controlled laboratory conditions and as a result of

 occupational  exposure.  In a study by Stewart et al. (1970)» five

 male  subjects were exposed to 100 ppm (678 mg/m3)  tetrachloro-

 ethylene  for  7 hours/day  on  5 consecutive days (approximately

 equivalent to 20 mg/kg/day;  see Appendix).  Subjects were
 monitored with respect to blood chemistry, ability to perceive

 tetrachloroethylene odor, pulmonary function, performance  levels

 on behavioral/neurological tests, and asked to report on a

 variety of subjective complaints.  Odor perception decreased over

 time  during the course of the week.  Perception at the beginning

 of each day decreased faster as the week progressed.  After 3

 hours of  exposure  on  the  first day, 3 or 8 subjects were unable

 to respond normally to a  modified Romberg test, but were able to

 overcome  this inability with greater mental effort.  Subjective

 complaints during  the five days included mild eye, nose and

throat irritation,  lightheadedness, mild frontal headache,

sleepiness, and some  difficulty in speaking.  These complaints

decreased over the  course of the study week.  Normal readings

were obtained for  all other  tests.  In follow-up studies,  the


 authors concluded that prolonged exposure to  100 ppm  (678 mg/m3)

 had no consistent adverse effects on  performance in these

 behavioral tests (Stewart et al.,  1974,  1977).

      Other studies of human experimental exposure  indicate  the

 ability of tetrachloroethylene  to cause  eye irritation  and  CNS

 effects such as  dizziness at concentrations of 100 to 600 ppm

 (695  to 4,100 mg/m3;  Rowe et al.,  1952).  Accidental industrial

 exposure to higher concentrations (exact concentrations unknown)

 produce more serious  CNS  effects and  hepatotoxicity  (Stewart

 et  al.,  1961; Hake and Stewart,  1977).

      Observations in  Other Species

      Neurotoxicity:   Severe ataxia and anesthesia  in mice and

 rats  have  been observed at lethal  concentrations of tetrachloro-

 ethylene  (NTP, 1985).   Less severe effects have been studies in

 experimental  animals  with the use  of  behavioral tests,  but

 available  studies indicate that  effects  on the liver or kidney

 occur at lower exposure levels.  Goldberg (1964) exposed rats to

 1,500 ppm  (10,200 mg/m3)  and  2,300 ppm (15,600 mg/m3)  tetra-

 chloroethylene for 2  weeks,  4 hours/day, 5 days/week.   At 2,300

ppm (15,600 mg/m3), ataxia and diminished escape avoidance

response was  observed.  No effects were  seen  at exposure levels

of 1.500 ppm  (10,200  mg/m3).  Savolainen et al., (1977)  exposed

rats to 200 ppm  (1,360  mg/m3), 6 hours/day for 4 days.  Little if

any impairment over controls  was observed.


      Hepatotoxicity;   The  hepatotoxicity of tetrachloroethylene

 in  experimental  animals  has  been  studied in greater detail than

 its behavioral/CNS  effects.  Acute effects in mice have been

 observed  at  concentrations as  low as 200 ppm (1,360 mg/m3; Kylin

 et  al,, 1963} or oral  doses  as low as 100 mg/kg  (Schumann et al.,

 1980).  Hepatotoxicity from  exposure to concentrations as low as

 100 ppm(678  mg/m3; NTP,  1985)  has been observed after chronic

 exposure; data on chronic  ingestion of tetrachloroethylene are


      Kylin et al. (1963) observed reversible hepatotoxic effects

 (fatty degeneration) in  mice exposed to 200 ppm  (1,360 mg/m3) for

 4 hours (approximately equivalent to 160 mg/kg/day, see

Appendix).   Other acute  studies demonstrate hepatotoxic effects

at  higher concentrations (Rowe et al., 1952).

      The subchronic and  chronic effects of tetrachloroethylene

inhalation have  been described in several studies, including NTP

 (1985), Carpenter (1937),  Rowe et al. (1952), and Mazza (1972).

     Hepatotoxic  effects were  observed in the NTP  (1985) 13-week

range finding study.   Rats and mice were exposed to

concentrations of 100, 200,  400, 800 and 1,600 ppra (678, 1,360,

2,710, 5,420 and  10,800  mg/m3)  for 6 hours/day,  5 days/week for

13 weeks (approximately  equivalent to 160 to 2,600 mg/kg/day

(mice) and 66 to  1,600 mg/kg/day  (rats); see Appendix).  Liver

lesions (infiltration, necrosis and bile stasis) were observed in


 mice exposed to concentrations  of  400 ppm  (2,710 mg/m3) dose

 groups.   Effects on the kidneys were also  observed within this

 dose range and are described below.

      Carpenter (1937)  exposed rats of both sexes to 70 ppm  (475

 mg/m3), 230  ppm 1,560  mg/m3)  or 470 ppm (3,190 mg/m3),  8

 hours/day 5  days/week  for  150 days; approximately equal to  62,

 200  and 410  mg/kg/day,  see Appendix).  Animals in the highest

 dose group exhibited hepatic and renal congestion and swelling.

 At the middle dose,  congestion  was only observed in the kidney.

 No significant changes were seen at the lowest dose.

      Chronic effects on the liver  and kidney were also observed
 in the NTP (1985)  inhalation bioassay.  Rats were exposed to

 concentrations of  200  and  400 ppm  (approximately equal to 130 and

 260  mg/kg/day;  see Appendix)  for 6 hours/day, 5 days/week for

 103  weeks.   Effects  on the kidney  were observed in all treated

 groups and are described below.  Hepatotoxicity was observed in

 all  treated  male mice  and  female mice in the high dose group.

The  effects  observed include increased incidences of degeneration

 (Vacuolation,  infiltration,  pigmentation,  and hyperplasia)

necrosis and nuclear inclusions.

     In contrast to  these  findings, Rowe et al. (1952) observed

no toxic effects in  rats,  rabbits  or monkeys exposed to 400 ppm

 (2,710 mg/m3) tetrachloroethylene  for 7 hours/day,  5 days/week

for 179 days.  Guinea pigs  exposed to the  same regiment at  100,


 200,  300 and 400 ppm (678,  1,360,  2,030 and 2,710 mg/m3) showed a

 dose  dependent increase in  liver weight and fatty infiltration of

 the liver when exposed over 236  days  (Rowe et al.,  1952).

 Similar hepatotoxic effects were observed in mice exposed  to

 200 ppm (1,360 mg/m3),  4 hours/day, 5 days/week for 8 months

 (Kylin  et al.,  1965;  approximately equal to 160 mg/kg/day;  see

 Appendix).   Effects have also  been observed in rabbits, but at

 higher  concentrations.  Mazza  (1972) exposed rabbits to 2,790 ppm

 (18,900 mg/m3),  approximately  equal to 840 rag/kg (see Appendix),

 for 4 hours/day,  5  days/week for 45 days and observed changes in

 serum levels  of glutamic-oxaloacetic transaminase (SCOT),

 glutamic-pyruvic transaminase  (SGPT) and glutamide  dehydrogenase


     Hepatotoxic effects have  also been observed as a result of

 oral exposure.   Studies of  acute oral exposure to tetrachloro-

 ethylene indicate that  doses of  4,000 mg/kg or greater are lethal

 to  experimental animals (Wenzel  and Gibson, 1951,* Smyth et al./

 1969).   A variety of  hepatotoxic effects have been  demonstrated

 at  lower doses.   Fujii  (1975)  found elevated serum  enzyme  levels

 in  rabbits exposed  to 2,186  mg/kg.  Vaino et al. (1976) studied

microsomal enzymes  in vitro  after  in vivo exposure  of rats to

tetrachloroethylene in  olive oil via gavage (2.6 uunol/kg  [429

mg/kg]).  Recovery  of some microsomal enzyme activities

 (benzpyrene hydroxylase and  p-nitrcanisole O-demethylase)  per

gram liver (wet weight) were significantly lower than controls.

      Mice  appear  to  be  mores  sensitive to the effects of tetra-

 chloroethylene  exposure than  rats.  Schumann et al.  (1980)

 administered  tetrachloroethylene  in corn oil to both rats and

 mice  via gavage for  11  consecutive days at doses of 100, 250, 500

 and 1,000  mg/kg.   Histopathological changes including

 centrilobular hepatocellular  swelling and increased liver weight

 were  observed in  all treated  mice; rats were more resistant, with

 toxicity being  apparent only  at the highest does.

      Similar  hepatotoxic effects were observed in mice after

 subchronic exposure.  In a study by Buben and O1Flaherty (1985),

 male  Swiss-Cox  mice  were exposed to tetrachloroethylene in corn

 oil via gavage  at doses of 1, 20, 100, 200, 500, 1,000, 1.500 and
 2,000 rag PCE/kg 5 days/week for 6 weeks.  Liver toxicity was

 evaluated  by  several parameters including liver weight/body

 weight ratio, hepatic triglyceride concentration, serum GEP and

 SGPT  activity,  hepatic  DNA content, histopathological evaluation

 and hepatic dry weight/wet weight ratios.  All parameters

 indicated  liver toxicity at high doses.  Increased liver

 triglycerides were first observed in mice treated with 100 mg/kg.

 Liver weight/body weight ratios were significantly different from

 controls for the  100  mg/kg group, and slightly higher than

controls in the 20 mg/kg group.

     Lifetime oral exposure to tetrachloroethylene was shown to

cause liver and kidney  toxicity in two separate studies (NCI,

1977;  NTP,  1983).   In the NCI study, Osborne-Mendel rats and


 B6C3F1  mice  were  exposed to  PCS  in corn oil via gavage for 5

 days/week  for  78  weeks at does of 471 to 949 mg/kg  (rats) and  386

 to  1,072 mg/kg (mice).   In addition to hepatocarcinogenic

 effects, toxic nephropathy was observed in all treatment groups

 for both species.   In the NTP study, female B6C3F1 mice were

 exposed to tetrachloroethylene in corn oil (25, 50, 100 or 200

 mg/kg)  5 days/week  for 103 weeks.  This report had not been

 audited as of  June,  1985.

     Renal Toxicitv:   Renal  toxicity from tetrachloroethylene

 exposure via inhalation has  been demonstrated in rabbits, rats

 and mice.  Brancaccio et al.  (1971) exposed rabbits to 2,280 ppm

 (15,500 mg/m3, approximately equivalent to 680 mg/kg)  for 4

 hours/day, 5 days/week for 45 days.  Decreases in glomerular

 filtration,  renal plasma flow and maximal tubular excretion were

 observed.  In  the NTP (1985) 13-week range finding study, rats

 and  mice were  exposed to concentrations of 0, 100, 200, 400, 800

 and  1,600  ppm  (0, 678,  1,360, 2,710, 5,420 and 10,800 mg/m3).

 Renal toxicity was not  observed in rats, but mice exposed to

 concentrations of 200 ppm (1,360 mg/m3; equivalent to about 320

mg/kg/day; see Appendix)  or greater exhibited karyomegaly of the

tubular epithelium.

     Carpenter  (1937)  exposed rats of both sexes to

concentrations of 70  ppra (475 mg/m3),  230 ppm (1,560 mg/m3), and

470 ppm (3,190 mg/m3) for 8 hours/day,  5 days/week for 150 days;

approximately  equal to  62, 200 and 410 mg/kg/day.  At the two


 highest doses,  the  kidney  showed  increased secretion, cloudiness,

 swelling and  desquamation?  the  spleen was congested and showed an

 increase in pigment content.  Renal toxicity was also observed in

 the NTP tetrachloroethylene bioassay  (1985) in which male and

 female  rats and mice were  exposed to tetrachloroethylene for 6

 hours/day, 5  days/week  for 103  weeks.  An increased incidence of

 tubular cell  karyomegaly was observed for all treatment groups

 (200 and 400  ppm for rats,  approximately equivalent to 130 and

 260 mg/kg/day;  100  and  2OO  ppm  for mice, approximately equivalent

 to the  120 and  240  mg/kg/day; see Appendix).

     Other Effects;  Reproductive and developmental effects have

 been shown to result from  exposure of rats and mice to
 tetrachloroethylene.  Pregnant  rats and mice exposed to 300 ppm

 (2,000  mg/m3)  for 7  hours/day on  days 6 through 15 of gestations

 (approximately  equivalent  to doses of 230 mg/kg [rats] and 560

mg/kg [mice]; see Appendix).  Rats had twice the number of

resorptions per implantation compared with controls, while mouse

pups exhibited  significant  subcutaneous edema, delayed skull

ossification  and split  sternebrae (Schwetz et al., 1975).

     Study Selection for Quantification of Noncarcinogenic


     The entire data base on tetrachloroethylene must be

evaluated before appropriate studies can be selected as the basis

for One-day,  Ten-day, Longer-term or Lifetime HA values.  The


 CMS,  hepatic and renal toxicity  of  tetrachloroethylene are of

 primary concern.   Although some  data  are available on human

 exposures to tetrachloroethylene, these data were not used as  the

 basis for HA values.   From the available data, it is not possible

 to judge the most sensitive toxic endpoint  in man.  The

 qualitative  CMS  effects observed subsequent to controlled

 inhalation exposure (Stewart et  al.,  1970) were not used as the

 basis for quantitation due to the subjective nature of the

 effects and  the  difficulty in extrapolating between inhaled and

 ingested doses.

      The renal toxicity observed after chronic Ingestion of

 tetrachloroethylene by rats and  mice  (NCI,  1977,* NTP, 1983) is of
 concern.   However,  the most sensitive endpoint of toxicity

 identified from  acute  and  subacute  ingestion of tetrachloro-

 ethylene by  laboratory animals appears to be hepatotoxicity in

 the mouse (Schumann et al.,  1980; Buben and O'Flaherty, 1985).

      Derivation  of  Health  Advisory Values

      Health Advisories (HAs)  are generally determined for

exposures  of One-day,  Ten-days,  Longer-term (approximately 7 year

exposure)  and Lifetime if  adequate data are available which

identify  a sensitive noncarcinogenic endpoint of toxicity.  The

HAs for noncarcinogenic toxicants are derived using the following


          fNOAEL OR LOAEL)  X (BW)
        =    (UP)  x (_ L/day)      =
         NOAEL or LOAEL = No-  or  Lowest-Observed-Adverse-Effect-
                          Level in rog/kg bw/day.

                     BW = assumed body weight of protected
                          individual  (10 kg for a child and  70 kg
                          for  an  adult).

                  UF(s)  = uncertainty factor  (10, 100 or up  to
                          1,000), based on nature and quality of
 One-day  Health Advisory

     The available  studies  were  not considered sufficient  for

 derivation of  a One-day HA.  The Ten-day HA of 2.0 mg/L  is

 recommended as a  conservative  estimate of 1-day exposure.

 Ten-day  Health Advisory

     Hepatotoxicity in  mice exposed to tetrachloroethylene was

 selected as the basis for calculating the Ten-day HA value.

     Schumann  et  al.  (1980) administered tetrachloroethylene in

 corn oil to rats  and mice via  gavage for 11 consecutive  days at

 doses of 0, 100,  500 and 1,000 mg/kg.  For mice,

 histopathological changes including increased liver weights were

 observed in all treated animals.  The lowest does, 100 mg/kg/day,

 represents the  LOAEL for the study.  This value is consistent

with the  estimated  LOAEL of 220  rag/kg/day for mice exposed to 200


ppm  for  4 hours  (Kylin,  1963,-  see Appendix).  Applying an

uncertainty  factor  of  1,000 may be overly conservative.

     Buben and O1Flaherty  (1985) treated mice with doses ranging

from 20  to 2,000  mg/kg»  5  days/week for 6 weeks and observed a

slight increase  in  liver weight in mice treated with 20 mg/kg;  at

100  mg/kg, increases were  significantly different from controls.

On this  basis, a  dose  of 20 mg/kg was identified as a NOAEL and

100  mg/kg was identified as a  LOA1L.

     Basing  the Ten-day  HA on  the NQAEL of 20 mg/kg with an

uncertainty  factor  of  100  is consistent with the protection of

humans from  the CNS effects observed by Stewart et al.  (1980) at

100  ppm  for  7 hours (approximately 16 mg/kg; see Appendix).  The

value was calculated as  follows:
          _„  -,_„ „,    (20 ma/ka/davl  (10 ken   „ n    /T
          Ten-day HA =     (ioo)  (1 L/day)     = 2•° mg/L
     20 mg/kg/day = NOAEL based on the absence of hepatotoxicity
                    in mice.

            10 kg = assumed body weight of a child.

           100 kg = uncertainty factor, chosen in accordance with
                    EPA or NAS/ODW guidelines for use with  a
                    NOAEL from an animal study.

          1 L/day = assumed daily water consumption of a child.

Longer-term Hea1thAdvisory

     The study by Buben  and 0'Flaherty was also selected as the

basis for the Longer-term HA.  Lifetime careinogenicity assays

were not selected because of the high doses used  (NCI, 1977; NTP,

1985).  The NOAEL of  20  mg/kg/day and LOAEL of 100 mg/kg/day

identified in this study are consistent with the estimates of

LOAELs from chronic inhalation studies.  A LOAEL of 63 mg/kg/day

was estimated from chronic exposure of guinea pigs to 100 ppm for

7 hours/day (Rowe et  al., 1952; see Appendix), and a LOAEL of 160

mg/kg/day from mice exposed to 200 ppm for 4 hours (Kylin, 1965).

The Longer-term HAs ,for  the child and adult were calculated as


     For a child:
                  HA = •*•*•*


     20 mg/kg/day — NOAEL based on the absence of hepatotoxic
                    effects in mice.

              5/7 = Factor to convert 5 day/week exposure to
                    daily exposure,

            10 kg = Assumed body weight of a child.

              100 = Uncertainty factor, chosen in accordance with
                    EPA or NAS/ODW guidelines for use with a
                    NOAEL from an animal study.

          1 L/day = Assumed daily water consumption of a child.

      For  an adult:
                   HA .
      20 rag/kg/day  =  NOAEL  based on the absence of hepatotoxic
                     effects  in mice.

              5/7  =  Factor to convert 5 day/week exposure to
                     daily  exposure,

             70 kg  =  Assumed  body weight of an adult.

              100  =  Uncertainty factor, chosen in accordance with
                     EPA or NAS/ODW guidelines for use with a
                     NOAEL  from an animal study.

           2  L/day  =  Assumed  daily water consumption of an adult.
Derivation ofReference Doseand the Drinking Water Equivalent


     No suitable chronic oral or lifetime oral studies were

located in the literature to serve as the basis for the Lifetime

HA.  NOAELs were not identified in the NCI  (1977) study in which

LOAELs were identified at high doses (386 mg/kg/day, mice; 471

mg/kg/day, rats).  The NTP  (1983) study in which lower doses were

tested has not been validated.

     Approximate NOAELs and LOAELs calculated from chronic and

lifetime inhalation studies give less conservative estimates of

toxic doses than the six-week oral study of Buben and O1Flaherty

 (1985).   LOAEL estimates  of  63 mg/kg/day for guinea pigs exposed

 to  100 ppm,  7  hours,day (Rows et al.,  1952), 200 mg/kg/day  for

 rats  exposed to 230  ppm for  7 hour/day (Carpenter, 1937) and

 160 mg/kg/day  for  mice  exposed to 100  ppm for 6 hour/day  (NTP,

 1985) are consistent with the NOAEL  of 20 mg/kg/day and LOAEL of

 100 mg/kg/day  identified  in  the study  by Buben and O1Flaherty.

 In  this  study,  mice  were  treated with  doses of 20 to 2,000

 mg/kg/day, 5 days/week  for 6 weeks.  A slight increase in liver

 weight was observed  at  20 mg/kg; at  100 mg/kg, liver weight and

 hepatic  triglyceride levels  were significantly increased over

 controls.  Using the NOAEL of 20 mg/kg/day and an uncertainty

 factor of 1,000 consistent with the  use of data from less than

 lifetime studies,  the RfD and DWEL were calculated as follows:
             RfD .                      - 0.0143 mg/kg/day


     20 mg/kg/day = NOAEL.

              5/7 = Factor to convert 5 day/week exposure to
                    daily exposure.

            1,000 = Uncertainty factor, chosen in accordance with
                    EPA or NAS/ODW guidelines for use with a
                    NOAEL from an animal study of less-than-
                    lifetime duration.
     The DWEL for tetrachloroethylene based on noncarcinogenic

effects and assuming 100% exposure from drinking water is

calculated as follows:

     DWEL = (0.0143 mq/kq/day)  (70 kq)

      The  estimated  excess upper bound cancer risk associated with

 lifetime  exposure to  drinking water containing tetrachloroethy-

 lene  at 0.5 mg/L is approximately 1 x 10~3.

     Tetrachloroethylene was tested for carcinogenic potential in

 B6C3F1 mice and Fischer 344 rats in the NCI Bioassay Program

 (NCI, 1977),  In those bioassays, the test compound, containing a

 small amount of stabilizer, was administered in oil by gavage 5

 days/week for 78 weeks.  Under the experimental conditions

 employed in the studies, it was shown that tetrachloroethylene

 caused a significant  increase in the incidence of hepatocellular

 carcinomas in both  sexes of mice at both dose levels when

 compared with the untreated and vehicle control groups.  In the

 rats, there appeared  to be no significant increased incidence of

 neoplastic lesions  at any site.  The implications of these

 results must be tempered by the fact that, among the rats, there

were high incidences  of respiratory disease in all groups, high

 incidences of toxic nephropathy in the tetrachloroethylene groups

and a higher mortality rate among the treated groups than the

control groups.  For  a variety of reasons, it was decided that

the bioassay would be repeated.

      On the basis  of the data  reported  in the NCI bioassay

 published in 1977,  IARC (1979) concluded that there is limited

 evidence to state  that it is a carcinogen in the mouse.

 Chemicals which  fall into this category or classification by this

 Agency  are usually there for two reasons.  Firstly, the

 experimental data  may be restricted such that it is not possible

 to  determine a causal relationship between exposure and

 development of a lesion.   Secondly, certain neoplasms, such as

 lung  adenomas and  hepatomas in mice, are considered by some

 investigators to be of lesser  significance than tumors of other

 types occurring  at other sites.  In addition, some chemicals for

 which there is limited evidence of carcinogenicity in animals

 also  have been studied in humans, with, in general, inconclusive

 results.   While  there is some  evidence  for increased risk of

 urinary  tract cancer in dry cleaner works, there is insufficient

 evidence to demonstrate or refute a carcinogenic hazard for

 tetrachloroethylene alone.  EPA concludes that the human evidence

 for tetrachloroethylene is inadequate to develop a more

 definitive  conclusion.

     An  additional  inhalation  bioassay  was conducted by the NTP

 in which  rats were  exposed to  200 and 400 ppm (1,360 and

 2,710 Mg/m3) and mice to  100 and 200 ppm (678 and 1,360 pg/m3)

tetrachloroethylene (NTP,  1985).  Statistically significant

increases in mononuclear  cell  leukemia  were observed to have an

increased incidence of  hepatocellular carcinoma.  In addition, a

statistically significant  increase in the incidence of renal


 adenomas/carcinomas  (combined) was observed for male mice in the

 high  dose  group.   Based on  this and previous studies, it can be

 concluded  that  there is sufficient evidence of carcinogenicity in

 animals on exposure  to  tetrachloroethylene.

      Controversy  exists over the classification of tetrachloro-

 ethylene because  different  interpretations can be given to either

 the bioassay  data on tetrachloroethylene or to the cancer

 guidelines (51  FR 33992).   EPA recommended that "sufficient"

 evidence of carcinogenicity existed based on positive findings of

 carcinogenicity in two  species with multiple tumor sites, and via

 two routes of administration.  Using the same data, the

 Halogenated Organic  Solvent Subcommittee of EPA's Science
 Advisory Board  concluded that the evidence was "inadequate," and

 suggested  a classification  of Group C:  possible human carcinogen

 (U.S. EPA, 1987).

     A major  difference between the analysis of the data by the

 subcommittee  and  that of the Agency (U.S. EPA, 1986) was the

 interpretation  of the data  on the tumor incidence in rats in the

 1985 NTP inhalation  bioassay.  Concerning the finding of

 increased  renal tumors  in rats, the subcommittee questioned the

diagnosis  of  neoplasia,  and objected to the statistical analysis

in which a significant  increase was observed only when adenomas

and carcinomas were  combined for statistical analysis.  The

subcommittee  also  questioned the finding of raononuclear cell

leukemia in rats.   EPA  has  included preleukemic stages for


 statistical analysis of the results.  The committee raised

 questions concerning the method  of  staging and also questioned

 the diagnosis of the tumor.   The subcommittee agreed that

 tetrachloroethylene caused an increase  in mouse liver tumors, but

 they questioned the relevance of this tumor type to man.

      EPA has carefully considered these questions; many are

 similar  to questions arising for other  compounds.  For example,

 the question of mouse liver tumors  is discussed in the cancer

 guidelines (51 FR 33992).   Although uncertainty exists,

 sufficient understanding of the  pathology of renal neoplasia and

 mononuclear cell leukemia exists to make reasonable judgments on

 these issues,  have confidence in the diagnosis of these tumor

 types, and make reasonable decisions on methods of statistical

 analysis.   Combining adenomas/carcinomas is a valid method for

 analyzing renal tubular cell neoplasia  and is consistent with the

 cancer guidelines  and the work of McConnell et al. (1986).  The

 guidelines do  not  specifically mention  staging leukemia, but

 preleukemic stages do not  need to be included in the analysis to

 obtain a  significant tumor increase in  rats.  Therefore, it can

 be concluded that  this bioassay  gives positive evidence of

 carcinogenicity in a second  species (U.S. EPA, 1986).

     The  role  of tetraehloroethylene metabolites in the

manifestation  of toxicity  including carcinogenicity cannot be

 ignored.   The  available  information indicates that there is no

 reason  to  believe  that  qualitative differences in the metabolism

 of  tetrachloroethylene  among various animal species exists.

     Tetrachloroethylene  is metabolized by two metabolic

 pathways:   oxidative pathway dependent upon cytochrome P 450 and

 the conjugative pathway involving glutathione.  The major

 metabolite of  oxidative pathway is trichloroacetic acid which is

 excreted in urine.  Some  of the intermediates in the

 trichloroacetic acid pathway possess cytotoxic and genotoxic


     The conjugative pathway, a multistep glutathione dependent

 pathway —  the so-called  cysteine conjugate fl-lyase pathway is

 toxicologically important even though it is minor route of

 disposition of tetrachloroethylene.  In this pathway, haloalkene,

 i.e., tetrachloroethylene, forms hepatic glutathione S-conjugate

 and the resulting conjugate(s) {glutathione, cysteine or N-

 acety(cystein S-conjugate) is transferred to the kidney where it

 is bioactivated by 6-lyase.  There is evidence that this pathway

 is responsible for the  nephrotoxicity, mutagenicity and possible

 nephrocarcinogenicity of  chloroalkenes including

tetrachloroethylene (Monks et al., 1990).

     Quantification of  Carcinogenic Effects

     Using methodology  described in detail elsewhere (51 FR

33992),  the EPA's Carcinogen Assessment Group (CAG) has


 calculated  estimated  incremental excess upper bound cancer risk

 associated  with exposure  to  tetrachloroethylene in ambient water,

 extrapolating  from  data obtained in the 1977 NCI bioassay in mice

 with this compound  (NCI,  1977).  GAG employed a linearized,

 non-threshold  multistage  model to estimate the upper bound of the

 excess cancer  rate  that would occur at a specific exposure level

 for a 70 kg adult ingesting  two liters of water and 6.5 grams of

 fish and seafood (fish factor) every day over a 70-year lifespan.

     The National Academy of Sciences  (NAS, 1977, 1980) and EPA's

 Carcinogen  Assessment Group  (Anderson, 1983) have calculated

 estimated upper bound incremental excess cancer risks associated

 with the consumption  of tetrachloroethylene via drinking water
 alone by mathematical extrapolation from the high-dose animal

 studies.  Each group  employed the linearized, non-threshold

 multistage  model, extrapolating from data obtained in the 1977

 NCI bioassay in mice.

     In all three instances, a range of tetrachloroethylene

 concentrations was  computed  that would be estimated to increase

 the risk by one excess cancer per million  (10~6) ,  per hundred

 thousand (10~5) and per ten thousand (10~4)  population  over  a  70-

year lifetime  assuming daily consumption of 2 liters of water by

a 70-kg adult  at the  stated  exposure level.  The ranges of

concentrations  are  summarized in Table 1.

     The NCI bioassay  also was the basis for the upper bound  unit

risk derivation,  i.e.,  the risk associated with 1 ng/L drinking

water or 1 /jg/m3  air (U.S. EPA, 1985).  The upper bound risk

associated with exposure  to  1 jLtg/L water was estimated to  be

1.5 x 10~6; concentrations corresponding to risks of 10"4,  10~5

and 10~6 were derived by extrapolation {Table 1).


     The World Health  Organization has recommended a tentative

guideline value of 10  Mg/L for tetrachloroethylene in drinking

water, based on carcinogenic properties (WHO, 1984) .

     The National Academy of Sciences  (NAS, 1980) calculated

24-hour and 7-day SNARLs.  The 24-hour SNARL was 172 mg/L,  based

on a 490 nig/kg LOAEL following i.p. administration, a 100-fold

uncertainty factor, and a 70-kg adult drinking 2 L/day of

drinking water.  A 7-day SNARL of 24.5 mg/L was calculated by

dividing the 24-hour SNARL by seven.

                             Table  1

      Estimated  tetrachloroethylene  concentrations causing excess
                  Cancer  risks of  1CT4, 1CT5  and 10"6
Excess Tetrachloroethylene concentrations (/Ltg/L)
Lifetime Basis for concentration estimates
Riska CAGb
10~4 90.0
10"5 9,0
10~6 0.9
1.5 x 10~6
65.8 350 (66.7)
6.6 35 (6.7)
0.7 3.5 (0.7)
       a  Assumes 2 L of water consumed/day by 70-kg adult over  a
          lifetime; number represents upper bound.
       b  U.S. EPA, 1980.  Includes "fish factor," assumed daily
          consumption of 6.5 grams of contaminated fish and
       c  Anderson, 1983.
       d  NAS, 1977, 1980.
       e  U.S. EPA, 1985.  Based on linear extrapolation from
          risk estimate based on concentrations of 1

     The recommended HA values are listed below:

               One-day             2.0 mg/L
               Ten-day             2.0 mg/L
               Longer-term  (child) 1.4 mg/L
               Longer-term  (adult) 5.0 mg/L

     A DWEL of 500 /Lig/L was calculated from which a lifetime HA

value could be derived.  The estimated excess upper bound cancer

risk associated with lifetime exposure to drinking water

containing tetrachloroethylene at 500 ^g/L is approximately

1 x 10~3.   This estimate is derived from extrapolations using the

linearized, multistage model.


 Anderson,  E.L.   1983.   Draft memo  to Frederic A Eidsness, Jr.,
      entitled "Latest  Cancer Risk  Rate Estimates."  March 22.

 Brancaccio,  A,,  V.  Mazza and R.  DiPaolo.   1971  Renal  function  in
      experimental  tetrachloroethylene poisoning.  Folia Med.
      (Naples),   54:233-237.

HBuben,  J.A.  and E.  O1Flaherty.   1985.  Delineation of  the role  of
      metabolism in the hepatotoxicity of trichloroethylene  and
      perchloroethylene;   A dose-effect study.  Tox. Appl. Pharm.

 Carpenter, C.P.  1937.   The  chronic toxicity of tetrachloro-
      ethylene.   J.  Ind.  Hyg.  Toxicol.  19:323-326.

 Federal Register.   1986.   Guidleines for carcinogen risk assess-
      ment.   51(85):33392-34003.  September 24.

 Fujii,  T.  1975.  The variation  in  the liver function of rabbits
      after administration of chlorinated hydrocarbons.  Jap. J.
      Ind. Health.   17:81-88.

 Goldberg, M.E.,  H.E. Johnson, U.C. Pozzani and H.F. Smyth,  Jr.
      1964.   Effect of  repeated  inhalation  of vapors of industrial,
      solvents on animal  behavior.  I.  Evaluation of nine solvent
      vapors  on  pole-climb performance in rats.  Am. Ind. Hyg.
      ASSOC.  J.   25:369-375.

 Hake, C.L. and  R.D. Stewart.  1977.  Human exposure to tetra-
      chloroethylene:   Inhalation and skin  contact.  Environ.
      Health  Perspect.   21:377-401,

 IARC.   1979.   International  Agency for Research on Cancer.  IARC
      monographs  on the  evaluation  of the carcinogenic  risk  of
      chemicals  to  man.   Some monomer, plastic and synthetic
      elastomes  and acrolein.  19:377-401.

 Kylin,  B., H. Reichard,  I. Sumegi  and S. Yllner.  1963.
      Hepatotoxicity of  inhaled  trichloroethylene, tetrachloro-
      ethylene and  chloroform.   Single exposure.  Acta  Pharmacol.
      Toxicol.   20:16-26.

 Kylin,  B., I. Sumegi and  S.  Yllner.  1965.  Hepatotoxicity  of
      inhaled trichloroethylene  and tetrachloroethylene.  Long-
      term exposure.  Acta Pharmacol. Toxicol.  22:379-385.

Mazza, V.  1972.   Enzymatic  changes in experimental tetrachloro-
      ethylene poisoning.   Folia  Med.  55(9-10):373-381.

 Monks,  T.J.,  M.W.  Anders, W.  DeKant, J.L. Stevens, S.S. Lau and
      P.J. Van Bladeren.   1990.  Contemporary issues in
      toxicology:   Glutathione conjugate mediated toxicities.
      Toxicol.  Appl.  Pharmacol.  106:1-19,

 NAS.   1977.   National Academy of Sciences.  Drinking water and
      Health.   Volume 1.   National Academy Press.  Washington, DC.

 NAS.   1980.   National Academy of Sciences.  Drinking Water and
      Health.   Volume 3.   National Academy Press.  Washington, DC.

 NCI.   1977.   National Cancer  Institute.  Bioassay of tetrachloro-
      ethylene for  possible carcinogenicity.  DHEW Publication No.
      NIH 77-813, U.S. Department of HEW, PHS, National Institute
      of Health, National  Cancer Institute PB-272 950, NTIS.

 NTP.   1983.   Bioassay on  tetrachloroethylene in female B6C3F1
      mice.  Draft.

 NTP.   1985.   NTP technical report on the toxicology and carcino-
      genesis  studies on tetrachloroethylene  (perchloroethylene).
      NTP, Research Triangle Park, NC.

 Rowe, V.K., D.D. McCollister, H.C. Spencer, E.M. Adams and D.D.
      Irish.   1952.
 Savolainen, H., P. Pfaffli, M. Tengen and H. Vainio.  1977.
      Biochemical and behavioral effects of inhalation exposure to
      tetrachloroethylene  and  dichloromethane.  J. Neuropathol.
      Exp. Neurol.  36(6):941-949.

 Schumann, A.M., J.F. quast and P.G. Watanabe.  1980.  The
      pharmacokinetics and macromolecular interactions of per-
      chloroethylene  in mice and rats as related to oncogenicity.
      Toxicol. Appl.  Pharmacol.  55:207-219.

 Schwetz, B.A., B.K.J. Leong and P.J. Gehring.  1975.  The effect
      of maternally inhaled trichloroethylene, perchloroethylene,
     methyl chloroform, and methylene chloride on embryonal and
      fetal development in mice and rats.  Toxicol. Appl.
     Pharmacol.  32:84-96.

Smyth, H.F., Jr., C.S. Weil,  J.S. West and C.P. Carpenter.  1969.
     An exploration  of joint  toxic action:  Twenty-seven
      industrial chemicals intubated in rats in all possible
     pairs.  Toxicol. Appl. Pharmacol.  14:340-347.

Stewart, R.D., H.H.  Gay, D.S. 'Erley, C.L. Hake and A.W. Schaffer.
     1961.   Human exposure to tetrachloroethylene vapor.  Arch.
     Environ. Health.  20:516-522.

 Stewart,  R.D.,  E.D.  Barretta,  B.C.  Dowd  and T.R. Torkelson.
      1970.   Experimental human exposure  to tetrachloroethylene.
      Arch.  Environ.  Health.   20:224-229.

 Stewart,  R.D.,  C.L.  Hake,  H.V.  Forster,  A.J. Lebrum, J.E.
      Peterson and A.  Wu.   1974.   Tetrachloroethylene:
      Development of  a biological  standard for the  industrial
      worker by breath analysis.   Report  No. NIOSH-MCOW-ENVM-PCE-
      74-6.   National Institute for  Occupational Safety  and

 Stewart,  R.D.,  C.L.  Hake,  A.  Wu,  J. Kalbfleisch, P.E. Newton,
      S.K. Marloro and M.V.  Salama.  1977.  Effects of perchloro-
      ethylene/drug interaction on behavior and neurological
      function.   Final report.   National  Institute  for
      Occupational Safety and  Health.  April,

 U.S.  EPA.   U.S.  Environmental Protection Agency.   1980.  Water
      quality criteria documents:  Notice of availability.  Office
      of Water.   Federal  Register  45:79318-79379.

 U.S.  EPA.   U.S.  Environmental Protection Agency.   1985.  Health
      assessment  document for  tetrachloroethylene
      (perchloroethylene).   Office of Health and Environmental
 U.S.  EPA.   U.S.  Environmental Protection Agency.   1986.  Addendum
      to the Health Assessment Document for Tetrachloroethylene
      (Perchloroethylene).   Office of Health and Environmental
      Assessment.   External  Review Draft.  April.

 U.S.  EPA.   U.S.  Environmental  Protection Agency.   1987.  Science
      Advisory Board's Environmental Health Committee, Halogenated
      Organics Subcommittee  Report.  Memo from N. Nelson and R.A.
      Griesemer to Lee M. Thomas,  January 27, 1987.

 U.S.  EPA.   U.S.  Environmental  Protection Agency.   1990.  Draft
      document.   Response to issues  and data submissions on
      tetrachloroethylene  (perchloroethylene).

Vainio, H.M., G.  Parkki  and J. Marniemi.  1976.  Effects of
      aliphatic chlorohydrocarbons on drug metabolizing  enzymes in
      rat liver in vitro.  Xenobiotica. 6:559-604.

Wenzel, D.G., and R.D. Gibson.  1951.  A study of the toxicity
      and anthelminthic activity of  n-butylidene chloride.  J.
      Pharm. Pharmacol. 3:169-176.

WHO.  1984.  World Health Organization.  Guidelines for Drinking
     Water Quality.   Volume I.  Geneva.  ISBN 9241541687.

              Estimation of absorbed dose based on inhalation exposure
Approx ,
Species (kg)
Human 70,0
Guinea 0,50

Rat 0.25

Mouse 0.025

Approx. Hours Approx.
minute Exposed dose
volume [ PCB ]
(1/min) (ppm)
10.0 100
0.222 100

0.132 200
0.024 100


Stewart et
Rowe et al .
Rowe et al.
Rowe et al .
Rowe et al.
NTP, 1985
NTP, 1985
Carpenter ,
NTP, 1985
NTP, 1985
NTP, 1985
NTP, 1985
NTP, 1985

al., 1977
, 1952
, 1952
, 1952
, 1952
et al., 1977; NTP, 1985
et al., 1977j NTP, 1985



0.742 2,280


160   Kylin, 1963, 1965

680   Brancaccio et al., 1971

840   Mazza, 1972
»„       rtetrachloroethvlenefrnq/m311 f Iung_vol(m5/hri 1 fTlme(hr/davi l.f 50%  absorption!
 Dose =	— 		—	^	—	
                                       body weight (kg)
       [tetrachloroethylene(mg/nr} ] =  (ppm) x  (6.78 mg/m  - ppm)
       [lung vol.